Appendix D.3 Chronic RELs and toxicity summaries using the previous version of the Hot 

Spots Risk Assessment guidelines (OEHHA 1999) 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Acrylonitrile.................................................................................................................................. 11 

Ammonia....................................................................................................................................... 19 

Benzene......................................................................................................................................... 24 

Beryllium and Beryllium Compounds.......................................................................................... 35 

1,3-Butadiene................................................................................................................................ 48 

Cadmium and Cadmium Compounds........................................................................................... 59 

Carbon Disulfide........................................................................................................................... 66 

Carbon Tetrachloride .................................................................................................................... 81 

Chlorinated Dibenzo-p-Dioxins and Chlorinated Dibenzofurans................................................. 90 

Chlorine....................................................................................................................................... 106 

Chlorine Dioxide......................................................................................................................... 114 

Chloroform.................................................................................................................................. 124 

Chromium, Hexavalent (Soluble Compounds)........................................................................... 138 

Cresol Mixtures........................................................................................................................... 157 

1,4-Dichlorobenzene................................................................................................................... 165 

1,1-Dichloroethylene .................................................................................................................. 170 

Diethanolamine ........................................................................................................................... 177 

N,N-Dimethylformamide............................................................................................................ 184 

1,4-Dioxane................................................................................................................................. 189 

Epichlorohydrin .......................................................................................................................... 196 

1,2-Epoxybutane ......................................................................................................................... 200 

Ethyl Chloride............................................................................................................................. 204 

Ethylbenzene............................................................................................................................... 208 

Ethylene Dibromide .................................................................................................................... 214 

Ethylene Dichloride .................................................................................................................... 226 

Ethylene Glycol .......................................................................................................................... 234 

Ethylene Glycol Monoethyl Ether.............................................................................................. 240 

Ethylene Glycol Monoethyl Ether Acetate................................................................................. 248 

Ethylene Glycol Monomethyl Ether........................................................................................... 253 

Ethylene Glycol Monomethyl Ether Acetate.............................................................................. 260 

Ethylene Oxide............................................................................................................................ 264 

Fluorides including Hydrogen Fluoride..................................................................................... 270 

Glutaraldehyde............................................................................................................................ 286 

n-Hexane..................................................................................................................................... 292 

Hydrazine.................................................................................................................................... 303 

Hydrogen Chloride...................................................................................................................... 309 

Hydrogen cyanide ....................................................................................................................... 313 

Hydrogen Sulfide........................................................................................................................ 321 

Isophorone................................................................................................................................... 329 

Isopropanol ................................................................................................................................. 337 

Maleic Anhydride ....................................................................................................................... 348 

Methanol ..................................................................................................................................... 354 

METHYL BROMIDE ................................................................................................................ 360 

Methyl Chloroform..................................................................................................................... 366 

Methyl Isocyanate....................................................................................................................... 372 

Methyl t-Butyl Ether................................................................................................................... 383 

Appendix D3 2 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Methylene Chloride .................................................................................................................... 389 

4,4′-Methylene Dianiline ............................................................................................................ 399 

Methylene Diphenyl Isocyanate ................................................................................................. 407 

Naphthalene ................................................................................................................................ 413 

Nickel and Nickel Compounds: Nickel Oxide............................................................................ 420 

Phenol ......................................................................................................................................... 429 

Phosphine.................................................................................................................................... 437 

Phosphoric Acid.......................................................................................................................... 446 

Phthalic Anhydride ..................................................................................................................... 453 

Propylene .................................................................................................................................... 460 

Propylene Glycol Monomethyl Ether......................................................................................... 466 

Propylene oxide .......................................................................................................................... 471 

Selenium and Selenium Compounds.......................................................................................... 476 

SILICA (CRYSTALLINE, RESPIRABLE)............................................................................... 486 

Styrene ........................................................................................................................................ 534 

Sulfuric Acid............................................................................................................................... 551 

Toluene ....................................................................................................................................... 560 

2,4- and 2,6-Toluene Diisocyanate ............................................................................................. 574 

Trichloroethylene........................................................................................................................ 581 

Triethylamine.............................................................................................................................. 590 

Vinyl Acetate .............................................................................................................................. 596 

Xylenes ....................................................................................................................................... 602 

Appendix D3 3 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Acrylonitrile 

(107-13-1) 

; 5 Respiratory 

system 

 

Ammonia 

(7664-41-7) 

  200 Respiratory 

system 



Benzene 

(71-43-2) 

  60 Hematopoietic 

system; 

development; 

nervous system 

 

Beryllium 

(7440-41-7) 

and beryllium 

compounds 

; 0.007 0.002 Respiratory 

system; 

immune system 

; 

Butadiene 

(106-99-0) 

20 Reproductive 

system 

 

Cadmium 

(7440-43-9) & 

cadmium 

compounds 

; 0.02 0.0005 Kidney; 

respiratory 

system 

; 

Carbon tetrachloride 

(56-23-5) 

; 40 Alimentary 

system; 

development; 

nervous system 

 

Carbon disulfide 

(75-15-0) 

800 Nervous 

system; 

reproductive 

system 

; 

Chlorinated dioxins 

(1746-01-6) & 

dibenzofurans 

(5120-73-19) 

  0.00004 1 x 10-8 Alimentary 

system (liver); 

reproductive 

system; 

development; 

endocrine 

system; 

respiratory 

system; 

hematopoietic 

system 

 

Chlorine 

(7782-50-5) 

 0.2 Respiratory 

system 

 

Appendix D3 4 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Chlorine dioxide 

(10049-04-4) 

 0.6 Respiratory 

system 

 

Chlorobenzene 

(108-90-7) 

; 1000 Alimentary 

system; kidney; 

reproductive 

system 

 

Chloroform 

(67-66-3) 

 ; 300 Alimentary 

system; kidney; 

development 

 

Chloropicrin 

(76-06-2) 

; 0.4 Respiratory 

system 

 

Chromium 

hexavalent: soluble 

except chromic 

trioxide 

; 0.2 0.02 Respiratory 

system 

 

Chromic trioxide 

(as chromic acid 

mist) 

; 0.002 Respiratory 

system 

; 

Cresol mixtures 

(1319-77-3) 

; 600 Nervous system 

Dichlorobenzene 

(1,4-) 

(106-46-7) 

; 800 Nervous 

system; 

respiratory 

system; 

alimentary 

system; kidney 

 

Dichloroethylene 

(1,1) (75-35-4)  

; 70 Alimentary 

system 

 

Diesel Exhaust*  5 Respiratory 

system 

 

Diethanolamine 

(111-42-2) 

 3 Cardiovascular 

system; nervous 

system 

 

Dimethylformamide 

(N,N-) 

(68-12-2) 

 80 Alimentary 

system ; 

respiratory 

system 

; 

Appendix D3 5 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Dioxane (1,4-) 

(123-91-1) 

  3,000 Alimentary 

system; kidney; 

cardiovascular 

system 

 

Epichlorohydrin 

(106-89-8) 

; 3 Respiratory 

system; eyes 

 

Epoxybutane (1,2-) 

(106-88-7) 

 20 Respiratory 

system; 

cardiovascular 

system 

 

Ethylbenzene 

(100-41-4) 

 2,000 Development; 

alimentary 

system (liver); 

kidney; 

endocrine 

system 

 

Ethyl chloride 

(75-00-3) 

  30,000 Development; 

alimentary 

system 

 

Ethylene dibromide 

(106-93-4) 

; 0.8 Reproductive 

system 

; 

Ethylene dichloride 

(107-06-2) 

; 400 Alimentary 

system (liver) 

 

Ethylene glycol 

(107-21-1) 

400 Respiratory 

system; kidney; 

development 

 ; 

Ethylene glycol 

monoethyl ether 

(110-80-5)  

  70 Reproductive 

system; 

hematopoietic 

system 

 

Ethylene glycol 

monoethyl ether 

acetate 

(111-15-9) 

  300 Development 

Ethylene glycol 

monomethyl ether 

(109-86-4) 

  60 Reproductive 

system 

 

Ethylene glycol 

monomethyl ether 

acetate (110-49-6)  

  90 Reproductive 

system 

 

Appendix D3 6 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Ethylene oxide 

(75-21-8) 

; 30 Nervous system 

Fluoride including 

Hydrogen Fluoride 

13 F 

14 HF 

0.04 Bone and teeth; 

respiratory 

system 

; 

Glutaraldehyde 

(111-30-8) 

; 0.08 Respiratory 

system 

 

Hexane (n-) 

(110-54-3) 

7000 Nervous system 

Hydrazine 

(302-01-2) 

; 0.2 Alimentary 

system; 

endocrine 

system 

 

Hydrogen chloride 

(7647-01-0) 

  9 Respiratory 

system 

 

Hydrogen cyanide 

(74-90-8) 

  9 Nervous 

system; 

endocrine 

system; 

cardiovascular 

system 

; 

Hydrogen sulfide 

(7783-06-4) 

  10 Respiratory 

system 

 

Isophorone 

(78-59-1) 

 2000 Development; 

liver 

 

Isopropanol 

(67-63-0) 

 7,000 Kidney; 

development 

 

Maleic anhydride 

(108-31-6) 

; 0.7 Respiratory 

system 

 

Methanol 

(67-56-1) 

  4,000 Development 

Methyl bromide 

(74-83-9) 

  5 Respiratory 

system; nervous 

system; 

development 

 

Methyl chloroform 

(71-55-6) 

  1,000 Nervous system 

Appendix D3 7 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Methyl isocyanate 

(624-83-9) 

1 Respiratory 

system; 

reproductive 

system 

 

Methyl t-butyl ether 

(1634-04-4) 

 8,000 Kidney; eyes; 

alimentary 

system (liver) 

 

Methylene chloride 

(75-09-2) 

  400 Cardiovascular 

system; nervous 

system 

 

Methylene dianiline 

(4,4'-) (101-77-9) 

; 20 Eyes; 

alimentary 

system 

(hepatotoxicity) 

 

Methylene Diphenyl 

Isocyanate 

(101-68-8) 

 0.7 Respiratory 

system 

 

Naphthalene 

(91-20-3) 

  9 Respiratory 

system 

 

Nickel & 

compounds 

(except nickel oxide) 

  0.05 0.05 Respiratory 

system; 

hematopoietic 

system 

 

Nickel oxide 

(1313-99-1) 

 0.1 ? Respiratory 

system; 

hematopoietic 

system 

 

Phenol 

(108-95-2) 

  200 Alimentary 

system; 

cardiovascular 

system; kidney; 

nervous system 

 

Phosphine 

(7803-51-2) 

; 0.8 Respiratory 

system; 

alimentary 

system; nervous 

system; kidney; 

hematopoietic 

system 

Appendix D3 8 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Phosphoric acid 

(7664-38-2) 

7 Respiratory 

system 

 

Phthalic anhydride 

(85-44-9) 

; 20 Respiratory 

system 

; 

Propylene 

(115-07-1) 

 3,000 Respiratory 

system 

 

Propylene glycol 

monomethyl ether 

(107-98-2) 

 7,000 Alimentary 

system (liver) 

 

Propylene oxide 

(75-56-9) 

 ; 30 Respiratory 

system 

 

Selenium and 

selenium compounds 

(other than hydrogen 

selenide) 

; 20 0.005 Alimentary 

system; 

cardiovascular 

system; nervous 

system 

; 

Silica (crystalline, 

respirable)  

3 Respiratory 

system 

; 

Styrene 

(100-42-5) 

 ; 900 Nervous system ; 

Sulfuric acid 

(7664-93-9) 

 1 Respiratory 

system 

 

Tetrachloroethylene* 

(perchloroethylene) 

(127-18-4) 

; 35 Kidney; 

alimentary 

system (liver) 

 

Toluene 

(108-88-3) 

; 300 Nervous 

system; 

respiratory 

system; 

development 

 

Toluene 

diisocyanates (2,4- 

& 2,6-) 

; 0.07 Respiratory 

system 

; 

Trichloroethylene 

(79-01-6) 

 ; 600 Nervous 

system; eyes 

; 

Triethylamine 

(121-44-8) 

200 Eyes 

Vinyl acetate 

(108-05-4) 

 200 Respiratory 

system 

 

Appendix D3 9 Table of values 
Noncancer chronic Reference Exposure Levels determined using previous methodology 

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment 

Guidelines (OEHHA 1999)

Substance 

(CAS #)

Listed in 

CAPCOA 

(1993) 

Chronic 

Inhalation 

REL 

(μg/m3

) 

Oral REL 

(mg/kg 

Body 

Weight) 

Hazard Index 

Target Organs 

Human 

Data 

Xylenes 

(m-, o-, p-)  

; 700 Nervous 

system; 

respiratory 

system 

; 

*These peer-reviewed values were developed under the Toxic Air Contaminant (TAC) Program 

mandated by AB1807. 

Appendix D3 10 Table of values 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

ACRYLONITRILE 

(Acrylonitrile monomer, cyanoethylene, propenenitrile, 2-propenenitrile, VCN, vinyl cyanide.) 

CAS Number: 107-13-1 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 5 μg/m3 (2 ppb) 

Critical effect(s) Degeneration and inflammation of nasal 

epithelium in rats 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1994) 

Description Clear, colorless to pale yellow liquid 

(technical grades) 

Molecular formula C3H3N 

Molecular weight 53.1 g/mol 

Density 0.81 g/cm3

 @ 25oC 

Boiling point 77.3o

C 

Melting point −82o

C 

Vapor pressure 100 torr @ 23o

C 

Solubility Soluble in isopropanol, ethanol, ether, 

acetone, and benzene 

Conversion factor 1 ppm = 2.17 mg/m3 @ 25 o

C 

III. Major Uses or Sources 

Acrylonitrile is produced commercially by propylene ammoxidation, in which propylene, 

ammonia, and air are reacted by catalyst in a fluidized bed. Acrylonitrile is used primarily as a 

co-monomer in the production of acrylic and modacrylic fibers. Uses include the production of 

plastics, surface coatings, nitrile elastomers, barrier resins, and adhesives. It is also a chemical 

intermediate in the synthesis of various antioxidants, pharmaceuticals, dyes, and surface-active 

agents. Formerly, acrylonitrile was used as a fumigant for food commodities, flour milling, and 

bakery food processing equipment (HSDB, 1994). The annual statewide industrial emissions 

from facilities reporting under the Air Toxics Hot Spots Act in California based on the most 

recent inventory were estimated to be 3948 pounds of acrylonitrile (CARB, 2000). US EPA 

(1993) reported a mean ambient air concentration of acrylonitrile at four urban locations in the 

U.S.of 0.66 μg/m3

. 

Appendix D3 11 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IV. Effects of Human Exposure 

Many occupational epidemiology studies have investigated retrospectively the morbidity and 

mortality of acrylonitrile exposed workers. An increased incidence of lung cancer was 

associated with acrylonitrile exposure. No significant excess mortality has been observed for 

any noncarcinogenic endpoint. One early cross-sectional study (Wilson et al., 1948) observed 

multiple deleterious effects in synthetic rubber manufacturing workers acutely exposed (20 to 45 

minutes) to various concentrations of acrylonitrile (16 to 100 ppm, 34.7 to 217 mg/m3

). Mucous 

membrane irritation, headaches, feelings of apprehension, and nervous irritability were observed 

in the majority of workers. Other less common symptoms observed included low-grade anemia, 

leukocytosis, and mild jaundice. These effects were reported to subside with cessation of 

exposure. Human volunteers exposed for a single 8 hour period to acrylonitrile vapors exhibited 

no deleterious CNS effects at concentrations ranging from 5.4 to 10.9 mg/m3 (2.4 to 5.0 ppm) 

(Jakubowski et al., 1987). 

A cross-sectional study (Sakurai et al., 1978) found no statistically significant increases in 

adverse health effects in chronically exposed workers (minimum 5 years) employed at 6 acrylic 

fiber factories (n = 102 exposed, n = 62 matched controls). Mean acrylonitrile levels ranged 

from 0.1 to 4.2 ppm (0.2 to 9.1 mg/m3) as determined by personal sampling. Although not 

statistically significant, slight increases in reddening of the conjunctiva and pharynx were seen in 

workers from the plant with the highest mean levels (4.2 ppm arithmetic mean). However, this 

study has limitations, including small sample size and examiner bias, since the medical examiner 

was not blind to exposure status. The time-weighted average exposure of the group 

occupationally exposed to 4.2 ppm (9.1 mg/m3) acrylonitrile can be calculated as: TWA = 9.1 

mg/m3 x (10/20) m3/day x 5 days/7 days = 3 mg/m3. This level is comparable to the LOAEL 

(HEC) of 2 mg/m3

 derived by the U.S. EPA from the animal study of Quast et al. (1980). 

Czeizel et al. (1999) studied congenital abnormalities in 46,326 infants born between 1980 and 

1996 to mothers living within a 25 km radius of an acrylonitrile factory in Nyergesujfalu, 

Hungary. Ascertainment of cases with congenital abnormalities was based on the Hungarian 

Congenital Abnormality Registry plus review of pediatric, pathology and cytogenetic records. 

Particular attention was paid to indicators of germinal mutations (sentinel anomalies, Down’s 

syndrome, and unidentified multiple congenital abnormalities) and to indicators of teratogens 

(specific pattern of multiple congenital abnormalities). Three congenital abnormalities: pectus 

excavatum in Tata, 1990-1992 (OR = 78.5, 95%CI = 8.4-729.6), undescended testis in 

Nyergesujfalu between 1980 and 1983 (8.6, 1.4-54.3) and in Esztergom, 1981-1982 (4. 2, 1.3-

13.5) and clubfoot in Tata, 1980-1981 (5.5, 1.5-20.3) showed significant time-space clusters in 

the study area. The risk of undescended testis decreased with increasing distance from the 

factory. An unusual increase for the combination of oral cleft and cardiac septal defects was 

seen in multimalformed babies in Tatabanya in 1990. Unfortunately there were no data on levels 

of acrylonitrile or any other exposure. 

V. Effects of Animal Exposure 

Appendix D3 12 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Quast et al. (1980) exposed Sprague-Dawley rats (100/sex/ concentration) 6 hours/day, 5 

days/week for 2 years to concentrations of 0, 20, or 80 ppm acrylonitrile vapors (0, 43, or 174 

mg/m3

). A statistically significant increase in mortality was observed in the first year among 80 

ppm exposed rats (male and female). Additionally, the 80 ppm exposed group had a significant 

decrease in mean body weight. Two tissues, the nasal respiratory epithelium and the brain, 

exhibited treatment-related adverse effects due to acrylonitrile exposure. Proliferative changes in 

the brain glial cells (i.e., tumors and early proliferation suggestive of tumors) were significantly 

increased in the 20 ppm (8/100) and 80 ppm (20/100) females versus female controls (0/100), 

and in the 80 ppm males (22/99) versus male controls (0/100). Noncarcinogenic, 

extrarespiratory effects were observed in the nasal turbinate epithelium at both exposure 

concentrations, 20 and 80 ppm (see table below). Thus the LOAEL was 20 ppm. No treatmentrelated effects in the olfactory epithelium, trachea, or lower respiratory epithelium were observed 

at either concentration. 

Effects of acrylonitrile reported by Quast et al. (1980) 

Effect Sex 0 ppm 20 ppm 80 ppm 

Respiratory epithelium hyperplasia in the nasal turbinates Male 0/11 4/12 10/10* 

Hyperplasia of the mucous secreting cells Male 0/11 7/12* 8/10* 

Focal inflammation in the nasal turbinates Female 2/11 6/10 7/10* 

Flattening of the respiratory epithelium of the nasal turbinates Female 1/11 7/10* 8/10* 

Lung: pneumonia, consolidation, atelectasis, or edema Male 14/100 27/100* 30/100* 

Lung: pneumonia, consolidation, atelectasis, or edema Female 7/100 2/100 7/100 

* statistically significant difference from controls (p<.05) 

Maltoni and associates exposed Sprague-Dawley rats (30/sex/concentration) to 0, 5, 10, 20, or 40 

ppm acrylonitrile vapor for 5 days/week over 52 weeks, and at 60 ppm for 4 to 7 days, 5 

days/week for 104 weeks (Maltoni et al., 1977; Maltoni et al., 1988). Histopathologic 

examinations were performed, including on lungs, brain, kidney, and liver. No noncarcinogenic 

effects were reported. 

Gagnaire et al. (1998) studied motor and sensory conduction velocities (MCV and SCV, 

respectively) and amplitudes of the sensory and motor action potentials (ASAP and AMAP) of 

the tail nerve in male Sprague-Dawley rats during chronic treatment with acrylonitrile. (Four 

other unsaturated aliphatic nitriles were also given orally to other rats.) Rats were given doses of 

12.5, 25, and 50 mg/kg of acrylonitrile once a day, 5 days per week for 12 weeks. Rats were also 

exposed by inhalation to 25, 50, and 100 ppm of acrylonitrile vapors for 6 h/day, 5 days per 

week, for 24 weeks and neurophysiological examinations were carried out. After oral 

acrylonitrile, animals developed behavioral sensitization characterized by salivation, locomotor 

hyperactivity, and moderately intense stereotypies. Rats dosed with 50 mg/kg developed 

hindlimb weakness associated with decreases in sensory conduction velocity (SCV) and in the 

amplitude of the sensory action potential (ASAP). Rats exposed to acrylonitrile by inhalation 

exhibited time- and concentration-dependent decreases in motor conduction velocity (MCV), 

SCV, and ASAP, which were partially reversible after 8 weeks of recovery. The authors 

concluded that the nervous system of the rat appears to be a target following either oral or 

inhalation exposures of acrylonitrile. The NOAEL by inhalation for 24 weeks was 25 ppm. 

Appendix D3 13 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Changes in electrophysiological parameters after 24 wks of exposure (Gagnaire et al., 1998) 

Acrylonitrile MCV (m/sec) SCV (m/sec) AMAP (mvolts) ASAP (μvolts) 

0 ppm 42.9 ± 0.9a 53.3 ± 1.0 17.8 ± 1.2 186 ± 8 

25 ppm 41.6 ± 0.8 50.5 ± 0.8* 16.1 ± 0.8 164 ± 11 

50 ppm 38.1 ± 0.9** 49.1 ± 0.5*** 15.7 ± 1.0 159 ± 5* 

100 ppm 38.5 ± 1.2** 48.4 ± 1.0*** 17.4 ± 0.9 133 ± 11*** 

a

 Mean ± SEM; * p<0.05; ** p<0.01; ***p<0.001 

In a developmental study, Murray et al. (1978) exposed rats to acrylonitrile vapors at 0, 40 ppm 

(87 mg/m3

), or 80 ppm (174 mg/m3) for 6 hours/day during gestational days 6 to 15. In the 80 

ppm exposed group, significant increases in fetal malformations were observed including short 

tail, missing vertebrae, short trunk, omphalocoele, and hemivertebra (Murray et al., 1978). No 

differences in implantations, live fetuses, or resorptions were seen in the exposed (40 and 80 

ppm) versus the control group. Maternal toxicity was observed as decreased body weight at both 

exposure levels. After adjustment to continuous exposure, this study identified a developmental 

NOAEL of 10 ppm and a LOAEL of 20 ppm (with maternal toxicity). 

Saillenfait et al. (1993) studied the developmental toxicity of eight aliphatic mononitriles in 

Sprague-Dawley rats after inhalation exposure for 6 hr/day during days 6 to 20 of gestation. The 

range of exposure levels for acrylonitrile was 12, 25, 50, and 100 ppm; group sizes were 20-23 

females. Embryolethality was observed after exposure to 25 ppm (54 mg/m3

) acrylonitrile in the 

presence of overt signs of maternal toxicity. Fetal weights were significantly lower at 25 ppm. 

Thus 12 ppm (26 mg/m3) is a NOAEL for developmental toxicity using this study design. 

Appendix D3 14 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VI. Derivation of Chronic Reference Exposure Level 

Study Quast et al., 1980 

Study population Sprague-Dawley rats (100/sex/concentration) 

Exposure method Discontinuous whole-body inhalation 

exposures (0, 20, or 80 ppm) 

Critical effects Degeneration and inflammation of nasal 

respiratory epithelium; hyperplasia of 

mucous secreting cells

LOAEL 20 ppm 

NOAEL Not observed 

BMC05 1.5 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Average experimental exposure 0.27 ppm for BMC05 (1.5 x 6/24 x 5/7) 

Human equivalent concentration 0.067 ppm (gas with extrathoracic respiratory 

effects; RGDR = 0.25 based on MV = 

0.33 m3

/day, SA(ET) = 11.6 cm2) 

Exposure duration 2 years 

LOAEL uncertainty factor Not needed in the BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure 

level 

0.002 ppm (2 ppb; 0.005 mg/m3; 5 μg/m3) 

Sprague-Dawley rats (100/sex/concentration) were exposed 6 hours/day, 5 days/week for 2 years 

to 0, 20, or 80 ppm acrylonitrile (0, 43, and 174 mg/m3

, respectively). Significant degenerative 

and inflammatory changes were observed in the respiratory epithelium of the nasal turbinates at 

both exposure concentrations (20 and 80 ppm). This treatment-related irritation of the nasal 

mucosa appeared in the 20 ppm exposed male rats as either epithelial hyperplasia of the nasal 

turbinates, or as hyperplasia of the mucous secreting cells. In the 20 ppm exposed females it 

appeared as either focal inflammation in the nasal turbinates or flattening of the respiratory 

epithelium of the nasal turbinates. In 80 ppm exposed rats the effects were more severe, 

including suppurative rhinitis, hyperplasia, focal erosions, and squamous metaplasia of the 

respiratory epithelium. No treatment-related effects in the olfactory epithelium, trachea, or lower 

respiratory system were observed at either concentration. This study identified a LOAEL for 

pathological alterations in the respiratory epithelium of the extrathoracic region of the respiratory 

tract of 20 ppm (43 mg/m3

). The U.S. EPA (1994) based its RfC of 2 μg/m3 on the same study 

but included a Modifying Factor (MF) of 10 for database deficiencies. The criteria for use of 

modifying factors are not well specified by U.S. EPA. Such modifying factors were not used by 

OEHHA. 

OEHHA used a benchmark dose approach to determine the chronic REL for acrylonitrile. The 

cumulative gamma distribution model in the U.S. EPA's BMDS sotware was individually fit to 

the data on respiratory epithelium hyperplasia in the nasal turbinates in males, hyperplasia of the 

Appendix D3 15 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

mucous secreting cells in males, focal inflammation in the nasal turbinates in females, and 

flattening of the respiratory epithelium of the nasal turbinates in females. The resulting BMC05

values (1.27, 1.33, 2.18, 1.35) were averaged to yield a value of 1.5 ppm. The RGDR adjustment 

and appropriate uncertainty factors were applied as indicated in the above table and resulted in a 

chronic REL of 5 μg/m3. 

For comparison, Gagnaire et al. (1998) found a NOAEL for nervous system effects at 24 weeks 

of 25 ppm, which is equivalent to a continuous exposure of 4.5 ppm. Use of the default RGDR 

of 1 for systemic effects, a subchronic UF of 3, an interspecies UF of 3, and an intraspecies UF 

of 10 results in an estimated REL of 45 ppb (100 μg/m3). We were unable to derive a BMC from 

the neurotoxicity data due partly to the tendency of the animals in the 100 ppm group to yield 

values for two of the four endpoints measured closer to the controls than those in the 50 ppm 

group. 

As another comparison, Saillenfait et al. (1983) found a 12 ppm (26 mg/m3) NOAEL for fetal 

weight reduction (6 h/d exposure). This is equivalent to a continuous exposure of 3 ppm (on 

days 6 to 20 of gestation). Use of the default RGDR of 1 for systemic effects, an interspecies UF 

of 3, and an intraspecies UF of 10 results in an estimated REL of 100 ppb (200 μg/m3

). 

Finally, after adjustment to continuous exposure, Murray et al. (1978) identified a developmental 

NOAEL, adjusted to continuous exposure, of 10 ppm and a LOAEL of 20 ppm (with maternal 

toxicity at both levels). Use of the default RGDR of 1 for systemic effects, an interspecies UF of 

3, and an intraspecies UF of 10 results in an estimated REL of 30 ppb (70 μg/m3). 

VII. Data Strengths and Limitations for Development of the REL

Significant strengths in the chronic REL for acrylonitrile include (1) the availability of chronic 

inhalation exposure data from a well-conducted study with histopathological analysis and (2) the 

demonstration of a dose-response relationship. Major uncertainties are (1) the lack of adequate 

human exposure data, (2) the lack of a NOAEL in the 2 year study, (3) lack of inhalation 

bioassay in a second species, and (4) lack of reproductive data for inhalation exposures when an 

oral study showed adverse reproductive effects 

When assessing the health effects of acrylonitrile, its carcinogenicity must also be assessed. 

VIII. Potential for Differential Impacts on Children's Health

The chronic REL is considerably lower than the comparison estimate based on developmental 

effects. Although neurotoxicity, an endpoint which is often associated with increased sensitivity 

of younger animals or humans, was evaluated as one of the alternative endpoints, the comparison 

reference level for this end point in adults was more than an order of magnitude higher that the 

REL based on histological changes in the upper respiratory tract. It is therefore considered that 

the REL is likely to be adequately protective of infants and children. 

Appendix D3 16 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IX. References

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Czeizel AE, Hegedus S and Timar L. 1999. Congenital abnormalities and indicators of germinal 

mutations in the vicinity of an acrylonitrile producing factory. Mutat. Res. 427(2):105-123 

Gagnaire F, Marignac B, and Bonnet P. 1998. Relative neurotoxicological properties of five 

unsaturated aliphatic nitriles in rats. J. Appl. Toxicol. 18(1):25-31. 

HSDB. 1994. Hazardous Substances Data Bank. TOMES® Vol. 20. Denver, CO: Micromedex, 

Inc. 

Jakubowski M, Linhart I, Pielas G, and Kopecky J. 1987. 2-Cyanoethylmercapturic acid 

(CEMA) in the urine as a possible indicator of exposure to acrylonitrile. Br. J. Ind. Med. 44:834- 

840. 

Maltoni C, Ciliberti A, and Di Maio V. 1977. Carcinogenicity bioassays on rats of acrylonitrile 

administered by inhalation and by ingestion. Med. Lav. 68(6):401-411. 

Maltoni C, Ciliberti A, Cotti G, and Perino G. 1988. Long-term carcinogenicity bioassays on 

acrylonitrile administered by inhalation and by ingestion to Sprague-Dawley rats. Ann. NY 

Acad. Sci. 534:179-202. 

Murray FJ, Schwetz BA, Nitschke KD, John JA, Norris JM, and Gehring PJ. 1978. 

Teratogenicity of acrylonitrile given to rats by gavage or by inhalation. Food Cosmet. Toxicol. 

16(6):547-552. 

Quast JF, Schwetz DJ, Balmer MF, Gunshow TS, Park CN, and McKenna MJ. 1980. A two-year 

toxicity and oncogenicity study with acrylonitrile following inhalation exposure of rats. 

Toxicology Research Laboratory. Midland, MI: Dow Chemical Co. 

Saillenfait AM, Bonnet P, Guenier JP and de Ceaurriz J. 1993. Relative developmental toxicities 

of inhaled aliphatic mononitriles in rats. Fundam. Appl. Toxicol. 20(3):365-375. 

Sakurai H, Onodera T, Utsunomiya T, Minakuchi H, Iwai H, and Matsumura H. 1978. Health 

effects of acrylonitrile in acrylic fibre factories. Br. J. Ind. Med. 35:219-225. 

United States Environmental Protection Agency (U.S. EPA) (1993). Kelly TJ RMPASCCL. 

Ambient Concentration Summaries for Clean Air Act Title III Hazardous Air Pollutants. U.S. 

EPA Contract No. 68-D80082. 

Appendix D3 17 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

U.S. EPA 1994. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) Database. Reference concentration (RfC) for acrylonitrile. Available online at 

http://www.epa.gov/ngispgm3/iris 

Wilson RH, Hough GV and McCormick WE. 1948. Medical problems encountered in the 

manufacture of American-made rubber. Ind. Med. 17:199-207.

Appendix D3 18 Acrylonitrile 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

AMMONIA 

(Anhydrous ammonia; aqueous ammonia) 

CAS Registry Number: 7664-41-7 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 200 μg/m3

 (300 ppb) 

Critical effect(s) Pulmonary function tests or subjective 

symptomatology in workers 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (From HSDB, 1994; 1999) 

Description Colorless gas 

Molecular formula NH3

Molecular weight 17.03 g/mol 

Density 0.7710 g/L @ 0°C 

Boiling point -33.35° C 

Vapor pressure 7510 torr @ 25°C 

Solubility Soluble in water, alcohol, and ether 

Conversion factor 1 ppm = 0.71 mg/m3

III. Major Uses or Sources 

This strongly alkaline chemical is widely used in industry as a feed stock for nitrogen-based 

chemicals such as fertilizers, plastics and explosives (ATSDR, 1990). Ammonia is also used as a 

refrigerant. The general public is exposed by off-gasing from cleaning solutions containing 

aqueous ammonia. Household ammonia solutions contain 5-10% ammonia in water while 

industrial strength can be up to 28%. The annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 21,832,909 pounds of ammonia (CARB, 1999). 

IV. Effects of Human Exposures 

Comparisons were made between 52 workers and 31 control subjects in a soda ash plant for 

pulmonary function and eye, skin and respiratory symptomatology (Holness et al., 1989). The 

pulmonary function tests included FVC (forced vital capacity – the total amount of air the 

subject can expel during a forced expiration), FEV1 (forced expiratory volume in one second), 

FEF50 (forced expiratory flow rate at 50% of the FVC) and FEF75 (forced expiratory flow rate at 

Appendix D3 19 Ammonia 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

75% of the FVC). Age, height, and pack-years smoked were treated as covariates for the 

comparisons. The workers were exposed on average for 12.2 years to mean (time-weighted 

average) ammonia concentrations of 9.2 ppm (6.4 mg/m3) ± 1.4 ppm, while controls were 

exposed to 0.3 ppm (0.21 mg/m3

) ± 0.1 ppm. No differences in any endpoints (respiratory or 

cutaneous symptoms, sense of smell, baseline lung function, or change in lung function over a 

work shift at the beginning and end of a workweek) were reported between the exposed and 

control groups.

Groups of human volunteers were exposed to 25, 50, or 100 ppm (0, 17.8, 35.5, or 71 mg/m3

) 

ammonia 5 days/week for 2, 4, or 6 hours/day, respectively, for 6 weeks (Ferguson et al., 1977). 

Another group of 2 volunteers was exposed to 50 ppm ammonia for 6 hours/day for 6 weeks. 

Group Exposure Week 1 Week 2 Week 3 Week 4 Week 5 Week 6 

A ppm NH3

hours 

25 

2 

50 

4 

100 

6 

25 

2 

50 

4 

100 

6 

B ppm NH3

hours 

50 

6 

50 

6 

50 

6 

50 

6 

50 

6 

50 

6 

C ppm NH3

hours 

100 

6 

50 

4 

25 

2 

25 

6 

50 

4 

100 

2 

Pulmonary function tests (respiration rate, FVC and FEVl) were measured in addition to 

subjective complaints of irritation of the eyes and respiratory tract. The difficulty experienced in 

performing simple cognitive tasks was also measured, as was pulse rate. There were reports of 

transient irritation of the nose and throat at 50 or 100 ppm. Acclimation to eye, nose, and throat 

irritation was seen after two to three weeks (in addition to the short-term subjective adaption). 

No significant differences between subjects or controls on common biological indicators, in 

physical exams, or in performance of normal job duties were found. After acclimation, 

continuous exposure to 100 ppm, with occasional excursions to 200 ppm, was easily tolerated 

and had no observed effect on general health. 

V. Effects of Animal Exposures 

Rats were continuously exposed to ammonia at 0, 25, 50, 150, or 250 ppm (0, 18, 36, 107, or 

179 mg/m3) ammonia for 7 days prior to intratracheal inoculation with Mycoplasma pulmonis, 

and from 28 to 42 days following M. pulmonis exposure (Broderson et al., 1976). All exposures 

to ammonia resulted in significantly increased severity of rhinitis, otitis media, tracheitis, and 

pneumonia characteristic of M. pulmonis infection, therefore 25 ppm was a LOAEL in this 

subchronic study. Exposure to 250 ppm ammonia alone resulted in nasal lesions (epithelial 

thickening and hyperplasia) which were not like those seen in M. pulmonis-infected rats. 

The growth of bacteria in the lungs and nasal passages, and the concentration of serum 

immunoglobulin were significantly increased in rats exposed to 100 ppm (71 mg/m3

) ammonia 

over that seen in control rats (Schoeb et al., 1982). 

Appendix D3 20 Ammonia 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Guinea pigs (10/group) and mice (20/group) were continuously exposed to 20 ppm (14.2 mg/m3) 

ammonia for up to 6 weeks (Anderson et al., 1964). Separate groups of 6 guinea pigs and 21 

chickens were exposed to 50 ppm and 20 ppm ammonia for up to 6 and 12 weeks, respectively. 

All species displayed pulmonary edema, congestion, and hemorrhage after 6 weeks exposure, 

whereas no effects were seen after only 2 weeks. Guinea pigs exposed to 50 ppm ammonia for 6 

weeks exhibited enlarged and congested spleens, congested livers and lungs, and pulmonary 

edema. Chickens exposed to 200 ppm for 17-21 days showed liver congestion and slight 

clouding of the cornea. Anderson and associates also showed that a 72-hour exposure to 20 ppm 

ammonia significantly increased the infection rate of chickens exposed to Newcastle disease 

virus, while the same effect was observed in chickens exposed to 50 ppm for just 48 hours. 

Coon et al. (1970) exposed groups of rats (as well as guinea pigs, rabbits, dogs, and monkeys) 

continuously to ammonia concentrations ranging from 40 to 470 mg/m3. There were no signs of 

toxicity in 15 rats exposed continuously to 40 mg/m3

 for 114 days or in 48 rats exposed 

continuously to 127 mg/m3 for 90 days. Among 49 rats exposed continuously to 262 mg/m3

 for 

90 days, 25% had mild nasal discharge. At 455 mg/m3

 50 of 51 rats died. Thus 127 mg/m3 (179 

ppm) is a subchronic NOAEL for upper respiratory effects in rats. Coon et al. (1970) also found 

no lung effects in 15 guinea pigs exposed continuously to 40 mg/m3 (28 ppm) ammonia for 114 

days.

VI. Derivation of Chronic Reference Exposure Level 

Study Holness et al., 1989 (supported by Broderson et 

al., 1976) 

Study population 52 workers; 31 controls 

Exposure method Occupational inhalation 

Critical effects Pulmonary function, eye, skin, and respiratory 

symptoms of irritation 

LOAEL 25 ppm (Broderson et al., 1976) (rats) 

NOAEL 9.2 ppm (Holness et al., 1989) 

Exposure continuity 8 hours/day (10 m3

/day occupational inhalation 

rate), 5 days/week 

Exposure duration 12.2 years 

Average occupational exposure 3 ppm for NOAEL group (9.2 x 10/20 x 5/7) 

Human equivalent concentration 3 ppm for NOAEL group 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 10 

Inhalation reference exposure level 0.3 ppm ( 300 ppb; 0.2 mg/m3

; 200 μg/m3) 

The Holness et al. (1989) study was selected because it was a chronic human study and was 

published in a respected, peer-reviewed journal. It is also the only chronic study available. The 

USEPA (1995) based its RfC of 100 μg/m3

 on the same study but included a Modifying Factor 

Appendix D3 21 Ammonia 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

(MF) of 3 for database deficiencies. The criteria for use of modifying factors are not well 

specified by U.S. EPA. Such modifying factors were not used by OEHHA. 

For comparison with the proposed REL of 200 μg/m3 based on human data, we estimated RELs 

from 2 animal studies. (1) Anderson et al. (1964) exposed guinea pigs continuously to 50 ppm 

(35 mg/m3

) ammonia for 6 weeks and observed pulmonary edema. Use of an RGDR of 0.86 and 

a cumulative uncertainty factor of 3000 (10 for use of a LOAEL, 10 for subchronic, 3 for 

interspecies, and 10 for intraspecies) resulted in a REL of 10 μg/m3

. Staff note that the nearly 

maximal total uncertainty factor of 3000 was used in this estimation. (2) Coon et al. (1970) 

exposed rats continuously to 127 mg/m3 ammonia for 90 days and saw no signs of toxicity. Use 

of an RGDR(ET) of 0.16 for nasal effects (observed in rats exposed to higher levels of ammonia 

in Broderson et al. (1976)) and a cumulative uncertainty factor of 100 (3 for subchronic, 3 for 

interspecies, and 10 for intraspecies) resulted in a REL of 200 μg/m3

. 

VII. Data Strengths and Limitations for Development of the REL 

Significant strengths in the ammonia REL include (1) the availability of long-term human 

inhalation exposure data (Holness et al., 1989), (2) the demonstration of consistent effects in 

experimentally exposed human volunteers following short-term exposures (Ferguson et al., 

1977), and (3) reasonable consistency with animal data (Coon et al., 1970). 

Major areas of uncertainty are (1) the lack of a NOAEL and LOAEL in a single study, (2) a lack 

of animal data with chronic exposure and histopathological analyses, and (3) difficulties in 

estimated human occupational exposures. The overall database for this common chemical is 

limited. 

VIII. References 

ATSDR. 1990. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Ammonia. Atlanta, GA: ATSDR, U.S. Public Health Service. 

Anderson DP, Beard CW, and Hanson RP. 1964. The adverse effects of ammonia on chickens 

including resistance to infection with Newcastle disease virus. Avian. Dis. 8:369379. 

Broderson JR, Lindsey JR, and Crawford JE. 1976. The role of environmental ammonia in 

respiratory mycoplasmosis of rats. Am. J. Pathol. 85(1):115-130. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Coon RA, Jones RA, Jenkins LJ Jr, and Siegel J. 1970. Animal inhalation studies on ammonia, 

ethylene glycol, formaldehyde, dimethylamine, and ethanol. Toxicol. Appl. Pharmacol. 

16(3):646-655. 

Ferguson WS, Koch WC, Webster LB, and Gould JR. 1977. Human physiological response and 

adaptation to ammonia. J. Occup. Med. 19(5):319-326. 

Appendix D3 22 Ammonia 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

HSDB. 1994. Hazardous Substances Data Bank. TOMES®. Vol .20. Denver, CO: Micromedex, 

Inc. 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Holness DL, Purdham JT, and Nethercott JR. 1989. Acute and chronic respiratory effects of 

occupational exposure to ammonia. Am. Ind. Hyg. Assoc. J. 50(12):646-650. 

Schoeb TR, Davidson MK, and Lindsey JR. 1982. Intracage ammonia promotes growth of 

Mycoplasma pulmonis in the respiratory tract of rats. Infect. Immun. 38:212-217. 

U.S. EPA. 1995. U. S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration (RfC) for ammonia. 

Appendix D3 23 Ammonia 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

BENZENE 

(Benzol; Benzole; Cyclohexatriene) 

CAS Registry Number: 71-43-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 60 μg/m3 (20 ppb)

Critical effect(s) Lowered red and white blood cell counts in 

occupationally exposed humans 

Hazard index target(s) Hematopoietic system; development; nervous 

system 

II. Physical and Chemical Properties (HSDB, 1994; 1999) 

Description Colorless liquid 

Molecular formula C6H6

Molecular weight 78.1 g/mol 

Density 0.879 g/cm3

 @ 25° C 

Boiling point 80.1°C 

Vapor pressure 100 torr @ 26.1°C 

Solubility Soluble in ethanol, chloroform, ether, carbon 

disulfide, acetone, oils, and glacial acetic 

acid; slightly soluble in water 

Conversion factor 1 ppm = 3.2 mg/m3

 @ 25° C 

III. Major Uses or Sources 

Benzene has been widely used as a multipurpose organic solvent. This use is now discouraged 

due to its high toxicity, including carcinogenicity. Present uses include use as a raw material in 

the synthesis of styrene, phenol, cyclohexane, aniline, and alkyl benzenes in the manufacture of 

various plastics, resins, and detergents. Syntheses of many pesticides and pharmaceuticals also 

involve benzene as a chemical intermediate (HSDB, 1994). The tire industry and shoe factories 

use benzene extensively in their manufacturing processes. Annual demand in the U.S. was 

estimated to be 6 million tons in 1990 (HSDB, 1994). Benzene exposure also occurs as a result 

of gasoline and diesel fuel use and combustion (Holmberg and Lundberg, 1985). In 1996, the 

latest year tabulated, the statewide mean outdoor monitored concentration of benzene was 

approximately 0.7 ppb (CARB, 1999a). Annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 750,364 pounds of benzene (CARB, 1999b). (This does not include the 

large amount of benzene emitted by mobile sources.) 

Appendix D3 24 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Human Exposure 

The primary toxicological effects of chronic benzene exposure are on the hematopoeitic system. 

Neurological and reproductive/developmental toxic effects are also of concern at slightly higher 

concentrations. Impairment of immune function and/or various anemias may result from the 

hematotoxicity. The hematologic lesions in the bone marrow can lead to peripheral 

lymphocytopenia and/or pancytopenia following chronic exposure. Severe benzene exposures 

can also lead to life-threatening aplastic anemia. These lesions may lead to the development of 

leukemia years after apparent recovery from the hematologic damage (DeGowin, 1963). 

Kipen et al. (1988) performed a retrospective longitudinal study on a cohort of 459 rubber 

workers, examining the correlation of average benzene exposure with total white blood cell 

counts taken from the workers. These researchers found a significant (p < 0.016) negative 

correlation between average benzene concentrations in the workplace and white blood cell 

counts in workers from the years 1940-1948. A reanalysis of these data by Cody et al. (1993) 

showed significant decreases in RBC and WBC counts among a group of 161 workers during the 

1946-1949 period compared with their pre-exposure blood cell counts. The decline in blood 

counts was measured over the course of 12 months following start of exposure. During the 

course of employment, workers who had low monthly blood cell counts were transferred to other 

areas with lower benzene exposures, thus potentially creating a bias towards non-significance or 

removing sensitive subjects from the study population. Since there was a reported 75% rate of 

job change within the first year of employment, this bias could be highly significant. In addition, 

there was some indication of blood transfusions used to treat some “anemic” workers, which 

would cause serious problems in interpreting the RBC data, since RBCs have a long lifespan in 

the bloodstream. The exposure analysis in this study was performed by Crump and Allen 

(1984). The range of monthly median exposures was 30-54 ppm throughout the 12-month 

segment examined. Despite the above-mentioned potential biases, workers exposed above the 

median concentrations displayed significantly decreased WBC and RBC counts compared with 

workers exposed to the lower concentrations using a repeated measures analysis of variance. 

Tsai et al. (1983) examined the mortality from all cancers and leukemia, in addition to 

hematologic parameters in male workers exposed to benzene for 1-21 years in a refinery from 

1952-1978. The cohort of 454 included maintenance workers and utility men and laborers 

assigned to benzene units on a “regular basis”. Exposures to benzene were determined using 

personal monitors; the median air concentration was 0.53 ppm in the work areas of greatest 

exposure to benzene. The average length of employment in the cohort was 7.4 years. The 

analysis of overall mortality in this population revealed no significant excesses. Mortality from 

all causes and from diseases of the circulatory system was significantly below expected values 

based on comparable groups of U.S. males. The authors concluded the presence of a healthy 

worker effect. An internal comparison group of 823 people, including 10% of the workers who 

were employed in the same plant in operations not related to benzene, showed relative risks for 

0.90 and 1.31 for all causes and cancer at all sites, respectively (p < 0.28 and 0.23). A subset of 

303 workers was followed for medical surveillance. Up to four hematological tests per year 

were conducted on these workers. Total and differential white blood cell counts, hemoglobin, 

Appendix D3 25 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

hematocrit, red blood cells, platelets and clotting times were found to be within normal (between 

5% and 95% percentile) limits in this group. 

Collins et al. (1997) used routine data from Monsanto’s medical/industrial hygiene system to 

study 387 workers with daily 8-hour time-weighted exposures (TWA) averaging 0.55 ppm 

benzene (range = 0.01 – 87.69 ppm; based on 4213 personal monitoring samples, less than 5% of 

which exceeded 2 ppm). Controls were 553 unexposed workers. There was no increase in the 

prevalence of lymphopenia, an early, sensitive indicator of benzene toxicity, among exposed 

workers (odds ratio = 0.6; 95% confidence interval = 0.2 to 1.8), taking into account smoking, 

age, and sex. There also was no increase in risk among workers exposed 5 or more years (odds 

ratio = 0.6; 95% confidence interval = 0.2 to 1.9). There were no differences between exposed 

and unexposed workers for other measures of hematotoxicity, including mean corpuscular 

volume and counts of total white blood cells, red blood cells, hemoglobin, and platelets. 

Rothman et al. (1996) compared hematologic outcomes in a cross-sectional study of 44 male and 

female workers heavily exposed to benzene (median = 31 ppm as an 8-hr TWA) and 44 age and 

gender-matched unexposed controls from China. Hematologic parameters (total WBC, absolute 

lymphocyte count, platelets, red blood cells, and hematocrit) were decreased among exposed 

workers compared to controls; an exception was the red blood cell mean corpuscular volume 

(MCV), which was higher among exposed subjects. In a subgroup of 11 workers with a median 

8 hr TWA of 7.6 ppm (range = 1-20 ppm) and not exposed to more than 31 ppm on any of 5 

sampling days, only the absolute lymphocyte count was significantly different between exposed 

workers and controls (p = 0.03). Among exposed subjects, a dose response relationship with 

various measures of current benzene exposure (i.e., personal air monitoring, benzene metabolites 

in urine) was present only for the total WBC count, the absolute lymphocyte count, and the 

MCV. Their results support the use of the absolute lymphocyte count as the most sensitive 

indicator of benzene-induced hematotoxicity. 

An examination of 32 patients, who were chronically exposed to benzene vapors ranging from 

150 to 650 ppm for 4 months to 15 years, showed that pancytopenia occurred in 28 cases. Bone 

marrow punctures revealed variable hematopoeitic lesions, ranging from acellularity to 

hypercellularity (Aksoy et al., 1972). 

Central nervous system disorders have been reported in individuals with pancytopenia following 

chronic occupational benzene exposure to unknown concentrations for an average length of time 

of 6 years (Baslo and Aksoy, 1982). 

Runion and Scott (1985) estimated a composite geometric mean benzene concentration in 

various workplaces containing benzene to be 0.1 ppm (0.32 mg/m3

) (geometric standard 

deviation = 7.2 ppm, 23.3 mg/m3). This estimate was based on samples collected by industrial 

hygienists between the years 1978 and 1983. 

V. Effects of Animal Exposure

Appendix D3 26 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

A number of animal studies have demonstrated that benzene exposure can induce bone marrow 

damage, changes in circulating blood cells, developmental and reproductive effects, alterations 

of the immune response, and cancer. With respect to chronic toxicity, hematological changes 

appear to be the most sensitive indicator. 

Wolf et al. (1956) studied the effects of repeated exposure to benzene in rabbits (80 ppm, 175 

total exposures), rats (88 ppm, 136 total exposures) and guinea pigs (88 ppm, 193 total 

exposures). The observed effects included leukopenia, increased spleen weight, and histological 

changes to the bone marrow. Hematologic effects, including leukopenia, were observed in rats 

exposed to mean concentrations of 44 ppm (143 mg/m3) or greater for 5-8 weeks (Deichmann et 

al., 1963). Exposure to 31 ppm (100 mg/m3) benzene or less did not result in leukopenia after 3-

4 months of exposure. Snyder et al. (1978) exposed Sprague-Dawley rats and AKR/J mice to 

300 ppm benzene, 6 hours/day, 5 days/week for life. Lymphocytopenia, anemia and decreased 

survival time were observed in both species. Cronkite et al. (1982) exposed male mice to 400 

ppm benzene, 6 hours/day, 5 days/week for 9.5 weeks and observed depressed bone marrow 

cellularity, decreased stem cell count, and altered morphology in spleen colony-forming cells. 

Mice have been shown to be more sensitive than rats or rabbits to the hematologic and leukemic 

effects of benzene (Sabourin et al., 1989; IARC, 1982). Sabourin et al. (1988) showed that 

metabolism of benzene to the toxic hydroquinone, muconic acid, and hydroquinone glucuronide 

was much more prevalent in the mouse than in rats, whereas the detoxification pathways were 

approximately equivalent between the two species. 

A study on the chronic hematological effects of benzene exposure in C57 Bl/6 male mice (5-6 

per group) showed that peripheral lymphocytes, red blood cells and colony-forming units (CFUs) 

in the bone marrow and spleen were significantly decreased in number after treatment with 10 

ppm (32.4 mg/m3) benzene for 6 hours/day, 5 days/week for 178 days (Baarson et al., 1984). 

Inhalation of 0, 10, 31, 100, or 301 ppm (0, 32.4, 100.4, 324, or 975 mg/m3) benzene for 6 

hours/day for 6 days resulted in a dose-dependent reduction in peripheral lymphocytes, and a 

reduced proliferative response of B- and T-lymphocytes to mitogenic agents in mice (Rozen et 

al., 1984). In this study, total peripheral lymphocyte numbers and B-lymphocyte proliferation to 

lipopolysaccharide were significantly reduced at a concentration of 10 ppm (32.4 mg/m3). The 

proliferation of T-lymphocytes was significantly reduced at a concentration of 31 ppm 

(100.4 mg/m3

). 

Male and female mice (9-10 per group) exposed to 100 ppm (324 mg/m3

) benzene or greater for 

6 hours/day, 5 days/week for 2 weeks showed decreased bone marrow cellularity and a reduction 

of pluripotent stem cells in the bone marrow (Cronkite et al., 1985). The decrease in marrow 

cellularity continued for up to 25 weeks following a 16-week exposure to 300 ppm (972 mg/m3

) 

benzene. Peripheral blood lymphocytes were dose-dependently decreased with benzene 

exposures of greater than 25 ppm (81 mg/m3

) for 16 weeks, but recovered to normal levels 

following a 16-week recovery period. 

Ward et al. (1985) exposed 50 Sprague-Dawley rats and 150 CD-1 mice of both sexes to 0, 1, 10, 

30, or 300 ppm benzene, 6 hours/day, 5 days/week for 13 weeks. Serial sacrifices were 

Appendix D3 27 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

conducted at 7, 14, 28, 56, and 91 days. No hematological changes were found for mice and rats 

at 1, 10, or 30 ppm in this study. Significant increases in mean cell volume and mean cell 

hemoglobin values and decreases in hematocrit, hemoglobin, lymphocyte percentages, and 

decreases in red cell, leukocyte and platelet counts were observed in male and female mice at 

300 ppm. The changes were first observed after 14 days of exposure. Histological changes in 

mice included myeloid hypoplasia of the bone marrow, lymphoid depletion in the mesenteric 

lymph node, increased extramedullary hematopoiesis in the spleen, and periarteriolar lymphoid 

sheath depletion. Effects were less severe in the rats. 

Aoyama (1986) showed that a 14-day exposure of mice to 50 ppm (162 mg/m3) benzene resulted 

in a significantly reduced blood leukocyte count. 

The NTP (1986) conducted a bioassay in F344 rats and B6C3F1 mice of benzene by corn oil 

gavage. Doses were 0, 25, 50, and 100 mg/kg-day for females and 0, 50, 100, and 200 mg/kgday for males. Dose-related lymphocytopenia and leukocytopenia were observed in both species 

in all dosed groups but not controls. Mice exhibited lymphoid depletion of the thymus and 

spleen and hyperplasia of the bone marrow. 

Cronkite et al. (1989) exposed CBA/Ca mice to 10, 25, 100, 300, 400 and 3000 ppm benzene 6 

hours/day, 5 days/week for up to 16 weeks. No effects were observed at the 10 ppm level. 

Lymphopenia was observed in the 25 ppm exposure group. Higher concentrations of benzene 

produced dose-dependent decreases in blood lymphocytes, bone marrow cellularity, spleen 

colony-forming units, and an increased percentage of CFU-S in S-phase synthesis. 

Farris et al. (1997) exposed B6C3F1 mice to 1, 5, 10, 100, and 200 ppm benzene for 6 hr/day, 5 

days/week, for 1, 2, 4, or 8 weeks. In addition some animals were allowed to recover from the 

exposure. There were no significant effects on hematopoietic parameters from exposure to 10 

ppm benzene or less. Exposure to higher levels reduced the number of total bone marrow cells, 

progenitor cells, differentiating hematopoietic cells, and most blood parameters. The replication 

of primitive progenitor cells was increased. The authors suggested that this last effect, in 

concert with the genotoxicity of benzene, could account for the carcinogenicity of benzene at 

high concentrations. 

Reproductive and developmental effects have been reported following benzene exposure. Coate 

et al. (1984) exposed groups of 40 female rats to 0, 1, 10, 40, and 100 ppm (0, 3.24, 32.4, 129.6, 

or 324 mg/m3

) benzene for 6 hours/day during days 6-15 of gestation. In this study, teratologic 

evaluations and fetotoxic measurements were done on the fetuses. A significant decrease was 

noted in the body weights of fetuses from dams exposed to 100 ppm (324 mg/m3). No effects 

were observed at a concentration of 40 ppm (129.6 mg/m3

). 

Keller and Snyder (1986) reported that exposure of pregnant mice to concentrations as low as 

5 ppm (16 mg/m3

) benzene on days 6-15 of gestation (6 hr/day) resulted in bone-marrow 

hematopoietic changes in the offspring that persisted into adulthood. However, the 

hematopoietic effects (e.g. bimodal changes in erythroid colony-forming cells) in the above 

study were of uncertain biological significance. In a similar later study, Keller and Snyder 

(1988) found that exposure of mice in utero to 20 ppm (64 mg/m3

) benzene on days 6-15 of 

gestation resulted in neonatal suppression of erythropoietic precursor cells and persistent, 

Appendix D3 28 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

enhanced granulopoiesis. This effect was considered significant bone-marrow toxicity by the 

authors. No hematotoxicity was seen in this study at 10 ppm (32 mg/m3

). 

An exposure of 500 ppm (1,600 mg/m3

) benzene through days 6-15 gestation was teratogenic in 

rats while 50 ppm (160 mg/m3

) resulted in reduced fetal weights on day 20 of gestation. No fetal 

effects were noted at an exposure of 10 ppm (Kuna and Kapp, 1981). An earlier study by 

Murray et al. (1979) showed that inhalation of 500 ppm benzene for 7 hours/day on days 6-15 

and days 6-18 of gestation in mice and rabbits, respectively, induced minor skeletal variations in 

the absence of maternal toxicity. Red and white blood cell counts in the adults of either species 

were measured by Murray et al. (1979) but were not significantly different from control animals. 

However, fetal mouse hematological effects were not measured. 

Tatrai et al. (1980) demonstrated decreased fetal body weights and elevated liver weights in rats 

exposed throughout gestation to 150 mg/m3

 (47 ppm). 

VI. Derivation of Chronic Reference Exposure Level (REL)

Study Tsai et al. (1983) 

Study population 303 Male refinery workers 

Exposure method Occupational exposures for 1-21 years 

Critical effects Hematological effects 

LOAEL Not observed 

NOAEL 0.53 ppm 

Exposure continuity 8 hr/day (10 m3

 per 20 m3 day), 5 days/week 

Exposure duration 7.4 years average (for the full cohort of 454); 

 32% of the workers were exposed for more 

than 10 years 

Average occupational exposure 0.19 ppm 

Human equivalent concentration 0.19 ppm 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 10 

Inhalation reference exposure level 0.02 ppm (20 ppb; 0.06 mg/m3

; 60 μg/m3) 

Staff identified Tsai et al. (1983) as the most appropriate study for a chronic REL derivation. 

The authors examined hematologic parameters in 303 male workers exposed to benzene for 1-21 

years in a refinery from 1952-1978. Follow-up success was 99.3% in the entire cohort of 359. A 

total of approximately 1400 samples for hematological tests and 900 for blood chemistry tests 

were taken between 1959 and 1979. Exposures to benzene were determined using personal 

monitors. Data consisting of 1394 personal samples indicated that 84% of all benzene samples 

were less than 1 ppm; the median air concentration of benzene was 0.53 ppm in the work areas 

of greatest exposure to benzene (“benzene related areas”, for example, production of benzene 

Appendix D3 29 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

and cyclohexane and also of cumene). The average length of employment in the cohort was 7.4 

years. Mortality from all causes and from diseases of the circulatory system was significantly 

below expected values based on comparable groups of U.S. males. The authors concluded the 

presence of a healthy worker effect. An analysis using an internal comparison group of 823 

people, including 10% of the workers who were employed in the same plant in operations not 

related to benzene, showed relative risks for 0.90 and 1.31 for all causes and cancer at all sites, 

respectively (p < 0.28 and 0.23). Total and differential white blood cell counts, hemoglobin, 

hematocrit, red blood cells, platelets and clotting times were found to be within normal (between 

5% and 95% percentile) limits in this group. Although the exposure duration averaged only 7.4 

years, the study was considered to be chronic since 32% of the workers had been exposed for 

more than 10 years. 

VII. Data Strengths and Limitations for Development of the REL 

Both the animal and human databases for benzene are excellent. Although the study by Tsai et 

al. (1983) is a free-standing NOAEL, the endpoint examined is a known sensitive measure of 

benzene toxicity in humans. In addition, the LOAEL for the same endpoint in workers reported 

by Cody et al. (1993) help form a dose-response relationship and also yield an REL which is 

consistent with that derived from Tsai et al. (1983). The study by Cody et al. (1993), since it 

failed to identify a NOAEL and was only for a period of 1 year, contained a greater degree of 

uncertainty in extrapolation to a chronic community Reference Exposure Level. The recent 

results of Collins et al. (1997) that included a NOAEL of 0.55 ppm and of Rothman et al. (1996) 

that included a LOAEL of 7.6 ppm are consistent with those of Tsai et al. Therefore the study 

by Tsai et al. (1983) was used as the basis for the chronic REL for benzene. 

In the Cody et al. (1993) study, significant hematological effects, including reduced RBC and 

WBC counts, were observed in 161 male rubber workers exposed to median peak concentrations 

(i.e., only the peak concentrations for any given exposure time were reported) of 30-54 ppm or 

more for a 12-month period during 1948. The 30 ppm value was considered a 1-year LOAEL 

for hematological effects. In this rubber plant, workers who had blood dyscrasias were excluded 

from working in the high benzene units. Furthermore, individual workers having more than a 

25% decrease in WBC counts from their pre-employment background count were removed from 

the high benzene units and placed in other units with lower benzene concentrations. Sensitive 

individuals therefore could have been excluded from the analysis. The 30 ppm value is the low 

end of the range of median values (30-54 ppm) reported by Crump and used in the Kipen et al.

(1988) and Cody et al. (1993) studies. An equivalent continuous exposure of 10.7 ppm can be 

calculated by assuming that workers inhaled 10 m3

 of their total 20 m3 of air per day during their 

work-shift, and by adjusting for a normal 5 day work week. Application of uncertainty factors 

for subchronic exposures, estimation of a NOAEL, and for protection of sensitive subpopulations 

(10 for each) results in an REL of 0.01 ppm (10 ppb; 30 μg/m3

). This is comparable to the REL 

based on Tsai et al. (1983). 

Ward et al. (1996) determined a relationship between occupational exposures to benzene and 

decreased red and white cell counts. A modeled dose-response relationship indicated a 

possibility for hematologic effects at concentrations below 5 ppm. However, no specific 

measures of the actual effects at concentrations below 2 ppm were taken, and the Tsai et al. 

(1983) data were not considered in their analysis. The purpose of this study was to characterize 

Appendix D3 30 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

the trend for effects at low concentrations of benzene. A NOAEL or LOAEL was not identified 

in the study. The selection of a NOAEL of 0.53 ppm is therefore not inconsistent with the results 

of the Ward et al. (1996) study. 

The human data presented by Tsai and associates were selected over animal studies because the 

collective human data were considered adequate in terms of sample size, exposure duration, and 

health effects evaluation. 

For comparison with the REL of 20 ppb based on human data, we estimated a REL based on the 

chronic inhalation study in mice by Baarson et al. (1984), which showed that bone-marrow 

progenitor cells were markedly suppressed after intermittent exposures (6 hr/day, 5 days/week) 

to 10 ppm benzene for 6 months. An extrapolation of this value to an equivalent continuous 

exposure resulted in a concentration of 1.8 ppm. Application of an RGDR of 1 for a systemic 

effect and uncertainty factors of 3 and 10 for inter- and intraspecies variability, and 10 for 

estimation of a NOAEL from the LOAEL would result in an REL of 6 ppb (20 μg/m3). The 

Farris et al. (1997) 8 week study indicated a LOAEL of 100 ppm and a NOAEL of 10 ppm for 

hematological effects. Application of an RGDR of 1 and UFs of 10 for subchronic, 3 for 

interspecies and 10 for intraspecies extrapolation (total UF = 300) also resulted in an estimated 

REL of 6 ppb, in reasonable agreement with the proposed REL of 20 ppb. One could also 

crudely approximate an inhalation REL from the oral NTP bioassay where a dose of 25 mg/kgday was associated with hematological effects. The concentration approximately equivalent to a 

25 mg/kg dose for a 70 kg human breathing 20 cubic meters per day is 27 ppm. Assuming this is 

a LOAEL and applying an RGDR of 1 for systemic effects, a 3 fold UF for extrapolation to 

humans, a 10-fold UF for LOAEL to NOAEL extrapolation and a 10-fold UF for intraspecies 

extrapolation yields a REL of 90 ppb. There are a number of uncertainties to this approach, yet 

it comes within a factor of 5 of the proposed REL based on human studies. 

VIII. References 

Aksoy M, Dincol K, Erdem S, Akgun T, and Dincol G. 1972. Details of blood changes in 32 

patients with pancytopenia associated with long-term exposure to benzene. Br. J. Ind. Med. 

29:56-64. 

Aoyama K. 1986. Effects of benzene inhalation on lymphocyte subpopulations and immune 

response in mice. Toxicol. Appl. Pharmacol. 85:92-101. 

Baarson KA, Snyder CA, and Albert RE. 1984. Repeated exposure of C57Bl/6 mice to inhaled 

benzene at 10 ppm markedly depressed erythrocyte colony formation. Toxicol. Lett. 20:337-342. 

Baslo A, and Aksoy M. 1982. Neurological abnormalities in chronic benzene poisoning. A study 

of six patients with aplastic anemia and two with preleukemia. Environ. Res. 27:457-465. 

CARB. 1999a. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Benzene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm 

Appendix D3 31 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CARB. 1999b. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Coate WB, Hoberman AM, and Durloo RS. 1984. Inhalation teratology study of benzene in rats. 

In: Advances in Modern Environmental Toxicology, Vol. VI. Applied Toxicology of Petroleum 

Hydrocarbons. MacFarland HN, ed. Princeton, NJ: Princeton Scientific Publishers, Inc. 

Cody RR, Strawderman WW, and Kipen HM. 1993. Hematologic effects of benzene. J. Occup. 

Med. 35(8):776-782. 

Collins JJ, Ireland BK, Easterday PA, Nair RS, Braun J. 1997. Evaluation of lymphopenia 

among workers with low-level benzene exposure and the utility of routine data collection. J. 

Occup. Environ. Med. 39(3):232-237. 

Cronkite EP, Inoue T, Carten AL, Miller ME, Bullis JE, and Drew RT. 1982. Effects of benzene 

inhalation on murine pluripotent stem cells. J. Toxicol. Environ. Health 9(3):411-421. 

Cronkite EP, Drew RT, Inoue T, and Bullis JE. 1985. Benzene hematotoxicity and 

leukemogenesis. Am. J. Ind. Med. 7:447-456. 

Cronkite EP, Drew RT, Inoue T, Hirabayashi Y and Bullis JE. 1989. Hematotoxicity and 

carcinogenicity of inhaled benzene. Environ. Health Perspect. 82:97-108. 

Crump K, and Allen B. 1984. Quantitative estimates of risk of leukemia from occupational 

exposure to benzene. Occupational Safety and Health Administration; Docket H-059B. [As cited 

in: Cody RR, Strawderman WW, and Kipen HM. 1993. Hematologic effects of benzene. J. 

Occup. Med. 35(8):776-782.] 

DeGowin RL. 1963. Benzene exposure and aplastic anemia followed by leukemia 15 years later. 

J. Am. Med. Assoc. 185(10):748-751. 

Deichmann WB, MacDonald WE, and Bernal E. 1963. The hematopoietic tissue toxicity of 

benzene vapors. Toxicol. Appl. Pharmacol. 5:201-224. 

Farris GM, Robinson SN, Gaido, KW, Wong BA, Wong VA, Hahn WP, and Shah RS. 1997. 

Benzene-induced hematoxicity and bone marrow compensation in B6C3F1 mice. Fundam. Appl. 

Toxicol. 36:119-129. 

HSDB. 1994. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(CD-ROM version). Micromedex, Inc., Denver, CO (edition expires 11/31/94). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Holmberg B, and Lundberg P. 1985. Benzene: Standards, occurrence, and exposure. Am. J. Ind. 

Med. 7:375-383. 

Appendix D3 32 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IARC (International Agency for Research on Cancer). 1982. IARC Monographs on the 

Evaluation of the Carcinogenic Risk of Chemicals to Humans: Some Industrial Chemicals and 

Dyestuffs. Volume 29. Lyon: IARC. pp. 95-148. 

Keller KA, and Snyder CA. 1986. Mice exposed in utero to low concentrations of benzene 

exhibit enduring changes in their colony forming hematopoietic cells. Toxicology 42:171-181. 

Keller KA, and Snyder CA. 1988. Mice exposed in utero to 20 ppm benzene exhibit altered 

numbers of recognizable hematopoietic cells up to seven weeks after exposure. Fundam. Appl. 

Toxicol. 10:224-232. 

Kipen HM, Cody RP, Crump KS, Allen BC, and Goldstein BD. 1988. Hematologic effects of 

benzene: A thirty-five year longitudinal study of rubber workers. Toxicol. Ind. Health. 4(4):411-

430. 

Kuna RA, and Kapp RW. 1981. The embryotoxic/teratogenic potential of benzene vapor in rats. 

Toxicol. Appl. Pharmacol. 57:1-7. 

Murray FJ, John JA, Rampy LW, Kuna RA, and Schwetz BA. 1979. Embryotoxicity of inhaled 

benzene in mice and rabbits. Am. Ind. Hyg. Assoc. J. 40:993-998. 

NTP, National Toxicology Program 1986. NTP Technical Report on the Toxicology and 

Carcinogenesis Studies of Benzene (CAS No. 71-43-2) in F344/N Rats and B6C3F1 Mice 

(Gavage Studies). U.S. Department of Health and Human Servcies, Public Health Service, 

National Institutes of Health NIH Publication No. 86-2545. 

Rothman N, Li GL, Dosemeci M, Bechtold WE, Marti GE, Wang YZ, Linet M, Xi LQ, Lu W, 

Smith MT, Titenko-Holland N, Zhang LP, Blot W, Yin SN, Hayes RB. 1996. Hematotoxocity 

among Chinese workers heavily exposed to benzene. Am. J. Ind. Med. 29(3):236-246. 

Rozen MG, Snyder CA, and Albert RE. 1984. Depressions in B- and T-lymphocyte mitogeninduced blastogenesis in mice exposed to low concentrations of benzene. Toxicol. Lett. 20:343-

349. 

Runion HE, and Scott LM. 1985. Benzene exposure in the United States 1978-1983: An 

overview. Am. J. Ind. Med. 7:385-393. 

Sabourin PJ, Bechtold WE, Birnbaum LS, Lucier G, and Henderson RF. 1988. Differences in the 

metabolism and disposition of inhaled [3H]benzene by F344/N rats and B6C3F1 mice. Toxicol. 

Appl. Pharmacol. 94:128-140. 

Sabourin PJ, Bechtold WE, Griffith WC, Birnbaum LS, Lucier G, and Henderson RF. 1989. 

Effect of exposure concentration, exposure rate, and route of administration on metabolism of 

benzene by F344 rats and B6C3F1 mice. Toxicol. Appl. Pharmacol. 99:421-444. 

Appendix D3 33 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Snyder CA, Goldstein BD, Sellakamur A, Wolman S, Bromberg I, Erlichman MN, and Laskin S. 

1978. Hematotoxicity of inhaled benzene to Sprague-Dawley rats and AKR mice at 300 ppm. J. 

Toxicol. Environ. Health 4:605-618. 

Tatrai E, Ungvary GY, Hudak A, Rodics K, Lorincz M, and Barcza GY. 1980. Concentration 

dependence of the embryotoxic effects of benzene inhalation in CFY rats. J. Hyg. Epidem. 

Micro. Immunol. 24(3):363-371. 

Tsai SP, Wen CP, Weiss NS, Wong O, McClellan WA, and Gibson RL. 1983. Retrospective 

mortality and medical surveillance studies of workers in benzene areas of refineries. J. Occup. 

Med. 25(9):685-692. 

Ward CO, Kuna RA, Snyder NK, Alsaker RD, Coate WB, and Craig PH. 1985. Subchronic 

inhalation toxicity of benzene in rats and mice. Am. J. Ind. Med. 7:457-473. 

Ward E, Hornburg R, Morris J, Rinsky R, Wild D, Halperin W, and Guthrie W. 1996. Risk of 

low red or white blood cell count related to estimated benzene exposure in a rubberworker cohort 

(1940-1975). Am. J. Ind. Med. 29:247-257. 

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Appendix D3 34 Benzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY

BERYLLIUM AND BERYLLIUM COMPOUNDS 

(beryllium-9; glucinium; glucinum; beryllium metallic) 

CAS Registry Number: 7440-41-7 

(beryllium oxide; beryllia; beryllium monoxide) 

CAS Registry Number: 1304-56-9 

(beryllium hydroxide; beryllium hydrate; beryllium dihydroxide)

CAS Registry Number: 13327-32-7 

(beryllium sulfate; sulfuric acid; beryllium salt)

CAS Registry Number: 13510-49-1 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.007 μg Be/m3

Critical effect(s) Beryllium sensitization and chronic beryllium 

disease in occupationally exposed humans 

Hazard index target(s) Respiratory system; immune system 

Oral reference exposure level 0.002 mg/kg-day 

Critical efect Small intestinal lesions in dogs 

Hazard index target(s) Gastrointestinal tract/liver 

II. Physical and Chemical Properties Summary (ATSDR, 1993) 

 

Metallic 

beryllium 

Beryllium 

oxide 

Beryllium 

hydroxide 

Beryllium sulfate 

Description Solid gray, 

hexagonal 

structure 

White light, 

amorphous 

powder 

White 

amorphous 

powder or 

crystalline 

Colorless tetragonal 

crystals 

Molecular formula Be BeO Be(OH)2 BeSO4

Molecular weight 9.012 g/mol 25.01 g/mol 43.03 g/mol 105.07 g/mol 

Solubility Insoluble in water ⎜Soluble 

Conversion factor Not applicable 

Appendix D3 35 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

III. Major Uses and Sources 

Beryllium is a metallic element mined as bertrandite and beryl mineral ores. As the lightest 

structural metal, beryllium is used in the space, aircraft, and nuclear industries in a variety of 

components including aircraft disc brakes, x-ray transmission windows, vehicle optics, nuclear 

reactor neutron reflectors, fuel containers, precision instruments, rocket propellants, navigational 

systems, heat shields, and mirrors. In addition to the four species listed, there are many other 

beryllium-containing compounds, including other salts, ores, and alloys (see, e.g., CRC, 1994). 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 2279 pounds of 

beryllium (CARB, 2000). 

Beryllium alloys, especially the hardest alloy beryllium copper, are used in electrical equipment, 

precision instruments, springs, valves, non-sparking tools, and in molds for injection-molded 

plastics for automotive, industrial, and consumer applications. Beryllium oxide is used in hightechnology ceramics, electronic heat sinks, electrical insulators, crucibles, thermocouple tubing, 

and laser structural components. Other beryllium compounds, including the chloride, nitrate, 

fluoride, and sulfate, are utilized as chemical reagents or generated from the refining of 

beryllium-containing ores. 

Beryllium is naturally emitted into the atmosphere by windblown dusts and volcanic particles. 

However, the major emission source is the combustion of coal and fuel oil, which releases 

beryllium-containing particulates and ash. Other beryllium-releasing industrial processes include 

ore processing, metal fabrication, beryllium oxide production, and municipal waste incineration 

(ATSDR, 1993). Beryllium also occurs in tobacco smoke (0-0.0005 μg/cigarette) (Smith et al., 

1997). 

IV. Effects of Human Exposure 

The respiratory tract is the major target organ system in humans following the inhalation of 

beryllium. The common symptoms of chronic beryllium disease (CBD) include shortness of 

breath upon exertion, weight loss, cough, fatigue, chest pain, anorexia, and overall weakness. 

Most studies reporting adverse respiratory effects in humans involve occupational exposure to 

beryllium. Exposure to soluble beryllium compounds is associated with acute beryllium 

pneumonitis (Eisenbud et al., 1948). Exposure to either soluble or insoluble beryllium 

compounds may result in obstructive and restrictive diseases of the lung, called chronic 

beryllium disease (berylliosis) (Cotes et al., 1983; Johnson, 1983; Infante et al., 1980; Kriebel et 

al., 1988a; Metzner and Lieben, 1961). The total number of beryllium-related disease cases has 

declined since the adoption of industrial standards (Eisenbud and Lisson, 1983; ATSDR, 1993). 

Historically, beryllium pneumonitis has been associated with occupational concentrations over 

0.1 mg Be/m3

, primarily as beryllium sulfate or beryllium fluoride (Eisenbud et al., 1948). The 

atmospheric concentrations related to chronic beryllium disease have been more difficult to 

define, in part due to the lack of individual exposure estimates, especially in the studies derived 

Appendix D3 36 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

from the berylliosis case registries (Infante et al., 1980; Lieben and Metzner, 1959). However, 

Infante and associates (1980) reported significantly increased mortality due to non-neoplastic 

respiratory disease in beryllium-exposed workers, and noted one case of chronic berylliosis in a 

worker following seven years exposure to < 2 μg Be/m3. In a 30-year follow-up study of 146 

beryllium-exposed workers, Cotes et al. (1983) identified seven cases of chronic beryllium 

related disease. All the cases were exposed to beryllium oxide or hydroxide, but in a wide range 

of retrospectively estimated doses (over 3000 samples from 1952 to 1960). The estimated 

average daily exposure did not exceed 2 μg/m3

 for the ten site/process classifications, but 318 

samples did exceed 2 μg Be/m3

 (and 20 samples were greater than 25 μg Be/m3). No 

atmospheric samples were available after 1963, even though the exposure occurred through 

1973. The LOAEL for occupationally induced berylliosis observed in this study was estimated 

from uncertain exposure data to be less than 2 μg Be/m3

. 

One cross-sectional study (Kriebel et al., 1988a; Kriebel et al., 1988b) estimated beryllium 

exposure levels for 309 workers originally surveyed in 1977, with a median duration of exposure 

of 17 years (range 2 to 39 years). Historic plant levels were estimated to be as high as 100 μg 

Be/m3, and, even as late as 1975, some job classifications exceeded 10 μg Be/m3. The workers’ 

median cumulative exposure was 65 μg Be/m3

-year (range 0.1 to 4400 μg Be/m3-years); the 

median lifetime exposure estimate was 4.3 μg/m3 (range 0.01 to 150 μg/m3). Spirometric 

measurement of pulmonary function, chest x-rays, and arterial blood gas measurements were 

collected. Decrements in lung function, as defined by forced vital capacity (FVC) and forced 

expiratory volume in one second (FEV1), were associated with cumulative exposure up to 20 

years prior to the health survey, even in workers with no radiographic abnormalities. Differences 

in alveolar-arterial oxygen gradient were associated with cumulative exposure in the 10 years 

prior to the study. These endpoints give a LOAEL of 39 μg/m3-years (geometric mean 

cumulative exposure) for decrements in pulmonary function and changes in arterial blood gases. 

Non-occupational beryllium-related chronic disease has been reported in individuals residing in 

the vicinity of beryllium manufacturing industries (Eisenbud et al., 1949; Metzner and Lieben, 

1961). An early cross-sectional study (Eisenbud et al., 1949) described 11 cases of nonoccupational berylliosis after x-ray and clinical examination of approximately 10,000 residents 

near a beryllium fabrication facility in Lorain, Ohio. Ten of the cases resided within 3/4 mile of 

the plant (up to 7 years duration), and five cases resided within 1/4 mile. The authors estimated a 

1% disease incidence within 1/4 mile (500 individuals). Atmospheric sampling in 1947 

identified an average level of 0.2 μg Be/m3

 at 1/4 mile decreasing to 0 μg Be/m3 at 10 miles, but 

samples varied up to 100 fold over the 10 week sampling period. Utilizing current and historical 

exposure estimates based on discharge, process, inventory, and building design changes, this 

study estimated a chronic LOAEL in the range of 0.01 to 0.1 μg Be/m3

 for continuous exposure 

to beryllium compounds, based on the development of chronic berylliosis. 

Metzner and Lieben (1961) also reported 26 cases of chronic berylliosis in a population of 

approximately 100,000, living within 7 miles of a refining and alloy fabrication plant (duration 6 

to 19 years). Neighborhood exposure assessment conducted over 14 months during 1958 and 

1959 identified a mean level of 0.0155 μg Be/m3

, with 10% of the samples registering over 0.03 

μg Be/m3

. Limited measurements conducted earlier at the site were higher (1.0 to 1.8 μg Be/m3

in 1953 and 0.91 to 1.4 μg Be/m3 in 1954). 

Appendix D3 37 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Chronic beryllium disease appears to involve a cell-mediated immune response, especially 

granulomatous reactions found in the lungs of sensitive individuals. Humans exposed to 

beryllium compounds have demonstrated increased T-cell activity (in vitro) and histological 

abnormalities of the lymph nodes (Cullen et al., 1987; Johnson, 1983). Johnson (1983) 

described granuloma of lymph nodes and chronic interstitial pneumonitis in a small number of 

beryllium metal handling machinists (LOAEL = 4.6 μg Be/m3

). A second study identified 

granulomatous lung lesions, scarred lung tissue, and breathing difficulties in workers from a 

precious metal refining facility exposed to a mixture of beryllium and other metals (Cullen et al., 

1987). Also, altered proliferative responses of lymphocytes obtained by bronchoalveolar lavage 

indicated increased T-cell activity in vitro. Cullen et al. (1987) reported a mean exposure level 

of 1.2 μg Be/m3

 (range = 0.22 – 43 μg/m3). USEPA (1998) and ATSDR (2000) considered 

0.52 μg Be/m3 to be the LOAEL for CBD from this study since this was the average 

concentration in the furnace area where 4 of the 5 CBD cases worked. 

Sensitization to beryllium, as measured by the beryllium lymphocyte proliferation test (BeLPT), 

can occur in the absence of chronic beryllium disease (Kreiss et al., 1989). The authors hoped 

that the identification of sensitized individuals without disease might prevent clinical disease, 

presumably by removing the individuals from exposure to beryllium. Some beryllium-sensitized 

individuals progress to having clinical disease (Newman et al., 1992). Data obtained from a 

four-year survey conducted at beryllium-copper alloy manufacturing factories in Japan (Yoshida 

et al., 1997) indicated that the T cells of workers continuously exposed to more than 0.01 μg 

Be/m3 were activated and that the cell-mediated immune (CMI) response was promoted. The 

BeLPT in workers exposed to less than 0.01 μg Be/m3 was unaffected. 

Genetic influences on development of CBD have been identified. CBD is associated with the 

allelic substitution of glutamic acid for lysine at position 69 in the HLA-DPB1 protein (Richieldi 

et al., 1993). Up to 97% of CBD patients may have the Glu69 marker, but only 30-45% of 

beryllium-exposed, unaffected individuals carry the same marker. Because CBD occurs in only 

1-6% of exposed workers, Glu69 is not likely to be the only genetic factor influencing the 

development of CBD. Changes in other sequences of the HLA-DPB1 gene and in the copy 

number of Glu69 are also involved (Wang et al., 1999). 

The Rocky Flats Environmental Technology Site in Colorado is part of the U.S. Department of 

Energy nuclear weapons complex. Operations using Be began in 1953, Be production operations 

began in 1957, and the first case of CBD was diagnosed in a machinist in 1984. Exposures could 

have occurred during foundry operations, casting, shearing, rolling, cutting, welding, machining, 

sanding, polishing, assembly, and chemical analysis operations. Since 1991, 29 cases of CBD 

and 76-78 cases of beryllium sensitization have been identified (Stange et al., 1996). Several 

cases appear to have had only minimal Be exposure, since the employees were in administrative 

functions, not primary beryllium operations. Personal air monitoring devices used over a period 

of 4 years showed a breathing zone level of 1.04 μg Be/m3. ATSDR (2000) considered 1.04 μg 

Be/m3

 to be the LOAEL for this study. A recent case-control study of workers at Rocky Flats 

(Viet et al., 2000) suggested that exposures of workers to lower Be levels might lower the future 

incidence of CBD, but not necessarily the incidence of sensitivity to Be. 

Appendix D3 38 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Kreiss et al. (1996) investigated the prevalence of beryllium sensitization in relation to work 

process and beryllium exposure measurements in a beryllia ceramics plant that had operated 

since 1980. In 1992 they interviewed 136 employees (97.8% of the workforce), ascertained 

beryllium sensitization with the beryllium lymphocyte proliferation blood test (BeLPT), and 

reviewed industrial hygiene measurements. Eight employees were beryllium-sensitized (5.9%); 

six of the eight had granulomatous disease based on transbronchial lung biopsy. Machinists had 

a Be sensitization rate of 14.3% compared to 1.2% among other employees. Machining 

operations (drilling, dicing, centerless grinding, and/or surface grinding) had significantly higher 

general area and breathing zone measurements than other work processes during the time in 

which most beryllium-sensitized cases had started machining. Daily weighted average estimates 

of exposure for matching processes also exceeded estimates for other work processes in that time 

period (median daily weighted average = 0.9 μg/m3

). Daily weighted averages for the machining 

process accounted for the majority of exceedances of the 2.0 μg/m3 OSHA Permissible Exposure 

Limit (PEL); 8.1% of machining daily weighted averages were above the PEL. The LOAEL 

from this study was 0.55 μg/m3, the median exposure of the sensitized workers. 

The facility was again surveyed in 1998 after some attempts were made to lower exposure to 

beryllium (Henneberger et al., 2001). The investigators separated the workers into 77 long-term 

workers hired before the 1992 screening and 74 short-term workers hired after 1992. Among 20 

short-term workers exposed to the lowest mean Be level (0.05 to 0.19 μg/m3), two showed Be 

sensitivity by the BeLPT test. Thus a fraction of workers appears to be exquisitely sensitive to 

beryllium. 

Based on a review of this and other occupational studies Wambach and Tuggle (2000) have 

suggested that the workplace standard of 2 μg/m3

 be lowered to 0.1 μg/m3. Some workers might 

still be sensitized to beryllium at this level (Yoshida et al., 1997). 

V. Effects of Animal Exposure 

Three chronic studies, two in rats (Vorwald and Reeves, 1959; Reeves et al., 1967) and one in 

guinea pigs (Reeves et al., 1970), observed adverse inflammatory and proliferative respiratory 

changes following inhalation exposure to beryllium compounds. Vorwald and Reeves (1959) 

observed inflamed lungs and fibrosis in rats exposed to 0.006 mg Be/m3

 (as BeO) for an 

unspecified duration. A later study exposed Sprague-Dawley CD rats for 72 weeks (7 hr/d, 5 

d/wk) to 34.25 μg Be/m3 from BeSO4 (Reeves et al., 1967). Gross and histological changes 

observed in exposed versus unexposed rats included increased lung weight, inflamed lungs, 

emphysema, arteriolar wall thickening, granulomas, fibrosis, and proliferative responses within 

the alveoli (LOAEL = 34.25 μg Be/m3). Guinea pigs were exposed to 0, 3.7, 15.4, or 29.3 μg 

Be/m3

 (from the sulfate) for 6 hours/day, 5 days/week for up to 1 year (Reeves et al., 1970). 

Respiratory alterations observed in the beryllium-exposed groups included increased 

tracheobronchial lymph node and lung wet weights, interstitial pneumonitis, and granulomatous 

lesions. These adverse respiratory effects were observed in all the beryllium dosed groups and 

indicated a chronic inhalation LOAEL of 3.7 μg Be/m3

. 

Appendix D3 39 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Wagner et al. (1969) exposed monkeys, rats, and hamsters to 0.21 and 0.62 mg Be/m3

 as fumes 

from bertrandite or beryl ore, respectively, for 6 hours/day, 5 days/week for up to 17 months. 

Exposed animals displayed severe effects, including (1) bronchial lymphocytic infiltrates, 

abscesses, consolidated lobes, and granulomatous lesions after exposure to 0.21 mg Be/m3

 from 

bertrandite ore, and (2) inflamed lungs, fibrosis, and granuloma after exposure to 0.62 mg Be/m3

from beryl ore. Lung inflammation was observed in the exposed monkeys, and a few 

granulomatous lung lesions were observed in the hamsters after similar exposure conditions (up 

to 23 months). 

Immunological effects have been observed in a few subchronic studies (Schepers, 1964; 

Schepers et al., 1957; Stiefel et al., 1980). Schepers (1964) exposed monkeys (Macacus 

mullata) to three soluble forms of beryllium (BeF2, BeSO4, BeHPO4) daily for 6 hours/day over 

7 to 30 days. Increased lung weight, inflammation, emphysema, and fibrosis of the lung were 

observed after 17 days at 0.198 mg Be/m3

 (as BeSO4). Histological examination found pleuritis, 

congestion, emphysema, consolidation, and edema of the lung. Immunological effects were seen 

as hyperplasia of the lymph nodes typical of immune activation after 7 to 18 days exposure to 

either 0.198 or 0.184 mg Be/m3

 as the sulfate or fluoride. A subchronic inhalation study reported 

immunological effects as increased, beryllium-specific stimulation of T-lymphocytes in vitro

from Wistar rats and guinea pigs exposed daily (6 hours/day) over 10 weeks (LOAEL = 0.5 

mg/m3) (Stiefel et al., 1980). However, a subchronic inhalation study in Wistar and Sherman 

rats (Schepers et al., 1957) observed multiple lung alterations including granulomas (LOAEL = 

35 μg Be/m3) but did not find any accompanying immunological effects after 30 days 

discontinuous exposure (5-6 d/wk, 4-8 hr/d) to beryllium fumes from BeSO4. 

Appendix D3 40 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VI. Derivation of Chronic Reference Exposure Levels 

Derivation of Inhalation Reference Exposure Level 

Key study Kreiss et al., 1996 

Study population 8 beryllium-sensitized workers among 136 employees 

in a beryllia ceramics plant 

Exposure method Workplace 

Critical effects Beryllium sensitization (chronic beryllium disease) 

LOAEL 0.55 μg/m3

 (median exposure of sensitized workers) 

NOAEL Not observed 

Exposure continuity Workplace 

Average experimental exposure 0.2 μg/m3

 for LOAEL group (0.55 x 10/20 x 5/7) 

Human equivalent concentration 0.2 μg/m3

Exposure duration 6.1 years (5 mo – 10 yr) 

LOAEL uncertainty factor 10 (low incidence but serious, irreversible chronic 

disease) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 (sensitized may not be only sensitive subpopulation) 

(see below) 

Cumulative uncertainty factor 30 

Inhalation chronic REL 0.007 μg/m3

Supportive study Eisenbud et al. (1949) 

Study population Approximately 10,000 individuals within 2 miles of a 

beryllium manufacturing plant 

Exposure method Environmental exposure 

Critical effects Pulmonary berylliosis in 11 residents 

LOAEL 0.03 μg/m3

 (geometric mean of range of measured 

exposures associated with berylliosis of 0.01 to 

0.1 μg/m3

) 

NOAEL Not observed 

Exposure continuity Continuous 

Average exposure Estimated to be approximately 0.3 μg/m3

 (historical 

exposures estimated to be 10-fold higher than 

measured values) for LOAEL group 

Human equivalent 

concentration 

0.3 μg/m3 for LOAEL group 

Exposure duration Up to 7 years 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 100 

Inhalation chronic REL 0.003 μg/m3 

Appendix D3 41 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

U.S. EPA (1998) developed an RfC of 0.02 μg/m3

 based on beryllium sensitization and 

progression to chronic beryllium disease (CBD) identified by Kreiss et al. (1996). The Kreiss et 

al. (1996) occupational exposure study identified a LOAEL for beryllium sensitization in 

workers of 0.55 μg/m3

 (median of average exposure concentrations of the 8 Be sensitized 

workers). The Eisenbud et al. (1949) study, which U.S. EPA used as a co-principal study and 

which in U.S. EPA’s opinion used relatively insensitive screening methods, suggested a NOAEL 

of 0.01-0.1 μg/m3

 in community residents living near a beryllium plant. U.S. EPA used the 

LOAEL from the Kreiss et al. (1996) study for the operational derivation of the RfC, because the 

screening method used in the Eisenbud et al. (1949) study was considered to be less sensitive 

than the method used in the Kreiss et al. (1996) study. The LOAEL was time adjusted to 0.2 

μg/m3

, then a total UF of 10 was used to obtain the RfC of 0.02 μg/m3. The UF of 10 was 

comprised of a UF of 3 to account for the sensitive nature of the subclinical endpoint (beryllium 

sensitization) and a database UF of 3 to account for the poor quality of exposure monitoring in 

the Kreiss et al. and Eisenbud et al. studies. Poor exposure monitoring was also a problem in 

other epidemiology studies that assessed the incidence of beryllium sensitization. The U.S. EPA 

did not explicitly apply a LOAEL to NOAEL uncertainty factor. Thus implicitly the factor is 1. 

OEHHA prefers to use the methodology for assignment of UFs, which is described in OEHHA 

(2000) and used in our derivation of the REL for beryllium, including use of a LOAEL to 

NOAEL Uncertainty Factor of 10. Since chronic beryllium disease (CBD) is serious, chronic, 

disabling, usually irreversible, and often fatal (Newman et al., 1997), it is difficult to justify use 

of a LOAEL to NOAEL factor of only 3. OEHHA has not used database deficiency UFs since 

the criteria for use of such factors are not well specified by U.S. EPA. The people who get CBD 

are likely that part of the population who are by nature more sensitive to beryllium, for example 

those with the human leukocyte antigen (HLA) class II marker HLA-DP Glu69 (Richeldi et al., 

1993; Saltini et al., 1998). Although it is likely that the effects are seen in a "sensitive 

subpopulation," OEHHA applied an intraspecies uncertainty factor (UFH). OEHHA used an 

intermediate UFH of 3, since 1) there may be other population factors involved in being sensitive, 

such as immature lungs, and 2) all the diseased were initially healthy adult workers. 

For comparison the LOAEL from guinea pigs of 3.7 μg Be/m3

 (Reeves et al., 1970) is equivalent 

to a continuous exposure of 0.66 μg/m3. Division by UFs of 10 for intraspecies, 10 for 

interspecies (since HEC adjustments are not available yet for guinea pigs), and 10 for use of a 

LOAEL results in a REL of 0.0007 μg/m3

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the inhalation chronic REL for beryllium is the use of human data from 

persons occupationally exposed. The major uncertainties are the lack of a NOAEL observation 

in the key study, the lack of long-term exposure data, the difficulty of estimating exposures, and 

the lack of chronic exposure data. 

Appendix D3 42 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VIII. Potential for Differential Impacts on Children's Health

No evidence to support a differential effect of beryllium on infants or children was found in the 

literature. However, children have developed beryllium disease from metal brought home on the 

parents' work clothes and by living near a facility using beryllium. Unfortunately the number of 

children and their ages were not published (Eisenbud et al., 1948). 

Derivation of Chronic Oral Reference Exposure Level 

In addition to being inhaled, airborne beryllium can settle onto crops and soil and enter the body 

by ingestion. Thus an oral chronic reference exposure level for beryllium is also required for 

conducting Air Toxics Hot Spots risk assessments. 

Study Morgareidge et al., 1976 

Study population Male and female dogs (5/sex/group) 

Exposure method Diet containing 0, 1, 5, 50 or 500 ppm Be as 

beryllium sulfate tetrahydrate 

Critical effects Small intestinal lesions 

LOAEL 500 ppm 

NOAEL 50 ppm (1.2 mg/kg bw-day) 

Exposure continuity Continuous 

Exposure duration Up to 3 years, 4 months 

Average experimental exposure 1.2 mg/kg bw-day (males, 1.1; females, 1.3) 

BMD05 0.244 mg/kg-day 

LOAEL uncertainty factor Not needed in BMD approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 10 

Intraspecies factor 10

Cumulative uncertainty factor 100 

Oral reference exposure level 0.002 mg/kg-day 

Morgareidge et al. (1976) conducted a long-term feeding study in which beagle dogs (aged 8 to 

12 mo) were fed diets (for 1 h per day) containing 0, 5, 50, or 500 ppm Be for 172 weeks. The 

500 ppm group was terminated at 33 weeks because of overt signs of toxicity, and an additional 

group was added to the study and fed a diet containing 1 ppm Be (for 143 weeks). The 1, 5, 50, 

and 500 ppm concentrations corresponded to doses of 0.023, 0.12, 1.1, and 12.2 mg/kg-day for 

males and 0.029, 0.15, 1.3, and 17.4 mg/kg-day for females. All animals in the 500 ppm group 

showed fairly extensive erosive (ulcerative) and inflammatory lesions in the gastrointestinal 

tract. These occurred predominantly in the small intestine and to a lesser extent in the stomach 

and large intestine, and were considered treatment related. All animals with stomach or large 

intestinal lesions also had lesions in the small intestine, except for one animal (whose stomach 

lesions were very localized and not very severe). Lesions in the small intestine (4/5 males and 

5/5 females) were considered to be treatment-related and included desquamation of the 

epithelium, edema, fibrin thrombi, acute inflammation, subacute/chronic inflammation, necrosis 

and thinning/atrophy of the epithelium, and ulceration. High-dose animals also showed 

Appendix D3 43 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

moderate to marked erythroid hypoplasia of the bone marrow, which the authors also considered 

treatment related. (Bile stasis and vasculitis in the liver, acute inflammation in the lymph nodes, 

and kidney occurring in these animals was attributed to a likely systemic bacterial invasion 

through the damaged intestinal mucosa.) In the 50 ppm group, one female dog, which died after 

70 weeks of treatment, showed gastrointestinal lesions, which were less severe, but occurred in 

the same locations and appeared to be the same types of lesions as those in dogs administered 

500 ppm. The observation that beryllium is poorly absorbed by the gastrointestinal tract (Owen, 

1990; ATSDR, 2000) probably explains why lesions were not seen outside the gastrointestinal 

tract. In addition the predominance of lesions in the small intestine may have been partly due to 

precipitation of beryllium phosphate there due to the slightly alkaline pH (Reeves, 1965). Thus 

500 ppm was a LOAEL and 50 ppm was a NOAEL (statistically) for gastrointestinal lesions. 

USEPA used the same study to derive its RfD of 0.002 mg/kg-day. The U.S. EPA stated its 

confidence in the RfD as: study - medium; database – low to medium, and RfD - low to medium. 

USEPA used a BD10 approach and included a database UF of 3. OEHHA used a BD05 approach 

(specifically a Weibull model in the USEPA's BMDS software) and did not include a database 

UF since the criteria for use of modifying factors such as this are not well specified by U.S. EPA. 

However, the final value for the oral chronic REL was the same as the USEPA's RfD. 

This RfD and the oral REL are limited to soluble beryllium salts. Data on the teratogenicity or 

reproductive effects of beryllium are limited. Beryllium has been reported to produce terata and 

increased mortality in chick embryos. 

When assessing the health effects of beryllium, its carcinogenicity must also be assessed. 

IX. References 

ATSDR. 1993. Agency for Toxic Substances and Disease Registry. U.S. Department of Health 

and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry. 

Toxicological Profile for Beryllium. TP-92/04. Atlanta, GA: ATSDR. April 1993. 

ATSDR. 2000. Agency for Toxic Substances and Disease Registry. U.S. Department of Health 

and Human Services, Public Health Service, Agency for Toxic Substances and Disease Registry. 

Toxicological Profile for Beryllium (update). Dreaft for Public Comment. Atlanta, GA: ATSDR. 

September 2000. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Cotes JE, Gilson JC, McKerrow CB, and Oldham P. 1983. A long-term follow-up of workers 

exposed to beryllium. Br. J. Ind. Med. 40:13-21. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Appendix D3 44 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Cullen M, Kominsky J, Rossman M, Cherniack M, Rankin J, Balmes J, Kern J, Daniele R, 

Palmer L, Naegel G, McManus K, and Cruz R. 1987. Chronic beryllium disease in a precious 

metal refinery. Clinical epidemiologic and immunologic evidence for continuing risk from 

exposure to low level beryllium fume. Am. Rev. Respir. Dis. 135:201-208. 

Eisenbud M, and Lisson J. 1983. Epidemiological aspects of beryllium-induced nonmalignant 

lung disease: A 30 year update. J. Occup. Med. 25:196-202. 

Eisenbud M, Berghout CF, and Steadman L. 1948. Environmental studies in plants and 

laboratories using beryllium: the acute disease. J. Ind. Hyg. Toxicol. 30:282-285. 

Eisenbud M, Wanta R, Dustan C, Steadman L, Harris W, and Wolf B. 1949. Nonoccupational 

berylliosis. J. Ind. Hyg. Toxicol. 31:282-294. 

Henneberger PK, Cumro D, Deubner DD, Kent MS, McCawley M and Kreiss K. 2001. 

Beryllium sensitization and disease among long-term and short-term workers in a beryllium 

ceramics plant. Int. Arch. Occup. Environ. Health. 74(3):167-176. 

Infante PF, Wagoner JK, and Sprince NL. 1980. Mortality patterns from lung cancer and 

nonneoplastic respiratory disease among white males in the beryllium case registry. Environ. 

Res. 21:35-43. 

Johnson NR. 1983. Beryllium disease among workers in a spacecraft-manufacturing plant - 

California. MMWR. 32:419-425. 

Kreiss K, Newman LS, Mroz MM, and Campbell PA. 1989. Screening blood test identifies 

subclinical beryllium disease. J. Occup. Med. 31(7):603-608. 

Kreiss K, Mroz MM, Newman LS, Martyny J, and Zhen B. 1996. Machining risk of beryllium 

disease and sensitization with median exposures below 2 micrograms/m3

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30(1):16-25. 

Kriebel D, Sprince NL, Eisen EA, Greaves I, Feldman H and Greene R. 1988a. Beryllium 

exposure and pulmonary function: A cross sectional study of beryllium workers. Br. J. Ind. Med. 

45:167-173. 

Kriebel D, Sprince NL, Eisen EA, and Greaves I. 1988b. Pulmonary function in beryllium 

workers: Assessment of exposure. Br. J. Ind. Med. 45:83-92. 

Lieben J, and Metzner F. 1959. Epidemiological findings associated with beryllium extraction. 

Am. Ind. Hyg. Assoc. 20:494-499. 

Metzner F, and Lieben J. 1961. Respiratory disease associated with beryllium disease - A case 

study. J. Occup. Med. 3:341-345. 

Appendix D3 45 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Morgareidge K, Cox GE and Gallo MA. (1976) Chronic feeding studies with beryllium in dogs. 

Food and Drug Research Laboratories, Inc. Submitted to the Aluminum Company of America, 

Alcan Research & Development, Ltd., Kawecki-Berylco Industries, Inc., and Brush-Wellman, 

Inc. 

Newman LS, Mroz MM, Schumacher B, Daniloff E, and Kreiss K. 1992. Beryllium sensitization 

precedes chronic beryllium disease. Am. Rev. Resp. Dis. (Suppl) 145:A324. 

Newman LS, Lloyd J, and Daniloff E. 1996. The natural history of beryllium sensitization and 

chronic beryllium disease. Environ. Health Perspect. 104S(5):937-943. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov

Owen BA. 1990. Literature-derived absorption coefficients for 39 chemicals via oral and 

inhalation routes of exposure. Regul. Toxicol. Pharmacol. 11(3):237-252. 

Reeves AL. 1965. The absorption of beryllium from the gastrointestinal tract. Arch. Environ. 

Health. 11(2):209-214. 

Reeves AL, Deitch D, and Vorwald AJ. 1967. Beryllium carcinogenesis. I. Inhalation exposure 

of rats to beryllium sulfate aerosol. Cancer Res. 27:439-443. 

Reeves AL, Swanborg RH, Busby EK, and Krivanek ND. 1970. In: Target Organ Toxicology 

Series: Immunotoxicology and Immunopharmacology. Dean J, Luster M, Munson A, and Amos 

H, eds. New York, NY: Raven Press, Inc. pp. 441-456. 

Richeldi L, Sorrentino R, and Saltini C. 1993. HLA-DPB1 glutamate 69: a genetic marker of 

beryllium disease. Science 262(5131):242-244. 

Saltini C, Amicosante M, Franchi A, Lombardi G and Richeldi L. 1998. Immunogenetic basis of 

environmental lung disease: lessons from the berylliosis model. Eur Respir J. 12(6):1463-75. 

Schepers GW. 1964. The biological action of beryllium: Reaction of the monkey to inhaled 

aerosols. Ind. Med. Surg. 33:1-16. 

Schepers GW, Durkhan TM, Delahunt AB, and Creedon FT. 1957. The biological action of 

inhaled beryllium sulfate. A preliminary chronic toxicity study in rats. AMA Arch. Ind. Health 

15:32-58. 

Schroeder HA, and Mitchner M. 1975. Life-term studies in rats: Effects of aluminum, barium, 

beryllium and tungsten. J. Nutr. 105:421-427. 

Appendix D3 46 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Smith CJ, Livingston SD and Doolittle DJ. 1997. An international literature survey of "IARC 

Group I carcinogens" reported in mainstream cigarette smoke. Food Chem. Toxicol. 35(10-

11):1107-1130. 

Stange AW, Hilmas DE, and Furman FJ. 1996. Possible health risks from low level exposure to 

beryllium. Toxicology 111(1-3):213-224. 

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beryllium. Arch. Toxicol. 45:81-92. 

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Database. Available online at http://www.epa.gov/iris 

Viet SM, Torma-Krajewski J, and Rogers J. 2000. Chronic beryllium disease and beryllium 

sensitization at Rocky Flats: a case-control study. AIHAJ. 61(2):244-254. 

Vorwald A, and Reeves A. 1959. Pathologic changes induced by beryllium compounds. Arch. 

Indust. Health 19:190-199. 

Wagner W, Groth D, Holtz J, Madden G, and Stokinger H. 1969. Comparative chronic inhalation 

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Wambach PF, and Tuggle RM. 2000. Development of an eight-hour occupational exposure limit 

for beryllium. Appl. Occup. Environ. Hyg. 15(7):581-587 

Wang Z, White PS, Petrovic M, Tatum OL, Newman LS, Maier LA, and Marrone BL. 1999. 

Differential susceptibilities to chronic beryllium disease contributed by different Glu69 HLADPB1 and -DPA1 alleles. J. Immunol. 163(3):1647-1653. 

Yoshida T, Shima S, Nagaoka K, Taniwaki H, Wada A, Kurita H, and Morita K. 1997. A study 

on the beryllium lymphocyte transformation test and the beryllium levels in working 

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Appendix D3 47 Beryllium and Beryllium Compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY

1,3-BUTADIENE 

(butadiene; buta-1,3-diene; biethylene; bivinyl; divinyl; vinylethylene) 

CAS Registry Number: 106-99-0 

I. Chronic Toxicity Summary

Inhalation reference exposure level 20 μg/m3

 (8 ppb)

Critical effect(s) Increased incidence of ovarian atrophy in mice 

Hazard index target(s) Female reproductive system 

II. Physical and Chemical Properties Summary (HSDB, 2000; CRC, 1995) 

Description Colorless gas 

Molecular formula C4H6

Molecular weight 54.09 g/mol 

Boiling point −4.4o

C 

Melting point −108.9o

C 

Vapor pressure 910 torr at 20o

C 

Solubility Very slightly soluble in water (735 mg/L); 

soluble in ethanol, ether, acetone, benzene and 

organic solvents 

Conversion factor 1 ppm = 2.21 mg/m3

 at 25oC 

III. Major Uses and Sources

1,3-Butadiene is a major commodity product of the petrochemical industry, usually produced as 

a by-product of ethylene. The majority of 1,3-butadiene is used in the production of styrenebutadiene rubber copolymers (SBR). Other applications include use as a polymer component for 

polybutadiene, hexamethylene diamine, styrene-butadiene latex, acrylonitrile-butadiene-styrene 

(ABS) resins, chloroprene and nitrile rubbers. A variety of industrial syntheses use 1,3-

butadiene resins (AB as a chemical intermediate, such as in the production of adiponitrile (a 

nylon precursor), captan and captofol fungicides, ethylidene norbornene and sulfolane, boron 

alkyls, and hexachlorobutadiene. Additionally, 1,3-butadiene is found in automobile exhaust, 

gasoline vapor, fossil fuel incineration products, and cigarette smoke (HSDB, 2000). In 1996, 

the latest year tabulated, the statewide mean outdoor monitored concentration of 1,3-butadiene 

was approximately 0.2 ppb (CARB, 1999). The South Coast Air Quality Management District 

(SCAQMD, 2000) detected ambient levels of 1,3-butadiene ranging from 0.1 to 0.8 ppb at 10 

stationary monitors placed throughout the South Coast Air Basin. The annual statewide 

industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in California 

Appendix D3 48 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

based on the most recent inventory were estimated to be 20,846 pounds of 1,3-butadiene (CARB, 

2000). 

IV. Effects of Human Exposure

An early occupational study reported complaints of irritation of eyes, nasal passages, throat, and 

lungs in rubber manufacturing workers following acute exposure to unknown levels of 1,3-

butadiene (Wilson, 1944). Additional symptoms reported included coughing, fatigue, and 

drowsiness; however, all symptoms ceased on removal from the exposure. 

Studies on the chronic effects of 1,3-butadiene have been centered in the styrene-butadiene 

rubber manufacturing industry, which uses large quantities of 1,3-butadiene, and in the 1,3-

butadiene monomer industry. One retrospective epidemiological study reported an increase in 

overall mortality, emphysema, and cardiovascular diseases (chronic rheumatic and 

arteriosclerotic heart disease) among rubber workers (McMichael et al., 1976). Two other 

occupational studies (Divine and Hartman, 1996; Matanoski et al., 1990) indicated that the 

standardized mortality ratio for deaths from arteriosclerotic heart disease was elevated (∼1.4-1.8) 

among black workers in the 1,3-butadiene rubber industry. Other occupational studies have 

described the potential for adverse hematological effects due to butadiene exposure (Checkoway 

and Williams, 1982; McMichael et al., 1975). A survey of workers at a styrene-butadiene rubber 

plant revealed slightly lower levels (but within normal range) of red blood cells, hemoglobin, 

platelets, and neutrophils in exposed (mean = 20 ppm) versus unexposed workers (Checkoway 

and Williams, 1982). And 1,3-butadiene has been implicated in hematopoietic malignancies 

among styrene-butadiene rubber workers at levels lower than 20 ppm (McMichael et al., 1975). 

Since the workers in these studies were exposed to mixtures of chemicals, the specific 

contribution of butadiene to the adverse respiratory and hematopoietic effects remains unclear. 

V. Effects of Animal Exposure

The few available chronic animal inhalation studies have focused on the potential 

carcinogenicity of 1,3-butadiene. The National Toxicology Program (NTP) has sponsored two 

chronic inhalation studies in B6C3F1 mice (NTP, 1984; Melnick et al., 1990; NTP, 1993), while 

Hazelton Laboratories Europe (HLE) Ltd. conducted a chronic inhalation study in SpragueDawley rats (HLE, 1981; Owen et al., 1987; Owen and Glaister, 1990). 

The two B6C3F1 mice inhalation studies sponsored by NTP (Huff et al., 1985; Melnick et al., 

1990; NTP, 1984; NTP, 1993), although focused on carcinogenicity, identified other adverse 

chronic effects. The earlier NTP (1984) study in mice administered 0, 625 or 1250 ppm 1,3-

butadiene for 6 hours/day, 5 days/week for up to 61 weeks. Nonneoplastic changes observed 

were elevated testicular and ovarian atrophy at both doses (625 and 1250 ppm); liver necrosis in 

male mice at both doses and in female mice at 1250 ppm; and nonneoplastic lesions in the nasal 

cavity at 1250 ppm. At the highest dose, adverse changes in the nasal cavity included chronic 

inflammation, fibrosis, cartilaginous metaplasia, osseous metaplasia, and atrophy of the sensory 

epithelium. No nasal or respiratory lesions were seen in the controls. This study identified a 

chronic LOAEL of 625 ppm for gonadal atrophy in both sexes. 

Appendix D3 49 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

The later NTP study (Melnick et al., 1990; NTP, 1993) used lower exposure concentrations of 

1,3-butadiene (0, 6.25, 20, 62.5, 200 or 625 ppm) administered 6 hours/day, 5 days/week for up 

to 2 years. Two-year survival was significantly decreased in mice exposed to 20 ppm and 

greater, primarily due to chemical-related malignant neoplasms. Increased incidences of nonneoplastic lesions in exposed mice included bone marrow atrophy, gonadal atrophy (testicular, 

ovarian and uterine), angiectasis, alveolar epithelial hyperplasia, forestomach epithelial 

hyperplasia, and cardiac endothelial hyperplasia. Gonadal atrophy was observed at 200 ppm and 

625 ppm for males and at 6.25 ppm and higher for females. Bone marrow toxicity (regenerative 

anemia) was seen at 62.5 ppm and higher. This study identified a chronic LOAEL of 6.25 ppm 

for reproductive toxicity, and a NOAEL of 200 ppm and a LOAEL of 625 for non-neoplastic 

hematotoxic effects. 

Table 1. Reproductive system atrophy and 2 year survival (NTP, 1993) 

Butadiene 

 (ppm) 

Female 

survival 

Atrophy of 

ovary 

Atrophy of 

uterus 

Male 

survival 

Atrophy of 

testicle 

 0 37/50 4/49 1/50 35/50 1/50 

 6.25 33/50 19/49 0/49 39/50 3/50 

 20 24/50 32/48 1/50 24/50 4/50 

 62.5 11/50 42/50 1/49 22/50 2/48 

200 0/50 43/50 8/50 4/50 6/49 

625 0/80 69/79 41/78 0/70 53/72 

The U.S. EPA (1985) reviewed data from a 2-year chronic inhalation toxicity study sponsored by 

the International Institute of Synthetic Rubber Producers (IISRP) at Hazelton Laboratories 

Europe, Ltd (1981) on Sprague-Dawley rats exposed to 0, 1000 or 8000 ppm 1,3-butadiene. 

Results from the study were also reported later by Owen et al. (1987; 1990). Minor clinical 

effects, including excessive eye and nose secretions plus slight ataxia, were observed between 2 

and 5 months in rats exposed to 8000 ppm 1,3-butadiene. Alterations in organ weight were also 

observed in this high exposure group. A dose-related increase in liver weights was observed at 

both the 52-week interim kill and at study termination. Absolute and relative kidney weight was 

also significantly increased and associated with nephrosis. No reproductive organ atrophy was 

reported in this rat study; however, tumors were found in reproductive tissues (Owen et al., 

1987). 

Penn and Snyder (1996a,b) exposed cockerels (young male chickens) to 0 or 20 ppm 1,3-

butadiene 6 hr/day, 5 days/week for 16 weeks to study arteriosclerotic plaque development. The 

cockerel is a sensitive animal model for studying the effects of environmental arteriosclerotic 

plaque-promoting agents. Plaque frequency and location were not affected. However, plaque 

sizes were significantly larger in 1,3-butadiene-treated cockerels than in controls. 

The U.S. EPA (1985) described another secondary report, that of Miller (1978), which reviewed 

a group of Russian studies of subchronic 1,3-butadiene exposure in rats. One study (reported by 

Ripp in 1967) continuously exposed rats to relatively lower concentrations of 0.45, 1.4 or 13.5 

ppm. At 13.5 ppm, blood cholinesterase was elevated, blood pressure was lowered, and motor 

Appendix D3 50 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

activity was decreased. Histopathological changes reported at 0.45 ppm were congestion in the 

spleen and hyperemia and leukocyte infiltration of cardiac tissue. Alterations in lung tissue 

noted at 1.4 and 13.5 ppm included atelectasis, interstitial pneumonia, and emphysema. No other 

studies used such low exposure levels or measured such endpoints. Unfortunately, the specific 

research methods and results for this study are unavailable for direct review and comparison. 

A series of reproductive and developmental toxicity studies undertaken by U.S. EPA was 

summarized by Morrissey et al. (1990). In developmental toxicity studies, pregnant female rats 

and mice were exposed to 0, 40, 200, or 1000 ppm 1,3-butadiene for 6 hrs/day on days 6-15 of 

gestation. In rats, maternal body weight gain and extra-gestational body weight gain was 

reduced at the highest exposure. However, no evidence of developmental toxicity was observed. 

In mice, maternal body weight gain and extra-gestational body weight gain were reduced at 200 

and 1000 ppm. Gravid uterine weight was reduced at 1000 ppm. Fetal and placental weights 

were reduced in an exposure-dependent manner with reduced male fetal body weight reaching 

statistical significance at 40 ppm and above. In the sperm head morphology assay and the 

dominant lethality study, groups of male mice were exposed to 200, 1000, and 5000 ppm 1,3-

butadiene for 5 consecutive days. Concentration-related small increases in the percentages of 

abnormal sperm heads were observed, but were statistically significant only at the two highest 

exposures. Dominant lethal effects were observed only in the first two weeks following 

exposure. At week 1, the percentage of dead implants/total implants was increased only at 1000 

ppm, and the percentage of females with > 2 dead implants was increased at 200 and 1000 ppm. 

The number of dead implants/pregnancy was increased beginning at 1000 ppm at week 1, and 

200 and 1000 ppm at week 2. While not strongly concentration dependent, the dominant 

lethality results are consistent with an adverse effect of 1,3-butadiene on more mature cells 

(spermatozoa and spermatids). 

An acute and subchronic (10 week) study identified male-mediated F1 effects in mice exposed to 

12.5 or 1250 ppm 1,3-butadiene for 6 hours/day, 5 days/week (Anderson et al., 1996). An 

additional group of mice were also exposed to 6250 ppm 1,3-butadiene in the acute study. 

Meaningful toxic effects were not observed in the acute study and no reproductive parameters 

were affected in either study. In the 10-week study, 1250 ppm (2762.4 mg/m3

) resulted in a 

statistically significant reduction in the number of implantations, an induction of dominant lethal 

mutations, an increased incidence of early and late deaths, and an increase in abnormalities. The 

lower level of 12.5 ppm (27.63 mg/m3

) resulted in an increase of late deaths and fetal 

abnormalities. 

A follow-up of the Anderson et al. (1996) dominant lethality study exposed male mice to 12.5 or 

125 ppm 1,3-butadiene under the same subchronic exposure conditions (Brinkworth et al., 1998). 

A statistically significant increase in early deaths was observed at 125 ppm. The incidences of 

late deaths, dead fetuses, and abnormalities were elevated at 125 ppm but were not statistically 

significant. Testicular DNA damage, as detected by the Comet assay, was observed at 125 ppm. 

Further dominant lethality studies in rodents by the same research group exposed male mice to 

12.5, 65, and 130 ppm 1,3-butadiene 6 hr/day, 5 days/week for four weeks (Anderson et al., 

1998). Groups of male rats were also exposed to 65, 400, and 1250 ppm 1,3-butadiene 6 hr/day, 

5 days/week for 10 weeks. In mice, a statistically significant increase in early deaths was 

Appendix D3 51 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

observed at 65 and 130 ppm but was not dose-related. Male-mediated effects in rats were not 

observed at any exposure level. 

Pacchierotti et al. (1998) investigated 1,3-butadiene-induced toxic effects on spermatogenic cell 

stages and first-cleavage embryos. Exposure of male mice to 130, 500, and 1300 ppm 1,3-

butadiene 6 hr/day for 5 days did not result in an increase of unfertilized oocytes after pairing 

with untreated females. However, statistically significant increases of cytogenetic aberrations in 

first-cleavage embryos were observed in the first mating week in mice exposed to 500 and 1300 

ppm, and in the second mating week in mice treated with 1300 ppm. Treatment-related effects 

on differentiating spematogonia were shown by a concentration-dependent decrease of round 

spermatids occurring 21 days after exposure, and confirmed 7 days later by a similar decrease of 

elongated spermatids. Testis weight was significantly reduced at all doses tested, 21 days after 

the end of exposure. A dose-dependent increase of variant sperm with single-stranded DNA 

content was observed 28 days after exposure, and attained statistical significance at 1300 ppm. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study NTP (1993) 

Study population B6C3F1 mice (70/sex/group) 

Exposure method Discontinuous inhalation (0, 6.25, 20, 62.5, 200, 

625 ppm) over 2 years 

Critical effects Increased incidence of ovarian atrophy 

LOAEL 6.25 ppm 

NOAEL Not observed 

BMC05 1.40 ppm 

Exposure continuity 6 hr/d, 5 d/wk 

Exposure duration 103 weeks 

Average experimental exposure 0.25 ppm for BMC05 

(1.40 ppm x 6/24 hr/day x 5/7 days/week) 

Human equivalent concentration 0.25 ppm (gas with systemic effects, based on 

RGDR = 1.0 using default assumption that 

lambda (a) = lambda (h)) 

LOAEL uncertainty factor Not needed in the BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 8 ppb (0.008 ppm; 0.02 mg/m3

; 20 μg/m3) 

The chronic REL for butadiene is based on an increased incidence of ovarian atrophy in mice. 

Characteristically, affected females had no evidence of oocytes, follicles, or corpora lutea. 

Significant reproductive toxicity was observed in both sexes of mice at the interim 9-month, 

interim 15-month, and 2-year study termination as gonadal atrophy (NTP, 1993). Testicular 

atrophy was induced in male B6C3F1 mice at 625 ppm or above in this principal study and in a 

previous study (NTP, 1984). In female mice exposed for 9-months, ovarian atrophy was 

Appendix D3 52 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

observed at 200 and 625 ppm (442 or 1381 mg/m3, respectively). After 15 months, ovarian 

atrophy was observed at exposure levels of 20 ppm (44.2 mg/m3

) and above. In mice exposed 

for up to 2 years (103 weeks), the incidence of ovarian atrophy increased at all exposure 

concentrations relative to controls, which establishes a chronic LOAEL of 6.25 ppm (13.81 

mg/m3

) for reproductive toxicity. 

Presentation of the ovarian atrophy data in quantal form (see Table 1) allows the use of the 

benchmark concentration (BMC) approach to determine the REL. A log-normal probit analysis 

(U.S. EPA, National Center for Environmental Assessment, benchmark dose software, version 

1.20) using only the control group and the log-dose of the three lowest butadiene exposure 

groups provided the lowest chi-square value (i.e., the best line fit to the data points). The 

proportion of mice developing ovarian atrophy in the two highest exposure groups did not 

increase appreciably with increasing exposure concentration, and therefore, deviated from the 

log-normal probit plot. The significantly shortened survival rate in these two groups may be one 

reason for this deviation. Another possible cause is that a relatively resistant subgroup of mice 

(to ovarian atrophy) is revealed at the two highest doses following 2-year exposure to 1,3-

butadiene. Thus, it may be biologically plausible to remove these resistant subgroups when 

using a BMC approach. The maximum likelihood estimate (MLE) for a 5% response was 1.53 

ppm. The resulting 95% lower confidence limit at the MLE provided a BMC05 of 1.40 ppm. A 

BMC05 is considered to be similar to a NOAEL in estimating a concentration associated with a 

low level of risk. 

The mouse ovary is more sensitive to butadiene’s epoxide metabolites than the rat ovary. Doerr 

et al. (1996) administered butadiene monoepoxide (BMO) or butadiene diepoxide (BDE) 

intraperitoneally to female B6C3F1 mice and Sprague-Dawley rats for 30 days and found that 

BMO and BDE exhibited a greater ovotoxic potential in the mice compared to the rats. Dahl et

al. (1991) reported that, for equivalent inhalation exposures, the concentrations of total butadiene 

metabolites in blood were 5-50 times lower in the monkeys than in the mice and 4-14 times 

lower than in the rats. People may be more like the monkey than the mouse or the rat in their 

formation of epoxides from butadiene. In vitro metabolism studies with human liver tissue 

present conflicting results regarding whether humans would be more like rats or mice in forming 

epoxide metabolites (Bond et al., 1996; Duescher and Elfarra, 1994). The considerable degree of 

interindividual variability in human samples was a reason given for the inconsistencies. Several 

pharmacokinetic models (Sweeney et al., 1997; reviewed by Himmelstein et al., 1997) have been 

developed to adjust for species differences in pharmacokinetics. However, an interspecies 

pharmacodynamic adjustment for this ovarian atrophy endpoint with butadiene is still needed. 

Therefore OEHHA staff use an interspecies uncertainty factor of 3 to account for 

pharmacodynamic differences between mice and women. 

Christian (1996) has postulated that it may be inappropriate to develop health-protective values 

for 1,3-butadiene based on 2-year ovarian atrophy in mice because the mice are beyond their 

normal reproductive age. It was suggested that the 15-month evaluation of ovarian atrophy 

conducted by the NTP (1993) would be a better indicator of reproductive risk. However, 

OEHHA staff believes that butadiene-induced ovarian atrophy represents a toxic manifestation in 

an organ system. The fact that it occurs in a reproductive organ is immaterial for the 

development of a chronic REL. Nonetheless, a comparison REL based on the 15-month interim 

Appendix D3 53 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

evaluation for ovarian atrophy can be estimated. Quantal data at the 15-month interim 

evaluation shows that no mice developed ovarian atrophy (0/10) in the control group or at the 

lowest exposure. Ovarian atrophy was observed in 1/10, 9/10, 7/10, and 2/2 mice at the 20, 62.5, 

200, and 625 ppm exposure groups, respectively. A log-normal probit analysis (U.S. EPA, 

National Center for Environmental Assessment, benchmark dose software draft, beta version 

1.1b) based on the 15-month ovarian atrophy data provided an MLE of 8.12 ppm and a BMC05

of 3.08 ppm. Following adjustment for exposure continuity (6/24 hr/day, 5/7 days/wk) to 0.55 

ppm and dividing by a total UF of 30 (3 for interspecies variability and 10 for intraspecies 

variability), a REL of 20 ppb (40 μg/m3

) was attained. 

Another comparison to the proposed REL can be made using the dominant lethality study of 

Anderson et al. (1998). Early fetal deaths were observed at 65 and 125 ppm, but not 12.5 ppm. 

An earlier dominant lethality study (Anderson et al., 1996) indicated that early deaths may occur 

at 12.5 ppm but the toxicological effect could not be repeated at this concentration in subsequent 

studies. The average exposure duration at the NOAEL is 3.125 ppm (12.5 ppm x 6 hr/24 hr). 

Use of an RGDR of 1 and a cumulative uncertainty factor of 30 (3 for interspecies and 10 for 

intraspecies) resulted in a REL of 0.1 ppm (0.2 mg/m3). Since the endpoint is a function of 

exposure during sperm maturation, no subchronic UF was used. The U.S. EPA had observed 

developmental toxicity in fetal rats (reduced male fetal body weight ) at 40 ppm (Morrissey et 

al., 1990). However, unlike the Anderson et al. (1998) study, a NOAEL was not determined. 

Recent studies have implicated 1,3-butadiene in accelerating arteriosclerotic plaque development 

in cockerels (Penn and Snyder, 1996a,b), although no animal studies in mammals have 

implicated 1,3-butadiene in this disease. The worker study by McMichael et al. (1976) observed 

a slight increase in mortality from arteriosclerosis among all rubber workers. But more recent 

mortality studies in the rubber industry found no association or found an actual mortality 

decrement from arteriosclerosis and other circulatory diseases when compared to a reference 

population, suggesting a ‘healthy worker’ effect (Divine and Hartman, 1996; Matanoski et al., 

1990; Sathiakumar et al., 1998). 

When mortality among rubber workers was adjusted for race, two studies found that black rubber 

workers had a small, although statistically significant, increased mortality from arteriosclerosis 

compared to the black male U.S. population (Divine and Hartman, 1996; Matanoski et al., 1990). 

But a larger study of black workers in the rubber industry found no association between 

circulatory diseases, which includes arteriosclerosis, and mortality (Sathiakumar et al., 1998). 

Weaknesses in these worker analyses include relatively small cohort sizes, the bias of having 

racial information on all deaths and not on all living workers, the lack of racial data on some 

workers (up to 15% of cohort), and the lack of complete or specific work histories of the 

subjects. Also, black men of certain age groups are known to have an increased standardized 

mortality ratio for arteriosclerotic (ischemic) heart disease compared to white men (CDC, 2000). 

Limited data, conflicting worker mortality results, and lack of underlying mechanisms of action 

prevent the use of these findings in 1,3-butadiene REL development. However, there clearly is a 

need for further animal and epidemiological studies to determine if there is a true association 

between 1,3-butadiene exposure and arteriosclerotic diseases. 

Appendix D3 54 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VII. Data Strengths and Limitations for Development of the REL

The major strength of the 1,3-butadiene REL is the observation of a dose-response effect in a 

well-conducted lifetime inhalation exposure study. The major weaknesses are the lack of 

adequate human health effects and metabolism data and the lack of a NOAEL observation in the 

key study. 

VIII. References

Anderson D, Edwards AJ, Brinkworth MH, and Hughes JA. 1996. Male-mediated F1 effects in 

mice exposed to 1,3-butadiene. Toxicology 113:120-127. 

Anderson D, Hughes JA, Edwards AJ, and Brinkworth MH. 1998. A comparison of malemediated effects in rats and mice exposed to 1,3-butadiene. Mutat. Res. 397:77-84. 

Bond JA, Himmelstein MW, Seaton M, Boogaard P, and Medinsky MA. 1996. Metabolism of 

butadiene by mice, rats, and humans: a comparison of physiologically based toxicokinetic model 

predictions and experimental data. Toxicology 113:48-54. 

Brinkworth MH, Anderson D, Hughes JA, Jackson LI, Yu T-W, and Nieschlag E. 1998. Genetic 

effects of 1,3-butadiene on the mouse testis. Mutat. Res. 397:67-75. 

CARB. 1999. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 1,3-

Butadiene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CDC. 2000. Centers for Disease Control and Prevention. National Center for Health Statistics. 

Available online at http://www.cdc.gov/nchs/datawh/statab/unpubd.htm 

Checkoway H, and Williams TM. 1982. A hematology survey of workers at a styrene-butadiene 

synthetic rubber manufacturing plant. Am. Ind. Hyg. Assoc. J. 43(3):164-169. 

CRC. 1995. CRC Handbook of Chemistry and Physics, 76th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Christian MS. 1996. Review of reproductive and developmental toxicity of 1,3-butadiene. 

Toxicology 113:137-143. 

Dahl AR, Sun JD, Birnbaum LS, Bond JA, Griffith WC Jr, Mauderly JL, Muggenburg BA, 

Sabourin PJ, Henderson RF. 1991. Toxicokinetics of inhaled 1,3-butadiene in monkeys: 

comparison to toxicokinetics in rats and mice. Toxicol. Appl. Pharmacol. 110(1):9-19. 

(Published erratum appears in Toxicol. Appl. Pharmacol. 1992. 116(1):152.) 

Appendix D3 55 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Divine BJ and Hartman CM. 1996. Mortality update of butadiene production workers. 

Toxicology 113:169-181. 

Doerr JK, Hollis EA, and Sipes IG. 1996. Species difference in the ovarian toxicity of 1,3-

butadiene epoxides in B6C3F1 mice and Sprague-Dawley rats. Toxicology 113(1-3):128-136. 

Duescher RJ and Elfarra AA. 1994. Human liver microsomes are efficient catalysts of 1,3-

butadiene oxidation: evidence for major roles by cytochromes P450 2A6 and 2E1. Arch. 

Biochem. Biophys. 311(2):342-349. 

Himmelstein MW, Acquavella JF, Recio L, Medinsky MA, and Bond JA. 1997. Toxicology and 

epidemiology of 1,3-butadiene. Crit. Rev. Toxicol. 27(1):1-108. 

HLE. 1981. Hazelton Laboratories Europe, Ltd. 1,3-Butadiene: The toxicity and carcinogenicity 

of butadiene gas administered to rats by inhalation for approximately 24 months. Prepared for 

the International Institute of Synthetic Rubber Producers, New York, NY. Unpublished. [as 

reported in U.S. EPA, 1985]. 

HSDB. 2000. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD. 

Available online through Toxicology Data Network at http://toxnet.nlm.nih.gov 

Huff JE, Melnick RL, Solleveld HA, Haseman JK, Powers M, and Miller RA. 1985. Multiple 

organ carcinogenicity of 1,3-butadiene in B6C3F1 mice after 60 weeks of inhalation exposure. 

Science 227(4686):548-549. 

Matanoski GM, Santos-Burgoa C, and Schwartz L. 1990. Mortality of a cohort of workers in the 

styrene-butadiene polymer manufacturing industry (1943-1982). Environ. Health Perspect. 

86:107-117. 

McMichael AJ, Spirtas R, and Kupper KK. 1975. Solvent exposure and leukemia among rubber 

workers: An epidemiologic study. J. Occup. Med. 17:234-239. 

McMichael AJ, Spirtas, R, Gamble JF, and Tousey PM. 1976. Mortality among rubber workers. 

Relationship to specific jobs. J. Occup. Med. 18(3):178-185. 

Melnick RL, Huff J, Chou BJ, and Miller R. 1990. Carcinogenicity of 1,3-butadiene in 

C57BL/6xC3H F1 mice at low exposure concentrations. Cancer Res. 50(20):6592-6599. 

Miller LM. 1978. Investigation of selected potential environmental contaminants: Butadiene and 

its oligomers. Philadelphia, PA: Franklin Research Center. [as reported in U.S. EPA, 1985]. 

Morrissey RE, Schwetz BA, Hackett PL, Sikov MR, Hardin BD, McClanahan BJ, Decker JR, 

and Mast TJ. 1990. Overview of reproductive and developmental toxicity studies of 1,3-

butadiene in rodents. Environ. Health Perspect. 86:79-84. 

Appendix D3 56 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

NTP. 1984. U.S. National Toxicology Program. Toxicology and Carcinogenesis Studies of 1,3-

Butadiene (CAS No.106-99-0) in B6C3F1 Mice (Inhalation Studies) (TR No. 288) Research 

Triangle Park, NC: National Institute of Environmental Health Sciences. 

NTP. 1993. U.S. National Toxicology Program. Toxicology and Carcinogenesis Studies of 1,3-

Butadiene (CAS No.106-99-0) in B6C3F1 Mice (Inhalation Studies) (TR No. 434). Research 

Triangle Park, NC: National Institute of Environmental Health Sciences. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov

Owen PE, Glaister JR, Gaunt IF, and Pullinger DH. 1987. Inhalation toxicity studies with 1,3-

butadiene. 3. Two-year toxicity/carcinogenicity studies in rats. Am. Ind. Hyg. Assoc. J. 48:407-

413. 

Owen PE, and Glaister JR. 1990. Inhalation toxicity and carcinogenicity of 1,3-butadiene in 

Sprague-Dawley rats. Environ. Health Perspect. 86:18-25. 

Pacchierotti F, Tiveron C, Ranaldi R, Bassini B, Cordelli E, Leter G, and Spanò M. 1998. 

Reproductive toxicity of 1,3-butadiene in the mouse: cytogenetic analysis of chromosome 

aberrations in first-cleavage embryos and flow cytometric evaluation of spermatogonial cell 

killing. Mutat. Res. 397:55-66. 

Penn A and Snyder CA. 1996. Butadiene inhalation accelerates arteriosclerotic plaque 

development in cockerels. Toxicology 113:351-354. 

Penn A and Snyder CA. 1996. 1,3 Butadiene, a vapor phase component of environmental 

tobacco smoke, accelerates arteriosclerotic plaque development. Circulation 93:552-557. 

Sathiakumar N, Delzell E, Hovinga M, Macaluso M, Julian JA, Larson R, Cole P, Muir DCF. 

1998. Mortality from cancer and other causes of death among synthetic rubber workers. Occup. 

Environ. Med. 55:230-235. 

SCAQMD. 2000. South Coast Air Quality Management District. MATES-II. Multiple Air 

Toxics Exposure Study in the South Coast Air Basin. Final Report. Diamond Bar, CA: 

SCAQMD. March. p. 3-10. 

Sweeney LM, Schlosser PM, Medinsky MA, and Bond JA. 1997. Physiologically based 

pharmacokinetic madeling of 1,3-butadiene, 1,2-epoxy-3-butene, and 1,2:3,4-diepoxybutane 

toxicokinetics in mice and rats. Carcinogenesis 18:611-625. 

U.S. EPA. 1985. U.S. Environmental Protection Agency. Mutagenicity and Carcinogenicity 

Assessment of 1,3-Butadiene. Publication No. EPA/600/8-85/004F. Washington, DC: U.S. 

Environmental Protection Agency, Office of Health and Environmental Assessment. 

Appendix D3 57 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Wilson RH. 1944. Health hazards encountered in the manufacture of synthetic rubber. JAMA 

124(11):701-703. 

Appendix D3 58 Butadiene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CADMIUM AND CADMIUM COMPOUNDS 

CAS Registry Number: 7440-43-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.02 μg/m3 (respirable) 

Critical effect(s) Kidney effects (proteinuria) and respiratory 

effects (reduction in forced vital capacity and 

reduction in peak expiratory flow rate) in 

occupationally exposed humans 

Hazard index target(s) Kidney; respiratory system 

II. Physical and Chemical Properties (ATSDR, 1993) 

 

Description Blue-white solid 

Molecular formula Cd 

Molecular weight 112.41 g/mol 

Density 8.642 g/cm3

 @ 20oC 

Boiling point 765oC (CRC, 1994) 

Melting point 320.9oC 

Vapor pressure 1 torr @ 394oC 

Conversion factor Not applicable 

III. Major Uses or Sources

The production of nickel-cadmium batteries is currently the primary use of cadmium (ATSDR, 

1993). Cadmium, a by-product of zinc- and sulfide-ore processing, is also used for metal plating 

and in pigments and plastics. The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 3672 pounds of cadmium (CARB, 2000). 

IV. Effects of Human Exposure

Pulmonary and renal function were examined in three worker groups: women with less than 20 

years of exposure [group E1]; men with less than 20 years of exposure [group E2], and men with 

more than 20 years of exposure [group E3] (Lauwerys et al., 1974). Exposed groups were 

matched to control groups in terms of age, body size, cigarettes smoked per day, duration of 

smoking, and duration of employment. Although urine cadmium concentrations were 

significantly elevated, the subjects in E1 did not exhibit pulmonary function changes or 

Appendix D3 59 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

proteinuria indicative of renal impairment. The workers in E1 had been exposed for a mean of 

4.08 years to 31 μg/m3 total cadmium (1.4 μg/m3 respirable cadmium). The 27 workers in E2 

had been exposed for a mean of 8.6 years to 134 μg/m3 total cadmium (88 μg/m3 respirable 

cadmium). The blood and urinary cadmium levels of these workers were also significantly 

elevated compared to matched controls. Glomerular proteinuria was observed in 15% of the 

workers in E2 and in 68% of workers in E3. The 22 workers of E3 had been exposed for a mean 

of 27.8 years to 66 μg/m3 total cadmium (21 μg/m3 respirable cadmium). Significantly increased 

levels of cadmium were observed in the blood and urine, and workers in E3 also exhibited 

significant decreases in some measures of pulmonary function (forced vital capacity, forced 

expiratory volume in one second, and peak expiratory flow rate). This study identifies the 

kidney as the key target organ of chronic cadmium exposure. For respirable cadmium, this study 

indicates a LOAEL of 21 μg/m3 for workers exposed for 28 years and a NOAEL of 1.4 μg/m3

for workers exposed for 4 years. 

A study of 82 cadmium exposed workers reports the time-weighted cumulative exposure index 

(TWE) and cadmium body burden determined in vivo (Ellis et al., 1985). Evidence of renal 

dysfunction (usually elevated urinary β2-microglobulin) was consistently observed when the 

worker’s liver cadmium burden exceeded 40 ppm and the time-weighted cumulative exposure 

index exceeded 400-500 μg years/m3. 

A detailed investigation of renal function in 75 male cadmium-exposed workers identified 

significant increases in urinary excretion of several low- and high molecular weight proteins, 

including β2-microglobulin, and significant decreases in renal reabsorption of calcium, urate, and 

phosphate compared to controls (Mason et al., 1988). Exposures, which ranged from 36 to 

600 μg/m3, were determined from background or personal exposure measurements made 

between 1964 and 1983, or were estimated. A time-weighted cumulative exposure index (TWE) 

was determined for each subject. A two phase linear regression model was applied to the data to 

identify inflection points for each biochemical parameter. The biochemical indicators most 

highly correlated to exposure were urinary retinol binding protein and urinary β2-microglobulin. 

Of these, the most sensitive parameter, urinary β2-microglobulin, demonstrated an inflection 

point at 1108 μg years/m3 with a 95% lower confidence limit of 509 μg years/m3. The endpoint 

selected is indicative of defects in tubular reabsorption of proteins. 

Diminished sensitivity of smell has also been observed in cadmium exposed workers (Rose et 

al., 1992). Cadmium body burden, β2-microglobulin levels, and olfactory function were 

measured in a group of 55 male workers exposed to cadmium fumes in a brazing operation. A 

group of 15 control workers was also tested. Exposed workers exhibited high urinary cadmium 

levels, tubular proteinuria, and a significant, selective defect in odor detection threshold. 

V. Effects of Animal Exposure 

Interstitial infiltration of lymphocytes and leukocytes and hyaline casts were observed in the 

kidneys of rabbits following exposure to 6.5 mg/m3

 cadmium-iron dust for 3 hours per day, 

21 days per month for 9 months (Friberg, 1950). Proteinuria was observed in the majority of 

exposed rabbits by the fourth month of exposure. Increased lung weights and emphysema were 

Appendix D3 60 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

also observed. The trachea and nasal mucous membranes exhibited chronic inflammatory 

changes (not specified) and lymphocyte infiltration. The kidney contained the greatest 

concentration of cadmium. This study also exposed a group of rabbits to 9.1 mg/m3

cadmium-iron dust for 3 hours per day, 23 days per month, for 7 months. Two rabbits in this 

group died from acute pneumonia at one month, and one rabbit was terminated at 3 months of 

exposure. Findings at necropsy were similar, although more severe than those observed in 

rabbits exposed to 6.5 mg/m3. Chronic bronchitis and hyperplasia of the bronchiolar epithelium 

were observed in the higher dose group in addition to the findings previously noted. 

Male and female rats were exposed to 0, 0.3, 1.0, or 2.0 mg Cd/m3 (as CdCl2) 6 hours per day, 5 

days per week for a total of 62 exposures (Kutzman et al., 1986). Rapid, shallow breathing and 

marked weight loss were observed in the highest dose group; all animals in this group died 

within the first 45 days of exposure. A dose-dependent increase in lung weight was observed in 

the remaining dose groups and a statistically significant increase in lung collagen and elastin was 

observed in rats exposed to 1.0 mg/m3. Pathological changes noted in the terminal bronchioles 

include flattening and hyperplasia of type II cells, and infiltration of macrophages, mononuclear 

cells, and polymorphonuclear leukocytes. Proliferation of fibroblasts with deposition of collagen 

was also noted. 

Male rats were exposed continuously to 0, 30, or 90 μg Cd/m3 cadmium oxide (CdO) dust for up 

to 18 months (Takenaka et al., 1990). Animals exposed to 30 μg/m3 were sacrificed at 6 and 18 

months of exposure. Although some rats in the high dose group were terminated after 6 months 

of exposure, the remaining rats were terminated after 7 months due to increased mortality and 

were not included in the study. Inflammation and hyperplasia of the alveolar epithelium 

occurred in animals of both groups after 6 months of exposure with more marked changes 

observed in the high dose group. Abnormal proliferation of the epithelium was observed in the 

low dose group following 18 months of exposure. Lung tumors observed in both dose groups 

were characterized as being duration dependent. 

Appendix D3 61 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VI. Derivation of Chronic Reference Exposure Levels (REL) 

Derivation of Chronic Inhalation Reference Exposure Level

Study Lauwerys et al., 1974 

Study population Humans (22 exposed men and 22 unexposed 

men in LOAEL group; 31 exposed women 

and 31 non-exposed women in NOAEL 

group) 

Exposure method Occupational exposures 

Critical effects Kidney effects - proteinuria in 68% of LOAEL 

group 

Respiratory effects – reduction in forced vital 

capacity (FVC), forced expiratory flow in 1 

second (FEV1); reduction in peak expiratory 

flow rate 

LOAEL 21 μg/m3

 respirable cadmium 

NOAEL 1.4 μg/m3

 respirable cadmium 

Exposure continuity Assumed to be 5 days/week for 8 hours/day 

during which 10 m3

 air is breathed 

Average occupational exposure 0.5 μg/m3

 for NOAEL group (1.4 x 10/20 x 5/7) 

Human equivalent concentration 0.5 μg/m3

 for NOAEL group 

Exposure duration Average of 4.1 years (1 to 12 years) for NOAEL 

group 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.02 μg/m3

VII. Data Strengths and Limitations for Development of the REL 

This evaluation of a chronic REL for cadmium is strengthened by being based on a human 

exposure study of workers exposed to cadmium for periods of 1 to over 20 years. The exposed 

group was matched to a control group in terms of age, body size, cigarettes smoked per day, 

duration of smoking, and duration of employment. The factory process was unchanged over the 

study period suggesting that exposures may have remained relatively constant over time. 

Significant areas of uncertainty include an incomplete knowledge of the past exposures over the 

full study interval and the relatively small number of subjects in the study. 

A similar evaluation of the LOAEL group led to an alternate estimate for an inhalation reference 

exposure level of 0.05 μg/m3

. The LOAEL group had an average occupational exposure of 

5.0 μg/m3 and an average exposure duration of 27.8 years (21 to 40 years). Default uncertainty 

Appendix D3 62 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

factors included a 10-fold LOAEL uncertainty factor and a 10-fold intraspecies uncertainty 

factor (UF). 

For comparison, using data presented by Ellis and associates (1985) and Mason and associates 

(1993) correlating human cumulative exposures (in terms of μg-years/m3) and renal tubular 

protein reabsorption, a LOAEL of 500 μg-years/m3 was predicted. This correlates to 7 μg/m3

over 70 years. A time-weighted exposure to account for continuous exposure rather than 40 hour 

per week occupational exposure is 1.7 μg/m3

. Applying a 10-fold LOAEL uncertainty factor and 

a 10-fold intraspecies uncertainty factor results in a REL value of 0.02 μg/m3

, the same value 

obtained using the Lauwerys et al. data. U.S. EPA has not published an RfC for cadmium. 

In addition to being inhaled, airborne cadmium can settle onto crops and soil and enter the body 

by ingestion. Thus an oral chronic reference exposure level for cadmium is also required. We 

propose adopting the U.S. EPA RfD as the chronic oral REL. 

Derivation of Chronic Oral Reference Exposure Level (U.S. EPA RfD)

Study U.S. EPA, 1985 

Study population Humans 

Exposure method Food and drinking water 

Critical effects Significant proteinuria 

LOAEL Not observed 

NOAEL 0.005 mg/kg bw-day 

Exposure continuity Chronic

Exposure duration Up to lifetime 

Average exposure 0.005 mg/kg bw-day 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies factor 10

Cumulative uncertainty factor 10 

Oral reference exposure level 0.0005 mg/kg bw-day 

The oral REL is the U.S. EPA’s Reference Dose (RfD) (U.S. EPA, 1996). A concentration of 

200 μg cadmium (Cd)/gm wet human renal cortex is the highest renal level not associated with 

significant proteinuria (U.S. EPA, 1985). A toxicokinetic model is available to determine the 

level of chronic human oral exposure (NOAEL) which results in 200 μg Cd/gm wet weight 

human renal cortex. The model assumes that 0.01% of the Cd body burden is eliminated per day 

(U.S. EPA, 1985). Assuming 2.5% absorption of Cd from food or 5% from water, the 

toxicokinetic model predicts that the NOAEL for chronic Cd exposure is 0.005 and 0.01 mg 

Cd/kg/day from water and food, respectively (i.e., levels which would result in 200 μg Cd/gm 

wet weight human renal cortex). Thus, based on an estimated NOAEL of 0.005 mg Cd/kg/day 

for Cd in drinking water and an UF of 10, an RfD of 0.0005 mg Cd/kg/day (water) was 

calculated; an equivalent RfD for Cd in food is 0.001 mg Cd/kg/day. 

Appendix D3 63 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Cd is unusual in relation to most, if not all, of the substances for which an oral RfD has been 

determined in that a vast quantity of both human and animal toxicity data are available. The RfD 

is based on the highest level of Cd in the human renal cortex (i.e., the critical level) not 

associated with significant proteinuria (i.e., the critical effect). A toxicokinetic model has been 

used to determine the highest level of exposure associated with the lack of a critical effect. Since 

the fraction of ingested Cd that is absorbed appears to vary with the source (e.g., food vs. 

drinking water), it is necessary to allow for this difference in absorption when using the 

toxicokinetic model to determine an RfD. 

The uncertainty factor of 10 is used to account for intrahuman variability to the toxicity of this 

chemical in the absence of specific data on sensitive individuals. No modifying factor was used. 

U.S. EPA stated its confidence in the RfD as: Study - Not applicable; Data Base - High; and RfD 

– High. The choice of NOAEL does not reflect the information from any single study. Rather, it 

reflects the data obtained from many studies on the toxicity of cadmium in both humans and 

animals. These data also permit calculation of pharmacokinetic parameters of cadmium 

including absorption, distribution, metabolism, and elimination. All this information considered 

together gives high confidence in the data base. High confidence in the RfD follows as well. 

VIII. References 

ATSDR. 1993. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Cadmium. Prepared under contract No. 205-88-0606 for: US Department of Health and Human 

Services, Public Health Service. Atlanta, GA: ATSDR. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Ellis KJ, Cohn SH, and Smith TJ. 1985. Cadmium inhalation exposure estimates: Their 

significance with respect to kidney and liver cadmium burden. J. Toxicol. Environ. Health 

15:173-187. 

Friberg L. 1950. Health hazards in the manufacture of alkaline accumulators with special 

reference to chronic cadmium poisoning. Acta Med. Scand. (Suppl. 240) 138:1-124. 

Lauwerys RR, Buchet JP, Roels HA, Brouwers J, and Stanescu D. 1974. Epidemiological survey 

of workers exposed to cadmium. Arch. Environ. Health 28:145-148. 

Kutzman RS, Drew RT, Shiotsuka RN, and Cockrell BY. 1986. Pulmonary changes resulting 

from subchronic exposure to cadmium chloride aerosol. J. Toxicol. Environ. Health 17:175-189. 

Appendix D3 64 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Mason HJ, Davison AG, Wright AL, Guthrie CJG, Fayers PM, Venables KM, et al. 1988. 

Relations between liver cadmium, cumulative exposure, and renal function in cadmium alloy 

workers. Br. J. Ind. Med. 45:793-802. 

Rose CS, Heywood PG, and Costanzo RM. 1992. Olfactory impairment after chronic 

occupational cadmium exposure. J. Occup. Med. 34(6):600-605. 

Takenaka S, Glaser U, Oldiges H, and Mohr U. 1990. Morphological effects of Cd-aerosols on 

the rat lung. Toxicol. Environ. Chem. 27:163-172. 

U.S.EPA. 1985. United States Environmental Protection Agency. Drinking Water Criteria 

Document on Cadmium. Final draft. Office of Drinking Water. Washington, DC: U.S. EPA. 

U.S. Environmental Protection Agency. 1996. Integrated Risk Information System (IRIS) 

Database. Cadmium. 

Appendix D3 65 Cadmium and Cadmium compounds 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

CHRONIC TOXICITY SUMMARY 

CARBON DISULFIDE 

(carbon bisulfide; carbon sulfide; dithiocarbonic anhydride) 

CAS Registry Number: 75-15-0 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 800 μg/m3 (300 ppb) 

Critical effect(s) CNS/PNS (reduction in motor nerve conduction 

velocities in occupationally-exposed humans) 

Hazard index target(s) Nervous system; reproductive system 

II. Physical and Chemical Properties Summary (HSDB, 1995; CRC, 1994) 

Description Clear, colorless or faintly yellow liquid 

Molecular formula CS2

Molecular weight 76.14 g/mol 

Boiling point 46.5o

C 

Melting point −111.5 oC 

Vapor pressure 297 torr @ 20o

C 

Solubility Slightly soluble in water (2.94 g/L); miscible 

in anhydrous methanol, ethanol, ether, 

benzene, chloroform, and carbon 

tetrachloride 

Conversion factor 3.1 mg/m3

 per ppm at 25oC 

III. Major Uses and Sources

The most prominent industrial use of carbon disulfide is in the production of viscose rayon 

fibers. Carbon disulfide is also used in the production of carbon tetrachloride and cellophane, 

and, as a solvent for rubber, sulfur, oils, resins, and waxes. In the past, carbon disulfide was used 

in soil fumigation and insect control in stored grain. Industrial processes that produce carbon 

disulfide as a by-product include coal blast furnaces and oil refining (HSDB, 1995). Carbon 

disulfide is also a breakdown product of metam sodium. The annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California based on the 

most recent inventory were estimated to be 1562 pounds of carbon disulfide (CARB, 2000). 

Appendix D3 66 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

IV. Effects of Human Exposure

A primary target of carbon disulfide (CS2) toxicity is the nervous system. The major neurotoxic 

action of carbon disulfide is the development of mental disturbances. These include change of 

personality, irritability, and forgetfulness, often with accompanying neurophysiological and 

neuropathological changes after prolonged exposure. Such changes include decreased peripheral 

nerve impulse conduction, motor and/or sensory neuropathies, cerebral or cerebellar atrophy, and 

neuropsychological organic changes (Aaserud et al. 1988, 1990, 1992; Foa et al., 1976; Hirata et 

al. 1992; Ruijten et al. 1990, 1993). Alterations in behavioral indices have historically been 

associated with high levels of CS2, often in excess of 20 ppm (Foa et al. 1976; Hanninen et al., 

1978). 

Studies have identified alterations in the nerve conduction of workers chronically exposed to 

lower CS2 levels (Hirata et al., 1992a; Johnson et al., 1983; Ruijten et al., 1990; Ruijten et al., 

1993). A cross-sectional study of Japanese spinning workers identified alterations in the central 

nervous system as measured by brain stem auditory evoked potential (BAEP) (Hirata et al., 

1992). The latencies of the three main BAEP components increased significantly in 

workersexposed to CS2 for more than 20 years when compared to controls. CS2 exposures 

ranged from 3.3 to 8.2 ppm (mean = 4.76 ppm). Ruijten et al. (1993) identified mild 

presymptomatic nerve impairment (decreased conduction velocities and response amplitudes) in 

44 CS2-exposed workers with an average cumulative exposure range from 192 to 213 ppm-year 

(mean duration = 26.1 years). 

A NIOSH occupational study evaluated the effects of CS2 on the peripheral nervous system. 

Johnson et al. (1983) identified a significant dose related reduction in the maximum motor nerve 

conduction velocities (MCV) in the calves and ankles of male viscose rayon workers exposed to 

high (median = 7.6 ppm) CS2 levels versus a comparison group exposed to low concentrations 

(median = 0.2 ppm). The workers were all employed in artificial fiber production in the same 

plant. Since these reduced MCVs were still within the normal range, the authors considered the 

measured difference an indication of minimal neurotoxicity. The mean exposure concentration 

for all exposed workers (n = 145) ranged from 0.6 to 16 ppm (mean = 7.6 ppm) with a mean 

duration of 12.1 years. This study identified a chronic LOAEL of 7.6 ppm for minor 

neurological effects (decreased peroneal nerve MCV and sural nerve conduction velocity). 

Another epidemiological study evaluated a group of 111 Belgian viscose rayon factory workers 

exposed to 4 to 112 mg/m3

 CS2 (time-weighted average 1 to 40 mg/m3) (Vanhoorne et al., 1995). 

Among four categories of cumulative exposure (0, 1 to 300, 301 to 600, and greater than 600 

mg/m3

-years), a clear dose-response effect was observed for reduced mean peroneal motor nerve 

conduction velocities in both fast and slow fibers. Unfortunately, the data are incompletely 

reported, and the mean duration of exposure is not given. Subgroups of workers whose 

exposures ever exceeded 10 ppm (n=64) and never exceeded 10 ppm (n=30) each showed 

significantly reduced fibular nerve motor conduction velocities compared with non-exposed 

workers. 

Vascular atherosclerotic changes are also considered a major effect of chronic carbon disulfide 

exposure. Several occupational studies have demonstrated an increase in the mortality due to 

Appendix D3 67 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

ischemic heart disease in CS2 exposed workers (Hernberg et al., 1970; MacMahon and Monson, 

1988; Tiller et al., 1968; Tolonen et al., 1979). A 2.5-fold excess in mortality from coronary 

heart disease in workers exposed to CS2 was first reported by Tiller et al. (1968). A subsequent 

prospective study by Hernberg et al. (1970) found a 5.6-fold increased risk in coronary heart 

disease mortality and a 3-fold increased risk of a first nonfatal myocardial infarction in CS2

exposed workers. 

Male workers (n=177) in a Polish fiber plant were exposed to CS2 for an average of 14 years 

(range of 5 to 38 years). Controls were 93 healthy male workers from other factories that did not 

use carbon disulfide. Carbon disulfide exposed workers had higher rates (42%) of 24-hour 

electrocardiographic abnormalities than non-exposed workers (24%, p=0.006) (Bortkiewicz et 

al., 2001). The most common abnormalities were ventricular extrasystoles and repolarization 

disturbances, the latter occurring most often in workers with the longest CS2 exposures. Longterm blood pressure monitoring did not reveal any differences between exposed and control 

groups. 

Male workers in a Belgian viscose rayon factory (n=85) were estimated by personal active 

sampling to have exposures of 2 to 32 mg/m3

 CS2. Controls were 37 non-exposed workers from 

factories that did not use CS2. Exposed workers had reduced common carotid artery distensibility 

as measured with ultrasound sonography, while the carotid artery compliance coefficient was not 

significantly affected. Also, blood pressures and cholesterol levels were not significantly 

different than observed among control workers (Kotseva et al., 2001a). Differences in carotid 

artery distensibility remained significant after adjustment for age, smoking, alcohol 

consumption, ethnicity, body mass index, heart rate, and systolic blood pressure. 

Egeland et al. (1992) and Vanhoorne et al. (1992) have reported that human exposure to CS2 for 

more than one year causes increases in biochemical changes often associated with cardiovascular 

disease - diastolic blood pressure, low density lipoprotein cholesterol, and apolipoproteins A1 

and B. Egeland et al. (1992) used cross sectional data on 165 CS2-exposed workers (245 

controls) collected in 1979 by Fajen et al. (1981). The affected workers were exposed for at 

least 1 year in a viscose rayon factory to an estimated median TWA (8-hour) of 7.6 ppm. The 

Egeland et al. (1992) study indicated that modest CS2 exposure (range = 3.4 to 5.1 ppm, median 

= 4.1 ppm) was associated with increased low density lipoprotein cholesterol (LDLc), the type of 

increase associated with atherosclerotic heart disease. No significant differences were seen 

between controls and the low CS2 exposed group (range = 0.04 to 1.02 ppm, median = 1.00 

ppm). Study NOAEL and LOAEL for increased LDLc and diastolic blood pressure were thus 

1.0 ppm and 4.1 ppm , respectively. Vanhoorne et al. (1992) identified increased LDLcholesterol, apolipoprotein B, systolic and diastolic blood pressure as indicative of an increased 

coronary risk in workers from a Belgian viscose rayon factory (115 exposed and 76 controls). 

CS2 concentrations ranged from 1 to 36 ppm. Duration of exposure was not indicated. Even 

though these biochemical changes were observed, no significant increases in cardiovascular 

disease, such as angina, myocardial infarction, or ischemia, were determined by ECG changes. 

Workers (n=141) with a minimum of 1 year employment in viscose rayons factories were 

compared with 141 age and gender-matched plastic industry workers. Current exposures were 

estimated as 1 to 30 mg/m3 (03 to 10 ppm). Exposed workers were categorized as group 1 or 

Appendix D3 68 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

group 2, with cumulative exposures of less than or greater than 100 mg/m3-years, respectively. 

Group 2 (p<0.001) but not group 1 workers had increased mean total cholesterol (5.3 and 4.5 

mmol/l) compared with controls (4.6 mmol/l) (Kotseva, 2001b). 

CS2 causes reproductive toxicity in both males and females. Lancranjan et al. (1969), 

Lancranjan (1972), Cirla et al. (1978), and Wagar et al. (1983) studied male reproductive effects 

of occupational exposure to CS2 and showed significant adverse effects on spermatogenesis, 

levels of serum FSH and LH, and libido; these effects persisted in 66% of the workers subject to 

follow-up. Zhou et al. (1988) investigated pregnancy outcomes and menstrual disturbances in 

265 women occupationally exposed to CS2 in five facilities and 291 controls. The CS2-exposed 

women had a significantly higher incidence of menstrual disturbances versus the control group 

(overall 34.9% vs. 18.2%). CS2 levels varied between the five facilities (exposure category 

means of low = 3.1 mg/m3, intermediate = 6.5 mg/m3, and high = 14.8 mg/m3), but all workers 

from these CS2 facilities had significantly higher incidences of menstrual disturbance. 

Irregularity of menstruation was the most common disturbance, followed by abnormal bleeding. 

No evidence was observed to indicate an adverse effect on the term and outcome of pregnancy. 

An abstract of an epidemiological study of birth defects among female workers occupationally 

exposed to CS2, was reported by Bao et al. (1991). Exposures were at rayon factories in four 

Chinese provinces and began at least 6 months prior to pregnancy and continued during 

pregnancy. An increased rate of birth defects (2.6% vs. 1.3%) among 682 exposed women was 

noted compared to 745 women in the control group. The most common defects were congenital 

heart defects, inguinal hernia, and CNS defects. However, there was no significant difference in 

birth defects between those with estimated exposures greater than 10 mg/m3 compared to those 

with lower exposures. There were no differences in rates of stillbirth, low birth weight, or 

neonatal or perinatal deaths among any of the groups. 

 

The possibility of determining LOAEL and/or NOAEL values for the major CS2-related adverse 

effects from epidemiology studies, which predominately use workers from the viscose rayon 

industry, is limited. The limitations include incomplete historical exposure measurements, 

concurrent exposure to other chemicals (including hydrogen sulfide or methylene chloride), lack 

of personal exposure determinations, and a high variability of individual exposures due to 

decreases of plant CS2 concentrations over time. 

V. Effects of Animal Exposure

Studies investigating the potential for CS2 toxicity in animals have usually been limited by 

intermediate or subchronic duration (less than 1 year) and a lack of multiple dose or exposure 

groups. The neuropathologic changes consistently observed in rodents following CS2 exposure 

include axonal swelling, demyelination, swelling at neuromuscular junctions, muscle atrophy 

and degeneration, damage to terminal axons, and nerve fiber breakdown (Clerici and Fechter, 

1991; Colombi et al. 1981; Eskin et al., 1988; Jirmanova and Lukas, 1984; Maroni et al., 1979; 

Szendzikowski et al., 1973). These adverse effects have been observed over a range of 

Appendix D3 69 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

exposures (250 to 800 ppm), but few studies have attempted to establish a dose response for this 

CS2-induced neurotoxicity. 

In a 90 day subchronic inhalation study, Sprague-Dawley and Fischer 344 rats exposed 

discontinuously (6 hours/day, 5 days/week) to CS2 developed morphological alterations in nerves 

including axonal swelling and myelin degradation (Gottfried et al., 1985). This study established 

a subchronic NOAEL of 50 ppm and a LOAEL of 300 ppm for morphological changes in nerves. 

A longer inhalation study in Wistar rats observed impairment in the conduction velocity of the 

sciatic and tibial nerves after 6 and 12 months of intermittent exposure to 289 ppm CS2 (LOAEL 

of 289 ppm) (Knobloch et al., 1979). 

In a 13-week subchronic study, male and female F344 rats inhaled 0, 50, 500, or 800 ppm CS2

discontinuously (6 h/day, 5 days per week) (Sills et al.,1998). Development of distal axonopathy 

in the muscular branch of the posterior tibial nerve (MBPTN) and spinal cord was examined. 

After 13 weeks, giant swollen axons were observed with thin myelin sheaths as well as some 

degenerated and regenerated axons. Axonal swelling was noted in the spinal cords of rats 

exposed to 500 or 800 ppm CS2. In the 800 ppm group, additional axonal swelling was observed 

in the muscular branch of the posterior tibial nerve, . Neurofilament deposits were found in 

swollen axons in the spinal cord and MBPTN. The NOAEL for axonal swelling was 50 ppm. 

Wronska-Nofer (1973) showed a positive relationship between the level of triglycerides, the rate 

of cholesterol synthesis, and CS2 exposure in Wistar rats exposed to 0, 73.8, 160, 321, or 546 

ppm CS2 for 5 hours/day, 6 days/week over 8 months. This study found a subchronic LOAEL of 

73.8 ppm for disturbances in lipid metabolism (increase in serum cholesterol and serum 

triglycerides). 

Lewis et al. (1999) investigated the capacity of CS2 to induce arterial fatty deposits by itself, and 

its ability to enhance the rate of fatty deposit formation induced by a high fat diet. Groups of 20 

female C57BL/6 mice were exposed to 0, 50, 500, or 800 ppm CS2 by inhalation. Half the 

animals in each group were placed on an atherogenic high fat diet and half on a control diet. 

Mice were necropsied after 1, 4, 8, 12, 16, or 20 weeks of exposure, and the rates of fatty deposit 

formation under the aortic valve leaflets were evaluated. Exposure of mice on the control diet to 

500 and 800 ppm CS2 induced a small but significant increase in the rate of fatty deposit 

formation over non-exposed controls. In the animals on the high fat diet there was marked 

enhancement of the rate of fatty deposit formation in mice exposed to 500 and 800 ppm over the 

animals on the high fat diet alone. In addition, there was a small but significant enhancement in 

mice exposed to 50 ppm over the rate of fatty deposit formation induced by the high fat diet 

alone. Thus CS2 is atherogenic at high concentrations and in conjunction with other risk factors, 

CS2 at relatively low concentrations can enhance atherogenesis in mice. Fifty ppm is thus the 

study LOAEL. 

Hepatic toxicity has also been induced in rats exposed to relatively high doses of CS2, usually 

following pretreatment with liver inducers such as phenobarbital. Bond et al. (1969) showed that 

high doses of CS2 to rats produced an increase in periportal liver fat, and decreases in hepatic 

cytochrome P450 content and in microsomal mixed function oxidase (MFO) activity. After 

phenobarbital induction, exposed rats exhibited more severe hepatoxicity characterized by 

hydropic degeneration and necrosis. Other hepatotoxic effects seen after CS2 exposures greater 

Appendix D3 70 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

than 400 ppm include increases in relative liver weight (Sokal, 1973), stimulation of liver 

microsomal lipid peroxidation (Wronska-Nofer et al., 1986), and decreases in hepatic cholesterol 

synthesis (Simmons et al., 1988). 

The 24-hr lethal ip LD50 values for CS2 were estimated in 1-, 5-, 10-, 20-, 30- and 40-day-old rats 

(sample size not specified) (Green and Hunter, 1985). 1-day-old rats (LD50 583 mg/kg, ip) were 

about 3-times more susceptible than 20-day-old rats (LD50 1545 mg/kg, ip). 

14C- and 35S-labelled CS2 was given ip to 1-, 5-, 10-, 20-, 30-, and 40-day-old rats (Snyderwine 

and Hunter, 1987). Thirty- and forty-day-old rats (sample size not reported) metabolized 

significantly more CS2 to CO2 and expired significantly less CS2 than 1- to 20-day-old rats. 

Twenty-four hr after administration, up to 13 times more 35S -label (radioactivity per g of tissue) 

were present in organs from 1-day-old rats than in similar organs from 40-day-old rats. The 

study does not specifically address the toxicological implications of the metabolic differences, 

and did not include fully mature animals. However, inability to detoxify CS2 would lead to 

higher tissue concentrations and thus, potentially, increased toxicity. 

The metabolite responsible for CS2 hepatotoxicity is believed to be reactive sulfur atoms that 

covalently bind to cellular macromolecules (Dalvi, 1988). Similarly, the correlation between 

increased lethality (Green and Hunter, 1985) and increasing binding of 35S –label (Snyderwine 

and Hunter, 1987) in younger CS2 -exposed animals is consistent with a role for reactive sulfur. 

Neurotoxicity of CS2 results from the formation of thiourea lysine cross-links between 

neurofilament proteins (DeCaprio et al., 1992; Valentine et al., 1997; Erve et al., 1998). 

New Zealand white rabbits (24 per group) inhaled 0, 60, 100, 300, 600 or 1200 ppm CS2 for 6 

h/d on gestation days 6 to 18 (Pathology Associates, 1991). Developmental toxicity (NOAEL = 

300 ppm; 930 mg/m3) was noted at concentrations lower than those associated with significant 

maternal toxicity (NOAEL = 600 ppm; 1860 mg/ m3

) (Pathology Associates, 1991). The adults 

did have some slight hematological changes at the 600 ppm level, but the authors questioned the 

biological significance of these marginal findings. Reduced fetal body weights were noted at 

600 and 1200 ppm. Cumulative malformations were increased in the 1200 (3720 mg/m3

) but not 

600 ppm group, though there were no significant increases in any specific malformation in any 

group. Maternal effects at 1200 ppm included decreased body weight, ataxia, wheezing, and 

tremors. In an initial range-finding study, exposure to 3000 ppm was associated with significant 

lethality. 

Rats were exposed to 100 mg/m3

 (32 ppm) for 4 hr/d on gestation days 7 and 8, and the embryos 

explanted to culture medium at day 9.5. Growth of explants of 10 treated and 17 control 

embryos was monitored for 44 hours. CS2 at this concentration induced growth retardation in 

treated embryos relative to controls (Zhao et al., 1997). 

In a two-generation study, Tabacova et al. (1983) exposed pregnant Albino rats (30-32 pregnant 

females per group) to 0.03, 10, 100, or 200 mg/m3 (0.01, 3, 32, or 64 ppm) CS2. The two highest 

dose levels were both teratogenic and maternally neurotoxic. There were no significant adverse 

effects in the F1 generation at the 2 low dose levels. However, significant increases in 

teratogenicity were found in the F2 generation at 10 mg/m3

, as well as increased postnatal 

Appendix D3 71 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

neurological effects including hypoactivity, mild ataxia and gait disturbances, hind-limb 

weakness, spinning and tremor (Tabacova et al., 1983). While the overall rate of malformations 

(club foot, hydrocephalus, microcephalus, generalized edema) exhibited a dose-response trend, 

with increased effects in the F2 generation, the specific malformations exhibited a less-consistent 

pattern. For example, while club foot was the predominant malformation in the F1 fetuses 

(occurring at 100 and 200 mg/m3); much lower rates of club foot were noted in the F2 generation 

(including none in the 200 mg/m3 group). Limitations of the study include a lack of information 

on chemical purity and exposure methods, lack of concurrent controls, lack of clear doseresponse trend, and incomplete reporting on the statistical significance of reported behavioral 

effects. 

Wistar albino rats (32 animals per group) were exposed to 50, 100, or 200 mg/m3 (16, 32, or 64 

ppm) CS2 for 8 hours per day throughout gestation. There were no statistically significant results 

in the 50 mg/m3

 group. In the 100 and 200 mg/m3 groups, there were statistically significant 

increases in reduced fetal body weights, and reduced post natal body weights for 21 days, which 

subsequently disappeared. There was an increase in external malformations (hydrocephalus, 

club foot, and tail deformations) at the two higher doses (Tabacova et al., 1978). 

Behavioral effects were examined in the offspring of Lati:CFY rats (8 per group) exposed to 0, 

10, 700, or 2000 mg/m3 CS2 (3, 230, or 640 ppm) for 6 hours per days over days 7 to 15 of 

gestation. The two high doses caused significant perinatal mortality. Avoidance conditioning 

was tested using a bell as a conditional stimulus prior to an electric shock. The animals learned 

to avoid the shock by jumping onto a pole at the sound of the bell. The latency to jump onto the 

pole and errors were measured as a means to evaluate avoidance conditioning in the treated 

versus control animals. The authors reported that there was a dose-related change in avoidance 

conditioning among male pups over the first 15 days (Lehotsky et al., 1985). While the 

magnitude of the effect on avoidance conditioning was greater at all doses relative to controls, 

and at 2000 mg/m3

 compared with 700 mg/m3, the effect was virtually identical between the 10 

and 700 mg/m3. This lack of dose-response effect raises some question about the significance of 

this finding. 

Effects of low (0.03 and 10 mg/m3

; 0.01 and 3 ppm) prenatal exposures (8 hours per day 

throughout gestation) of CS2 were studied in Wistar albino rats. No congenital malformations or 

significant prenatal effects were found in the 9-11 litters evaluated at each dose. Mortality 

during postnatal days 10 through 21 was increased in the 10 mg/m3

 group. Delays in the 

development of visual and auditory function were reported in the higher dose group (Tabacova 

and Balabaeva, 1980). There was no mention of maternal toxicity in this study. 

Several other studies yielded either no teratogenic effects or effects only at maternally toxic 

exposures. Saillenfait et al. (1989) exposed rats via inhalation to 0, 100, 200, 400, or 800 ppm 

CS2 for 6h/d during days 6-20 of gestation. Lower exposures (100 or 200 ppm; 310 or 620 

mg/m3

) were not associated with maternal toxicity or adverse effects on the developing embryo 

or fetus. Higher concentrations (400 or 800 ppm; 1240 or 2480 mg/m3) yielded a significant 

reduction of maternal weight gain as well as reductions of fetal body weight and a low incidence 

of club foot. Significant increases in unossified sternebrae were reported following 800 ppm 

(2480 mg/m3

) exposures. Nemec et al. (1993) reported no teratogenicity or maternal, 

Appendix D3 72 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

developmental, or reproductive toxicity among pregnant CD rats and their offspring following 

exposure to 125 or 250 ppm (388 or 775 mg/m3

) from 2 weeks prior to mating through gestation 

day 19. At 500 ppm, dams had decreased body weight gain and food consumption; decreased 

litter viability but no teratogenic effects were noted. CS2 was not found to be teratogenic or 

embryotoxic following intraperitoneal administration to rats on days 1-15 of gestation (Beliles et 

al., 1980; Hardin et al., 1981). No significant effects were noted in animal inhalation exposures 

(20 to 40 ppm; 62 to 125 mg/m3

 CS2) with either rats on days 1-19 of gestation or rabbits on days 

1-24 of gestation. 

VI. Derivation of Chronic Reference Exposure Level 

Study Johnson et al. (1983) 

Study population 145 occupationally exposed workers and 212 

comparison workers 

Exposure method Discontinuous occupational inhalation exposures 

(mean of 7.6 ppm and range of 0.6 to 16 ppm) 

Critical effects Reduction in motor nerve conduction velocities 

(decreased peroneal nerve MCV and sural 

nerve SVC) 

LOAEL 7.6 ppm 

NOAEL Not observed 

Exposure continuity 8 hr/day, 5 days/week 

Average occupational exposure 2.7 ppm for LOAEL group (7.6 x 10/20 x 5/7) 

Benchmark concentration (BMC05) 6.86 ppm (continuity-weighted exposure of 2.54 

ppm) 

Human equivalent concentration 2.54 ppm for BMC05 (6.86 x 10/20 x 5/7) 

Exposure duration Mean of 12.1 years (SD 6.9 years) 

Subchronic uncertainty factor 1 

LOAEL uncertainty factor Not needed in BMC approach 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 10 

Inhalation reference exposure level 0.3 ppm (300 ppb; 0.8 mg/m3

; 800 μg/m3) 

A benchmark dose analysis was performed on the peroneal MCV data. The NIOSH exposure 

data were regrouped into 8 geometrically spaced dose groups (Table 1). 

Appendix D3 73 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

Table 1. Peroneal MCV data used for benchmark dose modeling 

Exposure 

(ppm-months) Subjects 

Peroneal MCV (m/s) 

Mean Std. Dev. 

3.8 (2 - 6) 32 45.9 3.8 

13.1 (6 - 16) 61 45.8 5.8 

26.5 (16-44) 140 45.4 5.2 

77.3 (44-122) 17 45.6 4.6 

197 (122 – 336) 17 46.8 3.6 

619 (336 – 929) 54 43.9 4.2 

1428 (929-2563) 61 42.9 5.8 

3997 (2563 – 7075) 19 38.6 5.1 

Model fitting was conducted with U.S. Environmental Protection Agency BMDS Benchmark 

Dose Software, Version 1.3. Four continuous data models were compared: linear, polynomial 

(v. 2.1), power (v. 2.1) and hill (v. 2.1) models. All four models adequately fit the data set (Table 

2). 

Table 2. Benchmark dose modeling results 

Model 

MLE05 

(ppm-mo) 

BMC05 

(ppm-mo) p value 

Linear 1245 1005 0.84 

Polynomial 1100 736 0.78 

Hill 1092 670 0.65 

Power 1245 1005 0.58 

The BMC05 from the best-fitting linear model was used. An occupational BMC05 of 6.9 ppm was 

derived by dividing the 1005 ppm-month value by the average exposure duration of 145 months 

(12.1 years). The time-weighted average value was thus 2.5 ppm (6.9 ppm x 10/20 x 5/7). 

Appendix D3 74 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

36

38

40

42

44

46

48

0 500 1000 1500 2000 2500 3000 3500 4000

CS2 Exposure (ppm-months)

Linear Model (0.95 Confidence Level)

BMC05 MLE05

Peroneal

MCV

(m/s)

The U.S. EPA (1995) based its RfC of 700 μg/m3 on the same study but used a BMC10 and 

included a Modifying Factor (MF) of 3 for database deficiencies. The criteria for use of 

modifying factors are not well specified by U.S. EPA. Such modifying factors are not used by 

OEHHA. In addition OEHHA prefers use of a BMC05 since in practice it tends to be closer to 

the NOAEL while the BMC10 is often closer to the LOAEL (OEHHA, 2000). 

For comparison, 50 ppm was a 13 week NOAEL in rats for axonal swelling (Sills et al., 1998). 

The equivalent continuous exposure is 8.9 ppm. Use of an RGDR of 1, an interspecies UF of 3, 

a subchronic UF of 3, and an intraspecies UF of 10 results in a REL of 90 ppb. 

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the REL for carbon disulfide are the use of human data, the observation 

of a dose-response effect, and the duration of exposures. The major uncertainties are the poor 

quantitation of actual exposure magnitude over time and the limited nature of the health effects 

studies which have been conducted. 

VIII. Potential for Differential Impacts on Children's Health 

The data available on the developmental toxicity of carbon disulfide are equivocal. Several 

studies reported that adverse developmental effects are only noted with exposures exceeding 100 

ppm, while Tabacova and Balabaeva (1980) and Lehotsky et al. (1985) reported transient effects 

The data available on the developmental toxicity of carbon disulfide are equivocal. Several 

studies reported that adverse developmental effects are only noted with exposures exceeding 100 

ppm, while Tabacova and Balabaeva (1980) and Lehotsky et al. (1985) reported transient effects 

Appendix D3 75 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels May 2002

at levels as low as 10 mg/m3 (3 ppm). The results of these two studies are not consistent with 

the database as a whole. While further research into behavioral effects of low concentrations of 

CS2 would better clarify the risks associated with such exposures, no adverse effects have been 

reported at concentrations below the REL of 800 μg/m3 (300 ppb). 

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Appendix D3 80 Carbon Disulfide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CARBON TETRACHLORIDE 

(carbon chloride; carbon tet; freon 10; halon-104; methane tetrachloride; necatrine; 

tetrachlorocarbon; tetrachloromethane; tetraform; tetrasol; univerm)

CAS Registry Number: 56-23-5 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 40 μg/m3 (6 ppb)

Critical effect(s) Increased liver weight and hepatic fatty 

infiltration in guinea pigs

Hazard index target(s) Alimentary system; development 

(teratogenicity); nervous system 

II. Physical and Chemical Properties (HSDB, 1995; CRC, 1994) 

Description Colorless liquid 

Molecular formula CCl4

Molecular weight 153.8 g/mol 

Density 1.59 g/cm3

 @ 20°C 

Boiling point 76.7°C 

Melting point −23°C 

Vapor pressure 91.3 torr @ 20°C 

Solubility Soluble in acetone, ethanol, benzene, carbon 

disulfide, slightly soluble in water 

Conversion factor 1 ppm = 6.3 mg/m3

 @ 25°C 

III. Major Uses or Sources

Carbon tetrachloride was formerly used for metal degreasing and as a dry-cleaning fluid, fabricspotting fluid, fire-extinguisher fluid, grain fumigant and reaction medium (DeShon, 1979). 

Carbon tetrachloride is used as a solvent for the recovery of tin in tin-plating waste and in the 

manufacture of semiconductors. It is used in petrol additives, refrigerants, metal degreasing, and 

as a catalyst in the production of polymers. Carbon tetrachloride is also used as a chemical 

intermediate in the production of fluorocarbons and some pesticides (HSDB, 1995). In 1996, the 

latest year tabulated, the statewide mean outdoor monitored concentration of carbon tetrachloride 

was approximately 0.08 ppb (CARB, 1999a). The annual statewide industrial emissions from 

facilities reporting under the Air Toxics Hot Spots Act in California based on the most recent 

inventory were estimated to be 8781 pounds of carbon tetrachloride (CARB, 2000). 

Appendix D3 81 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposure

Kazantzis et al. (1960) evaluated 17 employees of a quartz processing factory who were 

occupationally exposed to 45-100 ppm (284-630 mg/m3

) carbon tetrachloride (CCl4) vapor. 

Fifteen of the 17 workers complained of symptoms including nausea, anorexia, vomiting, 

flatulence, epigastric discomfort or distention, depressive symptoms, headache or giddiness for 

up to 4 months prior to the evaluation. A week after CCl4 concentrations were reduced to 0-

9 ppm with control measures, workers were symptom-free. 

V. Effects of Animal Exposure 

Adams et al. (1952) chronically exposed albino Wistar rats, guinea pigs, albino rabbits and 

rhesus monkeys to 0, 5, 10, 25, 50, 100, 200 and 400 ppm CCl4 for varying duration. For each 

exposure group, two control groups were devised (unexposed and air-exposed controls) 

consisting of animals similar in age, sex, weight and number. The 2 control groups responded 

similarly to the experimental protocol. 

In the 100, 200 and 400 ppm exposure groups (Adams et al., 1952), mortality was excessive with 

moderate to severe liver cirrhosis and other various pathological changes in all the species tested. 

Fifteen male and 15 female rats were exposed to 50 ppm CCl4 134 times for 187 days. They 

experienced decreased body weight gain and liver weight increase as well as moderate fatty 

degeneration and slight to moderate liver cirrhosis. Females showed kidney weight increase and 

four rats showed slight to moderate swelling of the kidney tubular epithelium. Guinea pigs (8 

males and 8 females; 143 exposures in 200 days) showed depressed growth in the first two 

weeks, enlarged livers, moderate fatty degeneration and liver cirrhosis, and increased levels of 

liver total lipids, neutral fat, esterified cholesterol and plasma prothrombin clotting time. 

The rabbit group of 2 males and 2 females, which underwent 155 exposures to 50 ppm in 216 

days, showed slightly depressed growth and increased kidney weights, prolonged plasma 

prothrombin clotting time, and moderate fatty degeneration and cirrhosis of the liver. 

No change was seen in the group of 2 male monkeys exposed 198 times to 50 ppm in 277 days 

(Adams et al., 1952). One monkey experienced depressed weight gain compared to the other 

monkey and the controls, but no other adverse effects were seen with respect to organ weights, 

tissue examination, total liver lipid, blood urea nitrogen, blood non-protein nitrogen, serum 

phosphatase, plasma prothrombin clotting time, phospholipid, neutral fat, and free esterified 

cholesterol. 

At 25 ppm CCl4, 15 male and 15 female rats were exposed 137 times for 191 days. Early growth 

depression in males was observed, although final body weights did not significantly differ from 

the controls. Significant liver weight increase and slight to moderate fatty degeneration 

occurred. Liver lipid content was nearly twice the level of the controls and esterified cholesterol 

was five times that of the controls. For this exposure, phospholipid and neutral fat were not 

measured. Five male guinea pigs were exposed 133 times over 185 days and 5 female guinea 

pigs were exposed 93 times over 126 days. Symptoms included growth depression, liver weight 

Appendix D3 82 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

increase, increased plasma prothrombin clotting time, slight to moderate fatty degeneration, 

twice the level of the control total liver lipid, and five times the control level of esterified 

cholesterol. After 178 exposures to 25 ppm over 248 days, rabbits (2 per sex) showed increased 

liver weights and slight to moderate liver cirrhosis and fatty degeneration. 

Twenty male and 20 female rats were exposed 136 times over a period of 192 days to 10 ppm 

CCl4. These rats exhibited increase in liver weight, slight to moderate fatty degeneration and 

total lipid, neutral fat and esterified cholesterol levels that were twice the control levels. Guinea 

pigs (8 male and 8 female), which were exposed 139 times over 197 days, experienced liver 

weight increase, slight to moderate fatty degeneration without cirrhosis, and increased levels of 

total lipid, neutral fat, and esterified cholesterol. In an additional group of 18 male rats exposed 

13 times to 10 ppm, slight fatty degeneration was seen as early as 17 days. Two male and two 

female rabbits tolerated the same regimen as the guinea pigs and showed no symptoms as a result 

of the exposure. Sixteen additional guinea pigs developed hepatic changes after 12 exposures in 

16 days. 

Twenty-five male and 23 female rats, exposed 145 times over 205 days to 5 ppm CCl4, had no 

adverse effects. Nine male and nine female guinea pigs exposed 143 times over 203 days 

showed a statistically significant increase in the liver weights (females only), but only slightly 

higher liver lipid content. No additional histopathological effects were seen at this level of 

exposure. 

In a more recent study, Prendergast et al. (1967) exposed 15 Long-Evans or Sprague-Dawley 

rats, 15 guinea pigs, 3 rabbits, 2 dogs, and 3 monkeys 30 times to a concentration of 515 +39 

mg/m3 (81.7 ppm) carbon tetrachloride (CCl4) 8 hours a day, 5 days a week, for 6 weeks. (This 

intermittent exposure is equivalent to a continuous exposure to 123 mg/m3

.) Additionally, two 

90 day continuous exposure studies were conducted. One study exposed 15 rats, 15 guinea pigs, 

2 rabbits, 2 dogs and 3 monkeys to 61+5.2 mg/m3 CCl4 and the other exposed 15 rats, 3 rabbits, 

2 dogs and 3 monkeys continuously to 6.1+0.3 mg/m3 CCl4 in inhalation chambers. Control 

groups consisted of 304 rats, 314 guinea pigs, 34 dogs, 48 rabbits and 57 monkeys. All the 

animals’ weights were recorded prior to the study, at monthly intervals throughout the study, and 

at the conclusion of the study. 

During the 6 week study, one monkey died following the 7th exposure, and 3 guinea pigs died 

following the 20th, 22nd, and 30th exposures, respectively. Monkeys, guinea pigs, dogs and 

rabbits all exhibited weight loss. A high percentage of mottled livers was seen in all species 

except dogs. Histopathologic examination of the lungs and livers showed morphological 

changes in all the animals exposed to CCl4 (most prominently the guinea pigs). The guinea pigs 

were the most sensitive species displaying discolored lungs, fatty livers, bile duct proliferation, 

fibrosis, focal inflammatory cell infiltration, hepatic cell degeneration and regeneration, early 

portal cirrhosis, and alteration of lobular structure. Hepatic lipid content in the guinea pigs was 

35.4+10.7% compared to the control value of 11.0+3.6%. Alterations of liver lipid content were 

also observed, to a lesser extent, in the other four species; the most severe alteration occurred in 

the rats, less severe alteration in rabbits and dogs, and the least severe in the monkeys. 

Appendix D3 83 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

During the 61 mg/m3 (9.7 ppm) CCl4 continuous exposure study, 3 guinea pigs died (one each 

after 47, 63, and 71 days). All the monkeys were emaciated and experienced hair loss. 

Depressed body weight increases were seen in all exposed animals compared to the controls. 

Autopsies showed enlarged and/or discolored livers in a high percentage (not given) of monkeys, 

guinea pigs, rabbits, and rats. Rats and guinea pigs showed hepatic fatty acid changes, and a 

moderate reduction in succinic dehydrogenase activity was also evident in guinea pigs. Varying 

but lesser degrees of these changes were also seen in the other species tested. 

The low concentration of 6.1 mg/m3

 (1 ppm) CCl4 was attained by diluting the CCl4 to 10% of 

the above concentration with n-octane, resulting in a solution of 6.1 mg/m3 CCl4 in 61 mg/m3 of 

n-octane (Prendergast et al., 1967)). The level of n-octane used was shown to be nontoxic by an 

n-octane control, which yielded no effects. (The current TLV for n-octane is 1400 mg/m3

 (300 

ppm) (ACGIH, 1992).) No animals died during this study, and no signs of toxicity were noted. 

All exposed animals except the rats showed reduced weight gain when compared to the controls, 

and all species exhibited nonspecific inflammatory lung changes. Guinea pig liver lipid contents 

and serum urea nitrogen concentrations were similar to the control values. In several animals 

there were some nonspecific inflammatory changes in the liver, kidney and heart, but the authors 

did not attribute these to the chemical exposure. There was no other observed hematologic or 

histopathologic toxicity at this level. 

Shimizu et al. (1973) exposed groups of 4 female Sprague-Dawley rats to 10, 50 and 100 ppm of 

CCl4 vapor for 3 hours a day, 6 days a week for up to 6-8 weeks. The rats were terminated two 

days after the last inhalation. Accumulation of CCl4 occurred in the adipose tissue and was 

measured after 1 and 3 weeks of exposure. For the 10 ppm group, accumulation was gradual, 

reaching a level of 1/3 the amount found in the 50 ppm group after 6 weeks. A slight increase of 

triglycerides in the liver (6.2-6.4 mg/g) was observed in the 10 ppm group, but no control group 

was used for comparison. 

The intermittent exposure caused a more pronounced and higher number of change indices to 

occur (34 as opposed to the 17 change indices of the monotonous regimen), indicating a greater 

intensity of liver damage. Changes included a significant decrease in hippuric acid synthesis, 

presence of mitochondrial enzymes (glutamate dehydrogenase and ornithine carbonyl 

transferase) in the blood (indicating severe damage to hepatocytes), significant increase in 

cytoplasmic enzyme activity, and a decrease in the level of cytochrome P-450 in liver tissue. 

The effects seen in the monotonous group were the same variety as those in the intermittent 

group, but were less intense. The content of CCl4 in the blood was similar for both the 

intermittent and monotonous exposure groups. Another test was performed over a period of 27 

days varying the regimen, and therefore the concentration, of intermittent exposure while 

keeping the TWA level of CCl4 stable. Increasing the concentration threefold or fivefold with 

five 10 minute peaks did not potentiate the toxic effects. Varying the regimen tenfold to five 5-

minute peaks (peak exposure 402 mg/m3 (63.8 ppm)) with a time weighted average exposure of 

6.5 ppm (41+1 mg/m3

) did, however, result in more severe liver damage. 

Sakata et al. (1987) exposed 10-15 male Sprague-Dawley rats to <10 ppm CCl4 vapor for 15 

minutes a day, twice a week for 8 weeks. All the rats had chronic liver damage involving 

Appendix D3 84 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

nodular liver surfaces and extensive fibrosis. Researchers also found similar results in rats after 

8 weeks of subcutaneous injections of 0.1 mL of 50% CCl4 solution in olive oil twice a week. 

Ideura et al. (1993) exposed male Wistar rats to CCl4 vapor for 7 minutes, 3 times a week for 6-

10 weeks (concentration unspecified). Six experimental groups of 4-5 rats were used, two of 

which were exposed for 10 weeks, another two for 6 weeks, and two unexposed control groups. 

Following the last exposures, rats were injected with varying amounts of endotoxin (1.0 mL 

lipopolysaccharide (LPS)). The rats were sacrificed 24 hours after the injection and processed 

for histological examination. Examination of the rats’ left kidneys and livers revealed liver 

cirrhosis with destruction of normal structure and massive ascites retention after 10 weeks of 

exposure as compared to the controls. Those exposed for 6 weeks exhibited an increase in 

fibrous tissue. The control groups displayed normal liver structure. Researchers found that rats 

previously resistant to endotoxin became susceptible following CCl4 exposure, which was 

manifested as induced acute renal tubular necrosis in cirrhotic rats. 

Yoshimura et al. (1993) performed a similar experiment to that of Ideura et al. (1992) by 

exposing male Wistar rats for 6 (5 rats) and 10 weeks (5 rats) to 99% CCl4 vapor for 3 minutes a 

day. A control group of 5 rats was given phenobarbitone for 10 weeks. After 24 hours following 

the final exposure, rats were injected with endotoxin. Six weeks of CCl4 exposure caused liver 

fibrosis with bridging fibrosis, while 10 weeks of exposure to CCl4 caused liver cirrhosis and 

destruction of the normal liver architecture. 

Pregnant rats were exposed to 0, 300, or 1000 ppm (0, 1938, or 6460 mg/m3) carbon 

tetrachloride for 7 hours/day on days 6-15 of gestation (Schwetz et al., 1974). Significant fetal 

growth retardation, measured by decreased crown-rump length and body weight, was observed in 

the offspring of the exposed groups (n = 22 litters) compared with controls (n = 43 litters). 

Subcutaneous edema was observed in the 300 ppm group but not in the 1000 ppm group. 

Sternebral anomalies were observed in the 1000 ppm group. 

Appendix D3 85 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Effects of Chronic CCl4 Exposure (Adams et al., 1952) 

Species Concentration (ppm) Group 

size 

Endpoint 

Exposure 

scenario 

(days exposed/ 

experiment 

length) 

Rats 

(male) 

50 ppm 15 liver damage: fatty 

degeneration and cirrhosis; 

growth depression 

134/187 

Rats 

(female) 

50 ppm 15 same effects as males with 

the addition of increased 

kidney weight 

134/187 

Guinea pigs 50 ppm 16 liver damage: fatty 

degeneration and cirrhosis; 

growth depression 

143/200 

Rabbits 50 ppm 4 enlarged kidney; liver 

damage: fatty degeneration 

and cirrhosis; growth 

depression 

155/216 

Monkeys 50 ppm 2 one experienced growth 

depression 

198/277 

Rats 25 ppm 30 liver damage; early growth 

depression 

137/191 

Guinea pigs 

(male) 

25 ppm 5 liver damage: fatty 

degeneration; growth 

depression 

133/185 

Guinea pigs 

(female) 

25 ppm 5 liver damage: fatty 

degeneration; growth 

depression 

93/126 

Rabbits 25 ppm 4 liver damage: fatty 

degeneration; and cirrhosis 

178/248 

Rats 10 ppm 40 liver damage: fatty 

degeneration 

136/192 

Guinea pigs 10 ppm 16 liver damage: fatty 

degeneration 

139/197 

Rats 5 ppm 48 no adverse effects 145/205 

Guinea pigs 

(male) 

5 ppm 9 no adverse effects 143/203 

Guinea pigs 

(female) 

5 ppm 9 liver damage 143/203 

Appendix D3 86 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Data from Guinea Pigs and Rats Exposed to 5 ppm CCl4 for 7 Months (Adams et al., 1952) 

 Å----- g organ weight/g body weight---Æ

Group n BW (g) Lung Heart Liver Kidneys 

Rats, male 

 Unexposed controls 11 336 0.65 0.32 2.38 0.65 

 Air-exposed controls 16 322 0.62 0.31 2.25 0.66 

 5 ppm CCl4 13 336 0.62 0.31 2.23 0.65 

Rats, female 

 Unexposed controls 14 204 0.86 0.38 2.41 0.73 

 Air-exposed controls 17 209 0.76 0.37 2.76 0.76 

 5 ppm CCl4 18 214 0.81 0.38 2.58 0.73 

Guinea pigs, male 

 Air-exposed controls 7 695 0.79 0.27 3.07 0.63 

 5 ppm CCl4 8 669 0.82 0.27 3.14 0.65 

Guinea pigs, female 

 Air-exposed controls 9 611 0.81 0.27 2.58 0.59 

 5 ppm CCl4 6 636 0.78 0.26 2.82* 0.57 

* p =0.004 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Adams et al. (1952) 

Study population 9 male and 9 female guinea pigs 

Exposure method Discontinuous whole-body inhalation 

Critical effects Increase in liver weight and liver lipid content 

in females 

LOAEL 5 ppm 

NOAEL Not observed 

Exposure continuity 7 hours/day, 5 days/week 

Average experimental exposure 1.0 ppm 

Human equivalent concentration 1.7 ppm (gas with systemic effects, based on 

RGDR = 1.7 for lambda (a) : lambda (h) 

(Gargas et al. 1989)) 

Exposure duration 143 exposures over 203 days (7.3 months) 

LOAEL uncertainty factor 3 (mild effect; only in one sex of one species) 

Subchronic uncertainty factor 3 (7.3 mo/6 yr guinea pig life-span = 10.1%) 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.006 ppm (6 ppb; 40 μg/m3

; 0.04 mg/m3) 

Of the 2 adequate chronic inhalation studies available on CCl4, the Adams et al. (1952) study 

was chosen over the Prendergast et al. (1967) study as the key reference for the carbon 

tetrachloride chronic REL. The Adams et al. (1952) experiment was conducted over a longer 

Appendix D3 87 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

duration. In addition, the Adams study contained more specific endpoints of liver damage that 

were consistent with the mechanism of carbon tetrachloride toxicity. Both studies resulted in 

hepatic effects with exposed rats appearing less sensitive than the affected monkeys or guinea 

pigs. 

For comparison, conversion of the oral U.S. EPA RfD value of 0.7 μg/kg/day to an equivalent 

inhalation value by route-to-route extrapolation yields an inhalation REL estimate of 2.5 μg/m3. 

As another comparison, if the 6.1 mg/m3

 continuous exposure in Prendergast et al. (1967) is a 

NOAEL (for rats), the resulting REL estimate would be 60 μg/m3

. If the 6.1 mg/m3 continuous 

exposure is a mild LOAEL, the resulting REL estimate would be 20 μg/m3. 

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the REL for carbon tetrachloride are the chronic exposure study used and 

the target tissue affected. The major uncertainties are the lack of human data, the lack of a 

NOAEL observation, the small sample sizes used, and the lack of comprehensive multiple dose 

studies. 

VIII. References

Adams E, Spencer H, Rowe V, McCollister D, and Irish D. 1952. Vapor toxicity of carbon 

tetrachloride determined by experiments on laboratory animals. Arch. Indust. Hyg. Occup. Med. 

6(1):50-66. 

ACGIH. 1992. American Conference of Governmental and Industrial Hygienists. Documentation 

of the Threshold Limit Values and Biological Exposure Indices. 6th ed. Vol. I. Cincinnati, OH: 

ACGIH. 

CARB. 1999. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Carbon tetrachloride. Available online at http://www.arb.ca.gov/aqd/toxics.htm 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

DeShon ND. 1979. Carbon Tetrachloride. Kirk-Othmer Encyclopedia of Chemical Technology, 

3rd ed. Grayson M, and Eckroth D, eds. New York: John Wiley and Sons, Inc. 5:704-714. [as 

cited by Santodonato J. 1985. Monograph on human exposure to chemicals in the workplace: 

Carbon tetrachloride; PB86-143377; SRC-TR-84-1123.] 

Appendix D3 88 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Gargas ML, Burgess RJ, Voisard DE, Cason GH, and Andersen ME. 1989. Partition coefficients 

of low-molecular-weight volatile chemicals in various liquids and tissues. Toxicol. Appl. 

Pharmacol. 98(1):87-99. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(CD-ROM version). Denver, CO: Micromedex, Inc. (edition expires 7/31/95). 

Ideura T, Yoshimura A, Shirai M, Taira T, and Koshikawa S. 1993. Endotoxin induced acute 

tubular necrosis in cirrhotic rats. Scand. J. Urol. Nephrol. 27:433-439. 

Kazantzis G, Bomford RR, and Oxon DM. 1960. Dyspepsia due to inhalation of carbon 

tetrachloride vapour. Lancet 360-362. 

Prendergrast J, Jones R, Jenkins L, and Siegel J. 1967. Effects on experimental animals of longterm inhalation of trichloroethylene, carbon tetrachloride, 1,1,1-trichloroethane, 

dichlorodifluoromethane, and 1,1-dichloroethylene. Toxicol. Appl Pharmacol 10:270-289. 

Sakata T, Watanabe N, Hobara N, and Nagashima H. 1987. Chronic liver injury in rats by carbon 

tetrachloride inhalation. Bull. Environ. Contam. Toxicol. 38:959-961. 

Santodonato J. 1985. Monograph on human exposure to chemicals in the workplace: Carbon 

tetrachloride; PB86-143377; SRC-TR-84-1123. NTIS. 

Schwetz BA, Leong BK, and Gehring PJ. 1974. Embryo - and fetotoxicity of inhaled carbon 

tetrachloride, 1,1-dichloroethane and methyl ethyl ketone in rats. Toxicol. Appl. Pharmacol. 

28:452-464. 

Shimizu Y, Nagase C, and Kawai K. 1973. Accumulation and toxicity of carbon tetrachloride 

after repeated inhalation in rats. Ind. Health 11:48-54. 

Yoshimura A, Ideura T, Shirai M, Tiara T, Iwasaki S, Kitaoka T, and Koshikawa S. 1993. The 

distribution of 3

H-labeled endotoxin in the kidney of liver cirrhotic rats. Scand. J. Urol. Nephrol. 

27:295-299. 

Appendix D3 89 Carbon Tetrachloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CHLORINATED DIBENZO-p-DIOXINS and 

CHLORINATED DIBENZOFURANS 

(INCLUDING 2,3,7,8-TETRACHLORODIBENZO-P-DIOXIN) 

 (Polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans 

(PCDFs) including 2,3,7,8-Tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) which is the 

principal congener of concern based on toxicity)

CAS Registry Number: 1746-01-6 (TCDD); 5120-73-19 (TCDF) 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.00004 μg/m3

 (40 pg/m3) 

Oral reference exposure level 1 x 10-8 mg/kg/day (10 pg/kg/day) 

Critical effect(s) Increased mortality, decreased weight gain, 

depression of erythroid parameters, increased 

urinary excretion of porphyrins and deltaaminolevulinic acid, increased serum 

activities of alkaline phosphatase, gammaglutamyl transferase and glutamic-pyruvic 

transaminase, gross and histopathological 

changes in the liver, lymphoid tissue, lung and 

vascular tissues in rats. 

Hazard index target(s) Alimentary system (liver); reproductive system; 

development; endocrine system; respiratory 

system; hematopoietic system 

II. Physical and Chemical Properties (HSDB, 1995; 1999) 

Description All are white crystalline powders at 25° C. 

Molecular Formula C12H4Cl4O2 (TCDD) 

Molecular Weight 321.97 g/mol (TCDD) 

Density 1.827 g/ml (estimated for TCDD) 

Boiling Point 412.2°C (estimated for TCDD) 

Melting Point 305-306°C (TCDD) 

Vapor Pressure 1.52 x 10-9 torr at 25°C (TCDD) 

Solubility In water: 19.3 ng/L at 22°C (TCDD) 

Log Kow 

(octanol/water partition coefficient)

6.15-7.28 (6.8 for TCDD) 

Log Koc 

(organic-carbon distribution coefficient)

6.0-7.39 

Henry’s Law Constant 8.1 x 10-5 ATM-m3

/mol 

Appendix D3 90 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

III. Major Uses and Sources 

The chlorinated dioxins and furans are generated as by-products from various combustion and 

chemical processes. PCDDs are produced during incomplete combustion of chlorine containing 

wastes like municipal solid waste, sewage sludge, and hospital and hazardous wastes. Various 

metallurgical processes involving heat, and burning of coal, wood, petroleum products and used 

tires for energy generation also generate PCDDs. Chemical manufacturing of chlorinated 

phenols (e.g., pentachlorophenol), polychlorinated biphenyls (PCBs), the phenoxy herbicides 

(e.g., 2,4,5 T), chlorinated benzenes, chlorinated aliphatic compounds, chlorinated catalysts and 

halogenated diphenyl ethers are known to generate PCDDs as a by-product under certain 

conditions. While manufacture of many of these compounds and formulations has been 

discontinued in the United States, continued manufacture elsewhere in the world combined with 

use and disposal of products containing PCDD by-products results in the inadvertent release of 

PCDDs into the environment. Industrial and municipal processes in which naturally occurring 

phenolic compounds are chlorinated can produce PCDDs; the best example is chlorine bleaching 

of wood pulp in the manufacture of paper products. Additionally, municipal sewage sludge has 

been documented to occasionally contain PCDDs and PCDFs. Annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California based on the 

most recent inventory were estimated to be 0.123 pounds of 2,3,7,8-TCDD, 0.244 pounds of 

1,2,3,4,7,8-hexachlorodibenzodioxin and lesser amounts of other polychlorinated dibenzodioxins 

and dibenzofurans (CARB, 1999). 

IIIa. 2,3,7,8 Tetrachlorodibenzo-p-dioxin Toxic Equivalents 

2,3,7,8-Tetrachlorodibenzo-p-dioxin is considered the most potent congener of the 

polychlorinated dibenzo-p-dioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) 

families of compounds. Potency of PCDD and PCDF congeners correlates with the binding 

affinity to the cytosolic Ah receptor. Structure activity studies have demonstrated that optimal 

biological activity and Ah-receptor binding requires congeners with a planar conformation and 

chlorines at the corners of the molecule at the 2,3,7,8 positions (Poland and Knutson, 1982; Safe, 

1986). Chlorines at both ortho positions in these molecules (i.e., positions 1 and 9) sterically 

hinder a planar conformation that lessens the congeners’ biological activity. Thus only 15 of 210 

different PCDDs and PCDFs congeners possess significant biological activity based on chlorines 

in the 2,3,7,8 positions and some degree of planar conformation (Safe, 1986; U.S. EPA 1989). 

These include two tetrachloro-congeners: 2,3,7,8-tetrachlorodibenzo-p-dioxin and 2,3,7,8- 

tetrachlorodibenzofuran; three pentachloro congeners: 1,2,3,7,8-pentachlorodibenzo-p-dioxin, 

1,2,3,7,8-pentachlorodibenzofuran, and 2,3,4,7,8-pentachlorodibenzofuran; seven hexachloro 

congeners: 1,2,3,4,7,8 or 1,2,3,6,7,8 or 1,2,3,7,8,9-hexachlorodibenzo-p-dioxins and 

hexachlorodibenzofurans and 2,3,4,6,7,8-hexachlorodibenzofuran; and three heptachloro 

congeners: 1,2,3,4,6,7,8-heptachlorodibenzo-p-dioxin, 1,2,3,4,6,7,8-heptachlorodibenzofuran 

and 1,2,3,4,7,8,9-heptachlorodibenzofuran (U.S. EPA, 1989). The structures of the dibenzo-pdioxins and dibenzofurans along with their numbering schemes are shown in Figure 1. Toxic 

equivalents are calculated relative to the most potent congener, 2,3,7,8-tetrachlorodibenzo-pdioxin, and are determined based on structure activity studies examining relative affinity for the 

Appendix D3 91 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Ah receptor as well as on relative toxicity of different congeners. Values for the international 

system of toxic equivalents are provided in Table 1 (U.S. EPA, 1989). 

Table 1. International Toxic Equivalency Factors (I-TEFs) for PCDDs and PCDFs Chlorinated 

in the 2,3,7, and 8 Positions. (U.S. EPA 1989.) 

Compound1,2 I-TEF

Mono-, Di-, and Tri-CDDs and CDFs 0 

TetraCDD

2,3,7,8-substituted 

Others 

1.0 

0 

PentaCDD 

2,3,7,8-substituted 

Others 

0.5 

0 

HexaCDD 

2,3,7,8-substituted 

Others 

0.1 

0 

HeptaCDD

2,3,7,8-substituted 

Others 

0.01 

0 

OctaCDD 0.001 

TetraCDF

2,3,7,8 

Others

0.1 

0 

PentaCDF

1,2,3,7,8-PentaCDF 

2,3,4,7,8-PentaCDF 

others 

0.05 

0.5 

0 

HexaCDF 

2,3,7,8-substituted 

Others 

0.1 

0 

HeptaCDF 

2,3,7,8-substituted 

Others 

0.01 

0 

OctaCDF 0.001 

1

 CDD designates chlorinated dibenzo-p-dioxin 

2

 CDF designates chlorinated dibenzofuran

Appendix D3 92 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Figure 1. Structures of the Dibenzo-p-dioxins and Dibenzofurans

O 

O 

O 

6 

7

9 

8

1 

2 

3 

4 

4 

3 

2 

1 

8

9 

7

6 

D ibenz o-p-dioxin

D ib e n z o f u r an

IV. Effects of Human Exposure 

The information available on possible chronic toxic effects in humans is complicated by the 

relative insensitivity of epidemiological studies, the limited ability of case studies of exposed 

individuals to establish cause and effect relationships, the heterogeneous nature of human 

populations, the broad spectrum of exposures to other toxic agents in the human environment, 

and the episodic exposure of many of the exposed human populations which have been studied 

(e.g., Seveso, Italy). As a result, a limited number of effects have been associated with exposure 

to dioxins in humans. The meaning of these effects in terms of toxicity in most cases remains to 

be clarified. The majority of information comes from cross-sectional medical studies. 

Chloracne is the most widely recognized effect of exposure to 2,3,7,8-TCDD and TCDD-like 

PCDDs and PCDFs. Chloracne is a persistent condition, which is characterized by comedones, 

keratin cysts and inflamed papules and is seen after acute and chronic exposure to various 

chlorinated aromatic compounds (Moses and Prioleau, 1985). Other dermal effects include 

hyperpigmentation and hirsutism or hypertrichosis (Jirasek et al., 1974; Goldman, 1972; Suskind 

et al., 1953; Ashe and Suskind, 1950); both appear to resolve themselves more quickly over time 

than chloracne, making them more of an acute response rather than a chronic response (U.S. 

EPA, 1994a). Epidemiological data available for 2,3,7,8-TCDD have not allowed a 

determination of the threshold dose required for production of chloracne (U.S. EPA, 1994b). 

Case studies suggest that there may be a relationship between 2,3,7,8-TCDD exposure and 

hepatomegaly (Reggiani, 1980; Jirasek et al., 1974; Suskind et al., 1953; Ashe and Suskind, 

1950) and hepatic enzyme changes (Mocarelli et al., 1986; May, 1982; Martin 1984; Moses et 

al., 1984). Nevertheless, cross sectional epidemiological studies of trichlorophenol (TCP) 

production workers (Suskind and Hertzberg., 1984; Bond et al., 1983; Moses et al., 1984; 

Calvert et al. 1992), Vietnam veterans (Centers for Disease Control Vietnam Experience Study, 

1988; Roegner et al., 1991) and Missouri residents (Webb et al., 1989; Hoffman et al., 1986) 

Appendix D3 93 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

found little evidence for an association between exposure and hepatomegaly suggesting that this 

is not a chronic response. There is a consistent pattern of increased levels of serum gamma 

glutamyl transferase in populations exposed to 2,3,7,8-TCDD which is presumably of hepatic 

origin (Mocarelli, 1986; Caramaschi et al., 1981, May, 1982; Martin, 1984; Moses et al., 1984; 

Calvert et al., 1992; Centers For Disease Control Vietnam Experience Study, 1988). Two cross 

sectional studies have associated diabetes and elevated fasting serum glucose levels with 

relatively high serum 2,3,7,8-TCDD levels (Sweeney et al., 1992; Roegner et al., 1991). 

However other studies provided mixed results (Moses et al., 1984; Centers for Disease Control 

Vietnam Experience Study, 1988; Ott et al., 1993). TCDD has been associated with effects on 

reproductive hormonal status in males. The likelihood of abnormally low testosterone levels was 

2 to 4 times greater in individuals with serum 2,3,7,8-TCDD levels above 20 pg/ml (Egeland et 

al. 1994) and increased serum levels of luteinizing hormone and follicle stimulating hormone 

have been documented (Egeland et al., 1994). A number of other effects have been reported that 

were either not seen as chronic effects or effects seen long term in only one population of 

exposed persons. These include elevated liver enzymes (aspartate aminotransferase and alanine 

aminotransferase), pulmonary disorders, neurologic disorders, and changes in porphyrin 

metabolism and kidney disorders (U.S. EPA, 1994c). Areas in which there is presently 

insufficient information to draw solid conclusions include effects on the circulatory system, 

reproductive effects, immunological effects, effects on metabolism and handling of lipids, and on 

thyroid function (U.S. EPA, 1994c). Recent findings in Rhesus monkeys have shown 2,3,7,8-

TCDD to cause endometriosis (Reier et al., 1993) and epidemiological studies are currently 

underway to determine if there is an association between TCDD exposure and endometriosis in 

human populations exposed by the Seveso accident. 

Potential effects of a toxicant on normal fetal development include fetal death, growth 

retardation, structural malformations and organ system dysfunction. Evidence for all four of 

these responses has been seen in human populations exposed to dioxin-like compounds. In 

these poisoning episodes populations were exposed to a complex mixture of halogenated 

aromatic hydrocarbons contained within PCBs, PCDFs and PCDDs mixtures thus limiting the 

conclusions that could be drawn from the data. In the Yusho and Yu-Cheng poisoning episodes, 

human populations consumed rice oil contaminated with PCBs, PCDFs and PCDDs. Yu-Cheng 

women experienced high perinatal mortality in hyperpigmented infants born to affected mothers 

(Hsu et al. 1985). This occurred in women with overt signs of toxicity (chloracne) (Rogan, 

1982) and Rogan notes that, when there is no sign of toxicity in the mother, the likelihood of 

fetotoxicity appears to lessen considerably in the infants. Signs of toxicity from dioxin like 

compounds were absent in infants born to mothers apparently not affected in the Seveso, Italy 

and Times Beach, Missouri, incidents (Reggiani, 1989; Hoffman and Stehr-Green, 1989), which 

supports Rogan’s conclusion. There was an increased incidence of decreased birth weight in 

infants born to affected mothers in the Yusho and Yu-Cheng incidents suggesting fetal growth 

retardation (Wong and Huang, 1981; Law et al., 1981; Lan et al., 1989; Rogan et al., 1988). The 

structural malformation, rocker bottom heel, was observed in Yusho infants (Yamashita and 

Hayashi, 1985) making this malformation a possible result of exposure to dioxin-like 

compounds. Nevertheless, it is unknown if these compounds produce malformations in humans. 

Evidence for possible organ system dysfunction in humans comes from a study of Yu-Cheng 

children which found that children exposed in utero experienced delays in attaining 

developmental milestones, and exhibited neurobehavioral abnormalities (Rogan et al., 1988) 

Appendix D3 94 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

suggesting involvement of CNS function. Dysfunction of dermal tissues is noted in exposed 

infants of the Yusho and Yu-Cheng incidents and is characterized by hyperpigmentation of the 

skin, fingernails, and toenails, hypersecretion of the meibomian glands, and premature tooth 

eruption (Taki et al., 1969; Yamaguchi et al., 1971; Funatsu et al., 1971; Wong and Huang, 

1981; Hsu et al., 1985; Yamashita and Hayashi, 1985; Rogan et al., 1988; Rogan, 1989; Lan et 

al., 1989). 

V. Effects of Animal Exposure 

The toxicity to laboratory animals encompasses a number of areas including changes in energy 

metabolism manifested as wasting syndrome, hepatotoxicity, effects on tissue of epithelial 

origin, various endocrine effects, effects on vitamin A storage and use, immune system effects 

and reproductive and developmental toxicity. The limited number of chronic studies available 

do not examine all these endpoints. Therefore subchronic exposures are included here in order to 

provide a more complete coverage of potential chronic toxic effects of these compounds. 

Wasting syndrome is one of the most broadly occurring toxic effects. The wasting syndrome is 

characterized by loss of adipose tissue and lean muscle mass and is produced in all species and 

strains tested, but there are difference in sensitivity (U.S. EPA 1994d; Peterson et al., 1984; Max 

and Silbergeld, 1987). Numerous studies have not yet established the mechanism of wasting 

syndrome (U.S. EPA, 1994e). Hepatotoxicity is also seen in all species tested, but there is 

considerable variation in species sensitivity (U.S. EPA, 1994d). TCDD induces hyperplasia and 

hypertrophy of liver parenchymal cells. Morphological and biochemical changes in the liver 

include increased SGOT and SGPT, induction of microsomal monooxygenases and proliferation 

of the smooth endoplasmic reticulum, porphyria, increased regenerative DNA synthesis, 

hyperlipidemia, hyperbilirubinemia, hyperchloesterolemia, hyperproteinemia, degenerative and 

necrotic changes, mononuclear cell infiltration, multinucleated giant hepatocytes, increased 

numbers of mitotic figures, and parenchymal cell necrosis (U.S. EPA, 1994d; WHO/IPCS, 

1989). Epithelial effects seen include chloracne (rabbit ear and the hairless mouse) (Jones and 

Krizek, 1962; Schwetz et al., 1973) and hyperplasia and/or metaplasia of gastric mucosa, 

intestinal mucosa, the urinary tract, the bile duct and the gall bladder (U.S. EPA 1994f). TCDD 

exposure results in endocrine like effects including epidermal growth factor like effects such as 

early eye opening and incisor eruption in the mouse neonate (Madhukar et al., 1984), 

glucocorticoid like effects such as involution of lymphoid tissues (U.S. EPA, 1994g; Sunahara et 

al., 1989), alteration in thyroid hormone levels and in some cases thyroid hormone like effects 

(WHO/IPCS, 1989; Rozman et al., 1984), decreases in serum testosterone and 

dihydrotestosterone (Mittler et al., 1984; Keys et al., 1985; Moore and Peterson, 1985), and 

changes in arachidonic acid metabolism and prostaglandin synthesis (Quilley and Rifkind, 1986; 

Rifkind et al., 1990). TCDD is known to decrease hepatic vitamin A storage (Thunberg et al., 

1979). TCDD and other dioxin like PCDDs and PCDFs are potent suppressors of both cellular 

and humoral immune system function, characteristically producing thymic involution at low 

doses and involution of other lymphoid tissues at higher doses (U.S. EPA 1994h). 

In animal studies there is a large body of information available documenting both developmental 

and reproductive toxicity of 2,3,7,8-TCDD and other PCDDs and PCDFs. These compounds are 

Appendix D3 95 Chlorinated dibenzo-p-dioxins 
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acutely toxic to early life stages of fish and birds with fish being most sensitive (LD50 of 0.4 

μg/kg for rainbow trout sac fry eggs and LD50 of 34 ng/kg for lake trout eggs); some species of 

birds are also relatively sensitive (LD50 of 0.25 μg/kg for chicken eggs) (Peterson et al., 1993). 

2,3,7,8-TCDD has been documented to increase the incidence of prenatal mortality in a number 

of species of laboratory animals including the Rhesus monkey, Guinea pig, rabbit, rat, hamster, 

and mouse (Peterson et al., 1993). Exposure to 2,3,7,8-TCDD during gestation produces a 

characteristic set of fetotoxic responses in most laboratory animals which includes: thymic 

hypoplasia, subcutaneous edema, and decreased growth (Peterson et al., 1993). More species 

specific responses include cleft palate formation in the mouse at doses below maternal toxicity 

(Moore et al., 1973; Smith et al., 1976; Couture et al., 1990), intestinal hemorrhage in the rat 

(Sparschu et al., 1971), hydronephrosis in the mouse and hamster (Moore et al., 1973; Smith et 

al., 1976; Couture et al., 1990; Birnbaum et al., 1989; Olson et al., 1990), and extra ribs in the 

rabbit (Giavini et al., 1982). Female rats have also been found to be affected by perinatal 

exposure to 2,3,7,8-TCDD with clefting of the clitoris, incomplete or absent vaginal opening and 

a smaller vaginal orifice after a dose of 1 μg/kg to the mother on day 15 of gestation (Gray et al., 

1993). 

A number of effects on adult reproductive function are seen in male animals exposed in utero to 

2,3,7,8-TCDD. TCDD reduces plasma androgen levels in the adult male rat and perinatal 

exposure decreases spermatogenesis, spermatogenic function and reproductive capability, 

feminizes male sexual behavior, and feminizes male gonadotrophic function (LH secretion) 

(Mably et al., 1991; Mably et al., 1992a,b,c). Evidence suggests that these effects are the result 

of impaired sexual differentiation of the CNS, which in male rats is dependent on exposure of the 

developing brain to testosterone. 

There are numerous studies detailing the effects of the PCDDs, PCDFs and other dioxin like 

compounds, however a large number of these studies were conducted as either acute or 

subchronic exposures, studies in which it is unlikely that body burdens had reached steady state 

levels. Detailed below are three chronic studies that were considered in the setting of a chronic 

toxicity exposure level. 

The most definitive study of chronic toxicity in rats is that of Kociba et al. (1978). This study 

involved the administration of 2,3,7,8-TCDD in the diet at doses of 1 ng/kg/day, 10 ng/kg/day, 

and 100 ng/kg/day to groups of 50 male and 50 female Sprague Dawley rats for two years. A 

group of 86 male and 86 female rats received diet with solvent vehicle alone and served as 

controls. The following observations (excluding carcinogenic effects) were seen at the 100 

ng/kg/day dose: increased mortality, decreased weight gain, depressed erythroid values, 

increased urinary excretion of porphyrins and delta-aminolevulinic acid, and increased serum 

activities of alkaline phosphatase, gamma-glutamyl transferase, and glutamic-pyruvic 

transaminase. Histopathologic changes were noted in the liver, lymphoid tissue, respiratory and 

vascular tissues. The primary ultrastructural change in the liver was proliferation of the rough 

endoplasmic reticulum. At the 10 ng/kg/day dose the severity of toxic symptoms was less than 

that of the 100 ng/kg/day dose and included increased urinary excretion of porphyrins in females 

as well as liver and lung lesions. The 1 ng/kg/day dose produced no discernible significant toxic 

effects. Interpretation of this study by the authors was that the 1 ng/kg/day dose was a NOAEL. 

Appendix D3 96 Chlorinated dibenzo-p-dioxins 
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Two chronic toxicity studies are available in the mouse. The first is a one year study conducted 

by Toth et al. (1979) using male Swiss mice administered weekly oral doses of 7, 700, and 7000 

ng/kg/day. In this study 2,3,7,8-TCDD administration resulted in amyloidosis and dermatitis in 

0 of 38 control animals, 5 of 44 animals receiving 7 ng/kg/day, 10 of 44 animals receiving 700 

ng/kg/day and 17 of 43 animals receiving 7,000 ng/kg/day. The other study was from the NTP 

1982 gavage study (NTP, 1982) in B6C3F1 mice. This study employed groups of 50 male and 

50 female mice. The males received doses of 0, 10, 50, and 500 ng/kg/week by gavage for two 

years while female mice received doses of 0, 40, 200, and 2000 ng/kg/week by gavage for two 

years. No adverse effects were seen at the lowest doses tested in each sex, which correspond to 

NOAELs of approximately 1.4 and 6 ng/kg/day for males and females, respectively. Neither 

chronic toxicity study in mice reported data on enzyme activity. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Kociba et al. (1978) 

Study population Sprague-Dawley rats of both sexes (50/treatment 

group/sex) 

Exposure method Continuous dietary exposure starting at seven 

weeks of age for 2 years 

Critical effects Increased mortality, decreased weight gain, 

depression of hematologic measures, 

increased urinary excretion of porphyrins and 

delta-aminolevulinic acid, increased serum 

activities of alkaline phosphatase, gammaglutamyl transferase and glutamic-pyruvic 

transaminase, gross and histopathological 

changes in the liver, lymphoid tissue, lung and 

vascular tissues 

Observed LOAEL 210 ppt in diet (0.01 μg/kg/day)

Observed NOAEL 22 ppt in diet (0.001 μg/kg/day)

Exposure continuity Continuous exposure via the diet 

Exposure duration 2 years 

Subchronic uncertainty factor 1 

LOAEL uncertainty factor 1 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Oral reference exposure level 10 pg/kg/day 

Route-to-route extrapolation 3,500 μg/m3

 per mg/kg/day 

Inhalation reference exposure level 40 pg/m3

 (0.00004 μg/m3) 

The data available for chronic toxic effects in humans have a number of limitations. Some 

studies did not determine the body burden of compounds necessary to estimate dose.; The Yusho 

and Yu-Cheng poisoning episodes have uncertainty because exposure was to complex mixtures 

of halogenated aromatic hydrocarbons rather than to individual congeners. And epidemiological 

Appendix D3 97 Chlorinated dibenzo-p-dioxins 
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studies and case studies have limitations in determining cause and effect relationships. 

Therefore, an animal study was chosen for determination of a NOAEL/LOAEL. The study 

chosen for use was that of Kociba et al. (1978), based on the duration of the study (2 years), the 

number of animals employed (50 per treatment group per sex), testing of both sexes, a dose 

range, which spanned from an apparent NOAEL to severe hepatic effects including carcinogenic 

effects, a complete histopathological examination of all organ systems, examination of urinary 

excretion of porphyrins and delta-aminolevulinic acid, and determination of serum activities of 

alkaline phosphatase, gamma-glutamyl transferase, and glutamic-pyruvic transaminase. The 

elevation of human serum values for gamma-glutamyl transferase is one of the consistently seen 

chronic responses in exposed human populations and reflects changes in liver biochemistry. 

Thus the examination of markers of liver toxicity also altered in animal models of chronic 

toxicity make the Kociba study an appropriate choice for detecting potential chronic toxic effects 

of 2,3,7,8-TCDD in humans. The NOAEL in the Kociba et al. (1978) study was determined to 

be 1 ng/kg body weight/day. For the purposes of determining the REL the 1 ng/kg/day dose was 

considered to be a NOAEL based upon the observations of Kociba et al. (1978). 

VII. Data Strengths and Limitations for Development of the REL

NOAELs from a number of other studies compare favorably with the 1 ng/kg/day NOAEL. 

These include the NOAEL from the NTP (1982) study in B6C3F1 mice and the NOEL for 

enzyme induction in rats and marmosets calculated by Neubert (1991) of 1 ng/kg. Furthermore 

the 1 ng/kg/day NOAEL is lower than the LOAELs observed by Toth et al. (1979) of 7 

ng/kg/day in mice and by Schantz et al. (1978) of 2.3 ng/kg/day in rhesus monkeys. Current 

exposure assessments for 2,3,7,8-TCDD and other dioxin-like compounds including the PCBs, 

PCDDs, and PCDFs estimate that the average daily background dose in the U.S. is 3-6 pg 

TEQ/kg/day (U.S. EPA 1994i) also placing the REL close to background exposures. The REL 

of 10 pg/kg/day should be protective of chronic effects on liver function and avoid significant 

increases in exposure over the background level of human exposure. 

The strengths of the inhalation REL include the availability of chronic exposure data from a 

well-conducted study with histopathological analysis, the observation of a NOAEL, and the 

demonstration of a dose-response relationship. Major areas of uncertainty are the lack of 

adequate human exposure data and the lack of chronic inhalation exposure studies. 

Appendix D3 98 Chlorinated dibenzo-p-dioxins 
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Safe SH. 1986. Comparative toxicology and mechanism of action of polychlorinated dibenzo-pdioxins and dibenzofurans. Annu. Rev. Pharmacol. Toxicol. 26: 371-398. 

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1973. Toxicology of chlorinated dibenzo-p-dioxins. Environ. Health Perspect. 5: 87-99. 

Schantz SL, Barsotti DA, and Allen JR. 1978. Toxicological effects produced in nonhuman 

primates chronically exposed to fifty parts per trillion 2,3,7,8-tetrachlorodibenzo-p-dioxin 

(TCDD). Toxicol. Appl. Pharmacol. 48(1): A180. 

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in CF-1 mice. Toxicol. Appl. Pharmacol. 38: 517-523. 

Sparschu GL, Dunn FL, and Rowe VK. 1971. Study of the teratogenicity of 2,3,7,8-tetrachlorodibenzop-dioxin in the rat. Food Cosmet. Toxicol. 9: 405-412. 

Sunahara GI, Lucier G, McCoy Z, Bresnick EH, Sanchez ER, and Nelson KG. 1989. 

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surveys at Monsanto Chemical Co., Nitro, West Virginia, 1953. Cincinnati, OH: Department of 

Environmental Health, University of Cincinnati (unpublished). 

Suskind RR, and Hertzberg VS. 1984. Human health effects of 2,4,5-T and its toxic 

contaminants. JAMA 251:2372-2380. 

Appendix D3 103 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Sweeney MH, Hornung RW, Wall DK, Fingerhut MA, and Halperin WE. 1992. Prevalence of 

diabetes and increased fasting serum glucose in workers with long-term exposure to 2,3,7,8-

tetrachlorodibenzo-p-dioxin. Presented at: 12th International Symposium on Dioxins and Related 

Compounds; August 24-28; Tampere, Finland. 

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women and their fetuses. Fukuoka Acta Med. 60: 471-474 (Japan). 

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549. 

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chlorinated dibenzo-p-dioxins and dibenzofurans (CDDs and CDFs) and 1989 update. 

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(TCDD) and Related Compounds. Office of Health and Environmental Assessment, Office of 

Research and Development, United States Environmental Protection Agency, Washington, D.C. 

Vol 3:9-8 to 9-12. 

U.S. EPA. 1994a. ibid. Vol 2:7-107. 

U.S. EPA. 1994b. ibid. Vol 2:7-101. 

U.S. EPA. 1994c. ibid Vol 2:7-238. 

U.S. EPA. 1994d. ibid Vol 1:3-17. 

U.S. EPA. 1994e. ibid Vol 1:3-14. 

U.S. EPA. 1994f. ibid Vol 1:3-6. 

U.S. EPA. 1994g. ibid Vol 1:3-25. 

U.S. EPA. 1994h. ibid Vol 1:3-4-1. 

U.S. EPA. 1994i. ibid Vol 3:9-86. 

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known body levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin. J. Toxicol. Environ. Health 28: 183-

193. 

Appendix D3 104 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

WHO/IPCS. 1989. World Health Organization/International Programme on Chemical Safety. 

Polychlorinated dibenzo-p-dioxins and dibenzofurans. Environmental Health Criteria 88. 

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Japanese). 

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retardation and possible alteration in calcium metabolism. Environ. Health Perspect. 59: 41-45. 

Appendix D3 105 Chlorinated dibenzo-p-dioxins 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CHLORINE 

CAS Registry Number: 7782-50-5 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.2 μg/m3

 (0.08 ppb) 

Critical effect(s) Hyperplasia in respiratory epithelium in female 

rats

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1995; 1999 except as noted) 

Description Yellow/green gas 

Molecular formula Cl2

Molecular weight 70.906 (Weast, 1989) 

Density 2.9 g/L @ 25°C and 1 ATM 

Boiling point -34.04° C 

Vapor pressure 5 atm @ 10.3°C; 5830 torr @ 25°C 

Solubility Slightly soluble in water 

(310 mL per 100 mL water at 10° C; 

 1.46 g per 100 mL water at 0° C) 

Conversion factor 1 ppm = 2.9 mg/m3

 @ 25° C 

III. Major Uses and Sources

In an industrial setting, chlorine is widely used as an oxidizing agent in water treatment and 

chemical processes. Chlorine is also used to disinfect swimming pool water. Chlorine gas is 

sometimes used at large public pools while household pools typically use hypochlorite solutions. 

Chlorine is an integral part of the bleaching process of wood pulp in pulpmills, although chlorine 

dioxide is replacing this use of chlorine. Chlorine as sodium hypochlorite is commonly used as a 

household cleaner and disinfectant (HSDB, 1995). The annual statewide industrial emissions 

from facilities reporting under the Air Toxics Hot Spots Act in California based on the most 

recent inventory were estimated to be 244,955 pounds of chlorine (CARB, 1999).

IV. Effects of Human Exposure

Shi and associates (1990) evaluated 353 workers from a diaphragm cell chlorine chemical plant. 

The workers ranged in age from 23-52 years with an average of 42.4 years. Two groups were 

compiled with respect to the workers’ length of exposure in years. Group A consisted of 220 

workers who were employed/ exposed for 10-25 years. Group B consisted of 133 workers 

Appendix D3 106 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

employed for less than 10 years. Both groups of workers were exposed to a range of 2.60-

11.0 mg/m3 (0.37-1.75 ppm) chlorine. The control group’s average age was 39.7 years (ranging 

from 26-55 years), and it consisted of 192 workers not exposed to chlorine, but working within 

the same plant. For all the groups, respiratory symptoms and smoking habits were evaluated as 

well as clinical examinations, ENT examinations, chest x-rays and pulmonary function tests. 

Groups A and B showed 3-8 times higher incidence of upper airway complaints than the control 

workers. Current smokers in groups A and B experienced the highest incidence of pulmonary 

symptoms and group A workers had a higher prevalence of rhino-pharyngeal signs than the 

control workers. Abnormalities in chest x-rays were seen in 8.6% of group A workers and in 

2.8% of group B workers, compared to 2.3% of the control workers. Groups A and B showed 

significantly impaired pulmonary function in tests of V50/H and FEF25-75 (forced expiratory flow 

between 25 and 75% of forced vital capacity (FVC), the total amount of air the subject can expel 

during a forced expiration) - compared with the control group, and group A showed reduced 

FEV1 (forced expiratory volume in 1 second) results compared to the control group. 

Kennedy et al. (1991) compared 321 pulpmill workers (189 of whom were exposed to chlorine 

or chlorine dioxide “gassings”) to a control group of 237 rail yard workers in similar working 

conditions but not exposed to chlorine (79% and 84% respective participation rates). The 

workers had been employed for an average of 13 years at the pulpmill and 12.7 years at the rail 

yard. Chlorine gas and chlorine dioxide levels were measured together over a 4 week period 

during mainly a 12 hour shift. Time weighted averages (TWA) were <0.1 ppm, with the highest 

of <0.1-0.3 ppm. A significantly higher prevalence of wheezing was seen in pulpmill workers 

(both smokers and nonsmokers) who had reported more than one episode of chlorine “gassing” 

as compared to the rail yard workers and pulpmill workers with no chlorine gas exposure. More 

airflow obstruction was observed in exposed workers in spite of their nonsmoking and exsmoking status, correlating to significantly lower average values for MMF (maximal midexpiratory flow) and for the FEV1 to FVC ratio. Comparison of pulpmill workers exposed to 

chlorine and /or chlorine dioxide with those pulpmill workers not exposed, suggests that chronic 

respiratory health impairment is associated with exposure to chlorine and/or chlorine dioxide. 

These researchers hypothesized that after the first high exposure incident, an inflammatory 

response occurred in small airways and that this reaction did not resolve in those workers who 

were continuously or repeatedly exposed to the irritant. It was also suggested that chronic 

airflow obstruction caused by repeated minor exposures led to chronic respiratory disability in 

some of the workers. 

Patil et al. (1970) evaluated the exposure of 332 male diaphragm cell workers to 0.006-1.42 ppm 

chlorine gas (a range with a time-weighted average of 0.146 +0.287; most workers were exposed 

to less than 1 ppm). A control group consisting of 382 workers from 25 representative chlorine 

manufacturing plants was also studied. Both groups were comprised of men between the ages of 

19-69 with a mean age of 31.2 +11.0 years. Physical examinations (blood and urine analysis, 

chest x-rays and electrocardiograms) were conducted, in most cases, within the first six months 

of the study year. At two month intervals, each plant was surveyed and chlorine levels were 

determined. Exposed employees were grouped according to job classification. Researchers 

found the average number of exposure years for the study group to be 10.9 + 2.8 years and 

concluded that the exposure level had no correlation to the number of years exposure. Ninetyeight of the 332 workers were found to have abnormal teeth and gums, but no dose-response 

Appendix D3 107 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

relationship was concluded. Similarly, no dose-response relationships were shown with the 

symptoms of sputum production, cough, dyspnea, history of frequent colds, palpitation, chest 

pain, vital capacity, maximum breathing capacity and forced expiratory volume. Any 

deterioration in pulmonary function was shown to be age related. Of the 332 exposed workers, 

9.4% experienced abnormal EKGs. 8.5% of the control group showed the same abnormalities, 

but this difference was not significant. Above 0.5 ppm, an increase appeared in the incidence of 

fatigue. No neurological defects developed and there was no noted prolonged anoxia as a result 

of the chlorine exposure. Also, no consistent gastrointestinal trouble or abnormal incidence of 

dermatitis was found. Exposed workers showed elevated white blood cell counts and decreased 

hematocrit values compared to the control group. 

Bherer et al. (1994) conducted a follow up study of the Quebec pulp mill research done by 

Courteau and associates over a time interval of 18-24 months after the incidents of repeated 

exposures. Fifty-eight of the original 289 exposed workers from the moderate to high risk group 

were studied for developing reactive airways dysfunction syndrome (RADS). Workers at a 

moderate risk were defined as having shortness of breath after their most significant exposure, 

but not at the time of the initial study by Courteau et al. Moderate risk workers also had a record 

of other significant medical conditions and/or were 50 years of age or older. High risk workers 

were defined as those experiencing shortness of breath that continued one month after the 

exposure and/or abnormal lung sounds. Ninety percent of the follow up group completed 

questionnaires which revealed a 91% incidence of respiratory symptoms. Spirometry 

assessments and methacholine inhalation tests were conducted on 51 of the 58 workers. Twentythree percent of the 58 workers still experienced bronchial obstruction and 41% continued to 

have bronchial hyper-responsiveness. Lower baseline FEV1 was seen in those with a lower 

PC20, and 52% of these workers showed an FEV1 < 80% predicted. 

Enarson et al. (1984) compared 392 pulpmill workers exposed to chlorine (unspecified duration) 

to a comparable group of 310 rail yard workers living in the same community, but not exposed to 

chlorine. In the pulpmill areas surveyed that predominantly had significant chlorine gas levels 

(machine room and bleach plant), workers were exposed to either an average of 0.02 ppm or 

0.18 ppm Cl2 respectively. Of the machine room workers, 23.2% experienced a cough as did 

32.8% of those in the bleach plant, compared to 22.3% of the control rail yard workers. Chest 

tightness occurred in 31.5% of the machine room workers and 39.6% of the bleach plant workers 

as compared to 21.3% of the control. Only data from Caucasian subjects were reported. 

Chester et al. (1969) evaluated 139 workers occupationally exposed to <1 ppm chlorine for an 

unspecified duration. Fifty-five of the 139 workers were exposed to additional accidental high 

concentrations of chlorine, which were severe enough to require oxygen therapy. Ventilation 

was affected by chlorine inhalation, with a decrease in the maximal midexpiratory flow (MMF). 

Smokers in this group had significantly reduced FVC, FEV1 and MMF compared to 

nonsmokers. Fifty-six of the 139 subjects showed abnormal posteroanterior chest films, 49 of 

which had parenchyma and/or hilar calcifications consistent with old granulomatous disease and 

11 of which had multiple, bilateral and diffuse calcifications. Researchers suggest that the first 

ventilation function affected in obstructive airway disease is MMF. 

Appendix D3 108 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

V. Effects of Exposure to Animals 

Wolf et al. (1995) exposed male and female B6C3F1 mice and F344 rats to chlorine gas 

concentrations of 0 ppm, 0.4 ppm, 1.0 ppm and 2.5 ppm. The exposures were carried out for 104 

weeks at 6 hr/day 3 days/week for female rats and 6 hr/day 5 days/week for mice and male rats. 

Based on previous studies, the authors determined that female rats could not tolerate 5 days/week 

exposure to chlorine. Each treatment group contained 320 male and 320 female mice. The rats 

were studied in groups of 70, yielding 280 per gender per species. For the first 13 weeks of 

observation, body weights and clinical observations were noted weekly, and for the remainder of 

the study, they were recorded once every two weeks. After 52 weeks, 10 rats were euthanized 

and autopsied. Organ weights were recorded, and hematologic and clinical chemistry parameters 

were determined. These same measurements were performed on all of the surviving mice and 

rats at the conclusion of the 104 weeks. Male mice exposed to 1.0 and 2.5 ppm Cl2 showed 

decreased weight gain compared to controls while only female mice exposed to 2.5 ppm Cl2

showed decreased weight gain. Male rats showed decreased weight gain at all levels of exposure 

while female rats showed the same result at only 1.0 and 2.5 ppm Cl2 exposures. Various 

nonneoplastic nasal lesions were seen in all the airway epithelial types in the nose and at all 

levels of exposures for both species. These lesions were evaluated against background lesions 

found in the control animals. A statistically significant incidence of fenestration was seen in all 

three exposure concentrations of Cl2. Statistically significant responses were seen in the 

traditional and respiratory epithelial regions of all exposed rats and mice. Statistically significant 

damage to olfactory epithelium occurred in all exposed rats and female mice and also in the 1.0 

and 2.5 ppm exposed groups of male mice. 

Klonne et al. (1987) exposed 32 male and female rhesus monkeys to chlorine gas for one year to 

measured concentrations of 0, 0.1, 0.5, and 2.3 ppm Cl2. These monkeys were exposed to 

chlorine for 6 hours/day, 5 days/week. The monkeys were evaluated periodically on the basis of 

body weight, electrocardiograms, neurologic examinations, pulmonary function, hematologic 

parameters, serum chemistry, urinalysis, and blood gas and pH levels. Results were compared 

to the same test measurements recorded prior to the study. No significant difference was seen in 

body weight at any point in the experiment. Ocular irritation (tearing, rubbing of the eyes, 

reddened eyes) was observed after 6 weeks of exposure in the 2.3 ppm group. No exposurerelated differences were seen in neurologic examinations, electrocardiograms, clinical chemistry, 

urinalysis, hematology or blood gas levels. Also, no exposure-related changes were observed in 

the parameters of ventilation distribution. Pulmonary function evaluations yielded a statistically 

significant trend for increasing pulmonary diffusing capacity and distribution of ventilation 

values for males and females in the 2.3 ppm exposure group. Both males and females of the 2.3 

ppm group exhibited statistically significant increased incidence of respiratory epithelial 

hyperplasia. A mild form of the lesions was also seen in the 0.5 ppm group, 0.1 ppm group 

(females only) and one male in the control group. Two parasitic infections occurred, affecting 

the respiratory tract and resulting in 11 monkeys housing parasites and/or ova. Additionally, 16 

monkeys displayed histologic changes characteristic of the presence of the parasites. However, 

the parasitic induced lesions were not associated with lesions in the respiratory epithelium. 

Appendix D3 109 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Wolf et al., 1995 

Study population Female F344 rats (70 per group) 

Exposure method Discontinuous whole-body inhalation exposure 

(0, 0.4, 1.0 or 2.5 ppm) 

Critical effects Upper respiratory epithelial lesions (see 

following table) 

LOAEL 0.4 ppm 

NOAEL Not established 

BMC05 0.14 ppm 

Exposure continuity 6 hours/day, 3 days/week (MWF) 

Average experimental exposure 0.015 ppm 

Human equivalent concentration 0.0024 ppm (gas with 

 extrathoracic respiratory effects, 

 RGDR = 0.16 based on BW = 229 g, 

 MV = 0.17 L/min, SA(ET) = 15 cm2) 

Exposure duration 2 years 

LOAEL uncertainty factor (not needed in the BMC approach) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.08 ppb (0.20 μg/m3

) 

A benchmark dose analysis was performed using a log-normal probit analysis (Tox-Risk, version 

3.5; ICF-Kaiser Inc., Ruston, LA) of the female rat data. Using the data for glandular epithelial 

eosinophilic proteinaceous accumulation (see Table 1 below) to derive the BMC05 resulted in a 

3-fold lower value than the LOAEL of 0.4 ppm, or BMC05 = 0.14 ppm. (Adequate benchmark 

dose estimates could not be obtained for the other nasal lesions due to high background rates and 

shallow dose-response relationships.) A BMC05 is considered to be similar to a NOAEL in 

estimating a concentration associated with a low level of risk. 

The Wolf et al. (1995) study of mice and rats was chosen as the key reference for the chlorine 

chronic REL for several reasons. First, the duration of the experiment was for a full lifetime of 

two years. Second, the sample sizes were large (280 per sex per species). Finally, appropriate 

sensitive endpoints of respiratory epithelial damage were examined. The mice and male rats 

were exposed to chlorine for 6 hours/day, 5 days/week, but the female rats were only exposed for 

3 days/week as the authors observed the females to be more sensitive than the males. Table 1 

shows the histological findings of the female rats. Statistically significant results (p < 0.05) were 

seen for all the tissues at 0.4 ppm chlorine exposure and above. 

Appendix D3 110 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Table 1. Female Rat Epithelial Lesions following Chronic Chlorine Exposure 

 (based on Table 5 of Wolf et al., 1995) 

Tissues 0 ppm 0.4 ppm 1.0 ppm 2.5 ppm 

Goblet cell 

hyperplasia 

3/70 (4%) 50/70 (71%) 63/70 (90%) 64/70 (91%) 

Respiratory 

epithelium 

eosinophilic 

proteinaceous 

accumulation 

49/70 (70%) 60/70 (85%) 59/70 (84%) 65/70 (93%) 

Glandular 

epithelium 

eosinophilic 

proteinaceous 

accumulation 

16/70 (23%) 28/70 (40%) 52/70 (75%) 53/70 (76%) 

Olfactory 

epithelium 

eosinophilic 

proteinaceous 

accumulation 

36/70 (52%) 64/70 (91%) 69/70 (99%) 69/70 (99%) 

The Wolf et al. (1995) study was chosen over the Klonne et al. (1987) monkey study for the 

following reasons: the monkeys were exposed for only one year of their total 35 year lifetime, 

and the sample sizes were considerably smaller (4 monkeys per sex per group) than the mouse 

and rat groups (280 per sex per species). Although the exposure durations differed between the 

two studies, the histological results were similar, differing only slightly in the region of 

occurrence. The monkeys displayed both tracheal and nasal lesions. Both the rodents and the 

monkeys showed upper respiratory epithelial lesions, thus suggesting that the rodents may be an 

appropriate model for humans. 

For comparison with the proposed REL of 0.08 ppb (0.2 μg/m3) using the BMC approach, we 

estimated a REL of 0.02 ppb (0.06 μg/m3

) based on the same rat study but using the NOAEL/UF 

approach with a LOAEL of 0.4 ppm divided by a total UF of 300 (10 for LOAEL, 3 for 

interspecies, and 10 for intraspecies) and the RGDR of 0.16. As another comparison, using 0.1 

ppm as a LOAEL for respiratory epithelial lesions in female monkeys, the LOAEL can be timeadjusted to an equivalent continuous value of 24 ppb. Applying a UFL of 3 for a mild effect, a 

UFS of 10 since it was only a 6 month study, an interspecies UF of 3 for monkeys, and an 

intraspecies UF of 10 results in an estimated REL of 0.02 ppb (0.06 μg/m3). 

The human studies were examined for possible use in the calculation of a REL. The studies were 

limited by very variable exposures (e.g., Patil et al. (1970)), the presence of serious adverse 

health effects in some workers (chest x-ray abnormalities in Shi (1990), abnormal teeth and 

gums in 98 of 332 workers in Paril et al. 1970)), exposure to other compounds such as chlorine 

dioxide (Kennedy et al. (1991)), multiple acute “gassings” with chlorine (Kennedy et al. (1991)), 

and absence of data on cigarette smoking, also a respiratory system irritant. As an illustration of 

what would be estimated, the study of Shi (1990) had a mean workplace exposure of 4.82 mg/m3

(1.7 ppm). This LOAEL was time adjusted to an equivalent continuous exposure of 1.72 mg/m3

, 

Appendix D3 111 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

then divided by an uncertainty factor of 100 (10 for use of a LOAEL and 10 for intraspecies 

variability) to yield a REL of 20 μg/m3 (7 ppb). However, the use of a LOAEL default 

uncertainty factor of 10 does not seem adequate for frank, possibly irreversible effects such as 

the chest x-ray abnormalities reported. There is currently no methodology to deal with such 

effects in REL development. 

Adequate benchmark dose estimates could not be obtained for the other nasal lesions due to high 

background rates and shallow dose-response relationships. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for chlorine include the availability of chronic multiple-dose 

inhalation exposure data from a recent (1995), well-conducted animal study with 

histopathological analysis. Major areas of uncertainty are the lack of adequate human exposure 

data, the lack of observation of a NOAEL, and limited reproductive toxicity data. 

VIII. References

Bherer L, Cushman R, Courteau JP, Quevillon M, Cote G, Bourbeau J, L’Archeveque J, Cartier 

A, and Malo JL. 1994. Survey of construction workers repeatedly exposed to chlorine over a 

three to six month period in a pulpmill: II. Follow up of affected workers by questionnaire, 

spirometry, and assessment of bronchial responsiveness 18 to 24 months after exposure ended. 

Occup. Environ. Med. 51:225-228. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Chang-Yeung M, Lam S, Kennedy SM, and Frew AJ. 1994. Persistent asthma after repeated 

exposure to high concentrations of gasses in pulpmills. Am. J. Respir. Crit. Care Med. 149:1676-

1680. 

Chester EH, Gillespie DG, and Krause FD. 1969. The prevalence of chronic obstructive 

pulmonary disease in chlorine gas workers. Am. Rev. Respir. Dis. 99(3):365-373. 

Courteau JP, Cushman R, Bouchard F, Quevillon M, Chartrand A, and Bherer L. 1994. Survey 

of construction workers repeatedly exposed to chlorine over a three to six month period in a 

pulpmill: I. Exposure and symptomatology. Occup. Environ. Med. 51(4):219-224. 

Donnelly SC, and Fitzgerald MX. 1990. Reactive airways dysfunction syndrome (RADS) due to 

chlorine gas exposure. Ir. J. Med. Sci. 159:275-277. 

Enarson D, Johnson A, Block G, Schragg K, Maclean L, Dybuncio A, Chan-Yeung M, and 

Grzybowski S. 1984. Respiratory health at a pulpmill in British Columbia. Arch. Environ. Health 

39:325-330. 

Appendix D3 112 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Gautrin D, Boulet LP, Boutet M, Dugas M, Bherer L, L’Archeveque J, Laviolette M, Cote J, and 

Malo JL. 1994. Is reactive airways dysfunction syndrome a variant of occupational asthma? J. 

Allergy Clin. Immunol. 93:12-22. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (edition expires 7/31/95). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Kennedy SM, Enarson DA, Janssen RG, and Chan-Yeung M. 1991. Lung health consequences 

of reported accidental chlorine gas exposures among pulpmill workers. Am. Rev. Respir. Dis. 

143:74-79. 

Klonne D, Ulrich C, Riley M, Hamm T, Morgan K, and Barrow C.1987. One-year inhalation 

toxicology study of chlorine in rhesus monkeys (Macaca mullata). Fundam. Appl. Toxicol. 

9:557-572. 

Patil L, Smith R, Vonwald A, and Mooney T. 1970. The health of diaphragm cell workers 

exposed to chlorine. Am. Ind. Hyg. Assoc. J. 31:678-686. 

Rea W, Ross G, Johnson A, Smiley R, Sprague D, Fenyves E, and Samandi N. 1989. 

Confirmation of chemical sensitivity by means of double-blind inhalant challenge of toxic 

volatile chemicals. Clin. Ecol. 6:113-118. 

Shi Z. 1990. Effects of long-term exposure to low concentration of chlorine on workers’ health. 

In: Sakurai H; Okazaki I; Omae K, eds. Occupational Epidemiology, Proceedings of the seventh 

international symposium on epidemiology in occupational health. Tokyo, Japan. 1989. Pp. 173-

177. 

Weast RC. 1989. Handbook of Chemistry and Physics, 69th edition, Boca Raton, FL: CRC Press 

Inc. 

Wolf DC, Morgan KT, Gross EA, Barrow C, Moss OR, James RA, and Popp JA. 1995. Twoyear inhalation exposure of female and male B6C3F1 mice and F344 rats to chlorine gas induces 

lesions confined to the nose. Fundam. Appl. Toxicol. 24:111-131. 

Appendix D3 113 Chlorine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CHLORINE DIOXIDE 

(anthium dioxide; alcide; chlorine oxide; chlorine peroxide; 

chloryl radical; doxcide 50) 

CAS Registry Number: 10049-04-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.6 μg/m3 (0.2 ppb) 

Critical effect(s) Vascular congestion and peribronchiolar edema; 

hemorrhagic alveoli and congested capillaries 

in the lung in rats 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1994; CRC, 1994) 

Description Yellow to red liquid or gas 

Molecular formula ClO2

Molecular weight 67.45 g/mol 

Density 1.642 g/cm3

 @ 0°C (liquid) 

Boiling point 9.9-11°C 

Melting point −59.5°C 

Solubility Soluble in water, alkaline and sulfuric acid 

solutions 

Conversion factor 1 ppm = 2.76 mg/m3

III. Major Uses or Sources 

Chlorine dioxide is used directly as a bleaching agent for cellulose, textiles, flour, leather, oils, 

and beeswax. It is also used in the purification of water and as a bactericide and antiseptic 

(HSDB, 1994). The annual statewide industrial emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 1136 

pounds of chlorine dioxide (CARB, 2000). 

IV. Effects of Human Exposures 

Case reports of human occupational exposure to chlorine dioxide have shown that 19 ppm was 

fatal to one worker and 5 ppm was definitely irritating (Elkins, 1959). Seven out of 12 workers 

exposed regularly to chlorine dioxide at levels generally below 0.1 ppm (0.28 mg/m3

) reported 

symptoms of ocular and respiratory irritation leading to slight bronchitis (Gloemme and 

A - 114 

Chlorine Dioxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Lundgren, 1957). However, the authors ascribed the bronchitis to occasional acute excursions of 

chlorine dioxide levels above 0.1 ppm due to technical problems such as equipment leakage. 

Concurrent exposure to chlorine and chlorine dioxide in pulp mill workers resulted in an increase 

in the reporting of subjective symptoms of irritation (Ferris et al., 1967). In this study, the 

chlorine dioxide concentrations ranged from trace levels to 0.25 ppm (0.69 mg/m3). No 

differences were found between these workers and controls by pulmonary function tests. 

V. Effects of Animal Exposures 

Eight rats (sex unspecified) were exposed for 5 hours/day, 5 days/week, for 2 months to 0 or 

1 ppm (2.8 mg/m3

) chlorine dioxide (Paulet and Debrousses, 1972). The number of control 

animals was not specified. Microscopic evaluation of the lungs revealed vascular congestion and 

peribronchiolar edema in all animals exposed to chlorine dioxide. The subchronic LOAEL for 

respiratory effects was therefore 1 ppm (2.8 mg/m3

). 

An earlier study by these researchers (Paulet and Debrousses, 1970) examined the effects of 

exposure to 2.5, 5, or 10 ppm chlorine dioxide for several hours/day for 30 days in rats and 

rabbits (n = 4-10 animals per group). Body weights, blood cell counts, and histopathological 

examination of the liver, lungs, and other tissues were measured in each group. At 10 ppm, nasal 

discharge, localized bronchopneumonia, and desquamated alveolar epithelium were observed. 

White and red blood cell counts were also increased with this exposure. Rats and rabbits 

exposed to 2.5 ppm for 7 hours/day for 30 days or for 4 hours/day for 45 days, respectively, 

showed significant respiratory effects, including hemorrhagic alveoli and inflammatory 

infiltration of the alveolar spaces. 

Rats exposed to 5, 10, or 15 ppm (13.8, 27.6, or 41.4 mg/m3) chlorine dioxide for 15 minutes, 2 

or 4 times/day, for 1 month showed an increase in congested lungs, nasal discharge, and 

catarrhous lesions of the alveoli beginning at 10 ppm (Paulet and Debrousses, 1974). No 

significant changes in these parameters were seen at 5 ppm. 

Dalhamn (1957) found that acute exposure to 260 ppm chlorine dioxide for 2 hours resulted in 

the death of 1 out of 4 rats. Five out of 5 rats died during exposures of 4 hours/day for 14 days. 

All exposed animals exhibited signs of respiratory distress and ocular discharge. No effects were 

seen in 5 rats exposed to 0.1 ppm for 5 hours/day, 7 days/week, for 10 weeks. Thus 0.1 ppm was 

a subchronic NOAEL. 

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Determination of Noncancer Chronic Reference Exposure Levels December 2000

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Paulet and Debrousses (1970, 1972) 

Study population Wistar rats (8 per exposure concentration) 

Exposure method Discontinuous whole-body inhalation 

 (0 or 1 ppm) 

Critical effects Vascular congestion; peribronchial edema in 

all animals; lung alveolar damage 

LOAEL 1 ppm (2.8 mg/m3

) 

NOAEL Not observed 

Exposure continuity 5 hours/day, 5 days/week 

Exposure duration 2 months (2/24 = 8.3% of lifetime) 

Average experimental exposure 0.15 ppm for LOAEL group (1 x 5/24 x 5/7) 

Human equivalent concentration 0.23 ppm for LOAEL group (gas with thoracic 

 respiratory effects, RGDR = 1.57 based on 

 MV = 0.17 m3

, SA(Th) = 3,460 cm2) 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 1,000 

Inhalation reference exposure level 0.0002 ppm (0.2 ppb, 0.0006 mg/m3

, 

0.6 μg/m3) 

The U.S. EPA (1995) based its RfC of 0.2 μg/m3 on the same study but included a Modifying 

Factor (MF) of 3 for database deficiencies. The criteria for use of modifying factors are not well 

specified by U.S. EPA. Such modifying factors were not used by OEHHA. In addition OEHHA 

assigned uncertainty factors according to its peer-reviewed, approved methodology (OEHHA, 

2000). 

OEHHA earlier developed a chronic REL for chlorine of 0.2 μg/m3

 (0.08 ppb) based on 

hyperplasia in respiratory epithelium in female rats. Based on chemical reactivity, the REL for 

chlorine dioxide might be expected to be lower than that for chlorine. However, there are much 

less toxicologic data available for chlorine dioxide than for chlorine. 

VII. Data Strengths and Limitations for Development of the REL

The REL for chlorine dioxide had uncertainties in all areas of concern. Thus the best available 

study was still limited by lack of multiple exposure concentrations, by the relatively short 

duration of exposures, and by the small number of animals examined. Adequate human health 

effects information is lacking, although it appears likely that the proposed REL would be 

protective of the effects reported in the single limited human study available. Other limitations 

were the lack of dose-response information and the lack of comprehensive data on multi-organ 

effects. 

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Determination of Noncancer Chronic Reference Exposure Levels December 2000

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dalhamn T. 1957. Chlorine dioxide: Toxicity in experimental animals and industrial risks. AMA 

Arch. Ind. Hyg. 15:101-107. 

Elkins HB. 1959. The Chemistry of Industrial Toxicology, 2nd ed. New York: Wiley and Sons, 

Inc. pp.89-90. 

Ferris BG Jr, Burgess WA, and Worcester J. 1967. Prevalence of chronic respiratory disease in a 

pulp mill and a paper mill in the United States. Br. J. Ind. Med. 24(1):26-37. 

Gloemme J, and Lundgren KD. 1957. Health hazards from chlorine dioxide. Arch. Ind. Health. 

16: 169-176. 

HSDB. 1994. Hazardous Substances Data Bank. TOMES®. Vol. 20. Denver, CO: Micromedex, 

Inc. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov

Paulet G, and Debrousses S. 1970. [On the action of ClO2 at low concentrations on laboratory 

animals.] Arch. Mal. Prof. Med. Trav. Secur. Soc. 31(3):97-106. (in French) 

Paulet G, and Debrousses S. 1972. [On the toxicology of chlorine dioxide.] Arch. Mal. Prof. 

Med. Trav. Secur. Soc. 33(1-2):59-61. (in French) 

Paulet G, and Debrousses S. 1974. [Action of a discontinuous exposure to chlorine dioxide on 

the rat.] Arch. Mal. Prof. Med. Trav. Secur. Soc. 35(9):797-803. (in French) 

U.S. EPA. 1994. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS), CD-ROM version. Denver, CO: Micromedex, Inc. (Edition expires 10/31/94). 

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Chlorine Dioxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY

CHLOROBENZENE 

(monochlorobenzene; benzene chloride; benzene monochloride; chlorbenzene; chlorbenzol; 

phenyl chloride) 

CAS Registry Number: 108-90-7 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 1000 μg/m3 (300 ppb)

Critical effect(s) Increased liver weights, hepatocellular 

hypertrophy, renal degeneration and 

inflammation, and testicular degeneration in 

rats 

Hazard index target(s) Alimentary system; kidney; reproductive system 

II. Physical and Chemical Properties Summary (HSDB, 1995; CRC, 1994)

Description Colorless, neutral liquid 

Molecular formula C6H5Cl 

Molecular weight 112.56 g/mol 

Boiling point 132o

C 

Melting point −45.2o

C 

Vapor pressure 11.8 torr at 25o

C 

Solubility Practically insoluble in water (0.049 g/100 ml); 

soluble in alcohol, benzene, chloroform, 

diethyl ether 

Conversion factor 1 ppm = 4.60 mg/m3

 at 25 oC 

III. Major Uses and Sources 

As one of the most widely used chlorinated benzenes, mono-chlorobenzene has been a major 

chemical for at least 50 years. It was historically important in the manufacture of chlorinated 

pesticides, especially DDT, and in the production of phenol and aniline. Monochlorobenzene’s 

principal current use is as a chemical intermediate in the production of chemicals such as 

nitrochlorobenzenes and diphenyl oxide. These chemicals are subsequently used in the 

production of herbicides, dyestuffs, and rubber chemicals. Additionally, monochlorobenzene is 

used as a solvent in degreasing processes (e.g., in metal cleaning operations), paints, adhesives, 

waxes and polishes (HSDB, 1995; NIOSH, 1993). The annual statewide industrial emissions 

from facilities reporting under the Air Toxics Hot Spots Act in California based on the most 

recent inventory were estimated to be 29,451 pounds of chlorobenzene (CARB, 2000). 

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Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposure 

Even though monochlorobenzene has been used industrially for many years, few epidemiologic 

and/or occupational studies have addressed the potential health status of workers chronically 

exposed to monochlorobenzene (NIOSH, 1993). A Russian occupational study (Rozenbaum et 

al., 1947, as reported by the U.S. EPA, 1988) describes multiple central nervous system effects, 

including headache, numbness, dizziness, cyanosis, hyperesthesia, and muscle spasms, after 

intermittent exposure over 2 years to monochlorobenzene in a mixed chemical environment. No 

specific exposure levels or histopathologic data were reported. 

Two small studies utilizing volunteers exposed to single doses of monochlorobenzene have 

reported central nervous system effects (Ogata et al., 1991; Tarkhova, 1965). An exposure 

chamber study of five volunteers exposed up to 60 ppm monochlorobenzene (276 mg/m3

) for a 

single 7 hour exposure described acute subjective symptoms such as drowsiness, headache, eye 

irritation, and sore throat (Ogata et al., 1991). One other human volunteer study described 

altered electrical activity of the cerebral cortex in four individuals exposed to 43.4 ppm 

monochlorobenzene vapors for 2.5 minutes (Tarkhova, 1965). 

V. Effects of Animal Exposure

No chronic inhalation studies have evaluated the toxicity of monochlorobenzene. Only a single, 

oral chronic carcinogenicity study (NTP, 1985) has evaluated the long-term adverse affects of 

monochlorobenzene administration. However, a few subchronic inhalation studies have 

demonstrated adverse effects on the liver, the kidney, and, to a lesser extent, blood parameters 

following monochlorobenzene exposure over a period of weeks or months (Dilley, 1977; John et 

al., 1984; Nair et al., 1987). 

One subchronic study evaluated Sprague-Dawley male rats and rabbits exposed to 0, 75, or 

200 ppm of monochlorobenzene for 7 hr/day, 5 days/week, for up to 24 weeks (Dilley, 1977). In 

rats, monochlorobenzene-related toxicity included increased absolute and relative (to brain- or 

body-weight) organ weights (especially the liver) after 11 and 24 weeks of exposure (LOAEL 75 

ppm). Male rabbits also demonstrated increases in liver weight after 24 weeks of exposure 

(LOAEL = 75 ppm). Some hematological changes were reported in rats including differences in 

platelet and reticulocyte counts between control and exposed animals; however, some changes 

observed at 11 weeks were variable and comparable to controls at 24 weeks (red blood cell 

count, hemoglobin, hematocrit, and white blood cell count). Pathological changes were observed 

in rats, with occasional focal lesions in the adrenal cortex, tubular lesions in the kidneys, and 

congestion in the liver and kidneys. 

Two other subchronic inhalation studies reported adverse organ effects following 

monochlorobenzene exposure in rats and rabbits (John et al., 1984; Nair et al., 1987). In the first 

study, John et al. (1984) reported increased liver weights in rats and rabbits following short-term 

(10 or 13 day, 6 hours/day) monochlorobenzene exposure (LOAEL = 590 ppm in rats and 210 

ppm in rabbits). Nair et al. (1987) exposed male and female Sprague-Dawley rats to 0, 50, 150, 

or 450 ppm monochlorobenzene vapors daily for 6 hours over 10-11 weeks prior to mating, and 

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Chlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

up to day 20 of gestation for 2 generations. Nair et al. found dose-related changes in the livers, 

kidneys, and testes in both generations of males (F0 and F1). Hepatotoxicity occurred as 

hepatocellular hypertrophy and increased liver weights (mean and absolute) at concentrations 

greater than 50 ppm (LOAEL = 150 ppm). At this concentration (150 ppm), renal changes 

included tubular dilation, interstitial nephritis, and foci of regenerative epithelium. Testicular 

degeneration of the germinal epithelium occurred in both generations of exposed males, but no 

chlorobenzene-induced adverse effects on reproductive performance or fertility were seen. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Nair et al. (1987) 

Study population Sprague-Dawley rats (30/sex/group) 

Exposure method Discontinuous inhalation exposures 

(0, 50, 150, and 450 ppm) 

Critical Effects Increases in absolute and relative liver weights 

(F0 and F1 both sexes), hepatocellular 

hypertrophy (F0 and F1 males), renal 

degeneration and inflammation (F0 and F1 both 

sexes), testicular degeneration (F0 and F1

males). 

LOAEL 150 ppm 

NOAEL 50 ppm 

Exposure continuity 6 hours/day, 7 days/week 

Exposure duration 11 weeks 

Average experimental exposure 13 ppm for NOAEL group (50 x 6/24) 

Human equivalent concentration 26 ppm (gas with systemic effects, based on 

RGDR = 2.0 for lambda (a) : lambda (h)) 

(Gargas et al., 1989) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.3 ppm (300 ppb; 1.0 mg/m3

, 1000 μg/m3) 

Of the three inhalation studies available (Dilley, 1977; John et al., 1984; Nair et al., 1987), the 

Nair et al. (1987) two generational developmental study was selected for identifying a NOAEL 

and LOAEL. It best presented the histopathology of the adverse effects, and demonstrated a 

dose response relationship for these effects (statistically significant increases in mean liver 

weights, incidence of renal changes, and testicular degeneration). 

Another subchronic inhalation study (Dilley, 1977) also observed increases in organ weights, 

including the liver, in rats after 11 and 24 weeks exposure to 75 and 250 ppm 

monochlorobenzene (LOAEL = 75 ppm), and in rabbits at 24 weeks. Similar adverse liver and 

kidney effects were found in subchronic oral bioassays (Kluwe et al., 1985; NTP, 1985). These 

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Chlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

include increases in liver weight and hepatocellular degeneration in rats (LOAEL = 125 

mg/kg/day) and mice (LOAEL = 250 mg/kg/day), and renal necrosis and degeneration in rats 

(LOAEL = 500 mg/kg/day) and mice (LOAEL = 250 mg/kg/day) after 13 weeks oral exposure to 

chlorobenzene. 

Uncertainty factors are appropriate due to the lack of chronic studies, both animal bioassay and 

human, and the limited number of subchronic inhalation studies, thereby requiring estimation of 

the chronic REL from this shorter term, single species study. The magnitude of interspecies 

variation remains unknown, as few species have been tested and human data for comparison are 

lacking. However, metabolic studies have demonstrated species variation in the urinary 

elimination of chlorobenzene metabolites (Ogata and Shimada 1983; Ogata et al., 1991; Yoshida 

et al., 1986). Humans metabolize and excrete chlorobenzene predominately as free and 

conjugated forms of 4-chlorocatechol and chlorophenols, while the main rodent urinary 

metabolite, p-chlorophenylmercapturic acid, is found in minor amounts (<0.5%). No 

information exists which identifies human subpopulations possibly susceptible to 

monochlorobenzene exposure. 

For comparison with the proposed REL, a REL can be derived from the 24 week LOAEL of 75 

ppm for liver effects (Dilley, 1977). The LOAEL is equivalent to a continuous exposure 

LOAEL of 15.6 ppm. Multiplying by the RGDR of 2 and dividing by a cumulative UF of 100 (3 

for LOAEL, 3 for interspecies and 10 for intraspecies) also yields an estimate of 300 ppb. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for chlorobenzene include the observation of a NOAEL, the 

availability of subchronic inhalation exposure data from a well-conducted study with 

histopathological analysis, and the demonstration of a dose-response relationship. Major areas of 

uncertainty are the lack of adequate human exposure data and limited reproductive toxicity data. 

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dilley JV. 1977. Final Report: Toxic evaluation of inhaled chlorobenzene (monochlorobenzene). 

Stanford Research Institute. Contract No. 210-76-0126. Prepared for: National Institute of 

Occupational Safety and Health, Division of Biomedical and Behavioral Sciences. Cincinnati, 

OH: NIOSH. NTIS No. PB-276623. 

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Chlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Gargas ML, Burgess RJ, Voisard DE, Cason GH, and Andersen ME. 1989. Partition coefficients 

of low-molecular-weight volatile chemicals in various liquids and tissues. Toxicol. Appl. 

Pharmacol. 98(1):87-99. 

HSDB. 1995. Hazardous Substances Data Bank. TOMES® Denver, CO: Micromedex, Inc. 

John JA, Hayes WC, Hanley TR, Johnson KA, Gushow TS, and Rao KS. 1984. Inhalation 

teratology study on monochlorobenzene in rats and rabbits. Toxicol. Appl. Pharmacol. 76:365-

373. 

Kluwe WM, Dill G, Persing R, and Peters A. 1985. Toxic responses to acute, subchronic, and 

chronic oral administrations of monochlorobenzene to rodents. J. Toxicol. Environ. Health 

15(6):745-767. 

Nair RS, Barter JA, Schroeder RE, Knezevich A, and Stack CR. 1987. A two-generation 

reproduction study with monochlorobenzene vapor in rats. Fundam. Appl. Toxicol. 9:678-686. 

NTP. 1985. National Toxicology Program. Toxicology and carcinogenesis studies of 

chlorobenzenes (CAS No. 108-90-7) in F344/N rats and B6C3F1 mice (gavage studies). U.S. 

Dept. of Health and Human Services, Public Health Service, National Institutes of Health, NTP 

TR 261. NIH Pub. No. 86-2517, NTIS PB86-144714. 

NIOSH. 1993. National Institute for Occupational Safety and Health. NIOH and NIOSH basis 

for an occupational health standard: Chlorobenzene. U.S. Department of Health and Human 

Services, Public Health Service, DHHS(NIOSH) Publication No. 93-102. 

Ogata M, and Shimada Y. 1983. Differences in urinary monochlorobenzene metabolites between 

rats and humans. Int. Arch Occup. Environ. Health 53:51-57. 

Ogata M, Taguchi T, Hirota N, Shimada Y, and Nakae S. 1991. Quantitation of urinary 

chlorobenzene metabolites by HPLC: Concentrations of 4-chlorocatechol and chlorophenols in 

urine and of chlorobenzene in biological specimens of subjects exposed to chlorobenzene. Int. 

Arch. Occup. Environ. Health 63:121-128. 

Rozenbaum ND, Blekh RS, and Kremneva SN. 1947. [Use of chlorobenzene as a solvent from 

the standpoint of industrial hygiene.] Gig. I Sanit. 12:21-24. (Russian) [as cited in U.S. EPA, 

1988.] 

Tarkhova LP. 1965. Materials for determining the maximum permissible concentration of 

chlorobenzol in atmospheric air. Hygiene and Sanitation 30:327-333. (Jerusalem, Israel Program 

for Scientific Translation available for NTIS.) 

U.S. EPA. 1988. United States Environmental Protection Agency. Health effects criteria 

document for chlorobenzene. Final draft. EPA 600/88-90/99. Washington DC: U.S. EPA. 

National Technical Information Service, PB89-192116. 

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Chlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Yoshida M, Sunaga M, and Hara I. 1986. Urinary metabolites levels in workers exposed to 

chlorobenzene. Ind. Health 24:255-258. 

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Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

CHLOROFORM 

(trichloromethane; formyl trichloride; methenyl trichloride; methyl trichloride) 

CAS Registry Number: 67-66-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 300 μg/m3

 (50 ppb) 

Critical effect(s) Liver toxicity (degenerative, foamy 

vacuolization, and necrosis) in rats; increased 

liver weights in male rats 

Kidney toxicity (cloudy swelling and nephritis) 

in rats 

Developmental toxicity 

Hazard index target(s) Alimentary system; kidney; teratogenicity 

II. Chemical Property Summary (HSDB, 1995; 1999; CRC, 1994) 

Description Colorless liquid 

Molecular formula CHCl3

Molecular weight 119.49 g/mol 

Boiling point 61.1o

C 

Melting point −63.6 oC 

Vapor pressure 197-200 torr @ 25 o

C 

Solubility Soluble in water (8220 mg/L); miscible in 

carbon tetrachloride, carbon disulfide, 

alcohols, benzene, ethers and oils 

Conversion factor 4.9 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Chloroform (CHCl3) is used in industry and laboratory settings as a solvent for adhesives, 

pesticides, fats, oils and rubbers. It is also used as a chemical intermediate in the synthesis of 

fluorocarbon 22, dyes, pesticides, and tribromomethane. Chloroform is produced as a byproduct 

of water, sewage, and wood pulp chlorination (HSDB, 1995). In 1996, the latest year tabulated, 

the statewide mean outdoor monitored concentration of chloroform was approximately 0.037 ppb 

(CARB 1999a). The annual statewide industrial emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 

79,949 pounds of chloroform (CARB 1999b). 

Appendix D3 124 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure 

Limited information is available regarding possible adverse health effects in humans following 

chronic inhalation of chloroform. However, historical clinical reports from patients who 

underwent chloroform anesthesia indicate that acute inhalation exposure affects the central 

nervous system, cardiovascular system, stomach, liver, and kidneys (Schroeder, 1965; Smith et 

al., 1973; Whitaker and Jones, 1965). Acute chloroform toxicity included impaired liver 

function (Smith et al., 1973), toxic hepatitis (Lunt, 1953; Schroeder, 1965), cardiac arrhythmia 

(Payne, 1981; Schroeder, 1965; Whitaker and Jones, 1965), and nausea (Schroeder, 1965; Smith 

et al., 1973; Whitaker and Jones, 1965), and caused central nervous system symptoms 

(Schroeder, 1965; Whitaker and Jones, 1965). Chronic inhalation studies are limited to a few 

occupational studies identifying the liver and the central nervous system as target organs 

(Challen et al., 1958; Li et al., 1993; Phoon et al., 1983; Bomski et al., 1967). 

Challen et al. (1958) investigated workers manufacturing throat lozenges with exposure to 

chloroform vapors estimated in the range 77 to 237 ppm with episodes of >1100 ppm. Workers 

reported symptoms of fatigue, dull-wittedness, depression, gastrointestinal distress, and frequent 

and burning micturition. No evidence of liver dysfunction was found based on thymol turbidity, 

serum bilirubin, and urine urobilinogen levels. 

Bomski et al. (1967) reported 17 cases of hepatomegaly in a group of 68 chloroform-exposed 

workers. Chloroform concentrations ranged from 2 to 205 ppm (duration 1 to 4 years). Three of 

the 17 workers with hepatomegaly had toxic hepatitis based on elevated serum enzymes. 

Additionally, 10 workers had splenomegaly. Workers exposed to chloroform had a 10-fold 

increased risk of contracting viral hepatitis compared to the general population. The study 

authors considered the chloroform induced liver toxicity as a predisposing factor for viral 

hepatitis, but the incidence of viral hepatitis in the workers is in itself a confounding factor. 

Phoon et al. (1983) described two outbreaks of toxic jaundice in workers manufacturing 

electronics equipment in Singapore. One plant had 13 cases of jaundice, initially diagnosed as 

viral hepatitis, in a work area with >400 ppm chloroform. Blood samples from workers (five 

with jaundice, four without symptoms) contained between 0.10 and 0.29 mg chloroform/100 mL. 

A second factory reported 18 cases of hepatitis, all from a work area utilizing chloroform as an 

adhesive. Two samplings indicated air levels of 14.4 to 50.4 ppm chloroform. Due to a lack of 

fever and hepatitis B surface antigen in the patients, the authors attributed the jaundice to 

chloroform exposure rather than viral hepatitis. 

More recently, Li et al.(1993) reported on 61 chloroform exposed workers from a variety of 

production factories. Exposure levels at 3 representative worksites varied widely, from 4.27 to 

147.91 mg/m3 (0.9 to 30 ppm) (119 samples), with 45% of the samples below 20 mg/m3. The 

exposed workers were subclassified for some studies according to exposure levels into group 1 

(mean level = 13.49 mg/m3 or 2.8 ppm) and group 2 (mean level = 29.51 mg/m3 or 6 ppm). 

Workers exposed to chloroform had slight liver damage indicated by higher (abnormal) levels of 

serum prealbumin (in group 2) and transferrin (in both groups) than those of control workers. 

Neurobehavioral functions were also affected, manifested as increases in scores of passive mood 

states and dose-related, negative changes in neurobehavioral testing. 

Appendix D3 125 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

These cross sectional studies are limited in their ability to establish chronic NOAEL/LOAEL 

values due to limited exposures, concurrent exposure to other chemicals, inadequate control 

groups and potential confounders. However, these studies indicate the potential for liver and 

central nervous system toxicity in humans exposed to chloroform via inhalation. 

V. Effects of Animal Exposure 

Exposure of experimental animals to chloroform for acute, subchronic or chronic durations 

results in toxicity to the liver and kidney, as well as to the respiratory and central nervous 

systems (USDHHS, 1993). The majority of chronic animal studies have used oral routes of 

chloroform administration (USDHHS, 1993), while only limited data are available on inhalation 

specific exposures. Both routes of exposure, however, appear to primarily affect the liver and 

kidney (Chu et al., 1982; Heywood et al., 1979; Jorgenson et al., 1985; Miklashevshii et al., 

1966; Munson et al., 1982; Roe et al., 1979; Larson et al., 1996; Templin et al., 1996; Torkelson 

et al., 1976). 

Larson et al. (1996) exposed female and male B6C3F1 mice to atmospheric concentrations of 0, 

0.3, 2, 10, 30, and 90 ppm chloroform 6 hr/day, 7 days/week for exposure periods of 4 days or of 

3, 6, or 13 consecutive weeks. Additional exposure groups were exposed for 5 days/week for 13 

weeks or for 5 days/week for 6 weeks and then examined at 13 weeks. Complete necropsy and 

microscopic evaluation revealed that chloroform treatment induced dose- and time-dependent 

lesions only in the livers and nasal passage of the female and male mice and in the kidneys of the 

male mice. Large increases in the liver cell labeling index were seen in the 90-ppm groups at all 

time points. The female mice were most sensitive. The no-observed-adverse-effect level 

(NOAEL) for induced hepatic cell proliferation was 10 ppm. The hepatic labeling indices in the 

5 days/week groups were about half of those seen in the 7 days/week groups and returned to the 

normal baseline in the 6-week recovery groups. The NOAEL for increased liver weight 

(normalized to body weight) was 10 ppm in male mice. Histologic changes and regenerative cell 

proliferation were induced in the kidneys of male mice at 30 and 90 ppm with 7 days/week 

exposures and also at 10 ppm with the 5 days/week regimen. Nasal lesions were transient and 

occurred only in mice exposed to 10, 30, or 90 ppm for 4 days. 

Templin et al. (1996) exposed male and female F-344 rats to airborne concentrations of 0, 2, 10, 

30, 90, or 300 ppm chloroform 6 hr/day, 7 days/week for 4 days or 3, 6, or 13 weeks. Additional 

groups were exposed 5 days/week for 13 weeks, or 5 days/week for 6 weeks and held until Week 

13. A “full-screen” necropsy identified the kidney, liver, and nasal passages as the only target 

organs. The primary target in the kidney was the epithelial cells of the proximal tubules of the 

cortex; significantly elevated increases in the cell labeling index were observed at concentrations 

of 30 ppm chloroform and above. However, only a marginal increase in the renal cell labeling 

index in the males was seen after exposures of 90 ppm, 5 days/week. Chloroform induced 

hepatic lesions in the midzonal and centrilobular regions with increases in the labeling index 

throughout the liver, but only at 300 ppm, an extremely toxic level. An additional liver lesion 

seen only at 300 ppm was numerous intestinal crypt-like ducts surrounded by dense connective 

tissue. Enhanced bone growth and hypercellularity in the lamina propria of the ethmoid 

turbinates of the nose occurred at the early time points at concentrations of 10 ppm and above. 

Appendix D3 126 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

At 90 days there was a generalized atrophy of the ethmoid turbinates at concentrations of 2 ppm 

(the lowest concentration tested) and above. 

Torkelson and associates (1976) exposed rats (12/sex/group), rabbits (2-3/sex/group), and guinea 

pigs (8-12/sex/group) for 7 hours/day, 5 days/week over 6 months to 0, 25, 50 or 85 ppm 

chloroform vapor. Dogs were exposed to 25 ppm chloroform, for 7 hours/day, 5 days/week for 6 

months. Dose and species-dependent pathological changes in the liver included mild to severe 

centrilobular granular degeneration, foamy vacuolization, focal necrosis, and fibrosis in both 

sexes of all species tested. Guinea pigs were the least sensitive and male rats the most sensitive 

to chloroform induced hepatotoxicity; the above adverse effects occurred at 25 ppm. Adverse 

kidney effects observed in all species included cloudy swelling of the renal tubular epithelium 

and interstitial and tubular nephritis. Pneumonitis was observed in the high (85 ppm) exposure 

groups of male rats, female guinea pigs, and male rabbits, and in the lower dose group of female 

rabbits (25 ppm). Clinical and blood parameters were also examined in rats and rabbits, but no 

alterations were attributable to chloroform exposure. 

Effects on average body weight, and relative liver and kidney weights of rats due to 

chloroform exposure 7 hours/day for 6 months (Torkelson et al., 1976) 

Sex Parameter 

Unexposed 

control Air control 25 ppm 50 ppm 85 ppm 

male survival 11/12 10/12 9/10 6/10 

avg. bw 343 356 305* 316 

liver 2.45 2.52 2.48 2.76* 

kidney 0.69 0.70 0.81* 0.84* 

male survival 8/12 12/12 9/12 

avg. bw 319 347 335 

liver 2.67 2.41 2.65 

kidney 0.75 0.70 0.83* 

female survival 10/12 9/12 10/10 10/10 

avg. bw 202 223 203 206 

liver 2.92 2.99 3.00 3.12 

kidney 0.82 0.81 0.95 1.06 

Female survival 10/12 12/12 12/12 

avg. bw 211 202 194 

liver 3.02 2.93 3.08 

kidney 0.83 0.84 0.94* 

* p< 0.05 

Appendix D3 127 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Torkelson et al.(1976) 

Study population Rats, unspecified strain (12/sex/group) 

Exposure method Discontinuous whole-body inhalation 

exposures (0, 25, 50, 85 ppm) 

Critical effects Pathological changes in liver (degenerative), 

and kidneys (cloudy swelling) 

LOAEL 25 ppm 

NOAEL Not observed 

Exposure continuity 7 hr/day for 5 days/week for 6 months 

Average experimental exposure 5.3 ppm for LOAEL group (25 x 7/24 x 5/7) 

Human equivalent concentration 15.9 ppm for LOAEL group (gas with 

systemic effects, based on RGDR = 3.0 

for lambda (a) : lambda (h) (Gargas et al.,

1989)) 

Exposure duration 6 months 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.05 ppm (50 ppb; 0.30 mg/m3

; 300 μg/m3) 

In the study of Torkelson and associates (1976) rats were the most sensitive species and guinea 

pigs the least sensitive to chloroform vapors. Though of subchronic duration, this inhalation 

study still exposed rats discontinuously for 25% of a lifetime (25.8 weeks/104 weeks/lifetime). 

Pathological changes were observed in both sexes of rat at 50 and 85 ppm (244 or 415 mg/m3

) 

and in male rats at 25 ppm (122 mg/m3) chloroform. These hepatic changes included mild to 

severe centrilobular granular degeneration, foamy vacuolization, focal necrosis, and fibrosis. 

Adverse effects in the kidney including cloudy swelling and nephritis were seen in all species 

tested at 25 ppm (122 mg/m3

) chloroform. 

An unexpected finding in animals was the generalized atrophy of the ethmoid turbinates of F344 

rats after a 90 day exposure at concentrations of 2 ppm chloroform and above (Templin et al., 

1996). Nasal lesions have also been reported in F344 rats given chloroform by gavage (Larson 

et al., 1995). This severe and extensive chloroform-induced olfactory mucosal degeneration in 

rats is not associated with detectable olfactory deficit (Dorman et al., 1997). As the basis of the 

REL we have used the more usual chloroform organ targets of liver and kidney. However, 

confirmation of nasal effects in other rat strains and other species may require reassessing the 

basis of the REL for chloroform. 

The human occupational studies have reported jaundice with or without alterations in liver 

enzymes at similar ambient concentrations: 2 to 204 ppm chloroform (10 to 995 mg/m3

) after at 

least 1 year (Bomski et al., 1967) and 14 to 400 ppm chloroform (68 to 1952 mg/m3

) after 6 

months or less (Phoon et al., 1983). The presence of jaundice and hepatitis in these 2 reports 

Appendix D3 128 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

made them questionable for use in developing a REL. In the Li et al. (1993) study the workers 

were exposed for an average of 7.8 years (range = 1-15 years) and the air concentrations ranged 

from 4.27 to 141.25 mg/m3

 with a geometric average of 20.46 mg/m3. The exposed workers 

were subdivided into higher (n=46) and lower (n=14) exposures, but the separation was not 

indicated for all results. If the lower exposure level of 2.8 ppm (13.49 mg/m3) is classified as a 

mild LOAEL based on a significant difference from controls in one type of neurobehavioral test, 

the exposure level can be time adjusted to an equivalent continuous exposure of 1 ppm, then 

divided by a LOAEL UF of 3 and an intraspecies UF of 10 to yield a REL of 30 ppb, in good 

agreement with the proposed REL of 50 ppb (300 μg/m3

) based on animals (rats). 

Chloroform is metabolized by the cytochrome P-450 dependent mixed function oxidase system, 

primarily in the liver, the respiratory epithelium, and the kidney. In the rat liver and kidneys, 

chloroform is metabolized to phosgene (Pohl et al., 1984). The hepatotoxicity and 

nephrotoxicity of chloroform is thought to be due largely to phosgene (Bailie et al., 1984). 

Individuals with concurrent exposure to certain chemical inducers of liver cytochrome P450 

activity, including barbiturates, may be at potentially greater risk of chloroform toxicity (Cornish 

et al., 1973). Others with possible higher sensitivity to chloroform include persons with 

underlying liver, kidney or neurological conditions. 

VII. Data Strengths and Limitations for Development of the REL

Strengths of the chronic REL for chloroform derive from the critical effect being found in the 

liver, a well-established site of chloroform toxicity. Limitations in the data include the lack of a 

NOAEL in the key study, the less than lifetime duration of the key study, and the limited number 

of chronic inhalation studies available. 

VIII. References 

Bailie MB, Smith JH, Newton JF, and Hook JB. 1984. Mechanism of chloroform nephrotoxicity. 

IV. Phenobarbital potentiation of in vitro chloroform metabolism and toxicity in rabbit kidneys. 

Toxicol. Appl. Pharmacol. 74:285-292. 

Bomski H, Sobolewska A, and Strakowski A. 1967. [Toxic damage to the livers of chemical 

plant workers by chloroform.] Int. Arch. Gewerbepathol. Gewerbehyg. 24:127-134. 

CARB. 1999a. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Chloroform. Available online at http:/www.arb.ca.gov/aqd/toxics.htm 

CARB. 1999b. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Challen P, Hickish D, and Bedford J. 1958. Chronic chloroform intoxication. Br. J. Ind. Med. 

15:243-249. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Appendix D3 129 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Chu I, Villeneuve DC, Secours VE, Becking GC, and Valli VE. 1982. Toxicity of 

trihalomethanes: I. The acute and subacute toxicity of chloroform, bromodichloromethane, 

chlorodibromomethane and bromoform in rats. J. Environ. Sci. Health. B17:205-224. 

Cornish HH, Ling B, and Barth M. 1973. Phenobarbital and organic solvent toxicity. Am. Ind. 

Hyg. Assoc. J. 34:487-492. 

Dorman DC, Miller KL, D'Antonio A, James RA, Morgan KT. 1997. Chloroform-induced 

olfactory mucosal degeneration and osseous ethmoid hyperplasia are not associated with 

olfactory deficits in Fischer 344 rats.Toxicology 122(1-2):39-50. 

Gargas ML, Burgess RJ, Voisard DE, Cason GH, and Andersen ME. 1989. Partition coefficients 

of low-molecular-weight volatile chemicals in various liquids and tissues. Toxicol. Appl. 

Pharmacol. 98(1):87-99 

Heywood R, Sortwell RJ, Noel PRB, Street AE, Prentice DE, Roe FJC, Wadsworth PF, and 

Worden AN. 1979. Safety evaluation of toothpaste containing chloroform. III. Long-term study 

in beagle dogs. J. Environ. Pathol. Toxicol. 2:835-851. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex Inc. 

Jorgenson TA, Meierhenry EF, Rushbrook CJ, Bull RJ, and Robinson M. 1985. Carcinogenicity 

of chloroform in drinking water to male Osborne-Mendel rats and female B6C3F1 mice. 

Fundam. Appl. Toxicol. 5:760-769. 

Larson JL, Templin MV, Wolf DC, Jamison KC, Leininger JR, Mery S, Morgan KT, Wong BA, 

Conolly RB, Butterworth BE. 1996. A 90-day chloroform inhalation study in female and male 

B6C3F1 mice: implications for cancer risk assessment. Fundam. Appl. Toxicol. 30(1):118-137. 

Larson JL, Wolf DC, Mery S, Morgan KT, Butterworth BE. 1995. Toxicity and cell proliferation 

in the liver, kidneys and nasal passages of female F-344 rats, induced by chloroform 

administered by gavage. Food Chem. Toxicol. 33(6):443-456. 

Li LH, Jiang XZ, Laing YX, Chen ZQ, Zhou YF, and Wang YL. 1993. Studies on the toxicity 

and maximum allowable concentration of chloroform. Biomed. Environ. Sci. 6(2):179-186. 

Lunt RL. 1953. Delayed chloroform poisoning in obstetric practice. Br. Med. J. 1:489-490. 

Miklashevshii VE, Tugarinova VN, Rakhmanina NL, and Yakovleva GP. 1966. Toxicity of 

chloroform administered perorally. Hyg. Sanit. 31:320-323. 

Munson AE, Sain LE, Sanders VM, Kauffmann BM, White KL, Page G, Barnes DW, and 

Borzelleca JF. 1982. Toxicology of organic drinking water contaminants: trichloromethane, 

Appendix D3 130 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

bromodichloromethane, dibromochloro-methane and tribromomethane. Environ. Health 

Perspect. 46:117-126. 

Payne JP. 1981. Chloroform in clinical anaesthesis. Br. J. Anesth. 53:11s-15s. 

Phoon W, Goh K, Lee L, Tan K, and Kwok S. 1983. Toxic jaundice from occupational exposure 

to chloroform. Med. J. Malaysia. 38(1):31-34. 

Pohl L, Gorge J, and Satoh H. 1984. Strain and sex differences in chloroform-induced 

nephrotoxicity. Different rates of metabolism of chloroform to phosgene by the mouse kidney. 

Drug Metab. Disp. 12(3):304-3-8. 

Roe FJC, Palmer AK, Worden AN, and Van Abbe NJ. 1979. Safety evaluation of toothpaste 

containing chloroform. I. Long-term studies in mice. J. Environ. Pathol. Toxicol. 2:799-819. 

Schroeder HG. 1965. Acute and delayed chloroform poisoning. Br. J. Anaesth. 37:972-975. 

Smith AA, Volpitto PP, Gramling ZW, DeVore MB, and Glassman AB. 1973. Chloroform, 

halothane, and regional anesthesia: A comparative study. Anesth. Analg. 52:1-11. 

Templin MV, Larson JL, Butterworth BE, Jamison KC, Leininger JR, Mery S, Morgan KT, 

Wong BA, Wolf DC. 1996. A 90-day chloroform inhalation study in F-344 rats: profile of 

toxicity and relevance to cancer studies. Fundam. Appl. Toxicol. 32(1):109-125. 

Torkelson T, Oyen F, and Rowe V. 1976. The toxicity of chloroform as determined by single and 

repeated exposure of laboratory animals. Am. Ind. Hyg. Assoc. J. 37:697-705. 

USDHHS. 1993. Toxicology Profile for Chloroform. U.S. Department of Health and Human 

Services. Public Health Service. Agency for Toxic Substance and Disease Registry. April 1993. 

Whitaker AM, and Jones CS. 1965. Report of chloroform anesthetics administered with a 

precision vaporizer. Anesth. Analg. 44:60-65.

Appendix D3 131 Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY 

CHLOROPICRIN 

(trichloronitromethane; nitrochloroform; nitrochloromethane) 

CAS Registry Number: 76-06-2 

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.4 μg/m3 (0.05 ppb) 

Critical effect(s) Nasal rhinitis and bronchiectasis in mice 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (from HSDB (1996) except as noted)

 

Description Colorless to faint yellow liquid 

Molecular formula CCl3NO2

Molecular weight 164.4 g/mol 

Boiling point 112°C 

Melting point −64°C (CRC, 1994) 

Vapor pressure 5.7 torr @ 0°C (Fries and West, 1921); 

3.2 kPa (24 torr) @ 25°C (Tomlin, 1994) 

Solubility 1.6 g/L water @ 25°C; 2.272 g/L water @ 0°C 

1.9 g/L water @ 20°C; miscible with benzene, 

ethanol, carbon disulfide, ether, carbon 

tetrachloride, acetone, methanol, acetic acid 

Conversion factor 6.72 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Chloropicrin is used primarily as a preplant soil fumigant against insects and fungi; it also kills 

weed and grass seeds when applied to soil. Chloropicrin is occasionally used as a fumigant in 

grain elevators and storage bins (HSDB, 1996). Chloropicrin is used as an indicator chemical in 

other fumigants such as methyl bromide because of its potent irritant properties. Chloropicrin 

was used in World War I as a chemical warfare agent because of its potent activity as a 

lachrymator. Chloropicrin has a minor use in the chemical synthesis of methyl violet. 

Chloropicrin can also form in drinking water as a result of chlorination processes (Duguet et al., 

1985; Merlet et al., 1985). The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 1507 pounds of chloropicrin (CARB, 2000). This does not include emissions 

from its major use as a preplant soil fumigant, either alone or in combination with other 

Appendix D3 132 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

fumigants, because agricultural field applications are not covered under the Air Toxics Hot Spots 

program. Approximately 3,630,000 lbs. of chloropicrin were used in agriculture in California in 

1999 (DPR, 2000). 

IV. Effects of Human Exposure 

No studies are available which describe toxic effects to humans from chronic exposure to 

chloropicrin. Human exposures to concentrations less than 1 ppm for very short periods of time 

are extremely irritating (ACGIH, 1992; Fries and West, 1921). The threshold of odor detection 

in humans is approximately 1 ppm (ACGIH, 1992). 

V. Effects of Animal Exposure 

Burleigh-Flayer and Benson (1995) conducted a chronic inhalation bioassay with CD rats (50-60 

per sex per dose) exposed discontinuously to 0 (air), 0.1, 0.5, or 1.0 ppm 99.6% pure 

chloropicrin vapor 6 hours/day for 5 consecutive days/week over 107 weeks. Clinical signs 

(such as hypoactivity and decreased startle response) were increased in both sexes, primarily at 

1.0 ppm. Increased mortality was noted in males at 0.5 and 1 ppm and in females at 1 ppm. 

Absolute and relative increased lung and liver weights and increased nasal rhinitis were reported 

in both sexes at the 1 ppm level. However, no effects were seen at 0.1 ppm. Thus this study 

yielded a NOAEL of 0.1 ppm (0.67 mg/m3) for chronic non-cancer effects in rats. 

Results from chronic inhalation of chloropicrin in rats (Burleigh-Flayer and Benson, 1995) 

Chloropicrin Lung wt., m Lung wt., f Rhinitis, m Rhinitis, f 

Mean 

survival, m

0 2.086 g 1.574 g 20/50 18/50 696 d 

0.1 ppm 2.089 g 1.464 g 24/50 17/50 669 d 

0.5 ppm 2.202 g 1.460 g 21/50 26/50 672 d* 

1.0 ppm 2.448 g 1.633 g 35/50** 23/50 647 d** 

*p<0.05; **p<0.01 

A similar study in mice (Burleigh-Flayer et al., 1995) resulted in the same NOAEL. CD-1 mice 

(50/sex/dose) were exposed to chloropicrin (99.6% pure) vapor at 0 (air), 0.1, 0.5, or 1.0 ppm for 

6 hours/day, 5 days/week for at least 78 weeks. Body weights and body weight gains were 

significantly decreased in both sexes at ≥ 0.5 ppm. Food consumption was decreased in males at 

1.0 ppm and in females at ≥ 0.5 ppm. Absolute and relative lung weights were increased in a 

dose-related manner in both sexes at ≥ 0.5 ppm. Changes in pathology observed 

macroscopically in the 1.0 ppm males included increased numbers of lung nodules and increased 

numbers of kidney cysts. In females lung masses and kidney cysts were seen at 0.5 ppm. 

Microscopic pathology changes included increased nasal cavity lesions (including serous 

exudate, hyaline epithelial inclusions, rhinitis, olfactory and epithelial atrophy) and lung lesions 

(including alveolar protein deposits, alveolar histiocytosis, hemorrhage, peribronchiolar 

lymphocytic infiltrate, bronchiectasis, bronchial submucosal fibrosis, peribronchiolar smooth 

muscle hyperplasia), in addition to kidney cysts at ≥ 0.5 ppm (CDPR, 2000). 

Appendix D3 133 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Results from chronic inhalation of chloropicrin in mice (Burleigh-Flayer et al., 1995) 

Chloropicrin Rhinitis, m Rhinitis, f Bronchiectasis, m Bronchiectasis, f 

0 6/50 3/50 0/50 0/50 

0.1 ppm 7/50 6/50 3/50 5/50 

0.5 ppm 17/50** 18/50** 28/50** 28/50** 

1.0 ppm 35/50** 32/50** 41/50** 44/50** 

**p<0.01 

Yoshida et al. (1987) exposed groups of 12 male Fischer 344 rats intermittently to 0, 0.37, 0.67, 

1.58, or 2.93 ppm chloropicrin vapor 6 h/day, 5 days/week for 13 weeks. Mean body weights 

were reduced in the highest 2 exposure groups, and red blood cell count, hematocrit, and 

hemoglobin concentration were significantly increased in the 2.93 ppm group. The treatmentrelated histological lesions reported were degeneration and necrosis of the bronchial and 

bronchiolar epithelia at 2.93 ppm and hypertrophy of these epithelia at 1.58 ppm. Thus the 

primary target organ was the respiratory tract and the subchronic NOAEL was 0.67 ppm (4.5 

mg/m3). (Eyelid closure and decrease in motor activity were seen in all exposure groups only 

during exposure. No morphological changes were seen at 0.67 ppm, so the authors deemed the 

behavior changes minor and not toxicologically important.) 

Male Swiss-Webster mice (group numbers ranging from 16-24) were exposed by inhalation to a 

single level of different sensory irritants including chloropicrin for 6 hours/day for 5 days; 

unexposed control groups had 8-10 mice (Buckley et al., 1984). The exposure level for 

chloropicrin was 7.9 ppm, which approximated the level sufficient to cause a 50% decrease in 

respiratory rate in mice (RD50) (Kane et al., 1979). Half the exposed mice and half the control 

animals were terminated immediately after the exposures and the other half 72 hours after the 

last exposure. All were examined for respiratory tract lesions. Body weights of chloropicrin 

exposed animals were reduced 10-25% below controls, but increased to normal levels during the 

recovery period. Nasal exudate and distention of the abdomen were observed. “Moderate” 

lesions, characterized by exfoliation, erosion, ulceration, or necrosis, were observed in the 

respiratory and olfactory epithelium, and minimal inflammation and squamous metaplasia were 

observed in the respiratory epithelium alone. Moderate to severe damage to the lower 

respiratory tract was described as “fibrosing peribronchitis and peribronchiolitis”. Exfoliation, 

hyperplasia, and squamous metaplasia were also noted. 

Condie et al. (1994) conducted a study of the toxicity of chloropicrin by oral exposure in 

Sprague-Dawley rats. Ten and ninety-day studies were conducted by dosing animals daily with 

chloropicrin in vehicle (corn oil) at a volume of 1 ml/kg. Groups of 10 rats/sex/group were 

dosed with 0, 10, 20, 40, and 80 mg/kg for the 10-day study and with 0, 2, 8, and 32 mg/kg for 

the 90-day study. Parameters examined included mortality, body weight, food and water 

consumption, hematology, serum clinical chemistry, and gross pathology and histology of 

organs. Only the high-dose group and the control group animals from the 90-day study were 

examined histopathologically. In the 90-day study, 6 males and 2 females in the 32 mg/kg dose 

group and 1 male and 3 females in the 8 mg/kg dose group died before the scheduled termination 

time. The authors noted signs of pulmonary complications (inflammation and congestion) in the 

dead animals. Previously, the animals had shown signs of respiratory distress, including 

Appendix D3 134 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

wheezing and dyspnea. The deaths were considered to be exposure related and most likely due 

to aspiration of chloropicrin. Among the survivors, mean body weight, hemoglobin levels, and 

hematocrits were significantly reduced in males in the 32 mg/kg dose group. Absolute thymus 

weights were reduced in female rats at 32 mg/kg, and female rats in the 8 mg/kg dose group 

showed decreased white blood cell count. Most animals in the 32 mg/kg dose group (>60%) 

showed histopathological changes in the forestomach including chronic inflammation, 

acantholysis, and hyperkeratosis. The authors considered the NOAEL to be 8 mg/kg/day. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Burleigh-Flayer and Benson (1995) 

Study population CD-1 mice (60 per sex per dose) 

Exposure method Discontinuous inhalation (0, 0.1, 0.5 or 1.0 ppm) 

Critical effects Nasal rhinitis; bronchiectasis 

LOAEL 0.5 ppm 

NOAEL 0.1 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 107 weeks 

BMC05 0.042 ppm 

Average experimental exposure 

0.0075 ppm at the BMC05 (0.042 x 6/24 x 5/7) 

Human equivalent concentration 0.0016 ppm at the BMC05 (gas with 

extrathoracic respiratory effects, RGDR = 0.21 

based on MV = 0.044 L/min and SA(ET) = 3 

cm

2

) 

LOAEL uncertainty factor not needed in the BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 (since RGDR adjustment was made) 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.05 ppb (0.4 μg/m3

) 

The data on bronchiecstasis incidence in male and female mice were combined and the chronic 

REL for chloropicrin was developed using the BMC approach. Of the several models tested, the 

Gamma MultiHit Model gave the best fit to the combined bronchiecstasis data (p = 0.9750). The 

MLE05 was 0.070 ppm and the BMC05 was 0.042 ppm. Use of time extrapolation to equivalent 

continuous exposure, an RGDR adjustment for the area of the respiratory tract affected, and a 

total uncertainty factor of 30 resulted in a chronic REL of 0.05 ppb (0.4 μg/m3

). 

The chronic study in mice (Burleigh-Flayer et al., 1995) yielded the same NOAEL of 0.1 ppm as 

the chronic study in rats (Burleigh-Flayer and Benson, 1995). Use of the mouse data with the 

NOAEL/UF approach led to a cREL estimate of 0.1 ppb. Use of the rat data yielded a chronic 

REL estimate of 0.2 ppb by the NOAEL/UF approach. 

Appendix D3 135 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

As another comparison, the study of Yoshida et al. (1987) found a NOAEL in rats of 0.67 ppm 

for intermittent exposure for 13 weeks. This is equivalent to a continuous exposure of 120 ppb. 

Use of an RGDR of 0.25 for rats and a total uncertainty factor of 100 (3 for subchronic, 3 for 

interspecies, and 10 for intraspecies) results in a REL estimate of 0.03 ppb (0.2 μg/m3

). 

VII. Data Strengths and Limitations for Development of the REL 

Significant strengths in the REL for chloropicrin include the duration of exposure (lifetime) in 

the key study, the multiple dose study design with adequate sample sizes, and the demonstration 

of a NOAEL in rats and mice. Major areas of uncertainty are the lack of adequate human 

exposure data, limited reproductive toxicity data, and the appropriateness of time extrapolation 

of concentrations that cause irritative effects such as rhinitis. 

VIII. Potential for Differential Impacts on Children's Health 

Chloropicrin is a respiratory irritant. Respiratory irritants often have steep dose-response curves. 

Thus use of the human intraspecies factor of 10 should result in a REL that adequately protects 

children. Exacerbation of asthma, which has a more severe impact on children than on adults, is 

a known response to some respiratory irritants. However, there is no direct evidence in the 

literature to quantify such a response to chloropicrin, or to quantify a differential effect of 

chloropicrin on infants or children. We are currently evaluating our risk assessment 

methodologies, in particular the intraspecies uncertainty factor (UFH), for adequacy in protecting 

infants and children. While we have not so far identified any indications that the currently used 

UFH of 10 might be less than adequate to protect infants and children, this possibility should be 

considered in evaluating any exposure situation involving chronic exposures of infants or 

children to chloropicrin. 

IX. References 

ACGIH. 1992. American Conference of Governmental Industrial Hygienists, Inc. 

Documentation of the threshold limit values and biological exposure indices. Sixth edition. 

Cincinnati, OH: ACGIH. 

Buckley LA, Jiang XZ, James RA, Morgan KT, and Barrow CS. 1984. Respiratory tract lesions 

induced by sensory irritants at the RD50 concentration. Toxicol. Appl. Pharmacol. 74:417-429. 

Burleigh-Flayer HD, and Benson CL. 1995. Chloropicrin: Vapor inhalation oncogenicity study 

in CD rats. Bushy Run Research Center, July 29, 1995. 

Burleigh-Flayer HD, Kintigh WJ, and Benson CL. 1995. Chloropicrin: Vapor inhalation 

oncogenicity study in CD-1 mice. Bushy Run Research Center, April 20, 1995. 

Appendix D3 136 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

CDPR. 2000. California Department of Pesticide Regulation. Review of Burleigh-Flayer et al. 

(1995) Chloropicrin: Vapor inhalation oncogenicity study in CD-1 mice. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Condie LW, Daniel FB, Olson GR, and Robinson M. 1994. Ten and ninety-day toxicity studies 

of chloropicrin in Sprague-Dawley rats. Drug Chem. Toxicol. 17:125-137. 

DPR. 2000. California Department of Pesticide Regulation. Summary of Pesticide Use Report 

Data – 1999. Sacramento: DPR. 

Duguet JP, Tsutsumi Y, Bruchet A, and Mallevialle J. 1985. Chloropicrin in potable water: 

conditions of formation and production during treatment processes. In: Water Chlorination: 

Chemistry, Environmental Impact and Health Effects. Jolley RL, Bull RJ, Davis WP, Katz S, 

Roberts MH, and Jacobs VA. (eds.) Chelsea, MI: Lewis Publishers, pp. 1201-1213. 

Fries AA, and West CJ. 1921. Chapter VIII. Chloropicrin. In: Chemical Warfare. First edition. 

New York, NY: McGraw-Hill Book Company, Inc. 

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 

7/31/96). 

Kane LE, Barrow CS, and Alarie Y. 1979. A short-term test to predict acceptable levels of 

exposure to airborne sensory irritants. Am. Ind. Hyg. Assoc. J. 40:207-229. 

Merlet N, Thibaud H, and Dore M. 1985. Chloropicrin formation during oxidative treatments in 

the preparation of drinking water. Sci. Total Environ. 47:223-228. 

Tomlin, C.D.S. (ed.). 1994. The Pesticide Manual - World Compendium. 10th ed. Surrey, UK: The British 

Crop Protection Council. p. 192. 

Yoshida M, Ikeda T, Iwasaki M, Ikeda M, Harada T, Ebino K, Tsuda S and Shirasu. 1987. 

Subchronic inhalation toxicity of chloropicrin vapor in rats. J. Pesticide Sci. 12:673-681. 

Appendix D3 137 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY 

CHROMIUM, HEXAVALENT 

(SOLUBLE COMPOUNDS) 

Molecular 

Formula 

Molecular 

Weight Synonyms 

CAS Registry 

Number 

CrO3 99.99 g/mol Chromic trioxide, chromium oxide, 

chromium trioxide, chromium (VI) 

oxide. (In acid aqueous solutions, 

exists as H2CrO4 – “chromic acid”) 

1333-82-0 

K2CrO4 194.20 g/mol Potassium chromate, dipotassium 

chromate, potassium (VI) chromate, 

dipotassium monochromate, 

chromate of potash 

7789-00-6 

Li2CrO4 129.87 g/mol Lithium chromate, chromium lithium 

oxide, chromic acid dilithium salt, 

lithium chromate (VI) 

14307-35-8 

Na2CrO4 161.97 g/mol Sodium chromate, chromic acid 

disodium salt, chromium disodium 

oxide, sodium chromate (VI), 

chromate of soda 

7775-11-3 

K2Cr2O7 294.20 g/mol Potassium dichromate, dichromic acid 

dipotassium salt, bichromate of 

potash 

7778-50-9 

Na2Cr2O7 261.96 g/mol Sodium dichromate, bichromate of 

sodium, dichromic acid disodium 

salt, chromium sodium oxide 

10588-01-9 

I. Chronic Toxicity Summary 

A. Soluble Hexavalent Chromium Compounds (except chromic trioxide) 

Inhalation reference exposure level 0.2 μg Cr(VI)/m3

 

Critical effect(s) Bronchoalveolar hyperplasia in lungs of rats 

Hazard index target(s) Respiratory system 

Oral reference exposure level 0.02 mg Cr(VI)/kg/day 

Critical effect(s) Red blood cell effects (decreased mean 

corpuscular volume (MCV) and mean 

corpuscular hemoglobin (MCH)) in mice 

Hazard index target(s) Hematopoietic system 

Appendix D3 138 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

B. Chromic Trioxide (as chromic acid mist) 

Inhalation reference exposure level 0.002 μg Cr(VI)/m3

 

Critical effect(s) Respiratory effects (nasal atrophy, nasal mucosal 

ulcerations, nasal septal perforations, transient 

pulmonary function changes) in human 

occupational study 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 2000; CRC, 1994) 

Description CrO3: dark red or brown crystals, flakes, or 

powder, exists as chromic acid (H2CrO4) in 

solution; 

K2CrO4, Na2CrO4: yellow crystals; 

K2Cr2O7, Na2Cr2O7: orange-red crystals; 

Li2CrO4: yellow crystalline powder 

Molecular formula See above 

Molecular weight See above 

Density CrO3: 2.70 g/cm3

 @ 25°C 

Boiling point CrO3: decomposes (temperature not available); 

K2Cr2O7: 500 °C with decomposition; 

Na2Cr2O7: 400 °C 

Melting point CrO3: 197 °C; 

K2CrO4: 975 °C; 

Na2CrO4: 792 °C; 

K2Cr2O7: 398 °C; 

Na2Cr2O7: 356.7 °C 

Vapor pressure Not applicable 

Solubility CrO3: soluble in water, ethyl alcohol, ethyl ether, 

sulfuric and nitric acid; 

K2CrO4, K2Cr2O7, Na2Cr2O7: soluble in water, 

insoluble in ethyl alcohol; 

Na2CrO4: soluble in water, slightly soluble in 

ethyl alcohol; 

Li2CrO4: soluble in water and ethyl alcohol 

Conversion factor Not applicable for particulates and mists 

III. Major Uses or Sources 

Hexavalent chromium (Cr(VI)) is considerably more toxic than trivalent chromium (Cr(III)), the 

form most commonly found naturally (ATSDR, 1993). Cr(VI) is generally produced by 

industrial processes. While more information is available on the toxicity of soluble Cr(VI) 

Appendix D3 139 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

compounds, information on poorly soluble Cr(VI) compounds has been included where 

applicable. In California, the major emission source of Cr(VI) results from the chrome plating 

industry (CARB, 1997). Chromic acid, used to electroplate metal parts, is the most common 

Cr(VI) compound produced in the U.S. (ATSDR, 1998). Chromic acid is also registered as a 

fungicide and pesticide in California for use in wood and lumber protection treatments (CDPR, 

1998). Chromic acid solutions used for this purpose in the most recent year of reporting (1998) 

was 71,109 lbs. Minute emissions of Cr(VI) may result from lead chromate in paint used for 

road striping and from coatings in the aerospace and auto refinishing industries, although uses of 

Cr(IV)-containing coatings by these industries in California are decreasing (CARB, 1997 and 

1988). Use of Cr(VI) as a corrosion inhibitor in cooling tower water is prohibited in California, 

and recently, in the remainder of the U.S. as well. Fuel combustion releases trace amounts of 

chromium (CARB, 1988). Most, if not all, of this emitted chromium is in the Cr(III) state. In 

the chromium ferroalloy industry, sodium chromate and dichromate can be produced from 

imported chromite (Cr(III)) ore. However, no such facilities in California have reported 

production or emission of these Cr(VI) compounds. 

Primary routes of potential human exposure to chromium compounds are inhalation, ingestion, 

and dermal contact. Exposure to chromic acid is most often in the form of a mist; exposure to 

other soluble forms of Cr(VI) is as components of aerosols or particulate matter. The physical, 

chemical, and potency differences between Cr(VI) dusts and chromic acid mists necessitated the 

development of separate RELs for each. Environmental exposures would most likely occur 

through exposure to Cr(VI) dusts (U.S. EPA, 1998). Cr(VI) may persist in water as watersoluble complex anions. However, any Cr(VI) settling in the soil or water is expected to be 

eventually reduced to Cr(III) by organic matter. The South Coast Air Quality Management 

District (SCAQMD, 2000) detected ambient levels of hexavalent chromium ranging from 0.0001 

to 0.0003 μg/m3 at 10 stationary monitors placed throughout the South Coast Air Basin. The 

annual statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California based on the most recent inventory were estimated to be 2311 pounds of 

hexavalent chromium (CARB, 2000). 

IV. Effects of Human Exposure 

Cr(VI) forms oxyanions at physiological pH (CrO4

−2

), which are quite similar to sulfate (SO4

-2) 

and phosphate (HPO4

-3) anions. Therefore, it is able to penetrate virtually every cell in the body 

because all cells transport sulfate and phosphate (Costa, 1997). Harmful effects are speculated to 

be related to the reduction of Cr(VI) to Cr(III) intracellularly when it crosses the cell membrane 

and forms complexes with intracellular macromolecules. Thus, Cr(VI) compounds have the 

potential to injure numerous organ systems. Toxicity following chronic Cr(VI) exposure has 

been reported in the respiratory tract, gastrointestinal system, eyes and conjunctiva, kidney, and 

hematopoietic system. Cr(VI) is corrosive and exposure to chromic acid mists may cause 

chronic skin ulcerations and upper respiratory lesions (U.S. EPA, 1998). In addition, allergic 

skin and respiratory reactions can occur with no relation to dose. 

Nasal tissue damage has been frequently observed in chromium plating workers exposed 

chronically to chromic acid mists (Bloomfield and Blum, 1928; Vigliani and Zurlo, 1955; 

Appendix D3 140 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Kleinfeld and Rosso, 1965; Gomes, 1972; Sorahan et al., 1998). However, workers in the 

chromate extraction and ferrochromium industry, exposed to particulates containing soluble 

Cr(VI) compounds, have also reported nasal lesions (Mancuso, 1951; Federal Security Agency, 

1953; Machle and Gregorius, 1948; Wang et al., 1994; Walsh, 1953). Other less frequent 

mucous membrane injuries have been reported in workers exposed to chromate dust and chromic 

acid including sinusitis, laryngitis, conjunctivitis, and oral ulcerations (Mancuso, 1951; Federal 

Security Agency, 1953; Johansen et al., 1994). Nasal lesions include perforated septum, 

ulcerated septum, nasal atrophy, nosebleed, and inflamed mucosa following exposure to air 

chromium levels of about 0.1 to 5.6 mg/m3

. Exposure duration, when reported, ranged from 2 

weeks to 25 years. However, there were problems in quantifying the effect for the above studies. 

The difficulties were primarily lack of adequate methods or data for determining exposure 

duration and/or exposure levels. The occupational studies summarized below provide the most 

reliable estimates of inhalation durations and concentrations resulting in chronic toxicity. 

Workers exposed to >2 μg/m3

 Cr(VI) as chromic acid exhibited an increased incidence of nasal 

atrophy, nasal mucosal ulcerations, and nasal septal perforations as compared to controls 

(Lindberg and Hedenstierna, 1983). Workers exposed to less than 2 μg/m3 (expressed as < 1.9 

μg/m3

) exhibited an increased incidence of irritated nasal mucosa and nasal atrophy compared to 

controls. The median exposure time of exposed workers was 2.5 years (range = 0.2-23.6 years). 

Frequency of throat and chest symptoms was similar to that of controls. The same study 

reported statistically significant decreases in forced expiratory volume in 1 second (FEV1), 

forced vital capacity (FVC), and mean forced expiratory flow during the middle of the FVC in 1 

second (FEF25-75) measurements taken on a Thursday afternoon as compared to those taken on a 

Monday morning in nonsmoking workers exposed to 2 μg/m3 Cr(VI) or more. Similar changes 

were observed in the smokers although only the difference in the FVC measured on a Thursday 

was statistically significant. No significant differences were observed between pulmonary 

function measurements of exposed and unexposed workers taken on a Monday morning (prior to 

a work week of exposure). Thus the authors infer that the observed pulmonary function changes 

are transient. 

Nasal lesions were observed in 35 of 37 chrome platers exposed to a mean breathing zone 

concentration of 7.1 μg/m3 (range = 1.4-49.3 μg/m3) total chromium for an average of 2.2 years 

(range = 1.2 weeks-11 years) (Cohen et al., 1974). Actual exposure to Cr(VI) averaged 2.9 

μg/m3 (range = 0.09-9.1 μg/m3). Workers employed more than one year had significantly 

greater nasal pathology than workers employed one year or less. Due to poor personal hygiene 

habits of the exposed workers, a ‘direct contact’ etiology may explain some of the nasal lesions. 

Urinary levels of β2-microglobulin in 24 chrome platers increased in dose-dependent fashion 

with increasing intensity of exposure to Cr(VI), indicating a nephrotoxic effect resulting from 

inhalation of Cr(VI) (Lindberg and Vesterberg, 1983). The 8-hr mean Cr(VI) levels ranged from 

2 to 20 μg/m3 and averaged 6 μg/m3. Total exposure times ranged from 0.1 to 26 years and 

averaged 5.3 years. Most of the 24 chrome workers had irritation symptoms of the airways. As 

a group, the chrome platers had significantly higher levels of urinary β2-microglobulin compared 

to a group of 27 referents. Comparison of 27 referents to a group of 27 ex-chrome-platers found 

no difference in urinary β2-microglobulin levels, even though seven of the ex-chrome-platers had 

a permanent perforation of their nasal septum (indicating past exposure to high levels of Cr(VI)). 

Appendix D3 141 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

There was no correlation between total exposure time and urinary β2-microglobulin levels. 

Urinary albumin levels remained unchanged in the Cr(VI)-exposed group. The results suggest 

that the nephrotoxic effects are reversible at the exposure levels studied. 

Gastritis and duodenal ulcers, in addition to ulceration and perforation of the nasal septum, were 

observed in chrome platers exposed to a mean breathing zone concentration of 4 μg/m3

 chromic 

acid for an average of 7.5 years (Lucas and Kramkowski, 1975). 

Male workers in the chromate and dichromate production industry, whose occupational 

exposures were 0.05-1.0 mg Cr(VI)/m3

 as chromium trioxide for a mean of 7 years, were 

reported to have elevated levels of low molecular weight proteins (retinol binding protein and 

tubular antigens) in the urine (Franchini and Mutti, 1988). The authors suggest that the presence 

of such proteins in the urine is an early indicator of kidney damage. 

The respiratory health of workers exposed to low levels of dusts containing Cr(VI) was 

investigated at a stainless steel production plant (Huvinen et al., 1996). The data were presented 

as total chromium exposure and Cr(VI) exposure. A combined total of 109 exposed workers in 

the furnace department (median Cr(VI) exposure approximately 0.075-0.45 μg/m3) and the steel 

smelting shop (average Cr(VI) exposure 0.5 μg/m3) was compared to a control group of 95 

workers that worked in the cold rolling mill. Total work exposure duration was 16.0 years 

(range: 8-26 years). No significant differences in lung function tests and radiological findings 

were observed between exposed and control workers. After controlling for age and smoking, no 

differences were observed for the prevalence of rhinitis, eye irritation, or respiratory symptoms 

between the two groups. 

In a study summarized by U.S. EPA (1998), oral ulcers, diarrhea, stomach ache, indigestion, 

leukocytosis and vomiting were reported among a group of 155 Chinese villagers exposed to 

contaminated well-water containing 20 mg/L Cr(VI) in 1965 (Zhang and XiLin, 1987). 

However, precise exposure concentrations, exposure durations, and confounding factors were not 

provided. A follow-up study to assess cancer mortality reported that the average Cr(VI) 

concentration in 1965 from 170 wells of the most impacted village was only 2.6 ppm, and 

maximum levels did not exceed 5 ppm (Zhang and Li, 1997). Non-cancer effects were not 

presented and the apparent discrepancy in water levels of Cr(VI) with the earlier study was not 

discussed. 

V. Effects of Animal Exposure

Exposure of C57BL/6 mice to 0 or 13 mg/m3 CaCrO4 dust (about 136 animals/sex/group) 5 

hr/day, 5 days/wk for life resulted in emphysema-like changes of the lung, ‘bronchiolarization’ 

of the alveoli, and epithelial necrosis, marked hyperplasia, and atrophy of the bronchi in treated 

mice (Nettesheim et al., 1971). Other non-cancer histopathological findings in exposed mice 

included atrophy of the lymph nodes, spleen, and liver, and occasional small ulcerations of the 

stomach and intestinal mucosa. Cessation of body weight gain in both sexes was observed 

following the sixth month of exposure to the chromate dust. 

Appendix D3 142 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Glaser et al. (1986) exposed 20 male Wistar rats/group to 25, 50, and 100 μg/m3 aerosolized 

sodium dichromate solution and to 100 μg/m3 of a pyrolyzed Cr(VI)/Cr(III) (3:2) oxide dust 

mixture 22-23 hr/day for 18 months. Observation in filtered air continued for another 12 months 

thereafter. A control group consisted of 40 rats. Mortality and body weights were unaffected by 

treatment. Lung chromium retention at the end of the study was 10-fold greater in rats exposed 

to the slightly water soluble chromium oxide mixture compared to high dose rats exposed to 

water-soluble sodium dichromate. No clinical signs of irritation were observed in any group. 

No hematological effects were noted in rats exposed to sodium dichromate. Rats exposed to the 

chromium oxide mixture had a significantly elevated white blood cell count at the 17th and 18th

month, and significantly elevated red blood cells, hematocrits, and hemoglobin levels at the 27th

month. Mean serum content of total immunoglobulin was significantly reduced in this group at 6 

months exposure. Significantly increased lung weights were observed in chromium oxideexposed rats, and for livers of sodium dichromate-exposed rats at the highest dose. Pigmentloaded macrophages were found in the sodium dichromate-exposed rats in a dose dependent 

manner and also in the chromium oxide group. Chromium oxide-exposed rats also developed 

focal thickened septa, partially combined with interstitial fibrosis and accumulation of 

eosinophilic substance in the alveolar lumens. The authors concluded that the hematological and 

pulmonary effects may be due to Cr-accumulation in the lungs and to depressed lung clearance 

function. 

Rats exposed to 200 μg/m3 Cr(VI) as aerosolized sodium dichromate by inhalation for 22 hours 

per day for 42 days exhibited decreased alveolar macrophage phagocytic activity; the lung 

clearance of inert iron oxide was significantly reduced in exposed rats compared to controls 

(Glaser et al., 1985). Increased alveolar macrophage activity and a significantly elevated 

antibody response to injected sheep red blood cells were observed in rats exposed to 25 or 50 

μg/m3 Cr(VI) for 22 hours per day for 28 days. Ninety day exposure under the same exposure 

protocol resulted in increased rat lung and spleen weights at 50, 100 and 200 μg/m3

, but not 25 

μg/m3 (Glaser et al., 1985). Histopathology of major organs was similar among all groups. 

Bronchoalveolar lavage fluid contained decreased macrophage cell counts above 25 μg/m3. 

Increased antibody response to injected sheep red blood cells was observed in all treatment 

groups, while alveolar macrophage activity was elevated at 25 and 50 μg/m3

, but was 

significantly reduced at 200 μg/m3

. 

A later experiment exposed male rats to 0, 50, 100, 200, or 400 μg Cr/m3 22 hours per day, 7 

days per week for 90 days (Glaser et al., 1990). Average measured concentrations were 0, 54, 

109, 204, and 403 μg Cr/m3, respectively. Subacute respiratory dyspnea and reduction in body 

weight gain were observed at the two highest exposures. Mean white blood cell count increased 

in a dose-dependent manner among treated rats, but returned to normal 30 days following 

cessation of exposure. Histopathological examination revealed histiocytosis (macrophage 

accumulation) in all treatment groups (Table 1). Bronchoalveolar lavage fluid (BALF) contained 

elevated levels of albumin, lactate dehydrogenase (LDH), and total protein in all exposed groups. 

Statistically significant elevations in these parameters were observed mainly in the 200 and 

400 μg/m3 exposure groups. At necropsy, a statistically significant increase in lung weight (g 

dry wt/kg body wt) was observed in rats exposed to 100, 200, and 400 μg/m3

 as compared to 

controls. Lung weights were still significantly elevated in the three highest exposure groups 30 

days following cessation of exposure. An analysis of the data (Malsch et al., 1994) determined a 

Appendix D3 143 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

benchmark dose (95% confidence interval with dose associated with a 10% elevation in the 

parameter) for each of these endpoints. The analysis also examined changes in lung and spleen 

weight reported in Glaser et al. (1985). The most sensitive endpoint was LDH in BALF. 

Table 1. Key bronchoalveolar lavage fluid (BALF) and histopathological findings after 90 

days exposure to sodium dichromate (Glaser et al., 1990). 

μg Cr/m3

Total Protein 

in BALFa

(mg/L) 

Albumin in 

BALF (mg/L) 

LDH 

in BALF 

 (U/L) 

Bronchoalveolar 

Hyperplasia 

Lung 

Histiocytosis 

Right lung 

dry weight 

(g/kg BW) 

0 226+30 77+13 29+5 0/10 2/10 0.44± 0.03 

50 396+79** 115+23** 34+3* 3/10 9/10 0.48±0.05 

100 326+35** 86+13 31+4 2/10 10/10 0.50±0.06* 

200 703+178** 117+20** 63+11** 3/10 9/10 0.55±0.04** 

400 975+246** 184+59** 83+17** 7/10 10/10 0.65±0.05** 

a All BALF parameters are mean + SD, n = 10/group 

* p < 0.05; ** p < 0.001: comparison of exposed groups vs. controls 

Cohen et al. (1998) investigated the immunotoxicologic effects of inhaled chromium by 

exposing F-344 rats (10/group/exposure duration) nose-only to 0 and 360 μg/m3 potassium 

chromate 5 hr/day, 5 days/week for 2 or 4 weeks. Exposed rats had greater levels of total 

recoverable cells, neutrophils, and monocytes in bronchopulmonary lavage compared to controls 

at 2 and/or 4 weeks. Pulmonary macrophages (PM) were reduced, although total PM levels 

remained unaffected. Four-week exposure to potassium chromate also resulted in modulated 

PM-inducible interleukins-1 and –6, and tumor necrosis factor-α, and increased PM basal nitric 

oxide production and interferon-γ-primed/zymosan-stimulated reactive oxygen intermediate 

production. 

Nasal septal perforation, hyperplastic and metaplastic changes in the larynx, trachea, and 

bronchus, and emphysema were observed in mice exposed two days per week for 12 months to 

CrO3 mist (Adachi, 1987; Adachi et al., 1986). Chromic acid concentrations were either 3.63 

mg/m3

 for 30 minutes per day or 1.81 mg/m3 for 120 minutes per day. An additional 20 mice 

exposed to 1.81 mg/m3

 were necropsied 6 months after the last exposure. Lesions of the nasal 

septum, trachea, and lungs were still evident in some mice. 

The investigators of the toxicity studies summarized below administered soluble Cr(VI) 

compounds to experimental animals by the oral route. 

Groups of eight male and eight female Sprague-Dawley rats were supplied with drinking water 

containing 0-11 ppm (0-11 mg/L) Cr(VI), as K2CrO4, for 1 year (Mackenzie et al., 1958). The 

control group (10/sex) received distilled water. A second experiment involved three groups of 

12 male and 9 female rats. One group was given 25 ppm (25 mg/L) Cr(VI); a second received 

25 ppm chromium in the form of chromic chloride; and the controls received distilled water. For 

rats treated with 0-11 ppm (in the diet), hematological determinations (red and white blood cell 

counts, differential white cell counts, and hemoglobin) were performed monthly, and tissues 

(livers, kidneys and femurs) were examined at 6 months and 1 year. Spleens were also examined 

Appendix D3 144 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

at 1 year. The 25 ppm groups (and corresponding controls) were examined similarly, except that 

no animals were killed at 6 months. No significant adverse effects were seen in appearance, 

weight gain, or food consumption, and there were no treatment-related effects regarding 

hematological parameters or other tissues in any treatment group. The rats receiving 25 ppm 

Cr(VI) showed an approximate 20% reduction in water consumption. This dose corresponds to 

2.4 mg Cr(VI)/kg/day based on actual body weight and water consumption data. An abrupt rise 

in tissue chromium concentrations was noted in rats treated with greater than 5 ppm. The 

authors stated that “apparently, tissues can accumulate considerable quantities of chromium 

before pathological changes result.” In the 25 ppm treatment groups, tissue concentrations of 

chromium were approximately 9 times higher for those treated with hexavalent chromium than 

for the trivalent group. 

Anwar et al. (1961) observed no significant effects in groups of female dogs (2/dose group) 

given 0, 0.45, 2.25, 4.5, 6.75, or 11.2 ppm Cr(VI) (as K2CrO4) in drinking water for 4 years. The 

calculated doses ranged from 0.012-0.30 mg/kg of Cr(VI). 

Numerous rodent studies have been recently undertaken to investigate the reproductive and 

developmental effects of Cr(VI) exposure via the drinking water (Trivedi et al., 1989; Junaid et 

al., 1995; Murthy et al., 1996; Junaid et al., 1996a; Junaid et al., 1996b; Kanojia et al., 1996; 

Elbetieha and Al-Hamood, 1997; Al-Hamood et al., 1998; Kanojia et al., 1998). Exposure 

concentrations ranged from 250 to 5000 ppm for durations as short as five days during gestation 

to as long as 3 months pre-gestational exposure. In general, the longer exposures resulted in 

more serious reproductive and developmental effects. 

Kanojia et al. (1998) administered 0, 250, 500, and 750 ppm potassium dichromate via drinking 

water to female Druckrey strain rats for 90 days prior to gestation. Based on daily water intake 

and final body weights, the estimated daily Cr(VI) intake was 33, 68, and 98 mg/kg-day, 

respectively. Ten to 15% mortality, hair loss, lethargy, aggressiveness and a significant 

reduction in body weight gain were observed in rats at the two highest doses. While not 

statistically significant, weight of the low dose rats were 32% lower than controls. All treated 

rats were acyclic at the end of the 90 day exposure period and an additional 15-20 days without 

Cr(VI) exposure were needed for the estrus cycle to start. Mating and fertility indexes decreased 

with increasing Cr(VI) intake. Ten rats/group were sacrificed on day 19 of gestation for 

fetotoxicity assessment. Significantly reduced fetal weight and increased pre- and postimplantation loss occurred at all dose levels. Gross and skeletal abnormalities in low dose 

fetuses included subdermal hemorrhagic patches, drooping wrists, and reduced caudal bone 

ossification. No gross visceral abnormalities were seen in treated groups. 

Administration of potassium dichromate to rats (Kanojia et al., 1996) and mice (Junaid et al., 

1996a) in drinking water at concentrations of 250, 500, and 750 ppm for 20 days prior to 

gestation resulted in increased post-implantation loss and decreased placental weight in both 

species at the lowest dose. Also at this dose level, decreased fetal weight and crown-rump length 

were observed in mice, and increased resorptions and decreased number of live fetuses were 

observed in rats. Gross and skeletal abnormalities were observed in both species beginning at 

the 500 ppm dose level. 

Appendix D3 145 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Groups of Sprague-Dawley rats (NTP, 1996a) and BALB/C mice (NTP, 1996b) were 

administered potassium dichromate in their diet at 0, 15, 50, 100, or 400 ppm for 9 weeks (24 

males and 48 females/species/group) followed by a recovery period of 8 weeks. Average Cr(VI) 

consumption for male/female rats were 1/1, 3/3, 6/7, and 24/28 mg/kg-day, respectively. 

Average Cr(VI) consumption for male/female mice were 3/5, 10/16, 21/34, and 92/137 mg/kgday, respectively. Six males and 12 females of both species were necropsied after 3, 6, or 9 

weeks of treatment or after the full recovery period. There was no treatment-related 

histopathology observed in kidneys, ovaries, and testes in either species. Hematological analysis 

revealed slight decreases in mean corpuscular volume (MCV) and mean corpuscular hemoglobin 

(MCH) at the highest dose in both species, which is indicative of iron deficiency. MCV and 

MCH were normal in these groups following the 8-week recovery period. Microscopic 

evaluation of the livers of mice noted cytoplasmic vacuolization of hepatocytes in treated 

animals beginning at 50 ppm. Also in mice, there was a slight decrease in mean body weights in 

the 400 ppm males (5-9%) and females (4%) and the 100 ppm females (2-4%) during the dosing 

periods. Feed consumption by mice was generally increased in all treated groups, particularly 

the 400 ppm males and females. During the recovery period, feed consumption was comparable 

across groups. 

The NTP (1997) investigated the potential reproductive toxicity of Cr(VI) in mice using the 

Reproductive Assessment by Continuous Breeding protocol. Groups of 20 male and female 

pairs of BALB/c mice (F0) were exposed to 0, 100, 200, and 400 ppm potassium dichromate in 

their diet during the continuous breeding phase (approximately 12 weeks). F1 generation litters 

received the same concentration of Cr(VI) in their diet as their F0 parents and were used for 

assessment of second generation reproductive toxicity at sexual maturity. There were no 

treatment-related changes in any of the reproductive parameters in this study. In F1 mice, the 

MCV was slightly decreased in males at the two highest doses, and slightly decreased in females 

in all dose groups. MCH and hemoglobin were slightly reduced in high dose males and high 

dose females, respectively. Mean body weights of the high dose F0 and F1 animals were slightly 

decreased, and mean food consumption in the F1 mice was elevated. Reduced mean absolute 

liver weights were observed in 400 ppm F0 mice of both sexes. The mean calculated doses were 

19.4, 38.6, and 85.7 mg/kg-day for F0 males and females and 22.4, 45.5, and 104.9 mg/kg-day 

for F1 males and females in the 100, 200, and 400 ppm dose groups, respectively. 

In an investigation of the spermatogenic and steroidogenic effects of Cr (VI), Chowdhury and 

Mitra (1995) administered 0, 20, 40, and 60 mg/kg-day sodium dichromate by oral gavage to 

male rats for 90 days. Reduced Leydig cell population, reduced body and testicular weight, and 

degeneration of testicular tissue was observed at the two highest doses. Biochemical measures 

of spermatogenic and steroidogenic impairment, including decreased testicular DNA, RNA, 

protein, serum testosterone, and 3β-∆5

-hydroxy steroid dehydrogenase (3β-∆5-HCH), were also 

reduced at the two highest doses. Only relatively small reductions in testicular protein, 3β-∆5

-

HCH, and serum testosterone were seen in the 20 mg/kg rats. 

Appendix D3 146 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

VI. Derivation of Chronic Reference Exposure Levels (RELs) 

A. Derivation of Chronic Inhalation Reference Exposure Level for Soluble Hexavalent 

Chromium Compounds other than Chromic Trioxide 

Study Glaser et al., 1990 

Study population Male Wistar rats (30 per group) 

Exposure method Discontinuous whole-body inhalation (0, 54, 

109, 204, or 403 μg Cr(VI)/m3 as sodium 

dichromate aerosol) 

Critical effects Bronchoalveolar hyperplasia 

LOAEL 50 μg/m3

 

NOAEL Not observed 

BMC05 12.50 μg/m3

Exposure continuity 22 hr/day, 7 days/week 

Exposure duration 90 days 

Average exposure 11.46 μg/m3

 Cr(VI) (12.50 x 22/24) 

Human equivalent concentration 24.47 μg/m3

 Cr(VI) (2.1355 [RDDR] x 11.46) 

LOAEL uncertainty factor Not needed in the BMC approach 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.2 μg/m3

 (0.0002 mg/m3) 

The study by Glaser et al. (1990) provides the best available inhalation data that demonstrate a 

dose-response relationship for various pulmonary toxicity endpoints. The BMC05 of 12.50 

μg/m3 was derived from quantal data for bronchoalveolar hyperplasia. The presence of 

bronchoalveolar hyperplasia in exposed rats is supported by other indicators of lung 

inflammation, including increased total protein, LDH, and albumin in BALF (see Table 1). A 

quantal-linear model analysis (U.S. EPA, National Center for Environmental Assessment, 

benchmark dose software, version 1.20) of the quantal data provided the most reasonable line fit 

and resulted in the lowest BMC05. A BMC05 is considered to be similar to a NOAEL in 

estimating a concentration associated with a low level of risk. Lung histiocytosis (macrophage 

accumulation) was present in nearly all exposed animals, but this quantal data set was only 

suitable for a NOAEL/LOAEL approach and was not considered as direct an indicator of lung 

injury as bronchoalveolar hyperplasia. 

Based on OEHHA methodology, a comparison REL developed using the NOAEL/LOAEL 

approach would yield 0.3 μg/m3. Adjustment of the LOAEL of 50 μg/m3 (a NOAEL was not 

observed) to the human equivalent concentration uses the same parameters as shown in the REL 

derivation above. However, a LOAEL UF of 3 is added to the existing UFs to result in a 

cumulative UF of 300. 

The U.S. EPA (1998) RfC of 0.1 μg/m3 is also based on data from Glaser et al. (1990), but 

derived a BMC10 (16 μg/m3

), as developed by Malsch et al. (1994), from continuous data of 

Appendix D3 147 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

LDH in BALF. Using a polynomial model provided by a different benchmark software package 

(THC, Clement International Corp., Ruston LA), increasing LDH concentration in BALF with 

increasing dose provided the lowest BMC10 among the various BALF endpoints. OEHHA is 

currently not developing BMCs for RELs based on continuous data. A BMC05 derived from 

quantal data and a BMC05 derived from continuous data may not have the same meaning. 

Conceivably, depending on the standard deviations of the data points, the BMC05 based on 

continuous data could still be above the statistically significant effect level. OEHHA believes 

that further evaluation of BMC’s based on continuous data is needed prior to their application to 

RELs. 

OEHHA and U.S. EPA also diverge on the assignment of the Subchronic UF. The Glaser et al. 

(1990) study indicated that chromium was still acculmulating in lung tissue at the end of 90 days. 

This evidence and the fact that the study did not investigate upper airway effects and other extrapulmonary effects led U.S. EPA to assign a subchronic UF of 10 (U.S. EPA, 1998). Based on 

OEHHA methodology, OEHHA used a subchronic UF of 3. In support of a UF of 3, the 18-

month sodium dichromate exposure study performed by Glaser et al. (1986), under similar 

exposure conditions used in the key 90-day study, did not find histopathological evidence of lung 

inflammation or major organ effects, or suggest severe chromium accumulation in exposed rats. 

However, BALF analysis was not performed in the chronic study. 

For comparison with the proposed REL, the occupational study by Huvinen et al. (1996) 

established a NOAEL of 0.5 μg/m3 for lack of pulmonary findings. However, this study is 

deficient for REL purposes due to the lack of a LOAEL. Unfortunately, other occupational 

studies suffered from lack of adequate methods or data for determining exposure duration and/or 

exposure levels. Use of an occupational time adjustment (10/20 m3 inhaled/day, 5/7 days/week) 

and an interspecies UF of 10 for the Huvinen et al. (1996) study would result in an estimated 

REL of 0.02 μg/m3. Average exposure duration was 16 years, so a subchronic UF of 1 was 

sufficient. 

B. Derivation of Chronic Inhalation Reference Exposure Level for CrO3 as Chromic Acid 

Appendix D3 148 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Study Lindberg and Hedenstierna, 1983 

Study population Human workers (100 exposed workers, 

119 unexposed controls) 

Exposure method Occupational exposure to chromic acid mist 

Critical effects Nasal atrophy, nasal mucosal ulcerations, nasal 

septal perforations, transient pulmonary 

function changes 

LOAEL 1.9 μg/m3 established as “low exposure” group 

(8-hr mean < 1.9 μg/m3

) 

NOAEL Not observed 

Exposure continuity 8 hr/day (10 m3

 per 20 m3 day), 5 days/week 

Exposure duration Mean of 2.5 years (range = 0.2 - 23.6 years) 

Average exposure 0.68 μg/m3

 Cr(VI) (1.9 x 10/20 x 5/7) 

Human equivalent concentration 0.68 μg/m3

 Cr(VI) 

LOAEL uncertainty factor 3 

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.002 μg/m3

 (0.000002 mg/m3) 

The occupational exposure study of Lindberg and Hedenstierna (1983) was selected as the best 

available human study. A 3-fold LOAEL to NOAEL uncertainty factor (UF) was applied due to 

the low incidence of nasal atrophy at the LOAEL (4 out of 19) and the apparent reversibility of 

the lesion at this exposure level. While Lindberg and Hedenstierna (1983) did not follow-up on 

any of the active cases of nasal ulcerations, which occurred only in workers in the ‘high 

exposure’ group, they did note that one worker, who exhibited nasal atrophy, had no visible nasal 

lesions 4 months after termination of exposure. 

U.S. EPA (1998) based its RfC of 0.008 μg/m3 for exposure to chromic acid mists and dissolved 

Cr(VI) aerosols on the same study but established the LOAEL at 2 μg/m3

 and applied a total UF 

of 90 (3 each for the LOAEL to NOAEL and subchronic to chronic extrapolation, and 10 for 

intraspecies extrapolation). It was unclear why U.S. EPA (1998) chose UFs of 3 for LOAEL and 

subchronic extrapolations. It was also unclear why the total uncertainty factor was 90, rather 

than 100, which would be obtained by following the usual convention (that the value for 

uncertainty factors of “3” is actually 3.16, the square root of 10, although it is usually only 

quoted to 1 significant figure). 

For comparison, a REL can be estimated from the Adachi et al. (1987) study in which mice were 

exposed to 1.81 mg/m3 chromic acid mist 2 hr/day, twice a week for 12 months. Lesions were 

observed in treated mice throughout the respiratory tract; a NOAEL was not determined. 

Application of the exposure continuity adjustment (2/24 hr/day x 2/7 days/week), an RDDR of 

2.26 (MMAD and sigma g roughly estimated at 5 and 3 μm, respectively), and a total UF of 300 

(10 for LOAEL to NOAEL, 3 for interspecies, and 10 for intraspecies) yields a REL of 0.3 

μg/m3

. 

Appendix D3 149 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

In addition to being inhaled, airborne hexavalent chromium can settle onto crops and soil and 

enter the body by ingestion. Thus, an oral chronic reference exposure level for soluble salts of 

metallic chromium(VI) is also required for assessing risks from stationary sources in the Air 

Toxics Hot Spots program. 

C. Derivation of Chronic Oral Reference Exposure Level for Chromium VI (Based on 

U.S. EPA RfD) 

Study Mackenzie et al., 1958 

Study population 8 male and 8 female Sprague-Dawley rats 

Exposure method Drinking water 

Critical effects No adverse effects seen 

LOAEL None 

NOAEL 2.4 mg/kg-day (converted from 25 mg/L of 

chromium as K2CrO4) 

Exposure continuity Continuous

Exposure duration 1 year 

Average experimental exposure 2.4 mg/kg-day (0.11 ppm Cr(VI)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10

Cumulative uncertainty factor 100 

Oral reference exposure level 0.02 mg/kg bw-day 

The oral REL (0.02 mg/kg bw-day) and U.S. EPA’s oral Reference Dose (RfD) of 0.003 mg/kgday (U.S. EPA, 1998) are based on the same study by MacKenzie et al. (1958). No adverse 

effects were reported at any dose in the study. The highest dose group (25 mg/L) was selected 

for derivation of the oral REL and RfD based on the reported body weight of the rat (0.35 kg) 

and the reported average daily drinking water consumption for the rat (0.035 L/day). Because a 

LOAEL was not observed in the primary study, the subchronic NTP studies provide supporting 

evidence to justify a REL based on MacKenzie et al. (1958). Cr(VI) was administered in the diet 

of rats for 9 weeks and a NOAEL of 6 mg/kg-day was observed for slightly depressed MCV and 

MCH values (NTP, 1996a). The LOAEL was 24 mg/kg-day. The NTP (1996b, 1997) also 

observed slightly depressed MCV and MCH values in mice, but at higher Cr(VI) concentrations. 

While the changes are small and may be a mild adverse effect at best, the NTP (1997) noted that 

decreased MCV and MCH are indicators of iron deficiency and suggested that an interaction 

between chromium and iron is altering erythrocyte formation. The liver effects noted in female 

mice in the 9 week study (NTP, 1996b) were not observed in the mouse reproductive study 

(NTP, 1997). Therefore, the toxicological significance of this finding is uncertain. 

U.S. EPA (1998) applied UFs of 3 for subchronic, 10 for intraspecies, 10 for interspecies, and a 

modifying factor of 3 (to account for concerns raised by the study of Zhang and XiLin (1987)) to 

the NOAEL for an RfD of 0.003 mg/kg-day. The criteria for use of modifying factors are not 

Appendix D3 150 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

well specified by U.S. EPA. Such modifying factors were not used by OEHHA. Because the 

exposure duration in the primary study was greater than 12% of the estimated lifespan of rats, 

OEHHA applied UF of 1 for extrapolation to chronic exposure. 

U.S. EPA stated its confidence in the RfD as: Study - Low; Data Base - Low; and RfD - Low. 

Confidence in the chosen study is low because of the small number of animals tested, the small 

number of parameters measured, and the lack of toxic effect at the highest dose tested. 

Confidence in the database is low because the supporting studies are of equally low quality, and 

teratogenic and reproductive endpoints are not well studied. Low confidence in the RfD follows. 

OEHHA notes that more reproduction/developmental studies have been published that support 

the RfD and oral REL since U.S. EPA published its findings (U.S. EPA, 1998). In general, these 

studies indicate that reproductive and developmental effects occur at doses greater than an order 

of magnitude above the NOAEL established by MacKenzie et al. (1958) and the NTP (1996a,b, 

1997). However, the dose levels used were relatively high such that a NOAEL was typically 

lacking. 

VII. Data Strengths and Limitations for Development of the REL

The major strength of the inhalation REL for chromic acid mist is the use of human data. The 

major uncertainties for this inhalation REL is the lack of controlled and quantified exposure data 

and the lack of a NOAEL in the key chromic acid study. 

The suitably thorough analysis of lower airway effects and the development of a BMC from 

continuous data are strengths for the Cr(VI) dust inhalation REL. Limitations include the lack of 

comprehensive data on multi-organ effects, the lack of chronic studies, the lack of upper airway 

analysis in the key study, and the lack of quantified exposure data in humans. The animal studies 

by Glaser et al. (1990, 1986) suggest that the lower respiratory airway is a primary target for 

Cr(VI) dusts. However, occupational studies (Mancuso, 1951; Federal Security Agency, 1953; 

Machle and Gregorius, 1948; Wang et al., 1994; Walsh, 1953) indicate that nasal lesions result 

from exposure to Cr(VI) dusts and may, in fact, be the most sensitive indicator of human toxicity 

resulting from exposure to soluble Cr(VI) dusts. However, this finding is attenuated by the fact 

that dermal exposure to chromic acid and Cr(VI) dusts due to poor hygienic practices of workers 

may overestimate the airborne concentrations necessary to result in nasal lesions. 

The major strength for the oral REL is the consistency of the doses resulting in NOAELs and/or 

LOAELs among the major and supporting studies. The major limitations for the oral REL, other 

than the ones noted above by U.S. EPA, are the lack of lifetime exposure studies in experimental 

animals and the lack of adequate oral human exposure data. 

VIII. References 

Appendix D3 151 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

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Al-Hamood MH, Elbetieha A, and Bataineh H. 1998. Sexual maturation and fertility of male and 

female mice exposed prenatally and postnatally to trivalent and hexavalent chromium 

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Chromium. Prepared by Syracuse Research Corporation under subcontract to Clement 

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ATSDR. 1998. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

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Anwar RA, Langham, FF, Hoppert CA, Alfredson BV, and Byerrum RU. 1961. Chronic toxicity 

studies. III. Chronic toxicity of cadmium and chromium in dogs. Arch. Environ. Health 3:456-

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Bloomfield JJ, and Blum W. 1928. Health hazards in chromium plating. Public Health Rep 

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CARB. 1988. California Air Resources Board. Proposed Hexavalent Chromium Control Plan. 

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CARB. 1997. California Air Resources Board. Toxic Air Contaminant Identification List 

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CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CDPR. 1998. California Department of Pesticide Regulation. Summary of Pesticide Use Report 

Data, 1998, Indexed by Chemical. http://www.cdpr.ca.gov/docs/pur/pur98rep/chmrpt98.pdf 

Chowdhury AR, and Mitra C. 1995. Spermatogenic and steroidogenic impairment after 

chromium treatment in rats. Indian J. Exp. Biol. 33:480-484. 

Appendix D3 152 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Cohen MD, Zelikoff JT, Chen L-C, and Schlesinger RB. 1998. Immunotoxicologic effects of 

inhaled chromium: Role of particle solubility and co-exposure to ozone. Toxicol. Appl. 

Pharmacol. 152:30-40. 

Cohen SR, Davis DM, and Kramkowski PE. 1974. Clinical manifestations of chromic acid 

toxicity: Nasal lesions in electroplate workers. Cutis 13:558-568. 

Costa M. 1997. Toxicity and carcinogenicity of Cr(VI) in animal models and humans. Crit. Rev. 

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CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

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Federal Security Agency. 1953. Health of workers in chromate producing industry. A study. U.S. 

Public Health Service Publication No. 192. Washington, DC. U.S. Government Printing Office. 

Franchini I, and Mutti A. 1988. Selected toxicological aspects of chromium (VI) compounds. 

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Glaser U, and Hochrainer H. 1988. Report on the subchronic inhalation toxicity of sodium 

dichromate in rats with attachment and cover letter dated 112089. Industrial Health Foundation 

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Glaser U, Hochrainer D, Kloppel H, and Oldiges H. 1986. Carcinogenicity of sodium dichromate 

and chromium (VI/III) oxide aerosols inhaled by male Wistar rats. Toxicology 42:219-232. 

Glaser U, Hochrainer D, Kloppel H, and Kuhnen H. 1985. Low level chromium (VI) inhalation 

effects on alveolar macrophages and immune functions in Wistar rats. Arch. Toxicol. 57:250-

256. 

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Cr(VI) with possible influence on its carcinogenic action. In: Environmental Hygiene II. 

Seemayer NO, Hadnagy W. (eds.) Berlin/New York: Springer-Verlag. 

Gomes ER. 1972. Incidence of chromium-induced lesions among electroplating workers in 

Brazil. Ind. Med. 41(12);21-25. 

HSDB. 2000. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD. 

Available online through Toxicology Data Network at http://toxnet.nlm.nih.gov 

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Respiratory health of workers exposed to low levels of chromium in stainless steel production. 

Occup. Environ. Med. 53:741-747. 

Appendix D3 153 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Johansen M, Overgaard E, and Toft A. 1994. Severe chronic inflammation of the mucous 

membranes in eyes and upper respiratory tract due to work-related exposure to hexavalent 

chromium. J. Laryngol. Otol. 108:591-592. 

Junaid M, Murthy RC, and Saxena DK. 1995. Chromium fetotoxicity in mice during late 

pregnancy. Vet. Hum. Toxicol. 37(4):320-323. 

Junaid M, Murthy RC, and Saxena DK. 1996a. Embryo- and fetotoxicity of chromium in 

pregestationally exposed mice. Bull. Environ. Contam. Toxicol. 57:327-334. 

Junaid M, Murthy RC, and Saxena DK. 1996b. Embryotoxicity of orally administered chromium 

in mice: Exposure during the period of organogenesis. Toxicol. Lett. 84:143-148. 

Kleinfeld M, and Rosso A. 1965. Ulcerations of the nasal septum due to inhalation of chromic 

acid mist. Ind. Med. Surg. 34:242-243. 

Kanojia RK, Junaid M, and Murthy RC. 1996. Chromium induced teratogenicity in female rat. 

Toxicol. Lett. 89:207-213. 

Kanojia RK, Junaid M, and Murthy RC. 1998. Embryo and fetotoxicity of hexavalent chromium: 

a long-term study. Toxicol. Lett. 95:165-172. 

Lindberg E, and Hedenstierna G. 1983. Chrome plating: symptoms, findings in the upper 

airways, and effects on lung function. Arch. Environ. Health 38(6):367-374. 

Linberg E, and Vesterberg O. 1983. Urinary excretion of proteins in chromeplaters, 

exchromeplaters and referents. Scand. J. Work. Environ. Health 9:505-510. 

Lucas JB, and Kramkowski RS. 1975. Health hazard evaluation/ toxicity determination report 

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Occupational Safety and Health. NTIS #PB 249 401. 

Machle W, and Gregorius F. 1948. Cancer of the respiratory system in the United States 

chromate-producing industry. Public Health Rep 63(35):1114-1127. 

MacKenzie RD, Byerrum RU, Decker CF, Hoppert CA, and Langham RF. 1958. Chronic 

toxicity studies. II. Hexavalent and trivalent chromium administered in drinking water to rats. 

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Murthy RC, Junaid M, and Saxena DK. 1996. Ovarian dysfunction in mice following chromium 

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Appendix D3 155 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Wang X, Qin Q, Xu X, Xu J, Wang J, Zhou J, Huang S, Zhai W, Zhou H, and Chen J. 1994. 

Chromium-induced early changes in renal function among ferrochromium-producing workers. 

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Appendix D3 156 Chloropicrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

CHRONIC TOXICITY SUMMARY 

CRESOL MIXTURES 

Compounds Synonyms CAS Reg. No. 

cresols cresylic acid; tricresol; hydroxytoluene; methylphenol 1319-77-3 

o-cresol 1-hydroxy-2-methylbenzene; 2-hydroxytoluene; 

2-methylphenol 

95-48-7 

m-cresol 1-hydroxy-3-methylbenzene; 3-hydroxytoluene; 

3-methylphenol 

108-39-4 

p-cresol 1-hydroxy-4-methylbenzene; 4-hydroxytoluene; 

4-methylphenol 

106-44-5 

I. Chronic Toxicity Summary

Inhalation reference exposure level 600 μg/m3 (100 ppb)

Critical effect(s) Neurotoxicity 

Hazard index target(s) Nervous system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994, unless otherwise noted)

Description Colorless in pure form; yellowish, brownishyellow, or pinkish liquid 

Molecular formula C7H8O 

Molecular weight 108.14 g/mol 

Boiling point 191.0°C (o-cresol) 

202°C (m-cresol) 

201.9°C (p-cresol) 

Melting point 29.8°C (o-cresol) 

11.8°C (m-cresol) 

35.5°C (p-cresol) 

Solubility Soluble in 50 parts water; miscible with alcohol, 

benzene, ether, glycerol, petroleum ether; 

soluble in vegetable oils, glycol 

Conversion factor 4.42 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources

Cresol compounds (mixtures of the ortho-, meta- and para-isomers) can be obtained from coal tar 

and petroleum or synthesized by sulfonation or oxidation of toluene (HSDB, 1995). Crude 

cresol (commercial grade) contains approximately 20% o-cresol, 40% m-cresol, and 30% pAppendix D3 157 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

cresol. Phenol and xylenols are present in small amounts as contaminants. Cresylic acid 

compounds are called cresol when the boiling point is below 204°C. 

Cresols have a wide variety of uses including the manufacture of synthetic resins, tricresyl 

phosphate, salicylaldehyde, coumarin, and herbicides. Cresols also serve as components of 

degreasing compounds in textile scouring and paintbrush cleaners as well as fumigants in 

photographic developers and explosives. Cresols also function as antiseptics, disinfectants, and 

parasiticides in veterinary medicine. An approximate breakdown of cresol and cresylic acid use 

is 20% phenolic resins, 20% wire enamel solvents, 10% agricultural chemicals, 5% phosphate 

esters, 5% disinfectants and cleaning compounds, 5% ore flotation, and 25% miscellaneous and 

exports. 

Any combustion process, which results in the generation of phenolic compounds (such as 

automobile exhaust or coal, wood, or trash smoke), may be a potential source of exposure to 

cresols. Cresols are also formed from the atmospheric photooxidation of toluene. However, 

under normal conditions low vapor pressure limits the inhalation hazard presented by cresols 

(HSDB, 1995). The annual statewide industrial emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 8407 

pounds of mixtures of cresols (cresylic acid), 3 pounds of m-cresol, and 3 pounds of o-cresol 

(CARB, 2000). 

IV. Effects of Exposures to Humans 

Brief exposure to 6 mg cresol/m3 resulted in irritation of the throat and nose, nasal constriction, 

and dryness in 8 of 10 subjects (Uzhdavini et al., 1972). 

Chemical burns may result from exposure to cresols (Pegg and Campbell, 1985). The lungs of 

humans exposed to cresols have shown signs of emphysema, edema, bronchopneumonia, and 

small hemorrhages (Clayton and Clayton, 1982). Skin contact has resulted in the development of 

white patches and blistering, eventually turning brown or black (Lefaux, 1968). Other reported 

effects include turbidity, inflammation, and fatty degeneration of the liver, nephritis, and 

hemorrhage of the epicardium and endocardium. An infant fatally exposed to ~20 ml of a 90% 

cresol solution dermally showed widespread edema of the internal organs, especially the brain 

and kidney (Green, 1975). The liver showed signs of centrilobular and midzonal necrosis. 

Chronic systemic poisoning by any route of exposure may produce symptoms of vomiting, 

dysphagia, salivation, diarrhea, loss of appetite, headache, fainting, dizziness, and mental 

disturbances (Sittig, 1981). Skin rash and discoloration may also result from prolonged or 

repeated exposure of the skin. Death may result from severe damage to the liver and kidneys. 

Oral poisoning has resulted in kidney problems (likely from the direct action of cresol) and 

pancreatitis (from constriction of the pancreatic ducts) (Klimkiewicz et al., 1974, as reported in 

HSDB, 1995). 

Appendix D3 158 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

V. Effects of Exposures to Animals

The effects of inhaled o-cresol were examined in several species (Uzhdavini et al., 1972, as 

reported in ATSDR, 1992 and U.S. EPA, 1982). Cats exposed for 30 minutes to 5-9 mg ocresol/m3

 showed signs of respiratory irritation as indicated by increased parotid gland 

secretions. Exposure of mice for 2 hrs/day for 1 month to 50 mg o-cresol/m3 did not have an 

effect on mortality, however, heart muscle degeneration and degeneration of nerve cells and glial 

elements were observed. 

Uzhdavini et al. (1972) exposed rats (both sexes, numbers not stated) by inhalation to 9.0 ± 0.9 

mg o-cresol/m3

, first for 2 months (6 hours/day, 5 days/week), then for 2 more months (4 

hours/day, 5 days/week). Endpoints examined in rats included elementary conditioned defensive 

reflex, white blood cell levels, bone marrow elements, and liver function (as indicated by 

increased susceptibility to hexobarbital narcosis). Both cresol-exposed and control animals 

showed some loss of the defensive reflex; the effect occurred in all exposed animals before the 

end of the second month and in control animals at later times. White blood cell counts were 

elevated in male animals, peaked at the end of the exposure period, and returned to normal one 

month after cessation of exposure. Exposed animals also showed a statistically significant 

change in the leukoid-to-erythroid ratio in the bone marrow. Liver toxicity was suggested by an 

extension in the duration of hexobarbital narcosis in treated animals. Although guinea pigs were 

similarly evaluated for changes in blood cell counts and ECG, scant reporting of experimental 

detail limits the usefulness of this portion of the study. 

NR rats were exposed by inhalation to 0.0052 or 0.05 mg tricresol/m3

 for 3 months (Kurliandskii 

et al., 1975; as described by U.S. EPA, 1982). The proportional composition of the compound 

was not specified. Effects observed in the high-dose group included decreased weight gain, 

increased central nervous system excitability, increased oxygen consumption, and histological 

changes in the lung and liver. Serum gamma-globulin levels were also reduced. No effects were 

observed in the low-dose group. Rats (6/group, sex unspecified) were also exposed for 24 hours 

to 0.01, 0.1, and 2.4 mg tricresol/m3

 with a control group of 6 rats for each exposure group. The 

absorption of neutral red dye by lung tissue was used as an indicator of protein denaturation in 

the tissue. Significantly increased dye absorption over control animals was observed at both 2.4 

and 0.1 mg tricresol/m3

. The degree of dye absorption in the low-dose group was not 

significantly increased over controls. 

In a 90-day subchronic toxicity study (U.S. EPA, 1986), 30 Sprague-Dawley rats/sex/dose were 

gavaged daily with 0, 50, 175, or 600 mg/kg/day p-cresol. Body and organ weights, food 

consumption, mortality, clinical signs of toxicity, and clinical pathology were evaluated. At 600 

mg/kg/day, o-cresol showed 47% combined mortality (9/30 males, 19/30 females), and a 30% 

reduction in body weight at week 1 and 10% at necropsy. Kidney-to-body weight ratio was 13% 

higher than that of the control value at the end of the study. CNS effects such as lethargy, ataxia, 

coma, dyspnea, tremor, and convulsions were seen within 15 to 30 minutes after dosing; but 

recovery occurred within 1 hour post-gavage. At 450 mg/kg/day, combined mortality was 10% 

(1/10 male, 1/10 female). In the 175 mg/kg/day group, two animals exhibited tremors on day 1 

of the study during the hour following gavage administration, and one of the two became 

comatose. At 50 mg/kg/day, no significant adverse effects were observed (USEPA, 1999a,b). 

Appendix D3 159 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

In a 90-day neurotoxicity study (U.S. EPA, 1987), 10 Sprague-Dawley rats/sex/dose were 

gavaged daily with o-cresol at 0, 50, 175, 450, or 600 mg/kg/day. In addition to the parameters 

evaluated above, various signs of neurotoxicity were monitored. The lowest dose of o-cresol 

caused clinical signs of CNS-stimulation post-dosing, such as salivation, rapid respiration, and 

hypoactivity; however, these symptoms were low in incidence and sporadic in nature. Higher 

doses of o-cresol (greater than 450 mg/kg/day) produced significant neurological events, such as 

increased salivation, urination, tremors, lacrimation, palpebral closure, and rapid respiration. 

Animals given high doses also showed abnormal patterns in the neurobehavioral tests. The 

NOAEL based on systemic toxicity was 50 mg/kg/day (USEPA, 1999a,b). 

Dermal exposure of rats to 1.0-1.7 ml cresol/kg body weight for 1-2 hours resulted in skin 

discoloration and death of the animals (Campbell, 1941). 

Exposure to high concentrations of toluene vapors, or to intravenous o-cresol, a toluene 

metabolite, at about 0.9 mg/min, caused excitation of the somatosensory evoked potential (SEP) 

and electroencephalograph (EEG) of Fischer 344 rats (Mattsson et al., 1989). Both substances 

induced an increase in EEG beta activity and caused a large increase in activity at 5 Hz. Toluene 

exposed rats were lightly anesthetized, while o-cresol rats were conscious but hyperreactive. 

When exposure was continued, both sets of rats had involuntary muscle movements and tremors. 

Neither benzoic acid and hippuric acid, also metabolites of toluene, caused neuroexcitation. The 

authors concluded that metabolically derived cresols are plausible candidates for the 

neuroexcitatory properties of toluene. 

In rat liver slices at equimolar concentrations, p-cresol was 5- to 10-times as toxic as the o- or misomers for cell killing (Thompson et al., 1994). p-Cresol rapidly depleted intracellular 

glutathione levels, while the o- and m-isomers depleted it to a lesser extent. p-Cresol was 

metabolized to a reactive intermediate which bound covalently to protein. The reaction was 

inhibited by N-acetylcysteine. 

The National Toxicology Program (NTP) sponsored reproductive toxicity tests of cresol isomers 

in Swiss CD-1 mice using the risk assessment by continuous breeding (RACB) protocol (Heindel 

et al., 1997a, 1997b). For o-cresol the exposure concentrations in the continuous cohabitation 

task were 0.05%, 0.2%, and 0.5% in feed (approximately 60, 220, and 550 mg/kg/day (Heindel 

et al., 1997a). At these doses o-cresol was not a reproductive toxicant. When a m-/p-cresol 

mixture was used at concentrations of 0.25, 1.0 and 1.5% in feed (approximately 370, 1500, and 

2100 mg/kg/day), the m/p mixture was a reproductive toxicant, since (1) fewer F1 pups per litter 

were produced, (2) both generations showed reduced pup weights, and (3) reproductive organs 

showed weight reductions. Unfortunately the responses were not dose-dependent and the 

mixture was judged not to be a selective reproductive toxicant. Oral gavage administration of o-, 

m-, or p-cresol, separately, in rats did not produce selective reproductive toxicity; i.e., for each of 

the cresol isomers, in the absence of parental toxicity, there was no reproductive toxicity. The 

NOEL for reproductive toxicity for each isomer was 175 mg/kg/day (Tyl 1989a, 1989b, 1989c). 

Appendix D3 160 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

VI. Derivation of Inhalation Chronic Reference Exposure Level 

Study U.S. EPA, 1987 

Study population Sprague-Dawley rats 

Exposure method Gavage at 0, 50, 175, 450, or 600 mg/kg-day 

Critical effects Decreased body weights and neurotoxicity (tremors, 

salivation, lacrimation, etc.) 

LOAEL 175 mg/kg-day 

NOAEL 50 mg/kg-day 

Exposure continuity Daily gavage 

Exposure duration 90 days 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 (90 day study) 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

U.S. EPA Reference Dose (RfD) 0.17 mg/kg/day 

Route-to-route extrapolation factor 3500 μg/m3

 per mg/kg/day 

Inhalation chronic REL 600 μg/m3

 (100 ppb) 

An RfD of 0.05 mg/kg/day was derived by the USEPA for both o-cresol and m-cresol (USEPA 

1998a, 1998b; listed as 2-methylphenol and 3-methylphenol). The RfD for p-cresol was 

withdrawn by the USEPA. U.S EPA used a subchronic uncertainty factor of 10 for a 90 day 

study in rats. In accordance with its approved methodology (OEHHA, 2000), OEHHA used a 

factor of 3. 

The available literature on the observed toxicity of cresol compounds and cresol mixtures to 

humans by inhalation indicates that at high concentrations these compounds are initially toxic 

due to their ability to cause chemical burns and are therefore of concern at the site of contact. In 

humans occupationally exposed, inhalation exposure is reported to cause respiratory effects 

including the development of pneumonia, pulmonary edema, and hemorrhage (Clayton and 

Clayton, 1982). Other case reports of cresol toxicity to humans are confounded by the presence 

of other compounds, such as phenol, formaldehyde, and ammonia (Corcos, 1939; NIOSH, 1974). 

The only quantitative information from inhalation exposures to humans, however, comes from 

acute exposure studies showing irritation at 6 mg cresol/m3

 (Uzhdavini et al., 1972, as reported 

in ATSDR, 1992). Toxic effects reported in animals include bone marrow and liver toxicity in 

rats from 4 month exposure to 9 mg cresol/m3

 (Uzhdavini et al., 1972, as reported in U.S. EPA, 

1982). Other animal studies have shown more systemic effects from inhalation exposure to 

cresols. Uzhdavini et al., 1972 reported cardiac and nerve cell degeneration in mice exposed for 

2 hour/day for 1 month to 50 mg o-cresol/m3

. Kurliandskii et al. (1975) (as reported in HSDB, 

1995) observed decreased weight gain with histological changes in the liver and lungs of rats 

exposed for 3 months to 0.05 mg tricresol/m3

. Although this study reports adverse effects at 

levels below those observed in the Uzhdavini et al. (1972) study, limited experimental detail 

precludes the use of these data in the development of the chronic REL. 

Appendix D3 161 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

The only useful inhalation data for the development of a chronic REL are those showing 

hematological toxicity to the bone marrow of rats exposed for 4 months to o-cresol (Uzhdavini et

al. (1972) as reported in U.S. EPA, 1982). These authors report a LOAEL of 9 mg tricresol/m3

. 

OEHHA staff decided not to use this study. (1) A complete translation from the original Russian 

was not available so that only the interpretations of others were available. (2) Some endpoints 

tested are not commonly used in toxicology. And (3) some of the results reported were unusual 

(e.g., elevation of white blood cells in male but not female rats). 

As noted above, the inhalation study conducted by Kurliandskii et al. (1975) suggests that 

adverse health effects occur in experimental animals at exposure levels considerably below those 

reported by Uzhdavini et al. (1972) (9 mg/m3 vs. 0.05 mg/m3). The report from which the lower 

level is drawn has limitations. Human subjects exposed briefly to levels below the LOAEL have 

reported respiratory irritation. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the REL for cresols include the use of measured exposure data of animals 

exposed over a significant fraction of their lifetime. Major areas of uncertainty are route-to-route 

extrapolation, the lack of chronic human data, and the paucity of reproductive and developmental 

toxicity studies. Additional inhalation studies of cresols will be useful. 

VIII. References 

ATSDR. 1992. Agency for Toxic Substances and Disease Registry. Toxicological profile for 

cresols: o-cresol, p-cresol, and m-cresol. Atlanta, GA: ATSDR. NTIS No. PB93-110732. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Campbell I. 1941. Petroleum cresylic acids. A study of their toxicity and the toxicity of cresylic 

disinfectants. Soap Sanitation Chem. 17:103. 

Chao J, Lin CT, and Chung TH. 1983. Vapor pressure of coal chemicals. J. Chem. Refer. Data 

12:1033-1063. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Clayton GD, and Clayton FE (eds). 1982. Patty’s Industrial Hygiene and Toxicology. Third, 

revised ed. New York: John Wiley Sons, p. 2133. 

Corcos A. 1939. [Contribution to the study of occupational poisoning by cresols]. Dissertation. 

Paris (Vigot Freres Editeurs) (in French). 

Appendix D3 162 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

Daubert TE, and Danner RD. 1985. Data compilation tables of properties of pure compounds. In: 

Design Institute for Physical Property Data. New York: American Institute for Chemical 

Engineers. 

Green MA. 1975. A household remedy misused - fatal cresol poisoning following cutaneous 

absorption (a case report). Med. Sci. Law 15:65-66. 

Heindel J, George J, Fail P, Izard MK, and Grizzle T. 1997a. o-Cresol. Environ. Health Perspect. 

105(Suppl 1):293-294. 

Heindel J, Izard MK, George J, Fail P, and Grizzle T. 1997b. m-/p-Cresol. Environ. Health 

Perspect. 105(Suppl 1):295-296. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Klimkiewicz E, Klinger M, Kurbiel A, and Soltys W. 1974. [Case of liquid lysol poisoning 

complicated by acute pancreatitis and acute renal failure]. Wiad. Lek. 27:1211-1214. 

Kurliandskii BA, Partsef DI, and Chernomorskii AR. 1975. [A method of determining the daily 

average maximum permissible concentration of tricresol in the atmosphere]. Gig. Sanit. 5:85-87. 

Lefaux R. (ed.) 1968. Practical Toxicology of Plastics. Cleveland: CRC Press, Inc. 

Mattsson JL, Albee RR, and Gorzinski SJ. 1989. Similarities of toluene and o-cresol 

neurotoxication in rats. Neurotoxicol. Teratol. 11:71-75. 

NIOSH. 1974. National Institute for Occupational Safety and Health. Health hazard evaluation / 

Toxicity determination report no. 73-5-110. Prepared by Cummins Northeastern Incorporated. 

Dedham, MA. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov

Pegg SP, and Campbell DC. 1985. Children’s burns due to cresol. Burns Incl. Therm. Inj. 

11(4):294-296. 

Sittig M. (ed) 1981. Handbook of Toxic and Hazardous Chemicals. Park Ridge, NJ: Noyes Data 

Corporation. 

Thompson DC, Perera K, Fisher R, and Brendel K. 1994. Cresol isomers: comparison of toxic 

potency in rat liver slices. Toxicol. Appl. Pharmacol. 125(1):51-58 

Appendix D3 163 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000 

Tyl RW. 1989a. Two Generation Reproduction Study of o-Cresol (CAS No. 95-48-7) 

Administered by Gavage to Sprague-Dawley Rats. Bushy Run Research Center, Export, PA. 

Project Report 51-614. 

Tyl RW. 1989b. Two Generation Reproduction Study of m-Cresol (CAS No. 108-39-4) 

Administered by Gavage to Sprague-Dawley Rats. Bushy Run Research Center, Export, PA. 

Project Report 51-634. 

Tyl RW. 1989c. Two Generation Reproduction Study of p-Cresol (CAS No. 106-44-5) 

Administered by Gavage to Sprague-Dawley Rats. Bushy Run Research Center, Export, PA. 

Project Report 52-512. 

U.S. EPA 1982. United States Environmental Protection Agency. Health effects assessment for 

cresols. Office of Research and Development, Office of Health and Environmental Assessment, 

Environmental Criteria and Assessment Office. EPA/540/1-86/050. Cincinnati, OH: US EPA. 

NTIS. PB86-134616. 

U.S. EPA. 1986. o, m, p-Cresol. 90-Day oral subchronic toxicity studies in rats. Office of Solid 

Waste. Washington, DC: U.S. EPA. 

U.S. EPA. 1987. o, m, p-Cresol. 90-Day oral subchronic neurotoxicity study in rats. Office of 

Solid Waste. Washington, DC: U.S. EPA. 

U.S. EPA. 1999a. 2-Methyphenol. Integrated Risk Information System (IRIS). 

U.S. EPA. 1999b. 3-Methylphenol. Integrated Risk Information System (IRIS). 

Uzhdavini ER, Astaf’yeva K, Mamayeva AA, and Bakhtizina GZ. 1972. [Inhalation toxicity of 

o-cresol]. Trudy Ufimskogo Nauchno-Isseldovatel’skogo Instituto Gigiyeny Profzabolevaniya 

7:115-119. (in Russian) 

Appendix D3 164 Cresol Mixtures 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

1,4-DICHLOROBENZENE 

(p-dichlorobenzene; di-chloricide; p-dichlorobenzol; Paradow; Paramoth; Parazene; pchlorophenyl chloride) 

CAS Registry Number: 106-46-7 

I. Chronic Toxicity Summary

Inhalation reference exposure level 800 μg/m3 (100 ppb) 

Critical effect(s) General effects (reduced body weights and food 

consumption) in rats 

CNS effects (tremors) in rats 

Respiratory/dermal effects (nasal and ocular 

discharge) in rats 

Liver effects (increased liver weight) in rats, and 

Kidney effects (increased kidney weight) in rats. 

Hazard index target(s) Nervous system; respiratory system; alimentary 

system; kidney 

II. Chemical Property Summary (HSDB, 1997; CRC, 1994)

Description White crystals, monoclinic prisms 

Molecular formula C6H4Cl2

Molecular weight 147.01 g/mol 

Boiling point 174o

C 

Melting point 52.7o

C 

Vapor pressure 10 torr @ 54.8o

C 

Solubility Soluble in chloroform, carbon disulfide, alcohol, 

ether, acetone, benzene 

Conversion factor 1 ppm = 6.0 mg/m3

 at 25oC 

III. Major Uses and Sources 

Commercial grade 1,4-dichlorobenzene (1,4-DCB) is available in the USA as a technical grade 

liquid, typically containing a small percentage (>0.1% by weight) of meta (1,3-DCB) and ortho 

(1,2-DCB) isomers; as a solution in solvent or oil suspension; or as crystalline material pressed 

into various forms (HSDB, 1997). Besides its role as an intermediate in the synthesis of various 

organics, dyes and pharmaceuticals, 1,4-dichlorobenzene is used as a space or garbage 

deodorizer for odor control. The insecticidal and germicidal properties of 1,4-dichlorobenzene 

are used to control fruit borers and ants, moths, blue mold in tobacco seed beds, and mildew and 

mold on leather or fabrics. In 1996, the latest year tabulated, the statewide mean outdoor 

Appendix D3 165 Dichlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

monitored concentration of 1,4-DCB was approximately 0.15 ppb (CARB, 1999). The annual 

statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California based on the most recent inventory were estimated to be 30,577 pounds of 

dichlorobenzene (CARB, 2000). 

IV. Effects of Human Exposure 

Case reports of human exposure to 1,4-DCB include malaise, nausea, hepatic manifestations 

(yellow atrophy and cirrhosis), proteinuria, bilirubinuria, hematuria, and anemia. A woman 

exposed to 1,4-DCB for 6 years developed central nervous system effects, including severe 

cerebellar ataxia, dysarthria, weakness in all limbs, and hyporeflexia (U.S. EPA, 1985). 

No epidemiologic studies of 1,4-DCB exposures were located. 

V. Effects of Animal Exposure 

Rats, rabbits and guinea pigs were exposed to 0, 96, 158, 341 or 798 ppm (0, 577, 950, 2050 or 

4800 mg/m3) 1,4-DCB by inhalation 7 hours/day, 5 days/week for 6-7 months (Hollingsworth et 

al., 1956). High dose animals showed marked tremors, weakness, loss of weight, eye irritation 

and unconsciousness. Liver and kidney changes included cloudy swelling and centrilobular 

cellular degeneration (liver). In another inhalation study in rats animals were exposed to 0, 75 or 

500 ppm (0, 451 or 3006 mg/m3) for 5 hours/day, 5 days/week for 76 weeks (Riley et al., 1980). 

The authors found increased kidney and liver weights in the high dose group. Thus 75 ppm was 

a NOAEL. Studies with oral exposure to 1,4-DCB, including the NTP (1987) chronic bioassay 

study (maximum dose of 300 mg/kg-day), have also found an increased incidence of renal and 

hepatic lesions (cellular degeneration and focal necrosis). 

Three inhalation reproductive studies, one in rabbits (Hayes et al., 1985), one in mice (Anderson 

and Hodge, 1976), and one in rats (Chlorobenzene Producers Assn., 1986), found minimal 

reproductive effects. In rabbits exposed on days 6-18 of gestation to 100, 300, and 800 ppm 1,4-

DCB, only the differences in percentage of implantations resorbed and in percentage of litters 

with resorptions were significantly increased and only in the 300 ppm group (Hayes et al., 1985). 

No reduction in reproductive performance was observed in mice exposed to 0, 75, 225, or 450 

ppm 1,4-DCB for 6 hours/day for 5 days (Anderson and Hodge, 1976). 

In a two-generation reproductive study (Chlorobenzene Producers Association, 1986), SpragueDawley rats P1 (28/sex/group) were exposed to 0, 50, 150 or 450 ppm (0, 301, 902, or 2705 

mg/m3

) of 1,4-DCB vapor, 6 hours/day, 7 days/week for 10 weeks, and then mated for 3 weeks. 

The second generation F1 weanlings were exposed to 1,4-DCB for 11 weeks and then mated. 

No developmental abnormalities were observed in pups examined. At 450 ppm significant 

decreases in live births, pup weights, and pup survival were seen in both the F1 and F2 

generations. Non-reproductive effects observed in the parental males in the 150 and 450 ppm 

groups included significantly increased liver and kidney weights. All dose levels caused hyaline 

droplet nephrosis in post-pubescent males; but this change was associated with the formation of 

Appendix D3 166 Dichlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

alpha-2u-globulin, an abnormality considered specific for male rats with no relative human 

significance (U.S. EPA, 1991). The Chlorobenzene Producers Association reproductive study 

was chosen by the U.S. EPA to derive the RfC. 

VI. Derivation of Chronic Reference Exposure Level 

Study Chlorobenzene Producers Association, 1986 

Study population Sprague-Dawley rats (28 rats/sex/group) 

Exposure method Discontinuous whole-body inhalation exposures 

(0, 50, 150 or 450 ppm) 

Critical effects Reduced body weights and food consumption; 

tremors; nasal and ocular discharge; increased 

liver and kidney weights 

LOAEL 150 ppm 

NOAEL 50 ppm 

Exposure continuity 6 hr/day for 7 days/week 

Average experimental exposure 13 ppm for NOAEL group (50 x 6/24) 

Human equivalent concentration 13 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

Exposure duration 10 weeks 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.1 ppm (100 ppb, 0.8 mg/m3

, 800 μg/m3) 

The chronic REL for 1,4-dichlorochlorobenzene is also the U.S. EPA RfC. OEHHA agrees with 

the U.S.EPA analysis. A 3-fold subchronic uncertainty factor (instead of 10) was used by U.S. 

EPA because of data suggesting limited progression of hepatic lesions (Riley et al., 1980). Ten 

weeks are also greater than 8% of a rat's two-year lifetime and thus in accord with OEHHA’s use 

of a subchronic UF of 3 (OEHHA, 2000). 

For comparison, Riley et al. (1980) found a chronic NOAEL of 75 ppm for kidney and liver 

effects in rats, which is equivalent to 11.2 ppm continuous exposure. Use of an RGDR of 1 and 

a total UF of 30 (3 for interspecies and 10 for intraspecies) results in a REL estimate of 0.4 ppm. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the REL for 1,4-dichlorochlorobenzene are the observation of a NOAEL 

and the demonstration of a dose-response relationship. The major uncertainties are the lack of 

human data and the lack of chronic, multiple-species health effects data. 

Appendix D3 167 Dichlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VIII. References

Anderson D, and Hodge MCE. 1976. Paradichlorobenzene: Dominant lethal study in the mouse. 

ICI Report No. CTL/P/296. November. 

CARB. 1999. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Para-Dichlorobenzene. Available online at http://www.arb.ca.gov/aqd/toxics.htm 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Chlorobenzene Producers Association. 1986. Parachlorobenzene: Two-generation reproduction 

study in Sprague-Dawley rats. Study 86-81-90605. MRID No. 411088-1. 

Hayes WC, Hanley TR, Gushow TS, Johnson KA, and John JA. 1985. Teratogenic potential of 

inhaled dichlorobenzenes in rats and rabbits. Fundam. Appl. Toxicol. 5(1): 190-202. 

HSDB. 1997. Hazardous Substances Data Bank. TOMES® Vol. 33. Denver, CO: Micromedex, 

Inc. (edition expires 7/31/97) 

Hollingsworth RL, Rowe VK, Oyen F, Hoyle HR, and Spencer HC. 1956. Toxicity of 

paradichlorobenzene: Determinations of experimental animals and human subjects. AMA Arch. 

Ind. Health. 14: 138-147. 

NTP. 1987. National Toxicology Program. Toxicology and carcinogenesis studies of 1,4-

dichlorobenzene in F344/N rats and B6C3F1 mice (gavage studies). NTP TR 319. NIH Publ. No. 

87-2575. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov 

Riley RA, Chart IS, Doss A, Gore CW, Patton D, and Weight TM. 1980. Para-dichlorobenzene: 

Long-term inhalation study in the rat. ICI Report No. CTL/P/447. August, 1980. 

U.S. EPA. 1985. U.S. Environmental Protection Agency. Health Assessment Document for 

Chlorinated Benzenes. EPA/600/8-84/015F. Prepared by the Office of Health and Environmental 

Assessment, Environmental Criteria and Assessment Office. Cincinnati, OH: U.S. EPA. 

U.S. EPA. 1991. U.S. Environmental Protection Agency. Alpha-2u-Globulin: Association with 

Chemically Induced Renal Toxicity and Neoplasia in the Male Rat. EPA/625/3-91/019F. 

Appendix D3 168 Dichlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Prepared for the Risk Assessment Forum, U.S. EPA, Washington, DC 20460. [as cited in U.S. 

EPA, 1994.] 

U.S.EPA. 1994. U.S. Environmental Protection Agency. 1994. Integrated Risk Information 

System (IRIS) Database. Reference concentration (RfC) for 1,4-Dichlorobenzene. Available 

online at http://www.epa.gov/ngispgm3/iris

Appendix D3 169 Dichlorobenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY

1,1-DICHLOROETHYLENE 

(DCE; 1,1-dichloroethene; VDC; vinylidene chloride) 

CAS Registry Number: 73-35-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 70 μg/m3 (20 ppb) 

Critical effect(s) Increased mortality; hepatic effects (mottled 

livers and increases in liver enzymes) in 

guinea pigs 

Hazard index target(s) Alimentary system 

II. Physical and Chemical Properties (HSDB, 1994; CRC, 1994)

Description Colorless liquid 

Molecular formula C2H2Cl2

Molecular weight 96.95 g/mol 

Boiling point 31.7o

C 

Melting point −122.5o

C 

Vapor pressure 500 torr @ 20o

C 

Solubility Soluble in water (2.5 g/L); miscible in organic 

solvents 

Conversion factor 3.97 μg/m3

 per ppb at 25 oC 

III. Major Uses and Sources

1,1-Dichloroethylene (1,1-DCE) is used in the production of polyvinylidene chloride copolymers 

(HSDB, 1994). 1,1-DCE containing copolymers include other compounds such as acrylonitrile, 

vinyl chloride, methacrylonitrile, and methacrylate. These copolymers are used in flexible 

packaging materials; as flame retardant coatings for fiber, carpet backing, and piping; as coating 

for steel pipes; and in adhesive applications. Flexible films for food packaging, such as SARAN 

and VELON wraps, use such polyvinylidene chloride copolymers. The annual statewide 

industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in California 

based on the most recent inventory were estimated to be 2458 pounds of vinylidene chloride 

(CARB, 2000). 

IV. Effects of Human Exposure 

Appendix D3 170 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Limited information exists regarding the human health effects following exposure to 1,1-DCE. 

A few case reports and mortality studies have reported hepatotoxicity and nephrotoxicity after 

repeated, low-level exposures (USEPA, 1976; Ott et al., 1976). However, these investigations 

were conducted in industrial settings with the possibility of mixed chemical exposures. In 

preliminary clinical findings reported by the EPA (1976), workers exposed to 1,1-DCE for 6 

years or less had a high incidence of hepatotoxicity, with liver scans and measurements of liver 

enzymes revealing 50% or greater loss in liver function in 27 of 46 exposed workers. 

Unfortunately, no follow-up study was reported. 

V. Effects of Animal Exposure 

Several studies have reported on the subchronic or chronic toxicity of 1,1-DCE in laboratory 

animals exposed either via oral or inhalation routes. The liver is the primary target organ of 1,1-

DCE toxicity following acute or chronic inhalation exposure. Such exposure is marked by both 

biochemical changes (alterations in serum enzyme levels) and histological changes (e.g., 

midzonal and centrilobular swelling, degeneration, and necrosis) (Gage, 1970; Lee et al., 1977; 

Plummer et al., 1990; Quast, 1976; Quast et al., 1986). Unfortunately, these longer-term studies 

used only one or two doses or a limited number of animals. 

Male and female rats exposed intermittently (6 hours/day, 5 days/week) to 125 or 200 ppm 1,1-

DCE over 30 days exhibited centrilobular fatty degeneration or hepatocellular necrosis (Quast 

1976, as cited by USDHHS, 1994). Two other studies identified hepatic changes in rats at lower 

concentrations of 1,1-DCE (6 hours/day, 5 days/week): cytoplasmic vacuolation after 30- or 90-

day exposure to 25 or 75 ppm 1,1-DCE (Balmer et al., 1976, as cited by USDHHS, 1994), and 

fatty changes after 6 months at 25 ppm 1,1-DCE (Quast et al., 1986). 

Laboratory animals appear less tolerant of continuous exposure (23-24 hours per day) than 

intermittent exposure. Beagle dogs exposed to 100 ppm 1,1-DCE for 8 hours/day, 5 days/week 

for 42 days had no evidence of hepatotoxicity, but continuous exposure to 48 ppm for 90 days 

caused liver changes (Prendergast et al., 1967). Similarly, monkeys continuously exposed to 48 

ppm for 90 days exhibited focal necrosis and hemosiderin deposition, while no liver toxicity was 

apparent following 42 days of intermittent exposure to 100 ppm 1,1-DCE (Prendergast et al., 

1967). Guinea pigs exposed to 1,1-DCE for 24 hours per day for 90 days (0, 5, 15, 25, or 48 

ppm) displayed mottled livers at 15 ppm, and increased liver enzyme levels (serum glutamicpyruvic transaminase (SGPT) and alkaline phosphatase (AP)) at 48 ppm. A NOAEL of 5 ppm 

based on liver changes (Prendergast et al., 1967) is indicated by the results. 

Data on continuously exposed guinea pigs from Prendergast et al. (1967) 

ppm 1,1-DCE (mg/m3) Survival 

Body weight 

change Liver AP SGPT 

 0 312/314 +69.0% 0.08±0.03 10±5 

 5 (20) 43/45 +58.6% 0.08±0.03 11±3 

 15 (61) 12/15 +55.3% Not reported Not reported 

 25 (101) 12/15 +74.0% Not reported Not reported 

 48 (191) 8/15 +50.3% 0.19±0.04 >70 

Appendix D3 171 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Additional adverse effects observed to a lesser extent in laboratory animals include respiratory 

and renal toxicity. Nephrotoxicity observed following chronic 1,1-DCE exposure included gross 

organ (increases in kidney weight) (Klimisch et al., 1979; Quast et al., 1986) and histological 

changes (tubular swelling, degeneration, and necrosis) (Klimisch et al., 1979; Lee et al., 1977; 

Prendergast et al., 1967). Continuous exposure of rats to 48 ppm 1,1-DCE for 90 days caused 

nuclear hypertrophy of the renal tubular epithelium (Prendergast et al., 1976). Mice exposed to 

25 ppm 1,1-DCE 4 hours/day, 4 or 5 days/week, for 52 weeks displayed severe tubular 

nephrotoxicity (Maltoni et al., 1985 as cited by USDHHS, 1994). Nasal irritation was observed 

in rats exposed to 200 ppm for 4 weeks (Gage 1970). But no respiratory effects were attributed 

to 1,1-DCE exposure in rats, monkeys, dogs, rabbits, or guinea pigs exposed to 100 ppm 

intermittently for 6 weeks (Prendergast et al., 1967) or in rats exposed to 75 ppm for 18 months 

(Quast et al., 1986). 

Toxicokinetic studies in laboratory animals have demonstrated that 1,1-DCE is readily absorbed 

and rapidly distributed following inhalation exposure (Dallas et al., 1983; McKenna et al., 

1978b). Following inhalation exposure to radioactively labeled 1,1-DCE, rats preferentially 

accumulate radioactivity in the kidney and liver (McKenna et al., 1978b; Jaeger et al., 1977). 

Glutathione (GSH) conjugation appears to be the major detoxification route for 1,1-DCE 

intermediates, and GSH-depleting experimental states, such as drugs and fasting, may tend to 

increase 1,1-DCE toxicity (Jaeger et al., 1977; McKenna et al., 1978; Reichert et al., 1978). One 

study greatly increased 1,1-DCE induced lethality and hepatotoxicity in rats by pretreatment with 

acetaminophen (Wright and Moore, 1991). 

Appendix D3 172 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Prendergast et al. (1967) 

Study population Guinea pigs (15 per group, except 45 animals in 20 

mg/m3 group) 

Exposure method Continuous whole body inhalation (0, 20, 61, 101, 

or 189 mg/m3

) 

Critical effects Increased mortality at 61, 101, and 189 mg/m3

; 

hepatic effects (mottled livers and increases in 

SGPT and AP enzymes) noted at 189 mg/m3

LOAEL 61 mg/m3

 (15 ppm) 

NOAEL 20 mg/m3

 (5 ppm) 

Exposure continuity Continuous 

Exposure duration 90 days 

Average experimental exposure 20 mg/m3

 for NOAEL group 

Human equivalent concentration 20 mg/m3

 for NOAEL group (gas with systemic 

effects, based on default assumption that RGDR 

= 1 using default assumption that lambda (a) = 

lambda (h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 10 (since guinea pig life-span is approx. 6 years) 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.07 mg/m3

 (70 μg/m3; 0.02 ppm; 20 ppb) 

The principal study (Prendergast et al., 1967) identified adverse hepatic and/or renal effects in 

rats (15 or 45/group), guinea pigs (15 or 45/group), dogs (2 or 6/group), and monkeys (3, 9, or 

21/group) exposed to inhaled 1,1-DCE. Continuous exposure to 1,1-DCE, 24 hours/day over 90 

days, demonstrated more severe effects than intermittent exposure, 6 hours/day, 5 days/week for 

6 weeks, in the species tested. Unlike the other available subchronic and chronic studies, this 

principal study included multiple exposure levels of 0, 5, 15, 25 and 48 ppm (0, 20, 61, 101, and 

189 mg/m3

). Mortality, hematologic and body weight data were well tabulated and presented in 

this study. Histopathologic evaluation was conducted on the heart, lung, liver, spleen and 

kidneys. Following continuous exposure, adverse hepatic effects included focal necrosis in 

monkeys (LOAEL = 189 mg/m3

, NOAEL = 101 mg/m3), in dogs (LOAEL = 189 mg/m3, 

NOAEL = 101 mg/m3

), and in rats (LOAEL = 189 mg/m3, NOAEL = 101 mg/m3); and altered 

lipid content and increases in SGPT and alkaline phosphatase in guinea pigs (LOAEL = 189 

mg/m3, NOAEL = 20 mg/m3). Additionally, renal alterations were observed in rats as nuclear 

hypertrophy in the tubular epithelium (LOAEL = 189 mg/m3

, NOAEL = 61 mg/m3). Monkeys 

exposed to 1,1-DCE also displayed a greater than 25% decrease in body weight (LOAEL 189 

mg/m3

, NOAEL 20 mg/m3). The subchronic study by Prendergast et al. (1967) was chosen over 

the chronic studies because of its better design, its use of continuous exposure, and its exhibition 

of toxic effects below the LOAELs reported in the other studies. 

Although limited in number, the other chronic and subchronic studies available consistently 

demonstrate adverse hepatic effects following 1,1-DCE exposure (Lee et al., 1977; Maltoni et 

Appendix D3 173 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

al., 1985; Plummer et al., 1990; Quast et al., 1986). Hepatocellular fatty change was observed in 

rats exposed to 25 ppm or 75 ppm 1,1-DCE intermittently (6 hrs/d, 5 d/wk) for 18 months. This 

mid-zonal fatty change was also observed at the 12-month interim sacrifice, but did not appear to 

progress in severity or incidence over time (Quast et al., 1986). A more severe hepatocellular 

necrosis and renal tubular necrosis were observed in mice exposed to 55 ppm 1,1-DCE 6 hr/d, 5 

d/week for 1 year (Lee et al., 1977). 

For comparison, Quast et al. (1986) determined a LOAEL of 25 ppm for liver effects of minimal 

severity in rats after 18 months exposure. Use of continuous time adjustment to 4.5 ppm, 

multiplication by an RGDR of 1, and division by a total UF of 100 (3 for LOAEL to NOAEL, 3 

for interspecies, and 10 for intraspecies) results in an estimate of 45 ppb (200 μg/m3). 

VII. Data Strengths and Limitations for Development of the REL 

Uncertainty factors are appropriate due to the limited number of subchronic and chronic 

inhalation studies (greater than 1 year duration) in laboratory animals. In addition, few industrial 

surveys and epidemiological studies are available on the adverse effects of 1,1-DCE in humans; 

these are limited by small sample size, short follow-up, and/or brief exposure periods. But this 

limited evidence does suggest an association between repeated exposure to 1,1-DCE and liver 

damage in humans (EPA, 1976), and the key study is an animal study which found adverse 

hepatic effects. No toxicokinetic data regarding the absorption, distribution, metabolism or 

excretion of 1,1-DCE in humans are available. 

VIII. References

Balmer MF, Smith FA, Leach LJ, and Yuile CL. 1976. Effects in the liver of methylene chloride 

inhaled alone and with ethyl alcohol. Am. Ind. Hyg. Assoc. J. 37(6):345-352. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dallas CE, Weir FW, Feldman S, Putcha L, and Bruckner JV. 1983. The uptake and disposition 

of 1,1-dichloroethylene in rats during inhalation exposure. Toxicol. Appl. Pharmacol. 68(1):140-

151. 

Gage JC. 1970. The subacute inhalation toxicity of 109 industrial chemicals. Br. J. Ind. Med. 

27:1-18. 

HSDB. 1994. Hazardous Substances Data Bank. TOMES®. Denver, CO: Micromedex. 

Appendix D3 174 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Jaeger RJ, Shoner LG, and Coffman LJ. 1977. 1,1-Dichloroethylene hepatotoxicity: Proposed 

mechanism of action and distribution and binding of [14C] radioactivity following inhalation 

exposure in rats. Environ. Health Perspect. 21:113-119. 

Klimisch HJ, Deckardt K, and Mitea D. 1979. [Subchronic toxicity of vinylidene chloride in rats 

after 6 weeks exposure.] BASF Aktiengesellschaft, Ludwigshafen: 31. (German). HSE 

Translation No. 9Q. [as cited in USDHHS, 1994.] 

Lee CC, Bhandari JC, Winston JM, House WB, Peters PJ, Dixon RL, and Woods JS. 1977. 

Inhalation toxicity of vinyl chloride and vinylidene chloride. Environ. Health Perspect. 21:25-32. 

Maltoni C, Lefemine P, Chieco P, Cotti G, and Patella V. 1985. Experimental research on 

vinylidene chloride carcinogenesis. In: Archives of Research on Industrial Carcinogenesis. Vol. 

3. Maltoni C, Mehlman MA, eds. Princeton, NJ: Princeton Scientific Publishers. [as cited in 

USDHHS, 1994.] 

McKenna MJ, Zempel JA, Madrid EO, and Gehring PJ. 1978. The pharmacokinetics of [14C] 

vinylidene chloride in rats following inhalation administration. Toxicol. Appl. Pharmacol. 

45:599-610. 

Ott MG, Fishbeck WA, Townsend JC, and Schneider EJ. 1976. A health study of employees 

exposed to vinylidene chloride. J. Occup. Med. 18:735-738. 

Plummer JL, de la Hall PM, Ilsley AH, Jenner MA, and Cousins MJ. 1990. Influence of enzyme 

induction and exposure profile on liver injury due to chlorinated hydrocarbon inhalation. 

Pharmacol. Toxicol. 67:329-335. 

Prendergast JA, Jones RA, Jenkins LJ, and Siegel J. 1967. Effects on experimental animals of 

long-term inhalation of trichloroethylene, carbon tetrachloride, 1,1,1-trichloroethane, 

dichlorodifluoromethane, and 1,1-dichloroethylene. Toxicol. Appl. Pharmacol. 10:270-289. 

Quast JF. 1976. Pathology report on male and female rats exposed to vinylidene chloride vapors 

for 6 hours per day, 5 days per week during a 30-day period. Toxicology Research Laboratory, 

Health and Environmental Sciences, Dow Chemical U.S.A., Midland, MI. [as cited in USDHHS, 

1994.] 

Quast JF, McKenna MJ, Rampy LW, and Norris JM. 1986. Chronic toxicity and oncogenicity 

study on inhaled vinylidene chloride in rats. Fundam. Appl. Toxicol. 6:105-144. 

Reichert D, Werner HW, and Henschler D. 1978. Role of liver glutathione in 1,1-

dichloroethylene metabolism and hepatotoxicity in intact rats and isolated perfused rat liver. 

Arch. Toxicol. 41:169-179. 

USDHHS. 1994. United States Department of Health and Human Services. Toxicological profile 

for 1,1-dichloroethene (update). U.S. Department of Health & Human Services. Public Health 

Service. Agency for Toxic Substances and Disease Registry. TP-93/07. 

Appendix D3 175 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

U.S. EPA. 1976. United States Environmental Protection Agency. Summary characteristics of 

selected chemicals of near-term interest. EPA-56014-76-004. Washington, DC: U.S. 

Environmental Protection Agency. [as cited in USDHHS, 1994.] 

Wright PB, and Moore L. 1991. Potentiation of the toxicity of model hepatotoxicants by 

acetaminophen. Toxicol. Appl. Pharmacol. 109:327-335. 

Appendix D3 176 1,1-Dichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

DIETHANOLAMINE 

(DEA; 2,2’-iminodiethanol; 2,2’-iminobisethanol; diethylolamine; 2,2’-aminodiethanol; 2,2’-

dihydroxydiethylamine) 

CAS Registry Number: 111-42-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 3 μg/m3

 ( 0.6 ppb) 

Critical effect(s) Laryngeal lesions in rats 

Hazard index target(s) Respiratory system; cardiovascular system 

II. Physical and Chemical Properties (Melnick and Thomaszewski, 1990; Dow, 1980; 

CRC, 1994) 

Description Colorless crystals 

Molecular formula C4H11NO2

Molecular weight 105.14 g/mol 

Density 1.097 g/cm3

 @ 20°C 

Boiling point 268.8°C 

Melting point 28°C 

Vapor pressure 0.00014 torr @ 25°C 

Solubility Soluble in alcohol, water, acetone 

Conversion factor 1 ppm = 4.3 mg/m3

 @ 25°C 

III. Major Uses and Sources 

Diethanolamine is used in the formation of soaps, emulsifiers, thickeners, wetting agents, and 

detergents in cosmetic formulations (Melnick and Thomaszewski, 1990; Knaak et al., 1997). It 

is used as a dispersing agent in some agricultural chemicals, as an absorbent for acidic gases, as a 

humectant, as an intermediate in the synthesis of morpholine, as a corrosion inhibitor, and as a 

component in textile specialty agents (Beyer et al., 1983). Diethanolamine is permitted in 

articles intended for use in production, processing, or packaging of food (CFR, 1981; cited in 

Melnick and Thomaszewski, 1990). It is also found in adhesives, sealants, and cutting fluids 

(Melnick and Thomaszewski, 1990). The annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 1520 pounds of diethanolamine (CARB, 2000).

Appendix D3 177 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IV. Effects of Chronic Exposures to Humans

There have been no controlled or epidemiological studies of chronic diethanolamine exposure in 

humans. There is a single case report of occupational asthma determined to be due to the 

patient’s handling of a cutting fluid containing diethanolamine (Piipari et al., 1998). Specific 

bronchial provocation tests were done with the cutting fluid containing DEA and with DEA 

aerosol at two concentrations (0.75 mg/m3 and 1.0 mg/m3) below the occupational limit of 2.0 

mg/m3. DEA caused asthmatic airway obstruction at both concentrations, but IgE-antibodies 

specific for DEA were not found. 

V. Effects of Exposures in Animals 

Diethanolamine replaces choline in phospholipids (Blum et al., 1972). DEA also reversibly 

inhibits phosphatidylcholine synthesis by blocking choline uptake and competing for utilization 

in the CDP-choline pathway (Lehman-McKeeman and Gamsky, 1999). Systemic toxicity occurs 

in many tissue types including the nervous system, liver, kidney, and blood system. 

Gamer et al. (1996) exposed groups of 26 Wistar rats (13 male and 13 female) head-nose to a 

liquid aerosol of DEA for six hours per working day for 90 days at target concentrations of 15, 

150, and 400 mg/m3. Three of each sex were used for whole animal perfusion studies and the 

remaining 20 animals were examined for pathology. The study found no functional or 

morphological evidence of neurotoxicity. Retardation of body weight increase was observed in 

animals exposed to high concentrations. No systemic effects occurred at the low dose, but 

systemic effects in the liver, kidney, male reproductive system, and red blood cell occurred in the 

high concentration dose group. In the mid-dose group, mild liver and kidney effects were 

present. Local irritation of the larynx and trachea was found in the high and mid dose groups; 

irritating laryngeal effects were also detected in the low dose group. Based on this study 15 

mg/m3 is a NOAEL for liver and kidney effects and a LOAEL for irritation of the larynx. The 

equivalent continuous exposure at the LOAEL is 2.7 mg/m3

 (15 x 6/24 x 5/7). 

 Incidence of laryngeal lesions (Gamer et al., 1996) 

Aerosolized 

diethanolamine 

Chronic 

inflammation of the 

larynx 

Squamous 

hyperplasia 

Focal squamous 

metaplasia of 

laryngeal epithelium at 

base of the epiglottis 

0 None* None None 

15 mg/m3 4/20 0/20 20/20 

150 mg/m3 20/20 13/20 20/20 

400 mg/m3 20/20 17/20 20/20 

* The report does not give control incidences. Assumed 0/20. 

In an abstract Hartung et al. (1970) reported that inhalation by male rats of 6 ppm (25.8 mg/m3

) 

DEA vapor 8 hours/day, 5 days/week for 13 weeks resulted in depressed growth rates, increased 

lung and kidney weights, and even some mortality. Rats exposed continuously for 216 hours 

(nine days) to 25 ppm (108 mg/m3

) DEA showed increased liver and kidney weights, elevated 

Appendix D3 178 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

blood urea nitrogen (BUN), and increased serum glutamate oxaloacetate transferase (SGOT), an 

indicator of liver damage (Hartung et al., 1970). In studies at lower DEA levels, Eastman Kodak 

(1967) exposed dogs, weanling and adult rats, and guinea pigs to 0.26 ppm (1.1 mg/m3

) DEA for 

90 days and found no pathology attributable to DEA. In a 45-day study with 0.5 ppm (2.2 

mg/m3) DEA they also found no pathology attributable to DEA except for a possible slight 

retardation in rat growth rate. 

Gamer et al. (1993) exposed groups of 25 pregnant Wistar rats on gestation days 6-15 to a (noseonly) liquid aerosol of DEA at 10, 50 and 200 mg/m3

. Maternal toxicity, indicated by vaginal 

hemorrhage in 8 of the dams on gestation day 14, and fetotoxicity, evidenced by a statistically 

significant (p<0.05) increased incidence of total fetal skeletal variations, were observed at 200 

mg/m3

. No teratogenic effects were seen at any level. Thus 50 mg/m3 was a NOAEL for 

maternal toxicity and for embryo-fetal effects. 

A 13-week drinking water study in rats (10 per sex per group) showed significant dosedependent hematological changes following exposure to DEA at all concentrations tested: 320, 

630, 1250, 2500, and 5000 ppm in males, and 160, 320, 630, 1250, and 2500 ppm in females. 

Hematological effects included decreased hemoglobin and mean corpuscular volume (Melnick et 

al., 1994a). Similar hematological changes were observed following daily topical treatment. In 

addition to the hematological effects, female rats also showed dose-dependent spinal cord and 

medullary demyelination beginning at a drinking water concentration of 1250 ppm DEA. Male 

rats displayed demyelination beginning at 2500 ppm. Female rats gained significantly less 

weight than controls beginning at 63 mg/kg/day topical treatment. In a companion drinking 

water study (Melnick et al., 1994b), mice (10 per sex per group) were exposed to concentrations 

of 0, 630, 1250, 2500, 5000, and 10,000 ppm DEA and displayed dose-dependent hepatotoxicity, 

nephrotoxicity, and cardiac toxicity. Daily topical treatment in a separate study resulted in skin 

lesions in mice. Significant hepatic toxicity was observed at all drinking water concentrations, 

and skin lesions were observed at all topical doses. 

Data from female rats exposed to diethanolamine by Melnick et al. (1994) 

Dose (ppm) 

mg/kg/day 

DEA 

consumed Survival 

Mean bw 

change (g) Hgb (g/dL) 

Mean cell 

volume 

Mean cell 

Hgb (pg) 

 0 0 10/10 120±6a 15.1±0.3 56±0.2 17.9±0.2 

 160 14 9/10 106±3 15.2±0.1 55±0.2** 17.8±0.1* 

 320 32 10/10 98±3** 13.8±0.1** 54±0.2** 17.7±0.1** 

 630 57 10/10 95±4** 13.0±0.1** 53±0.3** 17.2±0.1** 

1250 124 10/10 85±4** 11.3±0.2** 51±0.3** 16.7±0.1** 

2500 242 10/10 63±4** 10.50±.2** 49±0.2** 16.30±.1** 

a Values are means±SEM; * p<0.05 or ** p<0.01 versus control group 

Barbee and Hartung (1979a) found that repeated treatment of rats with 330 mg DEA/kg/day 

significantly inhibited formation of phosphatidyl choline and phosphatidyl ethanolamine in the 

liver as compared with control rats. In a subsequent study, Barbee and Hartung (1979b) noted 

changes in liver mitochondrial activity in rats (4 per group) following exposure to DEA in 

Appendix D3 179 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

drinking water for up to 5 weeks. Mitochondrial changes were observed at 42 mg/kg/day after 2 

weeks. 

Daily oral treatment of male rats with 0, 250, 500, or 750 mg/kg/day for 5 days, or 

100 mg/kg/day for 14 days resulted in reduced activities of the liver enzymes microsomal 

hydroxylase and N-demethylase (Foster et al., 1971). 

In a developmental study Marty et al. (1999) administered DEA cutaneously to pregnant CD rats 

during gestation days 6-15 at doses of 0, 150, 500, and 1500 mg/kg/day. Dams exhibited 

reduced body weight at the highest dose, skin irritation and increased kidney weights at both 500 

and 1500 mg/kg/day, and a slight microcytic anemia with abnormal red blood cell morphology at 

all 3 dose levels. The blood results are consistent with the results of topical application of DEA 

by Melnick et al. (1994b). Rat fetuses had increased incidences of six skeletal variations at 1500 

mg/kg/day. Lower doses were without effect on the fetuses. Marty et al. (1999) also 

administered DEA cutaneously to pregnant New Zealand White rabbits on days 6-18 of gestation 

at 0, 35, 100, and 350 mg/kg/day. Dams administered the highest dose exhibited various skin 

lesions, reduced food consumption, and color changes in the kidneys, but no hematological 

changes. Body weight gain was reduced at ≥ 100 mg/kg/day. There was no evidence of 

maternal toxicity at 35 mg/kg/day and no evidence of developmental toxicity in rabbits at any 

dose. Developmental toxicity was observed only in the rat and only at doses causing significant 

maternal toxicity, including hematological effects. Due to a dose discrepancy, the authors 

adjusted the no observable effect level (NOEL) for DEA developmental toxicity to 380 

mg/kg/day for rats. In rabbits, the embryonal/fetal NOEL was 350 mg/kg/day. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Gamer et al. (1996) 

Study population Wistar rats (male and female) 

Exposure method Inhalation 6 h/day, 5 d/wk 

Critical effects Chronic inflammation and squamous hyperplasia 

and metaplasia of the larynx 

LOAEL 15 mg/m3

 

NOAEL Not observed 

Exposure duration 90 days 

Average experimental exposure 2700 μg/m3

 for LOAEL group 

(15 mg/m3

 x 6h/24h x 5d/7d x 1000 μg/mg) 

LOAEL uncertainty factor 3 (see below) 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 1000 

Inhalation reference exposure level 3 μg/m3

 (0.6 ppb) 

No chronic inhalation studies with diethanolamine were located in the peer-reviewed literature. 

Thus the 90 day study by Gamer et al., which found a LOAEL of 15 mg/m3 for irritation of the 

Appendix D3 180 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

rat larynx, was used to derive the REL. All 20 of the rats in the 15 mg/m3 exposure group 

showed focal squamous metaplasia of the laryngeal epithelium at the base of the epiglottis, and 4 

of the 20 had inflammatory cells present in the larynx. The former lesion seemed to be very 

limited and did not justify use of the full LOAEL uncertainty factor of 10. 

For comparison, the BASF (1993) developmental study by the inhalation route found a LOAEL 

of 200 mg/m3

 DEA and a NOAEL of 50 mg/m3 for fetotoxic effects. The equivalent continuous 

exposure at the NOAEL is 12.5 mg/m3. Multiplying by an RGDR of 1 and dividing by an 

interspecies uncertainty factor (UFA) of 3 and an intraspecies uncertainty factor (UFH) of 10 

results in a REL estimate of 40 μg/m3

. 

As another comparison, the study by Melnick et al. (1994a) shows dose-dependent adverse 

hematological and CNS effects in rats exposed to DEA in drinking water. Similar systemic 

effects were observed following dermal exposure. The Melnick et al. subchronic study was of 

the longest duration and was the most comprehensive report of the systemic effects of DEA in 

the literature. However, portal-of-entry effects of DEA have not been examined and should be 

addressed in future studies since this compound has irritant properties. The data from female rats 

were used since females were more sensitive than males to the hematologic effects of DEA. The 

LOAEL was 160 mg/L, or 14 mg/kg-day based on water consumption rates. Dividing by a 

LOAEL UF of 3, a subchronic UF of 3, an interspecies UF of 10, and an intraspecies UF of 10 

(cumulative UF = 1000) results in a oral REL of 0.014 mg/kg-day. Using route-to-route 

extrapolation and assuming that a 70 kg person inhales 20 m3 of air per day leads to an inhalation 

REL estimate of 50 μg/m3 (10 ppb) DEA. 

VII. Data Strengths and Limitations for Development of the REL 

The diethanolamine database is relatively weak. Major areas of uncertainty are the lack of 

adequate human exposure data, the absence of a NOAEL in the major study, the lack of 

reproductive and developmental toxicity studies, and the lack of chronic inhalation, multiplespecies, health effects data. 

VIII. Potential for Differential Impacts on Children's Health

Since the proposed chronic REL of 3 μg/m3

 based on laryngeal effects is much lower than the 

comparison REL of 40 μg/m3 based on fetotoxic effects, the REL should adequately protect 

infants and children. Diethanolamine is a respiratory irritant and thus might exacerbate asthma, 

which has a more severe impact on children than on adults. The large uncertainty factor of 1000 

should protect against that potential hazard. However, there is no direct evidence in the literature 

to demonstrate that DEA exacerbates asthma or to quantify a differential effect of 

diethanolamine on the larynx or on other organs in infants and children. 

Appendix D3 181 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IX. References 

Barbee SJ, and Hartung R. 1979a. The effect of diethanolamine on hepatic and renal 

phospholipid metabolism in the rat. Toxicol. Appl. Pharmacol. 47:421-430. 

Barbee SJ, and Hartung R. 1979b. Diethanolamine-induced alteration of hepatic mitochondrial 

function and structure. Toxicol. Appl. Pharmacol. 47:431-440. 

Beyer KH Jr, Bergfeld WF, Berndt WO, Boutwell RK, Carlton WW, Hoffman DK, and 

Schroeter AL. 1983. Final report on the safety assessment of triethanolamine, diethanolamine, 

and monoethanolamine. J. Am. Coll. Toxicol. 2:183-235. 

Blum K, Huizenga CG, Ryback RS, Johnson DK, and Geller I. 1972. Toxicity of diethanolamine 

in mice. Toxicol. Appl. Pharmacol. 22:175-185. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CFR. 1981. Code of Federal Regulations. Title 21, Parts 175.105, 175.300, 176.170, 176.180, 

176.200, 176.210, 177.1680, 177.2600, 177.2800, 178.3910. U.S. Printing Office, Washington, 

D.C. [as cited in Melnick and Thomaszewski, 1990.] 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dow Chemical Co. 1980. Alkanolamines Handbook. Midland, MI: Dow Chemical Company. 

Eastman Kodak. Co. 1967. Health and safety studies for diethanolamine, Laboratory tests to 

determine effect of inhalation of two ethanolamines - diethanolamine (DEA), 

methylaminoethanol (MAE), Formulation 485K - Histological addendum to Final Report. TSCA 

8d Submission 86-890000205, Microfiche Number OTS0516742. Washington, D.C.: OPPT, 

U.S. EPA. 

Foster GV Jr, Hartung R, and Cornish HH. 1971. Inhibition of hepatic microsomal enzymes by 

N-substituted ethanolamines. Toxicol. Appl. Pharmacol. 19:847-855. 

Gamer AO, Hellwig J, and Hildebrand B. 1993. Study of the prenatal toxicity of diethanolamin 

in rats after inhalation. BASF Project No. 31RO233/90010. Ludwigshafen, FRG: BASF 

Gamer AO, Mellert W, Leibold E, Deckardt K, Kaufman W, and Hildebrand B. 1996. 

Diethanolamin – Subchronic inhalation toxicity and neurotoxicity study in Wistar rats. 90-day 

liquid aerosol exposure-. BASF Project No. 50I0075/93011. Heidelberg, FRG: BASF. 

Hartung R, Rigas LK, and Cornish HH. 1970. Acute and chronic toxicity of diethanolamine. 

Toxicol. Appl. Pharmacol. 17:308 (abstract). 

Appendix D3 182 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Knaak JB, Leung HW, Stott WT, Busch J and Bilsky J. 1997. Toxicology of mono-, di-, and 

triethanolamine. Rev. Environ. Contam. Toxicol. 49:1-86. 

Lehman-McKeeman LD, and Gamsky EA. 1999. Diethanolamine inhibits choline uptake and 

phosphatidylcholine synthesis in Chinese hamster ovary cells. Biochem. Biophys. Res. Commun. 

262(3):600-604. 

Marty MS, Neeper-Bradley TL, Neptun DA, and Carney EW. 1999. Developmental toxicity of 

diethanolamine applied cutaneously to CD rats and New Zealand White rabbits. Regul. Toxicol. 

Pharmacol. 30(3):169-181. 

Melnick RL, and Thomaszewski KE. 1990. Diethanolamine. In: Ethel Browning’s Toxicity and 

Metabolism of Industrial Solvents. 2nd ed. Buhler DR, and Reed DJ. (eds.) Vol 2: Nitrogen and 

Phosphorus Solvents. New York: Elsevier. pp. 401-410. 

Melnick RL, Mahler J, Bucher JR, Thompson M, Hejtmancik M, Ryan MJ, and Mezza LE. 

1994a. Toxicity of diethanolamine. 1. Drinking water and topical application exposures in F344 

rats. J. Appl. Toxicol. 14(1):1-9. 

Melnick RL, Mahler J, Bucher JR, Hejtmancik M, Singer A, and Persing R. 1994b. Toxicity of 

diethanolamine. 2. Drinking water and topical application exposures in B6C3F1 mice. J. Appl. 

Toxicol. 14(1):11-19. 

Piipari R, Tuppurainen M, Tuomi T, Mantyla L, Henriks-Eckerman ML,Keskinen H, and 

Nordman H. 1998. Diethanolamine-induced occupational asthma, a case report. Clin. Exp. 

Allergy 28(3):358-362. 

Appendix D3 183 Diethanolamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

N,N-DIMETHYLFORMAMIDE 

(N-formyldimethylamine) 

CAS Registry Number: 68-12-2 

I. Chronic Toxicity Summary

Inhalation reference exposure level 80 μg/m3 (30 ppb) 

Critical effect(s) Liver dysfunction and respiratory irritation in 

humans 

Hazard index target(s) Alimentary system, respiratory system 

II. Chemical Property Summary (HSDB, 1994) 

 

Description Colorless to very slightly yellow liquid 

Molecular formula C3H7NO 

Molecular weight 73.09 g/mol 

Boiling point 153°C 

Melting point −61°C 

Vapor pressure 3.7 torr @ 25°C 

Solubility Soluble in alcohol, ether, acetone, benzene, 

and chloroform; miscible with water 

Conversion factor 2.99 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Dimethylformamide (DMF) is primarily used as a solvent in the production of polyurethane 

products and acrylic fibers. It is also used in the pharmaceutical industry, in the formulation of 

pesticides, and in the manufacture of synthetic leathers, fibers, films, and surface coatings 

(Howard, 1993; Gescher, 1993; Redlich et al., 1988). DMF may be emitted to the environment 

as a result of its use in a variety of petrochemical industries (Howard, 1993). The annual 

statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California based on the most recent inventory were estimated to be 18,249 pounds of DMF 

(CARB, 2000). 

IV. Effects of Human Exposure 

Among 100 workers occupationally exposed to DMF for at least one year (mean exposure of 5 

years; range = 1-15 years), a statistically significant incidence of hepatic impairment, as 

indicated by elevated gamma-glutamyl transpeptidase levels and digestive disturbances, was 

Appendix D3 184 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

noted (Cirla et al., 1984). Other changes, that were not statistically significant, included 

increased SGOT and SGPT and enlarged livers. The mean time-weighted average concentration 

of DMF was 22 mg/m3

 (range = 8-58 mg/m3). Symptoms of irritation occurring only during 

work at statistically significantly higher incidences included watery eyes, dry throat, and 

coughing. Also, the exposed workers reported a reduced sense of smell and dry coughs at home 

with a statistically significant difference as compared to controls. Several of the DMF exposed 

workers also reported alcohol intolerance characterized by a disulfiram-type reaction (facial 

flushing and palpitations following alcohol ingestion). Alcohol consumption, a potential 

confounder, was controlled for in the study design. 

A similar study was conducted on workers who had been employed in an acrylic acid fiber plant 

for more than 5 years (Cantenacci et al., 1984). Concentrations to which the workers were 

exposed were characterized as either an 8-hour TWA of 18 mg/m3

 or an 8-hour TWA of 3 

mg/m3

. Measures of liver function including SGOT, SGPT, gamma-glutamyl transferase, and 

alkaline phosphatase levels were not significantly different between exposed and unexposed 

workers. However, the U.S. EPA cautions that because only 54 matched pairs of workers were 

examined, the power of this study was not high enough to reliably detect a difference in enzyme 

levels. 

Redlich et al. (1988) characterized a plant-wide outbreak of liver disease among workers in a 

factory coating fabric with polyurethane. Fifty-eight of 66 (88%) workers participated and each 

had standard liver screening function tests done at least once. At the work site DMF was being 

used in poorly ventilated areas without appropriate skin protection. No other major known 

hepatotoxic exposure was identified. Overall, 36 of 58 (62%) workers tested had elevations of 

either aspartate aminotransferase (AST) or alanine aminotransferase (ALT) levels. Enzyme 

abnormalities occurred almost exclusively in production workers (35 out of 46 abnormal). Only 

1 of 12 non-production workers showed elevations in enzyme levels (p < 0.0001). Serologic 

tests excluded known infectious causes of hepatitis in all but 2 workers. Changes, characteristic 

of liver injury, were confirmed by histologic examination of biopsy specimens from 4 workers. 

Improvement in liver enzyme abnormalities and symptoms in most patients were seen, after 

modification of work practices and removal of workers most severely affected from exposure. 

However, some patients showed persistent elevations of enzyme levels. No measurements or 

estimates of DMF exposure levels were reported. 

Wang et al. (1991) investigated the prevalence of liver injury associated with DMF exposure in 

183 of 204 (76%) employees of a synthetic leather factory by performing medical examinations, 

liver function tests, and creatine phosphokinase (CPK) determinations. Air concentrations were 

measured with personal samplers and gas chromatography. The concentration of DMF in air to 

which each worker was exposed was categorized as high (DMF exposure index 2: 25-60 ppm; 

75-180 mg/m3), medium (index 1: 10-40 ppm), and low (index 0: <10 ppm). High exposure 

concentrations were significantly associated with elevated alanine aminotransferase (ALT) levels 

(i.e., greater than or equal to 35 International Units/liter), a result that did not change after 

stratification by hepatitis B carrier status. Logistic regression analysis indicated that exposure to 

high DMF levels was associated with elevated ALT (p = 0.01), whereas hepatitis B surface 

antigen (HBsAg) was slightly but independently associated with elevated ALT (p = .07). 

Workers with normal ALT values had significantly higher mean ALT and aspartate 

Appendix D3 185 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

aminotransferase (AST) activities, especially among those who were not HBsAg carriers. A 

significant association existed between elevated CPK levels and exposure to DMF. However, an 

analysis of the CPK isoenzyme among 143 workers did not reveal any specific damage to 

muscles. Thus the authors ascribed the liver injury to DMF. 

U.S. EPA (1994) states that subjective evidence of liver toxicity, such as digestive impairment 

and alcohol intolerance, is often observed at exposures below those that cause clinical changes in 

liver enzymes. Thus, the symptoms may be more sensitive indicators of hepatic impairment. 

Three unexplained cases of small-for-date third trimester intrauterine deaths were observed in a 

group of women working as quality control analysts in the pharmaceutical industry (Farquhason 

et al., 1983). This represented a 30% stillbirth rate as compared with the average for the general 

population of about 0.26%. While the authors concluded that the occurrence of stillbirth in these 

women was not likely due to chance, the effects cannot be solely attributed to DMF because the 

women were exposed to other agents in addition to DMF. 

V. Effects of Animal Exposure 

Malley et al. (1994) exposed male and female Crl:CD rats and mice to 0, 25, 100, or 400 ppm 

DMF for 6 hr/day, 5 days/week for 18 months (mice) or 2 years (rats). No compound-related 

effects on clinical observations or survival were observed. Body weights of rats exposed to 100 

(males only) and 400 ppm were reduced, while body weights were increased in 400 ppm mice. 

No hematologic changes were observed in either species. Serum sorbitol dehydrogenase activity 

was increased in rats exposed to 100 or 400 ppm. DMF-related morphological changes were 

observed only in liver. Exposure of rats to 100 and 400 ppm produced increased relative liver 

weights, centrilobular hepatocellular hypertrophy, lipofuscin/hemosiderin accumulation in 

Kupffer cells, and centrilobular single cell necrosis (400 ppm only). In mice, increased liver 

weights (100 ppm males, 400 ppm both sexes), centrilobular hepatocellular hypertrophy, 

accumulation of lipofuscin/hemosiderin in Kupffer cells, and centrilobular single cell necrosis 

were observed in all exposure groups. These observations occurred in a dose-response fashion 

and were minimal at 25 ppm. No increase in hepatic cell proliferation was seen in mice or 

female rats. Slightly higher proliferation was seen in male rats exposed to 400 ppm at 2 weeks 

and 3 months but not at 12 months. Thus 25 ppm was a chronic NOAEL for both rats and mice. 

A developmental toxicity study using three species (mice, rabbits, and rats) and four routes of 

administration (oral, inhalation, dermal, and intraperitoneal) identified the rabbit as the most 

sensitive of the three species. Groups of 15 pregnant rabbits were exposed for 6 hours per day 

on days 8-20 of gestation to 50, 150, or 450 ppm (150, 449, or 1350 mg/m3

) DMF (Hellwig et 

al., 1991). Slight maternal toxicity, as indicated by non-statistically significant decreases in 

maternal body weight gain, was observed in the 450 ppm exposure group. An increased number 

of total malformations per litter was observed in the 450 ppm exposure group. Malformations 

observed at statistically higher incidences compared to controls included hernia umbilicalis, 

external variations, pseudoankylosis of the forelimbs, and skeletal variation and retardation. The 

authors conclude that there was a clear teratogenic effect in rabbits following maternal exposure 

to 450 ppm DMF and a marginal effect following exposure to 150 ppm DMF. A NOAEL of 

Appendix D3 186 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

50 ppm for fetal and maternal effects was reported. Inhalation exposure to 150 ppm was 

calculated by the authors to approximate a daily dose of 45 mg/kg/day, which coincides with 

previous work on this compound in this species. 

VI. Derivation of Chronic Reference Exposure Level 

Study Cirla et al., 1984; Catenacci et al., 1984 

Study population Occupationally exposed workers 

Exposure method Discontinuous inhalation exposures 

Critical effects Digestive disturbances and slight hepatic 

changes 

LOAEL 22 mg/m3

NOAEL Not observed 

Exposure continuity 8 hr/day (10 m3

/day), 5 days/week (assumed) 

Average occupational exposure 7.9 mg/m3

 for LOAEL group (22 x 10/20 x 5/7) 

Human equivalent concentration 7.9 mg/m3

Exposure duration 5 years (mean exposure duration) 

LOAEL uncertainty factor 3 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.08 mg/m3 (80 μg/m3

, 0.03 ppm, 30 ppb) 

The U.S. EPA (1994) based its RfC of 30 μg/m3 on the same study but included a Modifying 

Factor (MF) of 3 due to lack of reproductive toxicity data in the DMF database. The criteria for 

use of modifying factors are not well specified by U.S. EPA. Such modifying factors were not 

used by OEHHA. Intermediate uncertainty factors were used for LOAEL to NOAEL and 

subchronic to chronic extrapolation because of the mild nature of the effects observed and the 

less than chronic exposure duration. 

For comparison Hellwig et al. (1991) found a developmental NOAEL of 50 ppm in rabbits 

exposed 6 hours per day on gestation days 8-20, equivalent to continuous exposure of 12.5 ppm. 

Multiplication by an RGDR of 1 and division by a UF of 30 (3 for interspecies and 10 for 

intraspecies) results in a REL estimate of 400 ppb. The NOAEL of 25 ppm for rats and mice in 

the chronic study of Malley et al. (1994) leads to a REL estimate of 150 ppb. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the REL for N,N-dimethylformamide is the availability of human health 

effects data over several years of exposure. The major uncertainties are the difficulty in 

estimating exposure patterns and magnitude, the lack of a NOAEL observation, and the lack of 

complete reproductive and developmental toxicity data.

Appendix D3 187 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Catenacci G, Grampella D, Terzi R, Sala A, and Pollini G. 1984. Hepatic function in subjects 

exposed to environmental concentrations of DMF lower than the actually proposed TLV. G. Ital. 

Med. Lav. 6(3-4):157-158. 

Cirla AM, Pisati G, Invernizzi E, and Torricelli P. 1984. Epidemiological study on workers 

exposed to low dimethylformamide concentrations. G. Ital. Med. Lav. 6(3-4):149-156. 

Farquharson RG, Hall MH, and Fullerton WT. 1983. Poor obstetric outcome in three quality 

control laboratory workers. Lancet 1(8331):983-984. [cited in U.S. EPA, 1994]. 

Gescher A. 1993. Metabolism of N,N-dimethylformamide: key to the understanding of its 

toxicity. Chem. Res. Toxicol. 6(3):245-251. 

HSDB. 1994. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/94). 

Hellwig J, Merkle J, Klimisch HJ, and Jackh R. 1991. Studies on the prenatal toxicity of N,Ndimethylformamide in mice, rats and rabbits. Food Chem. Toxicol. 29(3):192-201. 

Howard PH. (ed) 1993. Handbook of Environmental Fate and Exposure Data for Organic 

Chemicals. Vol. IV: Solvents 2. Chelsea, MI: Lewis Publishers Inc. 

Malley LA, Slone TW Jr, Van Pelt C, Elliott GS, Ross PE, Stadler JC, and Kennedy GL Jr. 1994. 

Chronic toxicity/oncogenicity of dimethylformamide in rats and mice following inhalation 

exposure. Fundam. Appl. Toxicol. 23(2):268-279. 

Redlich C, Beckett WS, Sparer J, Barwick KW, Riely CA, Miller H, Sigal SL, Shalat SL, and 

Cullen MR. 1988. Liver disease associated with occupational exposure to the solvent 

dimethylformamide. Ann. Intern. Med. 108:680-686. 

U.S. EPA. 1994. U.S. Environmental Protection Agency. IRIS. 1994. Integrated Risk 

Information System (IRIS). Reference concentration for chronic inhalation exposure (RfC) for 

N,N,-Dimethylformamide. Available online at http://www.epa.gov/ngispgm3/iris 

Wang JD, Lai MY, Chen JS, Lin JM, Chiang JR, Shiau SJ, and Chang WS. 1991. 

Dimethylformamide-induced liver damage among synthetic leather workers. Arch. Environ. 

Health 46(3):161-166. 

Appendix D3 188 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

1,4-DIOXANE 

(Synonym: dihydro-p-dioxin, diethylene dioxide, p-dioxane, glycolethylene ether) 

CAS Registry Number: 123-91-1 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 3,000 μg/m3

 (800 ppb) 

Critical effects Liver, kidney, hematologic changes in rats 

Hazard index target(s) Alimentary system; kidney; circulatory system 

II. Chemical Property Summary (HSDB, 1995; 1999; CRC, 1994) 

Description Colorless liquid with a faint, pleasant odor 

Molecular formula C4H8O2

Molecular weight 88.10 g/mol 

Boiling point 101.5 o

C 

Melting point 11.8o

C 

Vapor pressure 37 torr @ 25o

C 

Solubility Miscible with water, aromatic solvents, and oils 

Kow 0.537 

Conversion factor 3.60 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

1,4-Dioxane (dioxane), a cyclic ether, is used as a degreasing agent, as a component of paint and 

varnish removers, and as a wetting and dispersion agent in the textile industry. Dioxane is used 

as a solvent in chemical synthesis, as a fluid for scintillation counting, and as a dehydrating agent 

in the preparation of tissue sections for histology (Grant and Grant, 1987; HSDB, 1995). The 

annual statewide emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California based on the most recent inventory were estimated to be 155,549 pounds of 1,4-

dioxane (CARB, 1999). 

IV. Effects of Human Exposure 

Dioxane is absorbed by all routes of administration (HSDB, 1995). In humans, the major 

metabolite of dioxane is β-hydroxyethoxyacetic acid (HEAA) and the kidney is the major route 

of excretion (Young et al., 1976). The enzyme(s) responsible for HEAA formation has not been 

studied, but data from Young et al. (1977) indicate saturation does not occur up to an inhalation 

exposure of 50 ppm for 6 hours. Under these conditions the half-life for dioxane elimination is 

Appendix D3 189 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

59 min (plasma) and 48 min (urine). Although physiologically based pharmacokinetic (PBPK) 

modeling suggests HEAA is the ultimate toxicant in rodents exposed to dioxane by ingestion, the 

same modeling procedure does not permit such a distinction for humans exposed by inhalation 

(Reitz et al., 1990). 

Several anecdotal reports have appeared in which adverse health effects due to chronic dioxane 

exposure are described. Barber (1934) described dioxane exposed factory workers, some of 

whom exhibited signs of liver changes, increased urinary protein and increased white blood cell 

counts, and some of whom died from apparent acute exposures. Although the kidney and liver 

lesions were considered manifestations of acute exposure, the author suggested a chronic 

component that was manifested by increased white blood cells. A case was reported in which a 

worker, who died following exposure by inhalation and direct skin contact to high (unspecified) 

dioxane levels, exhibited lesions in the liver, kidneys, brain and respiratory system, but the 

effects could not be easily separated from the effects due to high intake of alcohol (Johnstone, 

1959). 

In a German study (Thiess et al., 1976 / in German, described in NIOSH, 1977) 74 workers 

exposed to dioxane in a dioxane-manufacturing plant (average potential exposure duration - 25 

years) underwent evaluation for adverse health effects. Air measurements indicated dioxane 

levels varied from 0.01 to 13 ppm. Clinical evaluations were applied to 24 current and 23 

previous workers. Evidence of increased (i.e., abnormal) aspartate transaminase (also known as 

serum glutamate-oxalacetic transaminase or SGOT), alanine transaminase (serum glutamate 

pyruvate transaminase or SGPT), alkaline phosphatase, and gamma glutamyltransferase 

activities (liver function) was noted in these workers, but not in those who had retired. The 

indicators of liver dysfunction, however, could not be separated from alcohol consumption or 

exposure to ethylene chlorohydrin and/or dichloroethane. 

A follow-up mortality study was conducted on 165 chemical plant manufacturing and processing 

workers who were exposed to dioxane levels ranging from less than 25 to greater than 75 ppm 

between 1954 and 1975 (Buffler et al., 1978). Total deaths due to all causes, including cancer, 

did not differ from the statewide control group, but the data were not reanalyzed after removing 

the deaths due to malignant neoplasms. The study is limited by the small number of deaths and 

by the small sample number. The study did not assess hematologic or clinical parameters that 

could indicate adverse health effects in the absence of mortality. 

Yaqoob and Bell (1994) reviewed human studies on the relationship between exposure to 

hydrocarbon solvents - including dioxane - and renal failure, in particular rare 

glomerulonephritis. The results of their analysis suggest that such solvents may play a role in 

renal failure, but dioxane was not specifically discussed. Of interest to the discussion on chronic 

exposure to dioxane is the suggestion that the mechanism of the disease process involves local 

autoimmunity with decreased circulating white blood cells (see below). 

V. Effects of Animal Exposure 

In rats, the major metabolite of dioxane is HEAA, which is excreted through the kidneys (Braun 

and Young, 1977). Exposure to dioxane by ingestion results in saturation of metabolism above 

Appendix D3 190 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

100 mg/kg given in single dose. Saturation of metabolism was also observed as low as 10 mg/kg 

if dioxane was administered in multiple doses. Dioxane itself is not cleared through the kidney. 

A decrease in metabolic clearance with increasing dose (iv) has been interpreted as the saturation 

of metabolism at the higher doses (Young et al., 1978). 

For Sprague-Dawley rats, the metabolic fate of inhaled dioxane (head only exposure) was based 

on one air concentration (50 ppm). At this level, nearly all the dioxane was metabolized to 

HEAA since HEAA represented 99 percent of the total dioxane + HEAA measured. The plasma 

half-life for dioxane under these conditions was 1.1 hours. The absorption of dioxane through 

the inhalation pathway could not be exactly determined, because of a high inhalation rate (0.24 

liters/min), calculated on the basis of complete absorption (Young et al., 1978; U.S. EPA, 1988). 

Although the high inhalation rate could be dioxane-related, another explanation may be the stress 

incurred when the jugular veins were cannulated as part of the experiment. Extensive absorption 

by inhalation is also inferred from the high tissue/air partition coefficients (Reitz et al., 1990). 

Although the PBPK modeling suggests that in rat the parent dioxane is a better dose surrogate 

than HEAA for exposure by ingestion, the inhalation modeling did not use more than one 

inhalation dose. No studies were located on the biological or biochemical properties of HEAA 

or the properties of the enzyme(s) that are responsible for the transformation of dioxane into 

HEAA. 

Rats (Wistar) were exposed by inhalation to dioxane (111 ppm; 7 hours/day, 5 days/week) for 2 

years (Torkelson et al., 1974). Increased mortality and decreased body weight gains, compared 

to unexposed control rats, were not observed. Among the male rats, decreased blood urea 

nitrogen (kidney function), decreased alkaline phosphatase (cholestatic liver function), increased 

red blood cells, and decreased white blood cells were observed. According to the authors, 

exposure-related, non-cancerous tissue lesions were not observed during the 2-year period. 

In another inhalation study, rats were exposed to dioxane at levels of 0.15, 1.3, and 5.7 ppm 

(Pilipyuk et al., 1978). Frequency was not specified, but the duration is given as “90 successive 

days”. At the end of the 3-month exposure, increased SGOT activity at the two highest doses 

and increased SGPT activity at all doses were measured in the sera of the exposed rats. Rats 

exposed to the highest dose also exhibited increased urinary protein and chloride levels, each of 

which returned to control levels during an unspecified recovery period. Pilipyuk et al. (1978) 

also report changes in the minimum time (ms) required for an electric stimulus to result in 

excitation of extensor and flexor muscles. Although Pilipyuk et al. (1978) consider the changes 

to be a reflection of adverse effects due to exposure to dioxane, Torkelson et al. (1974) do not 

consider the hematologic and clinical changes of toxicologic importance. In particular, toxic 

manifestations are usually associated with increased blood urea nitrogen and alkaline 

phosphatase levels, whereas these levels decreased in the Torkelson et al. (1974) investigation. 

The reason for the discrepancies between the two studies, in particular the extremely low 

dioxane exposure levels in the Pilipyuk et al. (1978) study, is unknown. One explanation could 

be the purity of the dioxane used, which was not described in the latter study, although such 

contamination would be unlikely to account for the large difference in exposure levels. 

Appendix D3 191 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Kociba et al. (1974) exposed rats (Sherman) to dioxane by ingestion of drinking water for up to 

2-years. The drinking water levels were 0, 0.01, 0.1, and 1.0 percent, which were converted to 

daily intake according to measured rates of water consumption during exposure. Exposure to the 

highest level resulted in decreased body weight gain and increased deaths. According to the 

authors, exposure related hematologic changes did not occur. Histopathologic examination 

revealed evidence of regeneration of hepatic and kidney tissues in rats exposed to 1.0 or 0.1 

percent, but not in rats exposed to 0.01 percent dioxane. On the assumption of total absorption 

of dioxane from the gastrointestinal tract, the exposure levels in female and male rats is as 

follows: 0.01%-18 ppm/F, 9.3 ppm/M; 0.1% -144 ppm/F, 91 ppm/M. 

The teratogenic potential of dioxane was studied in rats (Giavini et al., 1985). Dioxane was 

administered by gavage at doses of 0, 0.25, 0.5, and 1.0 ml/kg-day, on gestation days 6-15, and 

observations continued through day 21. Dams exposed to the highest dose exhibited 

nonsignificant weight loss and a significant decrease in food consumption during the first 16 

days. During the remaining 5 days, food consumption increased, but the weight gain reduction 

in the presence of dioxane continued. At the 1.0 ml/kg-day dose, mean fetal weight and ossified 

sternebrae were also reduced. The inability to separate the developmental toxicity from maternal 

or embryotoxicity renders these data inconclusive as to the developmental toxicity of dioxane. If 

toxicity to the dam and/or embryo exists, the NOAEL for dioxane (based on density = 1.03 

gm/ml) is 517 mg/kg-day. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Torkelson et al. (1974) 

Study populations Rats 

Exposure method Discontinuous inhalation 

Critical effects No effects on liver, kidney, or hematologic 

function were noted in this study. Such 

dysfunctions, however, were observed in rats 

exposed to dioxane by ingestion (Kociba et al.

1974) and humans (Thiess, et al., 1976, 

described by NIOSH, 1977). 

LOAEL Not observed in inhalation studies

NOAEL 111 ppm 

Exposure continuity 7 h/d x 5 days/wk 

Average experimental exposure 23 ppm (111 x 7/24 x 5/7) 

Human equivalent concentration 23 ppm (gas with systemic effects, based on 

RGDR = 1.0 using default assumption that 

lambda (a) = lambda (h)) 

Exposure duration 2 years 

LOAEL uncertainty factor 1 

Subchronic exposure 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.8 ppm (800 ppb; 2.8 mg/m3

; 3000 μg/m3) 

Appendix D3 192 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

The lifetime rat inhalation study of Torkelson et al. (1974) is the only detailed inhalation study 

available in the literature. The Pilipyuk et al. (1977) study contains useful and consistent data, 

but the absence of necessary details prevents the use of these results for the determination of a 

chronic reference exposure level (REL). Although the ingestion study (Kociba et al., 1974) 

shows unequivocal toxic responses (liver and kidney) of the rat to dioxane by ingestion, 

exposure to 111 ppm by inhalation leads to equivocal results (Torkelson et al., 1974). In 

particular, serum markers for liver and kidney dysfunction decrease in value, whereas toxic 

responses are associated with increased levels. The lack of toxic hematologic endpoints 

observed in the ingestion study suggests that toxicity of dioxane may be route-of-exposure 

specific. Hematologic changes were also observed in the early worker study wherein changes in 

white blood cell count occurred (Barber, 1934), but the directions are different. The studies on 

humans and rodents therefore suggest inhalation of dioxane may lead to adverse biologic effects, 

but good dose-response data are not available. A partial explanation may lie in the doseresponse characteristic of the metabolism of dioxane, wherein toxicity may be a function of the 

saturation of metabolism. For inhalation, neither the point of saturation nor the mechanism has 

been established. Importantly, the end-point for dioxane chronic exposure may not be 

established. 

VII. Data Strengths and Limitations for Development of the REL

Although a free-standing NOAEL is not a desirable parameter to use for the development of a 

chronic REL, other studies support the conclusion that exposure to dioxane leads to adverse 

health effects. These observations have been documented among experimental animals (Kociba 

et al., 1974; Pilipyuk et al., 1977) and humans (Thiess et al., 1976, described in NIOSH, 1977). 

Until additional data from inhalation dose-response studies become available, a chronic REL 

based on the free-standing NOAEL is considered the best available. 

The strength of the REL for 1,4-dioxane is that it is based on a full lifetime study, with a large 

number of toxic endpoints and a good sample size. The weaknesses include use of a free 

standing NOAEL, the limited human data, and the lack of developmental studies. 

VIII. References 

Barber H. 1934. Haemorrhagic nephritis and necrosis of the liver from dioxane poisoning. Guy’s 

Hospital Report. 84:267-280. 

Braun WH, and Young JD. 1977. Identification of β-hydroxyethoxyacetic acid as the major 

urinary metabolite of 1,4-dioxane in the rat. Toxicol. Appl. Pharmacol. 39:33-38. 

Buffler PA, Wood SM, Suarez MS, and Kilian DJ. 1978. Mortality follow-up of workers 

exposed to 1,4-dioxane. J. Occup. Med. 20:255-259. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Appendix D3 193 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Giavini W, Vismara C, and Broccia ML. 1985. Teratogenesis study of dioxane in rats. Toxicol. 

Lett. 26:85-88. 

Grant R, and Grant C. 1987. Grant and Hackh’s Chemical Dictionary. R. Grant and C. Grant, 

eds. 5th ed. New York: McGraw-Hill Book Co. p. 189. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (edition expires 7/31/95). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Johnstone RT. 1959. Death due to dioxane? AMA Arch. Ind. Health. 20:445-447. 

Kociba RJ. 1974. Chronic toxicity study of dioxane in the drinking water of Sherman rats. 

Midland,MI: Dow Chemical Company. 

NIOSH. 1977. Criteria for a Recommended Standard. Occupational Exposure to Dioxane. 

National Institute for Occupational Safety and Health, Centers for Disease Control, Public Health 

Service, Department of Health Education and Welfare. Publication No. 77-226. 

Pilipyuk ZI, Gorban GM, Solomin GI, and Gorshunova AI. 1977. Toxicology of 1-4-dioxane. 

Space Biology and Aerospace Medicine. 11:70-74. (translated from Russian). 

Reitz RH, McCroskey PS, Park CN, Andersen ME, and Gargas ML. 1990. Development of a 

physiologically based pharmacokinetic model for risk assessment with 1,4-dioxane. Toxicol. 

Appl. Pharmacol. 105:37-54. 

Thiess AM, Tress E, Fleig I. 1976. [Industrial-medical investigation results in the case of 

workers exposed to dioxane.] (Ger.) Arbeitsmed. Sozialmed. Praventivmed. 11:35-46. 

Torkelson TR, Leong BKJ, Kociba RJ, Richter WA, and Gehring PJ. 1974. 1.4-Dioxane. II. 

Results of a 2-year inhalation study in rats. Toxicol. Appl. Pharmacol. 30:287-298. 

U.S. EPA (U.S. Environmental Protection Agency). 1988. Recommendations for and 

Documentation of Biological Values for Use in Risk Assessment. Chapter 4. Cincinnati, OH: 

United States Environmental Protection Agency. 

Yaqoob M, and Bell GM. 1994. Occupational factors and renal disease. Renal Failure. 16:425-

434. 

Young JD, Braun WH, and Gehring PJ. 1978. Dose-dependent fate of 1,4-dioxane in rats. J. 

Toxicol. Environ. Health. 4:709-726. 

Appendix D3 194 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Young JD, Braun WH, Gehring PJ, Horvath BS, and Daniel RL. 1976. 1,4-Dioxane and βhydroxyethoxyacetic acid excretion in urine of humans exposed to dioxane vapors. Toxicol. 

Appl. Pharmacol. 38:643-646. 

Young JD, Braun WH, Rampy LW, Chenoweth MB, and Blau GE. 1977. Pharmacokinetics of 

1,4-dioxane in humans. J. Toxicol. Environ. Health. 3:507-520. 

Appendix D3 195 Dimethylformamide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

CHRONIC TOXICITY SUMMARY 

EPICHLOROHYDRIN 

(1-chloro-2,3-epoxy-propane) 

CAS Registry Number: 106-89-8 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 3 μg/m3 (0.8 ppb) 

Critical effects Histological changes in nasal turbinates in rats 

Hazard index target(s) Respiratory system; eyes 

II. Physical and Chemical Properties (HSDB, 1997; CRC, 1994) 

Description Colorless liquid 

Molecular formula C3H5ClO 

Molecular weight 92.52 g/mol 

Density 1.181 g/cm3

 @ 20° C 

Boiling point 117° C 

Melting point −26°C 

Vapor pressure 13 torr @ 20° C 

Solubility Slightly soluble in water, soluble in most organic 

solvents 

Conversion factor 1 ppm = 3.78 mg/m3 @ 25° C 

III. Major Uses and Sources 

Epichlorohydrin is a major raw material used in the manufacture of epoxy and phenoxy resins. It 

is also used as a solvent and in the synthesis of glycerol. Other uses include that of insect 

fumigation and as a chemical intermediate for the formation of glycidyl acrylate derivatives such 

as those used in the formation of eyeglass lenses (HSDB, 1994). The annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California based on the 

most recent inventory were estimated to be 4841 pounds of epichlorohydrin (CARB, 2000).

IV. Effects of Exposures to Humans

Studies of male reproductive function have shown no evidence of decreased sperm counts in 

populations occupationally exposed to epichlorohydrin (Milby et al., 1981). 

Appendix D3 196 Epichlorohydrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

V. Effects of Exposures in Animals 

Rats were exposed for 136 weeks (6 hours/day, 5 days/week) to 0, 10, 30, or 100 ppm (0, 38, 

113, or 380 mg/m3

) epichlorohydrin (Laskin et al., 1980). Kidney damage in the form of renal 

tubular degeneration and dilatation was observed in rats exposed to 30 ppm or greater. The 

observation of severe inflammation in the nasal passages of 90% of the control animals, as well 

as in the treated animals, prevented comparison of this effect between the two groups. 

A subchronic exposure of rats to 9, 17, 27, 56, or 120 ppm (34, 64, 102, 212, or 454 mg/m3) for 6 

hours/day, 5 days/week for 11-19 exposures showed evidence of extrarespiratory effects. These 

included liver congestion and necrosis and tubular atrophy in the kidneys at the highest 

concentration (Gage, 1959). Lethargy and weight loss were observed at 56 ppm. 

A study on the effects of epichlorohydrin exposure for 10 weeks (6 hours/day, 5 days/week) on 

male and female fertility in rats and rabbits showed that male rats, exposed to 50 ppm (189 

mg/m3

), were significantly less fertile than controls, as measured by successful matings to 

unexposed females (John et al., 1979; 1983a). No histological changes were observed in the 

testes of the male rats at the end of exposure. No significant effects on fertility occurred in the 

exposed female rats. Degenerative changes in the nasal epithelium were observed in the female 

rats exposed to 25 ppm (94.5 mg/m3), and in both sexes at 50 ppm. 

A teratology study was carried out in rats and rabbits exposed to 0, 2.5, or 25 ppm (0, 9.5, or 95 

mg/m3

) epichlorohydrin 7 hours/day during the critical days of gestation. There were no 

significant differences between controls and treated animals in the incidence of developmental 

defects, in maternal toxicity, or in histopathology of the lungs, nasal turbinates, or trachea (John 

et al., 1983b). 

Mice and rats (10/sex/concentration/strain) were exposed to 0, 5, 25, or 50 ppm (0, 19, 95, or 190 

mg/m3) epichlorohydrin for 6 hours/day, 5 days/week for 90 days (Quast et al., 1979). Animals 

were observed for clinical signs of toxicity and were measured biweekly for body weight 

changes. Body weight measurements, clinical chemistry, hematology, and urinalysis were 

conducted. Gross and histopathological examinations were performed at the end of the 

experiment. Exposures of rats to 25 and 50 ppm epichlorohydrin resulted in inflammation, focal 

erosions, hyperplasia, and metaplasia in the nasal turbinates. No adverse effects were observed 

in rats exposed to 5 ppm (19 mg/m3

). Mice similarly showed focal erosion, hyperplasia and 

metaplasia in the epithelium of the nasal turbinates when exposed to 25 ppm epichlorohydrin or 

greater. 

Appendix D3 197 Epichlorohydrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VI. Derivation of Chronic Reference Exposure Level 

Study Quast et al. (1979) 

Study population Rats and mice (10 per sex per concentration) 

Exposure method Discontinuous whole-body inhalation 

Critical effects Inflammation, focal erosions, hyperplasia, and 

metaplasia in the nasal turbinates 

LOAEL 25 ppm (94.5 mg/m3

) 

NOAEL 5 ppm (19 mg/m3

) 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 90 days 

Average experimental exposure 0.89 ppm (5 x 6/24 x 5/7) 

Human equivalent concentration 0.083 ppm (gas with extrathoracic respiratory 

effects, RGDR = 0.093, based on MVa = 0.14 

m3

/day , MVh = 20 m3/day, SAa(ET) = 15 cm2), 

SAh(ET) = 200 cm2

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.0008 ppm (0.8 ppb; 0.003 mg/m3

; 3 μg/m3) 

The U.S. EPA (1994) based its RfC of 1 μg/m3 on the same study but used a subchronic UF of 

10 for a 90 day study instead of 3 (OEHHA, 2000). 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for epichlorohydrin include the availability of controlled 

exposure inhalation studies in multiple species at multiple exposure concentrations and with 

adequate histopathogical analysis and the observation of a NOAEL. Major areas of uncertainty 

are the lack of adequate human exposure data, the lack of chronic inhalation exposure studies, 

the limited reproductive toxicity data, and the small groups tested in the study. 

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Gage JC. 1959. The toxicity of epichlorohydrin vapour. Br. J. Ind. Med. 16:11-14. 

Appendix D3 198 Epichlorohydrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

HSDB. 1997. Hazardous Substances Data Bank. TOMES® Vol. 33. Denver, CO: Micromedex, 

Inc. (edition expires 7/31/97) 

John JA, Quast JF, Murray FJ, et al. 1979. The effects of inhaled epichlorohydrin on the semen 

of rabbits and on the fertility of male and female rats. Toxicology Research Laboratory, Dow 

Chemical, Midland, MI. July 1979. NIEHS Contract No. NO1-ES-2102 (Unpublished). 

John JA, Gushow TS, Ayres JA, Hanley TR, Quast JF, and Rao KS. 1983a. Teratologic 

evaluation of inhaled epichlorohydrin and allyl chloride in rats and rabbits. Fundam. Appl. 

Toxicol. 3:437-442. 

John JA, Quast JF, Murray FJ, Calhoun LG, and Staples RE. 1983b. Inhalation toxicity of 

epichlorohydrin: effects on fertility in rats and rabbits. Toxicol. Appl. Pharmacol. 68:415-423. 

/ 

Laskin S, Sellakumar AR, Kuschner M, Nelson N, La Mendola S, Rusch GM, Katz GV, Dulak 

NC, and Albert RE. 1980. Inhalation carcinogenicity of epichlorohydrin in noninbred SpragueDawley rats. J. Natl. Cancer Inst. 65(4):751-757. 

Milby TH, Whorton MD, Stubbs HA, Ross CE, Joyner RE, and Lipshultz LI. 1981. Testicular 

function among epichlorohydrin workers. Br. J. Ind. Med. 38:372-377. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov

Quast JF, Henck JW, Postma BJ, Schuetz DJ, and McKenna MJ. 1979. Epichlorohydrin - 

subchronic studies. I. A 90-day inhalation study in laboratory rodents. Toxicology Research 

Laboratory, Dow Chemical U.S.A., Midland, MI (unpublished). 

U.S. EPA. 1994. U.S. Environmental Protection Agency. Reference concentration for 

epichlorohydrin. Integrated Risk Information System (IRIS), CD-ROM version. Denver, CO: 

Micromedex, Inc. (Edition expires 10/31/94). 

Appendix D3 199 Epichlorohydrin 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

1,2-EPOXYBUTANE 

(1-butene oxide; 1,2-butene oxide; 1,2-butylene oxide; 1,2-epoxybutane; 2-ethyloxirane; 

ethylethylene oxide; NCI-C55527) 

CAS Registry Number: 106-88-7 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 20 μg/m3 (6 ppb) 

Critical effect(s) Degenerative lesions of the nasal cavity in mice 

Hazard index target(s) Respiratory system; cardiovascular system 

II. Physical and Chemical Properties (HSDB, 1997) 

Description Colorless liquid with disagreeable odor 

Molecular formula C4H8O 

Molecular weight 72.12 g/mol 

Density 0.837 g/cm3

 @ 17°C 

Boiling point 63.3°C 

Melting point Not available (CRC, 1994) 

Vapor pressure 176 torr @ 25°C 

Solubility Soluble in ethanol, ether, acetone, water 

Odor threshold Unknown 

Conversion factor 1 ppm = 2.95 mg/m3

III. Major Uses or Sources 

1,2-Epoxybutane is used as a chemical intermediate, acid scavenger, and stabilizer for 

chlorinated solvents (Reprotext, 1994). It is highly reactive, flammable, and undergoes 

exothermic polymerization reactions in the presence of acids, bases, and some salts. It is less 

volatile than ethylene oxide or propylene oxide (Reprotext, 1994). The annual statewide 

industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in California 

based on the most recent inventory were estimated to be 6105 pounds of 1,2-epoxybutane 

(CARB, 2000). 

IV. Effects of Human Exposure 

No human toxicological data were found for 1,2-epoxybutane.

Appendix D3 200 Epoxybutane 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

V. Effects of Animal Exposure 

F344/N rats (50/sex) were exposed to 0, 200, or 400 ppm EBU for 6 hours/day, 5 days/week for 

2 years (NTP, 1988). Survival was impaired and concentration-related increases of 

inflammation, respiratory epithelial hyperplasia, olfactory sensory epithelial atrophy, and 

hyperostosis of the nasal turbinate bone cavity were observed in male and female rats exposed to 

either concentration. 

B6C3F1 mice (50/sex) were exposed to 0, 50, or 100 ppm EBU for 6 hours/day, 5 days/week for 

2 years (NTP, 1988). Survival and body weight gain were reduced significantly at 100 ppm in 

both sexes. Significant concentration-related increases in incidence of chronic inflammation, 

epithelial hyperplasia, and erosion of the nasal cavity were noted in both sexes at either 

concentration. Increases in granulocytic hyperplasia and splenic hematopoiesis were noted at 

both concentrations in female mice. 

Number of mice with lesions in the nasal cavity and olfactory sensory epithelium (NTP, 1988) 

Sex Males Females 

EBU concentration 0 ppm 50 ppm 100 ppm 0 ppm 50 ppm 100 ppm 

Number of mice studied 49 49 50 50 50 48 

Nasal cavity 

 Chronic inflammation 0 33 40 0 39 44 

 Erosion 0 7 17 0 16 24 

 Regeneration 0 15 17 0 14 15 

 Epithelial hyperplasia 0 32 45 1 34 35 

 Squamous metaplasia 1 24 41 0 34 41 

 Squam. cell papilloma 0 0 1 0 0 0 

Olfactory sensory 

 epithelium – atrophy 0 13 32 0 25 35 

Male and female mice exposed to 800 ppm (2360 mg/m3

) EBU for 6 hours/day, 5 days/week, for 

13 weeks were listless after the first exposure (NTP, 1988). Animals from this group all died by 

the end of the 13-week exposure. Renal tubular necrosis, and thymic and splenic atrophy were 

seen in mice exposed to 800 ppm; decreased liver weights were observed following exposure of 

mice to 400 ppm (1180 mg/m3

) or more. Inflammation of the nasal turbinates was seen in 

female mice exposed to 100 ppm (295 mg/m3) or more. No inflammation was observed in 

controls. 

Miller et al. (1981) exposed rats and mice of either sex to 0, 75, 150, or 600 ppm (0, 221, 442, or 

1770 mg/m3) EBU 6 hours/day, 5 days/week, for 13 weeks. In this study, no treatment-related 

effects were noted except for histological lesions in the nasal mucosal epithelium and reduced 

specific gravity in the urine of rats treated with 600 ppm. 

Wolf (1961) observed increased lung weights in rats exposed to 800 ppm of a mixture of 

epoxybutane isomers. No increase in lung weight was seen at 400 ppm. 

Appendix D3 201 Epoxybutane 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Sikov et al. (1981) conducted experiments to determine the reproductive toxicity of EBU in rats 

and rabbits. Rats were exposed to 0, 250, or 1000 ppm (0, 738, or 2950 mg/m3

) 1,2-epoxybutane 

for 7 hours/day, 5 days/week for 3 weeks prior to gestation, or for 7 hours/day on days 1-19 of 

gestation. Maternal toxicity in the form of 10% weight loss was observed in rats exposed to 

1000 ppm. One death out of 42 occurred in the dams exposed to 1000 ppm. No adverse 

histological, reproductive, or developmental effects were seen at any concentration. Exposure of 

rabbits on days 1-24 of gestation to the same concentrations as in the rat experiment showed 

more severe effects at lower concentrations than those observed in rats. In the rabbits, 6 out of 

48 dams died during exposure to 250 ppm, and 14 out of 24 died at 1000 ppm. Extensive 

maternal mortality in this study prevented evaluation of the reproductive and developmental 

effects. 

VI. Derivation of Chronic Reference Exposure Level 

Study National Toxicology Program (NTP, 1988) 

Study population Rats and mice 

Exposure method Discontinuous inhalation to 0, 50, or 100 ppm 

EBU 

Critical effects Damage to the upper respiratory epithelium was 

observed in both species at all concentrations. 

Mice also showed an increased incidence of 

granulocytic hyperplasia and splenic 

hematopoiesis at both concentrations, possibly 

due to inflammation in the upper respiratory 

tract. 

LOAEL 50 ppm (mice) 

NOAEL Not observed 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 2 years 

Average experimental exposure 8.9 ppm for LOAEL group (50 x 6/24 x 5/7) 

Human equivalent concentration 1.8 ppm for LOAEL group (gas with extrathoracic 

respiratory effects, RGDR = 0.20, based on 

MVa = 0.06 m3

/day, MVh = 20 m3/day, 

SAa(ET) = 3.0 cm2, SAh(ET) = 200 cm2) 

LOAEL uncertainty factor 10 (high incidence of adverse effects) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.006 ppm (6 ppb; 0.02 mg/m3

; 20 μg/m3) 

The chronic REL is also the U.S. EPA RfC (U.S. EPA, 1994). OEHHA staff reviewed and 

agreed with U.S. EPA’s analysis of the data. 

Appendix D3 202 Epoxybutane 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for 1,2-epoxybutane include the availability of controlled 

exposure inhalation studies in multiple species at multiple exposure concentrations and with 

adequate histopathological analysis. Major areas of uncertainty are the lack of adequate human 

exposure data and the lack of observation of a NOAEL in the key study. 

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

HSDB. 1997. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (edition expires 7/31/97). 

Miller RR, Quast JF, Ayres JA, and McKenna MJ. 1981. Inhalation toxicity of butylene oxide. 

Fundam. Appl. Toxicol. 1(4):319-324. [as cited in: Integrated Risk Information System (IRIS). 

U.S. Environmental Protection Agency, Washington, DC. (CD-ROM version). Denver, CO: 

Micromedex, Inc. (Edition expires 10/31/94).] 

NTP. 1988. National Toxicology Program. Technical Report Series No. 329. Toxicology and 

Carcinogenesis Studies of 1,2-Epoxybutane (CAS No. 106-88-7) in F344 Rats and B6C3Fl 

Mice. U.S. Department of Health and Human Services, Public Health Service, National Institutes 

of Health. Research Triangle Park, NC: NTP. 

REPROTOX. 1994. The REPROTOX(R) System: Reproductive Reviews of Drugs, Chemicals, 

Physical and Biological Agents. (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. 

Sikov MR, Cannon WC, Carr DB, Miller RA, Montgomery LF, and Phelps DW. 1981. 

Teratologic assessment of butylene oxide, styrene oxide, and methyl bromide. Division of 

Biomedical and Behavioral Science, NIOSH. NIOSH/000993 14. NIOSH Technical Report No. 

81-124. 

U.S. EPA. 1994. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration (RfC) for 1,2-epoxybutane. Available online at 

http://www.epa.gov/ngispgm3/iris 

Wolf MA. 1961. Results of repeated exposures of laboratory animals to the vapors of butylene 

oxide(s) (Mixed isomers). Dow Chemical Biochemical Research Department. EPA/OTS 

Document No. 878211232. [as cited in: Integrated Risk Information System (IRIS). U.S. 

Environmental Protection Agency, Washington, D.C. (CD ROM version). Denver, CO: 

Micromedex, Inc. (Edition expires 10/31/94] 

Appendix D3 203 Epoxybutane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

ETHYL CHLORIDE 

(Chloroethane; monochloroethane; ether hydrochloric) 

CAS Registry Number: 75-00-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 30,000 μg/m3

 (10,000 ppb) 

Critical effect(s) Delayed fetal ossification in mice 

Hazard index target(s) Teratogenicity; alimentary system 

II. Physical and Chemical Properties (HSDB, 1995; 1999)

Description Colorless gas 

Molecular formula C2H5Cl 

Molecular weight 64.52 

Density 0.9214 g/cm3

 @ 0°C 

Boiling point 12.3 °C 

Melting point -138.7 °C 

Vapor pressure 1000 torr @ 20 °C 

Conversion factor 1 ppm = 2.64 mg/m3

 @ 25°C 

III. Major Uses or Sources 

Ethyl chloride has been used as a starting point in the production of tetraethyl lead and as a 

refrigerant, solvent and alkylating agent (HSDB, 1995). It is also used as a topical anesthetic 

(Clayton and Clayton, 1994). The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 291,300 pounds of ethyl chloride (CARB, 1999).

IV. Effects of Human Exposure 

Neurological symptoms have been observed in human case studies in instances of ethyl chloride 

abuse. Cerebellar-related symptoms including ataxia, tremors, speech difficulties, and 

hallucinations were observed in a 28-year old female who had sniffed 200-300 ml ethyl chloride 

off her sleeve daily for 4 months (Hes et al. 1979). The patient’s liver was enlarged and tender. 

Four weeks following cessation of exposure, all symptoms were absent.

Appendix D3 204 Ethyl chloride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

V. Effects of Animal Exposure 

Pregnant mice were exposed to 1300, 4000, or 13000 mg/m3 ethyl chloride in air for 6 hours per 

day on days 6-15 of gestation (Scortichini et al., 1986). No effects on fetal resorption rates, litter 

size, body weight or maternal health were observed. A statistically significant increase in the 

incidence of delayed ossification of the skull bones was observed in fetuses from the 

13,000 mg/m3 (4900 ppm) ethyl chloride exposed group. This skull effect was accompanied by 

a non-significant increased incidence of cervical ribs (a supernumerary rib is considered to be a 

malformation). No significant adverse effects were observed in fetuses from the 4000 mg/m3

(1500 ppm) exposure group. 

No significant adverse effects were observed in rats and mice exposed to 0 or 15,000 ppm ethyl 

chloride for 6 hours per day, 5 days per week for 102 weeks (rats) or 100 weeks (mice) (NTP, 

1989). At necropsy, a complete histopathologic examination (approximately 35 tissues) failed to 

identify evidence of non-cancer toxicity. The same study also exposed rats and mice to 2500, 

5000, 10,000 or 19,000 ppm ethyl chloride 6 hours per day, 5 days per week for 13 weeks. No 

exposure-related clinical signs of toxicity or histological changes were observed in exposed 

animals. Thus the subchronic NOAEL for mice and rats is 19,000 ppm, which is equivalent to a 

continuous exposure of 3400 ppm, and a free-standing chronic NOAEL is 15,000 ppm, which is 

equivalent to a continuous exposure of 2700 ppm (7100 mg/m3

). 

Increased relative liver weights and a slight increase in hepatocellular vacuolation were observed 

in mice exposed to 5000 ppm ethyl chloride 23 hours per day for 11 days (Landry et al., 1989). 

No effects were observed in mice exposed to 0, 250, or 1250 ppm ethyl chloride for the same 

period. 

Following acclimatization to an inhalation chamber, two groups of 10 female mice were exposed 

to 0 or 15,000 ppm (40,000 mg/m3

) ethyl chloride 6 hours per day for 2 weeks (Breslin et al., 

1988). Groups of five male mice were housed in each inhalation chamber to synchronize and 

promote regular cyclicity. The mean length of the estrous cycle in control mice remained 

constant at 4.5 days during both pre-exposure and exposure periods. Mice in the 15,000 ppm 

exposure group showed a 0.6 day increase in the mean cycle length during exposure (5.6 days) 

when compared to the pre-exposure period (5.0 days). The authors attribute this increase in 

estrous cycle length to a general stress response although they note that it does not preclude 

direct effects on neuroendocrine function. 

Appendix D3 205 Ethyl chloride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Reference Exposure Level 

Study Scortichini et al., 1986 

Study population Mice 

Exposure method Discontinuous whole-body inhalation (on days 

6-15 of gestation) 

Critical effects Delayed ossification of skull foramina 

LOAEL 13,000 mg/m3

NOAEL 4,000 mg/m3

Exposure continuity 6 hours per day 

Exposure duration Days 6-15 of gestation 

Average experimental exposure 1,000 mg/m3

 for NOAEL group (4000 x 6/24) 

Human equivalent concentration 1,000 mg/m3

 for NOAEL group (gas with 

systemic effects, based on RGDR = 1.0 using 

default assumption that lambda (a) = lambda 

(h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 30 mg/m3

 ( 30,000 μg/m3; 10 ppm; 10,000 

ppb) 

To develop the chronic REL OEHHA used the same study on which U.S. EPA based its RfC of 

10,000 μg/m3

. The REL is based on a developmental toxicity study. In accordance with U.S. 

EPA methodology, a time-weighted average concentration for the discontinuous exposure 

experiment is not used by U.S. EPA when the key effect is developmental toxicity. However, 

OEHHA prefers to make a time adjustment to equivalent continuous exposure because the 

chronic REL assumes continuous exposure. U.S. EPA also used a Modifying Factor (MF). The 

database deficiencies leading U.S. EPA to employ a modifying factor include the lack of a 

multigenerational reproductive study. The criteria for use of such modifying factors are not well 

described. Such MFs were not used by OEHHA. 

As a comparison to the proposed REL of 10 ppm, NTP (1989) found a free-standing NOAEL of 

15,000 ppm in rats and mice exposed to ethyl chloride for 6 hours per day, 5 days per week for 2 

years. Time adjusting to continuous exposure results in an adjusted NOAEL of 2679 ppm. 

Applying an RGDR of 1, a UFA of 3 and a UFH of 10 results in an estimated REL of 90 ppm. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for ethyl chloride include the availability of controlled 

exposure inhalation studies in multiple species at multiple exposure concentrations and with 

adequate histopathogical analysis, and the observation of a NOAEL. Major areas of uncertainty 

are the lack of adequate human exposure data, and the lack of a multigenerational reproductive 

study. 

Appendix D3 206 Ethyl chloride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VIII. References 

Breslin JW, Berdasco NM, Phillips JE, and Johnson KA. 1988. Ethyl Chloride (EtCl): Effects 

on Estrous Cycling in B6C3F1 Mice. Final Report with cover letter dated 11/21/88. Dow 

Chemical Company. EPA Document # 86-890000040. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Clayton GD, and Clayton FE. (eds.) 1994. Patty’s Industrial Hygiene and Toxicology. Vol II 

Part E. New York: John Wiley and Sons, Inc. pp. 4082-4087. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Micromedex, Inc., Denver, Colorado (Edition expires 1/31/95). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Hes JPh, Cohn DF, Streifler M. 1979. Ethyl chloride sniffing and cerebellar dysfunction (case 

report). Isr. Ann. Psychiatr. Relat. Discip. 17(2):122-125. [cited in U.S. EPA, 1995]. 

Landry TD, Johnson KA, Phillips JE, Weiss SK. 1989. Ethyl chloride: 11-Day continuous 

exposure inhalation toxicity study in B6C3F1 mice. Fundam. Appl. Toxicol. 13:516-522. 

NTP. 1989. National Toxicology Program. Toxicology and Carcinogenesis Studies of 

Chloroethane (Ethyl Chloride) (CAS No. 75-00-3) in F344/N Rats and B6C3F1 Mice. NTP, US 

Department of Health and Human Services, Public Health Service, National Institutes of Health. 

NTP Technical Report Number 346. 

Scortichini BH, Johnson KA, Momany-Pfruender JJ, Hanley TR. 1986. Ethyl Chloride: 

Inhalation Teratology Study in CF-1 Mice. Dow Chemical Company. EPA Document #86-

870002248. 

U.S. EPA (U. S. Environmental Protection Agency). 1995. Integrated Risk Information System 

(IRIS). Cincinnati, OH: Office of Health and Environmental Assessment, Environmental Criteria 

Assessment Office (CD-ROM Version). 

Appendix D3 207 Ethyl chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

ETHYLBENZENE 

(Phenylethane; NCI-C56393) 

CAS Registry Number: 100-41-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 2000 μg/m3

 (400 ppb) 

Critical effect(s) Liver, kidney, pituitary gland in mice and rats 

Hazard index target(s) Alimentary system (liver); kidney; endocrine 

system 

II. Physical and Chemical Properties (HSDB, 1994) 

Description colorless liquid 

Molecular formula C8H10

Molecular weight 106.16 g/mol 

Boiling point 136.2°C 

Melting point -95°C 

Vapor pressure 10 torr @ 25.9°C 

Density 0.867 g/cm3

 @ 20°C 

Solubility Soluble in ethanol and ether, low solubility in 

water (0.014 g/100 ml at 15°C) 

Conversion factor 1 ppm = 4.35 mg/m3

III. Major Uses or Sources 

Ethylbenzene is used as a precursor in the manufacture of styrene (HSDB, 1994). It is also used 

in the production of synthetic rubber, and is present in automobile and aviation fuels. It is found 

in commercial xylene (Reprotext, 1994). In 1996, the latest year tabulated, the statewide mean 

outdoor monitored concentration of ethylbenzene was approximately 0.4 ppb (CARB, 1999a). 

The latest annual statewide emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California, based on the most recent inventory, were estimated to be 161,846 pounds of 

ethylbenzene (CARB, 1999b). 

IV. Effects of Human Exposure 

Studies on the effects of workplace exposures to ethylbenzene have been complicated by 

concurrent exposures to other chemicals, such as xylenes (Angerer and Wulf, 1985). Bardodej 

Appendix D3 208 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

and Cirek (1988) reported no significant hematological or liver function changes in 200 

ethylbenzene production workers over a 20-year period. 

V. Effects of Animal Exposure 

Rats and mice (10/sex/group) were exposed to 0, 100, 250, 500, 750, and 1000 ppm (0, 434, 

1086, 2171, 3257, and 4343 mg/m3) ethylbenzene 6 hours/day, 5 days/week for 90 days (NTP, 

1988; 1989; 1990). Rats displayed significantly lower serum alkaline phosphatase in groups 

exposed to 500 ppm or higher. Dose-dependent increases in liver weights were observed in male 

rats beginning at 250 ppm, while this effect was not seen until 500 ppm in the females. An 

increase in relative kidney weights was seen in the 3 highest concentrations in both sexes. 

Minimal lung inflammation was observed in several of the treatment groups, but this 

phenomenon was attributed to the presence of an infectious agent rather than to ethylbenzene 

exposure. The mice in this study did not show any treatment-related effects except for elevated 

liver and kidney weights at 750 and 1000 ppm, respectively. 

Rats and mice were exposed to ethylbenzene (greater than 99% pure) by inhalation for 2 years 

(NTP, 1999; Chan et al., 1998). Groups of 50 male and 50 female F344/N rats were exposed to 

0, 75, 250, or 750 ppm, 6 hours per day, 5 days per week, for 104 weeks. Survival of male rats 

in the 750 ppm group was significantly less than that of the chamber controls. Mean body 

weights of 250 and 750 ppm males were generally less than those of the chamber controls 

beginning at week 20. Mean body weights of exposed groups of females were generally less 

than those of chamber controls during the second year of the study. In addition to renal tumors, 

the incidence of renal tubule hyperplasia in 750 ppm males was significantly greater than that in 

the chamber controls. The severity of nephropathy in 750 ppm male rats was significantly 

increased relative to the chamber controls. Some increases in incidence and severity of 

nephropathy were noted in all exposed female rats, but these were statistically significant only at 

750 ppm. 

Groups of 50 male and 50 female B6C3F1 mice were exposed to 0, 75, 250, or 750 ppm 

ethylbenzene by inhalation, 6 hours per day, 5 days per week, for 103 weeks. Survival of 

exposed mice was similar to controls. Mean body weights of females exposed to 75 ppm were 

greater than those of the chamber controls from week 72 until the end of the study. In addition to 

lung and liver tumors, the incidence of eosinophilic liver foci in 750 ppm females was 

significantly increased compared to that in the chamber controls. There was a spectrum of 

nonneoplastic liver changes related to ethylbenzene exposure in male mice, including syncytial 

alteration of hepatocytes, hepatocellular hypertrophy, and hepatocyte necrosis. The incidences 

of hyperplasia of the pituitary gland pars distalis in 250 and 750 ppm females and the incidences 

of thyroid gland follicular cell hyperplasia in 750 ppm males and females were significantly 

increased compared to those in the chamber control groups. Based on an evaluation of all the 

non-cancer data in mice and rats OEHHA staff selected 75 ppm as the NOAEL for the NTP 

(1999) study. 

Rats (17-20 per group) were exposed to 0, 600, 1200, or 2400 mg/m3 for 24 hours/day on days 7 

to 15 of gestation (Ungvary and Tatrai, 1985). Developmental malformations in the form of 

“anomalies of the uropoietic apparatus” were observed at the 2400 mg/m3 concentration. 

Appendix D3 209 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Skeletal retardation was observed in all exposed groups compared with controls. The incidence 

of skeletal abnormalities increased with higher concentrations of ethylbenzene. 

Rabbits exposed by these investigators to the same concentrations as the rats on days 7 to 15 of 

gestation, exhibited maternal weight loss with exposure to 1000 mg/m3 ethylbenzene. There 

were no live fetuses in this group for which abnormalities could be evaluated. No developmental 

defects were observed in the lower exposure groups. 

Rats (78-107 per group) and rabbits (29-30 per group) were exposed for 6 or 7 hours/day, 7 

days/week, during days 1-l9 and 1-24 of gestation, respectively, to 0, 100, or 1000 ppm (0, 434, 

or 4342 mg/m3) ethylbenzene (Andrew et al., 1981; Hardin et al., 1981). No effects were 

observed in the rabbits for maternal toxicity during exposure or at time of necropsy. Similarly, 

no effects were seen in the fetuses of the rabbits. The only significant effect of ethylbenzene 

exposure in the rabbits was a reduced number of live kits in the 1000 ppm group. A greater 

number and severity of effects were seen in rats exposed to 1000 ppm ethylbenzene. Maternal 

rats exposed to 1000 ppm exhibited significantly increased liver, kidney, and spleen weights 

compared with controls. Fetal rats showed an increase in skeletal variations at the 1000 ppm 

concentration, but the results of the 100 ppm exposure were not conclusive. 

Clark (1983) found no significant effects on body weight, food intake, hematology, urinalysis, 

organ weights or histopathology in rats (18 per group) exposed to 100 ppm (434 mg/m3

) 

ethylbenzene for 6 hours/day, 5 days/week, for 12 weeks. 

Degeneration of the testicular epithelium was noted in guinea pigs and a rhesus monkey exposed 

to 600 ppm (2604 mg/m3) for 6 months (Wolf et al., 1956). No effects were reported for female 

monkeys exposed to the same conditions. 

Cragg et al. (1989) exposed mice and rats (5/sex/group) to 0, 99, 382, and 782 ppm (0, 430, 

1659, and 3396 mg/m3) 6 hours/day, 5 days/week for 4 weeks. Some evidence of increased 

salivation and lacrimation was seen in the rats exposed to 382 ppm. No other gross signs of 

toxicity were observed. Both male and female rats had significantly enlarged livers following 

exposure to 782 ppm. Female mice also showed a significant increase in liver weight at this 

concentration. No histopathological lesions were seen in the livers of these mice. 

Dose-dependent induction of liver cytochrome P450 enzymes in rats by ethylbenzene was 

observed by Elovaara et al. (1985). Rats (5 per group) were exposed to 0, 50, 300, or 600 ppm 

(0, 217, 1302, or 2604 mg/m3) ethylbenzene for 6 hours/day, 5 days/week for 2, 5, 9, or 16 

weeks. Cytochrome P450 enzyme induction, and microscopic changes in endoplasmic reticulum 

and cellular ultrastructure were evident at all ethylbenzene concentrations by week 2, and 

persisted throughout the exposure. Liver weights were not elevated in these studies. 

Appendix D3 210 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of the Chronic Reference Exposure Level 

Study NTP, 1999; Chan et al., 1998 

Study population Male and female rats and mice (50 per group) 

Exposure method Discontinuous inhalation 

Critical effects Nephrotoxicity, body weight reduction (rats) 

hyperplasia of the pituitary gland; liver cellular 

alterations and necrosis (mice) 

LOAEL 250 ppm 

NOAEL 75 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 103 weeks. 

Average experimental exposure 13 ppm for NOAEL group 

Human equivalent concentration 13 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.4 ppm (400 ppb; 2 mg/m3

; 2,000 μg/m3) 

The REL is based on a lifetime toxicity/carcinogenesis study. The NOAEL for non-neoplastic 

effects in the study was 75 ppm, and the LOAEL was 250 ppm. Some shorter duration studies 

discussed above (e.g. NTP, 1988, 1989, 1990) identify higher concentrations as NOAELs, but 

the study used (NTP 1999) is the most recent available and is considered the most reliable for 

assessing chronic effects. 

U.S. EPA based its RfC on developmental toxicity studies in rats and rabbits (Andrew et al., 

1981; Hardin et al., 1981; U.S. EPA, 1994). The NOAEL in the studies was 100 ppm, and the 

LOAEL was 1000 ppm. In accordance with its methodology, U.S. EPA did not use a timeweighted average concentration for the discontinuous exposure experiment since the key effect 

was developmental toxicity. If OEHHA methodology is followed (which includes the timeweighted averaging of the exposure concentrations, and uncertainty factors of 3 (interspecies, 

with RGDR = 1) and 10 (intraspecies), this study would indicate a REL of 0.6 ppm (3 mg/m3

). 

The study by Ungvary and Tatrai (1985) reported a NOAEL of 600 mg/m3

 for developmental 

and maternal effects in several species. However, the reporting and general quality of this paper 

create less confidence in its results. 

For comparison to the proposed REL of 0.4 ppm, Clark (1983) found no significant effects in 

rats exposed to 100 ppm ethylbenzene 6 h/day, 5 d/week, for 12 weeks. This NOAEL can be 

time-adjusted to 18 ppm, then divided by a subchronic UF of 3, an interspecies UF of 3, and an 

intraspecies UF of 10 which results in a REL of 0.2 ppm. (The default value of 1 for RGDR was 

used). It appears that the proposed REL provides a sufficient margin of safety to provide 

Appendix D3 211 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

protection against the reported developmental effects (Andrew et al., 1981; Hardin et al., 1981; 

Ungvary and Tatrai, 1985) 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for ethylbenzene include the availability of controlled 

exposure inhalation studies in multiple species at multiple exposure concentrations and with 

adequate histopathogical analysis, and the observation of a NOAEL in lifetime chronic 

inhalation exposure studies. The major area of uncertainty is the lack of adequate human 

exposure data. 

VIII. References 

Andrew FD, Buschbom RL, Cannon WC, Miller RA, Montgomery LF, Phelps DW, et al. 1981. 

Teratologic assessment of ethylbenzene and 2-ethoxyethanol. Battelle Pacific Northwest 

Laboratory, Richland, WA. PB 83-208074, p. 108. [as cited in: U.S.EPA’s Integrated Risk 

Information System (IRIS) database. 1994.] 

Angerer J, and Wulf H. 1985. Occupational chronic exposure to organic solvents. XI. 

Alkylbenzene exposure of varnish workers: Effects on hematopoietic system. Int. Arch. Occup. 

Environ. Health. 56(4):307-321. [as cited in U.S.EPA’s Integrated Risk Information System 

(IRIS) database. 1994.] 

Bardodej Z, and Cirek A. 1988. Long-term study on workers occupationally exposed to 

ethylbenzene. J. Hyg. Epidemiol. Microbiol. Immunol. 32(1):1-5. [As cited in U.S.EPA’s 

Integrated Risk Information System (IRIS) database. 1994.] 

CARB. 1999a. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Ethyl Benzene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm 

CARB. 1999b. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Clark DG. 1983. Ethylbenzene hydroperoxide (EBHP) and ethyl benzene (EB): 12-week 

inhalation study in rats. (Group research report with attachments and cover sheet.) EPA OTS 

Public Files. Shell Oil Co. Document No. 86870001629. Fiche Number 0516206. [as cited in 

U.S.EPA’s Integrated Risk Information System (IRIS) database. 1994.] 

Chan PC, Haseman JK, Mahleri J, Aranyi C. 1998. Tumor induction in F344/N rats and B6C3F1 

mice following inhalation exposure to ethylbenzene. Toxicol. Lett. 99(1):23-32. 

Cragg ST, Clarke EA, Daly IW, Miller RR, Terrill JB, and Quellette RE. 1989. Subchronic 

inhalation toxicity of ethylbenzene in mice, rats, and rabbits. Fundam. Appl. Toxicol. 13(3):399-

408. [as cited in U.S.EPA’s Integrated Risk Information System (IRIS) database. 1994.] 

Appendix D3 212 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Elovaara E, Engstrom K, Nickels J, Aito A, and Vainio H. 1985. Biochemical and morphological 

effects of long-term inhalation exposure of rats to ethylbenzene. Xenobiotica. 15(4):299-308. [as 

cited in U.S.EPA’s Integrated Risk Information System (IRIS) database. 1994.] 

Hardin BD, Bond GP, Sikov MR, Andrew FD, Beliles RP, and Niemeier RW. 1981. Testing of 

selected workplace chemicals for teratogenic potential. Scand. J. Work Environ. Health. 7(suppl 

4): 66-75. 

HSDB. 1994. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(CD-ROM version) Denver, CO: Micromedex, Inc. (Edition expires 4/30/94). 

NTP. 1988. National Toxicology Program. Subchronic and chronic toxicity study of 

ethylbenzene. 90-Day subchronic study report on inhalation exposure of F344/N rats and 

B6C3F1 mice. Chicago, IL: IIT Research Institute. [as cited in U.S.EPA’s Integrated Risk 

Information System (IRIS) database. 1994.] 

NTP. 1989. National Toxicology Program. Chairperson’s report. Pathology Working Group 

(PWG) review of subchronic toxicity testing on ethylbenzene administered by inhalation in F344 

rats and B6C3FI mice. [as cited in U.S.EPA’s Integrated Risk Information System (IRIS) 

database. 1994.] 

NTP. 1990. National Toxicology Program. Draft NTP Technical Report on the Toxicity Studies 

of Ethylbenzene in F344 Rats and B6C3F1 Mice (inhalation Studies). NTP TOX 10, U.S. 

DHHS. [as cited in U.S.EPA’s Integrated Risk Information System (IRIS) database. 1994.] 

NTP. 1999. National Toxicology Program. Toxicology and Carcinogenesis Studies of 

Ethylbenzene (CAS No. 100-41-4) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). TR466. 

Reprotext ® database. (CD-ROM version) Denver, CO: Micromedex, Inc. (Edition expires 

4/30/94). 

Ungvary G, and Tatrai E. 1985. On the embryotoxic effects of benzene and its alkyl derivatives 

in mice, rats, and rabbits. Arch. Toxicol. Suppl. 8:425-430. 

U.S.EPA. 1994. Ethylbenzene. Integated Risk Information System (IRIS) on-line database. 

Wolf MA, Rowe VKL, McCollister DD, Hollingsworth RL, and Oyen F. 1956. Toxicological 

studies of certain alkylated benzenes and benzene. Arch. Ind. Health 14: 387-398. 

Appendix D3 213 Ethylbenzene 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

ETHYLENE DIBROMIDE 

(1,2-dibromoethane; dibromoethane; alpha, beta-dibromoethane; EDB; ethylene bromide; 

glycol bromide) 

CAS Registry Number: 106-93-4 

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.8 μg/m3 (0.1 ppb)

Critical effect(s) Decreased sperm count/ejaculate, decreased 

percentage of viable and motile sperm, 

increased semen pH, and increased 

proportion of sperm with specific 

morphological abnormalities in human 

males 

Hazard index target(s) Reproductive system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994) 

Description Colorless, heavy, nonflammable liquid with a 

mildly sweet, chloroform-like odor. 

Molecular formula C2H4Br2

Molecular weight 187.88 g/mol 

Boiling point 131-132°C 

Melting point 9.9°C 

Vapor pressure 0.11 torr at 20°C 

Solubility Slightly soluble in water (3400 mg/L water at 

20°C). Miscible with most organic solvents. 

Conversion factor 7.68 μg/m3

 per ppb at 25°C 

III. Major Uses and Sources

Ethylene dibromide (EDB) is used as a solvent for resins, gums, and waxes, and as a chemical 

intermediate in the synthesis of dyes and pharmaceuticals (HSDB, 1995). EDB was once widely 

used as a fumigant for the control of pests in the U.S. Because of concerns regarding its 

carcinogenicity, the agricultural uses of EDB were banned in 1983 (RECT, 1988). EDB was 

also commonly used as a gasoline additive to scavenge inorganic lead compounds. The 

transition to the use of lead-free gasoline has drastically curtailed the use of EDB in this country 

(REPROTOX, 1995). EDB is now used mainly in industry. EDB may be formed naturally in 

Appendix D3 214 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

the ocean as a result of macro algae growth. Exposure to the general population, via inhalation, 

may occur in the vicinity of industries and in industrial settings where this compound is 

manufactured and used. The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 1179 pounds of EDB (CARB, 2000).

IV. Effects of Human Exposures

Pharmacokinetic studies of EDB in humans could not be found in the literature. However, in 

vitro studies of EDB metabolism in human liver samples have been performed (Wiersma et al., 

1986). These experiments have shown that the enzyme systems known to metabolize EDB in 

rodent liver also metabolize EDB in the human liver. EDB was metabolized by human liver 

cytosolic glutathione S-transferases (GST), microsomal GST, and microsomal mixed function 

oxidases (MFO). MFO activity resulted in adducts irreversibly bound to protein, while GST 

activity was mostly responsible for adducts irreversibly bound to DNA. Rodent liver enzymes 

similarly activate EDB to metabolites that bind to cellular macromolecules. In human fetal liver 

(16-18 weeks gestation) cytosolic GST was also found to metabolize EDB with high efficiency 

(Kulkarni et al., 1992). Since detoxification via MFO activity may be limited at this stage of 

development, the results suggest that the human fetus and neonate may be at greater risk from 

EDB toxicity than adults. 

A study of mortality from cancer and respiratory diseases was conducted among 161 employees 

exposed to EDB in 2 production units operated from 1942 to 1969 and from the mid-1920s to 

1976, respectively (Ott et al., 1980). No apparent connection was found between mortality due 

to respiratory diseases and exposure to EDB, when compared to U.S. white male mortality 

figures. 

Due to the structural similarity of EDB to dibromochloropropane (DBCP), a known toxic agent 

in human male reproductive organs, a number of epidemiological studies concerning male 

reproduction and spermatogenesis were conducted. 

In a study of 59 employees exposed to EDB at the Ethyl Corporation plant in Magnolia, 

Arkansas, the sperm counts of the exposed men were divided into 2 groups depending on 

estimated exposure (Ter Haar, 1980). Twenty percent of the low exposure group (<0.5 ppm) had 

sperm counts below 40 million, whereas 42% of the high exposure group (0.5 to 5 ppm) had 

sperm counts below this figure. The sperm counts were intermediate between counts reported 

for 2 types of U.S. samples (for normal men). The observed births among the two exposure 

groups were found to be similar to the number of expected births. The author determined that 

EDB had no effect on sterility or reproduction in the workers. Weaknesses of this study include 

the small population of exposed workers and the lack of a concurrent unexposed control group. 

Taking these defects of the study into account, Dobbins (1987) concluded that the results provide 

evidence that EDB exposure between 0.5 and 5.0 ppm is associated with lower sperm counts. 

A comparison of observed marital fertility with expected fertility (based on U.S. fertility rates) 

was conducted among 297 men working at 4 U.S. plants that manufacture EDB (Wong et al., 

Appendix D3 215 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

1979). Fertility was 20% below expected for the four plants combined. This was largely due to 

one plant (plant D), which was 49% below the expected level. After omitting the incidence of 

vasectomies and hysterectomies among married couples, observed fertility was still 39% below 

the expected figure for plant D but was now no longer statistically significant. Exposure levels 

of EDB at plant D were not known but were estimated to be no more than 5 ppm. Later review 

determined that expected (control) levels of fertility and the power of the study were too low, 

resulting in the inability to identify a possible adverse effect (Dobbins, 1987). The lower fertility 

at plant D indicates that EDB has the potential to reduce fertility, but the extent of the reduction 

cannot be estimated from this study. Further treatment of the data by a method that uses the 

proper statistical adjustments of reproductive experience in the U.S. population (used as the 

control) suggests borderline significance for reduced fertility among the combined workers at the 

four plants (Wong et al., 1985). The fertility evaluation indicates that more in-depth 

epidemiologic or physiologic studies are needed. 

Semen analysis of 83 pineapple workers at two plantations was performed by Rogers and 

associates (1981). EDB-exposed workers were removed from each group and placed in a 

separate group. The remaining two groups of workers acted as control groups. Sperm counts, 

motility, and morphology were similar among the three groups. However, 43.8% of exposed 

workers had abnormally low counts (<40 million/ml), while abnormally low sperm counts of 

controls were 34.2% and 17.8%. Of the four exposed workers that had fertility tests done, all 

tested in the infertile range. Forty percent or less tested in the infertile range among the control 

groups. The results suggest that workers exposed to EDB had reduced sperm counts, but 

exposure levels were not known. 

Semen analysis among 46 men employed in the papaya fumigation industry was conducted to 

determine if EDB affected semen quality (Ratcliff et al., 1987; Schrader et al., 1987). Average 

duration of exposure was 5 years and the geometric mean breathing zone exposure to airborne 

EDB was 88 ppb (8 hr time weighted average) with peak exposures of up to 262 ppb. The 

comparison group consisted of 43 unexposed men from a nearby sugar refinery. Following 

consideration of confounding factors, statistically significant decreases in sperm count/ejaculate, 

the percentage of viable and motile sperm, and increases in the proportion of sperm with specific 

morphological abnormalities (tapered heads, absent heads, and abnormal tails) were observed 

among exposed men. Semen pH was significantly more alkaline than that of unexposed workers. 

Other measured sperm quality parameters were unchanged. This study suggests that EDB can 

result in reproductive impairment. However, no measurement of male fertility was conducted. 

In a study that examined similar indices of semen quality, 6 week exposure of 10 forestry 

workers to EDB (60 ppb time weighted average, with peak exposures of up to 2165 ppb) resulted 

in decreased semen volume and slower sperm velocity (Schrader et al., 1988). Six unexposed 

men were used as controls. The researchers suggest that short-term exposure to EDB results in 

decreased sperm velocity, while long-term exposure, as in the previous study of EDB-exposed 

papaya workers, results in sperm immotility and cell death. 

Appendix D3 216 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

V. Effects of Animal Exposures

EDB is readily and rapidly absorbed from the lung when breathed as a vapor, from the GI tract 

when taken orally, or through the skin when applied dermally (HSDB, 1995). In rats, the rate of 

absorption of EDB from the respiratory tract reached a plateau within 10 to 20 minutes following 

exposure to 75 ppm EDB for up to 2 hours (Stott and McKenna, 1984). About 58% of the EDB 

was absorbed. Intraperitoneal injection of [14C]EDB into guinea pigs resulted in the highest 

concentrations in liver, kidneys, and adrenals (Plotnick and Conner, 1976). Sixty-five percent of 

the dose was excreted as metabolites in urine, 3% in feces, and 12% excreted unchanged in 

expired air. In rats, the highest concentrations of [14C]EDB label were found in liver, kidney and 

spleen following an oral dose of 15 mg/kg body wt (Plotnick et al., 1979). Studies with rats have 

provided evidence that 2 pathways of metabolic bioactivation exist for EDB (RECT, 1988). The 

oxidative pathway yields the metabolite 2-bromo-acetaldehyde, which is associated with cell 

macromolecule binding and liver damage. The conjugative pathway principally yields 

glutathione products, such as S-(2-bromoethyl)-glutathione, which are mainly responsible for 

DNA binding and mutagenesis. In rats, orally administered EDB is excreted primarily in the 

urine as mercapturic acid derivatives (Jones and Edwards, 1968). The biologic half-life for 

elimination of [14C]EDB in rats is 5.1-5.6 hours (Watanabe et al., 1978) and less than 48 hours in 

mice and guinea pigs (HSDB, 1995). Besides the small amount irreversibly bound to cell 

macromolecules and DNA, EDB shows little, if any, bioaccumulation in mammalian systems. 

In a subchronic toxicity study of experimental animals, rats and guinea pigs were given EDB by 

oral administration for about 4 months (Aman et al., 1946). Body weights and mortality of 

animals at or below an average daily dose of 40-50 mg/kg body wt-day were unaffected. 

However, only one control animal/species was used, the dosing regimen was not well described, 

and pathologic examination was apparently not performed. 

Subchronic exposure of rats (20/sex/group) to 50 ppm EDB for as many as 63 seven-hour 

exposures in 91 days resulted in no significant change in body weights (Rowe et al., 1952). 

Liver and kidney weights were increased in both sexes while testis weights were decreased in 

males. Also, lung weights in males were elevated and spleen weights in females were decreased. 

Histopathological examination revealed no changes. Guinea pigs (8/sex/group) subjected to as 

many as 57 seven-hour exposures of 50 ppm EDB in 80 days exhibited reduced body weights. 

Organ weights were unchanged, but microscopic examination of the livers showed slight central 

fatty degeneration. In kidneys, slight interstitial congestion and edema with slight 

parenchymatous degeneration of the tubular epithelium were observed. Four rabbits exposed to 

59 seven-hour sessions at 50 ppm in 84 days showed no signs of adverse effects. Clinical signs 

of monkeys exposed to 50 ppm EDB (49 seven-hour exposures in 70 days) included an ill, 

unkempt appearance and nervousness. Slight central fatty degeneration in livers was observed, 

but pathology was not seen in other tissues. Exposure of the same four species to 25 ppm EDB 

for up to 220 days (145 to 156 seven-hour exposures) showed no signs of adverse effects. 

In a 13-week inhalation study, 5 Fischer 344 albino rats/group/sex and 10 B6C3F1 

mice/group/sex were exposed to 0, 3, 15, or 75 ppm EDB for 6 hr/day, 5 days/week (Reznik et 

al., 1980). At 75 ppm, rats and mice exhibited severe necrosis and atrophy of the olfactory 

epithelium in the nasal cavity. Squamous metaplasia, hyperplasia and cytomegaly of the 

Appendix D3 217 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

epithelium were also seen in nasal turbinates, larynx, trachea, bronchi, and bronchioles. Minor 

alterations were seen in the nasal cavity of only a few male and female rats at 15 ppm. No 

compound-related lesions were observed in the olfactory and respiratory epithelium at 3 ppm. 

No lesions were seen in other tissues at any dose. 

In another 13-week inhalation study, 40 male and 20 female CDF(F344) rats/group were exposed 

to 0, 3, 10, or 40 ppm EDB 6 hr/day, 5 days/week (Nitschke et al., 1981). Male rats in the 40 

ppm group exhibited decreased weight gain throughout most of the exposure period. However, 

reduced weight gain was never more than 6-8% below control levels. With the exception of 

decreased specific gravity of urine in females of the 40 ppm group, no treatment-related changes 

were observed in any rat group with respect to urinalysis, hematology, and clinical chemistry. At 

the end of 13 weeks, relative liver and kidney weights of males exposed to 40 ppm EDB were 

significantly elevated, while relative liver weights of females in the two highest exposure groups 

were significantly elevated. Absolute liver weight of females in the 40 ppm group was also 

significantly elevated. Histopathological examination revealed lesions primarily confined to the 

anterior sections of the nasal turbinates. Hyperplasia and nonkeratinizing squamous metaplasia 

of the respiratory epithelium were observed in nasal turbinates of rats exposed to 40 ppm EDB. 

Only slight epithelial hyperplasia of nasal turbinates was noted at 10 ppm. No treatment related 

effects were seen at 3 ppm. Livers of females in the 40 ppm group showed a slight increase in 

fat. After an 88 day recovery period, there was a reversion to normal of the nasal turbinates in 

all but one rat. 

In what was originally scheduled to be a lifetime exposure study, 50 Osborne-Mendel 

rats/group/sex and 50 B6C3F1 mice/group/sex were administered EDB 5 days/week by gastric 

lavage over a substantial portion of their life-span (NCI, 1978). Twenty untreated controls/sex 

and 20 vehicle controls/sex of each species were included in the study. Rats received initial 

doses of 80 and 40 mg/kg body wt-day for the first 17 weeks. Due to high mortality, dosing of 

high dose rats was discontinued for 13 weeks and resumed on week 30 at 40 mg/kg body wt-day. 

In week 42, all intubations of low and high dose rats ceased for 1 week followed by 4 weeks of 

dose administration. All surviving, treated male rats were necropsied in week 49; all surviving, 

treated female rats were necropsied in week 61. The resulting time-weighted average dose over 

the test period was 38 and 41 mg/kg body wt-day for low and high dose males, respectively, and 

37 and 39 mg/kg body wt-day for low and high dose females, respectively. Mice received initial 

doses of 120 and 60 mg/kg body wt-day. In weeks 11-13, high and low doses were increased to 

200 and 100 mg/kg body wt-day, respectively. Original dose levels were resumed after week 13. 

At week 40, administration of EDB was decreased to 60 mg/kg body wt-day for high dose mice. 

EDB administration was discontinued at week 54 with necropsy occurring at week 78 for males 

and high dose females. Low dose female mice were observed for 37 weeks after intubation 

ceased. The resulting time-weighted average dose over the test period was 62 and 107 mg/kg 

body wt-day for low and high dose mice, respectively. In rats, clinical signs by week 5 included 

reddened ears and hunched back in all treatment groups. By week 10, all treated rats had 

reduced body weights (>10%). Both female and male rats exhibited dose-dependent mortality. 

Many of the deaths occurred during or shortly after intubation, suggesting an acute toxic 

reaction. Pathology revealed hyperkeratosis and acanthosis of the forestomach in high dose 

males and females and in one low dose female. A small number of rats in both treatment groups 

showed adrenal cortex degeneration and peliosis of the liver (hepatitis). Dosed males showed 

Appendix D3 218 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

early development of testicular atrophy. In mice, dose-related body weight reduction and 

mortality were observed. Clinical signs included alopecia, thin, hunched appearance, soft feces 

and body sores. Hyperkeratosis and acanthosis of the forestomach were seen in high dose male 

and female mice. One incidence each of hyperkeratosis (in a female) and acanthosis (in a male) 

was seen at the low dose. Splenic changes were present in high dose mice and testicular atrophy 

was present in high dose males. 

In a long-term inhalation exposure study, F344 rats and B6C3F1 mice were exposed to 0, 10, or 

40 ppm EDB 6 hr/day, 5 days/week for up to 103 weeks (NTP, 1982). In male and female rats, 

the high dose groups had reduced body weights and increased mortality that began at about week 

60. The treatment-related non-neoplastic pathology included hepatic necrosis (both sexes), 

epithelial hyperplasia and suppurative inflammation throughout the respiratory system (both 

sexes), and nephropathy (males only). Toxic nephropathy and mineralization were also seen in 

high dose female rats. Testicular degeneration and atrophy occurred with greater frequency in 

exposed rats and may be related to observed testicular tumors. Spermatic granulomas were also 

more frequently seen in high-dose males. Degeneration of the adrenal cortex appeared to be 

dose-related in females, but only one incidence each was seen in low and high dose males. 

Increased incidence of retinal atrophy was observed in exposed females. In mice, body weights 

were reduced at the high dose in both males and females. Many of the high dose animals 

exhibited a progressive weakness of the limbs or body during the second year. Increased 

mortality occurred in a dose-related manner in females and was significantly greater in low dose 

males. Non-neoplastic pathology included epithelial hyperplasia throughout the respiratory 

system and serous and suppurative inflammation of the nasal cavity in exposed mice. In all male 

mice, the principal cause of death was urinary bladder inflammation. However, bladder 

epithelial hyperplasia was only seen in exposed animals. An increased incidence of suppurative 

inflammation of the prostate was present but was also seen in controls. Dose-related spleen 

hematopoiesis was observed in females. 

Another long-term inhalation study investigated the effects of 0 or 20 ppm EDB (7 hr/day, 5 

days/week) on 48 Sprague-Dawley rats/sex/group for 18 months (Wong et al., 1982). 

Significantly lower body weight gains (>10% difference from controls) occurred by the 15th

month in males, and by the 18th month in females. Significantly reduced food consumption was 

not apparent. Increased mortality rates in both sexes occurred beginning in the 12th month of 

EDB exposure. All hematological findings were within normal ranges. The only recorded nonneoplastic gross or microscopic finding was atrophy of the spleen in males, which may be related 

to tumor formation (hemangiosarcoma). The nasal cavity was not examined. 

In a study of the effect of EDB on sperm production in bulls (Isreal-Friesian breed), 4 calves 

were fed 2 mg/kg body wt-day for 12 months (Amir and Volcani, 1965). The bulls were then 

given EDB in gelatin capsules every other day for 2-4 months longer. EDB did not appear to 

affect the growth, health, and libido of the bulls. However, semen density and motility were 

significantly lower compared to untreated control bulls of the same age. Many abnormal 

spermatozoa were also present in treated bulls. A NOAEL for this effect was apparently not 

determined. Cessation of EDB administration resulted in normal sperm within 10 days to 3 

months. Further studies confirmed that EDB adversely affected sperm production without any 

other apparent effects on bulls (Amir and Volcani, 1967; Amir and Ben-David, 1973). However, 

Appendix D3 219 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

feeding rams 2-5 mg/kg body wt-day for 120 days did not result in any effect on sperm or on the 

health of the animal (Amir, 1991). 

Female B6C3F1 mice (10/group) were given 31.25, 62.5, or 125 mg/kg EDB by gastric lavage 5 

days/week for 12 weeks (Ratajzak et al., 1995). At the highest dose, EDB significantly 

prolonged intervals between estrus, decreased hemoglobin and hematocrit levels, and increased 

cholesterol, triglycerides, total protein, and albumin. The highest dose also caused an 

immunosuppressive effect by lowering the in vitro splenic lymphocyte response to T- and B-cell 

mitogens. 

In a developmental toxicity study, 15-17 pregnant Charles River CD rats and 17-19 pregnant CD 

mice were exposed to 0, 20, 38, and 80 ppm EDB by inhalation 23 hr/day during days 6 to 16 of 

gestation (Short et al., 1978). A significant increase in mortality occurred in adult rats exposed 

to 80 ppm EDB and in adult mice exposed to 38 and 80 ppm EDB. Mice exposed to the highest 

dose experienced 100% mortality. Reduced body weights and feed consumption occurred in 

both species at all doses tested. Fetal mortality was increased in rats at the highest dose and in 

mice at 38 ppm. Reduced fetal body weights occurred at 38 ppm in rats and at all exposure 

levels in mice. No anomalies were seen in rat fetuses. An increase in runts at 38 ppm and a 

dose-dependent increase in skeletal anomalies were observed among mouse fetuses. However, 

these anomalies were characteristic of delayed development and occurred at doses that adversely 

affected maternal welfare. Therefore, these effects are indicative of fetal toxicity rather than 

teratogenicity. 

Male reproductive toxicity of EDB has been evaluated in some other experimental animals. New 

Zealand white rabbits, dosed subcutaneously with 0, 15, 30, or 45 mg/kg body wt-day, showed 

adverse effects at the highest dose (Williams et al., 1991). Increased mortality, increased serum 

enzymes, and liver damage were observed at this dose level. With respect to sperm quality, 

sperm velocity, motility, and motion parameters were reduced at the highest dose. A dose 

related decrease in semen pH was also noted. However, male fertility and fetal structural 

development were unaffected. 

A dominant lethal assay in mice was negative following a single intraperitoneal injection of 100 

mg EDB/kg body wt (Barnett et al., 1992). Germ cell tests did not indicate that EDB was a germ 

cell mutagen in male mice. 

Appendix D3 220 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VI. Derivation of Chronic Reference Exposure Level (REL)

Study Ratcliff et al., 1987 

Study population 46 exposed men, 43 unexposed men; 89 total 

Exposure method Variable workplace breathing zone airborne 

exposure (88 ppb geometric mean 8-hour time 

weighted average (TWA) exposure with peak 

exposures up to 262 ppb) 

Critical effects Reproductive toxicity; decreased sperm 

count/ejaculate, decreased percentage of viable 

and motile sperm, increased semen pH, and 

increased proportion of sperm with specific 

morphological abnormalities (tapered heads, 

absent heads, and abnormal tails) in human 

males 

LOAEL 88 ppb 

NOAEL Not observed 

Exposure continuity 8 hr/day (10 m3

/day occupational inhalation 

exposure rate), 5 days/week

Exposure duration Average, 4.9 years (with standard deviation of 3.6 

years) 

Average experimental exposure 31 ppb for LOAEL group (88 x 10/20 x 5/7) 

Human equivalent concentration 31 ppb 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 3 

Interspecies factor 1 

Intraspecies factor 10

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.1 ppb (0.0008 mg/m3

, 0.8 μg/m3) 

The primary study by Ratcliff and associates (1987) found significant changes in sperm quality 

indices of papaya workers exposed to EDB vapors for an average of nearly 5 years. No other 

health effects were apparent. A level of EDB at which no toxicity was observed (NOAEL) was 

not determined. 

In addition to the primary study of Ratcliff et al. (1987), several other epidemiological studies 

together strongly suggest a correlation between EDB exposure and male reproductive toxicity 

(Ter Haar, 1980; Wong et al., 1979; Wong et al., 1985; Rogers et al., 1981; Schrader et al., 

1988). This lesion appears to occur in humans at concentrations at which other toxic effects are 

not seen. EDB also shares some structural similarity to dibromochloropropane (DBCP), a 

known reproductive toxicant in human males. The evidence for male reproductive toxicity of 

EDB is not as strong as that for DBCP, probably because EDB is not as potent as DBCP in 

producing this toxic effect. However, animal studies demonstrate testicular toxicity and the 

number of studies indicating a connection between male reproductive toxicity and EDB exposure 

cannot be ignored for the development of the REL. 

Appendix D3 221 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Chronic oral exposure of bulls to EDB results in similar toxic effects at low concentrations 

(equivalent to 0.9 ppm) without affecting the general health of the animal (Amir and Volcani, 

1965; Amir, 1991). However, the small sample size and the lack of a dose-response effect and 

an observed NOAEL limits the usefulness of this study. Long-term studies of EDB toxicity in 

other experimental animals also lack the determination of a NOAEL (NCI, 1978; NTP, 1982). 

Evidence of testicular atrophy was found in other long-term studies with experimental animals, 

but at concentrations that also produced toxic effects in other organ systems. 

For comparison with the proposed REL based on a human study, the NTP (1982) chronic 

inhalation study established a LOAEL (10 ppm) for liver, kidney, eyes, and the respiratory, male 

reproductive, and endocrine system in rats. A LOAEL was established in mice for mortality, 

spleen changes in females, and respiratory system toxicity. A NOAEL was not established for 

either species. Use of a time adjustment (6/24 hr/day, 5/7 day/week), an RGDR of 1, and a total 

uncertainty factor of 300 (an interspecies UF of 3, a LOAEL to NOAEL UF of 10, and an 

intraspecies UF of 10) resulted in an estimated REL of 6 ppb (50 μg/m3

). 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for ethylene dibromide include the use of human exposure 

data from workers exposed over a period of years, and the presence of the toxic endpoint (male 

reproductive system) in several experimental animal species. Major areas of uncertainty are the 

lack of observation of a NOAEL, the uncertainty in estimating occupational exposure, the 

potential variability in occupational exposure concentration, and the limited nature of the study 

(fertility was not actually tested). The database for chronic toxicity of EDB in experimental 

animals would be enhanced if the proper doses were chosen to determine a NOAEL. 

VIII. Potential for Differential Impacts on Children's Health

Little fetal toxicity was observed when pregnant rats and mice were exposed to 20 ppm EDB 

during gestation (Short et al., 1978). Thus the REL of 0.1 ppb should adequately protect infants 

and children. However, we do not know if adolescent boys would be more sensitive than men to 

this alkylating agent. Differences in metabolic capability between infants and older children and 

adults may result in either more or less toxicity of EDB. Both oxidative and conjugated 

metabolites are toxic. Infants may produce proportionately more conjugate than oxidized 

metabolite relative to adults. 

IX. References 

Aman J, Farkas L, and Ben-Shamai MH. 1946. Experiments on the use of ethylene dibromide as 

a fumigant for grain and seed. Annals Appl. Biol. 33:389-395. 

Amir D. 1991. The spermicidal effect of ethylene dibromide in bulls and rams. Molec. Repro. 

Develop. 28:99-109. 

Appendix D3 222 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Amir D, and Ben-David E. 1973. The pattern of structural changes induced in bull spermatozoa 

by oral or injected ethylene dibromide (EDB). Ann. Biol. Anim. Biochem. Biophys. 13(2):165-

170. 

Amir D, and Volcani R. 1965. Effect of dietary ethylene dibromide on bull semen. Nature 

206(4979):99-100. 

Amir D, and Volcani R. 1967. The effect of dietary ethylene dibromide (EDB) on the testes of 

bulls. Fertil. Steril. 18(1):144-148. 

Barnett LB, Lovell DP, Felton CF, Gibson BJ, Cobb RR, Sharpe DS, Shelby MD, and Lewis SE. 

1992. Ethylene dibromide: Negative results with the mouse dominant lethal assay and the 

electrophoretic specific-locus test. Mutat. Res. 282:127-133. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dobbins JG. 1987. Regulation and the use of “negative” results from human reproductive 

studies: The case of ethylene dibromide. Am. J. Ind. Med. 12:33-45. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/95). 

Jones AR, and Edwards K. 1968. The comparative metabolism of ethylene dimethane-sulphonate 

and ethylene dibromide. Experientia 24:1100-1101. 

Kulkarni AP, Edwards J, and Richards IS. 1992. Metabolism of 1,2-dibromoethane in the human 

fetal liver. Gen. Pharmacol., 23(1):1-5. 

NCI. 1978. National Cancer Institute. Bioassay of 1,2-dibromoethane for possible 

carcinogenicity. Technical Report Series No. 86, DHEW Publ. no. (NIH) 78-1336. 

Nitschke KD, Kociba RJ, Keyes DG, and McKenna MJ. 1981. A thirteen week repeated 

inhalation study of ethylene dibromide in rats. Fundam. Appl. Toxicol. 1:437-442. 

NTP. 1982. National Toxicology Program. Carcinogenesis bioassay of 1,2-dibromoethane in 

F344 rats and B6C3F1 mice (inhalation study). Technical Report Series No. 210, NIH Publ. no. 

82-1766. 

Ott MG, Scharnweber HC, and Langner RR. 1980. Mortality experience of 161 employees 

exposed to ethylene dibromide in two production units. Br. J. Ind. Med. 37:163-168. 

Appendix D3 223 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Plotnick HB, and Conner WL. 1976. Tissue distribution of 14C-labeled ethylene dibromide in the 

guinea pig. Res. Commun. Chem. Path. Pharmacol. 13(2):251-258. 

Plotnick HB, Weigel WW, Richards DE, and Cheever KL. 1979. The effect of dietary disulfiram 

upon the tissue distribution and excretion of 14C-1,2-dibromoethane in the rat. Res. Commun. 

Chem. Pathol. Pharmacol. 26(3):535-545. 

Ratajczak HV, Thomas PT, Gerhart J, and Sothern RB. 1995. Imunotoxicologic effects of 

ethylene dibromide in the mouse and their modulation by the estrus cycle. In Vivo 9:299-304. 

Ratcliff JM, Schrader SM, Steenland K, Clapp DE, Turner T, and Hornung RW. 1987. Semen 

quality in papaya workers with long term exposure to ethylene dibromide. Br. J. Ind. Med. 

44:317-326. 

RECT. 1988. Reviews of Environmental Contamination and Toxicology. Ethylene dibromide. 

104:115-129. 

REPROTOX. 1995. The REPROTOX(R) System: Reproductive Reviews of Drugs, Chemicals, 

Physical and Biological agents. Denver, CO: Micromedex, Inc. (Edition expires 7/31/95). 

Reznik G, Stinson SF, and Ward JM. 1980. Respiratory pathology in rats and mice after 

inhalation of 1,2-dibromo-3-chloropropane or 1,2 dibromoethane for 13 weeks. Arch. Toxicol. 

46:233-240. 

Rogers BJ, Fujita JS, Najita L, and Hale RW. 1981. Reduction of sperm concentration in a 

population exposed to ethylene dibromide (EDB). J. Androl. 2:35-36. 

Rowe VK, Spencer HC, McCollister DD, Hollingsworth RL, and Adams EM. 1952. Toxicity of 

ethylene dibromide determined on experimental animals. Arch. Ind. Hyg. Occup. Med. 6(2):158-

173. 

Schrader SM, Ratcliff JM, Turner TW, and Hornung RW. 1987. The use of new field methods of 

semen analysis in the study of occupational hazards to reproduction: The example of ethylene 

dibromide. J. Occup. Med. 29(12):963-966. 

Schrader SM, Turner TW, and Ratcliff JM. 1988. The effects of ethylene dibromide on semen 

quality: A comparison of short-term and chronic exposure. Repro. Toxicol. 2:191-198. 

Short RD, Minor JL, Winston JM, Seifter J, and Lee C. 1978. Inhalation of ethylene dibromide 

during gestation by rats and mice. Toxicol. Appl. Pharmacol. 46:173-182. 

Stott WT, and McKenna MJ. 1984. The comparative absorption and excretion of chemical 

vapors in the upper, lower, and intact respiratory tract of rats. Fundam. Appl. Toxicol. 4:594-

602. 

Appendix D3 224 Ethylene dibromide 
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Ter Haar G. 1980. An investigation of possible sterility and health effects from exposure to 

ethylene dibromide. In: Banbury Report 5-Ethylene Dibromide: A Potential Health Risk? Ames 

B, Infante P, and Reitz R. (eds). Cold Spring Harbor, NY: Cold Spring Harbor Laboratory. pp. 

167-188. 

Watanabe P, Young J, Schlachter M, Zempel J, and Karbowski R. 1978. Fate of inhaled ethylene 

dibromide in rats. Toxicol. Appl. Pharmacol. 45:224(abstract). 

Wiersma DA, Schnellmann RG, and Sipes IG. 1986. The in vitro metabolism and bioactivation 

of 1,2-dibromoethane (ethylene dibromide) by human liver. J. Biochem. Toxicol. 1(3):1-11. 

Williams J, Gladen BC, Turner TW, Schrader SM, and Chapin RE. 1991. The effects of ethylene 

dibromide on semen quality and fertility in the rabbit: Evaluation of a model for human seminal 

characteristics. Fundam. Appl. Toxicol. 16:687-700. 

Wong LCK, Winston JM, Hong CB, and Plotnick H. 1982. Carcinogenicity and toxicity of 1,2-

dibromoethane in the rat. Toxicol. Appl. Pharmacol. 63:155-165. 

Wong O, Utidjian HMD, and Karten VS. 1979. Retrospective evaluation of reproductive 

performance of workers exposed to ethylene dibromide (EDB). J. Occup. Med. 21(2):98-102. 

Wong O, Morgan RW, and Whorton MD. 1985. An epidemiologic surveillance program for 

evaluating occupational reproductive hazards. Am. J. Ind. Med. 7:295-306. 

Appendix D3 225 Ethylene dibromide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

ETHYLENE DICHLORIDE 

(1,2-dichloroethane) 

CAS Registry Number: 107-06-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 400 μg/m3 (100 ppb) 

Critical effect(s) Hepatotoxicity; elevated liver enzyme levels in 

serum of rats. 

Hazard index target(s) Liver 

II. Physical and Chemical Properties (HSDB, 2000; CRC, 1994)

Description Clear, colorless, oily liquid 

Molecular formula C2H4Cl2

Molecular weight 98.97 g/mol 

Density 1.2351 g/cm3

 @ 20oC 

Boiling point 57.4oC 

Melting point −96.9oC 

Vapor pressure 64 torr @ 20oC 

Solubility Slightly soluble in water (0.869 g/100 ml at 

20oC); miscible with alcohol; soluble in 

ordinary organic solvents 

Conversion factor 1 ppm = 4.05 mg/m3

 

III. Major Uses or Sources 

Ethylene dichloride (EDC) is used primarily in the production of vinyl chloride monomer 

(HSDB, 2000). It is also an intermediate in the manufacture of trichloroethane and 

fluorocarbons and is used as a solvent. In California, EDC is also used as a reactant carrier in the 

production of solid fuel (CARB, 1997). EDC was commonly used as a gasoline additive to 

scavenge inorganic lead compounds. The transition to the use of lead-free gasoline has 

essentially eliminated the use of EDC as a fuel additive in this country. EDC was also used as a 

soil fumigant but is no longer registered for this use on agricultural products in the United States. 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 24,935 pounds 

of ethylene dichloride (CARB, 2000). 

Appendix D3 226 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposure 

Toxicological data resulting solely from long-term exposure to EDC in humans are lacking. 

Nausea, vomiting, dizziness, and unspecified blood changes were reported in a study of workers 

exposed to levels of 10-37 ppm EDC (Brzozowski et al., 1954). Kozik (1957) reported adverse 

central nervous system and liver effects in workers occupationally exposed to concentrations of 

16 ppm EDC and below. Rosenbaum (1947) also reported nervous system effects in a study of 

100 Russian workers exposed for less than 5 years to concentrations of EDC less than 25 ppm. 

Immediately following a 30-minute exposure to an unknown concentration of EDC, a 

51 year-old male was somnolent and experienced vomiting (Nouchi et al., 1984). Delirious and 

trembling, the worker was admitted to the hospital 20 hours post-exposure. The liver was 

palpable, but serum liver enzymes were normal. The patient lapsed into a coma 3.5 hours 

following admission to the hospital. A marked elevation in serum liver enzymes was noted on 

the second day of hospitalization, 35 hours post-exposure. Multiple organ failure occurred on 

the fourth day of hospitalization and the patient died of arrhythmia. At autopsy, the lungs were 

congested and edematous. Diffuse degenerative changes were observed in the myocardium. 

Extensive centrilobular necrosis was observed in the liver, and acute centrilobular necrosis was 

observed in the kidney. Nerve cells in the brain, including Purkinje cells, appeared shrunken 

with pyknotic nuclei. The latency period for hepatotoxicity of approximately 20 hours suggests 

that metabolism of the compound yields the reactive agent (see below). 

V. Effects of Animal Exposure 

As with humans, the absorption and distribution of EDC in rats following ingestion or inhalation 

is rapid and complete (IARC, 1999). Metabolism in rats and mice is extensive with 85% of the 

metabolites appearing in urine. Metabolism occurs predominantly via two pathways, one 

catalyzed by cytochrome P450 and one by glutathione S-transferase. The direct conjugation with 

glutathione catalyzed by glutathione S-transferase may ultimately result in the putative alkyating 

agent (episulfonium ion) primarily responsible for toxicity and carcinogenicity. Evidence for 

DNA-damaging metabolites resulting via the P450 pathway exists (IARC, 1999). However, this 

pathway appears to be a minor route for toxic metabolite formation. 

Acute exposure in mice resulted in toxic effects similar to those seen in the human case study 

presented above, including liver and kidney damage (Francovitch et al., 1986). Acute EDC 

exposure exhibits a steep dose-response curve with respect to mortality. However, the long-term 

exposure studies were notable for the limited organ toxicity and mortality observed in 

comparison to acute studies (IARC, 1999). 

Male and female rats (50 per sex) were exposed to 50 ppm EDC 7 hours per day, 5 days per 

week for 2 years (Cheever et al., 1990). Absolute and relative liver weights were not 

significantly different from controls. Daily observations, gross pathology, and extensive 

histopathology revealed no differences from controls other than a slight increase in unspecified 

testicular lesions in the EDC group. Additional rats were exposed to 50 ppm EDC with 0.05% 

disulfiram (a non-carcinogen used extensively in the rubber industry and as a treatment 

Appendix D3 227 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

(Antabuse) for alcoholism) in the diet. Disulfiram treatment resulted in increased number of 

tumors, increased blood levels of EDC, and increased liver (primarily bile duct cysts) and kidney 

(chronic nephropathy) lesions. It was concluded that some pathways responsible for metabolism 

of EDC were inhibited by disulfiram, resulting in increased EDC blood levels and bioactivation 

to toxic metabolites via other metabolic pathways. 

Rats (8-10 per sex per group) were exposed to 0, 5, 10, 50, and 150-250 ppm EDC 7 hours per 

day, 5 days per week for up to 18 months (Spreafico et al., 1980). Serum chemistry 

measurements were taken after 3, 6, 12, and 18 months of exposure. Rats to be examined after 3, 

6 and 18 months of exposure were 3 months of age at the beginning of the experiment, and rats 

to be examined after 12 months of exposure were 14 months of age at the beginning of the 

experiment. Complete histological exams were conducted but non-cancer effects were not 

discussed. No consistent treatment-related changes in serum chemistry parameters were 

observed at 3, 6, or 18 months of exposure. However, rats exposed to higher levels of EDC for 

12 months exhibited changes in serum chemistry indicative of chronic liver damage, primarily 

increased alanine aminotransferase (ALT) levels at the two highest exposures. Lactate 

dehydrogenase (LDH) and aspartate aminotransferase (AST) levels were significantly decreased, 

but did not appear to be dose-related. γ-Glutamyl transpeptidase levels were elevated but at nonsignificant levels. Indicators of kidney toxicity included increased blood urea nitrogen levels in 

the 150 ppm group and increased uric acid levels at the two highest exposures. However, the 

control values for both of these parameters were significantly lower than that seen in rats tested 

at other times in this study. Thus, the toxicological significance is questionable. Cholesterol was 

reduced significantly at the higher exposure levels but the toxicological significance of this 

finding was unknown. The marked difference between serum chemistry parameters following 

12 months of exposure, compared to those following 3, 6, and 18 months of exposure, may be 

due to the considerable difference in the age of the rats at the start of exposure. This study 

identifies a 12-month LOAEL of 50 ppm and a NOAEL of 10 ppm in rats. 

A study examining the interaction between 1,2-dichloroethane and disulfiram (DSF) exposed rats 

to EDC concentrations of 150, 300, or 450 ppm 5 days per week for 30 days (Igwe et al., 1986a; 

Igwe et al., 1986b). Increased liver weights and increased 5-nucleotidase (5-NT) activity were 

observed in rats following exposure to 450 ppm EDC (the LOAEL for this study). This study 

also determined that the interaction between DSF and EDC greatly increased the toxicity of EDC 

(i.e., increased serum activities of SDH, APT, and 5-NT, bilateral testicular atrophy, periportal 

necrosis and cytoplasmic swelling of hepatocytes, and bile duct proliferation). Therefore, any 

person exposed to DSF either occupationally or therapeutically is likely to be more susceptible to 

the effects of EDC toxicity. 

Rats, rabbits, guinea pigs, dogs, cats, and monkeys were used in exposures ranging from 

approximately 100 to 1000 ppm EDC (Heppel et al., 1946). At the highest experimental 

concentration of 963 ppm, high mortality was observed in rats, rabbits, and guinea pigs 

following exposure 7 hours per day, 5 days per week for two weeks or less. At 963 ppm guinea 

pigs exhibited lacrimation and inactivity during exposure; pulmonary congestion was noted at 

autopsy. Rats exposed to this concentration exhibited degenerative proliferative changes in the 

renal tubular epithelium and splenitis. Pulmonary congestion and focal hemorrhage were also 

noted in 2 of 4 rats examined. While 4 of 6 cats exposed to this concentration survived until 

Appendix D3 228 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

sacrifice 11 weeks following termination of exposure, congestion and fatty infiltration of the 

liver were observed at necropsy. Due to high mortality in the rodents at the higher concentration, 

a subsequent experiment exposed rats and guinea pigs 7 hours per day, 5 days per week to 

100 ppm EDC for four months. No increase in mortality or effects on growth was observed in 

rats exposed to this concentration. The rats were successfully bred and their pups were exposed 

with the dams. No significant findings were observed upon gross and histological examinations 

of 10/39 exposed and 10 control rats. This study is severely limited by the methods used to 

determine the exposure concentration and by the lack of quantitative measurements of toxicity 

other than death. This study does, however, indicate that fatty infiltration of the liver is one 

indication of toxicity following multiple exposures to EDC. 

In developmental toxicity studies summarized by Zhao et al. (1997), rats were exposed to 0, 

24.8, and 207.6 mg/m3 (equivalent to 0, 6, and 51 ppm) EDC for 6 hr/day from two weeks before 

mating and throughout gestation. Statistically significant increases in pre-implantation loss and 

decreased male pup weights were observed at the highest dose. Gross skeletal and visceral 

malformations were not found. 

In a developmental study by Payan et al. (1995), Sprague-Dawley rats were exposed to 150, 200, 

250, or 300 ppm EDC for 6 hrs/day from day 6 to 20 of gestation. Maternal toxicity (reduced 

body weight gain; death of two females) was observed at the highest exposure. Statistically 

significant evidence of altered growth and teratogenic effects were not observed at any 

concentration. 

Rao et al. (1980) exposed rats and rabbits to 100 or 300 ppm EDC for 7 hr/day on days 6 through 

15 (rats) or 6 through 18 (rabbits) of gestation. Maternal toxicity (mortality) was observed in 

rabbits at 100 ppm, and both species at 300 ppm. One rat exhibited resorption of all 

implantations at the maternally-toxic dose. Otherwise, no fetotoxic or teratogenic effects were 

observed in either species. In a reproduction study, rats were exposed to 25, 75, or 150 ppm 

EDC 6 hr/day, 5 days/week for 60 days before breeding. Exposure following this period was 6 

hr/day, 7 days/week. Maternal animals were not exposed to EDC from gestational day 21 

through day 4 postpartum. EDC had no effect on reproduction over one generation within two 

litters. 

In a two-generation study conducted by Lane et al. (1982), ICR Swiss mice were administered 

30, 90, or 290 mg/L EDC in drinking water (equivalent to about 5, 15, or 50 mg/kg bw/day) 

starting five weeks before mating of the F0 generation. No treatment-related effects on fertility, 

gestation, viability, weight gain, or lactation indices were noted. EDC exposure did not result in 

teratogenic or dominant lethal effects. 

No gross or histopathological indications of hepato- or nephrotoxicity were observed in OsbornMendel rats (47 or 95 mg/kg bw/day, 5 days/week for both sexes) or B6C3F1 mice (97 or 195 

mg/kg bw/day, 5 days/week for males; 149 or 299 mg/kg bw/day, 5 days/week for females), 

which were given EDC via gavage for 78 weeks (NCI, 1978). However, rats of each sex and 

female mice had significantly reduced survival at the highest dose. 

Appendix D3 229 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

In a comparative study of the toxicity of EDC, Morgan et al. (1990) administered 0, 500, 1000, 

2000, 4000, and 8000 ppm in drinking water to several species of rats for 13 weeks. A 

statistically significant increase in kidney weight was observed in male and female Fischer 

344/N rats administered 1000 ppm or greater in drinking water. However, minimal histological 

damage was observed only in the kidney of female Fischer 344/N rats. A statistically significant 

decrease in body weight was observed in rats administered 8000 ppm. Significant decreases in 

absolute and relative kidney weight were observed in male and female rats administered 

concentrations of 1000 ppm EDC. A significant increase in relative liver weight was observed in 

male rats administered 2000 ppm EDC and greater and female rats administered 4000 ppm EDC 

and greater. Similar but less marked toxicity was observed in the Sprague-Dawley and OsborneMendel rats administered 1000 ppm. Additionally, rats were administered EDC in corn oil by 

gavage at doses of 0, 30, 60, 120, 240, and 480 mg/kg for 13 weeks (Morgan et al., 1990). Rats 

administered EDC by gavage exhibited high mortality in the higher dose groups. Statistically 

significant increases in kidney weights were observed in surviving male rats administered EDC 

and in female rats administered 120 or 240 mg/kg. However, no histological damage to the liver 

or kidney was observed. 

VI. Derivation of Chronic Reference Exposure Level (REL)

Study Spreafico et al., 1980. 

Study population Rats (8-10 per sex/group) 

Exposure method Discontinuous whole-body inhalation exposures 

(0, 5, 10, 50, or 150-250 ppm) 

Critical effects Significant elevation in liver enzymes 

Exposure duration 12 months 

Exposure continuity 7 hours/day, 5 days/week 

LOAEL 50 ppm 

NOAEL 10 ppm 

Average experimental exposure 2.1 ppm for NOAEL group ( 10 x 7/24 x 5/7) 

Human equivalent concentration 3.2 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.5 for lambda (a) : 

lambda (h)) (Gargas et al., 1989) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.1 ppm (100 ppb; 0.4 mg/m3

; 400 μg/m3) 

Cheever et al. (1990) and Spreafico et al. (1980) were the only chronic inhalation exposure 

studies found in the literature that presented non-cancer effects. No reproductive and 

developmental effects were observed in studies published in peer-reviewed journals. The study 

by Spreafico et al. (1980) was chosen for REL development based on the utilization of multiple 

exposure levels and the observation of a NOAEL and a LOAEL for liver effects. 

Appendix D3 230 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

The Agency for Toxic Substances and Disease Registry (ATSDR) calculated a chronic inhalation 

minimal risk level (MRL) for EDC of 0.2 ppm (ATSDR, 1994). The calculation was based on 

the study by Cheever et al. (1990), which determined a free-standing NOAEL of 50 ppm for lack 

of liver effects. A LOAEL was not determined. To derive the MRL, the ATSDR applied 

uncertainty factors (UFs) of 10 each for intraspecies and interspecies variability, and a modifying 

factor of 3 to account for database deficiencies, to the NOAEL of 50 ppm. The criteria for use of 

modifying factors are not well specified by ATSDR. Such modifying factors were not used by 

OEHHA. A continuity correction for discontinuous exposure was not applied. The resulting 

MRL was 0.2 ppm (0.7 mg/m3

). 

For comparison to the proposed REL, a REL developed by OEHHA based on the free-standing 

NOAEL of 50 ppm determined in rats by Cheever et al. (1990) would include a continuity 

correction (50 ppm x 7/24 x 5/7) resulting in an equivalent continuous level of 10.42 ppm.. 

Application of an RGDR = 1.5 and UFs of 3 for interspecies and 10 for intraspecies differences 

result in a REL of 0.5 ppm (2 mg/m3

). 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for ethylene dichloride include the availability of chronic 

inhalation exposure data, the relatively large number of exposure levels at lower concentrations 

(allowing for better elucidation of the dose-response relationship for hepatotoxicity), and the 

observation of a NOAEL. Major areas of uncertainty are the lack of adequate human exposure 

data, the small groups tested in the key study, and the lack of health effects data from multiple 

species. 

The small number of animals per group and the relatively modest clinical chemistry findings 

observed in the Spreafico et al. (1980) study may have resulted in false-positives, falsenegatives, and lack of clear dose-response relationships. Repeating the study in one or more 

experimental animal species with full histopathological examination of organs and a greater 

number of animals/dose would significantly enhance the chronic toxicity database for EDC. 

VIII. References 

ATSDR. 1994. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

1,2-Dichloroethane (Update). U.S. Department of Health and Human Services, Public Health 

Service. 

Brzozowski J, Czajka J, Dutkiewicz T, Kesy I, and Wojcik J. 1954. Work hygiene and the health 

condition of workers occupied in combating the Leptinotarsa decemlineata with HCH and 

dichloroethane. Med. Pr. 5: 89-98. [cited in U.S. EPA, 1985]. 

CARB. 1997. California Air Resources Board. Ethylene Dichloride. In: Toxic Air Contaminant 

Identification List Summaries. California Environmental Protection Agency, pp. 483-487. 

Appendix D3 231 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Cheever KL, Cholakis JM, El-Hawari AM, Kovatch RM, and Weisburger EK. 1990. Ethylene 

dichloride: The influence of disulfiram or ethanol on oncogenicity, metabolism, and DNA 

covalent binding in rats. Fundam. Appl. Toxicol. 14:243-261. 

Francovitch RJ, Schor NA, and George WJ. 1986. Effects of SKF-525A, phenobarbitol, and 3-

methylcholanthrene on ethylene dichloride toxicity following inhalation exposure. J. Am. Coll. 

Toxicol. 5(2):117-126. 

Gargas ML, Burgess RJ, Voisard DE, Cason GH, and Anderson ME. 1989. Partition coefficients 

of low-molecular weight volatile chemicals in various liquids and tissues. Toxicol. Appl. 

Pharmacol. 98:87-99. 

HSDB. 2000. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD. 

Available online through Toxicology Data Network at http://toxnet.nlm.nih.gov 

Heppel LA, Neal PA, Perrin TL, Endicott KM, and Porterfield VT. 1946. The toxicology of 1,2-

dichloroethane (ethylene dichloride). V. The effects of daily inhalations. J. Ind. Hyg. Toxicol. 

28(4):113-120. 

IARC. 1999. International Agency for Research on Cancer. Re-evaluation of some organic 

chemicals, hydrazine and hydrogen peroxide. IARC Monogr. Eval. Carcinog. Risks Hum. 

71(Pt.2):501-529. 

Igwe OJ, Que Hee SS, and Wagner WD. 1986a. Interaction between 1,2-dichloroethane and 

disulfiram. I. Toxicological effects. Fundam. Appl. Toxicol. 6:733-746. 

Igwe OJ, Que Hee SS, and Wagner WD. 1986b. Interaction between 1,2-dichloroethane and 

tetraethylthiuram disulfide (disulfiram). II. Hepatotoxic manifestations with possible mechanism 

of action. Toxicol. Appl. Pharmacol. 86:286-297. 

Kozik I. 1957. Problems of occupational hygiene in the use of dichloroethane in the aviation 

industry. Gig. Tr. Prof. Zabol. 1:31-38. [cited in U.S. EPA, 1985]. 

Lane RW, Riddle BL, and Borzelleca JF. 1982. Effects of 1,2-dichloroethane and 1,1,1-

trichloroethane in drinking water on reproduction and development in mice. Toxicol. Appl. 

Pharmacol. 63:409-421. 

Morgan DL, Bucher JR, Elwell MR, Lilja HS, and Krishna Murthy AS. 1990. Comparative 

toxicity of ethylene dichloride in F344/N, Sprague-Dawley and Osborne-Mendel rats. Food 

Chem. Toxicol. 28(12):839-845. 

Appendix D3 232 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

NCI. 1978. United States National Cancer Institute. Bioassay of 1,2-dichloroethane for possible 

carcinogenicity CAS No. 107-06-2 (NCI-CG-TR-55), Washington DC. 

Nouchi T, Miura H, Kanayama M, Mizuguchi O, and Takano T. 1984. Fatal intoxication by 1,2-

dichloroethane - a case report. Int. Arch. Occup. Environ. Health 54:111-113. 

Payan JP, Saillenfait AM, Bonnet P, Fabry JP, Langonne I, and Sabate JP. 1995. Assessment of 

the developmental toxicity and placental transfer of 1,2-dichloroethane in rats. Fundam. Appl. 

Toxicol. 28:187-198. 

Rao KS, Murray JS, Deacon MM, John JA, Calhoun LL, and Young JT. 1980. Teratogenicity 

and reproduction studies in animals inhaling ethylene dichloride. In: Banbury Report 5. Ethylene 

Dichloride: A Potential Health Risk? Ames B, Infante P, and Reitz R. (eds). Cold Spring Harbor, 

NY: Cold Spring Harbor Laboratory. pp. 149-166. 

Rosenbaum ND. 1947. Ethylene dichloride as an industrial poison. Gig. Sanit. 12(2): 17-21. 

[cited in U.S. EPA, 1985]. 

Spreafico F, Zuccato E, Marcucci F, Sironi M, Paglialunga S, Madonna M, and Mussini E. 1980. 

Pharmacokinetics of ethylene dichloride in rats treated by different routes and its long-term 

inhalatory toxicity. In: Banbury Report 5. Ethylene Dichloride: A Potential Health Risk? Ames 

B, Infante P, and Reitz R. (eds). Cold Spring Harbor, NY: Cold Spring Harbor Laboratory. pp. 

107-129. 

U.S. EPA. 1985. U.S. Environmental Protection Agency. Health Assessment Document for 1,2-

Dichloroethane (Ethylene Dichloride). U.S. EPA, Office of Research and Development, Office 

of Health and Environmental Assessment, Environmental Criteria and Assessment Office. 

Research Triangle Park, NC: U.S. EPA. 

Zhao S-F, Zhang X-C, Zhang L-F, Zhao S-S, Zhang F, Wang Q-F, Wang Y-L, and Bao Y-S. 

1997. The evaluation of developmental toxicity of chemicals exposed occupationally using 

whole embryo culture. Int. J. Dev. Biol. 41:275-282. 

Appendix D3 233 Ethylene dichloride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICOLOGY SUMMARY 

ETHYLENE GLYCOL 

(1,2-dihydroxyethane; 1,2-ethanediol) 

CAS Registry Number: 107-21-1 

I. Chronic Toxicity Summary 

Chronic reference exposure level 400 μg/m3

 (200 ppb)

Critical effects Respiratory irritation in human volunteers 

Hazard index target(s) Respiratory system; kidney; teratogenicity 

II. Physical and Chemical Properties (HSDB, 1996; 1999) 

Description Clear, colorless, odorless liquid 

Molecular formula C2H6O2

Molecular weight 62.07 g/mol 

Density 1.1088-1.1135 g/cm3

 @ 20° C 

Boiling point 197.6° C 

Melting point −13° C (CRC, 1994) 

Vapor pressure 0.06 torr @ 20°C; 0.092 torr @ 25°C 

Solubility Soluble in water and ethanol; slightly soluble in 

ether. Insoluble in benzene and petroleum 

ether. 

Conversion factor 1 ppm = 2.5 mg/m3 @ 25° C 

III. Major Uses and Sources 

Ethylene glycol is used as an antifreeze agent in cooling and heating systems (HSDB, 1996). It 

is used in hydraulic brake systems; as an ingredient in electrolytic condensers; as a solvent in the 

paint and plastics industries; and in inks for ball-point pens and printer’s inks. It is used in the 

manufacture of some synthetic fibers (Terylene and Dacron), and in synthetic waxes. It is used in 

some skin lotions and flavoring essences. Also, it is used in asphalt emulsion plants, in wood 

stains and adhesives, and in leather dyeing. It has been used as a de-icing fluid for airport 

runways. The annual statewide emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 66,636 pounds 

of ethylene glycol (CARB, 1999). 

Appendix D3 234 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure

Laitinen et al. (1995) found that 10 motor servicing workers had significantly higher urinary 

levels of ethylene glycol and ammonia, and decreased urinary glycosaminoglycan levels, 

compared with 10 controls. The ethylene glycol levels in air were undetectable in the workers’ 

breathing zones (i.e. below 1.9 ppm), therefore dermal absorption appeared to be the primary 

route of exposure. Because the dermal absorption rate is high, airborne ethylene glycol 

concentrations in workplaces likely underestimate the total exposure. 

In a study of 20 volunteer male prisoners in Alabama, 20 hour/day exposure to aerosolized 

ethylene glycol concentrations varying up to a mean of 20 ppm (49 mg/m3

) for 30 days was 

without effect (Wills et al., 1974). The actual concentrations measured in the exposure chamber 

were: 

Concentration of ethylene glycol in air, mg/m3

Days Low Higha Mean 

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

 1-7 3.6 75.0 37 

 8-14 18.8 44.8 29 

15-21 0.8 41.6 17 

22-28 3.5 49.2 23 

29-35 20.6 66.8 49 

36-37 14.4 39.0 31 

------------------------------------------------------------- a does not include the very high concentrations maintained for comparatively brief periods.

Respiratory irritation was noted after 15 minutes at an exposure concentration of 75 ppm (188 

mg/m3

), and became quickly intolerable at 123 ppm (308 mg/m3). No effects were observed in 

normal clinical chemistry, clinical serum enzyme levels for liver and kidney toxicity (including 

SGOT and serum alkaline phosphatase), hematotoxicity (including % hematocrit and gm 

hemogloin per 100 ml blood), or psychological responses (including simple reaction time, weight 

discrimination, and depth perception). The respiratory irritation at 75 ppm resolved soon after 

exposure with no long term effects noted after a 6-week follow-up period. 

V. Effects of Animal Exposure 

A chronic feeding study in rats and mice was conducted by DePass et al. (1986a). In this study, 

rats (130 per sex per group) and mice (80 per sex per group) were exposed to 0, 0.04, 0.2, or 1 

g/kg/day for up to 2 years. All male rats in the high dose group died by 475 days. A large 

number of effects were observed in this group, including: reduced body weight, increased water 

intake, increased blood urea nitrogen and creatinine, reduced erythrocyte counts, reduced 

hematocrit and hemoglobin, increased neutrophil count, and increased urine volume. Heart, 

kidney, lung, parathyroid, stomach, and other vascular mineralization and hyperplasia were 

observed histologically in the high dose group of the male rats. Female rats exhibited fatty 

changes and granulomas in the liver at the high dose. Liver effects were not reported for the 

Appendix D3 235 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

males. The NOAEL in rats for chronic oral ethylene glycol toxicity was 200 mg/kg/day. No 

effects were observed in mice. Therefore, the NOAEL for mice was 40 mg/kg/day. 

Coon et al. (1970) exposed groups of rats (as well as guinea pigs, rabbits, dogs, and monkeys) to 

ethylene glycol intermittently 8 hours/day, 5 days per week for 6 weeks (30 exposures) to 10 or 

57 mg/m3

 or continuously to 12 mg/m3 for 90 days. At 10 mg/m3 2 rabbits had conjunctivitis 

and liver changes were noted in a few animals of the other species. At 57 mg/m3

 no signs of 

toxicity were seen during the exposure. Nonspecific inflammatory changes were noted in some 

lungs and hearts of all species. A few livers also showed necrotic areas. Continuous exposure to 

12 mg/m3

 led to moderate to severe eye irritation in rats and rabbits. Edema in the rabbits led to 

eye closure. Two rats developed corneal opacities. All hematologic parameters and various 

enzymes assayed were within normal limits. At necropsy organs appeared normal. 

Histopathological analysis revealed inflammatory changes in the lungs of all species, but the 

controls also showed a lesser degree of inflammation. Several guinea pigs showed foci of 

inflammatory cells in the kidney. 

Mortality in Coon et al. (1970) Number died/number exposed

Ethylene 

 glycol 

mg/m3

Exposure 

duration 

Equivalent 

continuous 

concentration Rat Guinea pig Rabbit Dog Monkey 

0 (control) 90 days 0 4/123 0/73 0/12 0/12 0/8 

10±1 6 wk 2.4 0/15 0/15 0/3 0/2 0/2 

57±14 6 wk 13.6 0/15 0/15 0/3 0/2 0/2 

12±2 90 days 12.0 1/15 3/15 1/3 0/2 0/3 

Studies on the effects of inhaled ethylene glycol on reproduction and development of rats and 

mice were conducted by Tyl et al. (1995a, 1995b). In a study using whole-body exposure of rats 

and mice to ethylene glycol at analyzed concentrations of 0, 119, 888, or 2090 mg/m3 for 6 

hours/day on days 6-15 of gestation, mice were found to be the more sensitive species. Maternal 

toxicity in rats included a significant increase in absolute and relative liver weight at 2090 

mg/m3. No effects on weight gain, organ weights other than liver, fecundity, live fetuses per 

litter, or pre- or post-implantation loss were observed in rats. In addition, terata were not 

observed at any concentration. Reduced ossification in the humerus, zygomatic arch, and the 

metatarsals and proximal phalanges of the hindlimb was present in fetuses exposed to 888 or 

2090 mg/m3

. The NOAEL for maternal toxicity in rats was 888 mg/m3, while the NOAEL for 

fetotoxicity was 119 mg/m3

. 

In mice, reduced body weight and gravid uterine weight during and after the exposure were 

observed at the 888 and 2090 mg/m3

 concentrations. Increased nonviable implants per litter and 

reduced fetal body weights were also observed in groups exposed to 888 or 2090 mg/m3. 

External, visceral, skeletal, and total malformations were increased in the 888 and 2090 mg/m3

groups. The NOAEL for these effects in mice was 119 mg/m3

. 

A similar experiment in mice using nose-only exposures was conducted by these researchers 

(Tyl et al., 1995a) to determine the role of dermal absorption and/or ingestion on the effects 

Appendix D3 236 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

observed with the whole-body exposure. Nose-only exposures to ethylene glycol were for 6 

hours/day, on gestational days 6 through 15 at concentrations of 0, 500, 1000, and 2000 mg/m3

. 

The NOAEL for maternal effects (increased kidney weight) was 500 mg/m3, and the NOAEL for 

fetal toxicity (skeletal variations and fused ribs) was 1000 mg/m3

. Thus, secondary dermal 

and/or oral exposures appear to have contributed significantly to the developmental and maternal 

toxicity in mice exposed to ethylene glycol aerosol. The nose-only inhalation exposure study by 

Tyl et al. (1995a) was conducted in addition to the whole-body inhalation study since extensive 

adsorption of ethylene glycol onto the fur of the animals was demonstrated in the whole-body 

experiment. Normal grooming behavior would have resulted in significantly larger doses of 

ethylene glycol than that expected by inhalation only. 

A 3-generation study on the effects of ethylene glycol on reproductive performance and gross 

health of offspring in rats was conducted by DePass et al. (1986b). Rats were exposed orally to 

40, 200, or 1000 mg/kg/day ad libitum in the feed through 3 generations. No effects on pup 

survivability or pup body weight were observed. Total and viable implants were also not 

affected. Teratogenic effects were not examined in this study. 

Tyl et al. (1993) studied the reproductive and developmental effects of ethylene glycol in rabbits 

exposed by gavage on days 6 to 19 of gestation. Dams were exposed to 0, 100, 500, 1000, or 

2000 mg/kg/day. Exposure to 2000 mg/kg/day resulted in 42% mortality, and abortion or early 

delivery in 4 does. No evidence of embryotoxicity or teratogenicity was observed in the groups 

exposed to 1000 mg/kg/day or less. The NOAEL for maternal toxicity was determined to be 

1000 mg/kg/day. 

VI. Derivation of Chronic Reference Exposure Level 

Study Wills et al. (1974) 

Study population Human volunteer prisoners 

Exposure method Discontinuous whole-body inhalation 

Critical effects Respiratory tract irritation 

LOAEL 75 ppm 

NOAEL 20 ppm 

Exposure continuity 20 hours/day 

Exposure duration 30 days 

Average exposure 16.7 ppm for NOAEL group (20 x 20/24) 

Human equivalent concentration 16.7 ppm 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 10 

Interspecies factor 1 

Intraspecies factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.2 ppm (200 ppb; 0.4 mg/m3

; 400 μg/m3) 

The subchronic study by Wills et al. (1974) represents the only human inhalation data for 

ethylene glycol toxicity. The experiment showed a concentration-response relationship, with 

onset of irritation occurring at 188 mg/m3 and intense and intolerable irritation occurring at 308 

Appendix D3 237 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

mg/m3

. The volunteers were followed for 6 weeks without any apparent long-term effects from 

the exposures. Although the irritation experienced in the human subjects appears to be an acute 

phenomenon and not a cumulative lasting effect, the subchronic uncertainty factor of 10 was 

retained to protect against other systemic effects associated with ethylene glycol such as kidney 

damage which may occur over a long-term exposure. 

The chronic feeding study in rats by DePass et al. (1986a) showed significant chronic effects 

including reduced body weight, increased water intake, increased blood urea nitrogen and 

creatinine, reduced erythrocyte counts, reduced hematocrit and hemoglobin, increased neutrophil 

counts, increased urine volume, and reduced urine specific gravity and pH in rats exposed to a 

concentration of 1000 mg/kg/day. However, no effects were reported in mice. In contrast, 

reproductive and developmental toxicity studies in mice, rats, and rabbits have shown the mouse 

to be the most sensitive species for both terata and maternal toxicity endpoints (Tyl et al., 1995a; 

Tyl et al., 1993; Neeper-Bradley et al., 1995). In addition, the 3-generation reproductive toxicity 

study by DePass et al. (1986b) showed no significant effects on rat pup survival or body weight 

at concentrations up to 1000 mg/kg/day. However, developmental endpoints were not reported 

in this study. From the available data, the toxicity of ethylene glycol is apparently greatest in the 

maternal mouse. The estimated equivalent air concentrations (assuming a 70 kg human inhales 

20 m3/day) from the feed in the 3-generation study by DePass et al. (1986b) are 700 mg/m3 and 

3500 mg/m3 for the NOAEL and LOAEL, respectively. 

For comparison with the proposed REL of 400 μg/m3 based on a one month human study, the 

inhalation NOAEL of 48 ppm, obtained by Tyl et al. (1995) in mice discontinuously exposed for 

10 days on gestation days 6-15, was used to estimate a REL based on animal data. Use of a time 

adjustment from 6 to 24 hours/day, an RGDR of 1, an interspecies UF of 3, and an intraspecies 

UF of 10 resulted in an estimated REL of 0.4 ppm (1000 μg/m3) for ethylene glycol. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for ethylene glycol include the use of human exposure data, 

the use of controlled, nearly continuous inhalation exposures, the observation of a NOAEL, and 

the similar REL value estimated from an animal study. Major areas of uncertainty are the short 

length of the key study and the lack of chronic inhalation exposure studies in both animals and 

man (LaKind et al., 1999). 

VIII. References 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Coon RA, Jones RA, Jenkins LJ Jr, and Siegel J. 1970. Animal inhalation studies on ammonia, 

ethylene glycol, formaldehyde, dimethylamine, and ethanol. Toxicol. Appl. Pharmacol. 

16(3):646-655. 

Appendix D3 238 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

DePass LR, Garman RH, Woodside MD, Giddens WE, Maronpot RR, and Weil CS. 1986a. 

Chronic toxicity and oncogenicity studies of ethylene glycol in rats and mice. Fundam. Appl. 

Toxicol. 7:547-565. 

DePass LR, Woodside MD, Maronpot RR, and Weil CS. 1986b. Three-generation reproduction 

and dominant lethal mutagenesis studies of ethylene glycol in the rat. Fundam. Appl. Toxicol. 

7:566-572. 

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(CD-ROM version) Denver, CO: Micromedex, Inc. (Edition expires 4/30/96). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Laitinen J, Liesivuori J, and Savolainen H. 1995. Exposure to glycols and their renal effects in 

motor servicing workers. Occup. Med. 45(5):259-262. 

LaKind JS, McKenna EA, Hubner RP, and Tardiff RG. 1999. A review of the comparative 

mammalian toxicity of ethylene glycol and propylene glycol. Crit. Rev. Toxicol. 29(4):331-365. 

Neeper-Bradley TL, Tyl RW, Fisher LC, Kubena MF, Vrbanic MA, and Losco PE. 1995. 

Determination of a No-Observed-Effect level for developmental toxicity of ethylene glycol 

administered by gavage to CD rats and CD-1 mice. Fundam. Appl. Toxicol. 27:121-130. 

Tyl RW, Ballantyne B, Fisher LC, Fait TA, Dodd DE, Klonne DR, Pritts IM, and Losco PE. 

1995a. Evaluation of the developmental toxicity of ethylene glycol aerosol in CD-1 mice by 

nose-only exposure. Fundam. Appl. Toxicol. 27:49-62. 

Tyl RW, Ballantyne B, Fisher LC, Fait DL, Savine TA, Dodd DE, Klonne DR, and Pritts IM. 

1995b. Evaluation of the developmental toxicity of ethylene glycol aerosol in the CD rat and 

CD-1 mouse by whole-body exposure. Fundam. Appl. Toxicol. 24:57-75. 

Tyl RW, Price CJ, Marr MC, Myers CB, Seely JC, Heindel JJ, and Schwetz BA. 1993. 

Developmental toxicity evaluation of ethylene glycol by gavage in New Zealand white rabbits. 

Fundam. Appl. Toxicol. 20:402-412. 

Wills JH, Coulston F, Harris ES, McChesney EW, Russell JC, and Serrone DW. 1974. Inhalation 

of aerosolized ethylene glycol by man. Clin. Toxicol. 7:463-476. 

Appendix D3 239 Ethylene Glycol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY

ETHYLENE GLYCOL MONOETHYL ETHER 

(2-ethoxyethanol; EGEE) 

CAS Registry Number: 110-80-5 

I. Chronic Toxicity Summary

Inhalation reference exposure level 70 μg/m3

 (20 ppb) 

Critical effect(s) Testicular degeneration and decreased 

hemoglobin in rabbits 

Hazard index target(s) Reproductive system; hematopoietic system 

II. Chemical Property Summary (from HSDB, 1996; 1999)

 

Description Colorless liquid; sweet, pleasant, ether-like odor 

Molecular formula C4H10O2

Molecular weight 90.12 

Boiling point 135°C 

Vapor pressure 3.8 torr @ 20°C; 5.31 torr at 25°C 

Solubility Miscible with water and organic solvents 

Conversion factor 3.69 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources

Ethylene glycol monoethyl ether (EGEE) is a widely used solvent for nitrocellulose, dyes, inks, 

resins, lacquers, paints, and varnishes (HSDB, 1996). It is also a component of many cleaning 

agents, epoxy coatings, paints, hydraulic fluid, and is an anti-icing fuel additive in aviation. 

EGEE is also a chemical intermediate in the production of another solvent, ethylene glycol 

monoethyl ether acetate. The specific annual statewide industrial emissions of EGEE from 

facilities reporting under the Air Toxics Hot Spots Act in California based on the most recent 

inventory were estimated to be 443,748 pounds (CARB, 1999). (Many industries did not report 

emissions of specific glycol ethers. Thus there were also emitted 2,922,744 pounds of the 

general category glycol ethers, which can include EGEE.) 

IV. Effects of Human Exposure 

Sperm quality was examined in 37 workers exposed to EGEE by skin contact and/or inhalation 

in two buildings (Clapp et al., 1987; Ratcliffe et al., 1989). Exposure levels ranged from 

undetectable to 24 ppm with an average exposure level of 6 ppm in one building and 11 ppm in 

the other. A statistically significant difference in mean sperm count was observed between the 

Appendix D3 240 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

37 exposed male workers and 39 unexposed male workers. Semen volume and pH, viability, 

motility, velocity, and morphology were not significantly different between the two groups. The 

primary metabolite of EGEE, ethoxyacetic acid, was identified in the urine of exposed but not 

control workers. Both exposed and control subjects had significantly lower sperm counts than 

historical controls. Furthermore, members of both groups may have been exposed to other 

compounds including metals, solvents, heat, and vibration. 

Welch and Cullen (1988) evaluated shipyard painters exposed to ethylene glycol ethers (EGEE 

and EGME). Air concentrations at the workplace were estimated based on 102 samples over six 

shifts in Sparer et al. (1988). Time –weighted average (TWA) exposures to EGEE ranged from 

0 to 80.5 mg/m3

 with a mean of 9.9 mg/m3. TWA exposures to EGME ranged from 0 to 17.7 

mg/m3

 (mean = 2.6 mg/m3). The authors note that during the time period of measurement, 

painting activities were unusually low and previous NIOSH analyses indicated considerably 

higher exposures. Ninety-four painters and 55 controls answered a medical and environmental 

exposure questionnaire including work history and provided blood, urine, and in some cases 

semen samples. Mean hemoglobin levels, total cell counts and differential counts did not differ 

between exposed and control. However, the authors found that the lowest quartile of 

hemoglobin was mostly painters and the lowest polymorphonuclear leukocyte counts were in 

painters. Nine painters were considered anemic and five were considered granulocytopenic. The 

authors note that the absence of a significant difference in the group as a whole and the inability 

to detect a dose-response pattern in the exposed group make a strong conclusion unwarranted. 

Welch et al. (1988) evaluated the semen samples from the workers in the cohort from Welch and 

Cullen (1988). Sperm concentration, velocity, motility, morphology, morphometry, and viability 

were measured. Although not statistically significant, the measures of sperm count tended to be 

lower in the painters with a p = 0.10 for density and p = 0.11 for count. When nonsmokers were 

analyzed separately from smokers, the number of oligospermic painters was larger than that in 

controls at p = 0.05. There was no difference between controls and exposed men who were 

smokers. The authors state that although mean values of sperm count did not differ significantly 

between controls and exposed groups, biologically important differences were seen when the 

proportion of men with oligospermia was examined. The proportion of painters with 

azoospermia was 5% with only 1% expected based on other population surveys. The authors 

note that to create a dose-response model for an effect of glycol ethers on semen parameters 

would require description of the exposure of each individual 3 to 6 months prior to sampling. 

The painters moved frequently from one exposure area to the next, making exposure assessment 

particularly difficult in this cohort. 

Cullen et al. (1992) conducted a histopathologic analysis of the bone marrow and circulating 

blood cells in the workers previously examined in Welch et al. (1988). The objects of the study 

included : 1) to exclude other causes for granulocytopenia and depressed hemoglobin levels 

noted in some painters exposed to ethylene glycol ethers, 2) to determine if subclinical evidence 

of hematologic damage is present in healthy coworkers, and 3) to identify host or exogenous 

factors which may increase the risk of hematologic damage in glycol ether exposed painters. 

Workers were grouped as follows: Group I consisted of those painters that had anemia or 

granulocytopenia in the Welch and Cullen (1988) study; Group II consisted of exposed painters 

with normal hematology; Group III consisted of unexposed controls. A battery of hematologic 

Appendix D3 241 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

and biochemical parameters were measured and a questionnaire was completed to determine 

occupational exposure status, health status and drug and alcohol consumption. All hematologic 

parameters were normal in all groups. Tests of liver, renal, and thyroid function were normal in 

all groups. Bone marrow histology showed no differences between groups. One biochemical 

parameter, pyruvate kinase activity, was lower in Group I than Groups II and III (p = 0.05). 

Depression of red cell pyruvate kinase did not vary by race and was lower in every subject in 

Group I by more than one standard deviation. Low pyruvate kinase is the most consistent red 

cell enzyme defect noted in acquired hematologic disorders. 

V. Effects of Animal Exposure 

Sprague-Dawley rats (15/sex/group) and New Zealand white rabbits (10/sex/group) were 

exposed to 0, 25, 103, or 403 ppm EGEE by inhalation for 6 hours/days, 5 days/week, for 13 

weeks (Barbee et al., 1984). Animals were physically examined weekly and, at the end of the 

study, hematology, clinical chemistry, and histopathological examination were performed. No 

histopathological changes in the respiratory tract were found. Among rabbits, body weight was 

reduced in the high-dose group males and females. In the 25 ppm dose group, adrenal weight 

was reduced significantly among males, although this effect was not found to be dose-related. 

Among males in the high-dose group, testes weights were significantly reduced with a 

corresponding degenerative change to the seminiferous tubule epithelium. No effect on 

spermatogenic activity was found, however. Significant hematological effects observed at the 

high-dose included decreased hemoglobin, hematocrit, and erythrocyte count. 

Teratologic effects in pregnant rats from the inhalation of EGEE were reported (Tinston et al., 

1983a). The results of this study were presented in summary form (Doe, 1984). Wistar rats 

(24/group) were exposed to target concentrations of 0, 10, 50, or 250 ppm EGEE for 6 hours/day 

during gestational days 6-15 and the animals were sacrificed on day 21. Maternal toxicity was 

observed in the high-dose group with decreased hemoglobin, hematocrit, and mean corpuscular 

volume. Significant increases in preimplantation loss occurred in the 10 and 50 ppm dose 

groups, however the absence of this effect at 250 ppm indicated a poor dose-response, and 

because implantation occurred on the first day of exposure, the relatedness of the effect to 

exposure is in question. Post-implantation loss was also increased in the mid-dose group, 

however, no corresponding decrease in intrauterine death was observed in this group. Minor 

skeletal defects, particularly delayed ossification, were widely observed in the fetuses of mothers 

exposed to 250 ppm EGEE. Delayed ossification of the cervical vertebrae and sternebrae and the 

presence of extra ribs was significantly increased in both the 50 and 250 ppm dose groups. 

Teratologic effects on pregnant rabbits from inhalation exposure to EGEE were also reported 

(Tinston et al., 1983b; also summarized by Doe, 1984). Dutch rabbits (24/group) were exposed 

to 0, 10, 50, or 175 ppm EGEE for 6 hours/day during gestational days 6-18, with sacrifice 

occurring on gestational day 29. There were no indications of maternal toxicity or litter effects. 

A statistically significant increase in minor defects and skeletal variants was found in fetuses in 

the 175 ppm dose group. Other slightly increased incidences of defects in the lower dose groups 

alone, including extra ribs and partial ossification of the vertebrae, were not considered 

treatment-related. 

Appendix D3 242 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Behavioral teratogenic effects were examined in pregnant Sprague-Dawley rats (14 or 15/dose 

group) exposed to 0 or 100 ppm EGEE for 7 hours/day through gestational days 7-13 (early) or 

days 14-20 (late) (Nelson et al., 1981). No maternal toxicity was observed and fetal weights 

were unchanged, although mean gestational length was increased in rats exposed on gestational 

days 14-20. Six tests (ascent, rotorod, open field, activity wheel, avoidance conditioning, and 

operant conditioning) were selected to measure motor, sensory, and cognitive function at several 

stages of development. The offspring of the rats exposed during days 7-13 exhibited impaired 

performance on the rotorod test (a test of neuromuscular ability) and increased latency in an open 

field test (a test of exploratory activity) as compared to controls. The offspring of rats exposed 

during days 14-20 of gestation exhibited decreased activity on an activity wheel (a test of 

circadian activity). Also, avoidance conditioning revealed that these pups received shocks of a 

greater number and duration than controls. Neurochemical differences between the prenatally 

exposed and control pups were measured in newborns and in pups 21 days of age. In newborns 

from both EGEE-exposed groups, total brain norepinephrine was decreased. In 21-day old pups 

of both groups, norepinephrine and dopamine levels in the cerebrum were increased. Serotonin 

level was increased in the cerebrum of the late exposure group only. The authors concluded that 

there were behavioral and neurochemical alterations in offspring of rats following prenatal 

exposure to 100 ppm EGEE, however the study design was inadequate to detect gross teratologic 

anomalies. In a dose range-finding study, two sets of pregnant rats (3-4/group) were exposed 

during the gestational days 7-13 or 14-20 to 0, 200 (late group only), 300, 600, 900, or 1200 ppm 

EGEE for 7 hours/day. Increased fetal and pup mortality was observed in all groups exposed to 

EGEE. 

Behavioral and neurochemical effects on the offspring of pregnant S-D rats exposed to 0 or 200 

ppm EGEE on gestational days 7-13 were reported (Nelson et al., 1982a; Nelson et al., 1982b). 

Pregnancy duration was significantly increased in exposed dams. Significantly increased levels 

of norepinephrine and dopamine were observed in the 21-day old offspring of EGEE-exposed 

animals. Behavioral changes in pups of treated dams included decreased neuromotor ability and 

decreased activity. 

An investigation into teratologic effects of EGEE was conducted by exposing pregnant rats and 

rabbits to EGEE by inhalation on gestational days 0-19 (Andrew et al., 1981). Rats (37/group) 

were exposed to 0, 202, or 767 ppm EGEE for 7 hours/day. All fetuses were resorbed and 

maternal weight gain was reduced in the high-dose group. In the mid-dose group, a decrease in 

fetal weight and size (crown-rump length) was observed. Minor skeletal defects and variants and 

cardiovascular defects were increased in the mid-dose group. Rabbits (29/group) were exposed 

to 0, 16, or 617 ppm EGEE for 4 hours/day. Maternal weight gain and food intake were 

decreased in exposed animals. The incidence of fetal resorptions was increased in both the midand high-dose group animals. Major cardiovascular defects and minor skeletal defects (extra 

ribs, delayed ossification) were significantly increased in the mid-dose group. Andrew et al.

(1981) also examined reproductive effects by exposing female Wistar rats (37/group) to 1, 150, 

or 649 ppm EGEE 7 hours/day, 5 days/week for 3 weeks before mating with untreated males. 

No significant effects were observed. 

Appendix D3 243 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Reference Exposure Level 

Study Barbee et al., 1984 

Study population Rabbits 

Exposure method Discontinuous inhalation 

Critical effects Testicular degeneration and decreased 

 hemoglobin levels 

LOAEL 403 ppm 

NOAEL 103 ppm 

Exposure continuity 6 hr/day, 5 days/week 

Exposure duration 13 weeks 

Average experimental exposure 18.4 ppm (68 mg/m3

) for the NOAEL group 

Human equivalent concentration 18.4 ppm (68 mg/m3

) for the NOAEL group 

 (gas with systemic effects, based on RGDR = 

 1.0 using default assumption that lambda (a) = 

 lambda (h)) 

Subchronic uncertainty factor 10 

LOAEL uncertainty factor 1 

Interspecies factor 10 (see explanation below) 

Intraspecies factor 10 

Cumulative uncertainty factor 1000 

Inhalation reference exposure level 0.02 ppm (20 ppb, 0.07 mg/m3 , 70 μg/m3

) 

The reproductive effects observed in the subchronic inhalation study of Barbee et al. (1984) were 

determined by the US EPA (U.S. EPA, 1990) to be the most sensitive endpoints due to EGEE 

exposure and resulted in a reference concentration (RfC) of 0.2 mg/m3 (0.06 ppm). OEHHA 

staff concurred regarding the basis of the U.S. EPA RfC but differed in the application of the 

interspecies uncertainty factor. Reduced testes weight and testicular degeneration were found in 

rabbits exposed to EGEE at 403 ppm for 13 weeks. Changes in hematological parameters 

including decreased hemoglobin, hematocrit, and erythrocyte count were also observed at this 

dose. A gas:extrarespiratory effect ratio of 1.0 was used to calculate a human equivalency 

concentration (HEC) in the absence of information relating the effect in rabbits relative to 

humans. 

For a comparison with the proposed REL of 60 ppb (200 μg/m3) based on testicular 

degeneration, a REL can be calculated from the LOAEL of 202 ppm observed in the teratology 

study of Andrew et al. (1981). The 7 h exposure to 202 ppm is time-adjusted to a continuous 

exposure of 59 ppm. Using a RGDR of 1 for a systemic effect, a UFL of 10, a UFA of 3 and a 

UFH of 10 results in an estimated REL of 200 ppb (700 μg/m3

). Nelson et al. (1981) found a 

LOAEL of 100 ppm for neurobehavioral developmental toxicity in rats exposed 7 hours per day 

on days 7 to 13 of gestation. The equivalent continuous exposure is 29 ppm. Using an RGDR of 

1, a LOAEL UF of 10, an interspecies UF of 3, and an intraspecies UF of 10 results in a REL of 

100 ppb (400 μg/m3

). 

Although reproductive toxicity has been reported in male workers occupationally exposed to 

EGEE (Clapp et al., 1987; Ratcliffe et al., 1989), potential confounding factors, particularly 

Appendix D3 244 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

exposure to other compounds, make the study inadequate for the development of the reference 

exposure level. However, for another comparison with the proposed REL of 60 ppb, if only 

EGEE caused the adverse reproductive effect, use of a mean concentration between the 2 

buildings of 8 ppm for workplace exposure, extrapolation to an equivalent continuous exposure 

of 3 ppm, and division by 10 for a LOAEL (serious effect) and 10 for intraspecies variability 

result in a REL of 30 ppb (100 μg/m3

). 

Another comparison with the proposed REL of 60 ppb can be made using the study of Welch et 

al. (1988), who studied shipyard painters exposed to both EGEE and EGME. The authors 

examined the semen of 73 painters and 40 non-exposed shipyard employees. The men supplied 

demographic characteristics, medical conditions, personal habits, and reproductive history; 

underwent a physical examination; and provided a semen sample. An industrial hygiene survey 

showed that the painters were exposed to EGEE at a time-weighted average (TWA) 

concentration varying from 0 to 80.5 mg/m3

 (mean = 9.9 mg/m3, and to EGME at a TWA 

concentration varying from 0 to 17.7 mg/m3 (mean = 2.6 mg/m3). The painters had an increased 

prevalence of oligospermia and azoospermia and an increased odds ratio for a lower sperm count 

per ejaculate. (The results were controlled for smoking.) Adding the mean levels together 

results in a total glycol ether concentration of 12.5 mg/m3, which is equivalent to a continuous 

exposure of 4.5 mg/m3. Division by a UF of 10 for a LOAEL and by another of 10 for 

intraspecies variability results in a REL of 40 μg/m3 (10 ppb). A similar REL would be 

calculated using the report by Cullen et al. (1992) of depression in red cell pyruvate kinase 

among anemic and granulocytopenic painters. Since exposure in these studies was to both 

EGEE and EGME, and exposure assessment was made difficult by frequent job movement and 

other factors, these studies were not deemed suitable for developing a REL. However, the 

possibility that humans are more susceptible to EGEE toxicity is raised by the series of studies 

by Welch et al. (1988) and Welch and Cullen (1988) such that we have deviated from the RfC 

and opted to use an interspecies uncertainty factor of 10 rather than 3 as would usually be the 

case with an HEC adjustment. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for EGEE include the availability of subchronic inhalation 

exposure data from a well-conducted study with histopathological analysis, and the observation 

of a NOAEL. The observation in several studies noted above of both hematological 

abnormalities and sperm abnormalities in exposed workers, although difficult to use in a 

quantitative risk assessment, provide support for the REL developed from animals. In addition, 

several comparative calculations indicate that RELs based on other studies are generally in 

agreement with that based on Barbee et al. (1984). Major areas of uncertainty are the lack of 

adequate human exposure data and the lack of chronic inhalation exposure studies. 

Appendix D3 245 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VIII. References 

Andrew FD, Buschbom RL, Cannon WC, et al. 1981. Teratologic assessment of ethylbenzene 

and 2-ethoxyethanol. Prepared for the National Institute for Occupational Safety and Health 

under Contract 210-79-0037, Battelle Pacific Northwest Laboratories, Richland WA 99352. 

Microfiche #OTS0513150. 

Barbee SJ, Terrill JB, DeSousa DJ, and Conaway CC. 1984. Subchronic inhalation toxicology of 

ethylene glycol monoethyl ether in the rat and the rabbit. Environ. Health Perspect 57:157-163. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Clapp DE, Smallwood AW, Moseley C, and DeBord KE. 1987. Workplace assessment of 

exposure to 2-ethoxyethanol. Appl. Ind. Hyg. 2:183-187. 

Cullen MR, Rado T, Waldron JA, Sparer J, Welch LS. 1983. Bone marrow injury in 

lithographers exposed to glycol ethers and organic solvents used in multicolor offset and 

ultraviolet curing printng processes. Arch. Environ. Health 38(6):347-354. 

Cullen MR, Solomon LR, Pace PE, Buckley P, Duffy TP, McPhedran P, Kelsey KT, Redlich 

CA. 1992. Morphologic, biochemical, and cytogenetic studies of bone marrow and circulating 

blood cells in painters exposed to ethylene glycol ethers. Environ. Res. 59(1):250-264. 

Doe JE. 1984. Ethylene glycol monoethyl ether and ethylene glycol monoethyl ether acetate 

teratology studies. Environ. Health Perspect. 57:33-41. 

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Nelson BK, Brightwell WS, and Setzer JV. 1982a. Prenatal interactions between ethanol and the 

industrial solvent 2-ethoxyethanol in rats: maternal and behavioral teratogenic effects. 

Neurobehav. Toxicol. Teratol. 4:387-94. 

Nelson BK, Brightwell WS, Setzer JV, and O’Donohue TL. 1982b. Prenatal interactions 

between ethanol and the industrial solvent 2-ethoxyethanol in rats: Neurochemical effects in the 

offspring. Neurobehav. Toxicol. Teratol. 4:395-401. 

Nelson BK, Brightwell WS, Setzer JV, Taylor BJ, Hornung RW, and O’Donohue TL. 1981. 

Ethoxyethanol behavioural teratology in rats. Neurotoxicology 2:231-249. 

Ratcliffe JM, Schrader SM, Clapp DE, Halperin WE, Turner TW, and Hornung R.W. 1989. 

Semen quality in workers exposed to 2-ethoxyethanol. Br. J. Ind. Med. 46:399-406. 

Sparer J, Welch LS, McManus K, and Cullen MR. 1988. Effects of exposure to ethylene glycol 

ethers on shipyard painters: I. Evaluation of exposure. Am. J. Ind. Med. 14:497-507. 

Appendix D3 246 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Tinston DJ, Doe JE, Godley MJ, Head LK, Killick M, Litchfield MH, and Wickramaratne GA. 

1983a. Ethylene glycol monoethyl ether (EE): teratogenicity study in rats. Imperial Chemical 

Industries PLC, Report No. CTL/P/761 to Chemical Manufacturers Association. 

Tinston DJ, Doe JE, Thomas M, and Wickramaratne GA. 1983b. Ethylene glycol monoethyl 

ether (EE): inhalation teratogenicity study in rabbits. Imperial Chemical Industries PLC. Report 

No. CTL/P/776 to Chemical Manufacturers Association. 

U.S. EPA (United States Environmental Protection Agency). 1990. Documentation of the 

reference concentration for chronic inhalation exposure (RfC) for 2-ethoxyethanol. Integrated 

Risk Information System (IRIS on-line). Washington, DC: US EPA. 

Welch LS, Cullen MR. 1988. Effect of exposure to ethylene glycol ethers on shipyard painters: 

III. Hematologic effects. Am. J. Ind. Med. 14(5):527-236. 

Welch LS, Schrader SM, Turner TW, Cullen MR. 1988. Effects of exposure to ethylene glycol 

ethers on shipyard painters: II. Male reproduction. Am. J. Ind. Med. 14(5):509-526. 

Appendix D3 247 Ethylene Glycol Monoethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

ETHYLENE GLYCOL MONOETHYL ETHER 

ACETATE 

(EGEEA; 1-acetoxy-2-ethoxyethane; 2-ethoxyethanol acetate; 2-ethoxyethyl acetate; acetic acid, 

2-ethoxyethyl ester; beta-ethoxyethyl acetate; Cellosolve® acetate; ethoxy acetate; ethyl 

Cellosolve® acetate; Poly-solv® EE acetate; ethyl glycol acetate; oxitol acetate) 

CAS Registry Number: 111-15-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 300 μg/m3

 (60 ppb)

Critical effect(s) Teratogenicity and fetotoxicity in rabbits 

Hazard index target(s) Development 

II. Chemical Property Summary (HSDB, 1996) 

Description Colorless liquid 

Molecular formula C6H12O3 

Molecular weight 132.16 g/mol 

Boiling point 156°C 

Vapor pressure 2 torr @ 20°C 

Solubility Soluble in water (229 g/l at 20°C); sol. in alcohol, 

ether, acetone; miscible with olive oil, aromatic 

hydrocarbons 

Conversion factor 5.41 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Ethylene glycol monoethyl ether acetate (EGEEA) is used in automobile lacquers where it 

retards “blushing” and evaporation and imparts a high gloss (HSDB, 1996). It is also used as a 

solvent for nitrocellulose, oils, and resins and as a component of varnish removers and wood 

stains. EGEEA is also used in the treatment of textiles and leather. The annual specific 

statewide industrial emissions of EGEEA from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 66,851 pounds 

(CARB, 1999).

IV. Effects of Human Exposure 

Appendix D3 248 Ethylene Glycol Monoethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

No studies relating exposure to EGEEA to adverse health effects in humans were located in the 

literature. 

Ten male volunteers were exposed to EGEEA by inhalation. Five were exposed to 14, 28, and 

50 mg EGEEA/m3

 and five to 28 mg/m3 for 4 hours (Groeseneken et al., 1987a). Twenty-two 

percent of the absorbed dose was eliminated in the urine as ethoxyacetic acid within 42 hours. In 

another study, male volunteers exposed to EGEEA by inhalation under various conditions were 

found to eliminate some in the form of ethylene glycol monoethyl ether (EGEE) (Groeseneken et 

al., 1987b).

V. Effects of Animal Exposure 

Pregnant rabbits (24 or 25/group) were exposed to 0, 25, 100, or 400 ppm EGEEA by inhalation 

for 6 hours/day on gestational days 6-18 (Tinston et al., 1983; reviewed in Doe, 1984). The 

animals were killed on gestational day 29. Maternal effects (decreased weight gain, decreased 

food consumption, decreased hemoglobin) were observed in the high-dose group. The number 

of rabbits with total fetal resorptions was increased in the 400 ppm dose group, accompanied by 

a decrease in weight in surviving fetuses. A reduction in average fetal weight was also observed 

at 100 ppm EGEEA, but this effect may relate to the increased litter size among dams in this 

dose group. Evidence of teratogenicity was observed in the 400 ppm dose group, with increased 

major malformations of the vertebral column. Both 400 and 100 ppm EGEEA were found to be 

fetotoxic as indicated by retarded ossification. No statistically significant effects were observed 

in the 25 ppm dose group, although a single case of a major defect (kidney agenesis) was 

observed in both the 25 and 400 ppm EGEEA dose groups. 

Rats (10/sex/dose) and rabbits (2/sex/dose) were exposed for 4 hours/day, 5 days/week for 10 

months to 0 or 200 ppm EGEEA (Truhaut et al., 1979). Observation of body weight gain, 

hematology, clinical chemistry, and gross pathology revealed no toxic effects among treated 

animals. Among male rats and rabbits, “discrete lesions of tubular nephritis with clear 

degeneration of the epithelium with hyaline and granular tubular casts” were observed. Four 

hour exposure to 2000 ppm EGEEA resulted in transient hemoglobinuria and hematuria in 

rabbits (2/sex/dose), but not rats (10/sex/dose). No pathological lesions were observed following 

a 2 week observation period. 

Dogs were exposed to 600 ppm EGEEA for 7 hours/day for 120 days (Carpenter et al., 1956; 

Gingell et al., 1982). Hematological, clinical chemistry, and histopathological examination 

revealed no adverse effects. 

Pregnant rats and rabbits (24/group) were exposed to nominal concentrations of 0, 50, 100, 200 

or 300 ppm EGEEA by inhalation during gestational days 6-15 and sacrificed on gestational day 

21 (Union Carbide Corporation, 1984). Maternal effects in rats included increased absolute liver 

weights (all treated groups); increased relative liver weights, and decreased RBC count, 

hemoglobin, hematocrit, and RBC size (all but low-dose group); decreased food consumption, 

increased white blood cell count, and decreased platelet count (200 and 300 ppm groups). An 

increase in the number of non-viable implantations per litter was observed at 300 ppm and 

Appendix D3 249 Ethylene Glycol Monoethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

decreased average fetal body weight per litter was observed at 200 and 300 ppm EGEEA. 

Visceral and skeletal malformations were widely observed at both 200 and 300 ppm EGEEA. 

Among rabbits, maternal effects included decreased platelets (100, 200, and 300 ppm); decreased 

weight gain, decreased gravid uterine weight, increased number of dams with non-viable 

implants, and increased number of non-viable implants per litter (200 and 300 ppm); increased 

occult blood, increased mean corpuscular volume, decreased corpora lutea/litter and increased 

early resorptions/litter (300 ppm). Visceral and skeletal malformations were observed in the 

100, 200, and 300 ppm EGEEA dose groups. 

Pregnant rats were exposed to 0, 130, 390, or 600 ppm EGEEA for 7 hours/day on gestational 

days 7-15 (Nelson et al., 1984). Dams were sacrificed on day 20. Complete resorption of litters 

was observed at 600 ppm. Skeletal and cardiovascular defects and decreased fetal weight and 

fetal resorptions were observed at 390 ppm EGEEA. Reduced fetal weights were also observed 

at 130 ppm EGEEA. 

Ethylene glycol monoethyl ether acetate (0.35 ml = 2.6 mmole/treatment) or water was applied 

to the shaved skin of pregnant rats four times daily on days 7 to 16 gestation (Hardin et al., 

1984). EGEEA treated rats showed reduced body weight (from litter resorption) and 

significantly fewer live fetuses per litter. Litters from treated dams also showed significantly 

increased visceral malformations and skeletal variations.

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Tinston et al., 1983 

Study population Rabbits 

Exposure method Discontinuous inhalation exposure 

Critical effects Fetotoxicity 

LOAEL 100 ppm 

NOAEL 25 ppm 

Exposure continuity 6 hours/day, 7 days/week 

Exposure duration 13 days 

Average experimental exposure 6.2 ppm for NOAEL group (25 x 6/24) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies factor 10 

Intraspecies factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.06 ppm (60 ppb, 0.03 mg/m3

, 300 μg/m3) 

A review of the literature on the toxicity of EGEEA indicates that the most sensitive endpoint of 

toxicity is that seen in experimental animals showing developmental effects from inhalation 

exposure during pregnancy. There are no adequate data associating exposures in humans with 

toxic effects for the development of a chronic reference exposure level. Separate studies in 

animals have demonstrated developmental toxicity. Reduced fetal weights were observed in rats 

exposed to 130 ppm EGEEA on gestational days 7-15 (Nelson et al., 1984). Skeletal and 

Appendix D3 250 Ethylene Glycol Monoethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

cardiovascular defects were observed at the next higher dose of 390 ppm EGEEA, and all litters 

were resorbed in the high-dose group. Visceral and skeletal defects were observed in all but the 

low-dose group (50 ppm EGEEA) in the litters of rabbit dams exposed to EGEEA on gestational 

days 6-15 (Union Carbide Corporation, 1984). Fetotoxicity, as indicated by retarded bone 

development, was observed in all but the low-dose group (25 ppm EGEEA) in the litters of 

rabbit dams exposed on gestational days 6-18 (Tinston et al., 1983). The lowest dose levels 

showing developmental toxicity are those reported by Union Carbide Corporation (1984) and 

Tinston et al. (1983), with 100 ppm EGEEA showing developmental defects in the offspring of 

exposed dams. Since only the Tinston et al. (1983) study also showed an exposure level without 

effect (a NOAEL), this study has been selected for the development of the chronic REL. 

VII. Data Strengths and Limitations for Development of the REL

Strengths of the database for EGEEA include the large number of animal studies available. 

Limitations include the lack of any human data for exposures longer than 4 hours and the lack of 

sperm count studies, a critical effect for the related compounds, EGEE and EGME. However, 

the REL calculated is similar to that for EGEE which is based on testicular degeneration. 

VIII. References 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Carpenter CP, Pozzani UC, Weil CS, Nair JH, Keck GA, and Smyth HF. 1956. The toxicity of 

butyl cellosolve solvent. Arch. Ind. Health 14:114-31. 

Doe JE. 1984. Ethylene glycol monoethyl ether and ethylene glycol monoethyl ether acetate 

teratology studies. Environ. Health Perspect. 57:33-41. 

Gingell R, Boatman RJ, Bus J, Cawley TJ, Knaak JB, Krasavage WJ, Skoulis NP, Stack CR, and 

Tyler TR. 1982. Glycol ethers and other selected glycol derivatives. In: Patty’s Industrial 

Hygiene and Toxicology. Fourth ed. Clayton, G.D. and Clayton, F.E. (eds.). New York: John 

Wiley Sons. 

Groeseneken D, Veulemans H, Masschelein R, and Van Vlem E. 1987a. Ethoxyacetic acid: a 

metabolite of ethylene glycol monoethyl ether acetate in man. Br. J. Ind. Med. 44:488-493. 

Groeseneken D, Veulemans H, Masschelein R, and Van Vlem E. 1987b. Pulmonary absorption 

and elimination of ethylene glycol monoethyl ether acetate in man. Br. J. Ind. Med. 44:309-316. 

Hardin BD, Goad PT, and Burg JR. 1984. Developmental toxicity of four glycol ethers applied 

cutaneously to rats. Environ. Health Perspect. 57:69-74. 

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Appendix D3 251 Ethylene Glycol Monoethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Nelson BK, Setzer JV, Brightwell WS, Mathinos PR, Kuczuk MH, Weaver TE, and Goad PT. 

1984. Comparative inhalation teratogenicity of four glycol ether solvents and an amino 

derivative in rats. Environ. Health Perspect. 57:261-271. 

Tinston DJ, Doe JE, Killick M, and Thomas M. 1983. Ethylene glycol monoethyl ether acetate 

(EEAc): Inhalation teratogenicity study in rabbits. Report No. CTL/P/840. Conducted by the 

Imperial Chemical Industries PLC, Central Toxicology Laboratory, Alderley Park, Macclesfield, 

Cheshire, UK. 

Truhaut R, Dutertre-Catella H, Phu-Lich N, and Huyen VN. 1979. Comparative toxicological 

study of ethylglycol acetate and butylglycol acetate. Toxicol. Appl. Pharmacol. 51:117-27. 

Union Carbide Corporation. 1984. Bushy Run Research Center. Teratologic evaluation of 

Cellosolve acetate in Fischer 344 rats and New Zealand white rabbits following inhalation 

exposure. EPA Doc. No. FYI-AX-1184-0360 (Fiche No. 0000360-0). 

Appendix D3 252 Ethylene Glycol Monoethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY

ETHYLENE GLYCOL MONOMETHYL ETHER 

(EGME; 2-methoxyethanol; 1-hydroxy-2-methoxyethane; methyl cellosolve) 

CAS Registry Number: 109-86-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 60 μg/m3

 ( 20 ppb) 

Critical effect(s) Testicular toxicity in rabbits 

Hazard index target(s) Reproductive system 

II. Physical and Chemical Properties (HSDB, 1995) 

Description Colorless liquid 

Molecular formula C3H8O2

Molecular weight 76.09 

Density 0.965 g/cm3

 @ 20° C 

Boiling point 125°C 

Melting point -85.1°C 

Vapor pressure 6.2 torr @ 20°C 

Solubility Miscible with water, alcohol, benzene, 

 ether, acetone 

Conversion factor 1 ppm = 3.1 mg/m3

 @ 25°C 

III. Major Uses and Sources 

Ethylene glycol monomethyl ether (EGME) is used as a solvent for cellulose acetate and resins 

(HSDB, 1995) as well as a solvent in the semiconductor industry. It is also used in dyeing 

leather and in the manufacture of photographic film. EGME is used as an anti-freeze in jet fuels. 

Quick drying varnishes, enamels, nail polishes, and wood stains may also contain EGME. The 

specific annual statewide industrial emissions of EGME from facilities reporting under the Air 

Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 7398 

pounds (CARB, 1999). (Many industries did not report emissions of specific glycol ethers. 

Thus there were also emitted 2,922,744 pounds of the general category glycol ethers, which can 

include EGME.) 

IV. Effects of Human Exposure 

Human exposures to ethylene glycol monomethyl ether have been associated with hematological 

and neurological abnormalities. To determine whether employees potentially exposed to 

ethylene glycol monomethyl ether during manufacturing and packaging had a higher prevalence 

of anemia, leukopenia, or sterility than an in-plant comparison group, a cross-sectional study was 

Appendix D3 253 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

conducted. Blood samples on 65 of 97 potentially exposed and control white males, and semen 

samples from a subset of 15 were analyzed. No gross abnormalities or clinically meaningful 

differences in hematological or fertility indices were noted. Decreased testicular size was 

reported in workers (who were exposed to an 8-hour TWA concentration of 0.42 ppm EGME or 

less) but it was not statistically significant (Cook et al., 1982). 

Cullen et al. (1983) studied possible bone marrow toxicity of workplace substances including 

dipropylene glycol monomethyl ether, EGME, and various aliphatic, aromatic and halogenated 

hydrocarbons used for offset and ultraviolet cured multicolor printing. Evaluation of seven coworkers of a printer with aplastic anemia indicated normal peripheral blood, but bone marrow 

specimens demonstrated clear patterns of injury in three while the others had nonspecific signs of 

marrow effect. The authors could not assign the changes to known risk factors and concluded 

that further evaluation of possible bone marrow toxicity resulting from exposure to glycol ethers 

and ultraviolet curing printing processes was warranted. This was done to some extent in their 

studies on shipyard painters below. 

Welch and Cullen (1988) evaluated shipyard painters exposed to ethylene glycol ethers (EGEE 

and EGME). Air concentrations at the workplace were estimated based on 102 samples over six 

shifts in Sparer et al. (1988). Time –weighted average (TWA) exposures to EGEE ranged from 0 

to 80.5 mg/m3 with a mean of 9.9 mg/m3. TWA exposures to EGME ranged from 0 to 17.7 

mg/m3 (mean = 2.6 mg/m3). The authors note that during the time period of measurement, 

painting activities were unusually low and previous NIOSH analyses indicated considerably 

higher exposures. Ninety-four painters and 55 controls answered a medical and environmental 

exposure questionnaire including work history and provided blood, urine, and in some cases 

semen samples. Mean hemoglobin levels, total cell counts and differential counts did not differ 

between exposed and control. However, the authors found that the lowest quartile of 

hemoglobin was mostly painters and the lowest polymorphonuclear leukocyte counts were in 

painters. Nine painters were considered anemic and five were considered granulocytopenic. The 

authors note that the absence of a significant difference in the group as a whole and the inability 

to detect a dose-response pattern in the exposed group makes a strong conclusion unwarranted. 

Welch et al. (1988) evaluated the semen samples from the workers in the cohort from Welch and 

Cullen (1988). Sperm concentration, velocity, motility, morphology, morphometry, and viability 

were measured. Although not statistically significant, the measures of sperm count tended to be 

lower in the painters with a p = 0.10 for density and p = 0.11 for count. When nonsmokers were 

analyzed separately from smokers, the number of oligospermic painters was larger than that in 

controls at p = 0.05. There was no difference between controls and exposed men who were 

smokers. The authors state that although mean values of sperm count did not differ significantly 

between controls and exposed groups, biologically important differences were seen when the 

proportion of men with oligospermia was examined. The proportion of painters with 

azoospermia was 5% with only 1% expected based on other population surveys. The authors 

note that to create a dose-response model for an effect of glycol ethers on semen parameters 

would require description of the exposure of each individual 3 to 6 months prior to sampling. 

The painters moved frequently from one exposure area to the next, making exposure assessment 

particularly difficult in this cohort. 

Appendix D3 254 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Cullen et al. (1992) conducted a histopathologic analysis of the bone marrow and circulating 

blood cells in the workers previously examined in Welch et al. (1988). The objects of the study 

included : 1) to exclude other causes for granulocytopenia and depressed hemoglobin levels 

noted in some painters exposed to ethylene glycol ethers, 2) to determine if subclinical evidence 

of hematologic damage is present in healthy coworkers, and 3) to identify host or exogenous 

factors which may increase the risk of hematologic damage in glycol ether exposed painters. 

Workers were grouped as follows: Group I consisted of those painters that had anemia or 

granulocytopenia in the Welch and Cullen (1988) study; Group II consisted of exposed painters 

with normal hematology; Group III consisted of unexposed controls. A battery of hematologic 

and biochemical parameters were measured and a questionnaire was completed to determine 

occupational exposure status, health status and drug and alcohol consumption. All hematologic 

parameters were normal in all groups. Tests of liver, renal, and thyroid function were normal in 

all groups. Bone marrow histology showed no differences between groups. One biochemical 

parameter, pyruvate kinase activity, was lower in Group I than Groups II and III (p = 0.05). 

Depression of red cell pyruvate kinase did not vary by race and was lower in every subject in 

Group I by more than one standard deviation. Low pyruvate kinase is the most consistent red 

cell enzyme defect noted in acquired hematologic disorders. 

Reversible neurological symptoms (apathy, fatigue, decreased appetite) and macrocytic anemia 

were observed in a worker following occupational dermal and inhalation exposure to an average 

concentration of 35 ppm EGME for 1-1.5 years (Cohen, 1984). The worker was also exposed to 

methyl ethyl ketone and propylene glycol monomethyl ether at concentrations of 1-5 ppm and 

4.2-12.8 ppm, respectively. 

Hematologic effects were also reported in three women employed in a factory working with glue 

consisting of 70% acetone and 30% EGME (Larese et al., 1992). The women exhibited 

abnormally low white blood cell counts, relative lymphocytosis and macrocytosis. These 

hematological parameters returned to normal following cessation of exposure. 

Older case reports support findings of neurological and hematological toxicity following 

occupational exposure to EGME (Greenburg et al., 1938; Zavon, 1963; Parsons and Parsons, 

1938). 

V. Effects of Animal Exposure 

A concentration dependent decrease in testes weight was observed in male rabbits exposed to 30, 

100, or 300 ppm EGME 6 hours per day, 5 days per week for 13 weeks (Miller et al., 1983). 

Degenerative changes in the germinal epithelium were observed in male rabbits of all exposed 

groups, but were not statistically significant at 30 ppm. Two of five male rabbits exposed to 300 

ppm EGME died during the course of the study. Female rabbits were also exposed; two of five 

female rabbits exposed to 100 or 300 ppm EGME died during the course of the study. The 

animals died at different times of different causes and thus the authors were uncertain if the 

deaths were treatment related. Reduced body weight gain, pancytopenia (abnormal depression of 

all the cellular elements of the blood), and thymic atrophy were observed in rabbits of both sexes 

Appendix D3 255 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

exposed to 300 ppm EGME. No effects on the reproductive organs of the female rabbits were 

observed. 

In the same study (Miller et al., 1983) male and female rats were exposed to 30, 100, or 300 ppm 

EGME 6 hours per day, 5 days per week for 13 weeks. Moderate to severe degeneration of the 

germinal epithelium and seminiferous tubules was observed in male rats exposed to 300 ppm 

EGME. A significant decrease in body weight was observed in male rats exposed to 300 ppm 

and in female rats exposed to concentrations of EGME of 100 ppm or greater. Pancytopenia, 

lymphoid tissue atrophy, and decreased liver weights were observed in animals of both sexes 

exposed to the highest concentration. Also in the highest exposure group, mean values for total 

serum protein, albumin and globulins were lower than control values. 

Doe et al. (1983) designed a two-part study to provide a rapid assessment of the effect of glycol 

ethers on some aspects of reproduction in the rat. Exposure to EGME was by inhalation at 100 

and 300 ppm for 6 hr/day. First, pregnant females were exposed on Days 6 to 17 of gestation. 

Body weight gain was reduced in both groups. No litters were delivered in the 300-ppm group 

and only 9/20 rats in the 100-ppm group produced litters; the number, weight, and viability of the 

pups were reduced, but the pups appeared normal externally. Second, male rats were exposed 

for 10 days. There was a reduction in testicular weight accompanied by seminiferous tubular 

atrophy in the 300-ppm group. There were no effects at 100 ppm. Exposure at 300 ppm EGME 

caused significant reductions in white blood cell count, red blood cell count, hemoglobin 

concentration, hematocrit, and mean cell hemoglobin. 

More recent data point to the immune system as a key endpoint of EGME toxicity. A 

statistically significant dose-related decrease in thymus weight was observed both in male rats 

administered drinking water containing 2000 and 6000 ppm EGME (161 or 486 mg/kg/day) and 

in female rats administered drinking water containing 1600 and 4800 ppm EGME (200 or 

531 mg/kg/day) for 21 days (Exon et al., 1991). Histopathological examination revealed thymic 

atrophy and loss of demarcation between the cortex and medulla. Decreased spleen cell numbers 

were observed in female rats at both dose levels and male rats at the high dose level. Male rats 

in the high dose group exhibited a statistically significant decrease in body weight gain. 

Testicular effects were also observed in exposed male rats. 

Pregnant mice were exposed to 100, 150, or 200 mg/kg/day EGME on days 10-17 of gestation 

(Holladay et al., 1994). Thymic atrophy and inhibition of fetal thymocyte maturation were 

observed in EGME-treated offspring examined on day 18 of gestation. Also, the ability of the 

EGME-treated fetal mouse liver cells to repopulate the spleen of irradiated mice was 

significantly impaired as compared to that of control fetal mouse liver cells. 

Appendix D3 256 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Reference Exposure Level 

Study Miller et al., 1983; U.S. EPA, 1995 

Study population Rats and rabbits 

Exposure method Inhalation (0, 30, 100, or 300 ppm) 

Critical effects Decreased testes weight and degenerative 

changes in the testicular germinal 

epithelium. 

LOAEL 100 ppm 

NOAEL 30 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Average experimental exposure 5.4 ppm for NOAEL group 

Human equivalent concentration 5.4 ppm for NOAEL group (gas with 

systemic effects, based on RGDR = 1.0 

using default assumption that lambda (a) 

= lambda (h)) 

Exposure duration 13 weeks 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factors 300 

Inhalation reference exposure level 0.02 ppm (20 ppb; 0.06 mg/m3

; 60μg/m3) 

The REL is based on the same study on which U.S. EPA based its RfC. However, OEHHA 

declined to use a modifying factor because the criteria for use of such factors are not well 

described by U.S. EPA. However, since rabbits were the more sensitive species and live 6 years 

(312 weeks), a 13 week study in rabbits merits a subchronic UF of 10. 

A comparison with the proposed REL for EGME of 20 ppb (60 μg/m3) can be made using the 

occupational study of Welch et al. (1988) of the semen of shipyard painters exposed to both 

EGEE and EGME. The men supplied demographic characteristics, medical conditions, personal 

habits, and reproductive history; underwent a physical examination; and provided a semen 

sample. The painters were exposed to EGEE at a TWA concentration of 0 to 80.5 mg/m3

 (mean 

= 9.9 mg/m3

, and to EGME at a TWA concentration of 0 to 17.7 mg/m3 (mean = 2.6 mg/m3). 

The painters had an increased prevalence of oligospermia and azoospermia and an increased 

odds ratio for a lower sperm count per ejaculate compared to shipyard employees who were not 

painters. (The results were controlled for smoking.) Adding the mean exposure levels together 

results in a total glycol ether concentration (EGME + EGEE) of 12.5 mg/m3

, equivalent to a 

continuous exposure of 4.5 mg/m3. Division by a UF of 10 for a LOAEL and by another of 10 

for human intraspecies variability results in a REL of 40 μg/m3 (10 ppb), similar to the REL 

based on rabbits. Since exposure was primarily to EGEE with co-exposure to EGME, and 

exposure assessment ws difficult to quantify, this study was not deemed suitable for developing a 

REL. Nonetheless, the REL developed using this study is close in value to the proposed REL of 

20 ppb. 

Appendix D3 257 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for EGME include the availability of subchronic inhalation 

exposure data from a well-conducted study with histopathological analysis and the observation 

of a NOAEL. In addition, there are a number of human studies showing similar toxicological 

endpoints to those demonstrated in animal studies. Major areas of uncertainty are the lack of 

adequate human exposure data, and the lack of chronic inhalation exposure studies. 

VIII. References 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Cohen R. 1984. Reversible subacute ethylene glycol monomethyl ether toxicity associated with 

microfilm production: A case report. Am. J. Ind. Med. 6(6):441-446. [cited in U.S. EPA, 1995]. 

Cook RR, Bodner KM, Kolesar RC, Uhlman CS, Van Peenen PFD, Dickson GS, and Flanagan 

K. 1982. A cross-sectional study of ethylene glycol monomethyl ether process employees. Arch. 

Environ. Health 37(6):346-351. 

Cullen MR, Rado T, Waldron JA, Sparer J, Welch LS. 1983. Bone marrow injury in 

lithographers exposed to glycol ethers and organic solvents used in multicolor offset and 

ultraviolet curing printng processes. Arch Environ Health 38(6):347-354. 

Cullen MR, Solomon LR, Pace PE, Buckley P, Duffy TP, McPhedran P, Kelsey KT, Redlich 

CA. 1992. Morphologic, biochemical, and cytogenetic studies of bone marrow and circulating 

blood cells in painters exposed to ethylene glycol ethers. Environ. Res. 59(1):250-264. 

Doe JE, Samuels DM, Tinston DJ, and de Silva Wickramaratne GA. 1983. Comparative aspects 

of the reproductive toxicology by inhalation in rats of ethylene glycol monomethyl ether and 

propylene glycol monomethyl ether. Toxicol. Appl. Pharmacol. 15;69(1):43-47. 

Exon JH, Mather GG, Bussiere JL, Olson DP, and Talcott PA. 1991. Effects of subchronic 

exposure of rats to 2-methoxyethanol or 2-butoxyethanol: thymic atrophy and immunotoxicity. 

Fundam. Appl. Toxicol. 16(4):830-840. 

Greenburg L, Mayers MR, Goldwater JL, Burke WJ, and Moskovitz S. 1938. Health hazards in 

the manufacture of “fused collars.” I. Exposure to ethylene glycol monomethyl ether. J. Ind. 

Hyg. Toxicol. 20(2):134-147. [cited in U.S. EPA, 1995]. 

HSDB 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 

4/30/95). 

Appendix D3 258 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Holladay SD, Comment CE, Kwon J, Luster MI. 1994. Fetal hematopoietic alterations after 

maternal exposure to ethylene glycol monomethyl ether: Prolymphoid cell targeting. Toxicol. 

Appl. Pharmacol. 129:53-60. 

Larese F, Fiorito A, De Zotti R. 1992. The possible hematologic effects of glycol monomethyl 

ether in a frame factory. Br. J. Ind. Med. 49:131-133. 

Miller RR, Ayres JA, Young JT, and McKenna MJ. 1983. Ethylene glycol monomethyl ether. I. 

Subchronic vapor inhalation study with rats and rabbits. Fundam. Appl. Toxicol. 3:49-54. 

Parsons CE, and Parsons MEM. 1938. Toxic encephalopathy and “granulopenic anemia” due to 

volatile solvents in industry: Report of two cases. J. Ind. Hyg. Toxicol. 20(2):124-133. [cited in 

U.S. EPA, 1995]. 

Sparer J, Welch LS, McManus K, and Cullen MR. 1988. Effects of exposure to ethyelen glycol 

ethers on shipyard painters: I. Evaluation of exposure. Am J Ind Med 14:497-507. 

U.S. EPA. 1995. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS). Office of Health and Environmental Assessment, Environmental Criteria Assessment 

Office, Cincinnati, OH. 

Welch LS, Cullen MR. 1988. Effect of exposure to ethylene glycol ethers on shipyard painters: 

III. Hematologic effects.Am. J. Ind. Med. 14(5):527-236. 

Welch LS, Schrader SM, Turner TW, Cullen MR. 1988. Effects of exposure to ethylene glycol 

ethers on shipyard painters: II. Male reproduction. Am. J. Ind. Med. 14(5):509-526. 

Zavon MR. 1963. Methyl cellosolve intoxication. Am. Ind. Hyg. Assoc. J. 24:36-41. 

Appendix D3 259 Ethylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

ETHYLENE GLYCOL MONOMETHYL ETHER 

ACETATE 

(EGMEA; 2-methoxyethanol acetate; 2-methoxyethylester acetic acid; methyl glycol acetate; 

 methyl Cellosolve® acetate) 

CAS Registry Number: 110-49-6 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 90 μg/m3

 ( 20 ppb)

Critical effect(s) Reproductive (testicular) toxicity in rabbits 

(EGME) 

Hazard index target(s) Reproductive system 

II. Chemical Property Summary (HSDB, 1995)

Description Colorless liquid 

Molecular formula C5H10O3

Molecular weight 118.3 g/mol 

Boiling point 144-145°C 

Vapor pressure 2 torr @ 20°C 

Solubility Miscible with water, organic solvents, oils 

Conversion factor 4.83 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Ethylene glycol monomethyl ether acetate (EGMEA) is used as a solvent for nitrocellulose, 

cellulose acetate, and various other gums, resins, waxes, and oils (HSDB, 1995). It is also used 

in the semiconductor industry and in textile printing, photographic films, lacquers, and silkscreening inks. The annual specific statewide industrial emissions of EGMEA from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 3,060 pounds (CARB, 1999). 

IV. Effects of Human Exposure

Developmental defects have been described in the offspring of a mother who was occupationally 

exposed to EGMEA during pregnancy (Bolt and Golka, 1990). The mother was exposed during 

pregnancy by skin absorption and inhalation for approximately 1-4 hours/day to 1-2 liters of 

EGMEA. Her first child was born with congenital hypospadia, chordee, micropenis, and 

Appendix D3 260 Ethylene Glycol Monomethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

scrotum bifida and her second child (3 years later) was born with chordee, cryptorchidism, penile 

hypospadia and scrotum bifida. Both children had normal karyotypes. No estimates of exposure 

were made. 

A single case report described allergic dermatitis which may have developed from contact with 

EGMEA (Jordan and Dahl, 1971). A 58-year-old woman developed dermatitis on the nose 

possibly from contact with EGMEA on her eyeglasses. Ethylene glycol monoethyl ether acetate 

(EGEEA) was also present. 

V. Effects of Animal Exposure 

Cats, rabbits, guinea pigs, and mice were repeatedly exposed by inhalation for 8 hours daily to 

500 and 1000 ppm EGMEA (Gross, 1943; as described by Gingell et al., 1982). This exposure 

regimen was fatal to cats at 500 ppm EGMEA. Death occurred after the animals showed slight 

narcosis. Similarly, exposure to 1000 ppm EGMEA produced deaths among rabbits, guinea 

pigs, and mice within a few days. Kidney toxicity was observed in animals in both dose groups. 

Repeated 4- and 6-hour exposure of cats to 200 ppm EGMEA resulted in decreased “blood 

pigments” and red blood cell counts. 

The toxic effects of EGMEA were examined in male mice treated by gastric intubation 5 

days/week for 5 weeks with 0, 62.5, 125, 250, 500, 1000, or 2000 mg EGMEA/kg/day (Nagano 

et al., 1984). Dose-related testicular atrophy was observed at doses above 250 mg 

EGMEA/kg/day. Decreased white blood cell counts were observed in all EGMEA-exposed 

groups. 

EGMEA was readily converted in vitro to ethylene glycol monomethyl ether (EGME) by the 

nasal mucosal carboxylesterases of mice and rabbits (Stott and McKenna, 1985). The enzyme 

activity in the nasal mucosa was equal to that of the liver and greater than that of the kidney and 

lung. 

A concentration dependent decrease in testes weight was observed in male rabbits exposed to 30, 

100, or 300 ppm ethylene glycol monomethyl ether (EGME) 6 hours/day, 5 days/week for 13 

weeks (Miller et al., 1983). Degenerative changes in the germinal epithelium were observed in 

male rabbits of all exposed groups, but the changes were not statistically significant at 30 ppm. 

Two of five male rabbits exposed to 300 ppm EGME died during the course of the study. 

Female rabbits were also exposed; two of five female rabbits exposed to 100 or 300 ppm EGME 

died during the course of the study. Reduced body weight gain, pancytopenia (abnormal 

depression of all the cellular elements of the blood), and thymic atrophy were observed in rabbits 

of both sexes exposed to 300 ppm EGME. No effects on the reproductive organs of the female 

rabbits were observed. 

In the same study male and female rats were exposed to 30, 100, or 300 ppm EGME 6 hrs/day, 5 

days/week for 13 weeks. Moderate to severe degeneration of the germinal epithelium and 

seminiferous tubules was observed in male rats exposed to 300 ppm EGME. A significant 

decrease in body weight was observed in male rats exposed to 300 ppm and in female rats 

Appendix D3 261 Ethylene Glycol Monomethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

exposed to concentrations of EGME of 100 ppm or greater. Pancytopenia, lymphoid tissue 

atrophy, and decreased liver weights were observed in animals of both sexes exposed to the 

highest concentration. Also in the highest exposure group, mean values for total serum protein, 

albumin and globulins were lower than control values.

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Miller et al., 1983 (see below) 

Study population Rabbits 

Exposure method Discontinuous inhalation exposure (0, 30, 

100, or 300 ppm EGME) 

Critical effects Testicular effects 

LOAEL 100 ppm EGME 

NOAEL 30 ppm EGME 

Exposure continuity 6 hr/day, 5 days/week 

Exposure duration 13 weeks 

Average experimental exposure 5.4 ppm EGME for NOAEL group 

(30 x 6/24 x 5/7) 

Human equivalent concentration 5.4 ppm EGME for NOAEL group (gas with 

systemic effects, based on RGDR = 1.0 

using default assumption that lambda (a) = 

lambda (h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 10 

Interspecies factor 3 

Intraspecies factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.02 ppm (20 ppb, 0.06 mg/m3

, 60 μg/m3) 

EGME 

90 μg/m3 EGMEA (20 ppb) 

(60 x MWEGMEA / MWEGME) 

Data relating specific EGMEA exposure levels to toxicity in humans are not available for the 

development of a chronic REL. Data from experimental animals indicate that EGMEA is toxic 

to the hematopoietic and reproductive systems (Gross, 1943; Nagano et al., 1984), however 

good, quantitative data relating chronic exposure to toxicity are lacking. Because of evidence 

that EGMEA is readily converted to EGME by several organ systems (Stott and McKenna, 

1985) and since the scant data on EGMEA toxicity in animals indicate that the spectrum of 

toxicity of the two compounds is similar, the chronic REL was derived based upon the 

assumption of equimolar toxicity of EGMEA and EGME. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for EGMEA include the availability of subchronic inhalation 

exposure data from a well-conducted study of EGME as well as a number of supportive human 

studies on EGME showing the same toxicological endpoint, and the observation of a NOAEL. 

Appendix D3 262 Ethylene Glycol Monomethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Major areas of uncertainty are the assumption that EGMEA toxicity is comparable to that of 

EGME, the lack of adequate human exposure data, and the lack of chronic inhalation exposure 

studies. 

VII. References 

Bolt HM, and Golka K. 1990. Maternal exposure to ethylene glycol monomethyl ether acetate 

and hypospadia in offspring: a case report. Br. J. Ind. Med. 47:352-353. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Gingell R, Boatman RJ, Bus JS, Cawley TJ, Knaak JB, Krasavage WJ, Skoulis NP, Stack CR, 

and Tyler TR. 1994. Glycol ethers and other selected glycol derivatives. In: Patty’s Industrial 

Hygiene and Toxicology. Fourth ed. Vol. II, Part B. Clayton GD, and Clayton FE. (eds.). New 

York: John Wiley Sons. p. 2761. 

Gross E. 1943. In: Toxicology and Hygiene of Industrial Solvents. Lehmann KB, and Flury F. 

(eds). Baltimore: Williams and Wilkins. p. 290. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 

7/31/96). 

Jordan WP, and Dahl MV. 1971. Contact dermatitis to a plastic solvent in eyeglasses. Crosssensitivity to ethyl acetate. Arch. Dermatol. 104:524-528. 

Miller RR, Ayres JA, Young JT, and McKenna MJ. 1983. Ethylene glycol monomethyl ether. I. 

Subchronic vapor inhalation study with rats and rabbits. Fundam. Appl. Toxicol. 3:49-54. 

Nagano K, Nakayama E, Oobayashi H, Nishizawa T, Okuda H, and Yamazaki K. 1984. 

Experimental studies on toxicity of ethylene glycol alkyl ethers in Japan. Environ Health 

Perspect. 57:75-84. 

Stott WT, and McKenna MJ. 1985. Hydrolysis of several glycol ether acetates and acrylate esters 

by nasal mucosal carboxylesterase in vitro. Fundam. Appl. Toxicol. 5:399-404. 

Appendix D3 263 Ethylene Glycol Monomethyl Ether Acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

ETHYLENE OXIDE 

(oxirane, dimethylene oxide, epoxyethane) 

CAS Registry Number: 75-21-8 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 30 μg/m3 (18 ppb) 

Critical effect(s) Neurotoxicity in rats 

Hazard index target(s) Nervous system 

II. Physical and Chemical Properties (HSDB, 1995; CRC, 1994)

Description Colorless gas 

Molecular formula C2H4O 

Molecular weight 44.06 g/mol 

Density 1.80 g/L @ 25oC 

Boiling point 10.6oC 

Melting point −111.6oC 

Vapor pressure 1095 torr @ 20oC 

Conversion factor 1 ppm = 1.80 mg/m3

III. Major Uses or Sources 

The majority of all ethylene oxide (EtO) produced is used as a chemical intermediate in the 

production of various compounds including ethylene glycol, glycol ethers, and non-ionic 

surfactants (ATSDR, 1990). EtO is also used as a fumigant for food and cosmetics, and in 

hospital sterilization of surgical equipment and heat sensitive materials such as plastics. The 

annual statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California based on the most recent inventory were estimated to be 43,972 pounds of 

ethylene oxide (CARB, 2000).

IV. Effects of Human Exposure 

Ten hospital sterilizer workers were matched with controls and examined for physical and 

neuropsychological health (Estrin et al., 1990). The workers had operated sterilizers using 12% 

EtO and 88% Freon for an average of 5 years (range 0.5-10 years). Regular monitoring of 

workroom air had not been done. Measurements at the time of the study indicated 

concentrations of 15 ppm EtO or less. However, a second measurement showed an air 

concentration of 250 ppm EtO. A significantly greater percent of exposed workers exhibited a 

Appendix D3 264 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

bilateral reflex reduction in the ankle compared to controls. Nerve conduction tests did not 

identify significant differences between control and exposed workers, but a highly significant 

reduction (p = 0.009) in finger tapping speed was observed in exposed workers. The exposed 

group also performed more poorly on tests of spatial and visual abilities, and on tests of visual 

motor function. The results extended previous work by the same group (Estrin et al., 1987).

Cognitive impairment and personality dysfunction were observed more frequently in hospital 

workers chronically exposed to EtO, compared to a control group (Klees et al., 1990). A group 

of 22 hospital workers, who had been exposed to an 8-hour TWA of 4.7 ppm EtO for a mean of 

6.13 years (range 1-11 years), were matched with 24 control subjects. Neuropsychological 

function in the workers was classified as either normal or impaired on the basis of the 

questionnaires and of neuropsychological tests by 2 clinical psychologists (who were unaware of 

exposure status). (If the classification of the two clinicians did not agree, the subject was 

classified as “disagreement.” Disagreement occurred in 7/23 (30%) of the controls and 10/22 

(45%) of the exposed.) Exposed subjects were significantly more frequently classified as 

impaired (5/12) compared to controls (1/16) (χ2

 = 6.0861; p<0.05). The Klees et al. (1990) 

study cites several earlier case reports of EtO neurotoxicity.

Recent studies have identified hemoglobin adducts, sister chromatid exchanges, and other 

hematological effects as indicators of ethylene oxide exposure (Ribeiro et al., 1994; Sarto et al., 

1991). However, a recent study of 68 female workers from 9 hospitals in the U.S. and one in 

Mexico not only reports biological indicators of ethylene oxide exposure, but also provides a 

complete blood count with differential (Schulte et al., 1995). The workers were classified as 

low- or high-exposure based on a mean 8-hour time weighted average of 0.08 or 0.17 ppm EtO. 

The mean length of employment for workers from U.S. hospitals was 5.5 and 10 years for lowand high-exposure workers, respectively. The mean length of employment in low- and highexposure workers from the hospital in Mexico was 5.9 and 4.2 years, respectively. In workers 

from U.S. hospitals only, statistically significant decreases in hematocrit and hemoglobin were 

observed in high-exposure workers compared to low-exposure workers. Also, a statistically 

significant increase in lymphocytes and a significant decrease in neutrophils were observed in 

high-exposure workers compared to controls. In the workers from the hospital in Mexico, a 

significant relationship of EtO exposure and elevated neutrophil count was observed using 

regression. 

At least 2 epidemiological reports indicate a possible association of EtO exposure and 

spontaneous abortion. Hemminki et al. (1982) analyzed spontaneous abortions in Finnish 

hospital sterilizing staff using data from a postal questionnaire and from a hospital discharge 

register. The study included all sterilizing staff employed in Finnish hospitals in 1980; the 

controls were nursing auxiliaries. When the women were involved in sterilizing procedures 

during their pregnancies, the frequency of spontaneous abortion was 16.7% versus 5.6% for the 

non-exposed pregnancies. The independent analysis of spontaneous abortions using the hospital 

discharge register confirmed the findings. Thus two analyses suggested that EtO exposure may 

carry a risk of spontaneous abortion among sterilizing staff. 

More recently Rowland et al. (1996) sent questionnaires to 7,000 dental assistants (ages 18-39 

years) registered in California in 1987. Of these, 4,856 responded (69%). They analyzed 1,320 

Appendix D3 265 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

women whose most recent pregnancy was conceived while working full-time. Thirty-two 

reported exposure to EtO; unexposed dental assistants comprised the comparison group. Among 

exposed women, the age-adjusted relative risk (RR) of spontaneous abortion was 2.5 [95% (CI) 

= 1.0-6.3]. The RR for pre-term birth was 2.7 (95% CI = 0.8-8.8) and the RR for post-term birth 

was 2.1 (95% CI = 0.7-5.9). The RR of any of these adverse outcomes among exposed women 

was estimated to be 2.5 (95% CI = 1.0-6.1). These results also indicate a possible relationship of 

EtO and spontaneous abortion. 

V. Effects of Animal Exposure 

A 2 year inhalation bioassay exposed groups of 80 male rats to 0, 50, or 100 ppm EtO 7 hours 

per day, 5 days per week for 104 weeks (Lynch et al., 1984). Mean body weights were 

significantly lower and mortality was significantly higher in both exposure groups. 

Inflammatory lesions of the lung, nasal cavity, trachea, and inner ear were observed more 

frequently in EtO exposed rats. Skeletal muscle myopathy, consisting of atrophy and 

degeneration of skeletal muscle fibers, was observed more frequently in rats exposed to 100 ppm 

EtO compared to controls. Neoplastic changes were also observed in EtO exposed rats. 

Mice (30 per sex) were exposed to 0, 10, 50, 100, or 250 ppm EtO for 6 hours per day, 5 days 

per week, for 10 weeks (males) or 11 weeks (females) (Snellings et al., 1984). Neuromuscular 

screening was conducted, and samples of urine and blood were collected. A significantly greater 

percent of exposed mice exhibited abnormal posture during gait and reduced locomotor activity. 

A dose-response was observed for these effects, with significant changes at 50 ppm and greater. 

An abnormal righting reflex was observed in a significantly greater percent of mice exposed to 

100 ppm and above. Reduced or absent toe and tail pinch reflexes were observed in a 

significantly greater percent of mice exposed to 250 ppm EtO. Hematological changes observed 

in mice exposed to 250 ppm include slight, yet significant, decreases in red blood cell count, 

packed cell volume, and hemoglobin concentration. Absolute and relative spleen weights were 

significantly decreased in female mice exposed to 100 and 250 ppm and in male mice exposed to 

250 ppm EtO. A significant increase in relative liver weight was observed in female mice 

exposed to 250 ppm EtO. Male mice exhibited a significant decrease in body weight at 10, 50, 

and 250 ppm and a significant decrease in absolute testes weights at 50, 100, or 250 ppm EtO. 

This study indicates a subchronic NOAEL for neurological effects of 10 ppm EtO. 

In a study of the testicular effects of EtO, male rats were exposed to 500 ppm EtO 6 hours per 

day, 3 days per week for 2, 4, 6, or 13 weeks (Kaido et al., 1992). An awkward gait was 

observed in rats after 6-9 weeks of exposure. Although no significant changes in body weight 

were observed, a statistically significant dose-related decrease in testes weight was observed at 4, 

6, and 13 weeks of exposure. Progressive degeneration and loss of germ cells were also 

observed during the 13 week exposure. While severe loss of germ cells and marked 

morphological changes in remaining germ cells were observed at 6 weeks of exposure, some 

intact spermatids were observed at 13 weeks of exposure. This suggests that recovery of 

spermatogenesis occurred. 

Saillenfait et al. (1996) studied the developmental toxicity of EtO in pregnant Sprague-Dawley 

rats using inhalation exposure during gestation days 6 to 15. Two protocols were used: (1) 

Appendix D3 266 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

exposure for 0.5 hr once a day to 0, 400, 800, or 1200 ppm EtO; or (2) exposure for 0.5 hr three 

times a day to 0, 200, or 400 ppm EtO or to 0, 800, or 1200 ppm EtO. The second protocol 

caused fetal toxicity as indicated by reduced fetal weight at 800 ppm (the LOAEL for this 

endpoint) and at 1200 ppm, and overt maternal toxicity manifested as reduced body weight gain 

at 1200 ppm. No embryolethality or teratogenicity occurred in either exposure protocol. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Snellings et al., 1984 

Study population Male and female B6C3F1 mice 

Exposure method Inhalation chamber exposure to 0, 10, 50, 100, 

or 250 ppm ethylene oxide 

Critical effects Impaired neurological function 

LOAEL 50 ppm 

NOAEL 10 ppm 

Exposure continuity 6-hours/day, 5 days/week 

Exposure duration 10 weeks (males), or 11 weeks (females) 

Average experimental exposure 1.79 ppm (10 x 8/24 x 5/7) 

Human equivalent concentration 1.79 ppm ((gas with systemic effects, based on 

RGDR = 1.0 using default assumption that 

lambda (a) = lambda (h)) ) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 18 ppb (30 μg/m3

) 

Snellings et al. (1984) found a subchronic NOAEL of 10 ppm for neurological effects in mice. 

A neuromuscular screening test indicated that certain reflex responses and locomotor activities 

were altered in EtO-exposed animals. Human studies have also indicated neurological 

impairment in ethylene oxide exposed workers. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for ethylene oxide include the use of an animal study with 

both a LOAEL and a NOAEL and the use of an endpoint seen in both animals and humans. 

Major areas of uncertainty are the short time-frame of the key study, the lack of an appropriate 

human study, and the limited number of developmental toxicity studies. 

Appendix D3 267 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VIII. References 

ATSDR. 1990. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Ethylene Oxide. Prepared under contract No. 205-88-0608 for: US Department of Health and 

Human Services, Public Health Service, ATSDR. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Estrin WJ, Bowler RM, Lash A, and Becker CE. 1990. Neurotoxicological evaluation of hospital 

sterilizer workers exposed to ethylene oxide. Clin. Toxicol. 28(1):1-20. 

Estrin WJ, Cavalieri SA, Wald P, Becker CE, Jones JR, and Cone JE. 1987. Evidence of 

neurologic dysfunction related to long-term ethylene oxide exposure. Arch. Neurol. 44(12):1283-

1286. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex Inc. (Edition expires 7/31/95). 

Hemminki K, Mutanen P, Saloniemi I, Niemi ML, and Vainio H. 1982. Spontaneous abortions in 

hospital staff engaged in sterilising instruments with chemical agents. Br. Med. J. (Clin. Res. 

Ed.) 285(6353):1461-1463. 

Kaido M, Mori K, and Koide O. 1992. Testicular damage caused by inhalation of ethylene oxide 

in rats: Light and electron microscopic studies. Toxicologic Path. 20(1):32-43. 

Klees JE, Lash A, Bowler RM, Shore M, and Becker CE. 1990. Neuropsychological 

“impairment” in a cohort of hospital workers chronically exposed to ethylene oxide. Clin. 

Toxicol. 28(1):21-28. 

Lynch DW, Lewis TR, Moorman WJ, Burg JR, Groth DH, Khan A, Ackerman LJ, and Cockrell 

BY. 1984. Carcinogenic and toxicologic effects of inhaled ethylene oxide and propylene oxide in 

F344 rats. Toxicol. Appl. Pharmacol. 76:69-84. 

Ribeiro LR, Salvadori, DMF, Rios ACC, Costa SL, Tates AD, Tornquist M, and Natarajan AT. 

1994. Biological monitoring of workers occupationally exposed to ethylene oxide. Mutat. Res. 

313:81-87. 

Rowland AS, Baird DD, Shore DL, Darden B, and Wilcox AJ. 1996. Ethylene oxide exposure 

may increase the risk of spontaneous abortion, preterm birth, and postterm birth. Epidemiology 

1996;7(4):363-368. 

Appendix D3 268 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Sarto F, Tornquist MA, Tomanin R, Bartolucci GB, Osterman-Golkar SM, and Ehrenberg L. 

1991. Studies of biological and chemical monitoring of low-level exposure to ethylene oxide. 

Scand. J. Work Env. Health 17:60-64. 

Saillenfait AM, Gallissot F, Bonnet P, and Protois JC. 1996. Developmental toxicity of inhaled 

ethylene oxide in rats following short-duration exposure. Fundam. Appl. Toxicol. 34(2):223-227 

Schulte PA, Walker JT, Boeniger MF, Tsuchiya Y, and Halperin WE. 1995. Molecular, 

cytogenetic, and hematologic effects of ethylene oxide on female hospital workers. J. Occup. 

Environ. Med. 37(3):313-320. 

Snellings WM, Weil CS, and Maronpot RR. 1984. A subchronic inhalation study on the 

toxicologic potential of ethylene oxide in B6C3F1 mice. Toxicol. Appl. Pharmacol. 76:510-518. 

Appendix D3 269 Ethylene oxide 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

CHRONIC TOXICITY SUMMARY 

FLUORIDES including 

HYDROGEN FLUORIDE 

(hydrofluoric acid (aqueous solution); hydrogen fluoride (as a gas); 

fluoride salts (particulates or in solution)) 

CAS Registry Number: 7664-39-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 14 μg HF/m3 (17 ppb); 13 μg F/m3

Oral reference exposure level 0.04 mg/kg-day 

Critical effect(s) Skeletal fluorosis 

Hazard index target(s) Bone and teeth; respiratory system 

II. Physical and Chemical Properties of HF (HSDB, 1995; CRC, 1994) 

Description Colorless gas (HF), or as particulates 

Molecular formula HF 

Molecular weight 20.0 g/mol 

Density 0.83 g/L @ 25°C 

Boiling point 19.54°C 

Melting point −83.1°C 

Vapor pressure 400 torr @ 2.5°C 

Solubility Soluble in water and alcohol 

Conversion factor 1 ppm = 0.83 mg/m3

 @ 25°C 

III. Major Uses or Sources 

Hydrofluoric acid (HF) is a colorless, fuming liquid with a sharp, penetrating odor (Fairhall, 

1949). This acid is used in the glass etching, electronic, microelectronic, and petroleum refining 

and chemical industries (Bertolini, 1992). These industries use HF in the manufacture of such 

things as computer chips (an important industry in California), phosphate fertilizer, metal cans, 

plastics, refrigerant chemicals (fluorocarbons), inorganic chemicals, soaps and detergents, highoctane gasoline, and aircraft parts (Wohlslagel et al., 1976; Wing et al., 1991). HF is also used 

in commercial rust removal products. Another high profile use of HF in California has been as a 

catalyst in petroleum alkylation to make high-octane gasoline. HF is also a product of 

combustion of any F containing materials; as such, it is produced during structural fires. 

Appendix D3 270 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Sodium fluoride has been used as a topical and ingested anticaries agent due to its ability to 

harden tooth enamel during development. The optimal doses are not well established, but have 

been suggested to be approximately 0.080 mg/kg/day for 7 to 9 month old infants decreasing to 

0.034 mg/kg/day at 13 years of age (Shulman et al., 1995). A dose of 1.0 mg F ingested per day 

was reported to reduce dental caries 43%, and to be associated with a greatly increased rate of 

minor tooth mottling which caused no esthetic damage (Van Nieuwenhuysen and D'Hoore, 

1992). Many communities in California routinely add fluoride to the drinking water. The 

California Department of Health Services has adopted regulations that establish standards for the 

addition of F (CDHS, 2002). Any public water system using fluoridation must maintain F levels 

within the range established for its climate. The ranges vary according to average air 

temperatures, since people in cooler climates typically drink less water per day than people in 

warmer climates. Thus, in cooler areas, more F is required to provide the same dental benefit. 

For 2001-2002, F levels in San Francisco municipal water ranged from 0.65 to 1.1 ppm, while in 

Los Angeles the range was 0.44 to 0.83 ppm (CDHS, 2002). 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 48,221 pounds 

of fluorides and compounds, and 62,670 pounds of hydrogen fluoride (CARB, 2000). 

IV. Effects of Human Exposure 

The chronic exposure to fluorides, including HF, and the incidence of minimal osseous changes 

were studied in the workplace by Derryberry et al. (1963). In this study, the 8-hour timeweighted average fluoride exposure was calculated for the employment period of each of 74 

male workers (30 Caucasian, 44 African-American). The overall average fluoride exposure in 

these workers was measured as a time-weighted average of 2.81 mg F/m3

. In comparison, the 17 

workers within this group who had evidence of minimally increased bone density had an average 

fluoride exposure of 3.38 mg F/m3. The other workers were exposed to an average measured 

concentration of 2.64 mg F/m3

. In addition, urinary fluoride levels were greater in the 17 

individuals with greatest exposure compared to the remaining 57 workers (average = 5.18 mg 

F/L vs. 4.53 mg F/L). No differences between exposed and unexposed individuals were 

observed for gastrointestinal, cardiovascular, or hematologic systems, or in a physical exam. A 

statistically significant (p < 0.05) increase in the incidence of acute respiratory disease as 

determined from past medical histories was observed in fluoride-exposed individuals (19/74 vs. 

8/67 in controls); radiographic examination revealed a difference of lesser significance (p < 0.10) 

for pulmonary changes (11/74 vs. 4/67). No pulmonary function tests were reported.

An analysis of these data by OEHHA (see derivation section below) showed a statistically 

significant relationship between air fluoride and the minimal bone density increases. The raw 

data from the Derryberry et al. (1963) study are shown in Table 1. A Pearson correlation matrix 

of the variables measured in the Derryberry et al. study indicated that bone density was best 

correlated with mean air fluoride level, and to a lesser extent with the age of the individual. A 

log-logistic regression using the log air fluoride concentration as the independent variable 

showed a significant (p < 0.033) relationship between increasing air fluoride concentrations and 

probability of skeletal fluorosis. The parameters for the regression were β0 = −2.3468 (std. error 

Appendix D3 271 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

= 0.6462), and β1 = 1.1736 (std error = 0.5508); the odds ratio for the occurrence of skeletal 

fluorosis was 3.24. Years of exposure were not correlated with increased bone-density, 

according to a Pearson Correlation procedure (p = 0.63). Bone density has been shown to 

decrease with age after the age of 40 among normal, non-fluoride-exposed males (Runge et al., 

1979). As expected, age was very highly correlated with years exposed (p<0.00001). Therefore 

including years exposed in the dose-metric likely introduces a confounding variable (see 

discussion in Section VI.). In addition, Runge et al. (1979) found no association between years 

exposed and mineral content or bone width among 245 aluminum smelter workers exposed to 

2.75 or 3.2 mg F/m3

. For these reasons, years exposed were not used as the dose-metric for 

bone-density in this analysis. 

Although a threshold was not readily apparent from the logistic regression model, grouping the 

74 individuals by air fluoride exposure level into quintiles of 15 each with one group of 14, 

allowed for a comparison of group mean responses (Table 2). The 14 employees exposed to a 

time-weighted average concentration of 1.07 mg F/m3 did not exhibit bone density changes. An 

analysis of the grouped responses using a binomial distribution showed a probability of p = 

0.008 for obtaining 4/15 increased bone density observations in the 2.34 mg/m3

 group, and a 

probability of p = 0.047 for obtaining 3/15 positive observations in the 1.89 mg F/m3

 group. The 

1.89 mg F/m3 group was therefore considered a LOAEL for chronic skeletal fluorosis, and the 

1.07 mg/m3

 group was considered a NOAEL. The above probabilities assume that a chance 

occurrence is, at most, 1 in 18 of skeletal fluorosis or other cause leading to an abnormally dense 

x-ray in the general population. Since osteosclerosis is a rare condition that is associated with 

several types of hematological malignancies such as myeloid leukemia, the actual incidence of 

conditions leading to osteosclerosis is far below 1 in 18. This lends strong support to the 

consideration of 1.89 mg/m3 as a LOAEL for skeletal fluorosis. 

Appendix D3 272 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Table 1. Data on worker exposure to fluoride from Derryberry et al. (1963) 

Observation # ID Bone 

density 

Years 

exposed 

Urine max F 

(mg F/L) 

Urine 

min F 

(mg F/L)

Mean 

urinary F 

(mg F/L) 

Age 

(years) 

Air fluoride 

(mg/m3) 

OEHHA 

exposure 

grouping 

1 119 normal 18.5 43.0 2.8 14.7 58 8.16 5 

2 0 normal 8.4 24.7 5.3 9.6 42 3.19 4 

3 41 normal 15.8 35.0 2.5 9.1 35 3.29 4 

4 147 minimally 

increased 

9.6 17.1 2.1 8.9 60 5.98 5 

5 120 normal 16.7 20.5 3.4 8.6 55 3.29 4 

6 54 minimally 

increased 

17.0 44.0 4.0 8.6 56 7.73 5 

7 148 normal 10.5 14.0 3.7 8.4 41 8.32 5 

8 314 minimally 

increased 

14.4 22.7 1.7 8.3 56 3.24 4 

9 29 normal 17.0 18.2 2.5 7.7 50 2.60 3 

10 14 normal 14.3 19.4 2.1 6.3 46 2.33 3 

11 115 normal 15.2 18.5 1.4 6.3 38 2.11 3 

12 10 minimally 

increased 

10.3 22.0 2.3 6.1 38 2.72 4 

13 4 minimally 

increased 

7.1 7.7 2.0 5.7 54 3.22 4 

14 51 normal 14.9 42.0 0.8 5.6 46 3.18 4 

15 94 normal 16.2 15.4 3.3 5.5 56 5.12 5 

16 217 normal 7.1 7.1 2.6 5.3 42 2.54 3 

17 281 minimally 

increased 

7.8 8.6 1.1 5.2 36 3.79 4 

18 114 normal 10.4 13.2 2.8 5.2 38 7.66 5 

19 7 normal 7.8 9.1 2.2 5.1 43 2.91 4 

20 308 normal 11.9 6.7 3.5 5.1 44 1.89 2 

21 301 minimally 

increased 

15.2 9.5 2.5 5 36 2.56 3 

22 72 normal 25.9 13.7 2.1 4.9 55 5.55 5 

23 241 minimally 

increased 

17.0 10.0 1.9 4.9 46 4.48 5 

24 345 normal 10.5 7.1 2.0 4.9 47 1.49 1 

25 26 normal 16.4 12.2 0.5 4.7 39 2.41 3 

26 231 minimally 

increased 

16.3 8.2 2.8 4.6 62 1.88 2 

27 2 normal 24.7 8.9 2.1 4.6 46 3.53 4 

28 295 normal 14.5 10.7 0.9 4.6 44 2.07 3 

29 1 normal 8.9 5.9 2.4 4.5 30 1.92 2 

30 203 minimally 

increased 

18.2 6.8 1.6 4.4 43 2.66 3 

31 63 normal 16.2 7.4 2.0 4.3 55 3.90 5 

32 5 normal 4.5 11.5 1.9 4.3 43 1.12 1 

33 460 normal 12.5 6.1 1.6 4.3 60 2.13 3 

Appendix D3 273 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Observation # ID Bone 

density 

Years 

exposed 

Urine max F 

(mg F/L) 

Urine 

min F 

(mg F/L)

Mean 

urinary F 

(mg F/L) 

Age 

(years) 

Air fluoride 

(mg F/m3

) 

OEHHA 

exposure 

grouping 

34 249 minimally 

increased 

15.0 8.0 1.8 4.3 39 2.95 4 

35 3 normal 7.6 14.5 2.1 4.3 31 3.90 5 

36 322 normal 9.3 6.3 2.0 4.3 35 4.23 5 

37 8 minimally 

increased 

24.8 5.9 3.0 4.2 55 2.50 3 

38 3 normal 15.2 12.2 2.1 4.2 42 1.14 1 

39 309 normal 12.1 5.5 2.4 4.1 42 1.94 2 

40 36 normal 9.1 13.2 0.8 4.1 33 1.94 2 

41 45 normal 11.3 14.0 2.2 4.1 33 3.84 4 

42 70 normal 17.9 8.0 1.0 3.9 44 4.00 5 

43 250 minimally 

increased 

9.8 6.7 1.5 3.9 35 1.78 2 

44 38 normal 16.9 5.9 1.0 3.9 35 2.10 3 

45 200 minimally 

increased 

14.0 7.0 2.8 3.8 66 3.92 5 

46 183 normal 9.8 4.9 2.2 3.7 48 1.67 2 

47 32 normal 12.5 6.6 0.9 3.7 47 2.21 3 

48 25 normal 13.6 5.5 1.5 3.7 44 1.86 2 

49 21 normal 13.9 9.1 0.4 3.7 50 1.98 2 

50 304 normal 13.4 5.0 2.1 3.7 36 2.62 3 

51 132 normal 10.9 5.1 2.4 3.6 39 1.81 2 

52 6 minimally 

increased 

8.4 4.8 0.9 3.6 35 3.85 5 

53 244 normal 16.6 7.1 1.4 3.6 62 2.87 4 

54 30 normal 14.0 14.0 0.9 3.6 43 1.56 1 

55 88 minimally 

increased 

15.5 4.9 1.7 3.5 66 2.06 2 

56 227 normal 16.6 5.7 1.0 3.5 41 1.18 1 

57 271 normal 17.7 4.1 3.0 3.4 60 1.82 2 

58 19 normal 13.9 10.0 1.8 3.4 41 1.32 1 

59 190 normal 9.3 7.7 1.9 3.3 36 1.95 2 

60 258 normal 17.8 5.6 1.6 3.2 58 0.87 1 

61 278 normal 10.0 7.0 0.3 3.2 34 1.93 2 

62 331 normal 12.8 5.6 1.5 3.1 34 1.23 1 

63 91 normal 25.3 7.9 0.2 3.1 63 3.49 4 

64 342 normal 18.5 6.0 1.3 3 40 2.73 4 

65 261 normal 18.1 5.3 0.9 2.9 52 4.41 5 

66 291 normal 13.5 4.5 1.5 2.8 34 2.14 3 

67 149 normal 11.3 4.5 2.1 2.8 34 0.76 1 

68 2 normal 24.7 4.5 1.5 2.7 51 1.15 1 

69 4 normal 16.8 5.7 1.2 2.7 56 0.71 1 

70 109 normal 8.3 5.1 0.8 2.7 36 1.89 2 

71 242 normal 18.1 4.1 1.2 2.5 49 1.26 1 

Appendix D3 274 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Observation #ID Bone 

density

Years 

exposed

Urine max F 

(mg F/L)

Urine 

min F 

(mg F/L)

Mean 

urinary F 

(mg F/L)

Age 

(years)

Air fluoride 

(mg F/m3

)

OEHHA 

exposure 

grouping

72 179 normal 18.9 3.9 1.0 2.4 46 0.50 1 

73 325 minimally 

increased 

11.8 5.0 0.5 2.2 40 2.10 3 

74 159 normal 18.9 5.0 0.7 2.1 45 0.67 1 

Table 2. Grouped mean exposure

Exposure 

group 

Mean age 

± SD 

Mean air level 

mg F/m3 ± SD 

Number of 

responses 

Probability of difference 

from group 1* 

1 45.0 ± 7.0 1.07 ± 0.32 0/14** Not Applicable 

2 43.9 ± 11.2 1.89 ± 0.09 3/15*** 0.047 

3 43.0 ± 7.6 2.34 ± 0.23 4/15 0.008 

4 45.9 ± 9.8 3.22 ± 0.35 5/15 0.001 

5 48.5 ± 10.7 5.41 ± 1.72 5/15 0.001 

* Probability of obtaining result assuming a chance occurrence of abnormally dense x-ray of, at 

most, 1 in 18 individuals, using a binomial distribution (Systat for Windows v.5.05, 1994). 

** NOAEL 

*** LOAEL (p < 0.05) 

Largent et al. (1951) found a significant increase in bone density in the lower thoracic spine, 

with calcification extending into the lateral ligaments of 3 workers exposed for 17, 14, and 10 

years to HF (concentrations not estimated). 

A group of 74 men, who were occupationally exposed to unspecified concentrations of HF for an 

average of 2.7 years, reported occasions of upper respiratory irritation (Evans, 1940). Repeated 

chest X-rays over a 5-year period did not reveal any visible evidence of lung changes. The death 

rate of these workers from pneumonia and other pulmonary infections was the same as that of 

unexposed plant employees. 

There are various reports of asthma and related respiratory effects in pot room workers in the 

primary aluminum smelting industry. Exposure to fluoride (among other materials such as sulfur 

trioxide and polycyclic aromatic hydrocarbons) was measured as a possible index of exposures 

related to this condition (Seixas et al., 2000). However multiple exposures to respiratory irritants 

and other compounds which may affect immune response appear to be common in this work 

environment making it difficult to quantitatively relate the respiratory symptoms to inhaled HF 

or fluorides. 

Workers in a warehouse containing HF retorts experienced transitory hyperemia of the skin on 

their face and hands (Dale and McCauley, 1948). Twenty four of the 40 workers had definite 

changes in the thickness and number of trabeculae in the upper and lower jaw. 

Appendix D3 275 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Examinations of 107 pot room workers in two aluminum plants with airborne fluorides revealed 

22 subjects with limited motion of the dorsolumbar spine, compared with none in a control group 

of 108 workers with no history of exposure to fluorides (Kaltreider et al., 1972). In one plant, 76 

of 79 workers had increased bone density as measured by roentgenogram, with diagnosis of 

slight to moderate fluorosis. Moderate and marked fluorosis was observed after 15 years 

employment. The 8-hour time-weighted average fluoride content in these workplaces was 2.4 to 

6.0 mg/m3

. Balazova (1971) measured significant fluoride uptake and distribution in children 

living near an aluminum smelter but reported no incidence of fluorosis. 

No studies regarding the chronic irritant or respiratory effects of pure HF exposure in humans 

were available. 

Fluoride ion produced by various fluorocarbons has been associated with toxicity to human 

kidney collecting duct cells leading to sodium and water disturbances (Cittanova et al., 1996). 

Oral supplementation of greater than 0.1 mg F/kg body weight daily has been associated with 

enamel fluorosis in young children (Forsman, 1977). 

The Agency for Toxic Substances and Disease Registry (ATSDR, 2001) recently reviewed 

fluorides since they are found at hazardous waste sites which are candidates for remediation. 

The focus of this document was on oral exposure studies as that is the main concern for waste 

site remediation. 

V. Effects of Chronic Exposures to Animals 

Stokinger (1949) studied the subchronic effects of HF inhalation in several animal species. 

Animals (dogs, rabbits, rats, guinea pigs, and mice; 1 to 6 per group) were exposed to 0, 

7.2 mg/m3, or 25.1 mg/m3 6 hours/day, 6 days/week, for 30 days. Mortality, body weight, blood 

coagulation mechanisms, and gross pathology were measured. Exposure to 25.1 mg/m3

 HF for 

30 days resulted in degenerative testicular changes and ulceration of the scrotum in all 4 dogs 

and hemorrhage and edema in the lungs of 3 dogs. Pulmonary hemorrhage was also seen in 20 

of 30 rats, and 4 of 10 rabbits. Renal cortical degeneration was observed in 27 of 30 rats. All of 

the rats and mice at the 25.1 mg/m3

 concentration died. No mortality was observed in the other 

species tested. Blood fibrinogen levels were significantly increased in dogs, rats, and rabbits 

exposed to 25.1 mg/m3. Exposure to 7.2 mg/m3 HF resulted in pulmonary hemorrhage in 1 out 

of 5 dogs. No other significant effects were observed at the lower concentration. 

Shusheela and Kumar (1991) administered male rabbits 10 mg NaF/kg-bw per day orally for 18 

months (7 rabbits) or 29 months (3 rabbits), then studied the testis, epididymis, and vas deferens 

microscopically. After 29 months of F administration, the spermatogenic cells in the 

seminiferous tubules had degenerated and lacked spermatozoa. After both 18 and 29 months, 

cilia were lost from the epithelial cells lining the ductuli efferentes of the caput epididymidis. 

Stereocilia on the epithelial cells lining the vas deferens were also lost. In some regions of 

epithelia, the cell boundaries were not clear, and even appeared to be peeled off. Mucus droplets 

were abundant in the vas deferens of controls, but none were present in F treated rabbits. 

Appendix D3 276 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Spermatogenesis ceased sometime between 18 and 29 months. The authors concluded that 

ingestion of a high concentration of F has adverse effects (including infertility) on the male 

rabbit reproductive system. 

Ghosh et al. (2002) investigated the effects of NaF on steroidogenic and gametogenic activities 

in rat testes. Male Wistar rats were given 20 mg/kg/day NaF by gavage for 29 days. F treatment 

resulted in significantly lower relative wet weight of the testis, prostate, and seminal vesicle, 

decreased testicular delta(5),3beta-hydroxysteroid dehydrogenase (HSD) and 17beta-HSD 

activities, and significant lowering in plasma levels of testosterone. Epididymal sperm count 

was decreased significantly in F-treated rabbits and there were fewer mature luminal 

spermatozoa. Indicators of oxidative stress due to F included increased conjugated dienes in the 

testis, epididymis, and epididymal sperm pellet, and decreases of peroxidase and catalase in the 

sperm pellet. Thus F, at a dose encountered in drinking water in contaminated areas (at least of 

India), exerts an adverse effect on the male rat reproductive system. These effects on rats and 

rabbits (and dogs; see above) may be relevant to anecdotal reports of reproductive system 

malfunction in human chronic fluorosis. 

Parameter Control (n=6) NaF (n=6) p value 

Body weight, final (g) 127.00±3.75 122.00±5.10 

Testis, relative weight (%) 1.522±0.034 1.923±0.081 < 0.05 

Prostate, relative weight 0.297±0.043 0.148±0.014 < 0.05 

Seminal vesicles, rel. weight 0.448±0.025 0.174±0.027 < 0.05 

Testicular delta(5),3beta HSD ~28a ~24a < 0.05b

Testicular 17betaHSD ~29a ~24a < 0.05b

Plasma testosterone (ng/ml) ~2a ~1a < 0.05b

Epididymal sperm count 

(106/ml) 

7.02±0.17 3.70±0.57 < 0.05 

a

 approximate values based on reading Figures 2 and 3 of paper; b p values of authors 

Long et al. (2002) used ligand binding and Western blotting to study neuronal nicotinic 

acetylcholine receptors (nAChRs) in the brains of male and female Wistar rats ingesting 0.5 ppm 

(controls), 30 ppm, or 100 ppm F in their drinking water for 7 months. (All received 4 ppm F in 

their diet.) The brains of rats exposed to 100 ppm had significantly less binding sites for 

[

3

H]epibatidine, an analgesic agonist, but no change occurred at 30 ppm. Binding sites for 

[

125I]alpha-bungarotoxin, a competitive antagonist, were significantly decreased in the brains of 

rats exposed to both levels. The brain levels of the nAChR alpha4 subunit protein was 

significantly lowered by exposure to 100 ppm F. Alpha7 subunit protein was significantly 

decreased by both levels of F. No significant changes were seen in levels of the beta2 subunit 

protein. These nicotinic receptors have roles in learning and memory. Some of the effects were 

also seen in rat PC cells cultured for 48 h in up to 50 ppm F (Chen et al., 2003). The results may 

help to explain anecdotal reports of nervous system symptoms in human chronic fluorosis 

(Waldbott, 1978). 

NTP (1990) exposed F344/N rats and B6C3F1 mice of both sexes for two years to 0, 25, 100, 

and 175 ppm sodium fluoride (NaF) in their drinking water. NaF caused a dose dependent 

whitish discoloration of the teeth in both rats and mice. Male rats had an increased incidence of 

Appendix D3 277 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

tooth deformities and attrition. NaF increased the dysplasia of dentine in both rats and mice. At 

the highest dose (175 ppm), osteosclerosis of long bones was increased in female rats. There 

was also equivocal evidence of carcinogenic activity of NaF in male rats based on four 

osteosarcomas in dosed animals (Bucher et al., 1991). Other organ systems showed no dosedependent effects. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Derryberry et al. (1963) 

Study population 74 fertilizer plant workers (67 unexposed control 

subjects) 

Exposure method Occupational 

Critical effects Increased bone density (skeletal fluorosis) 

LOAEL 1.89 mg F/m3

 (1.98 mg HF/m3) 

NOAEL 1.07 mg F/m3

 (1.13 mg HF/m3) 

BMC05 0.37 mg F/m3

 (0.39 mg HF/m3) 

Exposure continuity 8 hours/day, 5 days/week 

Exposure duration 14.1 years (range = 4.5 to 25.9 years) 

Average exposure concentration 0.14 mg HF/m3

 (0.39 x 10/20 x 5/7) 

 or 0.13 mg F/m3 (0.37 x 10/20 x 5/7) 

Human equivalent concentration 0.14 mg HF/m3

 or 0.13 mg F/m3

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 10 

Inhalation reference exposure level 

 for F or HF 

0.013 mg F/m3 (13 μg /m3; 0.016 ppm; 16 ppb) or 

0.014 mg HF/m3 (14 μg /m3; 0.017 ppm; 17 

ppb) 

OEHHA’s analysis of the data in Derryberry et al. (1963) indicates a LOAEL of 1.89 mg/m3, 

and a NOAEL of 1.07 mg/m3

. A benchmark concentration (BMC05) of 0.37 mg/m3 was derived 

by fitting the probit model to the log dose in the U.S. EPA’s BMDS (version 1.3) software, for 

the individual mean air exposure data and incidence data in Table 1 above. Individuals in the 

highest dose group (group 5 in Table 2) were not included in the model, since none of the models 

fit this range of exposures well. Several other models produced reasonable fits to the data, but 

the probit model with log-transformed dose was selected since it produced a good fit not only by 

statistical criteria (p = 0.71) but also, as determined by inspection, it fit the low dose curve shape 

better than other models. This model also has the advantage of biological plausibility, in that, 

since lower doses of fluoride have a beneficial or nutritional effect, a threshold type of response 

for adverse effects is clearly expected. A graphical representation of the fit is shown in Figure 1. 

Adjusting for exposure continuity and utilizing an intraspecies uncertainty factor of 10 (UFH) 

results in a REL for F of 13 μg/m3

. 

Appendix D3 278 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Figure 1. 

Changes in bone density in association with fluoride exposure have been observed in several 

studies, and appear to be the most sensitive health effect for chronic exposure. The minimally 

increased bone density in the Derryberry study was significantly (p < 0.04, Fisher’s Exact Test) 

associated with “other osseous changes,” which reportedly included disc lesions, arthritis, and 

calcified ligaments. An increase in pulmonary changes in the workers with high bone density 

was marginally significant (p < 0.06) and included emphysema, fibrosis, and healed tuberculous 

lesions. Although dental fluorosis is a sensitive endpoint in many fluoride studies, the dental 

examinations of exposed workers in this study showed healthier teeth than in controls. The 

increased bone density observed was considered as indicating that adverse effects had occurred, 

based on the adverse effects associated with the increased density in the study, and on other 

research showing that increased bone density caused by fluoride exposure (75 mg sodium 

fluoride per day for four years) also leads to decreased bone strength and increased fragility 

(Riggs et al., 1990). Symptoms of abdominal pain, backache, restricted joint movement, and 

respiratory symptoms have been associated with airborne fluoride exposures and bone density 

increases in industrial settings (Zhiliang et al., 1987). 

The absorption of particulate and gaseous fluorides is reported to be similar (Collings et al., 

1951). Therefore, it would be expected that the effects on bone density would be similar 

regardless of the form of fluoride. 

Appendix D3 279 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

As noted in the study description, Derryberry et al. (1963) did not find a good correlation 

between years of exposure to fluoride and bone density change. OEHHA reexamined the 

original individual data and confirmed that the presence of bone density changes showed a better 

correlation with mean air fluoride concentration than with years of exposure, or with the product 

of the individual values of mean air fluoride concentration and years of exposure. However, the 

product of exposure concentration and time did show a consistent pattern of cumulative 

incidence suggesting a dose-response relationship for this parameter. An attempt to derive a 

benchmark value by fitting the probit model to the log of (exposure duration*concentration) and 

response (presence or absence of bone density change) did not result in an acceptable fit, so a 

BMDL05 could not be reported. However a maximum likelihood estimate of the benchmark 

(BMD05) was found to be 6.04 (mg F*years/m3), with exclusion of the three highest values that 

appeared to be outliers to the main distribution. If this value is divided by the mean exposure 

duration for the data set of 14.1 years, a benchmark exposure concentration of 0.43 mg F/m3

 is 

obtained. While this value is evidently less reliable than that obtained by fitting the mean 

exposure concentration, it is consistent with it, suggesting that, although other confounding 

factors related to age or duration prevent the demonstration of a relationship between the 

exposure/time integral and response in this data set, such a relationship probably does exist, as 

would be expected. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the key study for fluoride are the observation of health effects in a large 

group of workers exposed over many years, the availability of individual exposure estimates for 

each worker, and the identification of a NOAEL. The primary uncertainty in the study is the 

lack of a comprehensive health effects examination. Another source for concern is the 

potentially greater susceptibility of children to the effects of inhaled fluorides, considering the 

rapid bone growth in early years. 

Derivation of Chronic Oral REL 

In addition to being inhaled, airborne fluoride salts in particulate form can settle onto crops and 

soil and enter the body by ingestion. Thus an oral chronic reference exposure level (REL) for 

fluoride is also required in order to conduct a health risk assessment under the Air Toxics Hot 

Spots Act. California has developed a Public Health Goal (PHG) of 1 ppm (1,000 ppb) fluoride 

in drinking water (OEHHA, 1997). This level is intended to be an approximate year-round 

average. Thus it has properties similar to a chronic oral REL. (The PHG assumed that drinking 

water was the only source of fluoride since it was based on comparing communities with and 

without added fluoridation.) 

Appendix D3 280 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Study Dean, 1942; U.S. Public Health Service, 1991; 

National Research Council, 1993 

Study population Inhabitants of several U.S. cities 

Exposure method Drinking water 

Critical effects Dental fluorosis 

LOAEL 2 ppm 

NOAEL 1 ppm = 0.04 mg/kg-day* 

Exposure continuity Continuous 

Exposure duration Long-term 

Average experimental exposure 1 ppm = 0.04 mg/kg-day 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 1 (studies included children)

Cumulative uncertainty factor 1 

Oral reference exposure level 0.04 mg/kg-day 

 * based on the assumption that an 18 kg child drinks 720 ml of water per day (OEHHA, 2000). 

The PHG is based on a no-observed adverse-effect-level (NOAEL) of 1 mg/L for dental fluorosis 

in children (equivalent to 720 μg/day from drinking water for an 18 kg child drinking 40 ml/kg 

body weight/day of water). Moderate to severe dental fluorosis is rare when the drinking water 

fluoride level is near 1 mg/L, but begins to become significant at concentrations close to 2 mg/L. 

Since the study involved long term exposure to humans including children, a sensitive 

population, the cumulative uncertainty factor was 1. If one were to do a route-to-route 

extrapolation from this oral REL using the specific parameters for an 18 kg child breathing 4.2 

m3

/day, an equivalent inhalation REL would be about 170 μg/m3. Thus, the inhalation REL of 

13 μg/m3 based on the adult occupational data is likely to be protective of children. 

VIII. Potential for Differential Impacts on Children's Health 

The critical effect for inhalation exposures is skeletal fluorosis. Since infants' and children's 

skeletons are developing, they may be more sensitive to this effect. This applies with particular 

importance to the teeth, and it is established that excessive exposure to fluoride during the period 

of tooth development in infancy and childhood causes dental fluorosis (Dean, 1942; U.S. Public 

Health Service, 1991; NRC, 1993). The oral REL and the California PHG for fluoride in 

drinking water are based on dental fluorosis. Although the inhalation chronic REL proposed is 

based on a study in adults, the inhalation chronic REL (see section VI) is lower than that implied 

by the oral REL and PHG. Since the oral REL and PHG are based on exposures throughout life, 

including the pre-natal period, infancy, and childhood, it is reasonable to conclude that the 

proposed inhalation REL is generally protective of infants and children, barring some unknown 

difference in toxicity between the two routes of exposure. The ratio of the intake at the PHG 

level in drinking water is closer to the effect level than the default intraspecies uncertainty factor 

of 10; this is to be expected since children are a sensitive subpopulation for the dental fluorosis 

effect. 

Appendix D3 281 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Extensive interindividual variation in total fluoride intake (930.7 ± 391.5 μg/day) was recently 

documented for a small group (n = 11) of healthy German children ages 3 to 6 years 

(Haftenberger et al., 2001). Similar interindividual variation has also been reported for slightly 

younger children in Connersville (n = 14) and Indianapolis, Indiana (n = 29) and in San Juan, 

Puerto Rico (n = 11) (Rojas-Sanchez et al., 1999). Consideration should therefore be given to 

populations with exceptionally high fluoride intake due to locally elevated concentrations in 

drinking water, since some of these populations are already close to adverse effect levels of 

fluoride intake, and certain individuals in California experience dental fluorosis. For these 

individuals, even exposure to fluorides at the oral and/or inhalation RELs, which are acceptable 

in isolation, might be deleterious. The table below compares the data of Haftenberger et al. 

(2001) with recent estimates of F intake ranges in California (OEHHA, 1997). 

 Fluoride Intake (mg/day) 

F in drinking 

water 

(mg/L) 

F from 

drinking 

water 

F from 

food 

F from 

mouthwash 

F from a 

supplement 

Total F 

F from 

toothpaste 

Children 

 (OEHHA) 

 <0.3 0.1 - 0.3 0.1 – 0.5 0.2 – 1.2 0.1 - 0.5 0.5 1.0 – 3.0 

 0.7 - 1.2 0.7 - 1.2 0.2 – 1.2 0.1 - 0.5 0 1.1 – 4.6 

Haftenberger 

 0.25 (see 

food) 

0.20±0.12 0.27±0.18 No data 0 - 1.0 0.93±0.39

Adults 

 (OEHHA) 

 <0.3 0.2 – 0.6 0.3 – 1.0 0.02 – 0.15 0.2 – 1.0 0 0.7 – 2.8 

 0.7 - 1.2 1.4 – 2.4 0.3 – 3.4 0.02 – 0.15 0.2 – 1.0 0 1.9 – 7.0 

IX. References 

ATSDR. 2001. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Fluorides. Draft for Public Comment. Atlanta: ATSDR. 

Balazova G. 1971. Long-term effect of fluoride emission upon children. Fluoride 4(2):85-88. 

Benke G, Abramson M, Sim M. 1998. Exposures in the alumina and primary aluminum industry: 

an historical review. Ann. Occup. Hyg. 42(3):173-189. 

Bertolini JC. 1992. Hydrofluoric acid: A review of toxicity. J. Emerg. Med. 10:163-168. 

Bucher JR, Hejtmancik MR, Toft JD 2nd, Persing RL, Eustis SL, Haseman JK. 1991. Results 

and conclusions of the National Toxicology Program's rodent carcinogenicity studies with 

sodium fluoride. Int. J. Cancer 48(5):733-737. 

Appendix D3 282 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CDHS. 2002. California Department of Health Services. Public Water System Fluoridation. 

Available on-line at www.dhs.cahwnet.gov/ps/ddwem/Fluoridation/Fluoridation.htm 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Chen J, Shan KR, Long YG, Wang YN, Nordberg A, Guan ZZ. 2003. Selective decreases of 

nicotinic acetylcholine receptors in PC12 cells exposed to fluoride. Toxicology 183(1-3):235-

242. 

Cittanova ML, Lelongt B, Verpont MC, Geniteau-Legendre M, Wahbe F, Prie D, Coriat P, and 

Ronco PM. 1996. Fluoride ion toxicity in human kidney collecting duct cells. Anesthesiology 

84(2):428-435. 

Collings GH, Fleming RBL, and May R. 1951. Absorption and excretion of inhaled fluorides. 

A.M.A. Arch. Ind. Hyg. 4:585-590. 

Dale PP, and McCauley HB. 1948. Dental conditions in workers chronically exposed to dilute 

and anhydrous hydrofluoric acid. J. Am. Dent. Assoc. 37:131-140. 

Dean, H. 1942. The investigation of physiological effects by the epidemiological method. In: 

Fluoride and Dental Health, F. R. Moulton, ed. Washington, DC: American Association for 

Advancement of Science, p. 23. 

Derryberry OM, Bartholomew MD, and Fleming RBL. 1963. Fluoride exposure and worker 

health - The health status of workers in a fertilizer manufacturing plant in relation to fluoride 

exposure. Arch. Environ. Health 6:503-514. 

Evans EE. 1940. An X-ray study of effects of industrial gases upon the human lung. Radiology 

34:411-424. 

Fairhall L. 1949. Industrial Toxicology. Baltimore, MD: Williams and Wilkins. [as cited in 

Smyth HF. 1956. Improved communication - hygienic standards for daily inhalation. Am. Ind. 

Hyg. Assoc. 129-185.] 

Forsman B. 1977. Early supply of fluoride and enamel fluorosis. Scand. J. Dent. Res. 85(1):22-

30. 

Ghosh D, Das Sarkar S, Maiti R, Jana D, Das UB. 2002. Testicular toxicity in sodium fluoride 

treated rats: association with oxidative stress. Reprod. Toxicol. 16(4):385-90. 

Haftenberger M, Viergutz G, Neumeister V, Hetzer G. 2001. Total fluoride intake and urinary 

excretion in German children aged 3-6 years. Caries Res. 35(6):451-7 

Appendix D3 283 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

HSDB. 1995. Hazardous Substances Data Bank. TOMES® Denver, CO: Micromedex, Inc. 

Hodge HC. 1950. The concentration of fluorides in drinking water to give the point of minimum 

caries with maximum safety. J. Am. Dent. Assoc. 40:436. 

Kaltreider NL, Elder MJ, Cralley LV, and Colwell MO. 1972. Health survey of aluminum 

workers with special reference to fluoride exposure. J. Occup. Med. 14:531-541. 

Largent EJ, Bovard PG, and Heyroth FF. 1951. Roentgenographic changes and urinary fluoride 

excretion among workmen engaged in the manufacture of inorganic fluorides. Am. J. 

Roentgenol. Radium Ther. Nucl. Med. 65:42-48. 

Long YG, Wang YN, Chen J, Jiang SF, Nordberg A, Guan ZZ. 2002. Chronic fluoride toxicity 

decreases the number of nicotinic acetylcholine receptors in rat brain. Neurotoxicol. Teratol. 

24(6):751-7. 

NRC 1993. National Research Council. Health Effects of Ingested Fluoride. Washington, DC: 

National Academy Press. 

NTP). 1990. National Toxicology Program. NTP Technical Report on the Toxicology and 

Carcinogenesis Studies of Sodium Fluoride (CAS No. 7681-49-4) in F344/N Rats and B6C3F1 

Mice (Drinking Water Studies). NTP TR 393. Research Triangle Park, NC: National Toxicology 

Program. 

OEHHA. 1997. Office of Environmental Health Hazard Assessment. Public Health Goal for 

Fluoride in Drinking Water. Available online at http://www.oehha.ca.gov 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part IV. Exposure Assessment and Stochastic Technical 

Support Document. Available online at http://www.oehha.ca.gov 

Riggs BL, Hodgson WM, O’Fallon WM, Chao EY, Wahner HW, Muhs JM et al. 1990. Effect of 

fluoride treatment on the fracture rate of postmenopausal women with osteoporosis. New Engl. J. 

Med. 322:802-809. 

Rojas-Sanchez F, Kelly SA, Drake KM, Eckert GJ, Stookey GK, Dunipace AJ. 1999. Fluoride 

intake from foods, beverages and dentifrice by young children in communities with negligibly 

and optimally fluoridated water: a pilot study. Community Dent. Oral Epidemiol. 27(4):288-297. 

Runge H, Franke J, Geryk B, Hein G, Fengler F, Paul H, Bismarck M, and Schmidt CW. 1979. 

Bone mineral analysis in persons with long-time fluoride exposure. Fluoride 12(1):18-27. 

Seixas NS, Cohen M, Zevenbergen B, Cotey M, Carter S, Kaufman J 2000. Urinary fluoride as 

an exposure index in aluminum smelting. Amer. Ind. Hyg. Assoc. J. 61(1):89-94. 

Shulman JD, Lalumandier JA, and Grabenstein JD. 1995. The average daily dose of fluoride: a 

model based on fluid consumption. Pediatr. Dent. 17(1):13-18. 

Appendix D3 284 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels August 2003 

Stokinger HE. 1949. Toxicity following inhalation of fluorine and hydrogen fluoride. In: 

Pharmacology and Toxicology of Uranium Compounds. Voegtlin C, and Hodge HC. (eds) New 

York: McGraw-Hill Co., Inc. pp. 1021-1057. 

Susheela AK, Kumar A. 1991. A study of the effect of high concentrations of fluoride on the 

reproductive organs of male rabbits, using light and scanning electron microscopy. J. Reprod. 

Fertil. 92(2):353-60. 

U.S. EPA. 2001. United States Environmental Protection Agency. Fluorine (soluble fluoride). 

IRIS database. Available online at http://www.epa.gov/iris/subst/index.html. 

U.S. Public Health Service. 1991. Review of fluoride, benefits and risks, report of the ad hoc 

subcommittee on fluoride of the committee to coordinate environmental health and related 

programs. February 1991. 

Van Nieuwenhuysen JP, D'Hoore W. 1992. [Dental caries, fluoride tablets and enamel opacities]. 

Arch. Fr. Pediatr. 49(7):617-621. 

Waldbott GL. Fluoridation: The Great Dilemma. Lawrence, KS: Coronado Press, 1978. 

Wing JS, Brender JD, Sanderson LM, Perrotta DM, and Beauchamp RA. 1991. Acute health 

effects in a community after a release of hydrofluoric acid. Arch. Environ. Health 46(3):155-160. 

Wohlslagel J, DiPasquale LC, and Vernot EH. 1976. Toxicity of solid rocket motor exhaust: 

effects of HCl, HF, and alumina on rodents. J. Combustion Toxicol. 3:61-70. 

Zhiliang Y, Yihua L, Liansheng Z, and Zhengping Z. 1987. Industrial fluoride pollution in the 

metallurgical industry in China. Fluoride 20(3):118-125. 

Appendix D3 285 Fluorides 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

GLUTARALDEHYDE 

(1,5-pentanedial; 1,5-pentanedione; glutaric dialdehyde; Aldesen; Cidex; Sonacide) 

CAS Registry Number: 111-30-8 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.08 μg/m3 (0.02 ppb)

Critical effect(s) Squamous metaplasia of the respiratory epithelium 

in the nose of male and female mice 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1996; CRC, 1994; Chemfinder, 2000)

 

Description Colorless liquid/oil 

Molecular formula C5H8O2

Molecular weight 100.12 g/mol 

Boiling point 188°C (decomposes) (CRC, 1994) 

Melting point −6°C (Chemfinder, 2000) 

Solubility Soluble in water, alcohol, benzene 

Conversion factor 4.1 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources

Glutaraldehyde is a chemical frequently used as a disinfectant and sterilizing agent against 

bacteria and viruses (2% solution), an embalming fluid and tissue fixative, a component of 

leather tanning solutions, and an intermediate in the production of certain sealants, resins, dyes, 

and electrical products (HSDB, 1996). For commercial purposes, solutions of 99%, 50%, and 

20% are available. Glutaraldehyde is also an atmospheric reaction product of cyclohexene. The 

annual statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California based on the most recent inventory were estimated to be 29,603 pounds of 

glutaraldehyde (CARB, 2000). 

IV. Effects of Human Exposure

Evidence of the toxicity of glutaraldehyde to humans is limited to reports of occupational 

exposure from its use as a disinfectant and sterilizing agent. Frequently observed effects from 

exposure include skin sensitivity resulting in dermatitis, and irritation of the eyes and nose with 

accompanying rhinitis (Jordan et al., 1972; Corrado et al., 1986; Hansen, 1983; Wiggins et al., 

Appendix D3 286 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

1989). Occupational asthma has also been reported among workers repeatedly exposed to 

glutaraldehyde, particularly respiratory technologists who use glutaraldehyde as a sterilizing 

agent for endoscopes (Chan-Yeung et al., 1993; Stenton et al., 1994; Gannon et al., 1995). 

Quantitation of the exposure levels that led to glutaraldehyde sensitization was not available 

from the studies. 

V. Effects of Animal Exposure

The histopathology of the respiratory tract in rats and mice exposed to glutaraldehyde by 

inhalation was examined (Gross et al., 1994). F344 rats and B6C3F1 mice (20 animals of each 

sex and of each species at each exposure level for a total of 480 rodents) were continuously 

exposed to glutaraldehyde in recirculating exposure chambers at concentrations of 0, 62.5, 125, 

250, 500, or 1000 ppb glutaraldehyde for one day, 4 days, 6 weeks, or 13 weeks. At termination, 

respiratory tract tissue as well as duodenum and any gross lesions were collected and formalin 

fixed. Animals were treated with tritiated thymidine two hours before termination to evaluate 

cell replication in certain respiratory tract tissues. Respiratory tract tissue sections were made as 

follows: transverse sections of the nose and trachea, frontal section of the carina, and 

longitudinal section of the lung. Ten male and 10 female mice exposed to 1000 ppb and one 

female mouse exposed to 500 ppb group died during the course of the study. Two male and 3 

female rats exposed to 1000 ppb died during the course of the study. Histopathological 

examination of animals surviving to the end of the study entailed scoring the severity of the 

finding from “no response” to “very severe” response on a 0 to 5 scale. Unit length labeling 

index, the indicator of cell proliferation, was evaluated by autoradiography at two sites: the nasal 

vestibule and the dorsal atrioturbinate. 

Lesions in animals treated with glutaraldehyde appeared primarily in the anterior third of the 

nose. Lesions were apparently more increased in mice compared to rats due to some level of 

“background” non-suppurative lesions in the rats. Mice were considered devoid of background 

lesions. In the 13-week study, female mice were the most sensitive, with lesions averaging a 

score of 2 (mild and clear, but of limited extent and/or severity). The lesions were characterized 

as neutrophilic infiltration primarily in the squamous epithelium of the vestibule, with thickening 

of the epithelium leading to loss of the characteristic surface grooves. Both cell size and number 

were reported to be increased. Lesions were generally found to increase in nature and severity 

with increased time and level of exposure. Obstruction of the nasal vestibule was thought to 

account for the mortality of animals in the higher dose groups. In female mice at 13 weeks, all 

glutaraldehyde dose groups showed the accumulation of eosinophilic proteinaceous deposits in 

the respiratory epithelium of the maxilloturbinate margin. Examination of unit length labeling 

indices as a measure of growth showed significant increases in all treated groups of female mice. 

No evidence of exposure related lesions was found in the respiratory tract in the trachea, carina, 

bronchi, or lungs. 

Appendix D3 287 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

 Mean Subjective Pathology Scores for Nasal Lesions in Female Mice at 13 Weeks 

Glutaraldehyde 

Intraepithelial 

neutrophils 

Subepithelial 

neutrophils 

Squamous 

metaplasia 

0 ppb 0 0.4 0 

62.5 ppb 2.0 2.0 0 

125 ppb 2.4 2.8 0 

250 ppb 3.2 3.2 0 

500 ppb 2.8 2.8 0.5 

1000 ppb* -- -- -- 

*Animals exposed to 1000 ppb died early in the experiment. 

Greenspan et al. (1985) exposed male and female F-344 rats to 0, 0.3, 1.1 and 3.1 ppm 

glutaraldehyde and 0, 0.2, 0.63, and 2.1 ppm glutaraldehyde, respectively, in a 9-day study, and 

both sexes to 0, 21, 49, and 194 ppb glutaraldehyde in a 14 week study. Animal numbers were 

not specified. Exposures were conducted for 6 hours per day, 5 days per week. In the 9-day 

study, observations in the high and intermediate dose level groups included reduced body weight 

gain, inflammation of the nasal and olfactory mucosa, and sensory irritation. In the two highest 

doses of the 14-week study, statistically significant differences in body weight gain were 

observed as well as perinasal wetness. No histopathological indication of inflammation in 

olfactory or nasal mucosa was observed. 

Mice were exposed to 0, 0.3, 1.0, and 2.6 ppm glutaraldehyde vapors for 6 hours/day for 4, 9, or 

14 days (Zissu et al., 1994). These mice were killed immediately after the exposure period. 

Other groups exposed to 1.0 ppm for 14 days were killed after recovery periods of 1, 2, and 4 

weeks. After 4 days of exposure to the lowest dose, mice showed lesions in the respiratory 

epithelium of the septum, and the naso- and maxilloturbinates. After exposure to 1.0 ppm 

glutaraldehyde, lesions were still judged as severe after 2 weeks of recovery. 

A study comparing the effects of intra-nasally instilled glutaraldehyde and formaldehyde on rat 

nasal epithelium found inflammation, epithelial degeneration, respiratory epithelial hypertrophy, 

and squamous metaplasia in treated animals (St. Clair et al., 1990). Acute inhalation exposure to 

formaldehyde produced identical lesions. Ten-fold higher concentrations of instilled 

formaldehyde were required to produce the same effect as instilled glutaraldehyde. 

In a chronic study, NTP (1998, 1999) exposed groups of 50 male and 50 female F344/N rats to 0, 

250, 500, or 750 ppb glutaraldehyde vapor by inhalation for 6 h/day, 5 days/week, for 104 

weeks. Survival of 500 and 750 ppb female rats was less than that of the chamber controls. 

Mean body weights of all exposed groups of male rats and 500 and 750 ppb female rats were 

generally less than those of the chamber controls. Increased incidences of nonneoplastic nasal 

lesions occurred primarily within the anterior section of the nose in 500 and 750 ppb rats and to a 

lesser extent in 250 ppb rats. The more significant lesions included hyperplasia and 

inflammation of the squamous and respiratory epithelia and squamous metaplasia of the 

respiratory epithelium. Thus 250 ppb (1000 μg/m3

) is a chronic LOAEL for rats. 

In the same study NTP (1998, 1999) exposed groups of 50 male and 50 female B6C3F1 mice to 

0, 62.5, 125, or 250 ppb glutaraldehyde vapor by inhalation for 6 h/day, 5 days/week, for 104 

Appendix D3 288 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

weeks. Survival of exposed mice was similar to that of the chamber controls. Mean body 

weights of female mice exposed to 250 ppb were generally less than those of the controls. The 

incidence of inflammation of the nose was marginally increased in 250 ppb females. Incidences 

of squamous metaplasia of the respiratory epithelium were increased in 250 ppb males and 

females and 125 ppb females. Incidences of hyaline degeneration of the respiratory epithelium 

were increased in all exposed groups of females. Thus 62.5 ppb was a chronic LOAEL for 

female mice. 

 Incidence of Nasal Lesions in Female Mice exposed for 104 weeks 

Glutaraldehyde Inflammation 

Respiratory 

epithelium 

hyaline 

degeneration 

Respiratory 

epithelium 

squamous 

metaplasia 

0 ppb 6/50 16/50 7/50 

62.5 ppb 7/49 35/49 11/49 

125 ppb 13/50 32/50 16/50 

250 ppb 14/50 30/50 21/50 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study NTP 1998, 1999 

Study population Male and female F344 rats and B6C3F1 mice 

(50/sex/group) 

Exposure method Continuous inhalation exposure 

 (0, 62.5, 125, and 250 ppb in mice; 

 0, 250, 500, or 750 ppb in rats) 

Critical effects Respiratory epithelium squamous metaplasia 

LOAEL 62.5 ppb (female mice) 

NOAEL Not observed 

BMC05 20.5 ppb 

Exposure continuity 6 hr/day, 5 days/week 

Exposure duration 104 weeks 

Equivalent continuous exposure 3.7 ppb (20.5 x 6/24 x 5/7) 

Human equivalent concentration 0.62 ppb (gas with extrathoracic respiratory 

effects, RGDR = 0.17, BW = 28 g, MV = 

0.032 L/min, SA = 3 cm2

) 

LOAEL uncertainty factor not needed in BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.02 ppb (0.08 μg/m3

) 

Several studies indicate that the upper respiratory tract is a target for the toxicity of 

glutaraldehyde from inhalation exposure. Reports of toxicity to humans show that exposure can 

Appendix D3 289 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

lead to occupational asthma as well as cause irritation of the eyes and nose with accompanying 

rhinitis. Likewise, animals exposed to glutaraldehyde by the inhalation route show evidence of 

respiratory irritation with the induction of lesions of the anterior nasal cavities upon long-term 

exposure (Gross et al., 1994; Greenspan et al., 1985; NTP, 1998, 1999). The NTP (1998, 1999) 

study yielded a chronic LOAEL for female mice of 62.5 ppb. Gross et al. (1994) showed 

neutrophilic infiltration in the olfactory epithelium in the lowest dose exposure group. (Female 

mice exposed to 62.5 ppb also showed subepithelial neutrophilic infiltration.) This level was 

taken to be the subchronic LOAEL. This effect on the nasal epithelium was demonstrated to be 

both concentration- and exposure duration-dependent. 

A benchmark concentration was determined using EPA's version 1.20 BMC software and the 

dose-response data on respiratory epithelium squamous metaplasia in female mice. The quantallinear model gave an MLE05 of 31.24 ppb, a BMC05 of 20.51 ppb, and a p value of 0.9471. With 

the benchmark approach no LOAEL UF is needed. The study was a lifetime study so the 

subchronic UF is 1. An interspecies UF of 3 rather than 10 was used since an RGDR adjustment 

had been made. The default intraspecies UF of 10 was used so that the total UF was 30. The 

resulting chronic REL for glutaraldehyde is 0.02 ppb (0.08 μg/m3

). 

For comparison with the proposed REL, the study of Gross et al. (1994) used 62.5 ppb 

continuous exposure. Multiplying by the RGDR of 0.17 and dividing by a cumulative 

uncertainty factor of 300 (3 for a LOAEL, 3 for subchronic, 3 for interspecies, and 10 for 

intraspecies) results in a REL of 0.035 ppb (0.1 μg/m3). 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the inhalation REL for glutaraldehyde is the availability of controlled 

exposure inhalation studies in multiple species at multiple exposure concentrations and with 

adequate histopathogical analysis. Major areas of uncertainty are the lack of human data, the 

lack of reproductive and developmental toxicity studies, the lack of dermal sensitization studies, 

and the lack of observation of a NOAEL. 

VIII. References

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Chan-Yeung M, McMurren T, Catonio-Begley F, and Lam S. 1993. Occupational asthma in a 

technologist exposed to glutaraldehyde. J. Allergy Clin. Immunol. 91:974-978. 

Chemfinder, 2000. Available online at http://www.chemfinder.com 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Appendix D3 290 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Corrado OJ, Osman J, and Davies RJ. 1986. Asthma and rhinitis after exposure to 

glutaraldehyde. Hum. Toxicol. 5:325-328. 

Gannon PF, Bright P, Campbell M, O’Hickey SP, and Burge PS. 1995. Occupational asthma to 

glutaraldehyde and formaldehyde in endoscopy and x-ray departments. Thorax 50:156-159. 

Greenspan BJ, Ballantyne B, Fowler EH, and Snellings WM. 1985. Subchronic inhalation 

toxicity of glutaraldehyde. Toxicologist 5:29 (abstract). 

Gross EA, Mellick PW, Kari FW, Miller FJ, and Morgan KT. 1994. Histopathology and cell 

replication responses in the respiratory tract of rats and mice exposed by inhalation to 

glutaraldehyde for up to 13 weeks. Fundam. Appl. Toxicol. 23:348-362. 

Hansen KS. 1983. Glutaraldehyde occupational dermatitis. Contact Dermatitis 9:81-82. 

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Jordan WP, Dahl MV, and Albert HL. 1972. Contact dermatitis from glutaraldehyde. Arch. 

Dermatol. 105:94-95. 

National Toxicology Program (NTP). 1998. Toxicology and Carcinogenesis Studies of 

Glutaraldehyde (CAS NO. 111-30-8) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). 

TR-490. Board Draft. 

National Toxicology Program (NTP). 1999. Toxicology and Carcinogenesis Studies of 

Glutaraldehyde (CAS NO. 111-30-8) in F344/N Rats and B6C3F1 Mice (Inhalation Studies). 

TR-490. September 1999. NIH Publication No. 99-3980. Available online at 

http://ntp-server.niehs.nih.gov/htdocs/LT-studies/tr490.html 

St. Clair MBG, Gross EA, and Morgan KT. 1990. Pathology and cell proliferation induced by 

intra-nasal instillation of aldehydes in the rat: comparison of glutaraldehyde and formaldehyde. 

Toxicol. Pathol. 18:353-361. 

Stenton SC, Beach JR, Dennis JH, Keaney NP, and Hendrick DJ. 1994. Glutaraldehyde, asthma 

and work -- a cautionary tale. Occup. Med. 44:95-98. 

Wiggins P, McCurdy SA, and Zeidenberg W. 1989. Epistaxis due to glutaraldehyde exposure. J. 

Occup. Med. 31:854-856. 

Zissu D, Gagnaire F, and Bonnet P. 1994. Nasal and pulmonary toxicity of glutaraldehyde in 

mice. Toxicol. Lett. 71:53-62. 

Appendix D3 291 Glutaraldehyde 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

n-HEXANE 

(normal hexane) 

CAS Registry Number: 110-54-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 7000 μg/m3

 (2000 ppb)

Critical effect(s) Neurotoxicity; electrophysiological alterations in 

humans 

Hazard index target(s) Nervous system 

II. Physical and Chemical Properties (HSDB, 1999)

Description Colorless liquid, gas 

Molecular formula C6H14

Molecular weight 86.10 

Density 0.660 g/cm3

 @ 20° C 

Boiling point 68.95°C 

Melting point −95.3°C 

Vapor pressure 150 torr @ 25° C 

Solubility Insoluble in water; soluble in most organic 

 solvents; very soluble in alcohol 

Conversion factor 1 ppm = 3.52 mg/m3

 @ 25° C 

III. Major Uses or Sources 

n-Hexane is used in the extraction of vegetable oil from seeds such as safflower, soybean, cotton, 

and flax (HSDB, 1995). It is also used as a alcohol denaturant and as a paint diluent. The 

textile, furniture and leather industries use n-hexane as a cleaning agent. Many petroleum and 

gasoline products contain n-hexane. The annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 999,225 pounds of hexane (CARB, 1999).

Appendix D3 292 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure 

In an offset printing factory with 56 workers, symptomatic peripheral neuropathy was noted in 

20 of 56 (36%) workers, while another 26 (46%) had evidence of subclinical neuropathy (Chang 

et al., 1993). Reduced sensory action potentials; reduced motor action potentials; decreased 

motor nerve conduction velocity; and increased distal latency were found in most workers. 

Giant axonal swellings with accumulation of 10 nm neurofilaments, myelin sheath attenuation, 

and widening of nodal gaps were noted upon sural nerve biopsy of a severe case. Optic 

neuropathy and CNS impairment were not usually found. Personal air samples had 80 to 210 

ppm hexane (mean = 132 ppm), 20 to 680 ppm isopropanol (mean = 235 ppm), and 20 to 84 ppm 

(mean = 50 ppm) toluene. The workers worked 12 hours per day for 6 days per week. The mean 

duration of employment was 2.6 years, with a range of 1 month to 30 years. 

An epidemiologic study was performed on workers employed in a factory producing tungsten 

carbide alloys and exposed for an average of 6.2 years to solvent vapors consisting of an 8-hour 

time weighted average of 58 ppm (±41 ppm) n-hexane and 39 ppm (±30 ppm) acetone (Sanagi et 

al., 1980). Neurological examinations performed on both control and exposed workers 

examined cranial nerves, motor and sensory nerves, reflexes, coordination and gait. 

Neurophysiological and nerve stimulation studies were also performed. While no overt 

neurological abnormalities were noted, the mean motor nerve conduction velocity and residual 

latency of the exposed group were significantly decreased as compared to unexposed workers. 

The effects observed are consistent with other reports of n-hexane-induced peripheral 

neuropathy. The study reports a LOAEL of 58 ppm n-hexane. 

Polyneuropathy with subsequent development of muscular atrophy and paresthesia in the distal 

extremities was observed in workers exposed to between 500 and 1000 ppm n-hexane in a 

pharmaceutical plant (Yamada, 1967). 

A group of 15 industrial workers exposed to n-hexane in vegetable oil extracting and adhesive 

bandage manufacturing processes was examined for signs of neurotoxicity and ophthalmological 

changes (Raitta et al., 1978; Seppalainen et al., 1979). The workers (11 males and 4 females) 

had been exposed to hexane for 5 to 21 years (mean of 12 years). Ten healthy workers served as 

controls. Exposures were found to be variable; concentrations as high as 3000 ppm were found 

on some occasions, although exposure concentrations were usually well below 500 ppm. The 

authors concluded that the high short-term exposures, occurring occasionally for 1 to 2 hours at a 

time, could have been major factors in the effects observed. Visual evoked potentials (VEPs) 

were generally reduced among the exposed subjects and latencies tended to be increased 

(Seppalainen et al., 1979). Visual acuity, visual fields, intraocular pressure, and 

biomicroscopical findings were normal. Macular changes were noted in 11 and impaired color 

discrimination was found in 12 of the 15 subjects, largely in the blue-yellow spectrum (Raitta et 

al., 1978). 

Fifteen (25%) of 59 press proofing workers had polyneuropathy (Wang et al., 1986). All of the 

patients with polyneuropathy were regularly exposed to n-hexane, and there was a significant 

association between n-hexane concentration and prevalence of polyneuropathy. The ambient 

concentration of n-hexane of 190 ppm was found in one factory in which all six workers 

Appendix D3 293 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

developed polyneuropathy. Workers exposed to less than 100 ppm n-hexane who frequently 

worked overtime demonstrated significant decreases in motor nerve conduction velocities in 

median, ulnar, and peroneal nerves. Twelve of 13 workers who regularly slept in the factory had 

polyneuropathy compared to three (7%) of 46 employees who did not sleep in the factory. 

Ninety-three of 1662 Japanese workers were found to have polyneuropathy (Yamamura, 1969; 

Sobue et al., 1978). All of the workers developing polyneuropathy were employed in pasting 

with rubber cement containing 70% or more hexane and small amounts of toluene. The 

worksites were poorly ventilated and concentrations in workrooms were measured at between 

500 and 2500 ppm hexane. One patient developed numbness and weakness of the legs after 6 

months of exposure to hexane-based solvents. This patient was hospitalized for over a year until 

the muscle weakness and atophy improved enough to discharge the patient. 

Urinary 2,5-hexanedione concentrations were significantly higher in 35 male workers exposed to 

n-hexane than in an unexposed group (Karakaya et al., 1996). Significant decreases in serum 

IgG, IgM and IgA levels were also found, and a significant correlation was noted between 

urinary 2,5-hexanedione concentrations and serum Ig level of the exposed group. 

An association between n-hexane and parkinsonism has been proposed based on two case reports 

(Pezzoli et al., 1989; 1995). Regional striatal abnormalities of the nigrostriatal dopaminergic 

system and of glucose metabolism, observed with positron emission tomography studies, were 

considered distinct from those seen in idiopathic Parkinson's disease. 

Co-exposure to acetone increased the urinary concentrations of free and total 2,5-hexanedione 

(2,5-HD) in a study of 87 hexane-exposed workers (Cardona et al., 1996). Increased urinary 2,5-

HD is noted also with coexposure to hexane and methyl ethyl ketone (Ichihara et al., 1998). 

V. Effects of Animal Exposure 

Groups of 12 Sprague-Dawley (SD) rats inhaled n-hexane (0, 6, 26, or 129 ppm) for 6 hours/day, 

5 days/week for 26 weeks (Bio/dynamics, 1978). A second experiment from the same report 

involved inhalation exposures of SD rats for 26 weeks to 0, 5, 27, or 126 ppm hexane for 21 

hours/day, 7 days/week. There were no consistent dose-related differences between exposed and 

control animals, although small numbers of animals were involved and examinations were 

limited to physical observation, body weight, hematological parameters, clinical chemistry, and 

necropsy of spontaneous deaths. The highest concentration (126 ppm for 21 hours/day, 7 

days/week) was a NOAEL and represents a time-weighted average exposure of 110.2 ppm over 

the duration of the experiment. 

F-344 rats and B6C3F1 mice (50/sex/concentration/species) inhaled commercial hexane solvent 

(0, 900, 3000, or 9000 ppm) for 6 h/day, 5 days/week over 2 years (Daughtrey et al., 1999). No 

significant differences in mortality were noted between hexane-exposed and control groups. 

Small statistically significant reductions in body weight gain were noted in male and female rats 

inhaling 3000 ppm or more and in female mice inhaling 9000 ppm. Epithelial cell hyperplasia 

was increased in the nasoturbinates and larynx of exposed rats. 

Appendix D3 294 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Fischer 344 rats (5/sex/dose) inhaled >99.5% pure n-hexane (0, 3000, 6500, or 10,000 ppm) for 6 

hours/day, 5 days/week over 13 weeks (Cavender et al., 1984). No statistically significant 

differences were notes in food consumption, ophthalmologic examination, neurological function, 

or hematological or serum chemistry parameters in either males or females. Female body 

weights and clinical observations were unaltered by hexane treatment. The mean body weight 

gain of male rats in the 10,000-ppm group was significantly decreased compared with controls at 

4 weeks of exposure and thereafter. Axonopathy was noted in the tibial nerve of four of five 

male rats exposed to 10,000 ppm and in one of five male rats exposed to 6500 ppm. Axonopathy 

in the medulla was noted in one male rat exposed to 10,000-ppm. Males inhaling 10,000 ppm 

had slightly but significantly lower brain weights. No other adverse histopathological effects 

were reported. This study identifies a NOAEL for neurotoxicity of 3000 ppm, with an average 

experimental exposure of 540 ppm. 

B6C3F1 mice were exposed to 500, 1000, 4000, or 10,000 ppm n-hexane 6 hours per day, 5 days 

per week for 13 weeks or to 1000 ppm n-hexane for 22 hours per day, 5 days per week for 13 

weeks (Dunnick et al., 1989). Mild inflammatory, erosive and regenerative lesions in the 

olfactory and respiratory epithelium were observed in the nasal cavity of mice exposed to 1000 

ppm n-hexane and higher. “Minimal lesions” were noted in those mice exposed to 500 or 1000 

ppm n-hexane. Paranodal axonal swelling in the tibial nerve was observed in 6/8 mice exposed 

to 1000 ppm for 22 hours per day and in 6/8 mice exposed to 10,000 ppm for 6 hours per day. 

No such swelling was noted in neurohistological examination of the control animals; 

neurohistological examination was not performed in those animals exposed to 500 and 1000 ppm 

for 6 hours per day. A NOAEL for histological lesions of the nasal turbinates of 500 ppm nhexane was identified. Because neurohistological examinations were not performed in animals 

exposed to 500 or 1000 ppm (the NOAEL and LOAEL, respectively), the interpretation of the 

results from this study are seriously limited. 

Male SM-A strain mice (10/group) were exposed continuously to 0, 100, 250, 500, 1000, or 2000 

ppm commercial grade hexane (65 to 70% n-hexane with the remainder being other hexane 

isomers) for 6 days/week for 1 year (Miyagaki, 1967). Electromyography, strength-duration 

curves, electrical reaction time, and flexor/extensor chronaxy ratio, gait posture and muscular 

atrophy were studied. Increased complexity of NMU (neuromuscular unit) voltages during 

electromyographic analysis was noted in 0/6 controls, 1/6 in the 100 ppm group, 3/6 in the 250 

ppm group, 5/6 in the 500 ppm group, 3/3 in the 1000 ppm group, and 4/4 in the 2000 ppm 

group. A dose-related increase in incidence and severity of reduced interference voltages from 

muscles was noted in mice exposed to 250 ppm or more, but not in controls (0/6 examined) or in 

the 100 ppm group (0/6). Dose-related abnormal posture and muscle atrophy were noted at 250 

ppm or more. This study identifies a NOAEL of 100 ppm for neurotoxicity (68 ppm when 

adjusted for 67.5% n-hexane). 

Rats inhaling 400-600 ppm n-hexane developed peripheral neuropathy after forty-five days of 

exposure (Schaumburg and Spencer, 1976). Giant axonal swellings and fiber degeneration were 

observed in the central and peripheral nervous systems. The changes were most notable in tibial 

nerves and in the cerebellum, medulla and spinal cord. 

Appendix D3 295 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

A dose-dependent decrease in motor nerve conduction velocity and body weight gain was 

observed in rats exposed to 500, 1200, or 3000 ppm n-hexane for 12 hours per day, 7 days per 

week for 16 weeks (Huang et al., 1989). The neurotoxicity was significant in the two highest 

exposure groups; peripheral nerve degeneration, characterized by paranodal swellings and 

demyelination and remyelination in the myelinated nerve fibers, was observed and was more 

advanced in the highest exposure group. 

Available studies indicate that the neurotoxicity of n-hexane is potentiated by concurrent 

exposure to methyl ethyl ketone (Altenkirch et al., 1982). 

Acetone has also been shown to potentiate the neurotoxicity of hexane and 2,5-HD. Male rabbits 

administered acetone and 2,5-HD intravenously had decreased body clearance of 2,5-HD 

(Lagefoged and Perbellini, 1986). Male rats were treated for 6 weeks with 0.5% w/v 2,5-

hexanedione alone or in combination with 0.50% w/v acetone in the drinking water (Ladefoged 

et al., 1994). Acetone potentiated effects on open field ambulation, or rearing and on the rotarod 

test. Giant axonal swelling was greater in acetone administered animals. During a dose-free 10-

week recovery period, the acetone-supplemented group had less improvement in neurological 

parameters. Male Wistar rats were administered 0.5% w/v 2,5-hexanedione alone or in 

combination with 0.50% w/v acetone in the drinking water for 7 weeks (Lam et al., 1991). 

Effects on radial arm maze behavior, a "brain-swelling" reaction, and synaptosomal functions 

were noted with 2,5-HD and exacerbated with acetone coexposure. In another study of male rats 

using the same doses for 6 weeks, testis weight, testis tubuli diameter and fertility were reduced 

with 2,5-HD exposure and potentiated with acetone coexposure (Larsen et al., 1991). 

Pregnant rats were exposed to 200, 1000, or 5000 ppm n-hexane 20 hours per day on days 9-19 

of gestation (Mast et al., 1987). A statistically significant decrease in fetal body weight 

compared to controls was observed in male offspring following maternal exposure to 1000 and 

5000 ppm n-hexane. Maternal toxicity, indicated by decreased body weight gain, was observed 

in all exposure groups. 

Pregnants rats were exposed to hexane (0, 93.4, or 408.7 ppm) on days 6 through 15 of gestation 

(Litton Bionetics, 1979). There were no adverse effects noted in dams, and no hexane-induced 

teratogenicity, changes in sex ratio, embryotoxicity, or impaired fetal growth or development. 

Male New Zealand rabbits exposed to 3000 ppm n-hexane for 8 hours per day, 5 days per week 

for 24 weeks developed exposure-related lesions of the respiratory tract with the terminal 

bronchioles exhibiting the most characteristic damage (Lungarella et al., 1984). These changes 

were noted even after a 120-day recovery period. Clinical signs of ocular and upper respiratory 

tract irritation and respiratory difficulties (such as gasping, lung rales, mouth breathing) were 

observed throughout the study in exposed rabbits. 

Appendix D3 296 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level 

Three studies, an experimental study with mice (Miyagaki, 1967) and two occupational studies 

(Sanagi et al., 1980; Chang et al., 1993), were considered by OEHHA to be most informative and 

relevant to the derivation of a chronic REL. This was because these studies (1) evaluated the 

most sensitive endpoint (peripheral neuropathy) and (2) involved exposures over a significant 

fraction of a lifetime. While significant limitations may be noted for each of these studies 

individually, viewed collectively they provide a consistent view of the chronic inhalation toxicity 

of hexane and yield a stronger basis for deriving a chronic inhalation REL.

While the animal study has the disadvantage of introducing the uncertainty of interspecies 

differences, the limitations of the human studies were considered to be more significant. 

Specifically, both human studies were considered likely to overestimate effects of inhalation 

exposures to hexane. 

The Sanagi study, which U.S. EPA used as the basis of its RfC, may overestimate hexane effect 

because of a confounding coexposure to acetone, which is known to potentiate hexane 

neuropathy. The minimum effective acetone inhalation concentration for potentiating hexane 

neuropathy is unclear, as studies (Ladefoged et al., 1994; Lam et al., 1991; Larsen et al., 1991) 

have used orally administered acetone. The minimum effective acetone inhalation dose for 

potentiation of carbon tetrachloride hepatotoxicity in male Sprague-Dawley rats was 2500 ppm 

over 4 hours (Charbonneau et al., 1986). A dose of 0.5% acetone in human drinking water is 

comparable, assuming equal absorption, to an inhalation concentration of approximately 1400 

ppm (0.5% w/v x 2 L/day ÷ 2 m3

/day = 5 g/m3; 5 g/m3 x 1000 mg/g ÷ 3.52 mg/m3 per ppm = 

1400 ppm). As the acetone potentiating effects were all noted at higher exposures than are being 

considered in occupational studies and are at much higher concentrations than the REL itself, the 

significance of these findings is uncertain. 

In the Chang study, the workers were probably intermittently exposed to higher inhalation 

exposures than were estimated from ambient air sampling, and significant dermal exposures 

were also likely. Furthermore, coexposure to high levels of isopropanol and toluene, may have 

confounded the results, although CNS effects were not noted and these substances are not known 

to induce or potentiate peripheral neuropathy. 

As shown in Table 1, the human studies by Sanagi et al. (1980) and Chang et al. (1993) yield 7 

to 10-fold lower RELs than the Miyagaki study. In view of the likely overprediction of hexane 

risks from these studies, due to co-exposure to other materials which may potentiate the effects 

of hexane, these calculations may be viewed as generally supporting the 7000 μg/m3 REL. 

Appendix D3 297 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Key study Miyagaki (1967) 

Study population Male mice 

Exposure method Discontinuous inhalation 

Critical effects Peripheral neuropathy (electromyographic 

alterations; dose-related abnormal posture and 

muscle atrophy) 

LOAEL 250 ppm 

NOAEL 100 ppm 

Exposure continuity 24 hours/day, 6 days/week 

Exposure duration 1 year 

Average experimental exposure 57.9 ppm for LOAEL group 

 (100 ppm * 0.675 * 6/7) 

Human equivalent concentration 57.9 ppm (gas with systemic effects, based on 

default RGDR = 1 for lambda (a) = lambda 

(h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 2 ppm (2000 ppb; 7 mg/m3

; 7000 μg/m3) 

Table 1: Reference Exposure Levels (RELs) from Selected Human Studies 

Study Duration Effect LOAEL LOAEL NOAEL NOAEL total REL REL 

(ppm) (ppm) (ppm) (ppm) UF (ppb) (μg/m3)

(TWA) (TWA) 

Sanagi et al., 

1980 

6.2 years decreased motor nerve 

conduction velocity; 

increased residual latency 

58 20.7 Not observed 100a 200 700 

Chang et al., 

1993 

mean 2.6 

years: 

range 1 

month to 

12 years 

Symptomatic peripheral 

neuropathy; decreased 

motor nerve conduction 

velocity; increased 

residual latency; axonal 

swelling of sural nerve 

mean 

132: 

range 

80 - 210

83 Not observed 300b 300 1000 

a LOAEL uncertainty factor, 10; Intraspecies uncertainty factor, 10 b

 LOAEL uncertainty factor, 10; Subchronic uncertainty factor, 3; Intraspecies uncertainty factor, 10

The hexane exposure estimate was reduced for the Miyagaki data as the solvent used contained 

67.5% n-hexane. 

The average occupational exposure for the Chang study involving an unusual 72-hour work 

week was calculated by assuming that 12 hours of occupational exposures at an inhalation rate of 

20 L/min was followed by 4 hours of light work at 20 L/min and 8 hours of rest at 7.5 L/min. 

Using these assumptions an estimated 63% of daily inhaled air occurred at the workplace. 

Appendix D3 298 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

The Chang study found that the severity of effects was not correlated with the length of 

exposure, suggesting that (1) susceptibility may differ markedly between individuals and/or (2) 

shorter exceedances of the time-weighted average concentration might be significant. Thus the 

subchronic uncertainty factor was reduced to 3-fold. 

VII. Data Strengths and Limitations for Development of the REL 

There is a substantial database on the health effects of n-hexane in both humans and animals 

from which to derive a chronic reference exposure level. Some relevant studies are summarized 

in the table below. 

Study Species Exposure 

concentration 

Exposure 

regimen 

TWA 

from 

NOAELa

TWA 

from 

LOAELa

Sanagi et al. 

(1980) 

Humans 58 ppm (mean) 10 m3/d, 5 d/wk, 

6.2 yr (mean) 

None 20.7 ppm 

Chang et al. 

(1993) 

Humans 130 ppm (mean) 12 hr/d, 6 d/wk, 

2.6 yr (mean) 

None 83 ppm 

Miyagaki 

(1967) 

Male 

mice 

0, 100, 250, 500, 

1000, 2000 ppm 

Continuous, 6 

d/wk, 1 yr 

57.9 ppm 121 ppm 

Daughtrey et 

al. (1999) 

F344 

rats 

0, 900, 3000, 

9000 ppm 

6 hr/d, 5 d/wk, 2 

yr 

None 161 ppm 

Daughtrey et

al. (1999) 

B6C3F1 

mice 

0, 900, 3000, 

9000 ppm 

6 hr/d, 5 d/wk, 2 

yr 

None 161 ppm 

Dunnick et al. 

(1989) 

B6C3F1 

mice 

0, 500, 1000, 

4000, 10,000 ppm 

6 hr/d, 5 d/wk, 

13 wk 

89 ppm 179 ppm 

Huang et al. 

(1989) 

Wistar 

rats 

0, 500, 1200, 

3000 ppm 

12 hr/d, 7 d/wk, 

16 wk 

None 250 ppm 

Bio/dynamics 

(1978) 

SD rats 0, 5, 27, 126 ppm 21 hr/d, 7 d/wk, 

26 weeks 

110 ppm None 

Cavender et al. 

(1984) 

F344 

rats 

0, 3000, 6500, 

10,000 ppm 

6 hr/d, 5 d/wk, 

13 wk 

540 ppm 1160 ppm 

a

 The experimental exposure was extrapolated to an equivalent (time-weighted average 

 or TWA) continuous exposure. 

The major strengths of the REL for hexane include (1) the primary use of an animal study 

(Miyagaki, 1967) with controlled, nearly continuous chronic hexane exposures not confounded 

by coexposure to other solvents, which observed both a NOAEL and LOAEL; and (2) the results 

obtained from two different human studies (Sanagi, 1980; Chang et al., 1993) which were 

viewed as being generally consistent with the animal study based REL. 

There is uncertainty about interspecies as well as intraindividual differences in susceptibility to 

n-hexane peripheral neuropathy. In one study, controlled TWA exposures of 540 ppm (Cavender 

Appendix D3 299 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

et al., 1984) were not found to cause neuropathy in rats. Also human studies (especially that of 

Chang et al., 1993) have shown that some individuals develop peripheral neuropathy within 

months, whereas others remain symptom-free despite years of employment at the same 

occupation at the same workplace. 

OEHHA staff also estimated RELs from two other animal studies for comparison. In 

Bio/Dynamics (1978), 126 ppm for 21 hours/day, 7 days/week for 26 months was a NOAEL and 

represents a time-weighted average exposure of 110.2 ppm. Using an RGDR of 1 and a 

cumulative 30-fold uncertainty factor (3 for interspecies differences not accounted for by the 

RGDR method and 10-fold for intraspecies differences), a REL of 4 ppm (10,000 μg/m3

) was 

derived. Cavender et al. (1984) identified a NOAEL for neurotoxicity of 3000 ppm, with an 

average experimental exposure of 540 ppm. A REL based on this study, using an RGDR of 1 

and a 100-fold uncertainty factor (3 for subchronic (13 weeks) to chronic, 3 for interspecies, and 

10 for intraspecies) would be 5.4 ppm (19,000 μg/m3

). 

VIII. References 

Altenkirch H, Wagner HM, Stoltenburg G, and Spencer PS. 1982. Nervous system responses of 

rats to subchronic inhalation of N-hexane and N-hexane + methyl-ethyl-ketone mixtures. J. 

Neurol. Sci. 57:209-219. 

Bio Dynamics Inc. 1978. 26 Week Inhalation Toxicity Study of n-Hexane in the Rat. US EPA; 

Document #FYI-AX-1081-0137. 

Cardona A, Marhuenda D, Prieto MJ, Marti J, Periago JF, Sanchez JM. 1996. Behaviour of 

urinary 2,5-hexanedione in occupational co-exposure to n-hexane and acetone. Int Arch Occup 

Environ Health 68(2):88-93. 

Cavender FL, Casey HW, Salem H, Graham DG, Swenberg JA, Gralla EJ. 1984. A 13-week 

vapor inhalation study of n-hexane in rats with emphasis on neurotoxic effects. Fundam. Appl. 

Toxicol. 4(2 Pt 1):191-201. 

Chang CM, Yu CW, Fong KY, Leung SY, Tsin TW, Yu YL, Cheung TF, Chan SY. 1993. Nhexane neuropathy in offset printers. J. Neurol. Neurosurg. Psychiatry 56(5):538-542. 

Charbonneau M, Brodeur J, du Souich P, Plaa GL. 1986. Correlation between acetonepotentiated CCl4-induced liver injury and blood concentrations after inhalation or oral 

administration. Toxicol. Appl. Pharmacol. 30;84(2):286-294. 

Daughtrey W, Newton P, Rhoden R, Kirwin C, Haddock L, Duffy J, Keenan T, Richter W, 

Nicolich M. 1999. Chronic inhalation carcinogenicity study of commercial hexane solvent in F344 rats and B6C3F1 mice. Toxicol. Sci. 48(1):21-29. 

Dunnick JK, Graham DG, Yang RS, Haber SB, and Brown HR. 1989. Thirteen-week toxicity 

study of n-hexane in B6C3F1 mice after inhalation exposure. Toxicology 57(2):163-172. 

Appendix D3 300 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Huang J, Kato K, Shibate E, Sugimura K, Hisanaga N, Ono Y, Takeuchi Y. 1989. Effects of 

chronic n-hexane exposure on nervous system-specific and muscle-specific proteins. Arch. 

Toxicol. 63(5):381-385. 

Ichihara G, Saito I, Kamijima M, Yu X, Shibata E, Toida M, Takeuchi Y. 1998. Urinary 2,5-

hexanedione increases with potentiation of neurotoxicity in chronic coexposure to n-hexane and 

methyl ethyl ketone. Int. Arch. Occup. Environ. Health 71(2):100-104. 

Karakaya A, Yucesoy B, Burgaz S, Sabir HU, Karakaya AE. 1996. Some immunological 

parameters in workers occupationally exposed to n-hexane. Hum. Exp. Toxicol. 15(1):56-8. 

Ladefoged O, Perbellini L. 1986. Acetone-induced changes in the toxicokinetics of 2,5-

hexanedione in rabbits. Scand. J. Work Environ. Health 12(6):627-629. 

Ladefoged O, Roswall K, Larsen JJ. 1994. Acetone potentiation and influence on the 

reversibility of 2,5- hexanedione-induced neurotoxicity studied with behavioural and 

morphometric methods in rats. Pharmacol. Toxicol. 74(4-5):294-299. 

Lam HR, Larsen JJ, Ladefoged O, Moller A, Strange P, Arlien-Soborg P. 1991. Effects of 2,5-

hexanedione alone and in combination with acetone on radial arm maze behavior, the "brainswelling" reaction and synaptosomal functions. Neurotoxicol. Teratol. 13(4):407-412. 

Larsen JJ, Lykkegaard M, Ladefoged O. 1991. Infertility in rats induced by 2,5-hexanedione in 

combination with acetone. Pharmacol. Toxicol. 69(1):43-46. 

Litton Bionetics Inc. 1979. Teratology Study in Rats. n-Hexane. US EPA; Document #FYI-AX0183-0231. 

Lungarella G, Barni-Comparini I, and Fonzi L. 1984. Pulmonary changes induced in rabbits by 

long-term exposure to n-hexane. Arch. Toxicol. 55(4):224-228. 

Mast TJ, Decker JR, Clark ML, et al. 1987. Inhalation developmental toxicology studies: 

teratology study of n-hexane in rats: Final report. Iss. PNL-6453; Order No. DE88006812. p.208. 

Miyagaki H. 1967. Electrophysiological studies on the peripheral neurotoxicity of n-hexane. Jap. 

J. Ind. Health 9(12-23): 660-671 

Paulson GW, and Waylonis GW. 1976. Polyneuropathy due to n-hexane. Arch. Int. Med. 

136:880-882. [cited in U.S. EPA, 1995]. 

Pezzoli G, Barbieri S, Ferrante C, Zecchinelli A, Foa V. 1989. Parkinsonism due to n-hexane 

exposure. Lancet 2(8667):874. 

Pezzoli G, Antonini A, Barbieri S, Canesi M, Perbellini L, Zecchinelli A, Mariani CB, Bonetti 

A, Leenders KL. 1995. n-Hexane-induced parkinsonism: pathogenetic hypotheses. Mov. Disord. 

10(3):279-82. 

Appendix D3 301 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Raitta C, Seppalainen AN, Huuskonen MS. 1978. N-hexane maculopathy in industrial workers. 

Albrecht Von Graefes Arch. Klin. Exp. Ophthalmol. 209(2):99-110. Sanagi S, Seki Y, Sugimoto 

K, and Hirata M. 1980. Peripheral nervous system functions of workers exposed to n-hexane at a 

low level. Int. Arch. Occup. Environ. Health 47:69-79. 

Schaumburg HH, Spencer PS. 1976. Degeneration in central and peripheral nervous systems 

produced by pure n-hexane: an experimental study. Brain 99(2):183-192.Seppalainen AM, Raitta 

C, and Huuskonen MS. 1979. n-Hexane-induced changes in visual evoked potentials and 

electroretinograms of industrial workers. Electroencephal. Clin. Neurophysiol. 47:492-498. 

Sobue I, Iida M, Yamamura Y, and Takayanagui T. 1978. N-Hexane polyneuropathy. Int. J. 

Neurol. 11(4):317-330. 

U.S. EPA. 1999. U. S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS). Cincinnati, OH: Office of Health and Environmental Assessment, Environmental Criteria 

Assessment Office. 

Wang J-D, Chang Y-C, Kao K-P, Huang C-C, Lin C-C, and Yeh W-Y. 1986. An outbreak of nhexane induced polyneuropathy among press proofing workers in Taipei. Am. J. Ind. Med. 

10(2):111-118. 

Yamada S. 1967. Intoxication polyneuritis in the workers exposed to n-hexane. Jap. J. Ind. 

Health 9:651-659 [cited in U.S. EPA, 1994]. 

Yamamura Y. 1969. n-Hexane polyneuropathy. Folia Psychiatr. Neurol. Jap. 23(1):45-57. 

Appendix D3 302 Hexane 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

HYDRAZINE 

(diamine; diamide; nitrogen hydride; levoxine) 

CAS Registry Number: 302-01-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.2 μg/m3 (0.1 ppb)

Critical effect(s) Amyloidosis of the liver and thyroid in 

hamsters 

Hazard index target(s) Alimentary system; endocrine system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994)

 

Description Colorless, oily liquid or white crystals 

Molecular formula N2H4

Molecular weight 32.05 g/mol 

Boiling point 113.5°C (Merck, 1983; CRC, 1994) 

Melting point 2.0°C 

Vapor pressure 14.4 torr @ 25°C 

Solubility Miscible with water, methyl-, ethyl-, isobutyl 

alcohols; slightly miscible with 

hydrocarbons; insoluble in chloroform, 

ether 

Conversion factor 1.31 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Hydrazine is a highly reactive base and reducing agent. Its primary uses are as a high-energy 

rocket propellant, as a reactant in military fuel cells, in nickel plating, in the polymerization of 

urethane, for removal of halogens from wastewater, as an oxygen scavenger in boiler feedwater 

to inhibit corrosion, and in photographic development (Von Burg and Stout, 1991). Hydrazine 

was historically used experimentally as a therapeutic agent in the treatment of tuberculosis, 

sickle cell anemia, and non-specific chronic illnesses (Von Burg and Stout, 1991; Gold, 1987). 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 1664 pounds of 

hydrazine (CARB, 2000). 

Appendix D3 303 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposure 

One person was occupationally exposed to hydrazine at unknown levels once per week for a 

period of 6 months (Sotaniemi et al., 1971). The worker showed symptoms of conjunctivitis, 

tremors, and lethargy for 1-2 days following each exposure. Vomiting, fever, and diarrhea 

developed on the last day of exposure and progressed to abdominal pain and incoherence. The 

previously healthy 59-year old individual died three weeks after the last exposure. Evidence of 

tracheitis, bronchitis, heart muscle degeneration, and liver and kidney damage was found at 

autopsy. A single case report can not prove a cause and effect relationship between hydrazine 

exposures and the noted symptoms and death, but the repeated association between exposures 

and symptoms is highly suspicious. Liver toxicity is also associated with acute exposure to 

hydrazine. 

The only epidemiological studies of human hydrazine exposures found involve workers in a 

hydrazine manufacturing plant (Wald et al., 1984; Wald, 1985; Morris et al., 1995). Workers 

were exposed to various durations of at least 6 months between 1945 and 1972 and have been 

followed through 1992. The studies are based on a review of medical records. Only 78 of 427 

workers were believed to have had more than incidental exposure to hydrazine. Only cumulative 

mortality was reviewed. Health effects reported during or after hydrazine exposure were not 

examined. No increase in mortality was noted for lung cancer, other cancers, or causes other 

than cancer. However, these small studies have little power to detect increased mortality, and 

age of death was not examined. The authors reported that relative risks up to 3.5 could have 

gone undetected. 

Dermal sensitization has also been reported from repeated contact with hydrazine (Van Ketal, 

1964; Von Keilig and Speer, 1983; Wrangsjo and Martensson, 1986). 

V. Effects of Animal Exposure 

An inhalation study of the toxicity and carcinogenicity of hydrazine was conducted in cats, mice, 

hamsters, and dogs (Vernot et al., 1985). Various animal groups were exposed 6 hours/day, 5 

days/weeks for one year to concentrations of 0.05, 0.25, 1.0, and 5.0 ppm anhydrous hydrazine 

base. Exposed and controls groups were made up of the following animals: 100 Fischer 344 

rats/sex at 0.05, 0.25, 1.0, and 5.0 ppm hydrazine plus 150 rats/sex as controls; 400 female 

C57BL/6 mice at 0.05, 0.25, and 1.0 ppm hydrazine plus 800 female mice as controls; 200 male 

Golden Syrian hamsters at 0.25, 1.0, and 5.0 ppm hydrazine plus 200 male hamsters as controls; 

4 beagle dogs/sex at 0.25 and 1.0 ppm hydrazine plus 4 dogs/sex as controls. Animals were 

observed post-exposure for the following periods: 18 months for rats, 15 months for mice, 12 

months for hamsters, and 38 months for dogs. Animals were observed hourly during the 

exposure period and daily in the post-exposure period. 

No non-cancer toxic effects were observed in mice or dogs, with the exception of a single dog, 

exposed to 1.0 ppm hydrazine, which showed cyclic elevations in serum glutamic-pyruvic 

transaminase levels and, upon necropsy at 36 months post-exposure, showed liver effects 

described as “clusters of swollen hepatocytes that had highly vacuolated cytoplasm.” Of the 

Appendix D3 304 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

other species examined, hamsters showed toxicity at the lowest dose levels, particularly 

amyloidosis in various organs including liver, spleen, kidney, thyroid, and adrenal glands. An 

increased incidence of amyloidosis was seen at the lowest exposure level (0.25 ppm hydrazine) 

in the liver and thyroid (67/160 exposed vs. 42/180 control for the liver and 20/117 exposed vs. 

9/155 control in the thyroid; p ≤ 0.01 by Fisher’s exact test). This effect was found to be dose 

related. The incidence of hemosiderosis of the liver was also significantly increased in all 

exposed groups. Significantly increased incidences of toxic effects observed in the 1.0 and 5.0 

ppm hydrazine groups include amyloidosis of the spleen, kidney glomerulus, and adrenals 

glands, and lymphadenitis of the lymph nodes. Significantly increased toxic effects observed 

only in the highest dose group include amyloidosis of the kidney interstitium and thyroid, and 

senile atrophy of the testis. The authors note these effects appear to reflect accelerated changes 

commonly associated with aging in hamsters. 

 Incidence of Nonneoplastic Lesions in Male Hamsters (from Table 3 of Vernot et al.) 

Lesion Control 0.25 ppm 1.0 ppm 5.0 ppm 

Liver

 Amyloidosis 42/180 (23)* 67/160 (42)a 68/148 (46)a 79/159 (50)a

 Hemosiderosis 42/180 (23) 63/160 (39)a 77/148 (52)a 94/159 (59)a

 Bile duct hyperplasia 14/180 (8) 31/160 (19)a 28/148 (19)a 44/159 (28)a

 Biliary cyst 45/180 (25) 45/160 (28) 42/148 (28) 55/159 (35)b

Thyroid 

 Amyloidosis 9/155 (6) 20/117 (17)a 11/127 (9) 22/137 (16)a

Adrenal

 Amyloidosis 38/177 (22) 49/199 (32)b 52/141 (37)a 76/153 (50)a

* Incidence of lesion (% of animals with lesion) a

 Incidence significantly greater than control, p ≤ 0.01 

b Incidence significantly greater than control, 0.01 < p ≤ 0.05 

In the hydrazine exposed rats, effects were observed in the respiratory tract of exposed animals. 

Specifically, squamous metaplasia of the larynx, trachea, and nasal epithelium (males only) was 

observed in the highest dose group (5.0 ppm hydrazine). Inflammation was also observed in the 

larynx and trachea of rats exposed to 5.0 ppm hydrazine. Increased incidence of focal cellular 

change of the liver was observed in female mice at 1.0 and 5.0 ppm hydrazine. Other effects 

with increased incidence only in the high dose group include hyperplastic lymph nodes in 

females, endometriosis, and inflammation of the uterine tube. 

The toxic effects from inhalation of hydrazine over a six month period from both intermittent 

and continuous exposure scenarios were examined (Haun and Kinkead, 1973). Groups of 8 

male beagle dogs, 4 female rhesus monkeys, 50 male Sprague-Dawley rats, and 40 female ICR 

rats per dose group were continuously exposed to 0.2 or 1.0 ppm hydrazine or intermittently (6 

hours/day, 5 days/week) to 1.0 or 5.0 ppm hydrazine. A control group consisted of equal 

numbers of animals. The experimental design was such that each intermittent exposure group 

had a time-weighted-average matching continuous exposure group. Dose-related body weight 

reductions were observed in all treated groups as well as evidence of hepatic degeneration, fatty 

Appendix D3 305 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

deposition in the liver, central nervous system depression and lethargy, eye irritation, and 

anemia. 

Toxic effects from the exposure of rats, mice, and dogs to airborne hydrazine at levels of 0, 4.6, 

or 14 ppm intermittently for 6 months were reported (Comstock et al., 1954). Observed adverse 

effects included anorexia, irregular breathing, vomiting, fatigue, and emphysema in dogs; 

pulmonary congestion and emphysema in rats and mice; and lung and liver damage in rats. 

Lymphoid bronchial hyperplasia was observed in guinea pigs exposed to 2-6 ppm hydrazine for 

5 days/week for 19-47 days (Weatherby and Yard, 1955). 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Vernot et al., 1985 

Study population Hamster 

Exposure method Inhalation of 0, 0.25, 1, and 5 ppm 

Critical effects Amyloidosis and hemosiderosis of the liver; 

thyroid amyloidosis 

LOAEL 0.25 ppm 

NOAEL Not observed 

Exposure continuity 6 hour/day, 5 days/week 

Exposure duration 1 year 

Average experimental exposure 0.045 ppm for LOAEL group (0.25 x 6/24 x 5/7) 

Human equivalent concentration 0.045 ppm for LOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

LOAEL uncertainty factor 10 (low incidence above controls but serious 

adverse effects) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.0001 ppm (0.1 ppb, 0.0002 mg/m3

, 0.2 μg/m3 ) 

Vernot et al.(1985) present a thorough examination of chronic health effects from inhalation 

exposure to hydrazine. This study was chosen for the development of the chronic reference 

exposure level because (1) it was conducted with an adequate number of animals, (2) the 

critical/sensitive adverse effect (degenerative change in the liver in hamsters) showed a doseresponse relationship, and (3) the findings of this study support data found in studies by other 

groups. 

This study shows a dose-related increase in the incidence of amyloidosis and hemosiderosis in 

hamsters intermittently exposed by inhalation to levels of hydrazine greater than 0.25 ppm. Other 

effects noted at 0.25 ppm included weight depression during exposure, mineralization of the 

kidney, and amyloidosis of the thyroid. Haun and Kinkead (1973) have also noted lesions of the 

Appendix D3 306 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

liver in dogs, monkeys, and mice exposed continuously to 0.2 ppm hydrazine for 6 months by 

inhalation. Comstock et al. (1954) observed liver damage in groups of rats exposed to hydrazine 

vapors. The single case report of hydrazine inhalation toxicity in humans showed necrosis and 

degeneration of the liver (Sotaniemi et al., 1971). 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for hydrazine include the availability of chronic inhalation 

exposure data from a well-conducted study with histopathological analysis. Major areas of 

uncertainty are the lack of adequate human exposure data, the lack of reproductive and 

developmental toxicity studies, and the lack of observation of a NOAEL in the key study. 

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Comstock CC, Lawson LH, Greene EA, and Oberst FW. 1954. Inhalation toxicity of hydrazine 

vapor. Arch. Ind. Hyg. Occup. Med. 10:476-490. 

Gold J. 1987. Hydrazine sulfate: a current perspective. Nutr. Cancer 9:59-66. 

Haun CC, and Kinkead ER. 1973. Chronic inhalation toxicity of hydrazine. Proceedings of the 

4th Annual Conference on Environmental Toxicology. Wright-Patterson Air Force Base, OH. 

AMRL-TR-73-125:351-363. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Merck. 1983.The Merck Index. 10th ed. Windholz M. (ed.) Rahway, NJ: Merck & Co. 

Morris J, Densem JW, Wald NJ, Doll R. 1995. Occupational exposure to hydrazine and 

subsequent risk of cancer. Occup. Environ. Med. 52(1):43-45. 

Sotaniemi E, Hivonen J, Isomaki H, Takkunen J, and Kaila J. 1971. Hydrazine toxicity in the 

human. Report of a fatal case. Ann. Clin. Res. 3:30-33. 

Van Ketal WG. 1964. Contact dermatitis from a hydrazine derivative in a stain remover. Crosssensitization to apresoline and isoniazid. Acta Derm. Venereol. (Stockh.) 44:49-53. 

Appendix D3 307 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Vernot EH, MacEwen JD, Bruner RH, Haun CC, Kinkead ER, Prentice DE, Hall A, 

Schmidt RE, Eason RL, Hubbard GB, and Young JT. 1985. Long-term inhalation toxicity of 

hydrazine. Fundam. Appl. Toxicol. 5:1050-1064. 

Von Burg R, and Stout T. 1991. Toxicology Update: Hydrazine. J. Appl. Toxicol. 11:447-450. 

Von Keilig, W. and Speer U 1983. Occupational allergic contact dermatitis from hydrazine. 

Dermatosen in Beruf und Umwelt 31:25-27. 

Wald N, Boreham J, Doll R, and Bonsall J. 1984. Occupational exposure to hydrazine and 

subsequent risk of cancer. Br. J. Ind. Med. 41(1):31-34. 

Wald N. 1985. Hydrazine: epidemiological evidence. IARC Scientific Publication 65:75-80. 

Weatherby JH, and Yard AS. 1955. Observations on the subacute toxicity of hydrazine. Arch. 

Ind. Health 11:413. 

Wrangsjo K, and Martensson A. 1986. Hydrazine contact dermatitis from gold plating. Contact 

Dermatitis, 15:244-245. 

Appendix D3 308 Hydrazine 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

HYDROGEN CHLORIDE 

(Hydrochloric acid; anhydrous hydrogen chloride; muriatic acid) 

CAS Registry Number: 7647-01-0 

I. Chronic Reference Exposure Level 

Inhalation reference exposure level 9 μg/m3

 (6 ppb)

Critical effect(s) Hyperplasia of nasal mucosa, larynx, and trachea 

in rats 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1999) 

Description Colorless gas 

Molecular formula HCl 

Molecular weight 36.46 

Density 1.49 g/L @ 25° C 

Boiling point -84.9° C (HCl gas) 

Melting point -114.8° C (HCl gas) 

Solubility Soluble in water, alcohol, benzene, ether; 

insoluble in hydrocarbons 

Conversion factor 1 ppm = 1.49 mg/m3

 at 25°C 

III. Major Uses or Sources 

Hydrogen chloride (HCl) is used in the manufacture of vinyl chloride, fertilizers, dyes, artificial 

silk, and pigments for paints. It is also used in electroplating, soap refining, and leather tanning. 

Other consumers of HCl include the photographic, textile and rubber industries (HSDB, 1999). 

Hydrogen chloride is produced in large quantities during combustion of most materials and 

especially materials with a high chlorine content. Thus, HCl is a major product formed during 

the thermal decomposition of polyvinyl chloride, a commonly used plastic polymer (BurleighFlayer et al., 1985). It is also released in large quantities during the test firing of some rocket 

and missile engines (Wohlslagel et al., 1976). Since HCl is extremely hygroscopic, it generally 

exists as an aerosol in the ambient atmosphere. The annual statewide industrial emissions from 

facilities reporting under the Air Toxics Hot Spots Act in California based on the most recent 

inventory were estimated to be 2,570,888 pounds of HCl (CARB, 1999b).

Appendix D3 309 Hydrogen Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Human Exposure 

Few reports are available on the effects of chronic HCl exposure on humans. Bleeding of the 

nose and gums and ulceration of the mucous membranes was observed following repeated 

occupational exposure to HCl mist at high but unquantified concentrations (Stokinger, 1981). 

In another report, workers exposed to various mineral acids, including HCl, exhibited etching 

and erosion of the front teeth (Ten Bruggen Cate, 1968). Dental erosion was noted in 176 of 555 

(32%) workers examined between 1962 and 1964, and progressive erosion was reported in 66 of 

324 (20%) workers examined repeatedly. Rates of active erosion were highest (50%) in the most 

highly-exposed category (battery formation workers), intermediate (23%) in an intermediateexposure category (picklers), and low (7%) in a low-exposure category (other processes). Grade 

1 erosion (enamel loss) was noted in workers exposed for greater than 3 months; grade 2 erosion 

(loss of enamel and dentine) was noted after 2.5 to 5 years exposure; and grade 3 (loss of enamel 

and dentine with exposure of secondary dentine) was noted after six or more years of exposure. 

V. Effects of Animal Exposure 

Male Sprague-Dawley rats were exposed to 10 ppm HCl for 6 hours per day, 5 days per week 

over their lifetime (Sellakumar et al., 1985). No differences in body weights or survival were 

observed between 99 exposed and 99 control animals. Increased incidences of hyperplasia of the 

nasal mucosa (62/99 vs. 51/99), larynx (22/99 vs. 2/99), and trachea (26/99 vs. 2/99) were 

observed in exposed rats compared to air-exposed controls. 

A 90-day inhalation study using B6C3F1 mice and Sprague-Dawley and Fisher 344 rats exposed 

the animals (groups of 31 males and 31 females for each species and strain) to 10, 20, or 50 ppm 

HCl for 6 hours per day, 5 days per week over 90 days (Toxigenics, 1984). Several animals died 

during the study, though the deaths were not considered to be exposure related. A slight but 

significant decrease in body weight gain was reported in male and female mice and in male 

Fischer 344 rats in the high-exposure groups. No effect were noted in hematology, clinical 

chemistry, or urinalysis. Minimal or mild rhinitis was observed in both strains of rats 

Concentration- and time-related lesions were noted in the anterior portion of the nasal cavity of 

exposed rats. Cheilitis, eosinophilic globules in the nasal epithelium and accumulation of 

macrophages in the peripheral tissues were observed in mice of all exposed groups. This study 

thus observed a LOAEL for both mice and rats of 10 ppm. The U.S. EPA considered this study 

supportive of the portal-of-entry effects observed at 10 ppm in the lifetime rat study (USEPA, 

1999). Female rats (8-15/group) exposed to 302 ppm HCl for 1 hour either 12 days prior to 

mating or on day 9 of gestation exhibited severe dyspnea and cyanosis; the exposure was lethal 

to one-third of the exposed animals (Pavlova, 1976). Fetal mortality was significantly higher in 

rats exposed during pregnancy. Organ functional abnormalities observed in offspring exposed at 

2-3 months of age were reported to be similar to those observed in the exposed dams. 

Female rats were exposed to 302 ppm HCl for 1 hour prior to mating (GEOMET Technologies, 

1981). Exposure killed 20 to 30% of the rats. In rats surviving 6 days after exposure, a decrease 

in blood oxygen saturation was reported, as were kidney, liver, and spleen effects. Estrus cycles 

Appendix D3 310 Hydrogen Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

were also altered. In rats mated 12-16 days postexposure and killed on day 21 of pregnancy, a 

decrease in fetal weight, an increase in relative fetal lung weights, and reduced numbers of live 

fetuses were observed. 

Derivation of Chronic Reference Exposure Level 

Study Sellakumar et al., 1985 

Study population Sprague-Dawley rats (100 males) 

Exposure method Discontinuous whole-body inhalation (0 or 10 

ppm) 

Critical effects Hyperplasia of the nasal mucosa, larynx and 

trachea 

LOAEL 10 ppm 

NOAEL Not identified 

Exposure continuity 6 hours per day, 5 days per week 

Average experimental exposure 1.8 ppm for LOAEL group 

Human equivalent concentration 0.57 ppm (gas with extrathoracic respiratory 

effects, RGDR =0.32, based on rat 

MVa = 0.33 L/min, MVh = 13.8 L/min, 

SAa(ET) = 15 cm2

); Sah = 200 cm3) (U.S. 

EPA, 1994) 

Exposure duration Lifetime 

LOAEL uncertainty factor 3 (<30% incidence; mild effect) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Reference Concentration (RfC) 0.006 ppm (6 ppb; 0.009 mg/m3

; 9 μg/m3) 

Both extrathoracic and tracheobronchial effects have been associated with exposures to hydrogen 

chloride. The REL was based on extrathoracic effects as humans are predicted to be relatively 

more susceptible to the effects of hydrogen chloride in that region. An intermediate LOAEL 

factor was used as the effects were both mild and occurring at a low incidence at the dose tested. 

VII. Data Strengths and Limitations for Development of the REL 

The USEPA based its RfC of 7 μg/m3 on the same study. U.S. EPA evaluated this RfC as a 

having a low level of confidence because of (1) the use of only one dose; (2) limited toxicity 

evaluation; (3) the lack of reproductive toxicity data; and (4) the lack of chronic exposure studies 

(U.S. EPA, 1994). OEHHA agrees with this assessment. The database for chronic exposure to 

this common chemical is limited. 

Appendix D3 311 Hydrogen Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VIII. References 

Burleigh-Flayer H, Wong KL, and Alarie Y. 1985. Evaluation of the pulmonary effects of HCl 

using CO2 challenges in guinea pigs. Fundam. Appl. Toxicol. 5:978-985. 

CARB. 1999b. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

GEOMET Technologies, Inc. 1981. Hydrogen chloride: Report 4, Occupational Hazard 

Assessment. U.S. Department of Health and Human Services, NIOSH, Cincinnati, OH. NTIS 

PB83-105296. 

HSDB. 1999. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Pavlova TE. 1976. Disturbance of development of the progeny of rats exposed to hydrogen 

chloride. Bull. Exp. Biol. Med. 82(7):1078-1081. 

Sellakumar AR, Snyder CA, Solomon JJ, and Albert RE. 1985. Carcinogenicity of formaldehyde 

and hydrogen chloride in rats. Toxicol. Appl. Pharmacol. 81:401-406. 

Stokinger HE. 1981. Hydrogen chloride, HCl. In: Clayton GD, and Clayton FE, eds. Patty’s 

Industrial Hygiene and Toxicology. 3rd rev. ed. Volume 2B, Toxicology. New York: Wiley 

Interscience. pp.2959-2961. 

Ten Bruggen Cate H.J. 1968. Dental erosion in industry. Br. J. Ind. Med. 25:249-266 

Toxigenics Inc. 1984. 90-Day inhalation study of hydrogen chloride gas in B6C3F1 mice, 

Sprague-Dawley rats and Fischer-344 rats. Study conducted for CIIT, Research Triangle Park, 

NC. CIIT Docket No. 20915. 

U.S.EPA.1999. United States Environmental Protection Agency. Integrated Risk Information 

System (IRIS). (CD-ROM Version) Washington, D.C: US Environmental Protection Agency. 

U.S. EPA. 1990. United States Environmental Protection Agency. Interim Methods for 

Development of Inhalation Reference Concentrations. Environmental Criteria and Assessment 

Office, Office of Health and Environmental Assessment, Office of Research and Development, 

U.S. EPA, Research Triangle Park, NC. Review Draft.

Wohlslagel J, Di Pasquale LC, and Vernot EH. 1976. Toxicity of solid rocket motor exhaust of 

HCl, HF, and alumina on rodents. J. Combustion Toxicol. 3:61-70. 

Appendix D3 312 Hydrogen Chloride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

HYDROGEN CYANIDE 

(Formonitrile; hydrocyanic acid; prussic acid) 

CAS Registry Number: 74-90-8 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 9 μg/m3

 (8 ppb)

Critical effect(s) CNS effects, thyroid enlargement, and 

hematological disorders in workers 

Hazard index target(s) Nervous system; endocrine system; 

 cardiovascular system 

II. Physical and Chemical Properties (HSDB, 1999 ) 

Description Colorless liquid/gas 

Molecular formula HCN 

Molecular weight 27.03 

Boiling point 25.6 o

C 

Melting point -13.4 o

C 

Vapor pressure 630 torr @ 20o

C 

Solubility Miscible in water, alcohol; slightly soluble 

 in ether 

Conversion factor 1 ppm = 1.10 mg/m3 @ 25 o

C 

III. Major Uses or Sources

Hydrogen cyanide is used in a variety of syntheses including the production of adiponitrile (for 

nylon), methyl methacrylate, sodium cyanide, cyanuric chloride, chelating agents, 

pharmaceuticals, and other specialty chemicals. Manufacturing activities releasing hydrogen 

cyanide include electroplating, metal mining, metallurgy and metal cleaning processes. 

Additionally, hydrogen cyanide has some insecticide and fungicide applications (ATSDR, 1993). 

Fires involving some nitrogen-containing polymers, often found in fibers used in fabrics, 

upholstery covers, and padding, also produce hydrogen cyanide (Tsuchiya and Sumi, 1977). 

Another common source of hydrogen cyanide is cigarette smoke. Levels in inhaled mainstream 

cigarette smoke range from 10 to 400 μg per cigarette (U.S. brands); 0.6% to 27% (w/w) of these 

mainstream levels are found in secondary or sidestream smoke (Fiskel et al., 1981). The annual 

statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

Appendix D3 313 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

California based on the most recent inventory were estimated to be 188,665 pounds of hydrogen 

cyanide (CARB, 1999b). 

IV. Effects of Human Exposure

Occupational epidemiological studies investigating hydrogen cyanide exposure are complicated 

by the mixed chemical environments created by synthetic and metallurgic processes. However, 

several reports indicate that chronic low exposure to hydrogen cyanide can cause neurological, 

respiratory, cardiovascular, and thyroid effects (Blanc et al., 1985; Chandra et al., 1980; El 

Ghawabi et al., 1975). Although these studies have limitations, especially with incomplete 

exposure data, they also indicate that long-term exposure to inhaled cyanide produces CNS and 

thyroid effects. 

El Ghawabi et al. (1975) studied 36 male electroplating workers in three Egyptian factories 

exposed to plating bath containing 3% copper cyanide, 3% sodium cyanide, and 1% sodium 

carbonate. Breathing zone cyanide concentrations ranged from 4.2 to 12.4 ppm (4.6 to 13.7 

mg/m3

), with means from 6.4 to 10.4 ppm (7.1 to 11.5 mg/m3), in the three factories at the time 

of this cross-sectional study. The men were exposed for a duration of 5 to 10 years, except for 

one man with 15 years exposure. Twenty non-exposed male volunteers were used as controls. 

None of the subjects, controls or workers, currently smoked cigarettes. Complete medical 

histories were taken, and medical exams were performed. Urinary levels of thiocyanate (a 

metabolite of cyanide) were utilized as a biological index of exposure. Thyroid function was 

measured as the uptake of radiolabeled iodine, since thiocyanate may block the uptake of iodine 

by the thyroid leading to iodine-deficiency goiters. Frequently reported symptoms in the 

exposed workers included headache, weakness, and altered sense of taste or smell. Lacrimation, 

abdominal colic, and lower stomach pain, salivation, and nervous instability occurred less 

frequently. Increased blood hemoglobin and lymphocyte counts were present in the exposed 

workers. Additionally, punctate basophilia were found in 78% (28/36) of the exposed subjects. 

Twenty of the thirty six exposed workers had thyroid enlargements, although there was no 

correlation between the duration of exposure with either the incidence or the degree of 

enlargement. Thyroid function test indicated significant differences in uptake between controls 

and exposed individuals after 4 and 24 hours. Urinary excretion of thiocyanates correlated with 

the breathing zone concentrations of cyanides. Symptoms persisted in 50% of the dyspneic 

workers in a 10-month nonexposure follow up period. This study reported a LOAEL of 6.4 ppm 

(7.1 mg/m3

) for the CNS symptoms and thyroid effects. 

Another retrospective study (Blanc et al., 1985) examined 36 former silver-reclaiming workers 

with long-term exposure to hydrogen cyanide fumes. The authors found significant trends 

between the incidence of self-reported CNS symptoms during active employment (headache, 

dizziness, nausea, and bitter almond taste), the symptoms reported post-exposure, and a 

qualitative index of exposure retroactively defined by the investigators as low-, moderate-, or 

high-exposure through work histories. Some symptoms persisted for 7 months or more after 

exposure. None of the workers had palpable thyroid gland abnormalities, but clinical tests 

revealed decreases in vitamin B12 absorption and folate levels and statistically significant 

Appendix D3 314 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

increases in thyroid-stimulating hormone levels, which in combination with the CNS effects, 

suggest long-term adverse effects associated with cyanide exposure. 

Due to the systemic nature of the lesions produced by cyanide, orally ingested cyanide will likely 

result in injuries similar to that seen by inhalation exposure. Cassava root, a dietary staple in 

many tropical regions, contains cyanogenic glycosides such as linamarin which release cyanide 

(CN-

) when metabolized endogenously (Sharma, 1993; Kamalu, 1995). Consumption of 

insufficiently processed cassava roots over a period of time in combination with a protein 

deficient diet has been implicated in neurotoxic effects. One such neuropathy known as konzo 

results in nerve cell degeneration leading to a permanent but non-progressive spastic weakness of 

the legs and degeneration of corresponding corticospinal pathways (Tylleskar et al., 1992; TorAgbidye et al., 1999). The development of this sydrome is hypothesized to depend on (a) the 

amount and duration of exposure to dietary cyanide, and (b) the ability of the body to detoxify 

cyanide, a function that may vary with nutritional status. The endogenous conversion of cyanide 

to cyanate (OCN-) is thought to be a contributor to the neurotoxic symptoms, but other 

substances found in cassava flour have been implicated (Obidoa and Obasi, 1991; Tor-Agbidye 

et al., 1999; Kamalu, 1995). Tylleskar et al. (1992) determined daily cassava flour consumption 

at above 0.5 kg per adult in a konzo-affected, albeit malnourished, African population. Thus, the 

potential daily cyanide exposure was estimated to be 0.5-1 mmol (13-26 mg), which correlated 

well with urinary concentrations of the metabolite, thiocyanate. A similar daily cyanide intake 

via cassava ingestion was estimated at 15-31.5 mg (approximately 0.2-0.45 mg/kg) following a 

major outbreak of konzo in Mozambique (Casadei et al.,1984; Cliff et al., 1984). 

Other effects associated with cassava consumption include pancreatic diabetes, vitamin B12

deficiency and decreased iodine uptake (Sharma, 1993; Jansz and Uluwaduge, 1997). Cretinism 

in children, associated with a deficiency of dietary iodine, is worsened by eating cassava (Miller, 

1974). Excess thiocyanate due to cyanide metabolism results in a depressed uptake of iodine by 

the thyroid gland that may lead to symptoms of iodine deficiency, including goiter. A 

comparison of three villages in Ethiopia observed increased total goiter rate with increasing rate 

of cassava consumption (Abuye et al., 1998). Goiter was also more prevalent in females and in 

individuals under 20 years of age. In one village, the incidence of goiter increased following the 

introduction of cassava, indicating that cassava exacerbated pre-existing iodine deficiency. 

Urinary iodine levels of school children revealed marginal dietary consumption of iodine, but 

were within the normal range. However, low T4 and high TSH levels indicated insufficient 

iodine uptake by the thyroid gland due to cassava consumption. 

V. Effects of Animal Exposures 

There is little animal data for chronic inhalation exposure to hydrogen cyanide; only two 

subchronic studies were noted by U.S. EPA, one in rabbits (Hugod, 1979, 1981) and the other in 

dogs (Valade, 1952). Continuous exposure of rabbits to 0.5 ppm HCN (0.55 mg/m3) for either 1 

or 4 weeks produced no microscopically detectable morphological changes of the lungs, 

pulmonary arteries, coronary arteries or aorta. This study observed a subacute inhalation NOAEL 

for HCN in rabbits of 0.5 ppm (Hugod, 1979, 1981). Four dogs exposed to 50 mg/m3

 (45 ppm) 

hydrogen cyanide in a series of 30-minute inhalation periods conducted at 2-day intervals 

Appendix D3 315 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

demonstrated extensive CNS toxicity, including dyspnea and vomiting, with vascular and 

cellular CNS lesions identified post-mortem (Valade, 1952). 

Male Sprague-Dawley rats were administered potassium cyanide (0, 40, 80, or 160 mg KCN/kg 

bw-day) in the drinking water for 13 weeks (Leuschner et al., 1991). At the highest dose, blood 

cyanide concentrations were between 16 and 26 mmol CN-

/ml blood and thiocyanate ranged 

between 341 and 877 mmmol SCN-

/ml plasma. The high dose group exposure was reduced to 

140 mg/kg-day after 12 weeks because of decreased body weight gain, reduced drinking water 

consumption, and mortality in this group. 

Male New Zealand white rabbits (6 per group) were administered potassium cyanide in the diet 

over a 40 week experiment (Okolie and Osagie, 1999). The average cyanide intake was 36.5 

mg/day. Based on the growth data presented in the report, cyanide intake was estimated at 

approximately 20 mg/kg-day. The cyanide exposed group had higher feed consumption with 

reduced weight gain, and focal necrosis was noted in the liver and kidney. 

Male weanling rats (strain not identified, 10 animals per group) were administered potassium 

cyanide (1500 ppm) in the diet for 11.5 months (Philbrick et al., 1979). There were no deaths or 

overt signs of toxicity. There was a reduction in body weight gain in the exposed group. Myelin 

degeneration was noted in the spinal cord white matter of cyanide exposed animals. 

Kamalu (1993) fed groups of dogs (6/group; strain not specified) either a control diet containing 

rice as the carbohydrate source, a diet with cassava as a carbohydrate source, or a control diet 

containing NaCN, for 14 weeks. Both the cassava and NaCN diets were adjusted to release 10.8 

mg HCN/kg cooked food. Growth was depressed only in the dogs fed rice + NaCN. Plasma 

thiocyanate was significantly lower in dogs fed cassava compared to dogs fed rice + NaCN. 

These effects indicate that all the intact cyanogenic glycosides absorbed from cassava, primarily 

linamarin, was not hydrolyzed to HCN. However, evidence of liver inflammation and 

hemorrhage were observed only in the cassava fed dogs. Kidney, adrenal, myocardial, and 

testicular lesions were noted in both treated groups but were considered more severe in the 

cassava fed dogs. It was concluded that the lesions observed in the cassava fed dogs were not 

entirely due to cyanide. 

No information was found regarding developmental and reproductive effects in humans for any 

route of hydrogen cyanide exposure. No animal studies utilizing dermal exposure have been 

reported for either hydrogen cyanide or cyanide salts. Dietary studies of the high cyanogenic 

glycoside cassava diet have shown adverse effects, increased runting and decreased ossification 

in hamsters (Frakes et al., 1986), but not in rats fed cassava alone, or supplemented with 

potassium cyanide (Tewe and Maner, 1981). Hamsters with gestational cassava exposure did not 

display reproductive effects (Frakes et al., 1986). 

Appendix D3 316 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level 

Study El Ghawabi et al. (1975); U.S. EPA (1994)

Study population 36 male electroplating workers

Exposure method Discontinuous occupational inhalation exposures

Critical effects CNS effects, thyroid enlargement, and 

 hematological disorders

LOAEL 7.1 mg/m3

NOAEL Not observed

Exposure continuity 8 hr/day (10 m3

/day/20 m3/day), 5 days/week 

Average occupational exposure 2.5 mg/m3

 for LOAEL group 

Human equivalent concentration 2.5 mg/m3

 for LOAEL group 

Exposure duration 5 to 10 years (except one man for 15 years)

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 3

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.008 ppm (8 ppb, 0.009 mg/m3

, 9 μg/m3) 

The USEPA based its RfC of 3 μg/m3 on the same study but included a Modifying Factor (MF) 

of 3 for lack of chronic and multigenerational reproduction studies. The criteria for use of 

modifying factors are not well specified by U.S. EPA. Such modifying factors were not used by 

OEHHA. OEHHA used a 3-fold subchronic uncertainty factor because most workers were 

exposed for less than ten years (78%) and many were exposed for less than 5 years (39%).. 

Appendix D3 317 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

An alternative analysis was conducted using data from an animal ingestion study reporting 

effects at low cyanide concentrations: 

Study Jackson (1988)

Study population Miniature swine

Exposure method Daily oral administration of aqueous potassium 

cyanide

Critical effects Behavioral effects; decreased blood T3 and T4

LOAEL 0.4 mg/kg-day

NOAEL Not observed

Exposure continuity Apparently 7 days per week 

Average exposure 0.4 mg/kg-day (1.4 mg/m3

 for LOAEL group 

assuming 20 m3/day inhalation by a 70 kg 

person) 

Human equivalent concentration Not derived due to lack of species-specific data 

Exposure duration 24 weeks

LOAEL uncertainty factor 3 (minimal effects at lowest dose) 

Subchronic uncertainty factor 10 (based on assumed 27 year lifespan)

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 3,000 

Inhalation reference exposure level 0.0005 mg/m3

 (0.5 μg/m3; 0.0004 ppm; 0.4 ppb) 

This study reported neurobehavioural and thyroid effects at cyanide exposure levels (equivalent 

to 1.4 to 4.2 mg/m3) similar to that reported by El Ghawabi (2.5 mg/m3). However, as greater 

uncertainty factors are required for use of the animal study, a lower REL was derived. Use of a 

cross-route extrapolation also introduces uncertainty. Therefore the REL derived from the 

human data is more appropriate. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the RfC for hydrogen cyanide is the use of human health effects data. The 

major uncertainties are the lack of a NOAEL observation in the key study, the difficulty in 

estimating exposures, and the discontinuous and variable nature of the exposures. 

VIII. References

Abuye C, Kelbessa U, and Wolde-Gebriel S. 1998. Health effects of cassava consumption in 

South Ethiopia. East African Med. J. 75(3):166-170. 

ATSDR. 1993. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Cyanide. U.S. Dept. of Health and Human Services, Public Health Service. TP-92/09. 

Ansell M, and Lewis FAS. 1970. A review of cyanide concentrations found in human organs: A 

survey of literature concerning cyanide metabolism, ‘normal’, non-fatal and fatal body cyanide 

levels. J. Forensic Med. 17:148-155. 

Appendix D3 318 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Blanc P, Hogan M, Mallin K, Hryhorczuk D, Hessl S, and Bernard B. 1985. Cyanide 

intoxication among silver-reclaiming workers. JAMA 253:367-371. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Casadei E, Jansen P, Rodrigues A, Molin A, and Rosling H. 1984. Mantakassa: An epidemic of 

spastic paraparesis associated with chronic cyanide intoxication in a cassava staple area of 

Mozambique. 2. Nutritional factors and hydrocyanic acid content of cassava products. Bull. 

World Health Org. 62(3): 485-492. 

Chandra H, Gupta BN, Bhargave SH, Clerk SH, and Mahendra PN. 1980. Chronic cyanide 

exposure - a biochemical and industrial hygiene study. J. Anal. Toxicol. 4:161-165. 

Cliff J, Martelli A, Mondlane E, Molin A, and Rosling H. 1984. Mantakassa: An epidemic of 

spastic paraparesis associated with chronic cyanide intoxication in a cassava staple area of 

Mozambique. 1. Epidemiology and clinical and laboratory findings in patients. Bull. World 

Health Org. 62 (3): 477-484. 

El Ghawabi SH, Gaafar MA, El-Saharti AA, Ahmed SH, Malash KK, and Fares R. 1975. 

Chronic cyanide exposure: a clinical, radioisotope, and laboratory study. Br. J. Ind. Med. 32:215-

219. 

Fiskel J, Cooper C, and Eschenroeder A. 1981. Exposure and risk assessment for cyanide. 

EPA/440/4-85/008. NTIS PB85-220572. 

Frakes RA, Sharma RP, Willhite CC, and Gomez G. 1986. Effect of cyanogenic glycosides and 

protein content in cassava diets on hamster prenatal development. Fundam. Appl. Toxicol. 

7:191-198. 

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Hugod C. 1979. Effect of exposure to 0.5 ppm hydrogen cyanide singly or combined with 200 

ppm carbon monoxide and/or 5 ppm nitric oxide on coronary arteries, aorta, pulmonary artery, 

and lungs in the rabbit. Int. Arch. Occup. Environ. Health 44:13-23. 

Hugod C. 1981. Myocardial morphology in rabbits exposed to various gas-phase constituents of 

tobacco smoke. Artherosclerosis 40:181-190. 

Jackson LC. 1988. Behavioral effects of chronic sublethal dietary cyanide in an animal 

model: implications for humans consuming cassava (Manihot esculenta). Hum. Biol. 60(4):597-

614. 

Appendix D3 319 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Jansz ER and Uluwaduge DI. 1997. Biochemical aspects of cassava (Manihot esculenta Crantz) 

with special emphasis on cyanogenic glucosides – a review. J. Natn. Sci. Coun. Sri Lanka 

25(1):1-24. 

Kamalu BP. 1993. Pathological changes in growing dogs fed on a balanced cassava (Manihot 

esculenta Crantz) diet. Br. J. Nutr. 69(3):921-34. 

Kamalu BP. 1995. The adverse effects of long-term cassava (Manihot esculenta Crantz) 

consumption. Int. J. Food Sci. Nutr. 46(1):65-93. 

Leuschner J, Winkler A, Leuschner F. 1991. Toxicokinetic aspects of chronic cyanide exposure 

in the rat. Toxicol. Lett. 57(2):195-201. 

Miller RW. 1974. Susceptibility of the fetus and child to chemical pollutants. Science 184:812-

814. 

Obidoa O and Obasi SC. 1991. Coumarin compounds in cassava diets: 2. Health implications of 

scopoletin in gari. Plant Foods Hum. Nutr. 41(3):283-289. 

Okolie NP, Osagie AU. 1999. Liver and kidney lesions and associated enzyme changes induced 

in rabbits by chronic cyanide exposure. Food Chem. Toxicol. 37(7):745-750. 

Philbrick DJ, Hopkins JB, Hill DC, Alexander JC, Thomson RG. 1979. Effects of prolonged 

cyanide and thiocyanate feeding in rats. J. Toxicol. Environ. Health 5(4):579-592. 

Sharma RP. 1993. Cyanide containing foods and potential for fetal malformations. In: Dietary 

Factors and Birth Defects (RP Sharma, ed.), Pacific Division, AAAS, San Francisco, CA. 

Tewe OO, and Maner JH. 1981. Long-term and carry-over effect of dietary inorganic cyanide 

(KCN) in the life cycle performance and metabolism of rats. Toxicol. Appl. Pharmacol. 58:1-7. 

Tor-Agbidye J, Palmer VS, Lasarev MR, Craig AM, Blythe LL, Sabri MI, Spencer PS. 1999. 

Bioactivation of cyanide to cyanate in sulfur amino acid deficiency: Relevance to neurological 

disease in humans subsisting on cassava. Toxicol. Sci. 50(2):228-235. 

Tsuchiya Y, and Sumi, K. 1977. Thermal decomposition products of polyacrylonitrile. J. Appl. 

Polym. Sci. 21:975-980. 

Tylleskar T, Banea M, Bikangi N, Cooke RD, Poulter NH, and Rosling H. 1992. Cassava 

cyanogens and konzo, an upper motoneuron disease found in Africa. Lancet (N Am Ed). 

339(8787):208-211. 

U.S.EPA. 1999. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) Database. Reference concentration (RfC) for Hydrogen Cyanide. 

Valade MP. 1952. [Central nervous system lesions in chronic experimental poisoning with 

gaseous hydrocyanic acid.] Bull. Acad. Nat. Med. (Paris) 136:280-285. 

Appendix D3 320 Hydrogen Cyanide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

HYDROGEN SULFIDE 

(hydrogen sulphide; dihydrogen sulfide; dihydrogen monosulfide; 

sulfur hydride; sulfureted hydrogen; hydrosulfuric acid) 

CAS registry number: 7783-06-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 10 μg/m3

 ( 8 ppb)

Critical effect(s) Nasal histological changes in B6C3F1 mice 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1999) 

Description Colorless gas 

Molecular formula H2S 

Molecular weight 34.08 

Density 1.4 g/L @ 25° C (air = 1) (AIHA, 1991) 

Boiling point −60.7° C (CRC, 1994) 

Melting point −85.5oC (CRC, 1994) 

Vapor pressure 15,600 Torr @ 25oC 

Solubility Soluble in water, hydrocarbon solvents, ether, and 

ethanol 

Odor threshold 8.1 ppb (11 μg/m3

 ) (Amoore and Hautala, 1983) 

Odor description Resembles rotten eggs 

Conversion factor 1 ppm = 1.4 mg/m3

 @ 25° C 

III. Major Uses or Sources 

Hydrogen sulfide (H2S) is used as a reagent and an intermediate in the preparation of other 

reduced sulfur compounds (HSDB, 1999). It is also a by-product of desulfurization processes in 

the oil and gas industries and rayon production, sewage treatment, and leather tanning (Ammann, 

1986). The annual statewide industrial emissions from point sources at facilities reporting under 

the Air Toxics Hot Spots Act in California based on the most recent inventory were estimated to 

be 5,688,172 pounds of hydrogen sulfide (CARB, 1999). 

Appendix D3 321 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure

Although numerous case studies of acutely toxic effects of H2S exist, there is inadequate 

occupational or epidemiological information for specific chronic effects in humans exposed to 

H2S. 

Bhambhani and Singh (1991) showed that 16 healthy subjects exposed for short durations to 

5 ppm (7 mg/m3

) H2S under conditions of moderate exercise exhibited impaired lactate and 

oxygen uptake in the blood. Bhambhani and Singh (1985) reported that exposure of 42 

individuals to 2.5 to 5 ppm (3.5 to 7 mg/m3

) H2S caused coughing and throat irritation after 15 

minutes. 

In another study, ten asthmatic volunteer subjects were exposed to 2 ppm H2S for 30 minutes and 

pulmonary function was tested (Jappinen et al., 1990). All subjects reported detecting “very 

unpleasant” odor but “rapidly became accustomed to it.” Three subjects reported headache 

following exposure. No significant changes in mean FVC or FEV1 were reported. Although 

individual values for specific airway resistance (SRaw) were not reported, the difference 

following exposure ranged from −5.95% to +137.78%. The decrease in specific airway 

conductance, SGaw, ranged from -57.7% to +28.9%. The increase in mean SRaw and decrease in 

mean SGaw were not statistically significant. 

Kilburn and Warshaw (1995) investigated whether people exposed to sulfide gases, including 

H2S, as a result of working at or living downwind from the processing of "sour" crude oil 

demonstrated persistent neurobehavioral dysfunction. They studied thirteen former workers and 

22 neighbors (of a California coastal oil refinery) who complained of headaches, nausea, 

vomiting, depression, personality changes, nosebleeds, and breathing difficulties. Their 

neurobehavioral functions and a profile of mood states were compared to 32 controls (matched 

for age and educational level). The exposed subjects' mean values were statistically significantly 

different (abnormal) compared to controls for several tests (two-choice reaction time; balance (as 

speed of sway); color discrimination; digit symbol; trail-making A and B; immediate recall of a 

story). Their profile of mood states scores were much higher than those of controls. Visual 

recall was significantly impaired in neighbors, but not in the former workers. The authors 

concluded that neurophysiological abnormalities were associated with exposure to reduced sulfur 

gases, including H2S from crude oil desulfurization. 

Xu et al. (1998) conducted a retrospective epidemiological study in a large petrochemical 

complex in Beijing, China in order to assess the possible association between petrochemical 

exposure and spontaneous abortion. The facility consisted of 17 major production plants divided 

into separate workshops, which allow for the assessment of exposure to specific chemicals. 

Married women (n = 2853), who were 20-44 years of age, had never smoked, and who reported 

at least one pregnancy during employment at the plant, participated in the study. According to 

their employment record, about 57% of these workers reported occupational exposure to 

petrochemicals during the first trimester of their pregnancy. There was a significantly increased 

risk of spontaneous abortion for women working in all of the production plants with frequent 

exposure to petrochemicals compared with those working in nonchemical plants. Also, when a 

comparison was made between exposed and non-exposed groups within each plant, exposure to 

Appendix D3 322 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

petrochemicals was consistently associated with an increased risk of spontaneous abortion 

(overall odds ratio (OR) = 2.7 (95% confidence interval (CI) = 1.8 to 3.9) after adjusting for 

potential confounders). When the analysis was performed with the exposure information 

obtained from interview responses for (self reported) exposures, the estimated OR for 

spontaneous abortions was 2.9 (95% CI = 2.0 to 4.0). When the analysis was repeated by 

excluding those 452 women who provided inconsistent reports between recalled exposure and 

work history, a comparable risk of spontaneous abortion (OR 2.9; 95% CI = 2.0 to 4.4) was 

found. In analyses for exposure to specific chemicals, an increased risk of spontaneous abortion 

was found with exposure to most chemicals. There were 106 women (3.7% of the study 

population) exposed only to hydrogen sulfide, and the results for hydrogen sulphide (OR 2.3; 

95% CI = 1.2 to 4.4) were significant. No hydrogen sulfide exposure concentration was 

reported. 

Four workers were exposed for several minutes to concentrations of hydrogen sulfide sufficient 

to cause unconciousness. Four other workers were exposed chronically to H2S and developed 

lacrimation, eye irritation, nausea, vomiting, headache, sore throat, and skin irritation but 

retained conciousness as the result of a 150-minute release. Both groups were subjected to 

olfactory testing 2 to 3 years later (Hirsch and Zavala, 1999). Six of eight workers showed 

deficits in odor detection and identification, with the workers who had experienced 

unconciousness most severely affected in the followup tests. 

Three patients exposed acutely to unknown concentrations of hydrogen sulfide developed 

persistent cognitive impairment (Wasch et al,. 1989). While standard neurological and physical 

examinations were unremarkable, all three subjects had prolonged P-300 latencies and persistent 

neurological and neurobehavioral deficits. 

V. Effects of Animal Exposure 

Rats (Fischer and Sprague-Dawley, 15 per group) were exposed to 0, 10.1, 30.5, or 80 ppm (0, 

14.1, 42.7, or 112 mg/m3

, respectively) H2S for 6 hours/day, 5 days/week for 90 days (CIIT, 

1983a,b). Measurements of neurological and hematological function revealed no abnormalities 

due to H2S exposure. A histological examination of the nasal turbinates also revealed no 

significant exposure-related changes. A significant decrease in body weight was observed in 

both strains of rats exposed to 80 ppm (112 mg/m3

). 

In a companion study, the Chemical Industry Institute of Toxicology conducted a 90-day 

inhalation study in mice (10 or 12 mice per group) exposed to 0, 10.1, 30.5, or 80 ppm (0, 14.1, 

42.7, or 112 mg/m3

, respectively) H2S for 6 hours/day, 5 days/week (CIIT, 1983c). Neurological 

function was measured by tests for posture, gait, facial muscle tone, and reflexes. 

Ophthalmological and hematological examinations were also performed, and a detailed necropsy 

was included at the end of the experiment. The only exposure-related histological lesion was 

inflammation of the nasal mucosa of the anterior segment of the noses of mice exposed to 80 

ppm (112 mg/m3) H2S. Weight loss was also observed in the mice exposed to 80 ppm. 

Neurological and hematological tests revealed no abnormalities. The 30.5 ppm (42.5 mg/m3

) 

level was considered the NOAEL for histological changes in the nasal mucosa. (Adjustments 

were made by U. S. EPA to this value to calculate an RfC of 0.9 μg/m3

.) 

Appendix D3 323 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Fischer F344 rats inhaled 0, 1, 10, or 100 ppm hydrogen sulfide for 8 hours/day for 5 weeks 

(Hulbert et al, 1989). No effects were noted on baseline measurements of airway resistance, 

dynamic compliance, tidal volume, minute volume, or heart rate. Two findings were noted more 

frequently in exposed rats: (1) proliferation of ciliated cells in the tracheal and bronchiolar 

epithelium, and (2) lymphocyte infiltration of the bronchial submucosa. Some exposed animals 

responded similarly to controls to aerosol methacholine challenge, whereas a subgroup of 

exposed rats were hyperreactive to concentrations as low as 1 ppm. 

Male rats were exposed to 0, 10, 200, or 400 ppm H2S for 4 hours (Lopez et al., 1987). Samples 

of bronchoalveolar and nasal lavage fluid contained increased inflammatory cells, protein, and 

lactate dehydrogenase in rats treated with 400 ppm. Lopez and associates later showed that 

exposure to 83 ppm (116 mg/m3) for 4 hours resulted in mild perivascular edema (Lopez et al., 

1988). 

A study by Saillenfait et al. (1989) investigated the developmental toxicity of H2S in rats. Rats 

were exposed 6 hours/day on days 6 through 20 of gestation to 100 ppm hydrogen sulfide. No 

maternal toxicity or developmental defects were observed.. 

Hayden et al. (1990) exposed gravid Sprague-Dawley rat dams continuously to 0, 20, 50, and 75 

ppm H2S from day 6 of gestation until day 21 postpartum. The animals demonstrated normal 

reproductive parameters until parturition when delivery time was extended in a dose dependent 

manner (with a maximum increase of 42% at 75 ppm). Pups which were exposed in utero and 

neonatally to day 21 postpartum developed with a subtle decrease in time of ear detachment and 

hair development and with no other observed change in growth and development through day 21 

postpartum. 

Appendix D3 324 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic REL 

Study CIIT, 1983c 

Study population B6C3F1 mice (10-12 per group) 

Exposure method Discontinuous inhalation 

Critical effects Histopathological inflammatory changes in the 

nasal mucosa 

LOAEL 80 ppm (112 mg/m3

) 

NOAEL 30.5 ppm (42.5 mg/m3

) 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 90 days 

Average experimental exposure 5.4 ppm for NOAEL group (30.5 x 6/24 x 5/7) 

Human equivalent concentration 0.85 ppm (gas with extrathoracic respiratory 

effects, RGDR = 0.16, based on mouse 

MVa = 0.033 L/min; MVh = 13.8 L/min; 

SAa(ET) = 3.0 cm2

; Sah(ET) = 200 cm3) (U.S. 

EPA, 1994) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 8 ppb (10 μg/m3

) 

The adverse effects reported in chronic animal studies occur at higher concentrations than effects 

seen in acute human exposures. For example, human irritation was reported at concentrations of 

2.5-5 ppm for 15 minutes (Bhambhani and Singh, 1985), yet no effects on laboratory animals 

were observed at concentrations up to 80 ppm for 90 days. This suggests either that humans are 

more sensitive to H2S, or that the measurements in laboratory animals are too crude to detect 

subtle measures of irritation. However, the uncertainty factor and HEC attempt to account for 

these interspecies differences. 

VII. Data Strengths and Limitations for Development of the REL 

Hydrogen sulfide is the leading chemical agent causing human fatalities following inhalation 

exposures. Although lower concentration acute exposures have been quantitatively studied with 

human volunteers, the dose-response relationship for human toxicity due to hydrogen sulfide 

exposure is not known. Thus, a major area of uncertainty is the lack of adequate long-term 

human exposure data. Subchronic (but not chronic) studies have been conducted with several 

animal species and strains, and these studies offer an adequate basis for quantitative risk 

assessment. 

The strengths of the inhalation REL include the availability of controlled exposure inhalation 

studies in multiple species at multiple exposure concentrations, adequate histopathogical 

analysis, and the observation of a NOAEL. 

Appendix D3 325 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Hydrogen sulfide has a strong unpleasant odor. The threshold for detection of this odor is low, 

but shows wide variation among individuals. A level of 7 μg/m3

, based on a 30 minute 

averaging time, was estimated by a Task Force of the International Programme on Chemical 

Safety (IPCS) (1981) to not produce odor nuisance in most situations. On the other hand, the 

current California Ambient Air Quality standard for hydrogen sulfide, based on a 1 hour 

averaging time, is 42 μg/m3

 (30 ppb). 

Amoore (1985) analyzed a large number of reports from the scientific literature and found that 

reported thresholds for detection were log-normally distributed, with a geometric mean of 10 

μg/m3 (8 ppb). Detection thresholds for individuals were reported to be log-normally distributed 

in the general population, with a geometric standard deviation of 4.0, i.e. 68% of the general 

population would be expected to have a detection threshold for hydrogen sulfide between 2.5 and 

40 μg/m3 (2 and 32 ppb). Sources of variation included age, sex, medical conditions, and 

smoking. Training and alertness of the subject in performing the test also affected the results. 

Amoore (1985) drew attention to the difference between a detection threshold under laboratory 

conditions, and the levels at which an odor could be recognized, or at which it was perceived as 

annoying. Analysis of various laboratory and sociological studies suggested that a level at which 

an odor could be recognized was typically a factor of three greater than the threshold for 

detection, while the level at which it was perceived as annoying was typically a factor of five 

greater than the threshold. Annoyance was characterized both in terms of esthetic or behavioral 

responses, and by physiological responses such as nausea and headache. He therefore predicted 

that, although at 10 μg/m3 (the proposed REL) 50% of the general population would be able to 

detect the odor of hydrogen sulfide under controlled conditions, only 5% would find it annoying 

at this level. At 50 μg/m3, 50% would find the odor annoying. 

On this basis, the proposed REL of 10 μg/m3 (8 ppb) is likely to be detectable by many people 

under ideal laboratory conditions, but it is unlikely to be recognized or found annoying by more 

than a few. It is therefore expected to provide reasonable protection from odor annoyance in 

practice. However, this consideration cannot be entirely dismissed due to the wide interindividual variation in sensitivity to odors. Amoore (1985) also points out that many industrial 

operations generating hydrogen sulfide also generate organic thiol compounds with similar, but 

even more potent odors (e.g., methyl mercaptan, butyl mercaptan). Such compounds may in fact 

have detection thresholds as much as a hundred-fold lower than hydrogen sulfide, so even 

minute quantities have a powerful impact on odor perception. Because of the concurrent 

emission of these contaminants, the incidence of odor complaints near hydrogen sulfide emitting 

sites correlated poorly with the levels of hydrogen sulfide measured in the affected areas. 

Appendix D3 326 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VIII. References 

AIHA. 1991. American Industrial Hygiene Association. Emergency Response Planning 

Guideline for Hydrogen Sulfide. Set 6. Akron, HI: AIHA. 

Ammann HM. 1986. A new look at physiologic respiratory response to H2S poisoning. J. Haz. 

Mat. 13 :369-374. 

Amoore JE (1985). The perception of hydrogen sulfide odor in relation to setting an ambient 

standard. Olfacto-Labs, Berkeley, CA: prepared for the California Air Resources board. 

 

Amoore JE, and Hautala E. 1983. Odor as an aid to chemical safety: Odor thresholds compared 

with threshold limit values and volatilities for 214 chemicals in air and water dilution. J. Appl. 

Toxicol. 3(6):272-290. 

Bhambhani Y, and Singh M. 1985. Effects of hydrogen sulphide on selected metabolic and 

cardio-respiratory variables during rest and exercise. Report submitted to Alberta Worker’s 

Health and Safety and Compensation. June, 1985. 

Bhambhani Y, and Singh M. 1991. Physiological effects of hydrogen sulfide inhalation during 

exercise in healthy men. J. Appl. Physiol. 71:1872-1877. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CIIT. 1983a. Chemical Industry Institute of Toxicology. 90-Day vapor inhalation toxicity study 

of hydrogen sulfide in Fischer-344 rats. U.S. EPA, Office of Toxic Substances Public Files. 

Fiche number 0000255-0. Document number FYI-OTS-0883-0255. 

CIIT. 1983b. Chemical Industry Institute of Toxicology. 90-Day vapor inhalation toxicity study 

of hydrogen sulfide in Sprague-Dawley rats. U.S. EPA, Office of Toxic Substances Public Files. 

Fiche number 0000255-0. Document number FYI-OTS-0883-0255. 

CIIT. 1983c. Chemical Industry Institute of Toxicology. 90-Day vapor inhalation toxicity study 

of hydrogen sulfide in B6C3F1 mice. U.S. EPA, Office of Toxic Substances Public Files. Fiche 

number 0000255-0. Document number FYI-OTS-0883-0255. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Hannah RS, Hayden LJ, and Roth SH. 1989. Hydrogen sulfide exposure alters the amino acid 

content in developing rat CNS. Neurosci. Lett. 99(3):323-327. 

Hannah RS, and Roth SH. 1991. Chronic exposure to low concentrations of hydrogen sulfide 

produces abnormal growth in developing cerebral Purkinje cells. Neurosci. Lett. 122(2):225-228. 

Appendix D3 327 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Hayden LJ, Goeden H, and Roth SH. 1990. Growth and development in the rat during 

subchronic exposure to low levels of hydrogen sulfide. Toxicol. Ind. Health 6(3-4):389-

401.Hirsch AR, Zavala G. 1999. Long-term effects on the olfactory system of exposure to 

hydrogen sulphide. Occup. Environ. Med. 56(4):284-287. 

HSDB. 1999. Hazardous Substances Data Bank. U.S. National Library of Medicine, Bethesda, 

MD 20894. (http://sis.nlm.nih.gov/sis1/) 

Hulbert, W.C., M.G. Prior, P. Pieroni and Z. Florence. 1989. Hyperresponsiveness in rats after 5 

weeks exposure to hydrogen sulfide. Clin. Invest. Med. 12(4): B89. 

International Programme on Chemical Safety. 1981. Environmental Health Criteria 19. 

Hydrogen Sulfide. Geneva: World Health Organization. p. 41. 

Jappinen P, Vilkka V, Marttila O, Haahtela T. 1990. Exposure to hydrogen sulphide and 

respiratory function. Br. J. Ind. Med. 47(12):824-828. 

Kilburn KH, Warshaw RH. 1995. Hydrogen sulfide and reduced-sulfur gases adversely affect 

neurophysiological functions. Toxicol. Ind. Health 11:185-197. 

Lopez A, Prior M, Yong S, Albassam M, and Lillie L. 1987. Biochemical and cytological 

alterations in the respiratory tract of rats exposed for 4 hours to hydrogen sulfide. Fundam. Appl. 

Toxicol. 9:753-762. 

Lopez A, Prior M, Lillie L, Gulayets C, and Atwal O. 1988. Histologic and ultrastructural 

alterations in lungs of rats exposed to sublethal to lethal concentrations of hydrogen sulfide. Vet. 

Pathol. 25:376-384. 

Saillenfait A, Bonnet P, and DeCeaurriz J. 1989. Effects of inhalation exposure to carbon 

disulfide and its combination with hydrogen sulfide on embryonal and fetal development in rats. 

Toxicol. Lett. 48:57-66. 

U.S.EPA. 1999. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration (RfC) for hydrogen sulfide. 

Wasch HH, Estrin WJ, Yip P, Bowler R, Cone JE. 1989. Prolongation of the P-300 latency 

associated with hydrogen sulfide exposure. Arch. Neurol. 46(8):902-904. 

Xu X, Cho SI, Sammel M, You L, Cui S, Huang Y, et al. Association of petrochemical exposure 

with spontaneous abortion. Occup. Environ. Med. 1998;55(1):31-36. 

Appendix D3 328 Hydrogen Sulfide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY 

ISOPHORONE 

(1,1,3-trimethyl-3-cyclohexene-5-one; 3,5,5-trimethyl-2-cyclohexen-1-one; isoforon; 

isoacetophorone) 

CAS Registry Number: 78-59-1

I. Chronic Toxicity Summary

Inhalation reference exposure level 2,000 μg/m3 (400 ppb)

Critical effect(s) Developmental effects (reduced crown-rump 

length of female rat fetuses); 

hepatocytomegaly and coagulative necrosis 

of the liver in mice 

Hazard index target(s) Development; liver 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994; CARB, 1997) 

Description Water-clear liquid with a peppermint-like odor 

Molecular formula C9H14O 

Molecular weight 138.21 g/mol 

Boiling point 215.2°C 

Melting point −8.1°C 

Vapor pressure 0.44 torr at 25°C 

Solubility Slightly soluble in water (12,000 mg/L water at 

25°C); miscible in organic solvents. 

Conversion factor 5.65 μg/m3

 per ppb at 25°C 

III. Major Uses and Sources

Isophorone is used extensively as a solvent in some printing inks, paints, lacquers, adhesives, 

vinyl resins, copolymers, coatings, finishes, and pesticides, in addition to being used as a 

chemical intermediate (HSDB, 1995). Since this compound has many different applications, 

release to the environment may originate from a wide variety of industrial sources including iron 

and steel manufacturers, manufacturers of photographic equipment and supplies, automobile tire 

plants, and printing operations. Coal-fired power plants may also emit isophorone to the air. 

Although it is mostly a man-made compound, isophorone has been found to occur naturally in 

cranberries (ATSDR, 1989). Occupational exposure may occur by inhalation or dermal contact. 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 2809 pounds of 

isophorone (CARB, 2000). 

Appendix D3 329 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

IV. Effects of Human Exposures

No information is available concerning long-term exposure or pharmacokinetics of isophorone in 

humans. In occupational monitoring studies, the time-weighted average concentration in 

breathing zones and workplace air of a screening plant ranged from 8.3-23 ppm and from 3.5-

14.5 ppm, respectively (Samimi, 1982). Up to 25.7 ppm was detected in air of a silk screening 

printing plant in Pittsburgh, PA (Kominsky, 1983). The concentration in breathing zone samples 

from a decal manufacturing plant in Ridgefield, NJ was 0.7-14 ppm (Lee and Frederick, 1982). 

It was suspected that the reported eye and nose irritation of workers at the silk screening plant 

and at the decal manufacturing plant was the result of acute and subacute exposure to isophorone 

vapors. 

Workers exposed to 5-8 ppm (28-45 mg/m3

) of isophorone for one month complained of fatigue 

and malaise (NIOSH, 1978). When concentrations were reduced to 1-4 ppm, no adverse effects 

were reported. Acute exposure studies in humans (up to 400 ppm for 1 to 4 minutes) resulted in 

eye, nose and throat irritation, nausea, headache, and dizziness or faintness (Union Carbide, 

1963). Inhalation exposure for 15 minutes to 10 ppm isophorone produced only mild effects in 

human subjects while 25 ppm produced irritation to eyes, nose, and throat (Silverman et al., 

1946). 

V. Effects of Animal Exposures

Few reports have been published regarding the pharmacokinetics of isophorone in experimental 

animals. Isophorone was widely distributed in the major organs of the rat following 4 hour 

inhalation exposure to 400 ppm (ATSDR, 1989). Oral gavage of 4000 mg/kg body wt to rats 

and a rabbit also resulted in wide distribution of the chemical. The highest blood levels of 

isophorone were reached by 30 min in rabbits following oral gavage and had decreased 

dramatically by 21 hours, indicating rapid absorption and elimination of the chemical. 

Preliminary results of a pharmacokinetic study indicate that rats treated orally with 14Cisophorone excreted 93% of the radiolabel in the urine, expired air, and feces in 24 hours 

(ATSDR, 1989). The highest levels of 14C-isophorone were found in the liver, kidney, preputial 

gland, testes, brain, and lungs. Several metabolites were identified in the urine of orally dosed 

rats and rabbits, including 3-carboxy-5,5-dimethyl-2-cyclohexene-1-one, 3,5,5-

trimethylcyclohexanol, and some glucuronide conjugates (Dutertre-Catella et al., 1978). A 

portion of the chemical was excreted unchanged in expired air. 

In an early inhalation study, 10 Wistar rats/group and 10 guinea pigs/group, all of mixed sex, 

were exposed to 0, 25, 50, 100, 200, or 500 ppm isophorone 8 hr/day, 5 days/week for 6 weeks 

(Smyth et al., 1942). Increased mortality and reduced body weights were observed at 100 ppm 

and up in both species. However, eye and nose irritation was noted only at the highest dose. 

Minor changes in blood chemistry and histopathological changes in the kidney and lungs were 

noted in treated animals. However, later investigations determined that the isophorone used in 

this study was contaminated with appreciable amounts of compounds (Rowe and Wolf, 1963). 

Appendix D3 330 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Therefore, some of the adverse effects (i.e., the lung lesions) may have been due to the 

contaminants. The accuracy of the concentration data in the 1942 study is also questionable. 

No treatment-related histopathological lesions were found in lungs, livers, or kidneys of male 

and female rats exposed intermittently (6 hr/day, 5 days/week) to 37 ppm isophorone for 4 weeks 

compared to controls (Hazleton Labs, 1968; summarized by ATSDR, 1989). Histological 

examination was limited to 30% of the control and treated rats. Body weight gain, mean 

absolute liver weights, and mean liver-to-body weight ratios of treated rats were significantly 

reduced compared to controls. Slight variations in hematological findings were noted in treated 

rats (increased lymphocytes and hemoglobin content; decreased neutrophils) but were not 

considered different from controls. 

Rats (10/sex) were exposed to 500 ppm isophorone 6 hr/day, 5 days/week for up to 6 months 

(Dutertre-Catella, 1976; summarized by ATSDR, 1989). Irritation of eyes and nasal mucosa was 

observed. One female and three males in the treatment group died during the study, which was 

considered to be a treatment-related effect. But no exposure-related histopathological lung or 

liver lesions were observed compared to controls. Dutertre-Catella (1976) also exposed rats and 

rabbits (number per group per sex not stated) to 250 ppm isophorone 6 hr/day, 5 days/week for 

18 months (Dutertre-Catella, 1976). Irritation of eyes and nasal mucosa was observed in both 

species, but no deaths occurred in the treatment groups. Histopathological examination of the 

lungs and kidneys, urinalysis, and hematological analysis revealed no exposure-related changes 

in either species. However, cytoplasmic microvacuolization of hepatocytes was observed in both 

species (ATSDR, 1989). 

In a 90-day feeding study, 20 CFE albino rats/group/sex were given isophorone in their diet at 

concentrations of 0, 750, 1500, or 3000 ppm. Four beagle dogs/group/sex received isophorone in 

gelatin capsules at concentrations of 0, 35, 75, or 150 mg/kg body wt-day (AME, 1972a,b). High 

dose rats exhibited slightly reduced weight gain compared to controls (8-10%) for most of the 

study. Average weight gain among the exposure groups of beagle dogs remained essentially 

unchanged during the entire study. Urinalysis, hematology, and clinical chemistry indices found 

no treatment-related effects in the animals at either the interim or final toxicological 

examinations. Gross pathology and a limited histopathological examination observed no 

treatment-related effects in either species. Data on isophorone purity and possible loss of 

isophorone from rat diet due to vaporization were not presented. 

In the most comprehensive isophorone toxicity study to date, 50 F344/N rats/group/sex and 50 

B6C3F1 mice/group/sex were administered 0, 250 or 500 mg isophorone/kg body wt 5 

days/week by oral gavage (in corn oil) for 103 weeks (Bucher et al., 1986; NTP, 1986). Clinical 

signs of toxicity were not seen during the length of the study. However, several deaths in male 

and female rats at the high dose occurred early in the study. A steep decline in survival rate of 

high dose male rats occurred after week 90. Male and female rats and female mice in the high 

dose group exhibited only a slight decrease in body weight (<10%) compared to controls. A 13-

week range finding study for the 2-year study did not find compound-related lesions in the 

kidney (or any other organs) of rats and mice exposed up to 1000 mg/kg body wt-day. However, 

pathological examination of rats exposed to isophorone for 2 years revealed non-neoplastic 

lesions in the kidney. Increased mineralization of the collecting ducts in isophorone-exposed 

Appendix D3 331 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

male (but not female) rats was observed. This lesion was characterized by basophilic aggregates 

of mineral most often found in the medullary collecting ducts and occurred coincidentally with 

lesions of chronic nephropathy. Nephropathy was observed in almost half the female controls 

and nearly all the male controls. Isophorone exposure appeared to increase both the severity of 

nephropathy in low dose male rats and the incidence of nephropathy in dosed female rats, but the 

effects were not pronounced. However, the isophorone potentiation of nephropathy in rats may 

be due to ‘male rat-specific nephropathy’ and may not have any relevance to human exposure 

(Strasser et al. 1988). Other adverse effects in kidneys of isophorone-treated male rats include 

tubular cell hyperplasia (in a dose-related manner) and epithelial hyperplasia of the renal pelvis. 

In mice, an increased incidence of chronic focal inflammation was observed in the kidneys of 

males, but was not considered treatment-related. A dose-dependent increase in fatty 

metamorphosis occurred in the adrenal cortex of male rats, but the biological significance of this 

change is unknown. All isophorone-exposed male mice had an increased incidence of 

hepatocytomegaly and coagulative necrosis of the liver. However, treatment-related liver lesions 

were not observed in female mice. Increased incidence of hyperkeratosis of the forestomach was 

observed in dosed male and high dose female mice, but was probably not a relevant treatmentrelated effect. 

Published studies on possible reproductive effects of isophorone are lacking. An unpublished 

inhalation study conducted by a commercial laboratory (Bio/dynamics, 1984b) studied possible 

teratogenicity due to isophorone in rats or mice at inhaled doses up to 115 ppm. Groups of 22 

female rats and 22 female mice were exposed to 0, 25, 50, or 115 ppm isophorone (6 hr/day) on 

gestational days 6-15. Maternal toxicity in rats included dose-dependent alopecia and 

cervical/anogenital staining. Low body weights (7-8%) were occasionally observed in the 115 

ppm group. In mice, maternal toxicity was confined to slightly decreased weight (7-8%) on one 

day in the 115 ppm group. No significant differences were found in uterine implantations, fetal 

toxicity, and external and internal malformations among the animals. However, a slight, but 

significant, growth retardation in the form of decreased crown-rump length was present among 

the high dose fetal rats. Also, a slight, but insignificant, increase in extra ribs and/or rudimentary 

ribs was seen in rat and mouse fetuses at the highest dose. In a pilot study for this developmental 

toxicity investigation (12 females/species), exencephaly was observed in 1 rat and 1 mouse 

undergoing late reabsorption and in 2 live rat fetuses from dams exposed to 150 ppm isophorone 

on gestational days 6-15 (Bio/dynamics, 1984a). Exencephaly was not observed at any dose 

level in the primary study. 

Dutertre-Catella (1976) did not find adverse reproductive or developmental effects in rats 

exposed to 500 ppm isophorone (6 hr/day, 5 days/week) for 3 months before mating and 

throughout gestation (females only) as well. The pups were not examined for internal 

malformations so the study was incomplete for determination of developmental effects. 

Appendix D3 332 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

VI. Derivation of Chronic Reference Exposure Level (REL)

Study Bio/dynamics 1984a,b 

Study population 22 female mice/group, 22 female rats/group 

Exposure method Discontinuous whole body inhalation exposure 

during gestation (0, 25, 50, or 115 ppm) 

Critical effects Developmental effects (reduced crown-rump 

length of female rat fetuses); teratogenicity 

(exencephaly in fetal rats and mice) in range 

finding study at 150 ppm 

LOAEL 115 ppm for reduced crown-rump length of 

female rat fetuses 

NOAEL 50 ppm 

Exposure continuity 6 hr/day during gestation

Exposure duration Days 6-15 of gestation 

Average experimental exposure 12.5 ppm (50 x 6/24) 

Human equivalent concentration 12.5 ppm (gas with systemic effects, based on 

RGDR = 1.0 using default assumption that 

lambda (a) = lambda (h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.4 ppm (400 ppb, 2 mg/m3

, 2,000 μg/m3) 

The inhalation study by Bio/dynamics (1984a,b) presents data that indicate exposure during 

gestation may be the most sensitive indicator of non-neoplastic toxicity by isophorone. Exposure 

of pregnant rats to 115 ppm isophorone during gestation resulted in significant growth 

retardation of female rat fetuses (reduced crown-rump length). Exposure to 50 ppm isophorone, 

the NOAEL, produced no developmental effects. The authors had removed the two shortest 

female fetuses prior to statistical analysis. The result was that there was no significant difference 

in fetal growth retardation; therefore, this adverse effect is not significant. However, this 

selective culling before the statistical analysis is not scientifically appropriate in this case. In 

addition, the authors did not perform some of the scheduled fetal examinations. Otherwise, the 

growth retardation might have had even greater statistical significance. The pilot study 

(Bio/dynamics, 1984a) observed exencephaly in a few mouse and rat fetuses at 150 ppm. 

Exencephaly was not considered significant by the authors because it was not present in any 

fetuses of the primary study (Bio/dynamics, 1984b). However, exencephaly is included as a 

critical effect in this summary because it is considered a serious teratogenic effect that was 

present at a dose (150 ppm) only slightly higher than the LOAEL of the primary study (115 

ppm). Alopecia of adult female rats was observed in many of the exposed animals. However, 

this effect may be considered more of an acute dermal irritation than a chronic effect. In 

addition, cervical and anogenital staining seen in many exposed rats is not considered a chronic 

‘adverse’ effect. 

Appendix D3 333 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

For comparison with the proposed REL of 0.4 ppm, the inhalation LOAEL of 250 ppm for mild 

liver effects (Dutertre-Catella, 1976) in rats and rabbits intermittently exposed to isophorone for 

18 months was used to estimate a REL. Use of a time adjustment (6/24 x 5/7), an RGDR of 1, 

and a total UF of 100 (LOAEL to NOAEL = 3, interspecies = 3, and intraspecies = 10), also 

resulted in an estimated REL of 0.4 ppm. These results indicate that the REL will also protect 

against adverse liver effects. 

While the toxicological significance of this liver effect observed by Dutertre-Catella (1976) is 

unknown, the NTP (1986) study observed an increased incidence of hepatocytomegaly and 

coagulative necrosis of the liver in treated male mice, but not in female mice and rats, orally 

gavaged with isophorone. Using 250 mg/kg-day as a LOAEL for mice and dividing by a total 

UF of 1000 (10 each for LOAEL to NOAEL, 10 for interspecies, and 10 for intraspecies) results 

in an oral REL of 0.25 mg/kg-day. Multiplying the oral REL by 3,500 μg/m3

 per mg/kg-day for 

route-to-route extrapolation results in a chronic inhalation REL estimate of 900 μg/m3 (0.16 

ppm), which is in good agreement with the REL developed from Dutertre-Catella (1976) and 

Biodynamics (1984a,b). 

VII. Data Strengths and Limitations for Development of the REL 

The strength of the database for isophorone is the consistent lack of relevant severe 

histopathological effects in the chronic inhalation study (Dutertre-Catella, 1976) and in the oral 

gavage study (NTP, 1986). Weaknesses of the database for isophorone include the lack of 

human exposure data, the lack of comprehensive long-term inhalation studies, and the lack of 

published peer-reviewed reproductive/developmental studies. The lack of human data may be 

due to isophorone’s rather low potency for causing chronic, non-neoplastic, adverse effects. 

Inhalation of isophorone is a relevant route of exposure under occupational settings, but is most 

likely a minor route of exposure for the general population. Due to the insufficient 

characterization of the kidney and liver lesions in the oral gavage NTP study (Bucher et al, 1986; 

NTP, 1986) and the inhalation study (Dutertre-Catella, 1976), a comprehensive chronic study in 

rodent and non-rodent species would enhance the database for isophorone. 

VIII. Potential for Differential Impacts on Children's Health

Since the REL is based on a developmental study, it is expected to be adequately protective of 

infants and children. However, there is no direct evidence in the literature to quantify a 

differential effect of isophorone in children relative to adults. Isophorone occurs in cranberries 

and thus presumably in cranberry juice, which is often mixed with other fruit juices. Children 

tend to consume more fruit juice. However, isophorone as a Hot Spot emission is unlikely to be 

a multimedia chemical, and there is no evidence to suggest that normal dietary levels of 

isophorone are associated with adverse health effects. 

Appendix D3 334 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

IX. References 

AME. 1972a. Affiliated Medical Enterprises, Inc. 90-Day subchronic toxicity of isophorone in 

the rat (final report). Unpublished study performed by Affiliated Medical Enterprises, Inc. 

Princeton, NJ for Rohm and Haas Co. Philadelphia, PA. OTS 8d submission Doc. ID. 87812179, 

Microfiche No. 205975. 

AME. 1972b. Affiliated Medical Enterprises, Inc. 90-Day subchronic toxicity of isophorone in 

the dog (final report). Unpublished study performed by Affiliated Medical Enterprises, Inc. 

Princeton, NJ for Rohm and Haas Co. Philadelphia, PA. OTS 8d submission Doc. ID. 87812178, 

Microfiche No. 205975. 

ATSDR. 1989. Agency for Toxic Substances and Disease Registry. Toxicological profile for 

isophorone. U.S. Public Health Service. Atlanta, GA: ATSDR. PB90-180225. 

Bio/dynamics. 1984a. Inhalation teratology probe study in rats and mice. Project No. 323771. 

Unpublished study performed by Bio/dynamics Inc. East Millstone, NJ. OTS Section 4 

submission Doc. ID 40-8455042. Microfiche No. OTS0507219, pp. 1-33. 

Bio/dynamics. 1984b. Inhalation teratology study in rats and mice. Final Report 3223772. 

Unpublished study performed by Bio/dynamics Inc. East Millstone, NJ for Exxon Biomedical 

Science, East Millstone NJ. OTS Section 4 submission Doc. ID 40-855049. Microfiche No. OTS 

0507224, pp. 1-107. 

Bucher JR, Huff J, and Kluwe WM. 1986. Toxicological and carcinogenesis studies of 

isophorone in F344 rats and B6C3F1 mice. Toxicology 39:207-219. 

CARB. 1997. California Air Resources Board. Toxic Air Contaminant Identification List 

Summaries. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dutertre-Catella H. 1976. Contribution to the analytical toxicological and bio-chemical study of 

isophorone (in French). Thesis for doctorate in pharmacology, Universite Rene Descartes, Paris. 

[Cited in Joint Assessment of Commodity Chemicals, No. 110, Isophorone, ECETOC, Brussels, 

1989.] 

Dutertre-Catella H, Nguyen PL, Dang Quoc Q, and Truhaut R. 1978. Metabolic transformations 

of the 3,5,5-2-cyclohexene-1-one trimethyl (isophorone). Toxicol. Eur. Res. 1(4):209-216. 

Appendix D3 335 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels December 2001

Hazleton Labs. 1968. Assessment and comparison of subacute inhalation toxicities of three 

ketones. Final Report. Prepared by Hazleton Laboratories, Inc. Falls Church, VA for Exxon 

Chem Amers. Houston, TX. OTS 8d submission Doc ID. 878210935, Microfiche No. 206267. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/95). 

Kominsky JR. 1983. Health hazard determination report no. HE 78-107-563. Pittsburgh, PA: 

Swinston Company. 

Lee SA, and Frederick L. 1982. NIOSH health hazard evaluation report no. HHE80-103-827; 

NTIS PB82-189226. 

NIOSH. 1978. National Institute for Occupational Safety and Health. Occupational exposure to 

ketones: criteria for a recommended standard. U.S. Department of Health, Education, and 

Welfare. DHEW (NIOSH) Publication No. 78-173. 

NTP. 1986. National Toxicology Program. Toxicology and carcinogenesis studies of isophorone 

in F/344 rats and B6C3F1 mice. NTP TR 291. NIH Publication No. 86-2547. 

Rowe VK, and Wolf MA. 1963. Ketones. In: Industrial Hygiene and Toxicology, Second ed. 

Patty FA (ed.) New York: Interscience Publishers. p. 1764. 

Samimi B. 1982. Exposure to isophorone and other organic solvents in a screen printing plant. 

Am. Ind. Hyg. Assoc. J. 43(1):43-48. 

Silverman L, Schulte HF, and First MW. 1946. Further studies on sensory response to certain 

industrial solvent vapors. J. Ind. Hyg. Toxicol. 28(6):262-266. 

Strasser J Jr, Charbonneau M, Borgoff SJ, Turner MJ, and Swenberg JA. 1988. Renal protein 

droplet formation in male Fischer 344 rats after isophorone (IPH) treatment. Toxicologist 

8(1):136 (abstract). 

Smyth HF Jr, Seaton J, and Fischer L. 1942. Response of guinea pigs and rats to repeated 

inhalation of vapors of mesityl oxide and isophorone. J. Ind. Hyg. Toxicol. 24(3):46-50. 

Union Carbide Corporation. 1963. Toxicology Studies - Isophorone Summary Data Sheet. 

Industrial Medical Toxicology Dept. New York: Union Carbide. 

Appendix D3 336 Isophorone 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

ISOPROPANOL 

(2-propanol; dimethylcarbinol; isopropyl alcohol) 

CAS Registry Number: 67-63-0 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 7,000 μg/m3 (3000 ppb) 

Critical effect(s) Kidney lesions in mice and rats; fetal growth 

retardation and developmental anomalies in 

rats 

Hazard index target(s) Kidney; development 

II. Chemical Property Summary (HSDB, 1995)

Description Colorless liquid at room temperature (25°C) 

with a pleasant odor. Slightly bitter taste. 

Molecular formula C3H8O 

Molecular Weight 60.09 

Boiling point 82.5°C 

Vapor Pressure 44.0 torr at 25°C 

Solubility Miscible in water and most organic solvents; 

insoluble in salt solutions. 

Conversion factor 1 ppb = 2.45 μg/m3

 at 25°C 

III. Major Uses and Sources

Isopropanol is used as a solvent and in making many commercial products (HSDB, 1995). The 

annual production volume of isopropanol has been in excess of one billion pounds since 1956; it 

was ranked 50th among chemicals produced in the U.S. in 1994 (C&EN, 1995). Rubbing alcohol 

is a solution of 70% isopropanol in water. Specific uses and sources include: a component of 

antifreeze; a solvent for gums, shellac, essential oils, creosote and resins; extraction of alkaloids; 

component of quick drying oils and inks; component of denaturing alcohol; antiseptic for hand 

lotions; rubefacient; component of household products (after-shave lotions, cosmetics, etc.); the 

manufacture of acetone; deicing agent for liquid fuels; dehydrating agent and synthetic flavoring 

adjuvant. Isopropanol can enter the environment as emissions from its manufacture and use as a 

solvent. It naturally occurs as a plant volatile and is released during the microbial degradation of 

animal wastes. Human exposure will be both in occupational atmospheres and from use of 

consumer products containing isopropanol as a volatile solvent. An odor threshold has been 

estimated as 22 ppm (Amoore and Hautala, 1983), which is 7-fold higher than the chronic REL. 

Appendix D3 337 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot 

Spots Act in California based on the most recent inventory were estimated to be 525,826 pounds 

of isopropanol (CARB, 1999b). 

IV. Effects of Human Exposures 

Currently, there are no adequate chronic exposure data for isopropanol in humans. While 

isopropanol is not considered a dermal irritant, it is a defatting agent and can cause dermatitis 

with prolonged exposure to skin (IARC, 1977). A subacute study of daily oral intake of 

isopropanol (2.6 or 6.4 mg/kg body weight) by groups of 8 men for 6 weeks had no effect on 

blood cells, serum or urine and produced no subjective symptoms (Wills et al., 1969). A 

pharmacokinetic study of men occupationally exposed to isopropyl alcohol revealed that uptake 

occurs readily via the inhalation route; acetone is the major metabolite (Brugnone et al., 1983). 

Acetone was eliminated mainly by the lung but was also eliminated in the urine. 

V. Effects of Animal Exposures 

In metabolism studies with rats and mice, up to 92% of the administered dose (via i.v. or 

inhalation) of isopropanol was exhaled as acetone, CO2 and the unmetabolized alcohol (Slauter 

et al., 1994). Approximately 3-8% of the administered dose was excreted in urine as 

isopropanol, acetone, and a metabolite tentatively identified as isopropyl glucuronic acid. 

Isopropanol is readily absorbed from the GI tract and persists in the circulation longer than ethyl 

alcohol. Alcohol dehydrogenase oxidizes most isopropanol to acetone. Acetone may be further 

metabolized to acetate, formate, and finally CO2. In another metabolism study, the amount of 

acetone in the blood stream was found to be directly related to the air concentration of 

isopropanol (Laham et al., 1980). This finding indicated that the acetone metabolite could be 

used as a biochemical indicator of isopropanol exposure. 

Subchronic studies by Guseinov and Abasov (1982) and Baikov et al. (1974) reported changes in 

certain hematologic and clinical chemistry parameters, as well as increases in some organ 

weights. But the Environmental Protection Agency deemed these studies insufficient to 

reasonably predict subchronic toxicity of isopropanol (Burleigh-Flayer et al., 1994). Three 

different routes of exposure have been used by researchers for isopropanol toxicity studies: 

inhalation, oral gavage and presence in drinking water. The following subchronic and chronic 

studies exposed experimental animals to isopropanol by the inhalation route: 

Toxicological and neurobehavioral endpoints were investigated in rats and mice following 13-

week inhalation exposure (6 hr/day, 5 days/week) to 0, 100, 500, 1500 or 5000 ppm isopropanol 

(Burleigh-Flayer et al., 1994). In rats, clinical signs observed following exposures included 

swollen periocular tissue (females) at the highest dose and perinasal encrustation (males) at 500 

ppm and above. Narcosis was observed in a few animals of both species during exposure to 

5000 ppm and possibly 1500 ppm as well. However, the animals became tolerant to the narcotic 

effects of isopropanol after week 2. No neurobehavioral changes were observed in any 

parameters of the functional observational battery. However, increased motor activity was noted 

Appendix D3 338 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

at week 9 of exposure in female rats of the 5000 ppm group. After an initial drop in body weight 

gain in the first week of exposure at the high dose (5000 ppm), rats in the 1500 and 5000 ppm 

groups had significant increases in body weight gain and/or body weight throughout most of the 

exposure period. But only the 5000 ppm group had greater than 10% body weight gain 

compared to controls. Increases in body weight and body weight gain greater than 10% were 

also noted in female mice in the 5000 ppm group. Consistent clinical pathology changes 

included an increase in mean corpuscular volume (rats; female mice) and mean corpuscular 

hemoglobin (male rats; female mice) at the 5000 ppm exposure level. Other changes noted 

include a slight anemia in rats at week 6 only and a slight dehydration in female mice at the end 

of the study. Relative liver weight in rats was elevated no more than 8% in the 5000 ppm 

groups. However, a 10 and 21% increase in relative liver weight was observed in female mice at 

1500 and 5000 ppm, respectively. No gross lesions were observed in any organs. The only 

microscopic change observed was hyaline droplets within kidneys of all male rats. This change 

was not clearly concentration related, although it was most pronounced in the 5000 ppm group. 

In a follow-up inhalation study spanning the lifetime of rats and mice, Burleigh-Flayer et al. 

(1997) exposed four groups of animals, each consisting of 75 CD-1 mice/sex and 75 Fischer 344 

rats/sex, to 0, 500, 2500, or 5000 ppm isopropanol vapor. Of these, 55 mice/sex/group and 65 

rats/sex/group were exposed 6 hr/day, 5 days/week for at least 78 weeks (mice) or 104 weeks 

(rats). Transient signs of narcosis were observed at the higher doses. Increased mortality and a 

decreased mean survival time (577 days versus 631 days for controls) were noted for male rats in 

the 5000 ppm group. Increases in body weight and/or body weight gain were observed for both 

sexes of mice and rats from the 2500 and 5000 ppm groups throughout the study. Concentrationrelated increases in absolute and relative liver weight were observed for male and female mice. 

In addition, increased absolute and/or relative liver and kidney weight were observed for male 

and/or female rats from the 2500- and 5000-ppm groups. Urinalysis and changes in urine 

chemistry, indicative of impaired kidney function (i.e. decreased osmolality and increased total 

protein, volume, and glucose), were noted for male rats in the 2500 ppm group and for male and 

female rats in the 5000 ppm group. At necropsy, the most significant noncancer lesions in rats 

were observed in the kidney, and were associated with an exacerbation of spontaneous chronic 

renal disease. The kidney lesions noted with increased severity and/or frequency included 

mineralization, tubular dilation, glomerulosclerosis, interstitial nephritis, interstitial fibrosis, 

hydronephrosis, and transitional cell hyperplasia. The authors considered chronic renal disease 

to be the main cause of death for male and female rats exposed to 5000 ppm and to account for 

much of the mortality observed for male rats exposed to 2500 ppm. Unlike the subchronic study, 

anemia was not observed in rats in the chronic study. In mice, an increased incidence of seminal 

vesicle enlargement was observed grossly in males in the 2500 and 5000 ppm groups. 

Microscopically, the lesions in mice included an increased incidence of ectasia (dilation) of the 

seminal vesicles for male mice in the 2500 and 5000 ppm groups, minimal renal tubular 

proteinosis for male and female mice from all isopropanol groups, and renal tubular dilation for 

female mice in the 5000-ppm group. The seminal vesicle effects did not have any associated 

inflammatory or degenerative changes. The enlargement may have been the result of either 

increased secretion or decreased evacuation of the secretory product by these glands. 

Microscopic evaluation of the livers of rats and mice revealed no exposure-related lesions. 

In a 13-week behavioral/neurotoxicity study by the same investigators, the reproducibility and 

reversibility of increased motor activity in isopropanol-exposed female Fischer 344 rats was 

Appendix D3 339 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

investigated (Burleigh-Flayer et al. 1998). Rats were exposed to 0 or 5000 ppm isopropanol for 

6 hr/day, 5 days/week. Increased motor activity was characterized as the summation of 

ambulation, rearing and fine movements and was first observed 4 weeks following exposure to 

5000 ppm isopropanol. Reversibility of this effect was observed 2 days following cessation of 

exposure in a subgroup of rats exposed to isopropanol for only 9 weeks. In the subgroup 

exposed for 13 weeks, reversal of the increased motor activity did not occur until 2 weeks 

following cessation of exposure. However, complete reversibility of the time versus activity 

profile, or habituation curve, was not noted until 42 days following exposure to isopropanol for 

13 weeks.. Other effects included a significant increase in body weight and an increased 

incidence of swollen periocular tissue in isopropanol-exposed animals. 

In a study conducted to investigate neurochemical and behavioural effects, 20 male Wistar 

rats/group were exposed to 0 or 300 ppm isopropanol 6 hr/day, 5 days/week for up to 21 weeks 

(Savolainen et al., 1979). Enzyme activity of superoxide dismutase and azoreductase in 

cerebellar homogenate was decreased at week 20-21. Acid protease activity in glial cells was 

increased up to week 10. Open-field tests indicated sporadic changes in urination (10th week) 

and defecation (15th week). Isopropanol also appeared to depress caffeine stimulation activity at 

15 weeks. 

In a subchronic neurotoxicity study by Teramoto et al. (1993), motor and sensory nerve 

conduction velocity increased significantly following a 20-week exposure (8 hr/day, 5 

days/week) of Jcl-Wistar rats to 8000 ppm isopropanol. Low dose (1000 ppm) exposure had no 

effect on conduction velocity. Conduction velocities returned to normal following the end of 

exposure. The sex of the rats in this study was not specified. 

A developmental study in rats exposed pregnant dams (15/group) to 0, 3500, 7000 or 10,000 

ppm isopropanol 7 hr/day on gestation days 1-19 (Nelson et al., 1988). At the two highest 

exposure levels, feed intake (weeks 1 and 2 of exposure) and maternal body-weight gain were 

reduced. Narcosis was evident only at the 10,000 ppm level.Increased fetal resorptions and 

reduced fetal weights (59% of controls) occurred at the highest exposure level. Fetal weights 

were also significantly reduced (85% of controls) at 7000 ppm. A slight reduction in fetal 

weight (96% of controls) occurred at 3500 ppm but was significant in the sense that a dosedependent relationship in fetal weight reduction was present across all exposed groups. Skeletal 

malformations (primarily rudimentary cervical ribs) were seen only in the presence of maternal 

toxicity at the two highest exposure levels. No detectable teratogenic effects were observed in 

the 3500 ppm group. The authors noted that the developmental effects at 3500 ppm were 

considered very slight, indicating that this exposure level is close to the LOAEL for isopropanol. 

The following studies administered isopropyl alcohol to experimental animals by oral gavage: 

In a developmental study, pregnant (VAF)CD(SD) rats (25/group) were gavaged with either 0, 

400, 800 or 1200 mg/kg body wt-day of isopropanol daily on gestational days 6 through 15 (Tyl 

et al., 1994). In the same study, pregnant New Zealand white rabbits (15/group) were dosed 

orally with either 0, 120, 240 or 480 mg/kg body wt-day of isopropanol daily during gestational 

days 6 through 18. In rats, fetal body weight exhibited a linear downward trend with increasing 

dose and was significantly lower at the highest dose compared to controls. However, the fetal 

Appendix D3 340 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

body weight differences at each dose level was less than 10% of controls. Maternal weight gain 

during gestation was significantly reduced at the highest dose level. In rabbits, maternal weight 

gain and food consumption was reduced during gestation at 480 mg/kg body wt-day. Four 

rabbits died after dosing at this level. No differences were observed in reproduction indices or in 

fetal development. No teratogenic effects were seen in either species. 

In another developmental study performed to investigate neurotoxicity in rat pups, 64 timemated Sprague-Dawley rats/group were administered 0, 200, 700 or 1200 mg/kg body wt-day 

isopropanol by oral gavage from gestational day 6 through postnatal day 21 (Bates et al., 1994). 

One high-dose dam died on postnatal day 15, but there were no other clinical observations of 

effects on maternal weight, food consumption or gestation length. All fetal developmental 

indices were unaffected at the dose levels used. Developmental neurotoxicity, in the form of 

motor activity, auditory startle and active avoidance tests, was not found at any dose of 

isopropanol. 

In a multi-generation study carried out to investigate potential reproductive and developmental 

effects of isopropanol, Sprague-Dawley rats were administered 0, 100, 500 or 1000 mg/kg body 

wt-day of isopropanol by oral gavage (Bevan et al., 1995). P1 and P2 rats were dosed daily for 

10 weeks prior to mating, throughout the mating, and during the gestation and lactation period 

for the F1 and F2 litters, respectively. In adult rats, centrilobular hepatocyte hypertrophy and 

increased relative liver weight (>10%) was observed in P2 males at 1000 mg/kg. A general 

increase in absolute and relative liver and kidney weights was observed (less than 10% in most 

cases) in treated animals in both P1 and P2 generations. However, with the exception of 

hepatocellular hypertrophy in P2 males, no histopathological effects relevant to human risk were 

present. Reproductive effects due to isopropanol were not seen at any dose level. Statistically 

significant reduction of body weights (5-12%) in F1 and F2 offspring and increased mortality 

(14%) in F1 offspring were observed at the 1000 mg/kg dose level. 

The following toxicology studies administered isopropanol to experimental animals in drinking 

water: 

In a study designed to investigate neurotoxicological effects, 22 male SPF rats/group were 

administered isopropanol in drinking water at concentrations of 0, 1, 2, 3, or 5% (w/v) for 12 

weeks (Pilegaard and Ladefoged, 1993). Average daily intake of isopropanol was 0, 870, 1280, 

1680 and 2520 mg/kg body wt, respectively. Water intake and body weights were consistently 

lower at the two highest doses. Relative weights of liver, kidney and adrenals were increased in 

a dose-dependent manner. However, histopathology revealed no treatment-related changes in 

organs other than the male rat-specific kidney lesions. Evidence of astrogliosis, in the form of 

increased glial fibrillary acidic protein in dorsal hippocampus, was not found in exposed rats. 

In a 1-generation study, 10 Wistar-derived rats/sex/group were exposed to 0, 0.5, 1.25, 2.0, and 

2.5% isopropanol in water for up to 18 weeks (USEPA/OTS, 1986). The doses are equivalent to 

0, 325, 711, 1002, and 1176 mg/kg body wt-day, respectively, for males; to 0, 517, 1131, 1330, 

and 1335 mg/kg body wt-day, respectively, for females during the pre-mating phase; and to 0, 

1167, 2561, 2825, and 2722 mg/kg body wt-day, respectively, in females during the post-partum 

phase. Exposure periods were: 70 days pre-mating, plus 15 days during mating, plus 42 days for 

Appendix D3 341 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

males; 21 days pre-mating plus 15 days during mating, plus 21 days gestation, plus 21 days 

rearing in females; and 21 days for the F1 generation. At the highest two levels, body weights of 

males during the first two weeks were reduced and the body weights of females during the postpartum period were reduced. Water consumption and food ingestion were generally lower at the 

top three dose levels. The authors concluded that these effects were related to the unpalatability 

of drinking water containing isopropanol and not due to a toxic effect of the alcohol itself. 

Anemia was present in post-partum females. Red cell numbers were reduced in a dose-related 

manner at doses of 1.25% isopropanol or higher. Hematocrit was lower at the two highest doses 

while hemoglobin was lower at the highest dose. In males, mean cell volume was reduced at 

1.25% isopropanol or higher. Absolute and relative liver and kidney weights were higher in 

most exposure groups at 2.0% or higher in both sexes, but no relevant pathology was seen. 

Absolute liver weight of females was also higher in the 1.25% group. Fetal weight gain was 

depressed in a dose-related fashion in the 1.25% and higher groups. Mean pup weights and pup 

survival were lower than controls at the two highest doses. Fewer pups were born per animal in 

the 2.5% exposure group. A teratogenic examination was not performed on the pups. 

In a similar exposure study investigating the potential teratogenic effects of isopropanol, 20 

pregnant Wistar-derived rats/group were exposed to 0, 0.5, 1.25 or 2.5% of the alcohol in 

drinking water (equivalent to 0, 596, 1242 and 1605 mg/kg body wt-day) during gestational days 

6 to 16 (USEPA/OTS, 1992a,b). Water and feed consumption were reduced at the two highest 

doses while maternal body weight was significantly reduced at the highest dose. Fetal body 

weights were decreased in the two highest dose groups. Minor abnormalities and variants 

(reduced ossification of the skeleton) were present in fetuses of exposed groups in a dose-related 

manner. However, the authors concluded that the reduced fetal weights are probably a 

consequence of maternal growth retardation during the critical period of organogenesis. 

Similarly, the fetal abnormalities are probably due to small fetal size, related to slightly retarded 

development. Therefore, the study found no indication of teratogenesis. 

A multi-generation study performed in ‘white’ rats also observed reduced body weights in F1 

offspring (Lehman et al., 1945). Body weights of F2 offspring were the same as controls. The 

adult rats had imbibed an average of 1.9 ml/kg (1470 mg/kg body wt) of isopropanol per day in 

drinking water 80 days prior to mating. No other developmental or reproductive effects were 

seen. In the same study, several dogs were given 4% isopropanol in drinking water for 

approximately 7 months. Histopathology at the end of exposure revealed a decrease in the 

number of nephrons with hydropic changes and necrosis of some of the tubular epithelium. 

Some capillary hemorrhages were also noted in the brains of two of the dogs. Average daily 

dose of isopropanol imbibed by the dogs could not be determined from data provided in the 

report. 

Appendix D3 342 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Burleigh-Flayer et al. (1997) 

Study population Rats and mice 

Exposure method Discontinuous whole-body inhalation 

(0, 504, 2,509 or 5,037 ppm) 

Critical effects Kidney lesions in mice and rats 

LOAEL 2,509 ppm 

NOAEL 504 ppm 

Exposure continuity 6 hours/day, 5 days/week

Exposure duration 78 weeks in mice; 104 weeks in rats 

Average experimental exposure 90 ppm for NOAEL group (500 x 6/24 x 

5/7) 

Human equivalent concentration 90 ppm for NOAEL group (gas with 

systemic effects, based on RGDR = 

1.0 using default assumption that 

lambda (a) = lambda (h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies factor 10

Cumulative uncertainty factor 30 

Inhalation reference exposure level 3 ppm (3000 ppb, 7 mg/m3

, 7000 

μg/m3

) 

The Burleigh-Flayer et al. (1997) study was selected because it was a chronic study, was recent, 

and was published in a respected, peer-reviewed journal. While numerous subchronic studies 

have been performed, this was the only study that conducted lifetime animal exposures. In 

addition, the chronic kidney effects observed in rats and mice were not seen in the subchronic 

studies, indicating that chronic exposure is necessary for development of these lesions. 

The lesions observed in the kidneys of male rats in some of the studies described above is typical 

of a male rat-specific chronic renal disease and is not considered to be relevant to human risk 

assessment (Phillips and Cockrell, 1984; Beyer, 1992). However, the exacerbation of chronic 

renal disease in male and female rats, and the slight kidney damage observed in mice of both 

sexes following chronic isopropanol exposure indicates that the kidney is a sensitive indicator for 

nonneoplastic effects (Burleigh-Flayer et al., 1997). Suggestive evidence also exists for kidney 

damage in dogs following subchronic exposure to isopropanol in drinking water (Lehman et al., 

1945). 

Some isopropanol exposure studies noted increased liver and kidney weights in exposed animals 

but no observable relevant pathology. With particular relevance to the liver, this weight change 

may be considered to be more of a metabolic response rather than a toxic effect of the alcohol. 

The changes noted in the neurochemical and behavioural study by Savolainen et al. (1979) may 

have also been more of a metabolic response to the increased load of isopropanol. It is also 

possible that these changes reflected the development of tolerance. The changes in behavior 

Appendix D3 343 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

were small and unconvincing. This study would have benefited from additional dose levels to 

analyze for dose-response trends. 

Other possible sensitive indicators of isopropanol toxicity include blood chemistry changes and 

reduced fetal body weights. However, the blood chemistry findings were conflicting among the 

various studies that investigated this endpoint. Reduced fetal weights at doses below maternal 

body weight reductions were minor (<10% compared to controls), but consistent, suggesting that 

reduced fetal weights are a manifestation of isopropanol developmental toxicity. 

A comparative REL was calculated from the only reproduction/developmental study that utilized 

inhalation as the route of exposure (Nelson et al., 1988). Exposure of pregnant rats to 

isopropanol during gestation caused dose-dependent reduction in fetal body weights across all 

treatment groups, resulting in a LOAEL of 3500 ppm (average measured concentration = 3510 

ppm). A NOAEL was not observed for this effect. Skeletal malformations probably related to 

reduced fetal weight was observed at 7000 ppm and 10,000 ppm. The average exposure duration 

at the LOAEL for this study is 1024 ppm (7hr/24hr x 3510 ppm). Use of an RGDR of 1 and a 

cumulative uncertainty factor of 100 (3 for LOAEL to NOAEL, 3 for interspecies, and 10 for 

intraspecies) resulted in a REL of 10 ppm (25 mg/m3

). Since the endpoint is a function of 

exposure only during gestation, no subchronic to chronic UF was used. This developmental REL 

is within an order of magnitude of the chronic REL for kidney lesions, and therefore, is also 

considered to be a critical effect. 

The oral dose developmental studies by Tyl et al. (1994), Bevan et al. (1995), USEPA/OTS 

(1986), and USEPA/OTS (1992 a, b) provide supportive evidence that reduced fetal weights is a 

sensitive developmental endpoint. The USEPA/OTS (1992a,b) study provides supportive 

evidence for skeletal malformations in exposed rat fetuses. 

VII. Data Strengths and Limitations for Development of the REL 

Strengths of the database for isopropanol include availability of a well-conducted chronic study 

in two species, similar toxicological endpoints among different studies, and pharmacokinetic 

similarities between humans and experimental animals. Isopropanol is metabolized through a 

similar pathway to acetone and CO2. 

Weaknesses of the database for isopropanol include a lack of literature regarding chronic toxicity 

endpoints in humans. The deficiency of chronic toxicity cases in humans may be related to the 

relatively low chronic toxicity of isopropanol. Another weakness is that, while most 

developmental studies observed maternal and fetal effects, only one study was performed via the 

inhalation route. 

Appendix D3 344 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VIII. References

Amoore JE, and Hautala E. 1983. Odor as an aid to chemical safety: Odor thresholds compared 

with threshold limit values and volatiles for 2114 industrial chemicals in air and water dilution. J. 

Appl. Toxicol. 3(6):272-290. 

Baikov BK, Gorlova OE, Gusev MI, Novikov Yu V, Yudina TV, and Sergeev AN. 1974. 

Hygenic standardization of the daily average maximum admissible concentrations of propyl and 

isopropyl alcohols in the atmosphere. Gigiena I Sanitariia, 4:6-13. 

Bates HK, McKee RH, Bieler GS, Gardiner TH, Gill MW, Strother DE, and Masten LW. 1994. 

Developmental neurotoxicity evaluation of orally administered isopropanol in rats. Fundam. 

Appl. Toxicol. 22:152-158. 

Bevan C, Tyler T, Gardiner T, Kapp R, Andrews L, Beyer B. 1995. Two-generation 

reproduction toxicity study with isopropanol in rats. J. Appl. Toxicol. 15:117-223. 

Brugnone F, Perbellini L, Apostoli P, Bellomi M, and Caretta D. 1983. Isopropanol exposure: 

environmental and biological monitoring in a printing works. Br. J. Ind. Med., 40:160-168. 

Burleigh-Flayer H, Gill M, Hurley J, Bevan C, Gardiner T, Kapp R, Tyler T, Wright G. 1998. 

Motor activity effects in female Fischer 344 rats exposed to isopropanol for 90 days. J. Appl. 

Toxicol. 18:373-381. 

Burleigh-Flayer H, Garman R, Neptun D, Bevan C, Gardiner T, Kapp R, Tyler T, Wright G. 

1997. Isopropanol vapor inhalation oncogenicity study in Fischer 344 rats and CD-1 mice. 

Fundam. Appl. Toxicol. 36(2):95-111. 

Burleigh-Flayer HD, Gill MW, Strother DE, Masten LW, McKee RH, Tyler TR, Gardiner T. 

1994. Isopropanol 13-week vapor inhalation study in rats and mice with neurotoxicity evaluation 

in rats. Fundam. Appl. Toxicol. 23:421-428. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

C&EN. 1995. Chemical & Engineering News. Production by the U.S. chemical industry. M. 

Heylin, ed., American Chemical Society, Washington D. C., June 26:38-44. 

Guseinov VG, and Abasov DM. 1982. Toxicological characteristics of isopropyl alcohol after 

chronic exposure under experimental conditions. Azerb. Med. Zh. 58:53-57. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/95). 

IARC. 1997. IARC monographs on the evaluation of the carcinogenic risk of chemicals to man. 

Vol. 15. Lyon: IARC. pp. 223-243. 

Appendix D3 345 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Laham S, Potvin M, Schrader K, and Marino I. 1980. Studies on inhalation toxicity of 2-

propanol. Drug Chem. Toxicol. 3(4):343-360. 

Lehman AJ, Schwerma H, and Rickards E. 1945. Isopropyl alcohol acquired tolerance in dogs, 

rate of disappearance from the blood stream in various species, and effects on successive 

generation of rats. J. Pharmacol. Exper. Therap. 85:61-69. 

Nelson BK, Brightwell WS, MacKenzie-Taylor DR, Khan A, Burg JR, Weigel WW, and 

Goad PT. 1988. Teratogenicity of n-propanol and isopropanol administered at high inhalation 

concentrations to rats. Food Chem. Toxicol. 26(3):247-254. 

Phillips RD, and Cockrell B. 1984. Effect of certain light hydrocarbons on kidney function and 

structure in male rats. Adv. Mod. Environ. Toxicol, 7:89-106. 

Pilegaard K, and Ladefoged O. 1993. Toxic effects in rats of twelve weeks’ dosing of 2-

propanol, and neurotoxicity measured by densitometric measurements of glial fibrillary acidic 

protein in the dorsal hippocampus. in vivo 7:325-330. 

Savolainen H, Pekari K, and Helojoki H. 1979. Neurochemical and behavioural effects of 

extended exposure to isopropanol vapour with simultaneous ethanol intake. Chem.-biol. Interact. 

28:237-248. 

Slauter R, Coleman D, Gaudette N, McKee R, Masten L, Gardiner T, Strother D, Tyler R, and 

Jeffcoat A. 1994. Disposition and pharmacokinetics of isopropanol in F-344 rats and B6C3F1 

mice. Fundam. Appl. Toxicol. 23:407-420. 

Teramoto K, Wakitani F, Horiguchi S, Jo T, Yamamoto T, Mitsutake H, and Nakseko H. 1993. 

Comparison of the neurotoxicity of several chemicals estimated by the peripheral nerve 

conduction velocity in rats. Environ. Res. 62:148-154. 

Tyl RW, Masten LW, Marr MC, Myers CB, Slauter RW, Gardiner TH, Strother DE, McKee RH, 

and Tyler TR. 1994. Developmental toxicity evaluation of isopropanol by gavage in rats and 

rabbits. Fundam. Appl. Toxicol, 22:139-151. 

United States Environmental Protection Agency/Office of Toxic Substances. 1986. A pilot onegeneration study with isopropyl alcohol in rats (final report) with attachments and cover letter 

dated 013092. EPA/OTS Doc. #86-920000728. Prepared by British Industrial Biological 

Research Association. Submitted by Chemical Manufacturers Association. Fiche No. 

NTIS/OTS0535607. 

United States Environmental Protection Agency/Office of Toxic Substances. 1992a. A teratology 

study with isopropyl alcohol in rats. Letter submitting one enclosed teratology study and one 

enclosed supplementary study on isopropanol. EPA/OTS Doc. #86-920000727. Prepared by 

Appendix D3 346 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

British Industrial Biological Research Association. Submitted by Chemical Manufacturers 

Association. Fiche No. NTIS/OTS0535606. 

United States Environmental Protection Agency/Office of Toxic Substances. 1992b. 

Investigations supplementary to studies with isopropyl alcohol. Letter submitting one enclosed 

teratology study and one enclosed supplementary study on isopropanol. EPA/OTS Doc. #86-

920000727. Prepared by British Industrial Biological Research Association. Submitted by 

Chemical Manufacturers Association. Fiche No. NTIS/OTS0535604. 

Wills JH, Jameson EM, and Coulston F. 1969. Effects on man of daily ingestion of small doses 

of isopropyl alcohol. Toxicol. Appl. Pharmacol. 15:560-565. 

Appendix D3 347 Isopropanol 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

MALEIC ANHYDRIDE 

(2,5-furandione; cis-butenedioic anhydride; toxilic anhydride; maleic andride) 

CAS Registry Number: 108-31-6 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.7 μg/m3 (2.5 ppb)

Critical effect(s) Neutrophilic infiltration of the nasal epithelium; 

irritation of the respiratory system in rats, 

hamsters and monkeys 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1995)

Description Colorless or white solid 

Molecular formula C4H2O3

Molecular weight 98.06 g/mol 

Boiling point 202°C 

Melting point 52.8°C 

Vapor pressure 0.1 torr @ 25°C (AIHA, 1970) 

Solubility Soluble in water, ether, acetate, chloroform, 

dioxane; @ 25°C, 227 g/100 g acetone, 

112 g/100 g ethyl acetate, 52.5 g/100 g 

chloroform, 50 g/100 g benzene, 

23.4 g/100 g toluene, 19.4 g/100 g o-xylene, 

0.6 g/100 g CCl4, 0.25 g/100 g ligroin 

Conversion factor 4.0 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Maleic anhydride is used as a chemical intermediate in the synthesis of fumaric and tartaric acid, 

certain agricultural chemicals, resins in numerous products, dye intermediates, and 

pharmaceuticals (HSDB, 1995). It is also used as a co-monomer for unsaturated polyester resins, 

an ingredient in bonding agents used to manufacture plywood, a corrosion inhibitor, and a 

preservative in oils and fats. The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 7366 pounds of maleic anhydride (CARB, 2000). 

Appendix D3 348 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IV. Effects of Human Exposure 

In many occupational situations workers are exposed to mixtures of acid anhydrides, including 

maleic anhydride, phthalic anhydride, and trimellitic anhydride. For example, Barker et al. 

(1998) studied a cohort of 506 workers exposed to these anhydrides. In one factory, workers 

were exposed only to trimellitic anhydride, which has the lowest acceptable occupational 

exposure limit (40 μg/m3) of the three anhydrides. In that factory there was an increased 

prevalence of sensitization to acid anhydride and work related respiratory symptoms with 

increasing full shift exposure even extending down to levels below the current occupational 

standard. However, none of the workplaces had exposure only to maleic anhydride and a doseresponse relationship was not seen with mixed exposures. 

The following reports involve exposure only to maleic anhydride. 

There are several case reports describing asthmatic responses possibly resulting from exposure to 

maleic anhydride. An individual showed an acute asthmatic reaction after exposure to dust 

containing maleic anhydride (Lee et al., 1991). Workplace concentrations of maleic anhydride 

were 0.83 mg/m3 in the inspirable particulate mass and 0.17 mg/m3 in the respirable particulate 

mass. Bronchial provocation testing was performed with phthalic anhydride, lactose, and maleic 

anhydride. Exposure of this individual to maleic anhydride (by bronchial provocation testing) at 

0.83 mg/m3 and 0.09 mg/m3 in inspirable and respirable particulate mass, respectively, showed a 

response of cough, rhinitis, and tearing within two minutes. Within 30 minutes, rales developed 

in both lungs and peak flow rate decreased 55%. 

An individual occupationally exposed to maleic anhydride developed wheezing and dyspnea 

upon exposure (Gannon et al., 1992). After a period without exposure, two re-exposures both 

resulted in episodes of severe hemolytic anemia. There was no evidence of pulmonary 

hemorrhage. Radioallergosorbent testing showed specific IgE antibodies against human serum 

albumin conjugates with maleic anhydride, phthalic anhydride, and trimellitic anhydride, but not 

with tetrachlorphthalic anhydride. A critique of the Gannon et al. (1992) study by Jackson and 

Jones (1993) questions the relationship of maleic anhydride exposure to the onset of the anemia, 

since there were extended periods of exposure to maleic anhydride before symptoms appeared. 

Another case report described occupational asthma due to exposure to maleic anhydride (Guerin 

et al., 1980). 

Humans exposed to maleic anhydride showed respiratory tract and eye irritation at 

concentrations of 0.25 to 0.38 ppm (1 to 1.6 mg/m3

) maleic anhydride (Grigor’eva, 1964). No 

irritation was reported at 0.22 ppm maleic anhydride. 

V. Effects of Animal Exposure 

Short et al. (1988) chronically exposed CD rats (15/sex/group), Engle hamsters (15/sex/group), 

and rhesus monkeys (3/sex/group) to maleic anhydride by inhalation. Four groups of each 

species were exposed to concentrations of 0, 1.1, 3.3, or 9.8 mg/m3 maleic anhydride for 6 

Appendix D3 349 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

hours/day, 5 days/week, for 6 months in stainless steel and glass inhalation chambers. Solid 

maleic anhydride was heated to 53°C to generate vapors, which were then mixed with a stream 

of nitrogen. Chamber target levels were monitored by gas chromatography as total maleic 

(maleic anhydride plus maleic acid). No exposure-related increase in mortality occurred. Of the 

species examined, only rats showed significant changes in body weight during the course of the 

experiment, with reductions among males in the high-dose groups after exposure day 40 and a 

transient weight reduction from days 78-127 in the mid-dose group. All species exposed to any 

level of maleic anhydride showed signs of irritation of the nose and eyes, with nasal discharge, 

dyspnea, and sneezing reported frequently. No exposure-related eye abnormalities were 

reported. The severity of symptoms was reported to increase with increased dose. No doserelated effects were observed in hematological parameters, clinical chemistry, or urinalysis. No 

effects on pulmonary function in monkeys were observed. Dose-related increases in the 

incidence of hyperplastic change in the nasal epithelium occurred in rats in all exposed groups, 

and in hamsters in the mid- and high-dose groups. Neutrophilic infiltration of the epithelium of 

the nasal tissue was observed in all species examined at all exposure levels. All changes in the 

nasal tissues were judged to be reversible. The only other significant histopathological 

observation was slight hemosiderin pigmentation in the spleens of female rats in the high-dose 

group. 

Incidence of epithelial hyperplasia of the nasal mucosa in animals from Short et al. (1988) 

 Maleic anhydride (mg/m3) 0 0 1.1 1.1 3.3 3.3 9.8 9.8 

 Pathology grade Trace Mild Trace Mild Trace Mild Trace Mild 

Rat 

 Male 0/15 0/15 2/15 6/15 1/15 14/15 0/15 12/15 

 Female 0/15 0/15 6/15 5/15 4/15 10/15 0/15 14/15 

 Combined 0/30 11/30 24/30 26/30 

Hamster 

 Male 0/15 0/15 0/15 0/15 0/15 5/15 0/15 8/15 

 Female 0/15 0/15 0/15 0/15 4/15 4/15 1/15 4/15 

 Combined 0/30 0/30 9/30 12/30 

Monkey 

 Male 0/3 0/3 0/3 0/3 0/3 0/3 0/3 0/3 

 Female 0/3 0/3 0/3 0/3 0/3 0/3 0/3 0/3 

 Combined 0/6 0/6 0/6 0/6 

The teratogenicity and multigeneration reproductive toxicity of maleic anhydride were also 

investigated (Short et al., 1986). To evaluate teratogenicity, pregnant CD rats were treated orally 

with maleic anhydride in corn oil at concentrations of 0, 30, 90, or 140 mg/kg-day from 

gestational days 6-15. Animals were necropsied on gestational day 20. No statistically 

significant dose-related effects were observed in maternal weight gain, implantation, fetal 

viability, post-implantation loss, fetal weight, or malformations. Groups of 10 male rats and 20 

female rats/group (F0 animals) were orally treated with 0, 20, 55, or 150 mg/kg-day maleic 

anhydride in corn oil to study multigeneration reproductive toxicity. Animals within the same 

dose group were bred together after 80 days of treatment to produce two F1 generation animals 

(F1a and F1b) and animals from the F1 generation were interbred to produce two F2 generation 

animals (F2a and F2b). A significant increase in mortality was observed among both F0 and F1

Appendix D3 350 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

generation animals in the high-dose group. Total body weight was significantly reduced in 

animals in the high-dose group at Week 11 of exposure for the F0 generation males and females 

and at Week 30 of exposure in the F1 generation males. No consistent pattern of dose- or 

treatment-related effect on fertility, litter size, or pup survival was observed. Examination of F0

animals showed necrosis of the renal cortex in the high-dose group (60% of males and 15% of 

females). Absolute kidney weights were significantly increased in F1 females in the low- and 

mid-dose groups, although there was no histological correlate. No changes in organ weight or 

histology were observed in the F2 generation animals. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Short et al., 1988 

Study population Rats (15/sex/group), hamsters (15/sex/group), 

monkeys (3/sex/group) 

Exposure method Discontinuous inhalation exposure (0, 1.1, 3.3, 

or 9.8 mg/m3) 

Critical effects Neutrophilic infiltration of the nasal epithelium; 

epithelial hyperplasia; respiratory irritation 

LOAEL 1.1 mg/m3

NOAEL Not observed in rats 

BMC05 0.12 mg/m3

 for mild epithelial hyperplasia in rats 

(males and females combined) 

Exposure continuity 6 hr/day, 5 days/week 

Exposure duration 6 months 

Average experimental exposure 21 μg/m3

 for the BMC05 (0.12 x 6/24 x 5/7 x 

1000) 

Human equivalent concentration 21 μg/m3

 for the BMC05 (Due to the lack of 

aerosol particle size data for the critical study, 

a human equivalent concentration could not 

be developed using recommended methods of 

inhalation dosimetry.) 

LOAEL uncertainty factor not needed in benchmark approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 (see below) 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.7 μg/m3

 0.2 ppb) 

Short et al. (1988) examined the toxicity of maleic anhydride to rats, hamsters, and monkeys by 

the inhalation route of exposure. Dose- and exposure related effects, although mild and 

reversible, were observed at all exposure levels. Specifically, exposure to maleic anhydride 

vapors resulted in hyperplastic change in the nasal epithelium of rats and hamsters (obligate nose 

breathers). Neutrophilic infiltration of the nasal epithelium was observed in all three species at 

all levels of exposure. All species also showed signs of irritation at all exposure levels. The 

observation that acute maleic anhydride is a strong respiratory irritant to humans (ACGIH, 1992) 

Appendix D3 351 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

suggests that this is a valid endpoint of toxicity to humans as well. Human exposure at levels as 

low as ~1 mg/m3

 appears to trigger acute asthmatic reactions in sensitive individuals (Lee et al., 

1991). The histological changes observed by Short et al. occurring as a result of inhalation 

exposure to a known strong irritant such as maleic anhydride are considered to be the adverse 

effect of repetitive acute exposures, rather than a chronic response, in the development of the 

REL. 

The chronic REL was developed using the benchmark approach. The gamma model in the U.S. 

EPA's BMDS software yielded a BMC05 of 0.12 mg/m3 for mild epithelial hyperplasia in male 

and female rats combined. Because of the similarities among species and the inclusion of 

monkeys in the study, an interspecies uncertainty factor of 3, rather than 10, was used. Although 

there is no evidence of a toxic response similar to the development of asthma in animals, the 1.1 

mg/m3

 LOAEL from the animal studies of Short et al. (1988) results in a REL of 0.7 μg/m3

which should protect asthmatics from maleic and other anhydrides. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the REL for maleic anhydride are the availability of multiple-species, 

multiple-dose subchronic inhalation studies, and the observation of a mild effect LOAEL. The 

major uncertainties are the lack of human data and the lack of a NOAEL observation. 

VIII. Potential for Differential Impacts on Children's Health

Minimal teratogenic and reproductive adverse effects were seen at the lowest oral dose of maleic 

anhydride (20 mg/kg-day), given to rats during gestation (Short et al., 1986). This dose is 

equivalent to a person inhaling 70 mg/m3. Thus the chronic REL of 0.7 μg/m3 should protect 

children. Maleic anhydride is a respiratory irritant and an inducer of asthma. Exacerbation of 

asthma has a more severe impact on children than on adults. However, there is no direct 

evidence in the literature to quantify a differential effect of maleic anhydride in children. 

IX. References 

ACGIH. 1992. American Conference of Governmental Industrial Hygienists, Inc. 

Documentation of the threshold limit values and biological exposure indices. Sixth ed. 

Cincinnati, OH: ACGIH. 

AIHA. 1970. American Industrial Hygiene Association. Maleic anhydride. Am. Ind. Hyg. Assoc. 

J. 31:391-394. 

Barker RD, van Tongeren MJ, Harris JM, Gardiner K, Venables KM and Newman Taylor AJ. 

1998. Risk factors for sensitisation and respiratory symptoms among workers exposed to acid 

anhydrides: a cohort study. Occup. Environ. Med. 55(10):684-691. 

Appendix D3 352 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Gannon PFG, Burge PS, Hewlett C, and Tee RD. 1992. Haemolytic anaemia in a case of 

occupational asthma due to maleic anhydride. Br. J. Ind. Med. 49:142-143. 

Grigor’eva, K.V. 1964. Pollution of atmospheric air with maleic anhydride (Abstract in Chem. 

Abstr. (1966) 65:14319b). Gig. Sanit. 29:7-11. 

Guerin JC, Deschamps O, Guillot TL, Chavallion JM, and Kalb JC. 1980. A propos d’un cas 

d’asthme a l’anhydride maleique [A case of asthma due to maleic anhydride]. Poumon et le 

Coeur 36:393-395. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/95). 

Jackson J, and Jones AH. 1993. Correspondence: Haemolytic anaemia in a case of occupational 

asthma due to maleic anhydride. (with reply by P. Gannon). Br. J. Ind. Med. 50:191-192. 

Lee HS, Wang YT, Cheong TH, Tan KT, Chee BE, and Narendran K. 1991. Occupational 

asthma due to maleic anhydride. Br. J. Ind. Med. 48:283-285. 

Short RD, Johannsen FR, Levinskas GJ, Rodwell DE, and Schardein JL. 1986. Teratology and 

multigeneration reproduction studies with maleic anhydride in rats. Fundam. Appl. Toxicol. 

7:359-366. 

Short RD, Johannsen FR, and Ulrich CE. 1988. A 6-month multispecies inhalation study with 

maleic anhydride. Fundam. Appl. Toxicol. 10:517-524. 

Appendix D3 353 Maleic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

METHANOL 

(methyl alcohol, wood spirit, carbinol, wood alcohol, wood naphtha) 

CAS Registry Number: 67-56-1 

I. Chronic Toxicity Exposure Level

Inhalation reference exposure level 4,000 μg/m3 (3,000 ppb)

Critical effect(s) Increased incidence of abnormal cervical ribs, 

cleft palate, and exencephaly in mice 

Hazard index target(s) Teratogenicity 

II. Chemical Property Summary (HSDB, 1999; CRC, 1994) 

Description Colorless liquid 

Molecular formula CH3OH 

Molecular weight 32.04 g/mol 

Boiling point 64.6°C 

Melting point −97.6°C 

Vapor pressure 92 torr at 20°C 

Solubility Methanol is miscible with water, ethanol, ether 

and many other organic solvents. 

Conversion factor 1 ppm = 1.31 mg/m3

III. Major Uses and Sources 

Originally distilled from wood, methanol is now manufactured synthetically from carbon oxides 

and hydrogen. Methanol is used primarily for the manufacture of other chemicals and as a 

solvent. It is also added to a variety of commercial and consumer products such as windshield 

washing fluid and de-icing solution, duplicating fluids, solid canned fuels, paint remover, model 

airplane fuels, embalming fluids, lacquers, and inks. Methanol is also used as an alternative 

motor fuel (HSDB, 1999). The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 3,009,776 pounds of methanol (CARB, 1999b). 

IV. Effects of Human Exposure 

The majority of the available information on methanol toxicity in humans relates to acute rather 

than chronic exposure. The toxic effects after repeated or prolonged exposure to methanol are 

Appendix D3 354 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

believed to be qualitatively similar but less severe than those induced by acute exposure (Kavet 

and Nauss, 1990). These effects include CNS and visual disturbances such as headaches, 

dizziness, nausea and blurred vision. The role of formate, a metabolite of methanol, in chronic 

toxicity is unclear. 

In one study, symptoms of blurred vision, headaches, dizziness, nausea and skin problems were 

reported in teachers aides exposed to duplicating fluid containing 99% methanol (Frederick et 

al., 1984). Individual aides worked as little as 1 hr/day for 1 day a week to 8 hrs/day for 5 

days/wk. The workers’ total exposure duration was not mentioned. A dose-response 

relationship was observed between the self-reported amount of time spent at the duplicator and 

the incidence of symptoms. The concentrations of methanol in the breathing zones near the 

machines in 12 schools ranged from 485 to 4096 mg/m3

 (365 to 3080 ppm) for a 15 minute 

sample. 

Forty-five percent of duplicating machine operators experienced blurred vision, headache, 

nausea, dizziness and eye irritation (NIOSH, 1981). Air concentrations of methanol for 25 

minutes near the machines averaged 1330 mg/m3. 

Employees working in the proximity of direct process duplicating machines complained of 

frequent headaches and dizziness (Kingsley and Hirsch, 1954). Air concentrations of methanol 

ranged from 15 ppm (20 mg/m3

) to 375 ppm (490 mg/m3). 

Thirty young women, who had polished wood pencils with a varnish containing methanol, all 

experienced headaches, gastric disorders, vertigo, nausea and blurred vision (Tyson, 1912; as 

cited in NIOSH, 1976). 

None of the above studies specified the workers’ total duration of exposure. 

Ubaydullayev (1968) exposed 3 to 6 subjects to methanol vapor for short durations (40 minutes 

for some subjects and others for an unspecified amount of time). Electrical brain cortex reflex 

activity was significantly altered upon exposures to 1.17 mg/m3 (0.89 ppm) or 1.46 mg/m3 (1.11 

ppm). No effect was observed at 1.01 mg/m3

 (0.77 ppm). 

V. Effects of Animal Exposure

With the exception of non-human primates, the signs of methanol toxicity in commonly used 

laboratory animals are quite different from those signs observed in humans (Gilger and Potts, 

1955). The major effect of methanol in non-primates (rodents, dogs, cats, etc) is CNS depression 

similar to that produced by other alcohols. Metabolic acidosis and ocular toxicity are not 

observed. The differences in toxicity are attributed to the ability of non-primates to metabolize 

formate more efficiently than humans and other primates (Tephly, 1991). 

Two chronic studies have been conducted with monkeys. In one study, ultrastructural 

abnormalities of hepatocytes indicating alteration of RNA metabolism were observed in rhesus 

monkeys given oral doses of 3 to 6 mg/kg methanol for 3 to 20 weeks (Garcia and VanZandt, 

1969). In a study aimed at examining ocular effects, cynomolgous monkeys were exposed by 

Appendix D3 355 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

inhalation to methanol concentrations ranging from 680 mg/m3

 (520 ppm) to 6650 mg/m3 (5010 

ppm) for 6 hours per day, 5 days per week for 4 weeks (Andrews et al., 1987). No deaths 

occurred and no treatment-related effects were found upon histopathologic examination. 

However, Andrews et al. did not examine possible neurologic or reproductive effects which have 

been observed in other species at lower concentrations (see Sections IV and V).

Exposure to a mixture of methanol and other solvents has been associated with central nervous 

system birth defects in humans (Holmberg, 1979). However, because of mixed or inadequate 

exposure data, methanol is not considered a known human teratogen. 

In two separate studies in male rats, inhalation exposure to methanol ranging from 260 to 

13,000 mg/m3 for 6 to 8 hours per day for either 1 day or 1, 2, 4 or 6 weeks resulted in a 

significant reduction in testosterone levels (Cameron et al., 1984; Cameron et al., 1985). 

Ubaydullayev (1968) exposed rats (15 per group) to 0, 0.57, or 5.31 mg/m3

 methanol 

continuously for 90 days. Chronaxy ratios of flexor and extensor muscles were measured in 

addition to hematologic parameters and acetyl cholinesterase activity. No changes were apparent 

in the 0.57 mg/m3

 group. Effects observed in the 5.31 mg/m3 group included decreased blood 

albumin content beginning 7 weeks after exposure, slightly decreased acetylcholinesterase 

activity, decreased coproporphyrin levels in the urine after 7 weeks, and changes in muscle 

chronaxy. (Chronaxy is the minimum time an electric current must flow at a voltage twice the 

rheobase to cause a muscle to contract. The rheobase is the minimal electric current necessary to 

produce stimulation (Dorland, 1981). 

Pregnant rats were exposed by inhalation to methanol at concentrations ranging from 5000 to 

20,000 ppm for 7 hours per day on days 1-19 gestation, and days 7-15 for the highest dose group 

(Nelson et al., 1985). A dose-related decrease in fetal weight, an increase in extra or 

rudimentary cervical ribs, and urinary or cardiovascular defects were observed. Exencephaly 

and encephalocoele were observed in the 20,000 ppm dose group. The no-observed-adverseeffect level (NOAEL) was 5000 ppm. 

Pregnant mice were exposed to methanol vapors at concentrations ranging from 1000 to 

15,000 ppm for 7 hours per day on days 6-15 of gestation (Rogers et al., 1993). Increased 

embryonic and fetal death, including an increase in full-litter resorptions, was observed at 

7500 ppm and higher. Significant increases in the incidence of exencephaly and cleft palate 

were observed at 5000 ppm and higher. A dose-related increase in the number of fetuses per 

litter with cervical ribs (usually small ossification sites lateral to the seventh cervical vertebra) 

was observed at 2000 ppm and above. The NOAEL was 1000 ppm. 

Appendix D3 356 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Rogers et al. (1993) 

Study population Pregnant mice 

Exposure method Discontinuous inhalation, 7 hours/day on days 6-

15 of gestation 

Critical effects Abnormal cervical ribs, exencephaly, cleft palate 

LOAEL 5000 ppm 

NOAEL 1000 ppm 

Benchmark Concentration (BMC05) 305 ppm 

Exposure continuity 7 hr/day 

Exposure duration 10 days 

Average experimental exposure 89 ppm at BMC05 (305 ppm x 7/24) 

Human equivalent concentration 89 ppm atBMC05 (gas with systemic effects, 

based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

Subchronic uncertainty factor 1 (see below) 

LOAEL uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 3 ppm ( 3,000 ppb, 4 mg/m3

, 4,000 μg/m3) 

A NOAEL of 1000 ppm for developmental malformations was observed in mice exposed for 7 

hours/day on days 6 through 15 of gestation (Rogers et al., 1993). Although not a chronic study, 

the endpoint, teratogenicity, is a function of exposure only during gestation, especially in the 

case of a non-accumulating compound such as methanol. Therefore, an uncertainty factor to 

account for differences between subchronic and chronic exposures was not required. The 

investigators calculated maximum likelihood estimates (MLEs) using a log-logistic model for 

both 1% and 5% added risks above background. The most sensitive developmental toxicity 

endpoint was an increase in the incidence of cervical ribs. The MLE05 and BMC05 for cervical 

ribs were 824 ppm (1079 mg/m3

) and 305 ppm (400 mg/m3), respectively. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the REL for methanol are the observation of a NOAEL and the 

demonstration of a dose-response relationship. The major uncertainties are the lack of human 

data for chronic inhalation exposure, the lack of comprehensive, long-term multiple dose studies, 

and the difficulty in addressing reproductive short-term effects within the chronic REL 

framework. 

Appendix D3 357 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VIII. References 

Andrews LS, Clary JJ, Terrill JB, and Bolte HF. 1987. Subchronic inhalation toxicity of 

methanol. J. Toxicol. Environ. Health 20:117-124. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Cameron AM, Nilsen OG, Haug E, and Eik-Nes KB. 1984. Circulating concentrations of 

testosterone, luteinizing hormone and follicle stimulating hormone in male rats after inhalation 

of methanol. Arch. Toxicol. Suppl. 7:441-443.. 

Cameron AM, Zahlsen K, Haug E, Nilsen OG, and Eik-Nes KB. 1985. Circulating steroids in 

male rats following inhalation of n-alcohols. Arch. Toxicol. Suppl. 8:422-424 

Dorland. 1981. Dorland’s Illustrated Medical Dictionary. 26th ed. Philadelphia: W.B. Saunders. 

Frederick LJ, Schlulte PA, and Apol A. 1984. Investigation and control of occupational hazards 

associated with the use of spirit duplicators. Am. Ind. Hyg. Assoc. J. 45:51-55. 

Garcia JH, and VanZandt JP. 1969. Nucleolar segregation and chronic methanol intoxication. 

Proc. Electron Microsc. Soc. Am. 27:360 [as cited in Rowe and McCollister, 1978]. 

Gilger AP, and Potts AM. 1955. Studies on the visual toxicity of methanol: V. The role of 

acidosis in experimental methanol poisoning. Am. J. Ophthalmol. 39:63-86. 

Holmberg PC. 1979. Central nervous system defects in children born to mothers exposed to 

organic solvents during pregnancy. Lancet 2:177-179. 

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Kavet R, and Nauss KM. 1990. The toxicity of inhaled methanol vapors. Crit. Rev. Toxicol. 

21:21-50. 

Kingsley WH, and Hirsch FG. 1954. Toxicologic considerations in direct process spirit 

duplicating mahcines. Comp. Med. 6:7-8. 

Nelson BK, Brightwell WS, MacKenzie DR, Khan A, Burg JR, Weigel WW, and Goad PT. 

1985. Teratological assessment of methanol and ethanol at high inhalation levels in rats. 

Fundam. Appl. Toxicol. 5:727-736. 

NIOSH. 1976. National Institute for Occupational Safety and Health. Criteria document for 

methyl alcohol. Cincinnati, OH. 

NIOSH. 1981. National Institute for Occupational Safety and Health. Health Hazard Evaluation 

Report, HETA 81-177, PB82-194648.178-88 [as cited in Kavet and Nauss, 1990]. 

Appendix D3 358 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Rogers JM, Mole ML, Chernoff N, Barbee BD, Turner CI, Logsdon TR, and Kavlock JV. 1993. 

The developmental toxicity of inhaled methanol in the CD-1 mouse, with quantitative doseresponse modeling for estimation of benchmark doses. Teratology 47:175-188. 

Rowe RK, and McCollister SB. 1978. Alcohols. In: Patty’s Industrial Hygiene and Toxicology. 

Clayton GD,and Clayton FE (eds). New York: Wiley. pp. 4527-4541. 

Tephly TR. 1991. Minireview. The toxicity of methanol. Life Sci. 48:1031-1041. 

Tyson HH. 1912. Amblyopia from inhalation of methyl alcohol. Arch. Ophthalmol. 16:459-471 

[as cited in NIOSH, 1976]. 

Ubaydullayev R. 1968. A study of hygienic properties of methanol as an atmospheric air 

pollutant (translation by B.S. Levine). USSR Literature on Air Pollution and Related 

Occupational Diseases - A Survey. 17:39-45. 

Appendix D3 359 Methanol 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

METHYL BROMIDE 

(bromomethane; monobromomethane) 

CAS Registry Number: 74-83-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 5 μg/m3 (1 ppb) 

Critical effect(s) Histological lesions of the olfactory epithelium of 

the nasal cavity in rats 

Hazard index target(s) Respiratory system; nervous system; 

development 

II. Physical and Chemical Properties (HSDB, 1994) 

Description Colorless gas 

Molecular formula CH3Br 

Molecular weight 94.95 g/mol 

Density 3.89 g/L @ 25°C 

Boiling point 3.6°C 

Vapor pressure 1420 torr @ 20°C 

Solubility Soluble in ethanol, benzene, carbon disulfide, and 

1.75% (w/w) in water 

Odor threshold 20.6 ppm 

Odor description Sweetish odor 

Metabolites Methanol, bromide, 5-methylcysteine 

Conversion factor 1 ppm = 3.89 mg/m3

 @ 25° C 

III. Major Uses and Sources 

Methyl bromide (MeBr) was used historically as an industrial fire extinguishing agent and was 

introduced in the U.S. from Europe in the 1920s. Current uses of MeBr include the fumigation 

of homes and other structures for termites and other pests. Methyl bromide is also used to 

fumigate soil before planting and fruits and vegetables after harvest. In 1981, 6.3 million pounds 

of MeBr were reportedly used in California (Alexeeff and Kilgore, 1983). By 1991, its use had 

grown to 18.7 million pounds in the state (Cal/EPA, 1993). The annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California, based on 

the most recent inventory, were estimated to be 75,575 pounds of methyl bromide (CARB, 

1999). This does not include emissions of methyl bromide during its use as a pesticide. 

Appendix D3 360 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Human Exposure

Workers (n = 32) exposed to MeBr during fumigation of soil or structures were compared to a 

referent group of 29 workers not exposed to MeBr, but exposed to other fumigants (Anger et al., 

1986). Exposures to MeBr were not quantified. It was found that workers exposed to MeBr had 

a higher rate of neurological symptoms and performed less well on several behavioral tests. 

Several confounding factors were present in this study, including lack of adjustments for age, 

alcohol consumption, prescription medication, illegal drugs, education, or ethnic group between 

the exposed and the referent groups. 

V. Effects of Animal Exposure

The first experimental animal study on repeated MeBr exposures was carried out and reported by 

Irish and associates (1940). In this study, rats (135 per group), rabbits (104 per group), or female 

rhesus monkeys (13 per group) were exposed to 0, 17, 33, 66, 100, or 220 ppm (0, 66, 128, 256, 

388, or 853 mg/m3

) 7-8 hours/day, 5 days/week for 6 months or until the majority of the animals 

exhibited severe signs of toxicity. Mortality was seen in rats, guinea pigs, and monkeys at 100 

ppm. Rabbits began to die at 33 ppm. Severe effects, including paralysis, were seen after 

exposure to 66 ppm in rabbits and monkeys. None of the species exhibited adverse effects after 

exposure to 17 ppm. 

Kato and associates (1986) observed focal lesions in the brain and heart in rats (10-12 per group) 

after inhalation of 150 ppm (585 mg/m3) MeBr 4 hours/day, 5 days/week for 11 weeks. In 

another experiment, rats were exposed to 0, 200, 300, or 400 ppm (0, 777, 1160, or 1550 mg/m3) 

MeBr 4 hours/day, 5 days/week for 6 weeks. In this experiment, rats exposed to any 

concentration of MeBr exhibited coronary lesions, and exposures of 300 ppm or greater resulted 

in neurological dysfunction, including ataxia and paralysis. Testicular atrophy was noted in 6 of 

the 8 animals exposed to 400 ppm. 

Anger et al. (1981) determined that rabbits are more sensitive than rats to neurotoxicity of MeBr. 

In this study, rats or rabbits were exposed to 0 or 65 ppm (0 or 254 mg/m3

) MeBr for 7.5 

hours/day, 4 days/week, for 4 weeks. Nerve conduction velocity and eyeblink reflex were 

impaired in the rabbits but not rats exposed to 65 ppm MeBr. Similarly, rats did not exhibit 

neurological signs after exposure to 55 ppm (215 mg/m3

) MeBr for 36 weeks. Rabbits exposed 

to 26.6 ppm (104 mg/m3) did not display any neurological effects after 8 months exposure 

(Russo et al., 1984). 

In the studies of Reuzel and associates (1987, 1991), groups of 50 male and 60 female Wistar 

rats were exposed to 0, 3, 30, or 90 ppm methyl bromide (98.8%) for 6 hours per day, 5 days per 

week. Three groups of animals (10/sex/exposure level) were killed for observations at 14, 53, 

and 105 weeks of exposure. Body weight, hematology, clinical chemistry, and urinalyses were 

examined throughout the experiment in addition to histopathology and organ weights at time of 

necropsy. Exposures of males and females to 90 ppm resulted in reduced body weight. 

Exposure to 90 ppm also resulted in significant lesions in the heart in the form of cartilaginous 

metaplasia and thrombus in the males, and myocardial degeneration and thrombus in the 

Appendix D3 361 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

females. Exposure of males to 30 or 90 ppm resulted in a decrease in relative kidney weight. 

Histological changes in the nose, heart, esophagus, and forestomach were the principal effects of 

methyl bromide toxicity. At the lowest concentration (3 ppm), very slight degenerative changes 

in the nasal epithelium, and olfactory basal cell hyperplasia were noted in both sexes at 29 

months. Based on this study, a LOAEL of 3 ppm (11.7 mg/m3

) was determined. 

The National Toxicology Program (NTP) conducted a 13-week and a chronic study on the 

toxicology and carcinogenesis of methyl bromide in rats and mice (NTP, 1990). In the 13-week 

study, 18 rats/sex/group were exposed to 0, 30, 60, or 120 ppm (0, 117, 233, or 466 mg/m3

) 

MeBr 6 hours/day, 5 days/week. The mice were exposed to 0, 10, 20, 40, 80, or 120 ppm (0, 39, 

78, 155, 311, or 466 mg/m3

) 6 hours/day, 5 days/week. Hematological parameters and selected 

organ weights were measured in both species, in addition to histopathological changes. 

Pseudocholinesterase activity and neurobehavioral tests were conducted in the mice. Serious 

effects, including 58% body weight loss, 17% mortality and severe curling and crossing of the 

hindlimbs were observed in mice exposed to 120 ppm MeBr. Exposure of males to 40 ppm or 

higher resulted in significant effects on several hematological parameters, including decreased 

mean cell hemoglobin and increased red blood cell count. The only exposure-related histological 

effect was olfactory epithelial dysplasia and cysts in the rats of both sexes exposed to 120 ppm. 

A 6-week study in rats and mice (5 animals/sex/group) exposed to 0 or 160 ppm (0 or 

624 mg/m3

) showed high mortality rates, loss in body weight and histological changes in 

multiple organ systems including brain, kidney, nasal cavity, heart, adrenal gland, liver, and 

testes (NTP, 1990). 

An exposure of mice (86 animals/group) to 0, 10, 33, or 100 ppm (0, 38.8, 128, or 388 mg/m3) 

MeBr for 6 hours/day, 5 days/week, for 103 weeks was also conducted by NTP (1990). In this 

study, high mortality rates in both males and females in the 100 ppm group resulted in a 

discontinuation of exposure after 20 weeks. A low incidence of sternal dysplasia and a 

significant decrease in locomotor activity were noted in the 10 ppm group. 

A 5-day exposure of rats (10 animals/group) to 0, 90, 175, 250, or 325 ppm (0, 350, 680, 971, or 

1260 mg/m3

) resulted in lesions in the nasal olfactory sensory cells, the cerebellum and adrenal 

gland beginning at 175 ppm (Hurtt et al., 1987). Hurtt and Working (1988) later observed severe 

histological damage to the nasal epithelium following a single exposure to 90 or 200 ppm (351 or 

780 mg/m3

) MeBr. Olfactory function, measured by the ability to locate buried food, was 

impaired at the 200 ppm exposure. In this study, reduced testosterone and testicular glutathione 

levels were observed in the male rats exposed to 200 ppm, but no effects on spermatogenesis, 

sperm quality, or testes histopathology were noted. 

Sikov et al. (1981) examined the teratogenic potential of MeBr in rats and rabbits exposed to 0, 

20, or 70 ppm (0, 78, or 272 mg/m3) 7 hours/day, 5 days/week for 3 weeks during days 1-19 

(rats) or 1-24 (rabbits) of gestation. No maternal or fetal effects were observed in the rats, 

however, severe maternal neurotoxic effects were observed in the rabbits that resulted in 24/25 

deaths. In this study, no significant maternal or fetal effects were observed at a concentration of 

20 ppm. 

Appendix D3 362 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Another developmental toxicity study was conducted in rabbits by Breslin et al. (1990). In this 

study, rabbits were exposed to 0, 20, 40, or 80 ppm (0, 78, 156, or 312 mg/m3

) MeBr for 6 

hours/day on gestation days 6-19. Maternal toxicity was observed at 80 ppm and included 

reduced body weight gain and signs of neurotoxicity. In addition to the maternal effects 

observed, a significant increase in incidence of gall bladder agenesis and fused sternebrae were 

observed in the offspring exposed to 80 ppm. No adverse effects were observed at 40 ppm or 

lower concentrations. 

A 2-generation reproduction and developmental toxicity study on MeBr in rats was conducted by 

American Biogenics Corporation (1986). Groups of rats (25/sex/concentration) were exposed to 

0, 3, 30, or 90 ppm (0, 12, 117, or 350 mg/m3) MeBr 6 hours/day, 5 days/week during premating, 

gestation, and lactation through 2 generations. Significant decreases in body weight during the 

pre-mating period and at the end of the study were observed in the males exposed to 90 ppm. 

Although some adult organ weights were affected in the 90-ppm group, there was no evidence of 

histopathology in these organs. Neonatal body weights were decreased by exposure to 30 ppm. 

There was a decreased cerebral cortex width in the 90 ppm F1 group, reduced brain weight in 30 

ppm F1 females, and reduced fertility in the 30 and 90 ppm F2b groups. 

VI. Derivation of Chronic Reference Exposure Level 

Study Reuzel et al., 1987; 1991 

Study population Male and female Wistar rats (50 and 60 per group, 

respectively) 

Exposure method Discontinuous inhalation exposures 

 (0, 3, 30, or 90 ppm) over 29 months 

Critical effects Basal cell hyperplasia of the olfactory epithelium 

of the nasal cavity 

LOAEL 3 ppm 

NOAEL Not observed 

Exposure continuity 6 hr/day, 5 days/week 

Exposure duration 29 months 

Average experimental exposure 0.54 ppm for the LOAEL group 

Human equivalent concentration 0.12 ppm for the LOAEL group (gas with 

extrathoracic respiratory effects, RGDR = 0.23, 

based on MV = 0.03 m3

/min, SA = 11.6 cm2) 

LOAEL uncertainty factor 3 (20% extra risk of a mild effect) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.001 ppm (1 ppb, 0.005 mg/m3

, 5 μg/m3) 

The chronic REL for methyl bromide is also the U.S EPA RfC. 

Appendix D3 363 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the REL for methyl bromide are the use of a comprehensive, long-term, 

multiple dose study with large sample sizes, and the availability of supporting data including 

long-term studies in other species and reproductive and developmental studies. The major 

uncertainties are the lack of human data and the lack of a NOAEL observation for the critical 

effect. 

The California Department of Pesticide Regulation used a different approach that adjusts for 

respiration rate differences between humans and animals and which uses 10-fold uncertainty 

factors for interspecies differences, for intraspecies variability, and for a LOAEL to NOAEL 

extrapolation. Applying these factors to the same 3 ppm LOAEL results in a level for children 

and adults of 1 and 2 ppb (4 and 8 μg/m3

), respectively. 

VIII. References 

Alexeeff GV, and Kilgore WW. 1983. Methyl bromide. In: Residue Reviews. Vol. 88. New 

York: Springer-Verlag. pp. 102-153. 

American Biogenics Corporation. 1986. Two-generation reproduction study via inhalation in 

albino rats using methyl bromide. Final Report. American Biogenics Corporation Study 450-

1525, OTS 0515364 sponsored by the Methyl Bromide Panel. 

Anger WK, Moody L, Burg J, Brightwell WS, Taylor BJ, Russo JM, et al. 1986. 

Neurobehavioral evaluation of soil and structural fumigators using methyl bromide and sulfuryl 

fluoride. Neurotoxicology. 7(3):137-156. 

Anger WK, Setzer JV, Russo JM, Brightwell WS, Wait RG, and Johnson BL. 1981. 

Neurobehavioral effects of methyl bromide inhalation exposures. Scand. J. Work Environ. 

Health 7(Suppl. 4):40-47. 

Breslin WJ, Zablotny CL, Brabley GJ, and Lomax LG. 1990. Methyl bromide inhalation 

teratology study in New Zealand white rabbits. Toxicology Research Laboratory, Health and 

Environmental Sciences, The Dow Chemical Company, Midland, MI. Study No. K-000681-033. 

OTS Number 8EHQ-1189-0844 S. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Cal/EPA. May, 1993. California Environmental Protection Agency. Letter to Carol Browner, 

Administrator, U.S.EPA, on methyl bromide uses in California. 

HSDB. 1999. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Appendix D3 364 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Hurtt ME, Morgan KT, and Working PK. 1987. Histopathology of acute toxic responses in 

selected tissues from rats exposed by inhalation to methyl bromide. Fundam. Appl. Toxicol. 

9:352-365. 

Hurtt ME, and Working PK. 1988. Evaluation of spermatogenesis and sperm quality in the rat 

following acute inhalation exposure to methyl bromide. Fundam. Appl. Toxicol. 10(3):490-498. 

Irish DD, Adams EM, Spencer HC, and Rowe VK. 1940. The response attending exposure of 

laboratory animals to vapors of methyl bromide. J. Ind. Hyg. Toxicol. 22(6):218-230. 

Kato N, Morinobu S, and Ishizu S. 1986. Subacute inhalation experiment for methyl bromide in 

rats. Ind. Health. 24:87-103. 

NTP. 1990. National Toxicology Program. Toxicology and carcinogenesis studies of methyl 

bromide (CAS No. 74-83-9) in B6C3F1 mice (inhalation studies). NTP TR 385, NIH Publication 

No. 91-2840. 

Reuzel PGJ, Kuper CF, Dreef-van der Meulen HC, and Hollanders VMH. 1987. Chronic (29-

month) inhalation toxicity and carcinogenicity study of methyl bromide in rats. Report No. 

V86.469/221044. Netherlands Organization for Applied Scientific Research, Division for 

Nutrition and Food Research, TNO. EPA/OTS Document No. 86-8700001202. 

Reuzel PGJ, Dreef-van der Meulen HC, Hollanders VMH, Kuper CF, Feron VJ, and Van der 

Heijden CA. 1991. Chronic inhalation toxicity and carcinogenicity study of methyl bromide in 

Wistar rats. Food Chem. Toxicol. 29(1):31-39. 

Russo JM, Anger WK, Setzer JV, and Brightwell WS. 1984. Neurobehavioral assessment of 

chronic low-level methyl bromide exposure in the rabbit. J. Toxicol. Environ. Health. 14:247-

255. 

Sikov MR, Cannon WC, Carr DB, Miller RA, Montgomery LR, and Phelps DW. 1981. 

Teratologic assessment of butylene oxide, styrene oxide and methyl bromide. Battelle Pacific 

Northwest Laboratory, Richland, WA, for the National Institute for Occupational Safety and 

Health, Cincinnati, OH. 

U.S.EPA. 1999. Environmental Protection Agency. Integrated Risk Information System (IRIS) 

database. Reference concentration (RfC) for methyl bromide. 

Appendix D3 365 Methyl Bromide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY

METHYL CHLOROFORM 

(1,1,1-trichloroethane, methyltrichloromethane) 

CAS Registry Number: 71-55-6 

I. Chronic Toxicity Summary

Inhalation reference exposure level 1,000 μg/m3 (200 ppb)

Critical effect(s) Astrogliosis in the sensorimotor cortex (brain) of 

gerbils 

Hazard index target(s) Nervous system 

II. Chemical Property Summary (HSDB, 1999)

Description Colorless liquid 

Molecular formula C2H3Cl3

Molecular weight 133.42 g/mol 

Density 1.3376 g/cm3

 @ 20° C 

Boiling point 74.1° C 

Melting point -30.4° C 

Vapor pressure 127 torr @ 25° C 

Solubility Soluble in acetone, benzene, methanol, carbon 

tetrachloride 

Conversion factor 5.47 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Methyl chloroform is used as a solvent for adhesives and for metal degreasing (ACGIH, 1992). 

It is also used in the manufacture of vinylidene chloride and in textile processing and dry 

cleaning. The annual statewide industrial emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California, based on the most recent inventory, were estimated to be 

25,316,458 pounds of methyl chloroform (CARB, 1999a). Statewide monitored median and 

mean concentrations of methyl chloroform have been generally declining; decreasing from 0.8 or 

1.71 ppb in 1990 to 0.12 or 0.30 ppb in 1996 (CARB, 1999b). 

IV. Effects of Human Exposure 

A 44-year old woman was diagnosed with peripheral neuropathy following 18 months of 

occupational exposure to methyl chloroform in a solvent bath (House et al., 1994). There was no 

identified exposure to agents known to cause peripheral neuropathy, such as n-hexane or 

Appendix D3 366 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

trichloroethylene. The worker reported that she wore protective gloves and a respirator, both of 

which frequently leaked. Seven months following removal from exposure, the worker showed 

improved nerve conduction. 

Other case reports have identified the nervous system as a target of methyl chloroform toxicity in 

similar exposure scenarios. Three workers developed distal sensory neuropathy after working 

with methyl chloroform in a degreasing operation with repeated dermal exposure (Liss, 1988; 

Howse et al., 1989). Changes were observed in nerve conduction in the upper extremities 

accompanied by both axonopathy and myelonopathy. 

Twenty-eight workers with chronic exposure to high (but unquantified) concentrations of 1,1,1-

trichloroethane had significant deficits in memory, intermediate memory, rhythm, and speed 

based on the Luria-Nebraska Neuropsychological Battery (Kelafant et al., 1994). Deficits in 

vestibular, somatosensory, and ocular components of balance were noted. 

A 13-year-old male died after intentional inhalation of 1,1,1-trichloroethane (Winek et al., 1997). 

Autopsy findings included tissue congestion of lung, liver and kidney. 

Cardiac arrhythmia resulting from heightened cardiac sensitivity to epinephrine has been 

reported in several case reports of high acute inhalation exposures to methyl chloroform 

(ATSDR, 1990). There are case reports of arrhythmias persisting for two weeks or more after 

cessation of exposure to methyl chloroform (McLeod et al., 1987). 

An epidemiological study of workers chronically exposed to low levels of methyl chloroform 

(<250 ppm) found no changes in blood pressure, heart rate, or electrocardiogram (Kramer et al., 

1978). This study consisted of 151 workers who had been exposed for more than one year. No 

neurophysiological testing was done. 

Another study of 22 female workers exposed to methyl chloroform (plus 7 unexposed control 

workers) at concentrations ranging from 110-345 ppm in air for a mean of 6.7 years failed to 

identify neurotoxicity resulting from methyl chloroform exposure (Maroni et al., 1977). The 

examination included evaluation for neurologic symptoms, changes in nerve conduction, and 

psychomotor tests. 

Liver disease was observed in a worker exposed to methyl chloroform in a clothing factory 

screen printing room (Cohen and Frank, 1994). The worker was exposed for a total of 4 years 

before occupational exposure was identified as the cause of the liver disease. The worker 

sprayed an adhesive (containing 65% methyl chloroform, 25% propane and dimethyl ether, and 

10% inert ingredients) during which the worker reported often feeling dizzy or intoxicated. 

Three months following removal of the worker from exposure, liver function tests, although still 

abnormal, were significantly improved. Other case reports support these findings (Hodgson et 

al., 1989; Halevy et al., 1980). 

Six male volunteers were exposed to 35 and 350 ppm methyl chloroform for 6-hours on two 

separate occasions (Nolan et al., 1984). Absorption was determined to be 25% of the inhaled 

dose. Of the absorbed dose, 91% was excreted unchanged in the expired air. Although the odor 

Appendix D3 367 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

was perceptible for the duration of the exposure, no subjective symptoms were reported by the 

volunteers. 

V. Effects of Animal Exposure 

Gerbils (4/sex/dose plus 24 sex-matched control animals) were continuously exposed to 70, 210, 

or 1000 ppm methyl chloroform for 3 months (Rosengren et al., 1985). A 4-month (solvent-free) 

recovery period following exposure was included to evaluate “lasting or permanent changes.” 

Body weights were not changed significantly as a result of exposure. Brain weights in the 

animals in the 1000 ppm dose group were significantly decreased. Fibrillary astrocytes are 

formed in the brain in response to injury. Brain injury in methyl chloroform exposed gerbils was 

evaluated by detection of glial fibrillary acidic (GFA) protein, the main protein subunit of 

astroglial filaments. Increased levels of GFA protein were detected in the sensorimotor cerebral 

cortex of animals exposed to 210 or 1000 ppm methyl chloroform. 

A later study in gerbils examined the effects of a 3-month continuous exposure to 70 ppm methyl 

chloroform followed by a 4-month recovery period (Karlsson et al., 1987). DNA content was 

significantly decreased in three areas of the brain: posterior cerebellar hemisphere, anterior 

cerebellar vermis, and hippocampus. The authors contended that depressions in DNA content 

reflect decreased cell density. 

No evidence of peripheral neuropathy or other neurotoxicity was detected in rats exposed to 200, 

620, or 2000 ppm methyl chloroform 6 hours per day, 5 days per week for 13 weeks (Mattson et 

al., 1993). The study included a functional observational test battery and measured visual, 

somatosensory, auditory and caudal nerve-evoked potentials. Histopathology of the brain, spinal 

cord, peripheral nerves and limb muscles was also examined at the end of the 13-week exposure. 

Forty percent of all mice continuously exposed to 1000 ppm methyl chloroform for 14 weeks 

exhibited evidence of hepatocellular necrosis (McNutt et al., 1975). A statistically significant 

increase in liver weight per body mass was observed throughout the study. Electron microscopy 

revealed accumulation of triglyceride droplets in the centrilobular hepatocytes following one 

week of exposure to 1000 ppm methyl chloroform. After 4 weeks of exposure, cytoplasmic 

alterations in centrilobular hepatocytes included a loss of polyribosomes and increased smooth 

endoplasmic reticulum. Similar changes observed occasionally in hepatocytes from mice 

exposed to 250 ppm were not as dramatic. 

Mild hepatocellular changes were observed in rats exposed to 1500 ppm methyl chloroform 6 

hours per day, 5 days per week for 6, 12, and 18 months (Quast et al., 1988). At 24 months, 

these slight effects were no longer discernible due to confounding geriatric changes. No 

hepatocellular changes or other adverse effects were observed in rats exposed to 150 or 500 ppm 

methyl chloroform for up to 24 months. 

The developmental toxicity of inhaled methyl chloroform was studied in CD-1 mice. Mice were 

exposed on gestation days 12 through 17 to either 2000 ppm methyl chloroform for 17 hours per 

day or 8000 ppm methyl chloroform for 1 hour three times per day (Jones et al., 1996). There 

Appendix D3 368 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

were no effects on pregnancy outcome, but exposed pups has reduced weight gain, had poorer 

results on motor coordination tests and showed delays in negative geotaxis (orienting towards the 

top of a sloped screen). 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Rosengren et al. (1985) 

Study population Mongolian gerbils (4/sex/dose) 

Exposure method Whole-body inhalation exposure 

Critical effects Astrogliosis in the sensorimotor cerebral 

cortex 

LOAEL 210 ppm 

NOAEL 70 ppm 

Exposure continuity Continuous 

Average experimental exposure 70 ppm for NOAEL group 

Human equivalent concentration Not derived (species-specific data for 

gerbils unavailable to validate 

assumption of RGDR=1) 

Exposure duration 3 months 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.2 ppm (200 ppb; 1 mg/m3

; 1,000 

μg/m3

) 

VII. Data Strengths and Limitations for Development of the REL 

Case reports indicate that the nervous system and the liver are targets of the toxicity of methyl 

chloroform (House et al., 1994; Liss, 1988; Howse et al., 1989; Cohen and Frank, 1994). The 

largest of the epidemiological studies (Kramer et al., 1978; Maroni et al., 1977), however, did 

not identify adverse effects as a result of chronic methyl chloroform exposure. The Kramer et al.

(1978) study limited its evaluation to changes in blood pressure, heart rate, or electrocardiogram 

and exposure levels were only characterized as less than 250 ppm. Maroni et al. (1977) 

conducted their study among 22 women exposed occupationally to methyl chloroform levels as 

low as 110 ppm. Although the subjects were evaluated specifically for signs of neurotoxicity, 

the small sample size limits conclusions that can be drawn from their failure to identify adverse 

effects in this population. If no effects are associated with the exposures in the 2 studies (Kramer 

et al., 1978; Maroni et al., 1977), the REL predicted would be approximately 3 ppm. 

Data from animal studies generally support the findings of the case reports from human 

exposures. Both neurotoxicity and hepatotoxicity have been identified among animals exposed 

by inhalation to methyl chloroform. The adverse effect observed at the lowest level in these 

studies was the development of astrogliosis in the brains of gerbils exposed for 3 months to 210 

ppm methyl chloroform (Rosengren et al., 1985). A no-observed-adverse-effect-level (NOAEL) 

Appendix D3 369 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

in this study was 70 ppm methyl chloroform. A subsequent study identified a more subtle 

change in the brains of gerbils exposed similarly to 70 ppm methyl chloroform, with slightly 

decreased DNA content found in several discrete brain regions of exposed animals. However, 

the relationship between tissue DNA content and cell density as an indication of adverse effect in 

the brain was considered too tenuous for the development of a guidance value for chronic 

exposure to methyl chloroform. 

The major strengths of the REL for methyl chloroform are the observation of the NOAEL and 

the continuous subchronic exposure regimen. The major uncertainties are the lack of human 

exposure data, the lack of dose-response information, and the lack of comprehensive multi-organ 

effects data. 

VIII. References 

ACGIH. 1992. American Conference of Governmental Industrial Hygienists, Inc. 

Documentation of the Threshold Limit Values and Biological Exposure Indices. Sixth edition. 

Cincinnati, OH. 

ATSDR. 1990. Agency for Toxic Substances and Disease Registry. US Public Health Service. 

Toxicological Profile for 1,1,1-Trichloroethane. Prepared by Syracuse Research Corporation 

under contract to Clement Associates, Inc. under contract no. 205-88-0608. 

CARB (California Air Resources Board). 1999a. Air toxics emissions data collected in the Air 

Toxics Hot Spots Program CEIDARS Database as of January 29, 1999. 

CARB. 1999b. 1990-1996 Statewide Methyl Chloroform Summary, ppb. WWW document 

(http://www.arb.ca.gov/aqd/tcea/tastate.htm). 

Cohen C, and Frank AL. 1994. Liver disease following occupational exposure to 1,1,1-

trichloroethane: a case report. Am. J. Ind. Med. 26:237-241.Halevy J, Pitlik S, Rosenfeld J, and 

Eitan B-D. 1980. 1,1,1-Trichloroethane intoxication: a case report with transient liver and renal 

damage. Review of the literature. Clin. Toxicol. 16:467-72. 

Hodgson MJ, Heyl AE, and Van Thiel DH. 1989. Liver disease associated with exposure to 

1,1,1-trichloroethane. Arch. Intern. Med. 149:1793-1798. 

House RA, Liss GM, and Wills MC. 1994. Peripheral sensory neuropathy associated with 1,1,1-

trichloroethane. Arch. Environ. Health 49:196-199. 

Howse DC, Shanks GL, and Nag S. 1989. Peripheral neuropathy following prolonged exposure 

to methyl chloroform. Neurology, 39:242. (Abstract) 

HSDB. 1999. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Appendix D3 370 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Jones HE, Kunko PM, Robinson SE, Balster RL. 1996. Developmental consequences of 

intermittent and continuous prenatal exposure to 1,1,1-trichloroethane in mice. Pharmacol 

Biochem Behav 55(4):635-46. 

Karlsson J-E, Rosengren LE, Kjellstrand P, and Haglid KG. 1987. Effects of low-dose inhalation 

of three chlorinated aliphatic organic solvents on deoxyribonucleic acid in gerbil brain. Scand. J. 

Work Environ. Health 13:453-458. 

Kelafant GA, Berg RA, Schleenbaker R. 1994. Toxic encephalopathy due to 1,1,1-

trichloroethane exposure. Am J Ind Med 25(3):439-46. 

Kramer CG, Ott MG, Fulkerson JE, Hicks N, and Imbus HR. 1978. Health of workers exposed to 

1,1,1-trichloroethane: a matched pair study. Arch. Environ. Health 33:331-342. 

Liss G. 1988. Peripheral neuropathy in two workers exposed to 1,1,1-trichloroethane. JAMA 

260(15):2217. 

Maroni M, Bulgheroni C, Cassitto G, Merluzzi F, Gilioli R, and Foa V. 1977. A clinical, 

neurophysiological and behavioral study of female workers exposed to 1,1,1-trichloroethane. 

Scand. J. Work Environ. Health 3:16-22. 

Mattson JL, Albee RR, Lomax LG, Beekman MJ, and Spencer PJ. 1993. Neurotoxicologic 

examination of rats exposed to 1,1,1-trichlorethane vapor for 13 weeks. Neurotoxicol. Teratol. 

15:313-326. 

McLeod AA, Marjot R, and Monaghan MJ. 1987. Chronic cardiac toxicity after inhalation of 

1,1,1-trichloroethane. Br. Med. J. 294:727-729. 

McNutt NS, Amster RL, McConnell EE, and Morris F. 1975. Hepatic lesions in mice after 

continuous inhalation exposure to 1,1,1-trichloroethane. Lab. Invest. 32:642-654. 

Nolan RJ, Freshour NL, Rick DL, McCarty LP, and Saunders JH. 1984. Kinetics and metabolism 

of inhaled methyl chloroform (1,1,1-trichloroethane) in human volunteers. Fundam. Appl. 

Toxicol. 4:654-662. 

Quast JF, Calhoun LL, and Frauson LE. 1988. 1,1,1-Trichloroethane formulation: a chronic 

inhalation toxicity and oncogenicity study in Fischer 344 rats and B6C3F1 mice. Fundam. Appl. 

Toxicol. 11:611-625. 

Rosengren LE, Kjellstrand AA, and Haglid KG. 1985. Astrogliosis in the cerebral cortex of 

gerbils after long-term exposure to 1,1,1-trichloroethane. Scand. J. Work Environ. Health 

11:447-455. 

 Winek CL, Wahba WW, Huston R, Rozin L. 1997. Fatal inhalation of 1,1,1-trichloroethane.. 

Forensic Sci Int 87(2):161-165. 

Appendix D3 371 Methyl Chloroform 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

METHYL ISOCYANATE 

(MIC, CH3-N=C=O) 

CAS Registry Number: 624-83-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 1 μg/m3 (0.5 ppb)

Critical effects(s) Decreased weight gain and lung pathology at 

cessation of exposure in rats 

Hazard index target(s) Respiratory system; reproductive system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994) 

Description Colorless liquid 

Molecular formula C2H3NO 

Molecular weight 57.06 g/mol 

Boiling point 39.5o

C 

Melting point −45o

C 

Vapor pressure 348 torr @ 20o

C, 600 torr @ 30oC (Varma and 

Guest, 1993) 

Solubility 10 percent in water @ 15o

C 

Conversion factor 2.3 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources (Dave, 1985; U.S. EPA, 1986; HSDB, 1995) 

Methylisocyanate (MIC) is prepared industrially by reacting methylamine with phosgene, 

oxidizing monomethylformamide at high temperatures (> 550o

C), or heating metal 

methylisocyanates. Because of its high reactivity, MIC is used as an intermediate in organic 

synthesis, most notably in the production of carbamate based pesticides. Tobacco smoke from 

some brands of cigarettes also contains MIC (about 4 μg per cigarette). Workers exposed to the 

MIC 8-hour threshold limit value of 0.02 ppm (46 μg/m3

) are exposed to approximately 460 μg 

MIC in a workday. Based on the most recent inventory, the annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California were 

negligible (CARB, 2000). This does not include estimates of emissions of breakdown products 

from the use of metam sodium in agricultural applications. Use of metam sodium averaged 

15,400,000 pounds/year from 1995 to 1999. 

Appendix D3 372 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IV. Effects of Human Exposure 

Although occupational exposures to MIC have been documented (Varma, 1986), few known 

exposures to the general public have occurred. A major exposure occurred in Bhopal, India in 

December 1984. Because of the sudden, short-term release (30-45 minutes), no measurements 

occurred, but the air concentration was estimated as 13 ppm (Dave, 1985) to 100 ppm (Varma, 

1986). 

The chemical identity of the ultimate toxicant has not been unequivocally determined and may 

consist of more than one chemical species. Although the chemistry of MIC suggests that 

hydrolysis to methylamine and dimethylurea is rapid, such hydrolysis in moist air is probably 

slow, and the reaction with photochemically produced hydroxyl radical is also slow (chemical 

T1/2 about 3 months) (U.S. EPA, 1986). Brown et al. (1987) have shown that the 

alkylisocyanates (e.g., MIC) are relatively resistant (compared to the arylisocyanates) to 

hydrolysis in water. Hence, despite the high water reactivity of MIC, this compound could 

possibly persist in the environment for many days after an initial release. 

Within 5 days of the initial exposure to MIC at Bhopal, more than 2,000 deaths occurred (Dave, 

1985), while 4,000 more deaths were documented during the following decade (Lepkowski, 

1994). The initial symptoms among the population living near the MIC plant were irritation and 

difficulty in breathing (Varma, 1986). Blindness occurred in more than 10,000 exposed persons 

but later resolved in most cases (Andersson et al., 1990). The acute damage that led to death was 

mainly to the respiratory system, most likely pulmonary edema, bronchospasm, and electrolyte 

imbalance (Varma, 1986). However extrapulmonary damage, including tissue anoxia, 

gastrointestinal symptoms, and muscular weakness, were also observed (Dave, 1985). Within a 

year of the exposure, survivors continued to exhibit damage to the lung and eyes. Fibrosis of the 

lungs was seen in 30 percent of this group (Dave, 1985). 

Reproductive toxicity was observed among women exposed to MIC in Bhopal. Varma (1987) 

reported 43 percent unsuccessful pregnancies among 865 women who were pregnant at the time 

of the MIC release. Among the live births, 14 percent of the infants died within 30 days, 

whereas a death rate of only 3 percent for the same interval was recorded 2 years prior to the 

release. Bhandari et al. (1990) reported increased spontaneous abortions and neonatal deaths 

among exposed women who were pregnant at the time of exposure compared to a control group 

in another city. In the latter study, stillbirths and congenital malformations were similar in the 

exposed and non-exposed groups. 

Non-reproductive, non-pulmonary responses were evident in a group of exposed Bhopal 

residents, 3-years following exposure to the MIC vapors. Loss of vision and loss of visual acuity 

were more prominent among exposed residents than among unexposed people, and the losses 

appeared to be dose-dependent (Andersson et al., 1990). In this study, the surrogate for dose was 

extent of early deaths in a housing cluster. Similarly, cataracts were reported more often among 

the exposed than among the unexposed group. 

The lesions associated with lung damage may be expressed as pulmonary edema for immediate 

effects (Varma, 1986), and lesions associated with the bronchoalveolar area for long-term effects 

Appendix D3 373 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

(Dave, 1985, Varma, 1986). Vijayan et al. (1995) studied cellular components of 

bronchoalveolar lavage (BAL) and pulmonary function in Bhopal patients 1.3, 2.7, and 5.1 years 

after exposure to MIC. All had lived within 3-miles of the factory and all experienced acute 

respiratory and ophthalmic symptoms on the day of exposure. All were experiencing continued 

respiratory symptoms. Among the exposed people, decrements in forced vital capacity and 

forced expiratory volume (at 1-minute) were observed. In general, the decrements ranged from 

12 - 21 percent of predicted values, whereas the control group exhibited decrements of 2 - 4 

percent of the expected values. Analysis of the BAL revealed increases in total cells (all exposed 

groups), increased absolute numbers of macrophages (all exposure groups), decreased percentage 

of lymphocytes (2.7 and 5.1 year groups), and increased numbers and percentage of neutrophils 

(5.1 year group). These cell types are involved, through the secretion of various factors, in 

inflammatory and immunologic processes in the lung (Reiser and Last, 1986). The Vijayan et al.

(1995) study thus suggests long term damage to lung parenchyma among people who survived 

the initial acute effects of MIC exposure. 

In summary, humans exposed acutely by inhalation to MIC may experience long-term (as well as 

immediate) damage to pulmonary and extrapulmonary systems. The lung is probably the critical 

target organ for long-term effects from acute exposure, although adverse effects on other organs 

(e.g., eye, reproductive, and gastrointestinal) also exist. The late responses to the acute exposure 

suggest an immunological component, which could involve several systems including lung, eye, 

liver, and kidney. The chemical identity of the ultimate toxicant is unknown and may be more 

than one compound. 

Avashia et al. (1996) assessed pulmonary effects from long-term, low-level MIC for more than 

400 workers at a large chemical facility. Serial pulmonary function data, cigarette smoking 

histories, and industrial-hygiene measurements were available. Jobs were classified according to 

level of MIC exposure as none, low, moderate or high. Where work records were incomplete, 

exposures were based on the ratings of supervisors and coworkers. The frequency of pulmonary 

impairment was evaluated for the assumed four levels of exposure. No specific or consistent 

pulmonary impairment was evident. Unfortunately the report gave no quantitative classification 

of low, moderate or high MIC levels. 

V. Effects of Animal Exposure 

Experimental animal studies have been designed to address the experiences of the victims of the 

Bhopal disaster, in which the exposure has been described as acute because of the short duration 

(30-45 min). No studies were found that described exposure duration greater than 10 days. 

However, a chronic component to MIC exposure may exist as a result of slower rates of 

hydrolysis in air (compared to water), the presence of carbamylated hemoglobin in MIC-exposed 

people, and the change from edematous to inflammatory and/or fibrotic lesions with time. 

Further, a glutathione-dependent reversible MIC transport system has been suggested in 

experimental animals (see below). 

MIC is absorbed through the respiratory tract and distributed to non-respiratory organs in 

experimental animals. In an acute (30 min) inhalation exposure to a dose of 14C-MIC (labeled in 

Appendix D3 374 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

the isocyanate moiety) equivalent to one-LC50 (23 mg/L), rats accumulated protein-bound 

radioactivity (including carbamylated proteins) in brain, liver, kidney, and lung, but not in blood 

(Bhattacharya et al., 1988). Ferguson et al. (1988) exposed guinea pigs by inhalation to 0.47 

ppm 

14C-MIC (methyl group) for 6-hours. At the end of exposure, the label was found in arterial 

and venous blood, bile, and urine. At 2.7 days post-exposure, the label decreased to 2-7 percent. 

MIC was retained in the nasal-laryngeal area of the guinea pigs. 

MIC, like reactive isocyanates in general, can react with biological molecules containing amino, 

alcohol, or sulfhydryl groups, as well as with water. While hydrolysis in an aqueous 

environment, such as the lung, is theoretically possible, measurements show that alkyl 

isocyanates are relatively resistant (compared to arylisocyanates) to such hydrolysis (Brown et 

al. 1987). The absence of a role for MIC hydrolytic products, methylamine (MA) or 

dimethylurea (DMU), is also suggested by the work of Jeevaratnam and Sriramachari (1994) and 

Sriramachari et al. (1994). Inhalation (30 min) or subcutaneous exposure of rats to either 

hydrolytic product at levels equivalent to the LC50 or LD50 did not result in death. Similarly, 

neither methylamine nor dimethylurea duplicated the acute effects of respiratory necrosis and 

congestion. However, exposure to these hydrolytic products did lead to interstitial pneumonitis, 

an observation that suggests MA and/or DMU could lead to subsequent inflammatory responses 

if sufficient amounts are present. 

A role for methylamine in reproductive/developmental toxicity was investigated by Guest and 

Varma (1991). In a mouse study, pregnant dams were exposed to varying doses (intraperitoneal) 

of methylamine (as well as the di- and trimethyl compounds). Reproductive toxicity was not 

observed for methylamines. However, in cultured embryo experiments, decrements in crownrump length, yolk-sac diameter, head length, and embryo survival were observed. The 

concentrations were high (>0.75 mM) and the interpretation of the biological activity of 

methylamine in terms of inhalation exposure is difficult. 

MIC is a carbamylating intermediate; this is the basis for its use in the manufacture of carbamate 

based pesticides. In the same way, MIC should react with the appropriate functional groups of 

proteins, peptides, and nucleic acids. However, in vitro studies with cholinesterases show that 

such a reaction is not efficient (Brown et al., 1987), an observation which may be explained by 

the presence of protonated amino groups at physiological pH (Baillie and Slatter, 1991). 

A transport system for MIC via reduced glutathione (GSH) has been suggested by the discovery 

of the MIC-adduct, S-(N-methylcarbamoyl)glutathione (SMG), in the bile and the MIC-adduct 

of N-acetylcysteine (mercapturic acid, AMCC) in the urine of rats exposed to MIC by noninhalation routes (Pearson et al., 1990; Slatter et al., 1991). The reaction of MIC with GSH and 

with cysteine is reversible, and can provide a source of free MIC in the tissues (Baille and 

Slatter, 1991). Similar studies in experimental animals exposed to MIC by inhalation have not 

been reported. However, humans exposed by inhalation to N,N-dimethylformamide (H-C-(=O)-

N(CH3)2) excrete AMCC in urine (Mraz and Nohova, 1992). Hence a reversible MIC-transport 

system in animals, including humans, is possible, and the presence of high levels of GSH in 

human lavage fluid (Cantin et al., 1987) would permit the initiation of this mechanism. 

Appendix D3 375 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

The toxicity of the adduct SMG was tested in mouse embryo culture (Guest et al., 1992). Mouse 

embryos, at day 8 of gestation in vivo, were removed from their dams and cultured in the 

presence (and absence) of SMG. Dose-dependent (0.25 - 2 mM) decrements were observed for 

yolk sac diameter, crown-rump length, somite number, and protein content. Delayed DNA 

synthesis in the embryos and in yolk-sacs occurred in the presence of 0.25 mM SMG. Similar to 

the results obtained with methylamine, the SMG concentrations were high and the exposures 

were not by inhalation. However, the data show that a MIC metabolite, SMG, has toxic 

properties. In the presence of GSH (1 or 3 mM), the extent of the SMG-dependent toxicities was 

decreased. Such data demonstrate the reversibility of the binding between MIC and GSH. 

Three inhalation studies were identified in which experimental animals were exposed to more 

than one dose of MIC. Among these studies, two used exposure durations for more than one day 

(Dodd and Fowler, 1986; Mitsumori et al., 1987). Rats and mice were exposed by inhalation to 

0, 1.1, and 2.8 (female) or 3.0 (male) ppm MIC for 6 hr/day for a total of 4 days, and then 

followed during a 91-day post-exposure interval (Mitsumori et al., 1987). Among the rats, postexposure deaths occurred by 49 days (male) and 14 days (female) at the high dose. Among the 

mice, only 1 male mouse died at 16 days post-exposure. Reduced weight gain was observed 

among the female and male rats in the high dose group, prior to death, although the absolute 

weights were not different from the unexposed rats one day before the end of exposure. Among 

the mice, a slowed weight gain was observed at 3- and 6-days post exposure (male) and 1 day 

post exposure (female) at the high dose, but normal weight gain returned by 1 week following 

cessation of exposure. At 7 days post-exposure, microscopic changes were observed in the 

respiratory system among the high dose rats of both sexes. Between 8- and 27 days postexposure, increased lesions in the respiratory tract and also in liver, thymus, spleen, heart, and 

brain were observed at the high dose. Similar lesions were not observed in rats exposed to 1.1 

ppm MIC and followed to the 8-27 day post-exposure. Among survivors, the incidence of 

lesions decreased to control values by 91 days. Among the mice, treatment related changes in 

the respiratory tract were observed at the high dose at 7 days post-exposure. Between 28 and 91 

days, the lesions associated with the upper respiratory tract disappeared, whereas those 

associated with the major bronchi remained, although somewhat attenuated. These data suggest 

that the rat is more sensitive than the mouse to the effects of MIC. A LOAEL of 2.9 ppm is 

indicated, based on post-exposure decreased weight gain and respiratory tract changes in rats. 

Dodd and Fowler (1986) exposed rats to 0, 0.15, 0.6, and 3.1 ppm MIC for two 4-day sessions at 

6-hours/day and examined the animals within 1-day following exposure. The 2-cycle exposure 

included a 2-day recess from exposure. No deaths occurred at any MIC concentration during the 

exposure. Lesser weight gain occurred for rats in the 3.1 ppm groups, whereas weight among the 

rats in the 0.15 and 0.6 ppm MIC groups was indistinguishable from the air-exposed control 

animals. On exposure days 3 and 8, mean food consumption values in the high dose group were 

below those for the non-exposed group. At the time of termination, male rats exposed to 3.1 

ppm MIC exhibited a 38 percent increase in hemoglobin concentration and a 26 percent decrease 

(p<0.001) in oxygen saturation, compared to the unexposed rats (p<0.001). Such changes were 

not observed for the female rats exposed to 3.1 ppm or for rats of either sex exposed to 0.15 or 

0.6 ppm MIC. Absolute lung weights increased (p<0.001) in both sexes after exposure to 3.1 

ppm, compared to the control rats. Decreases in liver, kidney and testes absolute weights were 

observed in this exposure group, but the authors interpreted these data as a reflection of the body 

Appendix D3 376 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

weight losses. No weight changes were observed in rats exposed to 0.15 or 0.60 ppm MIC. 

Gross and microscopic lesions were observed in rats (female and male) exposed to 3.1 ppm, but 

not in rats exposed to 0, 0.15, or 0.6 ppm MIC. The microscopic lesions occurred in the 

respiratory tract and consisted of inflammation, epithelial necrosis, squamous metaplasia, and 

epithelial hyperplasia. These lesions extended into the bronchioles. These data suggest a 

NOAEL of 0.6 ppm MIC, based on weight gain loss, absolute lung weight, and lung 

histopathology in rats, immediately following cessation of exposure. 

Post-exposure changes in lung pathology also occurred in the rats surviving 3.1 ppm in the Dodd 

and Fowler (1986) study. The early lesions associated with inflammation, epithelial necrosis, 

squamous metaplasia, and epithelial hyperplasia extending to the bronchioles either decreased in 

severity or receded toward the upper respiratory tract by 85-days post-exposure. In males, the 

intraluminal and submucosal fibroplasia changed in appearance during this interval, due in part 

to the maturation of fibrous tissue. Mucous plugs were also seen in the terminal bronchioles and 

alveoli in some rats. The importance of this observation is the progressive character of MIC 

induced lung disease. Such progression may be difficult to follow at lower doses, if the times 

involved are of insufficient duration. 

Sethi et al. (1989) exposed rats by inhalation to 0, 0.21, 0.26, and 0.35 ppm MIC for 6 days at 

0.5 hr/day. Statistical evaluation was not presented. No post-exposure deaths were reported, 

although lethality was recorded for rats exposed to 3.5 and 35 ppm for only 10 minutes. 

Following the 0.5 hr × 6-day exposure, the weight gain declined in proportion to the exposure 

dose. At the lowest dose (0.21 ppm) the weight gain was 111 g after 91 days post-exposure, 

compared to a weight gain of 218 g during the same interval among the non-exposed rats. The 

absolute weights of the rats at the end of the exposure were not given. According to the 

narrative, inflammatory lesions of bronchopulmonary tissue were present; their extent increased 

with dose. A dose-response increase in markers of lung infection was present and suggests that 

the MIC exposed rats were more prone to infectious agents than were the unexposed animals. 

Non-specific lesions in liver and kidneys were also observed and appeared to be dose dependent, 

but the authors suggested that these effects could be a result of the lung infections. 

Fetotoxicity was observed in two experimental animal studies (Schwetz et al., 1987; Varma, 

1987). Among female mice exposed to 0, 1, or 3 ppm MIC during gestation days 14 - 17 for 6 

hr/day, an increased incidence of fetal deaths was observed at 1 ppm (Schwetz et al., 1987). At 3 

ppm, the average number of pups/litter decreased relative to the air-exposed controls. The dams 

were unaffected in terms of survival, body weight, or length of gestation. Non-gestational 

exposure (6 hr/day, 4 days) did not affect the number of pregnancies or the live litter sizes, 

suggesting that the fetotoxic effect may be specific to the female reproductive tract rather than a 

general attribute of systemic toxicity. Similarly, female mice exposed for 3 hours on gestation 

day 8 to 0, 2, 6, 9 or 15 ppm MIC gave birth to pups with decreased body weights at the lowest 

dose, although a good dose-response was not observed (Varma, 1987). At 9 or 15 ppm MIC, the 

surviving dams lost 75 - 80 percent of their fetuses. Maternal mortality and decreased skeletal 

lengths were also observed at 9 and 15 ppm. A distinction between maternally induced 

fetotoxicity and a direct effect on fetal health could not be made. Because the inhalation 

exposure to the dams occurred for only 3 hrs on one day, a chronic LOAEL is not suggested. 

Exposure of male rats to one dose of 3.2 mg/L for 8 minutes resulted in a 21 percent fertility rate 

Appendix D3 377 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

among the cohabited female rats within the day 8-14 period post-exposure compared to a fertility 

rate of 40% for controls; however, the rates increased after 15 days post-exposure (Agarwal and 

Bose, 1992). There was no evidence of fetotoxicity among the dams impregnated by the MICexposed male rats. Exposure of male and female mice to 0, 1, or 3 ppm MIC did not result in 

altered body weights, fertility, or litter size (Schwetz et al, 1987). The results suggest that 

exposures to MIC at doses that are not toxic to adult male or female (pregestational) mice or rats 

do not result in adverse reproductive outcomes. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Dodd and Fowler (1986) 

Study populations F344 rats 

Exposure method Inhalation (0, 0.15, 0.6, or 3.1 ppm) 

Critical effects Decreased weight gain and lung pathology 

immediately after cessation of exposure 

LOAEL 3.1 ppm 

NOAEL 0.6 ppm 

Exposure continuity 6 hours/day, 8 days/10 day experiment (2-cycles, 

with one 2-day recess from exposure) 

Exposure duration 10 days 

Average experimental exposure 0.12 ppm for the NOAEL group 

 (0.6 x 8/10 x 6/24) 

Human equivalent factor 0.15 ppm for the NOAEL group (gas with 

pulmonary respiratory effects, RGDR = 1.23, 

based on BW = 152 g, MV = 0.12 L/min, SA 

= 225 cm2) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.5 ppb (1 μg/m3

) 

Although the exposure was for only 10 days, the Dodd and Fowler (1986) study includes the 

longest exposure duration of the available investigations and also uses some of the lower 

exposure levels (down to 0.15 ppm). The microscopic findings of the respiratory tract were 

statistically analyzed, although an observation of the tabulated data at the four doses (0, 0.15, 

0.6, and 3.1 ppm) clearly shows a NOAEL of 0.6 ppm. Other endpoints with the same NOAEL 

were increased hemoglobin and increased absolute lung weights. The symptomatic ramifications 

of the increased hemoglobin are unknown, although similar increases were reported for humans 

exposed to MIC in Bhopal (Srivastava et al., 1988). The lung weight gain may be a reflection of 

the pathological changes seen in the microscopic studies. 

Decreased body weight gain was also seen in the experimental 4 day rat inhalation study of 

Mitsumori et al. (1987) (NOAEL = 1.1 ppm), except that the decrease in the latter study did not 

Appendix D3 378 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

occur until 1 and 3 days (female and male, respectively) post-exposure. The apparent 

discrepancy could be explained, in part, on the basis of the length of exposure, which was twice 

as long in the Dodd and Fowler (1986) study. However, the weight gain loss in the Dodd and 

Fowler (1986) study was initiated within one day of the start of exposure. 

The MIC chronic REL of 0.5 ppb is based on endpoints observed within 1 day of cessation of 

exposure. Post-exposure evaluation showed that, at a higher exposure level (3.1 ppm), 

progressive changes, including death, occurred. Post-exposure observations, however, were not 

reported at the 0.15 and 0.6 ppm MIC levels. The attribute of delayed MIC inhalation toxicity 

has also been observed in other experimental animals studies (Dodd and Fowler, 1986, 

Mitsumori et al., 1987). In the case of the human MIC exposure in Bhopal, India, death did not 

occur during the immediate 30 - 45 minute exposure, but exhibited a lag phase. A few deaths 

occurred during the first few hours, the maximum occurred at 2 - 3 days, and by the end of a 

week about 2500 deaths were documented (Dave, 1985; Varma, 1986; Varma and Guest , 1993), 

although Varma (1986) suggests that the immediate number may be closer to 5,000. One report 

suggests that during the intervening decade as many as 6,000 deaths may be attributed to the 

initial exposure in Bhopal (Lepkowski, 1994). Such information suggests that the presence of an 

adverse effect at the NOAEL of 0.6 ppm (Dodd and Fowler, 1986) might be possible if the rats 

were observed during an extended post-exposure interval. Experimental evidence is needed to 

test this hypothesis. 

Only one study was identified in which post-exposure observations were made on experimental 

animals exposed subchronically by inhalation to multiple doses of MIC. Mitsumori et al. (1987) 

exposed rats to 0, 1.1, and 2.8 (females) or 3.0 (males) ppm MIC for 6 hr/day for 4 days and 

observed the rats for 91 days. No deaths and no weight gain loss (in contrast to Dodd and 

Fowler, 1986) were present until the post-exposure period and were mainly observed in animals 

exposed at the high dose. Using a NOAEL of 1.1 ppm MIC, a chronic REL of 1.1 ppb (2.6 

μg/m3

) was derived. The REL based on the Mitsumori et al. (1987) study is similar to the REL 

based on immediate effects (Dodd and Fowler, 1986), and may indicate that the time of 

occurrence of exposure related effects may not be as important as the MIC air concentration. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for MIC include the availability of controlled exposure 

inhalation studies in multiple species at multiple exposure concentrations and with adequate 

histopathogical analysis and the observation of a NOAEL. Major areas of uncertainty are the 

lack of adequate human exposure data and the lack of chronic inhalation exposure studies. 

VIII. Potential for Differential Impacts on Children's Health

Since exposures to MIC at levels that are not toxic to adult male or female (pregestational) mice 

or rats do not result in adverse reproductive outcomes, the chronic REL of 1 μg/m3

 should 

adequately protect infants and children. MIC is a respiratory irritant and the developing 

respiratory system is more sensitive than that of adults. However, there is no direct evidence in 

the literature to quantify a differential effect of MIC on the respiratory system of infants and 

children. 

Appendix D3 379 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IX. References 

Agarwal DK, and Bose M. 1992. Inhalation toxicity of methylisocyanate: assessment of germ 

cell mutagenicity and reproductive effects in rats. Indian J. Exp. Biol. 30:504-508. 

Andersson N, Ajwani MK, Mahashabde S, Tiwari MK, Muir MK, Mehra V, Ashiru K, and 

Mackenzie DC. 1990. Delayed eye and other consequences from exposure to methyl isocyanate: 

93% follow up of exposed and unexposed cohorts in Bhopal. Br. J. Ind. Med. 47:553-558. 

Avashia B, Battigelli MC, Morgan WK and Reger RB. 1996. Effects of prolonged low exposure 

to methyl isocyanate. J. Occup. Environ. Med. 38(6):625-630. 

Baillie TA, and Slatter JG. 1991. Glutathione: a vehicle for the transport of chemically reactive 

metabolites in vivo. Acc. Chem. Res. 24:264-270. 

Bhandari NR, Syal AK, Kambo I, Nair A, Beohar V, Saxena NC, Dabke AT, Agarwal SS, and 

Saxena BN. 1990. Pregnancy outcome in women exposed to toxic gas at Bhopal. Indian J. Med. 

Res [B]. February. pp. 28-33. 

Bhattacharya BK, Sharma SK, and Jaiswal DK. 1988. In vivo binding of 

[1-14C]methylisocyanate to various tissue proteins. Biochem. Pharmacol. 37:2489-2493. 

Brown WE, Green AH, Cedel TE, and Cairns J. 1987. Biochemistry of protein-isocyanate 

interactions: a comparison of the effects of aryl vs. alkyl isocyanates. Environ. Health Perspect. 

72:5-11. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Cantin AM, North SL, Hubbard RC, and Crystal RG. 1987. Normal alveolar epithelial lining 

fluid contains high levels of glutathione. J. Appl. Physiol. 63:152-157. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Dave JM. 1985. The Bhopal methyl isocyanate (MIC) incident: an overview. In: Proceedings of 

an International Symposium, Highly Toxic Chemicals: Detection and Protection Methods. 

Schiefer, H.B. ed. Saskatoon, Saskatchewan, Canada. pp 1-38. 

Dodd DE, and Fowler EH. 1986. Methyl isocyanate subchronic vapor inhalation studies with 

Fischer 344 rats. Fundam. Appl. Toxicol. 7:502-522. 

Ferguson JS, Kennedy AL, Stock MF, Brown WE, and Alarie Y. 1988. Uptake and distribution 

of 14C during and following exposure to [14C} methyl isocyanate. Toxicol. Appl. Pharmocol. 

94:104-117. 

Appendix D3 380 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Guest I, Baille TA, and Varma DR. 1992. Toxicity of the methyl isocyanate metabolite S-(nmethylcarbamoyl)GSH on mouse embryos in culture. Teratology 46:61-67. 

Guest I, and Varma DR. 1991. Developmental toxicity of methylamines in mice. J. Toxicol. 

Environ. Health. 32:319-330. 

HSDB. 1995. Hazardous Substances Data Bank. TOMES® Vol. 25. Denver, CO: Micromedex, 

Inc. Expires 07/31/95. 

Jeevaratnam K, and Sriramachari S. 1994. Comparative toxicity of methyl isocyanate and its 

hydrolytic derivatives in rats. I. Pulmonary histopathology in the acute phase. Arch. Toxicol. 

69:39-44. 

Lepkowski W. 1994. Ten years later - Bhopal. Chemical and Engineering News. 19 December 

1994. 

Mitsumori K, Boorman GA, Gupta BN, and Bucher JR. 1987. Four-day repeated inhalation and 

recovery study of methyl isocyanate in F344 rats and B6C3F1 mice. Fundam. Appl. Toxicol. 

9:480-495. 

Mraz J, and Nohova H. 1992. Absorption, metabolism and elimination of N,Ndimethylformamide in humans. Int. Arch. Occup. Environ. Health. 64:85-92. 

Pearson PG, Slatter JG, Rashed MS, Han D-H, Grillo MP, and Baillie TA. 1990. S-(Nmethylcarbamoyl)glutathione: a reactive S-linked metabolite of methyl isocyanate. Biochem. 

Biophys. Res. Comm. 166:245-250. 

Reiser KM, and Last JA. 1986. Early cellular events in pulmonary fibrosis. Exp. Lung Res. 

10:331-355. 

Schwetz BA, Adkins B Jr, Harris M, Moorman M, and Sloane R. 1987. Methyl isocyanate: 

reproductive and developmental toxicology studies in mice. Environ. Health Perspect. 72:149-

152. 

Sethi N, Dayal R, and Singh RK. 1989. Acute and subacute toxicity study of inhaled methyl 

isocyanate in Charles Foster rats. Ecotoxicol. Environ. Saf. 18:68-74. 

Slatter JG, Rashed MS, Pearson PG, Han D-H, and Baillie TA. 1991. Biotransformation of 

methyl isocyanate in the rat. Evidence of glutathione conjugation as a major pathway of 

metabolism and implications for isocyanate-mediated toxicities. Chem. Res. Toxicol. 4:157-161. 

Sriramachari S, Rao GJ, Sharma VK, Jadhav RK, Saraf AK, and Chandra H. 1991. GC-NPD and 

GC-MS analysis of preserved tissue of Bhopal gas disaster: evidence of methyl carbamylation in 

post-mortem blood. Med. Sci. Law. 31:289-293. 

Appendix D3 381 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Srivastava RC, Gupta BN, Athar M, Behari JR, Dwivedi RS, Hasan SK, Bharti RS, Singh A, 

Misra M, and Ray PK. 1988. Effect of exposure to toxic gas on the population of Bhopal: Part III 

- Assessment of toxic manifestations in humans - hematological and biochemical studies. Ind. J. 

Exp. Biol. 26:165-172. 

U.S. EPA. 1986. United States Environmental Protection Agency. Health and environmental 

effects profile for methyl isocyanate. EPA/600/22. Environmental Criteria and Assessment 

Office, Office of Research and Development. Cincinnati OH: U.S. EPA. 

Varma DR. 1986. Anatomy of the methyl isocyanate leak on Bhopal. In: Hazard Assessment of 

Chemicals. Saxena J, ed. New York: Hemisphere Publishing Corp. pp. 233-299. 

Varma DR. 1987. Epidemiological and experimental studies on the effects of methyl isocyanate 

on the course of pregnancy. Environ. Health Perspect. 72:153-157. 

Varma DR, and Guest I. 1993. The Bhopal accident and methyl isocyanate toxicity. J. Toxicol. 

Environ. Health 40:513-529. 

Vijayan VK, Sankaran K, Sharma SK, and Misra NP. 1995. Chronic lung inflammation in 

victims of toxic gas leak at Bhopal. Resp. Med. 89:105-111 

Appendix D3 382 Methyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

METHYL t-BUTYL ETHER 

(MTBE; 2-methoxy-2-methylpropane; tert-butyl methyl ether; 

methyl 1,1dimethyl ether) 

CAS Registry Number: 1634-04-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 8000 μg/m3

 (2000 ppb) 

Critical effect(s) Nephrotoxicity, prostration, periocular swelling 

in Fischer 344 rats 

Hazard index target(s) Kidney; eyes; alimentary system 

II. Physical and Chemical Properties (HSDB, 1994) 

Description Colorless liquid 

Molecular formula C5Hl20 

Molecular weight 88.15 g/mol 

Density 0.7405 g/cm3

 @ 20°C 

Boiling point 55.2°C @ 760 mm Hg 

Vapor pressure 245 torr @ 20°C 

Solubility Soluble in alcohol, ether, and 5% soluble in 

water 

Conversion factor 1 ppm = 3.61 mg/m3

 @ 25° C; 

 3.67 mg/m3

 @ 20° C 

III. Major Uses or Sources 

Methyl t-butyl ether (MTBE) is used as a gasoline additive to improve octane ratings and reduce 

emissions of some pollutants, in industry to improve miscibility of solvents, and in clinical 

medicine to dissolve cholesterol gall stones (Yoshikawa et al., 1994). The annual statewide 

industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California,based on the most recent inventory, were estimated to be 215,182 pounds of MTBE 

(CARB, 1999). 

IV. Effects of Human Exposure 

Gasoline (with 10% MTBE) tanker drivers reported significantly higher fatigue at the end of the 

work week than before the work week, and those with longer exposure to gasoline with MTBE 

during the work week reported significantly higher fatigue than drivers with shorter exposure 

(Hakkola et al., 1997). 20% of drivers reported symptoms such as headache, dizziness, nausea, 

Appendix D3 383 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

and dyspnoea at the end of work week. No human chronic toxicity or chronic epidemiology 

information for MTBE without coexposure to gasoline was found. 

Ten healthy male volunteers undergoing light physical work were exposed to 5, 25, and 50 ppm 

MTBE vapor for 2 hours (Nihlen et al., 1998). While a solvent smell was noted at these 

concentrations, there were no consistent concentration-related effects on reported ocular or nasal 

irritation. The blockage index (a measure of nasal airway resistance) increased significantly after 

exposure but was not correlated with exposure concentration. 

V. Effects of Animal Exposure 

Male and female rats (50/sex/group) were exposed by inhalation for 6 hours/day, 5 days/week to 

mean concentrations of 0, 403, 3023, or 7977 ppm (0, 1453, 10,900, or 28,760 mg/m3

) MTBE 

for 24 months (Chun et al., 1992). Clinical signs, hematology, body weights and food 

consumption were monitored. Necropsy included measurements of organ weights and 

histopathology. Corticosterone levels were measured on 10 animals prior to sacrifice. Serum 

enzymes were not monitored. The NOAEL for several endpoints, including non-alpha-2μglobulin induced nephrotoxicity, increased relative liver and kidney weights and prostration in 

females, and periocular swelling in both sexes was 403 ppm (1453 mg/m3

).

Mice were exposed for 6 hours/day, 5 days/week for 18 months to MTBE concentrations of 0, 

402, 3014, or 7973 ppm (0, 111, 835, or 2208 mg/m3) (Burleigh-Flayer et al., 1992). The mice 

exposed to the highest concentration (7973 ppm) all exhibited ataxia. Prostration was also noted 

in 8 of 50 animals in this group. Liver weights were elevated in a concentration-dependent 

manner in the female mice but this change was not significant at the lowest concentration (402 

ppm). Kidney weights were elevated in the female mice exposed to 7973 ppm. At the highest 

concentration, a significant increase in hepatocellular hypertrophy and adrenal gland weight was 

detected in the male mice. Spleen weights were increased in the females exposed to the highest 

concentration. 

Moser et al. (1998) exposed female B6C3F1 mice to 7924 ppm (2195 mg/m3

) MTBE for 4 

months, or 7919 ppm (2194 mg/m3

) MTBE for 8 months; controls received plain air. Body 

weight increases for control and MTBE-exposed mice, respectively, were 57% and 37% at 4 

months and 79% and 45% at 8 months: the reduced weight gain in MTBE-exposed mice was 

significantly different from the controls at both time points. In MTBE-exposed mice, mean 

uterine weight was 83% reduced relative to controls at 4 and 8 months. Ovary weight was also 

reduced in exposed mice, the mean weight being 55% of control at 4 months and 51% of control 

at 8 months. Pituitary weights were decreased by 44% and 31% at 4 and 8 months, relative to 

controls. Disturbances of the estrus cycle and histological changes in the reproductive organs 

were also noted. Although the changes in organ weights and histology were suggestive of an 

anti-estrogenic effect of MTBE, serum estrogen levels were unaffected. No changes in estrogen 

receptor (ER) immunoreactivity in reproductive system tissues were observed. Experiments in 

vitro failed to demonstrate any inhibition of estradiol binding to ER by MTBE or its metabolites. 

No inhibition of ER by MTBE was detected, nor was there any inhibition of the induction of ER 

by estradiol. The authors concluded that the apparent anti-estrogenic effects of MTBE were not 

Appendix D3 384 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

mediated via the ER, and drew a parallel with the anti-estrogenic effects of dioxins and 

chlorinated biphenyls. 

Tests for histopathology in the respiratory tract, plasma corticosterone levels, motor activity and 

neurobehavioral endpoints were performed in rats exposed to MTBE at concentrations of 0, 797, 

3920, or 8043 ppm (0, 2877, 14151, or 29035 mg/m3

), 6 hours/day, 5 days/week for 13 weeks 

(Dodd and Kintigh, 1989). Of these endpoints, the most significant finding was an elevation in 

plasma corticosterone in the high dose group. This finding was consistent with the elevated 

adrenal weights reported by Burleigh-Flayer et al. (1992). A clear dose-response for neurotoxic 

effects in these rats was not established.Biles et al. (1987) reported a NOAEL of 300 ppm 

(1083 mg/m3) MTBE for decreased pup viability in rats exposed for 6 hours/day, 5 days/week 

for a total of 16 weeks. Animals exposed to 1240 ppm (4470 mg/m3

) or 2860 ppm (10,311 

mg/m3) MTBE exhibited slightly decreased pup survival. 

Neeper-Bradley (1991) exposed rats to 0, 402, 3019, or 8007 ppm (0, 111, 836, and 2218 mg/m3) 

MTBE over 2 generations. Exposures were for 6 hours/day, 5 days/week during the prebreeding 

period, and for 7 hours/day, 5 days/week during gestation and lactation. Parental effects of 

MTBE exposure were observed, including ataxia, blepharospasm, lack of startle reflex, and 

increased relative liver weights (F1 generation only). There were no histological changes in the 

organs from either parental generation. Reduced body weights were observed in the F1 and F2 

pups at the 3019 and 8007 ppm concentrations. Reduced survivability to postnatal day 4 was 

observed in the 8007 ppm group. No adverse effects were noted at the 403 ppm (111 mg/m3) 

concentration. 

In a developmental and reproductive toxicity study, Conaway and associates (1985) found no 

significant increases in maternal or fetal toxicity, nor in pregnancy rates or in any gross 

toxicologic parameter tested with pregnant rats or mice exposed during gestation to 

concentrations of MTBE up to 3300 ppm (11,897 mg/m3

). 

Maternal toxicity, in the form of hypoactivity and ataxia, was observed in pregnant mice exposed 

during gestation to 4076 ppm (14,690 mg/m3

) MTBE (Bushy Run Research Center, 1989a). 

Significant reductions in food intake and body weight gain were observed in dams exposed to 

8153 ppm (29,390 mg/m3

). Fetal body weight was significantly reduced in the 4076 ppm group, 

and there were significant increases in the incidences of skeletal variations and unossified 

phalanges in the 4076 and 8153 ppm groups. Pregnant rabbits exposed to similar concentrations 

during gestation showed no significant maternal or fetal toxicity or developmental toxicity up to 

a concentration of 8021 ppm (28,918 mg/m3

) (Bushy Run Research Center, 1989b). 

Appendix D3 385 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Chronic Reference Exposure Level 

Study Chun et al., 1992; Bird 1997 

Study population Male and female rats (50 per sex/group) 

Exposure method Discontinuous whole-body inhalation exposures 

(0, 403, 3023, or 7977 ppm) 

Critical effects Nephrotoxicity, increased liver and kidney 

weight, prostration and periocular swelling 

LOAEL 3023 ppm 

NOAEL 403 ppm 

Exposure continuity 6 hours per day, 5 days per week 

Exposure duration 24 months 

Average experimental exposure 72 ppm for the NOAEL group 

Human equivalent concentration 72 ppm for the NOAEL group (gas with 

systemic effects, based on RGDR = 1.0 using 

default assumption that lambda (a) = lambda 

(h)) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 2 ppm (2000 ppb, 8 mg/m3

, 8000 μg/m3) 

The USEPA (1995) based its RfC of 3000 μg/m3 on the same study but included a Modifying 

Factor (MF) of 3 for database deficiencies. The criteria for use of modifying factors are not well 

specified by U.S. EPA. Such modifying factors were not used by OEHHA.

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the REL for MTBE are the use of a comprehensive, long-term multiple 

dose study with large sample sizes and the observation of a NOAEL. The major uncertainty is 

the lack of human data. 

VIII. References 

Biles RW, Schroeder RE, and Holdsworth CE. 1987. Methyl tertiary butyl ether inhalation in 

rats: A single generation study. Toxicol. Ind. Health 3:519-534. 

Bird MG, Burleigh-Flayer HD, Chun JS, Douglas JF, Kneiss JJ, Andrews LS. 1997. 

Oncogenicity studies of inhaled methyl tertiary-butyl ether (MTBE) in CD-1 mice and F-344 

rats. J. Appl. Toxicol. 17 Suppl 1:S45-55. 

Appendix D3 386 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Burleigh-Flayer HD, Chun JS, and Kintigh WJ. 1992. Methyl tertiary butyl ether: Vapor 

inhalation oncogenicity study in CD- 1 mice (unpublished material). Prepared for the MTBE 

Committee by Bushy Run Research Center, Union Carbide Chemicals and Plastics Company, 

Inc. Docket No. OPTS-42098. 

Bushy Run Research Center, Union Carbide Chemicals and Plastics Company, Inc. 1989a. 

Developmental toxicity study of inhaled methyl tertiary butyl ether in CD- 1 mice (final report). 

TSCATS/403186. EPA/OTS No. FYI-OTS-0889-0689. [cited in U.S. EPA, 1995]. 

Bushy Run Research Center, Union Carbide Chemicals and Plastics Company, Inc. 1989b. 

Developmental toxicity study of inhaled methyl tertiary butyl ether in New Zealand White 

rabbits (final report). TSCATS/403186. EPA/OTS No. FYI-OTS-0889-0689. [cited in U.S. EPA, 

1995]. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Chun JS, Burleigh-Flayer HD, and Kintigh WJ. 1992. Methyl tertiary butyl ether: Vapor 

inhalation oncogenicity study in Fischer 344 rats (unpublished material). Prepared for the MTBE 

committee by Bushy Run Research Center, Union Carbide Chemicals and Plastics Company, 

Inc., Docket No. OPTS-42098. 

Conaway CC, Schroeder RE, and Snyder NK. 1985. Teratology evaluation of methyl tertiary 

butyl ether in rats and mice. J. Toxicol. Environ. Health 16(6):797-809. 

Dodd DE, and Kintigh WJ. 1989. Methyl tertiary butyl ether (MTBE): Repeated (13 week) vapor 

inhalation study in rats with neurotoxicity evaluation (unpublished material). Prepared for the 

MTBE Committee by Bushy Run Research Center, Union Carbide Chemicals and Plastics 

Company, Inc., TSCATS 403187. EPA/OTS No. FYI-OTS-08890689. 

Hakkola M, Honkasalo ML, Pulkkinen P. 1997. Changes in neuropsychological symptoms and 

moods among tanker drivers exposed to gasoline during a work week. Occup Med (Lond) 

47(6):344-8. 

HSDB. 1999. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Moser GJ, Wolf DC, Sar M, Gaido KW, Janszen D, Goldsworthy TL. 1998. Methyl tertiary 

butyl ether-induced endocrine alterations in mice are not mediated through the estrogen receptor. 

Toxicol. Sci. 41(1):77-87. 

Neeper-Bradley TL. 1991. Two-generation reproduction study of inhaled methyl tertbutyl ether 

in CD Sprague-Dawley rats (unpublished material). Bushy Run Research Center, Union Carbide 

Chemicals and Plastics Company, Inc. 

Appendix D3 387 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Nihlen A, Walinder R, Lof A, Johanson G. 1998. Experimental exposure to methyl tertiary-butyl 

ether. II. Acute effects in humans. Toxicol. Appl. Pharmacol. 148(2):281-287. 

U.S.EPA. 1999. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration (RfC) for methyl t-butyl ether (MTBE). 

Yoshikawa M, Arashidani K, Katoh T, Kawamoto T, and Kodama Y. 1994. Pulmonary 

elimination of methyl tertiary butyl ether after intraperitoneal administration in mice. Arch. 

Toxicol. 68:517-519. 

Appendix D3 388 Methyl t-Butyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

METHYLENE CHLORIDE 

(dichloromethane, methylene dichloride) 

CAS Registry Number: 75-09-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 400 μg/m3 (100 ppb)

Critical effect(s) Carboxyhemoglobin formation above 2% in 

human workers 

Hazard index target(s) Cardiovascular system; nervous system 

II. Physical and Chemical Properties (HSDB, 1999, except as noted) 

Description Colorless liquid 

Molecular formula CH2Cl2

Molecular weight 84.93 

Density 1.32 g/cm3

 @ 20° C (ACGIH, 1991) 

Boiling point 39.75° C 

Vapor pressure 400 torr @ 24.1° C 

Solubility Miscible with most organic solvents, slightly 

soluble in water (ACGIH, 1991) 

Conversion factor 1 ppm = 3.47 mg/m3 @ 25° C 

III. Major Uses and Sources 

Methylene chloride (MC) is used in paint and varnish remover, in aerosols as a cosolvent or 

vapor pressure depressant, and in solvent degreasing and metal cleaning. It is also used in 

plastics processing and in extraction of fats and oils from food products (HSDB, 1999). The 

annual statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California, based on the most recent inventory, were estimated to be 3,504,271 pounds of 

methylene chloride (CARB, 1999a). Both mean and maximum monitored ambient methylene 

chloride concentrations have decreased slightly between 1990 and 1996 (CARB, 1999b). 

Median and maximum concentrations were 1.09 and 11 ppb in 1990 and 0.66 and 5.6 ppb in 

1996. 

IV. Effects of Human Exposure

Effects of a controlled 2-hour inhalation exposure to MC included CNS depression at 

concentrations of 1000 ppm (3500 mg/m3

) or more and increased blood carboxyhemoglobin 

Appendix D3 389 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

(COHb) content at lower concentrations (500 ppm) due to metabolism of MC to carbon 

monoxide (Stewart et al., 1972). High levels of COHb can be found in the blood hours after 

exposure to methylene chloride, due to its partitioning into fat and its slow release into 

circulation with subsequent metabolism, leading to formation of carbon monoxide (Engstrom 

and Bjurstrom, 1977). In situations of chronic exposure, carbon monoxide toxicity is also of 

concern. Barrowcliff (1978) documented the case of an adult male who developed an unsteady 

gait, a peculiar dysarthria and a loss of memory. The man had worked with 15-50 liters of 

methylene chloride daily for 3 years in a poorly ventilated room while cleansing road materials. 

No natural disease could be found to explain his conditions and the effects were attributed to 

chronic carbon monoxide poisoning. 

Twelve women volunteer subjects were exposed to 0, 300, or 800 ppm methylene chloride for 4 

hours (Fodor and Winneke, 1971). Neurobehavioral vigilance was measured by auditory 

discrimination of intensity of certain sound pulses against a background of continuous white 

noise. A significant interactive effect between methylene chloride concentration and duration of 

exposure using 2-way ANOVA (p < 0.01) was found. 

Human erythrocytes enzymatically convert methylene chloride to formaldehyde in cell-culture 

experiments (Hallier et al., 1994). 

A subacute controlled exposure of eleven resting non-smokers to methylene chloride was 

conducted by DiVincenzo and Kaplan (1981a). The eleven subjects were exposed to 50, 100, 

150, or 200 ppm methylene chloride for 7.5 hours on 5 consecutive days. Exposure to all 

concentrations led to dose-dependent elevation in COHb concentrations in the blood and 

elevated exhaled CO. The peak blood COHb saturations were 1.9, 3.4, 5.3, and 6.8%, 

respectively, for the 50, 100, 150, and 200 ppm groups. 

Divencenzo and Kaplan (1981a) also measured COHb percentage in the blood of workers 

occupationally exposed to methylene chloride and a group of workers not exposed to methylene 

chloride. The 19 workers exposed to methylene chloride had mean blood COHb concentrations 

of 2.3% in the morning and 3.9% at the end of the work-shift. Ambient concentrations in the 

workplace were estimated from 57 samples, which ranged from 0 to 250 ppm, with a mean 

concentration of 40 ppm. Three exposed workers also wore monitors to estimate personal 

exposures. The time-weighted average exposure for these workers was 33 ppm. Controls (8 

subjects) had significantly lower mean blood COHb concentrations of 0.8% in the AM and 1.3% 

in the PM compared with the exposed workers. The length of employment of the exposed 

workers was not given. 

A companion study by DiVincenzo and Kaplan (1981b) showed that smoking and methylene 

chloride exposure result in an additive effect on COHb levels compared with levels in nonsmokers. Similarly, light, moderate or heavy exercise workloads resulted in higher COHb levels. 

Soden et al. (1996) showed a dose-response increase in carboxyhemoglobin levels in nonsmokers with increasing methylene chloride exposure in workers involved in triacetate fiber 

production. Carboxyhemoglobin levels ranged from 1.77% to 4% from exposures ranging from 

6.5 to 89.7 ppm, respectively. The number of employees in the study was not reported. 

Appendix D3 390 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Although animal studies have shown COHb-induced cardiovascular effects following MC 

exposure (Aviado et al., 1977), no data exist on this outcome in humans. However, studies of 

men with coronary artery disease and exercise-induced angina report a decrease in time to onset 

of exercise-induced angina following exposure to carbon monoxide (CO) at concentrations 

sufficient to result in blood COHb levels of about 2% (Kleinman et al., 1989; Allred et al., 

1989). A physiologically based pharmacokinetic model of MC and CO estimated that a 1-hour 

exposure to 340 ppm (1200 mg/m3

) MC at a ventilation rate of 9 liters/min would result in a peak 

blood COHb level of 2% (Andersen et al., 1991; Reitz, 1994). The California Ambient Air 

Quality Standard for CO is based on a blood COHb level of 2% (CARB, 1982). 

An epidemiological study of 751 male workers in the Eastman Kodak Company exposed to daily 

8-hour time-weighted average concentrations of 30-125 ppm methylene chloride for up to 30 

years was conducted by Friedlander and associates (1978). A control group of workers in 

production but not exposed to methylene chloride was used together with New York state cause 

and age-specific mortality rates. The follow-up period for these workers was 13 years, with 97% 

success. The studies did not indicate any increase in risk of death from circulatory disease, 

cancer, or other causes due to methylene chloride exposure. 

A study of female pharmaceutical workers in eight different factories exposed to a variety of 

organic solvents indicated that solvent exposure, and particularly methylene chloride exposure, 

resulted in an increase in spontaneous abortions (Taskinen et al., 1986). In all, 1795 pregnancies 

were followed, with 142 spontaneous abortions occurring. The odds ratio for methylene chloride 

exposure was 1.0 to 5.7 (average = 2.3; p < 0.06). There was a significant effect of exposure to 4 

or more solvents, compared with age-matched controls (p < 0.05). The concentrations of MC 

were not reported in the study. 

The U.S. Occupational Safety and Health Administration reduced its permissible exposure limits 

(PEL) for MC from 500 ppm to 25 ppm in 1997 (U.S. CFR, 1997). 

V. Effects of Animal Exposure 

Nitschke et al. (1988) found that a 2-year exposure to 0, 50, 200, or 500 ppm MC for 6 

hours/day, 5 days/week resulted in significant histopathologic lesions in the livers of rats 

exposed to 500 ppm. No significant adverse effects were observed at 200 ppm or lower. The 

predominant hepatocellular lesion was fatty vacuolization of hepatocytes. 

Female B6C3F1 mice inhaling 2000 ppm MC for 1 to 26 weeks had 40 to 60% lower cell 

turnover rates of bronchiolar cells compared with controls (Kanno et al., 1993). At this 

concentration no observable pathological changes were found in the lungs of MC exposed 

animals. 

A continuous exposure of mice (16 per group) to 100 ppm MC for 1, 2, 3, 4 or 10 weeks resulted 

in signifcant elevation in liver triglycerides beginning at 2 weeks and lasting throughout the 10-

week period (Weinstein and Diamond, 1972). Liver/body weight ratios were unaffected at any 

Appendix D3 391 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

time point. After 1 week, small fat droplets were apparent in centrilobular hepatocytes and a 

decrease in hepatic glycogen was also noted. Necrosis was not observed during the 10-week 

period, but fat droplet size increased and glycogen depletion persisted. 

 Male and female Sprague-Dawley rats and Golden Syrian hamsters inhaled methylene chloride 

(0, 500, 1500, or 3500 ppm) for 6 hr per day, 5 days a week over 2 years (Burek et al., 1984). 

The groups consisted of 129 rats per sex per concentration, and 107 to 109 hamsters per sex per 

concentration. Females rats inhaling 3500 ppm had an increased mortality rate while female 

hamsters inhaling 1500 or 3500 ppm had decreased mortality rates. Slight histopathological 

findings were noted in livers of rats exposed to 500, 1500, or 3500 ppm MC. Decreased 

amyloidosis was also found in livers and other organs of hamsters at each of the three MC 

concentrations. Overall, effects were more potent in rats compared with hamsters, which had 

fewer spontaneous age-related changes, decreased mortality (at least for females), and evidence 

of specific target organ toxicity was weak. Carboxyhemoglobin values were elevated in both rats 

and hamsters exposed to 500 ppm or more of MC, with the percentage increase greater in 

hamsters than in rats. 

Monkeys were observed to be more susceptible subjects for methylene chloride induced COHb 

than dogs upon 14-week subchronic continuous exposure to 25 or 100 ppm (Haun et al., 1972). 

At 25 ppm, approximately 1.5% COHb was reached in the 4 monkeys, compared to 

approximately 0.5% in 16 dogs. Monkeys exposed to 100 ppm MC had COHb levels of 

approximately 4% compared with 2% in the dogs. 

Oral ethanol pretreatment in rats has been shown to suppress the COHb formation characteristic 

of methylene chloride exposure through inhibition of biotransformation of methylene chloride 

(Glatzel et al., 1987). 

Gerbils (10/sex per group; 60 controls) exposed continuously to MC concentrations of 210, 350, 

or 700 ppm for a period of 3 months, with a 4-month follow-up period, showed irreversible 

cellular and biochemical changes in brain (Rosengren et al., 1986). A high mortality rate (19/20) 

was observed in the 700 ppm group, and this exposure was terminated after 7 weeks. The gerbils 

exposed to 350 ppm also had a high mortality rate (9/20) and this exposure was terminated after 

10 weeks. The gerbils exposed to 210 ppm had no premature mortality and the exposure 

continued for the full 3 months. Four months after termination of exposure, the animals in the 

350 and 210 ppm groups had significantly decreased brain DNA content in the hippocampus. 

The 350 ppm group exhibited elevated astroglial proteins in the frontal and sensory motor 

cerebral cortex, consistent with astrogliosis in these regions. In addition, the gerbils exposed to 

350 ppm MC had significantly decreased DNA in the cerebellar hemispheres. Complimentary 

studies by these investigators showed that the formation of carboxyhemoglobin did not increase 

in gerbils between the 210 and 350 ppm exposures, indicating that the metabolism of MC to CO 

is saturable at concentrations below those in the study. On the other hand, the neurotoxic brain 

biochemical alterations were significantly greater in gerbils exposed to 350 ppm as compared 

with the 210 ppm group, implying that carboxyhemoglobin induced cerebral hypoxia is not the 

major cause of MC-induced neurotoxicity in the brain. 

Rats (50 per sex per group) were exposed to 0, 1000, 2000, or 4000 ppm methylene chloride 6 

hours/day, 5 days/week for 102 weeks (NTP, 1986). Both sexes exhibited hemosiderin 

Appendix D3 392 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

pigmentation in the liver in a dose-dependent fashion, beginning with the 1000 ppm 

concentration. Squamous metaplasia of the nasal cavity was observed in female rats, and thyroid 

C-cell hyperplasia was observed in males exposed to 2000 ppm or greater. Kidney tubule 

degeneration (not otherwise specified) was increased at all exposure levels. 

Mice (50 per sex per group) exposed to 0, 2000, or 4000 ppm methylene chloride 6 hours/day, 5 

days/week for 102 weeks showed increased incidence of liver cytologic degeneration and splenic 

atrophy at 4000 ppm (males) (NTP, 1986). Male and female mice also had an increased 

incidence of kidney tubule casts (not otherwise specified) at 2000 ppm or greater, and significant 

testicular atrophy was observed in males at 4000 ppm. Female mice showed cytologic 

degeneration in the liver at 2000 ppm or greater, and ovarian atrophy at 2000 ppm or greater. 

A six month exposure to 5000 ppm MC of 8 guinea pigs for 7 hours/day, 5 days/week resulted in 

3 deaths; 2 showed moderate centrilobular fatty degeneration of the liver and extensive 

pneumonia at necropsy (Heppel et al., 1944). None of the 14 control animals died. Food 

consumption and body weight were lower in the exposed guinea pigs, compared with control 

pigs. One out of 12 rats died at this concentration, and the liver histology in this animal revealed 

multiple thrombi in renal vessels, associated with marked cortical infarction. By comparison, 

dogs and rabbits showed no signs of illness, nor were blood pressure or hematological values 

altered at the 5000 ppm concentration. At 10,000 ppm, 2 of 4 dogs showed moderate 

centrilobular congestion, narrowing of liver cell cords, and slight to moderate fatty degeneration. 

One of 2 monkeys revealed disseminated tuberculosis lesions, but no other histological 

alterations. Four out of 6 guinea pigs had moderate fatty degeneration of the liver at this 

concentration. 

The offspring of rats (10 dams per group) exposed during gestation to 0 or 4500 ppm methylene 

chloride exhibited altered rates of behavioral habituation to novel environments (Bornschein et 

al., 1980). This effect was observed beginning at 10 days of age but was still demonstrable in 

rats 150 days old. The authors concluded that elevated maternal COHb could have been a 

contributing factor in the developmental impairment. 

In a study of the effects of methylene chloride on estrous cycle and serum prolactin, groups of 15 

female rats were exposed to 0 or 3500 ppm for 6 hours/day for 15 to 19 consecutive days 

(Breslin and Landry, 1986). Males (15 per group) were exposed for 5 hours/day for 5 

consecutive days. Female rats exhibited decreased body weight and increases in the estrous 

cycle duration and in serum prolactin. Males did not show any significant effects on serum 

prolactin from methylene chloride exposure. 

Pregnant mice and rats were exposed to 0 or 1250 ppm MC 7 hours/day, on days 6 through 15 of 

gestation (Schwetz et al., 1975). Significantly elevated absolute liver weights were seen in 

maternal animals from both species. In addition, significantly increased incidences of delayed 

ossification of the sternebrae were seen in both species, compared to controls. 

Methylene chloride exposure of female rats before or during gestation to 4500 ppm resulted in 

elevated maternal liver weights and decreased birth weights of the offspring, but no terata or 

skeletal/soft tissue anomalies (Hardin and Manson, 1980). 

Appendix D3 393 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

A 2-generation reproduction test was conducted by Dow Chemical Company (Nitschke et al., 

1985) which showed no significant reproductive or developmental effects in rats exposed to 0, 

100, 500, or 1500 ppm MC 6 hours/day, 5 days/week, for 14 weeks. The exposure conditions 

were identical for the F0 and F1 generations. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study DiVincenzo and Kaplan (1981a) 

Study population 19 workers, 8 controls 

Exposure method Occupational inhalation exposure 

Critical effects Significantly elevated carboxyhemoglobin levels 

(> 2%) 

LOAEL 40 ppm (ambient workplace exposures averaged 

40 ppm with a range of 0 to 250 ppm); 

controls = 0 ppm 

NOAEL Not observed 

Exposure continuity 8 hours/day, 5 days/week 

Exposure duration Length of employment unspecified 

Average occcupational exposure 14 ppm for LOAEL group 

 (40 x 10/20 x 5/7) 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 1 (see following text for explanation) 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.1 ppm (100 pbb; 0.4 mg/m3

; 400 μg/m3) 

Workers were exposed to average measured concentrations of 40 ppm during the workday, and 

the personal monitors on 3 of the subjects indicated a 8-hour time-weighted average of 33 ppm 

over a 2-week period. The average COHb levels were 3.9% at the end of the work-shift. 

Elevated carboxyhemoglobin concentrations of above 2% are considered to aggravate angina in 

some individuals (CARB, 1982). In effect, 2% COHb can be considered a NOAEL for 

aggravation of angina. Therefore, the 33 ppm concentration was considered a LOAEL for the 

formation of greater than 2% COHb. The duration of the employment period was not specified. 

However, in the DiVincenzo and Kaplan (1981a) study, the levels of COHb did not appear to 

increase over a period of 5 days in experimental exposures using volunteers, therefore an 

uncertainty factor for subchronic exposure was not necessary. A number of factors contribute to 

the uncertainty in determining the degree of sensitivity to methylene chloride, including activity 

level, metabolic enzyme activity, age, and background COHb status (e.g., from smoking, etc.). 

The subchronic study by Haun et al. (1972) with monkeys reported a NOAEL of 25 ppm and a 

LOAEL of 100 ppm for 2% COHb formation following a 14-week exposure. These results are 

consistent with the LOAEL reported in the DiVincenzo and Kaplan study. However, the human 

occupational study likely contains less uncertainty, since the toxicokinetics of the effect, 

Appendix D3 394 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

including rate of formation of CO and thus COHb is metabolism-dependent, resulting in 

considerable potential interspecies differences. 

The study in hamsters by Burek et al.(1984) showed a LOAEL for elevated carboxyhemoglobin 

of 500 ppm. A time-weight average exposure and HEC of 89 ppm was calculated. Using a 10-

fold LOAEL uncertainty factor, a 3-fold interspecies uncertainty factor for residual uncertainty 

not accounted for in the HEC calculation, and a 10-fold intraspecies uncertainty factor, a REL of 

300 ppb or 1000 μg/m3

 was derived. Thus, the REL derived from the best available animal study 

is comparable to the 400 μg/m3

 REL derived from the best-available human study. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the key study (DiVincenzo and Kaplan, 1981a) used to derive the REL for 

methylene chloride is that human health effects were observed. The major uncertainties from the 

key study itself are the lack of a NOAEL observation, the difficulty in estimating exposures, and 

the discontinuous and variable nature of the exposures. 

The health effects database for methylene chloride includes, in addition to an adequate study of 

human occupational exposures (DiVincenzo and Kaplan, 1981a), an adequate lifetime inhalation 

exposure study in 2 species of laboratory animals (Burek et al., 1984). The REL values derived 

from these studies (400 μg/m3

 vs. 1,000 μg/m3) are comparable. That both the human and animal 

studies measured the same endpoint and arrived at similar conclusions is a circumstance that is 

rarely found but one that considerably increases the weight of evidence from which the REL was 

derived. The two studies complement each other, as the animal study involved controlled, 

measured exposures over a lifetime but introduces the uncertainty of predicting human health 

effects from animal observations, and the human study involved poorly characterized human 

exposures but lacks the uncertainty inherent in interspecies extrapolation. 

VIII. References 

Allred EN, Bleecker ER, Chaitman BR, Dahms TE, Gottlieb SO, Hackney JD, Hayes D, 

Pagano M, Selvester RH, Walden SM, and Warren J. 1989. Acute effects of carbon monoxide 

exposure on individuals with coronary artery disease. Health Effects Research Institute (HEI) 

Research Report No. 25., Cambridge, MA: HEI. 

ACGIH. 1991. American Conference of Governmental Industrial Hygienists. Documentation of 

Threshold Limit Values and Biological Exposure Indices. 6th edition Cincinnati, OH: ACGIH. 

Andersen ME, Clewell HJ III, Gargas ML, MacNaughton MG, Reitz RH, Nolan RJ, and 

McKenna MJ. 1991. Physiologically based pharmacokinetic modeling with dichloromethane, its 

metabolite, carbon monoxide, and blood carboxyhemoglobin in rats and humans. Toxicol. Appl. 

Pharmacol. 108:14-27. 

Appendix D3 395 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Aviado DM, Zakhari S, Watanabe T. 1977. Methylene chloride. In: Golberg L. (ed.) NonFluorinated Propellants and Solvents for Aerosols, Cleveland, OH: CRC Press, pp.19-45. 

Barrowcliff DF. 1978. Chronic carbon monoxide poisoning caused by methylene chloride 

paintstripper. Med. Sci. Law 18(4):238. 

Bornschein RL, Hastings L, and Manson JM. 1980. Behavioral toxicity in the offspring of rats 

following maternal exposure to dichloromethane. Toxicol. Appl. Pharmacol. 52:29-37. 

Breslin WJ, and Landry TD. 1986. Methylene chloride: Effects on estrous cycling and serum 

prolactin in Sprague-Dawley rats. Unpublished study by Dow Chemical U.S.A., Midland, MI 

48674. 

Burek JD, Nitschke KD, Bell TJ, Wackerle DL, Childs RC, Beyer JE, Dittenber DA, 

Rampy LW, and McKenna MJ. 1984. Methylene chloride: A two-year inhalation toxicity and 

oncogenicity study in rats and hamsters. Fundam. Appl. Toxicol. 4:30-47. 

CARB (California Air Resources Board). 1982. California Air Resources Board. California 

Ambient Air Quality Standards for Carbon Monoxide (Sea Level). Technical Support Document. 

Sacramento, CA: California Air Resources Board, Technical Support Division. 

CARB (California Air Resources Board). 1999a. Air toxics emissions data collected in the Air 

Toxics Hot Spots Program CEIDARS Database as of January 29, 1999. 

CARB (California Air Resources Board). 1999b. 1990-1996 Statewide Methylene Chloride 

Summary, ppb. WWW document (http://www.arb.ca.gov/aqd/dcm/dmstate.htm). 

DiVincenzo GD, and Kaplan CJ. 1981a. Uptake, metabolism, and elimination of methylene 

chloride vapor by humans. Toxicol. Appl. Pharmacol. 59:130-140. 

DiVincenzo GD, and Kaplan CJ. 1981b. Effect of exercise or smoking on the uptake, 

metabolism, and excretion of methylene chloride vapor. Toxicol. Appl. Pharmacol. 59:141-148. 

Engstrom J, and Bjurstrom R. 1977. Exposure to methylene chloride: content in subcutaneous 

adipose tissue. Scand. J. Work Environ. Health 3:215-224. 

Fodor GG, and Winneke H. 1971. Nervous system disturbances in men and animals 

experimentally exposed to industrial solvent vapors. Proc. 2nd Intl. Clean Air Congr. New York: 

Academic Press. pp. 238-243. 

Friedlander BR, Hearne T, and Hall S. 1978. Epidemiologic investigation of employees 

chronically exposed to methylene chloride. J. Occup. Med. 20:657-666. 

Glatzel W, Tietze K, Gutewort R, and Pankow D. 1987. Interaction of dichloromethane and 

ethanol in rats: Toxicokinetics and nerve conduction velocity. Alcohol. Clin. Exp. Res. 

11(5):450-452. 

Appendix D3 396 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Hallier E, Schroder KR, Asmuth K, Dommermuth A, Aust B, Goerens HW. 1994. Metabolism 

of dichloromethane (methylene chloride) to formaldehyde in human erythrocytes: influence of 

polymorphism of glutathione transferase theta (GST T1-1). Arch. Toxicol. 68:423-427. 

Hardin BD, and Manson JM. 1980. Absence of dichloromethane teratogenicity with inhalation 

exposure in rats.Toxicol Appl Pharmacol 1980 52(1):22-28. 

Haun CC, Vernot EH, and Darmer KI. 1972. Continuous animal exposure to low levels of 

dichloromethane. Aerospace Medical Research Laboratories, Wright-Patterson Air Force Base, 

Dayton, OH. AMRL-TR-72-130. 

HSDB. 1999. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. WWW database (http://sis.nlm.nih.gov/sis1/). 

Heppel LA, Neal PA, Perrin TL, Orr ML, and Porterfield VT. 1944. Toxicology of 

dichloromethane (methylene chloride). I. Studies on effects of daily inhalation. J. Ind. Hyg. 

Toxicol. 26(1):8-16. 

Kanno J, Foley JF, Kari F, Anderson MW, Maronpot RR. 1993. Effect of methylene chloride 

inhalation on replicative DNA synthesis in the lungs of female B6C3F1 mice. Environ Health 

Perspect 101 Suppl 5:271-276. 

Kleinman MT, Davidson DM, Vandigraff RB, Caiozzo VJ, and Whittenberger JL. 1989. Effects 

of short-term exposure to carbon monoxide in subjects with coronary artery disease. Arch. Env. 

Health 44(6):361-369. 

NTP. 1986. National Toxicology Program. Toxicology and carcinogenesis studies of 

dichloromethane (methylene chloride) in F344/N rats and B6C3F1 mice. NIH Publication 86-

2562. Reasearch Triangle Park, NC: U.S. Dept. of Health and Human Services. 

Nitschke KD, Eisenbrandt DL, and Lomax LG. 1985. Methylene chloride: Two -generation 

reproduction study in Fischer 344 rats. Midland, MI: Dow Chemical Company. February 5. 

Nitschke KD, Burek JD, Bell TJ, Kociba RJ, Rampy LW, and McKenna MJ. 1988. Methylene 

chloride: A 2-year inhalation toxicity and oncogenicity study in rats. Fundam. Appl. Toxicol. 

11:48-59. 

Reitz RH. 1994. Personal communication. August 5, 1994. 

Rosengren LE, Kjellstrand P, Aurell A, and Haglid KG. 1986. Irreversible effects of 

dichloromethane on the brain after long term exposure: a quantitative study of DNA and the glial 

cell marker proteins S-100 and GFA. Br. J. Ind. Med. 43:291-299. 

Appendix D3 397 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Schwetz BA, Leong KJ, and Gehring PJ. 1975. The effect of maternally inhaled 

trichloroethylene, perchloroethylene, methyl chloroform, and methylene chloride on embryonal 

and fetal development in mice and rats. Toxicol. Appl. Pharmacol. 32(1):84-96 

Soden KJ, Marras G, and Amsel J. 1996. Carboxyhemoglobin levels in methylene chlorideexposed employees. J. Occup. Environ. Med. 38(4):367-371. 

Stewart RD, Fisher TN, Hosko MJ, Peterson JE, Baretta ED, and Dodd HC. 1972. Experimental 

human exposure to methylene chloride. Arch. Env. Health 25(5):342-348. 

Taskinen H, Lindbohm ML, and Hemminki K. 1986. Spontaneous abortions among women 

working in the pharmaceutical industry. Br. J. Ind. Med. 43:199-205. 

U.S. Code of Federal Regulations (29 CFR). Final Rule for Occupational Exposure to Methylene 

Chloride. Federal Register 62(7):1494. 

Weinstein RS, and Diamond SS. 1972. Hepatotoxicity of dichloromethane (methylene chloride) 

with continuous inhalation exposure at a low dose level. Proceedings of the 3rd Annual 

Conference on Environmental Toxicology, 25, 26, and 27 October, 1972. Aerospace Medical 

Research Laboratory, Aerospace Medical Division, Air Force Systems Command, WrightPatterson Air Force Base, OH. 

Appendix D3 398 Methylene Chloride 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

4,4′-METHYLENE DIANILINE 

(MDA; 4,4’-diaminodiphenylmethane; 4,4’-diphenylmethanediamine; DAPM; dianilinemethane) 

CAS Registry Number: 101-77-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 20 μg/m3 (2 ppb) 

Critical effect(s) Ocular toxicity to the retinas of guinea pigs 

Hazard index target(s) Eyes; alimentary system (hepatotoxicity) 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994)

 

Description Colorless to pale yellow flakes; tan 

Molecular formula C13H14N2

Molecular weight 198.3 g/mol 

Boiling point 398-399°C 

Melting point 92.5°C 

Vapor pressure 1 torr @ 197°C 

Solubility Soluble in alcohol, benzene, ether; 273 g/100 g 

acetone; 0.1 g/100 g water @ 25°C 

Conversion factor 8.1 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

4,4’-Methylene dianiline (MDA) is synthesized by the reaction of aniline with formaldehyde. 

MDA’s major uses are as a chemical intermediate in the synthesis of certain isocyanates and 

polyurethane polymers, as a corrosion inhibitor, in the preparation of azo dyes, as a rubber 

preservative, and in the curing of epoxy resins and neoprene (HSDB, 1995; ACGIH, 1992). The 

annual statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California based on the most recent inventory were estimated to be 1133 pounds of MDA 

(and its dichloride) (CARB, 2000). 

IV. Effects of Human Exposure 

Several cases of human exposure to MDA have identified the compound as a hepatotoxicant 

which produces cholestatic jaundice (Kopelman et al., 1966; McGill and Motto, 1974; Williams 

et al., 1974; Bastian, 1984). Bastian (1984) described cases of acute hepatic illness in four 

workers exposed from laying floors using an epoxy resin base, which contained MDA as a 

curing agent. The workers were exposed via fumes and dusts in the air as well from hand 

Appendix D3 399 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

contact with powder and had worked with epoxy resins for periods ranging from one to 12 years. 

The level of exposure was not quantified. The workers initially reported to the hospital with 

symptoms of abdominal pain three days after the most recent exposure and all were discharged 

within four days. Two workers continued to show severe symptoms five days after the onset, 

with abdominal pain, jaundice, a tender liver, nausea, dyspnea, and muscular pain. Plasma 

bilirubin, alkaline phosphatase, and aspartate aminotransferase levels were elevated. Some 

symptoms did not subside until two months after the onset. One worker, after another exposure, 

experienced nausea, abdominal pain, and muscular pain. A second worker reported further 

symptoms of headache, tiredness, and decreased libido. 

Williams et al. (1974) reported symptoms in 6 of approximately 300 workers exposed to MDA 

by surface coating concrete walls with epoxy resins. Exposure probably occurred by inhalation, 

ingestion, and skin contact as a result of mixing powder containing MDA. Symptoms of clinical 

hepatitis in the 6 workers appeared two days to two weeks after beginning work; five of the six 

had elevated bilirubin levels, and one liver biopsy showed bile stasis. All the workers recovered 

completely after an unspecified time. 

McGill and Motto (1974) described hepatitis among 13 men who, over the course of 6 years, 

were occupationally exposed to MDA in the blending of epoxy resins used in the manufacture of 

insulating material. Among the 13 patients showing symptoms, all reported weakness, jaundice, 

and dark urine; 11 reported abdominal pain, nausea or vomiting, and anorexia; and over half 

reported fever, chills and/or headache. All the workers recovered within a 10 week period. 

After the first cases of hepatitis occurred, air sampling showed initial levels of MDA to be 0.1 

ppm in the work area. After additional cases of hepatitis occurred, measures were taken to 

reduce worker exposure, and air levels were reduced to as low as 0.0064 ppm. The authors 

concluded that percutaneous absorption was the likely major route of exposure in light of the fact 

that cases occurred in spite of measures taken to reduce air levels and there was evidence that 

significant hand contact with the compound occurred during the workday. Since the symptoms 

appeared within one to 18 days after “working intensively” with the compound and exposure 

routes were not clearly established, quantitation of exposure levels was considered difficult. 

The most well-known incident of MDA toxicity to humans resulted from ingestion of bread 

made with flour contaminated with MDA during transport (Kopelman et al., 1966a). Eighty-four 

persons showed symptoms of abdominal pain and some degree of jaundice. All patients had 

elevated serum alkaline phosphatase and glutamic oxaloacetic transaminase levels. Seventeen 

had serum bilirubin levels over 5 mg/100 ml. Liver biopsy was performed on 8 persons and 

evaluated in a separate study (Kopelman et al., 1966b). The primary finding was an unusual 

lesion described during the early course of the disease as portal zone cholangitis and later as 

centrilobular cholestasis with necrosis. The initial study reported that all but 2 patients had 

complete recovery within several weeks. However, a two year follow-up study of 14 individuals 

showed that 10 still had symptoms of some severity 7 to 23 months after initial onset including 

food intolerance, gastrointestinal disturbances, fatigue, and visual disturbances (Kopelman, 

1968). 

Human effects other than hepatotoxicity have been described including several cases of contact 

dermatitis and skin sensitization (LeVine, 1983; Van Joost et al., 1987; de Pablo et al., 1992; 

Appendix D3 400 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Bruynzeel and van der Wegen-Keijser, 1993). A case report of a man exposed to MDA with 

potassium carbonate and γ-butyrolactone by accidental ingestion has been described (Roy et al., 

1985). In addition to hepatitis and abnormal liver function, which persisted over 18 months, the 

patient developed a progressively worsening retinopathy described as a “malfunction of the 

retinal pigment epithelium” accompanied by diminished visual acuity. The patient improved 

after approximately 3 months, but after examination at 18 months had not completely recovered. 

Another report described the development of acute cardiomyopathy in addition to hepatitis in a 

worker exposed to a large quantity of MDA dust as the result of air filtration malfunction 

(Brooks et al., 1979). The patient showed an abnormal ECG and an elevated cardiac LDH 

isoenzyme profile, which returned to normal within one month of onset. 

V. Effects of Animal Exposure 

The carcinogenicity of MDA was investigated in F344/N rats and B6C3F1 mice (50/sex/dose 

group) administered in the drinking water at concentrations of 0, 150, and 300 ppm MDA 

(dihydrochloride) for 103 weeks (Lamb et al., 1986). A 14-day range finding study was also 

conducted with 5 animal/sex/species/dose group, with exposure levels of 0, 200, 400, 800, 1600, 

and 3200 ppm MDA. A 13-week subchronic study was conducted with 10 

animals/sex/species/dose group and exposure levels of 0, 25 (mice), 50, 100, 200, 400, and 800 

(rats) ppm MDA. Using body weight and drinking water values from the study, low and high 

daily doses in the chronic study were calculated to be 9 and 16 mg/kg-day for male rats, 10 and 

19 for female rats, 25 and 57 for male mice, and 19 and 43 for female mice. In the chronic 

study, survival was reduced among male mice treated with 300 ppm MDA. Final mean body 

weights were reduced in the 300 ppm dose group of female rats (-9%), male mice (-13%), and 

female mice (-16%). Among rats, non-cancer effects included follicular cysts and follicular-cell 

hyperplasia of the thyroid (significantly increased incidence in high-dose females; p<0.05 by 

Fisher’s exact test). In the liver, the incidence of fatty and focal cellular change was elevated in 

low-dose male and female rats and also in high dose male rats. Incidence of unspecified 

dilatation of the liver was also elevated in high-dose male rats. Increased incidence of kidney 

mineralization was found in male rats treated with 300 ppm MDA. Among mice, incidence of 

liver degeneration was elevated in males in both treatment groups and females in the high-dose 

group (p<0.01 by Fisher’s exact test). Incidence of kidney nephropathy was increased in male 

and female mice in both treatment groups and mineralization of the renal papilla was increased 

in both sexes in the high-dose group (p<0.01). From the 13-week study, the authors noted 

thyroid and bile duct effects in rats at 800 ppm MDA in water and in mice at 400 ppm MDA in 

water. 

Albino and pigmented guinea pigs were exposed to aerosols of methylene dianiline in 

polyethylene glycol 200 (PEG) in nose-only exposure chambers (Leong et al., 1987). Animals 

(8 of each strain) were exposed to a time-weighted average aerosol concentration of 0.44 g 

MDA/m3

 in air for 4 hours/day, 5 days/week for 2 weeks. Eight control animals were neither 

exposed to aerosol nor placed in the exposure chamber. Two weeks after the exposure period, 

animals were evaluated for dermal sensitization and irritation by challenge with 0.05 ml of 0, 2, 

20, and 200 mg MDA/ml in PEG for up to 24 hours. No evidence of dermal irritation or 

Appendix D3 401 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

sensitivity was found. Subsequently, the animals were also examined for pulmonary 

sensitization by challenge with aerosols containing 0.01 and 0.05 ml of 200 mg MDA/ml PEG. 

Lung insufflation pressures were measured as an indication of changes in lung distensibility. No 

evidence of pulmonary sensitization was found. After the pulmonary challenge, the animals 

were examined histopathologically, with emphasis on eye, lung, liver, and kidney toxicity. 

Ocular toxicity ranging from mild to more severe was observed in all MDA-treated animals, but 

in none of the control animals. Pigmented animals did not differ in sensitivity or effect 

compared to albino animals. Mild lesions were described as “retraction and thickening of the 

outer segments of the photoreceptor cells” while more severe effects included swelling “through 

the inner segments of the photoreceptor cells to the outer nuclear layer.” Some evidence of 

inflammatory cell infiltration was also noted and the pigmented epithelial layer was also 

degenerated. The authors conclude that the effects were attributable to MDA because no retinal 

lesions have been associated with exposure to the PEG vehicle. Furthermore, the inhalation 

exposures to MDA are the likely cause rather than the dermal and lung sensitization study 

exposures because these subsequent studies were conducted on control as well as treated 

animals. Pulmonary granulomas consisting of “an aggregate of macrophages surrounded by a 

thin mantle of lymphocytes” were found in 7 of the 16 MDA-exposed animals and one of the 8 

control animals (level of significance was not stated). Treated and control animals had a high 

background incidence of pulmonary lesions including slight to mild bronchitis. No liver or 

kidney effects were detected in treated animals. 

Nine purebred beagle dogs were treated orally (by capsule) with 70 mg “crude” (4 dogs) or 

“purified” (5 dogs) MDA in corn oil three days per week for a period ranging from 

approximately 3 to 7 years (Deichmann et al., 1978). No concurrent controls were included 

since untreated animals were regularly maintained in the laboratory. After 2 years, cystoscopic 

examination was performed at 15-month intervals. After 41⁄2 years, clinical chemistry tests were 

performed at 4 month intervals on 3 dogs from each group. Microscopic examination of urinary 

bladder, liver, heart, ovaries, uterus, and lymph nodes was performed on moribund animals or at 

the end of the experimental period (7 years, 2 months). Liver toxicity was noted in all the treated 

animals. Effects were described as fatty change, cell degeneration and necrosis, and lymphoid 

cell infiltration. One dog from each treatment group died from the toxic effect on the liver. The 

kidneys of four treated animals (two from each group) showed toxic effects including granuloma, 

glomerular nephritis, and congestion with cloudy swelling. Two dogs treated with “purified” 

and one dog treated with “crude” MDA showed toxicity to the spleen described as hemosiderosis 

and swelling with lymphocyte infiltration. 

Wistar rats (5/sex/dose) were treated orally with 0, 0.0083, and 0.083 g MDA/kg body weight in 

propylene glycol daily for 12 weeks (Pludro et al., 1969). Doses were 1% and 10% of the 

experimentally determined median lethal dose. No significant changes in body weight or 

hematological parameters were found, although serum albumin, β-globulin, and γ-globulin were 

elevated in animals in the 0.083 mg/kg dose group. The livers of all the animals in the high dose 

group showed signs of degeneration, including atrophy of the parenchyma and stromal 

hyperplasia in the portal areas. Also in this dose group, all animals showed hypertrophy of the 

lymphatic nodules of the spleen. In the low dose group, one animal showed a liver lesion and 

one a lesion in the spleen. 

Appendix D3 402 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Schoental (1968) treated rats (8/sex) with MDA in 25% aqueous ethanol by stomach tube. Rats 

were given 20 mg doses a total of 2-5 times over several weeks up to 71⁄2 months (frequency not 

specified). Animals showed necrosis of the liver and kidney and congestion and edema of the 

lungs. 

Visual toxicity was reported in 15 cats treated perorally with 25-100 mg MDA/kg body weight in 

a 1% aqueous suspension (Schilling von Canstatt et al., 1966). In four animals treated once with 

100 mg/kg, no blindness was reported. In all the other treated animals (four with one dose of 

100 mg/kg, two with one dose of 150 mg/kg, and two with three doses of 25 mg/kg and 3 doses 

of 50 mg/kg), blindness occurred within 8 days. Three of the eight recovered sight within 4 

days. Two other treated animals were examined microscopically, one treated with 25 and then 

50 mg/kg and one treated once with 200 mg/kg. The first was examined after 7 days and showed 

signs of granular degeneration of the rods and cones with some proliferation of the pigmented 

epithelium. The second was examined after 41⁄4 years and showed atrophy of the retinal 

neuroepithelium. The authors noted that no visual disturbances were found in other MDA 

treated experimental animals, including dog, rabbit, guinea pig, and rat. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Leong et al., 1987 

Study population Guinea pigs 

Exposure method Discontinuous inhalation exposure (nose only) of 

aerosols 

Critical effects Degeneration of retinal epithelium 

LOAEL 440 mg/m3

 (54 ppm) 

NOAEL Not observed 

Exposure continuity 4 hours/day, 5 days/week 

Exposure duration 2 weeks 

Average experimental exposure 52 mg/m3

 for LOAEL group (440 x 4/24 x 5/7) 

(6.4 ppm) 

Human equivalent concentration 52 mg/m3

 using the default assumption of RGDR 

= 1 for a gas with systemic effects 

LOAEL uncertainty factor 10 (incidence = 100%) 

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 3,000 

Inhalation reference exposure level 0.02 mg/m3 (20 μg/m3

; 0.002 ppm; 2 ppb) 

Two specific types of toxicity have been associated with exposure to MDA: hepatotoxicity and 

ocular toxicity. Several studies have demonstrated hepatotoxicity in experimental animals. The 

best study of long term toxicity of MDA was the report by Lamb et al. (1986). In addition to 

addressing the carcinogenicity of MDA, Lamb described non-cancer health effects, which 

resulted from lifetime exposure of two species, rats and mice, to MDA at two concentrations in 

the drinking water. The 150 ppm dose level was a LOAEL for fatty change and focal cellular 

Appendix D3 403 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

change to the livers of male and female rats as well as for liver degeneration in male mice. The 

corresponding effects were also observed in high-dose male rats and male mice. Nephropathy 

was observed in mice of both sexes at the 150 and 300 ppm. There is abundant evidence from 

both human and animal studies that MDA is hepatotoxic. Bastian (1984), Williams et al. (1974), 

and McGill and Motto (1974) reported hepatitis in people exposed by inhalation and dermal 

absorption routes. Kopelman et al. (1966a,b) demonstrated human hepatotoxicity from exposure 

by the oral route. However, limited data detailing exposure levels associated with adverse health 

effects in humans preclude the development of a chronic REL from studies in humans. 

The other toxic effect of potential concern from MDA exposure is ocular toxicity. Leong et al.

(1987) reported damage to the retinas of guinea pigs exposed for 2 weeks to MDA aerosols (0.44 

g/m3

 for 4 hr/day, 5 days/week; average experimental exposure = 52 mg/m3) by inhalation. 

Schilling von Canstatt et al. (1966) also reported blindness in cats treated orally with MDA. A 

single case of retinopathy and visual toxicity in humans was reported in a man who accidentally 

ingested MDA with potassium carbonate and γ-butyrolactone. The Leong et al. (1986) study 

was selected for the development of the chronic REL because, although conducted for a 

relatively short period of time, the study appears to address the most sensitive endpoint of 

toxicity by the most appropriate route of exposure (inhalation). The studies, which established 

the hepatotoxicity of MDA, were conducted by the oral route of exposure. 

As a comparison with the proposed REL, the study by Lamb et al. (1986) found a LOAEL of 9 

mg/kg-day for liver changes in male rats. Use of a LOAEL UF of 3, an interspecies UF of 10, 

and an intraspecies UF of 10 results in an oral chronic REL of 0.03 mg/kg-day. Use of route-toroute extrapolation with the assumption that a 70 kg person breathes 20 m3

 of air per day leads to 

an inhalation chronic REL estimate of 100 μg/m3. The proposed chronic REL based on Leong et

al. (1987) is lower by a factor of 5 than that obtaimed by using Lamb et al. (1986) and should be 

protective of hepatotoxicity. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for 4,4′-methylene dianiline include the availability of a 

controlled exposure inhalation study. Major areas of uncertainty are the lack of adequate human 

exposure data, the lack of chronic inhalation exposure studies, the lack of reproductive and 

developmental toxicity studies, and the lack of observation of a NOAEL. In addition the test 

animals were under additional stress due to the restraint used to obtain nose-only exposure, while 

the control animals were not restrained. Liver toxicity has been included as a potential critical 

effect because of uncertainty regarding the relative potency of this compound in causing liver 

toxicity in different species by different routes of exposure.

When assessing the health effects of methylene dianiline, its carcinogenicity must also be 

assessed. 

VIII. Potential for Differential Impacts on Children's Health

No evidence to support a differential effect of methylene dianiline on infants and children was 

found in the literature. 

Appendix D3 404 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IX. References 

ACGIH. 1992. American Conference of Governmental Industrial Hygienists, Inc. 

Documentation of the threshold limit values and biological exposure indices. Sixth ed. 

Cincinnati, OH: ACGIH. 

Bastian PG. 1984. Occupational hepatitis caused by methylenedianiline. Med. J. Aust. 141:533-

535. 

Brooks LJ, Neale JM, and Pieroni DR. 1979. Acute myocardiopathy following tripathway 

exposure to methylenedianiline. JAMA 242:1527-1528. 

Bruynzeel DP, and van der Wegen-Keijser MH. 1993. Contact dermatitis in a cast technician. 

Contact Dermatitis 28:193-194. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

de Pablo P, Ortiz J, Borrego L, Romero G, and Iglesias L. 1992. Allergic contact dermatitis from 

diaminodiphenylmethane in an ostomy bag. Contact Dermatitis 27:260. 

Deichmann WB, MacDonald WE, Coplan M, Woods F, and Blum E. 1978. Di-(4-aminophenyl)-

methane (MDA): 4-7 year dog feeding study. Toxicology 11:185-188. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Kopelman H. 1968. The Epping Jaundice after two years. Postgrad. Med. J. 44:78-81. 

Kopelman H, Robertson MH, Sanders PG, and Ash I. 1966a. The Epping Jaundice. Br. Med. J. 

1:514-516. 

Kopelman H, Scheuer PJ, and Williams R. 1966b. The liver lesion of the Epping Jaundice. 

Quebec Journal of Medicine 35:553-64. 

Lamb JC, Huff JE, Haseman JK, Murthy ASK, and Lilja H. 1986. Carcinogenesis studies of 

4,4’-methylenedianilinedihydrochloride given in drinking water to F344/N rats and B6C3F1

mice. J. Toxicol. Environ. Health 18:325-337. 

Leong BKJ, Lund JE, Groehn JA, Coombs JK, Sabatis CP, Weaver RJ, and Griffin RL. 1987. 

Retinopathy from inhaling 4,4’-methylenedianiline aerosols. Fundam. Appl. Toxicol. 9:645-658. 

Appendix D3 405 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

LeVine MJ. 1983. Occupational photosensitivity to diaminodiphenylmethane. Contact 

Dermatitis 9:488-490. 

McGill DB, and Motto JD. 1974. An industrial outbreak of toxic hepatitis due to 

methylenedianiline. N. Engl. J. Med. 291:278-282. 

Pludro G, Karlowski K, Mankowska M, Woggon H, and Uhde W-J. 1969. Toxicological and 

chemical studies of some epoxy resins and hardeners. I. Determination of acute and subacute 

toxicity of phthalic acid anhydride, 4,4’-diaminodiphenylmethane and of the epoxy resin: Epilox 

EG-34. Acta Pol. Pharmacol. 26:352-357. 

Roy CW, McSorley PD, and Syme IG. 1985. Methylene dianiline: a new toxic cause of visual 

failure with hepatitis. Hum. Toxicol. 4:61-66. 

Schilling von Canstatt B, Hofmann HT, Oettel H, and Zeller H. 1966. [Netzhautveränderungen 

der Katze bei der Vergiftung mit peroral oder perkutan verabreichten Chemikalien]. 

Verhandlungen der Deutschen Gesellschaft für Pathologie 50:429-535. 

Schoental R. 1968. Pathological lesions, including tumors, in rats after 4,4’-

diaminodiphenylmethane and gamma-butyrolactone. Isr. J. Med. Sci. 4:1146-58. 

U.S.EPA. 1994. United States Environmental Protection Agency. Methods for derivation of 

inhalation reference concentrations and applications of inhalation dosimetry. EPA/600/8-

90/066F. Office of Research and Development. Washington DC: U.S.EPA. 

Van Joost T, Heule F, and de Boer J. 1987. Sensitization to methylenedianiline and parastructures. Contact Dermatitis 16:246-248. 

Williams SV, Bryan JA, Burk JR, and Wolf FS. 1974. Toxic hepatitis and methylenedianiline 

[letter]. N. Engl. J. Med. 291:1256. 

Appendix D3 406 Methylene dianiline 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

METHYLENE DIPHENYL ISOCYANATE 

(diphenylmethane diisocyanate) 

CAS Registry Number: 101-68-8 

I. Chronic Reference Exposure Level 

Inhalation reference exposure level 0.7 μg/m3

Critical effect(s) Hyperplasia of the olfactory epithelium in rats 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1995)

Description Light yellow solid 

Molecular formula C15H10N2O2 (monomer) 

Molecular weight Variable (monomer = 250.27 g/mol) 

Density 1.197 g/cm3

 @ 70°C (monomer) 

Boiling point 196°C (monomer) 

Melting point 37°C (monomer) 

Vapor pressure 0.001 torr @ 40°C (monomer) 

Solubility Soluble in acetone, benzene, kerosene, and 

nitrobenzene (monomer) 

Conversion factor Monomer: 1 ppm = 10.2 mg/m3

 at 25°C; 

 Not applicable for polymer 

III. Major Uses or Sources 

Methylene diphenyl isocyanate (MDI) is used for bonding rubber to nylon. MDI is also used in 

the manufacture of lacquer coatings and in the production of polyurethane resins and spandex 

fibers (HSDB, 1995). It is often handled in a partially polymerized form (“MDI polymer”), 

which has a much lower vapor pressure than the monomer. The annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California based on the 

most recent inventory were estimated to be 30,398 pounds of MDI (CARB, 2000). 

IV. Effects of Human Exposure 

A 5-year occupational study of 107 workers from a polyurethane plastic manufacturing plant 

examined pulmonary function, respiratory symptoms, and smoking habits (Musk et al., 1982, 

1985). No significant changes in pulmonary function or respiratory symptoms were observed 

Appendix D3 407 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

when controlled for smoking. Mean MDI concentrations measured ranged from 0.0003 to 

0.0006 ppm. 

Significantly increased prevalence of asthma in female workers and of chronic bronchitis in male 

and female workers was observed following occupational exposure to low levels of MDI 

(<0.02 ppm) (Pham et al., 1988). Workers from two plants were grouped by job classification 

and evaluated in this study conducted in 1976; workers were grouped as unexposed (62 men, 21 

women), indirectly exposed (61 men, 56 women), or directly exposed (91 men, 27 women). 

Further characterization of the exposure groups was not presented. Decrements in pulmonary 

function (measured by VC, FEV1 and single-breath carbon monoxide diffusion tests) were 

observed in men in the direct and indirect exposure groups: decrements in men with a history of 

direct exposure to MDI were statistically significant. Workers were also grouped by duration of 

occupational exposure (<20 months, 20-60 months, >60 months). Workers with known (direct 

or indirect) occupational exposure to MDI for greater than 60 months exhibited statistically 

significant decrements in pulmonary function tests. The follow-up examination of this study 

describes data from male workers only. At the time of the 5-year follow-up, air levels had been 

reduced to below the maximum allowed air concentration of 0.005 ppm by a modification of the 

ventilation system. Statistically significant decrements in pulmonary function were observed 

again in workers with known direct occupational exposure to MDI. Workers who were exposed 

at the time of the 1976 study but had since been removed from exposure did not exhibit 

decrements in pulmonary function, leading the authors to conclude that the effects of low-level 

exposure to MDI are to some extent reversible. Flaws in study design, including lack of 

exposure characterization, attrition, and inclusion of asthmatics in cohorts, preclude a 

quantitative assessment of MDI exposure on lung function. 

An epidemiologic study of foundry workers reported more respiratory symptoms and 

significantly lower mean FEV1 and maximum mid-expiratory flow at 25-75% in exposed 

workers compared to controls (Johnson et al., 1985). However, MDI-exposed workers also had 

unquantified exposure to silica, metal dust, phenol formaldehyde, and a pyridine derivative 

precluding the evaluation of respiratory effects resulting from MDI exposure. 

A worker with 5 years occupational exposure and suspected MDI hypersensitivity was exposed 

continuously in a controlled chamber to 5 ppb for 15 minutes, then 10 ppb for 30 minutes, and 

20 ppb for 15 minutes (Marczynski et al., 1992). The worker had not been exposed to MDI in 

the workplace for 5 days prior to the test challenge. Exposure to MDI resulted in an immediate, 

moderate, asthmatic reaction associated with significant hypoxemia. 

IgG antibodies recognizing MDI-human serum albumin conjugates were detected in 4 of 5 MDIexposed workers (Aul et al., 1999). The levels of specific IgG antibodies were more elevated 

with polymeric MDI compared with monomeric MDI. 

A workplace death of a 39-year-old foundry worker was ascribed to occupational asthma 

induced by MDI exposure (Carnio et al., 1997). Postmortem pulmonary findings included 

epithelial desquamation, mucosal eosinophilic/neutrophilic infiltration, bronchial vessel 

dilatation, and edema and hypertrophy of smooth muscle. 

Appendix D3 408 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

V. Effects of Animal Exposure 

Rats were exposed to 0.2, 1.0, and 6.0 mg/m3 aerosolized MDI polymer 6 hours per day, 5 days 

per week for 24 months (Reuzel et al., 1990; 1994). Statistically significant increased incidences 

of basal cell hyperplasia, olfactory epithelial degeneration, alveolar duct epithelialization, 

localized alveolar bronchiolization, and adenomas were observed in male and female rats 

exposed to 6.0 mg/m3

 MDI. An accumulation of macrophages with yellow pigment was also 

noted in the lungs and mediastinal lymph nodes. Male rats exposed to this concentration also 

exhibited a statistically significant increase in the incidence of Bowman’s gland hyperplasia. 

Male rats exposed to 1 mg/m3

 MDI also exhibited statistically significant increased incidences of 

basal cell hyperplasia and Bowman’s gland hyperplasia. An accumulation of macrophages with 

yellow pigment was observed in the lungs of female rats and the lungs and mediastinal lymph 

nodes of male rats exposed to 1 mg/m3

. No adverse effects were noted in rats exposed to 

0.2 mg/m3

 MDI. 

Hyperplasia of the olfactory epithelium with MDI exposure (Reuzel et al., 1990; 1994) 

Males Females Combined 

Concentratio

n 

(mg/m3) 

Responders N Incidence Responders N Incidence Responders N Incidence 

0 14 60 0.23 4 60 0.067 18 120 0.15 

0.2 13 60 0.22 8 60 0.13 21 120 0.18 

1 26 60 0.43 8 60 0.13 34 120 0.28 

6 32 60 0.53 49 59 0.83 81 119 0.68 

Guinea pigs were exposed to 2 ppm MDI 3 hours per day for 5 days (Aizicovici et al., 1990). 

Qualitative immunostaining techniques indicated that MDI was localized in the respiratory tract. 

The spleen, lymph nodes, and thymus had very little staining. However, another study exposed 

guinea pigs to 4 ppb radiolabelled toluene diisocyanate (TDI) for 1-hour and found measurable 

radioactivity in extrathoracic tissues and body fluids (Kennedy et al., 1989). Therefore, there is 

a possibility that MDI may be transported to sites other than the respiratory tract, such as the 

ovaries and testes, following inhalation exposure. 

Gravid Wistar rats, Crl:(WI)BR, were exposed by whole-body inhalation to clean air (control) 

and to 1, 3, and 9 mg/m3

 MDI, respectively, for 6 hr per day from days 6 to 15 post conception 

(Buschmann et al., 1996). Rats were killed on day 20. The lung weights in the high-dose group 

were significantly increased compared to the sham-treated control animals. Treatment did not 

influence any other maternal and/or fetal parameters investigated (including maternal weight 

gain, number of corpora lutea, implantation sites, pre-and postimplantation loss, fetal and 

placental weights, gross and visceral anomalies, and degree of ossification). A slight but 

Appendix D3 409 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

significant increase in litters with fetuses displaying asymmetric sternebra(e) was observed after 

treatment with the highest dose. Although the relevance of an increase of this minor anomaly in 

doses which maternal toxicity is limited and within the limits of biological variability, a 

substance-induced effect in the high-dose group cannot be excluded with certainty. Thus, the 

authors reported a NOAEL of 3 mg/m3

 for embryotoxic effects. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Reuzel et al., 1990; 1994 

Study population Rats 

Exposure method Inhalation of polymeric aerosolized MDI (0, 0.2, 

1.0, and 6.0 mg/m3) 

Critical effects Hyperplasia of the olfactory epithelium 

LOAEL 1 mg/m3

NOAEL 0.2 mg/m3

Benchmark Concentration (BMC05) 0.25 mg/m3

 (95% lower confidence limit on 

concentration for a 5% incidence of response 

based on analysis of the combined male and 

female data with a linear model, the bestfitting of 6 models examined, p = 0.99) 

Study continuity 6 hours per day, 5 days per week 

Study duration 24 months 

Average experimental exposure 0.046 mg/m3

 for BMC05 group 

(0.25 x 6/24 x 5/7) 

Human equivalent concentration 0.020 mg/m3

 for BMC05 group (particle with 

extrathoracic respiratory effects, 

RDDR = 0.453, based on MMAD = 0.68 μm 

and sigma g = 2.93) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference concentration 0.7 μg/m3

The data of Reuzel et al. (1990, 1994) were examined with six quantal dose-response models 

(linear, log-normal, Weibull, logistic, quadratic, gamma) using USEPA BMDS 1.2. All models 

except the quadratic gave a good fit to the combined male and female data set. The linear model 

was selected as the best-fitting model. Possible differences between male and female 

susceptibility are suggested by the gender-specific data, although the significance of these 

differences is uncertain. 

USEPA used the same two studies and a BMC10 approach to develop an RfC of 0.6 μg/m3. Since 

USEPA used a 3-fold database uncertainty factor, their BMC10-based RfC is comparable to the 

BMC05-based OEHHA REL. 

Appendix D3 410 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

VII. Data Strengths and Limitations for Development of the REL 

Strengths of the REL for MDI include the use of a well-conducted, long-term inhalation study, 

the observation of a NOAEL, and the estimation of a benchmark concentration. A limitation of 

the REL is that it is based on data on exposures to MDI “polymer” which actually contains 

nearly 50% monomer. Monomers may in some cases be more toxic than polymers. Thus, 

effects of pure monomeric MDI may occur at concentrations somewhat lower than observed in 

the reported study on MDI polymer. However, the capacity of MDI polymer to induce 

immunologic sensitization is greater than that of MDI monomer (Aul et al., 1999). The relative 

potential of MDI monomer and polymer to induce hyperplasia of the olfactory epithelium is 

unknown. 

VIII. References 

Aizicovici S, Jin R, LaPietra D, Gottlieb F, and Karol MH. 1990. Use of immunohistochemistry 

to detect diphenyl methane 4,4-diisocyanate (MDI) in exposed guinea pigs. Toxicologist 

10(1):286. (Abstr. #1144). [cited in U.S. EPA, 1995]. 

Aul DJ, Bhaumik A, Kennedy AL, Brown WE, Lesage J, Malo JL. 1999. Specific IgG response 

to monomeric and polymeric diphenylmethane diisocyanate conjugates in subjects with 

respiratory reactions to isocyanates. J. Allergy Clin. Immunol. 103(5 Pt 1):749-755. 

Buschmann J, Koch W, Fuhst R, and Heinrich U. 1996. Embryotoxicity study of monomeric 

4,4'-methylenediphenyl diisocyanate (MDI) aerosol after inhalation exposure in Wistar rats. 

Fundam. Appl. Toxicol. 32(1):96-101. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Carino M, Aliani M, Licitra C, Sarno N, and Ioli F. 1997. Death due to asthma at workplace in a 

diphenylmethane diisocyanate-sensitized subject. Respiration 64(1):111-113. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 1/31/95). 

Johnson A, Chan-Yeng M, MacLean L, Atkins E, Dybuncio A, Cheng F, and Enarson D. 1985. 

Respiratory abnormalities among workers in an iron and steel foundry. Br. J. Ind. Med. 42(2):94-

100. [cited in U.S. EPA, 1995]. 

Kennedy AL, Stock MF, Alarie Y, Brown WE. 1989. Uptake and distribution of 14C during and 

following inhalation exposure to radioactive toluene diisocyanate. Toxicol. Appl. Pharmacol. 

100:280-292. [cited in U.S. EPA, 1995]. 

Appendix D3 411 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Marczynski B, Czuppon AB, Hoffarth HP, Marek W, and Baur X. 1992. DNA damage in human 

white blood cells after inhalative exposure to methylenediphenyl diisocyanate (MDI) - case 

report. Toxicol. Lett. 60:131-138. 

Musk AW, Peters JM, and Berstein L. 1985. Absence of respiratory effects in subjects exposed 

to low concentrations of TDI and MDI: A reevaluation. J. Occup. Med. 27(12):917-920. 

Musk AW, Peters JM, DiBerardinis L, and Murphy RLH. 1982. Absence of respiratory effects in 

subjects exposed to low concentrations of TDI and MDI. J. Occup. Med. 24(10):746-750. 

Pham QT, Teculescu D, Meyer-Bisch C, and Mur JM. 1988. Effects of chronic exposure to 

diisocyantes. Bull. Eur. Physiopathol. Respir. 23:561-564. 

Reuzel PGJ, Arts JHE., Lomax LG, Kuijpers MHM, Kuper CF, Gembardt C, Feron VJ, and 

Loser E. 1994. Chronic inhalation toxicity and carcinogenicity study of respirable polymeric 

methylene diphenyl diisocyanate (polymeric MDI) aerosol in rats. Fundam. Appl. Toxicol. 

22:195-210. 

Reuzel PGJ, Arts JHE, Kuijpers MHM, and Kuper CF. 1990. Chronic toxicity/carcinogenicity 

inhalation study of polymeric methylene diphenyl diisocyanate aerosol in rats (Final report). 

Prepared by Civo Institute for the International Isocyanate Institute. Report No. V88.122. [cited 

in U.S. EPA, 1995]. 

U.S. EPA. 2000. U. S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS). Office of Health and Environmental Assessment, Environmental Criteria Assessment 

Office, Cincinnati, OH: U.S. EPA. Available online at http://www.epa.gov/ngispgm3/iris 

Appendix D3 412 Methylene diphenyl isocyanate 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

NAPHTHALENE 

 (naphthene, NCI-C5290, albocarbon, dezodorator, moth balls, 

moth flakes, tar camphor, white tar, naphthalin, naphthaline) 

CAS Registry Number: 91-20-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 9 μg/m3

 (2 ppb) 

Critical effect(s) Respiratory effects (nasal inflammation, 

 olfactory epithelial metaplasia, respiratory 

 epithelial hyperplasia) in mice 

Hazard index target(s) Respiratory system, blood systems 

II. Physical and Chemical Properties (HSDB, 1995; 1999 except as noted) 

 

Description White crystalline powder; odor of mothballs 

Molecular formula C10H8

Molecular weight 128.6 g/mol 

Density 4.42 g/cm3

 @ 20oC 

Boiling point 218oC 

Melting point 80.5 oC 

Vapor pressure 0.078 Torr @ 25oC (Sonnenfeld et al., 

 1983); 0.10 Torr @ 27oC (CRC, 1994) 

Conversion factor 5.26 μg/m3

 per ppb at 25oC

III. Major Uses or Sources 

Naphthalene is a natural constituent of coal tar (approximately 11%) (HSBD, 1995). It is present 

in gasoline and diesel fuels. Naphthalene is used as a moth repellent, though this use is 

decreasing in favor of p-dichlorobenzene (HSDB, 1995). It has also been used in the 

manufacture of phthalic anhydride, phthalic and anthranilic acids, naphthols, naphthylamines, 1-

naphthyl-n-methylcarbamate insecticide, beta-naphthol, naphthalene sulfonates, synthetic resins, 

celluloid, lampblack, smokeless powder, anthraquinone, indigo, perylene, and hydronaphthalenes 

(NTP, 1992; HSDB, 1995). The statewide emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California, based on the most recent inventory, were estimated to be 

164,459 pounds of naphthalene (CARB, 1999). 

Appendix D3 413 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure 

Nine persons (eight adults and one child) were exposed to naphthalene vapors from several 

hundred mothballs in their homes. Nausea, vomiting, abdominal pain, and anemia were reported 

(Linick, 1983). Testing at one home following the incident indicated an airborne naphthalene 

concentration of 20 ppb (105 μg/m3

). Symptoms abated after removal of the mothballs. 

Workers occupationally exposed to naphthalene fumes or dust for up to five years were studied 

for adverse ocular effects (Ghetti and Mariani, 1956). Multiple pin-point opacities developed in 

8 of 21 workers. Vision did not appear to be impaired. 

Cataracts and retinal hemorrhage were observed in a 44 year old man occupationally exposed to 

powdered naphthalene, and a coworker developed chorioretinitis (van der Hoeve, 1906). 

Wolf (1978) reported that a majority of 15 persons involved in naphthalene manufacture 

developed either rhinopharyngolaryngitis and/or laryngeal carcinoma. 

Ingestion of naphthalene or p-dichlorobenzene mothballs is a frequent cause of accidental 

poisoning of children (Siegel and Wason, 1986). Infants exposed to naphthalene vapors from 

clothes or blankets have become ill or have died (U.S. EPA, 1990). The effects in infants have 

been associated with maternal naphthalene exposure during gestation (U.S. EPA, 1990). 

Deaths have been reported following ingestion of naphthalene mothballs. A 17-year old male 

ingested mothballs, developed gastrointestinal bleeding, hematuria, and coma, and died after five 

days (Gupta et al., 1979). A 30-year old female ingested 30 mothballs and died after five days 

(Kurz, 1987). 

Acute hemolytic anemia was reported among 21 infants exposed to naphthalene vapors from 

nearby mothball-treated materials (Valaes et al., 1963). Increased serum bilirubin, 

methemoglobin, Heinz bodies, and fragmented red blood cells were observed. Kernicterus was 

noted in eight of the children, and two of the children died. Ten of these children had a genetic 

deficiency in glucose-6-phosphate dehydrogenase. 

A 12-year old male ingested 4 g of naphthalene and 20 hours later developed hematuria, anemia, 

restlessness, and liver enlargement (Manchanda and Sood, 1960). The patient recovered after 8 

days. 

A 69-year old female developed aplastic anemia two months after several weeks exposure to 

naphthalene and p-dichlorobenzene (Harden and Baetjer, 1978). 

Appendix D3 414 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

V. Effects of Animal Exposure 

Male and female B6C3F1 mice were exposed to naphthalene (>99% pure) vapor for 6 hours per 

day, 5 days per week over 104 weeks (NTP, 1992). Concentrations used were 0 (150 mice), 10 

(150 mice), or 30 ppm (300 mice) naphthalene. (Table 1). Lesions were observed in the nose 

and lungs of exposed mice, including increased incidences of chronic nasal inflammation, 

olfactory epithelial metaplasia, and respiratory epithelial hyperplasia. 

Table 1. Incidence of respiratory tract lesions in mice (male and female combined) chronically 

exposed to naphthalene vapors (NTP, 1992). 

0 ppm 10 ppm 30 ppm 

Nasal inflammation 3/139 34/134 108/270 

Olfactory epithelial 

metaplasia 

0/139 131/134 269/270 

Respiratory epithelial 

hyperplasia 

0/139 131/134 269/270 

CD-1 mice were administered 5.3, 53, or 133 mg/kg/day naphthalene by gavage over 90 days 

(Shopp et al., 1984). The only effect noted was inhibition of aryl hydrocarbon hydroxylase 

activity. No increase in mortality or changes in body weight were noted. Reduced spleen 

weights were noted in females exposed to the highest dose. No changes were noted in serum 

enzyme levels or electrolytes. The researchers did not conduct a histopathological examination. 

B6C3F1 mice were administered 200 mg naphthalene/kg/day by gavage for 5 days per week 

over 13 weeks. No adverse effects were observed (U.S. EPA, 1990). 

Developmental effects of naphthalene ingestion in Sprague-Dawley CD rats was studied by 

Navarro and associates (1991). The lowest dose tested (50 mg/kg/day by gavage) was associated 

with signs of CNS depression for the first 3 days. Fetal growth, survival, and morphological 

development were not significantly affected at 450 mg/kg/day compared with control animals, 

although a trend toward decreased fetal weight and increased malformations was observed. 

Harris and associates (1979) intraperitoneally administered 395 mg/kg/day naphthalene to 

Sprague-Dawley rats over days 1 though 15 of gestation. Fetuses had a 50% increase in 

incidence in delayed cranial ossification and heart development. 

New Zealand white rabbits were given 0, 40, 200, or 400 mg/kg/day by gavage over days 6 

through 18 of gestation (U.S. EPA, 1986a). A dose-dependent increase in grooming, 

vocalization, aggression, diarrhea, dyspnea, and ocular and nasal discharge were noted at all 

doses. No statistically significant increase in malformations or developmental abnormalities was 

observed. 

Appendix D3 415 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Sprague-Dawley rats were administered 0, 100, 300, or 1000 mg/kg/day of naphthalene via 

dermal application (U.S. EPA, 1986b). No effects were reported at 100 or 300 mg/kg/day. At 

the high dose a slight decrease in testes weight was noted. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study NTP (1992) 

Study population B6C3F1 mice (75 or 150/group/sex)

Exposure method Discontinuous whole-body inhalation exposures 

 to 0, 10, or 30 ppm naphthalene vapor

Critical effects Nasal inflammation, olfactory epithelial 

 metaplasia, and respiratory epithelial 

 hyperplasia 

LOAEL 10 ppm (96% incidence for males and 100% 

 incidence for females) 

NOAEL Not observed 

Exposure continuity 6 hours/day for 5 days/week 

Average experimental exposure 1.8 ppm (10 ppm x 6/24 x 5/7) for LOAEL group 

Exposure duration 104 weeks 

Subchronic uncertainty factor 1 

LOAEL uncertainty factor 10 

Interspecies uncertainty factor 10 (see below) 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 1000 

Inhalation reference exposure level 0.002 ppm (2 ppb, 0.009 mg/m3

, 9 μg/m3) 

The NTP study was chosen for the REL derivation since it is the only available lifetime animal 

inhalation bioassay and because no adequate epidemiological studies of long-term human 

exposure are available. The study was judged to be of adequate study design. The complete lack 

of nasal effects among control animals and the nearly total effect among animals exposed at 2 

different concentrations strongly indicates a causal relationship between naphthalene exposure 

and nasal effects. The effects seen are consistent with those reported among exposed workers, 

who developed rhinopharyngolaryngitis or laryngeal carcinoma (Wolf, 1978). However, the 

hematological effects observed in humans have not been reported in laboratory animals, which 

raises the possibility that humans may be significantly more sensitive to naphthalene. 

The most important limitation of the study is that the lowest concentration tested caused adverse 

effects in most (≥96%) of the animals tested. Thus the study amply demonstrates the risk of 

lifetime exposures to 10 ppm, but is uninformative regarding the concentration-response 

relationship at lower concentrations. Only a general assumption can be drawn on the magnitude 

of uncertainty factor needed to predict a concentration at which adverse effects would most 

likely not be observed. Lacking specific guidance or relevant research for this situation, the 

default 10-fold factor was applied. U.S. EPA also used the NTP study to develop its RfC of 3 

μg/m3

 with slightly different assumptions and a cumulative uncertainty factor of 3000 (U.S. 

EPA, 2000). OEHHA followed the U.S. EPA precedent in using an intraspecies UF of 10 for 

Appendix D3 416 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

naphthalene, rather than using the HEC/RGDR approach. According to U.S. EPA (2000), 

because of its low water solubility and low reactivity, naphthalene-related effects on the nasal 

epithelium are expected to result following absorption of naphthalene and its metabolism to 

reactive oxygenated metabolites, not from direct contact. This is supported by data on 

naphthalene metabolism indicating that toxic effects on the respiratory tract are due to a 

naphthalene metabolite that may be formed either in the liver or in the respiratory tract. Necrosis 

of bronchial epithelial (Clara) cells in mice and necrosis of olfactory epithelium in mice, rats, 

and hamsters occur following intraperitoneal injection of naphthalene. The nasal effects from 

inhalation exposure to naphthalene were considered to be extra-respiratory effects of a category 

3 gas (U.S. EPA, 1994). The assumption is made that nasal responses in mice to inhaled 

naphthalene are relevant to humans; however, it is uncertain that the RfC for naphthalene based 

on nasal effects will be protective for hemolytic anemia and cataracts, the more well-known 

effects from naphthalene exposure in humans. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the REL for naphthalene include the large number of animals in the key study 

on which the REL is based and the 2 year length of the study. The limitations include the very 

high incidence of lesions at the lowest level tested in the key study, the absence of a NOAEL in 

the key study, the absence of other animal studies by the inhalation route, and the paucity of 

human data. 

VIII. References 

ATSDR. 1990. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Naphthalene and 2-Methylnaphthalene (Draft). Atlanta, GA: U.S. Public Health Service, U.S. 

Department of Health and Human Services. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. p. 6-109. 

Ghetti G, and Mariani L. 1956. [Eyes changes due to naphthalene]. Med. Lav. 47:533-538. 

[reviewed in ATSDR, 1995]. 

Gupta R, Singhal PC, Muthusethupathy MA, Malik AK, Chugh KS. 1979. Cerebral oedema and 

renal failure following naphthalene poisoning. J. Assoc. Physic. India 27:347-348. 

Harden RA, and Baetjer AM. 1978. Aplastic anemia following exposure to p-dichlorobenzene 

and naphthalene. J. Occup. Med. 20: 820-822. 

Harris J, Bond GP, and Niemeier RW. 1979. The effects of 2-nitropropane, naphthalene, and 

hexachlorobutadiene on fetal rat development. [Abstract]. Toxicol. Appl. Pharmacol. 48:A35. 

Appendix D3 417 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/95). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Kurz JM. 1987. Naphthalene poisoning. Critical care nursing techniques. Dimens. Crit. Care 

Nurs. 6:264-270. 

Linick M. 1983. Illness associated with exposure to naphthalene in mothballs - Indiana. MMWR 

32:34-35. 

Manchanda SS, and Sood SC. 1960. Accidental poisoning in children: with a case report of 

naphthalene poisoning. Ind. J. Child Health 9(2):113-119. 

Navarro H, Proce C, Marr M, and Myers C. 1991. Developmental toxicity evaluation of 

naphthalene administered by gavage to Sprague-Dawley rats on gestational days 6 through 15. 

Research Triangle Institute. Research Triangle Park, NC: National Toxicology Program. 

NTP. 1992. Toxicology & Carcinogenesis Studies of Naphthalene in B6C3F1 Mice. Technical 

Report Series No. 410. NIH Publication No. 92-3141. 

Shopp GM, White KL, and Hosapple M. 1984. Naphthalene toxicity in CD-1 mice: General 

toxicology and immunotoxicology. Fundam. Appl. Toxicol. 4:406-429. 

Siegel E, and Wason S. 1986. Mothball toxicity. Pediatr. Clin. North Am. 33:369-374. 

Sonnenfeld WJ, Zoller WH, and May WE. 1983. Dynamic coupled-column liquid 

chromatographic determination of ambient temperature vapor pressures of polynuclear aromatic 

hydrocarbons. Anal. Chem. 55(2):275-280. 

U.S.EPA. 1986a. U.S. Environmental Protection Agency. Developmental toxicity study in 

rabbits. PH 329-TX-001-85. 86-870000563. Office of Research and Development, Cincinnati, 

OH. 

U.S.EPA. 1986b. U.S. Environmental Protection Agency. Ninety day (sub-chronic) dermal 

toxicity study with naphthalene in albino rats. 65-870000565. Cincinnati, OH: Office of 

Research and Development. 

U.S.EPA. 1990. U.S. Environmental Protection Agency. Drinking Water Health Advisories for 

15 Volatile Organic Chemicals. U.S. EPA/ODW. NTIS No. PB90-259821. 

U.S. EPA. 1994. U.S. Environmental Protection Agency Methods for Derivation of Inhalation 

Reference Concentrations and Application of Inhalation Dosimetry. EPA/600/8-90/066F. Office 

of Research and Development. Washington, DC: U.S.EPA. 

Appendix D3 418 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

U. S. EPA. 2000. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS). Naphthalene. Available on-line at http://www.epa.gov/ngispgm3/iris

Valaes T, Doxiadis SA, and Fessas T. 1963. Acute hemolysis due to naphthalene inhalation. J. 

Pediatr. 63:904-915. 

van der Hoeve, J. 1906. [Chorioretinitis in humans from the effects of naphthalene] [in German]. 

Arch. Augenheilkd. 56:259-262. [reviewed in ATSDR, 1995]. 

Wolf O. 1978. Cancer of the larynx in naphthalene cleaner. Z. Gesamte. Hyg. 24(10):737-739. 

Appendix D3 419 Naphthalene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

NICKEL AND NICKEL COMPOUNDS: 

NICKEL OXIDE 

Molecular 

Formula 

Molecular 

Weight Synonyms 

CAS Registry Number 

Ni 59 elemental nickel 7440-02-0 

NiO 74.69 nickel oxide 1313-99-1 

NiCl2 129.6 

nickel chloride 

nickel dichloride 7718-54-9 

NiSO4 154.75 

nickel sulfate 

nickelous sulfate 7786-81-4 

NiCO3 118.7 

nickel carbonate 

carbonic acid nickel salt 3333-67-3 

Ni3S2 240.19 

nickel subsulfide 

trinickel disulfide 

heazlewoodite 12035-72-2 

I. Chronic Toxicity Summary 

A. Nickel and Nickel Compounds (except nickel oxide) 

Inhalation reference exposure level 0.05 μg Ni/m3

Critical effect(s) Lung, nasal epithelial and lymphatic pathology 

in male and female rats 

Hazard index target(s) Respiratory system; hematopoietic system 

B. Nickel Oxide 

Inhalation reference exposure level 0.10 μg Ni/m3

Critical effect(s) Lung and lymphatic pathology in male and 

female rats 

Hazard index target(s) Respiratory system; hematopoietic system 

Appendix D3 420 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

II. Physical and Chemical Properties (from HSDB, 1995) 

Description Ni metal: Silvery metal; NiCl2: deliquescent 

crystals (U.S.EPA, 1985) 

Molecular formula See above 

Molecular weight See above 

Density 8.9 g/cm3

 @ 20°C (Ni) 

Boiling point 2730°C (Ni) 

Vapor pressure Not applicable 

Solubility Elemental nickel, nickel subsulfide, and nickel 

oxide are insoluble in water, but are soluble 

in dilute nitric, hydrochloric, and sulfuric 

acids. The chloride and sulfate forms of 

nickel are water soluble. 

Conversion factor Not applicable for fumes and dusts 

III. Major Uses and Sources 

The most common airborne exposures to nickel compounds are to insoluble nickel compounds 

such as elemental nickel, nickel sulfide, and the nickel oxides from dusts and fumes. 

Contributions to nickel in the ambient air are made by combustion of fossil fuels, nickel plating, 

and other metallurgical processes. The most common oxidation state of nickel is the divalent 

(Ni2+) form (U.S.EPA, 1985). Elemental nickel is a malleable, silvery-white metal that is highly 

resistant to strong alkali. Because of its corrosion resistance, nickel is used in the production of 

stainless steel, permanent magnets, and other alloys that require resistance to extremes of 

temperature or stress (U.S.EPA, 1985). Nickel is also used in electroplating baths, batteries, 

textile dyes, and catalysts (U.S.EPA, 1985). Nickel dust or powder is flammable (CDTSC, 

1985). Due to its unique toxicological and physico-chemical properties, nickel carbonyl is not 

included in this summary. The annual statewide industrial emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California, based on the most recent inventory, were 

estimated to be 110,334 pounds of nickel (CARB, 1999). 

IV. Effects of Human Exposure

Several studies have indicated that occupational inhalation exposure to nickel aerosols can result 

in development of asthma specific to nickel. Davies (1986) found 3 cases of asthma among 53 

nickel-plating workers without a history of asthma prior to employment. Novey et al. (1983) 

described biphasic metal-specific bronchial responses in an individual metal-plating worker 

exposed to nickel and chromium salts. In another case, immunological studies conducted in a 

24-year old man showed nickel-specific antibodies in the serum after several weeks of working 

in a nickel-plating shop using nickel sulfate (McConnell et al., 1973). Dermatitis was observed 

on exposed areas of his skin, and pulmonary function, measured by FEV1 with and without 

isoproterenol challenge, was significantly impaired compared with a control subject and normal 

Appendix D3 421 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

values. Dyspnea, non-productive cough, chest-tightness, and wheezing were reported as 

symptoms by this worker during the work period. 

A group of 7 metal plating workers with occupational asthma were evaluated for atopy and 

pulmonary function challenge in response to inhalational challenge with nickel and other metals 

(Cirla et al., 1985). Three of the asthmatics tested positive for the presence of nickel-specific 

IgE antibodies. Positive reactions to skin testing with nickel were found in 3 of the asthmatic 

workers who also had dermatitis. Six out of the 7 asthmatics exhibited significantly decreased 

FEV1 (> 15%) when exposed to 0.3 mg/m3

 nickel sulfate for 30 minutes. Control challenges 

with other metal salts did not reveal similar deficits in FEV1. 

Although asthma has been described in the above studies, occupational inhalation of nickel dusts 

has not been found to be associated with pulmonary fibrosis (Muir et al., 1993). An 

occupational epidemiology report by Broder et al. (1989) found no significant effects on 

pulmonary function in relation to nickel exposure in a nickel smelter, however a healthy worker 

effect was observed in this study.

V. Effects of Animal Exposure 

Early studies on the chronic non-cancer effects of metallic nickel dust were complicated by early 

mortality and cancer in guinea pigs and rats (Hueper, 1958). 

A 2-year inhalation study of nickel oxide in rats and mice (65 per sex, per group) was conducted 

by the National Toxicology Program (NTP, 1994a). In the first study, rats were exposed to 0, 

0.62, 1.25, or 2.5 mg nickel oxide/m3

 (0, 0.5, 1.0, or 2.0 mg Ni/m3) 6 hours/day, 5 days/week for 

104 weeks. In addition to the carcinogenic effects of nickel oxide, a number of non-cancerous 

lesions were observed, particularly in the lungs. The incidence of inflammatory pigmentation in 

the alveoli was significantly greater in all exposed groups, compared to controls. The severity of 

the lesions reportedly increased with increasing exposure. Atypical alveolar hyperplasia was 

also seen in all exposed groups. Lymphoid hyperplasia in the bronchial lymph nodes was 

observed in males and females exposed to 1 mg Ni/m3

 or greater at 7 and 15 months and the 

incidence generally increased with increasing concentration at the end of the 2-year study. 

Females had an increased incidence of adrenal medullary hyperplasia at all exposures of nickel 

oxide. Body weights were significantly lower in the groups exposed to 2.0 mg Ni/m3

 for both 

sexes, and in males exposed to 1.0 mg Ni/m3. 

A companion study on nickel oxide in mice conducted by NTP showed similar lung 

inflammatory changes as seen in the rats, in addition to pigmentation of the alveolar region at all 

exposure concentrations, compared with controls (NTP, 1994a). The mice were exposed to 0, 

1.0, 2.0, or 3.9 mg Ni/m3. Bronchial lymph-node hyperplasia was also evident in all nickelexposed animals. Body weights were slightly but significantly lower in the 3.9 mg Ni/m3

 group, 

compared with controls. 

A continuous exposure of rats (20 - 40 per group) to 0, 60, or 200 μg Ni/m3

 as nickel oxide for 2 

years resulted in severe pulmonary damage and premature mortality so that carcinogenesis could 

not be evaluated (Glaser et al., 1986). Pulmonary alveolar proteinosis and septal fibrosis were 

Appendix D3 422 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

observed in the animals exposed to nickel. Only 1 rat per group survived the nickel exposures to 

the end of the experiment. 

A 2-year study on the effects of nickel subsulfide in rats and mice was conducted by NTP 

(1994b). Rats (52-53 per sex per group) were exposed to 0, 0.15, or 1 mg Ni3S2/m3 (0, 0.11, or 

0.73 mg Ni/m3) for 6 hours/day, 5 days/week for 104 weeks. Body weights were lowered in rats 

exposed to 0.73 mg Ni/m3

 compared with controls. Lung inflammation, alveolar hyperplasia, 

macrophage hyperplasia, and pulmonary fibrosis were observed with a significantly increased 

incidence at both nickel concentrations. Female rats exposed to nickel had significantly 

increased adrenal medullary hyperplasia. In addition to the pulmonary lesions, nasal 

inflammation and olfactory epithelial atrophy was observed in both sexes exposed to 0.73 mg 

Ni/m3. 

In the second phase of the NTP study (NTP, 1994b), mice were exposed to 0, 0.6, or 1.2 mg 

Ni3S2/m3 (0, 0.44, or 0.88 mg Ni/m3) for 6 hours/day, 5 days/week for 104 weeks. The same 

pathological lesions were observed in the lung and nasal passages as in the rats in the above 

study. These lesions were evident at both the 0.44 mg Ni/m3 and the 0.88 mg Ni/m3

concentrations. The adrenal medullary hyperplasia seen in female rats was not observed in the 

mice. 

An exposure of rats to either 0 or 0.97 mg Ni3S2/m3 (0 or 0.71 mg Ni/m3) for 6 hours/day, 5 

days/week for 78-80 weeks resulted in decreased body weight, hyperplasia, metaplasia, and 

neoplasia in the lungs due to Ni (Ottolenghi et al., 1974). 

The NTP (1994c) studied the chronic non-cancer and carcinogenic effects of nickel sulfate 

hexahydrate on rats and mice. Rats were exposed to 0, 0.12, 0.25, or 0.5 mg NiSO4/m3

 (0, 0.03, 

0.06, or 0.11 mg Ni/m3) for 6 hours/day, 5 days/week for 104 weeks. Chronic effects of nickel 

exposure in rats included inflammatory lesions in the lung, lung macrophage hyperplasia, 

alveolar proteinosis, and fibrosis, in addition to bronchial lymph node hyperplasia and nasal 

epithelial atrophy. The above effects were seen at exposures of 0.06 mg Ni/m3

 or greater. 

Mice were exposed to a similar regimen that included 0, 0.06, 0.11, and 0.22 mg Ni/m3

 as nickel 

sulfate hexahydrate (NTP, 1994c). Similar pulmonary, lymphatic and nasal changes were 

observed in the mice as with the rats. Fibrosis was not reported, but an increased incidence of 

interstitial infiltration and alveolar proteinosis were observed at exposures of 0.11 mg Ni/m3

 or 

greater. No clinical findings or hematological effects were observed, but body weights were 

significantly depressed in all groups of nickel-exposed female mice. The body weights of males 

were reduced only in the group exposed to 0.22 mg Ni/m3

. 

Rats and mice (10 per group) were exposed to nickel sulfate, nickel subsulfide, or nickel oxide 6 

hours/day, 5 days/week, for 13 weeks (Dunnick et al., 1989). Exposure-related increases in lung 

weight and histological lesions were observed in both species for all nickel exposures. 

Histological lesions included inflammatory changes, fibrosis, and alveolar macrophage 

hyperplasia. Nasal lesions were also observed in animals treated with nickel sulfate or nickel 

subsulfide. Lung weight changes were observed at exposures of 0.05 mg Ni/m3

 or greater in 

female rats. Macrophage hyperplasia in the alveolar region was observed at concentrations as 

Appendix D3 423 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

low as 0.02 mg Ni/m3. Additional inflammatory lesions in the lungs were observed at 0.1 mg 

Ni/m3

. 

A similar study by Haley et al. (1990) found that exposure of mice to nickel sulfate, nickel 

subsulfide, or nickel oxide resulted in various immunological effects. Mice were exposed to 0, 

0.11, 0.45, or 1.8 mg Ni/m3 as Ni3S2; 0.47, 2.0, or 7.9 mg Ni/m3 as NiO; and 0.027, 0.11, and 

0.45 mg Ni/m3

 as NiSO4 for 6 hours/day, 5 days/week for 13 weeks. Nickel exposures 

consistently decreased splenic antibody-forming cell (AFC) responses, with significant decreases 

occurring at 1.8 mg Ni/m3

 as nickel subsulfide. In contrast, AFC responses in the lungassociated lymph nodes were consistently increased, indicating a possible indirect influence of 

inflammatory mediators released in the lung on local lymph nodes. 

Rabbits (8 nickel exposed and 8 controls) exposed to 0.24 mg Ni/m3

 as nickel chloride 6 

hours/day, 5 days/week for 4 weeks exhibited significantly decreased macrophage lysozyme 

activity in pulmonary lavage fluid and in macrophage cultures, compared with control animals 

(Lundborg and Camner, 1984). Similar exposures of rabbits to chlorides of cadmium, cobalt, or 

copper did not reduce lysozyme activity.

VI. Derivation of Chronic Reference Exposure Level (REL) 

A. Nickel and Nickel Compounds (except nickel oxide) 

Study National Toxicology Program, 1994c 

Study population Male and female F344/N rats (52-53 per group) 

Exposure method Discontinuous inhalation 

Critical effects Pathological changes in lung, lymph nodes, and 

nasal epithelium: (1) active pulmonary 

inflammation, (2) macrophage hyperplasia, 

(3) alveolar proteinosis, (4) fibrosis, (5) lymph 

node hyperplasia, (6) olfactory epithelial 

atrophy 

LOAEL 60 μg Ni/m3

 (as nickel sulfate hexahydrate) 

NOAEL 30 μg Ni/m3

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 104 weeks 

Average experimental exposure 5.4 μg Ni/m3

 for NOAEL group (30 x 6/24 x 5/7) 

Human equivalent concentration 1.6 μg Ni/m3

 for NOAEL group males 

(particulate with respiratory effects, RDDR = 

0.29 based on MMAD = 2.5, sigma g = 1.26, 

male rat body weight = 380 g, SA(PU) = 0.34 

m

2

, DEP(PU) = 0.024) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.05 μg Ni/m3

Appendix D3 424 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Nickel Oxide 

Study National Toxicology Program, 1994c 

Study population Male and female F344/N rats (52-53 per group) 

Exposure method Discontinuous inhalation 

Critical effects Pathological changes in lung and lymph nodes: 

(1) active pulmonary inflammation, 

(2) lymph node hyperplasia 

Adrenal medullary hyperplasia (females) 

LOAEL 500 μg Ni/m3

 

NOAEL Not observed 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 104 weeks 

Average experimental exposure 89.5 μg Ni/m3

 for LOAEL group 

(500 x 6/24 x 5/7) 

Human equivalent concentration 30 μg Ni/m3

 for LOAEL group males (particulate 

with respiratory effects, RDDR = 0.29 based 

on MMAD = 2.5, sigma g = 1.26, male rat 

body weight = 380 g, SA(PU) = 0.34 m2, 

DEP(PU) = 0.024) 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.10 μg Ni/m3

The studies conducted by NTP (1994 a,b, & c) all showed similar non-carcinogenic effects in 

rats and mice, regardless of the form of nickel administered. It therefore appears that soluble and 

insoluble forms of nickel cause similar effects in rodents. The human epidemiological literature 

predominantly describes cancer mortality rates from occupational exposures to nickel 

compounds, but does not specifically examine non-cancer effects. However, it is clear from 

many case reports that allergies and dermatitis can occur in exposed workers. Hypersensitive 

reactions to nickel have not been quantitatively studied in humans or animals, therefore it is not 

possible to develop an REL based on immunological hypersensitivity at the present time. A host 

of subacute and subchronic animal studies have shown nickel to affect certain immunological 

responses unrelated to hypersensitivity, but the applicability of these results to chronic human 

exposures and responses involves considerable uncertainty. Furthermore, data show that nickel 

may precipitate onset of asthma in occupational settings. 

The results of the NTP studies and these dose response analyses support the speciation of nickel 

oxide for noncancer effects. The health effects data for nickel oxide indicate that its adverse 

pulmonary effects were less severe (absence of fibrosis, lower chronic lung inflammation 

severity scores) at higher doses than the pulmonary effects observed for nickel sulfate and nickel 

Appendix D3 425 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

subsulfide. The higher chronic REL value for nickel oxide of 0.1 μg/m3 reflects these dose 

response differences. Furthermore, while it is based upon a LOAEL, the lower severity of the 

adverse health effects at the LOAEL mitigates some of the uncertainty associated with use of a 

LOAEL rather than a NOAEL. OEHHA therefore concludes that 0.1 μg/m3

 is an appropriate 

REL for nickel oxide. However, in setting inhalation exposure RELs for groups of compounds, 

OEHHA uses the most sensitive strain, species, sex, chronic endpoint, and agent for each group 

of substances. Therefore, as the pulmonary toxicity of the relatively insoluble nickel subsulfide 

is greater than that of nickel oxide and closer to that of nickel sulfate, OEHHA proposes to use 

the chronic REL derived from nickel sulfate for all other nickel compounds. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL include the availability of controlled lifetime exposure 

inhalation studies in multiple species at multiple exposure concentrations and with adequate 

histopathogical analysis and the observation of a NOAEL. The major areas of uncertainty are 

the lack of adequate human exposure data and the lack of lifetime toxicity studies in any nonrodent species. 

In addition to being inhaled, airborne nickel can settle onto crops and soil and enter the body by 

ingestion. Thus an oral chronic refererence exposure level for nickel is also required. 

Derivation of Oral Chronic Reference Exposure Level 

Study Ambrose et al., 1976 

Study population Rats 

Exposure method Diet 

Critical effects Decreased body and organ weights 

LOAEL 1000 ppm (50 mg/kg-day) 

NOAEL 100 ppm (5 mg/kg-day) 

Exposure continuity Continuous 

Exposure duration Lifetime 

Average exposure 5 mg/kg-day 

Human equivalent concentration 5 mg/kg-day 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Oral reference exposure level 0.05 mg/kg-day 

The oral REL for nickel used the same study used for the U.S. EPA’s oral Reference Dose 

(RfD). U.S.EPA assumed that rat consumption of 1 ppm Ni in the feed resulted in a dose of 0.05 

mg/kg/day. An uncertainty factor of 10 was used for interspecies extrapolation and another of 10 

to protect sensitive human populations. An additional uncertainty factor of 3 was used by U.S. 

EPA to account for inadequacies in reproductive studies of nickel. OEHHA has not used such 

special uncertainty or modifying factor because the criteria for their use are not well presented. 

Appendix D3 426 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

In addition there is an extensive toxicologic database on nickel in general which includes studies 

on reproductive effects. 

VIII. References 

Ambrose AM, Larson PS, Borzelleca JF, and Hennigar GR, Jr. 1976. Long term toxicologic 

assessment of nickel in rats and dogs. J. Food Sci. Technol. 13: 181-187. 

Broder I, Smith JW, Corey P, Holness L. Health status and sulfur dioxide exposure of nickel 

smelter workers and civic laborers. J. Occup. Med. 31(4):347-353. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CDTSC. 1985. California Department of Toxic Substances Control. Review: Nickel and 

compounds. 11/15/85. 

Cirla AM, Bernabeo F, Ottoboni F, and Ratti R. 1985. Nickel induced occupational asthma: 

Immunological and clinical aspects. In: Brown SS, and Sunderman FW. (eds.) Progress in Nickel 

Toxicology. Boston, MA: Blackwell Sci. Pub. 

Davies JE. 1986. Occupational asthma caused by nickel salts. J. Soc. Occup. Med. 36:29-31. 

Dunnick JK, Elwell MR, Benson JM, Hobbs CH, Hahn FF, Haly PJ, Cheng YS, and Eidson AF. 

1989. Lung toxicity after 13-week inhalation exposure to nickel oxide, nickel subsulfide, or 

nickel sulfate hexahydrate in F344/N rats and B6C3F1 mice. Fundam. Appl. Toxicol. 12:584-

594. 

Glaser U, Hochrainer D, Oldiges H, and Takenaka S. 1986. Long-term inhalation studies with 

NiO and As2O3 aerosols in Wistar rats. Int. Congr. Ser. Excerpta Med. 676: 325-328. 

Haley PJ, Shopp GM, Benson JM, Cheng YS, Bice DE, Luster MI, Dunnick JK, and Hobbs CH. 

1990. The immunotoxicity of three nickel compounds following 13-week inhalation exposure in 

the mouse. Fundam. Appl. Toxicol. 15:476-487. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version) Denver, CO: Micromedex, Inc. (Edition expires 4/30/95). 

Hueper WC. 1958. Experimental studies in metal cancerigenesis. Arch. Pathol. 65:600-607. 

Lundborg M, and Camner P. 1984. Lysozyme levels in rabbit lung after inhalation of nickel, 

cadmium, cobalt, and copper chlorides. Environ. Res. 34:335-342. 

McConnell LH, Fink JN, Schlueter DP, and Schmidt MG. 1973. Asthma caused by nickel 

sensitivity. Ann. Int. Med. 78:888-890. 

Appendix D3 427 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Muir DC, Julian J, Jadon N, Roberts R, Roos J, Chan J, Maehle W, Morgan WK. Prevalence of 

small opacities in chest radiographs of nickel sinter plant workers. Br. J. Ind. Med. 50(5):428-

431. 

NTP. 1994a. National Toxicology Program. Technical Report on the Toxicology and 

Carcinogenesis Studies of Nickel Oxide in F344/N Rats and B6C3F1 Mice. NTP TR 451, NIH 

Publication No. 94-3363. U.S. Department of Health and Human Services. 

NTP. 1994b. National Toxicology Program. NTP Technical Report on the Toxicology and 

Carcinogenesis Studies of Nickel Subsulfide in F344/N Rats and B6C3F1 Mice. NTP TR 453, 

NIH Publication No. 94-3369. U.S. Department of Health and Human Services. 

NTP. 1994c. National Toxicology Program. NTP Technical Report on the Toxicology and 

Carcinogenesis Studies of Nickel Sulfate Hexahydrate in F344/N Rats and B6C3F1 Mice. NTP 

TR 454, NIH Publication No. 94-3370. U.S. Department of Health and Human Services. 

Novey HS, Habib M, and Wells ID. 1983. Asthma and IgE antibodies induced by chromium and 

nickel salts. J. Allergy 72(4):407-412. 

Ottolenghi AD, Haseman JK, Payne WW, Falk HL, and MacFarland HN. 1974. Inhalation 

studies of nickel sulfide in pulmonary carcinogenesis of rats. J. Natl. Cancer Inst. 54(5):1165-

1172. 

U.S. EPA. 1985. United States Environmental Protection Agency. Health Assessment Document 

for Nickel. EPA/600/8-83/012F, pp. 3-3. 

Appendix D3 428 Nickel 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

PHENOL 

(Carbolic acid, phenylic acid, phenyl hydroxide) 

CAS Registry Number: 108-95-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 200 μg/m3

 (50 ppb) 

Critical effect(s) Twitching, muscle tremors, neurological 

impairment; elevated serum liver enzymes in 

rats 

Hazard index target(s) Alimentary system; circulatory system; kidney; 

nervous system 

II. Physical and Chemical Properties (From HSDB, 1995, 1999; ATSDR, 1989) 

Description Colorless to light pink solid 

Molecular formula C6H5OH 

Molecular weight 94.11 g/mol 

Density 1.0576 g/cm3

 @ 20° C 

Boiling point 181.75° C 

Melting point 40.9° C 

Vapor pressure 0.3513 torr @ 25° C 

Odor threshold 40 ppb (150 μg/m3

) (Amoore and Hautala, 

1983) 

Solubility 86,000 ppm in water, very soluble in alcohol, 

carbon tetrachloride, acetic acid and liquid 

sulfur dioxide; soluble in chloroform, ethyl 

ether, carbon disulfide; slightly soluble in 

benzene 

Henry’s Law Constant 3.97 x 10-7 ATM-m3

 /mol (25 °C) 

Conversion factor 1 ppm = 3.85 mg/m3

III. Major Uses or Sources (HSDB, 1995) 

Phenol is obtained from coal tar and is widely used as a disinfectant for industrial and medical 

applications. It also serves as a chemical intermediate for manufacture of nylon 6 and other manmade fibers and for manufacture of epoxy and other phenolic resins and as a solvent for 

petroleum refining. Approximately half of the U.S. consumption is directly related to the housing 

and construction industries, in applications such as germicidal paints and slimicides. Phenol is 

present in the atmosphere as an emission from motor vehicles and as a photooxidation product of 

Appendix D3 429 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

benzene. The annual statewide industrial emissions from facilities reporting under the Air 

Toxics Hot Spots Act in California, based on the most recent inventory, were estimated to be 

234,348 pounds of phenol (CARB, 1999). 

IV. Effects of Human Exposures 

The information that is available concerning the health effects of phenol exposure to humans is 

almost exclusively limited to case reports of acute effects of oral exposure (Bruce et al., 1987), 

dermal exposure (Griffiths, 1973), or occupational exposures, including some exposure by 

inhalation (Dosemeci et al., 1991; Ohtsuji and Ikeda, 1972; Connecticut Bureau of Industrial 

Hygiene, undated). Data in animals are consistent with human data and show phenol to be well 

absorbed by oral, dermal, and inhalation routes of exposure. Severe chronic poisoning manifests 

in systemic disorders such as digestive disturbances including vomiting, difficulty swallowing, 

ptyalism (excess secretion of saliva), diarrhea, and anorexia (Bruce et al., 1987; Baker et al., 

1978). Phenol poisoning is associated with headache, fainting, vertigo, and mental disturbances 

(Bruce et al., 1987; Gosselin et al. 1984) which are likely symptoms of neurological effects well 

documented in animal studies. Ochronosis, or discoloration of the skin, and other dermatological 

disorders may result from dermal phenol exposure (Deichmann and Keplinger, 1962; Bruce et 

al., 1987). Several investigators (Truppman and Ellenby, 1979; Warner and Harper, 1985) have 

reported that the use of phenol in the surgical procedure of skin peeling can produce cardiac 

arrhythmias although specifics of dose received were not determined and would be expected to 

be high. 

Human exposure studies in which populations were exposed to phenol over longer periods of 

time (subchronic and chronic) are limited and have serious deficiencies including multiple 

chemical exposures, in many cases small size of exposed populations, and lack of information on 

dose received. 

Occupational studies make up the majority of subchronic/chronic studies available on human 

health effects associated with phenol exposure. Merliss (1972) described muscle pain and 

weakness of unknown etiology, enlarged liver, and elevated serum enzymes (LDH, GOT, and 

GPT) characteristic of liver damage in an individual with intermittent inhalation and dermal 

exposures to phenol, cresol and xylenol. Bruze (1986) noted that a number of phenolformaldehyde based resins are dermal irritants and contact sensitizers. Johnson et al. (1985) 

examined 78 iron and steel foundry workers with multiple chemical and aerosol exposures that 

included phenol and found more respiratory symptoms in the phenol exposed group. However, 

multiple exposure to diphenyl methane diisocyanate, formaldehyde, and silica containing 

aerosols prevented determination of the effects of phenol. Baj et al. (1994) examined twentytwo office workers exposed for six months via inhalation to a commercial product containing 

formaldehyde, phenol and chlorohydrocarbons. At the end of the six month period the indoor air 

of the workers contained 1,300 μg/m3

 of formaldehyde and 800 μg/m3 of phenol. The eight 

workers with the highest concentrations of phenol in their urine had decreased erythrocyte and Thelper lymphocyte numbers and increased numbers of eosinophils and monocytes compared to 

controls. The multiple chemical exposure of this study prevents concluding that these effects are 

attributable to phenol exposure. In a study of hospital workers Apol and Cone (1983) 

documented dermal effects in workers exposed to a number of chemicals including phenols 

Appendix D3 430 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

contained in disinfectants. This study however could not document any differences in urinary 

levels of phenol metabolites between control populations and exposed populations and could not 

assign any of the dermal effects seen to phenol or other substances in the work environment. 

Dosemeci et al. (1991) conducted a follow-up study to evaluate mortality in 14,861 workers in 

five manufacturing facilities producing or using phenol and formaldehyde. Arteriosclerotic heart 

disease, emphysema, disease of the digestive system, and cirrhosis of the liver were inversely 

related to the extent of phenol exposure. Due to multiple chemical exposures the effects of 

phenol alone could not be identified with any certainty. 

Baker et al. (1978) completed a study of 39 individuals exposed to drinking water contaminated 

with phenol for a period of 4-8 weeks. Doses of phenol were estimated to range between 

10 mg/day and 240 mg/day. Effects seen included increased incidence of diarrhea, mouth sores 

and irritation of the oral cavity. 

Two occupational studies are of note since they reported NOAELs. Workers exposed 

continuously for an unspecified period of time to an average air concentration of 4 ppm phenol 

experienced no respiratory irritation (Connecticut Bureau of Industrial Hygiene, undated). No 

adverse effects were reported among workers in a Bakelite factory who were exposed to levels of 

phenol up to 12.5 mg/m3 (3.3 ppm) (Ohtsuji and Ikeda, 1972). In this study urinary phenol 

levels were measured and were observed to return to pre-exposure levels within 16 hours after 

exposure indicating a relatively rapid clearance of phenol from the body that was confirmed in a 

study by Piotrowski (1971). Ohtsuji and Ikeda (1972) did not clearly indicate the number of 

workers sampled or the duration of exposure. 

V. Effects of Animal Exposures

In animal studies a number of subchronic and chronic studies employing oral and inhalation 

routes of exposure are available as well as shorter term studies using the dermal route of 

exposure. Responses observed in animal studies include: pulmonary damage (inhalation 

exposure), myocardial injury (inhalation and dermal exposure), liver damage (inhalation 

exposure), renal damage (inhalation exposure), neurological effects (inhalation exposure), 

developmental effects (oral exposure) and dermal effects (dermal exposure). Comparison of the 

three routes of exposure found that oral exposure was less effective at producing systemic toxic 

effects possibly due to the rapid metabolism of phenol to sulfate and glucuronide conjugates by 

the gastrointestinal tract. Comparison of health effects among studies using dermal, oral and 

inhalation routes of exposure finds that inhalation is a sensitive route of exposure for laboratory 

animals. 

Several subchronic inhalation studies of health effects from phenol exposure are available but no 

inhalation studies longer than 90 days could be identified. Deichmann et al. (1944) exposed 

guinea pigs, rats, and rabbits to concentrations of phenol between 26 and 52 ppm for 28-88 days 

depending on species. Guinea pigs exposed for 7 hours per day, five days per week, for four 

weeks, displayed signs of respiratory difficulty and paralysis primarily of the hind quarters, 

indicating neurological effects. Five of twelve animals exposed at this concentration died at 28 

days. At necropsy, extensive myocardial necrosis, lobular pneumonia, fatty degeneration of the 

Appendix D3 431 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

liver, and centrilobular hepatocellular necrosis were observed in all animals exposed at this level. 

Guinea pigs that were necropsied at 41 days also exhibited pulmonary inflammation, pneumonia, 

bronchitis, endothelial hyperplasia, and capillary thrombosis. Rabbits exposed at these same 

concentrations did not exhibit any signs of discomfort, but showed similar findings at necropsy 

at 88 days. Rats were less sensitive in this study with an apparent NOAEL of 26 ppm phenol for 

these effects. In this study, guinea pigs were the most sensitive species. Limitations of the 

Deichmann study include the range of exposure concentrations and the lack of a control group. 

Sandage (1961) exposed Sprague-Dawley rats, mice and rhesus monkeys for 90 days 

continuously to 5 ppm phenol. Sandage found no effects on pulmonary, cardiovascular, 

hematological, hepatic, or renal systems, thus defining free-standing NOAELs for these systemic 

effects in these species. Limitations of this study include absence of guinea pigs (previously 

identified as the most sensitive species in the Deichmann study) and lack of a demonstrated dose 

response to the effects of phenol. 

Dalin and Kristofferson (1974) examined the effects of phenol on the nervous system in rats 

exposed continuously for 15 days to a concentration of 26 ppm phenol and found muscle 

tremors, twitching and disturbances in walking rhythm and posture after 3-5 days exposure. 

After 15 days exposure, severe neurological impairment as measured by decreased performance 

on tilting plane test was found. The Dalin and Kristofferson (1974) study also documented 

elevated serum concentrations of LDH, GOT, GPT, and GDH indicative of liver damage in 

animals exposed to 26 ppm phenol continuously for 15 days. 

The NCI (1980) study of the carcinogenicity of phenol is the most complete chronic study using 

the oral route of exposure. Mice and rats were exposed for 103 weeks to concentrations of 

phenol in their drinking water of 100, 2500, 5000, and 10,000 ppm. NOAELs in the mouse of 

523 mg/kg/day (5000 ppm in drinking water) and NOAELs in the rat of 630 mg/kg/day 

(5000 ppm in drinking water) were observed for effects on the respiratory system, cardiovascular 

system, gastrointestinal system, hepatic system, renal system, and the brain based on histological 

examination of tissues. Male rats exposed to the 5000 ppm had a higher incidence of kidney 

inflammation (94%) than controls (74%). No tests of kidney function were performed in this 

study. 

Boutwell and Bosch (1959) reported on the results of a chronic study in mice involving skin 

painting of 1.2 mg phenol or 2.5 mg phenol for a 52 week period. A NOAEL of 1.2 mg/animal 

for a 52 week exposure for dermal effects was found. 

No multi-generational studies evaluating reproductive or developmental effects under chronic 

exposure conditions could be identified. Jones-Price et al. (1983a) reported that pregnant rats 

dosed orally with 0, 30, 60, and 120 mg/kg/day on gestation days 6-15 exhibited reduced fetal 

weight in a dose-related manner. However, no teratogenic effects or fetal deaths were observed. 

In a following study Jones-Price et al. (1983b) reported that pregnant mice dosed orally with 0, 

70, 140, and 280 mg/kg/day on gestation days 6-15 exhibited decreased maternal weight gain, 

tremors, and increased maternal mortality at the 280 mg/kg/day dose. In the fetus reduced 

growth, decreased viability, and increased incidence of cleft palate were seen at the 280 

mg/kg/day dose. 

Appendix D3 432 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Sandage, 1961; Dalin and Kristofferson, 1974 

Study population Mice, Sprague Dawley rats and rhesus monkeys 

Exposure method Continuous inhalation 

Critical effects Systemic effects including liver and nervous 

system effects 

LOAEL 26 ppm (Dalin and Kristofferson, 1974) 

NOAEL 5 ppm (Sandage, 1961) 

Exposure continuity Continuous 

Average exposure concentration 5 ppm for NOAEL group 

Human equivalent concentration 5 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

Exposure duration 90 days 

Subchronic uncertainty factor 3 

LOAEL uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.05 ppm (50 ppb; 0.2 mg/m3

 (200 μg/m3) 

No suitable human studies were available for use since exposures were short term or 

occupational in nature with insufficient ancillary information (e.g., duration of exposure) or did 

not determine dose. Of the three routes of exposure available, inhalation appears to be the most 

sensitive based on the number and intensity of systemic effects noted (Deichmann et al., 1944) 

relative to oral exposure (NCI, 1980). In support of this, ATSDR (1989) notes that the 

gastrointestinal tract has a large capacity to metabolize phenol to sulfate and glucuronide 

conjugates which appear likely to be less toxic than the parent compound, thus NOAELs derived 

from oral studies may not be applicable for other routes of exposure. The Deichmann et al.

(1944) study identified guinea pigs as the most sensitive species. However, this study had a 

number of serious deficiencies including absence of controls, significant variability in the 

concentrations of phenol used in their exposure, and exposure that was not continuous. Since 

alternative studies using guinea pigs could not be identified, the rat was chosen as an alternative 

species since the rat has the most similar metabolic profile for metabolism of phenol to that of 

humans (ATSDR, 1989; Capel et al., 1972). The Sandage (1961) study was chosen over other 

available studies since it was the longest in duration (90 days), had a continuous exposure, and 

evaluated three species (rats, mice, monkey). NOAELs determined in the Sandage study for 

systemic effects in all three species examined were 5 ppm, consistent with the idea that 5 ppm is 

a NOAEL for a number of species. Although this is a free-standing NOAEL, a subsequent study 

in rats indicated that nervous system and hepatic effects occur at a concentration of 26 ppm after 

several days (Dalin and Kristofferson, 1974). 

The 5.0 ppm standard for phenol in the workplace (ACGIH, 1988; OSHA, 1985; NIOSH, 1976) 

is considered protective of the health of workers exposed occupationally but does not consider 

Appendix D3 433 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

sensitive populations and is not for continuous exposure conditions. The workplace standard is 

consistent with reports indicating that no respiratory irritation occurred among workers exposed 

regularly to 4 ppm phenol (Connecticut Bureau of Industrial Hygiene, undated) and no adverse 

effects were mentioned among workers exposed to 3.3 ppm (Ohtsuji and Ikeda, 1972). Neither 

report was considered appropriate to be the basis of a REL. However, for the sake of 

comparison adjusting the reported NOAEL of 4 ppm to continuous exposure and dividing by an 

intraspecies uncertainty factor of 10 results in an estimated chronic REL of 140 ppb, in 

reasonable agreement with the proposed REL of 50 ppb. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the key study is the observation of a NOAEL from a continuous exposure 

study involving exposure of several different species. The primary uncertainties are the lack of 

adequate human health effects data, the lack of multiple concentration inhalation exposure 

studies demonstrating a dose-response relationship, the lack of animal studies longer than 90 

days, and the lack of studies with guinea pigs, which have previously been identified as a 

sensitive species for phenol. 

VIII. References 

ACGIH. 1988. American Conference of Governmental Industrial Hygienists. TLVs - Threshold 

Limit Values and Biological Exposure Indices for 1988-1989. Cincinnati, OH: ACGIH. 

Amoore JE, Hautala E. 1983. Odor as an aid to chemical safety: odor thresholds compared with 

threshold limit values and volatilities for 214 industrial chemicals in air and water dilution. 

J. Appl. Toxicol. 3(6):272-290. 

Apol AG, and Cone J. 1983. Bay Area Hospital, Coos Bay, Oregon. Health Hazard Evaluation 

Report. HETA 82-053-1263. National Institute of Occupational Safety and Health. 

ATSDR. 1989. Agency for Toxic Substances and Disease Registry. Toxicological Profile for 

Phenol. NTIS Report PB90-181249. Atlanta, Georgia: ATSDR. 

Baj Z, Majewska E, Zeman K, Pokoca L, Dworniak D, Paradowski M, and Tchorzewski H. 

1994. The effect of chronic exposure to formaldehyde, phenol and organic chlorohydrocarbons 

on peripherial blood cells and the immune system in humans. J. Investig. Allergol. Clin. 

Immunol. 4(4):186-191. 

Baker EL, Landrigan PJ, Bertozzi PE, Field PH, Basteyns BJ, Skinner HG. 1978. Phenol 

poisoning due to contaminated drinking water. Arch. Environ. Health 33:89-94. 

Boutwell RK, and Bosch DK. 1959. The tumor-promoting action of phenol and related 

compounds for mouse skin. Cancer Res. 19:413-424. 

Bruce RM, Santodonato J, and Neal MW. 1987. Summary review of the health effects associated 

with phenol. Toxicol. Ind. Health 3:535-568. 

Appendix D3 434 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Bruze M. 1986. Sensitizing capacity of 2-methylol phenol, 4-methylol phenol and 2,4,6-

trimethylol phenol in the guinea pig. Contact Dermatitis 14:32-38. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Capel ID, French MR, Millburn P, Smith RL, Williams RT. 1972. Fate of [14C]-phenol in 

various species. Xenobiotica 2:25-34. 

Connecticut Bureau of Industrial Hygiene. (Year not reported) unpublished data. [as cited by 

American Conference of Governmental Industrial Hygienists, 1984] 

Dalin NM, and Kristoffersson R. 1974. Physiological effects of a sublethal concentration of 

inhaled phenol on the rat. Ann. Zool. Fenn. 11:193-199. 

Deichmann WB, Kitzmiller KV, and Witherup BS. 1944. Phenol studies. VII. Chronic phenol 

poisoning, with special reference to the effects upon experimental animals of the inhalation of 

phenol vapor. Am. J. Clin. Pathol. 14:273-277. 

Deichmann WB, and Keplinger ML. 1962. Phenols and phenolic compounds. In: Industrial 

Hygiene and Toxicology, Patty FA (ed.) Vol II. Toxicology, 2nd ed., pp.1363-1408. New York: 

Interscience Publishers, Inc. 

Dosemeci M, Blair A, Stewart PA, Chandler J, and Trush MA. 1991. Mortality among industrial 

workers exposed to phenol. Epidemiology 2(3):188-193. 

Gosselin RE, Smith RP, Hodge HC, and Braddock JE. (1984) Phenol. In: Clinical Toxicology of 

Commerical Products. Baltimore: Williams and Wilkins. pp. III-345-346. 

Griffiths GJ. 1973. Fatal acute poisoning by intradermal absorption of phenol. Med. Sci. Law 

13:46-48. 

HSDB. 1995. Hazardous Substances Data Bank. On-line data base. Profile No. 113. 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Johnson A, Moira CY, Maclean L, Atkins E, Dybuncio A, Cheng F, and Enarson D. 1985. 

Respiratory abnormalities among workers in an iron and steel foundry. Br. J. Ind. Med. 42(2):94-

100. 

Jones-Price C, Ledoux TA, Reel JR, et al. 1983a. Teratologic evaluation of phenol (CAS No. 

108-95-2) in CD-1 rats. Research Triangle Park, NC: Research Triangle Institute. 

Appendix D3 435 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Jones-Price C, Ledoux TA, Reel JR, et al. 1983b. Teratologic evaluation of phenol (CAS No. 

108-95-2) in CD-1 mice. Laboratory study: September 18, 1980 to January 12, 1981. Research 

Triangle Park, NC: Research Triangle Institute. 

Merliss RR. 1972. Phenol marasmus. Occup. Med. 14:55-56 

NCI. 1980. National Cancer Institute. Bioassay of phenol for possible carcinogenicity. Technical 

Report Series No. NCI-CG-TR-203. Bethesda, MD: U.S. Department of Health and Human 

Services. 

NIOSH. 1976. National Institute for Occupational Safety and Health. Criteria for a 

recommended standard ... occupational exposure to phenol. DHEW Pub 76-196. 

Ohtsuji H, and Ikeda M. 1972. Quantitative relation between atmospheric phenol vapor and 

phenol in the urine of workers in Bakelite factories. Br. J. Ind. Med. 29:70-73. 

OSHA. 1985. Occupational Safety and Health Administration.Occupational standards 

permissible exposure limits. 29 CFR 1910.1000. 

Piotrowski JK. 1971. Evaluation of exposure to phenol: Absorption of phenol vapor in the lungs 

through the skin and excretion of phenol in urine. Br. J. Ind. Med. 28:172-178. 

Sandage C. 1961. Tolerance criteria for continuous inhalation exposure to toxic material. I. 

Effects on animals of 90-day exposure to phenol, CCl4 and a mixture of indole, skatole, H2S and 

methyl mercaptan. ASD technical report 61-519(I). Wright Patterson Air Force Base, OH: U.S. 

Air Force Systems Command, Aeronautical Systems Division. 

Truppman ES, and Ellenby JD. 1979. Major electrocardiographic changes during chemical face 

peeling. Plast. Reconstr. Surg. 63:44-48. 

Warner MA, and Harper JV. 1985. Cardiac dysrhythmias associated with chemical peeling with 

phenol. Anesthesiology 62:366-367. 

Appendix D3 436 Phenol 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

CHRONIC TOXICITY SUMMARY 

PHOSPHINE 

(hydrogen phosphide; phosphorus trihydride; Celphos; Phostoxin) 

CAS Registry Number: 7803-51-2 

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.8 μg/m3 (0.6 ppb) 

Critical effect(s) Decreased body weight gain in mice 

Hazard index target(s) Respiratory system; alimentary system; nervous 

system; kidney; hematopoietic system 

II. Chemical Property Summary (HSDB, 1995, except as noted) 

Description Colorless gas 

Molecular formula PH3

Molecular weight 34 g/mol 

Vapor density 1.17 (air = 1) 

Boiling point −87.7°C 

Vapor pressure 20 atm @ −3°C (Weast, 1980) 

Solubility 0.26 volumes in water @ 20°C; soluble in alcohol, 

ether (Sax and Lewis, 1989) 

Conversion factor 1.39 mg/m3

 per ppm at 25oC 

III. Major Uses and Sources 

Phosphine is used as an agricultural fumigant against insects and is among the most acutely toxic 

of the fumigant gases (HSDB, 1995). In its use as a fumigant, application of aluminum, 

magnesium, or zinc phosphide pellets generates phosphine gas upon exposure to moisture. 

Because of high volatility, phosphine residue dissipates from treated material upon ventilation. 

Inadequate sealing of materials during the course of treatment can result in unplanned 

environmental exposure. 

Phosphine is also used by the semiconductor industry as a chemical doping agent for electronic 

components (n-type semiconductors) (HSDB, 1995). Other minor sources/uses of phosphine are 

in chemical syntheses: specifically, in preparations of phosphonium halides, for polymerization 

initiation, and as condensation catalysts. The annual statewide industrial emissions from 

facilities reporting under the Air Toxics Hot Spots Act in California based on the most recent 

inventory were estimated to be 3312 pounds of phosphine (CARB, 1999). In 2000, more than 

120,000 pounds of phosphide compounds (including 119,519 pounds of aluminum phosphide 

and 1607 pounds of zinc phosphide) were applied in California agriculture (CDPR, 2001). In the 

Appendix D3 437 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

presence of water these phosphides break down to phosphine. However, the fraction emitted as 

phosphine is not known. 

IV. Effects of Exposures to Humans 

Toxicity among 22 workers intermittently exposed to phosphine levels of 0.17-2.11 ppm in air 

from fumigation activity ranging over 0.5 to 29 years (mean = 11.1 years) was evaluated (Misra 

et al., 1988). The subjects were interviewed within one day of fumigation activity and reported 

that symptoms subsided when phosphine was not in use. The most frequently reported 

symptoms include dyspnea (31.8%), headache (31.8%), chest tightness (27.3%), cough (18.2%), 

anorexia and epigastric pain (18.2%), finger paresthesia and numbness (13.6%), and giddiness, 

numbness, and lethargy (13.6%). No change in motor or sensory nerve conduction velocity was 

found. 

A similar spectrum of toxic effects among workers involved in grain storage at a seaboard 

terminal has been reported (Jones et al., 1964). Among 69 men exposed to breathing zone 

phosphine levels of 0-35 ppm for as long as 16 hours per day, the authors report symptoms of 

multiple origins. These include gastrointestinal (diarrhea, nausea, epigastric pain, vomiting), 

cardio-respiratory (chest tightness, dyspnea, pain in chest, palpitations, retrospinal pain), and 

central nervous (headache, dizziness, staggering gait) systems. Symptoms were reported to 

appear only at the time of exposure and apparently were reversible. 

In another report of chronic occupational exposure, authors cited the appearance of chronic 

bronchitis, anemia, and digestive disorders (Eichler, 1934). 

Most literature reports of human toxic health effects of phosphine, however, come from case 

reports of acute exposures. Some are suggestive of potential chronic toxicity endpoints because 

of the irreversible nature of the effect. In a case report of phosphine poisoning of 29 people 

exposed by inhalation on a grain freighter, pathological findings included evidence of urinary 

tract injury (occult blood), liver damage (bilirubinuria and increased SGPT, GGPT, and LDH), 

and myocardial damage (increased MB fraction of CPK, abnormal ECG) (Wilson et al., 1980). 

A-two-year old child who died as a result of the exposure showed myocardial necrosis with 

mononuclear infiltrates, pulmonary edema with damaged epithelia, pleural effusion, and an 

enlarged spleen. In another case report exposure of a 7-months pregnant, 24-year-old woman to 

aluminum phosphide from a nearby grain storage site was lethal (Garry et al., 1993). There was 

evidence of severe pulmonary edema, necrosis of individual hepatic cells, and anoxic change in 

Purkinje cells of the cerebellum. These reported deaths of a small child and a pregnant woman 

exposed together with individuals who survived exposure to phosphine suggest that there may be 

sensitive human subpopulations. In another case report of acute phosphine poisoning by 

inhalation, Schoonbroodt et al., (1992) observed necrosis of the nasal mucosa, delayed onset of 

pulmonary edema, and myocardial injury. Chopra et al. (1986) treated sixteen patients with 

aluminium phosphide poisoning during 1985. Findings included mucosal necrosis and cardiac 

abnormalities (due to hypoxemia). Renal failure (1/16), proteinuria (1/16), and increased blood 

transaminases (2/16) resulted from oral exposure to phosphine. The multi-organ involvement in 

toxicity suggests that phosphine is a broad-spectrum toxicant. 

Appendix D3 438 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

In a 1994 two deaths and three illnesses were reported due to phosphine fumigation of 

agricultural products in railroad cars (Perrotta et al., 1994). The annual report for the year 2000 

of the American Association of Poison Control Centers’ Toxic Exposure Surveillance System 

does not list fumigants as a separate category of pesticides (Litovitz et al., 2001). However, two 

of the case reports of fatalities, that were presented in abstract form, were due to aluminum 

phosphide. 

V. Effects of Exposures to Animals 

A subchronic inhalation toxicity study of phosphine was conducted in Balb-c mice (Barbosa et 

al., 1994). Twelve animals/sex/dose group were exposed for 6 hours/day, 5 days/week for 13 

weeks to 0, 0.3, 1.0, or 4.5 ppm phosphine. Non-cancer toxicity endpoints included reduction in 

weight gain and changes in relative organ weights of kidneys, lungs, liver, heart, brain and 

spleen. In the highest dose group, itching and scratching of the eyes, feet and tail, and decreased 

overall activity were observed. No diarrhea, loss of equilibrium, convulsions, seizures, or other 

neurological disturbances were noted. A dose-dependent decrease in total body weight gain was 

observed at all exposure levels with a greater effect observed in females (p<0.0001). Statistically 

significant decreases in relative organ weights (kidney, heart, and brain) were observed in males 

only at the 0.3 ppm exposure level (p<0.001). On the other hand, female mice showed increased 

relative organ weights (lungs, heart, and spleen) predominantly at higher doses (1.0 and 4.5 ppm; 

p<0.001). At 4.5 ppm phosphine absolute kidney and spleen weights were significantly 

increased in females (p<0.01). Increased frequencies of micronuclei in polychromatic 

erythrocytes from bone marrow and spleen were also seen at 4.5 ppm. This group also 

conducted a short-term repeated dose experiment. Six mice/sex/group were exposed to 5.5 ppm 

phosphine for 2 weeks (6 hrs/day, 5 days/wk). No statistically significant changes in weight gain 

were observed at the end of this exposure period. 

In another subchronic inhalation toxicity study, male and female Fischer 344 rats (10/sex/group) 

were exposed to levels of 0, 0.37, 1.0, and 3.1 ppm phosphine for 6 hours per day, 5 days per 

week, for 13 weeks (Newton et al., 1993). A higher dose group at 10 ppm was terminated 

prematurely (at 3 days) because of high mortality. A satellite group exposed to 5.1 ppm for 2 

weeks was terminated after 13 days recovery. Observations of overt toxicity and viability were 

made at the time of each exposure; body weight and food consumption were monitored weekly; 

ophthalmic examination was done the day before termination; and hematological and clinical 

chemistry indices were measured after 4 and 13 weeks. Postmortem examination included gross 

necropsy, with particular attention to orifices, the cranial cavity, surfaces of the brain and spinal 

cord, nasal cavity and sinuses, the thoracic, abdominal, and pelvic cavities and viscera, and the 

cervical tissue and organs. Histopathology was performed on 10% buffered formalinfixed/hematoxylin-eosin-stained tissues. Significant observations after 13 weeks of phosphine 

exposure included decreased hemoglobin, hematocrit, and erythrocytes in males in the 3.1 ppm 

dose group. Male rats in the 1 ppm dose group showed decreased weight gain. Increased 

incidence of small seminal vesicles was noted at 1 and 3.1 ppm, although no histological 

correlate was observed. Absolute and relative decreases in liver weight were observed in all 

exposed groups, but there was no evidence that this effect was dose-related. A significant 

Appendix D3 439 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

decrease in serum glutamic pyruvic transaminase (SGPT) was observed at 3.1 ppm, although the 

authors noted unusually high control levels. None of these effects were observed after the 4 

week recovery period. Other effects of a transient nature noted during the exposure include 

decreased weight gain in female rats at 1 ppm, decreased food consumption at 0.37 ppm in males 

and females, and increased blood urea nitrogen (BUN) at 3.1 ppm. Observations in the 10 ppm 

group necropsied after 3 days of exposure included decreased erythrocytes, increased alkaline 

phosphatase, and increased kidney weight with coagulative necrosis of the tubular epithelium of 

the outer cortex. In a subchronic study of CD male and female rats under similar conditions 

(exposure to 0, 0.3, 1, or 3 ppm phosphine 6 h/day, 5 days/week for 13 weeks), no neurotoxicity 

was observed (Schaefer et al., 1998). 

Newton et al. (1993) also examined developmental toxicity by exposing 24 pregnant female CDR

rats per group to 0, 0.03, 0.33, 2.8, 4.9, and 7.0 ppm phosphine. The highest dose group was 

terminated prematurely because of high mortality; all other animals were necropsied after 20 

days for evaluation of maternal and fetal toxicity. Maternal toxicity endpoints included weight 

of ovaries and uteri, number of corpora lutea, pregnancy, and implantation rate. Fetal toxicity 

was evaluated by weight, number, and location of fetuses and resorptions, visceral 

malformations and variations, and skeletal changes after alizarin staining. No statistically 

significant differences from control animals were observed for any parameter at any dose, with 

the exception of a change in mean number of resorption sites (p≤0.01), mean resorption/implant 

ratio (p≤0.05), and incidence of females with resorption (p≤0.05), all at 0.03 ppm only. In the 

absence of this effect at higher dose levels, these observations are not considered useful in 

establishing a low adverse effect level. 

A 35-day phosphine inhalation study was conducted exposing rats continuously to 0, 0.05, 0.2, 

1.5, and 8.0 mg/m3 phosphine (0, 0.036, 0.14, 1.1, and 5.8 ppm) in which hematological 

endpoints and histopathological changes of the lungs and kidneys were examined (Pazynich et 

al., 1984). Observations include a statistically significant change in erythrocytes (increase 

followed by a decrease at day 35) and decreased hemoglobin at the 0.05 and 0.2 mg/m3 dose 

levels, although the 1.5 mg/m3

 dose group did not show this change. Other significant changes, 

noted in the lowest dose group, included decreased peroxidase activity after 35 days exposure, 

decreased sulfhydryl group content in blood after 27 days, and decreased phagocytotic index 

after 21 days. Some histological changes were noted in the lungs, kidneys, and to a lesser extent, 

the liver, particularly in the higher dose groups, although the exact nature of the degenerative 

change is not well described. Unclear dose-response relationships and temporal aspects of the 

endpoints also make establishment of a low adverse effect level unreliable. 

Two rats were exposed to 20 ppm phosphine for 14 days (4 hours/day) (Waritz and Brown, 

1975). Animals were monitored for weight gain, and organs/tissues fixed in Bouin’s solution 

and stained with trichrome were examined histopathologically. There were no reported 

histopathological effects, although there was slightly reduced weight gain in the exposed 

animals. 

In a chronic study of phosphine (Newton et al., 1999), 60 male and female F344 rats per group 

were exposed via whole-body inhalation for 6 h/day, 5 days/wk for up to 104 wk to mean 

concentrations of 0, 0.3, 1, or 3 ppm phosphine. Three ppm (4.17 mg/m3

) was the maximum 

Appendix D3 440 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

exposure level because of lethality seen at the high exposure level (7 ppm = 9.73 mg/m3

) in 

previous repeat dose studies (Newton et al., 1993). Ten rats per sex per group were killed after 

52 weeks of exposure. Survivors were killed after 104 weeks of exposure. There were no 

phosphine-related effects seen on clinical observations, body weight, food consumption, 

hematology, clinical chemistry, urinalysis, or ophthalmology. There were no phosphine-related 

macroscopic findings or effect on absolute or relative organ weights. No histologic or 

morphologic alterations attributable to phosphine exposure were seen in the more than 40 organs 

and tissues examined. Under the conditions of this study, the authors found no treatment-related 

changes suggestive of a toxic or carcinogenic effect in rats following 52 weeks or 2 years of 

whole-body inhalation exposure to 0.3, 1, or 3 ppm phosphine. Thus 3 ppm is a chronic NOAEL 

for rats. 

VI. Derivation of Chronic Reference Exposure Level 

Study Barbosa et al., 1994 

Study population Balb-c mice (12 animals/sex/group) 

Exposure method Discontinuous whole body inhalation exposure 

(0, 0.3, 1, or 4.5 ppm) 

Critical effects Decrease in body weight gain, increase in relative 

organ weights; increase in micronuclei 

LOAEL 4.5 ppm 

NOAEL 1 ppm 

Exposure continuity 6 hr/day, 5 days/ week 

Average experimental exposure 0.178 ppm for NOAEL group (1 x 6/24 x 5/7) 

Human equivalent concentration 0.178 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

Exposure duration 13 weeks 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 10 (see below)

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Reference exposure level 0.0006 ppm (0.6 ppb; 0.0008 mg/m3

; 0.8 μg/m3) 

Appendix D3 441 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

Supportive study Newton et al., 1999 

Study population Rats (60 animals/sex/exposure level) 

Exposure method Discontinuous whole body inhalation 

exposure (0, 0.3, 1, or 3 ppm) 

Critical effects None 

LOAEL None detected 

NOAEL 3 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Average experimental exposure 0.53 ppm (3 x 6/24 x 5/7) 

Human equivalent concentration 0.53 ppm (gas with systemic effects, 

based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda 

(h)) 

Exposure duration 2 years 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 10 (see below)

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Reference exposure level 0.005 ppm (5 ppb, 0.007 mg/m3

, 7 μg/m3) 

Newton et al. (1999) found no treatment-related changes suggestive of a toxic effect in F344 rats 

following 52 weeks or 2 years of whole-body inhalation exposure to 0.3, 1, or 3 ppm phosphine. 

Thus in this study 3 ppm is a chronic NOAEL for rats. Three ppm was set as the maximum level 

because in an earlier subchronic study in rats Newton et al. (1993) found lethality at 7 ppm. 

However, the chronic results of Newton et al. (1999) differ from the subchronic results of 

Newton et al. (1993), in which at least transient effects were seen in the hematopoietic system 

after 13 weeks at 3.1 ppm. In a subchronic study in mice (Barbosa et al., 1994), 4.5 ppm 

phosphine was a LOAEL and 1 ppm was a NOAEL for decrease in body weight gain. The 

results of Barbosa et al. (1994) indicated that mice may be more sensitive than rats. Thus, it was 

selected as the key study, and decrease in body weight gain was selected as the critical effect. 

OEHHA has applied a subchronic uncertainty factor of 3 to account for the short duration of the 

Barbosa et al. (1994) study and an intraspecies uncertainty factor of 10 to account for human 

variability. Due to the general inconsistencies among the various studies in the database on 

phosphine, and in particular with the observation of mortality at 7 ppm in a short-term 

developmental study in rats (Newton et al., 1993), it was considered prudent to include the full 

interspecies uncertainty factor of 10 (even though the HEC adjustment procedure could be 

applied) to acknowledge the severity of effect in at least one comparison study, and the 

additional uncertainty associated with the apparent wide and unpredictable variability between 

species and between different studies in the same species (rats). This results in a cumulative 

uncertainty factor of 100 to be applied to the NOAEL of 1 ppm in the subchronic study by 

Barbosa et al. (1994) and a chronic REL for phosphine of 0.8 μg/m3

 (0.6 ppb). 

The U.S. EPA based its RfC of 0.3 μg/m3 on the Barbosa et al. (1994) study, an adequate 

subchronic animal study for the derivation of a REL, and included a Modifying Factor (MF) of 3 

Appendix D3 442 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

for database deficiencies (lack of multigenerational reproduction studies). The criteria for use of 

modifying factors are not well specified by U.S. EPA. USEPA used its default interspecies 

uncertainty factor of 10 for a 13 week study. 

The lack of adequate data on levels of chronic phosphine exposure to humans precludes 

development of a REL from human studies. The endpoint used in the determination of the REL 

(total body weight gain) showed a dose-related decrease with phosphine exposure in Balb-c 

mice. This endpoint is also consistent with that found by Newton et al. (1993), who noted dosedependent decreases in body weight gain in Fischer 344 rats after a 13 week exposure regimen at 

1 ppm, and Waritz and Brown (1975), who reported slightly decreased weight gain in rats 

exposed for 14 days to 20 ppm. Surprisingly Newton et al. (1999) did not find differences in 

body weight gain at either 1 ppm or 3 ppm and they did not comment on the discrepancy 

between their 2 reports. Although body weight changes or changes in food consumption were 

not addressed in human studies, the scant human data do relate phosphine exposure to a broad 

spectrum of toxic effects (gastrointestinal, cardio-respiratory, CNS). The decrease in weight 

gain found in the animal studies and reported changes in some relative organ weights (Barbosa et 

al., 1994) suggest systemic toxicity. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for phosphine include the availability of data on multiple 

inhalation exposure concentrations and the observation of a NOAEL in a lifetime animal study. 

Major areas of uncertainty are the lack of adequate human exposure data, the lack of 

reproductive and developmental toxicity studies, and the inconsistency of the dose-response 

relationship across rodent studies. 

VIII. Potential for Differential Impacts on Children's Health

Based on the lack of a dose-response in developmental toxicity caused by exposing pregnant 

female rats to 4.9 ppm phosphine (Newton et al., 1993) , the proposed REL of 0.8 μg/m3

 (0.6 

ppb) is likely to be protective of developing humans in utero. However, there is no direct 

evidence in the literature to quantify a differential effect of phosphine on infants and children. 

IX. References

Barbosa A, Rosinova E, Dempsey J, and Bonin AM. 1994. Determination of genotoxic and other 

effects in mice following short term repeated-dose and subchronic inhalation exposure to 

phosphine. Environ. Mol. Mutag. 24:81-88. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Appendix D3 443 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

CDRR. 2001. California Department of Pesticide Regulation. Summary of Pesticide Use Report 

Data 2000 Indexed by Chemical. Preliminary Data. Available on-line at 

http://www.cdpr.ca.gov/docs/pur/pur00rep/chmrpt00.pdf 

Chopra JS, Kalra OP, Malik R, Sharma R, and Chandna A. 1986. Aluminum phosphide 

poisoning: a prospective study of 16 cases in one year. Postgrad. Med. J. 62:1113-1115. 

Eichler O. 1934. Phosphine poisoning: chronic, occupational? Sammlung von Vergiftungsfällen, 

5:23. 

Garry VF, Good PF, Manivel JC, and Perl DP. 1993. Investigation of a fatality from 

nonoccupational aluminum phosphide exposure: measurement of aluminum in tissue and body 

fluids as a marker of exposure. J. Lab. Clin. Med. 122:739-747. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Jones AT, Jones RC, and Longley EO. 1964. Environmental and clinical aspects of bulk wheat 

fumigation with aluminum phosphide. Ind. Hyg. J. 25:376-379. 

Litovitz TL, Klein-Schwartz W, White S, Cobaugh DJ, Youniss J, Omslaer JC, Drab A, Benson 

BE. 2001. 2000 Annual report of the American Association of Poison Control Centers Toxic 

Exposure Surveillance System. Am. J. Emerg. Med. 19(5):337-395. 

Misra UK, Bhargava SK, Nag D, Kidwai MM, and Lal MM. 1988. Occupational phosphine 

exposure in Indian workers. Toxicol. Lett. 42:257-263. 

Newton PE, Schroeder RE, Sullivan JB, Busey WM, and Banas DA. 1993. Inhalation toxicity of 

phosphine in the rat: acute, subchronic, and developmental. Inhal. Toxicol. 5:223-239. 

Newton PE, Hilaski RJ, Banas DA, Wilson NH, Busey WM, Shaheen DG. 1999. A 2-year 

inhalation study of phosphine in rats. Inhal. Toxicol. 11(8):693-708. 

OEHHA, 1997. Draft Air Toxics Hot Spots Program Risk Assessment Guidelines. Part III. 

Technical Support Document for the Determination of Noncancer Chronic Reference Exposure 

Levels. 

OEHHA, 1999. Draft Air Toxics Hot Spots Program Risk Assessment Guidelines. Part III. 

Technical Support Document for the Determination of Noncancer Chronic Reference Exposure 

Levels. 

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Air Toxics Hot Spots 

Program Risk Assessment Guidelines. Part III. Technical Support Document for the 

Determination of Noncancer Chronic Reference Exposure Levels. Available on-line at 

http://www.oehha.ca.gov 

Appendix D3 444 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

Pazynich VM, Mazur IA, Podloznyi AV, Chinchevich VI, and Mandrichenko BE. 1984. 

Experimental substantiation and prediction of time-related maximum permissible concentration 

of phosphine in the air. Gig. Sanit. Jan(1):13-15. 

Perrotta D, Willis T, Salzman D et al. 1994. Deaths associated with exposure to fumigants in 

railroad cars – United States. MMWR 43(27):489-491. 

Sax NI, and Lewis RJ (eds). 1989. Dangerous Properties of Industrial Chemicals. 7th ed. New 

York: Van Nostrand Reinhold. 

Schaefer GJ, Newton PE, Gruebbel MM, Busey WM, and Shaheen DG. 1998. Acute and 

subchronic inhalation neurotoxicity of phosphine in the rat. Inhal. Toxicol. 10:293-320. 

Schoonbroodt D, Guffens P, Jousten P, Ingels J, and Grodos J. 1992. Acute phosphine 

poisoning? A case report and review. Acta Clin. Belg. 47:280-284. 

Waritz RS, and Brown RM. 1975. Acute and subacute inhalation toxicities of phosphine, 

phenylphosphine and triphenylphosphine. Am. Ind. Hyg. Assoc. J. 36:452-458. 

Weast RC (ed.). 1980. CRC Handbook of Chemistry and Physics. 61st ed. Boca Raton, FL: CRC 

Press. 

Wilson R, Lovejoy FHJr, Jaeger RJ, and Landrigan PL. 1980. Acute phosphine poisoning aboard 

a grain freighter. JAMA 244:148-150. 

Appendix D3 445 Phosphine 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

PHOSPHORIC ACID 

(Orthophosphoric acid) 

CAS Registry Number: 7664-38-2 

I. Chronic Toxicity Summary

Inhalation reference exposure level 7 μg/m3

 

Critical effect(s) Bronchiolar fibrosis of the respiratory tract in 

rats 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1995; 1999)

Description Clear syrupy liquid or unstable crystals; odorless 

Molecular formula H3PO4

Molecular weight 98 

Boiling point 213°C 

Melting point 42.35°C 

Vapor pressure 0.03 torr @ 20°C 

Solubility Very soluble in hot water; 548 g/100 ml cold 

water; soluble in alcohol 

Conversion factor 4.0 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Phosphoric acid has varied uses (HSDB, 1995). In manufacturing, it is a chemical intermediate 

or reagent in the production of numerous phosphate fertilizers, agricultural feeds, waxes, 

polishes, soaps, and detergents. It is added to foods as a preservative, acidifying agent, flavor 

enhancer, and clarifying agent. Phosphoric acid is also used in processes such as the coagulation 

of rubber latex, electropolishing, soil stabilization, and as a catalyst in the production of 

propylene and butene polymers, ethylbenzene, and cumene. By far, largest use of phosphoric 

acid comes in the production of fertilizers (>75%). The annual statewide industrial emissions 

from facilities reporting under the Air Toxics Hot Spots Act in California, based on the most 

recent inventory, were estimated to be 81,103 pounds of phosphoric acid (CARB, 1999). 

Airborne phosphoric acid can be produced by the hydrolysis of phosphorus oxides generated 

from either the spontaneous ignition of white phosphorus in air or the combustion of red 

phosphorus (Burton et al., 1982; US Department of Defense (US DOD), 1981). 

Appendix D3 446 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Human Exposures 

The toxic effects to 48 workers exposed (28 unexposed control workers) to oxidation products of 

phosphorus during the course of phosphorus production were reported (Hughes et al., 1962). 

Exposure duration ranged from 1 to 17 years. No differences were observed between exposed 

and control workers with respect to leukocyte count, an effect observed in acute intoxications, or 

hand bone density, an effect observed in experimentally exposed animals (Inuzuka, 1956). 

A prospective study of 131 workers exposed to several compounds including phosphoric acid, 

phosphorus pentoxide, fluorides and coal tar pitch in the air was conducted at an industrial 

refinery (Dutton et al., 1993). Mean duration of exposure (employment) was 11.4 years and the 

maximum exposure level measured was 2.23 mg/m3 (phosphorus pentoxide). Pulmonary 

function tests were performed annually over a 3 to 7 year period. No significant residual effect 

was found after adjusting for age and smoking status. 

V. Effects of Animal Exposures 

Two 13-week inhalation studies of the effects of exposure to the combustion products of 95% 

red phosphorus and 5% butyl rubber were conducted in male Sprague-Dawley rats, with the first 

group exposed to 0, 300, 750, or 1200 mg/m3 combustion products, and the second exposed to 0, 

50, 180, or 300 mg/m3 combustion products (Aranyi et al., 1988a; Aranyi et al., 1988b). Group 

numbers in the first study were 176, 84, 176, and 176, respectively. The second study used 40 

animals/group. Animals were exposed for 21⁄4 hours/day on 4 consecutive days/week. Control 

animals were exposed to filtered air only. Daily particle measurements showed MMADs of 

0.49-0.65 μm and σgs of 1.56-1.83. Fractional content of phosphoric acid in the aerosol was 71-

79%. Nineteen of the 176 animals in the 1200 mg/m3 dose group died of treatment related 

effects. Post-mortem examination of animals that died during the course of the study showed 

damage to the laryngeal mucosa, which was probably contributory to mortality. The two highest 

dose groups in the first study also showed decreased weight gain. Twelve animals from each 

dose group in the first study were examined histologically and neurobehavioral studies were 

conducted on other animals. Half the animals in the second study were examined strictly for 

toxic effects on the respiratory tract, with examination of the trachea, 2 sections of the nasal 

turbinates, and 5 lobes of the lung. Surviving animals in the high-dose study were observed to 

have moderate to severe fibrosis of the terminal bronchioles, with minimal severity of this effect 

in the animals in the low-dose study. The reported incidence of this lesion was 9/20 at 300 

mg/m3

, 4/20 at 180 mg/m3, and 0/20 at 50 mg/m3. Little to no involvement of pulmonary tissue 

was observed. 

The effects of acid aerosols (particularly sulfuric and phosphoric acid) were studied by U.S. EPA 

(1989). The respiratory tract was the primary target of toxicity resulting from the irritational 

effect of the acid on the tissues of the larynx and trachea. The nature of the effect was dependent 

upon the aerosol particle size, duration of exposure, and the hygroscopic character of the acid. 

Sprague-Dawley rats were exposed to the smoke and combustion products of white phosphorus 

in felt pellets at 192.5 (18 animals/sex), 589 (24 animals/sex), or 1161 mg/m3

 (34 males, 43 

Appendix D3 447 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

females) phosphoric acid equivalents for 15 minutes/day, 5 days/week, for 13 weeks (US 

Department of Defense (US DOD), 1981). Control animals numbering half the size of the 

treated groups were exposed to air only. Groups of animals were sacrificed at 6 and 13 weeks, 

and 4 weeks post-exposure. Endpoints examined included: hematology, clinical chemistry, 

gross- and histo-pathology, ECG, pulmonary function, and behavior. Of the animals in the 

highest dose group, 56% died as a result of exposure, with the only other death occurring in the 

control group. Findings were restricted to effects on the respiratory system, with tracheitis and 

laryngitis incidences of 2/35, 32/47, and 28/31among surviving animals in the three dose groups. 

In the post-exposure examination, bronchiolitis occurred with a frequency of 0/12, 5/24, and 

6/16in the three dose groups. 

The toxicity of the combustion products of 95% amorphous red phosphorus and 5% polyvinyl 

butyral BL18 to female Wistar rats, Porton-strain mice, and guinea pigs was reported (Marrs et 

al., 1989). Rats (50/group), mice (100/group), and guinea pigs (42-48/group) were exposed to 

concentrations of 0, 16, or 128 mg/m3

 for 1 hour/day, 5 days/week for 36 weeks (mice) or 40 

weeks (rats and guinea pigs), with an examination conducted at 19 months or when animals 

appeared unhealthy. All groups, including controls, showed high mortality. Mice showed 

accumulation of alveolar macrophages with incidences of 2/41, 9/37, and 9/22 in the control, 

low-, and high-dose groups, respectively. Guinea pigs appeared to be particularly intolerant to 

the effects of the smoke. 

Female rabbits and rats (10/group) were examined for acute toxic effects of smoke generated by 

the combustion of either 95% red phosphorus / 5% butyl rubber (Smoke I) or 97% red 

phosphorus / 3% butadiene styrene (Smoke II) (Marrs, 1984). Animals were exposed for 30 

minutes and examined one and 14 days later. Smoke I produced inflammation of the larynx and 

trachea in rats at 1 day with some inflammation still observed at 14 days. Tracheal inflammation 

was also reported in rabbits exposed to Smoke I. Four of the rats exposed to Smoke II died 

within the first day, with severe pulmonary congestion observed in the animals. 

One hour exposure to the combustion products of 95% red phosphorus / 5% butyl rubber (plus 

1% mineral oil) produced epiglottal deformation, laryngeal edema, and laryngeal and tracheal 

lesions in rats (Burton et al., 1982). A four-hour exposure produced more severe effects of a 

similar nature plus some hemorrhaging. 

Rats (number unspecified) exposed to 150-160 mg/m3

 elemental phosphorus for 30 minutes/day 

for 60 days were examined for toxic effects (Inuzuka, 1956). Limb bone abnormalities were 

noted and effects included delayed ossification, widening of the epiphysis, and abnormal axial 

development. 

Two studies have addressed the reproductive and developmental toxicity from exposure to the 

combustion products of white phosphorus and felt for 15 minutes/day during gestational days 6-

15 in rats (24/group) (US Department of Defense (US DOD), 1981; US Department of Defense 

(US DOD), 1982). Fetal effects included increased incidence of some visceral variations and 

hypoplasia of the xiphoid process although data were incompletely reported. Another study, 

which exposed dams 3 weeks prior to mating, throughout gestation, and through lactation and 

males for 10 weeks prior to and during mating, showed decreased pup body weight, 24-hour and 

Appendix D3 448 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

21-day survival, and lactation. An oral study in which elemental phosphorus was administered 

to male and female rats by gavage in corn oil showed no statistically significant effects 

(Condray, 1985). 

VI. Derivation of the Chronic Reference Exposure Level 

Study Aranyi et al., 1988a 

Study population Male Sprague-Dawley rats (40-176/group) 

Exposure method Discontinuous whole body inhalation 

Critical effects Bronchiolar fibrosis of the respiratory tract 

LOAEL 180 mg/m3

 

NOAEL 50 mg/m3

 

BMC05 64 mg/m3

Exposure continuity 21⁄4 hours/day, 4 days/week 

Exposure duration 13 weeks 

Average experimental exposure 2.7 mg/m3

 for NOAEL group (estimated as 3.5 

mg/m3

 at BMC05) 

Human equivalent concentration 2.2 mg/m3

 at BMC05 (particle with respiratory 

effects, RDDR = 0.63) (3.5 x 0.63) 

LOAEL uncertainty factor 1 (BMC05 assumed to be similar to NOAEL) 

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Reference exposure level 0.007 mg/m3

 (7 μg/m3) 

OEHHA has used the same study which U.S. EPA used in the development of its Reference 

Concentration (RfC) of 10 μg/m3

. The U.S. EPA has used a benchmark dose methodology for 

the derivation of the RfC for phosphoric acid from the toxicity data in the Aranyi et al. (1988) 

study (U.S. EPA, 1995). The RfC is restricted to “aerosols of phosphoric acid and phosphorus 

oxidation products and does not apply to elemental phosphorus or other forms of phosphorus, 

such as phosphorus salts”. 

The U.S. EPA, using the Weibull model, estimated the lower 95% confidence level bound on the 

maximum likelihood estimate (MLE = 150 mg/m3

) resulting in 10% incidence of lesions in the 

tracheo-bronchiolar region to be 100 mg/m3 (the BMC10). The U.S. EPA considered 10% 

incidence level to be a correlate to the NOAEL, based on a precedent in the analysis of data with 

developmental toxicity endpoints (Allen et al., 1994; Faustman et al., 1994). After correction for 

exposure continuity, a regional deposited dose ratio (RDDR) for the tracheobronchial region of 

0.64 was applied due to the availability of data concerning the growth and deposition of 

phosphoric acid aerosol particles in humans and the similarities in the effects of phosphoric and 

better-characterized sulfuric acid aerosols. Key assumptions in the generation of this factor 

include: (1) the lowest σg of 1.56 μm cited in the study was used in the calculation; (2) geometric 

rather than aerodynamic diameter approximations were used; (3) particles of this size reach the 

deposition / lesion site (bronchioles); 4) these hygroscopic particles become more uniform with 

Appendix D3 449 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

growth; and (5) particle growth is similar in humans and rodents. An uncertainty factor of 10 

was applied because of the subchronic duration of the study. A factor of 3 was applied for 

interspecies extrapolation in light of the fact that some correction for human equivalency was 

made with the RDDR. Finally, a factor of 10 was applied for protection of potentially sensitive 

human subpopulations. The resulting RfC for phosphoric acid is 0.01 mg/m3. 

OEHHA uses a BMC05 for development of acute Reference Exposure Levels (OEHHA, 1999; 

Fowles et al., 1999). OEHHA staff believe that the BMC05 is more likely to approximate a 

NOAEL than a BMC10 since 5% is closer than 10% to the lower end of average risk levels 

associated with a NOAEL (Leisenring and Ryan, 1992). A BMC05 is more likely to represent a 

value close to the limit of most studies to detect an effect, and is therefore more like a NOAEL. 

In contrast, a BMC10 is more likely to represent a LOAEL since it is usually in the detectable 

range of responses. In the specific case of phosphoric acid the BMC10 of 100 mg/m3 was twice 

the NOAEL of 50 mg/m3

. The BMC05 was calculated to be 64 mg/m3, much closer to the 

NOAEL. Use of the BMC05 results in a chronic REL of 7 μg/m3

. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for phosphoric acid include the availability of subchronic 

inhalation exposure data from a well-conducted study with histopathological analysis and the 

observation of a NOAEL. Major areas of uncertainty are the lack of adequate human exposure 

data, the lack of chronic inhalation exposure studies, and the discontinuous nature of exposures 

(only 2 1/4 hours per day). 

The Aranyi et al. (1988a) study represents the most adequate study for the quantitative 

evaluation of the toxicity of phosphoric acid. It was conducted with a large number of animals 

with multiple doses, produced good dose-response data, and examined likely targets of toxicity 

(respiratory system) of smoke generated from the combustion of phosphorus and butyl rubber. 

Uncertainties associated with these data, however, include that (1) the study used combustion 

products of phosphorus rather than phosphoric acid itself, (2) the total exposure time was 

relatively short and discontinuous over the duration of the experiment, and (3) only one 

species/strain/sex was studied. 

VIII. References 

Allen BC, Kavlock RJ, Kimmel CA, and Faustman EM. 1994. Dose-response assessment for 

developmental toxicity. II. Comparison of generic benchmark dose estimates with no observed 

adverse effect levels. Fundam. Appl. Toxicol. 23:487-495. 

Aranyi C, Henry MC, Vana SC, and Iverson WO. 1988a. Effects of multiple intermittent 

inhalation exposures to red phosphorus/butyl rubber obscurant smokes in Sprague-Dawley rats. 

Inhal. Toxicol. 1:65-78. 

Appendix D3 450 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Aranyi C, Vana S, Bradof J, and Sherwood R. 1988b. Effects of inhalation of red phosphorusbutyl rubber combustion products on alveolar macrophage responses in rats. J. Appl. Toxicol. 

8:393-398. 

Burton F, Clark M, Miller R, and Schirmer R. 1982. Generation and characterization of red 

phosphorus smoke aerosols for inhalation exposure of laboratory animals. Am. Ind. Hyg. Assoc. 

J. 43:767-72. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Condray J. 1985. Elemental yellow phosphorus one-generation reproduction study in rats. IR-82-

215. IRD No. 401-189. St. Louis, MO: Monsanto Company. 

Dutton CB, Pigeon MJ, Renzi PM, Feustel PJ, Dutton RE, and Renzi GD. 1993. Lung function 

in workers refining phosphorus rock to obtain elementary phosphorus. J. Occup. Med. 35:1028-

33. 

Faustman EM, Allen BC, Kavlock RJ, and Kimmel CA. 1994. Dose-response assessment for 

developmental toxicity. I. Characterization of database and determination of no observed effect 

levels. Fundam. Appl. Toxicol. 23:478-86. 

Fowles JR, Alexeeff GV, Dodge D. 1999. The use of benchmark dose methodology with acute 

inhalation lethality data. Regul. Toxicol. Pharmacol. 29(3):262-278. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 

7/31/96). 

HSDB. 1999. Hazardous Substances Data Bank. Available online at http://sis.nlm.nih.gov 

Hughes JPW, Baron R, Buckland DH, et al. 1962. Phosphorus necrosis of the jaw: a present-day 

study. Br. J. Ind. Med. 19:83-99. 

Inuzuka H. 1956. A study on the experimentally-created chronic poisoning of yellow phosphorus 

to bones. Igaku Kenkyu, 26:2848-61. 

Leisenring W, Ryan L. 1992. Statistical properties of the NOAEL. Regul. Toxicol. Pharmacol. 

15(2 Pt 1):161-171. 

Marrs T. 1984. Histological changes produced by exposure of rabbits and rats to smokes 

produced from red phosphorus. Toxicol. Lett. 21:141-6. 

Marrs T, Colgrave H, Edginton J, Rice P, and Cross N. 1989. The toxicity of a red phosphorus 

smoke after repeated inhalation. J. Haz. Mat. 22:269-82. 

Appendix D3 451 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

OEHHA. 1999. Air Toxics Hot Spots Program Risk Assessment Guidelines. Part I. Technical 

Support Document for Determination of Acute Reference Exposure Levels for Airborne 

Toxicants. Oakland, CA: Office of Environmental Health Hazard Assessment, Air Toxicology & 

Epidemiology Section. Available online at http://www.oehha.ca.gov 

U.S. DOD. 1981. US Department of Defense. AD-B058-048. Available from Defense Technical 

Information Center, Alexandria, VA 22314 (703) 274-7633. 

U.S. DOD. 1982. US Department of Defense. AD-A118-682. Available from Defense Technical 

Information Center, Alexandria, VA 22314 (703) 274-7633. 

U.S. EPA. April 1989. An acid aerosols issue paper: health effects and aerometrics. Prepared by 

the Office of Health and Environmental Assessment, Environmental Criteria and Assessment 

Office. EPA/600/8-88/005F. Research Triangle Park, NC: U.S. EPA. 

U.S. EPA. 1995. United States Environmental Protection Agency. Documentation of the 

reference concentration for chronic inhalation exposure (RfC) for phosphoric acid. Integrated 

Risk Information System (IRIS on-line). Washington, DC: U.S. EPA. 

Appendix D3 452 Phosphoric Acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

PHTHALIC ANHYDRIDE 

 (1,3-isobenzofurandione; phthalic acid anhydride) 

CAS Registry Number: 85-44-9 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 20 μg/m3

Critical effect(s) Eye and respiratory irritation, asthma, and 

bronchitis in occupationally exposed workers 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994) 

Description White or pale yellow crystals 

Molecular formula C8H4O3

Molecular weight 148.11 g/mol 

Boiling point 295°C 

Melting point 130.8°C 

Vapor pressure 5.14 × 10-4 torr @ 25°C; 1 torr @ 96.5°C 

Solubility Soluble in 162 parts water, 125 parts carbon 

disulfide; soluble in hot benzene 

Conversion factor 1 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

The primary use of phthalic anhydride (PA) is as a chemical intermediate in the production of 

plastics from vinyl chloride. Phthalate esters, which function as plasticizers, are derived from 

phthalic anhydride. Phthalic anhydride has another major use in the production of polyester 

resins and other minor uses in the production of alkyd resins used in paints and lacquers, certain 

dyes (anthraquinone, phthalein, rhodamine, phthalocyanine, fluorescein, and xanthene dyes), 

insect repellents, and urethane polyester polyols. It has also been used as a rubber scorch 

inhibitor and retarder (HSDB, 1995; National Cancer Institute (NCI), 1979). The annual 

statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California based on the most recent inventory were estimated to be 11,442 pounds of phthalic 

anhydride (CARB, 2000). 

Appendix D3 453 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposure 

Symptoms in workers exposed to phthalic anhydride by inhalation in two plants (A and B) 

manufacturing alkyd and unsaturated polyester resins were studied (Nielsen et al., 1988). Two 

groups of exposed workers were identified in each plant. One group worked directly loading the 

reactors from bags of phthalic anhydride (“heavy” exposure – 35 workers) and the other group 

was involved with “other work” which led to “low” exposure (25 workers). Mean employment 

times for the “heavy” and “low” exposure groups were 13.3 and 11.9 years, respectively. Timeweighted average air concentrations for workers from the loading of PA was 6.1 (range: 1.8-

14.9) and 6.8 mg PA/m3

 (range: 1.5-17.4) in plants A and B, respectively. Similar exposure 

levels in both plants led to pooling of data. The exposure duration of the “heavy” group was 

estimated at approximately 30 minutes two times a day, corresponding to the time of loading, 

and resulted in a full-day time weighted exposure estimate of 0.4 mg PA/m3. For those engaged 

in “other work” exposure levels were estimated at < 0.1 mg PA/m3

 (the limit of detection). 

Other chemicals in use in smaller amounts included maleic anhydride, isophthalic anhydride, and 

trimellitic anhydride. Comparison of symptom incidence between the “heavy” and “low” 

exposure groups included conjunctivitis (46% vs. 20%), rhinitis (40% vs. 20%), 

rhinoconjunctivitis (17% vs. 12%), asthma (17% vs. 0%), and chronic bronchitis (17% vs. 4%). 

Serum antibodies were measured in both groups of workers and compared to 22 nonexposed 

workers (employed at a food processing factory). The only significantly changed level was an 

increase in specific IgG in the “heavy” exposure group. A correlation was also noted between 

specific IgG level and exposure level, although not all individuals with elevated specific IgG 

reported symptoms. 

In a study conducted at another plant manufacturing alkyd and/or unsaturated polyester resins, 

serum immunoglobulins and lung function were examined in 23 workers exposed to phthalic 

anhydride and 18 control subjects (Nielsen et al., 1991). Estimated exposure levels were 6.6 mg 

PA/m3

 (range: 1.5-17) (Nielsen et al., 1988). Workers were examined for sensitization to PA 

and other allergens and possible development of small airways disease. Among the exposed 

workers, there was significantly increased reporting of conjunctivitis and rhinoconjunctivitis. 

One worker showed an asthmatic response to anhydrides. No significant differences in lung 

function tests were observed between exposed and unexposed groups. 

Symptoms in workers occupationally exposed to PA during the course of producing alkyd and/or 

polyunsaturated polyester resins were described (Wernfors et al., 1986). Exposure estimates of 

breathing zone PA levels ranged from 3 to 13 mg/m3

 for workers engaged directly with the 

handling of PA. In other areas the estimated level was <0.3 mg/m3

. The study examined 48 

workers who were employed at the time of the study and 70 former employees who responded to 

a survey of symptoms related to exposure. No unexposed control group was included in the 

study. Workers who were employed for at least two months reported symptoms of rhinitis 

(28%), chronic bronchitis (11%), and asthma (28%). Among a subset of 11 workers with 

asthma, 3 had positive skin tests for PA sensitivity. Bronchial provocation tests with 6 or 0.5 

mg/m3 PA for 5 or 10 minutes were positive in 2 workers. 

Appendix D3 454 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

V. Effects of Animal Exposure 

Male albino rats (6/treatment group) were exposed to phthalic anhydride vapors at 0, 0.02, 0.2, 

and 1 mg/m3 continuously for 45 days (Protsenko, 1970). After a two week recovery period the 

testes were examined for spermatozoa motility time as well as for ascorbic acid, dehydroascorbic 

acid, and nucleic acid content. Motility time was defined as the time it took for spermatozoa to 

cease motion completely under microscopic examination. Spermatozoa motility time was 

decreased ~50% in the 1 mg/m3

 dose group and ~25% in the 0.2 mg/m3 dose group. Significant 

decreases in ascorbic acid and dehydroascorbic acid levels were found in animals exposed to 0.2 

and 1.0 mg/m3 phthalic anhydride, and dehydroascorbic acid levels were decreased in the 

0.02 mg/m3 dose group. At 1 mg/m3, RNA levels and combined RNA and DNA levels were 

significantly increased over controls. No significant changes were observed in the 0.02 mg/m3

dose group. 

Five and six female Hartley guinea pigs were exposed to 0.05-0.2 mg/m3 and 0.6-6 mg/m3

phthalic anhydride dust, respectively, for 3 hours/day for 5 consecutive days (Sarlo and Clark, 

1992). Exposures were expressed as ranges due to difficulty in regulating dust levels in the 

chambers. Sampling of dust showed particles were 65-80% < 10 μm diameter and had a mean 

mass diameter of 5.8-9.8 μm. Eight control animals were exposed to filtered air only. Two 

weeks after the last exposure, animals were challenged for 30 minutes with aerosolized PAguinea pig serum albumin conjugate. All animals in the “high” dose group showed immediate 

bronchoconstriction and transiently increased respiratory rate. Animals in this dose group also 

showed elevated IgG antibody titers. No detectable increase in antibody levels was found in the 

“low” dose group. 

Type I hypersensitivity was examined in female Hartley guinea pigs exposed to phthalic 

anhydride dust (Sarlo et al., 1994). Two groups of 8 animals were exposed to 0.5 or 1.0 mg/m3

, 

and two groups of 16 animals were exposed to 0 (filtered air only) or 5.0 mg/m3 phthalic 

anhydride dust (respirable size – 5 μm) in stainless steel chambers for 3 hours/day for 5 

consecutive days. Groups of 8 animals from the control and 5 mg/m3 groups were challenged 

after a two week recovery period for 30 minutes with 5.0 mg/m3

 phthalic anhydride dust. 

Respiratory data were collected using a plethysmograph from 30 minutes before the exposure to 

60 minutes after the exposure. No significant difference (defined as a change of 3 standard 

deviations from the same parameter in the control animals) in respiration rate or plethysmograph 

pressures was found between the exposed and unexposed animals. Eight animals in each of the 

four exposure groups were also challenged after two weeks of recovery with 2.0 mg/m3

aerosolized PA-guinea pig serum albumin (GPSA) conjugate as described above. Respiratory 

rate was increased in 4/8 of the high-dose group animals and 1/8 of the low-dose animals. 

Plethysmograph pressures were increased in 3/8 animals in the high-dose group and one animal 

each in the low- and mid-dose groups. Serum IgG antibodies to PA-GPSA were elevated in all 

exposed animal groups and the effect showed a dose-response. Passive cutaneous anaphylaxis 

testing for anti-phthalic anhydride-GPSA IgG1a immunoglobulins showed positive results for 

3/8, 1/8, and 5/8 animals in the 0.5, 1.0, and 5.0 mg/m3

 dose groups, respectively. Results in 

control animals were not described. Three of eight animals in the highest dose group had >189 

hemorrhagic foci in their lungs. No control animal had more than 2 such foci. No foci were 

Appendix D3 455 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

observed in animals challenged with albumin conjugate. Serum IgG titer correlated with the 

presence of these foci. 

Slavgorodskiy (1969) studied the toxicity of phthalic anhydride to animals from inhalation 

exposure. Sixty white male rats (strain not reported; group distribution not stated, but presumed 

to be 15 animals/treatment group) were exposed in 100 L chambers to 0, 0.18, 0.54, and 1.52 mg 

PA/m3

 aerosol continuously for 70 days. General condition and behavior, body weight, motor 

chronaxy of flexor and extensor muscles (every 10 days), cholinesterase activity (every two 

weeks), and hematological parameters were monitored during the course of the study. 

(Chronaxy is the minimum time for which a current must flow, at a voltage twice the minimal 

current necessary to produce muscle stimulation, in order to cause a muscle to contract.) No 

changes in body weight or behavior were observed in the treated animals. In animals in the highdose group, the chronaxy ratio of flexors and extensors differed from the controls beginning on 

day 31 of exposure and continued until two weeks after exposure ceased. Significantly 

decreased whole blood cholinesterase activity occurred in the high- and mid-dose groups, with 

the change occurring after 42 days of exposure. An increase in thrombocyte count occurred in 

the high- and mid-dose groups after 70 days of exposure, but returned to normal during the twoweek recovery period. Thus, 0.18 mg/m3

 PA appears to be a NOAEL in this study. 

A chronic feeding study was conducted with phthalic anhydride in rats and mice to evaluate the 

carcinogenicity of the compound (National Cancer Institute (NCI), 1979). F344 rats 

(50/sex/dose group plus 20/sex control animals) were treated with diet containing 0, 7500, or 

15,000 ppm phthalic anhydride for 105 weeks (which corresponds to approximately 0, 300, and 

600 mg/kg-day, assuming that food consumption is 4% body weight/day). Animals were 

monitored for changes in body weight and for survival, and, upon death or the end of the study, 

were examined histopathologically. The only group showing significantly lower body weights 

was male rats in the high-dose group after week 13. No significant change in mortality was 

observed. Adverse non-cancer effects observed in the dosed groups, but not in the control 

animals, included “arched back, rough hair coat, ulceration, and corneal opacity”, however, 

incidences were described as “low”. No significant histopathological effects were found to be 

associated with exposure to phthalic anhydride. B6C3F1 mice (50/sex/dose group plus 20/sex 

control animals) were initially treated with diet containing 0, 25,000, or 50,000 ppm phthalic 

anhydride (approximately 0, 3000, and 6000 mg/kg-day, assuming that food consumption is 12% 

body weight/day). Because of excessive weight loss after week 32, exposure levels were 

reduced during the course of the study such that the time-weighted average exposure for males 

was 16,346 and 32,692 ppm and for females was 12,019 and 24,038 ppm phthalic anhydride. 

Evaluation of toxicity was conducted at 104 weeks as with the rats. Mean body weight was 

reduced in male and female mice in a dose-related manner. No other significant treatmentrelated adverse effects were observed in the mice. 

Pregnant female CD-1 mice (10/dose group) were treated intraperitoneally with phthalic 

anhydride in 0.5%(w/v) carboxymethyl cellulose solution on gestational days 8-10 (Fabro et al., 

1982). Dosing was variable, beginning within the 95% confidence limits of the LD01 and 

progressing geometrically downward until no effect was observed. Animals were terminated on 

Day 18 and examined for teratogenic effects including fetal viability and number, resorption, and 

gross malformations. The 95% lower confidence limit on the dose producing teratogenicity 

Appendix D3 456 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

(grossly observable malformations and fetal internal malformations) in 5% and 50% of animals 

were 0.40 and 1.37 mmol/kg-day (59 and 203 mg/kg-day), respectively. 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Neilsen et al. (1988; 1991) 

Study population 23 occupationally-exposed workers 

Exposure method Discontinuous occupational inhalation 

exposures 

Critical effects Increased incidence of conjunctivitis, rhinitis, 

asthma, and chronic bronchitis 

LOAEL 6.5 mg/m3

 (mean of 6.1 and 6.8) 

NOAEL Not observed 

Exposure continuity 8 hours/day, 5 days/week 

Exposure duration Mean of 13.3 years 

Average experimental exposure 2.3 mg/m3

 for LOAEL group 

 (6.5 mg/m3 × 10/20 × 5/7) 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.02 mg/m3

 (20 μg/m3) 

Adverse effects were demonstrated to occur in humans occupationally exposed to phthalic 

anhydride in the workplace over long periods of time (Nielsen et al., 1988). The symptoms 

reported primarily affected the respiratory system, with increased incidence of rhinitis, 

rhinoconjunctivitis, asthma, and chronic bronchitis. Conjunctivitis was also reported in exposed 

workers. Specific anti-PA IgG was significantly elevated compared to a non-exposed group. 

Increased incidences of rhinoconjunctivitis, conjunctivitis, or chronic bronchitis have also been 

reported in workers exposed to similar levels of PA dust (Nielsen et al., 1991; Wernfors et al., 

1986). In these reports, adverse effects were clearly observed at the exposure level reported 

(6.5 mg PA/m3

; full-day time weighted exposure of 0.4 mg PA/m3). Although symptoms were 

reported by Nielsen (1988) in the lower exposure level group, the significance is not clear since a 

true control group (unexposed workers) was not included in the symptomatology section of the 

study. The low exposure group’s level of exposure was less than the detection limit for phthalic 

anhydride cited in the study, and this group was considered as a control group. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for phthalic anhydride include the use of human exposure 

data from workers exposed over a period of years. Major areas of uncertainty are (1) the 

uncertainty in estimating exposure, (2) the potential variability in exposure concentration, (3) the 

Appendix D3 457 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

potential low exposures of the group considered as controls, (4) potential confounding by 

exposures to other chemicals, (5) the limited nature of the study, (6) the lack of reproductive and 

developmental toxicity studies, and (6) the lack of observation of a NOAEL in the key study. 

Another area of uncertainty is the apparent 10-fold greater sensitivity to bronchoconstriction 

from PA exposure in guinea pigs (a model for human asthmatics) in comparison to 

occupationally exposed workers. 

The study in rats by Protsenko (1970) identified a LOAEL of 0.2 mg/m3

 and a NOAEL of 0.02 

mg/m3

 for decreased sperm motility. However, this result from 1970 has not been verified or 

further explored in more recent toxicological or epidemiological studies. The small sample size 

of 6/group further weakens confidence in this result. Therefore, the study in workers by Nielson 

et al. (1988, 1991) was chosen as the basis for the REL for PA. 

VIII. References 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Fabro S, Shull G, and Brown NA. 1982. The relative teratogenic index and teratogenic potency: 

proposed components of developmental toxicity risk assessment. Teratog. Carcinog. Mutag. 

2:61-76. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. 

NCI. 1979. National Cancer Institute. Bioassay of Phthalic Anhydride for Possible 

Carcinogenicity. Springfield, VA: NTIS. PB 293-594. 

Nielsen J, Bensryd I, Almquist H, Dahlqvist M, Welinder H, Alexandersson R, and Skerfving S. 

1991. Serum IgE and lung function in workers exposed to phthalic anhydride. Int. Arch. Occup. 

Environ. Health 63:199-204. 

Nielsen J, Welinder H, Schütz A, and Skerfving S. 1988. Specific serum antibodies against 

phthalic anhydride in occupationally exposed subjects. J Allergy Clin. Immunol. 82:126-133. 

Protsenko EI. 1970. [Effect of phthalic anhydride on gonads]. Gig. Sanit. 35:127-130. [in 

Russian] 

Sarlo K, and Clark ED. 1992. A tier approach for evaluating the respiratory allergenicity of low 

molecular weight chemicals. Fundam. Appl. Toxicol. 18:107-114. 

Appendix D3 458 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Sarlo K, Clark ED, Ferguson J, Zeiss CR, and Hatoum N. 1994. Induction of type I 

hypersensitivity in guinea pigs after inhalation of phthalic anhydride. J. Allergy Clin. Immunol. 

94:747-756. 

Slavgorodskiy LP. 1969. [Biological effect and hygienic evaluation of pollution of atmospheric 

air with phthalic anhydride.] Akademiya Meditsinskikh Nauk SSSr 10:86-96. [in Russian] 

Wernfors M, Nielsen J, Schütz A, and Skerfving S. 1986. Phthalic anhydride-induced 

occupational asthma. Int. Arch. Allergy Immunol. 79:77-82. 

Appendix D3 459 Phthalic anhydride 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

PROPYLENE 

(1-propene; 1-propylene; propene; methylethene; methylethylene) 

CAS Registry Number: 115-07-1 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 3,000 μg/m3

 (2,000 ppb) 

Critical effect(s) Squamous metaplasia (males and females), 

epithelial hyperplasia (females only), and 

inflammation (males only) of the nasal cavity 

in Fischer 344/N rats 

Hazard index target(s) Respiratory system 

II. Chemical and Physical Properties (HSDB, 1995; CRC, 1994) 

Description Colorless gas; practically odorless. 

Molecular formula C3H6

Molecular weight 42.08 

Boiling point −47.6 °C 

Melting point −185.2°C 

Vapor pressure 8690 torr at 25°C 

Solubility Soluble in alcohol and ether. 

Conversion factor 1.72 μg/m3

 per ppb at 25°C 

III. Major Uses and Sources (HSDB 1995) 

Propylene is produced primarily as a by-product of petroleum refining and of ethylene 

production by steam cracking of hydrocarbon feedstocks. Propylene is a major chemical 

intermediate. The most important derivatives of chemical and polymer grade propylene are 

polypropylene, acrylonitrile, propylene oxide, isopropanol and cumene. Use of polypropylene in 

plastics (injection moulding) and fibers (carpets) accounts for over one-third of U.S. 

consumption. It is also used in the production of synthetic rubber and as a propellant or 

component in aerosols. In 1994, propylene was ranked seventh among the top 50 chemicals 

produced domestically (C&EN, 1995). In the environment, propylene occurs as a natural 

product from vegetation. It is also a product of combustion of organic matter (biomass burning, 

motor vehicle exhausts and tobacco smoke) and is released during production and use. The most 

probable route of exposure to humans is by inhalation. Propylene has been detected in the 

atmosphere over both metropolitan (2.6 to 23.3 ppb) and rural (0.007 to 4.8 ppb) areas (Cox et 

al., 1976; Leonard et al., 1976). The annual statewide emissions from facilities reporting under 

Appendix D3 460 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

the Air Toxics Hot Spots Act in California, based on the most recent inventory, were estimated 

to be 696,350 pounds of propylene (CARB, 1999). 

IV. Effects of Human Exposures 

No data were available on the absorption, distribution or excretion of propylene in humans. 

However, hemoglobin (Hb) adducts of the metabolite intermediate propylene oxide have been 

used to monitor the internal dose of propylene (Tornqvist and Ehrenberg, 1990). The 

background level of the 2-hydroxylpropyl adduct to the N-terminal valine of hemoglobin was 

found to be about 2 pmol/g Hb. This was estimated to be equivalent to smoking 10 cigarettes per 

day; cigarette smoking is a source of propylene. Occupational exposure to propylene at 1 ppm 

(1.72 mg/m3

) was assumed to be associated with an increment of 5 pmol/g Hb (Kautiainen and 

Tornqvist, 1991). 

No data were available on the chronic effects of propylene in humans. 

V. Effects of Animal Exposures 

In rats and mice, most propylene inhaled into the lungs is exhaled again and does not reach the 

blood to become systemically available (Golka et al., 1989; Svensson and Osterman-Golkar, 

1984). Once absorbed, a major route of metabolism for propylene is through the cytochrome P450 system to propylene oxide, a known carcinogen in experimental animals. Cytochrome P450 enzymes in both the liver and nasal epithelium (Maples and Dahl, 1991) can convert 

propylene to its toxic metabolite. However, in rats, propylene metabolism becomes increasingly 

saturated at concentrations above 50 ppm (86 mg/m3

) in the atmosphere (Golka et al., 1989), 

which limits the amount of propylene oxide produced. Therefore, the amount of absorbed 

propylene may not reach high enough levels in classical long-term inhalation studies (Quest et 

al., 1984) to show positive carcinogenic or serious chronic effects. 

The only chronic toxicity investigation found for propylene was a comprehensive 2-year study in 

F344/N rats and B6C3F1 mice (Quest et al., 1984; NTP, 1985). Groups of 50 rats and 50 mice of 

each sex were exposed to concentrations of 0, 5000, and 10,000 ppm for 6 hr/day, 5 days/week, 

for 103 weeks. (Mean daily concentrations were 0, 4985, and 9891 ppm, respectively, for the rat 

study; and 0, 4999, and 9957 ppm, respectively, for the mouse study.) In exposed rats, 

treatment-related chronic effects were observed in the nasal cavity. In female rats, epithelial 

hyperplasia occurred in the high dose group and squamous metaplasia occurred in both dosage 

groups. In male rats, squamous metaplasia was seen only in the low dose group, but both dosage 

groups had inflammatory changes characterized by an influx of lymphocytes, macrophages and 

granulocytes into the submucosa and granulocytes into the lumen (see below). Nasal lesions 

were not observed in mice. The inflammatory lesions were more severe in the high dose group. 

Very mild focal inflammation was observed in the kidneys of treated mice but the relationship to 

propylene exposure was unclear. No other treatment-related effects, including clinical signs, 

mortality, mean organ and body weights, and histopathology, were observed. 

Appendix D3 461 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Incidences of epithelial changes in nasal cavities of rats (Table 2 from Quest et al., 1984) 

Observation Control 5000 ppm 10,000 ppm 

Epithelial hyperplasia 

 Male 2/50 (4%) 2/50 (4%) 5/50 (10%) 

 Female 0/49 (0%) 4/50 (8%) 9/50 (18%)* 

Squamous metaplasia 

 Male 2/50 (4%) 19/50 (38%)* 7/50 (14%) 

 Female 0/49 (0%) 15/50 (30%)* 6/50 (12%)* 

Inflammation 

 Male 11/50 (22%) 21/50 (42%)* 19/50 (38%) 

 Female 8/49 (16%) 10/50 (20%) 13/50 (26%) 

* Significantly (p < 0.05) higher than control values 

In a long-term carcinogenicity study, Sprague-Dawley rats and Swiss mice (100-120 

animals/group/sex) were exposed by inhalation to 0, 200, 1000 and 5000 ppm propylene 7 

hr/day, 5 days/week, for 104 weeks (rats) or 78 weeks (mice) (Ciliberti et al., 1988). No body 

weight differences were observed between treated and control animals of either species. 

Mortality was reported to be slightly increased in male rats in the 1000 and 5000 ppm groups and 

in male mice in the 5000 ppm group, but numerical values of mortality were not presented in the 

report. Therefore, it is assumed that mortality differences were insignificant. Other possible 

general body system or nonneoplastic effects were not reported and assumed to have not been 

investigated. 

Appendix D3 462 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Quest et al., 1984 

Study population 50 rats/group/sex, 300 total. 

Exposure method Discontinuous whole body inhalation exposure 

(0 or 4,985 or 9,891 ppm). 

Critical effects Respiratory system; squamous metaplasia (males 

and females), epithelial hyperplasia (females 

only), and inflammation (males only) of the 

nasal cavity 

LOAEL 4,985 ppm (8,570 mg/m3

) 

NOAEL Not observed 

Exposure continuity 6 hr/day, 5 days/week

Exposure duration 2 years 

Average experimental exposure 890 ppm for LOAEL group (4985 x 6/24 x 5/7) 

Human equivalent concentration 190 ppm (gas with extrathoracic respiratory 

effects, RGDR = 0.21, based on BW = 305 g, 

MV = 0.21 L/min, SA(ET) = 15 cm2) 

LOAEL uncertainty factor 3 (low severity) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies factor 10

Cumulative uncertainty factor 100 

Inhalation reference exposure level 2 ppm (2,000 ppb, 3 mg/m3

, 3,000 μg/m3) 

VII. Data Strengths and Limitations for Development of the REL

Strengths of the propylene REL include the availability of a long-term, controlled exposure study 

in large groups of experimental animals that included extensive histopathological analyses. 

Lifetime exposure of rats and mice to propylene resulted in adverse effects in the nasal cavity of 

rats at both exposure levels. Therefore, a NOAEL was not observed. However, the effects 

observed were mild. 

Other weaknesses of the database for propylene include the lack of lifetime toxicity studies in 

any non-rodent species. Also, no long-term human toxicity or epidemiology studies were 

located in the literature. Human pharmacokinetic studies to compare with experimental animal 

pharmacokinetic studies were absent. Another uncertainty is the lack of reproductive and 

developmental toxicity studies. A comprehensive multi-generation study in an experimental 

animal species would enhance the development of a propylene REL. 

Appendix D3 463 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VIII. References 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

C&EN. 1995. Chemical & Engineering News. Production by the U.S. chemical industry. M. 

Heylin, ed. American Chemical Society, Washington DC, June 26:38-44. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Ciliberti A, Maltoni C, and Perino G. 1988. Long-term carcinogenicity bioassays on propylene 

administered by inhalation to Sprague-Dawley rats and Swiss mice. Ann. N.Y. Acad. Sci. 

534:235-245. 

Cox RA, Derwent RG, and Sandalls FJ. 1976. Some air pollution measurements made at 

Harwell, Oxfordshire, during 1973-1975, AERE-R 8324, Harwell, U.K. Atomic Energy 

Research Establishment, Environmental and Medical Sciences Division. 

Golka K, Peter H, Denk B, and Filser JG. 1989. Pharmacokinetics of propylene and its reactive 

metabolite propylene oxide in Sprague-Dawley rats. Arch. Toxicol. Suppl. 13: 240-242. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (edition expires 11/31/95). 

Kautiainen A, and Tornqvist M. 1991. Monitoring exposure to simple epoxides and alkenes 

through gas chromatographic determination of hemoglobin adducts. Int. Arch. Occup. Environ. 

Health 63:27-31. 

Leonard MJ, Fisher EL, Brunelle MF, and Dickinson JE. 1976. Effects of the motor vehicle 

control program on hydrocarbon concentrations in the Central Los Angeles atmosphere. J. Air 

Pollut. Contr. Assoc. 26:359-363. 

Maples KR, and Dahl AR. 1991. Blood levels of propylene oxide during propylene inhalation 

and effect on hepatic and nasal cytochrome P-450 concentrations. Drug Metab. Dispos.

19(4):835-837. 

NTP. 1985. National Toxicology Program. Toxicology and carcinogenesis studies of propylene 

in F344/N rats and B6C3F1 mice. Technical Report Series NTP-TR-272, NIH publication no. 86-

2528. 

Quest JA, Tomaszewski JE, Haseman JK, Boorman GA, Douglas JF, and Clarke WJ. 1984. 

Two-year inhalation toxicity study of propylene in F344/N rats and B6C3F1 mice. Toxicol. Appl. 

Pharmacol. 76:288-295. 

Appendix D3 464 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Svensson K, and Osterman-Golkar S. 1984. Kinetics of metabolism of propene and covalent 

binding to macromolecules in the mouse. Toxicol. Appl. Pharmacol. 73:363-372. 

Tornqvist M, and Ehrenberg L. 1990. Approaches to risk assessment of automotive engine 

exhausts. In: Vainio H, Sorsa M, and McMicheal AJ, eds. Complex Mixtures and Cancer Risks. 

IARC Scientific Publications No. 104. Lyon: IARC, pp. 272-287. 

Appendix D3 465 Propylene 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

PROPYLENE GLYCOL MONOMETHYL ETHER 

(1-Methoxy-2-propanol; 1-methoxypropanol; Propapsol solvent M) 

CAS Registry Number: 107-98-2 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 7,000 μg/m3

 (2000 ppb) 

. 

Critical effect(s) Liver effects in rats 

Hazard index target(s) Alimentary system (liver) 

II. Physical and Chemical Properties (HSDB, 1995)

Description Colorless liquid 

Molecular formula C4H10O2

Molecular weight 90.14 

Density 0.962 g/cm3

 @ 20° C 

Boiling point 118-118.5°C 

Melting point -96.7°C 

Vapor pressure 11.8 torr @ 25°C 

Solubility Soluble in water, methanol, ether, and other 

organic solvents 

Conversion factor 1 ppm = 3.69 mg/m3

 at 25° C 

III. Major Uses or Sources 

Propylene glycol monomethyl ether (PGME) is used as a solvent for cellulose, acrylics, dyes, 

inks and stains (HSDB, 1995). Thus, the primary use of PGME is in lacquers and paints. Use of 

PGME is anticipated to increase due to its low systemic toxicity. The annual specific statewide 

industrial emissions of PGME from facilities reporting under the Air Toxics Hot Spots Act in 

California, based on the most recent inventory, were estimated to be 205,769 pounds (CARB, 

1999). (Many industries did not estimate emissions of specific glycol ethers so that in 1998 there 

were also emitted 2,922,744 pounds of the general category glycol ethers, some of which may be 

PGME.) 

Appendix D3 466 Propylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Exposures on Humans 

No reports or studies of human toxicity following chronic exposure to PGME were located in the 

literature. Slight eye irritation was reported by two of six human volunteers exposed to 100 ppm 

PGME for 2 hours (Stewart et al., 1970). These subjects were exposed for a total of 3 1/2 hours 

during which no decrement in visual acuity, coordination, neurological responses or reaction 

time was measured. The same experiment exposed 23 subjects to 250 ppm PGME. After 15 to 

30 minutes of exposure, 8/23 reported eye irritation and 3/23 reported throat irritation; 

lacrimation was observed in 3/23 subjects. Three subjects each reported one of the following 

symptoms: irritation, headache, and nausea. While the subjects frequently reported the odor to 

be objectionable upon first entering the chamber, the odor was usually undetectable by the end of 

the exposure. Clinical chemistry and urinalysis completed following exposure was not altered as 

compared to pre-exposure measurements. 

V. Effects of Exposures on Animals 

Male and female rats (10 per sex per concentration) and rabbits (7 per sex per concentration) 

were exposed by inhalation to 300, 1000, or 3000 ppm PGME 5 hours per day, 5 days per week 

for 13 weeks (Landry et al., 1983). Relative liver weights were statistically significantly higher 

than controls in both male and female rats exposed to 3000 ppm PGME. Hepatocellular 

hypertrophy was observed upon histopathologic examination of high dose females. The authors 

conclude that these effects are the result of physiologic adaptation rather than a manifestation of 

toxicity. The key observation in this study was sedation of rats and rabbits exposed to 3000 ppm 

PGME. The sedative effects were no longer apparent after 1-2 weeks of exposure. 

Similar findings of mild CNS depression were observed by Hanley et al. (1984). Pregnant rats 

and rabbits were exposed to 500, 1500, or 3000 ppm PGME 6 hours per day either days 6-15 or 

days 6-18 of gestation, respectively. During the first 4-5 days of exposure, rats in the 3000 ppm 

PGME exposure group were lethargic and moderately ataxic. Statistically significant decreases 

in food consumption and maternal body weight gain were also observed during this period. A 

statistically significant increase in the incidence of delayed sternebral ossification was observed 

in the 3000 ppm exposure group. Rabbits exposed to 3000 ppm exhibited mild lethargy during 

the first 1-2 days of exposure with rapid post-exposure recovery. Overall maternal weight gain 

during the exposure (days 6-18 of gestation) was statistically significantly lower than controls. 

No significant effect on fetal birth weight or on pup survival indices (e.g., proportion of pups 

surviving to day 3 post-delivery) was noted following exposure of pregnant rats to 200 or 

600 ppm PGME 6 hours per day on days 6-17 of gestation (Doe et al., 1983). Male rats were 

exposed to 200 or 600 ppm PGME 6 hours per day for 10 consecutive days. No significant 

effects on testicular weight or pathology were observed. 

Increased liver and kidney weights were observed in male and female rats (10 per sex per 

concentration) following exposure to 6000 ppm for 7 hours per day, for 81 exposures over a 

114-day period (Rowe et al., 1954). No histopathological abnormalities were observed at 

necropsy. 

Appendix D3 467 Propylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Ciezlak et al. (1998) evaluated the potential chronic toxicity/oncogenicity and the response of 

liver and kidney tissue of Fischer 344 rats to propylene glycol monomethyl ether (PGME) at 

targeted vapor concentrations of 0, 300, 1000 or 3000 ppm. Groups of 50 male and female rats 

per sex were whole-body exposed under dynamic airflow conditions for 6 hours/day, 5 

days/week for up to 2 years. Parameters evaluated included the general appearance and 

demeanor of animals, in-life body weights, survival, hematology, urinalysis and clinical 

chemistry determinations, survival, selected organ weights, gross and microscopic pathologic 

changes and tumor incidence. (The metabolic and morphological bases for PGME-induced 

sedation, hepatic hypertrophy and renal toxicity were characterized in separate groups of male 

and female rats exposed to PGME for 6, 12 or 18 months. Hepatic enzyme induction and 

cellular proliferation, as well as renal cellular proliferation and accumulation of alpha2u-globulin 

(males only) in the kidneys, were conducted in these separate groups of animals.) 

PGME-induced sedation at 3000 ppm resolved in all animals during the second week of 

exposure in conjunction with the appearance of adaptive changes in the liver (cytochrome P450 

induction and hepatocellular proliferation). Cytochrome P450 (pentoxyresorufin Odemethylase) activities dropped to near control concentrations by week 52, coinciding with a 

return of sedation at 3000 ppm PGME. In male rats, the loss of metabolic adaptation was 

followed by a dose-related increase in altered hepatocellular foci after two years of exposure to 

1000 or 3000 ppm PGME. The kidney toxicity observed in male rats was confirmed 

immunohistochemically as an alpha2u-globulin nephropathy. No statistically-identified increases 

in tumors were observed in any tissue. The authors established a NOEL of 300 ppm PGME for 

the study. 

Ethylene glycol methyl ether (EGME), a structurally related compound, exerts considerable 

toxicity on the blood, thymus, testes, and developing fetus. The toxicity of EGME has been 

linked to its primary metabolite, methoxyacetic acid. Recent comparative toxicity and 

metabolism studies (Miller et al., 1983, Miller et al., 1984) indicate that the relatively low 

systemic toxicity exerted by PGME is due to its different metabolites. Following a single oral 

dose of PGME, the key urinary metabolites identified in rats were propylene glycol and the 

sulfate and glucuronide conjugate of PGME (Miller et al., 1983). 

Appendix D3 468 Propylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Reference Exposure Level 

Study Ciezlak et al., 1998 

Study population Fischer 344 rats (50/sex/concentration) 

Exposure method Discontinuous whole-body inhalation (0, 300, 

1000, or 3000 ppm) 

Critical effects Increased eosinophilic foci of altered hepatocytes 

LOAEL 1000 ppm 

NOAEL 300 ppm 

Exposure continuity 6 hours per day, 5 days per week 

Average experimental exposure 54 ppm for NOAEL group (300 x 6/24 x 5/7) 

Human equivalent concentration 54 ppm for NOAEL group (gas with systemic 

effects, based on RGDR = 1.0 using default 

assumption that lambda (a) = lambda (h)) 

Exposure duration 104 weeks 

LOAEL factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 2 ppm (2000 ppb, 7 mg/m3

, 7000 μg/m3) 

VII. Data Strengths and Limitations for Development of the REL

Strengths of the PGME RfC include the observation of a NOAEL and a LOAEL in the same 

study, and the availability of chronic exposure studies involving multiple concentrations. A 

major area of uncertainty is the lack of human data. 

VIII. References 

AIHA. 1989. American Industrial Hygiene Association. Odor thresholds for chemicals with 

established occupational health standards. Akron, OH: AIHA. 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Ciezlak FS, Crissman JW, Stott WT, and Corley RA. 1998. Propylene glycol monomethyl ether: 

a two-year vapor inhalation chronic toxicity/oncogenicity study and evaluation of hepatic and 

renal cellular proliferation, P450 enzyme induction and protein droplet nephropathy in Fischer 

344 rats. Dow Chemical Company, Midland, MI. 

Doe JE, Samuels DM, Tinston DJ, de Silva Wickramaratne GA. 1983. Comparative aspects of 

the reproductive toxicology by inhalation in rats of ethylene glycol monomethyl ether and 

propylene glycol monomethyl ether. Toxicol. Appl. Pharmacol. 69:43-47. 

Appendix D3 469 Propylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Gingell R, Boatman RJ, Bus JS, Cawley TJ, Knaak JB, Krasavage WJ, Skoulis NP, Stack CR, 

and Tyler TR. 1994. Chapter 31: Glycol ethers and other selected glycol derivatives. In: Patty’s 

Industrial Hygiene and Toxicology. Fourth Edition, Vol. 2, Part D. Clayton GD, and Clayton FE, 

eds. New York: John Wiley and Sons, Inc. 

Hanley TR, Calhoun LL, Yano BL, and Rao KS. 1984. Teratologic evaluation of inhaled 

propylene glycol monomethyl ether in rats and rabbits. Fundam. Appl. Toxicol. 4:784-794. 

HSDB. 1995. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc., (Edition expires 

1/31/95). 

IRIS. 1995. Integrated Risk Information System. (CD-ROM Version). Cincinnati, OH: US 

Environmental Protection Agency, Office of Health and Environmental Assessment, 

Environmental Criteria Assessment Office. 

Landry TD, Gushow TS, and Yano BL. 1983. Propylene glycol monomethyl ether: A 13-week 

inhalation toxicity study in rats and rabbits. Fundam. Appl. Toxicol. 3:627-630. 

Miller RR, Hermann EA, Langvardt PW, McKenna MJ, and Schwetz BA. 1983. Comparative 

metabolism and disposition of ethylene glycol monomethyl ether and propylene glycol 

monomethyl ether in male rats. Toxicol. Appl. Pharmacol. 67:229-237. 

Miller RR, Hermann EA, Young JT, Landry TD, and Calhoun LL. 1984. Ethylene glycol 

monomethyl ether and propylene glycol monomethyl ether: Metabolism, disposition and 

subchronic inhalation toxicity studies. Environ. Health Perspect. 57:233-239. 

Rowe VK, McCollister DD, Spencer HC, Oyen F, Hollingsworth RL, and Drill VA. 1954. 

Toxicology of mono-, di-, and tri-propylene glycol methyl ethers. Arch. Ind. Hyg. Occup. Med. 

9:509-525. 

Stewart RD, Baretta ED, Dodd HC, and Torkelson TR. 1970. Experimental human exposure to 

vapor of propylene glycol monomethyl ether: Experimental human exposure. Arch. Environ. 

Health 20:218-223. 

U.S. EPA. 1986. U.S. Environmental Protection Agency. Health and Environmental Effects 

Profile for Glycol Ethers. Prepared by the office of Health and Environmental Assessment, 

Environmental Criteria and Assessment Office, Cincinnati, OH for the Office of Solid Waste and 

Emergency Response, Washington, DC. EP/540/1-86/052. [cited in U.S. EPA, 1995]. 

Appendix D3 470 Propylene Glycol Monomethyl Ether 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

CHRONIC TOXICITY SUMMARY 

PROPYLENE OXIDE 

(1-,2-propylene oxide; methyl ethylene oxide; propene oxide) 

CAS Registry Number: 75-56-9 

I. Chronic Toxicity Summary

Inhalation reference exposure level 30 μg/m3

 (9 ppb) 

Critical effect(s) Degenerative and hyperplastic changes in the 

 respiratory epithelium of rats 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1994) 

Description Colorless liquid 

Molecular formula C3H6O 

Molecular weight 58.08 

Density 0.83 g/cm3

 @ 20° C 

Boiling point 34.23° C 

Melting point -112.13° C 

Vapor pressure 445 torr @ 20° C 

Solubility Soluble in water, miscible in acetone, benzene, 

 carbon tetrachloride, methanol, ether 

Conversion factor 2.38 mg/m3

 per ppm at 25° C 

III. Major Uses or Sources 

Propylene oxide is used as a fumigant such as in the sterilization of packaged foods. It is also 

used as a chemical intermediate in the production of propylene glycol and glycol ethers and as a 

solvent. Propylene oxide is used in the preparation of surfactants and oil demulsifiers (HSDB, 

1994). The annual statewide emissions from facilities reporting under the Air Toxics Hot Spots 

Act in California, based on the most recent inventory, were estimated to be 619,494 pounds of 

propylene oxide. 

Appendix D3 471 Propylene Oxide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

IV. Effects of Human Exposures 

Conclusive data regarding the effects of occupational exposure to propylene oxide were not 

located. 

An epidemiological study examining mortality among workers with exposure to asbestos and 

several chemicals, including propylene oxide, identified three deaths due to mesothelioma, a rare 

cancer associated with asbestos exposure, and a statistically significant increase in the number of 

deaths attributed to forms of heart disease other than ischemia and hypertension (Egedahl et al., 

1989). The latter finding was explained by the authors to be the result of differences in 

diagnostic accuracy between rural and urban, and primary and tertiary medical care settings. A 

statistically significant decrease in observed deaths was found for all respiratory cancers, cancer 

of the bronchus and lung, circulatory disease, digestive diseases, cirrhosis and other liver disease, 

and death due to accidents, poisonings, and violence. These observations may be partially 

attributed to a “healthy worker effect”. 

V. Effects of Animal Exposures 

Male and female rats were exposed for 124 or 123 weeks (respectively) to 30, 100 or 300 ppm 

propylene oxide 6 hours per day, 5 days per week (Kuper et al., 1988). Interim sacrifices were 

performed at 12, 18, and 24 months. Cumulative mortality was statistically significantly 

different from controls at 115 weeks in rats of both sexes exposed to 300 ppm propylene oxide. 

Cumulative mortality was also significantly different from controls at 119 weeks in female rats 

exposed to 100 ppm. However, a contributing factor to the increased mortality in female rats 

was the presence of mammary tumors. Atrophy of the olfactory epithelium and degenerative 

changes in the respiratory epithelium were observed in both male and female rats following 28 

months of exposure to 30, 100, or 300 ppm propylene oxide. Severe hyperplastic changes in the 

olfactory epithelium were observed in male and female rats following 28 months exposure to 

300 ppm propylene oxide. Mild hyperplastic changes were observed in the olfactory epithelium 

of female rats exposed to 100 ppm propylene oxide. 

Rats and mice were exposed to 200 and 400 ppm propylene oxide 6 hours per day, 5 days per 

week for 103 weeks (NTP, 1985). Survival in mice was adversely affected in all groups exposed 

to propylene oxide; a statistically significant decrease in survival was observed in male and 

female mice exposed to 400 ppm propylene oxide. Survival in rats was not adversely affected by 

propylene oxide exposure. Rats exhibited exposure-related increases in suppurative 

inflammation of the nasal cavity, epithelial hyperplasia and squamous metaplasia. 

Rats were exposed to 1500 ppm propylene oxide 6 hours per day, 5 days per week for 7 weeks 

(Ohnishi et al., 1988). After 3-4 weeks of exposure the rats exhibited an awkward gait; the rats 

were ataxic by the seventh week. Histopathological examination revealed axonal degeneration 

of myelinated fibers of the hindleg nerve and fasciculus gracilis indicating central-peripheral 

distal axonopathy. 

Appendix D3 472 Propylene Oxide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Eldridge et al. (1995) exposed male F344 rats to 0, 10, 20, 50, 150, or 525 ppm propylene oxide 

vapor for up to 4 weeks (with up to 4 weeks of recovery). Histopathology showed that the 

incidence and severity of respiratory epithelial hyperplasia increased with exposure time and 

regressed after termination of exposure, with complete recovery after 4 weeks. Cell proliferation 

(determined by bromodeoxyuridine incorporation) was elevated following 1 and 4 weeks of 

exposure, but decreased to control values after 1 week of recovery. Degeneration of the 

olfactory epithelium was found after 4 weeks of exposure with a decrease in incidence and 

severity after termination of exposure. Proliferation of olfactory epithelium was elevated during 

the 4-week exposure period and 1 week post-exposure and returned to control values after 4 

weeks of recovery. The authors report a 4-week NOAEL for propylene oxide effects in nasal 

epithelium of 50 ppm. 

Artificially inseminated rabbits were exposed to 500 ppm propylene oxide on days 1-19 or 7-19 

of gestation (Hardin et al., 1983). Maternal toxicity as indicated by a significant reduction in 

food intake and a significant decrease in maternal body weight gain was observed in both 

exposed groups. An increased number of resorptions per litter, with no change in total 

resorptions, was observed in rabbits exposed on days 1-19 of gestation. Sternebral and limb 

anomalies (considered minor by U.S. EPA and the authors) were significantly increased in the 

offspring of rabbits exposed on days 1-19 of gestation. 

The same study also reported similar findings in sperm-positive rats exposed to 500 ppm 

propylene oxide on either days 1-16 or 7-16 of gestation or daily for 3 weeks prior to mating and 

then daily on days 1-16 of gestation. Reproductive capacity was impaired in rats exposed prior 

to breeding; the number of corpora lutea, implantation sites, and live fetuses were reduced. 

Those dams exposed pregestationally to propylene oxide also exhibited more resorptions. 

Maternal toxicity as indicated by decreased food intake and decreased body weight gain was 

observed in all exposed rats. Significant reductions in fetal body weight and fetal crown-rump 

length were observed in all exposed groups. An increased incidence of wavy ribs and reduced 

ossification were observed in the offspring of rats exposed from days 1-16 of gestation. 

Harris et al. (1989) evaluated the developmental toxicity potential of propylene oxide in Fischer 

344 rats. Four groups of 25 mated female rats were exposed to 0, 100, 300, and 500 ppm for 6 

hours per day on gestation days 6 through 15. Cesarean sections were performed on all females 

on gestation day 20 and the fetuses were removed for morphological evaluation. Exposure to 

propylene oxide did not adversely affect survival, appearance, or behavior at any level. Maternal 

body weight gain and food consumption were reduced significantly at the 500 ppm level during 

exposure. Only one exposure-related effect was noted with respect to maternal water 

consumption, organ weights, cesarean section, or fetal morphological observations: increased 

frequency of seventh cervical ribs in fetuses at the maternally toxic exposure level of 500 ppm. 

Thus 300 ppm was considered the NOAEL. 

Appendix D3 473 Propylene Oxide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

VI. Derivation of Chronic REL (U.S. EPA Reference Concentration (IRIS, 1995))

Study Kuper et al., 1988 

Study population Rats (male and female) 

Exposure method Inhalation (0, 30, 100 or 300 ppm) 

LOAEL 30 ppm 

Critical effects Degenerative and hyperplastic changes in the 

 respiratory epithelium 

NOAEL Not observed 

Exposure continuity 6 hr/day for 5 days/week 

Exposure duration 124 weeks 

Average experimental exposure 5.4 ppm for LOAEL group (30 x 6/24 x 5/7) 

Human equivalent concentration 1.2 ppm for LOAEL group (gas with 

 extrathoracic respiratory effects, RGDR = 0.23, 

 based on MV = 0.3 m3/day, SA(ET) = 11.6 cm2) 

LOAEL uncertainty factor 3 (mild effects only observed during last 4 

 months of exposure) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.009 ppm (9 ppb, 0.03 mg/m3

, 30 μg/m3) 

VII. Data Strengths and Limitations for Development of the REL

The chronic REL is equivalent to the US EPA RfC. The major strength of the REL for 

propylene oxide is the use of a well-conducted, long-term, multi-concentration study with 

adequate histopathological analyses. Weaknesses include the lack of adquate human data and 

the lack of a chronic NOAEL observation. 

VIII. References 

CARB. 1999. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

Egedahl RD, Olsen GW, Coppock E, Young ML, Arnold IM. 1989. An historical prospective 

mortality study of the Sarnia Division of Dow Chemical Canada Inc., Sarnia, Ontario (1950-

1984).Can. J. Public Health 80(6):441-446. 

Eldridge SR, Bogdanffy MS, Jokinen MP, Andrews LS. 1995. Effects of propylene oxide on 

nasal epithelial cell proliferation in F344 rats. Fundam. Appl. Toxicol. 27(1):25-32. 

HSDB. 1994. Hazardous Substance Data Bank. National Library of Medicine, Bethesda, 

Maryland. (TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 

11/31/94). 

Appendix D3 474 Propylene Oxide 
Determination of Noncancer Chronic Reference Exposure Levels March 2000 

Hardin B, Niemeier R, Sikov M, and Hackett P. 1983. Reproductive-toxicologic assessment of 

the epoxides ethylene oxide, propylene oxide, butylene oxide, and styrene oxide. Scand. J. Work 

Environ. Health 9:94-102. 

Kuper CF, Reuzel GJ, and Feron VJ. 1988. Chronic inhalation toxicity and carcinogenicity study 

of propylene oxide in Wistar rats. Food Chem. Toxicol. 26(2):159-167. 

NTP. 1985. National Toxicology Program. Technical Report on the Toxicology and 

Carcinogenesis Studies of Propylene Oxide in F344/N Rats and B6C3F1 Mice (Inhalation 

Studies). NTP TR 267. U.S. Department of Health and Human Services, Public Health Service, 

National Institutes of Health. Research Triangle Park, NC: NTP. 

Ohnishi A, Yamamoto T, Murai Y, Hayashida Y, and Hori Tanaka I. 1988. Propylene oxide 

causes central-peripheral distal axonopathy in rats. Arch. Environ. Health 43(5):353-356. 

IRIS. 1995. Integrated Risk Information System (IRIS). US Environmental Protection Agency, 

Washington, D.C. (CD-ROM Version). 

Appendix D3 475 Propylene Oxide 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

SELENIUM AND SELENIUM COMPOUNDS 

(other than Hydrogen Selenide) 

Molecular 

Formula 

Synonyms Molecular 

Weight (g/mol)

CAS Reg. 

No. 

Se elemental selenium 78.96 7782-49-2 

SeO2 selenium dioxide; selenium oxide; 

selenious anhydride 

110.96 7446-08-4 

H2SeO3 selenious acid 128.97 7783-00-8 

SeOCl2 seleninyl chloride; selenium 

oxychloride; selenium oxichloric 

165.86 7791-23-3 

Na2SeO3 disodium selenite 263.01 10102-18-8 

Na2SeO4 disodium selenate 188.94 13410-01-0 

SeS selenium sulfide; sulfur selenide 111.02 7446-34-6 

I. Chronic Toxicity Summary

Inhalation reference exposure level 20 μg/m3

Oral reference exposure level 0.005 mg/kg/day (USEPA RfD)

Critical effect(s) Clinical selenosis 

Hazard index target(s) Alimentary system; cardiovascular system; 

nervous system 

Appendix D3 476 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

II. Chemical Property Summary (HSDB, 1995; Weast, 1980; Canady and Hodes, 1994; 

ACGIH, 1992) 

Description Se0

 crystal: metallic gray 

H2SeO4, Na2SeO3: white crystals 

H2SeO3, Na2SeO4: colorless crystals 

SeO2: lustrous crystals; yellow vapor 

SeS: yellow to orange powder 

Molecular formula see above 

Molecular weight see above 

Vapor pressure 0.001 torr @ 20°C 

Melting point SeO2: 340°C 

SeS: decomposes at 118-119°C 

Solubility Se0

: insoluble in water, alcohol; slightly soluble 

in CS2; soluble in ether 

H2SeO4: sol. in water; decomposes in alcohol 

H2SeO3: sol. in hot water, alcohol 

Na2SeO3: sol. in water 

Na2SeO4: 84 g/100 ml water at 35°C 

SeO2: 38.4 g/100 ml water at 14°C 

SeS: insoluble in water 

Conversion factor Se0

: not applicable (particulate) 

SeO2: 4.5 μg/m3 per ppb at 20°C 

III. Major Uses and Sources 

Selenium occurs in four valence states: selenates (Se6+), selenites (Se4+), selenides (Se2-), and 

elemental selenium (Se0) (Goyer, 1991) which include compounds formed with oxygen, sulfur, 

metals, and/or halogens. Selenium compounds are used in the glass industry as decolorizing 

agents and in the rubber industry as vulcanizing agents. Selenium compounds are also found in 

toning baths used in photography and xerography, and in insecticides and photoelectric cells. 

Selenious acid is a component of gun cleaning chemicals (Quadrani et al., 2000). Selenium 

sulfide is used in shampoos as an anti-dandruff agent. The most widely used selenium 

compound in industry is selenium dioxide (SeO2) which catalyzes reactions of organic 

compounds and is produced by the oxidation of selenium with nitric acid followed by 

evaporation or by burning selenium in oxygen (HSDB, 1995). The largest anthropogenic 

sources of atmospheric selenium are from the combustion of fossil fuels and the 

production/refining of copper; particulates are the primary expected form of the compound 

(National Academy of Sciences (NAS), 1976; U.S. EPA, 1984). The annual statewide industrial 

emissions from facilities reporting under the Air Toxics Hot Spots Act in California based on the 

most recent inventory were estimated to be 12,417 pounds of selenium and 4846 pounds of 

selenium sulfide (CARB, 1999). 

Selenium is an essential trace element in humans and other species; selenium deficiency leads to 

cardiomyopathy in humans (Goyer, 1991). For dietary intake, the National Research Council has 

Appendix D3 477 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

set a U.S. Recommended Daily Allowance (RDA) of 0.87 μg/kg (55-70 μg/person/day) 

(Subcommittee on the Tenth Edition of the RDAs, 1989). The average daily oral intake of 

selenium is 125 μg/person (U.S. EPA, 1991). Organic selenium compounds (e.g., dimethyl 

selenide) are known to occur as metabolites and as microbial degradation products in the 

environment. These compounds appear to have relatively low toxicity. 

IV. Effects of Human Exposures 

Acute occupational exposure to SeO2 resulted in bronchospasm, irritation of the upper 

respiratory passages, violent coughing, and gagging with nausea and vomiting (Wilson, 1962). 

The relationship between inhalation exposure to selenium and the presence of selenium in the 

urine was investigated in a five year study of workers at a selenium rectifying plant (Glover, 

1967). Workers were exposed to fumes and dusts of elemental red selenium, which, the author 

reported, is converted 80% to SeO2 in the presence of air. Average air concentrations of 

selenium were reported to be 3.6 mg/m3

 in grinding processes, 0.04 mg/m3 in annealing 

processes, and a range of averages of 0.23-0.87 mg Se/m3 in various “special” processes, e.g., 

punching, scraping, sorting, refining, and testing. The same author previously reported 

symptoms among selenium exposed workers including garlic-like odor of the breath, skin rashes, 

indigestion, and poorly-defined “socio-psychological” effects including lassitude and irritability 

(Glover, 1954). 

Clinical signs of toxicity were observed among a population exposed to high levels of selenium 

in soils and food supplies in China (Yang et al., 1983). Approximately half of 248 people in this 

region showed symptoms including hair and nail loss, discoloration and decay of the teeth, and 

CNS disturbances including pain and anesthesia of the extremities. Animals in the region were 

also affected, with hoof damage and horn sloughing reported in water buffalo, cattle, and pigs. 

Populations in low-, medium-, and high-selenium areas of China were later studied to associate 

the symptoms with selenium intake. Estimated daily intake for adults in these areas were 70, 

195, and 1438 μg Se for males and 62, 198, and 1238 μg for females, respectively (Yang et al., 

1989). Selenium intake was highly correlated with whole blood, breast milk, and 24-hour urine 

selenium levels. The authors also suggested the possibility of liver dysfunction as indicated by a 

delay in prothrombin time among persons with intake of 750-850 μg Se/day. More clearly 

recognized and characteristic clinical signs, however, were only observed in a group exposed to 

greater than 1261 μg Se/day and not among those exposed to less than 853 μg Se/day. Assuming 

a 55 kg body weight, these respective daily dose rates were 0.023 and 0.015 mg/kg-day. 

A population of 142 subjects in seleniferous areas of western South Dakota and eastern 

Wyoming was examined for signs of selenosis over a two-year period with monitoring of 

selenium levels in diet, whole blood, serum, urine, and toenails (Longnecker et al., 1991). 

Subjects completed health questionnaires, underwent physical examinations, provided blood 

samples for clinical assessment, and provided blood, urine, toenails, and duplicate-plate food 

collections for selenium analysis. About half of the 142 free-living subjects had selenium 

intakes greater than 2.54 mμmol/day (200 μg/day) (range 0.86-9.20 mμmol/day, or 68-724 

μg/day). Average intake among the population was estimated at 239 μg Se/day. No clinical 

Appendix D3 478 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

signs and no changes in hematological function, clinical chemistry, or liver function were 

observed in the population, even in subjects whose intake was as high as 9.20 mμmol/day (724 

μg/day). 

V. Effects of Animal Exposures 

Toxic effects from acute inhalation exposure to selenium dust were examined in rats, guinea 

pigs, and rabbits (Hall et al., 1951). Twenty female rats were exposed once for 8 hours to 

33 ± 10 mg Se/m3

. Many animals showed signs of pulmonary effects at both one week and 4 

weeks after exposure; however, no control group was included in the experiment with which to 

compare incidence. Similarly, six female rabbits and 10 male guinea pigs were exposed to the 

same level of selenium dust for four 4-hour periods every 48 hours (8 days total duration). The 

animals showed signs of interstitial pneumonitis at one week (2 animals of each species) and 

lung congestion and alveolar infiltration of large macrophages. 

Guinea pigs exposed one time to concentrations “less than 0.021 mg H2Se/L” (22 mg Se/m3 as 

hydrogen selenide) for 2, 4, or 8 hours exhibited difficulty breathing and a red-tinged discharge 

from the nose (Dudley and Miller, 1941). Mortality studies were conducted with guinea pigs 

(16/group) using the same exposure duration and selenium concentrations ranging from 1 to 43 

mg Se/m3. Fifty percent mortality was observed at 30 days among animals exposed once for 2 

hours to 12 mg Se/m3

. Mortality after 30 days was 50% among animals exposed once to 1 mg 

Se/m3 for 8 hours. Histopathological evaluation of guinea pigs exposed once for 4 hours to 8 mg 

Se/m3

 showed fatty change to the liver, pneumonia, lymphoid hyperplasia, and increased 

reticuloendothelial tissue in the spleen. These effects did not begin to resolve until more than 17 

days after the exposure. 

Several studies have addressed the toxicity of selenium compounds to animals when 

administered in either food or drinking water. Mice (50/group) treated with 0, 1, 4, or 8 ppm 

Na2SeO3 in drinking water over 50 weeks showed decreased growth rates at 8 ppm (Jacobs and 

Forst, 1981). The same group reported gross liver pathology in male mice treated by oral gavage 

for 3 days with 0.5 ml of 64 ppm Na2SeO3. Hamsters (8/sex/group) treated with 0.1 

(unsupplemented), 1, 5, 10, or 20 ppm Na2SeO3 in the diet for 42 days showed histopathological 

changes to the liver (Beems and van Beek, 1985). Rats (6-8/group) treated in the diet with SeS2, 

Na2Se, Na2SeO3, or Na2SeO4 showed increased relative liver weights and/or decreased body 

weight gain at 10 ppm (for each compound) over a 5 week exposure (Dausch and Fullerton, 

1993). A 13-week drinking water study of Na2SeO3, and Na2SeO4 in rats and mice showed 

increased mortality, decreased body weights, and histopathological changes to the kidneys in rats 

and decreased body weight and decreased water consumption in mice (Abdo, 1994). Decreased 

body weights were observed in rats treated for 6 weeks in drinking water with 2 ppm Na2SeO3 or 

Na2SeO4 (Palmer and Olson, 1974). 

Decreased percentage of live spermatozoa, altered sperm morphology, and decreased body 

weight gain were observed in rats (6/group) treated for 5 weeks with 2 ppm Na2SeO3 in the diet 

(Kaur and Parshad, 1994). Rats (7-12/group) exposed to 0, 4, 8, or 16 ppm Na2SeO3 in drinking 

Appendix D3 479 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

water for 240 days showed alterations in testicular LDH and β-glucuronidase activity at 4 ppm 

(Nebbia et al., 1987). 

Developmental toxicity endpoints were examined in hamsters (5-10/group) exposed by oral 

gavage on gestational day 8 to Na2SeO3 and Na2SeO4 at concentrations ranging from 0 - 110 

μmol/kg body weight (Ferm et al., 1990). Effects observed at 100 μmol Na2SeO3/kg included 

decreased fetal crown-rump length and increased percentage of abnormal litters. At 90 μmol 

Na2SeO4/kg, an increased percentage of abnormal litters was observed. Mice (10 or 14/group) 

treated with 0, 3, or 6 ppm Na2SeO3 in drinking water from 30 days pre-gestation through 

gestation showed altered estrus cycle length, decreased fetal growth, and a decreased number of 

ossified vertebrae in offspring (Nobunaga et al., 1979). 

VI. Derivation of Chronic Reference Exposure Level (REL) (for selenium and selenium 

compounds other than hydrogen selenide) 

Study Yang et al., 1989 

Study population 400 people in China 

Exposure method Low, medium, & high environmental levels of Se 

Critical effects Clinical selenosis (liver, blood, skin, CNS) 

LOAEL 0.023 mg/kg-day* (1.261 mg/day / 55 kg) 

NOAEL 0.015 mg/kg-day* (0.853 mg/day / 55 kg) 

Exposure continuity Continuous 

Exposure duration Lifetime 

Average experimental exposure 70, 195, and 1438 μg/day for adult males; 

62, 198, and 1238 μg/day for adult females 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies factor 1 

Intraspecies factor 3

Cumulative uncertainty factor 3 

Oral reference exposure level 0.005 mg/kg/day (USEPA RfD) 

Inhalation extrapolation factor 3,500 μg/m3

 per mg/kg-day 

Inhalation reference exposure level 20 μg/m3

*Factors: NOAEL (0.853 mg/day) and LOAEL (1.261 mg/day) calculated from regression 

analysis (log Y = 0.767 log X - 2.248, where Y = blood selenium and X = selenium intake) based 

upon the correlation (r = 0.962) between dietary selenium intake and blood selenium level for 

data showing incidence of clinical selenosis in adults based on an average adult body weight of 

55 kg. 

The inhalation chronic REL is based on the oral chronic REL, which is the same as the USEPA’s 

oral reference dose (RfD) (U.S. EPA, 1996). In addition to being inhaled, airborne selenium can 

settle onto crops and soil and enter the body by ingestion. Thus an oral chronic reference 

exposure level for selenium is also required for Air Toxics Hot Spots health risk assessments. 

The chronic inhalation REL was derived by route-to-route extrapolation of the RfD. The 

Appendix D3 480 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

principal study used for the REL/RfD was that of Yang et al. (1989). Yang et al. (1989), in a 

follow-up to an earlier study (Yang et al., 1983), studied a population of approximately 400 

individuals living in an area of China with unusually high environmental concentrations of 

selenium (Se). The subjects were evaluated for clinical and biochemical signs of Se intoxication. 

Three geographical areas with low, medium, and high selenium levels in the soil and food supply 

were chosen for comparison in the studies. The earlier study was conducted in response to 

endemic selenium intoxication in two separate areas with sample sizes of only 6 and 3. 

Comparisons were then made to a selenium-adequate area (n=8) and low-selenium area (n=13). 

The Yang et al. (1989) studies provide a much larger sample size and include additional analysis 

of tissue selenium levels. This allows a more accurate estimation of the dose-response 

relationship observed for selenium toxicity. Selenium levels in soil and approximately 30 typical 

food types commonly eaten by the exposed population showed a positive correlation with blood 

and tissue Se levels. The daily average Se intakes, based on lifetime exposure, were 70, 195, and 

1438 μg for adult males and 62, 198, and 1238 μg for adult females in the low-, medium- and 

high-selenium areas, respectively. Significant correlations, demonstrated between Se 

concentrations of various tissues, were used to estimate the minimal daily Se intake values that 

elicited various alterations in biochemical parameters indicative of possible Se-induced liver 

dysfunction (i.e., prolongation of clotting time and serum glutathione titer) and clinical signs of 

selenosis (i.e., hair or nail loss, morphological changes of the nails, etc.). In this manner, a 

marginal safe level of daily Se intake was estimated. Analysis of the results indicated that 

persistent clinical signs of selenosis were observed only in 5/349 adults, a potentially sensitive 

subpopulation. The blood selenium concentration in this group ranged from 1.054 to 1.854 mg/L 

with a mean of 1.346 mg/L. Clinical signs observed included the characteristic "garlic odor" of 

excess selenium excretion in the breath and urine, thickened and brittle nails, hair and nail loss, 

lowered hemoglobin levels, mottled teeth, skin lesions, and CNS abnormalities (peripheral 

anesthesia, acroparesthesia, and pain in the extremities). Alterations in the measured 

biochemical parameters occurred at dietary intake levels of 750-850 μg/day. These alterations 

were described as a delay in prothrombin time, i.e., increase in blood coagulation time and 

reduction in blood glutathione concentration. However, these indicators were poorly 

characterized and are not typically used as an index for clinical selenosis resulting from chronic 

exposure to selenium (NAS, 1989). Based upon the blood selenium levels shown to reflect 

clinical signs of selenium intoxication, a whole blood selenium concentration of 1.35 mg/L 

corresponding to 1.261 mg of daily selenium intake is indicative of the lowest correlative 

selenium intake causing overt signs of selenosis. The next lowest whole blood selenium 

concentration of 1.0 mg/L, corresponding to 0.853 mg selenium/day, produces no clinical signs 

of selenosis. The NOAEL for this study is 0.85 mg Se/day and the LOAEL is 1.26 mg Se/day. 

An intraspecies uncertainty factor of 3 was applied to the NOAEL to account for sensitive 

individuals. A full factor of 10 was not deemed necessary since similar NOAELs were identified 

in two moderately-sized human populations exposed to selenium levels in excess of the RDA 

throughout a lifetime without apparent clinical signs of selenosis. No modifying factor was 

applied by USEPA. OEHHA accepted the USEPA analysis. 

Route-to-route extrapolation assumes by default that a chemical is equally absorbed by the 

inhalation and the oral routes and that the first pass effect due to metabolism by the liver is not 

important for the chemical. The latter assumption is applicable to most metals. There are 

Appendix D3 481 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

limited data to evaluate the assumption of equal absorption across the gastrointestinal tract and 

the lungs. Limited data indicate that 60% (range = 44-100%) of ingested Se is absorbed by the 

gastrointestinal tract, while in one study 30% (single estimate) of inhaled selenium was deposited 

in the respiratory tract (Owen, 1990). Deposition is dependent on particle size. The available 

data are not adequate to depart from the default assumption. 

The USEPA stated its confidence in the RfD as: Study - Medium; Data Base - High; and RfD - 

High. Confidence in the chosen principal study is medium. Although this is a human 

epidemiological study in which a sizable population with sensitive subpopulations was studied, 

there are still several possible interactions that were not fully accounted for, e.g., fluoride intake 

and protein status. Also, except for clinical signs of selenosis there are no other reliable 

indicators, biochemical or clinical, of selenium toxicity. Confidence in the database is high 

because many animal studies and epidemiologic studies support the principal study. An 

additional human study with a freestanding NOAEL (Longnecker et al., 1991) provides support 

for the NOAEL identified in the principal study. Longnecker et al. (1991) found no effects at 

238 μg Se per day, which would equate to 0.004 mg/kg-day for a 55 kg person. Therefore, high 

confidence in the RfD is selected based upon support of the critical study and the high level of 

confidence in the database. 

There are insufficient data relating human inhalation exposure to selenium compounds to adverse 

health effects to use for the development of a chronic REL although toxicity has been reported 

from occupational exposure to gases of both H2Se and SeO2 (Buchan, 1947; Wilson, 1962). 

Experiments in animals have shown that H2Se is toxic following inhalation exposure, with 8-

hour exposures to concentrations as low as 1 mg H2Se/m3 causing “irritation sufficiently 

damaging to cause pneumonitis” and subsequently increasing 30-day mortality (Dudley, 1937; 

Dudley and Miller, 1941). Thus the selenium chronic REL is not meant to be applied to H2Se, 

which may be considerably more toxic than other selenium compounds. At this time there are 

inadequate data to develop a REL for H2Se. It is also not intended to be applied to organic 

metabolites of selenium. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the REL for selenium include its basis on a study with a large number of human 

subjects in a non-occupational setting that determined both a NOAEL and a LOAEL. The 

weaknesses include its basis on a route of exposure other than inhalation and its lack of 

applicability to hydrogen selenide, the most toxic selenium compound. 

VIII. Potential for Differential Impacts on Children's Health

The key study (Yang et al., 1989) included evaluation of children as young as one year old. 

Thus the chronic REL should be protective of infants and children. No adverse reproductive 

outcomes were reported, although only 400 people were studied. However, the inhalation REL 

is based on an oral REL of 0.005 mg/kg-day (0.06 μmol/kg-day). Ferm et al. (1990) did not find 

adverse effects on hamster development with Se doses below 34 μmol/kg. Thus the chronic 

REL should also be protective of infants and children. 

Appendix D3 482 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

IX. References 

Abdo KM. 1994. NTP (National Toxicology Program) technical report on toxicity studies of 

sodium selenate and sodium selenite (CAS No. 13410-01-0 and 10102-18-8) administered in 

drinking water to F344/N rats and B6C3F1 mice. PB94-215753. Springfield, VA: NTIS. 

ACGIH. 1992. American Conference of Governmental Industrial Hygienists, Inc. 

Documentation of the threshold limit values and biological exposure indices. Sixth ed. 

Cincinnati, OH: ACGIH. 

Beems RB, and van Beek L. 1985. Short-term (6-week) oral toxicity study of selenium in Syrian 

hamsters. Food Chem. Toxicol. 23:945-947. 

Buchan RF. 1947. Industrial selenosis: A review of the literature, report of five cases and a 

general bibliography. Occup. Med. 3:439-456. 

CARB. 1999. California Air Resources Board. Air toxics emissions data collected in the Air 

Toxics Hot Spots Program CEIDARS Database as of January 29, 1999. 

Canady RA, and Hodes CS. August, 1994. In: Draft Toxicological Profile for Selenium. Agency 

for Toxic Substances and Disease Registry, Public Health Service, US Department of Health and 

Human Services, Washington, DC: USDHHS. 

Dausch JG, and Fullerton FR. 1993. Increased levels of S-adenosylmethionine in the livers of 

rats fed various forms of selenium. Nutr. Cancer 20:31-39. 

Dudley HC. 1937. Toxicology of selenium. IV. Effects of exposure to hydrogen selenide. U.S. 

Public Health Reports 52:1217-1231. 

Dudley HC, and Miller JW. 1941. Toxicology of selenium. VI. Effects of subacute exposure to 

hydrogen selenide. J. Ind. Hyg. Toxicol. 23:470-477. 

Ferm VH, Hanlon DP, Willhite CC, Choy WN, and Book SA. 1990. Embryotoxicity and doseresponse relationships of selenium in hamsters. Reprod. Toxicol. 4:183-190. 

Glover JR. 1954. Some medical problems concerning selenium in industry. Trans. Assoc. 

Industr. Med. Offrs. 4:94-96. 

Glover JR. 1967. Selenium in human urine: a tentative maximum allowable concentration for 

industrial and rural populations. Ann. Occup. Hyg. 10:3-14. 

Goyer RA. 1991. Toxic effects of metals. In: Casarett and Doull’s Toxicology. The Basic 

Science of Poisons. Fourth ed. Amdur MO, Doull J, and Klaassen CD (eds.). New York: 

Pergamon Press, pp. 658-660. 

Hall RH, Laskin S, Frank P, Maynard EA, and Hodge HC. 1951. Preliminary observations on 

toxicity of elemental selenium. Arch. Ind. Hyg. Occup. Med. 4:458-464. 

Appendix D3 483 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Jacobs M, and Forst C. 1981. Toxicological effects of sodium selenite in Swiss mice. J. Toxicol. 

Environ. Health 8:587-598. 

Kaur R, and Parshad VR. 1994. Effects of dietary selenium on differentiation, morphology and 

functions of spermatozoa of the house rat, Rattus rattus L. Mutat. Res. 309:29-35. 

Longnecker MP, Taylor PR, Levander OA, Howe M, Veillon C, McAdam PA, Patterson KY, 

Holden JM, Stampfer MJ, Morris JS, and Willett WC. 1991. Selenium in diet, blood, and toenail 

in relation to human health in a seleniferous area. Am. J. Clin. Nutr. 53:1288-1294. 

NAS. 1976. National Academy of Sciences. Selenium. Committee on Medical and Biological 

Effects, Environmental Pollution Subcommittee on Selenium. Washington, DC: NAS. 

Nebbia C, Brando C, Burdino E, and Rasero R. 1987. Effects of the chronic administration of 

sodium selenite on rat testis. Res. Commun. Chem. Pathol. Pharmacol. 58:183-197. 

Nobunaga T, Satoh H, and Suzuki T. 1979. Effects of sodium selenite on methylmercury 

embryotoxicity and teratogenicity in mice. Toxicol. Appl. Pharmacol. 47:79-88. 

Owen BA. 1990. Literature-derived absorption coefficients for 39 chemicals via oral and 

inhalation routes of exposure. Regul. Toxicol. Pharmacol. 11(3):237-252. 

Palmer IS, and Olson OE. 1974. Relative toxicities of selenite and selenate in the drinking water 

of rats. J. Nutr. 104:306-314. 

Quadrani DA, Spiller HA and Steinhorn D. 2000. A fatal case of gun blue ingestion in a toddler. 

Vet. Hum. Toxicol. 42(2):96-98. 

Subcommittee on the Tenth Edition of the RDAs 1989. Food and Nutrition Board: 

Recommended Dietary Allowances. Tenth ed. Commission on Life Sciences, National Research 

Council. Washington, DC: National Academy Press. 

U.S. EPA. 1984. United States Environmental Protection Agency. Health effects assessment for 

selenium (and compounds). EPA/540/1-86/058. NTIS Doc. No. PB86-134699. Washington, DC: 

U.S. EPA. 

U.S. EPA. 1991. United States Environmental Protection Agency. Title 40, Code of Federal 

Regulations, Parts 141, 142, and 143. National Primary Drinking Water Regulations; Final Rule. 

Federal Register, 56:3538-3539. 

U.S. EPA. 1996. United States Environmental Protection Agency. Selenium. IRIS (Integrated 

Risk Information System). Available online at http://www.epa.gov/ngispgm3/iris 

Weast RC. (ed.) 1980. CRC Handbook of Chemistry and Physics. 61st ed. Boca Raton, FL: CRC 

Press. 

Appendix D3 484 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Wilson HM. 1962. Selenium oxide poisoning. N.C. Med. J. 23:73-75. 

Yang G, Wang S, Zhou R, and Sun S. 1983. Endemic selenium intoxication of humans in China. 

Am. J. Clin. Nutr. 37:872-881. 

Yang G, Zhou R, Yin S, and Gu L. 1989. Studies of safe maximal daily dietary selenium intake 

in a seleniferous area in China. I. Selenium intake and tissue selenium levels of the inhabitants. J. 

Trace Elem. Electrolytes Health Dis. 3:77-87. 

Appendix D3 485 Selenium 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

CHRONIC TOXICITY SUMMARY 

SILICA (CRYSTALLINE, RESPIRABLE) 

(silicon dioxide, quartz, tridymite, cristobalite) 

CAS Registry Number: 7631-86-9 

I. Chronic Toxicity Summary 

Inhalation Reference Exposure Level 3 μg/m3 [respirable, as defined occupationally by 

ACGIH (2004)/ISO (1995) 

Critical effect(s) Silicosis in miners and other workers 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 2001) 

Description Transparent crystals 

Molecular formula SiO2

Molecular weight 60.09 g/mol 

Density 2.65 g/cm3

 @ 0 °C (quartz) 

Melting point 1610 °C 

Boiling point 2230 °C (2503.20 °K) 

Vapor pressure 10 torr @ 1732 °C 

Solubility Practically insoluble in water or acids, except 

hydrofluoric acid; very slightly sol. in alkali. 

Conversion factor Not applicable 

In crystalline silica, the silicon and oxygen atoms are arranged in a definite regular pattern 

throughout the crystal. The characteristic crystal faces of a crystalline form of silica are the 

outward expression of this regular arrangement of the atoms (HSDB, 2001). This REL is meant 

to be applied only to particles of crystalline silica (quartz, cristobalite, tridymite), of respirable 

size, as defined by the occupational hygiene methods described by ACGIH (2004)/ISO (1995) 

which has a 50% cut-point at 4 μm particle aerodynamic diameter. This occupational definition 

of respirable differs from the environmental definition of respirable, which is PM10. (The 

occupational particle category “thoracic” has a 50% cut-point at 10 μm particle diameter 

(ACGIH, 2004) and the category “inhalable” has a 50% cut-point at 100 μm particle diameter 

(ACGIH, 2004).) 

III. Major Uses and Sources 

At least 11 chemically identical forms (polymorphs) have been described for crystalline silica. 

Alpha-quartz is the most abundant polymorph and constitutes 12% of the earth's crust (Elzea, 

1997). Silica is also found in the amorphous (non-crystalline) state. The amorphous silica in 

diatomaceous earth (composed mainly of the cell walls of diatoms) can be converted to the 

crystalline form cristobalite by heating to 1000-1100 oC (calcining). Silica is often associated 

Appendix D3 486 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

with silicates, which, in addition to silicon and oxygen, contain other metals such as iron, 

magnesium, aluminum, calcium, potassium, and sodium. 

The major uses of silica are in the manufacture of glass, abrasives, ceramics, and enamels, in 

scouring and grinding compounds, and in molds for castings. Silica is also used in decolorizing 

and purifying oils and petroleum products; as a clarifying agent; in filtering liquids; and in the 

manufacture of heat insulators, firebrick, and fire- and acid-proof packing materials. As 

diatomite (naturally occurring diatomaceous earth), silica is used as a filtration agent, as an 

abrasive, and as an industrial filler. Sources of ambient respirable crystalline silica in California 

include mines, quarries, diatomaceous earth calcining plants, sand blasting, and entrained fines 

(e.g., PM10) from surface soil. The annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 2,514,981 pounds of crystalline silica (CARB, 2001. The fraction, which is 

respirable as defined either occupationally or environmentally, is not known. 

Measurement of crystalline silica has evolved. Instrumentation has varied by country. In South 

Africa since the 1930s, dust was collected with a konimeter (Le Roux, 1970; Cherrie and Aitken, 

1999). A small volume of air (e.g., 5 cm3 captured in less than a second) was collected 

(impacted) onto a small area of a glass slide coated with adhesive. Total dust particles were 

counted and expressed as dust particles per cubic centimeter. Later, slides were heated to 500-

550 oC (ignition) to remove carbonaceous materials and immersed in hot 50% hydrochloric acid 

followed by a second ignition to remove acid-soluble materials. The remainder was mostly silica 

particles, which could be counted. The konimeter was superseded by the thermal precipitator, 

which also deposited particles onto glass but could sample larger air volumes at high flow rates 

(> 1 L/minute) for several hours. With time, particle counting was replaced by estimation of a 

particle’s surface area, initially by examining slides but more recently by an automated method 

(Kitto, 1960; 1970). 

In the United States the impinger method was used from 1922 until 1984 (Lippmann, 2001). Air 

was drawn into a trap containing fluid, particles in an aliquot of the fluid were counted under 

magnification, and concentrations were expressed as million particles per cubic foot of air 

sampled. Later, gravimetric analysis was introduced. Gravimetric analysis is dominated by the 

larger particles in any given size range. 

When it was realized that only a fraction of the dust was responsible for silicosis, respirable dust 

was collected onto filters using size-specific dust collectors, such as horizontal plate elutriators in 

South Africa and cyclones in the United States. The sizes of particles collected on the filter were 

a function of the apparatus used and the rate of airflow through the apparatus. Quartz dust was 

quantified by examining filters in an electron microscope with a specific X-ray diffraction beam 

absorbed by crystalline silica. The National Institute of Occupational Sciences and Health 

(NIOSH, 2003) has approved Method 7500, which uses one of three approved cyclones and a 5 

μm PVC membrane filter to sample, and X-ray diffraction to measure crystalline silica. The 

ARB has used Method 7500 in research projects. 

In order to harmonize respirable particulate sampling methodology in workers, an international 

agreement has been reached to use dust samplers that have a 50% cut point for particles of 4 μm 

aerodynamic diameter (ISO, 1995; ACGIH, 2004). 

Appendix D3 487 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Various attempts have been made to estimate the changes in silica levels in workplaces over time 

(e.g., Seixas et al., 1997 for diatomaceous earth facilities in California; Verma et al., 1989 for 

Ontario hard rock miners). However, although some conversion factors have been proposed, 

correlation between dust particle number in earlier studies, when dust concentrations were 

higher, and dust particle weight in the later studies, when the dust concentrations have been 

lowered, is imprecise so it is difficult to compare the earlier silica measurements with the more 

recent ones. 

IV. Effects of Human Exposures 

Inhalation of crystalline silica initially causes respiratory irritation and an inflammatory reaction 

in the lungs (e.g., Vallyathan et al., 1995). Acute exposures to high concentrations cause cough, 

shortness of breath, and pulmonary alveolar lipoproteinosis (acute silicosis). After chronic but 

lower workplace exposures to silica for six to sixteen years, the small airways become obstructed 

as measured by pulmonary function tests (e.g., decreased FEV1) in granite quarry workers (no 

measurement of silica levels reported; Chia et al., 1992). In a report on the hazards of exposure 

to crystalline silica, the American Thoracic Society (1997) stated: “Studies from many different 

work environments suggest that exposure to working environments contaminated by silica at 

dust levels that appear not to cause roentgenographically visible simple silicosis can cause 

chronic airflow limitation and/or mucus hypersecretion and/or pathologic emphysema.” Hnizdo 

and Vallyathan (2003) also concluded that “chronic levels of silica dust that do not cause 

disabling silicosis may cause the development of chronic bronchitis, emphysema, and/or small 

airways disease that can lead to airflow obstruction, even in the absence of radiological 

silicosis.” Fibrotic lesions associated with crystalline silica have also been found at autopsy in 

the lungs of granite workers who lacked radiological evidence of silicosis (Craighead and 

Vallyathan, 1980). 

Silicosis results from chronic exposure; it is characterized by the presence of histologically 

unique silicotic nodules and by fibrotic scarring of the lung. The histological progression of 

silicosis has been described as: (1) granuloma composed of histiocytic cells, collagen, and 

lymphocytes; (2) cellular fibrotic nodule with irregular collagen at the center and circular 

collagen at the periphery; (3) more mature nodule with acellular and avascular center; and 

(4) late mature nodule composed of dust and collagen including a calcified center (Green and 

Vallyathan, 1996). Lung diseases other than cancer associated with silica exposure include 

silicosis, tuberculosis/silicotuberculosis, chronic bronchitis, small airways disease, and 

emphysema (Oxman et al., 1993; Park et al., 2002; Hnizdo and Vallyathan, 2003; Balmes et al., 

2003). Silica exposure has been implicated in autoimmune diseases (rheumatoid arthritis, 

scleroderma, systemic lupus erythematosus) in gold miners and granite workers (Steenland and 

Goldsmith, 1995; Parks et al., 1999) and in the causation of kidney disease in some occupations 

(Goldsmith and Goldsmith, 1993; Stratta et al., 2001), possibly by an immune mechanism. 

At the cellular level, silica particles are engulfed in the lung by alveolar macrophages (AM). 

According to the generally assumed pathological model, the AM subsequently release various 

growth factors and reactive oxygen species (ROS; superoxide anion, hydrogen peroxide, 

hydroxyl radical) (Lapp and Castranova, 1993; Mossman and Churg, 1998; Ding et al., 2002). 

ROS and some growth factors (e.g., activator protein-1, platelet activating factor) are 

inflammatory and attract neutrophils to the site of inflammation, while other factors (fibronectin, 

Appendix D3 488 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

alveolar macrophage-derived growth factor) stimulate fibroblasts to proliferate and to make 

collagen. Since silica particles cannot be digested by the macrophage, the inflammatory process 

becomes chronic (frustrated phagocytosis). An increased silica burden leads to more foci of 

inflammation, nodule formation, and fibrosis. The internal process can continue after external 

exposure ends. Silica particles also enter into alveolar Type I epithelial cells (Churg, 1996), 

which can lead to cell death of Type I cells and to hypertrophy and proliferation of Type II 

epithelial cells to replace the Type I cells. The epithelial repair process is associated with a 

subsequent increase in collagen formation. 

The initial diagnosis of silicosis is often based on chest radiographs. Recent papers have used 

the 1980 classification by the International Labor Organization (ILO, 1980) to identify and 

classify silicosis into categories and subcategories of seriousness by comparison of patient 

radiographs with ILO-supplied reference radiographs taken at various stages of silicosis (Table 

1): 

Table 1. International Labor Organization categorization of silicosis (ILO, 1980). 

ILO Category Qualitative Description 

0/0 No small (up to 1 cm) silicotic opacities (nodules) are present 

0/1 Probably no nodules, but some areas of radiograph are suspect 

[possible silicosis] 

1/0 Small silicotic nodules are most likely present, but not certainly 

[probable silicosis] 

1/1 Small silicotic nodules are definitely present 

1/2 Small silicotic nodules are definitely present; other areas of the 

radiograph may indicate more advanced lesions including large 

opacities (> 1 cm), pleural thickening, etc. 

2/1, 2/2, 2/3, 3/2, 3/3 More advanced stages of silicosis/increasing certainty of the presence 

of lung abnormalities 

Some reports (e.g., Kreiss and Zhen, 1996; Hughes et al., 1998) use 1/0 (probable) as the basis of 

classification of silicosis, since many cases of silicosis are not detected by chest radiographs, yet 

silicotic nodules and other lesions are found at autopsy (Craighead and Vallayathan, 1980; 

Hnizdo et al., 1993). Other reports (e.g., Hnizdo and Sluis-Cremer, 1993) use the definite 1/1 as 

the lowest category indicating silicosis. Some disease is missed by radiography and is 

determined only by autopsy (Hnizdo et al., 1993). The ILO criteria are intended as an 

epidemiologic classification and comparison tool, not as a diagnostic classification on an 

individual basis. In occupational medicine practice, a group of tests is used to clinically 

diagnose silica-related lung disease including physical examination, X-rays, and high resolution 

computed tomography (CT) scans of the lung (e.g., Begin et al., 1991; Olivetti et al., 1993). 

A. Environmental silicosis 

Several studies have reported "environmental silicosis", cases where the silicosis occurs in the 

absence of an industry usually associated with the disease (reviewed by USEPA, 1996). In one 

of the stronger examples, Saiyed et al. (1991) investigated non-occupational pneumoconiosis in 

Ladakh, India, high in the western Himalayas where there are no mines or industries. Among 

449 randomly selected inhabitants of three villages, there were many cases of pneumoconiosis 

Appendix D3 489 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

associated with progressive massive fibrosis (nodules > 1 cm) and "egg shell" calcification of 

hilar glands. The prevalence of pneumoconiosis was 2.0% (3/150) in the village of Saboo, 

20.1% (31/149) in Shey, and 45.3% (68/150) in Chushot, and corresponded with the severity of 

dust storms and the presence or absence of chimneys in the kitchens (i.e., ventilated cooking). 

Without chimneys (Chushot), dust concentrations in kitchens averaged 7.5 mg/m3 during 

cooking periods. The free silica content of the dust storms was 60-70%. The authors suggested 

that exposure to free silica from dust storms and to soot from cooking with domestic fuels caused 

the pneumoconiosis. Perhaps the interaction of silica and soot led to the disease. Such 

exposures in this and other studies, such as Bar-Ziv and Goldberg (1974), might be considered to 

be non-industrial but occupational, since the subjects studied by Saiyed et al. (1991) were 

involved in the domestic work of cleaning and cooking (USEPA, 1996). In any case, the 

exposures were very high and thus similar to some occupational exposures. 

B. Occupational silicosis 

Several relatively recent reports have presented data that allow a quantitative relationship 

between occupational dust exposure and the development of silicosis in workers to be calculated. 

Hard rock miners in Ontario, Canada (Muir et al., 1989) 

Muir et al. (1989) examined the relationship between cumulative exposure to silica (free 

crystalline silica, specifically alpha-quartz) and the development of silicosis in 2109 male hard 

rock (uranium, gold, mixed metals) miners in Ontario, Canada. The miners began work between 

1940 and 1959 and were followed either until they ended their dust exposure or until December 

31, 1982 (whichever came first). Five X-ray readers examined chest radiographs; one or more 

readers identified 32 cases of silicosis, defined as ILO category 1/1 or greater with round 

opacities. All five readers agreed on only six cases, while 12 cases were identified by only one 

reader (Table 2). A Weibull model of the form 

R(x) =1 ─ exp[−(αx)β] (x ≥ 0, β > 0) 

gave the best fit to the data for cumulative risk R of silicosis as a function of cumulative 

exposure in units of (mg/m3

)-yr. In this model x is the cumulative exposure (lagged five years), 

α is the Weibull scale parameter, and β is the Weibull shape parameter (Table 2). Estimates of α

and β for each reader are given in Table II of Muir et al. (1989). 

Table 2. Silicosis Risk vs. Cumulative Respirable Silica in (mg/m3)-y (Table IV of Muir et al.) 

Reader Cases (n) 1% risk a 2% risk 5% risk 10% risk 

1 14 3.5 (2.4-5.1) 5.7 (3.9-8.4) 11.2 (6.8-18.2) 18.6 (9.9-35.0) 

2 24 2.7 (2.0-3.6) 4.1 (3.2-5.3) 7.1 (5.5-9.1) 10.9 (8.1-14.8) 

3 24 3.0 (2.3-3.9) 4.3 (3.4-5.3) 6.9 (5.6-8.5) 9.9 (7.8-12.7) 

4 14 3.7 (2.6-5.2) 5.6 (4.1-7.7) 9.8 (6.7-14.3) 15.1 (9.3-24.4) 

5 7 5.7 (4.0-8.0) 7.8 (5.5-11.0) 11.9 (7.8-18.3) 16.5 (9.7-28.2) 

Any reader 32 2.1 (1.6-2.9) 3.3 (2.6-4.2) 6.0 (4.8-7.5) 9.6 (7.3-12.5) 

At least 3 15 3.5 (2.5-4.9) 5.4 (4.0-7.3) 9.5 (6.6-13.6) 14.6 (9.3-23.2) 

All readers 6 6.1 (4.1-8.9) 8.5 (5.6-12.8) 13.2 (7.8-22.5) 18.7 (9.7-36.1) a

 In parentheses is the 95% confidence interval (CI) for each risk estimate. 

Appendix D3 490 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

The Ontario cohort gives the shallowest dose-response relationship for silicosis of the several 

cohorts examined (see Summary Table 15 below) due in part to the lack of follow-up of 

members who left the mines (either for another type of work or for retirement). Silicosis often 

develops after leaving employment (Hnizdo and Sluis-Cremer, 1993; Chen et al., 2001). In 

Hnizdo and Sluis-Cremer (1993), for more than half the cases of silicosis radiographic signs 

developed at an average of 7.4 years after mining exposure ended. In addition, some of the 

Ontario miners in the Muir et al. study may have changed to a less dusty job if their physician 

told them that their (annual) radiograph showed abnormalities. The lack of follow-up, leading to 

under-ascertainment of silicosis, is a serious limitation of this study. 

Gray iron foundry workers (Rosenman et al., 1996) 

Rosenman et al. (1996) evaluated 1,072 (96.8% males) current and retired workers in a Midwestern gray iron foundry, which produces engine blocks for the automotive industry. Medical 

records and silica exposure data were analyzed for those with at least 5 years of employment as 

of June 1991. Nearly half had worked at the foundry for 20 years. Sixty had radiographic 

evidence of pneumoconiosis (ILO categories 1/0 and greater). Twenty-eight workers had 

radiographs consistent with silicosis; of these 25 had simple silicosis and three had progressive 

massive fibrosis. The prevalence of radiographic changes consistent with silicosis increased 

with years at the foundry, work area, quantitative silica exposure, and cigarette smoking. In 

regard to quantitative silica exposure, the authors stated that 0.3-2.7% of workers at the OSHA 

standard (90-100 μg/m3

) were silicotic, as were 4.9-9.9% of workers above 100 μg/m3. After 

controlling for confounders, Rosenman et al. (1996) used a logistic regression analysis based on 

cumulative silica exposure to determine an odds ratio of 1.45 for developing a radiograph 

consistent with silicosis after 20 years of work at 100 μg/m3 and an odds ratio of 2.10 after 40 

years of work at 100 μg/m3 (Tables 3 and 4). This study probably underestimates risk due to 

lack of follow-up of the current workers. Although silica is not the only toxic chemical in a 

foundry, the unique nature of the silicotic nodule diminishes the likelihood of confounding by 

other exposures. 

Table 3. Silicosis risk based on Rosenman et al. data (Finkelstein, 2000) 

Cumulative silica exposure Prevalence of silicosis 

< 2 (mg/m3)-y 0.4% 

2-6 (mg/m3)-y 2.7% 

> 6 (mg/m3)-y 10% 

Appendix D3 491 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 4. Odds ratios for silicosis (from Table 8 of Rosenman et al.)a

Time-weighted 

average silica 

exposure (mg/m3

) 

20-year 

cumulative 

exposure 

[(mg/m3)-y] 

Odds ratio 

(95% C.I.) 

40-year 

cumulative 

exposure 

[(mg/m3)-y]

Odds ratio 

(95% C.I.) 

0.010 0.2 1.04 (1.02-1.15) 0.4 1.08 (1.05-1.11) 

0.025 0.5 1.10 (1.06-1.14) 1.0 1.20 (1.12-1.30) 

0.050 1.0 1.20 (1.12-1.30) 2.0 1.45 (1.25-1.68) 

0.075 1.5 1.32 (1.18-1.47) 3.0 1.74 (1.40-2.17) 

0.100 2.0 1.45 (1.25-1.68) 4.0 2.10 (1.15-2.82) 

0.150 3.0 1.74 (1.40-2.17) 6.0 3.04 (1.96-4.72) 

0.200 4.0 2.10 (1.56-2.82) 8.0 4.40 (2.45-7.93) 

0.300 6.0 3.04 (1.96-4.72) 12.0 9.24 (3.83-22.3) a Additional mean silica exposures, their calculated odds ratios, and 95% confidence intervals 

(C.I.) are given in the paper. 

Diatomaceous earth workers in California (Hughes et al., 1998; Park et al., 2002) 

Hughes et al. (1998) investigated 1,809 Caucasian male diatomaceous earth workers in Lompoc, 

California, who had at least one year of exposure to cristobalite between 1942 and 1987. The 

crystalline silica isomorph cristobalite is formed when the amorphous silica in diatomaceous 

earth is calcined at 1000-1100 oC. Quantitative estimates of dust exposure were made and 

published in the peer-reviewed literature by Seixas et al. (1997) based on 6395 air sampling 

records taken from 1948-1988. The average estimated respirable dust concentrations for 135 

jobs were 3.55 ± 1.25 mg/m3

 prior to 1949, 1.37 ± 0.48 mg/m3 from 1949-1953, 0.47 ± 0.16 

mg/m3 from 1954-1973, and 0.29 ± 0.10 mg/m3 from 1974-1988. The workers had periodic 

chest radiographs. Based on the median of radiographic readings by three independent readers, 

81 workers (4.5%) were judged to have opacities on chest radiographs (small opacities, ILO 

profusion ≥ 1/0, and/or large opacities). Age-adjusted relative risk of opacities increased 

significantly with cumulative exposure to crystalline silica. The concentration of respirable 

crystalline silica was an important determinant of risk after accounting for cumulative exposure. 

The workers were split into two categories: those exposed to < 0.50 mg/m3

 (or hired after 1950) 

and those exposed to > 0.50 mg/m3

 (or hired before 1950). The risk of opacities for a cumulative 

exposure to crystalline silica of 2.0 mg/m3

-yr is shown in Table 5. 

Table 5. Silica exposure and silicosis based on data of Hughes et al. (1998) 

Average crystalline silica exposure Cumulative risk of silicotic opacities 

< 0.50 mg/m3 (or hired after 1950) 1.1% 

> 0.50 mg/m3 (or hired before 1950) 3.7% 

The findings of Hughes et al. (1998) indicate an exposure-response relationship between 

cumulative exposure to crystalline silica as cristobalite and radiographic opacities. The 

relationship was substantially steeper among those exposed at the highest average concentrations 

of crystalline silica. The authors believe that the data do not support the regulatory assumption 

that cristobalite is more fibrogenic than quartz (i.e., prior to 2000 the occupational limit for 

cristobalite was half that for quartz), since at average silica levels comparable to other 

Appendix D3 492 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

epidemiologic studies quartz gave a higher incidence of silicosis than did cristobalite in this 

study. However, since radiography can under-diagnose silicosis, complete accounting for 

silicosis will require evaluation at autopsy. The ACGIH recently lowered the TLV for alphaquartz from 100 to 50 μg/m3

, so that it has the same TLV as cristobalite (ACGIH, 2000). 

Park et al. (2002) carried out a quantitative risk assessment, by Poisson regression methods, of 

the onset of silicosis among the diatomaceous earth workers in Lompoc. A linear relative risk 

model gave the best fit to the data. They estimated an excess lifetime risk for radiographic 

silicosis of 68-75 cases per thousand workers exposed to 50 μg/m3

 silica (cristobalite) for a 45 

year work-life, then living to age 85. At 1 μg/m3

 silica the excess lifetime risk was estimated to 

be 1.6 cases of lung disease other than cancer per thousand workers exposed (Table 6). 

Table 6. Excess lifetime risk of silicosis predicted by Park et al. (2002) 

Silica 

concentration 

(mg/m3

) 

45 year cumulative 

exposure in mg/m3

-y 

Radiographic 

silicosis - all workers 

Radiographic 

silicosis in workers 

with < 10 mg/m3-y 

0.001 0.045 6.2/1000* 1.6/1000 

0.005 0.225 17/1000 7.8/1000 

0.01 0.45 26/1000 16/1000 

0.02 1.8 39/1000 31/1000 

0.05 2.25 68/1000 75/1000 

0.1 4.5 100/1000 140/1000 

0.2 9 150/1000 260/1000 

* Excess risk estimates assume that workers were exposed to a constant silica concentration for 

up to 45 years (ages 20-65). Annual risks are accumulated up to age 85. 

White South African gold miners (Hnizdo and Sluis-Cremer, 1993) 

Hnizdo and Sluis-Cremer (1993) investigated silicosis risk retrospectively in a cohort of 2,235 

white male South African gold miners. Exposure estimates were made for nine separate 

occupational categories based on a special study of dust levels in these mines done by Beadle in 

the 1960s (Beadle, 1971). To compensate for the fact that the average hours working in dust 

ranged among the 9 categories from 4 hours for “other officials” to 8 hours for “shaft sinkers and 

developers,” exposure was “normalized” to 8-hour shifts. The workers had a minimum of 10 

years and an average of 24 years service from 1940 until the early 1970s. Dust levels were fairly 

constant during this period (see, e.g., Table 2 in Gibbs and DuToit (2002)). The miners had an 

annual chest radiograph while mining; they were followed until 1991 for radiographic signs of 

the onset of silicosis. An ILO category 1/1 (definite silicosis) or greater was selected to 

designate silicosis. Two independent readers initially read the chest films, but only the reader 

whose interpretations correlated better with autopsy results was used for additional analysis; the 

use of one reader is a limitation of the study. There were 313 miners (14% of the cohort) who 

developed radiographic signs of silicosis at an average age of 55.9 years. The latency period was 

largely independent of the cumulative dust exposure (CDE). In 57% of the silicotics, the 

radiographic signs developed at an average of 7.4 years after mining exposure ceased. The risk 

of silicosis determined by chest radiographs increased exponentially with cumulative dust dose. 

At the highest level of 15-(mg/m3

)-years CDE (approximately 37 years of gold mining at an 

Appendix D3 493 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

average respirable dust concentration of 0.4 mg/m3), the cumulative risk for silicosis reached 

77% as estimated by the accelerated failure time model using the log-logistic distribution (SAS 

Proc LIFEREG): 

CR(t) = 1 – {1/[1 + exp (-μ/σ) x t(1/σ)

]} 

where CR(t) = cumulative risk at time t, and μ (2.439) is the intercept and σ (0.2199) is the scale 

parameter estimated by SAS’s LIFEREG procedure. The authors concluded that the risk of 

silicosis was strongly dose-dependent, but that the latency period was largely independent of 

dose. The life table analysis (SAS Proc LIFETEST) below (Table 7) shows the number of 

miners who developed silicosis (“cases”), the number of miners considered by the authors to be 

at risk, and the risk per unit of CDE (also as calculated by the authors). In the table in column 1 

(in parentheses) are OEHHA’s determination of the mg/m3

–yr respirable silica exposure, based 

on Hnizdo and Sluis-Cremer’s estimate of 30% silica in the dust, and in column 4 is the total 

number of miners actually at each midpoint level of CDE or silica. The values in column 4 of 

Table 7 are the number of workers in the group with the temporally integrated dust exposure in 

column 1. 

Table 7. Life table results - Risk of silicosis per unit Cumulative Dust Exposure (CDE) 

(from Table IV of Hnizdo and Sluis-Cremer, 1993) 

Midpoint 

in (mg/m3)-y 

of CDE 

(silica) 

Cases of 

silicosis 

Number of 

workers at risk 

based on life 

table 

Number of 

workers 

remaining 

at this CDE 

midpoint 

“Risk/ 

unit 

CDE” 

Mean 

years in 

dust 

Mean 

dust conc. 

 (mg/m3) 

 1 (0.3) 0 2218 204 

 3 (0.9) 9 2014 474 0.002 20.5 0.17 

 5 (1.5) 48 1540 556 0.016 23.5 0.24 

 7 (2.1) 85 984 469 0.045 27.2 0.30 

 9 (2.7) 93 515 318 0.099 28.0 0.33 

11 (3.3) 53 197 142 0.156 29.4 0.38 

13 (3.9) 20 55 44 0.222 31.5 0.41 

15 (4.5) 5 11 11 0.227 37.0 0.42 

a

 CDE = Σ number of dusty shifts x mean mass respirable dust conc. x average number of hours 

spent underground / (270 shifts/year x 8 h/shift) 

A plot of risk of silicosis per unit of Cumulative Dust Exposure (CDE) versus the mid-point unit 

CDE, as given in Figure 1 of the Hnizdo and Sluis-Cremer report, and a plot of % silicosis 

among the workers actually exposed to a given level of silica (Figure 2), as determined by 

OEHHA staff, respectively, are given below. 

Appendix D3 494 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Figure 1. Risk of silicosis per unit CDE vs. CDE mid-point 

0

0.05

0.1

0.15

0.2

0.25

0 5 10 15 20

midpoint CDE (mg/cubic m-yr)

risk per unit of CDE

Figure 2. Percent silicosis among workers at each silica level 

0

10

20

30

40

50

012345

(mg/cubic m)-yr silica

% silicosis

Black South African gold miners (Churchyard et al., 2004; Murray et al., 1996)

Black migrant contract workers constitute a large majority (85 - 90%) of South African gold 

miners. In a cross sectional study, Churchyard et al. (2004) interviewed and took chest 

radiographs of 520 black gold miners (mean age = 46.7 years, range = 37.1 – 59.9) who were 

still mining (average service = 21.8 years, range 6.3-34.5). Two readers examined the 

radiographs. As in the Hnizdo and Sluis-Cremer study, silicosis was defined as an ILO (1980) 

profusion of ≥ 1/1. The mean respirable dust concentration was 0.37 mg/m3

 (0 - 0.70); the mean 

quartz concentration was 0.053 mg/m3

 (0 - 0.095). The prevalence of silicosis was determined to 

be 18.3% by one reader and 19.9% by the other (mean 19.1%) (Table 8). This included several 

workers with more serious silicosis as indicated by ILO profusions ≥ 2/1 (see Table 1). 

Appendix D3 495 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Significant trends were found between the prevalence of silicosis and: (1) length of service (OR 

= 1.69 per 5 years), (2) mean intensity of exposure (OR = 1.18 per 0.01 mg/m3

), and (3) 

cumulative exposure to quartz (OR = 3.2). The study confirms the large burden of silicosis 

among older black workers in this industry (see next paragraph). The burden is likely to worsen 

with continuous employment in dusty jobs. For this cohort the prevalence of silicosis will 

increase even if the miners stop mining immediately. If, as assumed by the authors, the dust 

levels during the working life of these black miners were constant, silicosis developed while they 

were exposed to a quartz level below the workplace limit of 0.100 mg/m3. 

Table 8. Silicosis in black gold miners (Churchyard et al., 2003; 2004) 

Cumulative quartz 

exposure in 

mg/m3

-yr 

Mid-point of 

cumulative quartz 

exposure 

Number in 

quintile* 

Cases of 

silicosis 

Percent 

silicosis 

0 – 0.80 0.4 103 11 10.7 

0.80 – 0.99 0.9 97 8 8.2 

0.99 – 1.24 1.12 103 18 17.5 

1.24 – 1.48 1.36 104 23 22.1 

1.48 –3.08 2.28 103 33 32.0 

 (Total) (510)** (93) (18.2) 

* Personal communication from Dr. J. teWaterNaude, December 2, 2004. 

** Ten of the 520 films were unreadable. 

Murray et al. (1996) analyzed data from 16,454 black South African gold miners dying from 

unnatural causes between 1975 and 1991 in order to study change in prevalence in silicosis and 

pulmonary tuberculosis (TB). TB prevalence increased from 0.9% in 1975 to 3.9% in 1991, 

while that for silicosis increased from 9.3% to 12.8%. The prevalence of both increased with age 

and duration of service. Silicosis was the most significant predictor of TB (OR = 1.78, CI = 1.27 

- 2.30, p = 0.0001). A highly significant trend for TB, for year of autopsy, remained after 

adjustment for other variables, such as age and duration of service (OR = 1.04, CI = 1.01 – 1.06, 

p = 0.0046). (Another 21,202 black gold miners died of natural causes during the study period.) 

Hong Kong granite workers (Ng and Chan, 1994) 

Ng and Chan (1994) investigated silicosis among 338 male workers, who had worked at least 

one year between 1967 and 1985 in two granite quarries in Hong Kong. Three readers examined 

the chest radiographs. Silicosis was defined as an ILO classification of at least 1/1 (for small 

rounded opacities) or greater, assigned by at least two of the three readers. Exposure was 

estimated for each worker based on job category and particle counts. Thirty-six workers (10.6%) 

were designated silicotic. Both a logistic and a linear model fit the data well. The study suffered 

because only about half of the previously employed granite workers were studied, which 

probably led to an underestimate of silicosis risk in at least the highest exposure category and 

maybe in others. The data are summarized in Table 9. 

Appendix D3 496 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 9. Silica exposure and silicosis in Ng and Chan (Finkelstein, 2000) 

Mean cumulative exposure (mg/m3)-y Prevalence of silicosis a

< 1 0% 

3.1 13% 

7.1 25% 

22 22% a

 rounded opacities determined by at least 2 of 3 readers (Table 3 of Ng and Chan) 

Gold miners in South Dakota (Steenland and Brown, 1995) 

Steenland and Brown (1995) studied a very large cohort (3330) of white male gold miners in 

South Dakota, who had worked at least 1 year underground between 1940 and 1965 (average = 9 

years underground). The mine dust contained on average 13% silica (range = 1-48%). A jobexposure matrix was created for full-time underground workers grouped into five categories. 

The authors estimated that most miners were exposed to a median silica level of 0.05 mg/m3, but 

that those hired before 1930 were exposed to a median level of 0.15 mg/m3. A total of 170 cases 

of silicosis (5.1% of the cohort) was determined from death certificates only (n = 128 cases), 

from two cross-sectional radiographic surveys in 1960 and 1976 (n = 29 cases; ILO category 1/1 

or greater), or from both (n = 13 cases). Unfortunately, only 25% of living cohort members were 

surveyed radiographically. The life-time risk of silicosis was less than 1% with a cumulative 

exposure under 0.5 mg/m3-years and increased to 68% to 84% for the highest cumulative 

exposure category (more than 4 (mg/m3)-years) (Table 10). 

Table 10. Risk of silicosis for cohort by cumulative exposure (Table 3, Steenland and Brown) 

Silica 

exposure in 

(mg/m3)-yrs: 

range 

(midpoint) 

Miners 

with 

silicosis 

Number 

entering 

exposure 

category 

(from life 

table) 

Number 

remaining 

at this 

exposure 

level 

Cumulative a

Risk 

Mean years 

of exposure 

Mean 

year first 

exposed 

0-0.2 (0.10) 5 3330 1530 0.002 2.9 1953 

0.2-0.5 (0.35) 5 1800 740 0.005 9.7 1948 

0.5-1.0 (0.75) 15 1060 376 0.017-0.022b 15.4 1942 

1.0-2.0 (1.50) 33 684 353 0.060-0.084b 13.2 1931 

2.0-3.0 (2.50) 44 331 206 0.167-0.245 b 18.8 1926 

3.0-4.0 (3.50) 42 125 73 0.403-0.534b 25.5 1921 

>4.0 26 52 52 0.678-0.844b 30.6 1914 

a

 Cumulative risk = 1−exp[−sum of (hazards * interval width)], where the hazards for each 

category of cumulative exposure are: 

no. cases/(width*(no. entering category − 0.5*no. cases − 0.5*no. withdrawals)) 

b

 Cumulative risk adjusted for age and calendar time (Steenland and Brown, 1995) 

The best predictor of disease was cumulative exposure ((mg/m3) – years), followed by duration 

of exposure (years), and then by average exposure (mg/m3

). Figure 1 of Steenland and Brown 

indicates that a plot of their data for silicosis risk versus cumulative silica exposure was similar 

Appendix D3 497 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

to a plot of the data of Hnizdo and Sluis-Cremer (1993). After adjustment for competing risks of 

death, Steenland and Brown estimate that a 45-year exposure to 90 - 100 μg/m3 silica would lead 

to a lifetime risk of silicosis for gold miners of 35% to 47%. A limitation of this study is the 

reliance on death certificates rather than on ILO interpretation of radiographs. In addition, no 

mention was made of validating the data on the death certificates. It was also not clear what, if 

any, autopsy data were available. A plot of silicosis incidence among the workers (as 

determined by OEHHA staff) actually exposed to the estimated level of silica is given in Figure 

3 below. An accompanying editorial (Wagner, 1995) commended the article for estimating both 

the risk of silicosis while working and the lifetime risk of silicosis resulting from exposure 

during work. 

Figure 3. % Silicosis vs. silica exposure in Steenland and Brown (see Table 10) 

0

10

20

30

40

50

60

70

01234

(mg/cubic m)-yr silica

% silicosis

Miners in Leadville, Colorado (Kreiss and Zhen, 1996) 

Kreiss and Zhen (1996) investigated the exposure-response relationships for silicosis among 134 

male miners over 40 years old in Leadville, Colorado. The men had been studied three years 

earlier in a random sample of respiratory disease in their community (Kreiss et al., 1989). Of 

100 dust-exposed miners, 32 had radiological profusions of small opacities of ILO category 1/0 

or greater at a mean of 36.1 years since their first silica exposure. Of miners with cumulative 

silica, exposures of 2 (mg/m3

)-years or less, 20% had silicosis while 63% of miners 

accumulating greater than 2 (mg/m3)-years had silicosis. Average silica exposure was also 

strongly associated with silicosis prevalence rates (Table 11). 

Table 11. Miners studied by Kreiss and Zhen (1996) 

Average silica exposure % silicotics 

0.025-0.05 mg/m3 13% (5/38) 

> 0.05-0.1 mg/m3 34% (15/44) 

> 0.1 mg/m3 75% (9/12) 

Cumulative silica exposure % silicotics 

≤ 2 (mg/m3)-y 20% (14/70) 

2 − 4 (mg/m3)-y 63% (15/24) 

Appendix D3 498 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Based on logistic regression models of the form R(x) = [1 + exp (−α−B'x]−1, Kreiss and Zhen 

concluded that the risk of silicosis was best predicted by elapsed time since last silica exposure 

together with either (1) cumulative silica exposure or (2) a combination of average silica 

exposure and duration of exposure. Exposure-response relationships were substantially higher 

using measured silica exposures (compared to using estimated silica exposures based on 

measured total dust exposures and assuming a constant silica proportion of dust). The risk of 

silicosis in this study is higher than in workforce studies having no follow-up of those leaving the 

mining industry (e.g., Muir et al., 1989) and in studies without job title-specific silica 

measurements (e.g., Hnizdo and Sluis-Cremer, 1993). However, the risk is comparable to 

several recent studies of exposure-response relationships for mining dust (e.g., Ng and Chan, 

1994; Steenland and Brown, 1995) (see Summary Table 15 below). A limitation relative to other 

studies is the small number of subjects (100) in the group. 

Chinese tin miners (Chen et al., 2001) 

Chen et al. (2001) found a clear exposure-response relationship between silica dust exposure and 

silicosis in a cohort of 3010 (2795 male and 215 female) miners employed for at least 1 year 

during the period 1960-1965 in any of four Chinese tin mines. No other diseases due to silica or 

tin were mentioned. Each cohort member was followed through 1994. Historical Chinese total 

dust (CTD) data were used to create a job exposure matrix for each facility, job title, and 

calendar year. The CTD data were converted to estimates of respirable crystalline silica for 

comparison with findings from other epidemiological studies of silicosis (including some of 

those above). Each miner's work history was abstracted from employment records. The 

diagnosis of silicosis was based on 1986 Chinese Roentgen diagnostic criteria for 

pneumoconiosis. The criteria classified silicosis as stages I-III, similar to an ILO classification 

of 1/1 or greater. Of the 3010 miners, 1015 (33.7%) were identified as silicotic (mean age = 48.3 

years, with a mean of 21.3 years after first exposure) (Table 12). Among the silicotics, 684 

(67.4%) developed silicosis after their tin mine exposure had ended (mean = 3.7 years after). 

The risk of silicosis was strongly related to cumulative exposure to silica. The Weibull 

distribution gave a very good fit to the data. The risk of silicosis was less than 0.1% when CTD 

was less than 10 (mg/m3

)-yr (= 0.36 (mg/m3)-yr of respirable crystalline silica). The risk of 

silicosis increased to 68.7% when CTD exposure was equal to 150 (mg/m3

)-yr (= 5.4 (mg/m3)-yr 

of respirable crystalline silica). Latency period was not correlated to the risk of silicosis or to 

cumulative dose. From their data, the authors predicted a 55% risk of silicosis for 45 years 

exposure to 0.1 mg/m3

 respirable crystalline silica, the workplace exposure limit (4.5-(mg/m3)-

years silica). Figure 4 plots the fraction of the workers in Chen et al. with silicosis (column 2 in 

Table 12 divided by column 4) exposed to a given level of silica (mid-point – in parentheses in 

column 1 of Table 12), as calculated by OEHHA staff. 

Appendix D3 499 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 12. Cumulative silicosis risk based on cumulative total dust (CTD) 

 (Table 5, Chen et al., 2001) 

Range of CTD 

exposure 

in (mg/m3

)-y/ 

(silica mid-point) 

Cases of 

silicosis 

(n) 

Workers 

entering 

category 

Workers at 

this level of 

CTD/silica 

Cumulative 

risk based 

on Weibull 

model 

Mean net 

exposure 

(years) 

Mean 

latency 

(years) 

<10 (0.18) 2 3010 333 0.001 2.2 14.7 

10-19.99 (0.54) 24 2677 334 0.010 5.3 21.3 

20-39.99 (1.08) 126 2343 626 0.070 9.3 22.0 

40-59.99 (1.80) 127 1717 429 0.145 11.9 21.5 

60-79.99 (2.52) 196 1288 386 0.285 9.9 20.3 

80-99.99 (3.24) 141 902 264 0.405 10.8 19.0 

100-149.99 (4.50) 244 638 417 0.663 13.1 20.4 

≥ 150 (≥ 5.4) 155 221 221 0.917 15.7 25.4 

Figure 4. Percent silicosis vs. silica level from Chen et al. 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

024

Silica (mg/cubic m-yr)

Fraction with silicosis

6

Industrial sand workers (McDonald et al., 2001; Hughes et al., 2001; Rando et al., 2001) 

McDonald et al. (2001) studied a cohort of 2670 men employed before 1980 for 3 years or more 

and followed through 1994 in one of nine North American sand-producing plants and in a large 

associated office complex (since most of the office employees had previously worked in the 

mines). They found 37 deaths due to silicosis and silicotuberculosis. The mean exposure of the 

cohort was 42 μg/m3

 silica (Rando et al., 2001). Odds ratios for silicosis mortality, determined 

using conditional multiple logistic regression (SAS software), were significantly related to 

cumulative silica exposure (Hughes et al., 2001) (Table 13). The odds ratios are in general 

agreement with those in the gray foundry workers of Rosenman et al. (1996) (Table 4). 

Appendix D3 500 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 13. Median cumulative silica exposure and odds ratio (Table 3 in Hughes et al., 2001) 

No lagging Lagged 15 yr 

Median 

exposure in 

(mg/m3

)-y Silicotics (n) 

Odds ratio a 

for mortality 

Median 

exposure in 

(mg/m3

)-y Silicotics (n) 

Odds ratio a,b 

for mortality 

 0.832 7 1.00 0.142 7 1.00 

 2.744 7 1.27 1.229 7 2.54 

 6.916 8 2.62 2.583 7 4.55 

12.084 7 2.13 7.990 8 5.16 a

 Matched odds ratio relative to lowest cumulative exposure category. Although labeled a cohort 

study, the data analysis compared cases of silicosis with non-silicotic controls. 

b

 Significant increasing trend across exposure categories (see Hughes et al. for more details) 

Ceramic workers (Cavariani et al., 1995; Legrand-Cattan et al. (1998)

Cavariani et al. (1995) investigated the incidence of silicosis among 2,480 men in the ceramics 

industry in central Italy. The workers were surveyed during the period 1974-1987 and followed 

through 1991 with annual chest radiographs. The cumulative risk of silicosis (ILO category 1/1 

or greater) was 48% after 30 years of employment. A multivariate Cox's proportional hazards 

model indicated that silicosis increased linearly up to the period of 25-29 years employment. A 

hazard risk of 14.6 was found comparing those with ≥ 30 years exposure to those employed 10 

years. Smoking significantly contributed to the model, but its role was unclear. 

Legrand-Cattan et al. (1998) examined the dose-response relationship in two French ceramic 

plants. A 1992 cross-sectional study included more than 200 silica-exposed workers. Three ILO 

certified B readers read chest radiographs. Silica was sampled in the airborne dust. The results 

are tabulated below (Table 14). 

Table 14. Silicosis in two French ceramic plants (Legrand-Cattan et al., 1998) 

Cumulative 

exposure to 

silica in 

(mg/m3 – years) 

Number of 

workers at this 

level 

Number with 

small opacities 

with ILO 

profusion ≥ 1/0 Percent 

< 0.35 50 2 4 

0.35 – 1.08 57 8 14 

1.09 – 1.77 55 11 20 

> 1.77 55 17 31 

 Total 217 38 (18) 

A dose response relationship is clear; the authors reported a p value of 0.002. However, the 

study is limited by the lack of follow-up of the workers. 

Appendix D3 501 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Slate workers (Glover et al., 1980; Saiyed et al., 1985; Saiyed and Bannerjee, 1985)

Slate contains calcium carbonate, iron oxides, silicates, amorphous silica, and crystalline silica. 

Glover et al. (1980) studied slate workers in North Wales. The respirable slate dust contained 

13-32% crystalline silica. In the study group were 725 current and former workers exposed only 

to slate dust, while the controls were 530 men from the same area who had never been exposed 

to dust. Pneumoconiosis was found in 239 slate workers (33 %), and 10% had degrees of 

pneumoconiosis (category 2 or higher using the 1971 ILO scheme) that would bring worker’s 

compensation. The prevalence of respiratory symptoms (cough, phlegm, dyspnea) was high. 

There was evidence of an effect of both simple and complicated pneumoconiosis on lung 

function (declines in FVC and FEV1) additional to the effect of age. The high prevalence (40-

50%) of radiological lesions suggested the presence of healed tubercular lesions in men over 55. 

Either pneumoconiosis or old tubercular lesions (or both) could account for the symptomatology 

and disability of the men. 

Saiyed et al. (1985) surveyed the slate-pencil industry in India. An industrial hygiene survey 

revealed very high levels of free silica (2-10 mg/m3), while a medical survey showed that 324 of 

593 workers (54.6%) had silicosis. Of these, 105 had “conglomerate” silicosis (progressive 

massive fibrosis, PMF). Some lung lesions were detectable after less than five years of exposure 

to slate dust. Saiyed and Bannerjee (1985) conducted a follow-up examination 16 months later. 

The progression of silicosis was very rapid, and a total of 23 workers had died during this period 

(mean age = 34.7 years; mean exposure = 11.9 years). The authors attributed the high mortality 

to high levels of silica leading to early onset of PMF. The progression of silicosis was related to 

the intensity and duration of dust exposure, and to the severity of silicosis found initially. 

Silicosis has been reported in other groups of slate workers in Norway (Bang and Suhr, 1998; 

Suhr et al., 2003) and in Germany (Mehnert et al., 1990). 

Silica particle size

Data on silica particle size in the various workplaces are limited. According to Witschi and Last 

(2001), silica particles with a diameter of 1 μm (range = 0.5 - 3 μm) appear to be the most 

fibrotic in humans. NIOSH (1974) reviewed the existing literature and found that in five 

diatomite plants the mean silica diameter was 1.1 μm (range = 0.5 - 2 μm). For nine potteries, 

the particle size was 1.2 μm. For 18 foundries, more than 90% of the particles were less than 3 

μm. The majority of particles to which shipyard sandblasters were exposed was also less than 3 

μm. In the Vermont granite sheds, 10 mppcf (million particles per cubic foot) granite dust were 

initially estimated to be equal to 0.1 mg/m3

 respirable quartz. Steenland and Brown (1995) used 

this estimate for silica in South Dakota gold mines. Assuming that the density of quartz is 2.65 

g/cm3

 and that the quartz particles are spherical, the data indicate that the particles have a 

diameter of 0.59 μm. NIOSH (1974) listed 0.94 μm as the median particle size in metal mines. 

No indication was given of the dispersion of the particle sizes around the average value. Davis et 

al. (1983) used the value of 10 mppcf in granite sheds as equal to 0.075 mg/m3

 silica. For that 

estimation, OEHHA staff calculated the particle diameter to be 0.53 μm. Thus, existing data 

indicate that the majority of silica in the workplace is respirable. In most of the occupational 

studies examined, the exposures were measured using a calibrated cyclone sampler similar to 

that recommended in the current NIOSH (2003) method. This allows collection of particles 

Appendix D3 502 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

primarily in the 0.5 – 5 μm range, with a collection efficiency profile intended to match the 

penetration of particles into the alveolar region of the human lung. In the case of the South 

African gold mine studies (Beadle, 1971; Page-Shipp and Harris, 1972; Hnizdo and SluisCremer, 1993), particle number was determined by an optical method selecting respirable 

particles (range of 0.5 to 5 μm). Thus, the risk estimates obtained from these studies refer to 

particles in the size range where penetration occurs into the respiratory region of the lung. This 

corresponds to the size range of particles thought to be responsible for silicosis. It differs from 

the definition of “respirable” particles (i.e. PM10) commonly used in environmental 

measurements, which refers to particles capable of penetrating anywhere in the lower respiratory 

tract (described as “thoracic” particles in occupational studies). 

Risk estimation for silicosis from epidemiologic studies

The data from the above studies have been used by a number of investigators (Finkelstein, 2000; 

Chen et al., 2001; Hughes, 1995) and by OEHHA staff to estimate percent silicosis based on 

cumulative silica exposure in units of (mg/m3

)-yr. The results are summarized in Table 15. 

In Table 15, more than 14,000 workers were studied, of whom approximately 12% were 

classified as silicotic. The 12% is likely an underestimate of the incidence of silicosis due to lack 

of follow-up by chest radiographs during life in some cohorts and to the lack of an autopsy after 

death. 

Appendix D3 503 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 15. Summary - Estimates of % silicosis based on cumulative silica exposure 

 in (mg/m3)-y

Study 

Population (number 

with silicosis) 

Exposure of 

2 (mg/m3

)-y 

Exposure of 

4 (mg/m3)-y 

Exposure of 

4.5 (mg/m3)-y 

Muir et al., 

1989 

2109 male Ontario 

hard rock miners 

(“15”) 

0.4a,c 1.2a,c 2b

Rosenman et

al., 1996 

1072 Midwestern 

foundry workers (28) 

2a 10a 3b

Graham et

al., 1991 

408 Vermont granite 

workers (35) 

~3c ─ ─

Hughes et 

al., 1998 

1809 white male 

diatomaceous earth 

workers (81) 

1.1 (low intensity)

3.7 (high intens.)a

 4 (low) 

12 (high)a ─

Park et al., 

2002 

2342 white male 

diatomaceous earth 

workers (80) 

~7e 13e 14e

Hnizdo & 

SluisCremer 1993 

2235 white male 

South African gold 

miners (313) 

 5a

10c

52a

60c

77b

Ng & Chan, 

1994 

338 male Hong Kong 

granite workers (36) 

6a 15a 15-20b

Steenland & 

Brown, 1995 

3330 male S. Dakota 

gold miners (170) 

8a 53a 70b

Kreiss & 

Zhen, 1996 

100 miners in 

Leadville, CO (32) 

11a 53a 92b

Chen et al., 

2001 

3010 Chinese tin 

miners (1015) 

14d 47d 55b

Churchyard 

et al., 2004 

510 black gold 

miners (93) 

~28 - - 

a

 From Table II of Finkelstein (2000) 

b

 From Table 6 of Chen et al. (2001) c

 From Tables 3 and 4 of Hughes (1995) 

d

 Interpolated by OEHHA staff from Fig. 2 of Chen et al. (2001). e

 Estimated by OEHHA staff from Table 4 of Park et al. (2002) 

f

 158 had an ILO reading ≥ 1/0, while 103 had an ILO reading ≥ 1/1. 

Appendix D3 504 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Determination of LOAEL and NOAEL for silicosis (Rice and Stayner, 1995) 

In another approach to the data, Rice and Stayner (1995) identified the NOAEL and LOAEL for 

silicosis in several studies (Table 16). The study of Hnizdo and Sluis-Cremer (1993) yielded 

both a LOAEL and a NOAEL. 

 Table 16. Estimates of NOAELs and LOAELs for silicosis (Rice and Stayner, 1995) 

Study Subjects NOAEL in μg/m3 LOAEL in μg/m3 

Davis et al., 1983 969 granite workers 67.5 

Hnizdo and SluisCremer, 1993 2235 gold miners 7 20 

1321 gold miners - 8a McDonald and 

Oakes, 1984 64 gypsum miners 35 49 

Muir et al., 1989 2109 gold miners Could not 

determine 

Could not 

determine 

Rice et al., 1986 888 dusty trade 

workers 

80-100 200-252 

a

 McDonald and Oakes (1984) considered this value to be only an approximation. 

Proposals to change the occupational exposure limit

Silicosis is still being diagnosed at death in workers who were supposed to be exposed to 

occupational levels of 50-100 μg/m3. Thus, there have been recommendations that the 

occupational exposure limit for respirable, crystalline silica (specifically alpha-quartz) be 

lowered from the current level of 100 μg/m3 to 50 μg/m3 (NIOSH, 1974; Rosenman et al., 1996; 

ACGIH, 1999; Finkelstein, 2000). In 2000, the ACGIH lowered its TLV for quartz from 100 to 

50 μg/m3. In 1986, WHO recommended that the occupational level be set at 40 μg/m3 (WHO, 

1986). Greaves (2000) recommended that the TLV be lowered to 10 μg/m3. Based on existing 

data Greaves (2000) estimated that at 10 μg/m3

 the incidence rate for ILO grade 1/0 silicosis 

would be less than 5%, while for grade 1/1 it would be less than 2%. Chen et al. (2001) 

recommended that the TLV be lowered to 5 μg/m3

. "If the lifetime risk of silicosis is to be under 

1 in 1000 (a criterion used by OSHA) for a lifetime exposure of 45 years, then the mean Chinese 

total dust concentration must be lower than 0.14 mg/m3 (or lower than 0.005 mg/m3 respirable 

crystalline silica)" (Chen et al., 2001). Mannetje et al. (2002) pooled data from six occupational 

cohorts. These included four groups discussed above: diatomaceous earth workers, Vermont 

granite workers, U.S. industrial sand workers, and South Dakota gold miners. Among them 170 

deaths from silicosis were reported. The estimated mortality risk from silicosis to age 65 after 45 

years of exposure at 100 μg/m3 silica was 13 per 1000, while the risk of death at 50 μg/m3 was 

estimated at 6 per 1000. Both estimates are above the 1 per 1000 risk acceptable to OSHA. 

Mannetje et al. also concluded that the occupational standards for silica should be lowered, but 

they did not specify a level. They further state that their estimates of silicosis mortality are 

probably underestimates due to exposure misclassification and to outcome misclassification, 

since deaths due to silicosis might have been coded to tuberculosis or chronic obstructive 

pulmonary disease. 

Appendix D3 505 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

C. Silica exposure and lung cancer in workers 

In 1997, IARC classified respirable crystalline silica in Class 1, a Known Human Carcinogen, 

based on occupational epidemiologic studies. However, chronic RELs are not based on cancer 

endpoints. Further, there is no approved cancer potency factor for silica. 

V. Effects of Animal Exposures 

Several papers have reported that freshly fractured quartz, which has increased surface activity, 

causes greater inflammation than "aged" quartz. Vallyathan et al. (1991) reported that “fresh” 

silica was 4.2-fold more potent than silica aged for 1-2 days in decreasing the membrane 

integrity of male rat macrophages; 50% more potent in activating hydrogen peroxide secretion 

by macrophages; and 4.6-fold more potent in stimulating cellular chemiluminescence. 

Vallyathan et al. (1995) reported that inhalation of 19.3 mg/m3 aged (for 2 months) quartz for 

five hours/day for 10 days by male Fischer 344 rats increased the number of cells recoverable by 

bronchoalveolar lavage (BAL) (Table 17). Aged quartz also gave histopathological evidence of 

increased pulmonary infiltrates, showed higher levels of biochemical markers of lung injury, 

increased lipid peroxidation, and increased the ability of pulmonary phagocytes to produce more 

oxygen radicals than air-exposed controls. These pulmonary responses were significantly more 

pronounced after inhalation of 22.4 mg/m3

 freshly fractured quartz. 

Table 17. Cells recovered in bronchoalveolar lavage from rats (Vallyathan et al., 1995) 

Cell type Room air Aged quartz Freshly fractured Fresh/aged 

Total cells 7.1±0.78* 9.3±1.2 20.4±2.2 2.2 

Macrophages 6.7±0.69 4.7±0.79 5.4±0.78 1.1 

Neutrophils ≥ 0.038 5.3±0.66 10.4±1.44 2.0 

Lymphocytes ≥ 0.038 1.7±0.25 3.6±0.27 2.1 

Red blood cells ≥ 0.038 1.7±0.26 6.0±0.57 3.5 

* Cell counts are in millions. Each value is the mean ± standard error of 5 rats. 

Burns et al. (1980) exposed female Balb/c mice for up to 39 weeks to 4.9 mg/m3 Min-U-Sil 

brand crystalline silica. By 24 weeks, silica-laden macrophages were present in the lungs. After 

39 weeks of exposure, silicotic lesions were seen in the lungs and adjacent lymph nodes (Table 

18). 

Davis et al. (1998) exposed mice to an aerosol of cristobalite silica (mass median aerodynamic 

diameter (MMAD) = 1.7 μm) for five hours/day in order to examine (1) the effects of exposure 

dose, (2) the evolution of disease over time, and (3) the variation in responses among strains. In 

C3H/HeN mice, incremental, cumulative exposure doses of cristobalite (10 mg/m3 for 8 days, 43 

mg/m3 for 9 days, and 70 mg/m3 for 12 days) caused (1) increased initial lung dust burden at 12 

to 16 weeks post-exposure, (2) progressively intense pathological responses, and (3) increased 

total lung collagen (as measured by hydroxyproline). 

The histopathological changes and total lung collagen increased with time after exposure. 

Silicosis was compared in four inbred strains of mice (BALB/c, C3H/HeN, MRL/MpJ, New 

Zealand Black) 16 weeks after aerosol inhalation exposure to cristobalite (70 mg/m3

, 5 

hours/day, 12 days). C3H/HeN mice had histopathological silicotic lesions, enlarged 

Appendix D3 506 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

intrapulmonary lymphoid tissue, and increased lung wet weight, increased bronchoalveolar 

lavage (BAL) recoverable macrophages, lymphocytes, and neutrophils, and increased total lung 

collagen (hydroxyproline analyses). BALB/c mice developed slight pulmonary lesions. 

MRL/MpJ mice showed prominent pulmonary infiltrates with lymphocytes. New Zealand Black 

(NZB) mice developed extensive alveolar proteinaceous deposits, inflammation, and fibrosis. 

The authors found both dose-time-response relationships and a substantial variation of responses 

among mouse strains to the high level, short duration exposure. 

At Brookhaven National Laboratory, groups of Fischer 344 rats were exposed to 0, 2, 10, and 

20 mg/m3

 Min-U-Sil brand silica (alpha-quartz) for six months (Kutzman, 1984a; as summarized 

by USEPA, 1996). Other groups of rats had the same exposure, but were allowed to "recover" in 

air for an additional 6 months (Kutzman, 1984b; as summarized by USEPA, 1996). Significant 

alterations in total lung weight, total lung collagen, total elastin per unit lung dry weight, and 

total protein per unit lung dry weight at 2 mg/m3

 silica and microscopic evidence of silicotic 

lesions at the higher silica levels indicated that 2 mg/m3

 was a LOAEL for silica effects. After 

six months in clean air, the silica-induced lesions appeared to worsen. 

Muhle et al. (1989) exposed groups of 50 male and 50 female rats to 1 mg/m3

 DQ12 quartz six 

hours/day, five days/week for 24 months. DQ12 contains 87% crystalline alpha-quartz, has a 

mass median aerodynamic diameter (MMAD) of 1.3 μm, and is 74% respirable. Moderate 

fibrosis was seen in 85 animals, slight fibrosis in 13, and very slight fibrosis in 1 (total rats with 

fibrosis = 99/100). Varying amounts of peribronchial granulomatous foci were noted in 95 rats. 

Muhle et al. (1998) reported lung fibrosis in hamsters exposed to 3 mg/m3 DQ12 silica. After 18 

months of exposure to DQ12 for 6 h/day, 5 days/week, all hamsters in the group of 15-19 

animals necropsied had very slight fibrosis. Approximately 100 silica-exposed animals were 

exposed for five more months to air only. Afterward 22.2% had very slight fibrosis, 68.7 % had 

slight fibrosis, and 1% had moderate fibrosis (i.e., more than 90/100 hamsters had lung fibrosis). 

No collagen measurements were reported. Thus, rats, mice, and hamsters show pulmonary 

fibrosis after crystalline silica exposure at and above 1 mg/m3. 

Wagner et al. (1968) exposed dogs up to 2.5 years, guinea pigs up to 18 months, and rats up to 2 

years for 6 hours/day, 5 days/week to 61% cristobalite (in calcined diatomaceous earth). Dust 

exposures were 2 and 5 million particles per cubic foot (mppcf), equivalent to 0.2 and 0.5 mg/m3

cristobalite (USEPA, 1996), with occasional excursions to 50 mppcf. No lung fibrosis was 

detected at these levels but all levels caused accumulation of inflammatory cells in the lung 

parenchyma. However, in dogs fibrotic nodules developed in the hilar lymph nodes with more 

nodules at 5 mppcf than at 2 mppcf. 

Scheuchenzuber et al. (1985) examined immunologic responses in Balb/c mice following 

inhalation of 1.954 mg/m3 silica for 150, 300, or 570 days. Mice exposed for 570 days were 

tested immediately post-exposure. Those exposed for 150 or 300 days were tested immediately 

or were rested for 30 or 150 days to allow for possible recovery from effects of dust inhalation. 

Silica inhalation suppressed the number of specific plaque-forming cells (PFC) in the spleen 

produced in response to aerosolized E. coli. After 570 days of inhalation, silica also reduced the 

ability of alveolar macrophages to phagocytize Staphylococcus aureus in vitro and impaired the 

ability to lyse allogeneic tumor cells (from mice other than Balb/c) in vitro. Silica inhalation did 

Appendix D3 507 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

not affect antibody-dependent cell-mediated cytotoxic and mitogenic responses by splenic 

lymphocytes. (Fibrosis was not an endpoint measured, but the effect level is similar to the 

LOAELs in other animal studies.) 

Table 18. Animal studies of silica inhalation analyzed by USEPA (1996) 

Study Species Duration a LOAEL 

Muhle et al., 1989 Rat 24 mo 1.0 mg/m3

Scheuchenzuber et al., 1985 Mice 150-570 d 2.0 

Burns et al., 1980 Mice 3-39 wk 4.9 

Kutzman, 1984a Rat 6 mo 2.0 

Kutzman, 1984b Rat 6 mo + 6 mo 

recovery 

2.0 

Wagner et al., 1986 Dog Up to 2.5 yr 0.2 a

 Inhalation exposure was generally for 6 h/day, 5 d/wk. 

Quartz has the ability to induce the generation of free radicals and to cause oxidative stress in 

tissues. Many substances that affect the quartz surface can modify this ability. Some of these 

modifiers could originate from other minerals, which exist together with quartz in nature. 

Donaldson and Borm (1998) proposed that the hazard posed by quartz may vary widely 

depending on the origin of the silica sample or on its contact with other chemicals/minerals. 

Such mechanistic data could assist in the interpretation of epidemiological studies such as those 

above. Experimentally their group found that DQ12 quartz, a European quartz standard which is 

often used in experimental studies of silica effects, is much more inflammatory in rat lung than 

respirable silica collected from two workplaces (Clouter et al., 2001). 

Humans appear to show adverse effects of silica exposure at lower levels than animals (compare 

LOAELs in Table 18 to LOAELs/NOAELs in Table 16). Rodents tend to be obligate nosebreathers and to have extensive nasal turbinates, which may result in less silica reaching the 

lower lung. For silica, results in animals may not be a good predictor of human effect levels. 

Appendix D3 508 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Key study Hnizdo and Sluis-Cremer, 1993 

Study population 2235 white South African gold miners 

Exposure method Workplace inhalation 

Critical effects Silicosis (313 miners) (14 %) 

LOAEL 3 mg/m3

-years CDE (9 miners with silicosis) 

NOAEL 2 mg/m3

-years CDE (0 miners with silicosis) or 

600 μg/m3-years silica (dust = 30% silica) 

BMCL01 2.12 (mg/m3

)–yr CDE or 0.636 (mg/m3)–yr silica 

Exposure continuity 8 h/day, 5 d/wk 

Exposure duration Average of 24 years dust exposure (10-39 years)

Average experimental exposure 235 μg/m3

-yr silica at BMC01

(636 x 10 m3/20 m3 x 270 shifts/365 days) 

235 μg/m3-yr/24 yr = 9.8 μg/m3

Human Equivalent Concentration (HEC) 9.8 μg/m3

 

LOAEL uncertainty factor Not needed in BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 3 

Inhalation Reference Exposure Level 3 μg/m3

 (based on 30% silica in mine dust) 

[respirable, as defined occupationally by 

ACGIH/ISO] 

First supportive study Steenland and Brown, 1995 

Study population 3330 S. Dakota gold miners 

Exposure method Workplace inhalation 

Critical effects Silicosis (170 miners) (5.1%) 

LOAEL 0-0.2 mg/m3

-years (5 miners with silicosis) 

NOAEL Not found 

BMCL01 0.34 (mg/m3

)–yr (see text below) 

Exposure continuity 8 h/day, 5 d/wk 

Exposure duration 3-36 years (average 9 years underground) 

Average experimental exposure 112 μg/m3

-y 

(340 x 10 m3

/20 m3 x 5 d/7 d x 48 wk/52 wk) 

112 μg/m3

-y/9 y = 12.4 μg/m3 

Human Equivalent Concentration (HEC) 12.4 μg/m3

LOAEL uncertainty factor Not needed in BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 3 

Inhalation Reference Exposure Level 4 μg/m3

 [respirable, as defined occupationally] 

Appendix D3 509 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Second supportive study Hughes et al., 1998 

Study population 1809 California diatomaceous earth workers 

Exposure method Workplace inhalation 

Critical effects Silicosis (81 workers) (4.5%) 

LOAEL > 1, ≤ 3 mg/m3

-years (17 workers with silicosis) 

NOAEL ≤ 1 mg/m3

-years (6 cases). (Six cases were 

observed, but Hughes et al. assigned the group a 

Relative Risk = 1 for silicosis.) 

Exposure continuity 8 h/day, 5 d/wk 

Exposure duration 1-45 years (mean = 11.5 years) 

Average experimental exposure ≤ 330 μg/m3

-y (1000 x 10/20 x 5/7 x 48/52) 

≤ 330 μg/m3-y/ 11.5years = ≤ 29 μg/m3

Human Equivalent Concentration (HEC) 29 μg/m3

LOAEL uncertainty factor 3 (authors’ NOAEL actually is a LOAEL) 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 10 

Inhalation Reference Exposure Level 3 μg/m3 [respirable, as defined occupationally] 

Third supportive study Chen et al. (2001) 

Study population 3010 Chinese tin miners 

Exposure method Workplace inhalation 

Critical effects Silicosis (1015 workers) (33.7 %) 

LOAEL 10-19.99 mg CTD/m3

-years (24 cases) 

NOAEL ≤ 10 mg CTD/m3

-years (2 cases) 

≤ 360 μg silica/m3 - years 

BMCL01 132 μg silica/m3

 - years 

Exposure continuity 8 h/day, 5 d/wk 

Exposure duration 2.2 years for NOAEL group 

Average experimental exposure 40 μg/m3

-y (132 x 10/20 x 5/7 x 48/52) 

40 μg/m3-y/2.2 years = 18 μg/m3

Human Equivalent Concentration (HEC) 18 μg/m3

LOAEL uncertainty factor Not needed in BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 3 

Inhalation Reference Exposure Level 6 μg/m3

 [respirable, as defined occupationally] 

Appendix D3 510 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Fourth supportive study Churchyard et al., 2004 

Study population 510-520 black South African gold miners 

Exposure method Workplace inhalation 

Critical effects Silicosis (93 cases)

LOAEL 0-0.80 mg/m3

-yr (11 cases) 

NOAEL Not identified 

BMCL05 0.673 (mg/m3

)-yr 

Exposure continuity 270 shifts/year 

Exposure duration 21.8 yr (6.3-34.5) 

Average experimental exposure 249 (μg/m3

)-yr (673 x 10/20 x 270shifts/365) 

249 (μg/m3)-yr/21.8 yr= 11.4 μg/m3

Human equivalent concentration (HEC) 11.4 μg/m3

LOAEL uncertainty factor Not needed in BMC approach 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 

Intraspecies uncertainty factor 

Cumulative uncertainty factor 

Inhalation Reference Exposure Level 

The study of 2235 white South African gold miners by Hnizdo and Sluis-Cremer (1993) not only 

determined a NOAEL of 2 (mg/m3)-yr CDE (600 μg/m3-yr silica), but also had sufficient doseresponse data for a BMC derivation. This study was powerful enough to detect a 1.9% incidence 

of silicosis (9 cases out of 474 exposed) at 0.9 mg/m3-yr silica (0/204 vs. 9/474, p = 0.064 by 

Fisher exact test, two-tailed). Because this incidence represents approximately the sensitivity 

limit of the data, and silicosis is a severe irreversible endpoint, the BMCL01 (i.e., the lower 

bound estimate of the concentration at which 1% of the population develops silicosis) was 

selected as the basis of the chronic REL. In benchmark analysis of chronic animal studies, 

BMCL05 is typically regarded by OEHHA as equivalent to a NOAEL. However, the power of 

this large-scale study is sufficient to demonstrate measurable responses below the 5% incidence 

level (which cannot then be logically considered a no-effect level). Furthermore, the endpoint 

measured in this epidemiological study is considered to be severe, since it represents the 

occurrence of clinically recognizable and irreversible disease, rather than an adverse 

physiological or biochemical response or a histopathological result seen at autopsy. 

1 

3 

3 

4 μg/m3 [respirable, as defined occupationally] 

Benchmark Concentration (BMC) models, developed by the USEPA (BMDS versions 1.3, 1.3.1, 

and 1.3.2), were fit to the human data in Hnizdo and Sluis-Cremer (1993) (Table 7 and Figure 2 

above). Fitting the probit model to the log dose of the Hnizdo and Sluis-Cremer (1993) data 

yielded an MLE01 of 2.45 (mg/m3

)–yr CDE and a BMCL01 of 2.12 (mg/m3)–yr CDE (χ

2

 = 0.64; 

p value for fit = 0.9957) (Figure 5, Table 19). (For comparison the BMCL05 was 3.73 (mg/m3

)–

yr CDE.) Fitting the logistic model to the same data yielded a BMCL01 of 1.73 (mg/m3)–yr CDE 

(χ

2

 = 2.71; p value for fit = 0.8446) (Table 19). The BMCL01 from these data is about the same 

as the apparent NOAEL. In general, a BMC is preferred to a NOAEL because the BMC takes 

into account all the dose response data in a study. The apparent NOAEL may be either above or 

below an actual effect level, depending on the study design and distribution of the data. 

Appendix D3 511 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Figure 5. Probit model fit to the log dose of the Hnizdo and Sluis-Cremer data. 

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0.8

0 2 4 6 8 10 12 14 16

Fraction Affected

Dose

Log Dose/Probit Model with 0.95 Confidence Level 

BMDL BMD

Probit

BMD Lower Bound

Table 19. Fits of benchmark models to the Hnizdo and Sluis-Cremer (1993) data 

BMDS Model MLE01 BMCL01 p value for fit 

Probit-log-dose 2.45 (mg/m3)–yr CDE 2.12 (mg/m3)–yr CDE 0.9957 

Logistic-log-dose 2.07 1.73 0.8446 

Multistage (n=2) 2.47 1.89 0.7213 

Quantal-quadratic 1.62 1.54 0.5017 

Probit 1.56 1.32 0.0079 

Logistic 1.48 1.28 0.0003 

Quantal-linear 0.37 0.34 0.0000 

For the estimate of 30% silica in the South African gold mine dust, Hnizdo and Sluis-Cremer 

(1993) relied on estimates for the years 1956-1960 by Beadle (Beadle and Bradley, 1970; 

Beadle, 1971). The original data, obtained by Corner House Laboratories for the South African 

Bureau of Mines, are partly presented by Beadle and Bradley (1970), but a more detailed 

presentation of exposures for various classes of workers is given by Page-Shipp and Harris 

Appendix D3 512 Silica 
Determ

Appendix D3 

ination of Noncancer Chronic Reference Exposure Levels February 2005 

513 Silica 

(1972). The latter paper also describes in some detail the methodology used to obtain the 

particle counts, and to convert those data into either respirable surface area or respirable mass 

values. Gibbs and Du Toit (2002) reviewed the data and methodology used by Hnizdo and 

Sluis-Cremer (1993) to estimate silica exposures of workers, which apparently depended on an 

unpublished analysis of the Corner House Laboratories’ data done by Du Toit in 1991. Gibbs 

and Du Toit state that the exact relationship between the observed particle counts and 

theoretically derived mass concentrations cannot be determined, but that the uncertainties in this 

conversion do not appear to be severe for the dust characteristics observed in the South African 

mines. They accept the estimates by Beadle and Bradley (1970) of the quartz percentages in the 

dust, i.e. 54% for incinerated and acid-washed dust and 30% for unmodified dust. 

However, Gibbs and Du Toit (2002) assert that Hnizdo and Sluis-Cremer (1993) incorrectly 

applied the 30% (total dust) silica content to figures for acid-treated dust in calculating the silica 

exposures of each occupational group. This contention is supported by the footnote to Table II 

in Hnizdo and Sluis-Cremer (1993) where the respirable dust concentration is described as 

“After heat and acid treatment”. In order to clarify this point, OEHHA reviewed the independent 

reporting of the underlying data by Page-Shipp and Harris (1972). For most occupational 

groups, the silica exposures (shown in Table 20) calculated from Appendix I of Page-Shipp and 

Harris (1972), using the 54% silica content appropriate for acid-washed dust, correspond more 

closely to those calculated by Hnizdo and Sluis-Cremer (1993) (applying the 30% quartz content 

to their reported “respirable dust concentrations,” i.e., the untreated dust), than to the modified, 

and higher, quartz exposures proposed by Gibbs and Du Toit (2002). For example, 113 exposure 

samples were taken for stopers. 
Determ

A

ination of Noncancer Chronic Reference Exposure Levels 

ppendix D3 514 

Occupation 

Shaft Sinkers Developers Stopers 

Assistant 

miners/ Trammers Shift Bosses

Other Officials 

Banks/ Skips 

Workers Near shafts 

Boilermakers 

A

Page-Shipp and Harris (1972) (Table III and Appendix I)Hours/shift (t) 7.70 8.00 7.80 7.70 5.20 4.00 7.50 6.50 6.30 Number of samples 10 37 113 157 43 106 33 34 41 RM x t 4.44 1.96 1.57 1.20 0.87 0.77 1.31 0.56 1.00 s.d. 3.94 1.59 1.00 0.93 0.71 0.53 1.38 0.57 0.71 Respirable Mass (RM) 0.58 0.25 0.20 0.16 0.17 0.19 0.17 0.09 0.16 Silica (54%) 0.31 0.13 0.11 0.08 0.09 0.10 0.09 0.05 0.09 

(after acid treat m t en )

Hnizdo and Sluis-Cremer (1993) (Table II)RM 0.48 0.37 0.27 0.30 0.30 0.13 0.10 0.19 Silica (30%) 0.14* 0.11* 0.08* 0.09* 0.09* 0.04 0.03 0.06 

(before acid treat m t en )

Gibbs and Du Toit (2002) (Table 4)RM 0.48 0.37 0.27 0.30 0.30 0.13 0.10 0.19 Silica (54%) 0.26 0.20 0.15 0.16 0.16 0.07* 0.05* 0.10*  (after acid treatment) Table 20. Estimates of silica exposures in mg/m3 for different occupational groups in South African gold m* denotes that value is equal to or closer to the value based on Page-Shipp and Harris 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

The last line of Appendix I of Page-Shipp and Harris (1972) gives a mean value for stopers of 

1.57 (mg/m3

)-hours respirable dust mass after acid treatment. Since the average work shift for 

stopers was 7.8 hours (Page-Shipp and Harris, 1972, Table III, last row), the average exposure 

level was 0.20 mg/m3

. If 54% of this were quartz, the quartz level would be 0.11 mg/m3. Table 

II of Hnizdo and Sluis-Cremer (1993) lists 0.37 mg/m3

 respirable dust for stopers. Thirty % of 

0.37 mg/m3

 equals 0.11 mg/m3, the same value reported by Page-Shipp and Harris. In Table 4 of 

Gibbs and Du Toit (2002) stopers are also reported to be exposed to 0.37 mg/m3

 respirable dust. 

If 54% were quartz, as Gibbs and Du Toit contend, the quartz level would be 0.2 mg/m3

. For 6 

of the 9 categories of workers comprising 83% of the samples taken the silica levels correspond 

more closely to values used by Hnizdo and Sluis-Cremer than to those suggested by Gibbs and 

Du Toit. 

Several more recent analyses of quartz content of South African mining rock have been reported 

(Table 21). Kielblock et al. (1997) give the overall silica content of the dust as 15% for the late 

1980s to early 1990s. Dr. Eva Hnizdo (personal communication, 2003), now with the U.S. 

National Institute of Occupational Safety and Health (NIOSH), provided a summary of various 

other estimates that have been made. “Past surveys indicate that the amount of airborne 

respirable dust in SA gold mines in 1980's and in 1970's was on average around 0.4 mg/m3

 with 

average quartz concentration of 0.08 mg/m3

” (about 20%). In a Ph.D. thesis submitted by the 

late R.E.G. Rendall (1999) on dust in the air of gold mines, the silica percentage averaged 22% 

during the period from 1964 to 1988. In summary, 

(1) Notwithstanding some apparent contradictions in the various accounts, the silica 

concentrations in air proposed by Hnizdo and Sluis-Cremer, based on the Corner House 

Laboratory data, are a reasonable contemporary estimate of the exposures experienced by 

the workers examined in the study by Hnizdo and Sluis-Cremer (1993). 

(2) Other, more recent estimates of percent silica in the mine dust were lower than the value 

of 30% used by Hnizdo and Sluis-Cremer (1993). Newer studies, which using more 

sophisticated methods to measure silica in the dust, indicate lower silica concentrations in 

the various occupational settings. Since dust levels in the mines were fairly constant for 

decades and quantification of silica was improving, 30% is more likely to be an 

overestimate than an underestimate of silica levels. 

(3) Analysis of the data of Page-Shipp and Harris (1972) by OEHHA staff indicated that 

Hnizdo and Sluis-Cremer (1993) used the correct silica content, despite an erroneous 

statement in a footnote to Table II of their paper1

. 

 

1

 Dr. Eva Hnizdo reviewed this analysis of the silica content of the dust and agrees with the 

assessment. (“I am very pleased that you studied carefully all the reports and came to the 

conclusion that our study was after all reasonably correct. Based on the Churchyard study and 

the measurements data I have seen in SA during the 1990s, I am also convinced that our results 

are reasonable estimates of the exposure of the cohort.” (Hnizdo, personal communication 

October 2004) 

Appendix D3 515 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Table 21. Estimates of respirable silica fraction of South African gold mine dust 

Authors Time 

frame 

% silica Number of 

samples 

Methods 

Beadle and Bradley, 

1970 

1958-

1967 

total dust: 

25.7%; 

gravimetric: 

28.5%; 

microscopy. 

acid-washed: 

54% 

142 grav; 

143 elect 

ppt 

gravimetric; 

precipitator + 

microscopy 

Hnizdo and SluisCremer (1993) 1956-

1960 

30% precipitator + 

microscopy 

Rendall 

(unpublished thesis) 

 Survey 1 

 Survey 2 

 Survey 3 

 Survey 4a 

1987-8 

1977 

1977 

1964-7 

17% 

20% 

17% 

22% 

588 

166 

90 

112 

gravimetric 

gravimetric 

gravimetric 

gravimetric 

Hnizdo (personal 

communication) 

1970-

1989 

20% 

Kielblock (1997) ~1990 15.08% Not stated 

Churchyard (2004) 2000-1 14.3% Gravimetric + 

X-ray 

diffraction 

In the first supportive study Steenland and Brown (1995) found five cases of silicosis in the 

lowest dose group of 0 – 0.2 (mg/m3

)-yr and considered the group to be a LOAEL (Table 10 

above). None of the BMDS models gave an acceptable fit at the p ≥ 0.05 level using six or seven 

silica levels. The closest was the quantal quadratic model (χ

2

 = 9.62; p = 0.0473), which resulted 

in a BMC01 for silica of 0.43 (mg/m3

)–yr using the six lowest levels of silica. In risk assessment, 

the highest dose or doses are often dropped in order to obtain an acceptable fit of the model to 

the data. This is reasonable with the benchmark approach since the highest doses should be least 

informative and the doses in the low dose region near the benchmark should be most informative 

for the benchmark concentration (USEPA, 1995; Filipsson et al., 2003). Fitting the probit model 

to the log dose of the five lowest silica levels from Steenland and Brown yielded a BMCL01 of 

0.34 (mg/m3

)–yr CDE (χ

2

 = 1.32; p value for fit = 0.5177). [For comparison, BMCL05 = 0.85 

(mg/m3

)–yr CDE.] Fitting the quantal quadratic model gave a BMCL01 of 0.45 (mg/m3)–yr (χ

2

 = 

3.36; p = 0.3395). Use of the BMC01 value of 0.34 (mg/m3

)–yr CDE from the log dose probit 

model resulted in a chronic REL estimate for crystalline silica of 4 μg/m3

. Steenland and Brown 

stated that “silicosis has no background rate for non-exposed populations that changes with age 

or calendar time” and thus they assumed that the five silicotics in the 0 – 0.2 (mg/m3

)-yr were 

exposed to silica in the mines. 

In a second supportive study, Hughes et al. (1998) found six cases of silicosis in the lowest 

exposure group of ≤ 1 mg/m3

-yr but considered that group to be a NOAEL, not a LOAEL. If the 

lowest exposure group is used as a NOAEL, a chronic REL of 10 μg/m3

 is calculated from the 

Appendix D3 516 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

data. Hughes et al. (1998) cite examples of possible non-occupational chest radiograph opacities 

(due, for example, to age or smoking) to explain the six cases in the lowest exposure group. 

However, due to the rarity of silicosis the six cases are biologically significant. OEHHA 

considers that the six cases may be work related, not cases of environmental or background 

silicosis. When a LOAEL to NOAEL UF of 3 is applied to the data of Hughes et al. (1998), the 

estimated REL is 3 μg/m3

. 

In a third supportive study, Chen et al. (2001) found two cases of silicosis in the lowest exposure 

group of ≤ 10 mg CTD/m3

-years and considered that exposure level to be a NOAEL. One of the 

advantages of the benchmark dose analysis is that a NOAEL/LOAEL controversy, such as the 

one above with the Hughes et al. (1998) data, does not impact the procedure. The chart of the 

Chen et al. data above (Figure 4) indicates that the dose response is linear at low doses. Fitting 

the probit model to the log dose of the four lowest data points yielded a BMC01 of 0.132 (mg/m3

) 

- yr CDE (χ

2

 = 2.19; p value for fit = 0.335). Use of five, six, or seven data points gave BMC01s 

of 0.14 to 0.17, but the p values were less than 0.1. For comparison, fitting the logistic model to 

the log dose of the four lowest data points yielded a BMCL01 of 0.093 (mg/m3

) - yr CDE (χ

2

 = 

4.86; p value for fit = 0.0879). An inhalation chronic Reference Exposure Level for crystalline 

silica of 6 μg/m3

 was estimated from the Chen et al. data. 

The fourth supportive study is that of black South African gold miners by Churchyard et al. 

(2003, 2004). A problem with this data set is the statistical “noise” in the lower exposure 

groups; e.g., the lowest exposure group has a higher incidence of silicosis (11/103) than the next 

group (8/97). This noise causes problems in estimating a low benchmark such as the BMCL01

used with the Hnizdo and Sluis-Cremer (1993) data and with data from Steenland and Brown 

(1995) and Chen et al. (2001). The calculation therefore uses a 5% BMCL of 0.673 (mg/m3

)-yr 

from the probit log dose model as the benchmark, which is reasonably well within the range of 

the reliably observed data and which does not differ too widely from the MLE05 estimate of 

0.955 (mg/m3

)-yr for that parameter. This BMCL05 is not strictly comparable to the BMCL01

calculated from the data of Hnizdo and Sluis-Cremer, but the concerns about the severity of the 

effect noted in the discussion of that derivation apply with equal or greater force here. 

On the other hand, the variability in the low-dose data in this study implies that an unobserved 

factor is affecting the data. All the models predict that the “background” is substantially (as 

much as 5 – 10%) above zero, which is intrinsically implausible for silicosis unless there is an 

unrecorded additional source of silica exposure. Possibly, there were occasional excursions in 

the exposure of the workers in less exposed jobs, which were not captured by the systematic 

assessments for these job classifications. Alternatively, perhaps the assignment of an assumed 

zero exposure value to “non-dusty” job classifications noted in the paper was in fact inaccurate 

for some individuals. There may also be some distortion of the curve resulting from the 

“binning” of the exposure categories; certainly, the bin widths hinder the attempt to calculate a 

BMDL01 in this case. In relation to the model fit, other models (including the quantal linear 

model, which emphasizes the likely more reliable incidences at higher dose levels) are consistent 

with the BMCL05 results from the log probit model used here. Finally, the total number of cases 

and controls examined is fewer than in Hnizdo and Sluis-Cremer’s study, which reduces the 

precision. 

Appendix D3 517 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

None of these issues can be resolved without recourse to the individual data, which were not 

available for this analysis, and may not be resolved even then. However, the derivation of an 

HEC of 11.2 μg/m3

 (and thus a comparison REL of 4 μg/m3) indicates at least that the results of 

the earlier analysis are unlikely to have underestimated the proper value of the REL. Although 

the uncertainties in the Churchyard et al. data prevent us from being more precise, we cannot 

eliminate the possibility that the REL should be set lower. However, the California ambient 

monitoring data, although subject to considerable uncertainty as to the relevant particle size 

distributions, suggest a plausible lower bound on the REL, which is consistent with our analysis 

of Hnizdo and Sluis-Cremer’s data. 

Other investigators have approached the possibility that some opacities on radiographs may be 

due to background influences such as age and smoking. In regard to smoking, Blanc and Gamsu 

(1988) reviewed the literature and concluded that smoking would not interfere with the 

determination of silicosis by the ILO system. Based on reading 1422 films of unexposed bluecollar workers, Castellan et al. (1985) stated that the use of the median result of 3 readers (the 

same number used by Hughes et al.) rarely results in interpreting a chest radiograph as ILO 

category ≥ 1/0 in workers who were not exposed to dust (and regardless of smoking status). 

The USEPA (1996) did a benchmark analysis with the Hnizdo and Sluis-Cremer (1993) data. 

They estimated that the lower bound for a 1% risk for silicosis (BMCL01) was 1.31 (mg/m3

)-yr, 

which by their methods is equivalent to a continuous, 70-year exposure to 6.7 μg/m3

 silica. 

However, USEPA did not do a formal Reference Concentration (RfC) derivation for silica by 

either the BMC/UF or NOAEL/UF approach. 

The key (Hnizdo and Sluis-Cremer, 1993) and supporting (Steenland and Brown, 1995; Hughes 

et al., 1998; Chen et al., 2001) studies were of human adults, nearly all males, who were 

presumably healthy, at least initially, since they were able to work. Thus there is need to protect 

the sensitive members of the population, especially children, in whose airways penetration of 

silica particles will be greater (Phalen et al., 1985; Schiller-Scotland et al., 1994; Oldham et al., 

1997; Bennett and Zeman, 1998). In addition, women may be more sensitive than men to the 

development of silicosis (Gerhardsson and Ahlmark, 1985; Katsnelson et al., 1986). The 

selection of three as the intraspecies uncertainty factor (UFH) was based on several 

considerations. 

(1) The workers who developed silicosis at low silica concentrations are by definition the 

most sensitive workers to silica-induced silicosis. Because of the large population of 

workers examined in these studies (more than 14,000), the sensitive individuals represent 

at least part of the range of sensitivity to be expected in the general population. This may 

justify reducing the UFH from the default value of 10. Since these workers did not 

include children, the elderly, or females (except for the 215 females in Chen et al.), some 

uncertainty related to inter-individual variability remains. Therefore, a UFH of 3 rather 

than 1 is chosen. 

(2) Mukherji et al. (1993) reported mean ambient silica levels (in PM10) at three locations in 

the northern part of Santa Barbara County, California (see the Appendix to this report). 

At Santa Maria (an urban site) the level was 2.3 μg/m3

; in Santa Ynez (a rural site) 0.6 

μg/m3; and in Buellton (a remote background site) 0.2 μg/m3 crystalline silica. Thus, use 

Appendix D3 518 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

of a human intraspecies uncertainty factor (UFH) of 10 with the data from the key study 

would result in an estimated chronic REL of 0.9 μg/m3

 (ACGIH method), a level in the 

range of ambient levels in California. Although the reported levels at the urban site may 

(according to the authors) have reflected some anthropogenic contributions such as 

disturbance and tracking of siliceous road dust, the rural and remote site values are 

apparently (perhaps conservatively) reflective of the natural background to which all 

California residents are exposed. (U.S. EPA (1996) found slightly higher average 

ambient levels of silica in PM10; this average may include some sites affected by 

disturbance and emissions.) There is no evidence that these background levels of silica 

are causing silicosis. On the other hand, silicosis in the general population is not a target 

for medical attention, and autopsy rates are very low, so the possibility of a low 

frequency of response at these levels cannot be entirely dismissed. On balance, it appears 

plausible that a REL of 3 μg/m3

 (benchmark + UFH = 3) would be protective of the 

general population. 

(3) The dose-response curve for silicosis due to inhalation of crystalline silica is steep, and 

an upward curvature of this dose response was seen in some studies (Figure 7-1 in 

USEPA, 1996). It is notable that, whereas exposures in the 1-3 μg/m3

 range are 

apparently without effect (based on the benchmark calculations and the California 

ambient background data), Rice and Stayner (1995) described a LOAEL for silicosis of 

8 μg/m3

 in gold miners (Table 16; based on data from McDonald and Oakes [1984]). 

This finding may partly reflect differences in physical state of the silica, and coexposures, but it might indicate that, although the chronic REL should be protective of 

public health, chronic exposures only moderately exceeding the REL may lead to 

clinically observable disease. 

The animal studies gave LOAELs for silica of 0.2 mg/m3

 in dogs and from 1 to 4.9 mg/m3 in 

rodents. After extrapolation to equivalent continuous time and application of LOAEL to 

NOAEL, interspecies, and intraspecies UFs, the estimated chronic RELs from animal data are all 

less than 1 μg/m3

. This reflects in part the greater uncertainty in extrapolating from animal 

studies to predicted human health effects. 

The silica particles of concern in the causation of silicosis are those of respirable size. California 

EPA defines ‘respirable’ as particles 10 μm or less MMAD. This reflects one usual type of 

sampler (for “PM10”) used for ambient air sampling in the general environment. The other usual 

type of environmental sampler, PM2.5, collects even smaller particles. There are differences in 

the size range distribution between a typical PM10 measuring device and the NIOSH type 

personal samplers, or other devices with similar size selection properties, used by the 

investigators in the epidemiological studies. The NIOSH-type samplers capture 50% of particles 

with a MMAD of 4 μm, and higher percentages of smaller particles. A smaller proportion of 

larger particles between 4 and 10 μm in aerodynamic diameter will also be collected. Figure 5, 

from Volume I of U.S. EPA’s Third External Review Draft of Air Quality Criteria for Particulate 

Matter (April 2002), includes particle penetration curves for PM10, PM2.5, and occupational 

samplers. 

Appendix D3 519 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

Figure 5. Size cut curves of particle penetration 

The NIOSH samplers are designed to mimic the size range of particles that reach into the 

bronchiolar and alveolar spaces (what the occupational community calls respirable). PM10

samplers are meant to capture particles that penetrate the entire length of the lower respiratory 

tree, including those that penetrate to the tracheobronchial and alveolar regions. Penetration (and 

therefore presumably deposition) by particle size is complex, and is dependent on the 

aerodynamic diameter, hygroscopicity, and electrostatic charge of the particles, and on a number 

of host factors including airway structure and geometry, as well as depth, rate, and mode of 

breathing (nasal vs. oronasal). The fractional penetration in the various regions of the respiratory 

tract is not linear with respect to size. Generally, though, larger particles impact higher in the 

respiratory tree (the extrathoracic and tracheobronchial regions), while smaller particles show 

greater penetration to the lower tracheobronchial and alveolar regions. There are a number of 

models of regional penetration and deposition in the respiratory tract, as well as some 

measurements. Chan and Lippmann (1980) showed peak alveolar deposition for particles about 

3 μm MMAD with deposition dropping above and below that. Their data and model indicate 

that tracheobronchial deposition rises rapidly above about 3 μm MMAD. Available data also 

Appendix D3 520 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

indicate significant inter-individual variability in fractional deposition. The ICRP (1994) model 

used in evaluating risk from radioactive particles indicates that total deposition in the respiratory 

tract for particles 3 μm in activity median thermodynamic diameter (AMTD) is about 0.78 with a 

regional deposition fraction of 0.077 for the alveolar region for a reference male worker during 

nasal breathing. The same model predicts a total deposition in the respiratory tract of 0.77 for 

10 μm AMTD particles and a deposition fraction of 0.024 in the alveolar region. Thus, many 

particles with a 10 μm MMAD get into the alveolar space. A smaller difference in regional 

penetration and deposition is predicted for mouth breathers. Therefore, if only the size range 

measured by the samplers used in the studies were considered, the measurement might 

underestimate the amount of silica that reaches the gas exchange regions of the lung, depending 

on the actual particle size distributions in the occupational studies and in the environments in 

which the REL is to be applied. Unfortunately, neither the occupational nor the environmental 

silica particle size distributions are known in detail; measurements have been reported only in 

terms of NIOSH sampler results or PM10 cutoff values. 

It is generally assumed that the silicosis is induced by that fraction of the silica that reaches the 

alveoli. Nevertheless, no actual data exonerate the coarser particles in the 4 - 10 μm range. A 

fraction of these particles can enter the bronchioles and alveoli. However, some data from South 

African gold mines indicate that more than 99% of the crystalline silica dust can be in the PM2.5

fraction (Sichel, 1957). Thus, the samplers used in the key study appear to be collecting the 

biologically relevant range of particles in that situation. 

In the absence of comprehensive data on the silicosis-inducing activity of different particle sizes, 

it is not possible to adjust the REL for different particle size distributions, which might be found 

in the general environment, or for different measurement methods. The REL is therefore 

specified as applicable to concentrations of particles having a size range (and reactivity) similar 

to those measured in the occupational studies [respirable as defined occupationally (ISO, 1995; 

NIOSH, 2003; ACGIH, 2004)]. Results obtained by other sampling methods would need to be 

corrected for any difference in size selectivity of the method used. Such a correction factor 

would be specific to the particle size distribution present at the site studied, so no general 

correction factors can be proposed. A more inclusive sampling procedure, such as that used for 

PM10, would overestimate the relevant exposure in any situation, and so would be inappropriate 

for precise risk quantification. However, PM10 would be useful as a screening method to 

establish that a particular situation is unlikely to present a hazard. For example, if the silica 

concentration in PM10 modeled at a receptor is less than the REL (3 μg/m3

), occupationally 

respirable silica will also be less than 3 μg/m3

, so a facility would not pose a risk due to silica at 

that receptor. If the silica concentration in PM2.5 modeled at a receptor is less than 3 μg/m3

 but 

PM10 is greater than 3 μg/m3

, further testing would be needed. If both PM2.5 and PM10 exceeded 

the REL, the chronic Hazard Index would exceed 1 to an undetermined extent, suggesting a need 

for risk management. More precise determination of the amount of material in the respirable size 

fraction for environmental samples may require further work on measurement methodology, 

since ISO (1995) and similar occupational methods have not been validated for the lower levels 

encountered in environmental samples. 

Appendix D3 521 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for silica include: 

(1) The availability of several long-term studies of inhalation in workers at varying 

exposure concentrations (see Summary Table 15 above), with adequate 

histopathological and radiological analysis, and with adequate follow-up. 

(2) The finding of a dose-response effect for silicosis in several of the studies (e.g., Hnizdo 

and Sluis-Cremer, 1993; Steenland and Brown, 1995; Chen et al., 2001). 

(3) The observation of a NOAEL in some studies including the key study (summarized by 

Rice and Stayner, 1995). 

(4) The power of the Hnizdo and Sluis-Cremer (1993) data to detect a small effect. 

Major areas of uncertainty are: 

(5) The limited follow-up of the cohort members in some studies (e.g., Muir et al., 1989; 

Rosenman et al., 1996) with consequent under-ascertainment of silicosis (even to the 

extent that such studies are useless for determining exposure-response). 

(6) The general underestimation of silicosis by radiography alone (Hnizdo et al., 1993), 

which results in higher, less health-protective chronic REL estimates. 

(7) The possible underreporting of silicosis where complete radiographic data and autopsy 

data are not available (Steenland and Brown, 1995). 

(8) The uncertainties in exposure estimation, especially when reconstructing historical 

levels of silica exposure (Seixas et al., 1997; Gibbs and Du Toit, 2002) including the 

variability in the estimates of percent quartz in the South African mine dust (Beadle, 

1971; Hnizdo and Sluis-Cremer, 1993; Kielblock et al., 1997; Gibbs and Du Toit, 2002; 

Hnizdo, personal communication) and when converting particle counts to mass. 

(9) The differences in percent silicosis in different studies at what were considered similar 

silica levels and similar exposure duration (see Summary Table 15 above). 

(10) The variability in toxicity of various forms of silica (e.g., freshly fractured vs. aged 

quartz; cristobalite vs. quartz) although all forms have toxicity (Table 17). 

(11) The limited information on silica particle size (including its variability) in the 

epidemiological studies, other than that the silica was respirable, and the variability in 

particle penetration and deposition as a function of particle size in the respiratory tract in 

the human population (e.g., Heyder et al., 1982; ICRP, 1994; Hattis et al., 2001). 

(12) The use of area samplers rather than personal samplers to estimate exposure, which 

usually results in an underestimation of silica exposure (Cherrie, 1999). 

Appendix D3 522 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

VIII. Potential for Differential Impacts on Children's Health 

Silica is a respiratory irritant and a modifier of immune function. Since the key study involved 

over 2000 men, some were likely to be more sensitive to silica than others. In addition, we used 

a benchmark of 1% adverse effect, rather than the usual 5 %. Thus, use of the human 

intraspecies uncertainty factor (UFH) of 3 should result in a REL that adequately protects most 

members of the general population. Exacerbation of asthma, which has a more severe impact on 

children than on adults, is a known response to some respiratory irritants. However, there is no 

data on such a response to silica in infants or children. The epidemiological studies used in the 

derivation of the REL did not include children. If children’s susceptibility were much greater 

than that of adults, it would be expected that clinical disease would be evident in children 

following exposures in the upper range of the respirable silica levels measured in ambient air in 

California. No such reports have been identified in the literature. There are no data on silica’s 

effects on the immune system of children. 

OEHHA is currently evaluating its risk assessment methodology, in particular the UFH, for its 

adequacy in protecting infants and children. Since children have smaller airways than adults and 

breathe more air on a body weight basis, penetration and deposition of particles in the airways 

and alveoli in children is likely greater than that in adults exposed to the same concentration 

(Phalen et al., 1985; Schiller-Scotland et al., 1994; Oldham et al., 1997; Bennett and Zeman, 

1998). 

Appendix D3 523 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

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Appendix D3 531 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

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Appendix D3 532 Silica 
Determination of Noncancer Chronic Reference Exposure Levels February 2005 

X. Appendix 

Particulate Levels of Interest for Exposure to Respirable Crystalline Silica Isomorphs 

150 μg/m3

 Federal 24 hour PM10 standard 

65 μg/m3

 Federal 24 hour PM2.5 standard 

50 μg/m3

 California 24 hour PM10 standard 

50 μg/m3

 Federal PM10 annual standard (chronic exposure) 

50 μg/m3

 8 hour TLV for quartz, cristobalite, and tridymite for workers (ACGIH Method) 

50 μg/m3

 estimated workplace LOAEL for silicosis from studies by Theriault et al. (“) 

20 μg/m3

 CA annual PM10 standard (chronic exposure) (arithmetic mean) 

15 μg/m3

 Federal annual PM2.5 standard (chronic exposure) 

12 μg/m3

 CA annual PM2.5 standard (chronic exposure) (arithmetic mean) 

12 μg/m3

 current silica TLV adjusted to equivalent continuous exposure 

(50 μg/m3

 x 8 h/24 h x 5 d/7d) (ACGIH) 

10 μg/m3

 TLV for silica proposed by Greaves (2000) (ACGIH) 

8 μg/m3

 current silica TLV further adjusted by 46/70 years occupational exposure 

 (ACGIH) 

8 μg/m3

 estimated high-end ambient crystalline silica level in US (USEPA, 1996) (PM10) 

6.7 μg/m3

 lower bound on 1% risk of silicosis estimated by USEPA (1996) (PM10) 

5 μg/m3

 TLV for silica proposed by Chen et al. (2001) (ACGIH) 

5 μg/m3

 “acceptable” ambient level for silica (10% of PM10) (USEPA, 1996) 

5 μg/m3

 RfC for diesel exhaust particulate, a respirable PM (PM2.5) 

3 μg/m3

 estimated average ambient exposure to crystalline silica (USEPA, 1996) (PM10) 

3 μg/m3

 draft silica chronic REL proposed by OEHHA (ACGIH) 

2.3 μg/m3

 (1.17-3.46; n=12)* silica level during 1989 in Santa Maria, CA (urban site) (PM10) 

0.6 μg/m3

 (0-1.44; n=16)* silica level during 1989 in Santa Ynez, CA (rural site) (“) 

0.2 μg/m3

 (0-1.15; n=18)* silica level during 1989 in Buellton, CA (remote background)(“) 

* mean, range, and number of crystalline silica measurements (Mukherji et al., 1993) 

Appendix D3 533 Silica 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

STYRENE 

(ethenylbenzene, phenylethylene, vinylbenzene) 

CAS Registry Number: 100-42-5 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 900 μg/m3

 (200 ppb) 

Critical effects(s) Neuropsychological deficits in humans as 

measured by memory and sensory/motor 

function tests 

Hazard index target(s) Nervous system 

II. Chemical Property Summary 

Description Colorless to slightly yellow liquid with sweet, 

floral odor (HSDB, 1999) 

Molecular formula C8H8

Molecular weight 104.16 

Boiling point 145.2 o

C 

Melting point -31o

C (HSDB, 1999) 

Vapor pressure 10 torr at 31o

C, polymerizes at 82oC and above 

(Weast, 1979) 

Solubility 310 μg/ml (Dean, 1985) 

Conversion factor 4.26 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

The major source of styrene is industrial synthesis in which ethylbenzene is the starting material 

(ATSDR, 1992). The major uses of styrene are in polystyrene manufacturing, the butadienestyrene rubber industry, and in the reinforced plastics industry (RPI) (WHO, 1983). Major nonstyrene contaminants in the butadiene-styrene rubber industry are butadiene, benzene, carbon 

disulfide, and trichloroethylene, whereas the main co-contaminants associated with the RPI are 

glass fibers and acetone (WHO, 1983). Environmental exposures to styrene may result from 

mainstream cigarette smoke (Newhook and Caldwell, 1993) and newly installed carpets 

containing a styrene-butadiene rubber latex adhesive (Hodgson et al., 1993). The Third National 

Health and Nutrition Examination Survey (NHANES) (Ashley et al., 1994) reported a mean 

blood styrene level among ≥ 600 individuals as 0.074 ppb. In 1996, the latest year tabulated, the 

statewide mean outdoor monitored concentration of styrene was less than 0.1 ppb (CARB, 

1999a). The annual statewide industrial emissions of styrene from facilities reporting under the 

Appendix D3 534 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Air Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 

2,365,873 pounds (1999b). 

IV. Effects of Human Exposure 

Chronic exposures to styrene (to be discussed below) result in central nervous system (CNS) and 

peripheral nervous system effects, although the latter are not as pronounced (ATSDR, 1992; 

Rebert and Hall, 1994; Murata et al., 1991). Irritation or discomfort of the upper respiratory tract 

resulting from styrene exposure has not been reported in long-term occupational studies 

(Foureman, 1994). However, sensory irritation and neurological impairment does occur in acute 

human studies at concentrations above 100 ppm (Stewart et al., 1968). The evidence for styrene 

induced hepatic changes is either negative or equivocal (ATSDR, 1992). Evidence for 

nephrotoxicity due to long-term occupational exposure is also negative or equivocal (ATSDR, 

1992; Verplanke and Herber, 1998; Kolstad et al.,1995). Some human studies suggest that 

chronic exposure to styrene results in reproductive effects, but the limited data are difficult to 

interpret because of the small sample numbers (Brown, 1991; Lindbohm, 1993). Immunologic 

alterations (e.g., altered phenotypic profiles among lymphocyte subsets, decreased natural killer 

cell activity, and decreased chemotaxis) have also been observed, but the limited data prevent 

quantitative interpretation (Bergamaschi et al., 1995; Governa et al., 1994). 

The CNS depressant effects of acute exposures to high styrene levels are probably mediated by 

the direct effect of the lipophilic, unmetabolized styrene on nerve cell membranes. Long-term 

effects of styrene exposure may result from the action of one or more metabolites of styrene 

(Savolainen, 1977; Mutti et al., 1988). In humans, styrene metabolism is initiated by 

cytochrome P450 (P450)-mediated oxidation of styrene to a reactive metabolite, styrene oxide. 

The reaction takes place in human liver and, to a minor extent, in lung (Nakajima et al., 1994). 

The P450 enzymes responsible for the epoxidation of styrene to stryene oxide are also found in 

human brain, but the brain isozymes have not been tested specifically with styrene as a substrate 

(Bhamre et al., 1993). Styrene may also be oxidized to styrene oxide by enzymes which share 

specific iron and porphyrin components with P450 and those that utilize active oxygen species 

(Belvedere et al., 1983; Tursi et al., 1983; Miller et al., 1992). 

The major end product of styrene metabolism in humans is urinary mandelic acid (MA) and 

phenylglyoxylic acid (PGA) (Bardodej and Bardodejova, 1970; Leibman, 1975; Guillemin and 

Bauer, 1979). Other pathways that may be present in other animals are either absent or are 

quantitatively negligible in humans, except when high styrene levels are encountered (Guillemin 

and Berode, 1979; Chakrabarti et al., 1993; Hallier et al., 1995). Confounders of the quantitative 

relationship between styrene exposure and urinary MA+PGA are the consumption of ethanol 

(Berode et al., 1986) and exposure to ethylbenzene (Bardodej and Bardodejova, 1970). An 

important consequence of ethanol related decreased levels of urinary mandelic acid is the 

potential underestimation of exposure to styrene (Guillemin and Bauer, 1979; Berode et al., 

1986). However, the urinary metabolite levels return to control values 4-5 hours after the 

ethanol consumption (Berode et al., 1986). 

Indicators of human styrene exposure include exhaled styrene, blood styrene, urinary MA, and 

urinary MA+PGA (Guillemin and Berode, 1988). Exposure to styrene by inhalation results in 89 

percent absorption (Guillemin and Berode, 1988). In the occupational studies that are the basis 

Appendix D3 535 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

for quantifying the relationship between chronic styrene exposure and health effects, end-of-shift 

or next-morning MA+PGA have been used. The next-morning measurements are more reflective 

of past exposures due to the high fat solubility of styrene (fat:blood partition coefficient = 94 

(Csanady et al., 1994)), the presence of a second, long biological half-life for MA = 25 hours., 

and a long biological half-life for PGA = 11 hours (Guillemin and Bauer, 1979). Following 

inhalation, the half-life for styrene is 41 minutes in blood (Wigaeus et al. 1983) and 32-46 hours 

in fat tissue (Perbellini et al., 1988). 

One postulated mechanism for the chronic non-cancer toxicity of styrene is the binding of the 

highly reactive styrene oxide to components of nervous tissue. Another postulated mechanism is 

an alteration in the levels of circulating catecholamines (e.g., dopamine) due to the binding of 

PGA to these biogenic amines (Mutti, 1993; Mutti et al., 1984a; Checkoway, 1994) and the 

subsequent changes in physiological functions that are under biogenic amine control. Although 

long-term exposures to styrene are associated with decrements in physiological functions, the 

exact mechanism(s) for these effects have not been clearly established (see reviews by ATSDR, 

1992; Mutti, 1993; Rebert and Hall, 1994). 

Kolstad et al. (1995) estimated excess deaths due to four major non-malignant disease groups for 

53,847 male workers in the Danish RPI. Low and high styrene exposures were based on 

companies with less than 50% (low) and those with 50% or more (high) employees involved 

with reinforced plastics. An internal comparison was made with workers unexposed to styrene 

to account for more similar activities and lifestyles. Statistically significant (p < 0.05) excess 

deaths due to pancreatitis and degenerative disorders of the myocardium and non-significant 

excess deaths due to degenerative diseases of the nervous system were observed. Nonsignificant excess deaths due to glomerulonephritis were also observed. 

Checkoway et al. (1994) described a cross-sectional study of 59 male boat plant workers 

exposed to <1 to 144 (mean = 37.2) ppm styrene. Monoamine oxidase B (MAO-B) activity in 

platelets was measured as an indicator of catecholamine metabolism. When the styrene exposed 

workers were divided into quartile exposures, a dose dependent decrease in MAO-B activity was 

observed after adjustments were made for age, smoking, alcohol and medication use. 

Female workers employed in the reinforced plastics industry (RPI) were studied for levels of 

substances associated with neuroendocrine function (Mutti et al., 1984a). Serum prolactin, 

thyroid stimulating hormone, human growth hormone, follicle stimulating hormone, and 

luteinizing hormone were measured in 30 women who were between the 5th and 15th day of the 

menstrual cycle. Exposure was based on the next-morning MA+PGA, and levels of the 

neuroendocrine substances were measured in venous blood samples taken the next morning 

before the start of work. On the basis of a relationship (not detailed in the report) between 

urinary metabolites and styrene air concentration, the authors estimated that the average styrene 

TWA/8 hr was about 130 ppm. Controls consisted of women factory workers living in the same 

area as the styrene-exposed women, but not knowingly exposed to styrene. After controlling for 

age and exposure time, the increased prolactin and thyroid stimulating hormone levels were 

correlated with the concentration of next-morning urinary MA+PGA, although only the 

increased prolactin levels were statistically significant.Numerous occupational studies have 

noted CNS disturbances in styrene-exposed workers. Decreased manual dexterity, increased 

reaction times, and/or abnormal vestibuloocular reflex (ability to track moving objects) were 

observed by Gotell et al. (1972), Gamberale et al. (1975), Lindstrom et al. (1976), Mackay and 

Kelman (1986), Flodin et al. (1989), Moller et al. (1990), and Cherry and Gautrin (1990) for air 

Appendix D3 536 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

styrene levels of about 12 ppm to more than 100 ppm. However, in each of these studies, there 

were difficulties in quantifying the effect. The difficulties included small sample size, unknown 

exposure duration, lack of concurrent control group, lack of dose-response data, and either 

unknown ethanol consumption or lack of adjustment for ethanol consumption. In the Cherry and 

Gautrin (1990) investigation, however, the authors determined that accounting for ethanol 

consumption did not reduce the correlation between increased reaction time and exposure. 

Decrements in other CNS functions were observed among workers in the well controlled studies 

of Fallas et al. (1992), Chia et al. (1994), and Mutti et al. (1984b). Fallas et al. (1992) studied 60 

male workers (average age = 29.5 years, average air styrene = 24.3 ppm). The styrene-exposed 

population was compared to non-exposed worker controls and matched for age, intellectual level, 

and ethnic origin. The results from a standardized test battery showed decrements in the aiming 

response and 22/60 styrene exposed workers exhibited increased reaction times compared to 7/60 

controls. Acquired color vision loss (dyschromatopsia) was also observed in the styrene-exposed 

workers compared to controls. Chia et al. (1994) also observed decrements in CNS function as 

defined by altered visual retention, audio-digit recognition, and digit recognition. However, a 

dose-response relationship did not exist. These workers also exhibited a statistically 

nonsignificant dose-dependent dyschromatopsia. 

In the most comprehensive occupational study to date on CNS effects of styrene exposure, Mutti 

et al. (1984b) assessed memory and sensory/motor function in a group of 50 male styreneexposed workers (average exposure = 8.6 years) and a control group of 50 manual workers. In 

addition to matching for age, sex, and educational level, a vocabulary test was included to match 

for general intelligence. Eligibility criteria included absence of metabolic, neurologic, or 

psychiatric disorders, limited ethanol intake, and limited cigarette usage. All subjects were 

instructed to avoid intake of alcohol and drugs for two days prior to testing. Styrene exposure 

was assessed from urinary MA+PGA levels the morning after the last workday in the week, 

followed immediately by participation in a battery of 8 neuropsychological tests designed to 

measure CNS function. The tests included reaction time, short and long term logic memory, 

short and long term verbal memory, digit-symbol association (using a reference code), block 

design (reproducing a displayed design using colored blocks), and embedded figures (timed 

identification of figures in Rey’s table). The mean + 2 SDs of the values found in the control 

group was set as the normal range limit for each neuropsychological test. The results were 

expressed as continuous and quantal data. Expressed as continuous data, styrene-exposed 

workers exhibited significantly poorer performances than controls in all tests, except in the digitsymbol test. Also, urinary metabolite concentration and duration of exposure were found to be 

significantly correlated with the scores of several tests. As a subgroup, workers with metabolite 

levels of up to 150 mmoles MA+PGA/mole creatinine (mean = 75 mmoles/mole creatinine + 33 

[SD], which is equivalent to a mean styrene concentration of 15 ppm) appeared to have no 

significant effects. The authors state that this level of urinary metabolites corresponds to a mean 

daily 8-hour exposure to air styrene of 25 ppm (106 mg/m3

). Based on greater urinary excretion 

of styrene metabolites, significantly poorer performances in four or more neuropsychological 

tests were recorded in the other three subgroups (150-299, 300-450, and > 450 mmoles MA + 

PGA/mole creatinine). 

. 

Appendix D3 537 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Mutti et al. (1984b) expressed the quantal data as the fraction of tested subjects who responded 

abnormally to ≥ 1, ≥ 2, and ≥ 3 tests (see Table 1). Positive dose-response relationships existed 

between intensity of styrene exposure (mmoles MA + PGA/mole creatinine) and abnormal 

scores, whether it was expressed as abnormal responses in at least one, at least two, or at least 

three neuropsychological tests. The chi-square test and validity calculations were performed by 

constructing 2 x 2 tables selecting different levels of urinary excretion of MA and PGA as a cutoff point. The highest values for chi-square and predictive validity were found when the cut-off 

of 150 mmol/mol creatinine was chosen, suggesting that the quantal isolation of the low dose 

subgroup from the next subgroup is appropriate. When the quantal data for the low dose 

subgroup were analyzed by OEHHA using the Fisher’s Exact Test, a significant level of 

abnormal responses were observed for >1 (p = 0.005) and >3 (p = 0.04) tests. The abnormal 

responses for >2 tests were statistically marginal (p = 0.06). For each of the remaining exposure 

groups, the p-values were <0.05. Unlike the assumptions made concerning the continuous data, 

quantal data results suggest that the low dose subgroup represents a LOAEL, and that a NOAEL 

is not available from the data. Mutti et al. (1984b) also expressed the data in a quantal three-way 

representation including prevalence (number of respondents for at least one, two or three 

abnormal tests), duration (years at work), and intensity (metabolite level). This representation 

revealed a positive correlation of neuropsychological deficits with duration as well as intensity. 

Table 1. Subjects Classified Positive on Neuropsychological Tests as a Function of Styrene 

Exposure a

. 

MA+PGA, mmoles per mole 

creatinineb

 

Total 

Subjects Number of Abnormal Tests 

> 1 > 2 > 3c

Controls 50 4/50 2/50 0/50 

< 150 (mean = 75 ± 33)d 14 6/14 3/14 2/14 

150-299 (mean = 216 ± 45) 9 6/9 5/9 3/9 

300 - 450 (mean = 367 ± 49) 14 10/14 7/14 5/14 

> 450 (mean = 571 ± 108) 13 11/13 8/13 6/13 

a

 Data from Table IV in Mutti et al. (1984b). 

b

 “Next-morning” styrene urinary metabolites. 

c

 The quantal grouping of the number of subjects that performed abnormally in >3 tests based 

on their styrene urinary metabolite concentrations, both shown in bold, were used in a 

benchmark concentration (BMC) analysis for the derivation of the REL (see Section VI 

below). 

d

 Based on Guillemin et al. (1982), a linear relationship exists for converting the urinary 

metabolite concentrations to ppm air styrene levels (4.97 mmoles MA+PGA/mole creatinine is 

equivalent to 1 ppm styrene). Thus, the mean styrene concentrations per group are 0, 15, 44, 74, 

Appendix D3 538 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

and 115 ppm.In addition to dyschromatopsia observed by Chia et al. (1994), Gobba and 

Cavalleri (1993) and Campagna et al. (1995) also reported this visual dysfunction among styrene 

workers in the reinforced plastics industry. Workers (n=36) exposed to an average of 16 ppm 

styrene exhibited significantly greater dyschromatopsia than controls, matched for age, ethanol 

consumption and tobacco smoking (Gobba and Cavalleri, 1993). Among the study population, 

only 1/36 styrene-exposed workers (compared to 16/36 controls) performed the test with 100 

percent accuracy. When a different group of styrene-exposed workers was tested, those exposed 

to > 50 ppm styrene exhibited greater dyschromatopsia than those exposed to ≤ 50 ppm, and 

within this group, a subset exhibited a similar decrement after returning from a one month 

vacation. In the Campagna et al. (1995) study, the test for dyschromatopsia was given to 81 

reinforced plastics industry workers (79 male and 2 female) exposed to 4.6, 10.1, and 88.8 ppm 

styrene (first quartile, median, and third quartile, respectively). No control group was used in 

this study. Statistical analysis revealed a correlation of color vision loss with exposure to styrene 

(defined as next-morning urinary mandelic acid), age, and ethanol consumption. 

Exposure to styrene may affect the peripheral nervous system (PNS). In a case report (Behari et 

al., 1986), a man working for 5 years with a photostat process that used styrene was diagnosed 

with peripheral neuropathy. However, in occupational studies, the relationship between 

exposure to styrene and PNS effects has been inconsistent (Triebig et al., 1985; Cherry and 

Gautrin, 1990). A major difficulty in understanding the potential for this relationship is the lack 

of knowledge about the appropriate surrogate for dose that leads to PNS disturbance (Murata et 

al., 1991). In one study, however, chronic exposure indices were developed which included 

work method, years at work, time spent laminating (source of high exposure), styrene air 

concentration, and end-of-shift urinary mandelic acid (Matikainen et al. (1993). Numbness in 

the extremities increased with the exposure index, although statistically the effect was marginally 

insignificant (p < 0.1). The styrene TWA/8 hr was 32 ppm for the 100 study subjects. 

Female reproductive toxicity has been inconsistently reported among humans (Brown, 1991; 

Lindbohm, 1993). These studies are difficult to interpret because of the high background rates of 

endpoints such as spontaneous abortion and menstrual disorders in combination with 

confounding exposures. In those studies that showed no reproductive effects due to styrene 

exposure, the power of the studies was low due to the small numbers of women. Hence the 

evidence for any adverse effects of exposure to styrene on female reproductive function is 

inconclusive. 

Male workers employed in the reinforced plastics industry were examined for effects on sperm 

chromatin structure and semen quality (Kolstad et al., 1999a) and time to pregnancy (Kolstad et 

al., 1999b). No indications of an exposure-response relationship were seen when individual 

changes in semen quality were related to the postshift urinary mandelic acid concentrations 

among 23 exposed workers. A weak increase in sperm DNA-susceptibility to in situ

denaturation as a function of mandelic acid concentration was indicated, but was within the 

interassay variability. No detrimental effect of styrene exposure was observed with regard to 

male fecundity among 188 exposed workers when compared to 353 unexposed workers. 

Immune system alterations were reported in a study conducted by Bergamaschi et al. (1995). 

Reinforced plastics industry workers (n=32 female/39 male, average age = 32 years, average 

exposure duration = 7 years) were compared with non-styrene exposed factory workers and 

matched for age, sex, tobacco use and ethanol consumption. Air styrene levels, among the 

different factories, varied between 10 - 50 ppm, and individual worker exposure was measured 

Appendix D3 539 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

by urinary metabolites the morning after the last shift (15 hours post-exposure). Among all 

workers in the study (median exposure = 16 ppm - according to the data of Guillemin et al.

(1982)), the proportion of 12/18 lymphocyte subsets and the prevalence of abnormal values of 

immunologic phenotypes for 11/18 subsets were statistically different from the controls (p < 

0.001 to < 0.05). When the workers were placed into three exposure groups (0, < 25 ppm, and > 

25 ppm styrene), dose-response relationships were observed for prevalences of abnormal 

responses for four lymphocyte subsets and, in the case of two subsets, abnormal responses were 

observed in the group exposed to < 25 ppm styrene. Natural killer cell activity (a lymphocyte 

function), measured in a different group of workers in the same study, was decreased compared 

to unexposed worker controls. The median exposure, given in terms of urinary metabolites, was 

calculated as 21 ppm based on the data of Guillemin et al. (1982). The data show that exposure 

of these workers to air styrene levels below 50 ppm, and probably at levels near 25 ppm, resulted 

in alterations of the immune system. 

Governa et al. (1994) observed reduced chemotactic responses of polymorphonuclear 

lymphocytes (PMNs) obtained from 21 styrene-exposed workers. However, the lack of exposure 

data prevents a quantitative assessment. In the same study, 0.1 - 0.6 mM styrene inhibited the 

chemotaxis of isolated healthy PMNs. 

V. Effects of Animal Exposure 

In a subchronic study, carried out under the auspices of NTP (NTP, 1992), mice and rats were 

exposed by inhalation to styrene vapors to establish a maximum tolerated dose for chronic 

studies. Mice were exposed to 0, 62.5, 125, 250, or 500 ppm styrene (6 hr/d, 5 d/wk, 13 wks). 

Among males deaths occurred in the 250 ppm group. Body weights among all exposed mice 

were lower than controls, and the difference was about 9 percent. Lung, olfactory epithelial, and 

forestomach lesions were observed in females and males. In females, degeneration of the 

adrenal gland cortex was observed. An effect not discussed in the chairperson’s report, but 

recorded in the original laboratory report, was an increased estrous cycle length among the 

female mice at all styrene doses. A LOAEL of 62.5 ppm is indicated by the olfactory epithelial, 

forestomach and respiratory tract lesions in mice of both sexes and for lesions in the adrenal 

cortex in the female mice. Rats were exposed to 0, 125, 250, 500, 1000, or 1500 ppm styrene (6 

hr/d, 5 d/wk, 13 wks). No deaths occurred, but reduced body weights were observed at the two 

highest doses. Lesions of the respiratory tract were observed at all dose levels. A LOAEL of 

125 ppm is therefore indicated for the rats. 

Rats were exposed by ingestion for 2-years to styrene in drinking water (0, 125, and 250 ppm). 

(The water solubility of styrene is 310 ppm.) The only effect was a styrene-related reduction in 

water consumption (Beliles et al., 1985). 

Kishi et al. (1995) carried out a developmental study on rat pups born to dams exposed by 

inhalation to styrene (0, 50, 300 ppm; 6-hr/d; gestation days 7-21). Although the small number 

of litters (n=2) at the 50 ppm dose prevented detailed statistical analysis, the data suggest that 

exposure of the dams to 50 ppm styrene resulted in deficits and delays in some motor and 

coordination abilities among the pups. Pups born to dams exposed to 300 ppm exhibited 

statistically significant increases in spontaneous activity and in the delay of some 

neurobehavioral functions. Many of the effects became diminished as the pups aged. 

Measurements of reproductive toxicity (maternal weight gain, length of gestation, number of live 

Appendix D3 540 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

births) did not change. Postnatal body weights were lower among the styrene-exposed pups, but 

the differences became less as the pups aged to 125-days. 

A follow-up developmental study by the same research group investigated neurochemical levels 

in rat pups born to dams exposed by inhalation to styrene (0, 50, 300 ppm; 6 hr/day on gestation 

days 6-20) (Katakura et al., 1999). Cerebrum weights of day 0 pups were significantly lower 

when compared to cerebrum weights of ad libitum fed animals, but not pair-fed animals. At the 

highest dose, occasional reductions in neuroamines, i.e. 5-hydroxytryptamine, homovanillic acid, 

and 5-hydroxyindoleacetic acid, were seen in various parts of the brains of rat pups compared to 

one or both control groups on day 0 and day 21. No reproductive or histopathological changes 

were seen. 

Rosengren and Haglid (1989) investigated whether long term inhalation exposure (three months) 

to styrene (90 and 320 ppm) could induce long lasting astroglial alterations in Sprague Dawley 

rats, traceable four months after exposure ceased. Styrene exposure at 320 ppm induced the 

alterations as shown by raised concentrations of the glial cell marker, glial fibrillary acidic 

protein (GFA), in the sensory motor cortex and in the hippocampus. GFA is the structural 

protein of the astroglial filaments. These filaments form after damage to the central nervous 

system from any cause. The authors concluded that exposure to styrene at moderate exposure 

levels induces regional, long lasting astroglial reactions that serve as an indicator of solvent 

induced brain damage. 

Mice, exposed acutely (14 days) by inhalation to 125 - 500 ppm styrene, exhibited decreased 

spleen / body weight, splenic hypocellularity, altered lymphocyte proportions among subsets, 

and increased proliferative response to mitogens (Corsini et al., 1994). Mice and rats, exposed 

by gavage to high levels of styrene (18, 27, 45 mg/kg - mouse; 118, 177, 294 mg/kg - rat) for 5 

days/week for 4 weeks, exhibited decreased resistance to encephalomyocarditis virus, 

Plasmodium berghie (a malaria parasite), and Nippostrongylus braseleinisi (a parasitic worm) 

(Dogra et al., 1992). 

Groups of 70 male and 70 female Charles River CD (Sprague-Dawley-derived) rats were 

exposed whole body to styrene vapor at 0, 50, 200, 500, or 1000 ppm 6 h/day 5 days/week for 

104 weeks (Cruzan et al., 1998). A battery of hematologic and clinical pathology examinations 

was conducted at 13, 26, 52, 78, and 104 weeks. Nine or 10 rats per sex per group were 

necropsied after 52 weeks of exposure and the remaining survivors were necropsied after 104 

weeks. Control and high-exposure rats received a complete histopathologic examination, while 

target organs, gross lesions, and all masses were examined in the other 3 groups. Styrene had no 

effect on survival in males, but females exposed to 500 or 1000 ppm had a dose-related 

increasein survival. Levels of styrene in the blood at the end of a 6-h exposure during week 95 

were proportional to exposure. Levels of styrene oxide in the blood of rats exposed to 200 ppm 

or greater styrene were proportional to styrene exposure concentration. The authors found no 

changes of toxicologic significance in hematology, clinical chemistry, urinalysis, or organ 

weights. Styrene-related non-neoplastic histopathologic changes were confined to the olfactory 

epithelium of the nasal mucosa. (The authors also found no evidence of cancer induction.) 

Groups of 70 male and 70 female CD-1 mice were exposed in whole body inhalation chambers 

to styene vapor concentrations of 0, 20, 40, 80, and 160 ppm 6 hrs/day, 5 days/week, over a 

period of up to 2 years (Huntingdon Life Sciences, 1998). Ten mice per sex per group were 

Appendix D3 541 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

necropsied after 52 and 78 weeks of exposure, and the remaining survivors necropsied after 104 

weeks. Due to increased mortality in female control mice, terminal sacrifice for this group 

occurred at 98 weeks. Two female mice exposed to 160 ppm styrene died during or immediately 

following the first week of exposure. Histopathology revealed liver necrosis that was a likely 

contributor to the deaths. Reduced body weight gain and increased food consumption were 

observed in male mice at the two highest exposure levels and in female mice at the highest 

exposure level. Both styrene monomer and styrene oxide in blood increased with exposure 

concentration. No changes of toxicologic significance in hematology, ophthalmology, clinical 

chemistry, urinalysis, or organ weights were noted. Styrene-related non-neoplastic 

histopathologic changes were seen in the lungs (bronchiolar-alveolar hyperplasia) and nasal 

olfactory epithelium (respiratory metaplasia, degeneration or necrosis, and changes to the 

underlying Bowman’s glands) from all exposure groups. The nasal lesions showed progression 

with time. Focal loss of bone from the turbinate was also seen more frequently as the study 

progressed. In addition, atrophy of the olfactory nerve fibers was present in mice at the three 

highest exposure concentrations. 

VI. Derivation of the Chronic Reference Exposure Level (BMC Approach)

. 

Study Mutti et al. (1984b) 

Study populations Human (occupational) 

Exposure method Inhalation 

Critical effects Central nervous system 

LOAEL 15 ppm 

NOAEL 

BMC05

Not established 

1.7 ppm 

Exposure continuity 8 hr/d (10 m3

 per 20 m3 day), 5 d/wk 

Exposure duration 8.6 years (average years at work) 

Average occupational exposure 0.61 ppm (1.7 x 10/20 x 5/7) 

Human equivalent concentration 0.61 ppm 

LOAEL uncertainty factor Not needed in the BMC approach 

Subchronic uncertainty factor 1 (average exposure 12.3% of lifetime) 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 3 

Cumulative uncertainty factor 3 

Inhalation reference exposure level 0.2 ppm (200 ppb; 0.9 mg/m3

; 900 μg/m3) 

The most relevant chronic noncancer effect due to styrene exposure is neurotoxicity. The Mutti 

et al. (1984b) occupational study presented convincing dose-response information and was well 

designed and executed in terms of experimental protocol and statistical evaluation, which 

included tests for false positive and false negative responses. While not all confounders could be 

ruled out (e.g., compensatory mechanisms, biorhythms, workers who leave because of styrene 

related illness), careful attention was paid to include eligibility criteria for the control group that 

correct for confounders unique for this population (e.g., limited ethanol intake, a control workforce not exposed to neurotoxic substances, and a test to allow a match for general intelligence). 

The use of urinary metabolites to measure exposure dose is based on the observation that the 

Appendix D3 542 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

next-morning urinary MA+PGA is directly related to the air level of styrene. The Guillemin et 

al. (1982) study provides the basis for the conversion of urinary MA+PGA levels to styrene 

exposure levels used by Mutti et al. (1984b). 

The quantal dose-response data by Mutti et al. (1984b) is applicable for use in a benchmark 

concentration (BMC) approach. The quantal grouping of the number of subjects that performed 

abnormally in >3 tests based on their urinary metabolite concentrations was chosen for a BMC 

analysis (see Table 1). Basing the BMC on abnormal responses to >3 tests reduces the 

complexity of multiple test comparisons and the potential for inappropriate comparison of 

different neuropsychological tests between control and exposure groups for statistical purposes. 

Also, the potential for false positive responses is reduced due to the zero background level of 

abnormal responses in the control group when the criteria are >3 abnormal tests. Using a lognormal probit analysis (Tox-Risk, version 3.5; ICF-Kaiser Inc., Ruston, LA) with the data 

(emphasized in bold typeface) in Table 1 (above) the maximum likelihood estimate (MLE) for a 

5% response was 4.0 ppm. The resulting 95% lower confidence limit at the MLE provided a 

BMC05 of 1.7 ppm. A BMC05 is considered to be similar to a NOAEL in estimating a 

concentration associated with a low level of risk. Following adjustment for exposure continuity 

(10 m3

 per 20 m3 day for 5 d/wk) and application of an UF of 3 to account for human 

intraspecies variability, a REL of 0.2 ppm (0.9 mg/m3

) was attained. For exposure data that 

utilizes healthy human subjects, the resulting BMC represents a less than 10% incidence in the 

general population. When combined with an UF of 3, as carried out above, the resulting REL 

will be protective of the vast majority of individuals. 

This analysis of the quantal data is supported by recognizing that, in a population of 50 subjects, 

individual test-specific effects that occur at low doses may not have been observed. If the 

criterion for abnormality is expressed in terms of CNS dysfunction, defined by all tests, the 

sensitivity of the testing procedure is increased and the low dose effects are more easily 

observed. The quantal data of Mutti et al. (1984b), i.e., the proportion of subjects responding 

abnormally to the tests, therefore provide a more sensitive approach to detecting low dose 

effects. Collasping a battery of test data to increase sensitivity may introduce the dilemma of 

multiple test comparisons, as noted above. However, OEHHA believes that a statistical method 

to correct for this, known as a Bonferroni correction, is unnecessary. The REL development is 

based on calculating a statistic of one effect of a complex of responses (or a syndrome) that 

results from CNS dysfunction, and not based on calculating a statistic for each test within the 

group of tests. The apparent global nature of the neurological syndrome resulting from longterm styrene exposure, in addition to basing the BMC on abnormal responses to >3 tests, should 

more than adequately address any concerns that may result from combining neurological test 

data. 

Applying NOAEL/LOAEL methodology to the Mutti et al. (1984b) quantal data yields an 

exposure value similar to that attained with the BMC approach. The LOAEL of 15 ppm is 

adjusted to an equivalent continuous exposure of 5.36 ppm (15 ppm x 10/20 m3

 x 5/7 d/wk). Use 

of a LOAEL UF of 3 and an intraspecies UF of 10 resulted in an estimated REL of 0.2 ppm (0.8 

mg/m3

). 

The U.S. EPA (1996) calculated a reference concentration (RfC) of 0.3 ppm (1 mg/m3

), which is 

slightly higher than the OEHHA-derived chronic REL of 0.2 ppm (0.9 mg/m3

). The RfC for 

styrene is also based on the findings of Mutti et al. (1984b), but utilized the continuous data for 

Appendix D3 543 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

development of the RfC and used standard NOAEL methodology for the RfC derivation. U.S. 

EPA (1996) established a NOAEL for the lowest exposure group (<150 MA+PGA mmole/mole 

creatinine; equivalent to < 25 ppm styrene). However, OEHHA staff believe that the use of the 

continuous data to establish a NOAEL overlooks the advantages of using the BMC approach 

using the quantal data. These advantages are that the BMC05 reflects the shape of the doseresponse curve and takes into account the number of subjects involved in the study. In addition, 

OEHHA staff evaluated the quantal data with the Fisher’s Exact Test and determined the 

probabilities of abnormal responses among the exposed subjects based on the unexposed subjects 

whose responses were assumed to be normal. At the lowest exposure, the probability that the 

proportion of subjects responding abnormally to >1 and >3 tests was within the expected range 

was p = 0.005 and p = 0.04, respectively, indicating that neuropsychological deficits due to 

styrene occur in the low dose subgroup. Thus, the quantal data indicate that a NOAEL was not 

established in this study. 

With regard to application of uncertainty factors, U.S. EPA (1996) applied a UF of 3 for 

intraspecies variability and a partial UF of 3 for lack of information on chronic studies because 

the critical study was considered intermediate, i.e., between subchronic and chronic duration 

(Foureman, 1994). OEHHA applied a UF of 1 because the mean exposure duration, 8.6 years, 

was greater than 12 percent of expected lifetime (8.6/70 = 12.3%). The U.S. EPA (1996) also 

included a modifying factor of 3 for database deficiencies. The criteria for use of modifying 

factors are not well specified by U.S. EPA. Such modifying factors were not used by OEHHA. 

In addition to the OEHHA and the U.S. EPA hazard assessments, the Agency for Toxic 

Substances and Disease Registry (ATSDR) also calculated a chronic inhalation minimal risk 

level (MRL) for styrene (ATSDR, 1992). The calculation was based on the same Mutti et al.

(1984b) worker study. ATSDR (1992) identified the lowest exposure group as a LOAEL and 

assigned an air styrene level of 25 ppm. To derive the MRL, ATSDR corrected the LOAEL for 

discontinuous exposure and applied uncertainty factors (UFs) for the use of a LOAEL and for 

intraspecies variability. The MRL was calculated as: 25 x (8/24 x 5/7) / 10 x 10 equal 0.06 ppm 

(ATSDR, 1992). The MRL was a factor of 3 different from the proposed REL. 

For comparison, chronic exposure levels for styrene can be developed from chronic inhalation 

studies in rats (Cruzan et al., 1998) and mice (Huntingdon Life Sciences, 1998). The mice were 

more sensitive to the styrene vapors than were rats, and a LOAEL of 20 ppm was identified 

based on lesions in various organs in both sexes. The adjustment factor for discontinuous 

exposure is (6/24 x 5/7) = 0.18. The uncertainty factors are: 10 for intraspecies variability, 3 for 

interspecies sensitivity, and 10 for adjustment from a LOAEL to a NOAEL. The resultant 

exposure level is (20 ppm x 0.18) / 300 which equals 0.01 ppm or 10 ppb (40 μg/m3

). Besides 

the different toxic endpoints between the chronic mouse exposure study and human occupational 

studies, the well designed human study of Mutti et al. (1984b) is preferable for REL 

development because it does not introduce the uncertainties associated with interspecies 

extrapolations. 

The NOAEL of 50 ppm from the chronic rat study of Cruzan et al. (1998) may be adjusted to an 

equivalent continuous exposure of 8.9 ppm. Use of an RGDR of 1, an interspecies UF of 3, and 

an intraspecies UF of 10 resulted in an estimated REL of 300 ppb (1300 μg/m3

) for styrene. 

Appendix D3 544 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the REL for styrene include the excellent database available on styrene effects 

and the availability of a suitable human study for use as the key study. Limitations include the 

lack of direct exposure data and selection bias. Although a NOAEL was not observed in the key 

study, the BMC05 is considered to be similar to a NOAEL in estimating a concentration 

associated with a low level of risk. 

Use of urinary metabolite concentrations to indirectly determine styrene exposure, while an 

accepted approach, still introduces another level of uncertainty in the hazard assessment. In 

addition, potential absorption of styrene via dermal exposure in the reinforced plastics industry 

has not been addressed and may overestimate the air concentration determined by urinary 

metabolite levels. However, unlike air levels, the presence of urinary metabolites of styrene 

gives an unequivocal indication that an individual has been exposed to styrene. At the present 

time, a system does not exist to obtain direct exposure information, although a recent report 

suggests a methodology is being developed (Jensen et al., 1995). 

A potential bias in the key study was the finding that general intelligence, as measured by the 

vocabulary test, appeared to be negatively correlated with both age and exposure intensity. This 

finding suggests that age may also be a factor in poor neuropsychological test scores of highly 

exposed subgroups. Another source of uncertainty is that the reinforced plastics industry, from 

which the workers in the Mutti et al. (1984) study were taken, is characterized by a large 

turnover of highly exposed workers (Wong, 1990; Kogevinas et al., 1993). This possible 

selection bias may result in more sensitive workers leaving employment while more tolerant 

workers remain. 

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styrene. Scand. J. Work Environ. Health 16(3):189-194. 

Murata K, Araki S, and Yokoyama K. 1991. Assessment of the peripheral, central, and 

autonomic nervous system function in styrene workers. Am. J. Ind. Med. 20:775-784. 

Mutti A. 1993. Mechanisms and biomarkers of solvent-induced behavioral and neuroendocrine 

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Pollutants. Travis CC, ed., New York: Plenum Press. pp 183-199. 

Appendix D3 548 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Mutti A, Falzoi M, Romanelli MC, Bocchi MC, Ferroni C, and Frandhini I. 1988. Brain 

dopamine as a target for solvent toxicity: effects of some monocyclic aromatic hydrocarbons. 

Toxicology 49:77-82. 

Mutti A, Mazzucchi A, Rustichelli P, Frigeri G, Arfini G, and Franchini I. 1984b. Exposureeffect and exposure-response relationships between occupational exposure to styrene and 

neurophychological functions. Am. J. Ind. Med. 5:275-286. 

Mutti A, Vescovi PP, Falzoi M, Arfini G, Valenti G, and Franchini I. 1984a. Neuroendocrine 

effects of styrene on occupationally exposed workers. Scand. J. Work Environ. Health 10:225-

228. 

Nakajima T, Elovaara E, Gonzalez FJ, Gelboin HV, Raunio H, Pilkonen O, Vainio H, and 

Aoyama T. 1994. Styrene metabolism by cDNA-expressed human hepatic and pulmonary 

cytochromes P450. Chem. Res. Toxicol. 7:891-896. 

Newhook R, and Caldwell I. 1993. Exposure to styrene in the general Canadian population. 

IARC Scientific Publ. 127:27-33. 

NTP. 1992. Pathology working group’s chairperson’s report. 13-Week Toxicity Study of Styrene 

(CO2200B) in F344 Rats and B6C3F1 Mice by Inhalation. Report dated 28 April 1992. 

Perbellini L, Mozzo P, Turri PV, Zedde A, and Brugnone F. 1988. Biological exposure index of 

styrene suggested by a physiologico-mathematical model. Int. Arch. Occup. Environ. Health 

60:187-193. 

Rebert CS, and Hall TA. 1994. The neuroepidemiology of styrene: a critical review of 

representative literature. Crit. Rev. Toxicol. 24(Suppl 1):S57-S106. 

Rosengren LE, Haglid KG. 1989. Long term neurotoxicity of styrene. A quantitative study of 

glial fibrillary acidic protein (GFA) and S-100. Br. J. Ind. Med. 46(5):316-320. 

Savolainen H. 1977. Some aspects of the mechanisms by which industrial solvents produce 

neurotoxic effects. Chem.-Biol. Interact. 18:1-10. 

Stewart, RD, Dodd HC, Baretta ED, and Schaffer AW. 1968. Human exposure to styrene vapor. 

Arch. Environ. Health. 16:656-662. 

Treibig G, Schaller K-H, and Valentin H. 1985. Investigations on neurotoxicity of chemical 

substances at the workplace. VII. Longitudinal study with determination of nerve conduction 

velocities in persons occupationally exposed to styrene. Int. Arch. Occup. Environ. Health. 

56:239-247. 

Tursi F, Samaia M, Salmoma M, and Belvedere G. 1983. Styrene oxidation to styrene oxide in 

human erythrocytes catalyzed by oxyhemoglobin. Experientia. 39:593-594. 

U.S. EPA. 1996. Integrated Risk Information System (IRIS). US Environmental Protection 

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CO: Microdex, Inc. (Edition expires 07/31/96). 

Weast RC. 1979. CRC Handbook of Chemistry and Physics. Vol. 60. Weast RC, and Astle MJ, 

eds. Boca Raton, FL: CRC Press. 

Wigaeus E, Lof A, Bjurstrom R, and Nordqvist MB. 1983. Exposure to styrene. Scand J. Work 

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Appendix D3 549 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

WHO. 1983. Environmental Health Criteria 26: Styrene. Finland: World Health Organization. 

Wilson HK, Robertson SM, Waldron HA, and Gompertz D. 1983. Effect of ethanol on the 

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40:75-80. 

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Appendix D3 550 Styrene 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

SULFURIC ACID 

(dithionic acid; pyrosulphuric acid) 

CAS Registry Number: 7664-93-9 

II. I. Chronic Toxicity Summary 

Inhalation reference exposure level 1 μg/m3

Critical effect(s) Bronchiolar epithelial hyperplasia, and thickening 

of the bronchial walls in monkeys 

Hazard index target(s) Respiratory system 

III. Physical and Chemical Properties (HSDB, 1995; CRC, 1994; CARB, 1997) 

Description Colorless liquid 

Molecular formula H2SO4

Molecular weight 98.1 g/mol 

Density 1.84 g/cm3

 @ 15° C 

Boiling point 330±0.5°C (100%) 

Melting point 10.36°C (100%) 

Vapor pressure <0.001 torr @ 25° C; 1 torr @ 145.8° C 

Solubility Soluble in water 

Conversion factor Not applicable 

III. Major Uses or Sources 

Sulfuric acid is a strong acid used as an intermediate in the synthesis of linear alkylbenzene 

sulfonation surfactants used in dyes, in petroleum refining, for the nitration of explosives, in the 

manufacture of nitrocellulose, in caprolactam manufacturing, as the electrolyte in lead-acid 

batteries, and as a drying agent for chlorine and nitric acid. Sulfuric acid is formed in the 

atmosphere from sulfur dioxide, from sulfur trioxide, and from oleum (a combination of sulfur 

trioxide and sulfuric acid used industrially). The annual statewide industrial emissions from 

facilities reporting under the Air Toxics Hot Spots Act in California based on the most recent 

inventory were estimated to be 4460 pounds of sulfuric acid (CARB, 1999). 

IV. Effects of Human Exposures 

Workers in the lead battery industry showed etching and erosion of the teeth after 4 months 

exposure to an average concentration of 0.23 mg/m3

 H2SO4 (Gamble et al., 1984). Dental 

erosion increased in a dose-dependent manner with longer duration of exposure. 

Appendix D3 551 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

A study of 33 storage battery plant workers exposed to H2SO4 concentrations as high as 35 

mg/m3

 showed a greater group mean decrease in FEVl across the time of their work shift 

compared to workers who were not exposed to sulfuric acid (El-Saddik et al., 1972). The 

salivary pH of the sulfuric acid exposed workers, a qualitative measure of acid exposure, was 

lower than the controls during the course of the work shift. 

OEHHA recently reviewed the California Ambient Air Quality Standard (CAAQS) for sulfates 

(25 μg/m3

 for 24 hours) to see if it adequately protects children (OEHHA, 2000). The report was 

peer-reviewed by the Air Quality Advisory Committee. The report indicates that H+ itself may 

play a role in the effects seen in epidemiological studies of sulfate air pollution. Controlled acute 

inhalation studies in humans and laboratory animals of pH neutral or nearly neutral sulfate salts 

(e.g., ammonium sulfate) (Utell et al., 1983; Lippman et al., 1987; Schlesinger et al., 1990), even 

at relatively high concentrations, do not produce the effects reported from epidemiologic studies 

of sulfates (asthma exacerbation, bronchoconstriction, decrements in lung function) that might be 

expected from short-term excursions. The controlled exposure studies show that sulfate aerosols 

containing strong acids, such as sulfuric acid and, to a lesser extent, ammonium bisulfate, 

produce functional and structural changes in healthy subjects consistent with those observed in 

epidemiological studies. A working hypothesis is that H+

 is a causal factor for adverse human 

health effects (e.g., see Lippmann and Thurston, 1996) and that, among the commonly measured 

particulate matter (PM) indices, SO4

=

 is the best surrogate metric for H+. 

A large number of epidemiologic studies have been conducted showing that elevated levels of 

several air pollutants, including acid aerosols, sulfur and nitrogen oxides, and particulate sulfates 

are correlated with an increased prevalence of pulmonary disease (U.S. EPA, 1989; OEHHA 

2000). Elevated sulfate levels (1.6 ppb or 6.6 μg/m3

) have been associated with statistically 

significant decrements in FVC and FEV1 in a cohort of Canadian children (Stern et al., 1989). 

Further analysis of these data led Bates and Sitzo (1989) to conclude that H2SO4 was the most 

likely cause for the pulmonary changes observed. Similarly, Ostro et al. (1989) reported a 

statistical association between asthma-related symptoms reported by 209 asthmatics and sulfate 

and acidity levels in ambient air in Denver. Delfino et al. (1997) found that ambient H+ was 

associated with emergency room visits by children for respiratory symptoms in a study in 

Montreal. Additionally, Damokosh et al. (1993) in a follow-up analysis of the 6-City study 

suggested associations between average H+ concentration and chronic bronchitic symptoms. 

The relative odds of bronchitic symptoms with the highest acid concentration (58 nmoles/m3

H+)_ versus the lowest concentration (16 nmoles/m3

) was 2.4 (95% CI:1.9 to 3.2). Furthermore 

in a study of children in 24 U.S. and Canadian communities (Dockery et al., 1996) in which the 

analysis was adjusted for the effects of gender, age, parental asthma, parental education, and 

parental allergies, bronchitic symptoms were confirmed to be significantly associated with 

strongly acidic PM (OR= 1.66; 95% CI 1.11-2.48). It was also found that FVC and FEV1 were 

lower in locales with high particle acidity (Raizenne et al., 1996). Gwynn et al. (2000) reported 

an association between both H+ and sulfate particles and respiratory hospital admissions and 

mortality in Buffalo, NY. Acidic sulfates may act to increase the toxicity of particles by 

enhancing the availability of metals present in the particles to generate reactive oxygen species in 

the respiratory epithelium. This may account for some of the effects seen in these 

epidemiological studies and makes it difficult to use these studies as a basis for a Reference 

Exposure Level for sulfuric acid. The relationship between the effect levels observed in these 

Appendix D3 552 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

studies and the proposed REL is discussed in the section below on the potential for differential 

impacts on children's health. 

The occupational standard for sulfuric acid is based on a study in human subjects by Amdur et 

al. (1952). In their study, 22 healthy male subjects were exposed to 0, 0.35, 0.4, 0.5, 1, 2, or 5 

mg/m3

 for 5-15 minutes. The odor, taste, and irritation threshold was 1 mg/m3. Since the basis 

for this standard is an acute exposure, it is not useful in determining a chronic non-cancer REL 

for sulfuric acid. A review of chronic human exposures to sulfuric acid and resulting 

carcinogenicity outcomes can be found in IARC (1992). However, none of the studies in that 

review examined non-cancer endpoints. 

Sulfuric acid and oleum (supersaturated anhydrous sulfuric acid with varying concentrations of 

free sulfur trioxide) are absorbed as salts of sulfate anion (SO4

2-), and are excreted as organic 

sulfates, neutral sulfur, or neutral sulfur compounds such as sulfur-containing amino acids. The 

low systemic toxicity of these metabolites is likely of secondary importance to the irritation 

caused by the inhaled acid. 

V. Effects of Animal Exposures

An exposure of 9 cynomolgus monkeys per group to H2SO4 concentrations of 0, 0.38, 0.48, 2.43, 

and 4.79 mg/m3

 continuously for 78 weeks resulted in dose-dependent adverse histological 

changes in lung and bronchiolar epithelial and parenchymal tissue in addition to a dosedependent decrease in blood oxygenation (Alarie et al., 1973). In the animals exposed to 

0.38 mg/m3

, significant bronchiolar epithelial hyperplasia was observed in 5 of 9 animals; 

thickening of the bronchiolar walls was observed in 3 of 9 animals. A slight focal bronchial 

epithelial hyperplasia was present in 4 of the 9 animals. One animal died after 4 weeks exposure 

to 0.38 mg/m3

. Although signs of pulmonary edema and cardiac hypertrophy were found, the 

cause of death was not determined. 

Respiratory system effects of H2SO4 exposure in monkeys (Alarie et al., 1973) 

H2SO4 

(g/m3) 

Particle

size 

MMD 

Bronchiolar epithelial 

hyperplasia 

Incidence – severity 

Thickening of walls of 

respiratory bronchioles 

Incidence – severity 

Increase in thickness of 

alveolar walls 

Incidence – severity 

0 0/9 0/9 0/9 

0.38 2.15 5/8 – slight 3/8 - slight 0/8 

0.48 0.54 0/8 0/8 0/8 

2.43 3.60 8/8 – moderate 8/8 – moderate 8/8 – moderate 

4.79 0.73 8/8 – moderate to severe 8/8 – moderate to severe 0/8 

Alarie et al. (1973) also exposed groups of 50 guinea pigs of each sex to 0, 0.08, or 0.1 mg/m3

H2SO4 continuously for 52 weeks. The group exposed to 0.1 mg/m3

 also received larger sized 

particulates than the 0.08 mg/m3

 group (2.78 μm vs. 0.84 μm, respectively). No exposure related 

effects were observed in the animals exposed to 0.08 mg/m3

, whereas exposure of 0.1 mg/m3

resulted in decreased body weights in the female guinea pigs. No other histological changes in 

any organs were observed at the end of the 52-week study. 

Appendix D3 553 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Rabbits (4 per group) were exposed to 250 μg/m3

 H2SO4 1 hour/day, 5 days/week for 4, 8, or 12 

months. They showed significantly increased bronchoconstriction upon acetylcholine challenge 

after 8 and 12 months exposure, compared with a control group of 4 animals that received no 

H2SO4 (Gearhart and Schlesinger, 1986, 1988). Mucociliary clearance was also impaired by 

H2SO4 exposure and did not improve 3 months after cessation of exposure. A decline in 

dynamic lung compliance was observed after 12 months exposure. There was no evidence of 

inflammatory cell infiltration in the lungs of the exposed animals. 

In guinea pigs, significantly slower and irregular breathing patterns were noted when the animals 

had inhaled albumin followed by 30 minute exposures to H2SO4 at 1.91 mg/m3

 twice per week 

for 5 weeks (Kitabatake et al., 1979). Similarly, when guinea pigs were exposed to 2.49 mg 

H2SO4/m3

 for 4 hours/day, 6 days/week for 4 weeks, in vitro lung histamine release was 

significantly enhanced following heterogeneous albumin inhalation, compared to control animals 

unexposed to albumin (Fujisawa et al., 1986; Iguchi et al., 1986). In guinea pigs, sulfuric acid 

caused significantly greater lung function changes when adsorbed on the surface of zinc oxide 

particles as compared with pure sulfuric acid (Amdur and Chen, 1989). An exposure to 

24 μg/m3

 sulfuric acid, layered on zinc oxide, produced significant reductions in lung function 

when followed by a brief exposure to 0.15 ppm ozone (Chen et al., 1991). 

A chronic exposure of beagle dogs to an average concentration of 889 μg/m3

 H2SO4 for 21 

hours/day over a 620 day period resulted in increased expiratory resistance, reduced carbon 

monoxide diffusing capacity, reduced total and residual lung volume, and decreased lung and 

heart weights (Lewis et al., 1973). 

In apparent contrast to the above studies, rats and guinea pigs exposed to H2SO4 at 10 mg/m3

 for 

6 hours/day, 5 days/week for 6 months exhibited no adverse histologic changes in lung tissue. 

Lung function measurements were not reported in this study (Cavender et al., 1978). 

Mice inhaled sulfuric acid mist at a concentration of 1.4 mg/m3

 in combination with a carbon 

particle mixture (1.5 mg/m3

) for 3 hours/day, 5 days/week for up to 20 weeks. The exposure 

resulted in significant alterations in specific antibody titer (decreased IgG, Ig2a, IgM; increased 

IgG2b), depression of primary splenic antibody response, and decreased resistance to respiratory 

infection as measured by mortality and survival time compared to controls (Fenters et al., 1979). 

There are no reliable studies indicating that sulfuric acid is a developmental or reproductive 

toxicant. In the absence of massive overexposure leading to maternal acidemia, H2SO4 will be 

neutralized in the maternal circulation and is unlikely to reach the fetus. 

Appendix D3 554 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VI. Derivation of Chronic Reference Exposure Level 

Study Alarie et al., 1973 

Study population Cynomolgus monkeys (5 males and 4 females per 

group or vice versa) 

Exposure method Continuous inhalation exposures (0, 380, 480, 

2400, or 4800 μg/m3

) for 78 weeks 

Critical effects Significantly increased bronchial epithelial 

hyperplasia and bronchial thickening 

LOAEL 380 μg/m3

NOAEL Not observed 

Exposure continuity The exposure was continuous during the 

experiment. 

Exposure duration 78 weeks 

Average experimental exposure 380 μg/m3

 for the LOAEL group 

Human equivalent concentration 380 μg/m3

 

LOAEL uncertainty factor 3 (slight effects) 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 3 (non-human primate) 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Reference exposure level 1 μg/m3

The study by Alarie et al. (1973) identified a LOAEL for chronic exposure to sulfuric acid of 

380 μg/m3

. The principal uncertainties of this study are the small sample size of the test groups 

and the absence of an observed NOAEL. A lower chronic LOAEL for bronchial reactivity is 

presented by Gearhart and Schlesinger (1986, 1988) for rabbits (250 μg/m3

). This study was not 

selected as the basis of the REL because Gearhart and Schlesinger used only a single 

concentration of sulfuric acid, exposed the animals only for 1 hour per day for 5 days/week, used 

only 4 animals per group, and measured effects over the course of up to 12 months. The 

predominant weakness in the rabbit study, however, was the extreme discontinuity of the 

exposures (1 hour/day, 5 days/week), which would have necessitated use of a very large 

continuity adjustment. For these reasons, in addition to obvious physiological and genetic 

similarity arguments, the study in monkeys by Alarie et al. (1973) was felt to be more 

appropriate as the basis for the chronic REL for sulfuric acid. Alarie et al. (1975) determined a 

NOAEL for sulfuric acid in monkeys of 0.1 mg/m3

. However, other particulate matter (fly ash) 

was also present during the exposure. The Alarie et al. (1973) report provides data from 

exposure to sulfuric acid alone. 

A free-standing NOAEL for histological changes in 100 guinea pigs exposed continuously for 1 

year to 0.08 mg/m3 was reported by Alarie et al. (1973). Guinea pigs respond to high 

concentrations of sulfuric acid by occasional laryngeal spasms that appear similar to a human 

asthmatic attack (Silbaugh et al., 1981; Amdur and Chen, 1989). As a result, guinea pigs are 

thought to be sensitive models for the acute effects of sulfuric acid. For chronic effects of 

sulfuric acid on the lung, monkeys are likely a suitable model due to their physiological and 

structural similarites to humans. 

Appendix D3 555 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

For comparison, a chronic REL based on the guinea pig free-standing NOAEL of 0.08 mg/m3

 in 

animals exposed continuously for one year (Alarie et al., 1973) would be 0.8 μg/m3

. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the study on sulfuric acid is the use of health effects observations from 

continuous long-term exposures to a primate. The major weaknesses are the lack of adequate 

human health effects data and the lack of a NOAEL observation. 

VIII. Potential for Differential Impacts on Children's Health

There are no reliable studies indicating that sulfuric acid is a developmental or reproductive 

toxicant. Children are likely to be at greater risk from long-term exposures because their bodies 

are growing, and their developmental processes, especially in the lung, may well be impacted by 

air pollution exposures. Elevated sulfate levels (1.6 ppb or 6.6 μg/m3

) have been associated with 

statistically significant decrements in FVC and FEV1 in a cohort of Canadian children (Stern et 

al., 1989). The chronic REL for sulfuric acid of 1 μg/m3

 is below the level associated with those 

decrements in pulmonary function. However, in a study of moderately to severe asthmatic 

children (ages 7-13) (Thurston et al., 1997), a sensitive subpopulation for sulfate effects, 

approximately 1 μg/m3 was the lowest level of ambient sulfate measured. The mean daily 

morning to afternoon peak airflow change, the use of beta-agonist medication, and the number of 

chest symptoms versus sulfate concentration in these children extrapolated linearly down to 1 

μg/m3. Thurston et al. (1997) also examined earlier data from Ontario (Burnett et al., 1994) on 

respiratory admissions to hospitals, and concluded that the sulfate threshold of effects, if it exists, 

lies below 5 μg/m3

, perhaps at about 2 μg/m3. It should be noted that the sulfate and hydrogen 

ion effects are difficult to disentangle from each other and from the effcts of other PM 

constituents. The chronic REL of 1 μg/m3

 appears to have a relatively low margin of safety with 

respect to the epidemiological studies, but these observations are consistent with the proposed 

REL of 1 μg/m3

 since asthmatic children appear to be the critically sensitive human population 

for exposure to sulfuric acid (or sulfate). 

IX. References 

Alarie Y, Busey WM, Krumm AA, Ulrich CE, and Va V. 1973. Long-term continuous exposure 

to sulfuric acid mist in cynomolgus monkeys and guinea pigs. Arch. Environ. Health 27:16-24. 

Alarie YC, Krumm AA, Busey WM, Urich CE, and Kantz RJ. 1975. Long-term exposure to 

sulfur dioxide, sulfuric acid mist, fly ash, and their mixtures. Results of studies in monkeys and 

guinea pigs. Arch. Environ. Health 30(5):254-262. 

Amdur MO, and Chen LC. 1989. Furnace-generated acid aerosols: Speciation and pulmonary 

effects. Environ. Health Perspect. 79:147-150. 

Appendix D3 556 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Amdur MO, Silverman L, and Drunker P. 1952. Inhalation of sulfuric acid mist by human 

subjects. Arch. Ind. Hyg. Occup. Med. 6:305-313. 

Bates DV, and Sitzo R. 1989. The Ontario air pollution study: Identification of the causative 

agent. Environ. Health Perspect. 79:69-72. 

Burnett RT, Dales RE, Raizenne ME, Krewski D, Summers PW, Roberts GR, Raad-Young M, 

Dann T, and Brook J. 1994. Effects of low ambient levels of ozone and sulfates on the 

frequency of respiratory admissions to Ontario hospitals. Environ Res 65:172-194. 

CARB. 1997. California Air Resources Board. Toxic Air Contaminant Identification List 

Summaries. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Cavender FL, Singh B, and Cockrell BY. 1978. The effects in rats and guinea pigs from six 

months exposures to sulfuric acid mist, ozone, and their combination. J. Environ. Pathol. 

Toxicol. 2:485-492. 

Chen LC, Miller PD, Lam HF, Guty J, and Amdur MO. 1991. Sulfuric acid-layered ultrafine 

particles potentiate ozone-induced airway injury. J. Toxicol. Environ. Health 34:337-352. 

Damokosh AI, Spengler JD, Dockery DW, Ware JH, and Speizer FE. 1993. Effects of acidic 

particles on respiratory symptoms in 7 US communities. Am Rev Respir Dis 147:A632. 

Delfino RJ, Murphy-Moulton AM, Burnett RT, Brook JR, and Becklake MR. 1997. Effects of air 

pollution on emergency room visits for respiratory illnesses in Montreal, Quebec. Am J Respir 

Crit Care Med 155:568-576. 

Dockery DW, Cunningham J, Damokosh AI, Neas LM, Spengler JD, Koutrakis P, Ware JH, and 

Speizer FE 1993. Health effects of acid aerosols on North American children: respiratory 

symptoms. Environ Health Perspect 104:500-505. 

El-Saddik YM, Osman HA, and El-Gazzar RM. 1972. Exposure to sulfuric acid in 

manufacturing of storage batteries. J. Occup. Med. 14:224-226. 

Fenters JD, Bradof JN, Aranyi C, Ketels K, Ehrlich R, and Gardner DE. 1979. Health effects of 

long-term inhalation of sulfuric acid mist-carbon particle mixtures. Environ. Res. 19:244-257. 

Fujisawa T, Iguchi K, and Uchida Y. 1986. Effects of exposure to sulfuric acid mist on the 

induction of experimental asthma in guinea pigs. Japan J. Allergol. 35(2):137-144. 

Appendix D3 557 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Gamble J, Jones W, Hancock J, and Meckstroth R. 1984. Epidemiologic environmental study of 

lead acid battery workers. III. Chronic effects of sulfuric acid on the respiratory system and teeth. 

Environ. Res. 35:30-52. 

Gearhart JM, and Schlesinger RB. 1986. Sulfuric acid-induced airway hyperresponsiveness. 

Fundam. Appl. Toxicol. 7:681-689. 

Gearhart JM, and Schlesinger RB. 1988. Response of the tracheobronchial mucociliary clearance 

system to repeated irritant exposure: effect of sulfuric acid mist on function and structure. Exp. 

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Gwynn RC, Burnett RT, and Thurston GD. 2000. A time-series analysis of acidic particulate 

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HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

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IARC. 1992. International Agency for Research on Cancer. IARC Monographs on the 

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41-130 

Iguchi K, Fujisawa T, and Uchida Y. 1986. Effects of exposure to sulfuric acid mist on the 

induction of experimental asthma in guinea pigs. Japan J. Allergol. 35(6):402-408. 

Kitabatake M, Imai M, Kasama K, Kobayashi I, Tomita Y, and Yoshida K. 1979. Effects of air 

pollutants on the experimental induction of asthma attacks in guinea pigs: Sulfuric acid mist and 

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Lippmann M, Thurston GD. 1996. Sulfate concentrations as an indicator of ambient particulate 

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Ostro B, Lipsett M, Wiener M, and Selner JC. 1989. A panel study of the effect of acid aerosols 

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Raizenne M, Neas LM, Damokosh AI, Dockery DW, Spengler JD, Koutrakis P, Ware JH, and 

Speizer FE. 1996. Health effects of acid aerosols on North American children: pulmonary 

function. Environ Health Perspect 104: 506-514. 

Schlesinger RB, Chen LC, Finkelstein I, and Zelikoff JT. 1990. Comparative potency of inhaled 

acidic sulfates: speciation and the role of hydrogenion. Environ Res 52:210-224. 

Silbaugh SA, Mauderly JL, Macken CA. 1981. Effects of sulfuric acid and nitrogen dioxide on 

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EPA. 

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sulfuric acid aerosols in asthmatics: an exposure-response relationship. Am Rev Resp Dis 

128:444-450. 

Appendix D3 559 Sulfuric acid 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

TOLUENE 

(Methyl benzene; methyl benzol; phenyl methane; toluol) 

CAS Registry Number: 108-88-3 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 300 μg/m3

 (70 ppb) 

Critical effect(s) Neurotoxic effects (decreased brain [subcortical 

limbic area] weight, altered dopamine receptor 

binding). 

Hazard index target(s) Nervous system; respiratory system; teratogenicity 

II. Physical and Chemical Properties (HSDB (1999) except as noted) 

Description Colorless liquid 

Molecular formula C7H8

Molecular weight 92.13 g/mol 

Density 0.8661 g/cm3

 @ 20°C 

Boiling point 110.6 ° C (CRC, 1994) 

Melting point −94.9° C (CRC, 1994) 

Vapor pressure 28.1 torr @ 25°C (U.S. EPA, 1984) 

Solubility miscible in most organic solvents 

Conversion factor 1 ppm = 3.76 mg/m3

 @ 25°C 

III. Major Uses or Sources 

Toluene occurs naturally as a component of crude oil and is produced in petroleum refining and 

coke oven operations; toluene is a major aromatic constituent of gasoline (HSDB, 1999). It is 

used in household aerosols, nail polish, paints and paint thinners, lacquers, rust inhibitor, 

adhesives and solvent based cleaning agents. Toluene is also utilized in printing operations, 

leather tanning and chemical processes. Benzene and other polycyclic aromatic hydrocarbons 

are common contaminants of toluene. Toluene is considered a sentinel chemical for benzene in 

the context of air and water sample monitoring. In 1996, the latest year tabulated, the statewide 

mean outdoor monitored concentration of toluene was approximately 2.2 ppb. For 1998, annual 

statewide industrial emissions of toluene from facilities reporting under the requirements of the 

Air Toxics Hot Spots Act in California were estimated to be 5,176,626 pounds (CARB, 1999). 

Note that this estimate is for stationary sources, and does not include emissions from mobile 

sources. 

Appendix D3 560 Toluene 
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IV. Effects of Human Exposures 

Neurological Effects 

Most studies reporting adverse effects due to chronic toluene exposures involve either toluenecontaining solvent abuse or occupational exposure to toluene. Solvent abusers are generally 

exposed to higher levels of toluene than are workers. A continuum of neurotoxic effects ranging 

from frank brain damage to degraded performance on psychometric tests which roughly track 

exposure levels has been observed. 

Solvent abusers

Chronic toluene abuse has been shown to cause permanent changes in brain structure (loss of 

grey and white matter differentiation; cerebral, cerebellar and brainstem atrophy) which 

correlated with brain dysfunction as measured by magnetic resonance imaging (MRI) and 

brainstem auditory evoked response (BAER) evaluations (Caldemeyer et al., 1996; Filley et al., 

1990; Ikeda and Tsukagoshi, 1990; Rosenberg et al., 1988a; Rosenberg et al., 1988b; 

Yamanouchi et al., 1995; reviewed by Agency for Toxic Substances and Disease Registry 

(ATSDR), 1999). 

Eleven chronic solvent (spray lacquer; ≈ 60% toluene, 10% dichloromethane) abusers were 

examined using MRI and BAER tests (Rosenberg et al., 1988b). Neurological abnormalities 

were seen in four of 11 subjects and included brainstem, cerebellar, cognitive and pyramidal 

findings. Brain MRIs were abnormal in three of 11 subjects and indicated the occurrence of 

diffuse cerebral, cerebellar, and brainstem atrophy and loss of differentiation between the gray 

and white matter throughout the CNS. BAERs were abnormal in five of 11 individuals. All 

three individuals with abnormal MRI scans also had abnormal neurological examinations and 

BAERs. However, two of five individuals with abnormal BAERs had normal neurological 

examinations and MRI scans. The authors suggested that BAERs may detect early CNS injury 

from toluene inhalation, even at a time when neurological examination and MRI scans are 

normal. 

 

Two subjects of a group of 22 hospitalized solvent abusers (primarily abusing toluene-based 

solvents) demonstrated decreases in intelligence quotient (IQ) as measured by the comparison of 

tests administered before the commencement of solvent abuse with tests administered during 

hospitalization for long-term solvent abuse (Byrne et al., 1991). 

Filley et al. (1990) studied 14 chronic toluene abusers using MRI and neuropsychological 

evaluations. The neuropsychological testing indicated that three patients functioned normally, 

three were in a borderline range, and eight were impaired. Independent analyses of white matter 

changes on MRI demonstrated that the degree of white matter abnormality was strongly 

correlated (p < 0.01) with neuropsychological impairment. The authors concluded that dementia 

in toluene abuse appears to be related to the severity of cerebral white matter involvement. 

Six chronic toluene abusers were examined using MRI by Caldemeyer et al. (1996). All patients 

examined demonstrated white matter atrophy and T2 hyperintensity (T2: “Spin-spin” relaxation 

time; a time constant that reflects the rate at which protons stop rotating in phase with each other 

because of the local magnetic fields of adjacent nuclei; OTA, 1984), and five of six demonstrated 

Appendix D3 561 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

T2 hypointensity of the basal ganglia and thalami. The authors noted a correlation between the 

severity of white matter degeneration and degree of neurological dysfunction. However, there 

was no correlation between the severity of imaged white matter changes and the presence of T2 

hypointensity or duration of toluene abuse. Additionally, no definite clinical evidence of damage 

to the basal ganglia and thalami was found despite the MR imaging finding of T2 hypointensity.. 

Ungar et al. (1994) developed a physical bilayered model of dipalmitoylphosphatidylcholine 

(DPPC) and toluene, and subjected DPPC control and toluene-mixed bilayers to MRI. T1 (T1: 

“Spin-lattice” relaxation time; a time constant that reflects the rate at which excited protons 

exchange energy with the surrounding environment; OTA, 1984) and T2 were measured as a 

function of toluene and lipid concentrations. Measurements of the DPPC-toluene model 

indicated that toluene-containing lipid bilayers substantially shortened T2 and had little effect on 

T1. By comparison, DPPC alone had little effect on either T1 or T2. The authors believe that 

these results suggest that partitioning of toluene into the lipid membranes of cells in cerebral 

tissue may be responsible for the hypointensity of basal ganglia noted on T2-weighted MR 

images of brains of toluene abusers. 

Occupational exposure

Solvent workers exposed to 42.8 ppm toluene (estimated as a time-weighted average) for an 

average duration of 6.8 years reported a significantly greater incidence of sore throat, dizziness 

and headache than controls; the sore throat and headache incidence demonstrated a rough doseresponse (Yin et al., 1987). 

Orbaek and Nise (1989) examined the neurological effects of toluene on 30 rotogravure printers, 

33-61 years of age (mean 50), employed at two Swedish printing shops for 4-43 years (median 

29) in 1985. Mean exposure levels at the two printing shops were 43 and 157 mg/m3

 of toluene, 

respectively; however, before 1980 the exposure levels had exceeded 300 mg/m3

 in both shops. 

The authors noted that rotogravure printing provides an occupational setting with practically pure 

toluene exposure. Comparisons were made to a reference group of 72 men aged 27-69 (mean 

47). The alcohol consumption of both the workers and referents was also determined (< 200 

g/week or > 200 g/week). Neurological function in the workers and referents was evaluated 

using interviews and psychometric testing; the results from each of the two printing shops were 

pooled. The printers reported statistically significantly higher occurrences of fatigue (60%), 

recent short-term memory problems (60%), concentration difficulties (40%), mood lability 

(27%), and other neurasthenic symptoms. The printers also scored significantly worse than 

referents in a number of psychometric tests, including synonym, Benton revised visual retention 

and digit symbol tests, even after adjustment for age. For all comparisons, tests of interaction 

between the effects of toluene exposure and alcohol consumption were not statistically 

significant. 

A battery of neurobehavioral tests was performed in 30 female workers exposed to toluene 

vapors in an electronic assembly plant (Foo et al., 1990). The average number of years worked 

was 5.7 ± 3.2 for the exposed group and 2.5 ± 2.7 years for the controls. Study subjects did not 

smoke tobacco or drink alcohol, were not taking any medications, and had no prior history of 

central or peripheral nervous system illness or psychiatric disorders. The exposed group of 

workers inhaled a time-weighted average (TWA) of 88 ppm (330 mg/m3

) toluene while the 

Appendix D3 562 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

control workers inhaled 13 ppm (49 mg/m3

). A significant decrease in neurobehavioral 

performance was observed in the exposed workers in 6 out of 8 tests. Irritant effects were not 

examined, and concurrent exposures to other chemicals were not addressed. In this study, 88 

ppm was considered a LOAEL for central nervous system effects. However, the workers 

designated by the authors to be controls did not comprise a true control group, since they were 

exposed to 13 ppm toluene. This may have resulted in an underestimation of the effects of 

exposure to 88 ppm toluene. Similar effects were noted in a follow-up study by Boey et al.

(1997). 

Abbate et al. (1993) evaluated alterations induced in the auditory nervous system by exposure to 

toluene in a group of rotogravure workers. A sample of 40 workers of normal hearing ability 

was selected from a group of 300 workers who were apparently in good health but were 

professionally exposed to toluene (12 – 14 years exposure, 97 ppm average exposure, exposure 

assessment not described). They were subjected to an adaptation test utilizing a BAER 

technique with 11 and 90 stimulus repetitions a second. The results were compared with an age 

and sex-matched control group not professionally exposed to solvents. A statistically significant 

alteration in the BAER results was noted in the toluene-exposed workers with both 11 and 90 

stimuli repetitions. The authors suggested that these results can be explained as a tolueneinduced effect on physiologic stimulus conduction mechanisms, even in the absence of any 

clinical sign of neuropathy. Furthermore, this effect could be observed in the responses of the 

entire auditory system, from peripheral receptors to brainstem nuclei. 

A group of 49 printing-press workers occupationally exposed to toluene for approximately 21.6 

years was studied by Vrca et al. (1997). Toluene exposure levels were determined from blood 

toluene and urinary hippuric acid levels, and were estimated to range from 40-60 ppm. No 

control group was used. Brain evoked auditory potential (BEAP; similar to BAER) and visual 

evoked potential (VEP) measurements were performed on a Monday morning after a 

nonworking weekend. There was a significant increase in the latencies of all the BEAP waves 

examined, except for P2 waves, as well as in the interpeak latency (IPL) P3-P4, while IPL P4-P5 

decreased significantly with the length of exposure. No correlation was noted between the 

amplitude of BEAP waves and the length of exposure. The amplitude but not the latency of all 

the VEPs examined decreased significantly with the length of exposure. 

 

The effects of acute and chronic toluene exposure on color vision were studied in a group of 

eight rotogravure printing workers (Muttray et al., 1999). The workers had been employed as 

printers for an average of 9.8 years. The color vision acuity of the workers before and after an 

acute toluene exposure (28 – 41 minutes in duration, concentration 1115 – 1358 mg/m3

) was 

evaluated using the Farnsworth panel D-15 test, the Lanthony desaturated panel D-15 test, and 

the Standard Pseudoisochromatic Plates part 2. A control group of 8 unexposed workers was 

also tested. Acute toluene exposure had no effect on color vision. Print worker performance 

prior to acute toluene exposure (chronic effects) was similar to controls on the Farnsworth panel 

D-15 and Standard Pseudoisochromatic Plates part 2 tests. Print worker performance on the 

Lanthony desaturated panel D-15 test was worse than that of controls (median scores of 1.18 and 

1.05 for exposed and controls (higher number indicates degraded performance), respectively, 

but not significantly (p = 0.06). The authors noted that the small number of subjects limited the 

statistical power of the study. 

Appendix D3 563 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Zavalic et al. (1998) examined the effects of chronic occupational toluene exposure on color 

vision using a group of 45 exposed workers (mean toluene exposure concentration = 120 ppm) 

and 53 controls. Color vision was evaluated using the Lanthony desaturated panel D-15 test; test 

scores were age and alcohol consumption-adjusted. Color vision was significantly impaired in 

toluene-exposed workers (p < 0.0001) compared to controls. It was also observed that there was 

no significant difference between test scores on Monday morning (prework) and Wednesday 

morning. The authors stated that the effect of toluene on color vision can be chronic and that the 

possible recovery period is longer than 64 hours. 

Hepatic Effects 

Greenburg et al. (1942) reported liver enlargement in 32 of 106 (30.2%) painters employed in an 

aircraft factory compared to 7% in a control group. However, there was some exposure to other 

solvents (ethanol, ethyl acetate, butyl acetate) and paint ingredients such as zinc chromate. 

Liver toxicity has been reported in toluene solvent abusers (Fornazzari et al., 1983). Eight of 24 

solvent abusers demonstrated abnormal results in three liver function tests; however, the tests 

used were not specified. The test parameters returned to normal after two weeks of toluene 

abstinence, suggesting that any liver damage caused by toluene abuse in those patients was not 

long lasting. 

A cross-sectional study by Boewer et al. (1988) showed no significant correlation between 

toluene exposure and the levels of serum enzymes (serum aspartate aminotransferase (ASAT), 

alanine aminotransferase (ALAT), γ-glutamyltransferase (GGT)) considered to be indicators of 

hepatic damage. In another cross-sectional study of 289 printing workers exposed to less than 

200 ppm for 8 hours/day, 8 workers had significantly elevated serum enzymes (ALT/AST ratio, 

mean = 1.61) potentially indicative of liver damage. In each case, liver biopsy indicated a mild 

pericentral fatty change (Guzelian et al., 1988). However, the mean toluene exposure 

concentration was not reported (only an upper bound), and no control group was included in the 

study. 

V. Effects of Animal Exposures 

Neurotoxic Effects 

Sprague-Dawley rats (15/sex/group) were exposed to 0, 100, or 1481 ppm toluene for 6 

hours/day, 5 days/week for 26 weeks (API, 1981). Neurohistopathological examinations were 

conducted in 3-5 rats/sex/group at weeks 9, 18, and 27. No significant treatment-related effects 

were reported. The study usefulness was limited because there were no other 

neurohistopathological examinations or organ weight measurements conducted on the animals. 

Forkman et al. (1991) studied the potential neurotoxicity of toluene inhalation exposure (3700 

mg/m3

 (1000 ppm), 21 hours/day, 5 days/week for 4 weeks) in male Sprague-Dawley rats. The 

rats were either trained in behavior meant to be performance tested and then exposed to toluene, 

or exposed and then trained. The rats were then subjected to several behavioral tests, including 

an operant test with baseline performance and extinction, motor coordination, and exploratory 

Appendix D3 564 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

activity. All tests were performed from 11 to 35 days after the end of the exposure. Exposure of 

trained rats to toluene resulted in a significantly different overall test performance when 

compared to controls. Rats trained after toluene exposure also had test performances different 

from controls, but the difference was not statistically significant. 

Rats exposed to toluene concentrations of 1000 ppm or 100 ppm, 6 h/day, 5 days/week, for 3 or 

6 months, respectively, demonstrated statistically significant decreased motor function as 

measured by degraded performance (approximately 60% and 65% of control at 1000 and 100 

ppm toluene, respectively) on a rotarod performance test and decreases in spontaneous motor 

activity (approximately 62% of control at 100 ppm toluene) (Korsak et al., 1992). 

von Euler et al. (1993) studied the effects of subchronic toluene inhalation exposure (80 ppm, 4 

weeks, 5 days/week, 6 hours/day) on spatial learning and memory, dopamine-mediated 

locomotor activity and dopamine D2 agonist binding in rats. Spatial learning (postexposure days 

3-6) and memory (postexposure day 14) were tested using a water maze. Spontaneous and 

apomorphine-induced locomotor activity was evaluated on postexposure day 17. Effects on 

binding parameters of the dopamine D2 agonist S(-)[N-propyl- 3

H(N)]-propylnorapomorphine 

([H]NPA) were determined using membrane preparations of the neostriatum of the rat brain. 

Toluene exposure caused a statistically significant impairment in spatial learning and memory. 

Toluene also significantly increased apomorphine-induced locomotion and motility but not 

rearing. Spontaneous locomotion, motility and rearing were not affected by toluene. Toluene 

exposure significantly increased the BB

max and KD values for [ H]NPA binding. These results 

indicate that subchronic toluene exposure of rats to toluene causes persistent deficits in spatial 

learning and memory, a persistent increase in dopamine-mediated locomotor activity and an 

increase in the number of dopamine D

3

2 receptors in the neostriatum. 

Male rat exposure to toluene (0, 40, 80, 160 or 320 ppm, 4 weeks, 6 hours/day, 5 days/week) , 

followed by a postexposure period of 29-40 days, resulted in decreased brain wet weights of the 

caudate-putamen (trend test for dose-response significant at p < 0.05) and subcortical limbic 

areas (trend test for dose-response significant at p < 0.01; significantly less than controls (p < 

0.001) at concentrations of 80 ppm and higher) (Hillefors-Berglund et al., 1995). Toluene 

exposure also significantly altered dopamine receptor activity (trend test for dose-response) as 

indicated by decreased IC50 (inhibition constant) (significantly less than controls (p < 0.05) at 80 

ppm), KH (inhibition constant for high-affinity receptor sites), KL (inhibition constant for lowaffinity receptor sites), and RH% (high-affinity receptor site specific binding) values for 

dopamine competitive inhibition of [3

H]raclopride-binding in the caudate-putamen. Toluene 

exposure did not significantly affect the wet weights of the whole brain, serum prolactin levels, 

the KD (disassociation constant) or the BB

max (maximal specific binding) values of [ H]raclopridebinding in the caudate-putamen and the subcortical limbic area, or the effect of dopamine on IC

3

50

values at [ H]raclopride-binding sites in the subcortical limbic area. Exposure to xylene or 

styrene (80 and 40 ppm, respectively; 4 weeks, 6 h/day, 5 days/week) followed by a 

postexposure period of 26-32 days had no effect on the parameters described above. The authors 

concluded that long-term exposure to low concentrations of toluene (≥ 80 ppm), but not xylene 

(80 ppm) or styrene (40 ppm), leads to persistent increases in the affinity of dopamine D

3

2 agonist 

binding in the rat caudate-putamen. The authors also suggested that the enhancement of 

Appendix D3 565 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

apomorphine-induced locomotor activity seen after toluene exposure by von Euler et al. (1993) 

may be related to the increased D2 agonist activity described above (IC50, KH, KL values). 

Respiratory Effects 

A study of the chronic effects of toluene in rats (5-20 animals per group) exposed for 106 weeks 

to 0, 30, 100, or 300 ppm (0, 113, 375, or 1125 mg/m3

) toluene showed no treatment-related 

effects on histopathology of major organs, including the nasal turbinates (CIIT, 1980). In this 

study, the nasal histopathology examination sampling may have been inadequate to demonstrate 

the nasal lesions reported by the NTP (1990). 

Rats (20 per group) exposed for 2 years to 0, 600, or 1200 ppm (0, 2261, or 4523 mg/m3

) toluene 

6.5 hours/day, 5 days/week for 103 weeks were examined for hematological and 

histopathological effects in addition to gross observations of toxicity (NTP, 1990). Significant 

erosion of the olfactory epithelium was observed in male rats while degeneration of the 

respiratory and nasal epithelium was observed in both sexes at 600 ppm. 

Mice were exposed chronically to 0, 120, 600, or 1200 ppm (0, 452, 2261, or 4523 mg/m3

) 

toluene 6.5 hours/day, 5 days/week, for 2 years (NTP, 1990). The only treatment-related effect 

was a significant increase in the number of animals with hyperplasia of the bronchial epithelium 

in the 1200 ppm exposure group. 

Reproductive and Developmental Toxicity 

Reproductive toxicity to maternal rats was observed during exposure to 1500 ppm toluene, 24 

hours/day on days 9 to 14 of gestation (Hudak and Ungvary, 1978). Two dams out of 19 died 

during exposure. Fetuses from the 1500 ppm group showed increased incidence of sternebral 

alterations, extra ribs and missing tails. The same exposure on days 1 through 8 of gestation 

resulted in 5 deaths out of 14 dams. Fetuses in this regimen showed increased incidence of 

hydrocephaly and growth retardation compared to controls. A third regimen that exposed 

maternal rats to 1000 ppm on days 1 through 21 of gestation resulted in no maternal deaths or 

toxicity, and an increase in the incidence of skeletal variations in the fetuses. When exposed to 

1500 ppm continuously, maternal mice died within 24 hours of exposure whereas exposure to 

500 ppm had no apparent effect. Examination of the fetal mice showed significant growth 

retardation in the 500 ppm group. 

A 2-generation study of the effects of 0, 100, 500, or 2000 ppm (0, 377, 1885, or 7538 mg/m3

) 

toluene in rats (males, 10-40 per group; females, 20-80 per group) was done by the American 

Petroleum Institute (API)(1985). Rats were exposed for 6 hours/day, 7 days/week for 80 days 

and a 15 day mating period. The mated females were then exposed to the same concentrations 

during days 1-20 of gestation and days 5-20 of lactation. After weaning, the F1 pups were 

exposed 80 times to the appropriate exposure level and then randomly mated to members of the 

same exposure group. The F1 generation showed significantly decreased body weight which 

persisted throughout lactation. No effects were observed on histopathology. No data were 

presented for the F2 generation. 

Da Silva et al. (1990) exposed rats and hamsters to 0 or 800 mg/m3

 toluene for 6 hours/day on 

gestation days 14-20 (rats), or days 6-11 (hamsters). Exposed rats demonstrated a significant 

Appendix D3 566 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

exposure-related decrease in birth weight compared with controls. In addition to low birth 

weight, the number of live pups was significantly lower in the 800 ppm group. No deficits in 

any parameter were noted in the hamsters. In this study, no neurobehavioral effects were noted in 

the offspring. 

Hass et al. (1999) exposed rats to 0 or 1200 ppm toluene for 6 h per day from day 7 of pregnancy 

until day 18 postnatally. Developmental and neurobehavioral effects in the offspring were 

investigated using a test battery including assessment of functions similar to those in the 

proposed Organization for Economic Cooperation and Development (OECD) Testing Guidelines 

for Developmental Neurotoxicity Study (physical development, reflex development, motor 

function, motor activity, sensory function, and learning and memory). The exposure did not 

cause maternal toxicity or decreased offspring viability. However, lower birth weight, delayed 

development of reflexes, and increased motor activity in the open field was noted in the exposed 

offspring. The exposed female offspring had poorer scores on a Morris water maze test (they 

took longer to locate a hidden platform after platform relocation) at the age of 3.5 months 

indicating impaired cognitive function. The difference was not related to impaired swimming 

capabilities since swim speeds were similar to control values. The authors stated that exposure to 

1200 ppm toluene during brain development caused long-lasting developmental neurotoxicity in 

rats. 

Toluene has been listed under Proposition 65 as being known to the State of California to cause 

reproductive toxicity (OEHHA, 1999). Its NSRL is 7,000 micrograms per day. 

Appendix D3 567 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Hillefors-Berglund et al. (1995); supported by Orbaek 

and Nise (1989), Foo et al. (1990) 

Study population Male Sprague-Dawley rats 

Exposure method Inhalation 

Critical effects Decreased brain (subcortical limbic area) weight, 

 altered dopamine receptor (caudate-putamen) binding 

LOAEL 80 ppm 

NOAEL 40 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 4 weeks, followed by 29-40 days recovery 

Average experimental exposure 7 ppm (40 × 6/24 hours × 5/7 days) 

Human equivalent concentration 7 ppm (gas with systemic effects, based on RGDR = 

1.0 using default assumption that λa = λh

Subchronic uncertainty factor 10 

Interspecies uncertainty factor 1 (see below) 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.07 ppm (70 ppb; 0.3 mg/m3

; 300 μg/m3) 

Supportive human study Foo et al., 1990 

Study population 30 female workers in an electronic assembly plant 

Exposure method Occupational inhalation 

Critical effects Neurobehavioral deficits in 6 out of 8 tests 

LOAEL 88 ppm 

NOAEL Not observed 

Exposure continuity 10 m3

/day occupational inhalation rate, 

5 days/week 

Average occupational exposure 31.4 ppm (88 ppm x 10/20 x 5/7) 

Exposure duration 5.7 + 3.2 years (exposed group); 

2.5 + 2.7 years (controls) 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 300 

Inhalation reference exposure level 0.1 ppm (100 ppb; 0.4 mg/m3

; 400 μg/m3) 

The critical animal study (Hillefors-Berglund et al., 1995) used to derive an REL for toluene 

describes adverse neurological effects in rats after a well characterized inhalation exposure to 

toluene. The study results contain both a LOAEL and a NOAEL. Decreased brain (subcortical 

limbic area) weight and altered dopamine receptor binding compared to controls were noted at 

the NOAEL, but the changes were not statistically significant; this suggests that if a threshold for 

Appendix D3 568 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

adverse neurological effects exists in this study, it would be at or below the observed NOAEL. 

The study LOAEL for altered dopamine receptor binding agrees qualitatively with results from 

similar studies (von Euler et al., 1994). Additionally, toluene-induced neurotoxicity has been 

described in many studies by a variety of endpoints in both animals and humans (ATSDR, 1999). 

The adverse neurotoxic effects associated with toluene exposure in the rat study by HilleforsBerglund et al. (1995), decreased brain (subcortical limbic area) weight and altered dopamine 

receptor binding, occur in areas of the rat brain that are structurally and functionally similar to 

brain areas (basal ganglia, thalami) of some human toluene abusers that demonstrate MRI 

alterations (T2 hypointensity). The altered MRI parameters may be the result of the partitioning 

of toluene into the lipid membranes of brain cells (Ungar et al., 1994). Table 1 lists several 

Reference Exposure Levels (RELs) calculated from the most sensitive animal and human 

neurotoxicity studies available. These RELs are also protective for other adverse endpoints, such 

as respiratory tract damage and teratogenicity. 

Table 1: Reference Exposure Levels (RELs) from Selected Neurotoxicity Studies 

Study Duration Effect LOAEL LOAEL NOAEL NOAEL total REL REL 

(ppm) (ppm) (ppm) (ppm) UFa (ppb) (μg/m3)

(TWA) (TWA) 

VonEuler et 

al. (1988) 

4 weeks rat: altered brain dopamine 

receptor binding 

80 14.3 1000 14 54 

Orbaek and 

Niseb (1989) 

29 years human: impairment on 

neuropsychometric tests 

11.2 - 41 4 - 14.6 100 40 - 

146

150 - 

551 

Foo (1990) 5.7 years human: neurobehavioral 

tests 

88 31.4 300 105 394 

Korsak (1992) 6 months rat: impaired motor 

function 

100 17.9 100 179 671 

HilleforsBerglund 

(1995) 

4 weeks rat: decreased brain 

(subcortical limbic area) 

weight; altered brain 

dopamine receptor binding

80 14.3 40 7.1 100 71 271 

LOAEL: Lowest Observable Effect Level; NOAEL: No Observable Effect Level 

REL: Reference Exposure Levels; TWA: time-weighted average 

a: Uncertainty Factors used to derive RELs

VonEuler et al. (1988) LOAEL to NOAEL UF = 10, subchronic to chronic UF = 10, animal to human UF = 

1, intraspecies variability = 10; total UF = 1000. 

Orbaek and Nise (1989) LOAEL to NOAEL UF = 10, intraspecies variability = 10; total UF = 100

Foo et al. (1990) LOAEL to NOAEL UF = 10, subchronic to chronic UF = 3, intraspecies variability 

= 10; total UF = 300 

Korsak et al. (1992) LOAEL to NOAEL UF = 10, animal to human UF = 1, intraspecies variability = 10; 

total UF = 100. 

Hillefors-Berglund et al. 

(1995) 

subchronic to chronic UF = 10, animal to human UF = 1, intraspecies variability = 

10; total UF = 100.

b: Pooled psychometric data from two printing plants with different toluene concentrations (11.2 and 41 ppm) were 

used to determine significant neurotoxic effects by Orbaek and Nise (1989). The range of RELs derived from that 

study lists the upper and lower bounds for risk associated with the pooled population exposures. ATSDR (1999) 

used the Orbaek and Nise (1989) study data, assuming an exposure concentration of 11.2 ppm, to derive a chronic 

inhalation minimal risk level (MRL). 

Appendix D3 569 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

If both human and animal adverse effect data on a chemical are available, OEHHA prefers to use 

the human data to develop a REL when possible. However, the study by Hillefors-Berglund et 

al. (1995) provides data (decreased brain [subcortical limbic area] weight and altered brain 

dopamine receptor binding) which are specific and sensitive measures of neurotoxicity that 

would not be obtainable in human studies. In contrast, the psychometric tests used to generate 

the neurotoxicity data in the human occupational exposure studies described above tend to be 

less sensitive and suffer from greater measurement uncertainty. Additionally, the HilleforsBerglund et al. (1995) study has better exposure characterization than the human occupational 

exposure studies. Nonetheless, the human studies are useful in supporting the derivation of the 

REL for toluene. Ordinarily, an interspecies uncertainty factor of 3 would be applied, in addition 

to the human equivalent concentration calculation, to reflect the uncertainty associated with 

extrapolating from animals to humans. However, in this case the uncertainty in the interspecies 

extrapolation is reduced by the availability of human epidemiological data with generally 

consistent effect levels, after appropriate duration corrections. Based on comparison of the data 

in both animals and humans, it appears that a REL of 271 μg/m3

 (rounded to 300 μg/m3 in the 

final derivation) would protect exposed humans from experiencing chronic neurotoxic effects. 

VII. Data Strengths and Limitations for Development of the REL 

The major strength of the REL for toluene is the use of an animal study with accurate exposure 

characterization and both LOAEL and NOAEL observations for an effect (neurotoxicity), 

supported by observations from other animal and human studies. A weakness is the uncertainty 

in predicting human health risk from animal adverse effect data. However, this is mitigated by 

the availability of human data showing effect levels that are, after appropriate corrections, 

broadly consistent with the animal data. 

VII. References 

Abbate C, Giorgianni C, Munao F, and Brecciaroli R. 1993. Neurotoxicity induced by exposure 

to toluene. An electrophysiologic study. Int. Arch. Occup. Environ. Health 64:389-92. 

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von Euler G, Ogren SO, Li XM, Fuxe K, and Gustafsson JA. 1993. Persistent effects of 

subchronic toluene exposure on spatial learning and memory, dopamine-mediated locomotor 

activity and dopamine D2 agonist binding in the rat. Toxicology 77:223-32. 

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Appendix D3 573 Toluene 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY 

2,4- and 2,6-TOLUENE DIISOCYANATE 

(2,4- and 2,6-TDI; 2,4- and 2,6-diisocyanato-1-methylbenzene; 2,4- and 2,6-diisocyanatoluene) 

CAS Registry Number: 584-84-9 or 26471-62-5 (mixture) 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 0.07 μg/m3 (0.01 ppb)

Critical effect(s) Decreased lung function in occupationally 

exposed workers 

Hazard index target(s) Respiratory system 

II. Chemical Property Summary (HSDB, 1995; CRC, 1994)

Description Colorless to pale yellow liquid 

Molecular formula C9H6N2O2

Molecular weight 174.15 g/mol 

Boiling point 2,4-TDI: 251°C 

Melting point 2,4-TDI: 20.5°C 

2,6-TDI: 18.3°C 

Vapor pressure 2,4-TDI: 0.008 torr @ 20°C 

Solubility Miscible with ether, acetone, benzene, carbon 

tetrachloride, chlorobenzene, diglycol 

monomethyl ether, kerosene, olive oil, alcohol; 

soluble in ethyl acetate 

Conversion factor 7.1 μg/m3

 per ppb at 25oC 

III. Major Uses and Sources 

Commercial toluene diisocyanate is comprised of approximately 80% 2,4-TDI and 20% 2,6-TDI. 

TDI is used in the manufacture of polyurethane foams, elastomers, and coatings (HSDB, 1995; 

Howard, 1989). It is also used in the manufacture of floor and wood finishes, lacquers, foam 

plastics, polyurethane foam coated fabrics, and insulation materials (HSDB, 1995; Howard, 

1989; Duncan et al., 1962). Emissions of TDI to the atmosphere can occur during production, 

handling, and processing of polyurethane foam (Howard, 1989) and coatings. The annual 

statewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in 

California based on the most recent inventory were estimated to be 13,223 pounds of toluene 

diisocyanates, 35,663 pounds of toluene-2,4-diisocyanate, and 754 pounds of toluene-2,6-

diisocyanate (CARB, 1999). 

Appendix D3 574 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

IV. Effects of Human Exposures 

Diem et al. (1982) conducted a prospective study beginning in 1973 of 277 male workers 

involved in the production of TDI. The study examined pulmonary function, with nine 

examinations conducted over a five year period. A large group of workers (168) with no 

previously reported TDI exposure was examined 6 months prior to TDI production in the plant to 

provide baseline pulmonary function measurements. Personal sampling by continuous tape 

monitors provided exposure levels, but was not used until 2 years after the study was initiated. 

Sampling information resulted in a division of the workers into two groups: those exposed to 

levels below 68.2 ppb-months (which reflects the level of exposure of a worker for the entire 5 

year duration in the low-exposure area (geometric mean = 1.1 ppb)) and those above this level. 

The arithmetic mean exposure level for the non-smokers was 1.9 ppb TDI in the high-exposure 

group and 0.9 ppb TDI in the low-exposure group (calculated by Hughes, 1993). The higher 

exposure group was further limited to those individuals who showed a normal FEV1 to height 

ratio. Data were analyzed by the maximum likelihood weighted regression approach (Diem and 

Liukkonen, 1988). Both FEV1 and forced expiratory flow (25-75%) [FEF (25-75%)] among 

workers who never smoked were found to be significantly reduced in the high-exposure group 

(n = 21) compared to the low-exposure group (n = 35). Categorizing workers based on time 

spent at exposure levels above 20 ppb demonstrated a significant difference in FEV1 and 

FEF(25-75%) and this effect was also observed among current smokers. Among low-exposure 

workers, a smoking effect was observed, with smokers showing a significant decline in FEV1. 

A similar longitudinal study of lung function was conducted among workers exposed to TDI 

during the course of polyurethane foam production (Jones et al., 1992). Participants (181 males 

and 46 females) were required to have 3 or more spirometric examinations over the 5 year study 

period. Exposure of males was evaluated by personal monitors and resulted in arithmetic mean 

low exposure levels of 0.3, 0.4, and 0.4 ppb TDI for never-smokers, ex-smokers, and current 

smokers, respectively. Among workers with high-level exposure, mean TDI levels were 

reported to be 1.3, 1.2, and 1.2 ppb for never-, ex-, and current smokers, respectively. Stepwise 

multiple linear regression methods (excluding asthmatics) were used in evaluating the data 

(Diem and Liukkonen, 1988). No relationship between TDI exposure and change in lung 

function was observed, although the prevalence of chronic bronchitis was significantly 

associated with exposure. 

A longitudinal study of 780 workers exposed to TDI in the production of polyurethane foam was 

also conducted (Bugler et al., 1991; unpublished). Exposure levels were established using 

continuous-tape personal monitoring devices. The mean exposure level was 1.2 ± 1.1 (SD) ppb 

TDI among 521 workers and 0.3 ± 0.18 ppb TDI in the control group. Another control group 

who handled cold urethane products had an 8 hour time-weighted average exposure of 0.6 ppb 

TDI. No significant longitudinal changes in FEV1 were found after regression analysis, although 

FEV1 decline was high among the control group. Exposure levels among the different groups 

were close, limiting the power of the study to detect changes. Approximately 3% of the 780 

workers showed signs of TDI sensitization and, of these, over 80% were in the group exposed to 

1.2 ppb. 

Appendix D3 575 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Meta-analysis of the three data sets (Jones et al., 1992; Bugler et al., 1991; Diem et al., 1982) 

showed that the difference in significance among the findings of each of the studies could have 

been due to chance. The change in the probability density for the decline in FEV1 shifted in the 

same direction for all data sets and the smoker/non-smoker slope difference became less 

meaningful with the data set combination (Hasselblad, 1993). 

Another toxicological area of concern with exposure to TDI is the development of sensitization, 

resulting in a well-documented condition known as “isocyanate asthma” of either immediate or 

delayed-type onset (Moscato et al., 1991). The level of exposure required to either develop or 

trigger a sensitization reaction is not well documented, however. Weaknesses of studies showing 

pulmonary effects of TDI exposure include use of area sampling vs. breathing-zone 

measurement of exposure, poor statement of criteria for evaluating hypersensitivity, and the 

presence of other compounds in the environment which may influence lung function. 

V. Effects of Animal Exposures 

Mice were exposed to TDI concentrations ranging from 0.007 to 1.18 ppm for 3 hours/day for 5 

days consecutively (Sangha and Alarie, 1979); decreased respiratory rate was observed in groups 

exposed to levels higher than 0.023 ppm TDI. Groups of four mice were also exposed to 0.031 

and 0.250 ppm TDI for 3 hours/day for 3 days. Lesions of the external nares and respiratory 

epithelium were observed in the high dose group. 

Female guinea pigs were exposed to 0.12, 0.36, 0.61, 0.96, and 10.00 ppm TDI (head-only) for 3 

hours/day for 5 consecutive days (short protocol) or to 0.02 ppm TDI (whole body) plus controls 

for 6 hours/day, 5 days/week for 70 days (long protocol). The animals showed decreased 

respiration rate two hours into exposure at levels above 0.12 ppm TDI and had a cytophilic 

antibody response at 0.96 ppm and above (Karol, 1983). All animals exposed to 10 ppm died. 

Dermal sensitivity was evident among animals in the short protocol down to 0.12 ppm TDI. No 

antibody response or dermal sensitivity developed in the animals exposed to 0.02 ppm TDI in the 

long protocol. 

Similarly, guinea pigs (8 females) were exposed head only to 1.40 ppm TDI for 3 hours/day for 4 

days (no control group). In a second exposure regimen, animals (n = 24) were exposed to 0.02 

ppm TDI for 6 hours/day, 4 days/week for 70 days (whole body) including a control group 

(n = 8) exposed to room air in a similar manner (Wong et al., 1985). Half the animals (4/8) 

exposed to 1.40 ppm TDI showed pulmonary hypersensitivity (measured on days 37 and 38) and 

all developed TDI-specific IgE antibodies, whereas none of the animals in the 0.02 ppm TDI 

group showed either of these effects. Histopathological effects in the 1.40 ppm TDI group 

included interstitial inflammation, pleural thickening, and peripheral lymphoid hyperplasia. 

Interstitial inflammation was noted in 2/24 animals exposed to 0.02 ppm TDI. 

SD rats and CD-1 mice were exposed to 0.05 or 0.15 ppm TDI for 6 hours/day, 5 days/week for 

2 years (Loeser, 1983; nasal histopathology reported by Owen, 1984). Among female rats at 

both dose levels and male rats at the high dose level, histopathological effects observed included 

necrotic rhinitis, metaplasia, and inflammation of the respiratory epithelium. Female animals 

Appendix D3 576 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

showed dose-dependent increases in incidence and severity of this effect. Similar lesions were 

reported in mice, although they were not well characterized. 

Reproductive toxicity of TDI was evaluated in a two-generation study conducted in rats (Tyl and 

Neeper-Bradley, 1989). Weanling rats (28/sex/dose) were exposed to 0, 0.020, 0.079, and 0.290 

ppm TDI for 6 hours/day, 5 days/week, for 10 weeks, at which time the animals were randomly 

mated. Exposure of the females continued through gestation (excepting gestational day 20 

through the fourth day postpartum), and exposure of the males continued only until the delivery 

of the F1 generation. Weanlings in the F1 generation were exposed in a manner similar to the 

parental (P0 ) generation and bred after weaning to produce the F2 generation. Body weights 

were significantly reduced among animals of both sexes in the highest dose group and weight 

gain was reduced among males in the highest dose group. Effects on the respiratory system in 

the P0 generation animals included rhinitis of the epithelium in the two highest dose groups of 

both male and female animals. Hyperplasia of the respiratory epithelium was also increased in 

the high dose groups of both sexes among P0 animals. Among males in the F1 generation, the 

incidence of rhinitis was significantly increased at all exposure levels and the incidence of 

submucosal lymphoid infiltrates of the larynx and trachea was increased in the highest dose 

group. F2 generation animals showed reduced pup weight and weight gain during the lactation 

period in the two highest dose groups. 

Developmental toxicity of TDI was evaluated by exposing pregnant Sprague-Dawley rats 

(25/group) for 6 hours/day on gestational days 6-15 to 0, 0.021, 0.120, or 0.48 ppm TDI (Tyl, 

1988). Reduced maternal body weight, decreased food consumption, and rales occurred among 

the dams in the 0.48 ppm TDI dose group. A significant fetal effect, a statistically significant 

increase in a specific skeletal malformation, was reported in the highest dose group. 

VI. Derivation of the Chronic Reference Exposure Level 

Study Diem et al., 1982 

Study population Human TDI production workers (n = 168) 

Exposure method Occupational inhalation exposure 

Critical effects Decreased lung function 

LOAEL 0.014 mg/m3

 (1.9 ppb) 

NOAEL 0.006 mg/m3

 (0.9 ppb) (non-smokers) 

Exposure continuity 8 h/day (10 m3

/day occupational exposure), 5 d/wk 

Exposure duration 5 years 

Average occupational exposure 0.002 mg/m3

 for NOAEL group 

(0.006 x 10/20 x 5/7) 

Human equivalent concentration 0.002 mg/m3

 for NOAEL group 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 3 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.00007 mg/m3

 (0.07 μg/m3; 0.01 ppb) 

Appendix D3 577 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

The chronic REL is equivalent to the U.S. EPA RfC. OEHHA agreed with the U.S. EPA 

analysis and the selection of Diem et al. (1982) as the most appropriate study to use for the REL. 

The rationale for selection of this study is as follows. This study presented evidence of a decline 

in lung function, as indicated by decrements in FEV1, among workers involved in TDI 

production. Other factors supporting its quality include: 

(1) the absence of other confounding compounds in the work environment, 

(2) the establishment of baseline lung function prior to exposure to TDI, 

(3) a “parallel internal comparison” of study groups for lung function, 

(4) an appropriate statistical analysis which took into account interindividual variability, 

(5) breathing zone measurement of TDI (although commenced 2 years into the study), and 

(6) a smoking effect on lung function. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the chronic REL for TDI are the use of human exposure data from 

workers exposed over a period of years and the observation of a NOAEL. The major 

weaknesses are the uncertainty in estimating exposure, the potential variability in exposure 

concentration, and the limited nature of the study that focused on lung effects. 

VIII. References 

Bugler J, Clark RL, Hill ID, and McDermott M. 1991. The acute and long-term respiratory 

effects of aromatic di-isocyanates. A five year longitudinal study of polyurethane foam workers. 

Study sponsored by the British Rubber Manufacturers Association, the International Isocyanate 

Institute, and the Health and Safety Executive, UK. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Diem JE, Jones RN, Hendrick DJ, Glindmeyer HW, Dharmarajan V, Butcher BT, Salvaggio JE, 

and Weill H. 1982. Five-year longitudinal study of workers employed in a new toluene 

diisocyanate manufacturing plant. Am. Rev. Respir. Dis. 126:420-428. 

Diem JE, and Liukkonen JR. 1988. A comparative study of three methods for analyzing 

longitudinal pulmonary function data. Stat. Med. 7:19-28. 

Duncan B, Scheel LD, Fairchild EJ, Killens R, and Graham S. 1962. Toluene diisocyanate 

inhalation toxicity: pathology and mortality. Am. Ind. Hyg. Assoc. J. 23:447-456. 

Appendix D3 578 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Hasselblad V. 1993. Analysis of studies of workers exposed to toluene diisocyanate. Report 

prepared for U.S. EPA, Environmental Criteria and Assessment Office, Research Triangle Park, 

NC. 

Howard PH. (ed) 1989. Handbook of environmental fate and exposure data for organic 

chemicals. Vol. I. Large production and priority pollutants. Chelsea, MI: Lewis Publishers, Inc. 

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM Version). Denver, CO: Micromedex, Inc. (Edition expires 7/31/96). 

Hughes J. April and November, 1993. Memoranda from Janet Hughes, Tulane Medical Center, 

to Mark Greenburg, U.S. EPA. 

Jones RN, Rando RJ, Glindmeyer, Foster TA, Hughes JM, O’Neil CE, and Weill H. 1992. 

Abnormal lung function in polyurethane foam producers: weak relationship to toluene 

diisocyanate exposures. Am. Rev. Respir. Dis. 148:671-677. 

Karol MH. 1983. Concentration-dependent immunologic response to toluene diisocyanate (TDI) 

following inhalation exposure. Toxicol. Appl. Pharmacol. 68:229-241. 

Loeser E. 1983. Long-term toxicity and carcinogenicity studies with 2,4-/2,6-toluene 

diisocyanate (80/20) in rats and mice. Toxicol. Lett. 15:71-81. 

Moscato G, Dellabianca A, Vinci G, Candura SM, and Bossi MC. 1991. Toluene diisocyanateinduced asthma: clinical findings and bronchial responsiveness in 113 exposed subjects with 

work-related respiratory symptoms. J. Occup. Med. 33:720-725. 

Owen PE. 1984. The toxicity and carcinogenicity to rat of toluene diisocyanate vapour 

administered by inhalation for a period of 113 weeks. Addendum report. Volume 2. Hazelton 

Laboratories Europe. 

Sangha GK, and Alarie Y. 1979. Sensory irritation by toluene diisocyanate in single and repeated 

exposures. Toxicol. Appl. Pharmacol. 50:533-547. 

Tyl RW. 1988. Developmental toxicity study of inhaled toluene diisocyanate vapor in CD 

(Sprague-Dawley) rats. Union Carbide, Bushy Run Research Center. Revised project report 50-

592. 

Tyl RW, and Neeper-Bradley TL. 1989. Two-generation reproduction study of inhaled toluene 

diisocyanate in CD (Sprague-Dawley) rats. Union Carbide, Bushy Run Research Center. Project 

No. 86-33-90704. 

U.S. EPA. 1995. United States Environmental Protection Agency. Documentation of the 

reference concentration for chronic inhalation exposure (RfC) for 2,4-/2,6-toluene diisocyanate 

mixture. Integrated Risk Information System (IRIS on-line). Washington, DC: U.S. EPA. 

Available online at http://www.epa.gov/ngispgm3/iris 

Appendix D3 579 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels December 2000

Wong KL, Karol MH, and Alarie Y. 1985. Use of repeated CO2 challenges to evaluate the 

pulmonary performance of guinea pigs exposed to toluene diisocyanate. J. Toxicol. Environ. 

Health 15:137-148. 

Appendix D3 580 Toluene diisocyanate 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY

TRICHLOROETHYLENE 

(trichloroethylene; 1,1-2-trichloroethylene, 1,1-dichloro-2-chloroethylene, acetylene trichloride, 

and ethylene trichloride) ) 

CAS Registry Number: 79-01-6 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 600 μg/m3

 (100 ppb) 

Critical effect(s) Neurotoxicological effects (drowsiness, fatigue, 

headache) and eye irritation in workers. 

Hazard index target(s) Nervous system; eyes 

II. Physical and Chemical Properties (Fan, 1988; CRC, 1994) 

Description Colorless liquid/vapor; sweetish, chloroform-like

 odor 

Molecular formula C2HCl3

Molecular weight 131.4 

Density 1.47 g/cm3

 @ 20°C 

Boiling point 87.2 °C 

Melting point −84.7°C 

Vapor pressure 77 torr @ 25°C 

Vapor density 4.5 (air = 1) 

Solubility Soluble in alcohol, ethers, petroleum distillates 

 and other halogenated solvents

Conversion factor 1 ppm = 5.37 mg/m3

 @ 25° C

III. Major Uses or Sources

Trichloroethylene was once used as an extractant in food processing and has been used as an 

anesthetic and analgesic for medical purposes (Waters et al. 1977). Currently, it is widely used 

as a solvent in the industrial degreasing of metals, with secondary solvent uses in adhesive paint 

and polyvinyl chloride production (U.S. EPA, 1985). Trichloroethylene is used as a solvent in 

the textile industry, as a solvent for adhesives and lubricants, and as a low-temperature heat 

transfer fluid (IARC, 1979). Trichloroethylene is also implemented in the manufacturing of 

pesticides and other chemicals (Feldman, 1979). In 1996, the latest year tabulated, the statewide 

mean outdoor monitored concentration of trichloroethylene was approximately 0.035 ppb 

(CARB, 1999a). The annual statewide emissions from facilities reporting under the Air Toxics 

Hot Spots Act in California based on the most recent inventory were estimated to be 176,908 

pounds (CARB, 1999b)

Appendix D3 581 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

IV. Effects of Human Exposure 

An occupational study of trichloroethylene (TCE) vapor emissions in a pump room was 

conducted by Vandervort and Polnkoff (1973). Workers were an average age of 40 and had been 

employed for an average of 8 years. For 11-day shift workers, individual 8 hour time weighted 

average (TWA) TCE exposure concentrations were extrapolated from two area samples; these 

averages ranged from 170-420 mg/m3

 (32-78 ppm). Nineteen workers (including the 11 workers 

whose work areas were sampled) completed a questionnaire and reported the following 

symptoms: 73% eye irritation, 70% drowsiness, 58% heart palpitations, 58% cough, 53% 

weakness and 52% dizziness. About half of the 19 exposed workers reported that consumption 

of small amounts of alcohol outside of work resulted in changes of skin color and severe 

intoxication. One worker of the 19 reported no adverse effects from the occupational exposure. 

Nine control workers experienced none of the above symptoms. Urine samples from the 19 

exposed and 9 unexposed workers were collected before and after the work shift and examined 

for the TCE metabolites trichloroacetic acid (TCA) and trichloroethanol (TRI). TRI levels 

ranged from 4-260 mg/l and TCA levels ranged from 4-197 mg/l. Results of the urine assays 

showed a range of TCE metabolite concentrations and, therefore, confirmed that the workers 

were exposed to a variety of concentrations in their environments. 

Nomiyama et al. (1977) examined 36 trichloroethylene workers, of which 9 males and 12 

females were occupationally exposed to a constant concentration of trichloroethylene (TCE) and 

18 males were exposed to variable concentrations (duration of exposure unspecified). The 

control group consisted of 6 males and 10 females who were of similar educational, sociologic 

and economic status to the trichloroethylene workers. Researchers used urinary excretion of 

TCE metabolites as an indicator of the level of TCE exposure in the working environment; total 

excreted trichloro-compounds of 100 mg in 4 hours corresponded to 100 ppm TCE present in the 

working environment (Bardodej, 1958; Medek, 1958). Of the 36 exposed workers, 5 were 

exposed to 0-25 ppm; 14 were exposed to 25-50 ppm; 6 were exposed to 50-100 ppm; 8 were 

exposed to 100-150 ppm; and 3 were exposed to 150-200 ppm TCE. In the low exposure group, 

workers experienced mucous membrane irritation in the eyes, nose and throat, in addition to 

drowsiness, fatigue and headache. These symptoms were persistent through the higher 

concentration exposures with an increase in eye irritation, headache, fatigue, and nasal 

obstruction above 100 ppm TCE. Increases in rhinorrhea and drowsiness were seen above 50 

ppm TCE exposure. 

Kimmerle and Eben (1973) exposed 4 human subjects (3 males and 1 female) to a subacute 

regimen of 48 + 3 ppm trichloroethylene (TCE) for 4 hours a day over a period of 5 days. Levels 

of TCE and the metabolites trichlororethanol (TRI) and trichloroacetic acid (TCA) were 

determined. Trichloroethanol-blood levels were elevated immediately after exposure, and 

detection of trichloroethanol occurred up to 7 days after the last exposure. TCE-blood 

concentration increased slightly over the 5 days. Levels of urinary excreted trichloroethanol, as 

well as the TCA-concentration, increased throughout the study, with the female showing a 

significantly higher excretion of TCA. Levels of TCA were detected up to 12 days after the final 

exposure. 

Appendix D3 582 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Okawa and Bodner (1973) studied the occupational exposure of 24 electrical plant workers to 

trichloroethylene (TCE). The plant worker group consisted of 22 males and 2 females ranging in 

age from 21-52 years old. Environmental samples of TCE were collected over three days and 

yielded varying concentrations of TCE related to the task performed in certain areas (duration of 

exposure unspecified). Spray booth operators were exposed to an average of 25.3 ppm TCE (13-

40 ppm range) in addition to averages of 15.2 ppm n-propyl acetate (NPA) and 6 ppm toluene 

(TOL). Workers involved in washing board units were exposed to an average value of 39 ppm 

(6-82 ppm range) TCE. Although the workers wore respiratory protection during the washing 

procedure, the overall average of airborne TCE in this area was 48.3 ppm. In the testing area of 

the plant, researchers report that the amounts of toluene and n-propyl acetate were insignificant. 

Here, TCE levels were an average of 24.4 ppm (range = 8-44 ppm). The solder machine 

operators were exposed to an average of 44.0 ppm TCE (range = 23-87 ppm) with no NPA or 

TOL present. During the cleaning of the soldering machines, TCE levels rose to an average of 

70.5 ppm (range = 30-106 ppm). Concentrations were only at these elevated levels for 20-30 

minutes a day. Researchers note that although other agents were used in the work area, TCE was 

the only chemical found in significant amounts throughout the work area and that the levels of 

NPA and TOL were insignificant. An analysis of urinary TCE metabolites indicated that the 

workers were exposed to a time weighted average concentration of <50 ppm TCE. Three of the 

24 workers reported that they were unaffected by their working conditions, but the most 

prominent complaints consisted of 70.8% workers experiencing nausea, 54.2% headache, 33.3% 

dizziness, 25.0% fatigue, 25% nose and throat irritation, and 20.8% eye irritation. Workers 

reported that these symptoms were alleviated hours after leaving the work environment. 

Researchers collected 8 hour urine samples from 20 of the workers and from 9 controls and 

analyzed them for TCE metabolites. Results of urinary analysis showed that the controls had 

exposure to an unspecified amount of TCE. TCA levels in exposed workers were elevated from 

that of the controls and correlated to the different exposures in specific work areas. 

Phoon et al. (1984) reported on 5 cases of Stevens-Johnson syndrome (erythema multiforme 

major) with liver involvement which followed exposure to TCE. In two cases, reactions to the 

exposure began with a fever followed by an itchy rash on the face spreading over the body. 

Lesions were observed on the face, arms and in the mouth. Liver function tests were abnormal. 

One of the two developed jaundice with hepatomegaly. Case #3 developed a similar reaction 

after 5 weeks of exposure to 216-912 mg/m3

 TCE (40-170 ppm) as did case #5 after two weeks 

of exposure to 370 mg/m3

 TCE (69 ppm). Case #4 involved a 39 year old man exposed to 

<50 mg/m3

 TCE (< 9.3 ppm) for three weeks who developed the characteristic rash, lesions and 

jaundice with slight hepatomegaly. Upon returning to work over the next three weeks, he 

developed generalized erythrodermia and facial oedema, hepatosplenomegaly and liver failure 

with septicemia from which he died 14 days later. 

Stewart et al. (1974) studied the effects of subacute trichloroethylene (TCE) exposure in 

combination with alcohol consumption. Seven men exposed to 200 ppm TCE ingested 1 quart of 

beer or 90 ml of 100-proof vodka and developed red blotches on their faces 30-40 minutes later. 

These lesions enlarged with time until they reached a peak intensity, whereupon they faded. One 

subject experienced facial flush with the consumption of alcohol for three weeks after the last 

TCE exposure, while another showed flushing six weeks after the last exposure. 

Appendix D3 583 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

V. Effects of Animal Exposure

Kjellstrand et al. (1983) studied the effects of both intermittent and continuous exposures of 

various concentrations of trichloroethylene on male and female mice over a period of 30 days. 

The concentrations used range from 37-3600 ppm, and 7 of the 14 groups were continuously or 

intermittently exposed to lower concentrations of 37, 75, 150, 225 and 300 ppm TCE. 

Continuous exposure studies were conducted over a period of 30 days for exposure groups of 37, 

75, 150 and 300 ppm TCE. All groups consisted of 10 males and 10 females (except the 37 ppm 

group, consisting of 20 males and 20 females) and were compared to identical groups of airexposed controls. Liver weights increased in a non-linear fashion as the concentration level of 

TCE increased. All groups exhibited statistically significant increases in liver weights as 

compared to the controls. In both the 37 and the 75 ppm groups, the increase in females was less 

than in males. No increase in spleen weight was detected at either the 37 or 75 ppm exposure 

level. At the 37 ppm level, a slight increase in plasma butyrylcholinesterase (BuChE) activity 

(not statistically significant) was also detected. A significant increase in kidney weight was seen 

in the male 75 ppm group and was more pronounced with increasing concentration. Male mice 

in the 75 ppm group also showed statistically significant increases in BuChE activity. In the 150 

ppm group, male and female liver weight increases were statistically significant and of equal 

magnitude. A statistically significant increase was seen in the BuChE activity of the 150 ppm 

male mice. It was not until female mice were exposed to 300 ppm, that they showed slight 

increase in BuChE activity, while the males increased 3.5 times the controls. Liver weight 

increases for the 300 ppm group were close to the maximum with females showing greater 

increase than the males. Ten male and 10 female mice were continuously exposed to 150 ppm 

TCE for 30 days, but then allowed a 120 day rehabilitation period. Following rehabilitation, 

liver weights returned to levels comparable to the controls. The elevated BuChE activity 

returned to a normal level. No significant effects were seen after the period of rehabilitation. A 

continuous study was performed on 10 male and 10 female mice for 120 days at an exposure 

level of 150 ppm TCE. No further increase in liver weight occurred beyond the level reached in 

the 30 day study. Body weight gain was slightly decreased, and the same level of BuChE 

activity was seen as in the 30 day exposure. The intermittent study consisted of 30 days 

exposure to 225 ppm TCE for 16 hours a day, 7 days a week. A significant increase was seen in 

the BuChE activity of male mice, while females did not exhibit an increase in BuChE activity. 

Both males and females showed statistically significant increases in liver weight. Kidney weight 

increased in the same manner as in the continuous exposures. The authors noted that 

“extrapolation of the concentration-effect curve suggests that both liver weight and BuChE 

activities are influenced at still lower concentration.” 

Briving et al. (1986) examined neurotoxicity as a result of chronic trichloroethylene (TCE) 

inhalation exposure. Two groups of gerbils (6 in each group) were exposed to 50 or 150 ppm 

TCE for a period of 12 months. Two equivalent groups were used as controls. Two areas of the 

brain were specifically observed, the hippocampus and the posterior part of the cerebellar 

vermis. These discrete brain areas were previously shown to be sensitive towards chlorinated 

aliphatic solvents (Haglid et al., 1981). Following exposure, gerbils were decapitated and 

measurements were made of total free tissue amino acids as well as high-affinity uptake and 

release of 3H-aminobutyric acid (GABA) and 14C-glutamate. A significant increase in 

Appendix D3 584 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

glutathione was seen in the hippocampus of the 150 ppm gerbils, but amino acid levels were not 

significantly affected. In the posterior part of the cerebellar vermis, glutamate and GABA 

accumulation levels increased in a dose-dependent manner, with significant increases seen at 

both 50 and 150 ppm TCE. Evaluation of the hippocampus revealed no significant changes. The 

authors suggest that the stimulation of transport functions for GABA and glutamate may be 

triggered by the presence of the TCE metabolite, trichloroethanol. Therefore, the levels of 

GABA and glutamate are indicative of the amount of trichloroethanol from TCE in the brain. 

Kligerman et al. (1994) exposed 20 male CD rats to 0, 5, 50, or 500 ppm trichloroethylene 

(TCE) for 6 hours a day, over a period of 4 days. Groups at each concentration consisted of 5 

rats. One of the cytogenetic effects measured was peripheral blood lymphocytes (PBLs), 

abnormal with regard to sister chromatid exchanges. Also analyzed, were the cell cycle, bone 

marrow micronuclei in polychromatic erythrocytes (MN-PCEs/1000) and micronuclei in 

cytochalasin B-blocked binucleated cells (MN-BN/1000). The 5 ppm and 500 ppm exposure 

groups showed a decrease (not statistically significant) in cell cycle. In addition, the 50 ppm 

group exhibited a statistically significant decrease in cell cycle. For all concentrations, there was 

an overall increase in the PCE percentage. The number of PCEs with micronuclei also rose with 

the increasing concentrations of 50 ppm and 500 ppm TCE (not statistically significant due to 

high control values). The researchers conclude that the resulting increase of MN in exposed rats 

is indicative of aneuploidy induction as opposed to chromosomal breakage, and that the lack of 

chromosome aberrations corresponds to spindle effects such as aneuploid induction. Concurrent 

results of increased levels of leukocyte aneuploidy were also found by Konietzko et al. (1978) in 

degreasing workers occupationally exposed to TCE. 

Haglid et al. (1981) continuously exposed gerbils to 60 ppm or 320 ppm trichloroethylene (TCE) 

for 3 months. Following the exposure period, gerbils were maintained for 4 months in TCE-free 

conditions in order to observe any restoration of neuronal function. Both of the exposed groups 

as well as the control group consisted of six pairs of males and females. Brain samples were 

collected from the gerbils after the 4 month non-exposure period and used for determination of 

DNA and proteins. In order to determine areas of the brain that were sensitive to TCE, 

researchers examined biochemical and morphological changes in the hippocampus, the posterior 

part of the cerebellar vermis, and the brain stem. In addition to the biochemical tests, the 

cerebellum, brain stem, and cerebrum of two gerbils from each group, including the control, 

were used for neuropathological examination. Brain tissue from 2 gerbils in the control group 

and the 320 ppm group were examined under the electron microscope. No difference was seen 

in the body and brain weights of the exposed gerbils compared with controls. A slight but 

significant increase in soluble proteins was detected in the frontal cerebral cortex of the 60 ppm 

group, and a more significant elevation was seen in the visual cerebral cortex of both the 60 ppm 

and 320 ppm groups. In the 60 ppm group, a slight but significant decrease was seen in the 

soluble proteins of the sensory-motor cortex. Both groups exhibited significant decreases in 

levels of soluble proteins in the hippocampus, the brain stem, and in the posterior part of the 

cerebellar vermis. Soluble protein levels in the cerebellar hemisphere and anterior part of the 

vermis of gerbils in both exposed groups did not differ from the controls. The 320 ppm group 

showed significantly increased DNA levels in the posterior part of the sensory motor cortex and 

cerebellar vermis. The glial cytoplasmic protein (S 100 fraction) level of the 60 ppm group was 

decreased in the frontal and visual cerebral cortex, but increased in the posterior part of the 

Appendix D3 585 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

cerebellar hemisphere and the sensory-motor cortex. However, only a slight decrease of S 100 

protein was observed in the visual cerebral cortex of 320 ppm exposed gerbils. The most notable 

S 100 increase occurred in the hippocampus, brain stem and the posterior part of the cerebellar 

vermis, indicating that either the glial cells were directly affected or that damage to surrounding 

neuronal cells caused an indirect response. There was an increase in DNA in the posterior part 

of the cerebellar vermis in the exposed gerbils, suggesting that TCE induced astroglial cell 

mitosis. Light microscopy revealed shrinkage of cell bodies and axon swelling occurred in 

various parts of the brain. The electron microscopy performed on control and 320 ppm brain 

tissues revealed increased levels of filament bundles in the cytoplasm of some Purkinje and 

Golgi cell perikarya, lysosomes, myelin bodies and lipid containing lysosomal structures in the 

exposed gerbils. Unique arrangements of filament bundles were seen in Purkinje and Golgi cell 

dendrites of the exposed group. A significant decrease in the number of microtubules was 

observed as well as a decrease in the number of synaptic vesicles in the granular layer. Also, the 

granular layer had decreased maximal nerve cell surface area. Nerve cells were affected by the 

exposure as several types were reduced in size with fewer organelles and more lysosomes and 

myelin bodies. Many axons and dendrites had reduced numbers of microtubles, and there were 

filament bundles observed that were not present in the controls. Lysosomal structures were 

increased in the synaptic terminals. 

Kimmerle and Eben (1973) performed a subchronic study on 20 male rats for a period of 14 

weeks. Rats were exposed to a mean concentration of 55.0 +4 ppm trichloroethylene (TCE) for 

8 hours a day, 5 days a week. The control group consisted of 20 rats who in similar inhalation 

chambers under similar conditions to that of the exposed rats. Ten exposed rats were analyzed 

for TCE metabolite excretion on a daily basis. Blood levels of trichloroacetic acid (TCA), 

trichloroethanol (TRI) and chloral hydrate (CH) were measured during the 2nd, 3rd, 4th, 6th, 9th 

and 14th weeks. Weekly measurements of body weights were recorded. Macroscopic 

examinations were performed on the thyroid gland, heart, lungs, liver, kidneys, testes and adrenal 

glands. Hematological evaluations, liver function tests, and renal function tests were also 

conducted following exposure. Urinary levels of TRI varied individually among the rats, but a 

continuous increase in TRI was observed through the 10th week. TCA levels remained fairly 

constant throughout the duration of the experiment. TCE was not detectable in the blood or the 

tissues of exposed rats. Although liver and renal function tests did not reveal abnormalities, 

there was an increase in the liver weights of the exposed rats. The weights of the other organs 

examined were similar to the controls. 

Norpoth et al. (1974) observed an increase in liver cytochrome P450 activity in 9 rats exposed to 

50 ppm trichloroethylene for 28 days, compared with 9 control rats. 

Appendix D3 586 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level 

Study Vandervort and Polnkoff (1973) 

Study population 19 workersand 9 controls 

Exposure method Discontinuous occupational inhalation exposure 

Critical effects Drowsiness, fatigue, headache, and eye irritation 

LOAEL 32 ppm (170 mg/m3

) in the heavy assembly area 

NOAEL Not observed 

Exposure continuity 8 hours a day (10 m3

/day occupational inhalaton 

rate), 5 days a week 

Exposure duration 8 years 

Average occupational exposure 11.4 ppm for LOAEL group (32 x 10/20 x 5/7) 

Human equivalent concentration 11.4 ppm for LOAEL group 

LOAEL uncertainty factor 10 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Inhalation reference exposure level 0.1 ppm (100 ppb; 0.6 mg/m3

; 600 μg/m3) 

The Vandervort and Polnkoff (1973) study accounted for 8 years of human occupational 

exposure to TCE vapors. Sensitive, non-specific neurotoxicological endpoints were exhibited by 

a majority of those workers exposed. Although the time-weighted averages (TWAs) included a 

wide range of concentrations, the TWA of 32 ppm (170 mg/m3

) was shown to contribute to the 

high incidence (52 - 73%) of adverse effects experienced by the workers. Many of the 

symptoms reported by the workers may have been due to short-term fluctuations in the 

concentrations in the workplace. The symptoms were not reported separately for the various 

TWAs, therefore, the lowest TWA (32 ppm) was chosen as a LOAEL. Uncertainty includes the 

small number of workers studied, the limited extent of the effects mentioned, and the lack of a 

NOAEL. Strengths include the use of human data, the demonstration of a dose-response 

relationship, and exposure estimates correlated with urinary excretion measurements. 

This study was the best chronic account of the non-carcinogenic effects of TCE on humans, but 

several other studies show similar results. Nomiyama et al. (1977) found similar endpoints of 

drowsiness, fatigue and eye irritation in 36 workers occupationally exposed to trichloroethylene. 

Okawa et al. (1973) also saw non-specific neurological endpoints in 24 electrical plant workers 

who were similarly exposed to TCE. 

For comparison with the proposed REL of 100 ppb based on human studies, the LOAEL of 50 

ppm trichloroethylene obtained by Briving et al. (1986) in gerbils exposed continuously for 12 

months was used to estimate a REL based on animal data. Use of a LOAEL UF of 3, a 

subchronic UF of 1, an interspecies UF of 10, and an intraspecies UF of 10 resulted in an 

estimated REL of 200 ppb for trichloroethylene. 

Appendix D3 587 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for trichloroethylene include the use of human exposure data 

from workers exposed over a period of years. Major areas of uncertainty are the lack of 

reproductive and developmental toxicity studies and the lack of observation of a NOAEL. 

VIII. References 

Bardodej Z, and Krivucova M. 1958. Determination of trichloroethyl alcohol (urochloralic acid) 

as an exposition test on workers with trichloroethylene (in Czech.). Cs. Hyg. 3:268-272. [as cited 

by Nomiyama and Nomiyama, 1977]

Briving C, Jacobson I, Hamberger A, Kjellstrand P, Haglid K, and Rosengren L. 1986. Chronic 

effects of perchloroethylene and trichloroethylene on the gerbil brain amino acids and 

glutathione. Neurotoxicology 7(1):101-108. 

CARB. 1999a. California Air Resources Board. Toxics Air Quality Data. Substance Chooser. 

Trichloroethylene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm 

CARB. 1999b. Air toxics emissions data collected in the Air Toxics Hot Spots Program 

CEIDARS Database as of January 29, 1999. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Fan A. 1988. Trichloroethylene: Water contamination and health risk assessment. Rev. Environ. 

Contam. Toxicol. 101:55-92. 

Feldman R. 1979. Intoxications of the nervous system. Handbook of Clinical Neurology 36:457-

464. [as cited by Juntunen 1986] 

Haglid K, Briving C, Hansson H, Rosengren L, Kjellstrand, Stavron D, Swedin U, and Wronski 

A. 1981. Trichloroethylene: Long-lasting changes in the brain after rehabilitation. 

Neurotoxicology 2:659-673. 

IARC. 1979. Trichloroethylene. IARC Monographs on the Evaluation of the Carcinogenic Risk 

of Chemicals to Humans. Vol. 20. Lyon:IARC, pp. 545-572. 

Juntunen J. 1986. Occupational toxicology of trichloroethylene with special reference to 

neurotoxicity. New Concepts and Developments in Toxicology. PL Chambers, P Gehring, and F 

Sakai, eds. Amsterdam: Elsevier. pp. 189-200. 

Kimmerle G, and Eben A. 1973. Metabolism, excretion and toxicology of trichloroethylene after 

inhalation: 1. Experimental exposure on rats. Arch. Toxicol 30:115-126. 

Appendix D3 588 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Kimmerle G, and Eben A. 1973. Metabolism, excretion and toxicology of trichloroethylene after 

inhalation: 2. Experimental human exposures. Arch. Toxicol. 30:127-138. 

Kjellstrand P, Holmquist B, Alm P, Kanje M, Romare S, Jonsson I, Mansson L, and Bjerkemo 

M. 1983. Trichloroethylene: Further studies of the effects on body and organ weights and plasma 

butyrylcholinesterase activity in mice. Acta. Pharmacol. Toxicol. 53:375-384. 

Kligerman A, Bryant M, Doerr C, Erexson G, Evansky P, Kwanyuen P, and McGee J. 1994. 

Inhalation studies of the genotoxicity of trichloroethylene to rodents. Mutat. Res. 322:87-96. 

Konietzko H, Haberlandt W, Heilbronner H, Reill G, and Weichardt H. 1978. Cytogenetische 

Untersuchungen an Trichlorathylen-Arbeitern. Arch. Toxicol. 40:201-206. 

Medek V. 1958. The relation between trichloroethanol and trichloroacetic acid in the urine and 

trichloroethylene in the atmosphere. Pracov. Lek. 10:135-138. [as cited by Nomiyama and 

Nomiyama, 1977] 

Nomiyama K, and Nomiyama H. 1977. Dose-response relationship for trichloroethylene in man. 

Int. Arch. Occup. Environ. Health 39:237-248. 

Norpoth K, Witting U, and Springorum M. 1974. Induction of microsomal enzymes in the rat 

liver by inhalation of hydrocarbon solvents. Int. Arch. Arbeitsmed 33:315-321. 

Okawa M, and Bodner A.1973. NIOSH: Health hazard evaluation/toxicity determination. 

Western Electric Company, Inc. report 72-74. 

Phoon W, Chan M, Rajan V, Tan K, Thirumoorthy T, and Goh C. Stevens-Johnson syndrome 

associated with occupational exposure to trichloroethylene. Contact Dermatitis 10:270-276. 

Stewart R, Hake C, and Peterson J. 1974. “Degreasers’ Flush;” Dermal response to 

trichloroethylene and ethanol. Arch. Environ. Health 29:1-5. 

U.S. EPA. 1985. U.S. Environmental Protection Agency. Assessment of trichloroethylene as a 

potentially toxic air pollutant; proposed rule. Federal Register; Part IV 50(246):52422-52425. 

Vandervort R, and Polnkoff P. 1973. NIOSH: Health hazard evaluation/toxicity determination. 

Dunham-Bush, Inc. report 72-34. 

Waters E, Gerstner H, and Huff H. 1977. Trichloroethylene. I. An overview. J. Toxicol. Environ. 

Health 2:271-307. [as cited by Fan, 1988] 

Appendix D3 589 Trichloroethylene 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

CHRONIC TOXICITY SUMMARY 

TRIETHYLAMINE

(diethylaminoethane; ethanamine; N,N-diethylethanamine) 

CAS Registry Number: 121-44-8 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 200 μg/m3

 (40 ppb) 

Critical effect(s) Eye effects in rats and humans 

Hazard index target(s) Eyes 

II. Physical and Chemical Properties (Nelson and Bull, 1990, except as noted) 

Description Colorless, volatile liquid 

Molecular formula C6H15N 

Molecular weight 101.9 g/mol 

Density 0.726 g/cm3

 @ 25°C 

Boiling point 89.3°C 

Melting point −114.7°C (CRC, 1994) 

Vapor pressure 400 torr @ 31.5°C 

Odor threshold 480 ppb (Amoore and Hautala, 1983) 

Solubility soluble in acetone, benzene and chloroform 

Conversion factor 1 ppm = 4.14 mg/m3

 25°C 

III. Major Uses or Sources 

Triethylamine (TEA) is primarily used as a cross-linking catalyst in the production of 

polyurethane foam used in the manufacture of cores for metal castings (Albrecht and 

Stephenson, 1988). Triethylamine is also used as a catalyst for epoxy resins, and as a corrosion 

inhibitor for polymers (Nelson and Bull, 1990). TEA is one of the amines emitted from cattle 

feedlots (Mosier et al., 1973). In the gas phase TEA can react with nitric acid to form amine 

nitrates that become part of atmospheric particulates. The annual statewide industrial emissions 

from facilities reporting under the Air Toxics Hot Spots Act in California based on the most 

recent inventory were estimated to be 4152 pounds of triethylamine (CARB, 2000). 

IV. Effects of Human Exposures 

Acute, high level triethylamine exposures (20 mg/m3

 (4.8 ppm) for 8 hours) resulted in reversible 

ocular effects that included corneal swelling and halo vision in 4 out of 5 volunteer subjects 

(Akesson et al., 1988). Jarvinen et al. (1999) reported exposure chamber studies of ocular 

Appendix D3 590 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

responses to TEA in volunteer subjects, who were industrial workers exposed to TEA during the 

course of their normal jobs (but had good general and ocular health). Four people were exposed 

for 4 hours to 3.0, 6.5, or 40.6 mg/m3

 triethylamine. Corneal thickness was measured by 

ultrasonography, clinical observations were recorded using ocular microscopy, and statistical 

analysis of the size, density and distribution of corneal endothelial cells was performed by 

automated analysis of photographs. Visual acuity and contrast sensitivity were evaluated using 

test charts. After exposure to 40.6 mg/m3

 there was a marked edema in the corneal epithelium, 

and subepithelial microcysts. However, corneal thickness increased only minimally. Vision was 

blurred in all subjects and visual acuity and contrast sensitivity decreased in three of the four. 

After exposure to 6.5 mg/m3

 two subjects experienced symptoms (e.g., blurred vision), and 

contrast sensitivity decreased in three of the four. There were no symptoms or decreases in 

contrast sensitivity after exposure to 3.0 mg/m3

 triethylamine for 4 hours. 

A medical examination of 19 workers exposed at a polyurethane foam production plant to a 

time-weighted average concentration of 13 mg/m3

 (3.1 ppm) TEA showed reversible corneal 

edema in 5 workers (Akesson et al., 1986). Peak concentrations were up to twice the timeweighted average level. A questionnaire on self-reported symptoms of visual disturbances 

revealed repeated occurrences of temporary eye irritation and “foggy vision.” Small quantities 

of dimethylethanolamine, toluene diisocyanate, and methylene diphenyl isocyanate were also 

present in the workplace atmosphere. 

Jarvinen and Hyvärinen (1997) reported loss of visual acuity and contrast sensitivity in 41 

foundry workers (core makers) exposed to TEA. Concentrations of TEA were reported to have a 

mean of 46 mg/m3

 and a maximum of 486 mg/m3, but were highly variable with numerous large 

excursions above a background of about 20 mg/m3

 during a two-hour period of continuous 

monitoring. It is therefore difficult to determine an effect level for the observed symptoms. 

Jarvinen (1998) also reported that cold box core makers exposed to TEA had a somewhat 

increased incidence of mild headaches. 

V. Effects of Animal Exposures 

Lynch et al. (1990) exposed male and female Fischer 344 rats to triethylamine at concentrations 

of 0, 25, or 247 ppm (0, 103.4, or 1022.2 mg/m3

) for 6 hours/day, 5 days/week. Groups of rats 

were necropsied at approximately 30, 60, and 120 days of exposure. The last corresponds to an 

elapsed time of 28 weeks. Endpoints examined included gross and histopathological 

examination of all major organs, including the lungs, nasal passages, and eyes. Clinical enzyme 

and nitrogen levels (BUN, ALT, AST, CPK, and creatinine), and hematological values 

(hemoglobin, RBC count) were also measured. No gross or histological effects in any organ 

were observed in any group. Clinical and hematological parameters were unchanged with 

exposure. However, all rats exposed to 247 ppm TEA manifested irritation. “At 247 ppm TEA 

the rats kept their eyes closed and noses buried in their fur during the entire exposure period.” 

Thus 247 ppm is a LOAEL and 25 ppm is a NOAEL for eye and nose irritation in the rat. 

In a short-term study by the same authors for Virginia Chemicals (1987), necrotizing 

inflammation of the nasal cavity, metaplasia of the trachea, and thymic atrophy were observed 

Appendix D3 591 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

after exposure to 1000 ppm (4140 mg/m3

) triethylamine 6 hours per day for 10 days. Two of 

five males and one of five females died from pulmonary edema after the seventh day. Thymic 

atrophy was noted in 7 out of 10 animals, and all animals exhibited necrotizing inflammation in 

the nasal epithelium. 

Rabbits (6-12 per group), exposed to 48 or 100 ppm (199 or 414 mg/m3

) triethylamine for 7 

hours/day, 5 days/week, for 6 weeks, showed concentration-dependent pathology in the eyes, 

lungs, liver, kidney, and heart (Brieger and Hodes, 1951). The eyes showed multiple punctate 

erosions of the corneal epithelium, and corneal edema at 48 ppm. Lung lesions included 

thickening of vascular walls; liver lesions included parenchymal degeneration. Overall the 

lesions in the 48 ppm group were less severe than those seen in the 100 ppm group. No control 

animals were included in this study, nor were the incidences of histologic effects among the 

exposed animals reported. All animals did survive the exposures. The lesser effects at 48 ppm 

in the rabbit (compared to those at 100 ppm) are consistent with the findings of Lynch et al. 

(1990) where 25 ppm was a NOAEL in the rat for eye and nose irritation. 

A chronic 3-generation reproductive study in rats (10/sex/group) was inconclusive due to 

excessive mortality in controls (Davison et al., 1965). In this study, rats were exposed to 0, 2, or 

200 ppm triethylamine. The third generation of the 200 ppm group was changed to 500 ppm 

since no effects were noted in the 200 ppm group. Exposure of this group to 500 ppm resulted in 

decreased body weight and decreased water consumption. 

VI. Derivation of Reference Exposure Level 

Study Lynch et al., 1990; Brieger and Hodes, 1951 

Study population Rats; rabbits 

Exposure method Discontinuous whole-body inhalation 

Critical effects Eye irritation; lung and liver toxicity 

LOAEL 48 ppm (Brieger and Hodes, 1951) 

NOAEL 25 ppm (Lynch et al., 1990) 

Exposure continuity 6 or 7 hours/day, 5 days/week 

Exposure duration 28 weeks; 6 weeks 

Average experimental exposure 4.46 ppm for NOAEL group (25 x 6/24 x 5/7) 

Human equivalent concentration 4.46 ppm (18.5 mg/m3

) for NOAEL group 

LOAEL uncertainty facto 1 

Subchronic uncertainty factor 1 (NOAEL is based on a 28 wk study in rats) 

Interspecies uncertainty factor 10 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 100 

Reference exposure level 0.04 ppm (40 ppb; 200 μg/m3

) 

The U.S. EPA (1995) based its Reference Concentration (RfC) of 7 μg/m3

 (2 ppb) for 

triethylamine on Lynch et al. (1990) but included a Modifying Factor (MF) of 10 for “database 

deficiencies” - “lack of developmental and reproductive effects, and of appropriate data in a 

second species.” The criteria for use of modifying factors are not well specified by U.S. EPA. 

Appendix D3 592 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

Such modifying factors were not used by OEHHA. In addition OEHHA applied a subchronic 

UF of 1 since 24 male and 24 female rats in the NOAEL group were exposed to 25 ppm TEA for 

28 weeks, while USEPA used a subchronic UF of 10. U.S. EPA considered 247 ppm to be a 

NOAEL. However, the Lynch et al. (1990) study indicates that the animals closed their eyes and 

buried their noses in their fur, likely to prevent the irritant effects of TEA on their eyes and 

respiratory tract. Thus adverse effects occurred at 247 ppm, although they could be considered 

repeated acute effects. Brieger and Hodes (1951) observed adverse effects in the eyes, lungs, 

and livers of rabbits after six weeks of discontinuous exposure to TEA. Thus 48 ppm is a 

LOAEL in this study. 

For comparison, the five affected workers studied by Akesson et al. (1986) showed symptoms at 

12-13 mg/m3

 TEA, which is equivalent to 4.5 mg/m3 continuous exposure. (Other tasks were at 

4-5 mg/m3

 TEA or 1.6 mg/m3 continuous exposure.) Selection of a LOAEL UF of 3 (26% 

incidence of a reversible effect), a subchronic UF of 1 since the workers had been employed for 

9.7 years (range = 4-11), and an intraspecies UF of 10 results in an estimated REL of 200 μg/m3

based on human data. These workers experienced some short-term peak exposures to TEA and 

were also exposed to dimethylethanolamine (<0.1 mg/m3

), toluene diisocyanate, and methylene 

diphenyl isocyanate. 

VII. Data Strengths and Limitations for Development of the REL 

The major strengths of the triethylamine REL are the observation of a NOAEL in a controlled 

exposure experiment and finding of the same adverse effect, eye irritation, in humans and 

animals. The major weaknesses are the minimal amount of adequate human health effects 

information, the lack of dose-response data in a single experiment, and the lack of long-term 

exposure data. 

VIII. Potential for Differential Impacts on Children's Health

There is no direct evidence in the literature to quantify a differential effect of TEA in infants and 

children. However, it is a respiratory irritant and thus has the potential to exacerbate asthma. In 

addition, other alkylamines are known to be associated with occupational asthma (Bernstein et

al., 1999). There is some concern that TEA could have a similar effect. 

IX. References 

Akesson B, Bengtsson M, and Floren I. 1986. Visual disturbances by industrial triethylamine 

exposure. Int. Arch. Occup. Environ. Health 57:297-302. 

Akesson B, Vinge E, and Skerfving S. 1988. Pharmacokinetics of triethylamine and 

triethylamine-N-oxide in man. Toxicol. Appl. Pharmacol. 100:529-538. 

Appendix D3 593 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

Albrecht WN, and Stephenson RL. 1988. Health hazards of tertiary amine catalysts. Scand. J. 

Work Environ. Health 14:209-219. 

Amoore JE, and Hautala E. 1983. Odor as an aid to chemical safety: Odor thresholds compared 

with threshold limit values and volatilities for 214 chemicals in air and water dilution. J. Appl. 

Toxicol. 3(6):272-290. 

Bernstein IL, Chan-Yeung M, Malo J-L, Bernstein DI (eds.). 1999. Asthma in the Workplace. 

2nd ed. New York: Marcel Dekker, pp. 524-525. 

Brieger H, and Hodes WA. 1951. Toxic effects of exposure to vapors of aliphatic amines. AMA 

Arch. Ind. Hyg. Occup. Med. 3(3):287-291. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

CRC. 1994. CRC Handbook of Chemistry and Physics, 75th edition. Lide DR, ed. Boca Raton, 

FL: CRC Press Inc. 

Davison RR, Hood DW, and McMullen B. 1965. Toxicity of triethylamine to albino rats. 

Prepared for the Office of Saline Water through Texas A&M University (final report reference 

65-4F; Office of Saline Water Contract No. 14-01-0001-282). OTS # 303940. Doc. # 

86870000536. Fiche # 0513614. [as cited in: U.S.EPA’s Integrated Risk Information System 

(IRIS) database. 1995.] 

Jarvinen P. 1998. Headache and blood pressure among triethylamine-exposed foundry workers. 

Occup. Med. (Lond.). 48(2):113-117. 

Jarvinen P, Engstrom K, Riihimaki V, Ruusuvaara P, and Setala K. 1999. Effects of 

experimental exposure to triethylamine on vision and the eye. Occup. Environ. Med. 56(1):1-5. 

Jarvinen P, and Hyvarinen L. 1997. Contrast sensitivity measurement in evaluations of visual 

symptoms caused by exposure to triethylamine. Occup. Environ. Med. 54(7):483-6. 

Lynch DW, Moorman TR, Lewis P, Stober P, Hamlin R, and Scheuler RL. 1990. Subchronic 

inhalation of triethylamine vapor in Fischer-344 rats: Organ system toxicity. Toxicol. Ind. 

Health. 6(3/4):403-414. 

Mosier AR, Andre CE, Viets FG Jr. 1973. Identification of aliphatic amines volatilized from cattle feedyard. 

Environ. Sci. Technol. 7:642-644.

Nelson MA, and Bull RJ. 1990. Triethylamine. In: Ethel Browning’s Toxicity and Metabolism of 

Industrial Solvents. Vol. 2. Buhler DR, and Reed DJ (eds). Amsterdam: Elsevier, pp. 129-133. 

U.S.EPA. 1995. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration for chronic inhalation exposure (RfC) for 

triethylamine. 

Appendix D3 594 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels September 2002 

Virginia Chemicals. 1987. Pathologic findings in Fischer 344 rats exposed by inhalation to 

allylamine, ethylamine, diethylamine, and triethylamine with cover letter dated 042484. OTS # 

308080. Doc # 86-870000813. Fiche # 0515251. 

Appendix D3 595 Triethylamine 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

CHRONIC TOXICITY SUMMARY 

VINYL ACETATE 

(1-acetoxyethylene; acetic acid, vinyl ester; acetic acid, ethenyl ester; VAC; vinyl A monomer; 

ethenyl ethanoate) 

CAS Registry Number: 108-05-4 

I. Chronic Toxicity Summary 

Inhalation reference exposure level 200 μg/m3 (50 ppb) 

Critical effect(s) Nasal epithelial lesions in rats and mice 

Hazard index target(s) Respiratory system 

II. Physical and Chemical Properties (HSDB, 1994) 

Description Colorless liquid 

Molecular formula C4H6O2

Molecular weight 86.09 g/mol 

Density 0.932 g/cm3

 @ 20°C 

Boiling point 72.7° C 

Melting point −93.2°C 

Vapor pressure 115 torr @ 25°C 

Solubility Slightly soluble in water, soluble in ethane, acetone, 

chloroform; >10% soluble in ethanol and benzene 

Conversion factor 1 ppm = 3.52 mg/m3

 @ 25°C 

III. Major Uses and Sources 

The major use of vinyl acetate monomer is in the manufacture of polyvinyl and vinyl acetate 

copolymers, which are used in water-based paints, adhesives, paper coatings, and applications 

not requiring service at extreme temperatures (HSDB, 1994). It is also used in safety glass 

interlayers and in hair sprays (HSDB, 1994). In the atmosphere vinyl acetate breakdown can 

result in formation of acetaldehyde. The annual statewide industrial emissions from facilities 

reporting under the Air Toxics Hot Spots Act in California based on the most recent inventory 

were estimated to be 3855 pounds of vinyl acetate (CARB, 2000).

IV. Effects of Human Exposures

Deese and Joyner (1969) conducted an occupational study of 21 chemical workers with a mean 

length of employment of 15.2 years and exposed to a time-weighted average of 8.6 ppm (30.3 

mg/m3

) VA. No adverse effects were noted following chest x-ray, electrocardiogram, blood 

Appendix D3 596 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

chemistry, and urinalysis. The control group (sample size unspecified) consisted of workers in 

units not exposed to VA. Deese and Joyner (1969) also showed intolerable eye irritation in 3 out 

of 3 subjects exposed for an unspecified extended period of time to 21.6 ppm (76 mg/m3

) VA. 

Upper respiratory irritation was also experienced by a majority of 5 subjects. Odor was detected 

at 0.4 ppm (1.4 mg/m3

) in 3 out of 3 subjects. 

V. Effects of Animal Exposures 

A 104-week inhalation study in rats and mice (90/sex/group) was conducted using concentrations 

of 0, 50, 200, or 600 ppm (0, 176, 704, or 2113 mg/m3

) vinyl acetate (VA) (Owen, 1988). The 

study was later published by Bogdanffy et al. (1994). Exposures were for 6 hours/day, 5 

days/week. Histology was performed on all major organs. There was no mortality resulting 

from these exposures. A close examination of the effects of VA on the lung and nasal passages 

showed significant lesions in the nasal cavity, bronchi, and lungs of rats exposed to 600 ppm 

VA. Lesions included olfactory epithelial metaplasia/atrophy (see table below) and nest-like 

epithelial folds in the nasal cavity, exfoliation of bronchial epithelium, fibrous intraluminal 

projections in the bronchi, and pigmented histiocyte accumulation in the lungs. Body weight 

gain of rats was significantly decreased in the 600 ppm VA group. Rats treated with 200 ppm 

VA showed some evidence of epithelial atrophy and metaplasia in the nasal cavity. No effects 

were observed in the rats exposed to 50 ppm VA. 

 Number of male rats with olfactory epithelial atrophy (Bogdanffy et al. 1994) 

VA (ppm) N in group Very slight Slight Moderate Severe 

0 58 0 0 0 0 

50 59 1 2 0 0 

200 60 4 47*** 2 0 

600 60 0 7* 33*** 10*** 

* p<0.05; ***p<0.001 by Fisher’s pair-wise test compared to control group 

Mice also exhibited significant histological lesions in the respiratory tract following exposure to 

200 ppm VA or greater. The lesions included atrophy of the olfactory epithelium and 

submucosal gland. At 600 ppm, hyperplasia of the trachea was observed, in addition to 

exfoliation/flattening of the bronchial epithelium and decreased body weight gain. Relative 

brain and kidney weights were increased in the 600 ppm group at the end of the study, and 

absolute liver, heart and kidney weights were also significantly elevated. No adverse effects 

were observed in the 50 ppm group. 

A 13-week study on the effects of VA in mice was conducted by Owen (1980a). Mice 

(10/sex/concentration) were exposed to 0, 50, 200, or 1000 ppm (0, 176, 704, or 3520 mg/m3

) 

VA for 6 hours/day, 5 days/week for 13 weeks. A concentration-dependent increase in the 

incidence of diffuse rhinitis, beginning at the 200 ppm concentration, was detected using 

histopathological examination. Focal pneumonitis was observed in the 1000 ppm treatment 

group. No adverse effects were seen in the 50 ppm treatment group. An identical study in rats 

was also conducted by Owen (1980b). In this study, body weight gain was significantly reduced 

Appendix D3 597 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

in male and female rats exposed to 1000 ppm VA. An increase in the incidence of mild 

histiocytic alveolitis was observed in the 1000 ppm group. 

Irvine (1980) conducted a study on the developmental toxicity of VA in rats. Groups of 24 

pregnant female rats were exposed to 0, 52, 198, or 1004 ppm (0, 182, 696, or 3533 mg/m3

) VA 

for 6 hours/day on days 6-15 of gestation. Significant maternal toxicity, as measured by reduced 

weight gain from day 10 through day 15, was observed in animals exposed to 1004 ppm. 

Fetotoxicity, as measured by reduced crown-rump length, reduced body weight, and increased 

incidence of ossification defects in the sternebrae and occipital regions, was observed in the 1004 

ppm group. No maternal or fetal effects were seen at the lower two VA treatments. 

In another developmental toxicity study, groups of 23-24 Crl:CD(SD)BR rats were given 0, 200, 

1000, or 5000 ppm VA in drinking water or exposed 6 hr/day to 0, 50, 200, or 1000 ppm VA on 

gestation days 6-15 of gestation. The authors (Hurtt et al., 1995) estimated that the doses by 

both routes were approximately 0, 25, 100, or 500 mg/kg/day. VA in the drinking water 

produced no evidence of maternal or developmental toxicity at any dose. In the inhalation study, 

maternal toxicity was indicated by a reduction in weight gain of dams exposed to 1000 ppm. 

Fetal toxicity was evident by a significant decrease in mean fetal weight and mean crown-rump 

length in fetuses from the 1000-ppm group and by a significant increase in the incidence of 

minor skeletal alterations (especially delayed ossification) in fetuses from dams exposed to 1000 

ppm VA. These results indicated to the authors that VA is not uniquely toxic to the conceptus. 

The NOAEL was greater than 5000 ppm via the drinking water and 200 ppm by the inhalation 

route. 

Appendix D3 598 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VI. Derivation of Chronic Reference Exposure Level

Study Bogdanffy et al., 1994 

Study population Male and female Sprague-Dawley rats and CD-1 

mice (90/sex/group) 

Exposure method Discontinuous inhalation exposures 

 (0, 50, 200, or 600 ppm) over 104 weeks 

Critical effects Histological lesions of the nasal epithelium 

LOAEL 200 ppm 

NOAEL 50 ppm 

Exposure continuity 6 hours/day, 5 days/week 

Exposure duration 104 weeks 

Average experimental exposure 8.9 ppm for NOAEL group (50 x 6/24 x 5/7) 

Human Equivalent Concentration 

(HEC) 

1.4 ppm for NOAEL group (RGDR = 0.15 based 

on a gas with respiratory effects in both rats 

and mice) 

LOAEL uncertainty factor 1 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 3 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.05 ppm (50 ppb, 0.2 mg/m3

, 200 μg/m3) 

The chronic REL is the U.S. EPA RfC (U.S. EPA, 1995) for vinyl acetate. Acetaldehyde, a 

hydrolysis product of vinyl acetate, was present in the Owen (1988) study at a concentration of 

49 ppm (89 mg/m3

). The duration-adjusted concentration for acetaldehyde was 16 mg/m3, 

whereas the NOAEL for histological lesions in rats by Appleman et al. (1982) was 48.75 mg/m3

acetaldehyde. Therefore, the concentration of acetaldehyde was not considered to account for 

significant irritation in the Owen (1988) study. OEHHA accepted the U.S. EPA analysis. 

For comparison, Irvine (1980) obtained a NOAEL of 198 ppm for fetotoxicity in rats exposed 6 

hours/day on days 6-15 of gestation. This is equivalent to 50 ppm continuous exposure during 

development. Multiplying by an RGDR of 1 and dividing by a total UF of 30 (3 for interspecies 

and 10 for intraspecies) results in a REL estimate based on fetotoxicity of 1.7 ppm. The results 

of Hurtt et al. (1995) also yield an estimate of 1.7 ppm. 

VII. Data Strengths and Limitations for Development of the REL 

The strengths of the inhalation REL for vinyl acetate include the availability of controlled 

exposure lifetime inhalation studies in multiple species at multiple exposure concentrations and 

with adequate histopathological analysis, and the observation of a NOAEL. The major area of 

uncertainty is the lack of adequate human exposure data. 

Appendix D3 599 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

VIII. Potential for Differential Impacts on Children's Health

Since the chronic REL (0.05 ppm) is lower than the comparison estimate based on 

developmental effects (1.7 ppm), the REL is likely to be protective of children. However, there 

is no direct evidence in the literature to quantify a differential effect of vinyl acetate in infants 

and children relative to adults. 

IX. References 

Appleman LM, Woutersen RA, and Feron VJ. 1982. Inhalation toxicity of acetaldehyde in rats. I. 

Acute and subacute studies. Toxicology 23:293-307. 

Beems RB. 1988. Report No. V 88.133: Histopathology of the respiratory tract of mice used in a 

104-week inhalation study with vinyl acetate. (TNO-CIVO Institutes, April 1988). 

Bogdanffy MS, Dreef-van Der Meulen HC, Beems RB, Feron VJ, Cascier TC, Tyler TR, 

Vinegar MB, and Rickard RW. 1994. Chronic toxicity and oncogenicity inhalation study with 

vinyl acetate in the rat and mouse. Fundam. Appl. Toxicol. 23:215-229. 

CARB. 2000. California Air Resources Board. California Emissions Inventory Development and 

Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000. 

Deese DE, and Joyner RE. 1969. Vinyl acetate: A study of chronic human exposure. Am. Ind. 

Hyg. Assoc. J. 30:449-457. 

Dreef-van der Meulen HC. 1988. Report No. V 88.033/270836: Histopathology of the 

respiratory tract of rats used in a 104 week inhalation study with vinyl acetate: Revised version. 

(TNO-CIVO Institutes, October 1988). 

HSDB. 1994. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD 

(TOMES® CD-ROM version). Denver, CO: Micromedex, Inc. (Edition expires 11/31/94). 

Irvine LFH. 1980. Vinyl acetate oral and inhalation teratology studies in the rat. Report prepared 

by Hazelton Laboratories Europe Ltd., Harrogate, England for the Society of the Plastics 

Industry, Inc., New York. Report No. 2195-51/6&7. 

Owen PE. 1980a. Vinyl acetate: 3 month inhalation toxicity study in the mouse. Report prepared 

by Hazelton Laboratories Europe Ltd., Harrogate, England for the Society of the Plastics 

Industry, Inc., New York. Report No. 2303-51/5. 

Owen PE. 1980b. Vinyl acetate: 3 month inhalation toxicity study in the rat. Report prepared by 

Hazelton Laboratories Europe Ltd., Harrogate, England for the Society of the Plastics Industry, 

Inc., New York. Report No. 2286-51/5. 

Appendix D3 600 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels December 2001 

Owen PE. 1988. Vinyl acetate: 104 week inhalation combined chronic toxicity and 

carcinogenicity study in the rat and mouse. Report prepared by Hazelton Laboratories Europe 

Ltd., North Yorkshire, England, for the Society of the Plastics Industry, Inc., NY. Report No. 

4661-51/17a. December 1987. [as cited in U.S. EPA, 1995.] 

U.S.EPA. 1995. U.S. Environmental Protection Agency. Integrated Risk Information System 

(IRIS) database. Reference concentration (RfC) for vinyl acetate. 

Appendix D3 601 Vinyl acetate 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

CHRONIC TOXICITY SUMMARY 

XYLENES 

(Xylol or commercial xylenes (mixture of 60-70% m- and remaining percentage is mix of o- and 

p- xylenes), technical grade xylenes or mixed xylenes (20% o-xylene, 40% m-xylene, 20% pxylene, 20% ethyl benzene, and traces of toluene and C9 aromatics), o-xylene (1,2-

dimethylbenzene or 2-xylene), m-xylene (1,3-dimethylbenzene or 3-xylene), p-xylene (1,4-

dimethylbenzene or 4-xylene), also noted as methyltoluene, benzene-dimethyl, dimethylbenzene) 

CAS Registry Numbers.: 1330-20-7 (technical mixture of o-, p-, and m-xylene); 

95-47-6 (o-xylene); 108-38-3 (m-xylene); 106-42-3 (p-xylene) 

I. Chronic Toxicity Summary

Inhalation reference exposure level 700 μg/m3

 (200 ppb) (for technical or mixed 

xylenes or sum of individual isomers of xylene) 

Critical effect(s) CNS effects in humans; irritation of the eyes, nose, 

and throat 

Hazard index target(s) Nervous system; respiratory system 

II. Physical and Chemical Properties (ATSDR, 1995; HSDB, 1995; CRC, 1994) 

 

Description Colorless liquid 

Molecular formula C8H10

Molecular weight 106.16 g/mol 

Density 0.864 g/cm3

 @ 20°C(technical mixture); 

0.881 (o-); 0.860 (m-); 0.861 (p-) 

Boiling point 137-140°C @ 760 TorrHg (technical mixture); 

144.5 °C (o-); 139.1°C (m-); 138.3 °C (p-) 

Melting point −25.2 °C (o-); −47.8°C (m-); +13.2 °C (p-) 

Vapor pressure 6.6 torr (o-); 8.39 torr (m-); 8.87 torr (p-) all @ 

25°C. 

Solubility Practically insoluble in water; miscible with 

absolute alcohol, ether and many other organic 

solvents 

Conversion factor 1 ppb = 4.34 μg/m3

III. Major Uses or Sources

Mixtures of o-, p-, and m-xylenes are extensively used in the chemical industry as solvents for 

products including paints, inks, dyes, adhesives, pharmaceuticals, and detergents (HSDB, 1995). 

In the petroleum industry xylenes are used as antiknock agents in gasoline, and as an 

intermediate in synthetic reactions. Of the three isomers, p-xylene is produced in the highest 

Appendix D3 602 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

quantities in the U.S. for use in the synthesis of phthalic, isophthalic, and terephthalic acid used 

in manufacture of plastics and polymer fibers including mylar and dacron. In 1996, the latest 

year tabulated, the statewide mean outdoor monitored concentration of meta/para-xylene was 

approximately 1 ppb (CARB, 1999a). The annual statewide emissions from facilities reporting 

under the Air Toxics Hot Spots Act in California based on the most recent inventory were 

estimated to be 3,568,318 pounds of xylenes (CARB, 1999b). Also reported were speciated 

emissions of p-xylene - 51,203 pounds, of o-xylene - 34,573 pounds, and of m-xylene - 30,440 

pounds. (Xylenes are also present in motor vehicle exhaust.) 

IV. Effects of Human Exposure

Information on the toxicity of xylenes to humans is almost exclusively limited to case reports of 

acute exposures and studies of occupational exposures in which persons often inhaled a mixture 

of hydrocarbon solvents 8 hours per day, 5-6 days per week. These studies often have 

incomplete information on the airborne concentrations of xylene and other hydrocarbons. One 

study examining chronic effects in humans from inhalation of predominantly mixed xylenes was 

identified (Uchida et al., 1993) and one 4-week controlled exposure study examining the effects 

of p-xylene exclusively was identified (Hake et al., 1981). No studies examining the chronic 

effects of oral or dermal xylene exposure in humans were identified. 

Pharmacokinetic studies have documented the absorption of xylene in humans through 

inhalation, oral, and dermal routes of exposure. Approximately 60% of inspired xylene is 

retained systemically (Sedivec and Flek, 1979). The majority of ingested xylene (~90%) is 

absorbed into the systemic circulation (ATSDR, 1995). Xylene is also absorbed dermally; the 

rate of absorption of xylene vapor is estimated as 0.1-0.2% of that by inhalation (Riihimaki and 

Pfaffli, 1978). Loizou et al. (1999) exposed human volunteers to 50 ppm m-xylene for 4 hours 

and determined that the dermal route of exposure contributed 1.8% of the total body burden. 

Measurement of the rate of absorption through direct contact with the skin produced variable 

results ranging from 2 μg/cm2/min (Engstrom et al., 1977) to 75-160 μg/cm2/min (Dutkiewicz 

and Tyras, 1968). 

Xylene exposure has been associated with effects in a number of organ systems including the 

lungs, skin and eyes; neurological system; heart and gastrointestinal system; kidney; and 

possibly the reproductive system. 

Pulmonary effects have been documented in occupational exposures to undetermined 

concentrations of mixed xylenes (and other solvents) and include labored breathing and impaired 

pulmonary function (Hipolito 1980; Roberts et al., 1988). High levels of xylene exposure for 

short periods are associated with irritation of the skin, eyes, nose and throat (ATSDR, 1995). 

Chronic exposure to xylenes has been associated with eye and nasal irritation (Uchida et al., 

1993). 

The central nervous system is affected by both short term and long term exposure to high 

concentrations of xylene with: 100-200 ppm associated with nausea and headache; 200-500 ppm 

with dizziness, irritability, weakness, vomiting, and slowed reaction time; 800-10,000 ppm with 

Appendix D3 603 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

lack of muscle coordination, giddiness, confusion, ringing in the ears, and changes in sense of 

balance; and >10,000 ppm with loss of consciousness (HESIS, 1986). Other documented, 

neurological effects include impaired short term memory, impaired reaction time, performance 

decrements in numerical ability, and impaired equilibrium (dizziness) and balance (Carpenter et 

al., 1975; Dudek et al., 1990; Gamberale et al., 1978; Riihimaki and Savolainen, 1980; 

Savolainen and Linnavuo, 1979; Savolainen and Riihimaki 1981; Savolainen et al., 1979; 1984; 

1985). 

Chronic exposure to xylenes (with other hydrocarbons) has been associated with cardiovascular 

and gastrointestinal effects. Heart palpitations, chest pain, and abnormal electrocardiogram were 

noted (Hipolito, 1980; Kilburn et al., 1985) as were effects on the gastrointestinal system 

producing nausea, vomiting and gastric discomfort in exposed workers (Goldie, 1960; Hipolito, 

1980; Uchida et al., 1993; Klaucke et al., 1982; Nersesian et al., 1985). 

Results of studies of renal effects of xylene are mixed and come from case reports and 

occupational studies where multiple chemical exposures are common. The effects from 

subchronic exposure documented by Hake et al. (1981) and from chronic exposure documented 

by Uchida et al. (1993) did not include renal effects. However, Morley et al. (1970) found 

increased BUN and decreased creatinine clearance; Martinez et al. (1989) found distal renal 

tubular acidemia; Franchini et al. (1983) found increased levels of urinary β-glucuronidase; and 

Askergren (1981, 1982) found increased urinary excretion of albumin, erythrocytes, and 

leukocytes. 

Reproductive effects were documented by Taskinen et al. (1994) who found increased incidence 

of spontaneous abortions in 37 pathology and histology workers exposed to xylene and 

formaldehyde in the work place. The multiple chemical exposures and the small number of 

subjects in this study limit the conclusions that can be drawn as to reproductive effects of xylene 

in humans. 

No hematological effects have been identified in studies where exposure was to xylene only. 

Previous studies identifying hematological effects included known or suspected exposure to 

benzene (ATSDR, 1995; ECETOC, 1986). One series of case reports identified lowered white 

cell counts in two women with chronic occupational exposure to xylene (Hipolito, 1980; 

Moszczynsky and Lisiewicz, 1983; 1984), although they may also have had multiple chemical 

exposures. 

Groups of male volunteers (1 to 4 subjects/group) were exposed to p-xylene in a controlledenvironment chamber for 7.5, 3, or 1 hr/day, 5 days/week for 4-weeks (Hake et al., 1981). The 

p-xylene concentration was changed on a weekly basis starting at 100 ppm the first week, 

followed by 20 ppm, 150 ppm, and 100 ppm (average, with a range of 50 to 150 ppm) over 

subsequent weeks. In addition, groups of female volunteers (2 or 3/group) were exposed to 100 

ppm p-xylene for 7.5, 3, or 1 hr/day for 5 days. The volunteers acted as their own controls, with 

exposure to 0 ppm p-xylene occurring for two days (males) or one day (females) the week before 

and the week after the xylene exposures. No serious subjective or objective health responses, 

including neurological tests, cognitive tests and cardiopulmonary function tests were observed. 

Odor was noted, but the intensity decreased usually within the first hour of exposure. The 

Appendix D3 604 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

authors concluded that p-xylene may have a weak irritating effect on the soft tissues starting at 

100 ppm, but overall, the small sample size and high variability among the volunteers made all 

results difficult to interpret. 

The Uchida et al. (1993) study included a relatively large number of workers studied, exposure 

for an average of 7 years to xylenes predominately and a comprehensive set of medical 

examinations to document potential effects. A survey of 994 Chinese workers involved in the 

production of rubber boots, plastic coated wire and printing processes employing xylene solvents 

was carried out. The survey consisted of fitting individual workers with diffusive samplers for 

an 8 hour shift. At the end of the 8 hour shift the samplers were recovered for analysis of solvent 

exposure, and urine samples were collected for analysis of xylene metabolites. The following 

day workers answered a questionnaire concerning subjective symptoms, and blood and urine 

were collected for analysis. Out of this group of xylene-exposed workers, 175 individuals (107 

men and 68 women) were selected for further study and analysis based on completion of their 

health examinations and on results from diffusive samplers showing that xylene constituted 70% 

or more of that individual’s exposure to solvents in the workplace. The control population 

consisted of 241 (116 men and 125 women) unexposed workers from the same factories or other 

factories in the same region, of similar age distribution, of similar time in this occupation 

(average of 7 years), and having a similar distribution of alcohol consumption and cigarette 

usage. The xylene-exposed and unexposed groups were given health examinations which 

evaluated hematology (red, white, and platelet cell counts, and hemoglobin concentration), 

serum biochemistry (albumin concentration, total bilirubin concentration, aspartate 

aminotransferase, alanine aminotransferase, gamma-glutamyl transferase, alkaline phosphatase, 

leucine aminopeptidase, lactate dehydrogenase, amylase, blood urea nitrogen, creatinine), and 

subjective symptoms (survey of symptoms occurring during work and in the previous three 

months). 

Results of analysis of the diffusive samplers showed that workers were exposed to a geometric 

mean of 14.2 ± 2.6 ppm xylene (arithmetic mean of 21.3 ± 21.6 ppm). This was broken down 

into geometric means of 1.2 ppm o-xylene, 7.3 ppm m-xylene, 3.8 ppm p-xylene, 3.4 ppm ethyl 

benzene, and 1.2 ppm toluene. N-Hexane was rarely present and no benzene was detected. 

Analysis of data from the health examinations found no statistically significant difference 

(p<0.10) between hematology and serum biochemistry values for xylene-exposed and unexposed 

populations. The frequency of an elevated ratio of aspartate aminotransferase to alanine 

transferase and of elevated ratio of alkaline phosphatase to leucine aminopeptidase was 

significantly (p<0.01) higher in exposed men than in the control population of men. Results of 

the survey of subjective symptoms found differences in symptoms occurring during work and 

during a similar analysis over the preceding three month period, apparently related to effects on 

the central nervous system and to local effects on the eyes, nose and throat. The frequency of 

five symptoms experienced during work was significantly (p<0.01) elevated in either xyleneexposed men or women including: dimmed vision, unusual taste, dizziness, heavy feeling in the 

head, and headache. The frequency of four symptoms experienced during work were 

significantly (p<0.01) elevated in both men and women including irritation in the eyes, nasal 

irritation, sore throat, and floating sensation. Ten subjective symptoms occurring in the previous 

three months were significantly (p<0.01) elevated in exposed men and women including nausea, 

nightmare, anxiety, forgetfulness, inability to concentrate, fainting after suddenly standing up, 

Appendix D3 605 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

poor appetite, reduced grasping power, reduced muscle power in the extremities, and rough skin. 

Dose dependency appeared to exist for 3 subjective symptoms noted during work: irritation in 

the eyes, sore throat, floating sensation, and for one symptom occurring in the last three months, 

poor appetite. 

V. Effects of Animal Exposure

A limited number of chronic toxicity studies are available for xylene including two inhalation 

studies with o-xylene (Tatrai et al., 1981; Jenkins et al., 1970) and one oral chronic study with 

mixed xylenes (NTP, 1986). No chronic dermal studies could be identified. A spectrum of 

adverse effects has been documented in shorter term studies which potentially could occur with 

chronic exposure. These studies are presented here along with a brief description of the three 

chronic studies identified. Xylene affects a number of organ systems including the pulmonary 

system, the cardiovascular system, the gastrointestinal system, the hepatic system, the renal 

system, the dermis, and the eye, and it has numerous neurological effects and developmental 

effects. 

Animal data are consistent with human data in documenting respiratory effects from xylene 

exposure. Acute and subacute exposures in mice, rats, and guinea pigs have been associated 

with decreased metabolic capacity of the lungs; decreased respiratory rate; labored breathing; 

irritation of the respiratory tract; pulmonary edema; and pulmonary inflammation (Carpenter et 

al., 1975; De Ceaurriz et al., 1981; Elovaara et al., 1987; 1989; Furnas and Hine, 1958; Korsak 

et al., 1988; 1990; Patel et al., 1978; Silverman and Schatz, 1991; Toftgard and Nilsen, 1982). 

Limited evidence is available in animal studies for cardiovascular effects resulting from xylene 

exposure. Morvai et al. (1976; 1987) conducted two studies. The first study observed rats 

following acute and intermediate duration inhalation exposure to very high (unspecified) levels 

of xylene and recorded ventricular repolarization disturbances, atrial fibrillation, arrhythmias, 

occasional cardiac arrest and changes in electrocardiogram (Morvai et al., 1976). In a 

subsequent study morphological changes in coronary microvessels were seen in rats exposed to 

230 ppm xylene (isomer composition unspecified) (Morvai et al., 1987). However the chronic 

toxicity studies conducted by the National Toxicology Program (NTP, 1986) and by Jenkins et 

al. (1970), as well as other shorter term studies (Carpenter et al., 1975; Wolfe, 1988), have not 

identified histopathological lesions of the heart. 

Studies identifying adverse gastrointestinal effects, hematological effects, or musculoskeletal 

effects in animals were not identified. Studies reporting no hematological effects include 

Carpenter et al. (1975) (rats exposed to 810 ppm of mixed xylenes for 10 weeks, 5 days/week, 6 

hours/day and dogs exposed for 13 weeks to 810 ppm mixed xylenes, 5 days/week, 6 hours/day) 

and Jenkins et al. (1970) (rats, guinea pigs and dogs exposed for 6 weeks to 780 ppm o-xylene, 5 

days/week, 8 hours per day). Carpenter et al. (1975) and the NTP (1986) reported no effects on 

the musculoskeletal system. 

Hepatic effects have been documented after acute exposure to high concentrations of xylene 

(2,000 ppm) or subacute exposure to lower concentrations (345-800 ppm) of mixed xylene or 

individual isomers. These effects include increased cytochrome P-450 and b5 content, increased 

hepatic weight, increased liver to body weight ratios, decreased hepatic glycogen, proliferation 

Appendix D3 606 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

of endoplasmic reticulum, changes in distribution of hepatocellular nuclei, and liver degeneration 

(Bowers et al., 1982; Condie et al., 1988; Elovaara, 1982; Elovaara et al., 1980; Muralidhara and 

Krishnakumari 1980; Patel et al., 1979; Pyykko 1980; Tatrai and Ungvary, 1980; Tatrai et al., 

1981; Toftgard and Nilsen, 1981; 1982; Toftgard et al., 1981; Ungvary et al., 1980). 

Renal effects have been identified in studies with rats, guinea pigs, dogs, and monkeys exposed 

to 50-2,000 ppm of xylenes. These effects include increased cytochrome P-450 content and 

increased kidney to body weight ratios (Condie et al., 1988; Elovaara 1982; Toftgard and Nilsen, 

1982). Condie et al. (1988) also noted tubular dilation, atrophy, and increased hyaline droplets 

in the kidney of Sprague-Dawley rats administered 150 mg/kg/day orally of mixed xylenes. This 

response is consistent with early nephropathy. 

Xylene has been found to affect the dermis and eyes of animals. Hine and Zuidema (1970) 

found skin erythema and edema, epidermal thickening, and eschar formation in response to 

xylene exposure. Direct instillation of xylenes into the eyes of rabbits produces eye irritation 

(Hine and Zuidema, 1970; Smyth et al., 1962) 

Numerous neurological effects have been documented in response to acute and subchronic 

xylene exposures ranging from 100 to 2,000 ppm. This is consistent with effects on 

neurofunction documented in humans. These effects include narcosis, prostration, 

incoordination, tremors, muscular spasms, labored respiration, behavioral changes, hyperactivity, 

elevated auditory thresholds, hearing loss, and changes in brain biochemistry (Andersson et al., 

1981; Carpenter et al., 1975; De Ceaurriz et al., 1983; Furnas and Hine, 1958; Ghosh et al., 

1987; Gralewicz et al., 1995; Kyrklund et al., 1987; Molnar et al., 1986; NTP, 1986; Pryor et al., 

1987; Rank 1985; Rosengren et al., 1986; Savolainen and Seppalainen, 1979; Savolainen et al., 

1978; 1979a; Wimolwattanapun et al., 1987). 

Developmental effects have been documented in pregnant animals exposed to xylenes. ATSDR 

(1995) concluded that the body of information available for developmental effects is consistent 

with the hypothesis that xylene is fetotoxic and many of the fetotoxic responses are secondary to 

maternal toxicity. However, the ATSDR also observed that there was a large variation in the 

concentrations of xylene producing developmental effects and of those producing no 

developmental effects. The ATSDR thought that these differences were influenced by a number 

of factors (strain and species of animal, purity of xylene, method of exposure, exposure pattern 

and duration, etc.). The two most common test species have been the rat and the mouse. 

With respect to rats, Mirkova et al. (1983) exposed groups of pregnant rats (unspecified strain of 

white rats) to clean air or 2.3, 12, or 120 ppm of xylene (unspecified composition) for 6 h/day on 

days 1-21 of gestation. They reported increased postimplantation losses and fetotoxicity 

(reduced fetal weights) as well as a statistically increased incidence of visceral abnormalities 

(including ossification defects in bones of the skull) at xylene air concentrations of 12 ppm and 

above. The ATSDR has suggested that the Mirkova et al. (1983) study results may have been 

influenced by poor animal husbandry as indicated by the low conception rates and the high 

incidence of fetal hemorrhages seen in the controls. Hass and Jakobsen (1993) attempted to 

replicate the findings of Mirkova et al. (1983). Hass and Jakobsen (1993) exposed groups of 36 

pregnant Wistar rats to clean air or 200 ppm of xylene for 6 h/day on days 4-20 of gestation. 

Appendix D3 607 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

Unlike Mirakova et al. (1983), there was no sign of maternal toxicity and no decrease in fetal 

weights and no increase in soft-tissue or skeletal malformations. A large increase in the 

incidence of delayed ossification of the os maxillare of the skull, however, was observed (53% 

of experimental fetuses as opposed to 2% of the controls). Potential neurological/muscular 

changes measured as performance on a rotorod were also noted upon testing of 2-day-old rat 

pups. 

Ungvary et al. (1985) exposed CFY rats by inhalation to air concentrations of xylene (60 ppm, 

440 ppm, 800 ppm) for 24 h/day on days 7-15 of gestation. Maternal toxicity was described as 

moderate and dose-dependent. They observed weight retarded fetuses at all air concentrations. 

However, there was no increase in malformations, and an increase in minor anomalies and 

resorbed fetuses occurred only at the highest concentration. In a separate study investigating the 

interactions between solvents and other agents, Ungvary (1985) exposed CFY rats to either 140 

ppm or 440 ppm of xylene on days 10-13 of gestation and also reported increases for either 

condition in weight retarded and skeletal retarded fetuses without any increase in malformations. 

Hudak and Ungvary (1978) had earlier examined the effect of 230 ppm xylene (24 h/day, days 9-

14 of pregnancy) in the CFY rat and reported effects on skeletal development (e.g., fused 

sternebrae). In contrast to the other Ungvary findings, no effect on fetal weight was observed. 

Bio/dynamics (1983) conducted an inhalation exposure study in the rat (CrL-CD (SD) BR 

strain). Rats were exposed 6 h/day during premating, mating, gestation and lactation. Exposure 

concentrations were 0, 60, 250, and 500 ppm. Most measures for adverse effects on fetal 

development were not significantly increased. Mean fetal weights at the highest exposure level 

were lower than controls, but this difference was significant only for the female fetuses. These 

depressed weights were, however, still significant on day 21 of lactation. Other adverse effects 

(such as increased soft tissue and skeletal abnormalities, increased fetal resorptions) were not 

increased significantly at any of the test concentrations. 

Ungvary et al. (1980a) tested by inhalation the individual ortho, meta, and para isomers of 

xylene in the CFY rat. Pregnant rats were exposed 24 h/day on days 7 –14 of pregnancy to 35, 

350, or 700 ppm of each isomer. An increased incidence of weight retarded fetuses was 

observed for each isomer at the 700 ppm level, and for the ortho isomer at the 350 ppm level. 

Post implantation losses were increased only at the 700 ppm level in the para-xylene exposed 

group. Skeletal anomalies were increased only at the 700 ppm level for the meta and para 

isomers of xylene. Rosen et al. (1986) evaluated the effects of prenatal exposure to para-xylene 

in the rat. They exposed pregnant Sprague-Dawley rats by inhalation to either 800 ppm or 1600 

ppm of p-xylene from days 7-16 of gestation. Despite the high concentrations, no effects were 

seen on litter size or weight at birth or on the subsequent growth rates of the pups. 

Hass et al. (1995) examined postnatal development and neurobehavioral effects in rats following 

prenatal exposure to 0 or 500 ppm technical xylene 6 hr/day on gestation days 7-20 of 

pregnancy. Xylene exposure caused no signs of maternal toxicity and no difference in the 

number of live or dead fetuses. The mean birth weight in exposed litters was about 5% lower 

compared to control litters but the difference was not statistically significant. Body weights were 

similar between groups during the preweaning and postweaning period but lower absolute brain 

weights were observed in exposed animals. Exposed offspring showed a delay in the ontogeny 

of the air righting reflex and exhibited impaired performance in behavioral tests for neuromotor 

Appendix D3 608 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

abilities (Rotorod) and for learning and memory (Morris water maze). In a follow-up study 

under the same exposure conditions, exposed offspring exhibited impaired performances in the 

Morris water maze at 16, 28, and 55 weeks of age, although the difference was not statistically 

significant at 55 weeks (Hass et al., 1997). These data indicate that xylene exposure during 

development may cause long-lasting deficits on learning and memory in offspring. 

With respect to mice, Ungvary et al. (1985) exposed CFLP mice by inhalation to air 

concentrations of xylene (120 ppm, 230 ppm) for 24 h/day on days 7-15 of gestation. In the 

mouse, they observed increased incidences of weight-retarded fetuses and increased skeletal 

retarded fetuses at 230 ppm. Shigeta et al. (1983) exposed pregnant ICR mice to approximately 

0, 120, 230, 460, and 920 ppm of xylene in an exposure chamber for 6 h/day on days 6-12 of 

gestation. Shigeta et al. (1983) reported significant decreases in fetal weight in the 460 ppm and 

920 ppm dose groups only. There was no difference in the number of live or dead fetuses. 

Decreased weight gains and delayed development of body hair and teeth were observed at the 

920 ppm exposure level. Dose-response relations were reported for delayed ossification of the 

sternebrae. Marks et al. (1982) noted that 2060 mg/kg/day of mixed xylene administered orally 

is associated with cleft palate and decreased fetal weight in the mouse. 

Ungvary et al. (1985) also tested the individual ortho, meta, and para isomers of xylene at 120 

ppm in the CFLP mouse. Each isomer of xylene also increased the incidence of weight-retarded 

fetuses and skeletal retarded fetuses at 120 ppm. There was no increase in malformations. 

Of the three chronic studies available (Tatrai et al., 1981; Jenkins et al., 1970; NTP 1986) none 

comprehensively examined systemic effects. The study by Tatrai et al. (1981) exposed rats for 

one year, 7 days/week, 8 hours per day to 1096 ppm o-xylene. This exposure was a LOAEL for 

body weight gain in males and a NOAEL for hepatic effects in male rats. Jenkins et al. (1970) 

exposed rats, guinea pigs, squirrel monkeys, and beagle dogs for 90-127 days continuously to 78 

ppm of o-xylene. The study examined body weight gain; hematological parameters including 

white cell counts, red blood cell counts, and hematocrit; serum biochemistry including 

bromosulfophthalein retention, blood urea nitrogen, alanine aminotransferase, aspartate 

aminotransferase, alkaline phosphatase, and creatinine and liver function including alkaline 

phosphatase, tyrosine aminotransferase, and total lipids. No effects were observed in any of the 

parameters examined in this study. This study found a NOAEL for all effects examined of 78 

ppm o-xylene. The NTP (1986) study administered 0, 250, or 500 mg/kg/day doses of mixed 

xylene in corn oil by gavage 5 days/week for 103 weeks to groups of F344/N rats of both sexes, 

50 animals per group. B6C3F1 mice were treated in a similar manner but given 0, 500 or 1000 

mg/kg/day of mixed xylenes in corn oil by gavage. A complete histopathological examination of 

all tissues was made as well as determination of body weight gain. Based on histopathology of 

all organ systems, a NOAEL of 500 mg/kg/day was observed for rats and a NOAEL of 1000 

mg/kg/day was observed for mice. 

Appendix D3 609 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

VI. Derivation of Chronic Reference Exposure Level (REL) 

Study Uchida et al. (1993) 

Study population 175 xylene-exposed factory workers and control 

population of 241 factory workers 

Exposure method Inhalation 

Critical Effects Dose related increase in the prevalence of eye 

irritation, sore throat, floating sensation, and 

poor appetite. 

LOAEL 14.2 ppm (geometric mean of exposure 

concentrations) 

NOAEL Not applicable

Exposure continuity 8 hr/d (10 m3

/day occupational inhalation rate), 

5 d/wk 

Exposure duration Occupational exposure for an average of 7 years 

Average occupational exposure 5.1 ppm for LOAEL group (14.2 x 10/20 x 5/7) 

Human equivalent concentration 5.1 ppm for LOAEL group 

LOAEL uncertainty factor 3 

Subchronic uncertainty factor 1 

Interspecies uncertainty factor 1 

Intraspecies uncertainty factor 10 

Cumulative uncertainty factor 30 

Inhalation reference exposure level 0.2 ppm (200 ppb; 0.7 mg/m3

; 700 μg/m3) for 

mixed xylenes or for total of individual 

isomers 

A number of issues are important in considering the uncertainty associated with this REL. For 

ATSDR (1995) the animal and human toxicity data suggest that mixed xylenes and the different 

xylene isomers produce similar effects, although different isomers are not equal in potency for 

producing a given effect. Therefore exposure of workers to a mix of xylenes in the Uchida et al.

(1993) study would be expected to generate a similar spectrum of responses as exposure to single 

isomers, however the intensity of particular effects could be different. The use of a neurological 

endpoint for derivation of a REL is supported by the large number of inhalation and oral studies 

associating neurological effects with xylene exposure. ATSDR (1995) indicates that 

neurological effects are a sensitive endpoint. The observation that floating sensation is 

apparently related to dose further supports the concept that this subjective symptom related to 

neurological effects was due to xylene exposure. 

A UF of 3, rather than 10, was applied for the LOAEL to NOAEL extrapolation due to the 

generally mild adverse effects observed and the principally low incidence (<50%) of the effects. 

A factor of 1 was used for subchronic uncertainty. Although the average occupational exposure 

was only 7 years, there were 176 xylene-exposed workers of average age 29.7±9.0 years 

(arithmetic mean ±SD) for whom, according to the report, there had been essentially no change 

in workplace in their working life. Thus, many workers would likely have been exposed for 

more than 8.4 years, the cut-off point for chronic human exposure. Another issue is the use of 

diffusive samplers in the Uchida et al. (1993) study. These samplers provide a time weighted 

Appendix D3 610 Xylenes 
Determination of Noncancer Chronic Reference Exposure Levels April 2000 

average concentration of hydrocarbon and cannot indicate the maximum concentrations a worker 

is exposed to. It is unknown whether peak concentrations alter the response to xylenes in 

humans. 

For comparison with the proposed REL of 200 ppb based on human studies, (1) the free-standing 

NOAEL of 78 ppm o-xylene obtained by Jenkins et al. (1970) in rats and guinea pigs 

continuously exposed for 90 days was used to estimate a REL based on animal data. Use of an 

RGDR of 1, a subchronic UF of 3, an interspecies UF of 3, and an intraspecies UF of 10 result in 

a REL of 800 ppb for o-xylene for systemic effects. (2) Tatrai et al. (1981) found a free standing 

LOAEL of 1096 ppm o-xylene for body weight gain in male rats exposed every day for 8 hours. 

Time adjustment to continuous exposure and use of an RGDR of 1, a LOAEL UF of 3 for a mild 

effect, an interspecies UF of 3, and an intraspecies UF of 10 result in a REL of 4000 ppb. (3) 

Ungvary et al. (1985) exposed mice by inhalation continuously to 120 ppm or 230 ppm xylene 

for 24 h/day on days 7-15 of gestation. The LOAEL was 230 ppm and the NOAEL was 120 

ppm. No time adjustment is needed. Use of an RGDR of 1, a subchronic UF of 1, an 

interspecies UF of 3, and an intraspecies UF of 10 results in a REL of 4000 ppb for xylene for 

developmental effects. 

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for xylene include the use of human exposure data from 175 

workers exposed over a period of years. Major areas of uncertainty are the uncertainty in 

estimating exposure, the potential variability in exposure concentration, and the lack of 

observation of a NOAEL in the key study. 

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Appendix D3 619 Phosphine