Appendix D.3 Chronic RELs and toxicity summa ries using the previous version of the Hot Spots Risk Assessment gu idelines (OEHHA 1999)Noncancer chronic Reference Exposure Leve ls 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 valuesNoncancer chronic Reference Exposure Leve ls 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 199 9 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesNoncancer chronic Reference Exposure Leve ls determined using previous methodology

Table of Chronic RELs determined using the previous Hot Spots Risk Assessment
G uidelines (OEHHA 1999 )

Substance (CAS #) Listed in

CAPCOA

(1993) Chronic

Inhalation

REL ( μ g/m 3 ) 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 valuesDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# ACRYLONITRILE

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

CAS Number: 107-13-1

I. Chronic Toxicity Summar y

Inhalation reference exposure level 5 μg/m 3 (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 C 3 H 3 N Molecular weight 53.1 g/mol Density 0.81 g/cm 3 @ 25 o C Boiling point 77.3 o C Melting point

− 82 o C Vapor pressure 100 torr @ 23 o C Solubility Soluble in isopropanol , ethanol, ether, acetone, and benzene Conversion factor 1 ppm = 2.17 mg/m 3 @ 25 o C

III. Major Uses or Sources

Acrylonitrile is produced commercially by pr opylene ammoxidation, in which propylene, ammonia, and air are reacted by catalyst in a flui dized bed. Acrylonitrile is used primarily as a co-monomer in the production of acrylic and modacr ylic fibers. Uses include the production of plastics, surface coatings, nitrile el astomers, barrier resins, and adhesi ves. It is also a chemical intermediate in the synthesis of various antio xidants, 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 th e 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 μ g/m 3

.

Appendix D3 11 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

IV. Effects of Human Exposure

Many occupational epidemiology studies have inve stigated retrospectiv ely 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 ear ly 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 acr ylonitrile (16 to 100 ppm, 34.7 to 217 mg/m 3 ). Mucous membrane irritation, headaches, feelings of appr ehension, and nervous irritability were observed in the majority of workers. Other less comm on 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 concentr ations ranging from 5.4 to 10.9 mg/m 3 (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 wo rkers (minimum 5 years) employed at 6 acrylic fiber factories (n = 102 exposed, n = 62 matched c ontrols). Mean acrylo nitrile levels ranged from 0.1 to 4.2 ppm (0.2 to 9.1 mg/m 3 ) as determined by personal sampling. Although not statistically significant, slight increases in redd ening of the conjunctiva and pharynx were seen in workers from the plant with the highest mean le vels (4.2 ppm arithmetic mean). However, this study has limitations, including small sample size a nd 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/m 3 ) acrylonitrile can be ca lculated as: TWA = 9.1 mg/m 3 x (10/20) m 3 /day x 5 days/7 days = 3 mg/m 3 . This level is comparable to the LOAEL (HEC) of 2 mg/m 3 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 fact ory in Nyergesujfalu, Hungary. Ascertainment of cases with conge nital 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 mutati ons (sentinel anomalies, Down’s syndrome, and unidentified multiple congenital abnormalities) and to indicators of teratogens (specific pattern of multiple congenital abnorma lities). 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 increa sing distance from the factory. An unusual increase for the combination of oral cleft and cardi ac 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 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Quast et al. (1980) exposed Sprague-Daw ley 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/m 3 ). A statistically significant increase in mort ality was observed in the first year among 80 ppm exposed rats (male and female). Additi onally, the 80 ppm exposed group had a significant decrease in mean body weight. Two tissues, th e nasal respiratory epithelium and the brain,

exhibited treatment-related adverse effects due to acrylonitrile expos ure. Proliferative changes in the brain glial cells (i.e., tumors and early prolif eration suggestive of tumo rs) 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 ma le controls (0/100). Noncarcinogenic, extrarespiratory effects were observed in th e nasal turbinate epith elium 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 acrylonitr ile 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, atelectasi s, or edema Male 14/100 27/100* 30/100*
Lung: pneumonia, consolidation, atelect asis, 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/con centration) 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 a nd 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 ra ts.) 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 acr ylonitrile vapors for 6 h/day, 5 days per week, for 24 weeks and neurophysiological examinations were carried out. After oral acrylonitrile, animals developed behavioral sens itization characterized by salivation, locomotor hyperactivity, and moderately in tense stereotypies. Rats dosed with 50 mg/kg developed

hindlimb weakness associated with decreases in sensory conduction veloc ity (SCV) and in the amplitude of the sensory action potential (ASAP). Ra ts exposed to acrylo nitrile by inhalation exhibited time- and concentration-dependent de creases in motor conduction velocity (MCV), SCV, and ASAP, which were par tially reversible after 8 week s of recovery. The authors concluded that the nervous system of the rat ap pears to be a target fo llowing either oral or inhalation exposures of acryl onitrile. The NOAEL by inhala tion for 24 weeks was 25 ppm.

Appendix D3 13 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Changes in electrophysiological parameters af ter 24 wks of exposure (Gagnaire et al., 1998) Acrylonitrile MCV (m/sec) SCV (m/sec) AMAP (mvolts) ASAP ( μ volts)
0 ppm 42.9 ± 0.9 a 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 acryl onitrile vapors at 0, 40 ppm (87 mg/m 3 ), or 80 ppm (174 mg/m 3 ) for 6 hours/day during gestational days 6 to 15. In the 80 ppm exposed group, significant incr eases in fetal malformations we re observed including short tail, missing vertebrae, short trunk, ompha locoele, and hemivertebra (Murray et al. , 1978). No differences in implantations, live fetuses, or re sorptions 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 id entified a developmental NOAEL of 10 ppm and a LOAEL of 20 ppm (with maternal toxicity).

Saillenfait et al . (1993) studied the developmental toxic ity of eight aliphatic mononitriles in Sprague-Dawley rats after inhalati on 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 observe d after exposure to 25 ppm (54 mg/m 3 ) acrylonitrile in the presence of overt signs of matern al toxicity. Fetal weights were significantly lower at 25 ppm. Thus 12 ppm (26 mg/m 3 ) is a NOAEL for developmental to xicity using this study design.

Appendix D3 14 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

VI. Derivation of Chronic Reference Exposure Level

Study Quast et al. , 1980 Study population Sprague-Dawley rats (1 00/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

BMC 05 1.5 ppm Exposure continuity 6 hours/day, 5 days/week Average experimental exposure 0.27 ppm for BMC 05 (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 m 3 /day, SA(ET) = 11.6 cm 2 ) 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/m 3 ; 5 μg/m 3 )

Sprague-Dawley rats (100/sex/conc entration) were exposed 6 hours/ day, 5 days/week for 2 years to 0, 20, or 80 ppm acrylonitrile (0, 43, and 174 mg/m 3 , 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-re lated irritation of the nasal mucosa appeared in the 20 ppm exposed male rats as either epithelial hyp erplasia of the nasal turbinates, or as hyperplasia of the mucous secr eting cells. In the 20 ppm exposed females it appeared as either focal inflammation in the na sal turbinates or flattening of the respiratory epithelium of the nasal turbinates. In 80 ppm exposed rats the effects were more severe, including suppurative rhin itis, hyperplasia, focal erosions, and squamous metaplasia of the respiratory epithelium. No treatment-related effects in the olfactory epithelium, trachea, or lower respiratory system were observe d at either concentration. Th is study identified a LOAEL for pathological alterations in the re spiratory epithelium of the extrat horacic region of the respiratory tract of 20 ppm (43 mg/m 3 ). The U.S. EPA (1994) based its RfC of 2 μ g/m 3 on the same study but included a Modifying Factor (M F) of 10 for database deficienci es. The criteria for use of modifying factors are not well specified by U.S. EP A. Such modifying factors were not used by

OEHHA.

OEHHA used a benchmark dose approach to dete rmine 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 turbinat es in males, hyperplasia of the

Appendix D3 15 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

mucous secreting cells in males, focal inflamma tion in the nasal turbinates in females, and flattening of the respiratory epithelium of the na sal turbinates in females. The resulting BMC 05 values (1.27, 1.33, 2.18, 1.35) were averaged to yiel d a value of 1.5 ppm. The RGDR adjustment and appropriate uncertainty factors were applied as indicated in th e above table and resulted in a

chronic REL of 5 μ g/m 3

.

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 es timated REL of 45 ppb (100 μ g/m 3 ). We were unable to derive a BMC from the neurotoxicity data due part ly to the tendency of the anim als in the 100 ppm group to yield values for two of the four endpoints measured cl oser to the controls th an those in the 50 ppm

group.

As another comparison, Saillenfait et al . (1983) found a 12 ppm (26 mg/m 3 ) 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 effect s, an interspecies UF of 3, and an intraspecies UF of 10 re sults in an estimated REL of 100 ppb (200 μ g/m 3 ).

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 μ g/m 3 ).

VII. Data Strengths and Limitations for Development of the REL

Significant strengths in the chronic REL for acrylon itrile include (1) the availability of chronic inhalation exposure data from a well-conducted st udy with histopathological analysis and (2) the demonstration of a dose-response relationship. Ma jor 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 re productive data for inhalation exposures when an oral study showed advers e 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 associat ed with increased sensitivity of younger animals or humans, was evaluated as one of the alternative en dpoints, the comparison reference level for this end point in adults was mo re than an order of magnitude higher that the REL based on histological changes in the upper respir atory tract. It is ther efore considered that the REL is likely to be adequately protective of infants and children.

Appendix D3 16 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

IX. References

CARB. 2000. California Air Resources Board. Calif ornia Emissions Inventory Development and Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

Czeizel AE, Hegedus S and Timar L. 1999. Congen ital abnormalities and in dicators of germinal mutations in the vicinity of an acrylonitrile producing factor y. Mutat. Res. 427(2):105-123

Gagnaire F, Marignac B, and Bonnet P. 1998. Rela tive neurotoxicological properties of five unsaturated aliphatic nitriles in ra ts. 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 K opecky J. 1987. 2-Cyanoethylmercapturic acid (CEMA) in the urine as a possibl e indicator of exposure to acryl onitrile. Br. J. Ind. Med. 44:834-

840.

Maltoni C, Ciliberti A, and Di Maio V. 1977. Carc inogenicity 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. Mi dland, MI: Dow Chemical Co.

Saillenfait AM, Bonnet P, Guenier JP and de Ceau rriz J. 1993. Relative deve lopmental toxicities of inhaled aliphatic mononitriles in ra ts. 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 Agen cy (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 AcrylonitrileDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

U.S. EPA 1994. U.S. Environmen tal Protection Agency. Integrat ed Risk Information System (IRIS) Database. Reference concentration (R fC) 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 AcrylonitrileDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 NH 3 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/m 3

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 explos ives (ATSDR, 1990). Ammonia is also used as a refrigerant. The general public is exposed by off-gasing from cleani ng solutions containing aqueous ammonia. Household ammonia solutio ns contain 5-10% ammonia in water while industrial strength can be up to 28%. The annual st atewide industrial emis sions from facilities reporting under the Air Toxics Ho t 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 an d respiratory symptomatology (Holness et al. , 1989). The pulmonary function tests included FVC (forced vital capacity – the to tal amount of air the subject can expel during a forced expiration), FEV 1 (forced expiratory volume in one second),

FEF 50 (forced expiratory flow rate at 50% of the FVC) and FEF 75 (forced expiratory flow rate at

Appendix D3 19 AmmoniaDetermination of Noncancer Chronic Re ference 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 concentr ations of 9.2 ppm (6.4 mg/m 3 ) ± 1.4 ppm, while controls were exposed to 0.3 ppm (0.21 mg/m 3 ) ± 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 work week) 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/m 3 ) 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 NH 3 hours 25

2 50

4 100

6 25

2 50

4 100

6

B ppm NH 3 hours 50

6 50

6 50

6 50

6 50

6 50

6

C ppm NH 3 hours 100

6 50

4 25

2 25

6 50

4 100

2

Pulmonary function tests (resp iration rate, FVC and FEV l ) were measured in addition to subjective complaints of irritation of the eyes an d respiratory tract. The difficulty experienced in performing simple cognitive tasks was also measure d, as was pulse rate. There were reports of

transient irritation of the nose and throat at 50 or 100 ppm. Acclim ation 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 norma l job duties were found. After acclimation, continuous exposure to 100 ppm, with occasiona l 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 ammoni a at 0, 25, 50, 150, or 250 ppm (0, 18, 36, 107, or 179 mg/m 3 ) 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 seve rity of rhinitis, otitis media, tracheitis, and pneumonia characteristic of M. pulmonis infection, therefore 25 pp m was a LOAEL in this subchronic study. Exposure to 250 ppm ammonia al one resulted in nasa l lesions (epithelial thickening and hyperplasia) which were not like those seen in M. pulmonis -infected rats.

The growth of bacteria in the lungs and na sal passages, and the concentration of serum immunoglobulin were significantly increased in rats exposed to 100 ppm (71 mg/m 3 ) ammonia over that seen in control rats (Schoeb et al. , 1982).

Appendix D3 20 AmmoniaDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Guinea pigs (10/group) and mice (20/group) were continuously exposed to 20 ppm (14.2 mg/m 3 ) ammonia for up to 6 weeks (Anderson et al. , 1964). Separate groups of 6 guinea pigs and 21 chickens were exposed to 50 pp m and 20 ppm ammonia for up to 6 and 12 weeks, respectively. All species displayed pulmonary edema, conge stion, and hemorrhage af ter 6 weeks exposure,

whereas no effects were seen afte r only 2 weeks. Guinea pigs exposed to 50 ppm ammonia for 6 weeks exhibited enlarged and c ongested spleens, congested live rs 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 th e 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/m 3 . There were no signs of toxicity in 15 rats exposed continuously to 40 mg/m 3 for 114 days or in 48 rats exposed continuously to 127 mg/m 3 for 90 days. Among 49 rats exposed continuously to 262 mg/m 3 for 90 days, 25% had mild nasal discharge. At 455 mg/m 3 50 of 51 rats died. Thus 127 mg/m 3 (179 ppm) is a subchronic NOAEL for upper re spiratory effects in rats. Coon et al . (1970) also found no lung effects in 15 guinea pigs exposed continuously to 40 mg/m 3 (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 m 3 /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/m 3 ; 200 μg/m 3 )

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 μ g/m 3 on the same study but included a Modifying Factor

Appendix D3 21 AmmoniaDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

(MF) of 3 for database deficien cies. The criteria for use of modifying factors are not well specified by U.S. EPA. Such modifyi ng factors were not used by OEHHA.

For comparison with the proposed REL of 200 μ g/m 3 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/m 3 ) 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 intraspe cies) resulted in a REL of 10 μ g/m 3 . 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/m 3 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 intraspeci es) resulted in a REL of 200 μ g/m 3

.

VII. Data Strengths and Limitations for Development of the REL

Significant strengths in the a mmonia REL include (1) the ava ilability of long-term human inhalation exposure data (Holness et al ., 1989), (2) the demonstration of consistent effects in experimentally exposed huma n volunteers following short-term exposures (Ferguson et al.

, 1977), and (3) reasonable consiste ncy 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, a nd (3) difficulties in estimated human occupational exposures. The ov erall 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 inf ection 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 ra ts. 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 G ould JR. 1977. Human physiological response and adaptation to ammonia. J. Occup. Med. 19(5):319-326.

Appendix D3 22 AmmoniaDetermination of Noncancer Chronic Re ference 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. 198 9. Acute and chronic re spiratory 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. Inf ect. 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 AmmoniaDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 C 6 H 6 Molecular weight 78.1 g/mol Density 0.879 g/cm 3 @ 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/m 3 @ 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, ani line, and alkyl benzenes in the manufacture of various plastics, resins, and detergents. Synthe ses of many pesticides and pharmaceuticals also

involve benzene as a chemical in termediate (HSDB, 1994). The tire industry and shoe factories use benzene extensively in their manufacturi ng 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 combus tion (Holmberg and Lundberg, 1985). In 1996, the latest year tabulated, the st atewide mean outdoor monitored concentration of benzene was approximately 0.7 ppb (CARB, 1999a). Annual st atewide industrial emi ssions from facilities reporting under the Air Toxics Ho t 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 BenzeneDetermination of Noncancer Chronic Re ference 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 concer n at slightly higher concentrations. Impairment of immune functi on 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 ch ronic 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 fr om the hematologic damage (DeGowin, 1963).

Kipen et al. (1988) performed a retrospective long itudinal study on a cohort of 459 rubber workers, examining the correlation of averag e benzene exposure with total white blood cell

counts taken from the workers. These rese archers found a significant (p < 0.016) negative correlation between average benzene concentra tions 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-exposur e blood cell counts. The decline in blood

counts was measured over the course of 12 mont hs following start of exposure. During the course of employment, workers who had low mont hly blood cell counts were transferred to other areas with lower benzene exposures, thus potentia lly creating a bias towa rds non-significance or removing sensitive subjects from the study populati on. 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 bl ood transfusions used to treat some “anemic” workers, which would cause serious problems in interpreting th e 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 e xposures was 30-54 ppm throughout the 12-month segment examined. Despite the above-mentione d potential biases, workers exposed above the median concentrations displayed significantl y decreased WBC and RBC counts compared with workers exposed to the lower con centrations using a re peated 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 circulator y system was significantly below expected values based on comparable groups of U.S. males. Th e 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 no t 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 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

hematocrit, red blood cells, platel ets 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 Monsan to’s medical/industrial hygiene system to study 387 workers with daily 8-hour time-wei ghted exposures (TWA) averaging 0.55 ppm benzene (range = 0.01 – 87.69 ppm; based on 4213 pe rsonal 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 inte rval = 0.2 to 1.8), taking into account smoking, age, and sex. There also was no increase in ri sk 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 (med ian = 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 ab solute lymphocyte count as the most sensitive indicator of benzene-induced hematotoxicity.

An examination of 32 patients, who were chroni cally exposed to benzene vapors ranging from 150 to 650 ppm for 4 months to 15 years, showed that pancytopenia occurre d in 28 cases. Bone marrow punctures revealed variable hemat opoeitic lesions, ranging fr om acellularity to hypercellularity (Aksoy et al. , 1972).

Central nervous system disorders have been repo rted in individuals wi th pancytopenia following chronic occupational benzene expos ure to unknown concentrations fo r 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 be nzene to be 0.1 ppm (0.32 mg/m 3 ) (geometric standard deviation = 7.2 ppm, 23.3 mg/m 3 ). This estimate was based on samples collected by industrial hygienists between the years 1978 and 1983.

V. Effects of Animal Exposure

Appendix D3 26 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

A number of animal studies have demonstrated that benzene exposure can induce bone marrow damage, changes in circulating blood cells, deve lopmental and reproductive effects, alterations of the immune response, and cancer. With resp ect to chronic toxicit y, hematological changes appear to be the most sensitive indicator.

Wolf et al . (1956) studied the effects of repeated e xposure to benzene in rabbits (80 ppm, 175 total exposures), rats (88 ppm, 136 total e xposures) and guinea pigs (88 ppm, 193 total exposures). The observed effect s included leukopenia, increased spleen weight, and histological changes to the bone marrow. Hematologic effect s, including leukopenia, were observed in rats exposed to mean concentr ations of 44 ppm (143 mg/m 3 ) or greater for 5-8 weeks (Deichmann et al. , 1963). Exposure to 31 ppm (100 mg/m 3 ) benzene or less did not re sult 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 lif e. 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 st em 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 hydroqui none, muconic acid, and hydroquinone glucuronide was much more prevalent in the mouse than in rats, whereas the detoxi fication pathways were approximately equivalent between the two species.

A study on the chronic hematological effects of be nzene exposure in C57 Bl/6 male mice (5-6 per group) showed that periphera l lymphocytes, red blood cells a nd colony-forming units (CFUs) in the bone marrow and spleen we re significantly decreased in nu mber after treatment with 10 ppm (32.4 mg/m 3 ) 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/m 3 ) 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 lymphocyt e numbers and B-lymphocyte proliferation to lipopolysaccharide were signifi cantly reduced at a concen tration of 10 ppm (32.4 mg/m 3 ). The proliferation of T-lymphocytes was significan tly reduced at a concentration of 31 ppm (100.4 mg/m 3 ).

Male and female mice (9-10 per gr oup) exposed to 100 ppm (324 mg/m 3 ) benzene or greater for 6 hours/day, 5 days/week for 2 weeks showed d ecreased bone marrow cellu larity 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 follo wing a 16-week exposure to 300 ppm (972 mg/m 3 ) benzene. Peripheral blood lymphocytes were dose-dependently decreased with benzene exposures of greater than 25 ppm (81 mg/m 3 ) for 16 weeks, but recovered to normal levels following a 16-week recovery period.

Ward et al . (1985) exposed 50 Sprague-Dawley rats a nd 150 CD-1 mice of both sexes to 0, 1, 10, 30, or 300 ppm benzene, 6 hours/day, 5 days/wee k for 13 weeks. Serial sacrifices were

Appendix D3 27 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

conducted at 7, 14, 28, 56, and 91 days. No hemato logical 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 hematocr it, hemoglobin, lymphocyte percentages, and decreases in red cell, leukocyte and platelet counts were observe d in male and female mice at 300 ppm. The changes were first observed after 14 days of exposur e. Histological changes in mice included myeloid hypoplasia of the bone marro w, lymphoid depletion in the mesenteric lymph node, increased extramedullary hematopoiesis in the spleen, and pe riarteriolar 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/m 3 ) 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 ex hibited 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 exposur e group. Higher concentrations of benzene produced dose-dependent decreases in blood ly mphocytes, bone marrow cellularity, spleen colony-forming units, and an increased percentage of CF U-S in S-phase synthesis.

Farris et al . (1997) exposed B6C3F 1 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 so me 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 cell s, 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/m 3 ) 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 fr om dams exposed to 100 ppm (324 mg/m 3 ). No effects

were observed at a concen tration of 40 ppm (129.6 mg/m 3 ).

Keller and Snyder (1986) reported th at exposure of pregnant mice to concentrations as low as 5 ppm (16 mg/m 3 ) 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/m 3 ) benzene on days 6-15 of gestation resulted in neonatal suppression of erythropoietic pr ecursor cells and persistent,

Appendix D3 28 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

enhanced granulopoiesis. This effect was cons idered significant bone-marrow toxicity by the authors. No hematotoxicity was s een in this study at 10 ppm (32 mg/m 3 ).

An exposure of 500 ppm (1,600 mg/m 3 ) benzene through days 6-15 gestation was teratogenic in rats while 50 ppm (160 mg/m 3 ) 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, re spectively, induced minor skeletal variations in

the absence of maternal toxicity. Red and white blood cell count s in the adults of either species were measured by Murray et al. (1979) but were not si gnificantly different from control animals. However, fetal mouse hematologi cal effects were not measured.

Tatrai et al. (1980) demonstrated decreased fetal body wei ghts and elevated liver weights in rats exposed throughout gestation to 150 mg/m 3 (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 m 3 per 20 m 3 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/m 3 ; 60 μ g/m 3 )

Staff identified Tsai et al. (1983) as the most appropriate st udy 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 su ccess was 99.3% in the entire cohort of 359. A total of approximately 1400 samples for hemato logical tests and 900 fo r blood chemistry tests were taken between 1959 and 1979. Exposures to benzene were determined using personal monitors. Data consisting of 1394 personal samp les indicated that 84% of all benzene samples were less than 1 ppm; the median air concentra tion 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 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

and cyclohexane and also of cumene). The aver age length of employment in the cohort was 7.4 years. Mortality from all causes and from diseas es of the circulatory system was significantly below expected values based on comparable groups of U.S. males. Th e authors concluded the presence of a healthy worker effect. An an alysis 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 di fferential white blood cell counts, hemoglobin, hematocrit, red blood cells, platel ets 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 benz ene are excellent. Although the study by Tsai et al. (1983) is a free-sta nding 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 comm unity 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 work ers exposed to median peak concentrations (i.e., only the peak concentra tions 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 plan t, workers who had blood dyscrasias were excluded from working in the high benzene units. Furt hermore, individual workers having more than a 25% decrease in WBC counts from their pre-em ployment 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 Cr ump and used in the Kipen et al. (1988) and Cody et al. (1993) studies. An equivalent c ontinuous exposure of 10.7 ppm can be calculated by assuming that workers inhaled 10 m 3 of their total 20 m 3 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, estima tion of a NOAEL, and for protect ion of sensitive subpopulations (10 for each) results in an REL of 0.01 ppm (10 ppb; 30 μ g/m 3 ). 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 c ounts. A modeled dose-respon se relationship indicated a possibility for hematologic effects at concentr ations below 5 ppm. However, no specific

measures of the actual effects at concentr ations 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 BenzeneDetermination of Noncancer Chronic Re ference 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 pp m is therefore not incons istent 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 extrapolati on of this value to an equivalent continuous exposure resulted in a concentrati on of 1.8 ppm. Appli cation 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 w ould result in an REL of 6 ppb (20 μ g/m 3 ). The

Farris et al . (1997) 8 week study indicated a LOA EL of 100 ppm and a NOAEL of 10 ppm for hematological effects. Application of an R GDR 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 th e 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 me ters 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 th e 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 inhalati on on lymphocyte subpopulations and immune response in mice. Toxicol. Appl. Pharmacol. 85:92-101.

Baarson KA, Snyder CA, and Albert RE. 1984. Repeat ed exposure of C57Bl/6 mice to inhaled benzene at 10 ppm markedly depressed erythr ocyte colony formation. Toxicol. Lett. 20:337-342.

Baslo A, and Aksoy M. 1982. Neurological abno rmalities 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. T oxics Air Quality Data. Substance Chooser.

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

Appendix D3 31 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CARB. 1999b. Air toxics emissions data collect ed in the Air Toxics Hot Spots Program CEIDARS Database as of January 29, 1999.

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

In: Advances in Modern Environmental Toxico logy, Vol. VI. Applied Toxicology of Petroleum Hydrocarbons. MacFarland HN, ed. Princeton, NJ : Princeton Scientific Publishers, Inc.

Cody RR, Strawderman WW, and Kipen HM. 1993. He matologic 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 a nd the utility of routin e data collection. J. Occup. Environ. Med. 39(3):232-237.

Cronkite EP, Inoue T, Carten AL, Miller ME, Bu llis JE, and Drew RT. 1 982. Effects of benzene inhalation on murine pluripot ent 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 H ealth 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. Na tional Library of Medicine, Bethesda, MD (CD-ROM version). Micromedex, Inc., De nver, 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 BenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

IARC (International Agency for Research on Cancer). 1982. IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemical s 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 form ing 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 e xposure. 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 long itudinal study of rubber workers. Toxicol. Ind. Health. 4(4):411-

430.

Kuna RA, and Kapp RW. 1981. The embryotoxic/terat ogenic 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. NT P Technical Report on the Toxicology and Carcinogenesis Studies of Benzene (CAS No. 71-43-2) in F344/N Ra ts 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. Depr essions in B- and T-lymphocyte mitogeninduced blastogenesis in mice exposed to low co ncentrations of benzene. Toxicol. Lett. 20:343-

349.

Runion HE, and Scott LM. 1985. Benzene expos ure 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]ben zene by F344/N rats and B6C3F1 mice. Toxicol. Appl. Pharmacol. 94:128-140.

Sabourin PJ, Bechtold WE, Griffith WC, Bi rnbaum LS, Lucier G, and Henderson RF. 1989. Effect of exposure concentrati on, 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 BenzeneDetermination of Noncancer Chronic Re ference 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 ben zene inhalation in CFY rats. J. Hyg. Epidem. Micro. Immunol. 24(3):363-371.

Tsai SP, Wen CP, Weiss NS, Wong O, McCl ellan WA, and Gibson RL. 1983. Retrospective mortality and medical surv eillance 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 be nzene 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 esti mated benzene exposure in a rubberworker cohort (1940-1975). Am. J. Ind. Med. 29:247-257.

Wolf MA, Rowe VK, McCollister DD et al . 1956. Toxicological studi es od certain alkylated

benzenes and benzene. Ar ch. Ind. Health 14:387-398.

Appendix D3 34 BenzeneDetermination of Noncancer Chronic Re ference 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 μ g Be/m 3
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 BeSO 4 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 CompoundsDetermination of Noncancer Chronic Re ference 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-ra y transmission windows, ve hicle optics, nuclear reactor neutron reflectors, fuel containers, precision instruments, rocket propellants, navigational systems, heat shields, and mirrors. In additi on to the four species listed, there are many other beryllium-containing compounds, including other sa lts, ores, and alloys (see, e.g., CRC, 1994).

The annual statewide industrial emissions from facilities reporting unde r 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 bery llium copper, are used in electrical equipment, precision instruments, springs, valves, non-sparki ng 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 beryll ium 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 atmosphe re by windblown dusts and volcanic particles. However, the major emission source is the comb ustion of coal and fuel oil, which releases beryllium-containing particulates and ash. Other beryllium-releasi ng industrial processes include ore processing, metal fabricati on, beryllium oxide production, a nd municipal waste incineration (ATSDR, 1993). Beryllium also occurs in tobacco smoke (0-0.0005 μ g/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, anorex ia, and overall weakness. Most studies reporting adverse re spiratory effects in humans involve occupational exposure to beryllium. Exposure to soluble beryllium comp ounds is associated with acute beryllium pneumonitis (Eisenbud et al. , 1948). Exposure to either so luble or insoluble beryllium compounds may result in obstructive and restri ctive diseases of th e 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 num ber 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/m 3 , primarily as beryllium sulfat e 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 CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

from the berylliosis case registries (Infante et al. , 1980; Lieben and Metzner, 1959). However, Infante and associates (1980) re ported significantly increased mo rtality 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/m 3 . 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/m 3 for the ten site/proces s classifications, but 318 samples did exceed 2 μg Be/m 3 (and 20 samples were greater than 25 μg Be/m 3 ). No atmospheric samples were available after 1963, even though the exposure occurred through 1973. The LOAEL for occupationally induced beryl liosis observed in this study was estimated from uncertain exposure data to be less than 2 μg Be/m 3

.

One cross-sectional study (Kriebel et al. , 1988a; Kriebel et al. , 1988b) estimated beryllium exposure levels for 309 workers originally survey ed in 1977, with a median duration of exposure of 17 years (range 2 to 39 years). Historic plan t levels were estimated to be as high as 100 μg

Be/m 3 , and, even as late as 1975, some job classifications exceeded 10 μg Be/m 3 . The workers’ median cumulative exposure was 65 μg Be/m 3 -year (range 0.1 to 4400 μg Be/m 3 -years); the median lifetime exposure estimate was 4.3 μg/m 3 (range 0.01 to 150 μg/m 3 ). Spirometric measurement of pulmonary function, chest x-rays , and arterial blood gas measurements were collected. Decrements in lung function, as defi ned by forced vital capacity (FVC) and forced expiratory volume in one second (FEV 1 ), were associated with cumulative exposure up to 20 years prior to the health survey, even in work ers with no radiographic abnormalities. Differences in alveolar-arterial oxygen gradient were associated with cumula tive exposure in the 10 years prior to the study. These endpoi nts give a LOAEL of 39 μg/m 3 -years (geometric mean cumulative exposure) for decrements in pulmonary function and changes in arterial blood gases.

Non-occupational berylliu m-related chronic disease has been reported in individuals residing in the vicinity of beryllium ma nufacturing 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 clini cal examination of approximately 10,000 residents near a beryllium fabrication facility in Lorain, Oh io. Ten of the cases resi ded 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). Atmos pheric sampling in 1947 identified an average level of 0.2 μg Be/m 3 at 1/4 mile decreasing to 0 μg Be/m 3 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/m 3 for continuous exposure to beryllium compounds, based on the development of chronic berylliosis.

Metzner and Lieben (1961) also reported 26 cas es of chronic berylliosis in a population of approximately 100,000, living within 7 miles of a re fining and alloy fabrica tion plant (duration 6 to 19 years). Neighborhood exposure assessm ent conducted over 14 months during 1958 and 1959 identified a mean level of 0.0155 μg Be/m 3 , with 10% of the samples registering over 0.03

μ g Be/m 3 . Limited measurements conducted earlier at the site were higher (1.0 to 1.8 μ g Be/m 3
in 1953 and 0.91 to 1.4 μg Be/m 3 in 1954).

Appendix D3 37 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference 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 demonstrat ed 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 in terstitial pneumonitis in a small number of beryllium metal handling mach inists (LOAEL = 4.6 μg Be/m 3 ). A second study identified granulomatous lung lesions, scarred lung tissue, and breathing difficulties in workers from a precious metal refining facility exposed to a mi xture of beryllium and other metals (Cullen et al.

, 1987). Also, altered proliferativ e responses of lymphocytes obta ined by bronchoalveolar lavage indicated increased T-cell activity in vitro . Cullen et al. (1987) reported a mean exposure level of 1.2 μg Be/m 3 (range = 0.22 – 43 μ g/m 3 ). USEPA (1998) and ATSDR (2000) considered

0.52 μ g Be/m 3 to be the LOAEL for CBD from th is study since this was the average

concentration in the furnace area wh ere 4 of the 5 CBD cases worked.

Sensitization to beryllium, as measured by the be ryllium lymphocyte prolif eration test (BeLPT),

can occur in the absence of ch ronic beryllium disease (Kreiss et al ., 1989). The authors hoped

that the identification of sensitized individuals without dis ease might prevent clinical disease, presumably by removing the individuals from exposure to beryllium. Some beryllium-sensitized individuals progress to havi ng clinical disease (Newman et al ., 1992). Data obtained from a four-year survey conducted at beryllium-copper a lloy manufacturing factories in Japan (Yoshida

et al ., 1997) indicated that the T ce lls of workers continuously exposed to more than 0.01 μ g Be/m 3 were activated and that th e cell-mediated immune (CMI) response was promoted. The BeLPT in workers exposed to less than 0.01 μ g Be/m 3 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 th e 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 sequenc es of the HLA-DPB1 gene and in the copy

number of Glu69 are also involved (Wang et al ., 1999).

The Rocky Flats Environmental Tec hnology Site in Colorado is part of the U.S. Department of Energy nuclear weapons complex. Operations us ing Be began in 1953, Be production operations began in 1957, and the first case of CBD was diagnosed in a mach inist in 1984. Exposures could have occurred during foundry operations, casting, shearing, rolling, cutting, welding, machining, sanding, polishing, assembly, and chemical analys is 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 exposur e, 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 μ g Be/m 3 . ATSDR (2000) considered 1.04 μ g Be/m 3 to be the LOAEL for this study. A recent case-control study of workers at Rocky Flats (Viet et al ., 2000) suggested that exposure s 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 CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Kreiss et al . (1996) investigated the prevalence of bery llium sensitization in relation to work process and beryllium exposure measurements in a beryllia ceramics pl ant that had operated since 1980. In 1992 they interviewed 136 employ ees (97.8% of the workforce), ascertained beryllium sensitization with th e 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 base d on transbronchial lung biopsy. Machinists had a Be sensitization rate of 14.3% compared to 1.2% among other employees. Machining operations (drilling, dicing, centerl ess grinding, and/or su rface 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. Da ily 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 μ g/m 3 ). Daily weighted averages for the machining process accounted for the majority of exceedances of the 2.0 μ g/m 3 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/m 3 , 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 separate d 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 μ g/m 3 ), 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 occupati onal studies Wambach and Tuggle (2000) have suggested that the workplace standard of 2 μ g/m 3 be lowered to 0.1 μ g/m 3 . Some workers might still be sensitized to bery llium 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 inflammato ry and proliferative respiratory changes following inhalation exposure to bery llium compounds. Vorw ald and Reeves (1959) observed inflamed lungs and fibrosis in rats exposed to 0.006 mg Be/m 3 (as BeO) for an unspecified duration. A later study exposed Spra gue-Dawley CD rats for 72 weeks (7 hr/d, 5 d/wk) to 34.25 μg Be/m 3 from BeSO 4 (Reeves et al. , 1967). Gross and histological changes observed in exposed versus unexposed rats incl uded increased lung weight, inflamed lungs, emphysema, arteriolar wall thicke ning, granulomas, fibrosis, and proliferative re sponses within the alveoli (LOAEL = 34.25 μg Be/m 3 ). Guinea pigs were exposed to 0, 3.7, 15.4, or 29.3 μg

Be/m 3 (from the sulfate) for 6 hours/day, 5 days/week for up to 1 year (Reeves et al. , 1970). Respiratory alterations observed in the be ryllium-exposed groups included increased tracheobronchial lymph node and lung wet weights, interstitial pneumonitis, and granulomatous lesions. These adverse respirator y effects were observed in al l the beryllium dosed groups and indicated a chronic inhalation LOAEL of 3.7 μg Be/m 3

.

Appendix D3 39 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Wagner et al. (1969) exposed monkeys, rats, and hamsters to 0.21 and 0.62 mg Be/m 3 as fumes from bertrandite or beryl ore, respectively, fo r 6 hours/day, 5 days/week for up to 17 months. Exposed animals displayed severe effects, incl uding (1) bronchial lym phocytic infiltrates, abscesses, consolidated lobes, and granul omatous lesions after exposure to 0.21 mg Be/m 3 from bertrandite ore, and (2) inflamed lungs, fibros is, and granuloma after exposure to 0.62 mg Be/m 3
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 stud ies (Schepers, 1964; Schepers et al. , 1957; Stiefel et al. , 1980). Schepers (1964) exposed monkeys ( Macacus
mullata ) to three soluble forms of beryllium (BeF 2 , BeSO 4 , BeHPO 4 ) daily for 6 hours/day over 7 to 30 days. Increased lung weight, inflamma tion, emphysema, and fibrosis of the lung were observed after 17 days at 0.198 mg Be/m 3 (as BeSO 4 ). 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/m 3 as the sulfate or fluoride. A subchronic inhalation study reported immunological effects as increased, berylliu m-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/m 3 ) (Stiefel et al. , 1980). However, a subchronic inha lation study in Wistar and Sherman rats (Schepers et al. , 1957) observed multiple lung altera tions including granulomas (LOAEL = 35 μg Be/m 3 ) 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 BeSO 4

.

Appendix D3 40 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference 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/m 3 (median exposure of sensitized workers)

NOAEL Not observed Exposure continuity Workplace Average experimental exposure 0.2 μg/m 3 for LOAEL group (0.55 x 10/20 x 5/7) Human equivalent concentration 0.2 μg/m 3
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 μ g/m 3

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/m 3 (geometric mean of range of measured exposures associated with berylliosis of 0.01 to 0.1 μg/m 3 )

NOAEL Not observed Exposure continuity Continuous Average exposure Estimated to be approximately 0.3 μg/m 3 (historical exposures estimated to be 10-fold higher than measured values) for LOAEL group Human equivalent

concentration 0.3 μg/m 3 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/m 3

Appendix D3 41 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

U.S. EPA (1998) developed an RfC of 0.02 μg/m 3 based on beryllium sensitization and progression to chronic beryllium disease (CBD) identified by Kreiss et al . (1996). The Kreiss et al . (1996) occupational exposure st udy identified a LOAEL for beryllium sensitization in workers of 0.55 μg/m 3 (median of average exposure conc entrations 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/m 3 in community residents living near a beryllium plant. U.S. EPA used the

LOAEL from the Kreiss et al . (1996) study for the operational deri vation 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/m 3 , then a total UF of 10 was used to obtain the RfC of 0.02 μg/m 3 . The UF of 10 was comprised of a UF of 3 to account for the sensi tive nature of the subclinical endpoint (beryllium sensitization) and a database UF of 3 to account for the poor qua lity 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 inci dence 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 irreversib le, 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 da tabase 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 popul ation who are by nature more sens itive 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 th e effects are seen in a "sensitive subpopulation," OEHHA applied an intr aspecies uncertainty factor (UF H ). OEHHA used an

intermediate UF H of 3, since 1) there may be other populati on factors involved in being sensitive, such as immature lungs, and 2) all the dis eased were initially h ealthy adult workers.

For comparison the LOAEL from guinea pigs of 3.7 μg Be/m 3 (Reeves et al. , 1970) is equivalent to a continuous exposure of 0.66 μg/m 3 . Division by UFs of 10 for intraspecies, 10 for interspecies (since HEC adjustments are not ava ilable yet for guinea pigs), and 10 for use of a LOAEL results in a REL of 0.0007 μg/m 3

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 uncer tainties 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 CompoundsDetermination of Noncancer Chronic Re ference 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 ex posure 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)

BMD 05 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-te rm feeding study in which beagle dogs (aged 8 to 12 mo) were fed diets (for 1 h per day) cont aining 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) a nd inflammatory lesions in the gastrointestinal tract. These occurred predominantly in the smal l intestine and to a lesser extent in the stomach and large intestine, and were considered treatm ent related. All animals with stomach or large

intestinal lesions also had lesions in the sma ll 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 ul ceration. High-dose animals also showed

Appendix D3 43 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

moderate to marked erythroid hypoplasia of the bone marrow, whic h 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 attrib uted to a likely systemic bacterial invasion through the damaged intestinal mu cosa.) In the 50 ppm group, one female dog, which died after 70 weeks of treatment, showed gastrointestinal lesions, which were less se vere, 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 predominan ce of lesions in the small intestin e may have been partly due to precipitation of beryllium phosphate there due to th e slightly alkaline pH (Reeves, 1965). Thus 500 ppm was a LOAEL and 50 ppm was a NOAEL (sta tistically) for gastro intestinal 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 BD 10 approach and included a database UF of 3. OEHHA used a BD 05 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 chr onic 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. Bery llium 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 Di sease Registry. U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic S ubstances and Disease Registry. Toxicological Profile for Beryllium. TP-92/04. Atlanta, GA: ATSDR. April 1993.

ATSDR. 2000. Agency for Toxic Substances and Di sease Registry. U.S. Department of Health and Human Services, Public Health Service, Agency for Toxic S ubstances and Disease Registry. Toxicological Profile for Beryllium (update). Drea ft for Public Comment. Atlanta, GA: ATSDR. September 2000.

CARB. 2000. California Air Resources Board. Calif ornia Emissions Inventory Development and Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

Cotes JE, Gilson JC, McKerrow CB, and Oldha m 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 CompoundsDetermination of Noncancer Chronic Re ference 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 fu me. Am. Rev. Respir. Dis. 135:201-208.

Eisenbud M, and Lisson J. 1983. Epidemiological aspects of beryllium-induced nonmalignant lung disease: A 30 year upda te. 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. H yg. 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, Ke nt 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 worker s in a spacecraft-manufacturing plant -

California. MMWR. 32:419-425.

Kreiss K, Newman LS, Mroz MM, and Cam pbell PA. 1989. Screening blood test identifies subclinical beryllium disease. J. Occup. Med. 31(7):603-608.

Kreiss K, Mroz MM, Newman LS, Martyny J, a nd Zhen B. 1996. Machining risk of beryllium

disease and sensitization with medi an exposures below 2 micrograms/m 3 . Am. J. Ind. Med. 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 beryl lium workers. Br. J. Ind. Med.

45:167-173.

Kriebel D, Sprince NL, Eisen EA, and Greav es 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 diseas e associated with beryllium disease - A case study. J. Occup. Med. 3:341-345.

Appendix D3 45 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Morgareidge K, Cox GE and Gallo MA. (1976) Ch ronic feeding studies with beryllium in dogs. Food and Drug Research Laboratories, Inc. Subm itted 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 na tural history of beryllium sensitization and chronic beryllium disease. Environ. Health Perspect. 104S(5):937-943.

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce Exposure Levels. Available on-line at http://www.oehha.ca.gov

Owen BA. 1990. Literature-derived absorption co efficients 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 th e gastrointestinal tract. Arch. Environ. Health. 11(2):209-214.

Reeves AL, Deitch D, and Vorwald AJ. 1967. Bery llium carcinogenesis. I. Inhalation exposure of rats to beryllium sulfate aerosol. Cancer Res. 27:439-443.

Reeves AL, Swanborg RH, Busby EK, and Kriv anek ND. 1970. In: Target Organ Toxicology Series: Immunotoxicology and I mmunopharmacology. 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 glutam ate 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 be ryllium: Reaction of the monkey to inhaled aerosols. Ind. Med. Surg. 33:1-16.

Schepers GW, Durkhan TM, Delahunt AB, a nd 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 st udies in rats: Effects of aluminum, barium, beryllium and tungsten. J. Nutr. 105:421-427.

Appendix D3 46 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Smith CJ, Livingston SD and Doolittle DJ. 1997. An international literatu re 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. Possi ble health risks from low level exposure to beryllium. Toxicology 111(1-3):213-224.

Stiefel T, Schulze K, Zorn H, and Toelg G. 1980. Toxicokinetic and toxicodynamic studies of beryllium. Arch. Toxicol. 45:81-92.

U.S. Environmental Protection Agency. 1998. Integrated Risk Information System (IRIS) 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 cas e-control study. AIHAJ. 61(2):244-254.

Vorwald A, and Reeves A. 1959. Pathologic ch anges induced by beryllium compounds. Arch.

Indust. Health 19:190-199.

Wagner W, Groth D, Holtz J, Madden G, and St okinger H. 1969. Comparative chronic inhalation toxicity of beryllium ores bert randite and beryl, with production of pulmonary tumors by beryl. Toxicol. Appl. Pharmacol. 15:10-29.

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 beryl lium 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 te st and the beryllium levels in working environment. Ind. Health 35(3):374-379.

Appendix D3 47 Beryllium and Beryllium CompoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# 1,3-BUTADIENE

(butadiene; buta-1,3-diene; biethylene; bivinyl; di vinyl; vinylethylene)

CAS Registry Number: 106-99-0

I. Chronic Toxicity Summary

Inhalation reference exposure level 20 μg/m 3 (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 C 4 H 6 Molecular weight 54.09 g/mol Boiling point

− 4.4 o C Melting point

− 108.9 o C Vapor pressure 910 torr at 20 o 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/m 3 at 25 o C

III. Major Uses and Sources

1,3-Butadiene is a major commodity product of th e petrochemical industry, usually produced as a by-product of ethylene. The majority of 1,3-but adiene 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 indus trial 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 produc ts, and cigarette smoke (HSDB, 2000). In 1996, the latest year tabulated, th e statewide mean outdoor monitore d concentration of 1,3-butadiene was approximately 0.2 ppb (CARB, 1999). The Sout h Coast Air Quality Ma nagement 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 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

based on the most recent inventory were estimat ed 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 followi ng acute exposure to unknown levels of 1,3- butadiene (Wilson, 1944). Additional symptoms reported included coughing, fatigue, and drowsiness; however, all symptoms ceas ed on removal from the exposure.

Studies on the chronic effects of 1,3-butadiene have been cent ered in the styrene-butadiene rubber manufacturing industry, wh ich uses large quantities of 1,3-butadiene, and in the 1,3- butadiene monomer industry. On e retrospective epidemiological study reported an increase in overall mortality, emphysema, and cardiovasc ular diseases (chronic rheumatic and

arteriosclerotic hear t 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 fo r deaths from arteriosclerotic heart disease was elevated (

∼ 1.4-1.8) among black workers in the 1,3-bu tadiene rubber industry. Othe r occupational studies have described the potential for adverse hematologica l effects due to butadie ne exposure (Checkoway and Williams, 1982; McMichael et al. , 1975). A survey of workers at a styrene-butadiene rubber plant revealed slightly lower levels (but w ithin normal range) of re d 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 chem icals, the specific contribution of butadiene to the adverse respir atory and hematopoietic effects remains unclear.

V. Effects of Animal Exposure

The few available chronic animal inhalati on studies have focused on the potential carcinogenicity of 1,3-butadiene. The Nationa l Toxicology Program (NTP) has sponsored two

chronic inhalation studies in B6C3F 1 mice (NTP, 1984; Melnick et al. , 1990; NTP, 1993), while Hazelton Laboratories Europe (HLE) Ltd. conduct ed a chronic inhalation study in SpragueDawley rats (HLE, 1981; Owen et al. , 1987; Owen and Glaister, 1990).

The two B6C3F 1 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 earlie r NTP (1984) study in mice admi nistered 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 ova rian 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, adve rse changes in the nasa l cavity included chronic inflammation, fibrosis, cartilaginous metaplasia, osseous metaplas ia, 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 gon adal atrophy in both sexes.

Appendix D3 49 ButadieneDetermination of Noncancer Chronic Re ference 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) admi nistered 6 hours/day, 5 days/week for up to 2 years. Two-year survival was signifi cantly decreased in mice exposed to 20 ppm and greater, primarily due to chemical-related mali gnant neoplasms. Increased incidences of nonneoplastic lesions in exposed mice included bone marrow atrophy, gonadal atrophy (testicular, ovarian and uterine), angiectas is, alveolar epithelial hyperpla sia, forestomach epithelial hyperplasia, and cardiac endotheli al hyperplasia. Gonadal atr ophy was observed at 200 ppm and 625 ppm for males and at 6.25 ppm and higher for fe males. Bone marrow t oxicity (regenerative anemia) was seen at 62.5 ppm and higher. Th is 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-y ear chronic inhalation toxicity study sponsored by the International Institute of Synthetic Rubber Producers (IISRP) at Hazelton Laboratories Europe, Ltd (1981) on Sprague-Dawley rats e xposed 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-butad iene. Alterations in or gan weight were also observed in this high exposure grou p. A dose-related increase in liver weights was observed at both the 52-week interim kill and at study termina tion. Absolute and relative kidney weight was also significantly increased and associated wi th 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 (y oung male chickens) to 0 or 20 ppm 1,3- butadiene 6 hr/day, 5 days/week for 16 weeks to st udy arteriosclerotic plaque development. The cockerel is a sensitive animal model for studying the effects of environm ental arteriosclerotic plaque-promoting agents. Plaque frequency and location were not affect ed. However, plaque sizes were significantly larger in 1,3-butadie ne-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-butadi ene exposure in rats. One study (reported by Ripp in 1967) continuously expos ed rats to relatively lower concentrations of 0.45, 1.4 or 13.5 ppm. At 13.5 ppm, blood cholin esterase was elevated, blood pr essure was lowered, and motor

Appendix D3 50 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

activity was decreased. Histopa thological changes reported at 0.45 ppm were congestion in the spleen and hyperemia and leukocyt e infiltration of cardiac tissue. Alterations in lung tissue noted at 1.4 and 13.5 ppm included at electasis, interstitia l pneumonia, and emphysema. No other studies used such low exposure levels or measur ed 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 t oxicity studies undertak en by U.S. EPA was summarized by Morrissey et al . (1990). In developmen tal 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 ev idence of developmental toxicity was observed. In mice, maternal body weight gain and extra-ge stational body weight ga in 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, gr oups 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 st atistically significant on ly 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 in creased beginning at 1000 ppm at week 1, and 200 and 1000 ppm at week 2. While not strong ly concentration dependent, the dominant lethality results are consistent with an advers e effect of 1,3-butadiene on more mature cells (spermatozoa and spermatids).

An acute and subchronic (10 week ) study identified male-mediated F 1 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/m 3 ) 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/m 3 ) 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 s ubchronic 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 detect ed by the Comet assay, wa s observed at 125 ppm.

Further dominant lethality studies in rodents by the same resear ch 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 statisti cally significant increase in early deaths was

Appendix D3 51 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

observed at 65 and 130 ppm but was not dose-related. Male-mediate d effects in rats were not observed at any exposure level.

Pacchierotti et al . (1998) investigated 1,3-but adiene-induced toxic eff ects on spermatogenic cell stages and first-cleavage embryos. Expos ure 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 unfertili zed oocytes after pairing with untreated females. However, statistically significant increases of cy togenetic aberrations in first-cleavage embryos were observed in the fi rst mating week in mice exposed to 500 and 1300 ppm, and in the second mating week in mice trea ted 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 la ter by a similar decrease of elongated spermatids. Testis wei ght was significantly reduced at all doses tested, 21 days after the end of exposure. A dose-dependent increas e of variant sperm w ith 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 B6C3F 1 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

BMC 05 1.40 ppm Exposure continuity 6 hr/d, 5 d/wk Exposure duration 103 weeks Average experimental exposure 0.25 ppm for BMC 05 (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/m 3 ; 20 μg/m 3 )

The chronic REL for butadiene is based on an in creased 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 terminati on 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 mi ce exposed for 9-months, ovarian atrophy was

Appendix D3 52 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

observed at 200 and 625 ppm (442 or 1381 mg/m 3 , respectively). After 15 months, ovarian atrophy was observed at exposur e levels of 20 ppm (44.2 mg/m 3 ) and above. In mice exposed for up to 2 years (103 weeks), the incidence of ovarian atrophy increased at all exposure

concentrations relative to c ontrols, which establishes a chronic LOAEL of 6.25 ppm (13.81 mg/m 3 ) for reproductive toxicity.

Presentation of the ovarian atrophy data in quan tal form (see Table 1) allows the use of the benchmark concentration (BMC) approach to dete rmine 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 ova rian atrophy in the two high est exposure groups did not increase appreciably with increasing exposure concentration, and therefore, deviated from the log-normal probit plot. The signifi cantly shortened survival rate in these two groups may be one

reason for this deviation. Anot her possible cause is that a re latively resistant subgroup of mice (to ovarian atrophy) is reveal ed at the two highest doses following 2-year exposure to 1,3- butadiene. Thus, it may be biologically plausi ble to remove these resistant subgroups when using a BMC approach. The maximum likeli hood estimate (MLE) for a 5% response was 1.53 ppm. The resulting 95% lower confid ence limit at the MLE provided a BMC 05 of 1.40 ppm. A

BMC 05 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 butad iene 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 co mpared to the rats. Dahl et al . (1991) reported that, for equivale nt inhalation exposures, the con centrations 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 lik e the monkey than the mouse or the rat in their formation of epoxides from butadiene. In vitr o metabolism studies with human liver tissue present conflicting results regardi ng 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 atro phy 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 in appropriate to develop he alth-protective values for 1,3-butadiene based on 2-year ovarian atr ophy in mice because the mice are beyond their normal reproductive age. It was suggested th at the 15-month evalua tion of ovarian atrophy conducted by the NTP (1993) would be a better indicator of reproducti ve risk. However,

OEHHA staff believes that butadie ne-induced ovarian atrophy repres ents 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 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

evaluation for ovarian atrophy can be estimate d. 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 observe d 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 BMC 05 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 inte rspecies variability and 10 for intraspecies variability), a REL of 20 ppb (40 μ g/m 3 ) 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 observ ed at 65 and 125 ppm, but not 12.5 ppm.

An earlier dominant lethality study (Anderson et al ., 1996) indicated that ear ly deaths may occur at 12.5 ppm but the toxicological eff ect could not be repeated at th is concentration in subsequent studies. The average exposure duration at th e 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/m 3 ). 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 a ccelerating arteriosclerotic plaque development in cockerels (Penn and Snyder, 1996a,b), although no animal studies in mammals have implicated 1,3-butadiene in this dis ease. 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 circul atory diseases when compared to a reference population, suggesting a ‘healthy worker’ eff ect (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 significan t, increased mortality from arteriosclerosis compared to the black male U.S. populat ion (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 arte riosclerosis, and mortality (Sathiakumar et al ., 1998). Weaknesses in these worker analyses include re latively small cohort sizes, the bias of having

racial information on all deaths and not on all livi ng 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 arteriosclerot ic (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 studie s to determine if there is a true association between 1,3-butadiene exposure a nd arteriosclerotic diseases.

Appendix D3 54 ButadieneDetermination of Noncancer Chronic Re ference 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 da ta 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 F 1 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 expo sed to 1,3-butadiene. Mutat. Res. 397:77-84.

Bond JA, Himmelstein MW, Seaton M, Boogaar d P, and Medinsky MA. 1996. Metabolism of butadiene by mice, rats, and humans: a comparis on of physiologically based toxicokinetic model predictions and experimental data. Toxicology 113:48-54.

Brinkworth MH, Anderson D, Hughes JA, Jacks on LI, Yu T-W, and Nieschlag E. 1998. Genetic effects of 1,3-butadiene on the m ouse testis. Mutat. Res. 397:67-75.

CARB. 1999. California Air Resources Board. Toxi cs 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. Calif ornia Emissions Inventory Development and Reporting System. (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

CDC. 2000. Centers for Disease Co ntrol and Prevention. National Ce nter for Health Statistics. Available online at http://www.cd c.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, Gr iffith WC Jr, Mauderly JL, Muggenburg BA, Sabourin PJ, Henderson RF. 1991. Toxicokinetic s of inhaled 1,3-butadiene in monkeys: comparison to toxicokinetics in rats and mi ce. Toxicol. Appl. Pharmacol. 110(1):9-19. (Published erratum appears in Toxico l. Appl. Pharmacol. 1992. 116(1):152.)

Appendix D3 55 ButadieneDetermination of Noncancer Chronic Re ference 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 Spra gue-Dawley rats. Toxicology 113(1-3):128-136.

Duescher RJ and Elfarra AA. 1994. Human liver microsomes are effici ent catalysts of 1,3- butadiene oxidation: evidence for major ro les 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. Cr it. Rev. Toxicol. 27(1):1-108.

HLE. 1981. Hazelton Laboratories Eu rope, Ltd. 1,3-Butadiene: The toxicity and carcinogenicity of butadiene gas administered to rats by inha lation for approximately 24 months. Prepared for the International Institute of Synthetic R ubber Producers, New York, NY. Unpublished. [as reported in U.S. EPA, 1985].

HSDB. 2000. Hazardous Substances Data Bank. Nati onal Library of Medicine, Bethesda, MD. Available online through Toxicology Da ta 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 cohor t of workers in the styrene-butadiene polymer manufacturing indus try (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 T ousey 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 con centrations. Cancer Res. 50(20):6592-6599.

Miller LM. 1978. Investigation of selected potential environmenta l contaminants: Butadiene and its oligomers. Philadelphia, PA: Franklin Research Center. [a s 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 an d developmental toxi city studies of 1,3-

butadiene in rodents. Envi ron. Health Perspect. 86:79-84.

Appendix D3 56 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

NTP. 1984. U.S. National Toxicology Program. T oxicology and Carcinogenesis Studies of 1,3- Butadiene (CAS No.106-99-0) in B6C3F1 Mice (Inhalation St udies) (TR No. 288) Research Triangle Park, NC: National Institute of Environmental Health Sciences.

NTP. 1993. U.S. National Toxicology Program. T oxicology and Carcinogenesis Studies of 1,3- Butadiene (CAS No.106-99-0) in B6C3F1 Mice (Inhalation St udies) (TR No. 434). Research Triangle Park, NC: National Institute of Environmental Health Sciences.

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce 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 toxic ity 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-butad iene in the mouse: cytogene tic 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 va por 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 Mana gement 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 ra ts. Carcinogenesis 18:611-625.

U.S. EPA. 1985. U.S. Environmental Protecti on 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 ButadieneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Wilson RH. 1944. Health hazards encountered in the manufacture of synthetic rubber. JAMA 124(11):701-703.

Appendix D3 58 ButadieneDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (respirable) Critical effect(s) Kidney effects (protein uria) and respiratory effects (reduction in forc ed 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/cm 3 @ 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 us e 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 statew ide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in Califor nia 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 th an 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. Alt hough urine cadmium concentrations were significantly elevated, the subj ects in E1 did not exhibit pulmonary function changes or

Appendix D3 59 Cadmium and Cadmium compoundsDetermination of Noncancer Chronic Re ference 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 μ g/m 3 total cadmium (1.4 μ g/m 3 respirable cadmium). The 27 workers in E2 had been exposed for a m ean of 8.6 years to 134 μ g/m 3 total cadmium (88 μ g/m 3 respirable cadmium). The blood and urinary cadmium levels of these workers were also significantly elevated compared to matched controls. Glom erular proteinuria was observed in 15% of the
workers in E2 and in 68% of workers in E3. Th e 22 workers of E3 had been exposed for a mean of 27.8 years to 66 μ g/m 3 total cadmium (21 μ g/m 3 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 pulmo nary function (forced v ital capacity, forced expiratory volume in one second, and peak expira tory 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 μ g/m 3 for workers exposed for 28 years and a NOAEL of 1.4 μ g/m 3
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 β 2 -microglobulin) was consistently observed when the worker’s liver cadmium burden exceeded 40 ppm and the time-weighted cumulative exposure

index exceeded 400-500 μ g years/m 3

.

A detailed investigation of renal function in 75 male cadmium-exposed workers identified significant increases in urinary excretion of se veral low- and high molecular weight proteins, including β 2 -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 μ g/m 3 , were determined from background or personal exposure measurements made between 1964 and 1983, or were estimated. A tim e-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 urin ary retinol binding pr otein and urinary β 2 -microglobulin. Of these, the most sensitive parameter, urinary β 2 -microglobulin, demonstrated an inflection point at 1108 μ g years/m 3 with a 95% lower confidence limit of 509 μ g years/m 3 . The endpoint

selected is indicative of defects in tubular reabsorption of proteins.

Diminished sensitivity of smell has also b een observed in cadmium exposed workers (Rose et
al. , 1992). Cadmium body burden, β 2 -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 exhibite d 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 le ukocytes and hyaline casts were observed in the kidneys of rabbits following exposure to 6.5 mg/m 3 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 exposur e. Increased lung weights and emphysema were

Appendix D3 60 Cadmium and Cadmium compoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

also observed. The trachea and nasal mucous membranes exhibited chronic inflammatory changes (not specified) and lymphocyte infilt ration. The kidney contained the greatest concentration of cadmium. This study al so exposed a group of rabbits to 9.1 mg/m 3
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, a nd 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/m 3 . Chronic bronchitis and hyperplasi a of the bronchiolar epithelium were observed in the higher dose group in a ddition to the findings previously noted.

Male and female rats were exposed to 0, 0.3, 1.0, or 2.0 mg Cd/m 3 (as CdCl 2 ) 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-depe ndent increase in lung weight was observed in the remaining dose groups and a st atistically significan t increase in lung collagen and elastin was observed in rats exposed to 1.0 mg/m 3 . Pathological changes noted in the terminal bronchioles include flattening and hyperplasia of type II ce lls, and infiltration of macrophages, mononuclear cells, and polymorphonuclear leukocyt es. Proliferation of fibroblas ts with deposition of collagen

was also noted.

Male rats were exposed continuously to 0, 30, or 90 μg Cd/m 3 cadmium oxide (CdO) dust for up to 18 months (Takenaka et al. , 1990). Animals exposed to 30 μ g/m 3 were sacrificed at 6 and 18 months of exposure. Although some rats in th e 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 m onths of exposure with more marked changes observed in the high dose group. Abnormal prolifer ation 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 compoundsDetermination of Noncancer Chronic Re ference 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 - proteinu ria in 68% of LOAEL

group Respiratory effects – redu ction in forced vital capacity (FVC), forced expiratory flow in 1 second (FEV 1 ); reduction in peak expiratory

flow rate

LOAEL 21 μ g/m 3 respirable cadmium

NOAEL 1.4 μ g/m 3 respirable cadmium Exposure continuity Assumed to be 5 days/week for 8 hours/day during which 10 m 3 air is breathed Average occupational exposure 0.5 μ g/m 3 for NOAEL group (1.4 x 10/20 x 5/7) Human equivalent concentration 0.5 μ g/m 3 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 μ g/m 3

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 inco mplete knowledge of the past exposures over the full study interval and the relatively sm all 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 μ g/m 3 . The LOAEL group had an aver age occupational exposure of 5.0 μg/m 3 and an average exposure duration of 27.8 year s (21 to 40 years). Default uncertainty

Appendix D3 62 Cadmium and Cadmium compoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

factors included a 10-fold LOAEL uncertainty fa ctor and a 10-fold intr aspecies uncertainty factor (UF).

For comparison, using data presented by Ellis an d associates (1985) and Mason and associates (1993) correlating human cumulative exposures (in terms of μg-years/m 3 ) and renal tubular protein reabsorption, a LOAEL of 500 μg-years/m 3 was predicted. This correlates to 7 μg/m 3
over 70 years. A time-weighted exposure to acc ount for continuous exposure rather than 40 hour per week occupational exposure is 1.7 μg/m 3 . Applying a 10-fold LOAE L uncertainty factor and a 10-fold intraspecies uncertainty fact or results in a REL value of 0.02 μg/m 3 , 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 ex posure 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 μ g 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 expos ure (NOAEL) which results in 200 μ g 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 μ g Cd/gm wet weight human renal cortex). Thus, base d 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 fo r Cd in food is 0.001 mg Cd/kg/day.

Appendix D3 63 Cadmium and Cadmium compoundsDetermination of Noncancer Chronic Re ference 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 a nd animal toxicity data are available. The RfD is based on the highest level of Cd in the human renal cortex (i.e., th e critical level) not associated with significant protei nuria (i.e., the critical effect). A toxicokinetic model has been used to determine the highest level of exposure associ ated with the lack of a critical effect. Since the fraction of ingested Cd th at is absorbed appears to va ry 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 sensit ive individuals. No modifying factor was used.

U.S. EPA stated its confidence in the RfD as: St udy - 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 cadmi um in both humans and animals. These data also permit calculati on 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 Serv ice. 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 ki dney and liver cadmium burden. J. Toxicol. Environ. Health

15:173-187.

Friberg L. 1950. Health hazards in the manuf acture of alkaline accumulators with special reference to chronic cadmium poisoning. Acta Med. Scand. (S uppl. 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 ae rosol. J. Toxicol. Environ. Health 17:175-189.

Appendix D3 64 Cadmium and Cadmium compoundsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Mason HJ, Davison AG, Wright AL, Guth rie CJG, Fayers PM, Venables KM, et al. 1988. Relations between liver cadmium, cumulative e xposure, 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. Dr inking 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 compoundsDetermination 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/m 3 (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 CS 2 Molecular weight 76.14 g/mol Boiling point 46.5 o C Melting point

− 111.5 o C Vapor pressure 297 torr @ 20 o C Solubility Slightly soluble in wa ter (2.94 g/L); miscible in anhydrous methanol, ethanol, ether, benzene, chloroform, and carbon

tetrachloride

Conversion factor 3.1 mg/m 3 per ppm at 25 o C

III. Major Uses and Sources

The most prominent industrial use of carbon di sulfide is in the production of viscose rayon

fibers. Carbon disulfide is also used in the production of carbon tetr achloride and cellophane, and, as a solvent for rubber, sulfur, oils, resins, a nd waxes. In the past, carbon disulfide was used in soil fumigation and insect control in stored grain. Industrial pro cesses 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 me tam 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 DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

IV. Effects of Human Exposure

A primary target of carbon disulfide (CS 2 ) toxicity is the nervous system. The major neurotoxic

action of carbon disulfide is the development of mental distur bances. These include change of personality, irritability, and forgetfulness, often with accompanying neurophysiological and neuropathological changes after pr olonged exposure. Such changes include decreased peripheral nerve impulse conduction, motor and/or sensory neuropathies, cerebral orcerebellar atrophy, and neuropsychological orga nic 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 CS 2 , often in excess of 20 ppm (Foa et al. 1976; Hanninen et al.

, 1978).

Studies have identified alterations in the nerv e conduction of workers chronically exposed to

lower CS 2 levels (Hirata et al. , 1992a; Johnson et al. , 1983; Ruijten et al. , 1990; Ruijten et al.

, 1993). A cross-sectional study of Ja panese spinning workers identifi ed 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 BA EP components increased significantly in workersexposed to CS 2 for more than 20 years when compared to controls. CS 2 exposures ranged from 3.3 to 8.2 ppm (mean = 4.76 ppm). Ruijten et al. (1993) identified mild presymptomatic nerve impairment (decreased c onduction velocities and response amplitudes) in

44 CS 2 -exposed workers with an average cumulative exposure range from 192 to 213 ppm-year (mean duration = 26.1 years).

A NIOSH occupational study ev aluated the effects of CS 2 on the peripheral nervous system.

Johnson et al. (1983) identified a significant dose relate d reduction in the maximum motor nerve conduction velocities (MCV) in the calves and ankl es of male viscose rayon workers exposed to high (median = 7.6 ppm) CS 2 levels versus a comparison group exposed to low concentrations (median = 0.2 ppm). The workers were all employe d in artificial fiber production in the same plant. Since these reduced MCVs were still with in the normal range, the authors considered the

measured difference an indication of minimal ne urotoxicity. 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 identif ied a chronic LOAEL of 7.6 ppm for minor neurological effects (decreased peroneal ne rve 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/m 3 CS 2 (time-weighted average 1 to 40 mg/m 3 ) (Vanhoorne et al., 1995). Among four categories of cumulative exposure (0, 1 to 300, 301 to 600, and greater than 600 mg/m 3 -years), a clear dose-response effect was obser ved for reduced mean peroneal motor nerve

conduction velocities in both fast and slow fibers. Unfortunatel y, 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 conduc tion velocities compared with non-exposed

workers.

Vascular atherosclerotic changes are also considered a major eff ect of chronic carbon disulfide exposure. Several occupational studies have dem onstrated an increase in the mortality due to

Appendix D3 67 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

ischemic heart disease in CS 2 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 CS 2 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 myo cardial infarction in CS 2 exposed workers.

Male workers (n=177) in a Polish fiber plant were exposed to CS 2 for an average of 14 years (range of 5 to 38 years). Controls were 93 healthy male workers fr om other factories that did not use carbon disulfide. Carbon disulfide exposed workers had higher rates (42%) of 24-hour electrocardiographic abnormalities than non-expo sed workers (24%, p=0.006) (Bortkiewicz et al., 2001). The most common abnormalities were ve ntricular extrasystoles and repolarization disturbances, the latter occurring most often in workers with the longest CS 2 exposures. Longterm blood pressure monitoring did not reveal any differences between exposed and control

groups.

Male workers in a Belgian viscose rayon fact ory (n=85) were estimated by personal active sampling to have exposures of 2 to 32 mg/m 3 CS 2 . Controls were 37 non-exposed workers from

factories that did not use CS 2 . Exposed workers had reduced comm on carotid artery distensibility as measured with ultrasound sonography, while the carotid artery compliance coefficient was not significantly affected. Also, bl ood pressures and cholesterol le vels were not significantly different than observed among control workers (Kot seva et al., 2001a). Di fferences in carotid artery distensibility remain ed significant afte r 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 CS 2 for more than one year causes increases in biochemi cal changes often associat ed with cardiovascular disease - diastolic blood pressure, low density li poprotein cholesterol, an d apolipoproteins A1 and B. Egeland et al. (1992) used cross s ectional data on 165 CS 2 -exposed workers (245 controls) collected in 1979 by Fajen et al. (1981). The affected work ers 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 CS 2 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 CS 2 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, apolipoprote in B, systolic and diastolic blood pre ssure as indicative of an increased coronary risk in workers from a Belgian visc ose rayon factory (115 exposed and 76 controls).

CS 2 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, myocar dial infarction, or ischemia, were determined by ECG changes.

Workers (n=141) with a minimum of 1 year em ployment in viscose rayons factories were compared with 141 age and gender-matched plastic industry workers. Cu rrent exposures were estimated as 1 to 30 mg/m 3 (03 to 10 ppm). Exposed worker s were categorized as group 1 or

Appendix D3 68 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

group 2, with cumulative exposures of less than or greater than 100 mg/m 3 -years, respectively. Group 2 (p<0.001) but not group 1 workers had in creased mean total c holesterol (5.3 and 4.5 mmol/l) compared with controls (4.6 mmol/l) (Kotseva, 2001b).

CS 2 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 CS 2 and showed significant advers e effects on spermatogenesis, levels of serum FSH and LH, and libido; these eff ects persisted in 66% of the workers subject to follow-up. Zhou et al. (1988) investigated pregnancy outco mes and menstrual disturbances in 265 women occupationa lly exposed to CS 2 in five facilities and 291 controls. The CS 2 -exposed women had a significantly higher incidence of menstrual disturbances versus the control group (overall 34.9% vs. 18.2%). CS 2 levels varied between the five facilities (exposure category means of low = 3.1 mg/m 3 , intermediate = 6.5 mg/m 3 , and high = 14.8 mg/m 3 ), but all workers

from these CS 2 facilities had significantly higher incidences of menstrual disturbance. Irregularity of menstruation was the most comm on 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 bi rth defects among female workers occupationally exposed to CS 2 , 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 def ects (2.6% vs. 1.3%) among 682 exposed women was noted compared to 745 women in the control gr oup. 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 esti mated exposures greater than 10 mg/m 3 compared to those with lower exposures. There were no differences in rates of stillbirt h, low birth weight, or neonatal or perinatal deat hs among any of the groups.

The possibility of determining LOAEL and/or NOAEL values for the major CS 2 -related adverse effects from epidemiology studies, which pre dominately use workers from the viscose rayon industry, is limited. The limitations include in complete historical exposure measurements, concurrent exposure to other chem icals (including hydrogen sulfide or methylene chloride), lack of personal exposure determinations, and a high variability of individual exposures due to decreases of plant CS 2 concentrations over time.

V. Effects of Animal Exposure

Studies investigating the potential for CS 2 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 consis tently observed in rodents following CS 2 exposure include axonal swelling, demyelination, swellin g at neuromuscular junctions, muscle atrophy and degeneration, damage to terminal axons, an d 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 DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

exposures (250 to 800 ppm), but few studies have a ttempted to establish a dose response for this

CS 2 -induced neurotoxicity.

In a 90 day subchronic inhalation study, Sp rague-Dawley and Fischer 344 rats exposed discontinuously (6 hours/day, 5 days/week) to CS 2 developed morphological alterations in nerves including axonal swelling and myelin degr adation (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 af ter 6 and 12 months of intermittent exposure to 289 ppm CS 2 (LOAEL of 289 ppm) (Knobloch et al. , 1979).

In a 13-week subchronic study, male and fema le F344 rats inhaled 0, 50, 500, or 800 ppm CS 2 discontinuously (6 h/day, 5 days per week) (Sills et al .,1998). Development of distal axonopathy in the muscular branch of the posterior tibial ne rve (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 swel ling was noted in the spinal cords of rats exposed to 500 or 800 ppm CS 2 . In the 800 ppm group, additiona l 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 CS 2 exposure in Wistar rats exposed to 0, 73.8, 160, 321, or 546 ppm CS 2 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 th e capacity of CS2 to induce arteri al fatty deposits by itself, and its ability to enhance th e rate of fatty deposit formation induc ed 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 week s of exposure, and the rates of fatty deposit

formation under the aortic valve leaflets were ev aluated. Exposure of mice on the control diet to 500 and 800 ppm CS2 induced a small but significan t increase in the ra te of fatty deposit formation over non-exposed controls. In the an imals on the high fat diet there was marked enhancement of the rate of fatty deposit forma tion 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 conjunc tion with other risk factors, CS2 at relatively low concentra tions 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 CS 2 , usually following pretreatment with liver in ducers such as phenobarbital. Bond et al. (1969) showed that high doses of CS 2 to rats produced an increase in peripor tal 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. Othe r hepatotoxic eff ects seen after CS 2 exposures greater

Appendix D3 70 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

than 400 ppm include increases in relative liver weight (S okal, 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 LD 50 values for CS 2 were estimated in 1-, 5-, 10- , 20-, 30- and 40-day-old rats (sample size not specifie d) (Green and Hunter, 19 85). 1-day-old rats (LD 50 583 mg/kg, ip) were about 3-times more susceptible than 20-day-old rats (LD 50 1545 mg/kg, ip).

14 C- and 35 S-labelled CS 2 was given ip to 1-, 5-, 10-, 20-, 30-, and 40-day-old rats (Snyderwine and Hunter, 1987). Thirty- and forty-day-old ra ts (sample size not reported) metabolized significantly more CS 2 to CO 2 and expired significantly less CS 2 than 1- to 20-day-old rats. Twenty-four hr after administ ration, up to 13 times more 35 S -label (radioactivity per g of tissue) were present in organs from 1-da y-old rats than in similar orga ns 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 CS 2 would lead to higher tissue concentrations and thus , potentially, increased toxicity.

The metabolite responsible for CS 2 hepatotoxicity is believed to be reactive sulfur atoms that covalently bind to cellular m acromolecules (Dalvi, 1988). Sim ilarly, the correlation between increased lethality (Green and Hunter, 1985) and increasing binding of 35 S –label (Snyderwine and Hunter, 1987) in younger CS 2 -exposed animals is consistent with a role for reactive sulfur. Neurotoxicity of CS 2 results from the formation of th iourea lysine cro ss-links between neurofilament proteins (DeCap rio et al., 1992; Valentine et al., 1997; Erve et al., 1998).

New Zealand white rabbits (24 per gr oup) inhaled 0, 60, 100, 300, 600 or 1200 ppm CS 2 for 6 h/d on gestation days 6 to 18 (Pathology Associ ates, 1991). Developmental toxicity (NOAEL = 300 ppm; 930 mg/m 3 ) was noted at concentrations lower th an those associated with significant maternal toxicity (NOAEL = 600 ppm; 1860 mg/ m 3 ) (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 finding s. Reduced fetal body weights were noted at 600 and 1200 ppm. Cumulative malformations were increased in the 1200 (3720 mg/m 3 ) but not 600 ppm group, though there were no significant incr eases in any specific malformation in any group. Maternal effects at 1200 ppm included de creased 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/m 3 (32 ppm) for 4 hr/d on gestati on days 7 and 8, and the embryos explanted to culture medium at day 9.5. Grow th of explants of 10 treated and 17 control embryos was monitored for 44 hours. CS 2 at this concentration indu ced 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/m 3 (0.01, 3, 32, or 64 ppm) CS 2 . The two highest dose levels were both teratogenic and maternally neurot oxic. 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/m 3 , as well as increased postnatal

Appendix D3 71 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

neurological effects including hypoactivity, mild ataxia and gait disturbances, hind-limb weakness, spinning and tremor (T abacova et al., 1983). While the overall rate of malformations (club foot, hydrocephalus, microcephalus, gene ralized edema) exhibited a dose-response trend,

with increased effects in the F2 generation, the specific malformati ons exhibited a less-consistent pattern. For example, while club foot was the predominant malformation in the F1 fetuses (occurring at 100 and 200 mg/m 3 ); much lower rates of club foot were noted in the F2 generation (including none in the 200 mg/m 3 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 re porting on the statistical signif icance of reported behavioral

effects.

Wistar albino rats (32 animals per gr oup) were exposed to 50, 100, or 200 mg/m 3 (16, 32, or 64 ppm) CS 2 for 8 hours per day throughout gestation. Th ere were no statistically significant results in the 50 mg/m 3 group. In the 100 and 200 mg/m 3 groups, there were sta tistically significant increases in reduced fetal body weights, and re duced post natal body weights for 21 days, which subsequently disappeared. There was an increas e in external malformations (hydrocephalus, club foot, and tail deformations) at the two higher doses (Tab acova et al., 1978).

Behavioral effects were examined in the offspr ing of Lati:CFY rats (8 per group) exposed to 0, 10, 700, or 2000 mg/m 3 CS 2 (3, 230, or 640 ppm) for 6 hours per days over days 7 to 15 of gestation. The two high doses caused significan t perinatal mortality. Avoidance conditioning was tested using a bell as a conditional stimulus pr ior 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 condi tioning in the treated versus control animals. The authors reported th at there was a dose-relate d 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/m 3 compared with 700 mg/m 3 , the effect was virtually identical between the 10 and 700 mg/m 3 . This lack of dose-response effect raises some question about the significance of this finding.

Effects of low (0.03 and 10 mg/m 3 ; 0.01 and 3 ppm) prenatal e xposures (8 hours per day throughout gestation) of CS 2 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/m 3 group. Delays in the development of visual and auditory function we re reported in the higher dose group (Tabacova and Balabaeva, 1980). There was no mention of maternal toxi city in this study.

Several other studies yielded eith er no teratogenic effects or e ffects only at maternally toxic exposures. Saillenfait et al. (1989) exposed rats via inhalation to 0, 100, 200, 400, or 800 ppm

CS 2 for 6h/d during days 6-20 of gestation. Lower exposures (100 or 200 ppm; 310 or 620 mg/m 3 ) were not associated with maternal toxici ty or adverse effects on the developing embryo or fetus. Higher concentrations (400 or 800 ppm; 1240 or 2480 mg/m 3 ) yielded a significant reduction of maternal weight gain as well as re ductions of fetal body weight and a low incidence of club foot. Significant increases in unossi fied sternebrae were reported following 800 ppm (2480 mg/m 3 ) exposures. Nemec et al. (1993) repor ted no teratogenicity or maternal,

Appendix D3 72 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

developmental, or reproductive toxicity among pr egnant CD rats and their offspring following exposure to 125 or 250 ppm (388 or 775 mg/m 3 ) from 2 weeks prior to mating through gestation day 19. At 500 ppm, dams had decreased body we ight gain and food consumption; decreased litter viability but no teratoge nic effects were noted. CS 2 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 effect s were noted in animal inhalation exposures (20 to 40 ppm; 62 to 125 mg/m 3 CS 2 ) 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 gro up (7.6 x 10/20 x 5/7) Benchmark concentration (BMC 05 ) 6.86 ppm (continuity-weighted exposure of 2.54 ppm) Human equivalent concentration 2.54 ppm for BMC 05 (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/m 3 ; 800 μg/m 3 )

A benchmark dose analysis was performed on the peroneal MCV data. The NIOSH exposure data were regrouped into 8 geometrica lly spaced dose groups (Table 1).

Appendix D3 73 Carbon DisulfideDetermination 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. Envir onmental Protection Agency BMDS Benchmark Dose Software, Version 1.3. Four continuous da ta models were compared: linear, polynomial (v. 2.1), power (v. 2.1) and hill (v. 2.1) models. All four models adeq uately fit the data set (Table 2).

Table 2. Benchmark dose modeling results

Model MLE 05 (ppm-mo) BMC 05 (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 BMC 05 from the best-fitting linear model was used. An occupational BMC 05 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 valu e was thus 2.5 ppm (6.9 ppm x 10/20 x 5/7).

Appendix D3 74 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

36 38 40 42 44 46 48

0 500 1000 1500 2000 2500 3000 3500 4000

CS 2 Exposure (ppm-months) Linear Model (0.95 Confidence Level)

MLE 05 BMC 05 Peroneal
MCV (m/s)

The U.S. EPA (1995) based its RfC of 700 μ g/m 3 on the same study but used a BMC 10 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 BMC 05 since in practice it te nds to be closer to

the NOAEL while the BMC 10 is often closer to the LOAEL (OEHHA, 2000).

For comparison, 50 ppm was a 13 week NOAEL in rats for a xonal swelling (Sills et al ., 1998). The equivalent continuous exposure is 8.9 ppm. Us e of an RGDR of 1, an interspecies UF of 3, a subchronic UF of 3, and an intraspeci es 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 disulf ide are the use of huma n data, the observation of a dose-response effect, and the duration of ex posures. The major uncertainties are the poor quantitation of actual exposure ma gnitude 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 toxic ity of carbon disulfide are equivocal. Several studies reported that adverse de velopmental effects are only noted with exposures exceeding 100 ppm, while Tabacova and Balabaeva (1980) and Le hotsky et al. (1985) re ported transient effects The data available on the developmental toxic ity of carbon disulfide are equivocal. Several studies reported that adverse de velopmental effects are only noted with exposures exceeding 100 ppm, while Tabacova and Balabaeva (1980) and Le hotsky et al. (1985) re ported transient effects

Appendix D3 75 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

at levels as low as 10 mg/m3 (3 ppm). The resu lts 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 belo w the REL of 800 μg/m3 (300 ppb).

IX. References

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Aaserud O, Hommeren OJ, Tvedt B, Nakstad P, Mowe G, Efskind J, Russell D, Jorgensen EB, Nyberg-Hansen R, and Rootwelt K. 1990. Carbon disulfide exposure and neurotoxic sequelae among viscose rayon workers. Am. J. Ind. Med. 18(1):25-37.

Aaserud O, Russell D, Nyberg-Hansen R, Rootwe lt K, Jorgensen EB, Nakstad P, Hommeren OJ, Tvedt B, and Gjerstad L. 1992. Regional cerebral blood flow after long-term exposure to carbon disulfide. Acta Neurol. Scand. 85(4):266-271.

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Appendix D3 77 Carbon DisulfideDetermination of Noncancer Chronic Reference Exposure Levels May 2002

HSDB. 1995. Hazardous Substances Data Bank. (Edi tion of January 1995). National Library of Medicine, Bethesda, MD (TOMES ® CD-Rom Version). Denver, CO: Micromedex, Inc.

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Johnson BL, Boyd J, Burg JR, Lee ST, Xintar as C, and Albright BE. 1983. Effects on the peripheral nervous system of workers’ exposur e to carbon disulfide. Neurotoxicology 4(1)53-65.

Knobloch K, Stetkiewicz J, and Wronska-Nofe r T. 1979. Conduction velocity in the peripheral

nerves of rats with chronic carbon disul phide neuropathy. Br. J. Ind. Med. 36(2):148-152.

Kotseva K, Braeckman L, Duprez D, De Bac quer D, De Buyzere M, Van De Veire N, Vanhoorne M. 2001a. Decreased carotid artery distensi bility as a sign of ear ly atherosclerosis in viscose rayon workers. O ccup Med (Lond) 51(4):223-9.

Kotseva K. 2001b. Occupational exposure to low c oncentrations of carbon disulfide as a risk factor for hypercholesterolaemia. Int Ar ch Occup Environ Health 74(1):38-42.

Lancranjan I, Popescu HI, Klepsch I. 1969. Ch anges of the gonadic function in chronic carbon disulphide poisoning. Med. Lav. 60(10):566-571.

Lancranjan I. 1972. Alterations of spermatic li quid in patients chronically poisoned by carbon disulphide. Med. Lav. 63(1):29-33.

Lehotsky K, Szeberenyi JM, Ungvary G, Kiss A (1985). Behavioural e ffects of prenatal exposure to carbon disulfide and to aromatol in rats. Arch Toxicol Suppl 8:442-446.

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C57BL/6 mice to carbon di sulfide induces early lesions of athe rosclerosis and enhances arterial fatty deposits induced by a high fa t diet. Toxicol. Sci. 49(1):124-132.

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Appendix D3 80 Carbon DisulfideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# CARBON TETRACHLORIDE

(carbon chloride; carbon tet; freon 10; halon- 104; methane tetrachl oride; necatrine; tetrachlorocarbon; tetrachlorometha ne; tetraform; tetrasol; univerm)

CAS Registry Number: 56-23-5

I. Chronic Toxicity Summary

Inhalation reference exposure level 40 μ g/m 3 (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 CCl 4 Molecular weight 153.8 g/mol Density 1.59 g/cm 3 @ 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/m 3 @ 25°C

III. Major Uses or Sources

Carbon tetrachloride was formerly used for meta l degreasing and as a dr y-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. Ca rbon 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 stat ewide mean outdoor monitored con centration of carbon tetrachloride was approximately 0.08 ppb (CARB, 1999a). The annual statewide industrial emissions from facilities reporting under the Air Toxics Hot S pots Act in California based on the most recent inventory were estimated to be 8781 pounds of carbon tetrachloride (CARB, 2000).

Appendix D3 81 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference 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/m 3 ) carbon tetrachloride (CCl 4 ) vapor. Fifteen of the 17 workers complained of symptoms including nausea, anorexia, vomiting, flatulence, epigastric discomfort or distention, depressive symp toms, headache or giddiness for up to 4 months prior to the evaluation. A week after CCl 4 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 CCl 4 for varying duration. For each exposure group, two control groups were devi sed (unexposed and air-exposed controls) consisting of animals similar in age, sex, wei ght and number. The 2 control groups responded similarly to the expe rimental 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 ch anges in all the species tested.

Fifteen male and 15 female ra ts were exposed to 50 ppm CCl 4 134 times for 187 days. They experienced decreased body weight gain and liver weight increa se as well as moderate fatty degeneration and slight to moderate liver cirrhos is. 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 degenera tion and liver cirrhosis, a nd increased levels of liver total lipids, neutral fat, esterified c holesterol 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 fatt y 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 othe r adverse effects were seen w ith 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 CCl 4 , 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 modera te 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 expos ed 133 times over 185 days and 5 female guinea pigs were exposed 93 times over 126 days. Symp toms included growth de pression, liver weight

Appendix D3 82 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

increase, increased plasma prot hrombin 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

CCl 4 . These rats exhibited increase in liver weig ht, 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 e xposed 139 times over 197 days, experienced liver weight increase, slight to mode rate fatty degeneration without cirrhosis, and increased levels of total lipid, neutral fat, and esteri fied cholesterol. In an additi onal 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 th e guinea pigs and showed no symptoms as a result of the exposure. Sixteen additional guinea pigs developed hepatic change s after 12 exposures in 16 days.

Twenty-five male and 23 female rats, e xposed 145 times over 205 days to 5 ppm CCl 4 , had no

adverse effects. Nine male and nine female guinea pigs exposed 143 times over 203 days showed a statistically significant increase in th e liver weights (females only), but only slightly higher liver lipid content. No additional histopa thological 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/m 3 (81.7 ppm) carbon tetrachloride (CCl 4 ) 8 hours a day, 5 days a week, for 6 weeks. (This intermittent exposure is equivalent to a continuous exposure to 123 mg/m 3 .) Additionally, two 90 day continuous exposure studie s were conducted. One study e xposed 15 rats, 15 guinea pigs, 2 rabbits, 2 dogs and 3 monkeys to 61 + 5.2 mg/m 3 CCl 4 and the other exposed 15 rats, 3 rabbits, 2 dogs and 3 monkeys continuously to 6.1 + 0.3 mg/m 3 CCl 4 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 monthl y intervals throughout the study, and at the conclusion of the study.

During the 6 week study, one monkey died followi ng the 7th exposure, and 3 guinea pigs died following the 20th, 22 nd , 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 CCl 4 (most prominently the guinea pigs). The guinea pigs were the most sensitive species displaying discolor ed lungs, fatty livers, bile duct proliferation, fibrosis, focal inflammatory cell infiltration, he patic 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 li ver lipid content were also observed, to a lesser extent, in the other four species; the mo st severe alteration occurred in the rats, less severe alterati on in rabbits and dogs, and the least severe in the monkeys.

Appendix D3 83 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

During the 61 mg/m 3 (9.7 ppm) CCl 4 continuous exposure study, 3 gui nea pigs died (one each after 47, 63, and 71 days). All the monkeys we re emaciated and expe rienced 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 de hydrogenase activity was also ev ident 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/m 3 (1 ppm) CCl 4 was attained by diluting the CCl 4 to 10% of

the above concentration with n -octane, resulting in a solution of 6.1 mg/m 3 CCl 4 in 61 mg/m 3 of

n -octane (Prendergast et al., 1967)). The level of n -octane used was shown to be nontoxic by an

n -octane control, which yielde d no effects. (The current TLV for n-octane is 1400 mg/m 3 (300 ppm) (ACGIH, 1992).) No animal s died during this study, and no signs of toxicity were noted. All exposed animals except the rats showed reduced weight gain wh en compared to the controls, and all species exhibited nonspecifi c inflammatory lung changes. Gu inea pig liver lipid contents and serum urea nitrogen concentrations were simila r 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 Spra gue-Dawley rats to 10, 50 and 100 ppm of

CCl 4 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 inha lation. Accumulation of CCl 4 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 obs erved in the 10 ppm group, but no control group was used for comparison.

The intermittent exposure caused a more pronoun ced 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 si gnificant decrease in hip puric acid synthesis, presence of mitochondrial enzymes (glutamate dehydrogenase and ornithine carbonyl transferase) in the blood (indicat ing severe damage to hepatocytes), significant increase in cytoplasmic enzyme activity, and a decrease in th e 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 CCl 4 in the blood was similar for both the intermittent and monotonous exposure groups. A nother 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 CCl 4 stable. Increasing the concentra tion threefold or fivefold with five 10 minute peaks did not poten tiate the toxic effects. Varyi ng the regimen tenfold to five 5- minute peaks (peak exposure 402 mg/m 3 (63.8 ppm)) with a time wei ghted average exposure of 6.5 ppm (41 + 1 mg/m 3 ) did, however, result in more severe liver damage.

Sakata et al. (1987) exposed 10-15 male Spragu e-Dawley rats to <10 ppm CCl 4 vapor for 15 minutes a day, twice a week for 8 weeks. A ll the rats had chronic liver damage involving

Appendix D3 84 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

nodular liver surfaces and extensive fibrosis. Res earchers also found similar results in rats after 8 weeks of subcutaneous inje ctions of 0.1 mL of 50% CCl 4 solution in olive oil twice a week.

Ideura et al. (1993) exposed male Wistar rats to CCl 4 vapor for 7 minutes, 3 times a week for 6- 10 weeks (concentration unspecified). Six experime ntal 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 inje cted with varying amount s of endotoxin (1.0 mL lipopolysaccharide (LPS)). The ra ts were sacrificed 24 hours afte r the injection and processed for histological examination. Examination of th e rats’ left kidneys and livers revealed liver

cirrhosis with destruction of normal structure and massive asc ites retention af ter 10 weeks of exposure as compared to the controls. Those exposed for 6 weeks exhibited an increase in fibrous tissue. The control groups displayed norma l liver structure. Researchers found that rats previously resistant to endotoxi n became susceptible following CCl 4 exposure, which was

manifested as induced acute renal tu bular necrosis in cirrhotic rats.

Yoshimura et al. (1993) performed a similar expe riment to that of Ideura et al. (1992) by exposing male Wistar rats for 6 (5 rats) and 10 weeks (5 rats) to 99% CCl 4 vapor for 3 minutes a day. A control group of 5 rats was given phenoba rbitone for 10 weeks. After 24 hours following the final exposure, rats were inject ed with endotoxin. Six weeks of CCl 4 exposure caused liver fibrosis with bridging fibrosis, while 10 weeks of exposure to CCl 4 caused liver cirrhosis and

destruction of the norma l liver architecture.

Pregnant rats were exposed to 0, 300, or 1000 ppm (0, 1938, or 6460 mg/m 3 ) carbon tetrachloride for 7 hours/day on da ys 6-15 of gestation (Schwetz et al ., 1974). Significant fetal growth retardation, measured by decreased crow n-rump length and body weight, was observed in the offspring of the exposed groups (n = 22 litter s) 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 obs erved in the 1000 ppm group.

Appendix D3 85 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Effects of Chronic CCl 4 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 TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Data from Guinea Pigs and Rats Exposed to 5 ppm CCl 4 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 CCl 4 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 CCl 4 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 CCl 4 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 CCl 4 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 μ g/m 3 ; 0.04 mg/m 3 )

Of the 2 adequate chronic i nhalation studies available on CCl 4 , 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 TetrachlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

duration. In addition, the Adams study contained more specific endpoints of liver damage that

were consistent with the mechan ism of carbon tetrachloride toxici ty. 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 μ g/kg/day to an equivalent inhalation value by route-to-route extrapolati on yields an inhalation REL estimate of 2.5 μ g/m 3

. As another comparison, if the 6.1 mg/m 3 continuous exposure in Prendergast et al . (1967) is a NOAEL (for rats), the resulti ng REL estimate would be 60 μ g/m 3 . If the 6.1 mg/m 3 continuous exposure is a mild LOAEL, the resulting REL estimate would be 20 μ g/m 3

.

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the REL for carbon tetrachlo ride are the chronic exposure study used and the target tissue affected. The major uncertainti es 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 labo ratory animals. Arch. Indust. Hyg. Occup. Med. 6(1):50-66.

ACGIH. 1992. American Conference of Government al and Industrial Hygienists. Documentation of the Threshold Limit Values and Biological E xposure Indices. 6th ed. Vol. I. Cincinnati, OH:

ACGIH.

CARB. 1999. California Air Resources Board. T oxics Air Quality Data. Substance Chooser.

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

CARB. 2000. California Air Resources Board. Calif ornia 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-Oth mer 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 e xposure to chemicals in the workplace: Carbon tetrachloride; PB86-143377; SRC-TR-84-1123.]

Appendix D3 88 Carbon TetrachlorideDetermination of Noncancer Chronic Re ference 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 ti ssues. Toxicol. Appl. Pharmacol. 98(1):87-99.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (CD-ROM version). Denver, CO: Micromed ex, 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, car bon tetrachloride, 1,1,1-trichloroethane, dichlorodifluoromethane, and 1,1-dichloroethyl ene. 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. E nviron. Contam. Toxicol. 38:959-961.

Santodonato J. 1985. Monograph on human exposure to chemicals in th e workplace: Carbon tetrachloride; PB86-143377; SRC-TR-84-1123. NTIS.

Schwetz BA, Leong BK, and Gehring PJ. 1974. Embr yo - and fetotoxicity of inhaled carbon tetrachloride, 1,1-dichloroethane and methyl ethyl ketone in ra ts. Toxicol. Appl. Pharmacol.

28:452-464.

Shimizu Y, Nagase C, and Kawai K. 1973. Accu mulation 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 TetrachlorideDetermination of Noncancer Chronic Re ference 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 (P CDDs) and polychlorinated dibenzofurans (PCDFs) including 2,3,7,8-Tetrachlorodibenz o-p-dioxin (2,3,7,8-TCDD) which is the principal congener of co ncern 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/m 3 (40 pg/m 3 ) 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 ofporphyrins and deltaaminolevulinic acid, increased serumactivities 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 C 12 H 4 C l4 O 2 (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 K ow (octanol/water partition coefficient ) 6.15-7.28 (6.8 for TCDD)

Log K oc (organic-carbon distribution coefficient) 6.0-7.39

Henry’s Law Constant 8.1 x 10 -5 ATM-m 3 / mol

Appendix D3 90 Chlorinated dibenzo-p-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

III. Major Uses and Sources

The chlorinated dioxins and furans are genera ted as by-products from various combustion and chemical processes. PCDDs are produced during incomplete combustion of chlorine containing wastes like municipal solid wa ste, sewage sludge, and hospita l and hazardous wastes. Various metallurgical processes involving heat, and burni ng 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 ge nerate PCDDs as a by-product under certain

conditions. While manufactur e of many of these compounds and formulations has been

discontinued in the United Stat es, continued manufacture elsewher e in the world combined with use and disposal of products containing PCDD by-p roducts results in the inadvertent release of

PCDDs into the environment. Industrial and m unicipal processes in which naturally occurring phenolic compounds are chlorinated can produce PC DDs; the best example is chlorine bleaching of wood pulp in the manufacture of paper product s. 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 consid ered the most potent congener of the polychlorinated dibenzo-p-di oxins (PCDDs) and polychlorin ated 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’ biologi cal activity. Thus only 15 of 210 different PCDDs and PCDFs congeners possess signi ficant 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-tetrachlorodibenz o-p-dioxin and 2,3,7,8- tetrachlorodibenzofuran; three pentachloro congeners: 1,2,3,7,8-pentac hlorodibenzo-p-dioxin, 1,2,3,7,8-pentachlorodibenzofuran, and 2,3,4,7,8-pentach lorodibenzofuran; 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-hexachlo rodibenzofuran; 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 numbe ring 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-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Ah receptor as well as on relativ e toxicity of different congeners . Values for the international system of toxic equivalents are pr ovided 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.)

Compound 1,2 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-dioxinsDetermination of Noncancer Chronic Re ference 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 i b e n z o f u r a n

IV. Effects of Human Exposure

The information available on possible chronic to xic effects in humans is complicated by the relative insensitivity of epidemio logical studies, the limited ability of case studies of exposed

individuals to establish cause and effect rela tionships, the heterogeneous nature of human populations, the broad spectrum of exposures to ot her 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 effect s 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 chroni c exposure to various chlorinated aromatic compounds (Moses and Prio leau, 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 res ponse (U.S. EPA, 1994a). Epidemiological data av ailable for 2,3,7,8-TCDD have not allowed a determination of the threshold dose required fo r production of chloracn e (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 epidem iological 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 Di sease 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-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

found little evidence for an associ ation between exposure and hepatomegaly suggesting that this is not a chronic response. Ther e is a consistent pattern of in creased 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 Viet nam Experience Study, 1988). Two cross

sectional studies have associated diabetes a nd elevated fasting seru m 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 likeli hood of abnormally low testosterone levels was 2 to 4 times greater in indi viduals with serum 2,3,7,8-TCDD levels above 20 pg/ml (Egeland et al. 1994) and increased serum levels of luteiniz ing hormone and follicle stimulating hormone have been documented (Egeland et al., 1994). A number of other eff ects have been reported that

were either not seen as chr onic effects or effects seen l ong term in only one population of exposed persons. These include el evated liver enzymes (aspartate aminotransferase and alanine aminotransferase), pulmonary disorders, neur ologic disorders, and changes in porphyrin metabolism and kidney disorders (U.S. EPA, 19 94c). Areas in which there is presently

insufficient information to draw solid conclusi ons include effects on th e circulatory system, reproductive effects, immunological effects, effects on metabolism and handling of lipids, and on thyroid function (U.S. EPA, 1994c). Recent fi ndings 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 fe tal development include fetal death, growth retardation, structural malforma tions 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 contaminat ed 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 toxic ity (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 th e infants. Signs of t oxicity from dioxin like compounds were absent in infants born to mother s apparently not affected in the Seveso, Italy and Times Beach, Missouri, incidents (Reggian i, 1989; Hoffman and Stehr-Green, 1989), which supports Rogan’s conclusion. Ther e 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, wa s observed in Yusho infants (Yamashita and Hayashi, 1985) making this malformation a possible result of e xposure to dioxin-like compounds. Nevertheless, it is unknown if thes e compounds produce malformations in humans. Evidence for possible organ system dysfuncti on in humans comes from a study of Yu-Cheng children which found that children exposed in utero experienced de lays in attaining developmental milestones, and exhibited neurobehavioral abnormalities (Rogan et al. , 1988)

Appendix D3 94 Chlorinated dibenzo-p-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

suggesting involvement of CNS function. Dysfunc tion of dermal tissues is noted in exposed infants of the Yusho and Yu-Cheng incidents an d is characterized by hype rpigmentation of the skin, fingernails, and toenails, hypersecretion of the meibomia n 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, he patotoxicity, effects on tissue of epithelial origin, various endocrine effects, effects on vita min A storage and use, immune system effects and reproductive and developmental toxicity. Th e 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 oc curring toxic effects. The wasting syndrome is characterized by loss of adipose tis sue 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 speci es sensitivity (U.S. EPA, 1994d). TCDD induces hyperplasia and hypertrophy of liver parenchymal cells. Morpho logical 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 infiltrati on, multinucleated giant hepatocytes, increased numbers of mitotic figures, and parenchymal cell necrosis (U.S. EPA, 1994d; WHO/IPCS, 1989). Epithelial effects seen include chloracn e (rabbit ear and the hairless mouse) (Jones and Krizek, 1962; Schwetz et al. , 1973) and hyperplasia and/or me taplasia of gastric mucosa, intestinal mucosa, the urinary tract, the bile duct and the gall bladde r (U.S. EPA 1994f). TCDD exposure results in endocrine like effects including epidermal growth factor like effects such as early eye opening and incisor erupt ion 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 horm one levels and in some cases thyroid hormone like effects (WHO/IPCS, 1989; Rozman et al. , 1984), decreases in se rum testosterone and dihydrotestosterone (Mittler et al. , 1984; Keys et al. , 1985; Moore and Peterson, 1985), and changes in arachidonic acid metabolism and pr ostaglandin synthesis (Q uilley 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 suppre ssors of both cellular and humoral immune system function, characteri stically 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 availa ble 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-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

acutely toxic to early life stages of fish and birds with fish being most sensitive (LD 50 of 0.4 μg/kg for rainbow trout sac fry eggs and LD 50 of 34 ng/kg for lake trout eggs); some species of birds are also relatively sensitive (LD 50 of 0.25 μg/kg for chicken eggs) (Peterson et al. , 1993).
2,3,7,8-TCDD has been documented to increase the in cidence of prenatal mortality in a number of species of laboratory animals including the Rh esus 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 he morrhage in the rat (Sparschu et al. , 1971), hydronephrosis in th e 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 clefti ng 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 animal s exposed in utero to 2,3,7,8-TCDD. TCDD reduces plasma androgen leve ls in the adult male rat and perinatal exposure decreases spermatogenesis, sperma togenic function and reproductive capability,

feminizes male sexual behavi or, and feminizes male gonadotr ophic 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 de tailing the effects of the PCDDs, PCDFs and other dioxin like compounds, however a large number of these stud ies were conducted as either acute or subchronic exposures, studies in which it is unlik ely that body burdens had reached steady state

levels. Detailed below are three ch ronic studies that were consider ed in the setting of a chronic toxicity exposure level.

The most definitive study of chronic t oxicity in rats is that of Kociba et al. (1978). This study involved the administration of 2,3,7,8-TCDD in th e diet at doses of 1 ng/kg/day, 10 ng/kg/day, and 100 ng/kg/day to groups of 50 male and 50 fema le Sprague Dawley rats for two years. A group of 86 male and 86 female rats received di et with solvent vehicle alone and served as controls. The following observations (excludi ng 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 aci d, and increased serum activities of alkaline phosphatase, gamma-glutamyl transferase, and glutamic-pyruvic transaminase. Histopathologic ch anges were noted in the liver, lymphoid tissue, respiratory and vascular tissues. The primary ultrastructural ch ange in the liver was pr oliferation 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 a nd included increased urinary excr etion of porphyrins in females as well as liver and lung lesions. The 1 ng/kg/da y dose produced no discernible significant toxic effects. Interpretation of this study by the au thors was that the 1 ng/kg/day dose was a NOAEL.

Appendix D3 96 Chlorinated dibenzo-p-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Two chronic toxicity studies are available in th e mouse. The first is a one year study conducted by Toth et al. (1979) using male Swiss mice administer ed 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 recei ving 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 lo west 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 exposu re 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 ac id, increased serumactivities 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 μ g/m 3 per mg/kg/day Inhalation reference exposure level 40 pg/m 3 (0.00004 μg/m 3 )

The data available for chronic toxic effects in humans have a number of limitations. Some studies did not determine the body burden of co mpounds 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-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

studies and case studies have limitations in de termining 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 examinati on of all organ systems, examination of urinary excretion of porphyrins and delta-a minolevulinic acid, and determin ation of serum activities of alkaline phosphatase, gamma-glutamyl transferas e, 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 a nd reflects changes in liver biochemistry.

Thus the examination of markers of liver toxici ty also altered in anim al models of chronic toxicity make the Kociba study an appropriate ch oice 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 compar e 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 calcul ated 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 b ackground exposures. The REL of 10 pg/kg/day should be protectiv e of chronic effects on liver function and avoid significant increases in exposure over the b ackground level of human exposure.

The strengths of the inhalation REL include the availability of chronic exposure data from a well-conducted study with histopathological an alysis, the observation of a NOAEL, and the demonstration of a dose-response relationship. Major areas of uncertain ty are the lack of adequate human exposure data and the lack of chronic inhalation exposure studies.

Appendix D3 98 Chlorinated dibenzo-p-dioxinsDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

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U.S. EPA. 1994e. ibid Vol 1:3-14.

U.S. EPA. 1994f. ibid Vol 1:3-6.

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Appendix D3 105 Chlorinated dibenzo-p-dioxinsDetermination of Noncancer Chronic Re ference 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/m 3 (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 Cl 2 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/m 3 @ 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 wa ter. Chlorine gas is sometimes used at large public pools while hous ehold pools typically use hypochlorite solutions. Chlorine is an integral part of the bleaching process of wood pu lp 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 statewid e industrial emissions from facilities reporting under th e 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 work ers’ length of exposure in y ears. Group A consisted of 220 workers who were employed/ exposed for 10-25 years. Group B consisted of 133 workers

Appendix D3 106 ChlorineDetermination of Noncancer Chronic Re ference 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/m 3 (0.37-1.75 ppm) chlorine. The control gr oup’s average age was 39.7 years (ranging from 26-55 years), and it consisted of 192 worker s 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 complain ts than the control workers. Current smokers in groups A and B e xperienced the highest incidence of pulmonary symptoms and group A workers had a higher prev alence 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 fu nction in tests of V50/H and FEF 25-75 (forced expiratory flow
between 25 and 75% of forced vita l capacity (FVC), the total amount of air the subject can expel during a forced expiration) - compared with the control group, and gr oup A showed reduced

FEV 1 (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 ya rd workers in similar working conditions but not exposed to chlorine (79% a nd 84% respective partic ipation 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 av erages (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 mo re than one episode of chlorine “gassing” as compared to the rail yard workers and pulpm ill workers with no chlorine gas exposure. More airflow obstruction was observed in exposed wo rkers in spite of th eir nonsmoking and exsmoking status, correlating to si gnificantly lower average values for MMF (maximal midexpiratory flow) and for the FEV 1 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 th at after the first high exposur e incident, an inflammatory response occurred in small airway s and that this reaction did not resolve in those workers who were continuously or repeatedly exposed to the irritant. It wa s also suggested that chronic airflow obstruction caused by repeated minor exposur es led to chronic respiratory disability in

some of the workers.

Patil et al. (1970) evaluated the exposure of 332 ma le 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 fr om 25 representative chlorine manufacturing plants was also studied. Both gr oups were comprised of men between the ages of 19-69 with a mean age of 31.2 + 11.0 years. Physical examinati ons (blood and urine analysis, chest x-rays and electrocardiogram s) 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 leve l had no correlation to the numberof years exposure. Ninetyeight of the 332 workers were found to have abnormal teeth and gums, but no dose-response

Appendix D3 107 ChlorineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

relationship was concluded. Similarly, no dose- response relationships were shown with the symptoms of sputum production, cough, dyspnea, hi story of frequent colds, palpitation, chest pain, vital capacity, maximum breathing capac ity 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 a nd there was no noted prol onged 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 or iginal 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 shortnes s of breath after their mo st significant exposure,

but not at the time of the initial study by Courteau et al. Moderate risk work ers 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 experien cing shortness of breath that continued one month after the exposure and/or abnormal lung sounds. Nine ty percent of the follow up group completed questionnaires which revealed a 91% inciden ce of respiratory symptoms. Spirometry

assessments and methacholine inhalation tests we re 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 FEV 1 was seen in those with a lower

PC 20 , and 52% of these workers showed an FEV 1 < 80% predicted.

Enarson et al. (1984) compared 392 pulpmill workers exposed to chlorine (unspecified duration) to a comparable group of 310 rail yard workers li ving in the same commun ity, but not exposed to chlorine. In the pulpmill areas surveyed that pr edominantly 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 Cl 2 respectively. Of the mach ine 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 th e 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 seve re enough to require oxygen therapy. Ventilation was affected by chlorine inhalati on, with a decrease in the maxima l midexpiratory flow (MMF). Smokers in this group had significantly reduced FVC, FEV 1 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 calci fications. Researchers suggest that the first

ventilation function affected in ob structive airway disease is MMF.

Appendix D3 108 ChlorineDetermination of Noncancer Chronic Re ference 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 pp m, 1.0 ppm and 2.5 ppm. The e xposures were carried out for 104 weeks at 6 hr/day 3 days/week for female rats an d 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 cont ained 320 male and 320 female mice. The rats were studied in groups of 70, yielding 280 per ge nder 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 on ce 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 Cl 2 showed decreased weight gain compar ed to controls while only female mice exposed to 2.5 ppm Cl 2 showed decreased weight gain. Male rats showed decreased weight gain at all levels of exposure

while female rats showed the sa me result at only 1.0 and 2.5 ppm Cl 2 exposures. Various nonneoplastic nasal lesions were s een in all the airway epithelia l 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 Cl 2 . Statistically significant re sponses were seen in the traditional and respiratory epithelial regions of all exposed rats and mice. Statistically significant damage to olfactory epithelium o ccurred 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 Cl 2 . These monkeys were exposed to chlorine for 6 hours/day, 5 days/week. The monk eys were evaluated periodically on the basis of body weight, electrocardiograms, neurologic ex aminations, 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. Oc ular irritation (tearing, rubbing of the eyes, reddened eyes) was observed after 6 weeks ofexposure in the 2.3 ppm group. No exposurerelated differences were seen in neurologic exam inations, electrocardiograms, clinical chemistry, urinalysis, hematology or blood gas levels. Also , no exposure-related changes were observed in the parameters of ventilation distribution. Pulmona ry function evaluations yielded a statistically significant trend for increasing pulmonary diffu sing capacity and distri bution of ventilation values for males and females in the 2.3 ppm e xposure group. Both males and females of the 2.3 ppm group exhibited statistically significant increased incide nce of respiratory epithelial hyperplasia. A mild form of the lesions was also seen in th e 0.5 ppm group, 0.1 ppm group (females only) and one male in the control grou p. Two parasitic infections occurred, affecting the respiratory tract and resul ting in 11 monkeys housing parasite s 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 ChlorineDetermination of Noncancer Chronic Re ference 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-bo dy 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

BMC 05 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 cm 2 ) 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/m 3 )

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 ra t data. Using the data for glandular epithelial eosinophilic proteinaceous accumulation (see Table 1 below) to derive the BMC 05 resulted in a

3-fold lower value than th e LOAEL of 0.4 ppm, or BMC 05 = 0.14 ppm. (Adequate benchmark

dose estimates could not be obtaine d for the other nasal lesions due to high background rates and shallow dose-response relationships.) A BMC 05 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 durat ion of the experiment was for a full lifetime of two years. Second, the sample si zes were large (280 per sex per species). Finally, appropriate sensitive endpoints of respiratory epithelial damage were examin ed. The mice and male rats were exposed to chlorine for 6 hours/day, 5 days/w eek, 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 ChlorineDetermination of Noncancer Chronic Re ference 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 onl y one year of their total 35 year lifetime, and the sample sizes were considerably smalle r (4 monkeys per sex per group) than the mouse and rat groups (280 per sex per species). Alt hough the exposure durations differed between the two studies, the histological re sults were similar, differing onl y 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 μ g/m 3 ) using the BMC approach, we estimated a REL of 0.02 ppb (0.06 μ g/m 3 ) 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 UF L of 3 for a mild effect, a

UF S 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 μ g/m 3 ).

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-ra y 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 resp iratory system irritant. As an illustration of what would be estimated, the study of Shi ( 1990) had a mean workplace exposure of 4.82 mg/m 3
(1.7 ppm). This LOAEL was time adjusted to an equivalent continuous exposure of 1.72 mg/m 3

,

Appendix D3 111 ChlorineDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 dos e-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. Majo r areas of uncertainty are the l ack 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 wo rkers 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 respons iveness 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 pu lpmills. 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 work ers. 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 ch lorine over a three to six month period in a pulpmill: I. Exposure and symptomatol ogy. 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 ChlorineDetermination of Noncancer Chronic Re ference 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 s yndrome a variant of oc cupational asthma? J. Allergy Clin. Immunol. 93:12-22.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version). Denver, CO: Microm edex, 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 Ch an-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, Morg an K, and Barrow C.1987. One-year inhalation toxicology study of chlorine in rhesus monke ys (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 sens itivity 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. Occupa tional 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 Physic s, 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 B6C3 F1 mice and F344 rats to chlorine gas induces lesions confined to the nose. Fundam. Appl. Toxicol. 24:111-131.

Appendix D3 113 ChlorineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# CHLORINE DIOXIDE

(anthium dioxide; alcide; chlori ne oxide; chlorine peroxide; chloryl radical; doxcide 50)

CAS Registry Number: 10049-04-4

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.6 μ g/m 3 (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 ClO 2 Molecular weight 67.45 g/mol Density 1.642 g/cm 3 @ 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/m 3

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 purifica tion 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 mo st recent inventory were estimated to be 1136 pounds of chlorine dioxide (CARB, 2000).

IV. Effects of Human Exposures

Case reports of human occupati onal exposure to chlorine dioxid e have shown that 19 ppm was fatal to one worker and 5 ppm was definitely ir ritating (Elkins, 1959). Seven out of 12 workers exposed regularly to chlorine dioxide at levels generally below 0.1 ppm (0.28 mg/m 3 ) reported symptoms of ocular and respiratory i rritation leading to slight bronchitis ( Gloemme and

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Lundgren, 1957 ). However, the authors ascribed the bronchitis to occasional acute excursions of chlorine dioxide levels above 0.1 ppm due to te chnical problems such as equipment leakage. Concurrent exposure to chlorine and chlorine diox ide in pulp mill workers resulted in an increase in the reporting of subjective sy mptoms of irritation (Ferris et al. , 1967). In this study, the chlorine dioxide concentrations ranged from trace levels to 0.25 ppm (0.69 mg/m 3 ). No

differences were found between these workers and controls by pulmona ry 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/m 3 ) chlorine dioxide (Paulet and Debrou sses, 1972). The number of control animals was not specified. Micros copic evaluation of the lungs re vealed vascular congestion and peribronchiolar edema in all an imals exposed to chlorine dioxide. The subchronic LOAEL for respiratory effects was therefore 1 ppm (2.8 mg/m 3 ).

An earlier study by these researchers (Paulet a nd Debrousses, 1970) examined the effects of exposure to 2.5, 5, or 10 ppm chlori ne dioxide for several hours/day for 30 days in rats and rabbits (n = 4-10 animals per group). Body weights, blood ce ll counts, and histopathological examination of the liver, lungs, and other tissues were measured in each group. At 10 ppm, nasal discharge, localized bronchopneumonia, and des quamated alveolar epithelium were observed.

White and red blood cell counts we re 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, incl uding 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/m 3 ) chlorine dioxide for 15 minutes, 2 or 4 times/day, for 1 month showed an increas e in congested lungs, nasal discharge, and catarrhous lesions of the alveoli beginning at 10 ppm (Paulet and De brousses, 1974). No significant changes in these para meters 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. Fi ve out of 5 rats died during expo sures of 4 hours/day for 14 days. All exposed animals exhibited signs of respiratory distress and ocular discha rge. No effects were seen in 5 rats exposed to 0.1 ppm for 5 hours/da y, 7 days/week, for 10 weeks. Thus 0.1 ppm was

a subchronic NOAEL.

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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; pe ribronchial edema in all animals; lung alveolar damage

LOAEL 1 ppm (2.8 mg/m 3 )

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 m 3 , SA(Th) = 3,460 cm 2 ) 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/m 3

, 0.6 μ g/m 3 )

The U.S. EPA (1995) based its RfC of 0.2 μ g/m 3 on the same study but included a Modifying Factor (MF) of 3 for database deficiencies. The criteria for use of modify ing factors are not well specified by U.S. EPA. Such modifying factor s were not used by OEHH A. 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/m 3 (0.08 ppb) based on hyperplasia in respiratory epitheliu m in female rats. Based on chemical reactivity, the REL for chlorine dioxide might be expected to be lower th an that for chlorine. However, there are much less toxicologic data available for ch lorine 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 expos ure 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 a ppears likely that the proposed REL would be protective of the effects reporte d 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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VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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 e xperimental 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 Wor cester 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. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce Exposure Levels. Available on-line at http://www.oehha.ca.gov

Paulet G, and Debrousses S. 1970. [On the action of ClO 2 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. S ecur. 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: Micr omedex, Inc. (Edition expires 10/31/94).

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CHRONIC TOXICITY SUMMARY



# CHLOROBENZENE

(monochlorobenzene; benzene chloride; benzene monochloride; chlorben zene; chlorbenzol; phenyl chloride)

CAS Registry Number: 108-90-7

I. Chronic Toxicity Summary

Inhalation reference exposure level 1000 μ g/m 3 (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 C 6 H 5 Cl Molecular weight 112.56 g/mol Boiling point 132 o C Melting point

− 45.2 o C Vapor pressure 11.8 torr at 25 o 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/m 3 at 25 o C

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 historic ally important in the ma nufacture of chlorinated pesticides, especially DDT, and in the producti on of phenol and aniline. Monochlorobenzene’s principal current use is as a chemical intermed iate in the production of chemicals such as nitrochlorobenzenes and diphenyl oxide. Thes e chemicals are subsequently used in the production of herbicides, dyestuffs, and rubber ch emicals. Additionally, monochlorobenzene is used as a solvent in degreasing processes (e.g., in metal cleaning operation s), paints, adhesives, waxes and polishes (HSDB, 1995; NIOSH, 1993). The annual statewid e industrial emissions from facilities reporting under th e 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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IV. Effects of Human Exposure

Even though monochlorobenzene has been used in dustrially 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) describe s multiple central nervous system effects, including headache, numbness, dizziness, cyanos is, hyperesthesia, and muscle spasms, after intermittent exposure over 2 years to monochlorob enzene in a mixed chemical environment. No specific exposure levels or histop athologic 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/m 3 ) 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 activ ity of the cerebral cort ex in four individu als 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 monoc hlorobenzene. Only a single, oral chronic carcinogenicity st udy (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 ki dney, and, to a lesser ex tent, blood parameters following monochlorobenzene exposure over a peri od 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 (especi ally the liver) after 11 and 24 weeks of exposure (LOAEL 75 ppm). Male rabbits also demonstrated increa ses in liver weight after 24 weeks of exposure (LOAEL = 75 ppm). Some hematological changes were reported in rats including differences in platelet and reticulocyte count s between control and exposed an imals; however, some changes observed at 11 weeks were variable and compar able 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 ad renal cortex, tubular lesions in the kidneys, and congestion in the liver and kidneys.

Two other subchronic inhalation studies re ported 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 liv er weights in rats and ra bbits 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 fo r 6 hours over 10-11 weeks prior to mating, and

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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 (F 0 and F 1 ). Hepatotoxicity occurred as hepatocellular hypertrophy and increased liver we ights (mean and absolute ) at concentrations greater than 50 ppm (LOAEL = 150 ppm). At this concentration (150 ppm), renal changes included tubular dilation, interstitial nephritis, an d foci of regenerative epithelium. Testicular degeneration of the germinal ep ithelium occurred in both generations of exposed males, but no

chlorobenzene-induced adverse effects on reproductive performa nce 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 (F 0 and F 1 both sexes), hepatocellular hypertrophy (F 0 and F 1 males), renal degeneration and inflammation (F 0 and F 1 both sexes), testicular degeneration (F 0 and F 1 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/m 3 , 1000 μg/m 3 )

Of the three inhalation studies available (Dilley, 1977; John et al. , 1984; Nair et al. , 1987), the

Nair et al. (1987) two generational developmental st udy was selected for identifying a NOAEL and LOAEL. It best presented the histopatholog y of the adverse effects, and demonstrated a dose response relationship for these effects (sta tistically significant increases in mean liver weights, incidence of renal change s, and testicular degeneration).

Another subchronic inhalation st udy (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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include increases in liver we ight and hepatocellular degene ration in rats (LOAEL = 125 mg/kg/day) and mice (LOAEL = 25 0 mg/kg/day), and renal necros is 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 la ck 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 de monstrated species variation in the urinary

elimination of chlorobenzene metabo lites (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 hu man 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 LOAE L 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 in traspecies) 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 th e demonstration of a dose-response relationship. Major areas of uncertainty are the lack of ade quate human exposure data and lim ited reproductive toxicity data.

VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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 Biom edical and Behavioral Sciences. Cincinnati,

OH: NIOSH. NTIS No. PB-276623.

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Gargas ML, Burgess RJ, Voisard DE, Cason GH, and Andersen ME. 1989. Partition coefficients of low-molecular-weight volatile chemicals in various liquids and ti ssues. 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, Knezev ich A, and Stack CR. 1987. A two-generation reproduction study with monochlorobenzene vapor in rats. Fundam. A ppl. Toxicol. 9:678-686.

NTP. 1985. National Toxicology Program. Toxi cology and carcinogenesis studies of chlorobenzenes (CAS No. 108-90-7) in F344/N ra ts and B6C3F1 mice (gavage studies). U.S. Dept. of Health and Human Services, Public Hea lth Service, National Institutes of Health, NTP TR 261. NIH Pub. No. 86-2517, NTIS PB86-144714.

NIOSH. 1993. National Institute for Occupationa l Safety and Health. NIOH and NIOSH basis for an occupational health standard: Chlorobenz ene. 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 O ccup. Environ. Health 53:51-57.

Ogata M, Taguchi T, Hirota N, Shimada Y, and Nakae S. 1991. Quantitation of urinary chlorobenzene metabolites by HPLC: Concentrati ons of 4-chlorocatechol and chlorophenols in urine and of chlorobenzene in biological specimen s of subjects exposed to chlorobenzene. Int. Arch. Occup. Environ. Health 63:121-128.

Rozenbaum ND, Blekh RS, and Kr emneva SN. 1947. [Use of chlor obenzene as a solvent from the standpoint of industrial hygie ne.] Gig. I Sanit. 12:21-24. (Russi an) [as cited in U.S. EPA, 1988.]

Tarkhova LP. 1965. Materials for determining th e 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. Washi ngton DC: U.S. EPA. National Technical Information Service, PB89-192116.

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Yoshida M, Sunaga M, and Hara I. 1986. Urinar y metabolites levels in workers exposed to

chlorobenzene. Ind. Health 24:255-258.

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CHRONIC TOXICITY SUMMARY



# CHLOROFORM

(trichloromethane; formyl tric hloride; methenyl trichloride; methyl trichloride)

CAS Registry Number: 67-66-3

I. Chronic Toxicity Summary

Inhalation reference exposure level 300 μg/m 3 (50 ppb) Critical effect(s) Liver toxicity (degenerative, foamy vacuolization, and necros is) 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 CHCl 3 Molecular weight 119.49 g/mol Boiling point 61.1 o C Melting point

− 63.6 o C 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/m 3 per ppb at 25oC

III. Major Uses and Sources

Chloroform (CHCl 3 ) is used in industry and laboratory settings as a solvent for adhesives, pesticides, fats, oils and rubbers. It is also used as a chemical in termediate in the synthesis of fluorocarbon 22, dyes, pesticides, and tribromome thane. Chloroform is produced as a byproduct of water, sewage, and wood pulp chlorination (H SDB, 1995). In 1996, the latest year tabulated,

the statewide mean outdoor monitored concentr ation of chloroform was approximately 0.037 ppb (CARB 1999a). The annual statewide industrial em issions 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 ChloroformDetermination of Noncancer Chronic Reference Exposure Levels April 2000

IV. Effects of Human Exposure

Limited information is available regarding possi ble adverse health effects in humans following chronic inhalation of chloroform. However, historical clinical re ports 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 chlo roform toxicity included impaired liver function (Smith et al. , 1973), toxic hepatitis (Lunt, 1953; Sc hroeder, 1965), cardiac arrhythmia (Payne, 1981; Schroeder, 1965; Whitaker and Jo nes, 1965), and nausea (Schroeder, 1965; Smith

et al. , 1973; Whitaker and Jones, 1965), and caused central ne rvous system symptoms (Schroeder, 1965; Whitaker and Jones, 1965). Chr onic inhalation studies are limited to a few occupational studies identifying the liver and th e 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 manufactur ing 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, depr ession, 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 he patitis based on elevated serum enzymes. Additionally, 10 workers had splenomegaly. Wo rkers exposed to chloroform had a 10-fold increased risk of contracting viral hepatitis compared to the gene ral population. The study

authors considered the chloroform induced liver toxicity as a predis posing factor for viral hepatitis, but the incide nce of viral hepatitis in the workers is in itself a confounding factor.

Phoon et al. (1983) described two outbr eaks 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) contai ned 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 antig en in the patients, the author s attributed the jaundice to chloroform exposure rather than viral hepatitis.

More recently, Li et al.(1993) reported on 61 chlo roform exposed workers from a variety of production factories. Exposure levels at 3 repres entative 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 expos ure 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 ha d 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 al so affected, manifested as incr eases in scores of passive mood states and dose-related, negative cha nges in neurobehavioral testing.

Appendix D3 125 ChloroformDetermination 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 expo sure to other chemicals, inadequate control groups and potential confounders. However, thes e studies indicate the po tential for liver and central nervous system toxicity in humans exposed to chloroform via inhalation.

V. Effects of Animal Exposure

Exposure of experimental animals to chlorofo rm for acute, subchronic or chronic durations results in toxicity to the liver and kidney, as well as to the respirat ory and central nervous systems (USDHHS, 1993). The majority of chroni c 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, ho wever, 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 exposu re 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 chlorofo rm treatment induced dose- and time-dependent lesions only in the livers and nasa l 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 sens itive. 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 gr oups. 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/w eek 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 concen trations 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 kidne y, liver, and nasal passages as the only target organs. The primary target in the kidney was th e 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 increase s 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 surr ounded by dense connective tissue. Enhanced bone growth and hypercellula rity in the lamina propria of the ethmoid

turbinates of the nose occurred at the early time points at concen trations of 10 ppm and above.

Appendix D3 126 ChloroformDetermination 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 (1 2/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 pp m 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 vacuo lization, focal necrosis, and fibrosis in both sexes of all species tested. Guin ea pigs were the least sensitive and male rats the most sensitive to chloroform induced hepatotoxicity; the abov e adverse effects occurre d at 25 ppm. Adverse kidney effects observed in all species included cloudy swelling of the re nal tubular epithelium and interstitial and tubular nephritis. Pneumon itis 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 liv er 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 ChloroformDetermination 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 stra in (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/m 3 ; 300 μg/m 3 )

In the study of Torkelson and associates (1976) ra ts were the most sensitive species and guinea pigs the least sensitiv e to chloroform vapors. Though of subchronic duration, this inhalation study still exposed rats discontinuously for 25% of a lifetime (25.8 week s/104 weeks/lifetime). Pathological changes were obser ved in both sexes of rat at 50 and 85 ppm (244 or 415 mg/m 3 ) and in male rats at 25 ppm (122 mg/m 3 ) chloroform. These hepatic changes included mild to severe centrilobular granular de generation, foamy vacuolization, fo cal necrosis, and fibrosis. Adverse effects in the kidney including cloudy swe lling and nephritis were seen in all species tested at 25 ppm (122 mg/m 3 ) chloroform.

An unexpected finding in animals was the generali zed atrophy of the ethm oid turbinates of F344 rats after a 90 day exposure at concentrati ons 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 chlorofo rm-induced olfactory mucosal degeneration in

rats is not associated with dete ctable 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 stra ins and other species ma y require reassessing the

basis of the REL for chloroform.

The human occupational studies have reported ja undice with or without alterations in liver enzymes at similar ambient concentrations : 2 to 204 ppm chloroform (10 to 995 mg/m 3 ) after at least 1 year (Bomski et al. , 1967) and 14 to 400 ppm chloroform (68 to 1952 mg/m 3 ) after 6 months or less (Phoon et al. , 1983). The presence of jaundice and hepatitis in these 2 reports

Appendix D3 128 ChloroformDetermination 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 ai r concentrations ranged from 4.27 to 141.25 mg/m 3 with a geometric average of 20.46 mg/m 3 . The exposed workers were subdivided into higher (n=46) and lowe r (n=14) exposures, but the separation was not

indicated for all results. If the lower exposure level of 2.8 ppm (13.49 mg/m 3 ) is classified as a mild LOAEL based on a significant di fference 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 intraspeci es UF of 10 to yield a REL of 30 ppb, in good agreement with the proposed REL of 50 ppb (300 μg/m 3 ) 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 th e 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 barbit urates, may be at potentially greater risk of chloroform toxicity (Cornish et al., 1973). Others with possible higher sensit ivity 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 dur ation 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 a nd 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. T oxics Air Quality Data. Substance Chooser. Chloroform. Available online at http:/www.arb.ca.gov/aqd/toxics.htm

CARB. 1999b. Air toxics emissions data collect ed in the Air Toxics Hot Spots Program CEIDARS Database as of January 29, 1999.

Challen P, Hickish D, and Bedford J. 1958. Chr onic 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 ChloroformDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Chu I, Villeneuve DC, Secours VE, Beck ing GC, and Valli VE. 1982. Toxicity of

trihalomethanes: I. The acute and subacute t oxicity of chloroform, bromodichloromethane,

chlorodibromomethane and bromoform in rats. J. Environ. Sci. Health. B17:205-224.

Cornish HH, Ling B, and Barth M. 1973. Phenobarb ital 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 et hmoid hyperplasia are no t 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 ti ssues. Toxicol. Appl. Pharmacol. 98(1):87-99

Heywood R, Sortwell RJ, Noel P RB, Street AE, Prentice DE, Roe FJC, Wadsworth PF, and Worden AN. 1979. Safety evaluation of toothpast e 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 chloro form inhalation study in female and male B6C3F1 mice: implications for cancer risk a ssessment. Fundam. Appl . Toxicol. 30(1):118-137.

Larson JL, Wolf DC, Mery S, Morgan KT, Butte rworth 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 chlo roform. 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 pe rorally. 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 drin king water contaminants: trichloromethane,

Appendix D3 130 ChloroformDetermination 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 an aesthesis. 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 Ab be 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 chloro form poisoning. Br. J. Anaesth. 37:972-975.

Smith AA, Volpitto PP, Gramling ZW, DeVore MB, and Glassman AB. 1973. Chloroform, halothane, and regional an esthesia: A comparative study. Anesth. Analg. 52:1-11.

Templin MV, Larson JL, Butterworth BE, Jamis on KC, Leininger JR, Mery S, Morgan KT, Wong BA, Wolf DC. 1996. A 90-day chloroform i nhalation study in F-3 44 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 toxi city of chloroform as determined by single and repeated exposure of laboratory anim als. 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 chlo roform anesthetics ad ministered with a precision vaporizer. Anesth. Analg. 44:60-65.

Appendix D3 131 ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# CHLOROPICRIN

(trichloronitromethane; nitrochl oroform; nitrochloromethane)

CAS Registry Number: 76-06-2

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.4 μg/m 3 (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 CCl 3 NO 2 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 disulf ide, ether, carbon tetrachloride, acetone, methanol, acetic acid Conversion factor 6.72 μg/m 3 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). Chlo ropicrin 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 warf are 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 wate r as a result of chlo rination processes (Duguet et al.

, 1985; Merlet et al. , 1985). The annual statewide industr ial emissions from facilities reporting under the Air Toxics Hot Spots Act in Califor nia based on the most recent inventory were estimated to be 1507 pounds of chloropicrin (C ARB, 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 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

fumigants, because agricultural field applications are not covered under th e 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 toxi c 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 chr onic 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 incr eased in both sexes, primarily at 1.0 ppm. Increased mortality was noted in male s 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/m 3 ) for chronic non-cancer effects in rats.

Results from chronic inhalation of chloropicr in 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 sa me 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 week s. 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 lu ng 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 l ung nodules and increased numbers of kidney cysts. In females lung ma sses and kidney cysts were seen at 0.5 ppm. Microscopic pathology changes included increas ed nasal cavity lesi ons (including serous exudate, hyaline epithelial inclusions, rhinitis, ol factory and epithelial at rophy) and lung lesions (including alveolar protein de posits, alveolar histiocytosis, hemorrhage, peribronchiolar lymphocytic infiltrate, bronchiectasis, bronchial submucosal fibrosis, peribronchiolar smooth muscle hyperplasia), in a ddition to kidney cysts at ≥ 0.5 ppm (CDPR, 2000).

Appendix D3 133 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Results from chronic inhalation of ch loropicrin 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 in creased in the 2.93 ppm group. The treatmentrelated histological lesions re ported were degeneration and n ecrosis of the bronchial and bronchiolar epithelia at 2.93 ppm and hypertrophy of these ep ithelia at 1.58 ppm. Thus the primary target organ was the respiratory tr act and the subchronic NOAEL was 0.67 ppm (4.5 mg/m 3 ). (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 sensor y 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 (RD 50 ) (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 examin ed for respiratory tract lesions. Body weights of chloropicrin exposed animals were reduced 10-25% below contro ls, but increased to no rmal levels during the recovery period. Nasal exudate and distenti on of the abdomen were observed. “Moderate” lesions, characterized by exfolia tion, erosion, ulceration, or necr osis, were observed in the respiratory and olfactory epithelium, and minima l inflammation and squamous metaplasia were observed in the respiratory epit helium alone. Moderate to severe damage to the lower respiratory tract was described as “fibrosing peribronchitis and peribronchiolitis”. Exfoliation, hyperplasia, and squamous meta plasia 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 studie s were conducted by dosing animals daily with chloropicrin in vehicle (corn oil) at a volum e 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 incl uded mortality, body we ight, food and water consumption, hematology, serum clinical chemis try, and gross pathology and histology of organs. Only the high-dose group and the cont rol group animals from the 90-day study were examined histopathologically. In the 90-day st udy, 6 males and 2 females in the 32 mg/kg dose group and 1 male and 3 females in the 8 mg/kg do se group died before the scheduled termination time. The authors noted signs of pulmonary co mplications (inflammation and congestion) in the dead animals. Previously, the animals had shown signs of respirat ory distress, including

Appendix D3 134 ChloropicrinDetermination of Noncancer Chronic Re ference 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 surv ivors, 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 author s 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

BMC 05 0.042 ppm Average experimental exposure 0.0075 ppm at the BMC 05 (0.042 x 6/24 x 5/7) Human equivalent concentration 0.0016 ppm at the BMC 05 (gas with extrathoracic respirator y 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 μ g/m 3 )

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

MLE 05 was 0.070 ppm and the BMC 05 was 0.042 ppm. Use of time extrapolation to equivalent continuous exposure, an RGDR adjustment for th e area of the respirator y tract affected, and a total uncertainty factor of 30 resu lted in a chronic REL of 0.05 ppb (0.4 μ g/m 3 ).

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 ChloropicrinDetermination of Noncancer Chronic Re ference 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 unc ertainty factor of 100 (3 for subchronic, 3 for interspecies, and 10 for intraspecies) results in a REL estimate of 0.03 ppb (0.2 μ g/m 3 ).

VII. Data Strengths and Limitations for Development of the REL

Significant strengths in the REL fo r chloropicrin include the durat ion 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 irri tative effects such as rhinitis.

VIII. Potential for Differential Impacts on Children's Health

Chloropicrin is a respiratory irri tant. 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 chloropi crin, or to quantify a differential effect of chloropicrin on infants or children. We are currently evaluating our risk assessment methodologies, in particular the intr aspecies uncertainty factor (UF H ), for adequacy in protecting

infants and children. While we have not so far id entified any indications th at the currently used

UF H of 10 might be less than adequate to protect infants and children, this possibility should be considered in evaluating any exposure situati on involving chronic expos ures of infants or children to chloropicrin.

IX. References

ACGIH. 1992. American Conference of Gove rnmental 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 RD 50 concentration. Toxicol. Appl. Pharmacol. 74:417-429.

Burleigh-Flayer HD, and Benson CL. 1995. Chlor opicrin: Vapor inhalation oncogenicity study in CD rats. Bushy Run Res earch Center, July 29, 1995.

Burleigh-Flayer HD, Kintigh WJ, and Bens on CL. 1995. Chloropicrin: Vapor inhalation oncogenicity study in CD-1 mice. Bus hy Run Research Center, April 20, 1995.

Appendix D3 136 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CDPR. 2000. California Department of Pesticide Regulation. Review of Bu rleigh-Flayer et al. (1995) Chloropicrin: Vapor inhalation oncogenicity study in CD-1 mice.

CARB. 2000. California Air Resources Board. Calif ornia 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 Mallev ialle J. 1985. Chloropicrin in potable water: conditions of formation and production during tr eatment processes. In: Water Chlorination: Chemistry, Environmental Impact and Health Eff ects. Jolley RL, Bull RJ, Davis WP, Katz S, Roberts MH, and Jacobs VA. (eds.) Ch elsea, MI: Lewis Publishers, pp. 1201-1213.

Fries AA, and West CJ. 1921. Chapter VIII. Chlor opicrin. 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: Mi cromedex, 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 irritant s. Am. Ind. Hyg. Assoc. J. 40:207-229.

Merlet N, Thibaud H, and Dore M. 1985. Chloropi crin formation during oxidative treatments in the preparation of drinking wate r. Sci. Total Environ. 47:223-228.

Tomlin, C.D.S. (ed.). 1994. The Pe sticide Manual - World Compendium. 10th ed. Surrey, UK: The British

Crop Protection Council. p. 192.

Yoshida M, Ikeda T, Iwasaki M, Ikeda M, Ha rada 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 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# CHROMIUM, HEXAVALENT (SOLUBLE COMPOUNDS)

Molecular

Formula Molecular Weight Synonyms CAS Registry
Number

CrO 3 99.99 g/mol Chromic trioxide, chromium oxide, chromium trioxide, chromium (VI) oxide. (In acid aqueous solutions,

exists as H 2 CrO 4 – “chromic acid”) 1333-82-0

K 2 CrO 4 194.20 g/mol Potassium chromate, dipotassium chromate, potassium (VI) chromate, dipotassium monochromate, chromate of potash 7789-00-6

Li 2 CrO 4 129.87 g/mol Lithium chromate, chromium lithium oxide, chromic acid dilithium salt, lithium chromate (VI) 14307-35-8

Na 2 CrO 4 161.97 g/mol Sodium chromate, chromic acid disodium salt, chromium disodium oxide, sodium chromate (VI),

chromate of soda 7775-11-3

K 2 Cr 2 O 7 294.20 g/mol Potassium dichromate, dichromic acid dipotassium salt, bichromate of potash 7778-50-9

Na 2 Cr 2 O 7 261.96 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)/m 3
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 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

B. Chromic Trioxide (as chromic acid mist)

Inhalation reference exposure level 0.002 μg Cr(VI)/m 3
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 CrO 3 : dark red or brown crystals, flakes, or powder, exists as chromic acid (H 2 CrO 4 ) in solution;

K 2 CrO 4 , Na 2 CrO 4 : yellow crystals;

K 2 Cr 2 O 7 , Na 2 Cr 2 O 7 : orange-red crystals;

Li 2 CrO 4 : yellow crystalline powder Molecular formula See above Molecular weight See above Density CrO 3 : 2.70 g/cm 3 @ 25°C Boiling point CrO 3 : decomposes (temperature not available);

K 2 Cr 2 O 7 : 500 °C with decomposition;

Na 2 Cr 2 O 7 : 400 °C Melting point CrO 3 : 197 °C;

K 2 CrO 4 : 975 °C;

Na 2 CrO 4 : 792 °C;

K 2 Cr 2 O 7 : 398 °C;

Na 2 Cr 2 O 7 : 356.7 °C Vapor pressure Not applicable Solubility CrO 3 : soluble in water, ethyl alcohol, ethyl ether, sulfuric and nitric acid;

K 2 CrO 4 , K 2 Cr 2 O 7 , Na 2 Cr 2 O 7 : soluble in water, insoluble in ethyl alcohol;

Na 2 CrO 4 : soluble in water, slightly soluble in ethyl alcohol;

Li 2 CrO 4 : 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 mo re toxic than trivalen t 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 toxi city of soluble Cr(VI)

Appendix D3 139 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

compounds, information on poorly soluble Cr (VI) compounds has been included where applicable. In California, the major emission s ource of Cr(VI) results fr om the chrome plating industry (CARB, 1997). Chromic acid, used to electroplate meta l 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 Calif ornia for use in wood and lumber protection treatments (CDPR, 1998). Chromic acid solutions used for this purpos e in the most recent year of reporting (1998) was 71,109 lbs. Minute emissions of Cr(VI) may resu lt 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 wate r is prohibited in California, and recently, in the remainder of the U.S. as we ll. 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 chro mate and dichromate can be produced from imported chromite (Cr(III)) ore. However, no su ch facilities in Calif ornia 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. Envi ronmental exposures would most likely occur through exposure to Cr(VI) dusts (U.S. EPA, 19 98). Cr(VI) may persist in water as watersoluble complex anions. However, any Cr(VI) settl ing 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) detect ed ambient levels of hexavalent chromium ranging from 0.0001

to 0.0003 μ g/m 3 at 10 stationary monitors placed th roughout the South Coast Air Basin. The

annual statewide industrial emissions from fac ilities reporting under the Air Toxics Hot Spots

Act in California based on the most recent i nventory were estimated to be 2311 pounds of hexavalent chromium (CARB, 2000).

IV. Effects of Human Exposure

Cr(VI) forms oxyanions at physiological pH (CrO 4 − 2 ), which are quite similar to sulfate (SO 4 -2 ) and phosphate (HPO 4 -3 ) anions. Therefore, it is able to penetrate virtually every cell in the body because all cells transport sulfat e and phosphate (Costa, 1997). Harmful effects are speculated to

be related to the reduction of Cr(VI) to Cr(III) intracel lularly when it crosses the cell membrane and forms complexes with intracellular macrom olecules. Thus, Cr(VI) compounds have the potential to injure numerous organ systems. Toxicity following chroni c Cr(VI) exposure has been reported in the respirator y 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 uppe r respiratory lesions (U.S. EPA, 1998). In addition, allergic skin and respiratory reactions ca n occur with no relation to dose.

Nasal tissue damage has been frequently obs erved in chromium plating workers exposed chronically to chromic acid mists (Bloomfiel d and Blum, 1928; Vigliani and Zurlo, 1955;

Appendix D3 140 ChloropicrinDetermination of Noncancer Chronic Re ference 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 particulat es containing soluble Cr(VI) compounds, have also reported nasal lesi ons (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, la ryngitis, conjunctivitis, and oral ulcerations (Mancu so, 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/m 3 . 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 adequa te methods or data for determining exposure duration and/or exposure levels. The occupational studies summa rized below provide the most

reliable estimates of inhalati on durations and concentrations resulting in chronic toxicity.

Workers exposed to > 2 μg/m 3 Cr(VI) as chromic acid exhibited an increased incidence of nasal atrophy, nasal mucosal ulcerations, and nasal sept al perforations as compared to controls (Lindberg and Hedenstierna, 1983). Wo rkers exposed to less than 2 μg/m 3 (expressed as < 1.9 μg/m 3 ) exhibited an increased incide nce 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 signifi cant decreases in forced expiratory volume in 1 second (FEV 1 ), forced vital capacity (FVC), and mean forced e xpiratory flow during the middle of the FVC in 1 second (FEF 25-75 ) measurements taken on a Thursday afternoon as compared to those taken on a Monday morning in nonsmoking workers exposed to 2 μg/m 3 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 we re 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 in fer 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 μ g/m3 (range = 1.4-49.3 μ g/m3) total chromium for an average of 2.2 years (range = 1.2 weeks-11 years) (C ohen et al., 1974). Actual exposure to Cr(VI) averaged 2.9

μ g/m3 (range = 0.09-9.1 μ g/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 β 2 -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, 1 983). The 8-hr mean Cr(VI) levels ranged from

2 to 20 μ g/m 3 and averaged 6 μ g/m 3 . Total exposure times ranged from 0.1 to 26 years and averaged 5.3 years. Most of the 24 chrome work ers had irritation symptoms of the airways. As a group, the chrome platers had signifi cantly higher levels of urinary β 2 -microglobulin compared to a group of 27 referents. Comparison of 27 re ferents to a group of 27 ex-chrome-platers found no difference in urinary β 2 -microglobulin levels, even though seven of the ex-chrome-platers had a permanent perforation of their nasal septum (ind icating past exposure to high levels of Cr(VI)).

Appendix D3 141 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

There was no correlation between total exposure time and urinary β 2 -microglobulin levels. Urinary albumin levels remained unchanged in the Cr(VI)-exposed group. The results suggest that the nephrotoxic effects are reversib le at the exposure levels studied.

Gastritis and duodenal ulcers, in a ddition to ulceration and perfora tion of the nasal septum, were observed in chrome platers exposed to a m ean breathing zone concentration of 4 μg/m 3 chromic acid for an average of 7.5 years (Lucas and Kramkowski, 1975).

Male workers in the chromate and dichro mate production industry, whose occupational exposures were 0.05-1.0 mg Cr(VI)/m 3 as chromium trioxide for a mean of 7 years, were reported to have elevated levels of low molecu lar weight proteins (ret inol binding protein and tubular antigens) in the urine (F ranchini and Mutti, 1988). The au thors 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 μ g/m 3 ) and the steel smelting shop (average Cr(VI) exposure 0.5 μ g/m 3 ) 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 signifi cant differences in lung function tests and radiological findings were observed between exposed and control work ers. 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 c ontaining 20 mg/L Cr(VI) in 1965 (Zhang and XiLin, 1987). However, precise exposure con centrations, 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 mo st 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 discre pancy in water levels of Cr(VI) with the earlier study was not

discussed.

V. Effects of Animal Exposure

Exposure of C57BL/6 mi ce to 0 or 13 mg/m 3 CaCrO 4 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 histopathol ogical findings in exposed mice included atrophy of the lymph nodes, spleen, and liver, and occasional sma ll 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 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Glaser et al . (1986) exposed 20 male Wist ar rats/group to 25, 50, and 100 μ g/m 3 aerosolized

sodium dichromate solution and to 100 μ g/m 3 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 wei ghts 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 expo sed to sodium dichromate. Rats exposed to the chromium oxide mixture had a significantl y elevated white blood cell count at the 17 th and 18 th
month, and significantly elevated red blood cells, hematocrits, and hemoglobin levels at the 27 th
month. Mean serum content of total immunoglobulin was significantly reduced in this group at 6 months exposure. Significantly increased l ung weights were observed in chromium oxideexposed rats, and for livers of sodium dichroma te-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 w ith interstitial fibrosis and accumulation of eosinophilic substance in the alveolar lumens. Th e 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/m 3 Cr(VI) as aerosolized sodium dichromate by inhalation for 22 hours per day for 42 days exhibited decreased alve olar macrophage phagocyt ic 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 shee p red blood cells were observed in rats exposed to 25 or 50 μg/m 3 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/m 3 , but not 25 μg/m 3 (Glaser et al. , 1985). Histopathology of major or gans was similar among all groups. Bronchoalveolar lavage fluid contained d ecreased macrophage cell counts above 25 μg/m 3

. Increased antibody response to injected sheep red blood cells was observed in all treatment groups, while alveolar macrophage activ ity was elevated at 25 and 50 μg/m 3 , but was significantly reduced at 200 μg/m 3

.

A later experiment exposed male rats to 0, 50, 100, 200, or 400 μ g Cr/m 3 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 μ g Cr/m 3 , 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. Histopa thological examination reveal ed 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 ob served mainly in the 200 and

400 μ g/m 3 exposure groups. At necropsy, a statistica lly significant increase in lung weight (g dry wt/kg body wt) was observed in rats exposed to 100, 200, and 400 μg/m 3 as compared to controls. Lung weights were sti ll significantly elevated in th e three highest exposure groups 30 days following cessation of exposure. An analysis of the data (Malsch et al. , 1994) determined a

Appendix D3 143 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

benchmark dose (95% confidence interval with do se associated with a 10% elevation in the parameter) for each of these endpoints. The analys is also examined changes in lung and spleen weight reported in Glaser et al. (1985). The most sensitiv e endpoint was LDH in BALF.

Table 1. Key bronchoalveolar lavage fluid (BAL F) and histopathological findings after 90 days exposure to sodium dichromate (Glaser et al., 1990).

μg Cr/m 3 Total Protein

in BALF a (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 immunotoxicologi c effects of inhaled chromium by exposing F-344 rats (10/group/exposure duration) nose-only to 0 and 360 μ g/m 3 potassium chromate 5 hr/day, 5 days/week for 2 or 4 weeks. Exposed rats had greater levels of total recoverable cells, neutrophils, a nd monocytes in bronchopulmonary la vage compared to controls at 2 and/or 4 weeks. Pulmonary macrophages (PM) were reduced, although total PM levels remained unaffected. Four-week exposure to po tassium 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, hyperpla stic and metaplastic changes in the larynx, trachea, and bronchus, and emphysema were observed in mice e xposed two days per week for 12 months to

CrO 3 mist (Adachi, 1987; Adachi et al. , 1986). Chromic acid concen trations were either 3.63 mg/m 3 for 30 minutes per day or 1.81 mg/m 3 for 120 minutes per day. An additional 20 mice exposed to 1.81 mg/m 3 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 toxi city studies summarized below administered soluble Cr(VI) compounds to experimental animals by the oral route.

Groups of eight male and eight female Sprague-D awley rats were supplied with drinking water containing 0-11 ppm (0-11 mg/L) Cr(VI), as K 2 CrO 4 , for 1 year (Mackenzie et al. , 1958). The control group (10/sex) received di stilled water. A second experi ment involved three groups of 12 male and 9 female rats. One group was give n 25 ppm (25 mg/L) Cr(V I); 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), hemato logical 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 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

at 1 year. The 25 ppm groups (and corresponding c ontrols) were examined similarly, except that no animals were killed at 6 months. No signifi cant adverse effects were seen in appearance, weight gain, or food consumption, and there were no treatment-rela ted effects regarding hematological parameters or other tissues in a ny 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 actua l body weight and water consumption data. An abrupt rise

in tissue chromium concentrati ons was noted in rats treated with greater than 5 ppm. The authors stated that “apparently, tissues can accu mulate considerable quantities of chromium before pathological changes resu lt.” In the 25 ppm treatment gr oups, tissue concentrations of chromium were approximately 9 times higher for those treated with hexa valent 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 K 2 CrO 4 ) 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) exposu re 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 pp m 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 lo ss, 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 sacrifi ced 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 patc hes, drooping wrists, and reduced caudal bone ossification. No gross visceral abnor malities were seen in treated groups.

Administration of potassium di chromate 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-implantati on loss and decreased placental weight in both species at the lowest dose. Also at this dose le vel, decreased fetal weight and crown-rump length were observed in mice, and increased resorpti ons and decreased number of live fetuses were

observed in rats. Gross and sk eletal abnormalities were observe d in both species beginning at the 500 ppm dose level.

Appendix D3 145 ChloropicrinDetermination of Noncancer Chronic Re ference 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 r ecovery 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 recove ry period. There wa s no treatment-related histopathology observed in kidneys, ovaries, and testes in either sp ecies. Hematological analysis revealed slight decreases in mean corpuscula r volume (MCV) and mean corpuscular hemoglobin (MCH) at the highest dose in both species, whic h is indicative of iron deficiency. MCV and MCH were normal in these groups following th e 8-week recovery period. Microscopic evaluation of the livers of mice noted cytoplas mic vacuolization of he patocytes in treated animals beginning at 50 ppm. Also in mice, ther e 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 reco very period, feed consumption was comparable

across groups.

The NTP (1997) investigated the potential repr oductive toxicity of C r(VI) in mice using the Reproductive Assessment by Continuous Breeding pr otocol. Groups of 20 male and female pairs of BALB/c mice (F 0 ) were exposed to 0, 100, 200, and 400 ppm potassium dichromate in their diet during the continuous breedi ng phase (approximately 12 weeks). F 1 generation litters

received the same concentration of Cr(VI) in their diet as their F 0 parents and were used for assessment of second generation reproductive toxi city at sexual maturity. There were no treatment-related changes in any of the reproductive parameters in this study. In F 1 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 F 0 and F 1 animals were slightly decreased, and mean food consumption in the F 1 mice was elevated. Reduced mean absolute liver weights were observed in 400 ppm F 0 mice of both sexes. The mean calculated doses were 19.4, 38.6, and 85.7 mg/kg-day for F 0 males and females and 22.4, 45.5, and 104.9 mg/kg-day

for F 1 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 pop ulation, reduced body and te sticular 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 3 β

- ∆ 5 -hydroxy steroid de hydrogenase (3 β

- ∆ 5 -HCH), were also reduced at the two highest doses. Only relati vely small reductions in testicular protein, 3 β

- ∆ 5

- HCH, and serum testosterone were seen in the 20 mg/kg rats.

Appendix D3 146 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

VI. Derivation of Chronic Reference Exposure Levels (RELs)

A. Derivation of Chronic Inhalation Refe rence 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 μ g Cr(VI)/m 3 as sodium dichromate aerosol) Critical effects Bronchoalveolar hyperplasia

LOAEL 50 μ g/m 3

NOAEL Not observed

BMC 05 12.50 μ g/m 3

Exposure continuity 22 hr/day, 7 days/week Exposure duration 90 days Average exposure 11.46 μg/m 3 Cr(VI) (12.50 x 22/24) Human equivalent concentration 24.47 μg/m 3 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/m 3 (0.0002 mg/m 3 )

The study by Glaser et al . (1990) provides the best available inhalation data that demonstrate a dose-response relationship for various pul monary toxicity endpoints. The BMC 05 of 12.50 μg/m 3 was derived from quantal data for bronc hoalveolar 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, Na tional Center for Environmental Assessment, benchmark dose software, version 1.20) of the quant al data provided the most reasonable line fit

and resulted in the lowest BMC 05 . A BMC 05 is considered to be similar to a NOAEL in estimating a concentration associat ed 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/m 3 . Adjustment of the LOAEL of 50 μ g/m 3 (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 existi ng UFs to result in a

cumulative UF of 300.

The U.S. EPA (1998) RfC of 0.1 μ g/m 3 is also based on data from Glaser et al. (1990), but

derived a BMC 10 (16 μg/m 3 ), as developed by Malsch et al . (1994), from continuous data of

Appendix D3 147 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

LDH in BALF. Using a polynomial model provide d by a different benchmark software package ( THC , Clement International Corp., Ruston LA), in creasing LDH concentration in BALF with increasing dose provided the lowest BMC 10 among the various BALF endpoints. OEHHA is currently not developing BMCs for R ELs based on continuous data. A BMC 05 derived from quantal data and a BMC 05 derived from continuous data ma y not have the same meaning. Conceivably, depending on the standard de viations of the data points, the BMC 05 based on

continuous data could still be above the statistica lly 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 assi gnment of the Subchronic UF. The Glaser et al

. (1990) study indicated that chromi um was still acculmulating in l ung 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 s ubchronic 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 expos ure 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 o ccupational study by Huvi nen 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 inha led/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 e xposure 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 ChloropicrinDetermination of Noncancer Chronic Re ference 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/m 3 established as “low exposure” group (8-hr mean < 1.9 μg/m 3 )

NOAEL Not observed Exposure continuity 8 hr/day (10 m 3 per 20 m 3 day), 5 days/week Exposure duration Mean of 2.5 years (range = 0.2 - 23.6 years) Average exposure 0.68 μg/m 3 Cr(VI) (1.9 x 10/20 x 5/7) Human equivalent concentration 0.68 μg/m 3 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/m 3 (0.000002 mg/m 3 )

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 appa rent reversibility of the lesion at this exposure level. While Lindbe rg and Hedenstierna ( 1983) did not follow-up on any of the active cases of nasal ulcerations, wh ich occurred only in workers in the ‘high exposure’ group, they did note that one worker, who exhibited nasal atro phy, had no visible nasal lesions 4 months after termination of exposure.

U.S. EPA (1998) based its RfC of 0.008 μg/m 3 for exposure to chromic acid mists and dissolved Cr(VI) aerosols on the same study but established the LOAEL at 2 μg/m 3 and applied a total UF of 90 (3 each for the LOAEL to NOAEL and s ubchronic to chronic ex trapolation, and 10 for intraspecies extrapolation). It was unclear w hy U.S. EPA (1998) chose UFs of 3 for LOAEL and subchronic extrapolations. It wa s also unclear why the total uncer tainty 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, a lthough 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/m 3 chromic acid mist 2 hr/day, twice a week for 12 months. Lesions were observed in treated mice throughout the respir atory 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 a nd 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

μ g/m 3

.

Appendix D3 149 ChloropicrinDetermination of Noncancer Chronic Re ference 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 chroni c reference exposure leve l 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 K 2 CrO 4 ) 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 dos e 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 co nsumption for the rat (0.035 L/day). Because a LOAEL was not observed in the primary study, th e 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-da y was observed for slight ly 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 a dverse effect at best, th e NTP (1997) noted that

decreased MCV and MCH are indica tors of iron deficiency and s uggested 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 to xicological significance of th is finding is uncertain.

U.S. EPA (1998) applied UFs of 3 for subchronic, 10 for intrasp ecies, 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 crit eria for use of modi fying factors are not

Appendix D3 150 ChloropicrinDetermination of Noncancer Chronic Re ference 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 st udy was greater than 12% of th e estimated lifespan of rats, OEHHA applied UF of 1 for extr apolation to chronic exposure.

U.S. EPA stated its confidence in the RfD as: St udy - 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 endpoi nts are not well studied. Low confidence in the RfD follows.

OEHHA notes that more reproduction/developmenta l 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 re latively 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 chro mic acid mist is the use of human data. The major uncertainties for this inhalation REL is th e 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 i nhalation 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 quantif ied exposure data in humans. The animal studies by Glaser et al . (1990, 1986) suggest that th e lower respiratory airway is a primary target for Cr(VI) dusts. However, occupational studies (M ancuso, 1951; Federal Security Agency, 1953; Machle and Gregorius, 1948; Wang et al ., 1994; Walsh, 1953) indicate th at 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 concentrati ons necessary to result in nasal lesions.

The major strength for the oral REL is the consis tency 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 l ack of lifetime exposure studies in experimental animals and the lack of adequate oral human exposure data.

VIII. References

Appendix D3 151 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Adachi S, Yoshimura H, Katayama H, Takemoto K

. 1986. [Effects of chromium compounds on the respiratory system. IV. Long te rm inhalation of chromic acid mist in electroplating to ICR female mice.] Sangyo Igaku (Jpn. J. Ind. H ealth) 28:283-287. [cited in ATSDR, 1993].

Adachi S. 1987. [Effect of chromium compounds on the respiratory system. V. Long term inhalation of chromic acid mist in electropla ting by C57BL female mice and recapitulation on our experimental studies.] Sa ngyo Igaku (Jpn. J. Ind. Health) 29:17-33. [cited in ATSDR, 1993].

Al-Hamood MH, Elbetieha A, and Bataineh H. 1998. Sexual maturation and fertility of male and female mice exposed prenatally and postnata lly to trivalent and hexavalent chromium compounds. Reprod. Fertil. Dev. 10:179-183.

ATSDR. 1993. Agency for Toxic Substances and Disease Registry. Toxicological Profile for Chromium. Prepared by Syracuse Research Corporation under subcontract to Clement International Corporation under Contract No. 205-88-0608. Prepared for U.S. Department of Health and Human Services, Publ ic Health Service, ATSDR.

ATSDR. 1998. Agency for Toxic Substances and Disease Registry. Toxicological Profile for Chromium. Prepared by Sciences International u nder subcontract to Rese arch Triangle Institute under Contract No. 205-93-0606. Prepared for U.S. Department of Health and Human Services, Public Health Service, ATSDR.

Anwar RA, Langham, FF, Hoppert CA, Alfreds on BV, and Byerrum RU. 1961. Chronic toxicity studies. III. Chronic toxicity of cadmium a nd chromium in dogs. Arch. Environ. Health 3:456-

460.

Bloomfield JJ, and Blum W. 1928. Health hazards in chromium plating. Public Health Rep

43:2330-2351.

CARB. 1988. California Air Resources Board. Pr oposed Hexavalent Chromium Control Plan. Staff Report. Date of Release: January 1988.

CARB. 1997. California Air Resources Board. T oxic Air Contaminant Identification List (Summaries). Date of release: September 1997.

CARB. 2000. California Air Resources Board. Calif ornia Emissions Inventory Development and Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

CDPR. 1998. California Department of Pesticid e Regulation. Summary of Pesticide Use Report Data, 1998, Indexed by Chemical. http://www.c dpr.ca.gov/docs/pur/pur 98rep/chmrpt98.pdf

Chowdhury AR, and Mitra C. 1995. Spermatoge nic and steroidogenic impairment after chromium treatment in rats. Indian J. Exp. Biol. 33:480-484.

Appendix D3 152 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Cohen MD, Zelikoff JT, Chen L-C, and Schl esinger RB. 1998. Immunotoxicologic effects of inhaled chromium: Role of pa rticle solubility and co-expos ure 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 elec troplate workers. Cutis 13:558-568.

Costa M. 1997. Toxicity and carci nogenicity of Cr(VI) in animal models and humans. Crit. Rev. Toxicol. 27(5):431-442.

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Appendix D3 155 ChloropicrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

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Appendix D3 156 ChloropicrinDetermination of Noncancer Chronic Re ference 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-methylbe nzene; 2-hydroxytoluene; 2-methylphenol 95-48-7

m-cresol 1-hydroxy-3-methylbe nzene; 3-hydroxytoluene; 3-methylphenol 108-39-4

p-cresol 1-hydroxy-4-methylbe nzene; 4-hydroxytoluene; 4-methylphenol 106-44-5

I. Chronic Toxicity Summary

Inhalation reference exposure level 600 μg/m 3 (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 C 7 H 8 O 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, glycer ol, petroleum ether; soluble in vegetable oils, glycol Conversion factor 4.42 μg/m 3 per ppb at 25oC

III. Major Uses and Sources

Cresol compounds (mixtures of the ortho-, meta- a nd 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 approxima tely 20% o-cresol, 40% m-cresol, and 30% pAppendix D3 157 Cresol MixturesDetermination of Noncancer Chronic Re ference 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 herbicid es. Cresols also serve as components of degreasing compounds in textile scouring and pain tbrush cleaners as well as fumigants in photographic developers and explosiv es. Cresols also function as antiseptics, disinfectants, and parasiticides in veterinary medi cine. An approximate breakdown of cresol and cresylic acid use is 20% phenolic resins, 20% wire enamel solven ts, 10% agricultural chem icals, 5% phosphate esters, 5% disinfectants and cleaning compounds, 5% ore flotation, and 25% miscellaneous and
exports.

Any combustion process, which results in th e generation of phenolic compounds (such as automobile exhaust or coal, wood, or trash smok e), may be a potential source of exposure to

cresols. Cresols are also formed from the at mospheric 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 mo st recent inventory were estimated to be 8407 pounds of mixtures of cresols (c resylic 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/m 3 resulted in irritation of the th roat and nose, nasal constriction, and dryness in 8 of 10 subjects (Uzhdavini et al. , 1972).

Chemical burns may result from exposure to cr esols (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, even tually turning brown or blac k (Lefaux, 1968). Other reported effects include turbidity, inflammation, and fa tty degeneration of th e liver, nephritis, and hemorrhage of the epicardium and endocardium. An infant fatally exposed to ~20 ml of a 90% cresol solution dermally showed widespread edem a of the internal organs, especially the brain and kidney (Green, 1975). The liver showed si gns 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 di scoloration 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 MixturesDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

V. Effects of Exposures to Animals

The effects of inhaled o-cresol were ex amined 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/m 3 showed signs of respiratory irritation as indicated by incr eased parotid gland secretions. Exposure of mice for 2 hr s/day for 1 month to 50 mg o-cresol/m 3 did not have an effect on mortality, however, hear t muscle degeneration and degene ration of nerve cells and glial

elements were observed.

Uzhdavini et al. (1972) exposed rats (bot h sexes, numbers not st ated) by inhalation to 9.0 ± 0.9 mg o-cresol/m 3 , first for 2 months (6 hours/day, 5 days/week), then for 2 more months (4 hours/day, 5 days/week). Endpoints examined in ra ts included elementary conditioned defensive reflex, white blood cell levels, bone marrow elem ents, 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 e ffect occurred in all expo sed 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 re turned to normal one month after cessation of exposure. Exposed anim als also showed a statistically significant change in the leukoid-to-e rythroid ratio in the bone marrow. Liver toxicity was suggested by an

extension in the duration of hexoba rbital narcosis in treated anim als. Although guinea pigs were similarly evaluated for changes in blood cell coun ts and ECG, scant reporting of experimental

detail limits the usefulness of this portion of the study.

NR rats were exposed by inhala tion to 0.0052 or 0.05 mg tricresol/m 3 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 hi gh-dose group included de creased weight gain, increased central nervous system excitability, increased oxygen consumption, and histological changes in the lung and liver. Serum gamma-globulin levels were also redu ced. No effects were observed in the low-dose group. Rats (6/group, se x unspecified) were also exposed for 24 hours to 0.01, 0.1, and 2.4 mg tricresol/m 3 with a control group of 6 rats for each exposure group. The absorption of neutral red dye by l ung tissue was used as an indicat or of protein denaturation in the tissue. Significantly increased dye absorpti on over control animals was observed at both 2.4 and 0.1 mg tricresol/m 3 . 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 mort ality (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/da y 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 adve rse effects were observed (USEPA, 1999a,b).

Appendix D3 159 Cresol MixturesDetermination of Noncancer Chronic Re ference 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-do sing, 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 significan t neurological events, such as increased salivation, urination, tremors, lacrima tion, palpebral closure, and rapid respiration. Animals given high doses also showed abnormal patterns in the neurobehavioral tests. The NOAEL based on systemic toxicity wa s 50 mg/kg/day (USEPA, 1999a,b).

Dermal exposure of rats to 1.0-1.7 ml cresol/k g 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 in travenous 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 se ts of rats had involuntary mu scle movements and tremors. Neither benzoic acid and hippuric acid, also metabolites of toluene, caused neuroexcitation. The authors concluded that metabolically derive d cresols are plausible candidates for the neuroexcitatory properties of toluene.

In rat liver slices at equimolar concentrations, p-cr esol was 5- to 10-times as toxic as the o- or misomers for cell killing (Thompson et al ., 1994). p-Cresol rapidl y depleted intracellular glutathione levels, while the o- and m-isomers depleted it to a lesser extent. p-Cresol was

metabolized to a reactiv e intermediate which bound covalen tly to protein. The reaction was inhibited by N-acetylcysteine.

The National Toxicology Program (NTP) sponsored reproductive toxicity te sts 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 con centrations in the co ntinuous cohabitation task were 0.05%, 0.2%, and 0.5% in feed (a pproximately 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 re productive toxicant, since (1) fewer F 1 pups per litter were produced, (2) both generations showed re duced pup weights, and (3) reproductive organs showed weight reductions. Unfortunately th e responses were not dose-dependent and the mixture was judged not to be a selective reproductiv e toxicant. Oral gavage administration of o-, m-, or p-cresol, separately, in rats did not produc e selective reproductive t oxicity; i.e., for each of the cresol isomers, in the absence of parental toxicity, th ere was no reproductive toxicity. The NOEL for reproductive toxicity for each isom er was 175 mg/kg/day (Tyl 1989a, 1989b, 1989c).

Appendix D3 160 Cresol MixturesDetermination of Noncancer Chronic Re ference 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 ex trapolation factor 3500 μ g/m 3 per mg/kg/day

Inhalation chronic REL 600 μ g/m 3 (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-me thylphenol). The Rf D for p-cresol was withdrawn by the USEPA. U.S EPA used a s ubchronic uncertainty factor of 10 for a 90 day study in rats. In accordance with its appr oved 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 th at at high concentrations thes e compounds are initially toxic due to their ability to cause chemi cal burns and are therefore of concer n at the site of contact. In humans occupationally exposed, inhalation exposu re is reported to caus e respiratory effects including the development of pneumonia, pulmonary edema, and hemorrhage (Clayton and Clayton, 1982). Other case reports of cresol to xicity to humans are c onfounded by the presence of other compounds, such as phenol, formaldehyde , and ammonia (Corcos, 1939; NIOSH, 1974). The only quantitative information from inhalati on exposures to humans, however, comes from acute exposure studies showi ng irritation at 6 mg cresol/m 3 (Uzhdavini et al. , 1972, as reported in ATSDR, 1992). Toxic effects reported in an imals include bone marrow and liver toxicity in rats from 4 month exposure to 9 mg cresol/m 3 (Uzhdavini et al. , 1972, as reported in U.S. EPA, 1982). Other animal studies have shown more sy stemic 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/m 3 . Kurliandskii et al. (1975) (as reported in HSDB, 1995) observed decreased weight gain with histol ogical changes in the liver and lungs of rats exposed for 3 months to 0.05 mg tricresol/m 3 . Although this study repor ts 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 MixturesDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

The only useful inhalation data for the deve lopment 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). Thes e authors report a LOAEL of 9 mg tricresol/m 3

. OEHHA staff decided not to use th is study. (1) A complete translat ion from the original Russian was not available so that only th e interpretations of others were available. (2) Some endpoints tested are not commonly used in toxicology. A nd (3) some of the resu lts 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/m 3 vs. 0.05 mg/m 3 ). The report from which the lower level is drawn has limitations. Human subjects e xposed 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 cr esols include the use of measur ed exposure data of animals exposed over a significant fraction of their lifetime. Major areas of uncertainty are route-to-route extrapolation, the lack of chr onic human data, and the paucity of reproductive and developmental toxicity studies. Additional inhalati on studies of cresol s 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. Calif ornia 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 o ccupational poisoning by cresols]. Dissertation. Paris (Vigot Freres Editeurs) (in French).

Appendix D3 162 Cresol MixturesDetermination of Noncancer Chronic Re ference 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 - fa tal cresol poisoning following cutaneous absorption (a case report) . Med. Sci. Law 15:65-66.

Heindel J, George J, Fail P, Izard MK, and Gri zzle T. 1997a. o-Cresol. Environ. Health Perspect. 105(Suppl 1):293-294.

Heindel J, Izard MK, George J, Fail P, a nd 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: Micr omedex, Inc. (Edition expires 7/31/96).

Klimkiewicz E, Klinger M, Kurbiel A, and Soltys W. 1974. [Case of liquid lysol poisoning complicated by acute pancre atitis 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 tr icresol 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. Neur otoxicol. Teratol. 11:71-75.

NIOSH. 1974. National Institute for Occupational Sa fety and Health. Health hazard evaluation / Toxicity determination report no. 73-5-110. Prep ared by Cummins Northeastern Incorporated. Dedham, MA.

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce 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 Brende l K. 1994. Cresol isomers: comparison of toxic potency in rat liver slices. Toxi col. Appl. Pharmacol. 125(1):51-58

Appendix D3 163 Cresol MixturesDetermination of Noncancer Chronic Re ference 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 Ra ts. 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 Ra ts. 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 Ra ts. Bushy Run Research Center, Export, PA. Project Report 52-512.

U.S. EPA 1982. United States Environmental Prot ection Agency. Health effects assessment for cresols. Office of Research and Development, Office of Health and Environmental Assessment,

Environmental Criteria and A ssessment 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 st udies in rats. Office of Solid Waste. Washington, DC: U.S. EPA.

U.S. EPA. 1987. o, m, p-Cresol. 90-Day oral subchronic neurotoxic ity study in rats. Office of Solid Waste. Washington, DC: U.S. EPA.

U.S. EPA. 1999a. 2-Methyphenol. Integrat ed Risk Information System (IRIS).

U.S. EPA. 1999b. 3-Methylphe nol. Integrated Risk Information System (IRIS).

Uzhdavini ER, Astaf’yeva K, Mamayeva AA, and Bakhtizina GZ. 1972. [I nhalation toxicity of o-cresol]. Trudy Ufimskogo Nauchno-Isseldovate l’skogo Instituto Gigiyeny Profzabolevaniya 7:115-119. (in Russian)

Appendix D3 164 Cresol MixturesDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# 1,4-DICHLOROBENZENE

(p-dichlorobenzene; di-chloricide; p-dichlo robenzol; Paradow; Paramoth; Parazene; pchlorophenyl chloride)

CAS Registry Number: 106-46-7

I. Chronic Toxicity Summary

Inhalation reference exposure level 800 μg/m 3 (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 li ver 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 C 6 H 4 Cl 2 Molecular weight 147.01 g/mol Boiling point 174 o C Melting point 52.7 o C Vapor pressure 10 torr @ 54.8 o C Solubility Soluble in chloroform, carbon disulfide, alcohol, ether, acetone, benzene Conversion factor 1 ppm = 6.0 mg/m 3 at 25oC

III. Major Uses and Sources

Commercial grade 1,4-dichlorobenz ene (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 oi l suspension; or as crystalline material pressed into various forms (HSDB, 1997). Be sides its role as an intermedia te 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 late st year tabulated, the statewide mean outdoor

Appendix D3 165 DichlorobenzeneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

monitored concentration of 1,4-DCB was a pproximately 0.15 ppb (CARB, 1999). The annual

statewide industrial emissions from facilities re porting under the Air Toxics Hot Spots Act in

California based on the most recent invent ory 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 incl ude malaise, nausea, hepatic manifestations (yellow atrophy and cirrhosis), proteinuria, bili rubinuria, 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/m 3 ) 1,4-DCB by inhalation 7 hours/day, 5 days /week for 6-7 months (Hollingsworth et al. , 1956). High dose animals showed marked tremor s, weakness, loss of weight, eye irritation and unconsciousness. Liver and kidney cha nges 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/m 3 ) for 5 hours/day, 5 days/week for 76 weeks (Riley et al. , 1980). The authors found increased kidney and liver wei ghts in the high dose group. Thus 75 ppm was a NOAEL. Studies with oral exposure to 1,4-DC B, 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 dege neration and focal necrosis).

Three inhalation reproductive st udies, one in rabbits (Hayes et al. , 1985), one in mice (Anderson and Hodge, 1976), and one in rats (Chlor obenzene Producers Assn., 1986), found minimal reproductive effects. In rabbits exposed on da ys 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 incr eased and only in the 300 ppm group (Hayes et al. , 1985). No reduction in reproductive performance wa s 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 (Chlorob enzene Producers Association, 1986), SpragueDawley rats P1 (28/sex/gr oup) were exposed to 0, 50, 150 or 450 ppm (0, 301, 902, or 2705 mg/m 3 ) of 1,4-DCB vapor, 6 hours/da y, 7 days/week for 10 weeks, and then mated for 3 weeks. The second generation F1 weanlings were expos ed 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 we ights, 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 in creased 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 DichlorobenzeneDetermination of Noncancer Chronic Re ference 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 Associat ion 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-bo dy 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/m 3 , 800 μg/m 3 )

The chronic REL for 1,4-dichloroc hlorobenzene is also the U.S. EPA RfC. OEHHA agrees with the U.S.EPA analysis. A 3-fold subchronic uncert ainty 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 equiva lent to 11.2 ppm continuous exposure. Use of an RGDR of 1 and a total UF of 30 (3 for interspecies and 10 for in traspecies) 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-dichlo rochlorobenzene are the observation of a NOAEL and the demonstration of a dose-response relationsh ip. The major uncertainties are the lack of human data and the lack of chronic, multiple-species health effects data.

Appendix D3 167 DichlorobenzeneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

VIII. References

Anderson D, and Hodge MCE. 1976. Paradichloroben zene: Dominant lethal study in the mouse. ICI Report No. CTL/P/296. November.

CARB. 1999. California Air Resources Board. T oxics Air Quality Data. Substance Chooser.

Para-Dichlorobenzene. Available online at http://www.arb.ca.gov/aqd/toxics.htm

CARB. 2000. California Air Resources Board. Calif ornia 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 ra bbits. 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, Hoyl e HR, and Spencer HC. 1956. Toxicity of paradichlorobenzene: Determina tions of experimental animals and human subjects. AMA Arch.

Ind. Health. 14: 138-147.

NTP. 1987. National Toxicology Program. Toxico logy 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. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce 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 Protecti on Agency. Health Assessment Document for Chlorinated Benzenes. EPA/600/8-84/015F. Prepared by the Office of Health and Environmental Assessment, Environmental Criteria and Assess ment Office. Cincinnati, OH: U.S. EPA.

U.S. EPA. 1991. U.S. Environmental Protecti on Agency. Alpha-2u-Globulin: Association with Chemically Induced Renal Toxicity and Neopl asia in the Male Rat. EPA/625/3-91/019F.

Appendix D3 168 DichlorobenzeneDetermination of Noncancer Chronic Re ference 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 Protecti on 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 DichlorobenzeneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# 1,1-DICHLOROETHYLENE

(DCE; 1,1-dichloroethene; VD C; vinylidene chloride)

CAS Registry Number: 73-35-4

I. Chronic Toxicity Summary

Inhalation reference exposure level 70 μ g/m 3 (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 C 2 H 2 Cl 2 Molecular weight 96.95 g/mol Boiling point 31.7 o C Melting point

− 122.5 o C Vapor pressure 500 torr @ 20 o C Solubility Soluble in water (2.5 g/L); miscible in organic

solvents Conversion factor 3.97 μg/m 3 per ppb at 25 o C

III. Major Uses and Sources

1,1-Dichloroethylene (1,1-DCE) is used in the production of polyvinyliden e chloride copolymers (HSDB, 1994). 1,1-DCE containing copolymers include other compounds such as acrylonitrile, vinyl chloride, methacrylonitrile, and methacrylate. These copol ymers 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 ch loride 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 estimat ed to be 2458 pounds of vinylidene chloride (CARB, 2000).

IV. Effects of Human Exposure

Appendix D3 170 1,1-DichloroethyleneDetermination of Noncancer Chronic Re ference 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 hepatotoxi city and nephrot oxicity after repeated, low-level exposures (USEPA, 1976; Ott et al. , 1976). However, these investigations

were conducted in industrial set tings with the possibility of mixed chemical exposures. In preliminary clinical findings reported by th e EPA (1976), workers exposed to 1,1-DCE for 6 years or less had a high inciden ce of hepatotoxicity, with liver scans and measurements of liver enzymes revealing 50% or grea ter 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 toxi city 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 inhalati on exposure. Such exposure is marked by both biochemical changes (alterations in serum en zyme levels) and histological changes (e.g., midzonal and centrilobular swelling, degene ration, 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 de generation or hepatocellu lar 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 vacuolat ion 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 c ontinuous 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 deposit ion, 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 48ppm) displayed mottled livers at 15 ppm, and increased liver enzyme levels (serum glutamicpyruvic transaminase (SGPT) and alkaline phosp hatase (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 gui nea pigs from Prendergast et al . (1967)

ppm 1,1-DCE (mg/m 3 ) 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-DichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Additional adverse effects observed to a lesser exte nt in laboratory animals include respiratory and renal toxicity. Nephrotoxi city observed following chroni c 1,1-DCE exposure included gross organ (increases in kidney weight) (Klimisch et al. , 1979; Quast et al. , 1986) and histological changes (tubular swelling, degene ration, 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, ra bbits, 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 laborator y animals have demonstrated that 1,1-DCE is readily absorbed and rapidly distributed followi ng inhalation exposure (Dallas et al. , 1983; McKenna et al.

, 1978b). Following inhalation exposure to radio actively 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 de toxification 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-DichloroethyleneDetermination of Noncancer Chronic Re ference 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/m 3 group) Exposure method Continuous whole body i nhalation (0, 20, 61, 101, or 189 mg/m 3 ) Critical effects Increased mortality at 61, 101, and 189 mg/m 3 ; hepatic effects (mottled livers and increases in SGPT and AP enzymes) noted at 189 mg/m 3

LOAEL 61 mg/m 3 (15 ppm)

NOAEL 20 mg/m 3 (5 ppm) Exposure continuity Continuous Exposure duration 90 days Average experimental exposure 20 mg/m 3 for NOAEL group Human equivalent concentration 20 mg/m 3 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-sp an is approx. 6 years) Interspecies uncertainty factor 3 Intraspecies uncertainty factor 10 Cumulative uncertainty factor 300 Inhalation reference exposure level 0.07 mg/m 3 (70 μ g/m 3 ; 0.02 ppm; 20 ppb) The principal study (Prendergast et al. , 1967) identified adverse hepati c 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 othe r available subchronic and chronic studies, this principal study included multiple exposure le vels of 0, 5, 15, 25 and 48 ppm (0, 20, 61, 101, and 189 mg/m 3 ). Mortality, hematologic and body weight da ta were well tabulated and presented in this study. Histopathologic eval uation was conducted on the heart, lung, liver, spleen and kidneys. Following continuous exposure, adverse hepatic effect s included focal necrosis in monkeys (LOAEL = 189 mg/m 3 , NOAEL = 101 mg/m 3 ), in dogs (LOAEL = 189 mg/m 3

, NOAEL = 101 mg/m 3 ), and in rats (LOAEL = 189 mg/m 3 , NOAEL = 101 mg/m 3 ); and altered lipid content and increases in SGPT and al kaline phosphatase in guinea pigs (LOAEL = 189 mg/m 3 , NOAEL = 20 mg/m 3 ). Additionally, renal alterations were observed in rats as nuclear hypertrophy in the tubular epithelium (LOAEL = 189 mg/m 3 , NOAEL = 61 mg/m 3 ). Monkeys exposed to 1,1-DCE also displayed a greater than 25% decrease in body weight (LOAEL 189 mg/m 3 , NOAEL 20 mg/m 3 ). 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 a nd subchronic studies available consistently demonstrate adverse hepatic eff ects following 1,1-DCE exposure (Lee et al. , 1977; Maltoni et

Appendix D3 173 1,1-DichloroethyleneDetermination of Noncancer Chronic Re ference 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 in cidence over time (Quast et al. , 1986). A more severe hepatocellular

necrosis and renal tubular necrosis were obser ved 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 mont hs 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 μ g/m 3 ).

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 labo ratory animals. In addition, few industrial surveys and epidemiological studies are availabl e on the adverse effects of 1,1-DCE in humans; these are limited by small sample size, short foll ow-up, and/or brief exposure periods. But this limited evidence does suggest an association be tween repeated exposure to 1,1-DCE and liver damage in humans (EPA, 1976), and the key st udy is an animal study which found adverse hepatic effects. No toxicokinetic data rega rding 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 et hyl alcohol. Am. Ind. Hyg. Assoc. J. 37(6):345-352.

CARB. 2000. California Air Resources Board. Calif ornia 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. Toxico l. 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-DichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Jaeger RJ, Shoner LG, and Coffman LJ. 1977. 1,1- Dichloroethylene hepatotoxicity: Proposed

mechanism of action and dist ribution and binding of [ 14 C] radioactivity following inhalation exposure in rats. Environ. Health Perspect. 21:113-119.

Klimisch HJ, Deckardt K, and Mitea D. 1979. [Subchr onic toxicity of vinylid ene chloride in rats after 6 weeks exposure.] BASF Aktiengesell schaft, 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 ch loride 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: Pr inceton Scientific Publ ishers. [as cited in USDHHS, 1994.]

McKenna MJ, Zempel JA, Madrid EO, and Ge hring PJ. 1978. The pharmacokinetics of [14C] vinylidene chloride in rats following inhalati on administration. Toxicol. Appl. Pharmacol.

45:599-610.

Ott MG, Fishbeck WA, Townsend JC, and Schne ider EJ. 1976. A health study of employees exposed to vinylidene chloride. J. Occup. Med. 18:735-738.

Plummer JL, de la Hall PM, Il sley AH, Jenner MA, and Cousins MJ. 1990. Influence of enzyme induction and exposure profile on liver injury due to chlorinate d hydrocarbon inhalation.

Pharmacol. Toxicol. 67:329-335.

Prendergast JA, Jones RA, Jenkins LJ, and Sieg el J. 1967. Effects on experimental animals of long-term inhalation of trichloroethylene, carbon tetrachloride, 1,1,1-trichloroethane, dichlorodifluoromethane, and 1,1-dichloroethyl ene. Toxicol. Appl. Pharmacol. 10:270-289.

Quast JF. 1976. Pathology report on male and female rats exposed to vinylid ene chloride vapors for 6 hours per day, 5 days per week during a 30-day period. Toxicology Research Laboratory, Health and Environmental Sciences, Dow Chemi cal 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 ch loride 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 isolat ed perfused rat liver.

Arch. Toxicol. 41:169-179.

USDHHS. 1994. United States Depart ment of Health and Human Se rvices. Toxicological profile for 1,1-dichloroethene (update). U.S. Department of Health & Human Se rvices. Public Health Service. Agency for Toxic Substanc es and Disease Registry. TP-93/07.

Appendix D3 175 1,1-DichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

U.S. EPA. 1976. United States Environmental Pr otection Agency. Summary characteristics of

selected chemicals of near-term intere st. 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-DichloroethyleneDetermination of Noncancer Chronic Reference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# DIETHANOLAMINE

(DEA; 2,2’-iminodiethanol; 2,2’-i minobisethanol; diethylola mine; 2,2’-aminodiethanol; 2,2’- dihydroxydiethylamine)

CAS Registry Number: 111-42-2

I. Chronic Toxicity Summary

Inhalation reference exposure level 3 μ g/m 3 ( 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 C 4 H 11 NO 2 Molecular weight 105.14 g/mol Density 1.097 g/cm 3 @ 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/m 3 @ 25°C

III. Major Uses and Sources

Diethanolamine is used in the formation of soap s, 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 ch emicals, as an absorbent for acidic gases, as a humectant, as an intermediate in the synthesis of morpholine, as a corros ion inhibitor, and as a component in textile sp ecialty agents (Beyer et al. , 1983). Diethanolamine is permitted in articles intended for use in production, processi ng, or packaging of f ood (CFR, 1981; cited in Melnick and Thomaszewski, 1990). It is also f ound in adhesives, sealants, and cutting fluids (Melnick and Thomaszewski, 1990). The annual st atewide industrial emi ssions 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 DiethanolamineDetermination 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 cas e 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/m 3 and 1.0 mg/m 3 ) below the occupational limit of 2.0 mg/m 3 . 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 c holine 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 sy stem, 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 hour s per working day for 90 days at target concentrations of 15, 150, and 400 mg/m 3 . 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 neurotoxi city. 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 repr oductive 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 dos e group. Based on this study 15 mg/m 3 is a NOAEL for liver and kidney effects a nd a LOAEL for irritation of the larynx. The equivalent continuous exposur e at the LOAEL is 2.7 mg/m 3 (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/m 3 4/20 0/20 20/20

150 mg/m3 20/20 13/20 20/20 400 mg/m 3 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/m 3 ) DEA vapor 8 hours/day, 5 days/week for 13 weeks re sulted in depressed gr owth rates, increased lung and kidney weights, and even some mort ality. Rats exposed continuously for 216 hours (nine days) to 25 ppm (108 mg/m 3 ) DEA showed increased liver a nd kidney weights, elevated

Appendix D3 178 DiethanolamineDetermination of Noncancer Chronic Reference Exposure Levels December 2001

blood urea nitrogen (BUN), and increased serum gl utamate oxaloacetate transferase (SGOT), an indicator of liver damage (Hart ung et al., 1970). In studies at lower DEA levels, Eastman Kodak (1967) exposed dogs, weanling and adult ra ts, and guinea pigs to 0.26 ppm (1.1 mg/m 3 ) DEA for 90 days and found no pathology attributable to DEA. In a 45-day study with 0.5 ppm (2.2 mg/m 3 ) DEA they also found no pathology attributab le 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/m 3 . 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 inci dence of total fetal skeletal variations, were observed at 200 mg/m 3 . No teratogenic effects were s een at any level. Thus 50 mg/m 3 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 exposu re to DEA at all concentrations tested: 320, 630, 1250, 2500, and 5000 ppm in males, and 16 0, 320, 630, 1250, and 2500 ppm in females. Hematological effects included decreased hem oglobin and mean corpuscular volume (Melnick et al. , 1994a). Similar hematological changes were observed following daily topical treatment. In addition to the hematological effects, female ra ts 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 pp m. 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 disp layed dose-dependent hepatotoxicity, nephrotoxicity, and cardiac toxicit y. Daily topical treatment in a separate study resulted in skin lesions in mice. Significant he patic toxicity was observed at a ll drinking water concentrations,

and skin lesions were obse rved 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 wi th 330 mg DEA/kg/day significantly inhibited formation of phosphatidyl choline and phos phatidyl ethanolamine in the liver as compared with control rats. In a subsequent study, Barbee and Hartung (1979b) noted changes in liver mitochondrial act ivity in rats (4 per group) following exposure to DEA in

Appendix D3 179 DiethanolamineDetermination 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 cutane ously 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 ir ritation 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 topi cal application of DEA by Melnick et al . (1994b). Rat fetuses had in creased incidences of six skeletal variations at 1500 mg/kg/day. Lower doses were withou t 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 administer ed 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 hema tological effects. Due to a dose discrepancy, the authors adjusted the no observable effect level (NO EL) for DEA developmental toxicity to 380 mg/kg/day for rats. In rabbits, th e 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/m 3

NOAEL Not observed Exposure duration 90 days Average experimental exposure 2700 μ g/m 3 for LOAEL group (15 mg/m 3 x 6h/24h x 5d/7d x 1000 μ g/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 μ g/m 3 (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/m 3 for irritation of the

Appendix D3 180 DiethanolamineDetermination 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/m 3 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 fu ll LOAEL uncertainty factor of 10.

For comparison, the BASF (1993) developmenta l study by the inhalation route found a LOAEL of 200 mg/m 3 DEA and a NOAEL of 50 mg/m 3 for fetotoxic effects. The equivalent continuous exposure at the NOAEL is 12.5 mg/m 3 . Multiplying by an RGDR of 1 and dividing by an interspecies uncertainty factor (UF A ) of 3 and an intraspecies uncertainty factor (UF H ) of 10

results in a REL estimate of 40 μ g/m 3

.

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 comprehensiv e 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 m 3 of air per day leads to an inhalation

REL estimate of 50 μ g/m 3 (10 ppb) DEA.

VII. Data Strengths and Limitations for Development of the REL

The diethanolamine database is relatively weak. Major areas of uncerta inty 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 chr onic inhalation, multiplespecies, health effects data.

VIII. Potential for Differential Impacts on Children's Health

Since the proposed chronic REL of 3 μ g/m 3 based on laryngeal effects is much lower than the comparison REL of 40 μ g/m 3 based on fetotoxic effects, th e 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 evid ence in the literature

to demonstrate that DEA exacerbates asthma or to quantify a differential effect of diethanolamine on the larynx or on othe r organs in infants and children.

Appendix D3 181 DiethanolamineDetermination 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. T oxicol. Appl. Pharmacol. 47:421-430.

Barbee SJ, and Hartung R. 1979b. Diethanolamine-i nduced alteration of hepatic mitochondrial function and structure. Toxicol. Appl. Pharmacol. 47:431-440.

Beyer KH Jr, Bergfeld WF, Berndt WO, B outwell RK, Carlton WW, Hoffman DK, and Schroeter AL. 1983. Final report on the safety as sessment 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. Calif ornia Emissions Inventory Development and Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

CFR. 1981. Code of Federal Regula tions. 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 Handbo ok. 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 Numb er 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. Toxi col. 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, Deckar dt K, Kaufman W, and Hildebrand B. 1996.

Diethanolamin – Subchronic inhala tion toxicity and neurotoxicit y study in Wistar rats. 90-day liquid aerosol exposure-. BASF Project No. 50I0075/93011. Heidel berg, 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 DiethanolamineDetermination of Noncancer Chronic Reference Exposure Levels December 2001

Knaak JB, Leung HW, Stott WT, Busch J and Bi lsky J. 1997. Toxicology of mono-, di-, and

triethanolamine. Rev. Environ. Contam. Toxicol. 49:1-86.

Lehman-McKeeman LD, and Gamsky EA. 1999. Diet hanolamine inhibits choline uptake and phosphatidylcholine synthesis in Chinese hamste r 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. Diethanol amine. 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 wa ter and topical applica tion 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 applicati on 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 occu pational asthma, a cas e report. Clin. Exp. Allergy 28(3):358-362.

Appendix D3 183 DiethanolamineDetermination of Noncancer Chronic Re ference 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/m 3 (30 ppb) Critical effect(s) Liver dysfunction and resp iratory 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 C 3 H 7 NO
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/m 3 per ppb at 25oC

III. Major Uses and Sources

Dimethylformamide (DMF) is primarily used as a solvent in the pr oduction 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 petr ochemical industries (Howard, 1993). The annual

statewide industrial emissions from facilities re porting under the Air Toxics Hot Spots Act in

California based on the most recent inventor y 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 st atistically significant incidence of hepatic impairment, as indicated by elevated gamma-glutamyl transpep tidase levels and digestive disturbances, was

Appendix D3 184 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

noted (Cirla et al. , 1984). Other changes, that were not statistically si gnificant, included increased SGOT and SGPT and enlarged livers. The mean time-weighted average concentration of DMF was 22 mg/m 3 (range = 8-58 mg/m 3 ). Symptoms of irritation occurring only during work at statistically significantly higher inci dences included watery eyes, dry throat, and coughing. Also, the exposed worker s reported a reduced sense of smell and dry coughs at home with a statistically significant difference as compared to co ntrols. Several of the DMF exposed workers also reported alcohol in tolerance characterized by a di sulfiram-type reaction (facial flushing and palpitations following alcohol i ngestion). Alcohol consumption, a potential confounder, was controlled for in the study design.

A similar study was conducted on workers who had b een 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/m 3 or an 8-hour TWA of 3 mg/m 3 . Measures of liver function including S GOT, SGPT, gamma-glutamyl transferase, and alkaline phosphatase levels we re 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 e nough to reliably detect a difference in enzyme

levels.

Redlich et al . (1988) characterized a plant-wide outbr eak 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 func tion tests done at least once. At the work site DMF was being used in poorly ventilated areas without appropr iate skin protection. No other major known hepatotoxic exposure was identified. Overall, 36 of 58 (62%) workers test ed had elevations of either aspartate aminotransferase (AST) or al anine 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 infec tious causes of hepatitis in all but 2 workers. Changes, characteristic of liver injury, were confirmed by histologic exam ination 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 wo rkers most severely affected from exposure. However, some patients showed persistent eleva tions 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 leathe r factory by performing medical examinations, liver function tests, and creatin e phosphokinase (CPK) determinati ons. Air concentrations were measured with personal samplers and gas chroma tography. 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/m 3 ), medium (index 1: 10-40 ppm), and lo w (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 re sult that did not change after stratification by hepatitis B carrier status. Logistic regression analysis indicated that exposure to high DMF levels was associated with elevat ed 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 signi ficantly higher mean ALT and aspartate

Appendix D3 185 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

aminotransferase (AST) activitie s, especially among those who were not HBsAg carriers. A significant association existed betw een elevated CPK levels and exposure to DMF. However, an analysis of the CPK isoenzyme among 143 work ers did not reveal any specific damage to muscles. Thus the authors ascribed the liver injury to DMF.

U.S. EPA (1994) states that subj ective evidence of liver toxicit y, 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 tr imester intrauterine deaths were observed in a group of women working as quality control anal ysts in the pharmaceutical industry (Farquhason

et al. , 1983). This represente d a 30% stillbirth rate as compared with the average for the general population of about 0.26%. While the authors conclude d that the occurrence of stillbirth in these women was not likely due to chance, the effects ca nnot 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:C D rats and mice to 0, 25, 100, or 400 ppm DMF for 6 hr/day, 5 days/week for 18 months (m ice) 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 b ody 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 in creased relative liver weights, centrilobular hepatocellular hypertr ophy, lipofuscin/hemosiderin accumulation in Kupffer cells, and centrilobular si ngle cell necrosis (400 ppm only). In mice, increased liver weights (100 ppm males, 400 ppm both sexe s), centrilobular hepa tocellular hypertrophy, accumulation of lipofuscin/hemosiderin in Kupffe r cells, and centrilobular single cell necrosis were observed in all exposure groups. These ob servations occurred in a dose-response fashion and were minimal at 25 ppm. No increase in hepatic cell proliferati on 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 usin g three species (mice, rabbits, and rats) and four routes of administration (oral, inhalation, dermal, and intr aperitoneal) identified the rabbit as the most sensitive of the three species. Groups of 15 pr egnant 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/m 3 ) DMF (Hellwig et al. , 1991). Slight maternal toxici ty, as indicated by non-statistic ally significant decreases in maternal body weight gain, was observed in the 450 ppm exposure group. An increased number of total malformations per litter was observe d in the 450 ppm exposure group. Malformations observed at statistically higher incidences compared to contro ls included hernia umbilicalis, external variations, pseu doankylosis of the forelimbs, and skel etal variation and retardation. The authors conclude that there was a clear teratoge nic effect in rabbits following maternal exposure to 450 ppm DMF and a marginal effect followi ng exposure to 150 ppm DMF. A NOAEL of

Appendix D3 186 DimethylformamideDetermination of Noncancer Chronic Re ference 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 da ily dose of 45 mg/kg/day, which coincides with previous work on this co mpound 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/m 3

NOAEL Not observed Exposure continuity 8 hr/day (10 m 3 /day), 5 days/week (assumed) Average occupational exposure 7.9 mg/m 3 for LOAEL group (22 x 10/20 x 5/7) Human equivalent concentration 7.9 mg/m 3
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/m 3 (80 μg/m 3 , 0.03 ppm, 30 ppb)

The U.S. EPA (1994) based its RfC of 30 μ g/m 3 on the same study but included a Modifying Factor (MF) of 3 due to lack of reproductive toxicity data in th e DMF database. The criteria for use of modifying factors are not well specified by U.S. EPA. Su ch modifying factors were not used by OEHHA. Intermediate uncertainty factors were used for LOAEL to NOAEL and subchronic to chronic extrapolation because of th e 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, e quivalent 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 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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 DM F lower than the actually proposed TLV. G. Ital. Med. Lav. 6(3-4):157-158.

Cirla AM, Pisati G, Invernizzi E, and To rricelli P. 1984. Epidemiological study on workers exposed to low dimethylformamide concentr ations. 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-dimethylform amide: key to the understanding of its toxicity. Chem. Res. Toxicol. 6(3):245-251.

HSDB. 1994. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 11/31/94).

Hellwig J, Merkle J, Klimisch HJ, and Jackh R. 1991. Studies on the pren atal 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 dimethylform amide 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 w ith occupational exposure to the solvent dimethylformamide. Ann. Intern. Med. 108:680-686.

U.S. EPA. 1994. U.S. Environmental Prot ection Agency. IRIS. 1994. Integrated Risk Information System (IRIS). Reference concentrat ion for chronic inhalation exposure (RfC) for N,N,-Dimethylformamide. Available onlin e at http://www.epa.gov/ngispgm3/iris

Wang JD, Lai MY, Chen JS, Lin JM, Ch iang 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 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# 1,4-DIOXANE

(Synonym: dihydro-p-dioxin, diethylene dioxide, p-diox ane, glycolethylene ether)

CAS Registry Number: 123-91-1

I. Chronic Toxicity Summary

Inhalation reference exposure level 3,000 μ g/m 3 (800 ppb) Critical effects Liver, kidney, hematol ogic 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 C 4 H 8 O 2 Molecular weight 88.10 g/mol Boiling point 101.5 o C Melting point 11.8 o C Vapor pressure 37 torr @ 25 o C Solubility Miscible with water, aromatic solvents, and oils

Kow 0.537 Conversion factor 3.60 μ g/m 3 per ppb at 25 o C

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 ag ent in the textile industry. Dioxane is used as a solvent in chemical synthesi s, as a fluid for scintillation counting, and as a dehydrating agent in the preparation of tissue sections for hi stology (Grant and Grant, 1987; HSDB, 1995). The annual statewide emissions from facilities repo rting under the Air Toxi cs Hot Spots Act in

California based on the most recent inventor y 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 admini stration (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 cond itions the half-life for dioxane elimination is

Appendix D3 189 DimethylformamideDetermination of Noncancer Chronic Re ference 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 a ppeared 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. Altho ugh the kidney and liver

lesions were considered manifestations of acu te exposure, the author suggested a chronic component that was manifested by increased wh ite blood cells. A case was reported in which a worker, who died following exposure by inhalation a nd 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 ef fects 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 SG OT), alanine transaminase (serum glutamate pyruvate transaminase or SGPT), alkaline phosphatase, and gamma glutamyltransferase activities (liver function) was noted in these wo rkers, but not in those who had retired. The indicators of liver dysfunction, however, could no t be separated from alcohol consumption or exposure to ethylene chlorohydr in 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, bu t the data were not reanalyzed after removing the deaths due to malignant neoplasms. The st udy is limited by the small number of deaths and by the small sample number. The study did not a ssess hematologic or clinical parameters that

could indicate adverse health eff ects 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 diseas e process involves local autoimmunity with decreased circul ating white blood cells (see below).

V. Effects of Animal Exposure

In rats, the major metabolite of dioxane is HE AA, which is excreted through the kidneys (Braun and Young, 1977). Exposure to dioxane by ingestion results in saturation of metabolism above

Appendix D3 190 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

100 mg/kg given in single dose. Saturation of me tabolism 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 leve l, 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 thes e conditions was 1.1 hours. The absorption of dioxane through the inhalation pathway could not be exactly dete rmined, because of a high inhalation rate (0.24 liters/min), calculated on the basi s of complete absorption (Young et al. , 1978; U.S. EPA, 1988). Although the high inhalation rate co uld 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 inha lation modeling did not use more than one

inhalation dose. No studies we re located on the biological or biochemical properties of HEAA or the properties of the enzyme(s) that are re sponsible for the transfor mation 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 d ecreased body weight gains, compared to unexposed control rats, were not observe d. Among the male rats, decreased blood urea nitrogen (kidney function), decreased alkaline p hosphatase (cholestatic liv er function), increased red blood cells, and decreased wh ite blood cells were observed. According to the authors, exposure-related, non-cancerous tissue lesions were not observed duri ng 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, in creased SGOT activity at the two highest doses and increased SGPT activity at all doses were meas ured in the sera of the exposed rats. Rats exposed to the highest dose also exhibited increased urin ary protein and chlori de 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) re quired for an electric stimulus to result in

excitation of extensor and fle xor 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 tw o 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 latt er study, although such contamination would be unlikely to account for the large difference in exposure levels.

Appendix D3 191 DimethylformamideDetermination of Noncancer Chronic Re ference 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 duri ng exposure. Exposure to the highest level resulted in decrea sed body weight gain and increas ed deaths. According to the authors, exposure related hematologic change s did not occur. Histopathologic examination revealed evidence of regeneration of hepatic a nd kidney tissues in rats exposed to 1.0 or 0.1 percent, but not in rats exposed to 0.01 percen t dioxane. On the assumption of total absorption of dioxane from the gastrointestin al 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 consum ption 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 se parate 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/m 3 ; 3000 μg/m 3 )

Appendix D3 192 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

The lifetime rat inhala tion study of Torkelson et al. (1974) is the only de tailed inhalation study available in the literature. The Pilipyuk et al. (1977) study contains usef ul 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 lead s 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 leve ls. The lack of toxic hematologic endpoints observed in the ingestion study s uggests that toxicity of dioxa ne may be route-of-exposure specific. Hematologic changes were also observe d in the early worker study wherein changes in white blood cell count occurred (Bar ber, 1934), but the directions are different. The studies on humans and rodents therefore sugge st inhalation of dioxane may lead to adverse biologic effects, but good dose-response data are not available. A partial expl anation may lie in the doseresponse characteristic of the me tabolism of dioxane, wherein toxi city may be a function of the saturation of metabolism. For inhalation, neithe r 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 experi mental 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, a nd the lack of developmental studies.

VIII. References

Barber H. 1934. Haemorrhagic nephritis and necros is 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 ra t. 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 DimethylformamideDetermination of Noncancer Chronic Re ference 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. Tera togenesis study of dioxane in rats. Toxicol.

Lett. 26:85-88.

Grant R, and Grant C. 1987. Grant and Hackh’s Ch emical Dictionary. R. Grant and C. Grant, eds. 5th ed. New York: McGraw-Hill Book Co. p. 189.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version). Denver, CO: Microm edex, 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 dioxa ne? AMA Arch. Ind. Health. 20:445-447.

Kociba RJ. 1974. Chronic toxicity study of dioxa ne in the drinking water of Sherman rats. Midland,MI: Dow Chemical Company.

NIOSH. 1977. Criteria for a Recommended Sta ndard. Occupational Exposure to Dioxane. National Institute for Occupational Safety and Hea lth, Centers for Disease Control, Public Health Service, Department of Health Educa tion and Welfare. Publication No. 77-226.

Pilipyuk ZI, Gorban GM, Solomin GI, and Go rshunova 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-medi cal investigation resu lts in the case of workers exposed to dioxane.] (Ger.) Ar beitsmed. Sozialmed. Praventivmed. 11:35-46.

Torkelson TR, Leong BKJ, Kociba RJ, Richte r 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 Protecti on 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 factor s 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 DimethylformamideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Young JD, Braun WH, Gehring PJ, Horvath BS, and Daniel RL. 1976. 1,4-Dioxane and β

hydroxyethoxyacetic acid excretion in urine of hu mans 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. Toxi col. Environ. Health. 3:507-520.

Appendix D3 195 DimethylformamideDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (0.8 ppb) Critical effects Histological changes in na sal turbinates in rats Hazard index target(s) Respiratory system; eyes

II. Physical and Chemical Properties (HSDB, 1997; CRC, 1994)

Description Colorless liquid Molecular formula C 3 H 5 ClO Molecular weight 92.52 g/mol Density 1.181 g/cm 3 @ 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/m 3 @ 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 lens es (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 show n no evidence of decreased sperm counts in populations occupationally exposed to epichlorohydrin (Milby et al. , 1981).

Appendix D3 196 EpichlorohydrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

V. Effects of Exposures in Animals

Rats were exposed for 136 weeks (6 hours/da y, 5 days/week) to 0, 10, 30, or 100 ppm (0, 38, 113, or 380 mg/m 3 ) 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 pa ssages of 90% of the co ntrol animals, as well as in the treated animals, prevented compar ison 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/m 3 ) for 6 hours/day, 5 days/week for 11-19 exposures showed ev idence of extrarespiratory effects. These included liver congestion and n ecrosis and tubular atrophy in the kidneys at the highest concentration (Gage, 1959). Lethargy and weight loss were obs erved 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/m 3 ), were significantly less fertile than cont rols, 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 e ffects on fertility occurred in the exposed female rats. Degenerative changes in th e nasal epithelium were observed in the female rats exposed to 25 ppm (94.5 mg/m 3 ), 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/m 3 ) epichlorohydrin 7 hours/day dur ing the critical days of gestation. There were no significant differences between c ontrols 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/m 3 ) 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 examinat ions were performed at the end of the experiment. Exposures of rats to 25 and 50 ppm epichlorohydrin resulted in inflammation, focal erosions, hyperplasia, and metapl asia in the nasal turbinates. No adverse effects were observed in rats exposed to 5 ppm (19 mg/m 3 ). Mice similarly showed fo cal erosion, hyperplasia and metaplasia in the epithelium of the nasal turb inates when exposed to 25 ppm epichlorohydrin or
greater.

Appendix D3 197 EpichlorohydrinDetermination of Noncancer Chronic Re ference 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/m 3 )

NOAEL 5 ppm (19 mg/m 3 ) 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

m 3 /day , MVh = 20 m 3 /day, SAa(ET) = 15 cm 2 ), SAh(ET) = 200 cm 2
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/m 3 ; 3 μ g/m 3 )

The U.S. EPA (1994) based its RfC of 1 μ g/m 3 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 epichloro hydrin include the availability of controlled exposure inhalation studies in multiple species at multiple exposure conc entrations and with adequate histopathogical analysis and the obser vation of a NOAEL. Major areas of uncertainty are the lack of adequate huma n exposure data, the lack of ch ronic inhalation exposure studies, the limited reproductive toxicity data, and the small groups tested in the study.

VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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 epichlor ohydrin vapour. Br. J. Ind. Med. 16:11-14.

Appendix D3 198 EpichlorohydrinDetermination of Noncancer Chronic Re ference 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 fe male 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 ch loride in rats and rabbits. Fundam. Appl.

Toxicol. 3:437-442.

John JA, Quast JF, Murray FJ, Calhoun LG, a nd Staples RE. 1983b. Inha lation 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. Inhala tion carcinogenicity of epichl orohydrin 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. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce 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 r odents. Toxicology Research Laboratory, Dow Chemical U.S. A., Midland, MI (unpublished).

U.S. EPA. 1994. U.S. Environmental Protec tion 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 EpichlorohydrinDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# 1,2-EPOXYBUTANE

(1-butene oxide; 1,2-butene oxide; 1,2-butyl ene oxide; 1,2-epoxybut ane; 2-ethyloxirane; ethylethylene oxide; NCI-C55527)

CAS Registry Number: 106-88-7

I. Chronic Toxicity Summary

Inhalation reference exposure level 20 μ g/m 3 (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 C 4 H 8 O Molecular weight 72.12 g/mol Density 0.837 g/cm 3 @ 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/m 3

III. Major Uses or Sources

1,2-Epoxybutane is used as a chemical intermed iate, 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 (Repro text, 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 EpoxybutaneDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

V. Effects of Animal Exposure

F344/N rats (50/sex) were e xposed 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 hyperplasi a, 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 concentr ation-related increases in inci dence of chronic inflammation, epithelial hyperplasia, and erosi on of the nasal cavity were no ted in both sexes at either concentration. Increases in granulocytic hyperp lasia and splenic hematopoiesis were noted at

both concentrations in female mice.

Number of mice with lesions in the nasal ca vity and olfactory sensor y epithelium (NTP, 1988)

Sex Males Females

EBU concentration 0 ppm 50 pp m 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/m 3 ) EBU for 6 hours/day, 5 days/week, for 13 weeks were listless after the first exposure (N TP, 1988). Animals from this group all died by the end of the 13-week exposure. Renal tubular necrosis, and t hymic and splenic atrophy were seen in mice exposed to 800 ppm; decreased liv er weights were observed following exposure of mice to 400 ppm (1180 mg/m 3 ) or more. Inflammation of the nasal turbinates was seen in female mice exposed to 100 ppm (295 mg/m 3 ) 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/m 3 ) EBU 6 hours/day, 5 days/week, for 13 week s. In this study, no treatment-related effects were noted except for histological lesi ons 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 EpoxybutaneDetermination of Noncancer Chronic Re ference 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/m 3 ) 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 toxi city in the form of 10% weight lo ss was observed in rats exposed to 1000 ppm. One death out of 42 occurred in th e dams exposed to 1000 ppm. No adverse histological, reproductive, or deve lopmental effects were seen at any concentration. Exposure of rabbits on days 1-24 of gestation to the same c oncentrations 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, a nd 14 out of 24 died at 1000 ppm. Extensive maternal mortality in this study prevented ev aluation 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 (g as with extrathoracic respiratory effects, RGDR = 0.20, based on

MVa = 0.06 m 3 /day, MVh = 20 m 3 /day, SAa(ET) = 3.0 cm 2 , SAh(ET) = 200 cm 2 ) 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/m 3 ; 20 μg/m 3 )

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 EpoxybutaneDetermination of Noncancer Chronic Re ference 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 conc entrations and with adequate histopathological analysis . Major areas of uncertainty ar e 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. Calif ornia 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version). Denver, CO: Microm edex, 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):319-3 24. [as cited in: Integrated Ri sk Information System (IRIS). U.S. Environmental Protection Agency, Wash ington, DC. (CD-ROM version). Denver, CO: Micromedex, Inc. (Edition expires 10/31/94).]

NTP. 1988. National Toxicology Program. Techni cal 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 Servi ces, Public Health Servi ce, National Institutes of Health. Research Triangle Park, NC: NTP.

REPROTOX. 1994. The REPROTOX(R) System: Repr oductive 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, styr ene 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-epoxy butane. 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 Bi ochemical Research Department. EPA/OTS Document No. 878211232. [as cited in: Integrated Risk Information System (IRIS). U.S. Environmental Protection Agency, Washingt on, D.C. (CD ROM version). Denver, CO: Micromedex, Inc. (Edition expires 10/31/94]

Appendix D3 203 EpoxybutaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# ETHYL CHLORIDE

(Chloroethane; monochloroet hane; ether hydrochloric)

CAS Registry Number: 75-00-3

I. Chronic Toxicity Summary

Inhalation reference exposure level 30,000 μg/m 3 (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 C 2 H 5 Cl Molecular weight 64.52 Density 0.9214 g/cm 3 @ 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/m 3 @ 25 ° C

III. Major Uses or Sources

Ethyl chloride has been used as a starting poin t in the production of tetr aethyl 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 statewid e industrial emissions fr om facilities reporting under the Air Toxics Hot Spots Act in Califor nia based on the most recent inventory were estimated to be 291,300 pounds of et hyl 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 sy mptoms 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 expos ure, all symptoms were absent.

Appendix D3 204 Ethyl chlorideDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

V. Effects of Animal Exposure

Pregnant mice were exposed to 1300, 4000, or 13000 mg/m 3 ethyl chloride in air for 6 hours per day on days 6-15 of ge station (Scortichini et al. , 1986). No effects on fetal resorption rates, litter size, body weight or maternal heal th were observed. A statistically significant increase in the incidence of delayed ossification of the s kull bones was observed in fetuses from the 13,000 mg/m 3 (4900 ppm) ethyl chloride exposed group. This skull effect was accompanied by a non-significant increased incidence of cervical ribs (a supern umerary rib is considered to be a malformation). No significant adverse effect s were observed in fetuses from the 4000 mg/m 3
(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 hi stopathologic examination (approxi mately 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 ch anges were observed in exposed

animals. Thus the subchronic NOAEL for mice and rats is 19,000 ppm, whic h is equivalent to a continuous exposure of 3400 ppm, and a free-st anding chronic NOAEL is 15,000 ppm, which is equivalent to a continuous exposure of 2700 ppm (7100 mg/m 3 ).

Increased relative liver weights and a slight increase in hepatocel lular vacuolation were observed in mice exposed to 5000 ppm ethyl chlori de 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 chambe r, two groups of 10 female mice were exposed to 0 or 15,000 ppm (40,000 mg/m 3 ) 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-expos ure period (5.0 days). The aut hors attribute this increase in estrous cycle length to a gene ral stress response a lthough they note that it does not preclude

direct effects on neuroendocrine function.

Appendix D3 205 Ethyl chlorideDetermination of Noncancer Chronic Re ference 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/m 3

NOAEL 4,000 mg/m 3

Exposure continuity 6 hours per day Exposure duration Days 6-15 of gestation Average experimental exposure 1,000 mg/m 3 for NOAEL group (4000 x 6/24) Human equivalent concentration 1,000 mg/m 3 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/m 3 ( 30,000 μg/m 3 ; 10 ppm; 10,000 ppb)

To develop the chronic REL OEHHA used the sa me study on which U.S. EPA based its RfC of 10,000 μg/m 3 . The REL is based on a developmental toxicity study. In ac cordance with U.S. EPA methodology, a time-weighted average concentration for the discontinuous exposure experiment is not used by U.S. EPA when the ke y 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 modifyi ng factors are not well

described. Such MFs we re 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 exposur e results in an adjusted NOAEL of 2679 ppm. Applying an RGDR of 1, a UF A of 3 and a UF H 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 conc entrations and with adequate histopathogical analysis, and the obser vation of a NOAEL. Major areas of uncertainty are the lack of adequate human exposure data, and the lack of a multig enerational reproductive study.

Appendix D3 206 Ethyl chlorideDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

VIII. References

Breslin JW, Berdasco NM, Phillips JE, and Johns on KA. 1988. Ethyl Chloride (EtCl): Effects on Estrous Cycling in B6C3F1 Mice. Final Report with cover le tter 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Micromedex, Inc., Denve r, 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 chlo ride sniffing and cerebellar dysfunction (case report). Isr. Ann. Psychiatr. Relat. Disci p. 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. Toxi cology 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 H ealth Service, National Institutes of Health. NTP Technical Report Number 346.

Scortichini BH, Johnson KA, Momany-Pfruende r 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 En vironmental Assessment, Environmental Criteria Assessment Office (CD-ROM Version).

Appendix D3 207 Ethyl chlorideDetermination of Noncancer Chronic Re ference 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/m 3 (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 C 8 H 10 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/cm 3 @ 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/m 3

III. Major Uses or Sources

Ethylbenzene is used as a precursor in the manufact ure of styrene (HSDB, 1994). It is also used in the production of synthetic rubbe r, 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 ethylben zene was approximately 0.4 ppb (CARB, 1999a).

The latest annual statewide emissions from f acilities 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 EthylbenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

and Cirek (1988) reported no significant hemato logical or liver function changes in 200 ethylbenzene production work ers over a 20-year period.

V. Effects of Animal Exposure

Rats and mice (10/sex/group) were e xposed to 0, 100, 250, 500, 750, and 1000 ppm (0, 434, 1086, 2171, 3257, and 4343 mg/m 3 ) ethylbenzene 6 hours/day, 5 days/week for 90 days (NTP, 1988; 1989; 1990). Rats displayed significantly lower serum alka line phosphatase in groups exposed to 500 ppm or higher. Dose-dependent in creases in liver weights were observed in male rats beginning at 250 ppm, while this effect wa s not seen until 500 ppm in the females. An increase in relative kidney we ights 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 a ny treatment-related effects except for elevated liver and kidney weights at 750 and 1000 ppm, respectively.

Rats and mice were exposed to ethylbenzene (gr eater than 99% pure) by inhalation for 2 years (NTP, 1999; Chan et al ., 1998). Groups of 50 male and 50 fema le 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 expos ed 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 signi ficantly greater than that in the chamber controls. The severity of nephropathy in 750 ppm male rats was significantly

increased relative to the chamber controls. So me increases in incidence and severity of nephropathy were noted in all exposed female rats , but these were statisti cally 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 eo sinophilic liver foci in 750 ppm females was significantly increased compared to that in th e chamber controls. There was a spectrum of nonneoplastic liver changes related to ethylbenzen e exposure in male mice, including syncytial alteration of hepatocytes, hepa tocellular hypertro phy, and hepatocyte necrosis. The incidences of hyperplasia of the pituitary gland pars dista lis in 250 and 750 ppm fema les 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. Ba sed on an evaluation of all the

non-cancer data in mice and rats OEHHA staf f 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/m 3 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 apparatu s” were observed at the 2400 mg/m 3 concentration.

Appendix D3 209 EthylbenzeneDetermination of Noncancer Chronic Re ference 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/m 3 ethylbenzene. There were no live fetuses in this group for which abnor malities could be evaluated. No developmental

defects were observed in the lower exposure groups.

Rats (78-107 per group) and rabbits (29-30 pe r group) were exposed for 6 or 7 hours/day, 7 days/week, during days 1-l9 a nd 1-24 of gestation, respectively, to 0, 100, or 1000 ppm (0, 434, or 4342 mg/m 3 ) ethylbenzene (Andrew et al. , 1981; Hardin et al., 1981). No effects were observed in the rabbits for maternal toxicity dur ing exposure or at time of necropsy. Similarly,

no effects were seen in the fetuses of the rabb its. 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 ra ts exposed to 1000 ppm et hylbenzene. Maternal rats exposed to 1000 ppm exhibited significantly increased liver, kidne y, and spleen weights compared with controls. Fetal rats showed an increase in skeletal variations at the 1000 ppm concentration, but the resu lts of the 100 ppm exposur e 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/m 3 ) ethylbenzene for 6 hours/day, 5 days/week, for 12 weeks.

Degeneration of the testicular epithelium was not ed in guinea pigs and a rhesus monkey exposed to 600 ppm (2604 mg/m 3 ) 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/m 3 ) 6 hours/day, 5 days/week for 4 week s. 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/m 3 ) 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 weig hts were not elevated in these studies.

Appendix D3 210 EthylbenzeneDetermination of Noncancer Chronic Re ference 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/m 3 ; 2,000 μg/m 3 )

The REL is based on a lifetime toxicity/car cinogenesis study. The NOAEL for non-neoplastic effects in the study was 75 ppm, and the LOAE L was 250 ppm. Some shorter duration studies discussed above (e.g. NTP, 1988, 1989, 1990) identif y higher concentrations as NOAELs, but the study used (NTP 1999) is the most recent ava ilable and is considered the most reliable for assessing chronic effects.

U.S. EPA based its RfC on developmental toxi city 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 it s methodology, U.S. EPA did not use a timeweighted average concentration for the discontin uous exposure experiment since the key effect was developmental toxicity. If OEHHA met hodology is followed (which includes the timeweighted averaging of the expos ure concentrations, and uncertain ty factors of 3 (interspecies, with RGDR = 1) and 10 (intraspecies), this study would indicate a REL of 0.6 ppm (3 mg/m 3 ). The study by Ungvary and Tatrai (1985) reported a NOAEL of 600 mg/m 3 for developmental and maternal effects in several species. Howeve r, the reporting and genera l 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 subchroni c UF of 3, an interspecies UF of 3, and an intraspecies UF of 10 which resu lts 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 EthylbenzeneDetermination of Noncancer Chronic Re ference 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 conc entrations and with adequate histopathogical analysis, and th e observation of a NOAEL in lifetime chronic inhalation exposure studies. Th e major area of uncertainty is the lack of adequate human exposure data.

VIII. References

Andrew FD, Buschbom RL, Cannon WC, M iller 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: Eff ects 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 st udy on workers occupationally exposed to ethylbenzene. J. Hyg. Epidemio l. 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. T oxics Air Quality Data. Substance Chooser. Ethyl Benzene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm

CARB. 1999b. Air toxics emissions data collect ed in the Air Toxics Hot Spots Program CEIDARS Database as of January 29, 1999.

Clark DG. 1983. Ethylbenzene hydroperoxide (E BHP) 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 ethyl benzene. Toxicol. Lett. 99(1):23-32.

Cragg ST, Clarke EA, Daly IW, Miller RR, Terrill JB, and Quellette RE. 1989. Subchronic inhalation toxicity of et hylbenzene 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 EthylbenzeneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Elovaara E, Engstrom K, Nickel s J, Aito A, and Vainio H. 1985. Biochemical and morphological effects of long-term inhalation exposure of rats to ethylbenzene. Xenobi otica. 15(4):299-308. [as cited in U.S.EPA’s Integrated Risk In formation System (IRIS) database. 1994.]

Hardin BD, Bond GP, Sikov MR, Andrew FD, Be liles RP, and Niemeier RW. 1981. Testing of selected workplace chemicals for teratogenic poten tial. Scand. J. Work Environ. Health. 7(suppl 4): 66-75.

HSDB. 1994. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (CD-ROM version) Denver, CO: Micromed ex, Inc. (Edition expires 4/30/94).

NTP. 1988. National Toxicology Program. Subc hronic and chronic toxicity study of ethylbenzene. 90-Day subchroni c study report on inhalation exposure of F344/N rats and B6C3F1 mice. Chicago, IL: IIT Research Institu te. [as cited in U.S.EPA’s Integrated Risk Information System (IRIS) database. 1994.]

NTP. 1989. National Toxicology Program. Chai rperson’s report. Pathology Working Group (PWG) review of subchronic toxi city testing on ethylbenzene admi nistered 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 Ri sk Information System (IRIS) database. 1994.]

NTP. 1999. National Toxicology Program. Toxi cology and Carcinogenesis Studies of Ethylbenzene (CAS No. 100-41-4) in F344/N Rats and B6C3F1 Mi ce (Inhalation Studies). TR466.

Reprotext ® database. (CD-ROM version) Denve r, 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, Hollings worth RL, and Oyen F. 1956. Toxicological studies of certain alkylated benzenes a nd benzene. Arch. Ind. Health 14: 387-398.

Appendix D3 213 EthylbenzeneDetermination 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 μ g/m 3 (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 C 2 H 4 Br 2 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 μ g/m 3 per ppb at 25 ° C

III. Major Uses and Sources

Ethylene dibromide (EDB) is used as a solvent fo r 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 ED B 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 dibromideDetermination 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 in dustrial emissions from facilities reporting under the Air Toxics Hot Spots Act in Califor nia 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-transfer ases (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 metabol ites 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 activ ity 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 respirat ory diseases was conducted among 161 employees exposed to EDB in 2 production units operate d 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 me n were divided into 2 groups depending on estimated exposure (Ter Haar, 1980). Twenty pe rcent 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 c ounts 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, Dobbi ns (1987) concluded that the results provide evidence that EDB exposure between 0.5 and 5.0 pp m is associated with lower sperm counts.

A comparison of observed marital fertility with exp ected 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 dibromideDetermination 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 exp ected level. After omitting the incidence of vasectomies and hysterectomies among married c ouples, observed fertility was still 39% below the expected figure for plant D but was now no l onger statistically signifi cant. Exposure levels of EDB at plant D were not known but were estima ted to be no more than 5 ppm. Later review determined that expected (contro l) 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 tr eatment 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 redu ced fertility among the combined workers at the four plants (Wong et al. , 1985). The fertility evaluation indicates that more in-depth epidemiologic or physiol ogic studies are needed.

Semen analysis of 83 pineapple workers at two plantations was performed by Rogers and associates (1981). EDB-exposed workers we re removed from each group and placed in a separate group. The remaining two groups of wo rkers 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 expo sed workers that had fertility tests done, all tested in the infertile range. Forty percent or less tested in the infert ile 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 pa paya 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 geomet ric mean breathing zone exposure to airborne EDB was 88 ppb (8 hr time wei ghted 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 confoundi ng factors, statistically significant decreases in sperm count/ejaculate, the percentage of viable and motile sperm, and in creases 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 sl ower sperm velocity (Schrader et al. , 1988). Six unexposed men were used as controls. The researchers sugge st 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 dibromideDetermination of Noncancer Chronic Reference Exposure Levels December 2001

V. Effects of Animal Exposures

EDB is readily and rapidly absorbed from the l ung when breathed as a vapor, from the GI tract when taken orally, or through the skin when applie d dermally (HSDB, 1995). In rats, the rate of absorption of EDB from the respiratory tract r eached 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. Intraper itoneal injection of [ 14 C]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 [ 14 C]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. Th e conjugative pathway principally yields glutathione products, such as S -(2-bromoethyl)-glutathione, whic h are mainly responsible for DNA binding and mutagenesis. In rats, orally ad ministered EDB is excreted primarily in the urine as mercapturic acid derivatives (Jones and Edwards, 1968). The biologic half-life for elimination of [ 14 C]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 irre versibly 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 us ed, 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 fo r as many as 63 seven-hour exposures in 91 days resulted in no si gnificant 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 reveal ed 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 inte rstitial congestion an d edema with slight parenchymatous degeneration of th e tubular epithelium were observe d. Four rabbits exposed to 59 seven-hour sessions at 50 ppm in 84 days showed no signs of adverse eff ects. Clinical signs of monkeys exposed to 50 pp m EDB (49 seven-hour exposures in 70 days) included an ill, unkempt appearance and nervousness. Slight ce ntral fatty degeneration in livers was observed, but pathology was not seen in other tissues. E xposure of the same four species to 25 ppm EDB for up to 220 days (145 to 156 seven-hour exposur es) 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 se vere necrosis and atrophy of the olfactory epithelium in the nasal cavit y. Squamous metaplasia, hyperp lasia and cytomegaly of the

Appendix D3 217 Ethylene dibromideDetermination 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 olf actory and respiratory epithelium at 3 ppm.

No lesions were seen in other tissues at any dose.

In another 13-week inhalation st udy, 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 weig ht 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 ur ine in females of the 40 ppm group, no treatment-related changes were observed in any rat group with respect to ur inalysis, 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 rela tive liver weights of females in the two highest exposure groups were significantly elevated. Absolute liver we ight of females in the 40 ppm group was also significantly elevated. Histopathol ogical examination revealed lesions primarily confined to the

anterior sections of the nasal turbinates. Hyperplasia and nonke ratinizing squamous metaplasia of the respiratory epithelium were observed in nasal turbinates of rats exposed to 40 ppm EDB. Only slight epithelial hyperplasia of nasal turbin ates was noted at 10 ppm. No treatment related effects were seen at 3 ppm. Livers of female s in the 40 ppm group showed a slight increase in fat. After an 88 day recovery period, there was a reversion to normal of th e 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 we re 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 intubatio ns of low and high dose rats ceased for 1 week followed by 4 weeks of dose administration. All surviving, treated male ra ts 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 fe males, 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 discontinue d at week 54 with necropsy o ccurring 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 do se over the test period was 62 and 107 mg/kg body wt-day for low and high dose mice, respectivel y. 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-depe ndent mortality. Many of the deaths occurred during or shortl y after intubation, sugge sting an acute toxic reaction. Pathology revealed hype rkeratosis 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 degenerati on and peliosis of the liver (hep atitis). Dosed males showed

Appendix D3 218 Ethylene dibromideDetermination 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 includ ed 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 hyperker atosis (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 B6C3F 1 mice were exposed to 0, 10, or 40 ppm EDB 6 hr/day, 5 days/week for up to 103 w eeks (NTP, 1982). In male and female rats, the high dose groups had reduced body weights and in creased mortality that began at about week 60. The treatment-related non-neoplastic pathol ogy included hepatic necrosis (both sexes), epithelial hyperplasia and suppur ative inflammation throughout th e respiratory system (both sexes), and nephropathy (males onl y). Toxic nephropathy and mine ralization 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. Dege neration 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 weakne ss of the limbs or body during the second year. Increased mortality occurred in a dose-rela ted manner in females and was si gnificantly greater in low dose males. Non-neoplastic pathol ogy included epith elial hyperplasia through out 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 urin ary bladder inflammation. However, bladder epithelial hyperplasia was only seen in exposed animals. An in creased 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% di fference from controls) occurred by the 15 th
month in males, and by the 18 th month in females. Significantly reduced food consumption was not apparent. Increased mortality rates in both sexes occurred beginning in the 12 th 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 producti on in bulls (Isreal-Frie sian 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 fo r 2-4 months longer. EDB did not appear to affect the growth, health, and libido of the bu lls. However, semen density and motility were significantly lower compared to untreated cont rol bulls of the same age. Many abnormal spermatozoa were also present in treated bulls. A NOAEL for this eff ect was apparently not

determined. Cessation of EDB administration re sulted in normal sperm within 10 days to 3

months. Further studies confir med that EDB adversely affect ed sperm production without any other apparent effects on bulls (Amir and Vol cani, 1967; Amir and Ben-David, 1973). However,

Appendix D3 219 Ethylene dibromideDetermination of Noncancer Chronic Reference Exposure Levels December 2001

feeding rams 2-5 mg/kg body wt-day for 120 days di d 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 hema tocrit levels, and increased cholesterol, triglycerides, total protein, a nd albumin. The highest dose also caused an immunosuppressive effect by lowering the in vitro splenic lymphocyte res ponse 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 mort ality 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 occu rred at 38 ppm in rats and at all exposure

levels in mice. No anomalies were seen in ra t fetuses. An increase in runts at 38 ppm and a dose-dependent increase in skelet al anomalies were observed am ong mouse fetuses. However,

these anomalies were characterist ic of delayed development and o ccurred at doses that adversely

affected maternal welfare. Ther efore, these effects are indicative of fetal toxicity rather than teratogenicity.

Male reproductive toxicity of ED B has been evaluated in some ot her 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 foll owing a single intraper itoneal 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 dibromideDetermination 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 brea thing 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 m 3 /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/m 3 , 0.8 μ g/m 3 )

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 corr elation between EDB exposure a nd 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 occu r 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 st udies 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 dibromideDetermination 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 a ffecting the general h ealth 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 useful ness of this study. Long-term studies of EDB toxicity in other experimental animals also lack the de termination 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 LOAE L (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 μ g/m 3 ).

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for ethylene dibromide incl ude the use of human exposure data from workers exposed over a period of year s, 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 occupa tional exposure concentration, and the limited nature of the study (fertility was not actually tested). The databa se 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 obser ved when pregnant rats and mi ce were exposed to 20 ppm EDB during gestation (Short et al. , 1978). Thus the REL of 0.1 ppb s hould adequately protect infants and children. However, we do not know if adoles cent 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 toxi city of EDB. Both oxidative and conjugated metabolites are toxic. Infants may produce proportionately more conjugate than oxidized

metabolite relative to adults.

IX. References

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Amir D. 1991. The spermicidal effect of ethylen e dibromide in bulls and rams. Molec. Repro. Develop. 28:99-109.

Appendix D3 222 Ethylene dibromideDetermination of Noncancer Chronic Reference Exposure Levels December 2001

Amir D, and Ben-David E. 1973. The pattern of st ructural changes induced in bull spermatozoa by oral or injected et hylene dibromide (EDB). Ann. Biol . Anim. Biochem. Biophys. 13(2):165-

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Amir D, and Volcani R. 1965. Effect of diet ary ethylene dibromide on bull semen. Nature 206(4979):99-100.

Amir D, and Volcani R. 1967. The effect of diet ary ethylene dibromide (EDB) on the testes of

bulls. Fertil. Steril

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Barnett LB, Lovell DP, Felton CF, Gibson BJ, C obb 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

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FL: CRC Press Inc.

Dobbins JG. 1987. Regulation and the use of “n egative” results from human reproductive studies: The case of ethylene dibr omide. Am. J. Ind. Med. 12:33-45.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 11/31/95).

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fetal liver

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F344 rats and B6C3F 1 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 pr oduction units. Br. J. Ind. Med. 37:163-168.

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Plotnick HB, and Conner WL. 1976. Tissue distribution of 14 C-labeled ethylene dibromide in the guinea pig

. Res. Commun. Chem. Path. Pharmacol. 13(2):251-258.

Plotnick HB, Weigel WW, Richar ds DE, and Cheever KL. 1979. The effect of dietary disulfiram upon the tissue distribu tion and excretion of 14 C-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 modul ation by the estrus cy cle. In Vivo 9:299-304.

Ratcliff JM, Schrader SM, Steen land K, Clapp DE, Turner T, and Hornung RW. 1987. Semen quality in papaya workers with long term expos ure to ethylene dibromide. Br. J. Ind. Med.

44:317-326.

RECT. 1988. Reviews of Environmental Contam ination and Toxicology. Ethylene dibromide.

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REPROTOX. 1995. The REPROTOX(R) System: Repr oductive 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. Re spiratory 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 dibr omide (EDB). J. Androl. 2:35-36.

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. Arch. Ind. Hyg. Occup. Med. 6(2):158-

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Schrader SM, Ratcliff JM, Turner TW, and Hor nung RW. 1987. The use of new field methods of semen analysis in the study of occupational h azards 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 ch ronic exposure. Repro. Toxicol. 2:191-198.

Short RD, Minor JL, Winston JM, Seifter J, a nd Lee C. 1978. Inhalation of ethylene dibromide during gestation by rats and mice. T oxicol. Appl. Pharmacol. 46:173-182.

Stott WT, and McKenna MJ. 1984. The comparativ e absorption and excr etion of chemical vapors in the upper, lower, and intact respirat ory tract of rats. Funda m. Appl. Toxicol. 4:594-

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Ter Haar G. 1980. An investigation of possible st erility and health eff ects from exposure to ethylene dibromide. In: Banbury Report 5-Ethylene Dibromide: A Potential Health Risk? Ames B, Infante P, and Reitz R. (eds). Cold Spri ng 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 hu man 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. Toxi col. 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 surv eillance program for evaluating occupational reproductive hazards. Am. J. Ind. Med. 7:295-306.

Appendix D3 225 Ethylene dibromideDetermination of Noncancer Chronic Re ference 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/m 3 (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 C 2 H 4 Cl 2 Molecular weight 98.97 g/mol Density 1.2351 g/cm 3 @ 20oC Boiling point 57.4oC Melting point

− 96.9oC Vapor pressure 64 torr @ 20oC Solubility Slightly soluble in wa ter (0.869 g/100 ml at 20oC); miscible with alcohol; soluble in ordinary organic solvents Conversion factor 1 ppm = 4.05 mg/m 3

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 Californ ia, 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 tran sition 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 unde r the Air Toxics Hot Spots Act in California based on the most recent inventory were estimated to be 24,935 pounds of ethylene dichlori de (CARB, 2000).

Appendix D3 226 Ethylene dichlorideDetermination of Noncancer Chronic Re ference 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 ch anges 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 work ers 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 year s 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 di ed 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 ac ute centrilobular necrosis was observed in the kidney. Nerve cells in the brai n, including Purkinje cell s, appeared shrunken with pyknotic nuclei. The latency period for he patotoxicity of approximately 20 hours suggests that metabolism of the compound yields the reactive agen t (see below).

V. Effects of Animal Exposure

As with humans, the absorption and distribution of EDC in rats following i ngestion or inhalation is rapid and complete (IARC, 1999). Metabolism in rats and mice is exte nsive 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 di rect conjugation with glutathione catalyzed by glutathione S-transfer ase may ultimately result in the putative alkyating agent (episulfonium ion) primarily responsible for toxicity and carcinoge nicity. Evidence for DNA-damaging metabolites resultin g via the P450 pathway exists (IARC, 1999). However, this pathway appears to be a minor r oute for toxic metabolite formation.

Acute exposure in mice resulted in toxic effect s 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 e xposed 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 c ontrols. Daily observations, gr oss pathology, and extensive histopathology revealed no differences from controls other than a slight increase in unspecified testicular lesions in the EDC group. Additiona l 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 dichlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

(Antabuse) for alcoholism) in the diet. Disulfir am treatment resulted in increased number of tumors, increased blood levels of EDC, and increa sed 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 le vels 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 ag e at the beginning of the experiment, and rats to be examined after 12 months of exposure we re 14 months of age at the beginning of the experiment. Complete histological exams were conducted but non-can cer effects were not

discussed. No consistent treatment-related ch anges 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) leve ls at the two highest exposures. Lactate dehydrogenase (LDH) and aspartate aminotransfera se (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 toxic ity 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 si gnificantly lower than that seen in rats tested at other times in this study. Thus, the toxicologi cal significance is questio nable. Cholesterol was reduced significantly at the hi gher exposure levels but the toxi cological significance of this finding was unknown. The marked difference between serum chemistry parameters following 12 months of exposure, compared to those fo llowing 3, 6, and 18 months of exposure, may be

due to the considerable differen ce in the age of the rats at th e 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-dichlo roethane 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 wei ghts and increased 5-nucleotid ase (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 D SF 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 therap eutically is likely to be more susceptible to the effects of EDC toxicity.

Rats, rabbits, guinea pigs, dogs, cats, and monke ys 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, ra bbits, and guinea pigs following exposure 7 hours per day, 5 days per w eek for two weeks or less. At 963 ppm guinea pigs exhibited lacrimation and inactivity duri ng exposure; pulmonary congestion was noted at autopsy. Rats exposed to this concentration exhi bited degenerative proliferative changes in the renal tubular epithelium and sple nitis. 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 dichlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

sacrifice 11 weeks following termination of expo sure, congestion and fatty infiltration of the liver were observed at necropsy. Due to high mort ality in the rodents at the higher concentration, a subsequent experiment exposed rats and gui nea pigs 7 hours per day, 5 days per week to 100 ppm EDC for four months. No increase in mo rtality 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 ob served upon gross and hist ological examinations of 10/39 exposed and 10 control ra ts. This study is severely limited by the methods used to determine the exposure concentration and by the l ack of quantitative measur ements of toxicity other than death. This study does, however, indicate that fatty in filtration 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/m 3 (equivalent to 0, 6, and 51 ppm) EDC for 6 hr/day from two weeks before mating and throughout gestation. St atistically significant increases in pre-implantation loss and decreased male pup weights were observed at the highest dose. Gross sk eletal 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 obser ved at the highest expos ure. Statistically significant evidence of altered growth and tera togenic effects were not observed at any

concentration.

Rao et al . (1980) exposed rats and rabb its 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. Othe rwise, no fetotoxic or teratogenic effects were observed in either species. In a reproducti on study, rats were expos ed 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 fr om gestational day 21 through day 4 postpartum. EDC had no effect on reproduction over one ge neration 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 (equ ivalent to about 5, 15, or 50 mg/kg bw/day) starting five weeks before mating of the F 0 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 histopatholog ical indications of hepato- or neph rotoxicity were observed in OsbornMendel rats (47 or 95 mg/kg bw/day, 5 days/wee k 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 dichlorideDetermination of Noncancer Chronic Re ference 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 seve ral species of rats for 13 weeks. A statistically significant increase in kidney wei ght 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 ad ministered 8000 ppm. Significant decreases in absolute and relative kidney weight were obser ved in male and female rats administered concentrations of 1000 ppm EDC. A significant incr ease in relative liver weight was observed in male rats administered 2000 ppm EDC and great er 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 mort ality in the higher dose groups. Statistically significant increases in kidney we ights were observed in surviving male rats administered EDC

and in female rats administered 120 or 240 mg/kg. However, no hist ological 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-bo dy 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/m 3 ; 400 μ g/m 3 )

Cheever et al . (1990) and Spreafico et al. (1980) were the only chronic inhalation exposure

studies found in the literatur e that presented non-cancer e ffects. No reproductive and developmental effects were observed in studies published in peer-revie wed 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 dichlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

The Agency for Toxic Substances and Disease Re gistry (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 determin ed. To derive the MRL, the ATSDR applied uncertainty factors (UFs) of 10 each for intraspeci es 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 sp ecified by ATSDR. Such modify ing factors were not used by OEHHA. A continuity correction for disconti nuous exposure was not applied. The resulting MRL was 0.2 ppm (0.7 mg/m 3 ).

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 in terspecies and 10 for in traspecies differences result in a REL of 0.5 ppm (2 mg/m 3 ).

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation R EL for ethylene dichlo ride include the ava ilability of chronic inhalation exposure data, the relati vely large number of exposure le vels at lower concentrations (allowing for better elucidation of the dose-res ponse relationship for he patotoxicity), 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 re latively modest clinical chemistry findings observed in the Spreafico et al. (1980) study may have resulted in false-positives, falsenegatives, and lack of clear dos e-response relationships. Repea ting the study in one or more experimental animal species with full histopat hological examination of organs and a greater

number of animals/dose would si gnificantly 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 Hu man 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. Ethyl ene Dichloride. In: Toxic Air Contaminant

Identification List Summaries. California Environmental Protection Agency, pp. 483-487.

Appendix D3 231 Ethylene dichlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CARB. 2000. California Air Resources Board. Calif ornia 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 influe nce of disulfiram or ethanol on oncogenicity, metabolism, and DNA covalent binding in rats. F undam. Appl. Toxicol. 14:243-261.

Francovitch RJ, Schor NA, and George WJ. 1986. Effects of SKF-525A, phenobarbitol, and 3- methylcholanthrene on ethylene dich loride toxicity following inha lation 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 ti ssues. Toxicol. Appl.

Pharmacol. 98:87-99.

HSDB. 2000. Hazardous Substances Data Bank. Nati onal Library of Medicine, Bethesda, MD. Available online through Toxicology Da ta 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 dichlori de). V. The effects of daily i nhalations. J. Ind. Hyg. Toxicol. 28(4):113-120.

IARC. 1999. International Agency for Research on Cancer. Re-evalua tion 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. In teraction between 1,2-dichloroethane and disulfiram. I. Toxicological effect s. Fundam. Appl. Toxicol. 6:733-746.

Igwe OJ, Que Hee SS, and Wagner WD. 1986b. In teraction between 1,2-dichloroethane and tetraethylthiuram disulfide (disulfiram). II. Hepa totoxic manifestations with possible mechanism of action. Toxicol. Appl. Pharmacol. 86:286-297.

Kozik I. 1957. Problems of occupa tional hygiene in the use of dichloroetha ne 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-dic hloroethane and 1,1,1- trichloroethane in drinking water on reproducti on 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 dichlorideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

NCI. 1978. United States National Cancer Institute . Bioassay of 1,2-dichlo roethane for possible carcinogenicity CAS No. 107-06-2 (N CI-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 pl acental transfer of 1,2-dichloro ethane in rats. Fundam. Appl.

Toxicol. 28:187-198.

Rao KS, Murray JS, Deacon MM, John JA, Ca lhoun LL, and Young JT. 1980. Teratogenicity and reproduction studies in animals inhaling ethy lene dichloride. In: Banbury Report 5. Ethylene Dichloride: A Potential Health Risk? Ames B, Infa nte 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 dich loride in rats treated by diffe rent 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 Spri ng 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. EP A, 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 dichlorideDetermination 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 μ g/m 3 (200 ppb) Critical effects Respiratory irritati on 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 C 2 H 6 O 2 Molecular weight 62.07 g/mol Density 1.1088-1.1135 g/cm 3 @ 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 etha nol; slightly soluble in ether. Insoluble in benzene and petroleum

ether. Conversion factor 1 ppm = 2.5 mg/m 3 @ 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 ingredie nt in electrolytic condensers; as a solvent in the paint and plastics industries; and in inks for ball-point pens and prin ter’s inks. It is used in the manufacture of some synthetic fibe rs (Terylene and Dacron), and in s ynthetic waxes. It is used in some skin lotions and flavoring essences. Als o, it is used in asphalt emulsion plants, in wood stains and adhesives, and in le ather 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 GlycolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

IV. Effects of Human Exposure

Laitinen et al. (1995) found that 10 motor servicing wo rkers had significantly higher urinary levels of ethylene glycol and ammonia, a nd decreased urinary glyc osaminoglycan levels, compared with 10 controls. The ethylene glycol levels in air were undete ctable 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 absorp tion rate is high, air borne 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 vary ing up to a mean of 20 ppm (49 mg/m 3 ) for 30 days was without effect (Wills et al. , 1974). The actual concentrations measured in the exposure chamber

were:

Concentration of ethylen e glycol in air, mg/m 3
Days Low High a 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 exposur e concentration of 75 ppm (188 mg/m 3 ), and became quickly intolerable at 123 ppm (308 mg/m 3 ). 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 res ponses (including simple reaction time, weight discrimination, and depth perception). The respir atory irritation at 75 pp m 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 gr oup, including: reduced body we ight, increased water intake, increased blood urea nitrogen and creatin ine, 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 GlycolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

males. The NOAEL in rats for chronic oral et hylene glycol toxicity was 200 mg/kg/day. No

effects were observed in mice. Theref ore, 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/m 3 or continuously to 12 mg/m 3 for 90 days. At 10 mg/m 3 2 rabbits had conjunctivitis and liver changes were noted in a few anim als of the other species. At 57 mg/m 3 no signs of toxicity were seen during the exposure. Nonspe cific inflammatory changes were noted in some lungs and hearts of all species. A few livers also showed necrotic areas. Continuous exposure to 12 mg/m 3 led to moderate to severe eye irritation in rats and rabbits. Edema in the rabbits led to eye closure. Two rats devel oped corneal opacities. All hematologic parameters and various enzymes assayed were within normal limits. At necropsy organs appeared normal. Histopathological analysis reveal ed inflammatory changes in th e 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/m 3 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 inhale d ethylene glycol on reproduction and development of rats and mice were conducted by Tyl et al. (1995a, 1995b). In a study usi ng whole-body exposure of rats and mice to ethylene glycol at analyzed concentrations of 0, 119, 888, or 2090 mg/m 3 for 6 hours/day on days 6-15 of gestation, mice were f ound to be the more sensitive species. Maternal toxicity in rats included a si gnificant increase in absolute a nd relative liver weight at 2090 mg/m 3 . 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 a ddition, terata were not observed at any concentration. Reduced ossification in the hume rus, zygomatic arch, and the metatarsals and proximal phalanges of the hindl imb was present in fetuses exposed to 888 or 2090 mg/m 3 . The NOAEL for maternal toxi city in rats was 888 mg/m 3 , while the NOAEL for fetotoxicity was 119 mg/m 3

.

In mice, reduced body weight and gravid uterin e weight during and after the exposure were observed at the 888 and 2090 mg/m 3 concentrations. Increased nonviable implants per litter and reduced fetal body weights were also observed in groups exposed to 888 or 2090 mg/m 3

. External, visceral, skeletal, and total malformations were increased in the 888 and 2090 mg/m 3
groups. The NOAEL for these effects in mice was 119 mg/m 3

.

A similar experiment in mice using nose-only ex posures was conducted by these researchers (Tyl et al. , 1995a) to determine the role of dermal absorption and/or inge stion on the effects

Appendix D3 236 Ethylene GlycolDetermination 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/m 3

. The NOAEL for maternal effects (inc reased kidney weight) was 500 mg/m 3 , and the NOAEL for fetal toxicity (skeletal varia tions and fused ribs) was 1000 mg/m 3 . Thus, secondary dermal and/or oral exposures appear to have contributed significantly to the deve lopmental and maternal toxicity in mice exposed to ethy lene glycol aerosol. The nos e-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 be havior would have resulted in significantly larger doses of ethylene glycol than that e xpected by inhalation only.

A 3-generation study on the effects of ethylen e 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 repr oductive and developmental effects of ethylene glycol in rabbits exposed by gavage on days 6 to 19 of gesta tion. 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 fo r 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/m 3 ; 400 μ g/m 3 )

The subchronic study by Wills et al. (1974) represents the only human inhalation data for ethylene glycol toxicity. The experiment showed a concentra tion-response relationship, with onset of irritation occurring at 188 mg/m 3 and intense and intolerable irritation occurring at 308

Appendix D3 237 Ethylene GlycolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

mg/m 3 . The volunteers were followed for 6 weeks w ithout any apparent l ong-term effects from the exposures. Although the irrita tion 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 st udy in rats by DePass et al. (1986a) showed significant chronic effects including reduced body weight, increased wate r 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 e ffects 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 te rata and maternal toxicity endpoints (Tyl et al. , 1995a; Tyl et al. , 1993; Neeper-Bradley et al. , 1995). In addition, the 3-gene ration 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. Howe ver, developmental endpoints were not reported in this study. From the available data, the toxicity of ethylene glycol is ap parently greatest in the maternal mouse. The estimated equivalent ai r concentrations (assumi ng a 70 kg human inhales

20 m 3 /day) from the feed in the 3-generation study by DePass et al. (1986b) are 700 mg/m 3 and 3500 mg/m 3 for the NOAEL and LOAEL, respectively.

For comparison with the proposed REL of 400 μ g/m 3 based on a one month human study, the inhalation NOAEL of 48 ppm, obtained by Tyl et al . (1995) in mice disc ontinuously exposed for 10 days on gestation days 6-15, was used to estima te 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 μ g/m 3 ) for ethylene glycol.

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation R EL for ethylene glycol include the use of human exposure data, the use of controlled, nearly continuous inha lation 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 GlycolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

DePass LR, Garman RH, Woodside MD, Gidde ns 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, a nd Weil CS. 1986b. Three-generation reproduction and dominant lethal mutagenesis studies of ethylen e glycol in the rat. Fundam. Appl. Toxicol.

7:566-572.

HSDB. 1996. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (CD-ROM version) Denver, CO: Micromed ex, 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 propy lene 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 fo r 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, S eely JC, Heindel JJ, and Schwetz BA. 1993. Developmental toxicity evaluati on 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 GlycolDetermination of Noncancer Chronic Re ference 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/m 3 (20 ppb) Critical effect(s) Testicular degenera tion 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 C 4 H 10 O 2 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/m 3 per ppb at 25oC

III. Major Uses and Sources

Ethylene glycol monoethyl ether (E GEE) is a widely used solvent for nitrocellulose, dyes, inks, resins, lacquers, paints, and varn ishes (HSDB, 1996). It is al so 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 pr oduction of another solvent, ethylene glycol monoethyl ether acetate. The specific annual statewide industrial em issions of EGEE from facilities reporting under the Air Toxics Hot S pots Act in California based on the most recent inventory were estimated to be 443,748 pounds (C ARB, 1999). (Many industries did not report emissions of specific glycol ethers. T hus 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 signi ficant difference in mean sperm count was observed between the

Appendix D3 240 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

37 exposed male workers and 39 unexposed male workers. Semen volume and pH, viability, motility, velocity, and morphology were not signifi cantly 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 signif icantly lower sperm counts than

historical controls. Furtherm ore, members of both groups may have been exposed to other compounds including metals, so lvents, heat, and vibration.

Welch and Cullen (1988) evaluated shipyard painte rs exposed to ethylene glycol ethers (EGEE and EGME). Air concentrations at the wor kplace 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/m 3 with a mean of 9.9 mg/m 3 . TWA exposures to EGME ranged from 0 to 17.7 mg/m 3 (mean = 2.6 mg/m 3 ). The authors note that during the time period of measurement, painting activities were unusually low and prev ious 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 differen tial counts did not differ between exposed and control. However, th e authors found that th e 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 expos ed group make a strong conclusion unwarranted.

Welch et al . (1988) evaluated the semen samples from th e workers in the cohort from Welch and Cullen (1988). Sperm concentration, velocity, motility, morphology, morphometry, and viability were measured. Although not sta tistically 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 o ligospermic painters was larger than that in controls at p = 0.05. There was no difference be tween controls and exposed men who were smokers. The authors state that although mean va lues of sperm count di d not differ significantly between controls and exposed gr oups, biologically important diffe rences were seen when the proportion of men with oligospermia was exam ined. The proportion of painters with azoospermia was 5% with only 1% expected base d on other population surveys. The authors note that to create a dose-respon se model for an effect of glyc ol 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 expos ure area to the next, making exposure assessment particularly difficult in this cohort.

Cullen et al . (1992) conducted a histopathologic anal ysis of the bone marrow and circulating blood cells in the workers prev iously examined in Welch et al . (1988). The objects of the study included : 1) to exclude other causes for granulocytopenia an d 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 ri sk of hematologic damage in glycol ether exposed painters. Workers were grouped as follows: Group I consiste d of those painters that had anemia or granulocytopenia in the Welch and Cullen (1988) study; Group II c onsisted of exposed painters with normal hematology; Group III consisted of un exposed controls. A battery of hematologic

Appendix D3 241 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

and biochemical parameters were measured a nd a questionnaire was completed to determine occupational exposure status, hea lth status and drug and alcohol consumption. All hematologic parameters were normal in all groups. Tests of liver, renal, and thyroi d function were normal in all groups. Bone marrow histology showed no di fferences between groups. One biochemical parameter, pyruvate kinase activity, was lowe r in Group I than Groups II and III (p = 0.05). Depression of red cell pyruvate ki nase did not vary by race and was lower in every subject in Group I by more than one standard deviation. Lo w pyruvate kinase is th e most consistent red cell enzyme defect noted in ac quired 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 inhala tion for 6 hours/days, 5 days/week, for 13 weeks (Barbee et al. , 1984). Animals were physically exam ined weekly and, at the end of the study, hematology, clinical chemistry, and histop athological examination were performed. No histopathological changes in the respiratory tract were f ound. 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 male s, 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. Signif icant 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 presente d in summary form (Doe, 1984). Wistar rats (24/group) were exposed to ta rget 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, hema tocrit, and mean corpuscular volume. Significant increases in preimplant ation 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 intrauteri ne death was observed in this group. Minor skeletal defects, particularly de layed ossification, were widely obs erved 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 in creased 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). Du tch rabbits (24/group) were exposed to 0, 10, 50, or 175 ppm EGEE for 6 hours/day duri ng gestational days 6-18, with sacrifice occurring on gestational da y 29. There were no indications of mate rnal 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 sligh tly increased incidences of de fects in the lower dose groups alone, including extra ribs and partial ossifica tion of the vertebrae, were not considered

treatment-related.

Appendix D3 242 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Behavioral teratogenic effects were examined in pregnant Sprague-Daw ley rats (14 or 15/dose group) exposed to 0 or 100 ppm EGEE for 7 hours/day through gesta tional 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 ge stational length was increased in rats exposed on gestational days 14-20. Six tests (ascent, rotorod, open field, activity wheel, a voidance 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 neurom uscular ability) and incr eased latency in an open field test (a test of e xploratory activity) as compared to controls. The offspring of rats exposed during days 14-20 of gestation e xhibited 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 newbor ns and in pups 21 days of age. In newborns from both EGEE-exposed groups, total brain norep inephrine was decreased. In 21-day old pups of both groups, norepinephrine and dopamine levels in the cerebrum were increased. Serotonin

level was increased in th e cerebrum of the late exposure gro up only. The authors concluded that

there were behavioral and neur ochemical 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 onl y), 300, 600, 900, or 1200 ppm EGEE for 7 hours/day. Increased fetal and pup mort ality was observed in all groups exposed to

EGEE.

Behavioral and neurochemical effects on the offspr ing 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 exposi ng 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 skel etal defects and variants and

cardiovascular defects were incr eased in the mid-dose group. Ra bbits (29/group) were exposed to 0, 16, or 617 ppm EGEE for 4 hours/day. Mate rnal weight gain and food intake weredecreased in exposed animals. The incidence of fetal resorptions was increased in both the midand high-dose group animals. Major cardiovascula r defects and minor skeletal defects (extra ribs, delayed ossification) were significan tly increased in the mid-dose group. Andrew et al. (1981) also examined reproductive effects by e xposing female Wistar rats (37/group) to 1, 150, or 649 ppm EGEE 7 hours/day, 5 days/week for 3 w eeks before mating with untreated males. No significant effects were observed.

Appendix D3 243 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference 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 degenera tion 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/m 3 ) for the NOAEL group Human equivalent concentration 18.4 ppm (68 mg/m 3 ) 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/m 3 , 70 μg/m 3 )

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 referenc e concentration (R fC) of 0.2 mg/m 3 (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 de generation were found in rabbits exposed to EGEE at 403 ppm for 13 week s. Changes in hematological parameters including decreased hemoglobin, he matocrit, 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 informa tion relating the effect in rabbits relative to

humans.

For a comparison with the proposed REL of 60 ppb (200 μ g/m 3 ) based on testicular degeneration, a REL can be calculated from the LOAEL of 202 ppm obser ved 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 UF L of 10, a UF A of 3 and a

UF H of 10 results in an estimated REL of 200 ppb (700 μ g/m 3 ). Nelson et al . (1981) found a LOAEL of 100 ppm for neurobehavi oral developmental toxicity in rats exposed 7 hours per day on days 7 to 13 of gestation. The equivalent c ontinuous 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 μ g/m 3 ).

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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

exposure to other compounds, make the study inade quate 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, us e of a mean concentration between the 2 buildings of 8 ppm for workplace exposure, extr apolation to an equiva lent continuous exposure of 3 ppm, and division by 10 for a LOAEL (seri ous effect) and 10 for intraspecies variability result in a REL of 30 ppb (100 μ g/m 3 ).

Another comparison with the proposed REL of 60 ppb can be made using the study of Welch et al . (1988), who studied shipyard painters e xposed 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 conditi ons, personal habits, a nd reproductive history; underwent a physical examination; and provided a se men 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/m 3 (mean = 9.9 mg/m 3 , and to EGME at a TWA concentration varying from 0 to 17.7 mg/m 3 (mean = 2.6 mg/m 3 ). The painters had an increased prevalence of oligospermia and azoospermia and an increased odds ratio for a lower sperm count per ejaculate. (The results we re controlled for smoking.) Adding the mean levels together results in a total glycol et her concentration of 12.5 mg/m 3 , which is equivalent to a continuous exposure of 4.5 mg/m 3 . Division by a UF of 10 for a LOAEL and by another of 10 for intraspecies variability results in a REL of 40 μ g/m 3 (10 ppb). A similar REL would be calculated using the report by Cullen et a l. (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 EG EE 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 in clude the availability of subchronic inhalation exposure data from a well-conducted study with hi stopathological analys is, 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, pr ovide support for the REL developed from animals. In addition, several comparative calculations indicate that RELs based on ot her studies are generally in agreement with that based on Barbee et al . (1984). Major ar eas of uncertainty are the lack of adequate human exposure data and the lack of chronic inhalation exposure studies.

Appendix D3 245 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

VIII. References

Andrew FD, Buschbom RL, Cannon WC, et al. 1981. Teratologic assess ment of ethylbenzene and 2-ethoxyethanol. Prepared for the National Institute for Occupational Safety and Health under Contract 210-79-0037, Battelle Pacific No rthwest Laboratories, Richland WA 99352.

Microfiche #OTS0513150.

Barbee SJ, Terrill JB, DeSousa DJ, and Conawa y 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, Du ffy TP, McPhedran P, Kelsey KT, Redlich CA. 1992. Morphologic, biochemical, and cytogene tic 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 monoe thyl ether and ethylene glyc ol 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: Micr omedex, 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 Se tzer JV. 1982a. Prenatal intera ctions between ethanol and the industrial solvent 2-ethoxyethanol in rats: maternal and beha vioral teratogenic effects.

Neurobehav. Toxicol. Teratol. 4:387-94.

Nelson BK, Brightwell WS, Se tzer JV, and O’Donohue TL. 1982b. Prenatal interactions between ethanol and the industrial solvent 2-ethoxye thanol in rats: Neurochemical effects in the offspring. Neurobehav. Toxi col. 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, Halp erin 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Tinston DJ, Doe JE, Godley MJ, Head LK, Killi ck M, Litchfield MH, and Wickramaratne GA. 1983a. Ethylene glycol monoethyl ether (EE): tera togenicity 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 teratogeni city 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 chr onic 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 et hylene 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 painte rs: II. Male reproduction. Am. J. Ind. Med. 14(5):509-526.

Appendix D3 247 Ethylene Glycol Monoethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# ETHYLENE GLYCOL MONOETHYL ETHER
ACETATE

(EGEEA; 1-acetoxy-2-ethoxyethane; 2- ethoxyethanol acetate; 2-ethoxye thyl acetate; acetic acid, 2-ethoxyethyl ester; beta-etho xyethyl acetate; Cellosolve ® acetate; ethoxy acetate; ethyl

Cellosolve ® acetate; Poly-solv ® EE acetate; ethyl glycol a cetate; oxitol acetate)

CAS Registry Number: 111-15-9

I. Chronic Toxicity Summary

Inhalation reference exposure level 300 μg/m 3 (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 C 6 H 12 O 3 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/m 3 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 evaporati on and imparts a high gloss (HSDB, 1996). It is also used as a

solvent for nitrocellulo se, oils, and resins and as a com ponent 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 repor ting 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 AcetateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

No studies relating exposure to EGEEA to advers e 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/m 3 and five to 28 mg/m 3 for 4 hours (Groeseneken et al. , 1987a). Twenty-two percent of the absorbed dose was eliminated in the urine as ethoxyacetic aci d 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 ethylen e glycol monoethyl ether (EGEE) (Groeseneken et al. , 1987b).

V. Effects of Animal Exposure

Pregnant rabbits (24 or 25/group) were e xposed 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. Mate rnal 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 re duction 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 obser ved in the 400 ppm dose group, with increased major malformations of the vert ebral column. Both 400 and 1 00 ppm EGEEA were found to be fetotoxic as indicated by retard ed ossification. No statisticall y 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 path ology revealed no toxic effects among treated animals. Among male rats and rabbits, “discr ete lesions of tubular nephritis with clear degeneration of the epithelium w ith hyaline and granular tubular casts” were observed. Four hour exposure to 2000 ppm EGEEA resulted in transient hem oglobinuria and hematuria in rabbits (2/sex/dose), but not rats (10/sex/dose). No patholog ical lesions were observed following a 2 week observation period.

Dogs were exposed to 600 ppm EGEEA fo r 7 hours/day for 120 days (Carpenter et al. , 1956; Gingell et al. , 1982). Hematological, clinical chemis try, 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 duri ng gestational days 6-15 and sacrificed on gestational day 21 (Union Carbide Corporation, 1984). Maternal eff ects in rats included increased absolute liver weights (all treated groups); increased relative liv er 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 a nd 300 ppm groups). An increase in the number of non-viable implanta tions per litter was observed at 300 ppm and

Appendix D3 249 Ethylene Glycol Monoethyl Ether AcetateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

decreased average fetal body wei ght per litter was observed at 200 and 300 ppm EGEEA.

Visceral and skeletal malfor mations were widely observed at both 200 and 300 ppm EGEEA. Among rabbits, maternal effects included decreas ed 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 li tter (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 malf ormations 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 da y 20. Complete resorption of litters was observed at 600 ppm. Skelet al and cardiovascular defects and decreased fetal weight and fetal resorptions were observed at 390 ppm EGEEA. Reduced fe tal 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/m 3 , 300 μg/m 3 )

A review of the literatu re on the toxicity of EGEEA indicates that the most sensitive endpoint of toxicity is that seen in expe rimental 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 re ference exposure level. Separate studies in animals have demonstrated developmental toxicit y. Reduced fetal weights were observed in rats exposed to 130 ppm EGEEA on ge stational days 7-15 (Nelson et al. , 1984). Skeletal and

Appendix D3 250 Ethylene Glycol Monoethyl Ether AcetateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

cardiovascular defects were observed at the ne xt 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). Fe totoxicity, 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 repor ted by Union Carbide Corporation (1984) and

Tinston et al. (1983), with 100 ppm EGEEA showing develo pmental 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 longe r 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 EG EE 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 monoe thyl ether and ethylene glyc ol 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 ethe r 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. Envir on. Health Perspect. 57:69-74.

HSDB. 1996. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, Maryland (CD-ROM Version). Denver, CO: Micr omedex, Inc. (Edition expires 7/31/96).

Appendix D3 251 Ethylene Glycol Monoethyl Ether AcetateDetermination of Noncancer Chronic Re ference 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 Toxi cology Laboratory, Alderl ey 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 Res earch Center. Terato logic evaluation of

Cellosolve acetate in Fischer 344 rats and Ne w Zealand white rabbits following inhalation exposure. EPA Doc. No. FYI-AX-1184-0360 (Fiche No. 0000360-0).

Appendix D3 252 Ethylene Glycol Monoethyl Ether AcetateDetermination of Noncancer Chronic Re ference 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/m 3 ( 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 C 3 H 8 O 2 Molecular weight 76.09 Density 0.965 g/cm 3 @ 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/m 3 @ 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 solv ent in the semiconductor industry. It is also used in dyeing leather and in the manufacture of photographic film. EG ME 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 mo st recent inventory were estimated to be 7398 pounds (CARB, 1999). (Many industries did not re port 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 glyc ol monomethyl ether have been associated with hematological and neurological abnormalities. To determin e whether employees potentially exposed to ethylene glycol monomethyl et her 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 EtherDetermination of Noncancer Chronic Re ference 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 stat istically significant (Cook et al. , 1982).

Cullen et al . (1983) studied possible bone marrow toxi city of workplace substances including dipropylene glycol monomethyl ether, EGME, and various aliphatic, aromatic and halogenated hydrocarbons used for offset and ultraviolet cure d multicolor printing. Evaluation of seven coworkers of a printer with apla stic 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 possibl e bone marrow toxicity resulting from exposure to glycol ethers and ultraviolet curing printing pro cesses was warranted. This was done to some extent in their studies on shipyard painters below.

Welch and Cullen (1988) evaluated shipyard painte rs exposed to ethylene glycol ethers (EGEE and EGME). Air concentrations at the wor kplace 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 th at during the time period of measurement, painting activities were unusually low and prev ious 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 differen tial counts did not differ between exposed and control. However, th e authors found that th e 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 expos ed group makes a strong conclusion unwarranted.

Welch et al . (1988) evaluated the semen samples from th e workers in the cohort from Welch and Cullen (1988). Sperm concentration, velocity, motility, morphology, morphometry, and viability were measured. Although not sta tistically 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 o ligospermic painters was larger than that in controls at p = 0.05. There was no difference be tween controls and exposed men who were smokers. The authors state that although mean va lues of sperm count di d not differ significantly between controls and exposed gr oups, biologically important diffe rences were seen when the proportion of men with oligospermia was exam ined. The proportion of painters with azoospermia was 5% with only 1% expected base d on other population surveys. The authors note that to create a dose-respon se model for an effect of glyc ol 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 expos ure area to the next, making exposure assessment particularly difficult in this cohort.

Appendix D3 254 Ethylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Cullen et al . (1992) conducted a histopathologic anal ysis of the bone marrow and circulating blood cells in the workers prev iously examined in Welch et al . (1988). The objects of the study included : 1) to exclude other causes for granulocytopenia an d 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 ri sk of hematologic damage in glycol ether exposed painters. Workers were grouped as follows: Group I consiste d of those painters that had anemia or granulocytopenia in the Welch and Cullen (1988) study; Group II c onsisted of exposed painters with normal hematology; Group III consisted of un exposed controls. A battery of hematologic and biochemical parameters were measured a nd a questionnaire was completed to determine occupational exposure status, hea lth status and drug and alcohol consumption. All hematologic parameters were normal in all groups. Tests of liver, renal, and thyroi d function were normal in all groups. Bone marrow histology showed no di fferences between groups. One biochemical parameter, pyruvate kinase activity, was lowe r in Group I than Groups II and III (p = 0.05). Depression of red cell pyruvate ki nase did not vary by race and was lower in every subject in Group I by more than one standard deviation. Lo w pyruvate kinase is th e most consistent red cell enzyme defect noted in ac quired hematologic disorders.

Reversible neurological sympto ms (apathy, fatigue, decreased a ppetite) and macrocytic anemia were observed in a worker following occupationa l dermal and inhalation exposure to an average concentration of 35 ppm EGME for 1-1.5 years (C ohen, 1984). The worker was also exposed to methyl ethyl ketone and propylen e glycol monomethyl ether at concentrati ons 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 neurol ogical 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 wei ght was observed in male rabbits exposed to 30, 100, or 300 ppm EGME 6 hours per day, 5 da ys per week for 13 weeks (Miller et al. , 1983). Degenerative changes in the germinal epit helium 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. Fe male 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 we ight 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

exposed to 300 ppm EGME. No effects on the re productive 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 week s. 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 wa s 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 wei ghts 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 st udy to provide a rapid assessmen t of the effect of glycol ethers on some aspects of repr oduction in the rat. Exposure to EGME was by inhalation at 100 and 300 ppm for 6 hr/day. First, pregnant female s 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 lit ters; 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 eff ects at 100 ppm. Exposure at 300 ppm EGME caused significant reductions in white bl ood 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 a nd 6000 ppm EGME (161 or 486 mg/kg/day) and in female rats administered drinking wa ter containing 1600 and 4800 ppm EGME (200 or 531 mg/kg/day) for 21 days (Exon et al. , 1991). Histopathological exam ination 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 obs erved 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 spl een of irradiated mice was significantly impaired as compared to that of control fetal mouse liver cells.

Appendix D3 256 Ethylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference 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/m 3 ; 60μg/m 3 )

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 μ g/m 3 ) can be made using the occupational study of Welch et al . (1988) of the semen of shipya rd painters exposed to both EGEE and EGME. The men supplied demographic characteristics, medical conditions, personal habits, and reproductive histor y; underwent a physical exam ination; and provided a semen sample. The painters were exposed to EG EE at a TWA concentration of 0 to 80.5 mg/m 3 (mean = 9.9 mg/m 3 , and to EGME at a TWA concentration of 0 to 17.7 mg/m 3 (mean = 2.6 mg/m 3 ). The painters had an increased prevalence of o ligospermia 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 smoki ng.) Adding the mean exposure levels together results in a total glycol ether co ncentration (EGME + EGEE) of 12.5 mg/m 3 , equivalent to a continuous exposure of 4.5 mg/m 3 . Division by a UF of 10 for a LOAEL and by another of 10 for human intraspecies variab ility results in a REL of 40 μ g/m 3 (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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for EGME include the availa bility of subchronic inhalation exposure data from a well-conducted study with hi stopathological analys is 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 mono methyl ether toxicity associated with microfilm production: A case report. Am. J. I nd. 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, Du ffy TP, McPhedran P, Kelsey KT, Redlich CA. 1992. Morphologic, biochemical, and cytogene tic 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 Silv a 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-but oxyethanol: 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 et hylene glycol monomet hyl 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: Mi cromedex, Inc. (Edition expires 4/30/95).

Appendix D3 258 Ethylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference 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 possi ble 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. Fund am. Appl. Toxicol. 3:49-54.

Parsons CE, and Parsons MEM. 1938. Toxic encepha lopathy and “granulopenic anemia” due to

volatile solvents in i ndustry: Report of two cases. J. I nd. 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. Evalua tion 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 et hylene 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 painte rs: II. Male reproduction. Am. J. Ind. Med. 14(5):509-526.

Zavon MR. 1963. Methyl cellosolve intoxica tion. Am. Ind. Hyg. Assoc. J. 24:36-41.

Appendix D3 259 Ethylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference 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 μ g/m 3 ( 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 C 5 H 10 O 3 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/m 3 per ppb at 25oC

III. Major Uses and Sources

Ethylene glycol monomethyl ethe r acetate (EGMEA) is used as a solvent for nitrocellulose, cellulose acetate, and various othe r 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 spec ific statewide industrial emissi ons 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 c ongenital hypospadia, chordee, micropenis, and

Appendix D3 260 Ethylene Glycol Monomethyl Ether AcetateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

scrotum bifida and her second child (3 years late r) 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 eyeglasse s. Ethylene glycol monoethyl ether acetate (EGEEA) was also present.

V. Effects of Animal Exposure

Cats, rabbits, guinea pigs, and mice were repeated ly 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 pp m 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 atr ophy was observed at doses above 250 mg EGMEA/kg/day. Decreased white blood cell coun ts were observed in all EGMEA-exposed

groups.

EGMEA was readily converted in vitro to ethyl ene 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 wei ght was observed in male rabbits exposed to 30, 100, or 300 ppm ethylene glycol monomethyl et her (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 change s 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 expos ed; two of five female rabbits exposed to 100 or 300 ppm EGME died during the course of the study. Redu ced body weight gain, pancytopenia (abnormal depression of all the cellular elements of the bl ood), and thymic atrophy were observed in rabbits of both sexes exposed to 300 ppm EGME. No e ffects 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 ra ts 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 AcetateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

exposed to concentrations of EGME of 100 pp m or greater. Pancytopenia, lymphoid tissue atrophy, and decreased liver wei ghts 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/m 3 , 60 μg/m 3 )

EGME 90 μg/m 3 EGMEA (20 ppb) (60 x MW EGMEA / MW EGME )

Data relating specific EGMEA exposure levels to toxicity in humans ar e not available for the development of a chronic REL. Data from expe rimental 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 toxic ity of EGMEA and EGME.

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for EGMEA include the availa bility of subchronic inhalation exposure data from a well-conducted study of EG ME as well as a number of supportive human studies on EGME showing the same toxicologi cal endpoint, and the observation of a NOAEL.

Appendix D3 262 Ethylene Glycol Monomethyl Ether AcetateDetermination of Noncancer Chronic Re ference 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 m onomethyl 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, Kn aak JB, Krasavage WJ, Skoulis NP, Stack CR, and Tyler TR. 1994. Glycol ethers and other select ed glycol derivatives. In: Patty’s Industrial Hygiene and Toxicology. Fourth ed. Vol. II, Pa rt B. Clayton GD, and Clayton FE. (eds.). New York: John Wiley Sons. p. 2761.

Gross E. 1943. In: Toxicology and Hygiene of I ndustrial 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: Mi cromedex, Inc. (Edition expires 7/31/96).

Jordan WP, and Dahl MV. 1971. Contact dermatitis to a plastic solvent in eyeglasses. Crosssensitivity to ethyl acetat e. 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. F undam. Appl. Toxicol. 3:49-54.

Nagano K, Nakayama E, Oobayashi H, Nish izawa T, Okuda H, and Yamazaki K. 1984. Experimental studies on toxicity of ethylene gl ycol alkyl ethers in Japan. Environ Health Perspect. 57:75-84.

Stott WT, and McKenna MJ. 1985. Hydrolysis of severa l 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 AcetateDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 C 2 H 4 O 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/m 3

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 ethyle ne 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 ma terials such as plastics. The

annual statewide industrial emissions from fac ilities reporting under the Air Toxics Hot Spots

Act in California based on the most recent i nventory 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 opera ted 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. Measurem ents 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 oxideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

bilateral reflex reduction in the ankle compared to controls. Nerve conduction tests did not identify significant differences between contro l 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 spatia l and visual abilities, and on tests of visual motor function. The results extended pr evious 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 w ith 24 control subjects. Neuropsychological

function in the workers was classified as eith er normal or impaired on the basis of the questionnaires and of neuropsychol ogical tests by 2 clinical ps ychologists (who were unaware of exposure status). (If the classi fication 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 freque ntly classified as impaired (5/12) compared to controls (1/16) ( χ 2 = 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 worker s from 9 hospitals in the U.S. and one in Mexico not only reports biological indicators of ethylen e 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 ti me weighted average of 0.08 or 0.17 ppm EtO. The mean length of employment for workers fromU.S. hospitals was 5.5 and 10 years for lowand high-exposure workers, respec tively. The mean length of employment in low- and highexposure workers from the hospital in Mexico wa s 5.9 and 4.2 years, resp ectively. In workers from U.S. hospitals only, statistically significan t decreases in hematocrit and hemoglobin were observed in high-exposure workers compared to low-exposure work ers. 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 fr om the hospital in Mexico, a significant relationship of EtO exposure and elevated neutro phil count was observed using regression.

At least 2 epidemiological re ports indicate a possi ble association of EtO exposure and spontaneous abortion. Hemminki et al. (1982) analyzed spontaneous abortions in Finnish hospital sterilizing staff using data from a posta l questionnaire and from a hospital discharge register. The study included a ll 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 fr equency of spontaneous abortion was 16.7% versus 5.6% for the non-exposed pregnancies. The i ndependent analysis of spontane ous abortions using the hospital discharge register confirmed the findings. Thus two analyses suggested that EtO exposure may carry a risk of spontaneous abor tion among sterilizing staff.

More recently Rowland et al . (1996) sent questionnaires to 7, 000 dental assistants (ages 18-39 years) registered in Californi a in 1987. Of these, 4,856 respo nded (69%). They analyzed 1,320

Appendix D3 265 Ethylene oxideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

women whose most recent pregnancy was concei ved while working full-time. Thirty-two reported exposure to EtO; unexposed dental a ssistants 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 signi ficantly higher in both exposure groups. Inflammatory lesions of the lung, nasal cavity, trachea, and inner ear were observed more frequently in EtO exposed ra ts. Skeletal muscle myopathy, consisting of atrophy and degeneration of skeletal muscle fibers, was observed more frequen tly 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 postu re during gait and reduced locomotor activity. A dose-response was observed for thes e effects, with significant ch anges at 50 ppm and greater. An abnormal righting reflex was observed in a si gnificantly greater perc ent of mice exposed to 100 ppm and above. Reduced or absent toe a nd tail pinch reflexes were observed in a significantly greater perc ent 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 relati ve liver weight was obs erved 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 absolu te testes weights at 50, 100, or 250 ppm EtO. This study indicates a subchronic NOAEL for neurological effect s of 10 ppm EtO.

In a study of the testicular effects of EtO, ma le 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-relate d decrease in testes we ight was observed at 4, 6, and 13 weeks of exposure. Progressive dege neration 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 oxideDetermination of Noncancer Chronic Re ference 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 indicat ed 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 μ g/m 3 )

Snellings et al . (1984) found a subchronic NOAEL of 10 ppm for neurological effects in mice. A neuromuscular screening test indicated that ce rtain reflex responses and locomotor activities were altered in EtO-exposed animals. Huma n studies have also i ndicated neurological impairment in ethylene oxide exposed workers.

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation R EL for ethylene oxide include th e 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 oxideDetermination of Noncancer Chronic Re ference 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 H ealth Service, ATSDR.

CARB. 2000. California Air Resources Board. Calif ornia 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 expose d 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 ethyl ene oxide exposure. Arch. Neurol. 44(12):1283-

1286.

HSDB. 1995. Hazardous Substances Data Bank. Na tional 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 sterilis ing 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 worker s chronically exposed to ethylene oxide. Clin. Toxicol. 28(1):21-28.

Lynch DW, Lewis TR, Moorman WJ, Burg JR, Gr oth DH, Khan A, Ackerman LJ, and Cockrell BY. 1984. Carcinogenic and toxicolo gic 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 occupationa lly 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 oxideDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Sarto F, Tornquist MA, Tomani n R, Bartolucci GB, Osterman-Golkar SM, and Ehrenberg L. 1991. Studies of biological and chemical monitori ng of low-level exposure to ethylene oxide.

Scand. J. Work Env. Health 17:60-64.

Saillenfait AM, Gallissot F, Bonnet P, and Protoi s 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, Tsuchi ya Y, and Halperin WE. 1995. Molecular, cytogenetic, and hematologic effects of ethylen e 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 B6C3F 1 mice. Toxicol. Appl. Pharmacol. 76:510-518.

Appendix D3 269 Ethylene oxideDetermination of Noncancer Chronic Reference Exposure Levels August 2003

CHRONIC TOXICITY SUMMARY



# FLUORIDES including
HYDROGEN FLUORIDE

(hydrofluoric acid (aqueous soluti on); hydrogen fluoride (as a gas); fluoride salts (particulat es or in solution))

CAS Registry Number: 7664-39-3

I

. Chronic Toxicity Summary

Inhalation reference exposure level 14 μ g HF/m 3 (17 ppb); 13 μ g F/m 3
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/m 3 @ 25°C

III. Major Uses or Sources

Hydrofluoric acid (HF) is a co lorless, fuming liquid with a sh arp, penetrating odor (Fairhall, 1949). This acid is used in the glass etching, electronic, microele ctronic, and pe troleum refining and chemical industries (Bertolini, 1992). These industries use HF in the manufacture of such things as computer chips (an im portant industry in California), phosphate fertilizer, metal cans, plastics, refrigerant chemicals (f luorocarbons), inorganic chemical s, 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 FluoridesDetermination 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 approxima tely 0.080 mg/kg/day for 7 to 9 m onth 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 carie s 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 adop ted regulations that esta blish standards for the addition of F (CDHS, 2002). Any public water syst em using fluoridation must maintain F levels within the range established for its climate. The ranges vary acco rding 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 pr ovide the same dental benefit. For 2001-2002, F levels in San Francisco municipa l 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 unde r 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 inciden ce of minimal osseous changes were studied in the workplace by Derryberry et al. (1963). In this study, the 8-hour timeweighted average fluoride expos ure was calculated for the empl oyment 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/m 3 . In comparison, the 17 workers within this group who had evidence of mi nimally increased bone density had an average fluoride exposure of 3.38 mg F/m 3 . The other workers were exposed to an average measured concentration of 2.64 mg F/m 3 . In addition, urinary fluoride levels were greater in the 17 individuals with greatest expos ure compared to the remaining 57 workers (average = 5.18 mg F/L vs. 4.53 mg F/L). No differences betw een 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 inci dence of acute resp iratory disease as determined from past medical histories was obs erved in fluoride-exposed individuals (19/74 vs. 8/67 in controls); radiographic ex amination 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 be low) showed a statistically significant relationship between ai r 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 th e log air fluoride concentration as the independent variable showed a significant (p < 0.033) re lationship between increasing air fluoride concentrations and probability of skeletal fluorosis. The parameters for the regression were β 0 =

− 2.3468 (std. error

Appendix D3 271 FluoridesDetermination of Noncancer Chronic Reference Exposure Levels August 2003

= 0.6462), and β 1 = 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 am ong normal, non-fluoride-exposed males (Runge et al

., 1979). As expected, age was very highly correl ated with years exposed (p<0.00001). Therefore including years exposed in the dose-metric likely introdu ces 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 am ong 245 aluminum smelter workers exposed to 2.75 or 3.2 mg F/m 3 . For these reasons, years exposed we re 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 in to 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 con centration of 1.07 mg F/m 3 did not exhibit bone density changes. An analysis of the grouped responses using a binom ial distribution showed a probability of p = 0.008 for obtaining 4/15 increased bone de nsity observations in the 2.34 mg/m 3 group, and a probability of p = 0.047 for obtaining 3/15 pos itive observations in the 1.89 mg F/m 3 group. The 1.89 mg F/m 3 group was therefore consider ed a LOAEL for chronic skel etal fluorosis, and the 1.07 mg/m 3 group was considered a NOAEL. The above probabilities assume that a chance occurrence is, at most, 1 in 18 of skeletal fluoro sis or other cause leading to an abnormally dense x-ray in the general population. Si nce osteosclerosis is a rare c ondition that is associated with several types of hematological malignancies such as myeloid leukemia, the actual incidence of conditions leading to osteoscler osis is far below 1 in 18. This lends strong support to the consideration of 1.89 mg/m 3 as a LOAEL for skeletal fluorosis.

Appendix D3 272 FluoridesDetermination 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/m 3 ) 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 FluoridesDetermination 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/m 3 ) 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 FluoridesDetermination 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/m 3 ) 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/m 3 ± 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 ch ance occurrence of abnormally dense x-ray of, at most, 1 in 18 individuals, using a binomi al 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 occupa tionally exposed to unspecified concentrations of HF for an average of 2.7 years, reported occasions of uppe r respiratory irritation (Evans, 1940). Repeated chest X-rays over a 5-year period did not reveal any visible evidence of lu ng changes. The death rate of these workers from pneumonia and other pul monary 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 mate rials 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 re sponse appear to be common in this work environment making it difficult to quantitatively re late the respiratory symptoms to inhaled HF

or fluorides.

Workers in a warehouse containing HF retorts experienced transito ry hyperemia of the skin on their face and hands (Dale and Mc Cauley, 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 FluoridesDetermination of Noncancer Chronic Reference Exposure Levels August 2003

Examinations of 107 pot room workers in two alum inum plants with airborne fluorides revealed 22 subjects with limited motion of the dorsolumbar spine, compar ed 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 m easured by roentgenogram, with diagnosis of slight to moderate fluorosis. Moderate a nd marked fluorosis was observed after 15 years employment. The 8-hour time-weighted average fl uoride content in these workplaces was 2.4 to 6.0 mg/m 3 . Balazova (1971) measured significant fluoride uptake and dist ribution 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 associat ed with toxicity to human kidney collecting duct cells leading to s odium 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 Diseas e Registry (ATSDR, 2001) recently reviewed fluorides since they are found at hazardous wast e sites which are candidates for remediation. The focus of this document was on oral exposure st udies as that is the main concern for waste

site remediation.

V. Effects of Chronic Exposures to Animals

Stokinger (1949) studied the subchr onic 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/m 3 , or 25.1 mg/m 3 6 hours/day, 6 days/week, for 30 days. Mortality, body weight, blood coagulation mechanisms, and gross pathol ogy were measured. Exposure to 25.1 mg/m 3 HF for 30 days resulted in degenerative testicular chan ges 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 de generation was observed in 27 of 30 rats. All of the rats and mice at the 25.1 mg/m 3 concentration died. No mort ality was observed in the other species tested. Blood fibrinogen levels were significantly increas ed in dogs, rats, and rabbits exposed to 25.1 mg/m 3 . Exposure to 7.2 mg/m 3 HF resulted in pulmonary hemorrhage in 1 out of 5 dogs. No other significant effects we re 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 tes tis, epididymis, and vas deferens microscopically. After 29 months of F ad ministration, the spermatogenic cells in the seminiferous tubules had degenerated and lack ed spermatozoa. After both 18 and 29 months, cilia were lost from the epithelial cells lining th e 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 appear ed to be peeled off. Mucus droplets

were abundant in the vas defere ns of controls, but none were present in F treated rabbits.

Appendix D3 276 FluoridesDetermination 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 adve rse effects (including infertility) on the male rabbit reproductive system.

Ghosh et al . (2002) investigated the e ffects 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 dehydroge nase (HSD) and 17beta-HSD activities, and significant lowering in plasma levels of testosterone. Epididymal sperm count was decreased significantly in F-treated rabbi ts and there were fewer mature luminal spermatozoa. Indicators of oxidative stress due to F included increased co njugated dienes in the testis, epididymis, and epididymal sperm pellet, a nd 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 ra t 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 ~28 a ~24 a < 0.05 b

Testicular 17betaHSD ~29 a ~24 a < 0.05 b
Plasma testosterone (ng/ml) ~2 a ~1 a < 0.05 b
Epididymal sperm count (10 6 /ml) 7.02±0.17 3.70±0.57 < 0.05

a approximate values based on read ing Figures 2 and 3 of paper; b p values of authors

Long et al . (2002) used ligand binding and Wester n 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 [ 125 I]alpha-bungarotoxin, a competitive antagonist, we re significantly decreas ed in the brains of rats exposed to both levels. The brain levels of the nAChR alpha4 subunit protein was significantly lowered by exposur e to 100 ppm F. Alpha7 subun it protein was significantly decreased by both levels of F. No significant ch anges 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 culture d 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 mi ce of both sexes for two years to 0, 25, 100, and 175 ppm sodium fluoride (NaF) in their dr inking 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 FluoridesDetermination of Noncancer Chronic Reference Exposure Levels August 2003

tooth deformities and attrition. NaF increased the dys plasia of dentine in both rats and mice. At the highest dose (175 ppm), osteosclerosis of lo ng bones was increased in female rats. There was also equivocal evidence of carcinogenic acti vity 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 worker s (67 unexposed control subjects) Exposure method Occupational Critical effects Increased bone density (skeletal fluorosis)

LOAEL 1.89 mg F/m 3 (1.98 mg HF/m 3 )

NOAEL 1.07 mg F/m 3 (1.13 mg HF/m 3 )

BMC 05 0.37 mg F/m 3 (0.39 mg HF/m 3 ) 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/m 3 (0.39 x 10/20 x 5/7) or 0.13 mg F/m 3 (0.37 x 10/20 x 5/7) Human equivalent concentration 0.14 mg HF/m 3 or 0.13 mg F/m 3
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/m 3 (13 μ g /m 3 ; 0.016 ppm; 16 ppb) or 0.014 mg HF/m 3 (14 μ g /m 3 ; 0.017 ppm; 17 ppb)

OEHHA’s analysis of the data in Derryberry et al . (1963) indicates a LOAEL of 1.89 mg/m 3

, and a NOAEL of 1.07 mg/m 3 . A benchmark concentration (BMC 05 ) of 0.37 mg/m 3 was derived by fitting the probit model to the log dose in the U.S. EPA’s BM DS (version 1.3) software, for the individual mean air exposure data and inciden ce data in Table 1 above. Individuals in the highest dose group (group 5 in Tabl e 2) were not included in the m odel, 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 lo g-transformed dose was selected si nce it produced a good fit not only by statistical criteria (p = 0.71) but also, as dete rmined by inspection, it fit the low dose curve shape

better than other models. This model also has th e advantage of biological plausibility, in that,

since lower doses of fluoride have a beneficial or nutri tional effect, a threshold type of response for adverse effects is clearly expected. A graphical representation of the f it is shown in Figure 1. Adjusting for exposure continuity and utilizing an intraspecies uncertainty factor of 10 (UF H )

results in a REL for F of 13 μ g/m 3

.

Appendix D3 278 FluoridesDetermination of Noncancer Chronic Reference Exposure Levels August 2003

Figure 1.

Changes in bone density in asso ciation with fluoride exposure ha ve been observed in several studies, and appear to be the most sensitive heal th effect for chronic exposure. The minimally increased bone density in the Derryberry study wa s 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, fi brosis, and healed tuberculous lesions. Although dental fluorosis is a sensitiv e endpoint in many fluorid e studies, the dental examinations of exposed workers in this study sh owed healthier teeth th an in controls. The increased bone density observed was considered as indicating that adverse effects had occurred,

based on the adverse effects associated with th e increased density in the study, and on other research showing that increased bone densit y caused by fluoride exposure (75 mg sodium fluoride per day for four years) also leads to decreased bone st rength and increased fragility (Riggs et al. , 1990). Symptoms of abdominal pain, b ackache, restricted joint movement, and respiratory symptoms have been associated with airborne fluoride e xposures and bone density

increases in industria l settings (Zhiliang et al ., 1987).

The absorption of particulate and gaseous fluor ides 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 FluoridesDetermination 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 conf irmed 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 a nd years of exposure. However, the product of exposure concentrati on and time did show a consistent pattern of cumulative incidence suggesting a dose-respons e 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

BMDL 05 could not be reported. However a maxi mum likelihood estimate of the benchmark (BMD 05 ) was found to be 6.04 (mg F*years/m 3 ), with exclusion of the three highest values that appeared to be outliers to the ma in distribution. If this value is divided by the mean exposure duration for the data set of 14.1 years, a be nchmark exposure concentration of 0.43 mg F/m 3 is obtained. While this value is evidently less re liable than that obtained by fitting the mean exposure concentration, it is c onsistent with it, suggesting that, although other confounding factors related to age or duration prevent the demonstration of a rela tionship between the exposure/time integral and response in this data set, such a rela tionship 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 ar e the observation of hea lth effects in a large group of workers exposed over many years, the av ailability 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 exam ination. Another source for concern is the potentially greater susceptibility of children to the effects of inhaled fl uorides, considering the rapid bone growth in early years.

Derivation of Chronic Oral REL

In addition to being inhaled, airborne fluoride salt s in particulate form can settle onto crops and soil and enter the body by ingestion. Thus an or al chronic reference exposure level (REL) for fluoride is also required in order to conduct a he alth 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 ch ronic 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 FluoridesDetermination 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 dri nks 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 μ g/day from drinking water for an 18 kg child drinking 40 ml/kg body weight/day of water). Moderate to severe de ntal fluorosis is rare when the drinking water fluoride level is near 1 mg/L, but begins to become significant at concentra tions close to 2 mg/L. Since the study involved long term exposure to humans including children, a sensitive population, the cumulative uncertainty factor wa s 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

m 3 /day, an equivalent inhalation REL would be about 170 μ g/m 3 . Thus, the inhalation REL of

13 μ g/m 3 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 sk eletal 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 caus es 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 R EL (see section VI) is lo wer than that implied by the oral REL and PHG. Since the oral R EL and PHG are based on ex posures throughout life, including the pre-natal period, infancy, and ch ildhood, it is reasonable to conclude that the proposed inhalation REL is generally protectiv e of infants and child ren, 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 FluoridesDetermination of Noncancer Chronic Reference Exposure Levels August 2003

Extensive interindividual variation in total fl uoride intake (930.7 ± 391.5 μ g/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 variati on has also been reported for slightly younger children in Connersville (n = 14) and Indi anapolis, Indiana (n = 29) and in San Juan, Puerto Rico (n = 11) (Rojas-Sanchez et al ., 1999). Consideration shoul d therefore be given to populations with exceptionally high fluoride intake due to locally elevated concentrations in drinking water, since some of th ese populations are already close to adverse effect levels of fluoride intake, and certain individuals in Calif ornia experience dental fluorosis. For these individuals, even exposure to fl uorides at the oral and/or inha lation RELs, which are acceptable in isolation, might be deleterious. The tabl e 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

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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 FluoridesDetermination 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 Com pounds. 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 lig ht and scanning electr on microscopy. J. Reprod. Fertil. 92(2):353-60.

U.S. EPA. 2001. United States Environmental Pr otection 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 fluo ride, benefits and risks, report of the ad hoc

subcommittee on fluoride of the committee to co ordinate environmental health and related programs. February 1991.

Van Nieuwenhuysen JP, D'Hoore W. 1992. [Dental cari es, fluoride tablets a nd enamel opacities]. Arch. Fr. Pediatr. 49(7):617-621.

Waldbott GL. Fluoridation: The Great Dile mma. 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. Envi ron. 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 FluoridesDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

CHRONIC TOXICITY SUMMARY



# GLUTARALDEHYDE

(1,5-pentanedial; 1,5-pentanedi one; glutaric dialdehyde; Aldesen; Cidex; Sonacide)

CAS Registry Number: 111-30-8

I. Chronic Toxicity Summary

Inhalation reference exposure level 0.08 μg/m 3 (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 C 5 H 8 O 2 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/m 3 per ppb at 25oC

III. Major Uses and Sources

Glutaraldehyde is a chemical fr equently used as a disinfectant and sterilizing agent against bacteria and viruses (2% solution), an embalmi ng fluid and tissue fixa tive, 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 a tmospheric reaction product of cyclohexene. The

annual statewide industrial emissions from fac ilities reporting under the Air Toxics Hot Spots

Act in California based on the most recent i nventory 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 disinfec tant and sterilizing agent. Frequently observed effects from exposure include skin sensitivity resulting in derm atitis, 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 GlutaraldehydeDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

1989). Occupational asthma has also been re ported among workers re peatedly exposed to glutaraldehyde, particularly re spiratory 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 st udy. Histopathological examination of animals surviving to the end of the study entailed scoring the severity of the finding from “no response” to “very severe” resp onse on a 0 to 5 scale. Unit length labeling index, the indicator of ce ll proliferation, was evaluated by autora diography at two sites: the nasal

vestibule and the dorsal atrioturbinate.

Lesions in animals treated with glutaraldehyde a ppeared 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 lesi ons 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 a nd/or severity). The lesions were characterized as neutrophilic infiltration prim arily in the squamous epithelium of the vestibule, with thickening of the epithelium leading to loss of the characte ristic surface grooves. Bo th cell size and number were reported to be increased. Lesions were ge nerally found to increase in nature and severity with increased time and level of exposure. Obst ruction 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 accumula tion of eosinophilic proteinaceous deposits in the respiratory epithelium of the maxilloturbinate margin. Examination of unit length labeling indices as a measure of growth sh owed 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 GlutaraldehydeDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

Mean Subjective Pathology Scores for Na sal 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 glutar aldehyde, 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 h ours 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 a nd olfactory mucosa, and sensory irritation. In the two highest doses of the 14-week study, stat istically significant differences in body weight gain were observed as well as perinasal wetness. No histopathological indicati on of inflammation in olfactory or nasal mucosa was observed.

Mice were exposed to 0, 0.3, 1.0, and 2.6 ppm glut araldehyde vapors for 6 hours/day for 4, 9, or 14 days (Zissu et al. , 1994). These mice were killed im mediately 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 af ter 2 weeks of recovery.

A study comparing the effects of intra-nasally instilled glutaraldehyde and formaldehyde on rat nasal epithelium found inflammati on, epithelial degeneration, resp iratory 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 sa me 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 inha lation 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 th e chamber controls. Increased incidences of nonneoplastic nasal lesions occurred primarily within the anterior sect ion of the nose in 500 and 750 ppb rats and to a lesser extent in 250 ppb rats. The more si gnificant lesions in cluded hyperplasia and inflammation of the squamous and respiratory epithelia and squamous metaplasia of the respiratory epithelium. Thus 250 ppb (1000 μ g/m 3 ) 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 i nhalation for 6 h/day, 5 days/week, for 104

Appendix D3 288 GlutaraldehydeDetermination of Noncancer Chronic Re ference 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 ge nerally less than those of the controls. The incidence of inflammation of the nose was marginal ly 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 hyali ne degeneration of the respiratory epithelium were increased in all exposed groups of fe males. 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

BMC 05 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 cm 2 ) 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/m 3 )

Several studies indicate that the upper respir atory tract is a target for the toxicity of glutaraldehyde from inhalation e xposure. Reports of toxicity to humans show that exposure can

Appendix D3 289 GlutaraldehydeDetermination of Noncancer Chronic Re ference 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 glutar aldehyde by the inhalation route show evidence of respiratory irritation with the induction of lesi ons of the anterior na sal 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 ol factory epithelium in the lowest dose exposure group. (Female mice exposed to 62.5 ppb also showed subepithelial neutrophilic infiltrati on.) This level was

taken to be the subchronic LOAEL. This effect on the nasal epithelium was demonstrated to be both concentration- and expos ure 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 MLE 05 of 31.24 ppb, a BMC 05 of 20.51 ppb, and a p value of 0.9471. With the benchmark approach no LOAEL UF is need ed. The study was a lifetime study so the subchronic UF is 1. An interspecies UF of 3 ra ther 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 glut araldehyde is 0.02 ppb (0.08 μg/m 3 ).

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/m 3 ).

VII. Data Strengths and Limitations for Development of the REL

The major strength of the inhalation REL for glut araldehyde is the availability of controlled exposure inhalation studies in multiple species at multiple exposure conc entrations and with adequate histopathogical analysis. Major areas of uncertainty are the lack of human data, the lack of reproductive and developm ental toxicity studies, the lack of dermal sensitization studies,

and the lack of observation of a NOAEL.

VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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. Occupa tional asthma in a technologist exposed to gl utaraldehyde. J. Allerg y 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 GlutaraldehydeDetermination of Noncancer Chronic Re ference 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 formalde hyde in endoscopy and x-ray de partments. 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. Hi stopathology and cell replication responses in the respiratory trac t 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 7/31/96).

Jordan WP, Dahl MV, and Albert HL. 1972. Co ntact dermatitis from glutaraldehyde. Arch.

Dermatol. 105:94-95.

National Toxicology Program (NTP). 1998. T oxicology 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. T oxicology 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 pro liferation induced by

intra-nasal instillation of alde hydes in the rat: comparison of gl utaraldehyde 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 ta le. Occup. Med. 44:95-98.

Wiggins P, McCurdy SA, and Zeidenberg W. 1989. Epistaxis due to glutar aldehyde exposure. J. Occup. Med. 31:854-856.

Zissu D, Gagnaire F, and Bonnet P. 1994. Nasal a nd pulmonary toxicity of glutaraldehyde in

mice. Toxicol. Lett. 71:53-62.

Appendix D3 291 GlutaraldehydeDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (2000 ppb)

Critical effect(s) Neurotoxicity; electrophysio logical alterations in

humans Hazard index target(s) Nervous system

II. Physical and Chemical Properties (HSDB, 1999)

Description Colorless liquid, gas Molecular formula C 6 H 14 Molecular weight 86.10 Density 0.660 g/cm 3 @ 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/m 3 @ 25° C

III. Major Uses or Sources

n-Hexane is used in the extrac tion of vegetable oil from seeds such as safflower, soybean, cotton, and flax (HSDB, 1995). It is also used as a al cohol denaturant and as a paint diluent. The textile, furniture and leather i ndustries use n-hexane as a clean ing agent. Many petroleum and gasoline products contain n-hexane. The annual statewide industrial emissions from facilities reporting under the Air Toxics Ho t Spots Act in California based on the most recent inventory were estimated to be 999,225 pounds of hexane (CARB, 1999).

Appendix D3 292 HexaneDetermination of Noncancer Chronic Re ference 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 di stal 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 sura l nerve biopsy of a severe case. Optic neuropathy and CNS impairment were not usua lly found. Personal air samples had 80 to 210 ppm hexane (mean = 132 ppm), 20 to 680 ppm is opropanol (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, w ith a range of 1 month to 30 years.

An epidemiologic study was performed on work ers employed in a factory producing tungsten carbide alloys and exposed for an average of 6.2 years to solven t 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 perfor med 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 ot her reports of n-hexa ne-induced peripheral neuropathy. The study reports a LOAEL of 58 ppm n-hexane.

Polyneuropathy with subsequent development of muscular atrophy and pare sthesia 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-hexa ne in vegetable oil ex tracting and adhesive bandage manufacturing processes was examined fo r 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 concentrat ions were usually well below 500 ppm. The authors concluded that the high short-term exposur es, occurring occasionally for 1 to 2 hours at a time, could have been major factors in the eff ects observed. Visual e voked potentials (VEPs) were generally reduced among the exposed subj ects 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 proofi ng 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 foun d in one factory in which all six workers

Appendix D3 293 HexaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

developed polyneuropathy. Workers exposed to less than 100 ppm n-hexane who frequently worked overtime demonstrated significant decrea ses in motor nerve cond uction velocities in median, ulnar, and peroneal nerv es. Twelve of 13 workers who re gularly slept in the factory had polyneuropathy compared to three (7%) of 46 em ployees 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 he xane and small amounts of toluene. The worksites were poorly ventilated and concentrat ions 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 e nough to discharge the patient.

Urinary 2,5-hexanedione concentrations were sign ificantly 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, a nd 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 parkinsoni sm 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 wi th positron emission tomography studies, were

considered distinct from those seen in idiopathic Parkinson's disease.

Co-exposure to acetone increased the urinary con centrations 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 a nd 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 fo r 26 weeks to 0, 5, 27, or 126 ppm hexane for 21 hours/day, 7 days/week. There were no consistent dose-related di fferences between exposed and control animals, although small numbers of animals were involved and examinations were limited to physical observation, body weight, hema tological parameters, clinical chemistry, and necropsy of spontaneous deaths. The highe st 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/se x/concentration/species) inhale d 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 mort ality were noted between hexa ne-exposed and control groups. Small statistically signif icant reductions in body weight gain we re noted in male and female rats inhaling 3000 ppm or more and in female mice inhaling 9000 ppm. Epith elial cell hyperplasia

was increased in the nasoturbinat es and larynx of exposed rats.

Appendix D3 294 HexaneDetermination of Noncancer Chronic Re ference 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, opht halmologic examination, neurological function, or hematological or serum chemistry parameters in either males or females. Female body weights and clinical ob servations were unaltered by hexane treatment. The mean body weight gain of male rats in the 10,000-ppm group was sign ificantly decreased compar ed with controls at 4 weeks of exposure and thereafter . Axonopathy was noted in the ti bial 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 weight s. No other adverse histopathological effects were reported. This study identifies a NOAEL fo r neurotoxicity of 3000 ppm, with an average experimental exposure of 540 ppm.

B6C3F 1 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” we re noted in those mice exposed to 500 or 1000 ppm n-hexane. Paranodal axonal swelling in th e 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 an imals exposed to 500 and 1000 ppm for 6 hours per day. A NOAEL for histological le sions of the nasal tu rbinates of 500 ppm nhexane was identified. Because neurohistological examinations were not performed in animals exposed to 500 or 1000 ppm (the NOAEL and LOA EL, 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). Electromyogr aphy, strength-duration curves, electrical reaction time, and flexor/ext ensor chronaxy ratio, gait posture and muscular atrophy were studied. Increased complexity of NMU (neuromuscular unit) voltages during electromyographic analysis was noted in 0/6 co ntrols, 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 redu ced interference voltages from muscles was noted in mice exposed to 250 ppm or mo re, but not in controls (0/6 examined) or in the 100 ppm group (0/6). Dose-related abnormal pos ture and muscle atrophy were noted at 250 ppm or more. This study identifies a NOA EL of 100 ppm for neurotoxicity (68 ppm when adjusted for 67.5% n-hexane).

Rats inhaling 400-600 ppm n-hexane developed peripheral neuropat hy after forty-five days of exposure (Schaumburg and Spencer, 1976). Giant axonal swellings and fi ber degeneration were observed in the central and periphe ral nervous systems. The changes were most notable in tibial

nerves and in the cerebellum , medulla and spinal cord.

Appendix D3 295 HexaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

A dose-dependent decrease in motor nerve c onduction velocity and body weight gain was observed in rats exposed to 500, 1200, or 3000 pp m 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 myelin ated 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 et hyl ketone (Altenkirch et al. , 1982).

Acetone has also been shown to potentiate the ne urotoxicity of hexane a nd 2,5-HD. Male rabbits administered acetone and 2,5-HD intravenous ly 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 ambul ation, or rearing and on the rotarod test. Giant axonal swelling was gr eater in acetone administered animals. During a dose-free 10- week recovery period, the acetone-supplemente d group had less improvement in neurological parameters. Male Wistar rats were admini stered 0.5% w/v 2,5-hexa nedione alone or in
combination with 0.50% w/v acetone in the dri nking water for 7 weeks (Lam et al., 1991). Effects on radial arm maze behavior, a "brain-s welling" reaction, and synaptosomal functions were noted with 2,5-HD and exacerbat ed 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 significan t 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, indica ted by decreased body wei ght 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, embryotoxicit y, or impaired fetal growth or development.

Male New Zealand rabbits expos ed to 3000 ppm n-hexane for 8 hours per day, 5 days per week for 24 weeks developed exposure-related lesions of the respiratory trac t with the terminal bronchioles exhibiting the most ch aracteristic 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 HexaneDetermination of Noncancer Chronic Re ference 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 cons idered by OEHHA to be most informative and

relevant to the derivation of a chronic REL. This was because these st udies (1) evaluated the most sensitive endpoint (peripheral neuropathy) and (2) involved expo sures over a significant fraction of a lifetime. While significant limitations may be noted for each of these studies individually, viewed collectively th ey 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 uncertain ty of interspecies differences, the limitations of the human studies were considered to be more significant. Specifically, both human studies we re considered likely to overes timate effects of inhalation exposures to hexane.

The Sanagi study, which U.S. EPA used as the basi s of its RfC, may overestimate hexane effect because of a confounding coexposure to acetone , which is known to potentiate hexane neuropathy. The minimum effec tive acetone inhalation concentra tion 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. Th e minimum effective acetone inhalation dose for potentiation of carbon tetrachlori de 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 m 3 /day = 5 g/m 3 ; 5 g/m 3 x 1000 mg/g ÷ 3.52 mg/m 3 per ppm = 1400 ppm). As the acetone potentiating effects were all noted at higher ex posures than are being considered in occupational studies and are at mu ch higher concentrations than the REL itself, the significance of these fi ndings is uncertain.

In the Chang study, the workers were probably intermittently exposed to higher inhalation exposures than were estimated from ambient air sampling, and signifi cant dermal exposures were also likely. Furthermore, coexposure to hi gh 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 Mi yagaki 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 su pporting the 7000 μg/m 3 REL.

Appendix D3 297 HexaneDetermination of Noncancer Chronic Re ference 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/m 3 ; 7000 μg/m 3 )

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/m 3 ) (TWA) (TWA) Sanagi et al.

, 1980 6.2 years decreased motor nerve conduction velocity; increased residual latency 58 20.7 Not observed 100 a 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 300 b 300 1000

a LOAEL uncertainty factor, 10; Intr aspecies uncertain ty factor, 10

b LOAEL uncertainty factor, 10; Subchronic uncertain ty factor, 3; Intraspeci es 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 Ch ang study involving an unusual 72-hour work week was calculated by assuming that 12 hours of o ccupational 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 HexaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

The Chang study found that the severity of eff ects 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 e ffects 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

NOAEL a TWA from

LOAEL a

Sanagi et al

. (1980) Humans 58 ppm (mean) 10 m 3 /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 in clude (1) the primary us e of an animal study (Miyagaki, 1967) with controlled, nearly con tinuous chronic hexane exposures not confounded by coexposure to other solvents, which observe d both a NOAEL and LOAEL; and (2) the results obtained from two different human studies (S anagi, 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 HexaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

et al ., 1984) were not found to cause ne uropathy in rats. Also human studies (especially that of Chang et al ., 1993) have shown that so me 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 av erage exposure of 110.2 ppm. Using an RGDR of 1 and a cumulative 30-fold uncertainty factor (3 for in terspecies differences not accounted for by the RGDR method and 10-fold for intraspecies differences), a REL of 4 ppm (10,000 μg/m 3 ) 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 μ g/m 3 ).

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 + met hyl-ethyl-ketone mixtures. J.

Neurol. Sci. 57:209-219. Bio Dynamics Inc. 1978. 26 Week Inhalation Toxici ty 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 occupationa l 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 neur otoxic 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. Ne urol. Neurosurg. Psychi atry 56(5):538-542.

Charbonneau M, Brodeur J, du Souich P, Plaa GL. 1986. Correlation between acetonepotentiated CCl 4 -induced liver injury and blood concen trations 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 inhalati on 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 i nhalation exposure. Toxi cology 57(2):163-172.

Appendix D3 300 HexaneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, Maryland. WWW database (htt p://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-s pecific and muscle-spe cific proteins. Arch. Toxicol. 63(5):381-385.

Ichihara G, Saito I, Kamijima M, Yu X, Sh ibata 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-induc ed 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- hexanedi one-induced neurotoxicity studied with behavioural and morphometric methods in rats. Ph armacol. 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 ace tone on radial arm maze behavior, the "brainswelling" reaction and synaptosomal functi ons. Neurotoxicol. Teratol. 13(4):407-412.

Larsen JJ, Lykkegaard M, Ladefoged O. 1991. Infe rtility in rats induced by 2,5-hexanedione in combination with acetone. Pharmacol. Toxicol. 69(1):43-46.

Litton Bionetics Inc. 1979. Teratology Study in Ra ts. 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 th e peripheral neurotoxicity of n-hexane. Jap. J. Ind. Health 9(12-23): 660-671 Paulson GW, and Waylonis GW. 1976. Polyneur opathy 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 parkin sonism: pathogenetic hypotheses. Mov. Disord. 10(3):279-82.

Appendix D3 301 HexaneDetermination of Noncancer Chronic Re ference 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 f unctions of workers exposed to n-hexane at a

low level. Int. Arch. Occ up. Environ. Health 47:69-79.

Schaumburg HH, Spencer PS. 1976. Degeneration in central and peripher al 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 ch anges in visual evoked potentials and electroretinograms of industri al 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 En vironmental Assessment, Environmental Criteria

Assessment Office.

Wang J-D, Chang Y-C, Kao K-P, Huang C-C, Li n C-C, and Yeh W-Y. 1986. An outbreak of nhexane induced polyneuropathy among press proo fing workers in Taipei. Am. J. Ind. Med. 10(2):111-118.

Yamada S. 1967. Intoxication polyneuritis in th e 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 HexaneDetermination of Noncancer Chronic Re ference 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/m 3 (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 N 2 H 4 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, me thyl-, ethyl-, isobutyl alcohols; slightly miscible with hydrocarbons; insoluble in chloroform,

ether Conversion factor 1.31 μg/m 3 per ppb at 25oC

III. Major Uses and Sources

Hydrazine is a highly reactive ba se 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 oxyge n scavenger in boiler feedwater to inhibit corrosion, and in phot ographic development (Von Burg and Stout, 1991). Hydrazine was historically used experiment ally as a therapeutic agent in the treatment of tuberculosis, sickle cell anemia, and non-specific chronic il lnesses (Von Burg and Stout, 1991; Gold, 1987).

The annual statewide industrial emissions from facilities reporting unde r 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 HydrazineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

IV. Effects of Human Exposure

One person was occupationally exposed to hydraz ine 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 progre ssed to abdominal pain and incoherence. The previously healthy 59-year old i ndividual died three weeks after th e 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, bu t the repeated associa tion between exposures and symptoms is highly suspicious. Liver toxicity is also associated with acute exposure to hydrazine.

The only epidemiological studies of human hydr azine 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 le ast 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 in cidental exposure to hydrazine. Only cumulative mortality was reviewed. Health effects reported during or af ter 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 repo rted that relative risk s 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 carcinoge nicity 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 pl us 150 rats/sex as controls; 400 female C57BL/6 mice at 0.05, 0.25, and 1.0 ppm hydrazine pl us 800 female mice as controls; 200 male Golden Syrian hamsters at 0.25, 1.0, and 5.0 ppm hydr azine plus 200 male hamsters as controls; 4 beagle dogs/sex at 0.25 and 1.0 ppm hydrazine pl us 4 dogs/sex as controls. Animals were observed post-exposure for the fo llowing 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 cy clic elevations in serum glutamic-pyruvic transaminase levels and, upon necropsy at 36 months post-exposure, showed liver effects described as “clusters of swollen hepatocytes th at had highly vacuolated cytoplasm.” Of the

Appendix D3 304 HydrazineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

other species examined, hamsters showed toxic ity at the lowest dose levels, particularly amyloidosis in various organs including liver, sp leen, kidney, thyroid, and adrenal glands. An increased incidence of amyloidosis was seen at the lowest exposure le vel (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 th e liver was also significantly increased in all exposed groups. Significantly increased incidences of toxic e ffects observed in the 1.0 and 5.0 ppm hydrazine groups include amyloidosis of th e 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. Th e authors note these effects appear to reflect accelerated changes commonly associated with aging in hamsters.

Incidence of Nonneoplasti c Lesions in Male Hamsters (from Table 3 of Vernot et al .)

Lesion Control 0.25 ppm 1.0 ppm 5.0 ppm

L i v e r 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

A d r e n a l 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 gr eater 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 gr oup (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 hydr azine. 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 exam ined (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 contro l group consisted of equal numbers of animals. The experimental desi gn was such that each intermittent exposure group had a time-weighted-average matching conti nuous exposure group. Dose-related body weight

reductions were observed in all treated groups as well as eviden ce of hepatic degeneration, fatty

Appendix D3 305 HydrazineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

deposition in the liver, central nervous syst em depression and leth argy, eye irritation, and

anemia.

Toxic effects from the exposure of rats, mice, a nd 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 breat hing, vomiting, fatigue, and emphysema in dogs; pulmonary congestion and emphysema in rats a nd 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/m 3 , 0.2 μg/m 3 )

Vernot et al. (1985) present a thorough examination of chronic health effects from inhalation exposure to hydrazine. This st udy 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 li ver in hamsters) showed a doseresponse relationship, and (3) the findings of th is study support data found in studies by other

groups.

This study shows a dose-related increase in the in cidence 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 include d weight depression during expos ure, mineralization of the kidney, and amyloidosis of the thyroid. Haun and Kinkead (1973) have also noted lesions of the

Appendix D3 306 HydrazineDetermination of Noncancer Chronic Re ference Exposure Levels December 2000

liver in dogs, monkeys, and mice exposed contin uously 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 avai lability of chronic inhalation exposure data from a well-conducted study with hi stopathological analys is. 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. Calif ornia 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. Inha lation 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 Toxico logy. Wright-Patterson Air Force Base, OH.

AMRL-TR-73-125:351-363.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 7/31/96).

Merck. 1983.The Merck Index. 10 th 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. O ccup. 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 hydr azine derivative in a stain remover. Crosssensitization to apresoline and isoniazid. Acta Derm. Venereol. (Stockh.) 44:49-53.

Appendix D3 307 HydrazineDetermination of Noncancer Chronic Re ference 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 Youn g JT. 1985. Long-term inhalation toxicity of hydrazine. Fundam. Appl. Toxicol. 5:1050-1064.

Von Burg R, and Stout T. 1991. Toxicology Upda te: Hydrazine. J. Appl. Toxicol. 11:447-450.

Von Keilig, W. and Speer U 1983. Occupational a llergic contact derm atitis 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 evid ence. 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 HydrazineDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# HYDROGEN CHLORIDE

(Hydrochloric acid; anhydrous hy drogen chloride; muriatic acid)

CAS Registry Number: 7647-01-0

I. Chronic Reference Exposure Level

Inhalation reference exposure level 9 μg/m 3 (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/m 3 at 25°C

III. Major Uses or Sources

Hydrogen chloride (HCl) is used in the manufacture of vinyl chloride, fertil izers, dyes, artificial silk, and pigments for paints. It is also used in electroplating, soap refini ng, and leather tanning.

Other consumers of HCl include the photographic, textile and rubbe r 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 quan tities during the test fi ring 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 S pots 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 ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

IV. Effects of Human Exposure

Few reports are available on the effects of ch ronic HCl exposure on human s. 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 concen trations (Stokinger, 1981).

In another report, workers exposed to various mineral acids, including HC l, 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 worker s), intermediate (23%) in an intermediateexposure category (picklers), and low (7%) in a lo w-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 gr ade 3 (loss of enamel and dentine with exposure of sec ondary dentine) was noted after si x 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 compar ed to air-exposed controls.

A 90-day inhalation study using B6C3F1 mice a nd 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-expos ure groups. No effect were noted in hematology, clinical chemistry, or urinalysis. Minimal or mild rh initis 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 obser ved 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, 19 76). 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 simila r to those observed in the exposed dams.

Female rats were exposed to 302 ppm HCl fo r 1 hour prior to mati ng (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 ki dney, liver, and spleen effects. Estrus cycles

Appendix D3 310 Hydrogen ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

were also altered. In rats ma ted 12-16 days postexposure and k illed on day 21 of pregnancy, a decrease in fetal weight, an in crease in relative fetal lung weight s, 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 na sal 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

MV
a = 0.33 L/min, MV h = 13.8 L/min,

SA
a (ET) = 15 cm 2 ); Sa h = 200 cm 3 ) (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/m 3 ; 9 μg/m 3 )

Both extrathoracic and tracheobronchial effects ha ve been associated with exposures to hydrogen

chloride. The REL was based on extrathoracic e ffects 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 a nd 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 μ g/m 3 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 reproduc tive toxicity data; and (4) the l ack of chronic exposure studies (U.S. EPA, 1994). OEHHA agrees with this assessm ent. The database for chronic exposure to

this common chemical is limited.

Appendix D3 311 Hydrogen ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

VIII. References

Burleigh-Flayer H, Wong KL, a nd Alarie Y. 1985. Evaluation of the pulmonary effects of HCl using CO 2 challenges in guinea pigs. F undam. Appl. Toxicol. 5:978-985.

CARB. 1999b. Air toxics emissions data collect ed in the Air Toxics Hot Spots Program CEIDARS Database as of January 29, 1999.

GEOMET Technologies, Inc. 1981. Hydrogen ch loride: 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. Na tional Library of Medicine, Bethesda, Maryland. WWW database (htt p://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 Albe rt RE. 1985. Carcinogenicity of formaldehyde and hydrogen chloride in rats. To xicol. 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. e d. Volume 2B, Toxicology. New York: Wiley Interscience. pp.2959-2961.

Ten Bruggen Cate H.J. 1968. Dental erosi on in industry. Br. J. Ind. Med. 25:249-266 Toxigenics Inc. 1984. 90-Day i nhalation 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 Environmenta l Protection Agency. Interim Methods for Development of Inhalation Reference Concentr ations. Environmental Criteria and Assessment Office, Office of Health and Environmental Asse ssment, 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 ChlorideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# HYDROGEN CYANIDE

( Formonitrile; hydrocyani c acid; prussic acid)

CAS Registry Number: 74-90-8

I. Chronic Toxicity Summary

Inhalation reference exposure level 9 μg/m 3 (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 @ 20 o C Solubility Miscible in water, alcohol; slightly soluble

in ether Conversion factor 1 ppm = 1.10 mg/m 3 @ 25 o C

III. Major Uses or Sources

Hydrogen cyanide is used in a variety of synt heses including the producti on of adiponitrile (for nylon), methyl methacrylate, sodium cyanid e, cyanuric chloride, chelating agents, pharmaceuticals, and other specialty chemicals. Manufacturing activit ies releasing hydrogen cyanide include electroplati ng, metal mining, metallurgy and metal cleaning processes. Additionally, hydrogen cyanide has some insecticide and fungicide applications (ATSDR, 1993). Fires involving some nitrogen-containing polymer s, often found in fibers used in fabrics, upholstery covers, and padding, also produce hyd rogen cyanide (Tsuchiya and Sumi, 1977).

Another common source of hydrogen cyanide is cigarette smoke. Le vels in inhaled mainstream cigarette smoke range from 10 to 400 μg per cigare tte (U.S. brands); 0.6% to 27% (w/w) of these

mainstream levels are found in seco ndary or sidestream smoke (Fiskel et al. , 1981). The annual

statewide industrial emissions from facilities re porting under the Air Toxics Hot Spots Act in

Appendix D3 313 Hydrogen CyanideDetermination 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 st udies investigating hydrogen cyanide exposure are complicated by the mixed chemical environments created by s ynthetic and metallurgic processes. However, several reports indicate that chronic low expos ure 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 contai ning 3% copper cyanide, 3% sodium cyanide, and 1% sodium carbonate. Breathing zone cyanide concentr ations ranged from 4.2 to 12.4 ppm (4.6 to 13.7 mg/m 3 ), with means from 6.4 to 10.4 ppm (7.1 to 11.5 mg/m 3 ), in the three factories at the time of this cross-sectional study. The men were expo sed for a duration of 5 to 10 years, except for one man with 15 years exposure. Twenty non-expos ed male volunteers were used as controls. None of the subjects, controls or workers, currently smoked cigarette s. Complete medical histories were taken, and medical exams were performed. Urinar y levels of thiocyanate (a metabolite of cyanide) were utilized as a bi ological index of exposure. Thyroid function was measured as the uptake of radiolabeled iodine, since thiocyanate may bloc k the uptake of iodine by the thyroid leading to iodine-deficiency goi ters. Frequently reported symptoms in the exposed workers included headache, weakness, and a ltered sense of taste or smell. Lacrimation, abdominal colic, and lower stomach pain, saliv ation, and nervous inst ability occurred less frequently. Increased blood hemoglobin and ly mphocyte 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 w ith either the incidence or the degree of enlargement. Thyroid function test indicated significant difference s in uptake between controls and exposed individuals after 4 a nd 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 pe riod. This study reported a LOAEL of 6.4 ppm (7.1 mg/m 3 ) 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 fu mes. 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 retr oactively defined by the investigators as low-, moderate-, or high-exposure through work historie s. 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 CyanideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

increases in thyroid-stimulati ng hormone levels, which in comb ination with the CNS effects, suggest long-term adverse effects as sociated 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 inha lation exposure. Cassava ro ot, a dietary staple in many tropical regions, contains cy anogenic glycosides such as linamarin which release cyanide (CN - ) when metabolized endogenously (Sharma, 1993; Kamalu, 1995). Consumption of insufficiently processed cassava roots over a pe riod of time in combination with a protein deficient diet has been implicat ed in neurotoxic effects. One such neuropathy known as konzo results in nerve cell degeneration leading to a permanent but non-progressi ve spastic weakness of the legs and degeneration of corresponding corticospinal pathways (Tylleskar et al ., 1992; TorAgbidye et al ., 1999). The development of this sydrom e 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 nutriti onal 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 da ily cassava flour consumption at above 0.5 kg per adult in a konzo-affected, albeit malnourishe d, 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 (C asadei et al.,1984; Cliff et al., 1984).

Other effects associated with cassava consum ption include pancreatic diabetes, vitamin B 12 deficiency and decreased iodine uptake (Sharma, 1993; Jansz and Uluwaduge, 1997). Cretinism in children, associated with a deficiency of diet ary iodine, is worsened by eating cassava (Miller, 1974). Excess thiocyanate due to cyanide metabolis m 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 villag e, 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 gl and due to cassava consumption.

V. Effects of Animal Exposures

There is little animal data for chronic inha lation 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/m 3 ) for either 1 or 4 weeks produced no microscopically detect able morphological cha nges of the lungs, pulmonary arteries, coronary arteries or aort a. This study observed a subacute inhalation NOAEL for HCN in rabbits of 0.5 ppm (Hugod, 1979, 1981). Four dogs exposed to 50 mg/m 3 (45 ppm) hydrogen cyanide in a series of 30-minute inhalation periods conducted at 2-day intervals

Appendix D3 315 Hydrogen CyanideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

demonstrated extensive CNS toxicity, includ ing dyspnea and vomiting, with vascular and

cellular CNS lesions identifie d 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 (Leu schner 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 decrea sed 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 (P hilbrick et al., 1979). Th ere 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 carbohydr ate 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 fe d cassava compared to dogs fed rice + NaCN.

These effects indicate that all th e intact cyanogenic glycosides absorbed from cassava, primarily linamarin, was not hydrolyzed to HCN. Ho wever, 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 concl uded that the lesions observed in the cassava fed dogs were not entirely due to cyanide.

No information was found regarding developmen tal 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 effect s, increased runting a nd decreased ossification in hamsters (Frakes et al. , 1986), but not in rats fed cassava alone, or supplemented with potassium cyanide (Tewe and Maner, 1981). Hamste rs with gestational ca ssava exposure did not display reproductive effects (Frakes et al. , 1986).

Appendix D3 316 Hydrogen CyanideDetermination 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 occupation al inhalation exposures Critical effects CNS effects, thyroid enlargement, and hematological disorders

LOAEL 7.1 mg/m 3

NOAEL Not observed Exposure continuity 8 hr/day (10 m 3 /day/20 m 3 /day), 5 days/week Average occupational exposure 2.5 mg/m 3 for LOAEL group Human equivalent concentration 2.5 mg/m 3 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/m 3 , 9 μg/m 3 )

The USEPA based its RfC of 3 μ g/m 3 on the same study but include d 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. EP A. Such modifying factors were not used by

OEHHA. OEHHA used a 3-fold subchronic uncert ainty 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 CyanideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

An alternative analysis was conducted using da ta from an animal ingestion study reporting effects at low cyanide concentrations:

Study Jackson (1988) Study population Miniature swine Exposure method Daily oral administratio n of aqueous potassium cyanide Critical effects Behavioral effects; decreased blood T 3 and T 4 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/m 3 for LOAEL group assuming 20 m 3 /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/m 3 (0.5 μg/m 3 ; 0.0004 ppm; 0.4 ppb)

This study reported neurobehavioural and thyroid effects at cyanid e exposure levels (equivalent to 1.4 to 4.2 mg/m 3 ) similar to that reported by El Ghawabi (2.5 mg/m 3 ). However, as greater uncertainty factors are required fo r use of the animal study, a lower REL was derived. Use of a cross-route extrapolation also introduces uncer tainty. 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 hea lth effects data. The major uncertainties are the lack of a NOAEL observation in the key study, the difficulty in estimating exposures, and the discontinuous a nd variable nature of the exposures.

VIII. References

Abuye C, Kelbessa U, and Wolde-Gebriel S. 199 8. Health effects of cassava consumption in South Ethiopia. East Afri can 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 Se rvices, Public Health Service. TP-92/09.

Ansell M, and Lewis FAS. 1970. A review of cyan ide concentrations found in human organs: A survey of literature concerning cyanide meta bolism, ‘normal’, non-fatal and fatal body cyanide

levels. J. Forensic Med. 17:148-155.

Appendix D3 318 Hydrogen CyanideDetermination 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 factor s and hydrocyanic acid conten t 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, a nd 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 la boratory 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, radioiso tope, 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 pr enatal development. Fundam. Appl. Toxicol.

7:191-198.

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, Maryland. WWW database (htt p://sis.nlm.nih.gov/sis1/).

Hugod C. 1979. Effect of exposure to 0.5 ppm h ydrogen cyanide singly or combined with 200 ppm carbon monoxide and/or 5 ppm n itric oxide on coronary arteri es, aorta, pulmonary artery, and lungs in the rabbit. Int. Ar ch. Occup. Environ. Health 44:13-23.

Hugod C. 1981. Myocardial morphology in rabbits e xposed to various gas-phase constituents of

tobacco smoke. Artherosclerosis 40:181-190.

Jackson LC. 1988. Behavioral effects of chroni c sublethal dietary cyanide in an animal model: implications for humans consuming cassava (Manihot esculenta). Hum. Biol. 60(4):597-

614.

Appendix D3 319 Hydrogen CyanideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Jansz ER and Uluwaduge DI. 1997. Biochemical asp ects of cassava (Manihot esculenta Crantz) with special emphasis on cyanoge nic glucosides – a review. J. Natn. Sci. Coun. Sri Lanka 25(1):1-24.

Kamalu BP. 1993. Pathological changes in growi ng 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. Toxi cokinetic 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. Hea lth implications of scopoletin in gari. Plant F oods Hum. Nutr. 41(3):283-289.

Okolie NP, Osagie AU. 1999. Liver and kidney lesi ons and associated enzyme changes induced in rabbits by chronic cyanide exposur e. Food Chem. Toxicol. 37(7):745-750.

Philbrick DJ, Hopkins JB, Hill DC, Alexande r JC, Thomson RG. 1979. Effects of prolonged cyanide and thiocyanate feedi ng in rats. J. Toxicol. E nviron. Health 5(4):579-592. Sharma RP. 1993. Cyanide containing foods and pot ential 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 diet ary inorganic cyanide (KCN) in the life cycle performa nce 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 ca ssava. 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 diseas e found in Africa. Lancet (N Am Ed). 339(8787):208-211.

U.S.EPA. 1999. U.S. Environmen tal Protection Agency. Integrat ed Risk Information System (IRIS) Database. Reference concentr ation (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 CyanideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# HYDROGEN SULFIDE

(hydrogen sulphide; dihydrogen sulfide; dihydrogen monosulfide; sulfur hydride; sulfureted hy drogen; hydrosulfuric acid) CAS registry number: 7783-06-4

I

. Chronic Toxicity Summary

Inhalation reference exposure level 10 μ g/m 3 ( 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 H 2 S 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 μ g/m 3 ) (Amoore and Hautala, 1983) Odor description Resembles rotten eggs Conversion factor 1 ppm = 1.4 mg/m 3 @ 25° C

III. Major Uses or Sources

Hydrogen sulfide (H 2 S) is used as a reagent and an inte rmediate 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 ra yon production, sewage treatment, and leather tanning (Ammann, 1986). The annual statewide industrial emissions from point sources at faci lities 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 SulfideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

IV. Effects of Human Exposure

Although numerous case studies of acutely toxic effects of H 2 S exist, there is inadequate occupational or epidemiological information for specific chronic effects in humans exposed to

H 2 S.

Bhambhani and Singh (1991) showed that 16 healt hy subjects exposed for short durations to 5 ppm (7 mg/m 3 ) H 2 S under conditions of moderate exer cise 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/m 3 ) H 2 S caused coughing and thro at irritation after 15

minutes.

In another study, ten asthmatic volunteer subjects were exposed to 2 ppm H 2 S for 30 minutes and pulmonary function was tested (Jappinen et al. , 1990). All subjects re ported detecting “very unpleasant” odor but “rapidly became accustomed to it.” Three subjects reported headache following exposure. No significant changes in mean FVC or FEV 1 were reported. Although individual values for specif ic airway resistance (SR
aw ) were not reported, the difference following exposure ranged from

− 5.95% to +137.78%. The decrease in specific airway conductance, SG
aw , ranged from -57.7% to +28.9%. The increase in mean SR
aw and decrease in

mean SG
aw were not statistically significant.

Kilburn and Warshaw (1995) inve stigated whether people exposed to sulfide gases, including

H 2 S, as a result of working at or living downw ind from the processing of "sour" crude oil demonstrated persistent neurobeha vioral dysfunction. They studied thirteen former workers and 22 neighbors (of a California coastal oil refinery) who complained of headaches, nausea, vomiting, depression, personality changes, noseb leeds, and breathing difficulties. Their neurobehavioral functions and a pr ofile of mood states were comp ared to 32 controls (matched for age and educational level). The exposed subject s' mean values were st atistically significantly different (abnormal) compared to controls for several tests (two-choice reaction time; balance (as speed of sway); color discrimination; digit symb ol; trail-making A and B; immediate recall of a story). Their profile of mood st ates scores were much higher th an those of controls. Visual recall was significantly impaired in neighbors, bu t not in the former workers. The authors concluded that neurophysiological abnormalities were associated with expos ure to reduced sulfur gases, including H 2 S from crude oil desulfurization.

Xu et al . (1998) conducted a retrospective epidem iological study in a large petrochemical complex in Beijing, China in order to assess th e possible association between petrochemical exposure and spontaneous abortion. The facility consisted of 17 major production plants divided into separate workshops, which allow for the a ssessment 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 empl oyment at the plant, participat ed in the study. According to their employment record, about 57% of these workers reported occupational exposure to petrochemicals during the first trimester of thei r 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-expos ed groups within each plant, exposure to

Appendix D3 322 Hydrogen SulfideDetermination 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 wa s 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 inconsiste nt reports between recalled exposure and work history, a comparable risk of spontane ous abortion (OR 2.9; 95% CI = 2.0 to 4.4) was found. In analyses for exposure to specific chemi cals, 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, an d the results for hydrogen sulphide (OR 2.3; 95% CI = 1.2 to 4.4) were significant. No hydrogen sulfide exposur e 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 H 2 S and developed lacrimation, eye irritation, nausea, vomiting, headache, sore throat, and skin irritation but

retained conciousness as the re sult 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 identificat ion, with the workers who had experienced unconciousness most severely aff ected in the followup tests.

Three patients exposed acutely to unknown c oncentrations of hydrogen sulfide developed persistent cognitive impairment (Wasch et al ,. 1989). While standard neurological and physical examinations were unremarkable, all three subj ects had prolonged P-300 la tencies 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/m 3 , respectively) H 2 S for 6 hours/day, 5 days/week for 90 days (CIIT, 1983a,b). Measurements of neurological and he matological function revealed no abnormalities

due to H 2 S exposure. A histological examination of the nasal turbinates also revealed no significant exposure-related cha nges. A significant decrease in body weight was observed in both strains of rats exposed to 80 ppm (112 mg/m 3 ).

In a companion study, the Chemical Industry In stitute 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/m 3 , respectively) H 2 S 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. Th e 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/m 3 ) H 2 S. Weight loss was also observe d in the mice exposed to 80 ppm. Neurological and hematological tests revealed no abnormalities. The 30.5 ppm (42.5 mg/m 3 )

level was considered the NOAEL fo r histological changes in the nasal mucosa. (Adjustments were made by U. S. EPA to this value to calcul ate an RfC of 0.9 μ g/m 3 .)

Appendix D3 323 Hydrogen SulfideDetermination 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 baselin e 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 th e bronchial submucosa. Some exposed animals responded similarly to controls to aerosol methacholine challenge, whereas a subgroup of exposed rats were hyperreactive to co ncentrations as low as 1 ppm.

Male rats were exposed to 0, 10, 200, or 400 ppm H 2 S for 4 hours (Lopez et al. , 1987). Samples of bronchoalveolar and nasal lavage fluid contai ned 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/m 3 ) for 4 hours resulted in mild perivascular edema (Lopez et al.

, 1988).

A study by Saillenfait et al. (1989) investigated the de velopmental toxicity of H 2 S 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 H 2 S from day 6 of gestation until day 21 postp artum. 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 pp m). Pups which were exposed in utero and neonatally to day 21 postpartum developed with a s ubtle decrease in time of ear detachment and hair development and with no other observed ch ange in growth and development through day 21
postpartum.

Appendix D3 324 Hydrogen SulfideDetermination 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/m 3 )

NOAEL 30.5 ppm (42.5 mg/m 3 ) 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

MV
a = 0.033 L/min; MV h = 13.8 L/min;

SA
a (ET) = 3.0 cm 2 ; Sa h (ET) = 200 cm 3 ) (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 μ g/m 3 )

The adverse effects reported in chronic animal stud ies occur at higher concentrations than effects seen in acute human exposures. For example, hum an 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 s uggests either that humans are

more sensitive to H 2 S, or that the measurements in labor atory animals are too crude to detect subtle measures of irritation. However, the un certainty 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 agen t causing human fatalities following inhalation exposures. Although lower concentr ation 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 un certainty is the lack of adequate long-term human exposure data. Subchronic (but not chronic) st udies have been con ducted with several animal species and strains, and these studies o ffer 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 expos ure concentrations, adequate histopathogical analysis, and the observation of a NOAEL.

Appendix D3 325 Hydrogen SulfideDetermination 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 μ g/m 3 , 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 sta ndard for hydrogen sulfide, based on a 1 hour averaging time, is 42 μ g/m 3 (30 ppb).

Amoore (1985) analyzed a large number of reports from the scientific l iterature and found that reported thresholds for detecti on were log-normally distribute d, with a geometric mean of 10

μ g/m 3 (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 fo r hydrogen sulfide between 2.5 and

40 μ g/m 3 (2 and 32 ppb). Sources of variation included age, sex, medical conditions, and smoking. Training and alertness of th e subject in performing the test also affected the results.

Amoore (1985) drew attention to the difference between a detection th reshold under laboratory conditions, and the levels at which an odor could be recognized, or at which it was perceived as annoying. Analysis of various labo ratory and sociological studies suggested that a level at which an odor could be recognized was typically a f actor of three greater than the threshold for detection, while the level at which it was percei ved as annoying was typica lly a factor of five greater than the threshol d. Annoyance was characterized both in terms of esthetic or behavioral responses, and by physiological res ponses such as nausea and headach e. He therefore predicted that, although at 10 μ g/m 3 (the proposed REL) 50% of the ge neral population would be able to detect the odor of hydrogen sulfide under cont rolled conditions, only 5% would find it annoying

at this level. At 50 μ g/m 3 , 50% would find the odor annoying.

On this basis, the proposed REL of 10 μ g/m 3 (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 pr ovide reasonable protection from odor annoyance in practice. However, this consideration cannot be entirely dismissed due to the wide interindividual variation in sensitiv ity to odors. Amoore (1985) also points out that many industrial operations generating hydrogen sulfide also genera te organic thiol compo unds 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 hundr ed-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 th e levels of hydrogen sulfide meas ured in the affected areas.

Appendix D3 326 Hydrogen SulfideDetermination 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 physiolo gic respiratory response to H 2 S poisoning. J. Haz.

Mat. 13 :369-374.

Amoore JE (1985). The perception of hydrogen sulf ide 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 h ydrogen 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 sulf ide 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 T oxicology. 90-Day vapor inhalation toxicity study of hydrogen sulfide in Fischer-344 rats. U.S. EPA, Office of Toxi c Substances Public Files.

Fiche number 0000255-0. Document number FYI-OTS-0883-0255.

CIIT. 1983b. Chemical Industry Institute of Toxi cology. 90-Day vapor inhalation toxicity study of hydrogen sulfide in Sprague-Dawle y 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 T oxicology. 90-Day vapor inhalation toxicity study of hydrogen sulfide in B6C3F1 mice. U.S. EPA, Of fice 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. Hydroge n sulfide exposure alte rs 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 Pu rkinje cells. Neurosci. Lett. 122(2):225-228.

Appendix D3 327 Hydrogen SulfideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Hayden LJ, Goeden H, and Roth SH. 1990. Gr owth 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 effect s on the olfactory system of exposure to hydrogen sulphide. Occup. E nviron. 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 sulfid e. Clin. Invest. Med. 12(4): B89.

International Programme on Chemical Safe ty. 1981. Environmenta l Health Criteria 19. Hydrogen Sulfide. Geneva: World Health Organization. p. 41.

Jappinen P, Vilkka V, Marttila O, Haahte la 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. T oxicol. 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 e xposed 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 expos ed to sublethal to lethal concen trations 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 sulf ide on embryonal and fetal development in rats.

Toxicol. Lett. 48:57-66.

U.S.EPA. 1999. U.S. Environmen tal Protection Agency. Integrat ed Risk Information System (IRIS) database. Reference concentr ation (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 ex posure. 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. Occ up. Environ. Med. 1998;55(1):31-36.

Appendix D3 328 Hydrogen SulfideDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# ISOPHORONE

(1,1,3-trimethyl-3-cyclohexene- 5-one; 3,5,5-trimethyl-2-cycl ohexen-1-one; isoforon; isoacetophorone)

CAS Registry Number: 78-59-1

I. Chronic Toxicity Summary

Inhalation reference exposure level 2,000 μ g/m 3 (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 C 9 H 14 O 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 μ g/m 3 per ppb at 25 ° C

III. Major Uses and Sources

Isophorone is used extensively as a solvent in so me printing inks, paints, lacquers, adhesives, vinyl resins, copolymers, coatings, finishes, a nd pesticides, in additi on 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 industr ial sources including iron and steel manufacturers, manufact urers of photographic equipment a nd 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, isophor one has been found to occur naturally in cranberries (ATSDR, 1989). Occupa tional exposure may occur by inhalation or dermal contact.

The annual statewide industrial emissions from facilities reporting unde r 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 IsophoroneDetermination of Noncancer Chronic Re ference 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 screen ing plant ranged from 8.3-23 ppm and from 3.5- 14.5 ppm, respectively (Samimi, 1982). Up to 25.7 pp m 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 plan t in Ridgefield, NJ was 0.7-14 ppm (Lee and Frederick, 1982). It was suspected that the reported eye and nose ir ritation of workers at th e silk screening plant and at the decal manufacturing plant was the resu lt of acute and subacute exposure to isophorone

vapors.

Workers exposed to 5-8 ppm (28-45 mg/m 3 ) 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 orga ns 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 wi de 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 abso rption and elimination of the chemical. Preliminary results of a pharmacokinetic study indicate that rats treated orally with 14 Cisophorone excreted 93% of the radiolabel in th e urine, expired air, and feces in 24 hours (ATSDR, 1989). The highest levels of 14 C-isophorone were found in th e liver, kidney, preputial gland, testes, brain, and lungs. Several metabolites were identified in the urine of orally dosed rats and rabbits, includ ing 3-carboxy-5,5-dimethyl-2 -cyclohexene-1-one, 3,5,5- trimethylcyclohexanol, and some glucur onide 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/gro up and 10 guinea pigs/group, all of mixed sex, were exposed to 0, 25, 50, 100, 200, or 500 ppm is ophorone 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 chemistr y and histopathologica l changes in the kidney and lungs were

noted in treated animals. Howe ver, later investigations determ ined that the isophorone used in this study was contaminated with appreciable amounts of compounds (Rowe and Wolf, 1963).

Appendix D3 330 IsophoroneDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Therefore, some of the adverse effects (i.e ., the lung lesions) may ha ve been due to the contaminants. The accuracy of the concentratio n data in the 1942 study is also questionable.

No treatment-related histopatholog ical lesions were found in lungs , livers, or kidneys of male and female rats exposed intermittently (6 hr/d ay, 5 days/week) to 37 ppm isophorone for 4 weeks compared to controls (Hazleton Labs, 1968; su mmarized by ATSDR, 19 89). 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 ra tios of treated rats were significantly reduced compared to controls. Slight variations in hematological findings were noted in treated rats (increased lymphocytes and hemoglobin co ntent; 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 nasa l 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 micr ovacuolization 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 redu ced weight gain compared to c ontrols (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, he matology, and clinical chemistry indices found

no treatment-related effects in the animals at either the interim or final toxicological examinations. Gross pathology and a limite d histopathological examination observed no

treatment-related effects in ei ther species. Data on isophor one 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 administer ed 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 th e study. However, several deaths in male and female rats at the high dose occurred early in the study. A steep declin e 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 bo dy weight (<10%) compared to controls. A 13- week range finding study for the 2-year study did not find compound-rela ted 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. Increas ed mineralization of the colle cting ducts in isophorone-exposed

Appendix D3 331 IsophoroneDetermination of Noncancer Chronic Re ference 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 a nd occurred coincidentally with lesions of chronic nephropathy. Nephropathy was observed in almo st 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 inciden ce of nephropathy in dosed female rats, but the effects were not pronounced. However, the isoph orone potentiation of ne phropathy 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 hyperp lasia 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-re lated. A dose-dependent increase in fatty metamorphosis occurred in the adrenal cortex of ma le 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 in cidence of hyperkeratosis of the forestomach was observed in dosed male and high dose female mi ce, but was probably not a relevant treatmentrelated effect.

Published studies on possible re productive effects of isophorone are lacking. An unpublished inhalation study conducted by a commercial labo ratory (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 confin ed to slightly decreased weight (7-8%) on one day in the 115 ppm group. No significant differen ces were found in uterine implantations, fetal toxicity, and external and inte rnal 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 s light, but insignificant, increase in extra ribs and/or rudimentary

ribs was seen in rat and mouse fe tuses 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 fe tuses 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 IsophoroneDetermination of Noncancer Chronic Re ference 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/m 3 , 2,000 μ g/m 3 )

The inhalation study by Bio/dynamics (1984a,b) pr esents data that i ndicate 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 cr own-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 ad verse effect is not signi ficant. However, this selective culling before the statistical analysis is not scientifically appropri ate in this case. In addition, the authors did not perform some of the scheduled fetal examinations. Otherwise, the growth retardation might have had even greate r 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 c onsidered a serious teratogenic effect that was present at a dose (150 ppm) only slightly higher than the LO AEL 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 st aining seen in many exposed rats is not considered a chronic

‘adverse’ effect.

Appendix D3 333 IsophoroneDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

For comparison with the proposed REL of 0.4 pp m, the inhalation LOAEL of 250 ppm for mild liver effects (Dutertre-Catella, 1976) in rats and rabbits intermi ttently 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, inte rspecies = 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 observe d by Dutertre-Catella (1976) is unknown, the NTP (1986) study observed an increas ed incidence of hepatocytomegaly and coagulative necrosis of the liver in treated male mice, but not in female mice and rats, orally gavaged with isophorone. Usi ng 250 mg/kg-day as a LOAEL fo r 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-da y. Multiplying the oral REL by 3,500 μ g/m 3 per mg/kg-day for route-to-route extrapolation results in a chronic inhalation REL estimate of 900 μ g/m 3 (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/developmenta l 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 releva nt route of exposure under occ upational settings, but is most likely a minor route of exposure for the ge neral population. Due to the insufficient characterization of the kidney and liver lesi ons in the oral gava ge NTP study (Bucher et al , 1986; NTP, 1986) and the inhalation st udy (Dutertre-Catella, 1976), a co mprehensive 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 ad equately protective of infants and children. However, there is no dir ect evidence in the lite rature to quantify a differential effect of isophorone in children relati ve 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, isop horone as a Hot Spot emission is unlikely to be a multimedia chemical, and there is no evidence to suggest that norma l dietary levels of isophorone are associated with adverse health effects.

Appendix D3 334 IsophoroneDetermination of Noncancer Chronic Re ference 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 perfor med by Affiliated Medical Enterprises, Inc. Princeton, NJ for Rohm and Haas Co. Phila delphia, 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 performe d by Affiliated Medical Enterprises, Inc. Princeton, NJ for Rohm and Haas Co. Phila delphia, 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 Servi ce. Atlanta, GA: ATSDR. PB90-180225.

Bio/dynamics. 1984a. Inhalation teratology prob e study in rats and mice. Project No. 323771. Unpublished study performed by Bio/dynamics In c. 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. T oxicological and carcinogenesis studies of isophorone in F344 rats and B6C3 F1 mice. Toxicology 39:207-219.

CARB. 1997. California Air Resources Board. T oxic Air Contaminant Identification List

Summaries.

CARB. 2000. California Air Resources Board. Calif ornia 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. Contribut ion to the analytical toxicol ogical and bio-chemical study of isophorone (in French). Thesis for doctorate in pharmacology, Univer site Rene Descartes, Paris. [Cited in Joint Assessment of Commodity Chem icals, 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 (iso phorone). Toxicol. Eur. Res. 1(4):209-216.

Appendix D3 335 IsophoroneDetermination of Noncancer Chronic Re ference 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 11/31/95).

Kominsky JR. 1983. Health hazard determina tion 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 recomm ended standard. U.S. Departme nt of Health, Education, and Welfare. DHEW (NIOSH) Publication No. 78-173.

NTP. 1986. National Toxicology Program. Toxicol ogy and carcinogenesis studies of isophorone

in F/344 rats and B6C3F 1 mice. NTP TR 291. NIH Publication No. 86-2547.

Rowe VK, and Wolf MA. 1963. Ketones. In: In dustrial Hygiene and Toxicology, Second ed. Patty FA (ed.) New York: Interscience Publishers. p. 1764.

Samimi B. 1982. Exposure to isophorone and other or ganic 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 vapo rs. J. Ind. Hyg. Toxi col. 28(6):262-266.

Strasser J Jr, Charbonneau M, Borgoff SJ, Turn er 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. Res ponse 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 St udies - Isophorone Summary Data Sheet. Industrial Medical Toxicology De pt. New York: Union Carbide.

Appendix D3 336 IsophoroneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# ISOPROPANOL

(2-propanol; dimethylcarbi nol; isopropyl alcohol)

CAS Registry Number: 67-63-0

I. Chronic Toxicity Summary

Inhalation reference exposure level 7,000 μ g/m 3 (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. Sl ightly bitter taste. Molecular formula C 3 H 8 O 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 μ g/m 3 at 25 ° C

III. Major Uses and Sources

Isopropanol is used as a solvent and in maki ng many commercial products (HSDB, 1995). The annual production volume of isopropanol has been in excess of one billion pounds since 1956; it

was ranked 50 th among chemicals produced in the U.S. in 1994 (C&EN, 1995). Rubbing alcohol is a solution of 70% isopropanol in water. Spec ific uses and sources include: a component of antifreeze; a solvent for gums, shellac, essential o ils, creosote and resins; extraction of alkaloids; component of quick drying oils and inks; com ponent of denaturing alco hol; antiseptic for hand lotions; rubefacient; compone nt of household products (after-sha ve lotions, cosmetics, etc.); the manufacture of acetone; deicing ag ent 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), wh ich is 7-fold higher th an the chronic REL.

Appendix D3 337 IsopropanolDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

The annual statewide industrial emissions from facilities reporting unde r the Air Toxics Hot Spots Act in California based on the most recen t inventory were estimated to be 525,826 pounds of isopropanol (CARB, 1999b).

IV. Effects of Human Exposures

Currently, there are no adequate chronic expos ure 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 da ily 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 majo r 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 administer ed dose (via i.v. or inhalation) of isopropanol was exhaled as acetone, CO 2 and the unmetabolized alcohol (Slauter

et al. , 1994). Approximately 3-8% of the admi nistered dose was excreted in urine as isopropanol, acetone, and a metabolite tentatively identified as isopropyl glucuronic acid. Isopropanol is readily abso rbed from the GI tract and persists in the circulation longer than ethyl alcohol. Alcohol dehydrogenase oxidizes most isopropanol to acetone. A cetone may be further metabolized to acetate, formate, and finally CO 2 . In another metabo lism study, the amount of

acetone in the blood stream was found to be di rectly related to the air concentration of isopropanol (Laham et al. , 1980). This finding indicated that the acetone metabolite could be

used as a biochemical indicat or 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 Ag ency deemed these studies insufficient to reasonably predict subchr onic toxicity of isopropa nol (Burleigh-Flayer et al. , 1994). Three different routes of exposure have been used by researchers for isopropan ol toxicity studies: inhalation, oral gavage and pres ence in drinking water. The following subchronic and chronic studies exposed experiment al 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, clin ical signs observed follo wing exposures included swollen periocular tissue (females) at the highe st dose and perinasal encrustation (males) at 500 ppm and above. Narcosis was observed in a fe w animals of both species during exposure to 5000 ppm and possibly 1500 ppm as well. However, th e animals became tolerant to the narcotic effects of isopropanol after week 2. No neur obehavioral changes were observed in any parameters of the functional obs ervational battery. However, in creased motor activity was noted

Appendix D3 338 IsopropanolDetermination of Noncancer Chronic Re ference 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. Increas es 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 volu me (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 an d a slight dehydration in female mice at the end of the study. Relative liver we ight 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 w ithin kidneys of all male rats. This change was not clearly concentration related, a lthough 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 an imals, 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/wee k for at least 78 weeks (mice) or 104 weeks (rats). Transient signs of narcosis were observe d at the higher doses. Increased mortality and a decreased mean survival time (577 da ys versus 631 days for controls) were noted for male rats in the 5000 ppm group. Increases in body weight a nd/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 a nd/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 functio n (i.e. decreased osmolality and increased total protein, volume, and glucose), we re noted for male rats in th e 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, gl omerulosclerosis, interstitial nephritis, inters titial fibrosis, hydronephrosis, and transitional cell hyperplasia. The authors considered chronic renal disease

to be the main cause of death for male and fe male 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 st udy. 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 mi ce included an increased inciden ce 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 re nal tubular dilation for female mice in the 5000-ppm group. The seminal vesicle effects did not have any associated inflammatory or degenerative changes. The enla rgement 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 reveal ed no exposure-related lesions. In a 13-week behavioral/neurotoxicity study by th e same investigators, the reproducibility and reversibility of increased motor activity in isopropanol-exposed female Fischer 344 rats was

Appendix D3 339 IsopropanolDetermination of Noncancer Chronic Re ference 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 activ ity 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 isop ropanol for only 9 weeks. In the subgroup exposed for 13 weeks, reversal of the increa sed motor activity did not occur until 2 weeks following cessation of exposure. However, comple te reversibility of the time versus activity profile, or habituation curve, was not noted un til 42 days following exposure to isopropanol for 13 weeks.. Other effects included a significan t 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 eff ects, 20 male Wistar rats/group were exposed to 0 or 300 ppm isopropa nol 6 hr/day, 5 days/week for up to 21 weeks (Savolainen et al. , 1979). Enzyme activity of super oxide dismutase and azoreductase in cerebellar homogenate was decreased at week 20-21. Acid proteas e activity in glial cells was increased up to week 10. Open-field tests indi cated sporadic changes in urination (10th week) and defecation (15th week). Isopropanol also app eared 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 veloc ities 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 pre gnant dams (15/group) to 0, 3500, 7000 or 10,000 ppm isopropanol 7 hr/day on ge station days 1-19 (Nelson et al. , 1988). At the two highest exposure levels, feed intake (w eeks 1 and 2 of exposure) and ma ternal body-weight gain were reduced. Narcosis was evident only at the 10,00 0 ppm level.Increased fetal resorptions and reduced fetal weights (59% of c ontrols) occurred at the highest exposure level. Fetal weights were also significantly reduced (85% of contro ls) 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 we ight reduction was present across all exposed group s. Skeletal malformations (primarily rudimentary cervical ribs ) were seen only in the presence of maternal toxicity at the two highest expos ure levels. No detect able teratogenic effects were observed in the 3500 ppm group. The authors noted that th e developmental effects at 3500 ppm were considered very slight, indicati ng that this exposure level is close to the LOAEL for isopropanol.

The following studies administered isopropyl alcohol to experime ntal animals by oral gavage:

In a developmental study, pregnant (VAF)CD(SD) rats (25/group) were ga vaged with either 0, 400, 800 or 1200 mg/kg body wt-day of isopropanol da ily on gestational days 6 through 15 (Tyl

et al. , 1994). In the same study, pregnant New Z ealand white rabbits (1 5/group) were dosed orally with either 0, 120, 240 or 480 mg/kg body wt -day of isopropanol da ily during gestational days 6 through 18. In rats, fetal body weight ex hibited a linear downward trend with increasing dose and was significantly lower at the highest do se compared to controls. However, the fetal

Appendix D3 340 IsopropanolDetermination of Noncancer Chronic Re ference 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 obser ved 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 admi nistered 0, 200, 700 or 1200 mg/kg body wt-day isopropanol by oral gavage from gestationa l day 6 through postnatal day 21 (Bates et al. , 1994). One high-dose dam died on postnatal day 15, but th ere 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 use d. Developmental neurotox icity, in the form of motor activity, auditory startle and active a voidance 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, a nd during the gestation and lactation period for the F1 and F2 litters, respectively. In a dult rats, centrilobular hepatocyte hypertrophy and increased relative liver weight (>10%) was obs erved in P2 males at 1000 mg/kg. A general

increase in absolute and relative liver and kidne y weights was observed (less than 10% in most cases) in treated animals in both P1 and P2 ge nerations. However, with the exception of hepatocellular hypertrophy in P2 males, no histopat hological 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 obser ved at the 1000 mg/kg dose level.

The following toxicology studies administered isopropanol to expe rimental animals in drinking

water:

In a study designed to investig ate 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 in take and body weights were consistently lower at the two highest doses. Re lative 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 lesi ons. Evidence of astrogli osis, in the form of increased glial fibrillary acidic protein in dorsal hippocampus, was not found in exposed rats.

In a 1-generation study, 10 Wist ar-derived rats/sex/group we re exposed to 0, 0.5, 1.25, 2.0, and 2.5% isopropanol in water for up to 18 weeks (U SEPA/OTS, 1986). The doses are equivalent to 0, 325, 711, 1002, and 1176 mg/kg body wt-day, resp ectively, for males; to 0, 517, 1131, 1330, and 1335 mg/kg body wt-day, respectively, for fema les 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-mati ng, plus 15 days during mating, plus 42 days for

Appendix D3 341 IsopropanolDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

males; 21 days pre-mating plus 15 days during ma ting, plus 21 days gest ation, plus 21 days rearing in females; and 21 days for the F 1 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 author s concluded that these effects were related to the unpalatability of drinking water containing isoprop anol and not due to a toxic e ffect 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 hi gher 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 contro ls 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 inve stigating the potential teratoge nic effects of isopropanol, 20 pregnant Wistar-derived rats/group were e xposed 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 significan tly reduced at the highest dose. Fetal body weights were decreased in the two highest dose groups. Minor abnormalities and variants (reduced ossification of the skelet on) were present in fetuses of exposed groups in a dose-related manner. However, the authors concluded th at the reduced fetal weights are probably a consequence of maternal growth retardation du ring 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’ ra ts 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 othe r 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 necr osis 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 IsopropanolDetermination of Noncancer Chronic Re ference 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/m 3 , 7000

μ g/m 3 )

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 n ecessary 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 rena l 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 resp onse 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 developm ent of tolerance. The changes in behavior

Appendix D3 343 IsopropanolDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

were small and unconvincing. This study would ha ve benefited from additional dose levels to analyze for dose-response trends.

Other possible sensitive indicators of isopropano l 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 we ights 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/developmen tal 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 obser ved at 7000 ppm and 10,000 ppm. Th e average exposure duration at the LOAEL for this study is 1024 ppm (7hr/24h r x 3510 ppm). Use of an RGDR of 1 and a cumulative uncertainty factor of 100 (3 for LO AEL to NOAEL, 3 for interspecies, and 10 for intraspecies) resulted in a REL of 10 ppm (25 mg/m 3 ). Since the endpoint is a function of exposure only during gestation, no subchronic to chronic UF was us ed. This developmental REL is within an order of magnitude of the chronic REL for kidney le sions, 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 endpoint s among different studie s, and pharmacokinetic similarities between humans and experimental animals. Isopropanol is metabolized through a similar pathway to acetone and CO 2

.

Weaknesses of the database for isopropanol include a lack of literature rega rding chronic toxicity endpoints in humans. The deficien cy of chronic toxicity cases in humans may be related to the relatively low chronic toxicity of isopropa nol. Another weakness is that, while most developmental studies observed maternal and feta l effects, only one study was performed via the

inhalation route.

Appendix D3 344 IsopropanolDetermination of Noncancer Chronic Re ference 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 i ndustrial chemicals in ai r 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 ma ximum admissible concentr ations of propyl and isopropyl alcohols in the atmosphe re. Gigiena I Sanitariia, 4:6-13.

Bates HK, McKee RH, Bieler GS, Gardiner TH , Gill MW, Strother DE, and Masten LW. 1994. Developmental neurotoxicity eval uation 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 isopropa nol 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 is opropanol 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 inhalati on oncogenicity study in Fisc her 344 rats and CD-1 mice. Fundam. Appl. Toxicol. 36(2):95-111.

Burleigh-Flayer HD, Gill MW, Strother DE, Ma sten 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 & Engine ering News. Production by the U.S. chemical industry. M. Heylin, ed., American Chemical Soci ety, Washington D. C., June 26:38-44.

Guseinov VG, and Abasov DM. 1982. Toxicological ch aracteristics of isopropyl alcohol after chronic exposure under experime ntal conditions. Azerb. Med. Zh

. 58:53-57.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version). Denver, CO: Microm edex, 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 IsopropanolDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Laham S, Potvin M, Schrader K, and Mari no I. 1980. Studies on inha lation toxicity of 2- propanol. Drug Chem. Toxicol. 3(4):343-360.

Lehman AJ, Schwerma H, and Rickards E. 1945 . Isopropyl alcohol acqui red 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, M acKenzie-Taylor DR, Khan A, Burg JR, Weigel WW, and Goad PT. 1988. Teratogenicity of n -propanol and isopropanol admi nistered 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 effect s in rats of twelve weeks’ dosing of 2- propanol, and neurotoxicity measur ed by densitometric measurements of glial fibr illary 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 simu ltaneous ethanol intake. Chem.-biol. Interact.

28:237-248.

Slauter R, Coleman D, Gaudette N, McKee R, Mast en 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, Slaute r RW, Gardiner TH, Strother DE, McKee RH, and Tyler TR. 1994. Developmental toxicity evalua tion 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 alco hol in rats (final report) w ith 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/O ffice of Toxic Substances. 1992a. A teratology study with isopropyl alcohol in rats. Letter submitting one encl osed teratology study and one enclosed supplementary study on isopropa nol. EPA/OTS Doc. #86-920000727. Prepared by

Appendix D3 346 IsopropanolDetermination of Noncancer Chronic Re ference 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 supplemen tary study on isopropanol. EPA/OTS Doc. #86- 920000727. Prepared by British Industrial Biolog ical Research Association. Submitted by

Chemical Manufacturers Association. Fiche No. NTIS/OTS0535604.

Wills JH, Jameson EM, and Coulston F. 1969. Eff ects on man of daily ingestion of small doses of isopropyl alcohol. Toxicol. Appl. Pharmacol

. 15:560-565.

Appendix D3 347 IsopropanolDetermination of Noncancer Chronic Re ference 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/m 3 (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 C 4 H 2 O 3 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 CCl 4 , 0.25 g/100 g ligroin Conversion factor 4.0 μg/m 3 per ppb at 25oC

III. Major Uses and Sources

Maleic anhydride is used as a chemical intermedia te in the synthesis of fu maric 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 manuf acture plywood, a corrosi on inhibitor, and a preservative in oils and fats. The annual stat ewide industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in Califor nia based on the most recent inventory were estimated to be 7366 pounds of ma leic anhydride (CARB, 2000).

Appendix D3 348 Maleic anhydrideDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

IV. Effects of Human Exposure

In many occupational situations workers are expo sed to mixtures of acid anhydrides, including maleic anhydride, phthalic anhydride, and tr imellitic anhydride. For example, Barker et al

. (1998) studied a cohort of 506 work ers exposed to these anhydrid es. In one factory, workers were exposed only to trimellitic anhydride, which has the lowest acceptable occupational exposure limit (40 μ g/m 3 ) of the three anhydrides. In th at factory there was an increased prevalence of sensitization to acid anhydride and work related respiratory symptoms with increasing full shift exposure even extending do wn to levels below th e current occupational standard. However, none of the workplaces ha d exposure only to maleic anhydride and a doseresponse relationship was not seen with mixed exposures.

The following reports involve expos ure only to maleic anhydride.

There are several case reports de scribing 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 concentra tions of maleic anhydride were 0.83 mg/m 3 in the inspirable particulate mass and 0.17 mg/m 3 in the respirable particulate mass. Bronchial provocation test ing was performed with phthalic anhydride, lactose, and maleic anhydride. Exposure of this individual to male ic anhydride (by bronchial provocation testing) at 0.83 mg/m 3 and 0.09 mg/m 3 in inspirable and respirable part iculate mass, respectively, showed a response of cough, rhinitis, and t earing within two minutes. With in 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 an emia. There was no evidence of pulmonary hemorrhage. Radioallergosorbent testing show ed specific IgE antibodies against human serum albumin conjugates with maleic anhydride, phthal ic 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 occu pational asthma due to exposur e 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/m 3 ) 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 concen trations of 0, 1.1, 3.3, or 9.8 mg/m 3 maleic anhydride for 6

Appendix D3 349 Maleic anhydrideDetermination of Noncancer Chronic Re ference 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 mon itored by gas chromatography as total maleic (maleic anhydride plus maleic acid). No exposure- related increase in mortality occurred. Of the species examined, only rats show ed significant changes in body we ight during the course of the experiment, with reductions among males in th e 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 symp toms was reported to increase with increased dose. No doserelated effects were observed in hematological para meters, 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 ep ithelium 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 expos ure 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/m 3 ) 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, pr egnant 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 necropsie d on gestational day 20. No statistically significant dose-related effects were observed in maternal we ight gain, implantation, fetal viability, post-implantation loss, fetal weight, or malformations. Groups of 10 male rats and 20 female rats/group (F 0 animals) were orally treated w ith 0, 20, 55, or 150 mg/kg-day maleic anhydride in corn oil to study multigeneration repr oductive toxicity. Animals within the same dose group were bred together after 80 days of treatment to produce two F 1 generation animals (F 1a and F 1b ) and animals from the F 1 generation were interbred to produce two F 2 generation animals (F 2a and F 2b ). A significant increase in mortality was observed among both F 0 and F 1

Appendix D3 350 Maleic anhydrideDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

generation animals in the high-dose group. Tota l body weight was significantly reduced in animals in the high-dose group at Week 11 of exposure for the F 0 generation males and females and at Week 30 of exposure in the F 1 generation males. No consistent pattern of dose- or

treatment-related effect on fertil ity, litter size, or pup survival was observed. Examination of F 0 animals showed necrosis of the renal cortex in the high-dose group (60% of males and 15% of females). Absolute kidney weight s were significantly increased in F 1 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 F 2 generation animals.

VI. Derivation of Chronic Reference Exposure Level (REL)

Study Short et al. , 1988 Study population Rats (15/sex/group), ha msters (15/sex/group), monkeys (3/sex/group) Exposure method Discontinuous inhalation exposure (0, 1.1, 3.3, or 9.8 mg/m 3 ) Critical effects Neutrophilic infiltration of the nasal epithelium; epithelial hyperplasia; respiratory irritation

LOAEL 1.1 mg/m 3

NOAEL Not observed in rats

BMC 05 0.12 mg/m 3 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 μ g/m 3 for the BMC 05 (0.12 x 6/24 x 5/7 x 1000) Human equivalent concentration 21 μ g/m 3 for the BMC 05 (Due to the lack of aerosol particle size data for the critical study, a human equivalent concentration could not be developed using r ecommended 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/m 3 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- a nd exposure related effect s, although mild and reversible, were observed at all exposure leve ls. Specifically, exposure to maleic anhydride vapors resulted in hyperplastic change in the nasa l 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 anhydr ide is a strong respiratory ir ritant to humans (ACGIH, 1992)

Appendix D3 351 Maleic anhydrideDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

suggests that this is a valid endpoi nt of toxicity to hu mans as well. Human exposure at levels as low as ~1 mg/m 3 appears to trigger acute asthmatic re actions 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 male ic anhydride are consider ed 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 benchm ark approach. The gamma model in the U.S. EPA's BMDS software yielded a BMC 05 of 0.12 mg/m3 for mild epithelial hyperplasia in male

and female rats combined. Because of the si milarities among species and the inclusion of monkeys in the study, an interspecies uncertainty factor of 3, rath er than 10, was used. Although there is no evidence of a toxic response similar to the development of asth ma in animals, the 1.1 mg/m 3 LOAEL from the animal studies of Short et al. (1988) results in a REL of 0.7 μ g/m 3
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 anhydr ide are the availability of multiple-species, multiple-dose subchronic inhalation studies, and th e 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 we re 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/m 3 . Thus the chronic REL of 0.7 μ g/m 3 should protect children. Maleic anhydride is a respiratory irrita nt 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 qua ntify a differential e ffect of maleic anhydride in children.

IX. References

ACGIH. 1992. American Conference of Gove rnmental Industrial Hygienists, Inc.

Documentation of the threshold limit values and biological exposur e 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 respirat ory symptoms among workers exposed to acid anhydrides: a cohort study. Occ up. Environ. Med. 55(10):684-691.

Appendix D3 352 Maleic anhydrideDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CARB. 2000. California Air Resources Board. Calif ornia 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 ai r 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 asth ma due to maleic anhydride]. Poumon et le

Coeur 36:393-395.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 7/31/95).

Jackson J, and Jones AH. 1993. Correspondence: H aemolytic 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 anhydrideDetermination of Noncancer Chronic Reference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# METHANOL

(methyl alcohol, wood spirit, ca rbinol, wood alcohol, wood naphtha)

CAS Registry Number: 67-56-1

I. Chronic Toxicity Exposure Level

Inhalation reference exposure level 4,000 μ g/m 3 (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 CH 3 OH 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/m 3

III. Major Uses and Sources

Originally distilled from wood, methanol is now manufactured syntheti cally 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 cons umer products such as windshield washing fluid and de-icing solutio n, duplicating fluids, solid canned fuels, paint remover, model airplane fuels, embalming fluids, lacquers, and i nks. Methanol is also used as an alternative motor fuel (HSDB, 1999). The annual statewid e industrial emissions from facilities reporting under the Air Toxics Hot Spots Act in Califor nia 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 metha nol toxicity in humans relates to acute rather than chronic exposure. The toxic effects after repeated or prolonged exposure to methanol are

Appendix D3 354 MethanolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

believed to be qualitatively similar but less seve re than those induced by acute exposure (Kavet and Nauss, 1990). These effects include CNS an d visual disturbances such as headaches, dizziness, nausea and blurred vision. The role of formate, a meta bolite of methanol, in chronic toxicity is unclear.

In one study, symptoms of blurred vision, headach es, dizziness, nausea an d skin problems were reported in teachers aides expos ed 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 dur ation was not mentioned. A dose-response relationship was observed between the self-reported amount of tim e 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/m 3 (365 to 3080 ppm) for a 15 minute sample.

Forty-five percent of duplicating machine ope rators experienced bl urred vision, headache, nausea, dizziness and eye irrita tion (NIOSH, 1981). Air concen trations of methanol for 25 minutes near the machines averaged 1330 mg/m 3

. Employees working in the proximity of direct process duplicating machines complained of frequent headaches and dizziness (Kingsley and Hirsch, 1954). Ai r concentrations of methanol ranged from 15 ppm (20 mg/m 3 ) to 375 ppm (490 mg/m 3 ).

Thirty young women, who had polished wood pencils with a varnish containing methanol, all experienced headaches, gastric disorders, ver tigo, nausea and blurred vision (Tyson, 1912; as cited in NIOSH, 1976).

None of the above studies specified th e workers’ total dur ation of exposure.

Ubaydullayev (1968) exposed 3 to 6 subjects to me thanol vapor for short durations (40 minutes for some subjects and others for an unspecified amount of time). Electrical brain cortex reflex activity was significan tly altered upon exposures to 1.17 mg/m 3 (0.89 ppm) or 1.46 mg/m 3 (1.11 ppm). No effect was observed at 1.01 mg/m 3 (0.77 ppm).

V. Effects of Animal Exposure

With the exception of non-human primates, the si gns 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 metha nol in non-primates (rodents, dogs , cats, etc) is CNS depression similar to that produced by other alcohols. Me tabolic acidosis and oc ular toxicity are not observed. The differences in toxicity are attribut ed to the ability of non-primates to metabolize formate more efficiently than humans and other primates (Tephly, 1991).

Two chronic studies have been conducted w ith 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 meth anol 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 MethanolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

inhalation to methanol concen trations ranging from 680 mg/m 3 (520 ppm) to 6650 mg/m 3 (5010 ppm) for 6 hours per day, 5 days per week for 4 weeks (Andrews et al. , 1987). No deaths

occurred and no treatment-re lated effects were found upon hi stopathologic examination. However, Andrews et al. did not examine possible neurologic or reproductive effects which have been observed in other species at lower c oncentrations (see S ections 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 ma le rats, inhalation exposure to methanol ranging from 260 to 13,000 mg/m 3 for 6 to 8 hours per day for either 1 day or 1, 2, 4 or 6 weeks resulted in a significant reduction in test osterone levels (Cameron et al. , 1984; Cameron et al. , 1985).

Ubaydullayev (1968) exposed rats (15 per group) to 0, 0.57, or 5.31 mg/m 3 methanol continuously for 90 days. Chronaxy ratios of fl exor and extensor muscles were measured in addition to hematologic parameters and acetyl cho linesterase activity. No changes were apparent in the 0.57 mg/m 3 group. Effects observed in the 5.31 mg/m 3 group included decreased blood albumin content beginning 7 weeks after exposur e, slightly decrease d acetylcholinesterase activity, decreased coproporphyrin levels in the urine after 7 w eeks, and changes in muscle chronaxy. (Chronaxy is the minimum time an electri c current must flow at a voltage twice the

rheobase to cause a muscle to contract. The rheo base is the minimal electric current necessary to produce stimulation (Dorland, 1981). Pregnant rats were exposed by inhalation to methanol at con centrations ranging from 5000 to 20,000 ppm for 7 hours per day on days 1-19 gestati on, and days 7-15 for the highest dose group (Nelson et al. , 1985). A dose-related decrease in fe tal weight, an increase in extra or rudimentary cervical ribs, and ur inary or cardiovascular defect s 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 vapor s 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 in creases in the incidence of exencephaly and cleft palate were observed at 5000 ppm and higher. A dose-rela ted increase in the number of fetuses per litter with cervical ribs (usually small ossifica tion sites lateral to the seventh cervical vertebra) was observed at 2000 ppm and above. The NOAEL was 1000 ppm.

Appendix D3 356 MethanolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

VI. Derivation of Chronic Reference Exposure Level (REL)

Rogers et al. (1993) Study Study population Pregnant mice Exposure method Discontinuous inhalation, 7 hours/day on days 6- 15 of gestation Critical effects Abnormal cervical ribs, ex encephaly, cleft palate

LOAEL 5000 ppm

NOAEL 1000 ppm Benchmark Concentration (BMC 05 ) 305 ppm Exposure continuity 7 hr/day Exposure duration 10 days Average experimental exposure 89 ppm at BMC 05 (305 ppm x 7/24) Human equivalent concentration 89 ppm atBMC 05 (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/m 3 , 4,000 μg/m 3 )

A NOAEL of 1000 ppm for developmental malfor mations 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 metha nol. Therefore, an uncertainty factor to

account for differences between subchronic and chronic exposures was not required. The investigators calculated maximum likelihood esti mates (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 inci dence of cervical ribs. The MLE 05 and BMC 05 for cervical ribs were 824 ppm (1079 mg/m 3 ) and 305 ppm (400 mg/m 3 ), respectively.

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the REL for metha nol are the observation of a NOAEL and the demonstration of a dose-response relationship. Th e 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 MethanolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

VIII. References

Andrews LS, Clary JJ, Terrill JB, and Bolte HF. 1987. Subchronic inha lation 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-Ne s 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. Circ ulating steroids in male rats following inhalation of n-al cohols. Arch. Toxicol. Suppl. 8:422-424

Dorland. 1981. Dorland’s Illust rated Medical Dictionary. 26 th ed. Philadelphia: W.B. Saunders.

Frederick LJ, Schlulte PA, and Apol A. 1984. Inve stigation and control of occupational hazards associated with the use of spirit dupli cators. Am. Ind. Hyg. Assoc. J. 45:51-55.

Garcia JH, and VanZandt JP. 1969. Nucleolar se gregation 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 visu al toxicity of methanol: V. The role of acidosis in experimental methanol poisoning. Am. J. Ophthalmol. 39:63-86.

Holmberg PC. 1979. Central nervous system defect s in children born to mothers exposed to organic solvents during pr egnancy. Lancet 2:177-179.

HSDB. 1999. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, Maryland. WWW database (htt p://sis.nlm.nih.gov/sis1/).

Kavet R, and Nauss KM. 1990. The toxicity of i nhaled methanol vapors. Crit. Rev. Toxicol.

21:21-50.

Kingsley WH, and Hirsch FG. 1954. Toxicologic c onsiderations in dir ect process spirit duplicating mahcines. Comp. Med. 6:7-8.

Nelson BK, Brightwell WS, MacK enzie DR, Khan A, Burg JR, Weigel WW, and Goad PT. 1985. Teratological assessment of methanol and ethanol at high inhala tion levels in rats. Fundam. Appl. Toxicol. 5:727-736.

NIOSH. 1976. National Institute for Occupationa l 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 c ited in Kavet and Nauss, 1990].

Appendix D3 358 MethanolDetermination 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 properti es of methanol as an atmospheric air pollutant (translation by B.S. Levine). U SSR Literature on Air Pollution and Related Occupational Diseases - A Survey. 17:39-45.

Appendix D3 359 MethanolDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (1 ppb) Critical effect(s) Histological lesions of th e 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 CH 3 Br 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/m 3 @ 25 ° C

III. Major Uses and Sources

Methyl bromide (MeBr) was used historically as an industrial fire ex tinguishing 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 ot her pests. Methyl brom ide is also used to fumigate soil before planting and fruits and ve getables after harvest. In 1981, 6.3 million pounds of MeBr were reportedly used in California (Alexeeff and K ilgore, 1983). By 1991, its use had grown to 18.7 million pounds in the state (Cal/E PA, 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 met hyl bromide during its use as a pesticide.

Appendix D3 360 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

IV. Effects of Human Exposure

Workers (n = 32) exposed to MeBr during fumigati on 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 sy mptoms and performed less well on several behavioral tests. Several confounding factors were pr esent in this study, including lack of adjustments for age, alcohol consumption, prescripti on medication, illegal drugs, edu cation, 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), ra bbits (104 per group), or female rhesus monkeys (13 per group) were expos ed to 0, 17, 33, 66, 100, or 220 ppm (0, 66, 128, 256, 388, or 853 mg/m 3 ) 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 paraly sis, were seen after exposure to 66 ppm in rabbits and monkeys. None of the species exhibite d adverse effects after exposure to 17 ppm.

Kato and associates (1986) obser ved focal lesions in the brain and heart in rats (10-12 per group) after inhalation of 150 ppm (585 mg/m 3 ) 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/m 3 ) MeBr 4 hours/day, 5 days/week for 6 weeks. In this experiment, rats exposed to any concentration of MeBr exhibited coronary lesion s, and exposures of 300 ppm or greater resulted in neurological dysfunction, including ataxia and paralysis. Testic ular atrophy was noted in 6 of the 8 animals exposed to 400 ppm.

Anger et al. (1981) determined that rabbits are more sens itive than rats to neurotoxicity of MeBr. In this study, rats or rabbits were exposed to 0 or 65 ppm (0 or 254 mg/m 3 ) MeBr for 7.5 hours/day, 4 days/week, for 4 weeks. Nerve co nduction 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 e xposure to 55 ppm (215 mg/m 3 ) MeBr for 36 weeks. Rabbits exposed to 26.6 ppm (104 mg/m 3 ) 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, hemato logy, 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 hear t in the form of cartilaginous metaplasia and thrombus in the males, and myocardial degeneration and thrombus in the

Appendix D3 361 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

females. Exposure of males to 30 or 90 ppm re sulted 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 concentr ation (3 ppm), very slig ht degenerative changes in the nasal epithelium, and ol factory basal cell hyperplasia we re noted in both sexes at 29 months. Based on this study, a LOAEL of 3 ppm (11.7 mg/m 3 ) 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/m 3 ) 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/m 3 ) 6 hours/day, 5 days/week. Hematological parameters and selected organ weights were measured in both species, in addition to histopat hological changes. Pseudocholinesterase activity a nd neurobehavioral tests were c onducted in the mice. Serious effects, including 58% body weight loss, 17% mort ality and severe curli ng 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 e ffects on several hematological pa rameters, 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/m 3 ) showed high mortality rates, loss in body weight and histol ogical changes in multiple organ systems including brain, kidney, na sal 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/m 3 ) 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 fe males 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 lo comotor 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/m 3 ) resulted in lesions in the nasal olfactor y 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 fo llowing a single exposure to 90 or 200 ppm (351 or 780 mg/m 3 ) 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 eff ects on spermatogenesis, sperm quality, or testes hi stopathology were noted.

Sikov et al. (1981) examined the teratoge nic potential of MeBr in ra ts and rabbits exposed to 0, 20, or 70 ppm (0, 78, or 272 mg/m 3 ) 7 hours/day, 5 days/week for 3 weeks during days 1-19 (rats) or 1-24 (rabbits) of gestation. No matern al 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 significan t maternal or fetal effects were observed at a concentration of 20 ppm.

Appendix D3 362 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Another developmental toxicity stud y 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/m 3 ) MeBr for 6 hours/day on gestation days 6-19. Maternal toxicity was observed at 80 ppm and included reduced body weight gain and si gns of neurotoxicity. In addi tion to the maternal effects observed, a significant increase in incidence of ga ll bladder agenesis and fused sternebrae were observed in the offspring exposed to 80 ppm. No adverse effects were ob served at 40 ppm or

lower concentrations.

A 2-generation reproduction and developmental toxi city 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/m 3 ) MeBr 6 hours/day, 5 days/week during premating, gestation, and lactation through 2 generations. Significant decreas es 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, th ere 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 F 1 group, reduced brain weight in 30 ppm F 1 females, and reduced fertility in the 30 and 90 ppm F 2b 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 respirator y effects, RGDR = 0.23,

based on MV = 0.03 m 3 /min, SA = 11.6 cm 2 )

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/m 3 , 5 μ g/m 3 )

The chronic REL for methyl brom ide is also the U.S EPA RfC.

Appendix D3 363 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the REL for methyl brom ide are the use of a comprehensive, long-term, multiple dose study with large sample sizes, a nd the availability of supporting data including long-term studies in other species and reproduc tive and developmental studies. The major

uncertainties are the lack of human data and th e lack of a NOAEL observation for the critical

effect.

The California Department of Pesticide Regulati on used a different approach that adjusts for respiration rate differences be tween humans and animals and wh ich uses 10-fold uncertainty factors for interspecies differences, for intr aspecies variability, a nd 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 μ g/m 3 ), respectively.

VIII. References

Alexeeff GV, and Kilgore WW. 1983. Methyl bromide. In: Re sidue Reviews. Vol. 88. New York: Springer-Verlag. pp. 102-153.

American Biogenics Corporation. 1986. Two-ge neration reproduction stu dy via inhalation in albino rats using methyl bromide. Final Repor t. 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 usi ng methyl bromide and sulfuryl fluoride. Neurotoxicology. 7(3):137-156.

Anger WK, Setzer JV, Russo JM, Bri ghtwell WS, Wait RG, and Johnson BL. 1981. Neurobehavioral effects of methyl bromide i nhalation 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 Chemi cal 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 Prot ection Agency. Letter to Carol Browner, Administrator, U.S.EPA, on methyl bromide uses in California.

HSDB. 1999. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland. WWW database (htt p://sis.nlm.nih.gov/sis1/).

Appendix D3 364 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Hurtt ME, Morgan KT, and Working PK. 1987. Hi stopathology of acute toxic responses in selected tissues from rats exposed by inhalati on to methyl bromide. Fundam. Appl. Toxicol.

9:352-365.

Hurtt ME, and Working PK. 1988. Evaluation of sp ermatogenesis 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 br omide. 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. Toxicol ogy and carcinogenesis studies of methyl bromide (CAS No. 74-83-9) in B6C3F1 mice (inh alation 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 App lied Scientific Research, Division for Nutrition and Food Research, T NO. EPA/OTS Document No. 86-8700001202.

Reuzel PGJ, Dreef-van der Meulen HC, Holla nders VMH, Kuper CF, Feron VJ, and Van der Heijden CA. 1991. Chronic inhalation toxicity an d carcinogenicity study of methyl bromide in Wistar rats. Food Chem. Toxicol. 29(1):31-39.

Russo JM, Anger WK, Setzer JV, and Bright well 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 butylen e oxide, styrene oxide and met hyl bromide. Battelle Pacific Northwest Laboratory, Richland, WA, for the Nati onal Institute for Occupational Safety and Health, Cincinnati, OH.

U.S.EPA. 1999. Environmental Protection Agency. Integrated Risk Information System (IRIS)

database. Reference concentrati on (RfC) for methyl bromide.

Appendix D3 365 Methyl BromideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

C

HRONIC T OXICITY S UMMARY


# METHYL CHLOROFORM

(1,1,1-trichloroethane, methyltrichloromethane)

CAS Registry Number: 71-55-6

I. Chronic Toxicity Summary

Inhalation reference exposure level 1,000 μg/m 3 (200 ppb) Critical effect(s) Astrogliosis in the sensorim otor cortex (brain) of gerbils Hazard index target(s) Nervous system

II. Chemical Property Summary (HSDB, 1999)

Description Colorless liquid Molecular formula C 2 H 3 Cl 3 Molecular weight 133.42 g/mol Density 1.3376 g/cm 3 @ 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/m 3 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 vinylid ene chloride and in textile processing and dry cleaning. The annual statewide in dustrial emissions from faci lities reporti ng under the Air Toxics Hot Spots Act in California, based on th e 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 pe ripheral neuropathy following 18 months of occupational exposure to methyl chlo roform in a solvent bath (House et al. , 1994). There was no identified exposure to agents known to cause peripheral neuropathy, su ch as n-hexane or

Appendix D3 366 Methyl ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

trichloroethylene. The worker reported that sh e wore protective gloves and a respirator, both of which frequently leaked. Seven months followi ng 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 de veloped distal sensory neuropathy after working with methyl chloroform in a degreasing operati on with repeated derm al exposure (Liss, 1988;

Howse et al. , 1989). Changes were observed in ne rve 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 me mory, intermediate memory, rhythm, and speed

based on the Luria-Nebraska Neur opsychological 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 co ngestion 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). Th e 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 (containi ng 65% methyl chloroform, 25% pr opane and dimethyl ether, and 10% inert ingredients) during which the worker reported often feeling dizzy or intoxicated. Three months following removal of the worker fr om exposure, liver function tests, although still abnormal, were significantly improved. Othe r 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 ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

was perceptible for the duration of the exposur e, no subjective symptoms were reported by the

volunteers.

V. Effects of Animal Exposure

Gerbils (4/sex/dose plus 24 sex-matched contro l 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 si gnificantly as a result of e xposure. Brain weights in the animals in the 1000 ppm dose group were significan tly decreased. Fibri llary 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) prot ein, the main protein subunit of astroglial filaments. Increased le vels of GFA protein we re 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 br ain: posterior cerebellar hemisphere, anterior cerebellar vermis, and hippocampus. The author s contended that depressions in DNA content reflect decreased cell density.

No evidence of peripheral neuropathy or other ne urotoxicity 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 observa tional test battery an d measured visual, somatosensory, auditory and caudal nerve-evoked pot entials. Histopathology of the brain, spinal cord, peripheral nerves and limb mu scles 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 hepato cellular necrosis (McNutt et al. , 1975). A statistic ally significant increase in liver weight per body mass was obs erved throughout the study. Electron microscopy revealed accumulation of triglyceride droplets in the centrilobular hepatocytes following one week of exposure to 1000 ppm me thyl chloroform. After 4 w eeks of exposure, cytoplasmic alterations in centrilobular hepatocytes include d 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 conf ounding 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 ch loroform was studied in CD-1 mice. Mice were exposed on gestation days 12 thr ough 17 to either 2000 ppm methyl chloroform for 17 hours per day or 8000 ppm methyl chloroform fo r 1 hour three times per day (Jones et al ., 1996). There

Appendix D3 368 Methyl ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

were no effects on pregnancy outcome, but expo sed pups has reduced we ight gain, had poorer

results on motor coordination tests and showed dela ys 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/m 3 ; 1,000 μg/m 3 )

VII. Data Strengths and Limitations for Development of the REL

Case reports indicate that the ne rvous system and the liver are targ ets 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 chlo roform exposure. The Kramer et al. (1978) study limited its evaluation to changes in blood pressure, heart rate , or electrocardiogram and exposure levels were only characteri zed 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 eval uated 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 associ ated 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 hepatotoxici ty 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 th e brains of gerbils exposed for 3 months to 210 ppm methyl chloroform (Rosengren et al. , 1985). A no-observed-advers e-effect-level (NOAEL)

Appendix D3 369 Methyl ChloroformDetermination of Noncancer Chronic Re ference 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 simila rly to 70 ppm methyl chlo roform, with slightly

decreased DNA content found in several discrete br ain regions of exposed animals. However, the relationship between tissue DNA c ontent and cell density as an in dication 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 ch loroform are the observation of the NOAEL and the continuous subchronic exposure regimen. Th e major uncertainties are the lack of human exposure data, the lack of dose-response inform ation, and the lack of comprehensive multi-organ

effects data.

VIII. References

ACGIH. 1992. American Conference of Gove rnmental Industrial Hygienists, Inc.

Documentation of the Threshold Limit Values and Biological Exposure Indices. Sixth edition. Cincinnati, OH.

ATSDR. 1990. Agency for Toxic Substances and Di sease 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/a qd/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 tran sient 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. Arc h. Intern. Med. 149:1793-1798.

House RA, Liss GM, and Wills MC. 1994. Peripher al sensory neuropathy associated with 1,1,1-

trichloroethane. Arch. Environ. Health 49:196-199.

Howse DC, Shanks GL, and Nag S. 1989. Periph eral neuropathy followi ng prolonged exposure to methyl chloroform. Neurology, 39:242. (Abstract)

HSDB. 1999. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland. WWW database (htt p://sis.nlm.nih.gov/sis1/).

Appendix D3 370 Methyl ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Jones HE, Kunko PM, Robinson SE, Balster RL. 1996. Developmental consequences of intermittent and continuous prenatal exposur e to 1,1,1-trichloroethane in mice. Pharmacol Biochem Behav 55(4):635-46.

Karlsson J-E, Rosengren LE, Kjellstrand P, a nd Haglid KG. 1987. Effects of low-dose inhalation of three chlorinated aliphatic or ganic solvents on deoxyribonucleic aci d 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, a nd 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 work ers exposed to 1,1,1-tric hloroethane. JAMA 260(15):2217.

Maroni M, Bulgheroni C, Cassitto G, Merluz zi F, Gilioli R, and Foa V. 1977. A clinical, neurophysiological and behavioral study of female workers expos ed 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-trichloretha ne vapor for 13 weeks. Neurotoxicol. Teratol.

15:313-326.

McLeod AA, Marjot R, and Monaghan MJ. 1987. Chr onic cardiac toxicity after inhalation of 1,1,1-trichloroethane. Br. Med. J. 294:727-729.

McNutt NS, Amster RL, McConnell EE, and Morri s F. 1975. Hepatic lesions in mice after continuous inhalation exposure to 1,1,1-tr ichloroethane. 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-trichlor oethane) 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 oncogeni city study in Fischer 344 rats a nd 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-trichl oroethane. 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 ChloroformDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# METHYL ISOCYANATE

(MIC, CH 3 -N=C=O)

CAS Registry Number: 624-83-9

I. Chronic Toxicity Summary

Inhalation reference exposure level 1 μg/m 3 (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 C 2 H 3 NO Molecular weight 57.06 g/mol Boiling point 39.5 o C Melting point

− 45 o C Vapor pressure 348 torr @ 20 o C, 600 torr @ 30 o C (Varma and Guest, 1993) Solubility 10 percent in water @ 15 o C Conversion factor 2.3 μ g/m 3 per ppb at 25 o C

III. Major Uses and Sources (Dave, 1985; U.S. EPA, 1986; HSDB, 1995)

Methylisocyanate (MIC) is pr epared industrially by reacti ng methylamine with phosgene, oxidizing monomethylformamid e at high temperatures ( > 550 o 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 car bamate 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/m 3 ) are exposed to approximately 460 μ g MIC in a workday. Based on the most recent inventory, the annual statewide industrial emissions from facilities reporting under the Ai r 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 isocyanateDetermination of Noncancer Chronic Re ference 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 ha ve occurred. A major exposure occurred in B hopal, India in December 1984. Because of the sudden, short-term release (30-45 minutes), no measurements occurred, but the air concentration was estima ted 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 ra pid, such hydrolysis in moist air is probably slow, and the reaction with photoc hemically produced hydroxyl radi cal is also slow (chemical

T 1/2 about 3 months) (U.S. EPA, 1986). Brown et al. (1987) have shown that the alkylisocyanates (e.g., MIC) are re latively resistant (compared to the arylisocyanates) to hydrolysis in water. Hence, despite the hi gh 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 documente d during the following decade (Lepkowski, 1994). The initial symptoms among the population liv ing 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, broncho spasm, and electrolyte imbalance (Varma, 1986). However extrapul monary damage, including tissue anoxia, gastrointestinal symptoms, and muscular weakne ss, were also observed (Dave, 1985). Within a year of the exposure, survivors continued to exhib it 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 sa me 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 pre gnant 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 gr oup of exposed Bhopal residents, 3-years following exposure to the MIC va pors. Loss of vision a nd 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. Si milarly, 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 bronchoalve olar area for long-term effects

Appendix D3 373 Methyl isocyanateDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

(Dave, 1985, Varma, 1986). Vijayan et al. (1995) studied cellular components of bronchoalveolar lavage (BAL) and pulmonary fu nction in Bhopal patients 1.3, 2.7, and 5.1 years after exposure to MIC. All had lived within 3- miles of the factory a nd 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 obs erved. In general, the decrements ranged from 12 - 21 percent of predicted valu es, whereas the contro l 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 numbe rs of macrophages (all exposure groups), decreased percentage of lymphocytes (2.7 and 5.1 year groups), and in creased numbers and per centage of neutrophils (5.1 year group). These cell types are involved, through the secretion of various factors, in inflammatory and immunologic processes in th e 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 inhalatio n to MIC may experience long-term (as well as immediate) damage to pulmonary and extrapulmona ry systems. The lung is probably the critical target organ for long-term effects from acute ex posure, although adverse effects on other organs (e.g., eye, reproductive, and gastroin testinal) also exist. The late responses to the acute exposure suggest an immunological component , which could involve severa l 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. Seri al 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 supervis ors and coworkers. The frequency of pulmonary impairment was evaluated for the assumed four le vels of exposure. No specific or consistent pulmonary impairment was evident. Unfortunately the report gave no quan titative 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 expos ure has been described as acute because of the short duration (30-45 min). No studies were found that desc ribed 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 pres ence of carbamylated hemoglobin in MIC-exposed people, and the change from edematous to infla mmatory 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 re spiratory tract and di stributed to non-respiratory organs in experimental animals. In an acute ( 30 min) inhalation exposure to a dose of 14 C-MIC (labeled in

Appendix D3 374 Methyl isocyanateDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

the isocyanate moiety) equivalent to one-LC 50 (23 mg/L), rats accumulated protein-bound radioactivity (including carbamyl ated 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 14 C-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 labe l decreased to 2-7 percent.

MIC was retained in the nasal-la ryngeal area of the guinea pigs.

MIC, like reactive isocyanates in general, can react with biol ogical 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 resist ant (compared to aryl isocyanates) 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 subc utaneous exposure of rats to either hydrolytic product at levels equivalent to the LC 50 or LD 50 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 coul d lead to subsequent inflammatory responses if sufficient amounts are present.

A role for methylamine in reproductive/developme ntal toxicity was investigated by Guest and Varma (1991). In a mouse study, pregnant dams we re exposed to varying doses (intraperitoneal) of methylamine (as well as the di- and trimethyl compounds). Reproductive toxicity was not observed for methylamines. However, in cultu red 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 appropria te functional groups of proteins, peptides, and nu cleic acids. However, i n 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 glutathi one (GSH) has been suggested by the discovery of the MIC-adduct, S-(N-methylcarbamoyl)glutath ione (SMG), in the bile and the MIC-adductof 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(CH 3 ) 2 ) excrete AMCC in urine (Mraz and Nohova, 1992). Hence a reversible MIC-transport system in animals, including humans, is possibl e, and the presence of high levels of GSH in human lavage fluid (Cantin et al. , 1987) would permit the initi ation of this mechanism.

Appendix D3 375 Methyl isocyanateDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

The toxicity of the adduct SMG was te sted 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-depe ndent (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 occurre d in the presence of 0.25 mM SMG. Similar to the results obtained with methylamine, the SM G 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 demonstr ate the reversibility of the binding between MIC and GSH.

Three inhalation studies were id entified in which experimental animals were exposed to more than one dose of MIC. Among these studies, two us ed 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-e xposure 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 deat h, 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 norma l weight gain returned by 1 week following cessation of exposure. At 7 days post-exposur e, microscopic changes were observed in therespiratory system among the high dose rats of both sexes. Between 8- and 27 days postexposure, increased lesions in th e 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-expo sure. Among survivors, the incidence of lesions decreased to control values by 91 days . Among the mice, treatment related changes in the respiratory tract were obser ved at the high dose at 7 days post-exposure. Between 28 and 91 days, the lesions associated with the upper respiratory tract disa ppeared, whereas those associated with the major bronchi remained, alth ough 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 de creased weight gain and respir atory 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 ra ts in the 3.1 ppm groups, whereas weight among the rats in the 0.15 and 0.6 ppm MIC groups was i ndistinguishable 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 in crease in hemoglobin concentra tion 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 pp m 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 a nd testes absolute weights were observed in this exposure group, but the authors in terpreted these data as a reflection of the body

Appendix D3 376 Methyl isocyanateDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

weight losses. No weight changes were obser ved 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, immediatel y following cessation of exposure.

Post-exposure changes in lung pathology also occu rred in the rats surviv ing 3.1 ppm in the Dodd and Fowler (1986) study. The early lesions asso ciated with inflammati on, epithelial necrosis, squamous metaplasia, and epithe lial hyperplasia extending to the br onchioles 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 progres sive 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 evaluati on was not presented. No post- exposure deaths were reported, although lethality was recorded for rats expos ed to 3.5 and 35 ppm for only 10 minutes. Following the 0.5 hr × 6-day exposure, the weight gain de clined 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 sa me 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 pr esent; their extent increased with dose. A dose-response increase in markers of lung infection was pr esent and suggests that the MIC exposed rats were more prone to infec tious 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 re sult of the lung infections.

Fetotoxicity was observed in two expe rimental animal studies (Schwetz et al. , 1987; Varma, 1987). Among female mice exposed to 0, 1, or 3 ppm MIC during gesta tion 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/l itter decreased relativ e to the air-exposed controls. The dams were unaffected in terms of survival, body weight , or length of gestati on. Non-gestational exposure (6 hr/day, 4 days) did not affect the number of pregnancies or the live litter sizes, suggesting that the fetotoxic eff ect may be specific to the female reproductive tract rather than a general attribute of systemic toxicity. Sim ilarly, 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 fetuse s. Maternal mortality and decreased skeletal lengths were also observed at 9 and 15 ppm. A distinction between maternally induced fetotoxicity and a direct eff ect 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 isocyanateDetermination of Noncancer Chronic Re ference 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 incr eased after 15 days post-exposure (Agarwal and Bose, 1992). There was no evidence of fetotoxi city 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, fertilit y, or litter size (Schwetz et al , 1987). The results suggest that exposures to MIC at doses that are not toxic to a dult 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 ga in 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 cm 2 ) 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/m 3 )

Although the exposure was for only 10 days, the Dodd and Fowler (1986) study includes the longest exposure duratio n of the available investigations and also uses some of the lower exposure levels (down to 0.15 ppm). The microscopic findings of the re spiratory tract were statistically analyzed, although an observation of the tabulated da ta at the four doses (0, 0.15, 0.6, and 3.1 ppm) clearly shows a NOAEL of 0.6 pp m. Other endpoints with the same NOAEL were increased hemoglobin and in creased 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 la tter study did not

Appendix D3 378 Methyl isocyanateDetermination of Noncancer Chronic Re ference 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. Ho wever, 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 endpoi nts observed within 1 day of cessation of exposure. Post-exposure evaluation showed th at, at a higher exposure level (3.1 ppm), progressive changes, including de ath, occurred. Post-exposure observations, however, were not reported at the 0.15 and 0.6 ppm MIC levels. The at tribute of delayed MIC inhalation toxicity has also been observed in other experime ntal animals studies (Dodd and Fowler, 1986,

Mitsumori et al. , 1987). In the case of the human MIC e xposure 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 maximu m occurred at 2 - 3 days, and by the end of a week about 2500 deaths were documented (Dav e, 1985; Varma, 1986; Varma and Guest , 1993), although Varma (1986) suggests that the immediat e number may be closer to 5,000. One report suggests that during the intervening decade as ma ny as 6,000 deaths may be attributed to the initial exposure in Bhopal (Lepkowski, 1994). Such informati on suggests that the presence of an

adverse effect at the NOAEL of 0.6 ppm (Dodd and Fowler, 1986) mi ght 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 (f emales) 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/m 3 ) was derived. The REL based on the Mitsumori et al. (1987) study is si milar 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 incl ude 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 inhala tion 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 μ g/m 3 should adequately protect infants and children. MIC is a respiratory irrita nt 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 isocyanateDetermination of Noncancer Chronic Re ference 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, Maha shabde S, Tiwari MK, Muir MK, Mehra V, Ashiru K, and Mackenzie DC. 1990. Delayed eye and other conse quences from exposure to methyl isocyanate: 93% follow up of exposed and unexposed cohor ts in Bhopal. Br. J. Ind. Med. 47:553-558.

Avashia B, Battigelli MC, Morgan WK and Re ger RB. 1996. Effects of prolonged low exposure to methyl isocyanate. J. O ccup. Environ. Med. 38(6):625-630.

Baillie TA, and Slatter JG. 1991. Glutathione: a vehi cle for the transport of chemically reactive

metabolites i n vivo . Acc. Chem. Res. 24:264-270.

Bhandari NR, Syal AK, Kambo I, Nair A, Be ohar V, Saxena NC, Dabke AT, Agarwal SS, and Saxena BN. 1990. Pregnancy outcome in women expos ed to toxic gas at Bhopal. Indian J. Med. Res [B]. February. pp. 28-33.

Bhattacharya BK, Sharma SK, and Jaiswal DK. 1988. In vivo b inding of [1- 14 C]methylisocyanate to various tissue pr oteins. 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. En viron. Health Perspect.

72:5-11.

CARB. 2000. California Air Resources Board. Calif ornia Emissions Inventory Development and Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

Cantin AM, North SL, Hubbard RC, and Crysta l RG. 1987. Normal alveol ar epithelial lining fluid contains high levels of glut athione. 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 Ch emicals: Detection a nd Protection Methods. Schiefer, H.B. ed. Saskatoon, Saskatchewan, Canada. pp 1-38.

Dodd DE, and Fowler EH. 1986. Methyl isocyanate subchronic vapor in halation studies with

Fischer 344 rats. Fundam. Appl. Toxicol. 7:502-522.

Ferguson JS, Kennedy AL, Stock MF, Brown WE, and Alarie Y. 1988. Upta ke and distribution

of 14 C during and following exposure to [ 14 C} methyl isocyanate. T oxicol. Appl. Pharmocol.

94:104-117.

Appendix D3 380 Methyl isocyanateDetermination of Noncancer Chronic Re ference 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 toxici ty 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. Comparat ive toxicity of methyl isocyanate and its hydrolytic derivatives in rats. I. Pulmonary hi stopathology 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 isoc yanate 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. Ea rly 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 studie s in mice. Environ. Health Perspect. 72:149-

152.

Sethi N, Dayal R, and Singh RK. 1989. Acute and s ubacute toxicity stud y 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-med iated 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 me thyl carbamylation in post-mortem blood. Med. Sci. Law. 31:289-293.

Appendix D3 381 Methyl isocyanateDetermination of Noncancer Chronic Re ference 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 populati on 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 Crite ria and Assessment Office, Office of Research and Devel opment. Cincinnati OH: U.S. EPA.

Varma DR. 1986. Anatomy of the me thyl isocyanate leak on Bhopa l. 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 e ffects of methyl isocyanate on the course of pregnancy. Environ. Health Perspect. 72:153-157.

Varma DR, and Guest I. 1993. The Bhopal accident a nd 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 isocyanateDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# METHYL t -BUTYL ETHER

(MTBE; 2-methoxy-2-me thylpropane; tert-butyl methyl ether; methyl 1,1dimethyl ether)

CAS Registry Number: 1634-04-4

I. Chronic Toxicity Summary

Inhalation reference exposure level 8000 μ g/m 3 (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 C 5 H l2 0 Molecular weight 88.15 g/mol Density 0.7405 g/cm 3 @ 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/m 3 @ 25 ° C; 3.67 mg/m 3 @ 20 ° C

III. Major Uses or Sources

Methyl t-butyl ether (MTBE) is us ed as a gasoline additive to improve octane ratings and reduce emissions of some pollutants, in industry to im prove miscibility of solvents, and in clinical medicine to dissolve cholesterol gall stones (Yoshikawa et al. , 1994). The annual statewide industrial emissions from facilities reporti ng 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 fati gue at the end of the work week than before the work week, and t hose 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

and dyspnoea at the end of work week. No hu man chronic toxicity or chronic epidemiology

information for MTBE without coexposure to gasoline was found.

Ten healthy male volunteers undergoing light phys ical 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 concentr ation-related effects on reported ocular or nasal irritation. The blockage index (a measure of nasa l airway resistance) incr eased significantly after exposure but was not correlated with exposure co ncentration.

V. Effects of Animal Exposure

Male and female rats (50/sex/group) were expos ed 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/m 3 ) MTBE for 24 months (Chun et al. , 1992). Clinical signs, hematology, body weights and food consumption were monitored. Necropsy incl uded measurements of organ weights and histopathology. Corticosterone le vels were measured on 10 animal s prior to sacrifice. Serum enzymes were not monitored. The NOAEL for several endpoints, including non-alpha-2 μ

globulin induced nephrotoxicity, in creased relative liver and kidne y weights and prostration in females, and periocular swelling in both sexes was 403 ppm (1453 mg/m 3 ).

Mice were exposed for 6 hours/day, 5 days/week fo r 18 months to MTBE concentrations of 0, 402, 3014, or 7973 ppm (0, 111, 835, or 2208 mg/m 3 ) (Burleigh-Flayer et al. , 1992). The mice exposed to the highest concentration (7973 ppm) a ll exhibited ataxia. Prostration was also noted in 8 of 50 animals in this group. Liver weight s were elevated in a concentration-dependent manner in the female mice but this change was not significant at the lo west concentration (402 ppm). Kidney weights were elevat ed in the female mice exposed to 7973 ppm. At the highest concentration, a significant increase in hepato cellular 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 B6C3F 1 mice to 7924 ppm (2195 mg/m 3 ) MTBE for 4 months, or 7919 ppm (2194 mg/m 3 ) 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 re duced weight gain in MTBE-exposed mice was significantly different from the controls at bo th time points. In MTBE-exposed mice, mean uterine weight was 83% reduced re lative to controls at 4 and 8 m onths. 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 a nd histological changes in the reproductive organs were also noted. Although the changes in orga n weights and histology we re suggestive of an anti-estrogenic effect of MTBE, serum estrogen le vels were unaffected. No changes in estrogen receptor (ER) immunoreactivity in reproductive system tissues were observed. Experiments in

vitro failed to demonstrate any inhibition of estrad iol binding to ER by MTBE or its metabolites. No inhibition of ER by MTBE wa s detected, nor was there any inhi bition of the induction of ER by estradiol. The authors conclude d that the apparent anti-estrogenic effect s of MTBE were not

Appendix D3 384 Methyl t-Butyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

mediated via the ER, and drew a parallel with the anti-estrogenic e ffects of dioxins and chlorinated biphenyls.

Tests for histopathology in the resp iratory tract, plasma corticoste rone 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/m 3 ), 6 hours/day, 5 days/week for 13 weeks (Dodd and Kintigh, 1989). Of these endpoints, th e most significant finding was an elevation in plasma corticosterone in the high dose group. Th is 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/m 3 ) 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/m 3 ) or 2860 ppm (10,311 mg/m 3 ) MTBE exhibited slightly decreased pup survival.

Neeper-Bradley (1991) expos ed rats to 0, 402, 3019, or 8007 ppm (0, 111, 836, and 2218 mg/m 3 ) 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 gest ation and lactation. Parental effects of MTBE exposure were observed, including ataxia, blepharospasm, lack of startle reflex, and increased relative liver weights (F 1 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. Re duced survivability to postnatal day 4 was observed in the 8007 ppm group. No adverse e ffects were noted at the 403 ppm (111 mg/m 3 )

concentration.

In a developmental and reproductive toxicity study, Conaway and associates (1985) found no significant increases in maternal or fetal toxi city, 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/m 3 ).

Maternal toxicity, in the form of hypoactivity and ataxia, was obs erved in pregnant mice exposed during gestation to 4076 ppm (14,690 mg/m 3 ) MTBE (Bushy Run Research Center, 1989a). Significant reductions in food inta ke and body weight gain were observed in dams exposed to 8153 ppm (29,390 mg/m 3 ). Fetal body weight was significantly reduced in the 4076 ppm group, and there were significant increases in the inci dences of skeletal variations and unossified phalanges in the 4076 and 8153 ppm groups. Pregnant rabbits exposed to si milar concentrations during gestation showed no significa nt maternal or fetal toxicity or developmental toxicity up to a concentration of 8021 ppm (28,918 mg/m 3 ) (Bushy Run Research Center, 1989b).

Appendix D3 385 Methyl t-Butyl EtherDetermination of Noncancer Chronic Re ference 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-bo dy inhalation exposures (0, 403, 3023, or 7977 ppm)

Critical effects Nephrotoxicity, increased liver and kidney weight, prostration a nd 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/m 3 , 8000 μ g/m 3 )

The USEPA (1995) based its RfC of 3000 μ g/m 3 on the same study but included a Modifying Factor (MF) of 3 for database deficiencies. The criteria for use of modify ing factors are not well specified by U.S. EPA. Such modify ing 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 th e 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. T oxicol. Ind. Health 3:519-534.

Bird MG, Burleigh-Flayer HD, Chun JS, Douglas JF, Kneiss JJ, Andrews LS. 1997. Oncogenicity studies of inhale d 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 EtherDetermination of Noncancer Chronic Re ference 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, Uni on 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 i nhaled methyl tertiary butyl ethe r 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/OT S 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 materi al). Prepared for the MTBE committee by Bushy Run Research Center, Uni on 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. T oxicol. Environ. Health 16(6):797-809.

Dodd DE, and Kintigh WJ. 1989. Met hyl tertiary butyl ether (MTB E): Repeated (13 week) vapor inhalation study in rats with neurotoxicity eval uation (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, Pul kkinen P. 1997. Changes in neuropsychological symptoms and moods among tanker drivers exposed to gaso line during a work week. Occup Med (Lond) 47(6):344-8.

HSDB. 1999. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland. WWW database (htt p://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 reproducti on study of inhaled me thyl 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Nihlen A, Walinder R, Lof A, Johanson G. 1998. Expe rimental exposure to methyl tertiary-butyl

ether. II. Acute effects in humans. T oxicol. Appl. Pharmacol. 148(2):281-287.

U.S.EPA. 1999. U.S. Environmen tal Protection Agency. Integrat ed Risk Information System (IRIS) database. Reference concentration (RfC) for methyl t-butyl ether (MTBE).

Yoshikawa M, Arashidani K, Katoh T, Ka wamoto T, and Kodama Y. 1994. Pulmonary elimination of methyl tertiary butyl ether afte r intraperitoneal administration in mice. Arch.

Toxicol. 68:517-519.

Appendix D3 388 Methyl t-Butyl EtherDetermination of Noncancer Chronic Re ference 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/m 3 (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 CH 2 Cl 2 Molecular weight 84.93 Density 1.32 g/cm 3 @ 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/m 3 @ 25° C

III. Major Uses and Sources

Methylene chloride (MC) is used in paint and va rnish remover, in aerosols as a cosolvent or vapor pressure depressant, and in solvent degreasing and metal cl eaning. It is also used in plastics processing and in extrac tion of fats and oils from fo od products (HSDB, 1999). The

annual statewide industrial emissions from fac ilities 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 m ean and maximum monitored ambient methylene

chloride concentrations have decreased s lightly 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/m 3 ) or more and increased blood carboxyhemoglobin

Appendix D3 389 Methylene ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

(COHb) content at lower concentrations (500 ppm) due to metabolis m 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 par titioning into fat and its slow release into circulation with subsequent metabolism, lead ing to formation of carbon monoxide (Engstrom and Bjurstrom, 1977). In situations of chronic exposure, carbon monoxide toxicity is also of concern. Barrowcliff (1978) documen ted the case of an adult ma le who developed an unsteady gait, a peculiar dysarthria and a loss of memor y. The man had worked with 15-50 liters of methylene chloride daily for 3 years in a poorly ve ntilated room while clea nsing 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). Neurobehavior al 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 re sting 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 elevati on in COHb concentra tions 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 methyl ene chloride and a group of work ers 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 th e work-shift. Ambient concentrations in the workplace were estimated from 57 samples, wh ich ranged from 0 to 250 ppm, with a mean concentration of 40 ppm. Three exposed worker s also wore monitors to estimate personal exposures. The time-weighted average exposure fo r 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 exer cise workloads resulted in higher COHb levels.

Soden et al. (1996) showed a dose-response increase in carboxyhemoglobin levels in nonsmokers with increasing methylene chloride exposu re 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 st udy was not reported.

Appendix D3 390 Methylene ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Although animal studies have shown COHb-i nduced 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-indu ced angina report a decrea se 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 m odel of MC and CO estimated that a 1-hour exposure to 340 ppm (1200 mg/m 3 ) 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 associat es (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 o ccurring. The odds ratio for methylene chloride exposure was 1.0 to 5.7 (average = 2.3; p < 0.06). Th ere was a significant effect of exposure to 4 or more solvents, compared with age-matched c ontrols (p < 0.05). The concentrations of MC were not reported in the study.

The U.S. Occupational Safety and Health Administ ration reduced its permi ssible 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 exposur e to 0, 50, 200, or 500 ppm MC for 6 hours/day, 5 days/week resulted in significant hi stopathologic lesions in the livers of rats exposed to 500 ppm. No signifi cant adverse effects were observe d at 200 ppm or lower. The predominant hepatocellular lesion was fatty vacuolization of hepatocytes.

Female B6C3F1 mice inhaling 2000 ppm MC fo r 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 path ological 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). Li ver/body weight ratios were unaffected at any

Appendix D3 391 Methylene ChlorideDetermination of Noncancer Chronic Re ference 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 co ncentration, 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 or gans of hamsters at each of the three MC concentrations. Overall, effects were more poten t 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. Car boxyhemoglobin 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 conti nuous 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 e xposed to 100 ppm MC had COHb levels of approximately 4% compared with 2% in the dogs.

Oral ethanol pretreatment in rats has been s hown to suppress the COHb formation characteristic of methylene chloride exposure through inhibition of bi otransformation of methylene chloride (Glatzel et al. , 1987).

Gerbils (10/sex per group; 60 controls) exposed continuously to MC c oncentrations of 210, 350, or 700 ppm for a period of 3 months, with a 4- month follow-up period, showed irreversible

cellular and biochemical cha nges in brain (Rosengren et al. , 1986). A high mortal ity 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 hi gh 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 af ter 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 elevat ed astroglial proteins in the frontal and sensory motor cerebral cortex, consis tent 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 s howed that the formation of carboxyhemoglobin did not increase in gerbils between the 210 and 350 ppm exposures, indicating that the meta bolism of MC to CO

is saturable at concentrations below those in the study. On the other hand, the neurotoxic brain

biochemical alterations were si gnificantly greater in gerbils e xposed to 350 ppm as compared with the 210 ppm group, implying that carboxyhem oglobin 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 ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

pigmentation in the liver in a dose-depe ndent fashion, beginn ing 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 e xposed 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 inciden ce of liver cytologic de generation and splenic atrophy at 4000 ppm (males) (N TP, 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 pi gs for 7 hours/day, 5 da ys/week resulted in 3 deaths; 2 showed moderate centrilobular fa tty 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, co mpared 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, no r were blood pressure or hematological values altered at the 5000 ppm concen tration. 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 tuberc ulosis lesions, but no other histological alterations. Four out of 6 guinea pigs had mode rate fatty degeneration of the liver at this

concentration.

The offspring of rats (10 dams per group) expos ed 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 observe d 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 di d 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 addi tion, significantly increased incidences of delayed

ossification of the sternebrae were seen in both species, compared to controls.

Methylene chloride exposure of female rats be fore or during gestati on to 4500 ppm resulted in elevated maternal liver weights and decreased birth weights of the offspring, but no terata or

skeletal/soft tissue anomalie s (Hardin and Manson, 1980).

Appendix D3 393 Methylene ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

A 2-generation reproduction test was conducte d by Dow Chemical Company (Nitschke et al.

, 1985) which showed no significant re productive or developmental e ffects 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 F 0 and F 1 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/m 3 ; 400 μg/m 3 )

Workers were exposed to average measured co ncentrations of 40 ppm during the workday, and the personal monitors on 3 of the subjects indicated a 8-hour tim e-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 concentra tions of above 2% are consider ed to aggravate angina in some individuals (CARB, 1982). In effect , 2% COHb can be considered a NOAEL for aggravation of angina. Theref ore, 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 experime ntal exposures using vol unteers, therefore an uncertainty factor for subchronic exposure was not necessary. A number of factors contribute to the uncertainty in determining the degree of sens itivity to methylene chloride, including activity level, metabolic enzyme activ ity, age, and background COHb st atus (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 follo wing a 14-week exposure. These results are consistent with the LOAEL reported in the DiVi ncenzo and Kaplan study. However, the human occupational study likely contains less uncertainty, since the toxicokinetics of the effect,

Appendix D3 394 Methylene ChlorideDetermination of Noncancer Chronic Re ference 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 interspe cies uncertainty factor for residual uncertainty not accounted for in the HEC calculation, and a 10-f old intraspecies uncerta inty factor, a REL of 300 ppb or 1000 μg/m 3 was derived. Thus, the REL derived from the best available animal study is comparable to the 400 μg/m 3 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 observa tion, 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 Ka plan, 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/m 3 vs. 1,000 μg/m 3 ) are comparable. That both the human and animal studies measured the same endpoint and arrived at similar conclusi ons is a circumstance that is rarely found but one that considerably increase s the weight of evidence from which the REL was derived. The two studies complement each othe r, 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 hum an study involved poorly characterized human exposures but lacks the uncertainty inhere nt in interspecies extrapolation.

VIII. References

Allred EN, Bleecker ER, Chaitman BR, Da hms TE, Gottlieb SO, Hackney JD, Hayes D, Pagano M, Selvester RH, Walden SM, and Wa rren J. 1989. Acute effects of carbon monoxide exposure on individuals with corona ry artery disease. Health E ffects Research Institute (HEI) Research Report No. 25., Cambridge, MA: HEI.

ACGIH. 1991. American Conference of Government al Industrial Hygienists. Documentation of Threshold Limit Values and Biological Exposure Indices. 6th edition Ci ncinnati, OH: ACGIH.

Andersen ME, Clewell HJ III, Gargas ML, MacNaughton MG, Reitz RH, Nolan RJ, and McKenna MJ. 1991. Physiologically based pharmacoki netic modeling with di chloromethane, its metabolite, carbon monoxide, and blood carboxyhemogl obin in rats and humans. Toxicol. Appl.

Pharmacol. 108:14-27.

Appendix D3 395 Methylene ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Aviado DM, Zakhari S, Watanabe T. 1977. Meth ylene chloride. In: Go lberg 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. Be havioral toxicity in the offspring of rats following maternal exposure to dichloromethane. Toxicol. Appl. Pharmacol. 52:29-37.

Breslin WJ, and Landry TD. 1986. Methylene chlo ride: Effects on estrous cycling and serum prolactin in Sprague-Dawley rats. Unpublishe d 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 chlo ride: 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 Mono xide (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. Toxico l. Appl. Pharmacol. 59:130-140.

DiVincenzo GD, and Kaplan CJ. 1981b. Effect of exercise or smoking on the uptake, metabolism, and excretion of methylene chlori de 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 syst em disturbances in men and animals experimentally exposed to indus trial 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 ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Hallier E, Schroder KR, Asmuth K, Dommermu th A, Aust B, Goerens HW. 1994. Metabolism of dichloromethane (methylene chloride) to form aldehyde in human erythr ocytes: influence of polymorphism of glutathione transferase th eta (GST T1-1). Arch. Toxicol. 68:423-427.

Hardin BD, and Manson JM. 1980. Absence of dich loromethane teratogeni city with inhalation exposure in rats.Toxicol Appl Pharmacol 1980 52(1):22-28.

Haun CC, Vernot EH, and Darmer KI. 1972. Con tinuous animal exposure to low levels of dichloromethane. Aerospace Medical Research La boratories, Wright-Patterson Air Force Base, Dayton, OH. AMRL-TR-72-130.

HSDB. 1999. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland. WWW database (htt p://sis.nlm.nih.gov/sis1/).

Heppel LA, Neal PA, Perrin TL, Orr ML, and Porterfield VT. 1944. Toxicology of dichloromethane (methylene chlori de). I. Studies on effects of daily inhalation. J. Ind. Hyg. Toxicol. 26(1):8-16.

Kanno J, Foley JF, Kari F, Anderson MW, Mar onpot RR. 1993. Effect of methylene chloride inhalation on replicative DNA synt hesis in the lungs of female B6C3F1 mice. Environ Health Perspect 101 Suppl 5:271-276.

Kleinman MT, Davidson DM, Vandigraff RB, Ca iozzo VJ, and Whittenberger JL. 1989. Effects of short-term exposure to carbon monoxide in subj ects with coronary arte ry disease. Arch. Env. Health 44(6):361-369.

NTP. 1986. National Toxicology Program. Toxi cology 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 chlori de: Two -generation reproduction study in Fischer 344 rats. Midla nd, 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 t oxicity 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 afte r long term exposure: a quantita tive study of DNA and the glial cell marker proteins S-100 and GFA. Br. J. Ind. Med. 43:291-299.

Appendix D3 397 Methylene ChlorideDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Schwetz BA, Leong KJ, and Gehring PJ. 1975. The effect of maternally inhaled trichloroethylene, perchloroet hylene, 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. Carboxyhe moglobin 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). Fi nal 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 lo w dose level. Proceedings of the 3rd Annual Conference on Environmental Toxicology, 25, 26, and 27 October, 1972. Aerospace Medical Research Laboratory, Aerospace Medical Divi sion, Air Force Systems Command, WrightPatterson Air Force Base, OH.

Appendix D3 398 Methylene ChlorideDetermination of Noncancer Chronic Re ference 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/m 3 (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 C 13 H 14 N 2 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, benz ene, ether; 273 g/100 g acetone; 0.1 g/100 g water @ 25 ° C Conversion factor 8.1 μg/m 3 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, 19 95; ACGIH, 1992). The

annual statewide industrial emissions from fac ilities reporting under the Air Toxics Hot Spots

Act in California based on the most recent inve ntory were estimated to be 1133 pounds of MDA (and its dichloride) (CARB, 2000).

IV. Effects of Human Exposure

Several cases of human exposur e to MDA have identified th e compound as a hepatotoxicant which produces cholesta tic 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 dianilineDetermination of Noncancer Chronic Re ference 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 wo rkers initially reported to the hospital with symptoms of abdominal pain three days after th e most recent exposure and all were discharged within four days. Two workers continued to show severe sympto ms five days after the onset, with abdominal pain, jaundice, a tender liver, nausea, dyspnea, and muscular pain. Plasma bilirubin, alkaline phospha tase, and aspartate aminotransferase levels were elevated. Some symptoms did not subside until tw o months after the onset. One worker, after another exposure, experienced nausea, abdominal pain, and muscul ar pain. A second worker reported further symptoms of headache, tire dness, and decreased libido.

Williams et al. (1974) reported symptoms in 6 of a pproximately 300 workers exposed to MDA by surface coating concrete walls with epoxy resi ns. 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, a nd one liver biopsy showed bile stasis. All the workers recovered completely after an unspecified time.

McGill and Motto (1974) described hepatitis am ong 13 men who, over the course of 6 years, were occupationally exposed to MDA in the blendi ng 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 anor exia; and over half reported fever, chills and/or headache. All the workers rec overed within a 10 week period. After the first cases of hepatitis occurred, air sa mpling 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 absorp tion 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 th e compound occurred during the workday. Since the symptoms appeared within one to 18 days after “wor king intensively” with the compound and exposure routes were not clearly establ ished, quantitation of exposure le vels was considered difficult.

The most well-known incident of MDA toxicity to humans resulted from ingestion of bread

made with flour contaminated w ith 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 di sease 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. Howe ver, a two year follow-up study of 14 individuals showed that 10 still had symptoms of some seve rity 7 to 23 months after initial onset including food intolerance, gastrointestinal disturbances , fatigue, and visual disturbances (Kopelman, 1968).

Human effects other than hepatotoxicity have be en described including se veral cases of contact

dermatitis and skin sensitizat ion (LeVine, 1983; Van Joost et al. , 1987; de Pablo et al. , 1992;

Appendix D3 400 Methylene dianilineDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

Bruynzeel and van der Wegen-Keij ser, 1993). A case report of a man exposed to MDA with potassium carbonate and γ -butyrolactone by accidental i ngestion has been described (Roy et al.

, 1985). In addition to hepatitis and abnormal liv er function, which persisted over 18 months, the patient developed a progressively worsening re tinopathy described as a “malfunction of the retinal pigment epithelium” accompanied by dimi nished visual acuity. The patient improved after approximately 3 months, but after examinati on at 18 months had not completely recovered.

Another report described the development of acute cardiomyopathy in add ition to hepatitis in a worker exposed to a large quantity of MDA dus t as the result of ai r filtration malfunction (Brooks et al. , 1979). The patient showed an abnorma l 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 inve stigated in F344/N rats and B6C3F 1 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 subchr onic 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 drinki ng water values from the study, low and high daily doses in the chronic study were calculated to be 9 and 16 mg/kg-da y 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 tr eated 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 effect s included follicular cysts and follicular-cell hyperplasia of the thyroid (si gnificantly 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 hi gh 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. Amo ng mice, incidence of liver degeneration was elevated in males in bo th treatment groups and females in the high-dose group (p<0.01 by Fisher’s exact test). Inciden ce of kidney nephropathy was increased in male and female mice in both treatment groups and mine ralization 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/m 3 in air for 4 hours/day, 5 days/week for 2 w eeks. 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 dianilineDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

sensitivity was found. Subsequently, the an imals were also examined for pulmonary sensitization by challenge with aerosols cont aining 0.01 and 0.05 ml of 2 00 mg MDA/ml PEG. Lung insufflation pressures were measured as an i ndication of changes in lung distensibility. No evidence of pulmonary sensitiz ation was found. After the pulm onary challenge, the animals were examined histopathologica lly, with emphasis on eye, lung, liver, and kidney toxicity. Ocular toxicity ranging from mild to more seve re was observed in all M DA-treated animals, but in none of the control animals. Pigmented an imals 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 in cluded 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 eff ects 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 we re 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 signi ficance was not stated). Treate d and control animals had a high background incidence of pulmonary lesions includi ng slight to mild bronchitis. No liver or kidney effects were detected in treated animals.

Nine purebred beagle dogs were treated oral ly (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. Afte r 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 lym ph nodes was performed on moribund animals or at the end of the experimental period (7 years, 2 months). Liver toxi city was noted in all the treated animals. Effects were described as fatty ch ange, cell degeneration a nd 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 effe cts including granuloma, glomerular nephritis, and congestion with cloudy swelling. Two dogs treated with “purified” and one dog treated with “crude” M DA showed toxicity to the spl een described as hemosiderosis and swelling with lymphocyte infiltration.

Wistar rats (5/sex/dose) were treated or ally 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 dos e. 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, includ ing atrophy of the parenchyma and stromal hyperplasia in the portal areas. Also in this dose group, all an imals 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 dianilineDetermination of Noncancer Chronic Re ference 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 ove r 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 w ith 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 (f our with one dose of 100 mg/kg, two with one dose of 150 mg/kg, and tw o 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 at rophy of the retinal neuroepithelium. The authors noted that no visual disturbances were found in other MDA treated experimental animals, incl uding 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/m 3 (54 ppm)

NOAEL Not observed Exposure continuity 4 hours/day, 5 days/week Exposure duration 2 weeks Average experimental exposure 52 mg/m 3 for LOAEL group (440 x 4/24 x 5/7) (6.4 ppm) Human equivalent concentration 52 mg/m 3 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/m 3 (20 μg/m 3 ; 0.002 ppm; 2 ppb)

Two specific types of toxicity have been associat ed with exposure to MDA: hepatotoxicity and ocular toxicity. Several studies have demonstrated hepatotoxici ty 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, ra ts and mice, to MDA at two concentrations in the drinking water. The 150 ppm dose level was a LOAEL for fa tty change and focal cellular

Appendix D3 403 Methylene dianilineDetermination of Noncancer Chronic Re ference 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 an d 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) re ported 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 associat ed with adverse health effects in humans preclude the development of a chronic REL from studies in humans.

The other toxic effect of poten tial concern from MDA exposur e is ocular toxicity. Leong et al. (1987) reported damage to the retinas of guin ea pigs exposed for 2 weeks to MDA aerosols (0.44 g/m 3 for 4 hr/day, 5 days/week; average experimental exposure = 52 mg/m 3 ) 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 repor ted 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 appear s to address the most sensitive endpoint of toxicity by the most appropriate route of exposur e (inhalation). The stud ies, which established the hepatotoxicity of MDA, were c onducted by the oral route of exposure.

As a comparison with the propos ed 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 assump tion that a 70 kg person breathes 20 m 3 of air per day leads to

an inhalation chronic REL estimate of 100 μ g/m 3 . 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 la ck of reproductive and developmental toxicity studies, and the lack of observation of a NOAEL. In addition the test

animals were under additional stre ss due to the restrain t used to obtain nose-only exposure, while

the control animals were not restrained. Liver to xicity has been included as a potential critical effect because of uncertainty re garding the relative potency of this compound in causing liver toxicity in different species by different routes of exposure.

When assessing the health effects of methylen e dianiline, its carcinoge nicity 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 dianilineDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

IX. References

ACGIH. 1992. American Conference of Gove rnmental Industrial Hygienists, Inc.

Documentation of the threshold limit values and biological exposur e indices. Sixth ed. Cincinnati, OH: ACGIH.

Bastian PG. 1984. Occupational he patitis caused by me thylenedianiline. Med. J. Aust. 141:533-

535.

Brooks LJ, Neale JM, and Pieroni DR. 1979. Ac ute myocardiopathy following tripathway exposure to methylenedianiline. JAMA 242:1527-1528.

Bruynzeel DP, and van der Wegen-Keijser MH. 1 993. Contact dermatitis in a cast technician.

Contact Dermatitis 28:193-194.

CARB. 2000. California Air Resources Board. Calif ornia 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 Ig lesias L. 1992. Allergic contact dermatitis from diaminodiphenylmethane in an ostomy bag. Contact Dermatitis 27:260.

Deichmann WB, MacDonald WE, Coplan M, W oods F, and Blum E. 1978. Di-(4-aminophenyl)- methane (MDA): 4-7 year dog f eeding study. Toxicology 11:185-188.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 7/31/96).

Kopelman H. 1968. The Epping Jaundice after two years. Po stgrad. 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 B6C3F 1 mice. J. Toxicol. Environ. Health 18:325-337.

Leong BKJ, Lund JE, Groehn JA, Coombs JK, Sa batis CP, Weaver RJ, and Griffin RL. 1987. Retinopathy from inhaling 4,4’-methylenedianiline aerosols. Fundam. Appl. Toxicol. 9:645-658.

Appendix D3 405 Methylene dianilineDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

LeVine MJ. 1983. Occupational photosensitivi ty 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 harden ers. I. Determination of acute and subacute toxicity of phthalic acid anhydr ide, 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. Methyl ene dianiline: a new toxic cause of visual failure with hepatitis. Hum. Toxicol. 4:61-66.

Schilling von Canstatt B, Hofmann HT, Oette l H, and Zeller H. 1966. [Netzhautveränderungen der Katze bei der Vergiftung mit peroral oder perkutan ve rabreichten Chemikalien]. Verhandlungen der Deutschen Gesell schaft für Pathologie 50:429-535.

Schoental R. 1968. Pathological lesions, including tumors, in rats after 4,4’- diaminodiphenylmethane and gamma-butyrol actone. Isr. J. Med. Sci. 4:1146-58.

U.S.EPA. 1994. United States Environmental Pr otection Agency. Methods for derivation of

inhalation reference concentra tions and applications of i nhalation dosimetry. EPA/600/8-

90/066F. Office of Research and Deve lopment. Washington DC: U.S.EPA.

Van Joost T, Heule F, and de Boer J. 1987. Se nsitization to methylenedianiline and parastructures. Contact De rmatitis 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 dianilineDetermination of Noncancer Chronic Refe rence 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/m 3
Critical effect(s) Hyperplasia of the olfact ory epithelium in rats Hazard index target(s) Respiratory system

II. Physical and Chemical Properties (HSDB, 1995)

Description Light yellow solid Molecular formula C 15 H 10 N 2 O 2 (monomer) Molecular weight Variable (monomer = 250.27 g/mol) Density 1.197 g/cm 3 @ 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/m 3 at 25 ° C; Not applicable for polymer

III. Major Uses or Sources

Methylene diphenyl isocyanate (MD I) is used for bonding rubber to nylon. MDI is also used in the manufacture of lacquer coatings and in the production of polyuretha ne 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 th e 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, respirator y symptoms, and smoking habits (Musk et al. , 1982, 1985). No significant changes in pulmonary function or respir atory symptoms were observed

Appendix D3 407 Methylene diphenyl isocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

when controlled for smoking. Mean MDI concen trations measured ranged from 0.0003 to 0.0006 ppm.

Significantly increased prevalence of asthma in fe male 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, FEV 1 and single-breath carbon monoxi de diffusion tests) were observed in men in the direct and indirect expos ure groups: decrements in men with a history of direct exposure to MDI were sta tistically significant. Workers we re 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 gr eater than 60 months exhibited statistically significant decrements in pulmona ry 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. Statistica lly significant decrements in pul monary function were observed again in workers with known direct occupationa l exposure to MDI. Workers who were exposed at the time of the 1976 study but had since be en removed from exposure did not exhibit decrements in pulmonary function, leading the aut hors to conclude that th e effects of low-level exposure to MDI are to some extent reversib le. Flaws in study de sign, including lack of exposure characterization, attrit ion, 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 FEV 1 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, phe nol 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 album in conjugates were detected in 4 of 5 MDIexposed workers (Aul et al ., 1999). The levels of specific Ig G antibodies were more elevated with polymeric MDI compar ed with monomeric MDI.

A workplace death of a 39-year-old foundry work er was ascribed to occupational asthma induced by MDI exposure (Carnio et al ., 1997). Postmortem pulmonary findings included epithelial desquamation, mucosal eosinophilic/ne utrophilic infiltration, bronchial vessel dilatation, and edema and hype rtrophy of smooth muscle.

Appendix D3 408 Methylene diphenyl isocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

V. Effects of Animal Exposure

Rats were exposed to 0.2, 1.0, and 6.0 mg/m 3 aerosolized MDI polymer 6 hours per day, 5 days per week for 24 months (Reuzel et al. , 1990; 1994). Statistically signi ficant increased incidences of basal cell hyperplasia, olfact ory epithelial degeneration, alveolar duct epithelialization, localized alveolar bronchiolization, and adenomas were observed in male and female rats exposed to 6.0 mg/m 3 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 significan t increase in the incidence of Bowman’s gland hyperplasia. Male rats exposed to 1 mg/m 3 MDI also exhibited statistically significant increased incidences of basal cell hyperplasia and Bowman’s gland hyperp lasia. An accumulation of macrophages with yellow pigment was observed in the lungs of fema le rats and the lungs and mediastinal lymph nodes of male rats exposed to 1 mg/m 3 . No adverse effects were noted in rats exposed to 0.2 mg/m 3 MDI.

Hyperplasia of the olfactory epit helium with MDI exposure (Reuzel et al ., 1990; 1994)

Males Females Combined
Concentratio

n (mg/m 3 ) 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 lit tle staining. However, another study exposed guinea pigs to 4 ppb radiolabelled toluene diis ocyanate (TDI) for 1-hour and found measurable radioactivity in extr athoracic tissues and body fluids (Kennedy et al. , 1989). Therefore, there is a possibility that MDI may be tran sported to sites other than the respiratory tract, such as the ovaries and testes, followi ng 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/m 3 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 parame ters investigated (inclu ding maternal weight gain, number of corpora lutea, implantation si tes, pre-and postimplantation loss, fetal and placental weights, gross and visceral anomalies, and degree of ossification). A slight but

Appendix D3 409 Methylene diphenyl isocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

significant increase in litters with fetuses displa ying 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/m 3 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/m 3 ) Critical effects Hyperplasia of the ol factory epithelium

LOAEL 1 mg/m 3

NOAEL 0.2 mg/m 3
Benchmark Concentration (BMC 05 ) 0.25 mg/m 3 (95% lower confidence limit on concentration for a 5% incidence of response based on analysis of th e combined male andfemale 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/m 3 for BMC 05 group (0.25 x 6/24 x 5/7) Human equivalent concentration 0.020 mg/m 3 for BMC 05 group (particle with extrathoracic respir atory 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/m 3

The data of Reuzel et al . (1990, 1994) were examined with six quantal dose-response models (linear, log-normal, Weibull, logistic, quadratic, gamma) usi ng 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. Po ssible differences between male and female susceptibility are suggested by the gender-speci fic data, although the significance of these

differences is uncertain.

USEPA used the same two studies and a BMC 10 approach to develop an RfC of 0.6 μg/m 3 . Since

USEPA used a 3-fold database uncertainty factor, their BMC 10 -based RfC is comparable to the

BMC 05 -based OEHHA REL.

Appendix D3 410 Methylene diphenyl isocyanateDetermination of Noncancer Chronic Refe rence 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, l ong-term inhalation study, the observation of a NOAEL, and the estimation of a benchmark concentra tion. A limitation of the REL is that it is based on data on expos ures 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 con centrations 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 indu ce hyperplasia of the ol factory 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 guin ea pigs. Toxicologist 10(1):286. (Abstr. #1144). [cite d in U.S. EPA, 1995].

Aul DJ, Bhaumik A, Kennedy AL, Brown WE, Lesa ge J, Malo JL. 1999. Specific IgG response to monomeric and polymeric diphenylmethane d iisocyanate 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 Heinri ch U. 1996. Embryotoxicity study of monomeric 4,4'-methylenediphenyl diisocyanate (MDI) aerosol after inhalati on exposure in Wistar rats. Fundam. Appl. Toxicol. 32(1):96-101.

CARB. 2000. California Air Resources Board. Calif ornia 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 Io li 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Micromed ex, 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 work ers in an iron and steel foundr y. 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 di stribution of 14C during and following inhalation exposure to radioactive tolu ene diisocyanate. Toxicol. Appl. Pharmacol. 100:280-292. [cited in U.S. EPA, 1995].

Appendix D3 411 Methylene diphenyl isocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

Marczynski B, Czuppon AB, Hoffa rth 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 Mur phy 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. Phys iopathol. 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 MD I) 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 met hylene diphenyl diisocyanate ae rosol in rats (Final report). Prepared by Civo Institute for the Internationa l 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 isocyanateDetermination of Noncancer Chronic Reference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# NAPHTHALENE

(naphthene, NCI-C5290, albocarbon, dezodorator, moth balls, moth flakes, tar camphor, wh ite tar, naphthalin, naphthaline)

CAS Registry Number: 91-20-3

I. Chronic Toxicity Summary

Inhalation reference exposure level 9 μ g/m 3 (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 C 10 H 8 Molecular weight 128.6 g/mol Density 4.42 g/cm 3 @ 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/m 3 per ppb at 25oC

III. Major Uses or Sources

Naphthalene is a natural constitu ent 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 (HSD B, 1995). It has also been used in the manufacture of phthalic anhydride, phthalic and anthranilic acid s, naphthols, naphthylamines, 1- naphthyl-n-methylcarbamate insecticide, beta-napht hol, naphthalene sulfonates, synthetic resins, celluloid, lampblack, smokeless pow der, anthraquinone, indigo, perylene, and hydronaphthalenes (NTP, 1992; HSDB, 1995). The statewide emissi ons from facilities reporting under the Air Toxics Hot Spots Act in California, based on th e most recent inventory, were estimated to be 164,459 pounds of naphthalene (CARB, 1999).

Appendix D3 413 NaphthaleneDetermination 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 μ g/m 3 ). Symptoms abated after removal of the mothballs.

Workers occupationally exposed to naphthalene fu mes or dust for up to five years were studied for adverse ocular effects (Ghett i 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 develo ped chorioretinitis (van der Hoeve, 1906).

Wolf (1978) reported that a majority of 15 persons involved in na phthalene manufacture developed either rhinopharyngolaryng itis and/or laryngeal carcinoma.

Ingestion of naphthalene or p- dichlorobenzene mothballs is a frequent cause of accidental poisoning of children (Siegel and Wason, 1986). In fants 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 gestati on (U.S. EPA, 1990).

Deaths have been reported follo wing ingestion of naphthalene mothballs. A 17-year old male ingested mothballs, developed gastrointestinal bl eeding, 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 in fants exposed to naphthalene vapors from nearby mothball-treated materials (Valaes et al. , 1963). Increased serum bilirubin, methemoglobin, Heinz bodies, and fragmented re d blood cells were observed. Kernicterus was noted in eight of the children, and two of the chil dren 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 seve ral weeks exposure to naphthalene and p-dichlorobenzen e (Harden and Baetjer, 1978).

Appendix D3 414 NaphthaleneDetermination 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 w eeks (NTP, 1992). Concentrations used were 0 (150 mice), 10 (150 mice), or 30 ppm (300 mice) naphthalene. (Tab le 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 inhi bition 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 hi ghest dose. No changes were noted in serum enzyme levels or electrolytes. The researcher s did not conduct a hist opathological examination.

B6C3F1 mice were administered 200 mg naphtha lene/kg/day by gavage for 5 days per week

over 13 weeks. No adverse effect s were observed (U.S. EPA, 1990).

Developmental effects of naphthalene ingesti on 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 da ys. 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) in traperitoneally administered 395 mg/kg/day naphthalene to Sprague-Dawley rats over days 1 though 15 of gestation. Fetuses ha d a 50% increase in incidence in delayed cranial ossi fication 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 nasa l discharge were noted at all doses. No statistically significant increase in ma lformations or developmental abnormalities was

observed.

Appendix D3 415 NaphthaleneDetermination 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 eff ects 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/m 3 , 9 μ g/m 3 )

The NTP study was chosen for the REL derivation since it is the only avai lable 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 strongl y indicates a causal relationshi p between naphthalene exposure

and nasal effects. The effects seen are consis tent with those reported among exposed workers, who developed rhinopharyngolaryngitis or laryng eal carcinoma (Wolf, 1978). However, the hematological effects observed in humans have no t been reported in laboratory animals, which raises the possibility that hu mans 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 uninfor mative 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 conc entration at which adve rse 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 al so used the NTP study to develop its RfC of 3

μ g/m 3 with slightly different assumptions and a cumulative uncertainty factor of 3000 (U.S. EPA, 2000). OEHHA followed the U.S. EPA preceden t in using an intraspecies UF of 10 for

Appendix D3 416 NaphthaleneDetermination 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 reac tivity, naphthalene-related effects on the nasal epithelium are expected to result following ab sorption of naphthalene and its metabolism to reactive oxygenated metabolites, not from direct contact. This is supported by data on naphthalene metabolism indicati ng that toxic effects on the re spiratory 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 ep ithelium in mice, rats, and hamsters occur following intraperitoneal injec tion of naphthalene. The nasal effects from inhalation exposure to naphthalene were considered to be extra-re spiratory 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 human s; however, it is uncertain that the RfC for naphthalene based on nasal effects will be protective for hemoly tic 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 ke y study, the absence of a NOAEL in the key study, the absence of other animal studi es 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 pois oning. J. Assoc. Physic. India 27:347-348.

Harden RA, and Baetjer AM. 1978. Aplastic anem ia 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 NaphthaleneDetermination of Noncancer Chronic Reference Exposure Levels April 2000

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, 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 car e nursing techniques. Dimens. Crit. Care

Nurs. 6:264-270.

Linick M. 1983. Illness associated with exposur e to naphthalene in mothballs - Indiana. MMWR

32:34-35.

Manchanda SS, and Sood SC. 1960. Accidental pois oning 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 ga vage to Sprague-Dawley rats on gestational days 6 through 15. Research Triangle Institute. Research Tr iangle 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. F undam. 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 temper ature vapor pressures of polynuclear aromatic hydrocarbons. Anal. Chem. 55(2):275-280.

U.S.EPA. 1986a. U.S. Environmental Protecti on Agency. Developmental toxicity study in

rabbits. PH 329-TX-001-85. 86-870000563. Office of Re search and Development, Cincinnati,

OH.

U.S.EPA. 1986b. U.S. Environmen tal Protection Agency. Ninety day (sub-chronic) dermal toxicity study with naphthalene in albi no rats. 65-870000565. Cincinna ti, OH: Office of Research and Development.

U.S.EPA. 1990. U.S. Environmen tal 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. EP A/600/8-90/066F. Office of Research and Development. Washington, DC: U.S.EPA.

Appendix D3 418 NaphthaleneDetermination of Noncancer Chronic Reference Exposure Levels April 2000

U. S. EPA. 2000. U.S. Environmen tal 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. Acut e hemolysis due to naphthalene inhalation. J.

Pediatr. 63:904-915.

van der Hoeve, J. 1906. [Chorioretinitis in humans fr om the effects of naphthalene] [in German]. Arch. Augenheilkd. 56:259-262. [reviewed in ATSDR, 1995].

Wolf O. 1978. Cancer of the larynx in napht halene cleaner. Z. Gesamte. Hyg. 24(10):737-739.

Appendix D3 419 NaphthaleneDetermination of Noncancer Chronic Re ference 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

NiCl 2 129.6 nickel chloride

nickel dichloride 7718-54-9

NiSO 4 154.75 nickel sulfate

nickelous sulfate 7786-81-4

NiCO 3 118.7 nickel carbonate

carbonic acid nickel salt 3333-67-3

Ni 3 S 2 240.19 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 μ g Ni/m 3
Critical effect(s) Lung, nasal epithelial an d lymphatic pathology

in male and female rats Hazard index target(s) Respiratory system; hematopoietic system



## B. Nickel Oxide

Inhalation reference exposure level 0.10 μ g Ni/m 3
Critical effect(s) Lung and lymphatic pathology in male and

female rats Hazard index target(s) Respiratory system; hematopoietic system

Appendix D3 420 NickelDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

II. Physical and Chemical Properties (from HSDB, 1995)

Description Ni metal: Silvery metal; NiCl 2 : deliquescent crystals (U.S.EPA, 1985) Molecular formula See above Molecular weight See above Density 8.9 g/cm 3 @ 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, hydroc hloric, 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 inso luble nickel compounds such as elemental nickel, nickel sulfide, a nd the nickel oxides from dusts and fumes.

Contributions to nickel in the ambient air are ma de by combustion of fossil fuels, nickel plating, and other metallurgical processes. The most common oxidation stat e of nickel is the divalent
(Ni 2+ ) 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 al loys that require resistance to extremes of temperature or stress (U.S.EPA, 198 5). Nickel is also used in electroplating baths, batteries, textile dyes, and catalysts (U.S.EPA, 1985). Ni ckel dust or powder is flammable (CDTSC, 1985). Due to its unique toxicological and physic o-chemical properties, nickel carbonyl is not included in this summary. The annual statewid e industrial emissions fr om facilities reporting under the Air Toxics Hot Spots Act in Californi a, 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 th at 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 prio r to employment. Novey et al. (1983) described biphasic metal-specific bronchial resp onses in an individual metal-plating worker exposed to nickel and chromium salts. In a nother case, immunological studies conducted in a 24-year old man showed nickel-specific antibodie s in the serum after several weeks of working in a nickel-plating shop usi ng nickel sulfate (McConnell et al. , 1973). Dermatitis was observed on exposed areas of his skin, and pulmonary function, measured by FEV 1 with and without isoproterenol challenge, was significantly impair ed compared with a control subject and normal

Appendix D3 421 NickelDetermination of Noncancer Chronic Re ference 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 occupa tional asthma were evaluated for atopy and pulmonary function challenge in response to inhala tional challenge with nickel and other metals (Cirla et al. , 1985). Three of the asthmatics tested pos itive 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 th e 7 asthmatics exhibited significantly decreased

FEV 1 (> 15%) when exposed to 0.3 mg/m 3 nickel sulfate for 30 mi nutes. Control challenges

with other metal salts did not re veal similar deficits in FEV 1

.

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/m 3 (0, 0.5, 1.0, or 2.0 mg Ni/m 3 ) 6 hours/day, 5 days/week for 104 weeks. In addition to the carcinogenic effect s of nickel oxide, a nu mber 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. At ypical 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/m 3 or greater at 7 and 15 months and the incidence generally increased w ith increasing concentr ation at the end of the 2-year study.

Females had an increased incidence of adrenal me dullary hyperplasia at al l exposures of nickel oxide. Body weights were significantly lo wer in the groups exposed to 2.0 mg Ni/m 3 for both sexes, and in males exposed to 1.0 mg Ni/m 3

.

A companion study on nickel oxide in mi ce conducted by NTP showed similar lung inflammatory changes as seen in the rats, in addi tion 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/m 3 . Bronchial lymph-node hyperplasia was also evident in all nickelexposed animals. Body weights were sligh tly but significantly lo wer in the 3.9 mg Ni/m 3 group, compared with controls.

A continuous exposure of rats (20 - 40 per group) to 0, 60, or 200 μ g Ni/m 3 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 protei nosis and septal fibrosis were

Appendix D3 422 NickelDetermination of Noncancer Chronic Re ference 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 subs ulfide 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 Ni 3 S 2 /m 3 (0, 0.11, or 0.73 mg Ni/m 3 ) for 6 hours/day, 5 days/week for 104 weeks. Body weights were lowered in rats exposed to 0.73 mg Ni/m 3 compared with controls. Lung inflammation, alveolar hyperplasia, macrophage hyperplasia, and pulmonary fibrosis were observed with a significantly increased

incidence at both nickel concen trations. Female rats expose d to nickel had significantly increased adrenal medullary hyperplasia. In addition to the pulmonary lesions, nasal inflammation and olfactory epithelial atrophy was observed in both se xes exposed to 0.73 mg

Ni/m 3

.

In the second phase of the NTP study (NTP , 1994b), mice were exposed to 0, 0.6, or 1.2 mg

Ni 3 S 2 /m 3 (0, 0.44, or 0.88 mg Ni/m 3 ) for 6 hours/day, 5 days/week for 104 weeks. The same pathological lesions were observe d in the lung and nasal passages as in the rats in the above study. These lesions were evid ent at both the 0.44 mg Ni/m 3 and the 0.88 mg Ni/m 3
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 Ni 3 S 2 /m 3 (0 or 0.71 mg Ni/m 3 ) for 6 hours/day, 5 days/week for 78-80 weeks resu lted 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 a nd carcinogenic effects of nickel sulfate hexahydrate on rats and mice. Rats we re exposed to 0, 0.12, 0.25, or 0.5 mg NiSO 4 /m 3 (0, 0.03, 0.06, or 0.11 mg Ni/m 3 ) for 6 hours/day, 5 days/week for 104 weeks. Chronic effects of nickel exposure in rats included inflammatory le sions in the lung, lung macrophage hyperplasia, alveolar proteinosis, and fibros is, in addition to bronchial lymph node hyperplasia and nasal epithelial atrophy. The above effects we re seen at exposures of 0.06 mg Ni/m 3 or greater.

Mice were exposed to a similar regime n that included 0, 0.06, 0.11, and 0.22 mg Ni/m 3 as nickel sulfate hexahydrate (NTP, 1994c). Similar pu lmonary, lymphatic and nasal changes were

observed in the mice as with the rats. Fibrosis was not reported, but an increased incidence of

interstitial infiltr ation and alveolar proteinosis were observed at exposures of 0.11 mg Ni/m 3 or greater. No clinical findings or hematological effects were observed, but body weights were significantly depressed in all gr oups of nickel-exposed female mice. The body weights of males were reduced only in the group exposed to 0.22 mg Ni/m 3

.

Rats and mice (10 per group) were exposed to nickel sulfate, nickel subsul fide, 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 sp ecies for all nickel exposures. Histological lesions included inflammatory ch anges, 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 exposur es of 0.05 mg Ni/m 3 or greater in female rats. Macrophage hyperplasia in the alve olar region was observed at concentrations as

Appendix D3 423 NickelDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

low as 0.02 mg Ni/m 3 . Additional inflammatory lesions in the lungs were observed at 0.1 mg

Ni/m 3

.

A similar study by Haley et al. (1990) found that exposure of mi ce to nickel sulfate, nickel subsulfide, or nickel oxide resu lted in various imm unological effects. Mice were exposed to 0, 0.11, 0.45, or 1.8 mg Ni/m 3 as Ni 3 S 2 ; 0.47, 2.0, or 7.9 mg Ni/m 3 as NiO; and 0.027, 0.11, and 0.45 mg Ni/m 3 as NiSO 4 for 6 hours/day, 5 days/week for 13 weeks. Nickel exposures consistently decreased splenic antibody-forming cell (AFC) respons es, with significant decreases occurring at 1.8 mg Ni/m 3 as nickel subsulfide. In c ontrast, AFC responses in the lungassociated lymph nodes were cons istently 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/m 3 as nickel chloride 6 hours/day, 5 days/week for 4 weeks exhibited si gnificantly decreased macrophage lysozyme activity in pulmonary lavage fluid and in macr ophage 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 Comp ounds (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 μ g Ni/m 3 (as nickel sulfate hexahydrate)

NOAEL 30 μ g Ni/m 3

Exposure continuity 6 hours/day, 5 days/week Exposure duration 104 weeks Average experimental exposure 5.4 μ g Ni/m 3 for NOAEL group (30 x 6/24 x 5/7) Human equivalent concentration 1.6 μ g Ni/m 3 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 μ g Ni/m 3

Appendix D3 424 NickelDetermination of Noncancer Chronic Re ference 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 μ g Ni/m 3

NOAEL Not observed Exposure continuity 6 hours/day, 5 days/week Exposure duration 104 weeks Average experimental exposure 89.5 μ g Ni/m 3 for LOAEL group (500 x 6/24 x 5/7) Human equivalent concentration 30 μ g Ni/m 3 for LOAEL group males (particulate with respiratory eff ects, 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 10 Subchronic uncertainty factor 1 Interspecies uncertainty factor 3 Intraspecies uncertainty factor 10 Cumulative uncertainty factor 300 Inhalation reference exposure level 0.10 μ g Ni/m 3

The studies conducted by NTP ( 1994 a,b, & c) all showed simila r non-carcinogenic effects in rats and mice, regardless of the form of nickel ad ministered. It therefore appears that soluble and

insoluble forms of nickel cause similar effects in rodents. The human epidemiological literature predominantly describes cancer mortality ra tes from occupational exposures to nickel compounds, but does not specifically examine non-cance r effects. However, it is clear from many case reports that allergies and dermatitis ca n 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 re sults to chronic human exposures and responses involves c onsiderable uncertainty. Furtherm ore, data show that nickel may precipitate onset of asthma in occupational settings.

The results of the NTP studies and these dose re sponse analyses support the speciation of nickel

oxide for noncancer effects. The health effects da ta for nickel oxide indi cate 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 NickelDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

subsulfide. The higher chronic R EL value for nickel oxide of 0.1 μ g/m 3 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. OE HHA therefore concludes that 0.1 μ g/m 3 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, an d agent for each group of substances. Therefore, as the pulmonary toxic ity of the relatively inso luble nickel subsulfide is greater than that of nickel oxide and closer to that of nick el sulfate, OEHHA proposes to use

the chronic REL derived from nickel su lfate 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 th e lack of lifetime toxicity studies in any nonrodent species.

In addition to being inhaled, airborne nickel can settle onto crops a nd soil and enter the body by ingestion. Thus an oral chroni c refererence exposure level fo r 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 us ed for the U.S. EPA’s oral Reference Dose (RfD). U.S.EPA assumed that ra t consumption of 1 ppm Ni in th e feed resulted in a dose of 0.05 mg/kg/day. An uncertainty factor of 10 was used for interspecies extrapol ation and another of 10 to protect sensitive human populations. An additi onal 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 th e criteria for their use are not well presented.

Appendix D3 426 NickelDetermination of Noncancer Chronic Re ference 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 dioxi de exposure of nickel

smelter workers and civic laborer s. 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 aspect s. 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. So c. 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 exposu re to nickel oxide, nickel subsulfide, or nickel sulfate hexahydrate in F344/N rats a nd B6C3F1 mice. Fundam. Appl. Toxicol. 12:584-

594.

Glaser U, Hochrainer D, Oldiges H, and Take naka S. 1986. Long-term inhalation studies with

NiO and As 2 O 3 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version) Denver, CO: Microm edex, Inc. (Edition expires 4/30/95).

Hueper WC. 1958. Experimental studies in me tal 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 chlo rides. Environ. Res. 34:335-342.

McConnell LH, Fink JN, Schlueter DP, and Sc hmidt MG. 1973. Asthma caused by nickel sensitivity. Ann. Int. Med. 78:888-890.

Appendix D3 427 NickelDetermination of Noncancer Chronic Re ference 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 radiogra phs of nickel sinter plant work ers. 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 Nick el Subsulfide in F344/N Rats and B6C3F1 Mice. NTP TR 453, NIH Publication No. 94-3369. U.S. Departme nt of Health and Human Services.

NTP. 1994c. National Toxicology Program. NT P 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. Depa rtment 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, Fa lk 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 Pr otection Agency. Health Assessment Document for Nickel. EPA/600/8-83/012F, pp. 3-3.

Appendix D3 428 NickelDetermination 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/m 3 (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 C 6 H 5 OH Molecular weight 94.11 g/mol Density 1.0576 g/cm 3 @ 20° C Boiling point 181.75° C Melting point 40.9° C Vapor pressure 0.3513 torr @ 25° C

Odor threshold 40 ppb (150 μ g/m 3 ) (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-m 3 /mol (25 °C) Conversion factor 1 ppm = 3.85 mg/m 3

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 in termediate 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 app lications such as germicidal pain ts and slimicides. Phenol is present in the atmosphere as an emission from motor vehicles and as a photooxidation product of

Appendix D3 429 PhenolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

benzene. The annual statewide industrial emis sions from facilities reporting under the Air Toxics Hot Spots Act in California, based on th e 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 occupa tional exposures, including some exposure by inhalation (Dosemeci et al. , 1991; Ohtsuji and Ikeda, 1972; Connecticut Bureau of Industrial Hygiene, undated). Data in animals are consiste nt with human data and show phenol to be well absorbed by oral, dermal, and inhalation routes of exposure. Severe ch ronic poisoning manifests in systemic disorders such as digestive dist urbances including vom iting, difficulty swallowing, ptyalism (excess secretion of saliv a), diarrhea, and anorexia (Bruce et al. , 1987; Baker et al.

, 1978). Phenol poisoning is associated with headache, fainting, vertigo, an d mental disturbances (Bruce et al. , 1987; Gosselin et al. 1984) which are likely symptoms of neurological effects well documented in animal studies. Ochronosis, or disc oloration 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 p eeling 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 deficien cies 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 stud ies available on human health effects associated with phenol exposure. Merliss (1972) desc ribed muscle pain and weakness of unknown etiology, enlarged liver, and elevated serum enzymes (LDH, GOT, and GPT) characteristic of liver damage in an i ndividual with intermittent inhalation and dermal exposures to phenol, cresol and xylenol. Bruze (1986) not ed that a number of phenolformaldehyde based resins are dermal ir ritants 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 symp toms 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 vi a inhalation to a comme rcial product containing formaldehyde, phenol and chlorohydrocarbons. At th e end of the six month period the indoor air of the workers contained 1,300 μg/m 3 of formaldehyde and 800 μg/m 3 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 PhenolDetermination 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 th e work environment.

Dosemeci et al . (1991) conducted a follow-up study to ev aluate mortality in 14,861 workers in five manufacturing facilities pr oducing or using phenol and formal dehyde. Arteriosclerotic heart disease, emphysema, disease of the digestive sy stem, 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 iden tified 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. Dose s 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 sin ce they reported NOAELs. Workers exposed continuously for an unspecified period of time to an average air concentration of 4 ppm phenol experienced no respiratory irrita tion (Connecticut Bureau of Indus trial Hygiene, undated). No adverse effects were reported am ong workers in a Bakelite factory who were exposed to levels of phenol up to 12.5 mg/m 3 (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 employi ng oral and inhalation routes of exposure are available as well as sh orter term studies using the dermal route of exposure. Responses observed in animal st udies 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 sulfat e and glucuronide conjugates by the gastrointestinal tract. Co mparison 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 da ys 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 e xposed at this concentration died at 28 days. At necropsy, extensive myocardial necrosis , lobular pneumonia, fatty degeneration of the

Appendix D3 431 PhenolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

liver, and centrilobular hepatocellula r necrosis were observed in all animals exposed at this level. Guinea pigs that were necropsie d at 41 days also exhibited pulmonary inflammation, pneumonia, bronchitis, endothelial hyperplasi a, and capillary thrombosis. Rabbits exposed at these same

concentrations did not exhibit a ny 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 c oncentrations and the la ck of a control group.

Sandage (1961) exposed Sprague-Dawley ra ts, mice and rhesus monkeys for 90 days continuously to 5 ppm phenol. Sandage f ound 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 eff ects of phenol on the nervous system in rats exposed continuously for 15 days to a con centration 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 impa irment as measured by decreased performance on tilting plane test was found. The Dalin a nd 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 phe nol is the most comple te chronic study using the oral route of exposure. Mice and rats we re 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 respirator y 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 highe r incidence of kidney inflammation (94%) than controls (74%). No tests of kidney f unction were performed in this study.

Boutwell and Bosch (1959) reported on the result s 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 reprodu ctive or developmental effects under chronic exposure conditions could be identified. Jones-Price et al. (1983a) reported th at 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 tera togenic 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 e xhibited 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 incide nce of cleft palate were seen at the 280 mg/kg/day dose.

Appendix D3 432 PhenolDetermination 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/m 3 (200 μ g/m 3 )

No suitable human studies were available fo r 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 th at was not continuous. Since alternative studies using guinea pigs could not be identified, the ra t was chosen as an alternative species since the rat has the most similar metabolic profile for me tabolism 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 du ration (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 spec ies 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 hepati c effects occur at a con centration of 26 ppm after several days (Dalin a nd Kristofferson, 1974).

The 5.0 ppm standard for phenol in the wor kplace (ACGIH, 1988; OSHA, 1985; NIOSH, 1976) is considered protective of the health of workers exposed occ upationally but does not consider

Appendix D3 433 PhenolDetermination 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 resp iratory irritation occurred among workers exposed regularly to 4 ppm phenol (Connec ticut Bureau of Industrial Hygi ene, 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 obser vation 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-res ponse 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 Governme ntal Industrial Hygienists. TLVs - Threshold Limit Values and Biological Exposure I ndices for 1988-1989. Cinci nnati, OH: ACGIH.

Amoore JE, Hautala E. 1983. Odor as an aid to ch emical safety: odor thresholds compared with

threshold limit values and volatilities for 214 i ndustrial 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, Dwo rniak D, Paradowski M, and Tchorzewski H. 1994. The effect of chronic exposure to form aldehyde, phenol and orga nic 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. Summar y review of the health effects associated with phenol. Toxicol. Ind. Health 3:535-568.

Appendix D3 434 PhenolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Bruze M. 1986. Sensitizing cap acity 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. (Y ear not reported) unpublished data. [as cited by

American Conference of Governme ntal Industrial Hygienists, 1984]

Dalin NM, and Kristoffersson R. 1974. Physiological effects of a sublethal concentration of inhaled phenol on the ra t. 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 animal s of the inhalation of phenol vapor. Am. J. Clin. Pathol. 14:273-277.

Deichmann WB, and Keplinger ML. 1962. Phenol s 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. Mo rtality 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: W illiams and Wilkins. pp. III-345-346.

Griffiths GJ. 1973. Fatal acute poisoning by intr adermal absorption of phenol. Med. Sci. Law

13:46-48.

HSDB. 1995. Hazardous Substances Data Ba nk. 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 work ers in an iron and steel foundr y. Br. J. Ind. Med. 42(2):94-

100.

Jones-Price C, Ledoux TA, Reel JR, et al. 1983a. Teratologic evalua tion of phenol (CAS No. 108-95-2) in CD-1 rats. Research Triangl e Park, NC: Research Triangle Institute.

Appendix D3 435 PhenolDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Jones-Price C, Ledoux TA, Reel JR, et al. 1983b. Teratologic evalua tion of phenol (CAS No. 108-95-2) in CD-1 mice. Labor atory study: September 18, 1980 to January 12, 1981. Research Triangle Park, NC: Resear ch Triangle Institute.

Merliss RR. 1972. Phenol marasmus. Occup. Med. 14:55-56

NCI. 1980. National Cancer Institute. Bioassay of phenol for possible carc inogenicity. Technical Report Series No. NCI-CG-TR-203. Bethesda, MD : U.S. Department of Health and Human

Services.

NIOSH. 1976. National Institute for Occupa tional Safety and Health. Criteria for a recommended standard ... occupationa l exposure to phenol. DHEW Pub 76-196.

Ohtsuji H, and Ikeda M. 1972. Quantitative rela tion between atmospheric phenol vapor and phenol in the urine of workers in Bakeli te 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 phe nol: 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 continuou s inhalation exposure to toxic material. I. Effects on animals of 90-day exposure to phenol, CCl 4 and a mixture of indole, skatole, H 2 S and methyl mercaptan. ASD technical report 61-519(I). Wright Patte rson Air Force Base, OH: U.S. Air Force Systems Command, Aer onautical Systems Division.

Truppman ES, and Ellenby JD. 1979. Major electro cardiographic changes during chemical face peeling. Plast. Reconstr. Surg. 63:44-48.

Warner MA, and Harper JV. 1985. Cardiac dysrhythm ias associated with chemical peeling with phenol. Anesthesiology 62:366-367.

Appendix D3 436 PhenolDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 PH 3 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/m 3 per ppm at 25oC

III. Major Uses and Sources

Phosphine is used as an agricultural fumigant ag ainst 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 generate s phosphine gas upon exposure to moisture. Because of high volatility, phosphine residue dissip ates 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 i ndustry as a chemical doping agent for electronic components (n-type semiconductors) (HSDB, 1995) . Other minor sources/uses of phosphine are in chemical syntheses: specifi cally, in preparations of phosphoni um halides, for polymerization initiation, and as condensation catalysts. Th e annual statewide indus trial emissions from facilities reporting under the Air Toxics Hot S pots 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 (inclu ding 119,519 pounds of aluminum phosphide and 1607 pounds of zinc phosphide) were applied in California agriculture (CDPR, 2001). In the

Appendix D3 437 PhosphineDetermination of Noncancer Chronic Re ference 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 interm ittently exposed to phosphine levels of 0.17-2.11 ppm in air from fumigation activity ranging over 0.5 to 29 ye ars (mean = 11.1 years) was evaluated (Misra

et al. , 1988). The subjects were in terviewed 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 worker s 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, retros pinal pain), and central nervous (headache, dizziness, staggering gait) systems. Symptoms were reported to appear only at the time of exposur e and apparently were reversible.

In another report of chronic o ccupational exposure, authors c ited the appearance of chronic bronchitis, anemia, and digestive disorders (Eichler, 1934).

Most literature reports of hu man toxic health effects of phosphine, however, come from case reports of acute exposures. Some are suggestive of potential chronic toxi city 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 (bilir ubinuria 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 exposur e of a 7-months pregnant, 24-year-old woman to aluminum phosphide from a nearby grai n 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 reporte d 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 anothe r 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. Findi ngs 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 PhosphineDetermination of Noncancer Chronic Re ference Exposure Levels September 2002

In a 1994 two deaths and three illnesses were reported due to phosphine fumigation of agricultural products in ra ilroad 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 pest icides (Litovitz et al ., 2001). However, two of the case reports of fatalities, that were pres ented 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 re lative organ weights of kidneys , lungs, liver, heart, brain and spleen. In the highest dose gr oup, itching and scratching of the eyes, feet and tail, and decreased overall activity were observed. No diarrhea, lo ss of equilibrium, convulsions, seizures, or other neurological disturbances were noted. A dose-dependent decreas e 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 leve l (p<0.001). On the other hand, female mice showed increased relative organ weights (lungs, hear t, and spleen) predominantly at higher doses (1.0 and 4.5 ppm; p<0.001). At 4.5 ppm phosphine ab solute kidney and spleen weights were significantly increased in females (p<0.01). Increased fr equencies 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 experime nt. Six mice/sex/group were exposed to 5.5 ppm phosphine for 2 weeks (6 hrs/day, 5 days/wk). No st atistically signif icant 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 mortalit y. A satellite group exposed to 5.1 ppm for 2 weeks was terminated after 13 days recovery. Obse rvations 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, th e cranial cavity, surfaces of the brain and spinal cord, nasal cavity and sinuses, the thoracic, abdo minal, and pelvic cavities and viscera, and the cervical tissue and organs. Histopathology was performed on 10% buffered formalinfixed/hematoxylin-eosin-stained tissues. Signi ficant 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 not ed at 1 and 3.1 ppm, a lthough no histological

correlate was observed. Absolute and relative de creases 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 PhosphineDetermination of Noncancer Chronic Re ference Exposure Levels September 2002

decrease in serum glutamic pyruvic transamina se (SGPT) was observed at 3.1 ppm, although the authors noted unusually high contro l levels. None of these eff ects were observed after the 4 week recovery period. Other effects of a tran sient nature noted duri ng 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 (B UN) at 3.1 ppm. Observations in the 10 ppm group necropsied after 3 days of exposure includ ed decreased erythrocytes, increased alkaline phosphatase, and increased kidney weight with coagul ative 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 toxic ity by exposing 24 pregnant female CD R
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 mortalit y; all other animals were necropsied after 20 days for evaluation of maternal and fetal toxicity . Maternal toxicity e ndpoints included weight of ovaries and uteri, number of corpora lutea, pregnancy, and im plantation rate. Fetal toxicity was evaluated by weight, number, and locati on of fetuses and resorptions, visceral malformations and variations, and skeletal chan ges after alizarin stai ning. 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, th ese observations are not considered useful in establishing a low advers e effect level.

A 35-day phosphine inhalation st udy was conducted exposing rats continuously to 0, 0.05, 0.2, 1.5, and 8.0 mg/m 3 phosphine (0, 0.036, 0.14, 1.1, and 5.8 ppm) in which hematological endpoints and histopathological ch anges of the lungs and kidney s were examined (Pazynich et al. , 1984). Observations include a statistically significant change in er ythrocytes (increase followed by a decrease at day 35) and d ecreased hemoglobin at the 0.05 and 0.2 mg/m 3 dose levels, although the 1.5 mg/m 3 dose group did not show this cha nge. Other significant changes, noted in the lowest dose group, in cluded decreased peroxidase ac tivity after 35 days exposure, decreased sulfhydryl group cont ent in blood after 27 days, and decreased phagocytotic index after 21 days. Some histological changes were no ted 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-respon se 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, alt hough 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/m 3 ) was the maximum

Appendix D3 440 PhosphineDetermination of Noncancer Chronic Re ference Exposure Levels September 2002

exposure level because of lethality seen at the high exposure level (7 ppm = 9.73 mg/m 3 ) 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 ob servations, 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 f ound no treatment-related changes suggestive of a toxic or carcinogenic eff ect 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 gai n, 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/m 3 ; 0.8 μ g/m 3 )

Appendix D3 441 PhosphineDetermination of Noncancer Chronic Re ference 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/m 3 , 7 μ g/m 3 )

Newton et al . (1999) found no treatment-related changes s uggestive of a toxic effect in F344 rats following 52 weeks or 2 years of whole-body inha lation 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 lethal ity 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 tr ansient 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 wei ght gain was selected as the critical effect.

OEHHA has applied a subchronic uncer tainty factor of 3 to account for the short duration of the

Barbosa et al . (1994) study and an intraspecies uncerta inty factor of 10 to account for human variability. Due to the general inconsistencie s among the various studies in the database on phosphine, and in particular with the observati on of mortality at 7 ppm in a short-term developmental study in rats (Newton et al., 1993), it was considered pr udent to include the full interspecies uncertainty factor of 10 (even though the HEC adjustment procedure could be applied) to acknowledge the severity of eff ect in at least one comparison study, and the additional uncertainty associated with the appa rent wide and unpredicta ble variability between species and between different studi es 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 μ g/m 3 (0.6 ppb).

The U.S. EPA based its RfC of 0.3 μ g/m 3 on the Barbosa et al. (1994) study, an adequate subchronic animal study for the derivation of a R EL, and included a Modifying Factor (MF) of 3

Appendix D3 442 PhosphineDetermination of Noncancer Chronic Re ference Exposure Levels September 2002

for database deficiencies (lack of multigenerationa l 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 chr onic 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 c onsistent with that found by Newton et al. (1993), who noted dosedependent decreases in body weight gain in Fischer 344 rats afte r a 13 week exposure regimen at 1 ppm, and Waritz and Brown ( 1975), who reported slig htly decreased wei ght 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 ch anges 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, cardi o-respiratory, CNS). Th e decrease in weight gain found in the animal studies and reported ch anges 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 concentratio ns 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 Im pacts on Children's Health

Based on the lack of a dose-response in developm ental toxicity caused by exposing pregnant female rats to 4.9 ppm phosphine (Newton et al ., 1993) , the proposed REL of 0.8 μ g/m 3 (0.6 ppb) is likely to be protective of developing hum ans in utero. However, there is no direct evidence in the literature to quantify a differen tial effect of phosphine on infants and children.

IX. References

Barbosa A, Rosinova E, Dempsey J, and Boni n AM. 1994. Determination of genotoxic and other effects in mice following short term repeat ed-dose and subchronic inhalation exposure to phosphine. Environ. Mol. Mutag. 24:81-88.

CARB. 2000. California Air Resources Board. Calif ornia Emissions Inventory Development and Reporting System (CEIDARS). Data from Data Base Year 1998. February 12, 2000.

Appendix D3 443 PhosphineDetermination of Noncancer Chronic Re ference Exposure Levels September 2002

CDRR. 2001. California Department of Pesticide Regulation. Summary of Pesticide Use Report Data 2000 Indexed by Chemical. Prelim inary Data. Available on-line at http://www.cdpr.ca.gov/doc s/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 exposu re. J. Lab. Clin. Med. 122:739-747.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, Inc. (Edition expires 7/31/96).

Jones AT, Jones RC, and Longley EO. 1964. Environm ental and clinical aspects of bulk wheat fumigation with aluminum phosphide. Ind. Hyg. J. 25:376-379.

Litovitz TL, Klein-Schwartz W, White S, Coba ugh DJ, Youniss J, Omslaer JC, Drab A, Benson BE. 2001. 2000 Annual report of the American A ssociation 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 Progr am Risk Assessment Guidelines. Part III. Technical Support Document for the Determinat ion of Noncancer Chronic Reference Exposure

Levels.

OEHHA, 1999. Draft Air Toxics Hot Spots Progr am Risk Assessment Guidelines. Part III. Technical Support Document for the Determinat ion of Noncancer Chronic Reference Exposure

Levels.

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Ai r Toxics Hot Spots Program Risk Assessment Guidelines. Part III. Technical Support Document for the

Determination of Noncancer Chronic Refere nce Exposure Levels. Available on-line at http://www.oehha.ca.gov

Appendix D3 444 PhosphineDetermination of Noncancer Chronic Re ference 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. 7 th ed. New

York: Van Nostrand Reinhold.

Schaefer GJ, Newton PE, Gruebbel MM, Bu sey WM, and Shaheen DG. 1998. Acute and

subchronic inhalation neurotoxic ity of phosphine in the rat. Inhal. Toxicol. 10:293-320.

Schoonbroodt D, Guffens P, Jousten P, Inge ls J, and Grodos J. 1992. Acute phosphine poisoning? A case report and revi ew. Acta Clin. Belg. 47:280-284.

Waritz RS, and Brown RM. 1975. Acute and sub acute 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. 61 st ed. Boca Raton, FL: CRC

Press.

Wilson R, Lovejoy FHJr, Jaeger RJ, and Landr igan PL. 1980. Acute phosphine poisoning aboard a grain freighter. JAMA 244:148-150.

Appendix D3 445 PhosphineDetermination of Noncancer Chronic Re ference 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/m 3
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 H 3 PO 4 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/m 3 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. P hosphoric acid is also used in pr ocesses 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 phos phorus 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 AcidDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

IV. Effects of Human Exposures

The toxic effects to 48 workers exposed (28 unexpos ed control workers) to oxidation products of phosphorus during the course of phosphor us 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 obser ved in acute intoxications, or hand bone density, an effect observed in expe rimentally 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 wa s 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/m 3 (phosphorus pentoxide). Pulmonary function tests were performed annually over a 3 to 7 year period. No significant residual effect was found after adjusting fo r 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 Spra gue-Dawley rats, with the first group exposed to 0, 300, 750, or 1200 mg/m 3 combustion products, and the second exposed to 0, 50, 180, or 300 mg/m 3 combustion products (Aranyi et al. , 1988a; Aranyi et al. , 1988b). Group numbers in the first study were 176, 84, 176, a nd 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. Da ily particle measurements showed MMADs of

0.49-0.65 μ m and σ g s of 1.56-1.83. Fractional content of ph osphoric acid in the aerosol was 71- 79%. Nineteen of the 176 animals in the 1200 mg/m 3 dose group died of treatment related

effects. Post-mortem examination of animals th at died during the course of the study showed damage to the laryngeal mucosa, which was probab ly contributory to mortality. The two highest dose groups in the first study also showed decreas ed weight gain. Twelve animals from each dose group in the first study were examined hist ologically and neurobeha vioral studies were

conducted on other animals. Half the animals in the second study were examined strictly for toxic effects on the respiratory tract, with examin ation 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 repor ted incidence of this lesion was 9/20 at 300 mg/m 3 , 4/20 at 180 mg/m 3 , and 0/20 at 50 mg/m 3 . Little to no involvement of pulmonary tissue

was observed.

The effects of acid aerosols (particularly sulfur ic 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, dur ation of exposure, and the hygros copic character of the acid.

Sprague-Dawley rats were exposed to the sm oke and combustion products of white phosphorus in felt pellets at 192.5 (18 animals/se x), 589 (24 animals/sex), or 1161 mg/m 3 (34 males, 43

Appendix D3 447 Phosphoric AcidDetermination of Noncancer Chronic Re ference 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). Cont rol 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 onl y other death occurring in the control group. Findings were rest ricted to effects on the respirat ory system, with tracheitis and laryngitis incidences of 2/35, 32/47, and 28/31among surviving anim als 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 combusti on products of 95% amorphous re d phosphorus and 5% polyvinyl butyral BL18 to female Wistar rats, Porton-stra in 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/m 3 for 1 hour/day, 5 days/week for 36 weeks (mice) or 40 weeks (rats and guinea pigs), w ith an examination conducted at 19 months or when animals appeared unhealthy. All groups, including cont rols, showed high mortality. Mice showed accumulation of alveolar macrophages with incide nces of 2/41, 9/37, and 9/22 in the control, low-, and high-dose groups, respectively. Guinea pi gs 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) (M arrs, 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 inflammati on 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/m 3 elemental phosphorus for 30 minutes/day for 60 days were examined for toxic effect s (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 toxi city 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 Depa rtment of Defense (US DOD), 1981; US Department of Defense (US DOD), 1982). Fetal effects included increased incidence of some visc eral variations and hypoplasia of the xiphoid process although data were incompletely re ported. Another study, which exposed dams 3 weeks pr ior to mating, throughout gestat ion, and through lactation and males for 10 weeks prior to and during mati ng, showed decreased pup body weight, 24-hour and

Appendix D3 448 Phosphoric AcidDetermination of Noncancer Chronic Re ference 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/m 3

NOAEL 50 mg/m 3

BMC 05 64 mg/m 3

Exposure continuity 21⁄4 hours/day, 4 days/week Exposure duration 13 weeks Average experimental exposure 2.7 mg/m 3 for NOAEL group (estimated as 3.5 mg/m 3 at BMC 05 ) Human equivalent concentration 2.2 mg/m 3 at BMC 05 (particle with respiratory effects, RDDR = 0.63) (3.5 x 0.63) LOAEL uncertainty factor 1 (BMC 05 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/m 3 (7 μg/m 3 )

OEHHA has used the same study which U.S. EPA used in the development of its Reference Concentration (RfC) of 10 μ g/m 3 . 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 elem ental phosphorus or other forms of phosphorus, such as phosphorus salts”.

The U.S. EPA, using the Weibull model, estim ated the lower 95% confidence level bound on the maximum likelihood estimate (MLE = 150 mg/m 3 ) resulting in 10% incidence of lesions in the tracheo-bronchiolar region to be 100 mg/m 3 (the BMC 10 ). The U.S. EPA considered 10% incidence level to be a correlate to the NOAEL, ba sed 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 trache obronchial region of 0.64 was applied due to the availability of da ta concerning the grow th and deposition of phosphoric acid aerosol particles in humans and the similarities in the effects of phosphoric and

better-characterized sulfuric acid aerosols. Ke y assumptions in the generation of this factor include: (1) the lowest σ g of 1.56 μ m cited in the study was used in the calculation; (2) geometric rather than aerodynamic diameter approximations were used; (3) part icles of this size reach the deposition / lesion site (bronchiol es); 4) these hygroscopic partic les become more uniform with

Appendix D3 449 Phosphoric AcidDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

growth; and (5) particle growth is similar in hum ans 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 pr otection of potentially sensitive human subpopulations. The resulting RfC for phosphoric acid is 0.01 mg/m 3

.

OEHHA uses a BMC 05 for development of acute Reference Exposure Levels (OEHHA, 1999;

Fowles et al ., 1999). OEHHA staff believe that the BMC 05 is more likely to approximate a

NOAEL than a BMC 10 since 5% is closer than 10% to th e lower end of average risk levels associated with a NOAEL (Leisenring and Ryan, 1992). A BMC 05 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 BMC 10 is more likely to represent a LOAEL since it is usually in the detectable range of responses. In the specifi c case of phosphoric acid the BMC 10 of 100 mg/m 3 was twice the NOAEL of 50 mg/m 3 . The BMC 05 was calculated to be 64 mg/m 3 , much closer to the

NOAEL. Use of the BMC 05 results in a chronic REL of 7 μ g/m 3

.

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation REL for phosphoric acid include the availa bility 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 inhala tion exposure studies, and the disc ontinuous 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 num ber of animals with multiple doses, produced good dose-response da ta, 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 th an phosphoric acid itself, (2) the total exposure time was relatively short and discontinuous over the du ration of the experiment, and (3) only one species/strain/sex was studied.

VIII. References

Allen BC, Kavlock RJ, Kimmel CA, and Faus tman EM. 1994. Dose-response assessment for developmental toxicity. II. Comparison of gene ric benchmark dose estimates with no observed

adverse effect levels. Funda m. 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 ru bber obscurant smokes in Sprague-Dawley rats.

Inhal. Toxicol. 1:65-78.

Appendix D3 450 Phosphoric AcidDetermination of Noncancer Chronic Re ference 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 m acrophage 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 obtai n 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. Charac terization of database and dete rmination of no observed effect levels. Fundam. Appl. Toxicol. 23:478-86.

Fowles JR, Alexeeff GV, Dodge D. 1999. The us e 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: Mi cromedex, 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-c reated chronic poisoning of yellow phosphorus to bones. Igaku Kenkyu, 26:2848-61.

Leisenring W, Ryan L. 1992. Stat istical properties of the NOAEL. Regul. Toxicol. Pharmacol. 15(2 Pt 1):161-171.

Marrs T. 1984. Histological ch anges 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, a nd Cross N. 1989. The toxicity of a red phosphorus smoke after repeated inhalation. J. Haz. Mat. 22:269-82.

Appendix D3 451 Phosphoric AcidDetermination of Noncancer Chronic Re ference 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 Environmenta l Health Hazard Assessm ent, 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 ae rometrics. Prepared by

the Office of Health and Environmental Asse ssment, Environmental Criteria and Assessment

Office. EPA/600/8-88/005F. Research Triangle Park, NC: U.S. EPA.

U.S. EPA. 1995. United States Environmenta l Protection Agency. Documentation of the

reference concentration for chr onic inhalation exposure (RfC) for phosphoric acid. Integrated Risk Information System (IRIS on-line). Washington, DC: U.S. EPA.

Appendix D3 452 Phosphoric AcidDetermination of Noncancer Chronic Refe rence 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 μ g/m 3
Critical effect(s) Eye and respiratory irritation, asthma, and bronchitis in occupati onally exposed workers Hazard index target(s) Respiratory system

II. Chemical Property Summary (HSDB, 1995; CRC, 1994)

Description White or pale yellow crystals Molecular formula C 8 H 4 O 3 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/m 3 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 anot her major use in the production of polyester

resins and other minor uses in th e production of alkyd resins used in paints and lacquers, certain dyes (anthraquinone, phthalein, rhodamine, phtha locyanine, fluorescei n, 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 re porting 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 anhydrideDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

IV. Effects of Human Exposure

Symptoms in workers exposed to phthalic anhyd ride by inhalation in tw o plants (A and B) manufacturing alkyd and unsaturated polye ster resins were studied (Nielsen et al. , 1988). Two groups of exposed workers were id entified 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 group s were 13.3 and 11.9 years, respectively. Timeweighted average air concentrations for work ers from the loading of PA was 6.1 (range: 1.8- 14.9) and 6.8 mg PA/m 3 (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 tim es a day, corresponding to the time of loading, and resulted in a full-day time wei ghted exposure estimate of 0.4 mg PA/m 3 . For those engaged in “other work” exposure levels were estimated at < 0.1 mg PA/m 3 (the limit of detection).

Other chemicals in use in smaller amounts incl uded maleic anhydride, isophthalic anhydride, and trimellitic anhydride. Comparison of sympto m 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” exposur e group. A correlation was also noted between specific IgG level and exposure le vel, although not all individuals with elevated specific IgG reported symptoms.

In a study conducted at another plant manufacturi ng alkyd and/or unsaturated polyester resins, serum immunoglobulins and lung function were exam ined in 23 workers exposed to phthalic anhydride and 18 control subjects (Nielsen et al. , 1991). Estimated exposure levels were 6.6 mg

PA/m 3 (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 repo rting of conjunctivitis a nd rhinoconjunctivitis. One worker showed an asthmatic response to an hydrides. No significant differences in lung

function tests were observed betw een 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/m 3 for workers engaged directly with the handling of PA. In other areas the estimated level was <0.3 mg/m 3 . 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 sensitiv ity. Bronchial provocati on tests with 6 or 0.5 mg/m 3 PA for 5 or 10 minutes were positive in 2 workers.

Appendix D3 454 Phthalic anhydrideDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

V. Effects of Animal Exposure

Male albino rats (6/treatment group) were exposed to phtha lic anhydride vapors at 0, 0.02, 0.2, and 1 mg/m 3 continuously for 45 days (Protsenko, 1970). After a two week recovery period the testes were examined for spermatozoa motility ti me as well as for ascorbic acid, dehydroascorbic acid, and nucleic acid content. Motility time wa s 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/m 3 dose group and ~25% in the 0.2 mg/m 3 dose group. Significant

decreases in ascorbic acid and dehydroascorbic acid levels were found in animals exposed to 0.2 and 1.0 mg/m 3 phthalic anhydride, and dehydroascorbic acid levels were decreased in the 0.02 mg/m 3 dose group. At 1 mg/m 3 , RNA levels and combined RNA and DNA levels were significantly increased over controls. No signi ficant changes were observed in the 0.02 mg/m 3
dose group.

Five and six female Hartley guinea pigs were exposed to 0.05-0.2 mg/m 3 and 0.6-6 mg/m 3
phthalic anhydride dust, respectiv ely, for 3 hours/day for 5 consecu tive days (Sarlo and Clark, 1992). Exposures were expressed as ranges due to difficulty in regulating dust levels in the chambers. Sampling of dust show ed particles were 65-80% < 10 μ m diameter and had a mean

mass diameter of 5.8-9.8 μ m. Eight control animals were ex posed to filtered air only. Two weeks after the last exposure, animals were challenged for 30 minutes with aerosolized PAguinea pig serum albumin conjugate. All animal s in the “high” dose gr oup showed immediate bronchoconstriction and transiently increased respiratory rate. An imals in this dose group also showed elevated IgG antibody tite rs. No detectable increase in antibody levels was found in the “low” dose group.

Type I hypersensitivity was examined in fema le 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/m 3

, and two groups of 16 animals were expos ed to 0 (filtered air only) or 5.0 mg/m 3 phthalic anhydride dust (respirable size – 5 μ m) in stainless steel chambers for 3 hours/day for 5 consecutive days. Groups of 8 an imals from the control and 5 mg/m 3 groups were challenged after a two week recovery pe riod for 30 minutes with 5.0 mg/m 3 phthalic anhydride dust. Respiratory data were collected using a plethysmograph from 30 minutes before the exposure to 60 minutes after the exposure. No significant di fference (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 unexpos ed animals. Eight animals in each of the four exposure groups were also challenged after two weeks of recovery with 2.0 mg/m 3
aerosolized PA-guinea pig serum albumin (GPSA) conjugate as described above. Respiratory rate was increased in 4/8 of the high-dose gr oup 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 Ig G 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 IgG1 a immunoglobulins showed positive results for 3/8, 1/8, and 5/8 animals in the 0.5, 1.0, and 5.0 mg/m 3 dose groups, respectively. Results in

control animals were not descri bed. Three of eight animals in the highest dose group had >189 hemorrhagic foci in their lungs. No control anim al had more than 2 such foci. No foci were

Appendix D3 455 Phthalic anhydrideDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

observed in animals challenged with albumin c onjugate. 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 expos ed in 100 L chambers to 0, 0.18, 0.54, and 1.52 mg

PA/m 3 aerosol continuously for 70 days. Genera l condition and behavi or, 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 ex tensors differed from the controls beginning on day 31 of exposure and continued until two weeks after exposure ceased. Significantly decreased whole blood cholinesterase activity oc curred 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 afte r 70 days of exposure, but retu rned to normal during the twoweek recovery period. Thus, 0.18 mg/m 3 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 su rvival, 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 co at, ulceration, and corn eal opacity”, however,

incidences were described as “l ow”. No significant histopathol ogical effects were found to be associated with exposure to phthalic anhydride. B6C3F 1 mice (50/sex/dose group plus 20/sex control animals) were initia lly treated with diet contai ning 0, 25,000, or 50,000 ppm phthalic anhydride (approximately 0, 3000, and 6000 mg/kg-da y, assuming that food consumption is 12% body weight/day). Because of excessive weig ht loss after week 32, exposure levels were reduced during the course of the study such th at the time-weighted average exposure for males was 16,346 and 32,692 ppm and for females wa s 12,019 and 24,038 ppm phthalic anhydride. Evaluation of toxicity was conducted at 104 week s as with the rats. Mean body weight was

reduced in male and female mice in a dose-rela ted manner. No other significant treatmentrelated adverse effects were observed in the mice.

Pregnant female CD-1 mice (10/dose group) we re treated intraperitoneally with phthalic anhydride in 0.5%(w/v) carboxymethyl cellulose solution on gest ational days 8-10 (Fabro et al.

, 1982). Dosing was variable, beginning with in the 95% confidence limits of the LD 01 and progressing geometrically downward until no effect was observed. Animals were terminated on Day 18 and examined for teratogenic effects incl uding fetal viability and number, resorption, and gross malformations. The 95% lower confid ence limit on the dose produc ing teratogenicity

Appendix D3 456 Phthalic anhydrideDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

(grossly observable malformations and fetal inte rnal malformations) in 5% and 50% of animals were 0.40 and 1.37 mmol/kg-day (59 a nd 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/m 3 (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/m 3 for LOAEL group (6.5 mg/m 3 × 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/m 3 (20 μ g/m 3 )

Adverse effects were demonstrated to occur in humans occupationally exposed to phthalic anhydride in the workplace over l ong periods of time (Nielsen et al. , 1988). The symptoms reported primarily affected the respiratory system, with incr eased 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 rhinoconj unctivitis, conjunctivitis, or chr onic bronchitis have also been reported in workers exposed to simi lar levels of PA dust (Nielsen et al. , 1991; Wernfors et al.

, 1986). In these reports, advers e effects were clearly observed at the exposure level reported (6.5 mg PA/m 3 ; full-day time weighted exposure of 0.4 mg PA/m 3 ). Although symptoms were reported by Nielsen (1988) in the lower exposure level group, the sign ificance is not clear since a true control group (unexposed wo rkers) 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 gr oup was considered as a control group.

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalati on REL for phthalic anhydride incl ude the use of human exposure data from workers exposed over a period of year s. Major areas of un certainty are (1) the uncertainty in estimating exposure, (2) the potent ial variability in exposur e concentration, (3) the

Appendix D3 457 Phthalic anhydrideDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

potential low exposures of the group consider ed 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 mode l for human asthmatics) in comparison to occupationally exposed workers.

The study in rats by Protsenko (1970 ) identified a L OAEL of 0.2 mg/m 3 and a NOAEL of 0.02 mg/m 3 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 th e basis for the REL for PA.

VIII. References

CARB. 2000. California Air Resources Board. Calif ornia 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 rela tive teratogenic index and teratogenic potency: proposed components of developm ental toxicity risk assessmen t. Teratog. Carcinog. Mutag.

2:61-76.

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Micromedex, Inc.

NCI. 1979. National Cancer Institute. Bioa ssay of Phthalic Anhydride for Possible Carcinogenicity. Springfield, VA: NTIS. PB 293-594.

Nielsen J, Bensryd I, Almquist H, Dahlqvist M, Welinder H, Alexanderss on R, and Skerfving S. 1991. Serum IgE and lung function in workers expos ed 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 anhydr ide on gonads]. Gig. Sanit. 35:127-130. [in Russian]

Sarlo K, and Clark ED. 1992. A tier approach for ev aluating the respiratory allergenicity of low molecular weight chemicals. Fundam. Appl. Toxicol. 18:107-114.

Appendix D3 458 Phthalic anhydrideDetermination of Noncancer Chronic Refe rence 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 Me ditsinskikh Nauk SSSr 10:86-96. [in Russian]

Wernfors M, Nielsen J, Schütz A, and Sk erfving S. 1986. Phthalic anhydride-induced occupational asthma. Int. Arch. Allergy Immunol. 79:77-82.

Appendix D3 459 Phthalic anhydrideDetermination 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 μ g/m 3 (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 C 3 H 6 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 μ g/m 3 per ppb at 25 ° C

III. Major Uses and Sources (HSDB 1995)

Propylene is produced primarily as a by-produc t of petroleum refining and of ethylene production by steam cracking of hydrocarbon feed stocks. Propylene is a major chemical intermediate. The most important derivatives of chemical and polymer grade propylene are polypropylene, acrylonitrile, propyl ene oxide, isopropanol and cume ne. 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 producti on 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 inha lation. 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 emissi ons from facilities reporting under

Appendix D3 460 PropyleneDetermination of Noncancer Chronic Reference Exposure Levels April 2000

the Air Toxics Hot Spots Act in California, ba sed 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, distri bution or excretion of propylene in humans. However, hemoglobin (Hb) adducts of the meta bolite intermediate propylene oxide have been used to monitor the internal dose of pr opylene (Tornqvist and Ehrenberg, 1990). The background level of the 2-hydroxyl propyl adduct to the N-terminal valine of hemoglobin was found to be about 2 pmol/g Hb. This was estimat ed 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/m 3 ) was assumed to be associated with an increment of 5 pmol/g Hb (Kautiainen and Tornqvist, 1991).

No data were available on the chro nic 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 metabo lism 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 epith elium (Maples and Dahl, 1991) can convert propylene to its toxic metabolite. However, in rats, propylene metabolism becomes increasingly

saturated at concentratio ns above 50 ppm (86 mg/m 3 ) in the atmosphere (Golka et al. , 1989), which limits the amount of propylene oxide produ ced. Therefore, the amount of absorbed propylene may not reach high enough levels in cl assical long-term inha lation studies (Quest et al. , 1984) to show positive carcinogeni c or serious chronic effects.

The only chronic toxicity invest igation found for propylene was a comprehensive 2-year study in

F344/N rats and B6C3F 1 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 we re 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 cavit y. In female rats, epithelial hyperplasia occurred in the high dose group and s quamous metaplasia occurred in both dosage groups. In male rats, squamous metaplasia wa s 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 granulocyt es into the lumen (see below). Nasal lesions were not observed in mice. The inflammatory le sions 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 eff ects, including clinical signs, mortality, mean organ and body weight s, and histopathology, were observed.

Appendix D3 461 PropyleneDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Incidences of epithelial changes in nasa l 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) hi gher than control values

In a long-term carcinogenicity study, Sprague-Dawley ra ts and Swiss mice (100-120 animals/group/sex) were exposed by i nhalation 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 observe d between treated and control an imals of either species. Mortality was reported to be slightly increase d in male rats in th e 1000 and 5000 ppm groups and in male mice in the 5000 ppm group, but numerical va lues 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 PropyleneDetermination 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/m 3 )

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 cm 2 ) 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/m 3 , 3,000 μ g/m 3 )

VII. Data Strengths and Limitations for Development of the REL

Strengths of the propylene REL include the availa bility 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 result ed 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 incl ude the lack of lifetime toxicity studies in any non-rodent species. Also, no long-term hu man 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 comprehens ive multi-generation study in an experimental animal species would enhance the development of a propylene REL.

Appendix D3 463 PropyleneDetermination 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 & Engine ering News. Production by the U.S. chemical industry. M. Heylin, ed. American Chemical So ciety, 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-Dawl ey rats and Swiss mice. Ann. N.Y. Acad

. Sci.

534:235-245.

Cox RA, Derwent RG, and Sandalls FJ. 1976. So me air pollution measurements made at Harwell, Oxfordshire, during 1973-1975, AERE-R 8324, Harwell, U.K. Atomic Energy Research Establishment, Environmen tal and Medical Sc iences 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. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM version). Denver, CO: Microm edex, Inc. (edition expires 11/31/95).

Kautiainen A, and Tornqvist M. 1991. Monitori ng 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. Toxicol ogy and carcinogenesis studies of propylene

in F344/N rats and B6C3F 1 mice. Technical Report Series NTP-TR-272, NIH publication no. 86-

2528.

Quest JA, Tomaszewski JE, Haseman JK, B oorman GA, Douglas JF, and Clarke WJ. 1984. Two-year inhalation toxicity study of propylene in F344/N rats and B6C3F 1 mice. Toxicol

. Appl

. Pharmacol

. 76:288-295.

Appendix D3 464 PropyleneDetermination of Noncancer Chronic Reference Exposure Levels April 2000

Svensson K, and Osterman-Golkar S. 1984. Kineti cs 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 McMi cheal AJ, eds. Complex Mixtures and Cancer Risks.

IARC Scientific Publications No. 104. Lyon: IARC, pp. 272-287.

Appendix D3 465 PropyleneDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# PROPYLENE GLYCOL MONOMETHYL ETHER

(1-Methoxy-2-propanol; 1-methoxyp ropanol; Propapsol solvent M)

CAS Registry Number: 107-98-2

I. Chronic Toxicity Summary

Inhalation reference exposure level 7,000 μg/m 3 (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 C 4 H 10 O 2 Molecular weight 90.14 Density 0.962 g/cm 3 @ 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/m 3 at 25° C

III. Major Uses or Sources

Propylene glycol monomethyl ethe r (PGME) is used as a solven t for cellulose, acrylics, dyes, inks and stains (HSDB, 1995). Thus, the primary us e of PGME is in lacquers and paints. Use of PGME is anticipated to increase due to its low sy stemic toxicity. The an nual specific statewide

industrial emissions of PGME fr om facilities reporting under the Air Toxics Hot Spots Act in California, based on the most recent inventor y, 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 cate gory glycol ethers, some of which may be PGME.)

Appendix D3 466 Propylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference 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 irritati on was reported by two of six huma n volunteers exposed to 100 ppm PGME for 2 hours (Stewart et al. , 1970). These subjects were e xposed for a total of 3 1/2 hours during which no decrement in visual acuity, c oordination, neurological responses or reaction time was measured. The same experiment expos ed 23 subjects to 250 ppm PGME. After 15 to 30 minutes of exposure, 8/23 reported eye irri tation and 3/23 reported throat irritation; lacrimation was observed in 3/23 subjects. Three subjects each reporte d one of the following symptoms: irritation, headache, and nausea. While the subjects frequently reported the odor to be objectionable upon first entering the chamber, th e odor was usually undetectable by the end of the exposure. Clinical chemistry and urinalysis completed following expos ure was not altered as compared to pre-exposure measurements.

V. Effects of Exposures on Animals

Male and female rats (10 per sex per concentrat ion) and rabbits (7 per sex per concentration) were exposed by inhalation to 300, 1000, or 3000 pp m 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 ra ts exposed to 3000 ppm PGME. Hepatocellular hypertrophy was observed upon histopa thologic 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 longe r 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, respectivel y. 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 ga in 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 w ith rapid post-exposure recovery. Overall maternal weight gain during the exposure (days 6-18 of ge station) was statistically signifi cantly 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 da ys 6-17 of gestation (Doe et al. , 1983). Male rats were exposed to 200 or 600 ppm PGME 6 hours per da y for 10 consecutive days. No significant effects on testicular weight or pathology were observed.

Increased liver and kidney wei ghts 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Ciezlak et al. (1998) evaluated the potential ch ronic toxicity/oncogenic ity and the response of liver and kidney tissue of Fischer 344 rats to propylene glycol monome thyl 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 ev aluated included the ge neral appearance and demeanor of animals, in-life body weights, su rvival, hematology, urinalysis and clinical chemistry determinations, survival, selected or gan weights, gross and microscopic pathologic changes and tumor incidence. (The metabo lic and morphological bases for PGME-induced sedation, hepatic hypertrophy and rena l 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 cellu lar proliferation and accumulation of alpha 2u -globulin (males only) in the kidneys, were conducte d 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 contro l concentrations by week 52, coinciding with a return of sedation at 3000 ppm PGME. In ma le rats, the loss of metabolic adaptation was followed by a dose-related increase in altered hepa tocellular foci after two years of exposure to 1000 or 3000 ppm PGME. The kidney toxicity observed in male rats was confirmed immunohistochemically as an alpha 2u -globulin nephropathy. No sta tistically-identif ied increases in tumors were observed in any tissue. The authors established a NO EL of 300 ppm PGME for the study.

Ethylene glycol methyl ether (EGME), a struct urally related compound, exerts considerable toxicity on the blood, thymus, test es, 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 iden tified in rats were propylene glycol and the sulfate and glucuronide c onjugate of PGME (Miller et al. , 1983).

Appendix D3 468 Propylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference 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/m 3 , 7000 μg/m 3 )

VII. Data Strengths and Limitations for Development of the REL

Strengths of the PGME RfC include the obser vation 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 Associ ation. 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 glyc ol monomethyl ether: a two-year vapor inhalation chr onic toxicity/oncogenicity study and evaluation of hepatic and renal cellular proliferat ion, P450 enzyme induction and protei n droplet nephropathy in Fischer

344 rats. Dow Chemical Company, Midland, MI.

Doe JE, Samuels DM, Tinston DJ, de Silva Wi ckramaratne GA. 1983. Comparative aspects of the reproductive toxicology by inha lation in rats of ethylene gl ycol monomethyl ether and propylene glycol monomethyl ether. Toxicol. Appl. Pharmacol. 69:43-47.

Appendix D3 469 Propylene Glycol Monomethyl EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

Gingell R, Boatman RJ, Bus JS, Cawley TJ, Kn aak 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 eval uation of inhaled propylene glycol monomethyl et her in rats and rabbits. Funda m. Appl. Toxicol. 4:784-794.

HSDB. 1995. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland (TOMES ® CD-ROM Version). Denver, CO: Mi cromedex, Inc., (Edition expires 1/31/95).

IRIS. 1995. Integrated Risk Information Syst em. (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. Propylen e 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 a nd 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 toxici ty studies. Environ. Hea lth 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 gl ycol methyl ethers. Arch. Ind. Hyg. Occup. Med.

9:509-525.

Stewart RD, Baretta ED, Dodd HC, and Torkelso n TR. 1970. Experimental human exposure to vapor of propylene glycol monomethyl ether: Experimental human e xposure. Arch. Environ.

Health 20:218-223.

U.S. EPA. 1986. U.S. Environmental Protecti on Agency. Health and Environmental Effects Profile for Glycol Ethers. Prepared by the o ffice of Health and Environmental Assessment,

Environmental Criteria and Assessm ent Office, Cincinnati, OH for th e 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 EtherDetermination of Noncancer Chronic Re ference Exposure Levels March 2000

CHRONIC TOXICITY SUMMARY



# PROPYLENE OXIDE

(1-,2-propylene oxide; methyl et hylene oxide; propene oxide)

CAS Registry Number: 75-56-9

I. Chronic Toxicity Summary

Inhalation reference exposure level 30 μg/m 3 (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 C 3 H 6 O Molecular weight 58.08 Density 0.83 g/cm 3 @ 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/m 3 per ppm at 25° C

III. Major Uses or Sources

Propylene oxide is used as a fumiga nt 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 prepar ation of surfactants and oil demulsifiers (HSDB, 1994). The annual statewide emissions from fac ilities 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 OxideDetermination of Noncancer Chronic Re ference 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, identif ied three deaths due to mesothelioma, a rare

cancer associated with asbestos exposure, and a statistic ally 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 me dical care settings. A statistically significant d ecrease in observed deaths was f ound for all respiratory cancers, cancer of the bronchus and lung, circulator y disease, digestive diseases, ci rrhosis and other liver disease, and death due to accidents, poisonings, and viol ence. 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 (respectivel y) 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. Cumulativ e mortality was statis tically significantly

different from controls at 115 weeks in rats of both sexes exposed to 300 ppm propylene oxide. Cumulative mortality was also signi ficantly different from controls at 119 weeks in female rats exposed to 100 ppm. However, a contributing factor to the incr eased mortality in female rats was the presence of mammary tumors. Atrophy of the olfactory epithelium and degenerative changes in the respiratory epithelium were obser ved in both male and female rats following 28 months of exposure to 30, 100, or 300 ppm propylen e oxide. Severe hyperpla stic changes in the olfactory epithelium were obser ved in male and female rats following 28 months exposure to 300 ppm propylene oxide. Mild hyperplastic change s 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 surv ival was observed in male and female mice exposed to 400 ppm propylene oxide. Su rvival 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 ra ts exhibited an awkward gait; the rats were ataxic by the seventh wee k. Histopathological examinati on revealed axonal degeneration of myelinated fibers of the hindleg nerve and fasciculus gracilis indicating central-peripheral distal axonopathy.

Appendix D3 472 Propylene OxideDetermination of Noncancer Chronic Re ference 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 resp iratory epithelial hyperplasia in creased with exposure time and regressed after termination of expos ure, with complete re covery after 4 weeks. Cell proliferation (determined by bromodeoxyuridine incorporation) was elevated following 1 and 4 weeks of exposure, but decreased to cont rol values after 1 week of r ecovery. Degeneration of the olfactory epithelium was found af ter 4 weeks of exposure with a decrease in incidence and severity after terminatio n of exposure. Proliferation of olfactory epithelium was elevated during the 4-week exposure period and 1 week post-exposu re and returned to control values after 4 weeks of recovery. The author s report a 4-week NOAEL for pr opylene 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 indi cated 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 reso rptions per litter, with no change in total resorptions, was observed in rabb its exposed on days 1-19 of ge station. 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 ge station or daily for 3 weeks prior to mating and then daily on days 1-16 of gest ation. Reproductive capacity was impaired in rats exposed prior to breeding; the number of corpora lutea, impl antation sites, and live fetuses were reduced. Those dams exposed pregestationa lly to propylene oxide also e xhibited more resorptions. Maternal toxicity as indicated by decreased food intake and d ecreased body weight gain was observed in all exposed rats. Si gnificant reductions in fetal b ody weight and fetal crown-rump length were observed in all expo sed groups. An increased incide nce 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 toxici ty potential of propylene oxide in Fischer 344 rats. Four groups of 25 mated female ra ts were exposed to 0, 100, 300, and 500 ppm for 6 hours per day on gestation days 6 through 15. Cesa rean sections were performed on all females on gestation day 20 and the fetuses were remove d for morphological evaluation. Exposure to propylene oxide did not adversely aff ect survival, appearance, or beha vior at any level. Maternal body weight gain and food consum ption were reduced significan tly 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 t oxic exposure level of 500 ppm. Thus 300 ppm was considered the NOAEL.

Appendix D3 473 Propylene OxideDetermination of Noncancer Chronic Re ference 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 respirator y effects, RGDR = 0.23,

based on MV = 0.3 m 3 /day, SA(ET) = 11.6 cm 2 ) 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/m 3 , 30 μg/m 3 )

VII. Data Strengths and Limitations for Development of the REL

The chronic REL is equivalent to the US EP A RfC. The major strength of the REL for propylene oxide is the use of a well-conducte d, long-term, multi-con centration study with adequate histopathological analys es. 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 Do w Chemical Canada Inc., Sarnia, Ontario (1950- 1984).Can. J. Public Health 80(6):441-446.

Eldridge SR, Bogdanffy MS, Jokinen MP, A ndrews LS. 1995. Effects of propylene oxide on nasal epithelial cell proliferation in F344 ra ts. Fundam. Appl. Toxi col. 27(1):25-32.

HSDB. 1994. Hazardous Substance Data Bank. Na tional Library of Medicine, Bethesda, Maryland. (TOMES ® CD-ROM Version). Denver, CO: Mi cromedex, Inc. (Edition expires 11/31/94).

Appendix D3 474 Propylene OxideDetermination of Noncancer Chronic Re ference 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, a nd 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. T echnical Report on the Toxicology and Carcinogenesis Studies of Propylene Oxide in F344/N Rats and B6C3F 1 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 OxideDetermination of Noncancer Chronic Re ference 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

SeO 2 selenium dioxide; selenium oxide; selenious anhydride 110.96 7446-08-4

H 2 SeO 3 selenious acid 128.97 7783-00-8

SeOCl 2 seleninyl chloride; selenium oxychloride; selenium oxichloric 165.86 7791-23-3

Na 2 SeO 3 disodium selenite 263.01 10102-18-8

Na 2 SeO 4 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 μ g/m 3
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 SeleniumDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

II. Chemical Property Summary (HSDB, 1995; Weast, 1980; Canady and Hodes, 1994; ACGIH, 1992)

Description Se 0 crystal: metallic gray

H 2 SeO 4 , Na 2 SeO 3 : white crystals

H 2 SeO 3 , Na 2 SeO 4 : colorless crystals

SeO 2 : 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 SeO 2 : 340 ° C SeS: decomposes at 118-119 ° C Solubility Se 0 : insoluble in water, al cohol; slightly soluble

in CS 2 ; soluble in ether

H 2 SeO 4 : sol. in water; decomposes in alcohol

H 2 SeO 3 : sol. in hot water, alcohol

Na 2 SeO 3 : sol. in water

Na 2 SeO 4 : 84 g/100 ml water at 35 ° C

SeO 2 : 38.4 g/100 ml water at 14 ° C

SeS: insoluble in water Conversion factor Se 0 : not applicable (particulate)

SeO 2 : 4.5 μ g/m 3 per ppb at 20 ° C

III. Major Uses and Sources

Selenium occurs in four va lence states: selenates (Se 6+ ), selenites (Se 4+ ), selenides (Se 2- ), and elemental selenium (Se 0 ) (Goyer, 1991) which include compounds formed with oxygen, sulfur, metals, and/or halogens. Selenium compounds ar e used in the glass i ndustry as decolorizing agents and in the rubber industry as vulcanizing agents. Selenium com pounds are also found in toning baths used in photography and xerography, and in insecticides an d 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 (SeO 2 ) which catalyzes r eactions of organic compounds and is produced by the oxidation of selenium with nitric acid followed by evaporation or by burning selenium in oxyge n (HSDB, 1995). The largest anthropogenic sources of atmospheric selenium are from the combustion of fossil fuels and the production/refining of copper; pa rticulates are the primary e xpected 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 dietar y intake, the National Research Council has

Appendix D3 477 SeleniumDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

set a U.S. Recommended Daily Allowance (RDA) of 0.87 μ g/kg (55-70 μ g/person/day) (Subcommittee on the Tenth Edition of the RDAs, 1989). The average daily oral intake of

selenium is 125 μ g/person (U.S. EPA, 1991). Organic selenium compounds (e.g., dimethyl selenide) are known to occur as metabolites a nd as microbial degrad ation products in the environment. These compounds appear to have relatively low toxicity.

IV. Effects of Human Exposures

Acute occupational exposure to SeO 2 resulted in bronchospasm, irritation of the upper respiratory passages, violent coughing, and gagging with nausea and vomiting (Wilson, 1962).

The relationship between inhalati on exposure to selenium and the presence of selenium in the urine was investigated in a five year study of workers at a sele nium rectifying plant (Glover, 1967). Workers were exposed to fumes and dusts of elemental red selenium, which, the author reported, is converted 80% to SeO 2 in the presence of air. Average air concentrations of selenium were reported to be 3.6 mg/m 3 in grinding processes, 0.04 mg/m 3 in annealing processes, and a range of averages of 0.23-0.87 mg Se/m 3 in various “special” processes, e.g., punching, scraping, sorting, refining, and testing. The same au thor 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-, a nd high-selenium areas of China we re later studied to associate the symptoms with selenium intake. Estimated daily intake for adults in these areas were 70,
195, and 1438 μ g Se for males and 62, 198, and 1238 μ g for females, respectively (Yang et al.

, 1989). Selenium intake was highly correlated w ith whole blood, breast milk, and 24-hour urine

selenium levels. The authors al so suggested the possibility of liver dysfunction as indicated by a delay in prothrombin time among persons with intake of 750-850 μ g Se/day. More clearly recognized and characteristic cl inical signs, however, were only observed in a group exposed to greater than 1261 μ g Se/day and not among those exposed to less than 853 μ g 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, se rum, urine, and toenails (Longnecker et al. , 1991). Subjects completed health ques tionnaires, underwent physical examinations, provided blood samples for clinical assessment, and provided bl ood, urine, toenails, a nd duplicate-plate food collections for selenium analysis. About half of the 142 free-living subjects had selenium intakes greater than 2.54 m μ mol/day (200 μ g/day) (range 0.86-9.20 m μ mol/day, or 68-724

μ g/day). Average intake among th e population was estimated at 239 μ g Se/day. No clinical

Appendix D3 478 SeleniumDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

signs and no changes in hematological functi on, clinical chemistry, or liver function were observed in the population, even in subj ects whose intake was as high as 9.20 m μ mol/day (724

μ g/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/m 3 . 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 pe riods 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 inf iltration of large macrophages.

Guinea pigs exposed one time to con centrations “less than 0.021 mg H 2 Se/L” (22 mg Se/m 3 as hydrogen selenide) for 2, 4, or 8 hours exhibited difficulty breathing and a red-tinged discharge from the nose (Dudley and Miller, 1941). Mort ality studies were conducted with guinea pigs (16/group) using the same exposure duration and selenium concentrations ranging from 1 to 43 mg Se/m 3 . Fifty percent mortality was observed at 30 days among animals exposed once for 2 hours to 12 mg Se/m 3 . Mortality after 30 days was 50% among animals exposed once to 1 mg

Se/m 3 for 8 hours. Histopathological evaluation of guinea pigs exposed once for 4 hours to 8 mg

Se/m 3 showed fatty change to the liver, pne umonia, lymphoid hyperplasia, and increased reticuloendothelial tissue in the spleen. These e ffects did not begin to resolve until more than 17 days after the exposure.

Several studies have addresse d 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

Na 2 SeO 3 in drinking water over 50 weeks showed decr eased 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 Na 2 SeO 3 . Hamsters (8/sex/group) treated with 0.1 (unsupplemented), 1, 5, 10, or 20 ppm Na 2 SeO 3 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 SeS 2 , Na 2 Se, Na 2 SeO 3 , or Na 2 SeO 4 showed increased relative li ver weights and/or decreased body weight gain at 10 ppm (for each compound) ove r a 5 week exposure (Dausch and Fullerton, 1993). A 13-week drinking water study of Na 2 SeO 3 , and Na 2 SeO 4 in rats and mice showed increased mortality, decreased body weights, and hi stopathological changes to the kidneys in rats and decreased body weight and decreased wate r consumption in mice (Abdo, 1994). Decreased body weights were observed in rats treated fo r 6 weeks in drinking water with 2 ppm Na 2 SeO 3 or Na 2 SeO 4 (Palmer and Olson, 1974).

Decreased percentage of live spermatozoa, altered sperm mor phology, and decreased body weight gain were observed in rats (6/g roup) treated for 5 weeks with 2 ppm Na 2 SeO 3 in the diet (Kaur and Parshad, 1994). Rats (7-12/ group) exposed to 0, 4, 8, or 16 ppm Na 2 SeO 3 in drinking

Appendix D3 479 SeleniumDetermination of Noncancer Chronic Re ference 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 Na 2 SeO 3 and Na 2 SeO 4 at concentrations ranging from 0 - 110

μ mol/kg body weight (Ferm et al. , 1990). Effects observed at 100 μ mol Na 2 SeO 3 /kg included decreased fetal crown-rump length and increase d percentage of abnormal litters. At 90 μ mol

Na 2 SeO 4 /kg, an increased percentage of abnormal litters was observed. Mice (10 or 14/group) treated with 0, 3, or 6 ppm Na 2 SeO 3 in drinking water from 30 days pre-gestation through gestation showed altered estrus cycle length, decreased fetal grow th, 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 envir onmental 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 μ g/day for adult males;
62, 198, and 1238 μ g/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/m 3 per mg/kg-day Inhalation reference exposure level 20 μ g/m 3

*Factors: NOAEL (0.853 mg/day) and LOAEL (1.261 mg/day) calculate d from regression analysis (log Y = 0.767 log X - 2.248, where Y = bl ood selenium and X = selenium intake) based upon the correlation (r = 0.962) between dietary sele nium intake and blood selenium level for data showing incidence of clinical selenosis in adults base d on an average adult body weight of 55 kg.

The inhalation chronic REL is base d on the oral chronic REL, which is the same as the USEPA’s oral reference dose (RfD) (U.S. EPA, 1996). In a ddition to being inhaled, airborne selenium can settle onto crops and soil and enter the body by in gestion. Thus an oral chronic reference exposure level for selenium is also required for Ai r Toxics Hot Spots health risk assessments. The chronic inhalation REL was derived by rout e-to-route extrapolation of the RfD. The

Appendix D3 480 SeleniumDetermination of Noncancer Chronic Re ference 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 populati on of approximately 400 individuals living in an area of China with unusually high envi ronmental concentrations of selenium (Se). The subjects were evaluated for c linical 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-ade quate area (n=8) and low-selenium area (n=13). The Yang et al . (1989) studies provide a much larger sa mple 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 populat ion showed a positive correlation with blood and tissue Se levels. The daily average Se intakes, based on lifetime exposure, were 70, 195, and

1438 μ g for adult males and 62, 198, and 1238 μ g for adult females in the low-, medium- and high-selenium areas, respectively. Significant correlations, demonstrated between Se

concentrations of variou s tissues, were used to estimate the minimal daily Se intake values that elicited various alterations in biochemical para meters indicative of possible Se-induced liver dysfunction (i.e., prolongation of cl otting time and serum glutathione titer) and clinical signs of selenosis (i.e., hair or nail loss, morphological ch anges of the nails, etc.). In this manner, a marginal safe level of daily Se intake was estim ated. Analysis of the results indicated that persistent clinical signs of se lenosis were observed only in 5/349 adults, a potentially sensitive subpopulation. The blood selenium concentratio n in this group ranged from 1.054 to 1.854 mg/L with a mean of 1.346 mg/L. Clinical signs obser ved included the charact eristic "garlic odor" of excess selenium excretion in the breath and urine, thickened and brittle nails, hair and nail loss, lowered hemoglobin levels, mo ttled teeth, skin lesions, and CNS abnormalities (peripheral anesthesia, acroparesthesia, and pain in th e extremities). Alterations in the measured biochemical parameters occurred at dietary intake levels of 750-850 μ g/day. These alterations were described as a delay in prothrombin tim e, i.e., increase in blood coagulation time and reduction in blood glutathione concentration. However, th ese indicators were poorly characterized and are not typically used as an in dex for clinical selenosis resulting from chronic exposure to selenium (NAS, 1989). Based upon th e blood selenium levels shown to reflect clinical signs of selenium in toxication, a whole blood selenium concentration of 1.35 mg/L corresponding to 1.261 mg of daily selenium inta ke is indicative of th e lowest correlative selenium intake causing overt signs of selenos is. The next lowest whole blood selenium concentration of 1.0 mg/L, corre sponding to 0.853 mg selenium/d ay, 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 wa s applied to the NOAEL to account for sensitive

individuals. A full factor of 10 was not deemed necessary since simila r NOAELs were identified in two moderately-sized human populations exposed to selenium levels in excess of the RDA throughout a lifetime without apparent clinical si gns of selenosis. No modifying factor was applied by USEPA. OEHHA accepted the USEPA analysis.

Route-to-route extrapolation assumes by default th at a chemical is equa lly absorbed by the

inhalation and the oral routes a nd that the first pass effect due to metabolism by the liver is not important for the chemical. The latter assumpti on is applicable to most metals. There are

Appendix D3 481 SeleniumDetermination of Noncancer Chronic Re ference 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% (rang e = 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 depa rt 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 populat ion with sensitive subpopulations was studied, there are still several possible in teractions 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 clinic al, of selenium toxicity. Conf idence in the database is high because many animal studies and epidemiologi c studies support the principal study. An additional human study with a freestanding NOAEL (Longnecker et al., 1991) provides support

for the NOAEL identified in th e principal study. Longnecker et al . (1991) found no effects at

238 μ g 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 suppor t of the critical study and the high level of

confidence in the database.

There are insufficient data relating human inhala tion 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 H 2 Se and SeO 2 (Buchan, 1947; Wilson, 1962). Experiments in animals have shown that H 2 Se is toxic following inha lation exposure, with 8- hour exposures to concentra tions as low as 1 mg H 2 Se/m 3 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 H 2 Se, which may be considerably more toxic than othe r selenium compounds. At this time there are inadequate data to develop a REL for H 2 Se. 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 child ren as young as one year old. Thus the chronic REL should be protective of in fants 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 μ mol/kg-day). Ferm et al . (1990) did not find

adverse effects on hamster devel opment with Se doses below 34 μ mol/kg. Thus the chronic REL should also be protective of infants and children.

Appendix D3 482 SeleniumDetermination of Noncancer Chronic Re ference 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 B6C3 F1 mice. PB94-215753. Springfield, VA: NTIS.

ACGIH. 1992. American Conference of Gove rnmental Industrial Hygienists, Inc.

Documentation of the threshold limit values and biological exposur e 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 t oxics 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 se lenium. Nutr. Cancer 20:31-39.

Dudley HC. 1937. Toxicology of selenium. IV. Eff ects of exposure to hydrogen selenide. U.S. Public Health Reports 52:1217-1231.

Dudley HC, and Miller JW. 1941. To xicology of selenium. VI. Eff ects 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 concerni ng 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 populati ons. 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, D oull J, and Klaassen CD (eds.). New York: Pergamon Press, pp. 658-660.

Hall RH, Laskin S, Frank P, Maynard EA, a nd Hodge HC. 1951. Preliminary observations on toxicity of elemental selenium . Arch. Ind. Hyg. Occup. Med. 4:458-464.

Appendix D3 483 SeleniumDetermination of Noncancer Chronic Re ference Exposure Levels December 2001

HSDB. 1995. Hazardous Substances Data Bank. Na tional Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Microm edex, 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 diet ary selenium on differe ntiation, 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 Wille tt WC. 1991. Selenium in diet, blood, and toenail

in relation to human health in a seleni ferous area. Am. J. Clin. Nutr. 53:1288-1294.

NAS. 1976. National Academy of Sciences. Sele nium. Committee on Medi cal and Biological Effects, Environmental Pollution Subcom mittee 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. Effe cts of sodium selenite on methylmercury embryotoxicity and teratogenicity in mi ce. Toxicol. Appl. Pharmacol. 47:79-88.

Owen BA. 1990. Literature-derived absorption co efficients 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 fa tal case of gun blue ingestion in a toddler. Vet. Hum. Toxicol. 42(2):96-98.

Subcommittee on the Tenth Ed ition of the RDAs 1989. F ood 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 Pr otection 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 Pr otection Agency. Title 40, Code of Federal Regulations, Parts 141, 142, and 143. National Prim ary Drinking Water Regulations; Final Rule. Federal Register, 56:3538-3539.

U.S. EPA. 1996. United States Environmental Pr otection Agency. Selenium. IRIS (Integrated Risk Information System). Available on line at http://www.epa.gov/ngispgm3/iris

Weast RC. (ed.) 1980. CRC Handbook of Chemistry and Physics. 61 st ed. Boca Raton, FL: CRC

Press.

Appendix D3 484 SeleniumDetermination of Noncancer Chronic Re ference 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 SeleniumDetermination of Noncancer Chronic Refe rence 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/m 3 [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 SiO 2 Molecular weight 60.09 g/mol Density 2.65 g/cm 3 @ 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 sili con and oxygen atoms are arranged in a definite regular pattern throughout the crystal. The characteristic crysta l 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, crist obalite, 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 μ m particle aerodynamic diameter. This occupational definition of respirable differs from the environmental definition of respirable, which is PM 10 . (The occupational particle category “thor acic” has a 50% cut-point at 10 μ m particle diameter (ACGIH, 2004) and the category “inh alable” has a 50% cut-point at 100 μ m 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 h eating to 1000-1100 oC (calcining). Silica is often associated

Appendix D3 486 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

with silicates, which, in addi tion 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, abrasive s, ceramics, and enamels, in scouring and grinding compounds, and in molds for cas tings. 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 fi ller. Sources of ambient respirab le crystalline silica in California include mines, quarries, diatomaceous earth cal cining plants, sand blastin g, and entrained fines (e.g., PM10) from surface soil. The annual stat ewide 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 crystal line silica (CARB, 2001. The fraction, which is respirable as defined either occupatio nally 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 captu red in less than a s econd) was collected (impacted) onto a small area of a glass slide coat ed with adhesive. Total dust particles were counted and expressed as dust par ticles per cubic centimeter. Late r, slides were heated to 500- 550 oC (ignition) to remove carbonaceous materi als and immersed in ho t 50% hydrochloric acid followed by a second ignition to remove acid-soluble materials. The remainder was mostly silica particles, which could be counted. The konimete r was superseded by the thermal precipitator, which also deposited particles onto glass but coul d 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 sl ides but more recently by an automated method (Kitto, 1960; 1970).

In the United States the impinger method wa s 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 expresse d 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 du st was responsible for silicosis, respirable dust was collected onto filters using size-specific dust co llectors, 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 airflo w through the apparatus. Quartz dust was quantified by examining filters in an electron mi croscope 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 us es one of three approved cyclones and a 5

μ m PVC membrane filter to sample, and X-ray di ffraction 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 μ m aerodynamic diameter (ISO, 1995; ACGIH, 2004).

Appendix D3 487 SilicaDetermination of Noncancer Chronic Refe rence 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 f acilities 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 earl ier 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 compar e the earlier silica measurements with the more

recent ones.

IV. Effects of Human Exposures

Inhalation of crystalline silica initially causes re spiratory 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 sixt een years, the small airways become obstructed as measured by pulmonary function tests (e.g., de creased FEV1) in granite quarry workers (no measurement of silica levels repo rted; Chia et al., 1992). In a report on the hazards of exposure to crystalline silica, th e American Thoracic Society (1997) stat ed: “Studies from many different work environments suggest that exposure to working environments contaminated by silica at dust levels that appear not to cause roentge nographically visible simp le silicosis can cause chronic airflow limitation and/or mucus hypersec retion and/or pathologic emphysema.” Hnizdo and Vallyathan (2003) also concluded that “chr onic 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 crys talline silica have also been found at autopsy in the lungs of granite workers who lacked radiol ogical evidence of silic osis (Craighead and Vallyathan, 1980).

Silicosis results from chronic exposure; it is ch aracterized by the presence of histologically unique silicotic nodules and by fibr otic scarring of the lung. Th e histological progression of

silicosis has been described as : (1) granuloma composed of hi stiocytic cells, collagen, and lymphocytes; (2) cellular fibrotic nodule with irregular collagen at th e 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/silic otuberculosis, chronic bronchitis , small airways disease, and emphysema (Oxman et al., 1993; Park et al., 200 2; 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 pa rticles 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 protei n-1, platelet activating factor) are inflammatory and attract neutroph ils to the site of inflammation, while other factors (fibronectin,

Appendix D3 488 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

alveolar macrophage-derived growth factor) stimu late fibroblasts to proliferate and to make collagen. Since silica particles cannot be digested by the macropha ge, the inflammatory process becomes chronic (frustrated phagocytosis). An increased silica burden leads to more foci of inflammation, nodule formation, and fibrosis. The in ternal process can con tinue 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 hypertrop hy 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 silicosi s is often based on chest radiogra phs. Recent papers have used the 1980 classification by the In ternational Labor Organization (ILO, 1980) to identify and classify silicosis into categor ies and subcategories of seri ousness by comparison of patient radiographs with ILO-supplied refe rence 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 ar eas 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 defin itely present; other areas of the radiograph may indicate more adva nced 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 a nd 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 t ool, not as a diagnostic classification on an individual basis. In occupati onal medicine practice, a group of tests is used to clinically diagnose silica-related lung dis ease including physical examina tion, X-rays, and high resolution computed tomography (CT) scans of the lung (e .g., Begin et al., 1991; Ol ivetti et al., 1993).

A. Environmental silicosis

Several studies have reported "environmental sili cosis", 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-occupa tional pneumoconiosis in Ladakh, India, high in the western Himalayas wh ere there are no mines or industries. Among 449 randomly selected inhabitants of three vill ages, there were many cases of pneumoconiosis

Appendix D3 489 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

associated with progressive massive fibrosis ( nodules > 1 cm) and "egg sh ell" calcification of hilar glands. The prevalence of pneumoconios is was 2.0% (3/150) in the village of Saboo, 20.1% (31/149) in Shey, and 45.3% (68/150) in C hushot, and corresponded w ith the severity of dust storms and the presence or ab sence of chimneys in the kitchens (i.e., ventilated cooking). Without chimneys (Chushot), dus t concentrations in kitchens averaged 7.5 mg/m3 during cooking periods. The free silica content of the du st 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 occupationa l, 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 si milar to some occupational exposures.

B. Occupational silicosis

Several relatively recent reports have presente d data that allow a qua ntitative 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 al pha-quartz) and the development of silicosis in 2109 male hard rock (uranium, gold, mixed metals) miners in On tario, 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 read ers examined chest radiographs; one or more

readers identified 32 cases of s ilicosis, defined as ILO categor y 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[

− ( α x) β ] (x ≥ 0, β > 0)

gave the best fit to the data for cumulative ri sk R of silicosis as a function of cumulative exposure in units of (mg/m 3 )-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/m 3 )-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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

The Ontario cohort gives the shallowest dose-res ponse 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 fo r 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 ha lf 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) evalua ted 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 we re analyzed for those with at least 5 years of employment as of June 1991. Nearly half had worked at th e foundry for 20 years. Sixty had radiographic evidence of pneumoconiosis (ILO categories 1/ 0 and greater). Twenty-eight workers had radiographs consistent with sili cosis; of these 25 had simple s ilicosis and three had progressive massive fibrosis. The prevalence of radiographi c changes consistent with silicosis increased with years at the foundry, work area, quantitat ive silica exposure, and cigarette smoking. In regard to quantitative silica exposure, the auth ors stated that 0.3-2.7% of workers at the OSHA standard (90-100 μ g/m 3 ) were silicotic, as were 4.9-9.9% of workers above 100 μ g/m 3 . 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 af ter 20 years of work at 100 μ g/m 3 and an odds ratio of 2.10 after 40 years of work at 100 μ g/m 3 (Tables 3 and 4). This study prob ably underestimates risk due to lack of follow-up of the current workers. Alth ough 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/m 3 )-y 0.4% 2-6 (mg/m 3 )-y 2.7% > 6 (mg/m 3 )-y 10%

Appendix D3 491 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Table 4. Odds ratios for silico sis (from Table 8 of Rosenman et al .) a

Time-weighted average silica exposure (mg/m 3 ) 20-year
cumulative

exposure [(mg/m 3 )-y] Odds ratio (95% C.I.) 40-year
cumulative

exposure [(mg/m 3 )-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 calculat ed 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 fo rmed when the amorphous silica in diatomaceous earth is calcined at 1000-1100 oC. Quantitative estimates of dust exposure were made and published in the pe er-reviewed literature by Seixas et al . (1997) based on 6395 air sampling records taken from 1948-1988. The average estim ated respirable dust concentrations for 135 jobs were 3.55 ± 1.25 mg/m 3 prior to 1949, 1.37 ± 0.48 mg/m 3 from 1949-1953, 0.47 ± 0.16 mg/m 3 from 1954-1973, and 0.29 ± 0.10 mg/m 3 from 1974-1988. The workers had periodic chest radiographs. Based on the median of radi ographic readings by thre e independent readers, 81 workers (4.5%) were judged to have opacitie s 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 crystall ine silica. The concen tration of respirable crystalline silica was an important determinant of risk after accounting for cumulative exposure. The workers were split into two cat egories: those exposed to < 0.50 mg/m 3 (or hired after 1950) and those exposed to > 0.50 mg/m 3 (or hired before 1950). The risk of opacities for a cumulative exposure to crystalline silica of 2.0 mg/m 3 -yr is shown in Table 5.

Table 5. Silica exposure and silicosis based on data of Hughes et al . (1998)

Average crystalline silica exposure Cu mulative risk of silicotic opacities < 0.50 mg/m 3 (or hired after 1950) 1.1% > 0.50 mg/m 3 (or hired before 1950) 3.7%

The findings of Hughes et al. (1998) indicate an exposure- response relationship between cumulative exposure to crystallin e 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 quart z (i.e., prior to 2000 th e occupational limit for

cristobalite was half that fo r quartz), since at average sili ca levels comparable to other

Appendix D3 492 SilicaDetermination of Noncancer Chronic Refe rence 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 forsilicosis will require evaluation at autopsy. The ACGIH recently lowered the TLV for alphaquartz from 100 to 50 μ g/m 3 , so that it has the same TLV as cristobalite (ACGIH, 2000).

Park et al . (2002) carried out a quantit ative risk assessment, by Poisson regression methods, of the onset of silicosis among the diatomaceous eart h workers in Lompoc. A linear relative risk model gave the best fit to the data. They es timated an excess lifetime risk for radiographic silicosis of 68-75 cases per thousand workers exposed to 50 μ g/m 3 silica (cristobalite) for a 45 year work-life, then living to age 85. At 1 μ g/m 3 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/m 3 ) 45 year cumulative exposure in mg/m 3 -y Radiographic
silicosis - all workers Radiographic
silicosis in workers with < 10 mg/m 3 -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) in vestigated silicosis risk retr ospectively in a cohort of 2,235 white male South African gold miners. Exposur e 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-hou r shifts. The workers had a minimum of 10 years and an average of 24 years service from 19 40 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 we re followed until 1991 for radiographic signs of the onset of silicosis. An ILO category 1/1 (d efinite silicosis) or greater was selected to designate silicosis. Two indepe ndent readers initially read the chest films, but only the reader whose interpretations correlated better with auto psy results was used for additional analysis; the use of one reader is a limitation of the study. Th ere were 313 miners (1 4% of the cohort) who developed radiographic signs of si licosis at an average age of 55.9 years. The latency period was largely independent of the cumulative dust exposu re (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 in creased exponentially with cumulative dust dose. At the highest level of 15-(mg/m 3 )-years CDE (approximately 37 years of gold mining at an

Appendix D3 493 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

average respirable dust concentration of 0.4 mg/m 3 ), the cumulative risk for silicosis reached 77% as estimated by the accelerate d 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 LI FETEST) below (Table 7) shows the number of miners who developed silicosis (“cases”), the numb er of miners considered by the authors to be at risk, and the risk per unit of CDE (also as calc ulated by the authors). In the table in column 1 (in parentheses) are OEHHA’s determination of the mg/m 3 –yr respirable sili ca exposure, based
on Hnizdo and Sluis-Cremer’s estimate of 30% sili ca 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 silico sis per unit Cumulative Dust Exposure (CDE) (from Table IV of Hnizdo and Sluis-Cremer, 1993)

Midpoint in (mg/m 3 )-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/m 3 ) 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 respir able 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 Cumula tive Dust Exposure (CDE) versus the mid-point unit CDE, as given in Figure 1 of the Hnizdo and Sl uis-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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Figure 1. Risk of silicosis p er 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 wo rkers 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 la rge 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 ag e = 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/m 3 (0 - 0.70); the mean quartz concentration was 0.053 mg/m 3 (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 silicosi s as indicated by ILO profusions ≥ 2/1 (see Table 1).

Appendix D3 495 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Significant trends were found between the prevalen ce of silicosis and: (1 ) length of service (OR = 1.69 per 5 years), (2) mean intens ity of exposure (OR = 1.18 per 0.01 mg/m 3 ), 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. Fo r this cohort the preval ence of silicosis will increase even if the miners stop mining immedi ately. 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/m 3

. Table 8. Silicosis in black gold miners (Churchyard et al ., 2003; 2004)

Cumulative quartz exposure in mg/m 3 -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 S outh African gold miners dying from

unnatural causes between 1975 and 1991 in order to study change in preval ence 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 mo st 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 ag e 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 am ong 338 male workers, who had worked at least one year between 1967 and 1985 in two granite qua rries 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, a ssigned by at least two of the three readers. Exposure was estimated for each worker based on job category a nd 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Table 9. Silica exposure and silicosis in Ng and Chan (Finkelstein, 2000)

Mean cumulative exposure (mg/m 3 )-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 underg round workers grouped into five categories.

The authors estimated that most miners were exposed to a median silica level of 0.05 mg/m 3 , but

that those hired before 1930 were ex posed to a median level of 0.15 mg/m 3 . 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 survey s in 1960 and 1976 (n = 29 cases; ILO category 1/1 or greater), or from both (n = 13 cases). Unfo rtunately, 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/m 3 -years and increased to 68% to 84% for the highest cumulative exposure category (more than 4 (mg/m 3 )-years) (Table 10).

Table 10. Risk of silicosis for cohort by cumu lative exposure (Table 3, Steenland and Brown)

Silica exposure in (mg/m 3 )-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 0.017-0.022 b 15.4 1942 1.0-2.0 (1.50) 33 684 353 0.060-0.084 b 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 0.403-0.534 b 25.5 1921

>4.0 26 52 52 0.678-0.844 b 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 a nd calendar time (Steenland and Brown, 1995)

The best predictor of diseas e was cumulative exposure ((mg/m 3 ) – years), followed by duration of exposure (years), and then by average exposure (mg/m 3 ). Figure 1 of Steenland and Brown indicates that a plot of their da ta for silicosis risk versus cumu lative silica exposure was similar

Appendix D3 497 SilicaDetermination of Noncancer Chronic Refe rence 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 μ g/m 3 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 rath er than on ILO interpretation of radiographs. In addition, no mention was made of validating the data on the death certif icates. It was also not clear what, if any, autopsy data were availa ble. A plot of silicosis incidence among the workers (as determined by OEHHA staff) actually exposed to th e 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 life time 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 exposur e-response relationships for silicosis among 134 male miners over 40 years old in Leadville, Co lorado. The men had been studied three years earlier in a random sample of respirat ory disease in their community (Kreiss et al ., 1989). Of 100 dust-exposed miners, 32 had radiological profus ions of small opacities of ILO category 1/0 or greater at a mean of 36.1 y ears since their first s ilica exposure. Of mi ners with cumulative silica, exposures of 2 (mg/m 3 )-years or less, 20% had sili cosis while 63% of miners accumulating greater than 2 (mg/m 3 )-years had silicosis. Average silica exposure was also strongly associated with silicosi s prevalence rate s (Table 11).

Table 11. Miners studied by Kreiss and Zhen (1996)

Average silica exposure % silicotics 0.025-0.05 mg/m 3 13% (5/38) > 0.05-0.1 mg/m 3 34% (15/44) > 0.1 mg/m 3 75% (9/12) Cumulative silica exposure % silicotics

≤ 2 (mg/m 3 )-y 20% (14/70)

2

− 4 (mg/m 3 )-y 63% (15/24)

Appendix D3 498 SilicaDetermination of Noncancer Chronic Refe rence 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) cumu lative silica exposure or (2) a combination of average silica exposure and duration of exposure. Exposure-re sponse relationships were substantially higher using measured silica exposures (compared to using estimated silica exposures based on measured total dust exposures and assuming a cons tant silica proportion of dust). The risk of silicosis in this study is higher than in workfo rce studies having no follow- up of those leaving the mining industry (e.g., Muir et al ., 1989) and in studies withou t job title-specific silica measurements (e.g., Hnizdo and Sluis-Cremer, 1993) . However, the risk is comparable to

several recent studies of e xposure-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-r esponse relationship between silica dust exposure and silicosis in a cohort of 3010 (2795 male and 215 fe male) miners employed for at least 1 year during the period 1960-1965 in any of f our Chinese tin mines. No othe r diseases due to silica or

tin were mentioned. Each cohort member was followed through 1994. Hist orical Chinese total dust (CTD) data were used to create a job e xposure matrix for each facility, job title, and calendar year. The CTD data were converted to estimates of re spirable crystalline silica for comparison with findings from other epidemiologi cal 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 Ch inese 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 e nded (mean = 3.7 years after).

The risk of silicosis was str ongly related to cumulative expos ure to silica. The Weibull distribution gave a very good fit to the data. The risk of silico sis was less than 0.1% when CTD was less than 10 (mg/m 3 )-yr (= 0.36 (mg/m 3 )-yr of respirable crystalline silica). The risk of
silicosis increased to 68.7% when CTD exposure was equal to 150 (mg/m 3 )-yr (= 5.4 (mg/m 3 )-yr of respirable crystalline silica). Latency period was not correlated to the risk of silicosis or to cumulative dose. From their data, the authors pr edicted a 55% risk of silicosis for 45 years exposure to 0.1 mg/m 3 respirable crystalline silica, the workplace exposure limit (4.5-(mg/m 3 )- years silica). Figure 4 plots the fraction of the work ers 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Table 12. Cumulative silicosis risk ba sed on cumulative total dust (CTD) (Table 5, Chen et al ., 2001)

Range of CTD

exposure in (mg/m 3 )-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 empl oyed before 1980 for 3 years or more and followed through 1994 in one of nine North American sand-producing pl ants 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 silicotuberculos is. The mean exposure of the

cohort was 42 μ g/m 3 silica (Rando et al. , 2001). Odds ratios for sili cosis 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 SilicaDetermination of Noncancer Chronic Refe rence 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/m 3 )-y Silicotics (n) Odds ratio a for mortality Median exposure in (mg/m 3 )-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 cumula tive exposure category. Although labeled a cohort

study, the data analysis compared cases of silicosis with non- silicotic controls.

b Significant increasing trend acro ss 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. Th e workers were surveyed duri ng the period 1974-1987 and followed through 1991 with annual chest ra diographs. The cumulative risk of silicosis (ILO category 1/1 or greater) was 48% after 30 ye ars of employment. A multivaria te 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 f ound 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 radi ographs. 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/m 3 – 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 au thors reported a p value of 0.002. However, the study is limited by the lack of follow-up of the workers.

Appendix D3 501 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Slate workers (Glover et al ., 1980; Saiyed et al ., 1985; Saiyed and Bannerjee, 1985)

Slate contains calcium carbonate, iron oxides, sili cates, amorphous silica, a nd crystalline silica.

Glover et al . (1980) studied slate workers in North Wales. The resp irable slate dust contained 13-32% crystalline silica. In the study group we re 725 current and former workers exposed only to slate dust, while the contro ls were 530 men from the same area who had never been exposed

to dust. Pneumoconiosis was found in 239 sl ate workers (33 %), a nd 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 FEV 1 ) additional to the effect of age. The high prevalence (40- 50%) of radiological lesions suggest ed the presence of healed tube rcular lesions in men over 55. Either pneumoconiosis or old tubercular lesions (or both) co uld 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/m 3 ), 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 l ung lesions were detectable after less than five years of exposure to slate dust. Saiyed and Bannerjee (1985) condu cted a follow-up examination 16 months later. The progression of silicosis was ve ry rapid, and a total of 23 work ers 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 workpl aces are limited. According to Witschi and Last (2001), silica particles with a diameter of 1 μ m (range = 0.5 - 3 μ m) appear to be the most fibrotic in humans. NIOSH ( 1974) reviewed the existing litera ture and found that in five diatomite plants the mean silica diameter was 1.1 μ m (range = 0.5 - 2 μ m). For nine potteries, the particle size was 1.2 μ m. For 18 foundries, more than 90% of the particles were less than 3

μ m. The majority of particles to which shipyard sandblasters were exposed was also less than 3

μ m. In the Vermont granite sheds, 10 mppcf (m illion particles per cubic foot) granite dust were initially estimated to be equal to 0.1 mg/m 3 respirable quartz. Steen land and Brown (1995) used this estimate for silica in South Dakota gold mine s. Assuming that the density of quartz is 2.65 g/cm 3 and that the quartz particles are spherical, the data indicate that the particles have a

diameter of 0.59 μ m. NIOSH (1974) listed 0.94 μ m as the median particle size in metal mines. No indication was given of the di spersion 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/m 3 silica. For that estimation, OEHHA staff calculated th e particle diameter to be 0.53 μ m. Thus, existing data indicate that the majority of silica in the workpl ace is respirable. In most of the occupational studies examined, the exposures were measured using a calibrated cyclon e sampler similar to

that recommended in the current NIOSH (2003) method. This allows collection of particles

Appendix D3 502 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

primarily in the 0.5 – 5 μ m range, with a collection efficiency profile intended to match the penetration of particles into the alveolar regi on of the human lung. In the case of the South African gold mine studies (Beadle, 1971; Pa ge-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 μ m). Thus, the risk estimates obtained from these studies refer to particles in the size range where penetration occurs into the resp iratory region of the lung. This corresponds to the size range of particles thought to be responsible for silicosis. It differs from

the definition of “res pirable” particles ( i.e. PM 10 ) commonly used in environmental measurements, which refers to particles capable of penetrating anywhere in the lower respiratory tract (described as “thoracic” pa rticles 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/m 3 )-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 undere stimate 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

Table 15. Summary - Estimates of % silicosis based on cumulative silica exposure in (mg/m 3 )-y

Study Population (number with silicosis) Exposure of 2 (mg/m 3 )-y Exposure of 4 (mg/m 3 )-y Exposure of 4.5 (mg/m 3 )-y

Muir et al

., 1989 2109 male Ontario

hard rock miners (“15”) 0.4 a,c 1.2 a,c 2 b

Rosenman et al ., 1996 1072 Midwestern foundry workers (28) 2 a 10 a 3 b

Graham et al ., 1991 408 Vermont granite workers (35) ~3 c

─ ─

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) ~7 e 13 e 14 e

Hnizdo &

SluisCremer 1993 2235 white male South African gold miners (313) 5 a

10 c 52 a

60 c 77 b

Ng & Chan,

1994 338 male Hong Kong granite workers (36) 6 a 15 a 15-20 b

Steenland & Brown, 1995 3330 male S. Dakota gold miners (170) 8 a 53 a 70 b

Kreiss & Zhen, 1996 100 miners in Leadville, CO (32) 11 a 53 a 92 b

Chen et al

., 2001 3010 Chinese tin miners (1015) 14 d 47 d 55 b

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 st aff 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 SilicaDetermination of Noncancer Chronic Refe rence 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 st udies (Table 16). Th e study of Hnizdo and Slui s-Cremer (1993) yielded

both a LOAEL and a NOAEL.

Table 16. Estimates of NOAELs and LO AELs for silicosis (Rice and Stayner, 1995)

Study Subjects NOAEL in μ g/m 3 LOAEL in μ g/m 3

Davis et al ., 1983 969 granite workers 67.5

Hnizdo and SluisCremer, 1993 2235 gold miners 7 20

1321 gold miners

- 8 a 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 th is 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 μ g/m 3 . Thus, there have been recommendations that the occupational exposure limit for respirable, crys talline silica (specifica lly alpha-quartz) be

lowered from the current level of 100 μ g/m 3 to 50 μ g/m 3 (NIOSH, 1974; Rosenman et al ., 1996; ACGIH, 1999; Finkelstein, 2000) . In 2000, the ACGIH lowered its TLV for quartz from 100 to

50 μ g/m 3 . In 1986, WHO recommended that the occupational level be set at 40 μ g/m 3 (WHO, 1986). Greaves (2000) recommended that the TLV be lowered to 10 μ g/m 3 . Based on existing data Greaves (2000) estimated that at 10 μ g/m 3 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 μ g/m 3 . "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 mu st be lower than 0.14 mg/m 3 (or lower than 0.005 mg/m 3 respirable crystalline silica)" (Chen et al ., 2001). Mannetje et al . (2002) pooled data from six occupational cohorts. These included four groups discusse d 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 μ g/m 3 silica was 13 per 1000, while the risk of death at 50 μ g/m 3 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 stat e 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

C. Silica exposure and lung cancer in workers

In 1997, IARC classified respirable crystallin e silica in Class 1, a Known Human Carcinogen, based on occupational epidemiologic studies. Ho wever, chronic RELs are not based on cancer endpoints. Further, there is no approve d cancer potency factor for silica.

V. Effects of Animal Exposures

Several papers have reported th at freshly fractured quartz, whic h 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 ag ed for 1-2 days in decreasing the membrane integrity of male rat macrophages; 50% more po tent in activating hydro gen 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/m 3 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 histopa thological evidence of increased pulmonary infiltrates, showed higher le vels of biochemical markers of lung injury, increased lipid peroxidation, and increased the ab ility of pulmonary phago cytes to produce more oxygen radicals than air-exposed co ntrols. These pulmonary responses were significantly more pronounced after inhalation of 22.4 mg/m 3 freshly fractured quartz.

Table 17. Cells recovered in bronchoalve olar 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/m 3 Min-U-Sil brand crystalline silica. By 24 w eeks, 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 cr istobalite silica (m ass median aerodynamic diameter (MMAD) = 1.7 μ m) 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/m 3 for 8 days, 43 mg/m 3 for 9 days, and 70 mg/m 3 for 12 days) caused (1) increased initial lung dust burden at 12 to 16 weeks post-exposure, (2) pr ogressively intense pa thological responses, and (3) increased total lung collagen (as meas ured by hydroxyproline).

The histopathological changes a nd total lung collagen increased with time after exposure. Silicosis was compared in four inbred stra ins of mice (BALB/c, C3H/HeN, MRL/MpJ, New Zealand Black) 16 weeks after aerosol inhalation exposure to cr istobalite (70 mg/m 3 , 5 hours/day, 12 days). C3H/HeN mice had hist opathological silicotic lesions, enlarged

Appendix D3 506 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

intrapulmonary lymphoid tissue, and increase d lung wet weight, increased bronchoalveolar lavage (BAL) recoverable macrophages, lymphocyt es, and neutrophils, a nd 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 al veolar proteinaceous deposits , inflammation, and fibrosis. The authors found both dose-time-response relationshi ps and a substantial variation of responses among mouse strains to the high le vel, short duration exposure.

At Brookhaven National Laboratory, groups of Fischer 344 rats we re exposed to 0, 2, 10, and 20 mg/m 3 Min-U-Sil brand silica (al pha-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 coll agen, total elas tin per unit lung dry weight, and total protein per unit lung dry weight at 2 mg/m 3 silica and microscopic evidence of silicotic lesions at the higher silica le vels indicated that 2 mg/m 3 was a LOAEL for silica effects. After six months in clean air, the silica-i nduced lesions appeared to worsen.

Muhle et al . (1989) exposed groups of 50 male and 50 female rats to 1 mg/m 3 DQ12 quartz six hours/day, five days/week for 24 months. DQ12 c ontains 87% crystallin e alpha-quartz, has a mass median aerodynamic diameter (MMAD) of 1.3 μ m, and is 74% respirable. Moderate fibrosis was seen in 85 animals, s light fibrosis in 13, and very slight fibrosis in 1 (total rats with fibrosis = 99/100). Varying amounts of peribronchia l granulomatous foci were noted in 95 rats.

Muhle et al . (1998) reported lung fibrosis in hamsters exposed to 3 mg/m 3 DQ12 silica. After 18 months of exposure to DQ12 for 6 h/day, 5 da ys/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. Af terward 22.2% had very slight fibrosis, 68.7 % had slight fibrosis, and 1% had modera te fibrosis (i.e., more than 90/ 100 hamsters had lung fibrosis). No collagen measurements were reported. T hus, rats, mice, and hamsters show pulmonary fibrosis after crystalline sili ca exposure at and above 1 mg/m 3

.

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% cristoba lite (in calcined diatomaceous earth). Dust exposures were 2 and 5 million particles per c ubic foot (mppcf), equivalent to 0.2 and 0.5 mg/m 3
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 res ponses in Balb/c mice following inhalation of 1.954 mg/m 3 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 po ssible recovery from eff ects of dust inhalation. Silica inhalation suppressed the number of specific plaque-formi ng cells (PFC) in the spleen produced in response to aerosolized E. coli . After 570 days of inhalation, silica also reduced the ability of alveolar macr ophages 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

not affect antibody-dependent cell-mediated cy totoxic 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 i nhalation analyzed by USEPA (1996)

Study Species Duration a LOAEL

Muhle et al ., 1989 Rat 24 mo 1.0 mg/m 3

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 th e 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 interp retation 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 effect s, is much more inflammatory in rat lung than respirable silica collected fr om two workplaces (Clouter et al ., 2001).

Humans appear to show adverse effects of silica exposure at lowe r levels than animals (compare

LOAELs in Table 18 to LOAELs/NOAELs in Ta ble 16). Rodents tend to be obligate nosebreathers and to have extensive nasal turbinat es, 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 SilicaDetermination of Noncancer Chronic Refe rence 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/m 3 -years CDE (9 miners with silicosis)

NOAEL 2 mg/m 3 -years CDE (0 miners with silicosis) or

600 μ g/m 3 -years silica (dust = 30% silica)

BMCL 01 2.12 (mg/m 3 )–yr CDE or 0.636 (mg/m 3 )–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 μ g/m 3 -yr silica at BMC 01 (636 x 10 m 3 /20 m 3 x 270 shifts/365 days)

235 μ g/m 3 -yr/24 yr = 9.8 μ g/m 3
Human Equivalent Concentration (HEC) 9.8 μ g/m 3
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 μ g/m 3 (based on 30% silica in mine dust) [respirable, as defi ned 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/m 3 -years (5 miners with silicosis)

NOAEL Not found

BMCL 01 0.34 (mg/m 3 )–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 μ g/m 3 -y (340 x 10 m 3 /20 m 3 x 5 d/7 d x 48 wk/52 wk)

112 μ g/m 3 -y/9 y = 12.4 μ g/m 3
Human Equivalent Concentration (HEC) 12.4 μ g/m 3
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 μ g/m 3 [respirable, as defined occupationally]

Appendix D3 509 SilicaDetermination of Noncancer Chronic Refe rence 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/m 3 -years (17 workers with silicosis)

NOAEL ≤ 1 mg/m 3 -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 μ g/m 3 -y (1000 x 10/20 x 5/7 x 48/52)

≤ 330 μ g/m 3 -y/ 11.5years = ≤ 29 μ g/m 3
Human Equivalent Concentration (HEC) 29 μ g/m 3
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 μ g/m 3 [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/m 3 -years (24 cases)

NOAEL ≤ 10 mg CTD/m 3 -years (2 cases)

≤ 360 μ g silica/m 3 - years BMCL 01 132 μ g silica/m 3 - years Exposure continuity 8 h/day, 5 d/wk Exposure duration 2.2 years for NOAEL group Average experimental exposure 40 μ g/m 3 -y (132 x 10/20 x 5/7 x 48/52)

40 μ g/m 3 -y/2.2 years = 18 μ g/m 3
Human Equivalent Concentration (HEC) 18 μ g/m 3
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 μ g/m 3 [respirable, as defined occupationally]

Appendix D3 510 SilicaDetermination of Noncancer Chronic Refe rence 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/m 3 -yr (11 cases)

NOAEL Not identified

BMCL 05 0.673 (mg/m 3 )-yr Exposure continuity 270 shifts/year Exposure duration 21.8 yr (6.3-34.5) Average experimental exposure 249 ( μ g/m 3 )-yr (673 x 10/20 x 270shifts/365) 249 ( μ g/m 3 )-yr/21.8 yr= 11.4 μ g/m 3
Human equivalent concentration (HEC) 11.4 μ g/m 3
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/m 3 )-yr CDE (600 μ g/m 3 -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/m 3 -yr silica (0/204 vs. 9/474, p = 0.064 by Fisher exact test, two-tailed). Because this incidence represen ts approximately the sensitivity limit of the data, and silicosis is a se vere irreversible endpoint, the BMCL 01 ( 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,

BMCL 05 is typically regarded by OEHHA as equivale nt to a NOAEL. However, the power of this large-scale study is sufficien t to demonstrate measurable re sponses below the 5% incidence level (which cannot then be logically consider ed a no-effect level). Furthermore, the endpoint measured in this epidemiological study is considered to be severe, since it represents the occurrence of clinically recogni zable and irreversible diseas e, rather than an adverse physiological or biochemical response or a hi stopathological result seen at autopsy. 1

3

3

4 μ g/m 3 [respirable, as defined occupationally]

Benchmark Concentration (BMC) models, de veloped 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 MLE 01 of 2.45 (mg/m 3 )–yr CDE and a BMCL 01 of 2.12 (mg/m 3 )–yr CDE ( χ 2 = 0.64; p value for fit = 0.9957) (Figure 5, Ta ble 19). (For comparison the BMCL 05 was 3.73 (mg/m 3 )– yr CDE.) Fitting the logistic model to the same data yielded a BMCL 01 of 1.73 (mg/m 3 )–yr CDE ( χ

2 = 2.71; p value for fit = 0.8446) (Table 19). The BMCL 01 from these data is about the same as the apparent NOAEL. In general, a BMC is prefe rred to a NOAEL because the BMC takes into account all the dose respons e 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 SilicaDetermination of Noncancer Chronic Refe rence 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 16Fraction Affected

Dose 

### Lo g 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 MLE 01 BMCL 01 p value for fit

Probit-log-dose 2.45 (mg/m 3 )–yr CDE 2.12 (mg/m 3 )–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 Af rican gold mine dust, Hnizdo and Sluis-Cremer (1993) relied on estimates for the years 1956-1 960 by Beadle (Beadle and Bradley, 1970; Beadle, 1971). The original da ta, obtained by Corner House La boratories for the South African Bureau of Mines, are partly presented by Bead le 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 SilicaDeterm

Appendix D3 ination of Noncancer Chronic Refe rence 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 si lica exposures of workers, whic h apparently depended on an unpublished analysis of the Corner House Laborat ories’ data done by Du Toit in 1991. Gibbs

and Du Toit state that the exact relationshi p between the observed particle counts and theoretically derived mass concentr ations cannot be determined, but that the uncertainties in this conversion do not appear to be severe for the dus t characteristics observed in the South African mines. They accept the estimates by Beadle and Br adley (1970) of the quartz percentages in the dust, i.e. 54% for incinerated and acid-washed du st 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 conten tion is supported by the footnote to Table II in Hnizdo and Sluis-Cremer (1993) where the re spirable 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% sili ca content appropriate for acid- washed dust, correspond more closely to those calculated by Hnizdo and Sluis-Cr emer (1993) (applying th e 30% quartz content to their reported “respirable dust concentrations,” i.e., the untreated dust), than to the modified, and higher, quartz exposures propos ed by Gibbs and Du Toit (2002). For example, 113 exposure samples were taken for stopers.DetermAination of Noncancer Chronic Reference Exposure LevelsFebruary 2005ppendix D3514SilicaOccupationShaftSinkersDevelopersStopersAssistantminers/TrammersShift BossesOtherOfficialsBanks/SkipsWorkersNear shaftsBoilermakersOtherArtisansMiscellan-eousPage-Shipp and Harris (1972) (Table III and Appendix I)Hours/shift(t)7.708.007.807.705.204.007.506.506.305.707.20Number ofsamples1037113157431063334416111RM x t4.441.961.571.200.870.771.310.561.000.641.01s.d.3.941.591.000.930.710.531.380.570.710.510.79RespirableMass (RM)0.580.250.200.160.170.190.170.090.160.110.14Silica (54%)0.310.130.110.080.090.100.090.050.090.060.08(afteracidtreatm t)enHnizdo and Sluis-Cremer (1993) (Table II)RM0.480.370.270.300.300.130.100.190.19Silica (30%)0.14*0.11*0.08*0.09*0.09*0.040.030.060.06*(beforeacidtreatm t)enGibbs and Du Toit (2002) (Table 4)RM0.480.370.270.300.300.130.100.190.19Silica (54%)0.260.200.150.160.160.07*0.05*0.10*0.10(after acid treatment)Table 20. Estimates of silica exposures in mg/m3for different occupational groups in South African gold mines.*denotes that value is equal to or closer to the value based on Page-Shipp and HarrisDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

The last line of Appendix I of Page-Shipp and Ha rris (1972) gives a mean value for stopers of 1.57 (mg/m 3 )-hours respirable dust mass af ter 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/m 3 . If 54% of this were quartz, the quartz level would be 0.11 mg/m 3 . Table II of Hnizdo and Sluis-Cremer (1993) lists 0.37 mg/m 3 respirable dust for stopers. Thirty % of 0.37 mg/m 3 equals 0.11 mg/m 3 , the same value reported by Page-S hipp and Harris. In Table 4 of Gibbs and Du Toit (2002) stopers are also reported to be exposed to 0.37 mg/m 3 respirable dust. If 54% were quartz, as Gibbs and Du Toit contend, the quartz level would be 0.2 mg/m 3 . 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 conten t of the dust as 15% for the late 1980s to early 1990s. Dr. Eva Hnizdo (persona l 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 su rveys indicate that th e amount of airborne respirable dust in SA gold mines in 1980's and in 1970's was on average around 0.4 mg/m 3 with average quartz concentration of 0.08 mg/m 3 ” (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 contradic tions in the various accounts, the silica concentrations in air proposed by Hnizdo a nd Sluis-Cremer, based on the Corner House Laboratory data, are a reasona ble 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 silic a concentrations in the various occupational settings. Since dust le vels in the mines were fairly constant for decades and quantification of silica was im proving, 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 corr ect silica content, despite an erroneous statement in a footnote to Table II of their paper 1

.

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 studi ed carefully all the re ports and came to the conclusion that our study was after all reasona bly correct. Based on the Churchyard study and

the measurements data I have seen in SA duri ng the 1990s, I am also convinced that our results are reasonable estimates of the exposure of th e cohort.” (Hnizdo, personal communication October 2004)

Appendix D3 515 SilicaDetermination of Noncancer Chronic Refe rence 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 Steen land and Brown (1995) found five cases of silicosis in the lowest dose group of 0 – 0.2 (mg/m 3 )-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 BMC 01 for silica of 0.43 (mg/m 3 )–yr using the six lowe st levels of silica. In risk assessment, the highest dose or doses are often dropped in or der to obtain an acceptabl e 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 n ear the benchmark should be most informative

for the benchmark concentra tion (USEPA, 1995; Filipsson et al ., 2003). Fitting the probit model to the log dose of the five lowest silica le vels from Steenland a nd Brown yielded a BMCL 01 of 0.34 (mg/m 3 )–yr CDE ( χ 2 = 1.32; p value for fit = 0.5177). [For comparison, BMCL 05 = 0.85 (mg/m 3 )–yr CDE.] Fitting the quantal quadratic model gave a BMCL 01 of 0.45 (mg/m 3 )–yr ( χ 2 = 3.36; p = 0.3395). Use of the BMC 01 value of 0.34 (mg/m 3 )–yr CDE from the log dose probit model resulted in a chronic REL estimate for crystalline silica of 4 μ g/m 3 . 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/m 3 )-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/m 3 -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 μ g/m 3 is calculated from the

Appendix D3 516 SilicaDetermination of Noncancer Chronic Refe rence 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 biological ly significant. OEHHA considers that the six cases may be work rela ted, 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 μ g/m 3

.

In a third supportiv e study, Chen et al . (2001) found two cases of silico sis in the lowest exposure group of ≤ 10 mg CTD/m 3 -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 li near at low doses. Fitting the probit model to the log dose of the four lowest data points yielded a BMC 01 of 0.132 (mg/m 3 ) - yr CDE ( χ 2 = 2.19; p value for fit = 0.335). Use of fi ve, six, or seven data points gave BMC 01 s 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 BMCL 01 of 0.093 (mg/m 3 ) - yr CDE ( χ 2 = 4.86; p value for fit = 0.0879). An inhalation chronic Reference Exposure Level for crystalline

silica of 6 μ g/m 3 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 BMCL 01 used with the Hnizdo and Sluis-Cremer (1993) da ta and with data from Steenland and Brown (1995) and Chen et al . (2001). The calculation therefore uses a 5% BMCL of 0.673 (mg/m 3 )-yr from the probit log dose model as the benchmark, which is reasonably well within the range of the reliably observed data and which doe s not differ too widely from the MLE 05 estimate of 0.955 (mg/m 3 )-yr for that parameter. This BMCL 05 is not strictly comparable to the BMCL 01 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 eq ual 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 pr edict that the “background” is substantially (as much as 5 – 10%) above zero, which is intrinsicall y implausible for silicosis unless there is an

unrecorded additional source of si lica 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. Altern atively, perhaps the assignment of an assumed zero exposure value to “non-dusty” job classificati ons 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, th e bin widths hinder the attempt to calculate a

BMDL 01 in this case. In relation to the model fit, other models (including the quantal linear model, which emphasizes the likely more reliable in cidences at higher dose levels) are consistent

with the BMCL 05 results from the log probit model used he re. 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

None of these issues can be re solved 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 μ g/m 3 (and thus a comparison REL of 4 μ g/m 3 ) indicates at least that the results of the earlier analysis are unlikely to have underest imated 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 plausibl e lower bound on the REL, which is consistent with our analysis

of Hnizdo and Sluis-Cremer’s data.

Other investigators have approach ed the possibility that some op acities on radiographs may be due to background influences such as age and smoki ng. In regard to smoking, Blanc and Gamsu (1988) reviewed the literature and conclude d that smoking would not interfere with thedetermination of silicosis by the ILO system. Based on reading 1422 films of unexposed bluecollar workers, Castellan et al . (1985) stated that th e 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 fo r a 1% risk for silicosis (BMCL 01 ) was 1.31 (mg/m 3 )-yr, which by their methods is equivalent to a continuous, 70-year exposure to 6.7 μ g/m 3 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 su pporting (Steenland and Brown, 1995; Hughes

et al ., 1998; Chen et al ., 2001) studies were of human adu lts, nearly all males, who were presumably healthy, at least initia lly, since they were able to work . Thus there is need to protect the sensitive members of the population, especial ly 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 (Gerhardss on and Ahlmark, 1985; Katsnelson et al ., 1986). The

selection of three as the intras pecies uncertainty factor (UF H ) 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 silic osis. 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 UF H from the default value of 10. Since these workers did not include children, the elderly, or female s (except for the 215 females in Chen et al .), some uncertainty related to inter-individual variability remains. Therefore, a UF H of 3 rather

than 1 is chosen.

(2) Mukherji et al . (1993) reported mean ambi ent silica levels (in PM 10 ) at three locations in the northern part of Santa Barbara County, Calif ornia (see the Appendix to this report). At Santa Maria (an urban site) the level was 2.3 μ g/m 3 ; in Santa Ynez (a rural site) 0.6

μ g/m 3 ; and in Buellton (a remote background site) 0.2 μ g/m 3 crystalline silica. Thus, use

Appendix D3 518 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

of a human intraspecies uncertainty factor (UF H ) of 10 with the data from the key study

would result in an estimated chronic REL of 0.9 μ g/m 3 (ACGIH method), a level in the range of ambient levels in Ca lifornia. Although the reported le vels at the urban site may (according to the authors) have reflected some anthropogenic contributions such as disturbance and tracking of s iliceous road dust, the rural and remote site values are apparently (perhaps conservatively) reflectiv e of the natural background to which all California residents are expos ed. (U.S. EPA (1996) found slightly higher average

ambient levels of silica in PM 10 ; this average may include some sites affected by disturbance and emissions.) Th ere is no evidence that these background levels of silica are causing silicosis. On the other hand, silico sis in the general population is not a target for medical attention, and au topsy rates are very low, so the possibility of a low frequency of response at these le vels cannot be entirely dismisse d. On balance, it appears plausible that a REL of 3 μ g/m 3 (benchmark + UF H = 3) would be protective of the general population.

(3) The dose-response curve for silicosis due to in halation of crystalline silica is steep, and an upward curvature of this dose response wa s seen in some studies (Figure 7-1 in USEPA, 1996). It is notable th at, whereas exposures in the 1-3 μ g/m 3 range are apparently without effect (based on the benchmark calcu lations and the California ambient background data), Rice and Stayner (1 995) described a LOA EL for silicosis of

8 μ g/m 3 in gold miners (Table 16; based on data from McDonald and Oakes [1984]). This finding may partly refl ect differences in physical state of the silica, and coexposures, but it might indicate that, although the chronic REL shoul d be protective of public health, chronic exposures only mode rately exceeding the REL may lead to clinically observable disease.

The animal studies gave L OAELs for silica of 0.2 mg/m 3 in dogs and from 1 to 4.9 mg/m 3 in rodents. After extrapolation to equivalent continuous time and app lication of LOAEL to NOAEL, interspecies, and intraspecies UFs, the estimated chronic RELs from animal data are all

less than 1 μ g/m 3 . This reflects in part the greater un certainty 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 ‘respirab le’ as particles 10 μ m or less MMAD. This reflects one usual type of sampler (for “PM 10 ”) used for ambient air sampling in the general environment. The other usual type of environmental sampler, PM 2.5 , collects even smaller particle s. There are differences in the size range distributio n between a typical PM 10 measuring device and the NIOSH type personal samplers, or other de vices with similar size selec tion properties, used by the investigators in the epidemiological studies. The NIOSH-type samplers capture 50% of particles

with a MMAD of 4 μ m, and higher percentages of smaller particles. A smaller proportion of larger particles between 4 and 10 μ m in aerodynamic diameter will al so be collected. Figure 5,

from Volume I of U.S. EPA’s Third External Revi ew Draft of Air Quality Criteria for Particulate Matter (April 2002), includes part icle penetration curves for PM 10 , PM 2.5 , and occupational samplers.

Appendix D3 519 SilicaDetermination of Noncancer Chronic Refe rence 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 oc cupational community calls respirable). PM 10 samplers are meant to capture particles that penetrate the entire length of the lower respiratory tree, including those that penetrate to the tracheobronchial and al veolar regions. Penetration (and therefore presumably deposition) by particle size is complex, and is dependent on the aerodynamic diameter, hygroscopicity, and electrostatic charge of th e 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 pene tration in the various re gions of the respiratory

tract is not linear with respect to size. Generally, though, la rger particles imp act higher in the respiratory tree (the extrathoracic and tracheobronchial regions), while smaller particles show greater penetration to the lower tracheobronchial and alveolar regi ons. 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 depositi on for particles about

3 μ m MMAD with deposition dropping above and belo w that. Their data and model indicate that tracheobronchial deposition rises rapidly above about 3 μ m MMAD. Available data also

Appendix D3 520 SilicaDetermination of Noncancer Chronic Refe rence 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 μ m in activity median thermodynamic diam eter (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 μ m AMTD particles and a deposition fraction of 0.024 in the alveolar region. Thus, many particles with a 10 μ m 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 de tail; measurements have been reported only in terms of NIOSH sampler results or PM 10 cutoff values.

It is generally assumed that the silicosis is induced by that fracti on of the silica th at reaches the alveoli. Nevertheless, no actual data exone rate the coarser particles in the 4 - 10 μ m 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 PM 2.5 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 silico sis-inducing activity of di fferent particle sizes, it is not possible to adjust the REL for different particle size distribution s, which might be found in the general environment, or for different measurement methods. The REL is therefore specified as applicable to concentrations of pa rticles having a size range (and reactivity) similar to those measured in the occupa tional studies [respira ble 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 proced ure, such as that used for

PM 10 , would overestimate the relevant exposure in a ny situation, and so would be inappropriate for precise risk quantification. However, PM 10 would be useful as a screening method to establish that a particular situation is unlikely to present a hazard. Fo r example, if the silica

concentration in PM 10 modeled at a receptor is less than the REL (3 μ g/m 3 ), occupationally respirable silica will also be less than 3 μ g/m 3 , so a facility would not pose a risk due to silica at that receptor. If the silica concentration in PM 2.5 modeled at a receptor is less than 3 μ g/m 3 but

PM 10 is greater than 3 μ g/m 3 , further testing would be needed. If both PM 2.5 and PM 10 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhala tion 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 radi ological analysis, and with adequate follow-up.

(2) The finding of a dose-response effect for silico sis in several of the studies (e.g., Hnizdo and Sluis-Cremer, 1993; Steen land 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 memb ers 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 sili cosis by radiogra phy alone (Hnizdo et al ., 1993), which results in higher, less health -protective chronic REL estimates.

(7) The possible underreporting of s ilicosis where complete radiographic data and autopsy data are not available (S teenland and Brown, 1995).

(8) The uncertainties in exposure estimation, es pecially when reconstructing historical levels of silica exposure (Seixas et al., 1997; Gibbs and Du Toit, 2002) including the variability in the estimates of percent quart z in the South African mine dust (Beadle, 1971; Hnizdo and Sluis-Cr emer, 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 differe nt studies at what were considered similar silica levels and similar exposure dura tion (see Summary Table 15 above).

(10) The variability in toxicity of various forms of silica (e.g., freshly fractured vs. aged quartz; cristobalite vs. qu artz) 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 depositi on 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 SilicaDetermination of Noncancer Chronic Refe rence 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 sensitiv e to silica than others. In addition, we used a benchmark of 1% adverse effect, rather th an the usual 5 %. Thus, use of the human intraspecies uncertainty factor (UF H ) 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 in clude 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 rang e 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 a ssessment methodology, in particular the UF H , for its adequacy in protecting infants a nd children. Since children have smaller airways than adults and breathe more air on a body weight basis, penetra tion 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 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

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Witschi HR, Last JA. 2001. Toxic responses of the respiratory system. In: Casarett and Doull’s Toxicology. The Basic Science of Poisons. 6 th Ed. Klaassen CD (ed). New York: McGraw-Hill.

World Health Organization (W HO). 1986. Recommended health-based limits in occupational exposure to selected mineral dusts (silica, coal). Geneva, Swit zerland: World Health Organization, Technical Report Series 734.

Appendix D3 532 SilicaDetermination of Noncancer Chronic Refe rence Exposure Levels February 2005

X. Appendix Particulate Levels of Interest for Exposure to Respirable Crystalline Silica Isomorphs

150 μ g/m 3 Federal 24 hour PM 10 standard

65 μ g/m 3 Federal 24 hour PM 2.5 standard

50 μ g/m 3 California 24 hour PM 10 standard

50 μ g/m 3 Federal PM 10 annual standard (chronic exposure)

50 μ g/m 3 8 hour TLV for quartz, cristobalite, and tridymite for workers (ACGIH Method)

50 μ g/m 3 estimated workplace LOAEL for silicosis from studies by Theriault et al . (“)

20 μ g/m 3 CA annual PM 10 standard (chronic exposure) (arithmetic mean)

15 μ g/m 3 Federal annual PM 2.5 standard (chronic exposure)

12 μ g/m 3 CA annual PM 2.5 standard (chronic exposure) (arithmetic mean)

12 μ g/m 3 current silica TLV adjusted to eq uivalent continuous exposure (50 μ g/m 3 x 8 h/24 h x 5 d/7d) (ACGIH)

10 μ g/m 3 TLV for silica proposed by Greaves (2000) (ACGIH)

8 μ g/m 3 current silica TLV further adjusted by 46/70 years occupational exposure ( ACGIH)

8 μ g/m 3 estimated high-end ambient crystalline silica level in US (USEPA, 1996) (PM 10 )

6.7 μ g/m 3 lower bound on 1% risk of silicosis estimated by USEPA (1996) (PM 10 )

5 μ g/m 3 TLV for silica proposed by Chen et al . (2001) (ACGIH)

5 μ g/m 3 “acceptable” ambient leve l for silica (10% of PM 10 ) (USEPA, 1996)

5 μ g/m 3 RfC for diesel exhaust particulate, a respirable PM (PM 2.5 )

3 μ g/m 3 estimated average ambient exposure to crystalline silica (USEPA, 1996) (PM 10 )

3 μ g/m 3 draft silica chronic REL proposed by OEHHA (ACGIH)

2.3 μ g/m 3 (1.17-3.46; n=12)* silica level during 1989 in Santa Maria, CA (urban site) (PM 10 )

0.6 μ g/m 3 (0-1.44; n=16)* silica level during 1989 in Sant a Ynez, CA (rural site) (“)

0.2 μ g/m 3 (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 SilicaDetermination of Noncancer Chronic Re ference 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 μ g/m 3 (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 (H SDB, 1999) Molecular formula C 8 H 8 Molecular weight 104.16 Boiling point 145.2 o C Melting point -31 o C (HSDB, 1999) Vapor pressure 10 torr at 31 o C, polymerizes at 82 o C and above (Weast, 1979) Solubility 310 μ g/ml (Dean, 1985) Conversion factor 4.26 μ g/m 3 per ppb at 25 o C

III. Major Uses and Sources

The major source of styrene is i ndustrial synthesis in which ethylbe nzene 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 but adiene-styrene rubbe r industry are butadiene, benzene, carbon disulfide, and trichloroethylene, whereas the ma in co-contaminants associ ated with the RPI are glass fibers and acetone (WHO, 1983). Environmental exposures to styrene may result from mainstream cigarette smoke (Newhook and Ca ldwell, 1993) and newl y 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 monitore d concentration of styrene wa s less than 0.1 ppb (CARB, 1999a). The annual statewide industrial emissions of styrene from faciliti es reporting under the

Appendix D3 534 StyreneDetermination of Noncancer Chronic Re ference 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 belo w) result in central nervous system (CNS) and peripheral nervous system effects, although th e latter are not as pr onounced (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 concentra tions 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 resu lts in reproductive effects, but the limited data are difficult to interpret because of the small sample number s (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 lipoph ilic, 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 styren e 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 styr ene 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 Bard odejova, 1970; Leibman, 1975; Guillemin and Bauer, 1979). Other pathways that may be presen t 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 urin ary MA+PGA are the co nsumption of ethanol (Berode et al. , 1986) and exposure to ethylbenzene (B ardodej and Bardodejova, 1970). An important consequence of ethanol related decr eased levels of urinary mandelic acid is the potential underestimation of exposure to st yrene (Guillemin and Bauer, 1979; Berode et al.

, 1986). However, the urinary metabolite levels re turn to control values 4-5 hours after the ethanol consumption (Berode et al. , 1986).

Indicators of human styrene exposure include ex haled styrene, blood st yrene, urinary MA, and urinary MA+PGA (Guillemin and Berode, 1988). Expos ure to styrene by inhalation results in 89 percent absorption (Guillemin and Berode, 1988). In the occupationa l studies that are the basis

Appendix D3 535 StyreneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

for quantifying the relationship between chronic styr ene exposure and health effects, end-of-shift or next-morning MA+PGA have been used. The ne xt-morning measurements are more reflective of past exposures due to the high fat solubility of styrene (fat:b lood partition coefficient = 94 (Csanady et al. , 1994)), the presence of a second, long biol ogical half-life fo r MA = 25 hours., and a long biological half-life for PGA = 11 hour s (Guillemin and Bauer, 1979). Following inhalation, the half-life for styr ene 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 ner vous 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 am ine control. Although long-term exposures to styrene ar e associated with decrements in physiological functions, the exact mechanism(s) for these effects have not be en 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 a nd degenerative disorders of th e myocardium and non-significantexcess 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 st udy of 59 male boat plant workers exposed to <1 to 144 (mean = 37.2) ppm styren e. Monoamine oxidase B (MAO-B) activity in platelets was measured as an indicator of catech olamine 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 plas tics industry (RPI) were studied for levels of

substances associated with ne uroendocrine 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 th e 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 re lationship (not detailed in the report) between urinary metabolites and styrene ai r concentration, the authors estimated that the average styrene TWA/8 hr was about 130 ppm. Controls consiste d of women factory worker s living in the same area as the styrene-exposed women, but not knowingly exposed to st yrene. After controlling for age and exposure time, the increased prolactin and thyroid stimulating hormone levels were

correlated with the concentra tion 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 wo rkers. Decreased manua l dexterity, increased reaction times, and/or abnormal ve stibuloocular 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 StyreneDetermination of Noncancer Chronic Re ference 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-re sponse data, and either unknown ethanol consumption or lack of adjustment for ethanol consumption. In the Cherry and Gautrin (1990) investigation, however, the au thors determined that accounting for ethanol consumption did not reduce the correlation be tween increased reaction time and exposure.

Decrements in other CNS functions were observe d 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, averag e air styrene = 24.3 ppm). The styrene-exposed population was compared to non-exposed worker cont rols and matched for ag e, 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 lo ss (dyschromatopsia) was also ob served in the styrene-exposed workers compared to controls. Chia et al. (1994) also observed decr ements 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-depe ndent 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. Eligib ility criteria included absence of metabolic, neurologic, or psychiatric disorders, limited ethanol intake, an d limited cigarette usage. All subjects were

instructed to avoid intake of alcohol and drugs for two days pr ior to testing. Styrene exposure was assessed from urinary MA+PGA levels the mo rning after the last workday in the week, followed immediately by participation in a batt ery of 8 neuropsychological tests designed to

measure CNS function. The tests included reacti on 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 colo red 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 qua ntal data. Expressed as co ntinuous data, styrene-exposed workers exhibited significantly poore r performances than controls in all tests, except in the digitsymbol test. Also, urinary meta bolite 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 crea tinine (mean = 75 mmoles/mole creatinine + 33 [SD], which is equivalent to a mean styrene c oncentration of 15 ppm) appeared to have no significant effects. The authors st ate that this level of urinary me tabolites corresponds to a mean daily 8-hour exposure to air styrene of 25 ppm (106 mg/m 3 ). Based on greater urinary excretion of styrene metabolites, significantl y poorer performances in four or more neuropsychological

tests were recorded in the ot her three subgroups (150-299, 300- 450, and > 450 mmoles MA + PGA/mole creatinine).

.

Appendix D3 537 StyreneDetermination of Noncancer Chronic Re ference 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 te st and validity calculations were performed by constructing 2 x 2 tables selecti ng different levels of urinary excretion of MA and PGA as a cutoff point. The highest values fo r chi-square and predictive validity were found when the cut-off

of 150 mmol/mol creatinine was c hosen, suggesting that the quan tal 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 Fi sher’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 as sumptions made concerning the continuous data, quantal data results suggest that the low dos e 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 preval ence (number of respondents for at least one, two or three abnormal tests), duration (years at work), and intensity (metabol ite level). This representation revealed a positive correlation of neuropsychological deficits with duration as well as intensity.

Table 1. Subjects Classified Positive on Neuropsyc hological Tests as a Function of Styrene Exposure a

.

MA+PGA, mmoles per mole

creatinine b Total Subjects Number of Abnormal Tests

> 1 > 2 > 3 c

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 s ubjects that performed abnormally in >3 tests based on their styrene urinary metabolite concentra tions, 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 exis ts 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 st yrene concentrations per group are 0, 15, 44, 74,

Appendix D3 538 StyreneDetermination of Noncancer Chronic Re ference 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 visu al dysfunction among styrene workers in the reinforced plasti cs industry. Workers (n=36) e xposed to an average of 16 ppm styrene exhibited significantly greater dyschromat opsia than controls, ma tched for age, ethanol consumption and tobacco smoking (Gobba and Cava lleri, 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 styr ene-exposed workers was tested, those exposed to > 50 ppm styrene exhibited greater dys chromatopsia than those exposed to ≤ 50 ppm, and within this group, a subset e xhibited a similar decrement afte r returning from a one month vacation. In the Campagna et al. (1995) study, the test for dyschromatopsia was given to 81 reinforced plastics industry work ers (79 male and 2 female) exposed to 4.6, 10.1, and 88.8 ppm styrene (first quartile, median, a nd 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 ner vous 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 o ccupational 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 fo r this relationship is the lack of knowledge about the appropriate surrogate fo r dose that leads to PNS disturbance (Murata et al. , 1991). In one study, however, chronic expos ure indices were developed which included work method, years at work, time spent lamina ting (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 inconsis tently 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 incr ease in sperm DNA-susceptibility to in situ denaturation as a function of mandelic acid concentration was indicated, but was within the interassay variability. No detrim ental 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 report ed 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 w ith 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 StyreneDetermination of Noncancer Chronic Re ference 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 ly mphocyte subsets and the preval ence of abnormal values of immunologic phenotypes for 11/18 subs ets were statistically differe nt 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 re lationships were observed fo r prevalences of abnormal responses for four lymphocyte subsets and, in th e 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 exposur e, 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-expos ed workers. However, the lack of exposure data prevents a quantitative assessment. In th e 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 ou t under the auspices of NTP (N TP, 1992), mice and rats were exposed by inhalation to styrene vapors to es tablish a maximum tolera ted 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 gr oup. Body weights among all exposed mice were lower than controls, and the difference was a bout 9 percent. Lung, olfactory epithelial, and

forestomach lesions were observed in females a nd males. In females, degeneration of the adrenal gland cortex was observed. An effect not discussed in the ch airperson’s report, but recorded in the original laboratory report, wa s an increased estrous cycle length among the female mice at all styrene doses. A LOAEL of 62.5 ppm is indicated by th e olfactory epithelial, forestomach and respiratory tract lesions in mi ce 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 styr ene in drinking water (0, 125, and 250 ppm). (The water solubility of styrene is 310 ppm.) The only effect wa s 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 pp m; 6-hr/d; gestation days 7-21). Although the small number of litters (n=2) at the 50 ppm dose prevented detaile d 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 increas es in spontaneous activity and in the delay of some neurobehavioral functions. Many of the e ffects became diminished as the pups aged. Measurements of reproductive toxicity (maternal we ight gain, length of gestation, number of live

Appendix D3 540 StyreneDetermination of Noncancer Chronic Re ference 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 resear ch 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 br ains of rat pups compared to one or both control groups on day 0 and day 21. No reproductive or hi stopathological changes

were seen.
Rosengren and Haglid (1989) investigated whethe r long term inhalation exposure (three months) to styrene (90 and 320 ppm) could induce long las ting astroglial alterations in Sprague Dawley rats, traceable four months after exposure ce ased. 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 a nd in the hippocampus. GFA is the structural protein of the astroglial filame nts. These filaments form after damage to the central nervous system from any cause. The aut hors concluded that exposure to styrene at moderate exposure levels induces regional, long las ting astroglial reactions that serve as an indicator of solvent induced brain damage.

Mice, exposed acutely (14 days ) by inhalation to 125 - 500 ppm st yrene, exhibited decreased spleen / body weight, splenic hypocellularity, altered lymphocyt e proportions among subsets, and increased proliferative re sponse to mitogens (Corsini et al. , 1994). Mice and rats, exposed by gavage to high levels of styr ene (18, 27, 45 mg/kg - mouse; 118, 177, 294 mg/kg - rat) for 5 days/week for 4 weeks, exhibi ted 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 Rive r 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 a nd 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 re maining survivors were necropsied after 104 weeks. Control and high-exposure rats received a complete hist opathologic examination, while target organs, gross lesions, and all masses were ex amined in the other 3 groups. Styrene had no effect on survival in males, but females e xposed to 500 or 1000 ppm had a dose-related increasein survival. Levels of styrene in the bl ood 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 proporti onal to styrene exposure concentr ation. The authors found no changes of toxicologic significance in hematol ogy, clinical chemistry, urinalysis, or organ weights. Styrene-related non-neoplastic histopat hologic changes were conf ined to the olfactory epithelium of the nasal mucosa. (The author s 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/da y, 5 days/week, over a period of up to 2 years (Huntingdon Life Scienc es, 1998). Ten mice per sex per group were

Appendix D3 541 StyreneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

necropsied after 52 and 78 weeks of exposure, an d the remaining survivors necropsied after 104

weeks. Due to increased mortalit y 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. Histopath ology revealed liver necrosis that was a likely

contributor to the deat hs. Reduced body weight gain and increased food consumption were observed in male mice at the two highest exposur e levels and in female mice at the highest exposure level. Both styrene monomer and styr ene oxide in blood increased with exposure concentration. No changes of toxicologic sign ificance in hematology, ophthalmology, clinical chemistry, urinalysis, or organ weights we re noted. Styrene-related non-neoplastic histopathologic changes were s een in the lungs (bronchiolar-a lveolar 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 fr om the turbinate was also seen more frequently as the study progressed. In addition, atrophy of the olfactory nerve fibers wa s present in mice at the three highest exposure concentrations.

VI. Derivation of the Chronic Ref erence 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

BMC 05 Not established 1.7 ppm Exposure continuity 8 hr/d (10 m 3 per 20 m 3 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/m 3 ; 900 μg/m 3 )

The most relevant chronic noncancer effect due to styrene exposure is neurotoxicity. The Mutti

et al . (1984b) occupational study pres ented 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 incl ude 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 subs tances, and a test to allow a match for general intelligence). The use of urinary metabolites to measure expos ure dose is based on th e observation that the

Appendix D3 542 StyreneDetermination of Noncancer Chronic Re ference 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 conver sion 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 gro uping of the number of subjects that performed abnormally in > 3 tests based on their urin ary metabolite concentrati ons was chosen for a BMC analysis (see Table 1). Basing the BMC on a bnormal 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 posi tive responses is reduced due to the zero background level of abnormal responses in the control group when the cr iteria 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

BMC 05 of 1.7 ppm. A BMC 05 is considered to be similar to a NOAEL in estimating a

concentration associated with a low level of ri sk. Following adjustment for exposure continuity (10 m 3 per 20 m 3 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/m 3 ) 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 da ta 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 pro cedure is increased and the low dose effects are more easily observed. The quantal data of Mutti et al. (1984b), i.e., the proportio n of subjects responding abnormally to the tests, therefore provide a mo re 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. Howe ver, OEHHA believes that a statistical method to correct for this, known as a Bonferroni correction, is unnece ssary. 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 cal culating 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 m 3 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/m 3 ).

The U.S. EPA (1996) calculated a reference concentration (RfC) of 0.3 ppm (1 mg/m 3 ), which is slightly higher than the OEHHA-deri ved chronic REL of 0.2 ppm (0.9 mg/m 3 ). The RfC for styrene is also based on the findings of Mutti et al . (1984b), but utilized the continuous data for

Appendix D3 543 StyreneDetermination of Noncancer Chronic Re ference 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 lowe st exposure group (<150 MA+PGA mmole/mole creatinine; equivalent to < 25 ppm styrene). Ho wever, 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 BMC 05 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 lo w dose subgroup. Thus, the quantal data indicate that a NOAEL was not established in this study.

With regard to application of uncertainty f actors, U.S. EPA (1996) applied a UF of 3 for intraspecies variability and a part ial UF of 3 for lack of inform ation 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). Th e calculation was based on the same Mutti et al. (1984b) worker study. ATSDR ( 1992) identified the lowest e xposure 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 applie d uncertainty factors (UFs) fo r the use of a LOAEL and for intraspecies variability. The MR L 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 we re rats, and a LOAEL of 20 ppm was identified based on lesions in various organs in both sexe s. 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 μ g/m 3 ). Besides the different toxic endpoints between the chroni c mouse exposure study a nd human occupational studies, the well designed human study of Mutti et al. (1984b) is preferable for REL development because it does not introduce the un certainties 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 μ g/m 3 ) for styrene.

Appendix D3 544 StyreneDetermination of Noncancer Chronic Re ference 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 BMC 05 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 st yrene via dermal exposure in the reinforced plastics industry has not been addressed and may overestimate th e 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 be ing 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 correlat ed with both age and exposure intensity. This finding suggests that age may also be a factor in poor neuropsycho logical 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 wo rkers (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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Appendix D3 550 StyreneDetermination of Noncancer Chronic Refe rence 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/m 3
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 H 2 SO 4 Molecular weight 98.1 g/mol Density 1.84 g/cm 3 @ 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 us ed 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 trio xide, and from oleum (a combination of sulfur

trioxide and sulfuric acid used industrially). The annual stat ewide industrial emissions from facilities reporting under the Air Toxics Hot S pots 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 et ching and erosion of the teeth after 4 months exposure to an average concentration of 0.23 mg/m 3 H 2 SO 4 (Gamble et al. , 1984). Dental erosion increased in a dose-dependent ma nner with longer duration of exposure.

Appendix D3 551 Sulfuric acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

A study of 33 storage battery plant workers exposed to H 2 SO 4 concentrations as high as 35 mg/m 3 showed a greater group mean decrease in FEV l across the time of their work shift compared to workers who were not e xposed 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 Ambien t Air Quality Standard (CAAQS) for sulfates (25 μ g/m 3 for 24 hours) to see if it ade quately 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; Lipp man et al., 1987; Schlesinger et al., 1990), even at relatively high con centrations, 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 healt hy subjects consistent with those observed in epidemiological studies. A wo rking 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, SO 4 = is the best surrogate metric for H +

.

A large number of epidemiologic studies have be en conducted showing that elevated levels of several air pollutants, including ac id aerosols, sulfur and nitrogen oxides, and particulate sulfates

are correlated with an increas ed prevalence of pulmonary disease (U.S. EPA, 1989; OEHHA 2000). Elevated sulfate levels (1.6 ppb or 6.6 μ g/m 3 ) have been associated with statistically significant decrements in FVC and FEV 1 in a cohort of Canadian children (Stern et al. , 1989). Further analysis of these data led Bate s and Sitzo (1989) to conclude that H 2 SO 4 was the most likely cause for the pulmonary chan ges 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. De lfino 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 an alysis of the 6-City study suggested associations between average H+ concentration and chronic br onchitic symptoms. The relative odds of bronchitic symptoms with the highest acid concentration (58 nmoles/m 3
H+)_ versus the lowest concentration (16 nmoles/m 3 ) 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 c onfirmed to be significantly associated with strongly acidic PM (OR= 1.66; 95% CI 1.11-2. 48). It was also found that FVC and FEV 1 were lower in locales with high partic le acidity (Raizenne et al., 1996). Gwynn et al. (2000) reported an association between both H+ and sulfate pa rticles and respiratory hospital admissions and mortality in Buffalo, NY. Acidic sulfates may act to increase the toxicity of particles by enhancing the availability of meta ls present in the particles to ge nerate 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 acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

studies and the proposed REL is discussed in th e section below on the pot ential 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/m 3 for 5-15 minutes. The odor, taste, and irritation threshold was 1 mg/m 3 . 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 hum an exposures to sulfur ic 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 (supersat urated anhydrous sulfuric acid w ith varying concentrations of free sulfur trioxide) are absorbed as salts of sulfate anion (SO 4 2- ), and are excreted as organic sulfates, neutral sulfur, or neutral sulfur compoun ds 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 H 2 SO 4 concentrations of 0, 0.38, 0.48, 2.43, and 4.79 mg/m 3 continuously for 78 weeks resulted in dose-dependent adverse histological changes in lung and bronchiolar epithelial a nd parenchymal tissue in addition to a dosedependent decrease in blood oxygenation (Alarie et al. , 1973). In the animals exposed to 0.38 mg/m 3 , significant bronchiolar epith elial hyperplasia was observed in 5 of 9 animals; thickening of the bronchiolar wall s was observed in 3 of 9 animals. A slight focal bronchial epithelial hyperplasia was present in 4 of the 9 an imals. One animal died after 4 weeks exposure to 0.38 mg/m 3 . Although signs of pulmonary edema and cardiac hypertrophy were found, the

cause of death was not determined.

Respiratory system effects of H 2 SO 4 exposure in monkeys (Alarie et al ., 1973)

H

2 SO 4 (g/m 3 ) 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 pi gs of each sex to 0, 0.08, or 0.1 mg/m 3

H 2 SO 4 continuously for 52 weeks. The group exposed to 0.1 mg/m 3 also received larger sized particulates than the 0.08 mg/m 3 group (2.78 μ m vs. 0.84 μ m, respectively). No exposure related

effects were observed in the animals exposed to 0.08 mg/m 3 , whereas exposure of 0.1 mg/m 3
resulted in decreased body weights in the female guinea pigs. No other hi stological changes in any organs were observed at the end of the 52-week study.

Appendix D3 553 Sulfuric acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

Rabbits (4 per group) were exposed to 250 μ g/m 3 H 2 SO 4 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

H 2 SO 4 (Gearhart and Schlesinger, 1986, 1988). Mu cociliary clearance wa s also impaired by

H 2 SO 4 exposure and did not improve 3 months af ter cessation of exposure. A decline in dynamic lung compliance was observed after 12 mont hs exposure. There was no evidence of inflammatory cell infiltration in th e lungs of the exposed animals.

In guinea pigs, significantly slower and irregular breathing pattern s were noted when the animals had inhaled albumin followed by 30 minute exposures to H 2 SO 4 at 1.91 mg/m 3 twice per week for 5 weeks (Kitabatake et al. , 1979). Similarly, when guinea pigs were exposed to 2.49 mg

H 2 SO 4 /m 3 for 4 hours/day, 6 days/week for 4 weeks, in vitro lung histamine release was significantly enhanced following he terogeneous 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/m 3 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 μ g/m 3 H 2 SO 4 for 21 hours/day over a 620 day period resulted in incr eased expiratory resistance, reduced carbon monoxide diffusing capacity, reduced total and re sidual 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 H 2 SO 4 at 10 mg/m 3 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/m 3 in combination with a carbon particle mixture (1.5 mg/m 3 ) for 3 hours/day, 5 days/week for up to 20 weeks. The exposure resulted in significant alterations in specific antibody titer (decreased IgG, Ig 2a , IgM; increased IgG 2b ), depression of primary splenic antibody respons e, and decreased resi stance 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 leadi ng to maternal acidemia, H 2 SO 4 will be

neutralized in the maternal circulation and is unlikel y to reach the fetus.

Appendix D3 554 Sulfuric acidDetermination of Noncancer Chronic Refe rence 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 μ g/m 3 ) for 78 weeks

Critical effects Significantly increased bronchial epithelial hyperplasia and bronchial thickening

LOAEL 380 μ g/m 3

NOAEL Not observed Exposure continuity The exposure was continuous during the experiment. Exposure duration 78 weeks Average experimental exposure 380 μ g/m 3 for the LOAEL group Human equivalent concentration 380 μ g/m 3

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 μ g/m 3

The study by Alarie et al. (1973) identified a LOAEL for chr onic exposure to sulfuric acid of

380 μ g/m 3 . The principal uncertainties of this study are the small sample size of the test groups

and the absence of an observed NOAEL. A lowe r chronic LOAEL for bronchial reactivity is presented by Gearhart and Schlesin ger (1986, 1988) for rabbits (250 μ g/m 3 ). This study was not

selected as the basis of the REL because G earhart 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 effect s over the course of up to 12 months. The predominant weakness in the ra bbit study, however, was the ex treme discontinuity of the exposures (1 hour/day, 5 days/week), which w ould have necessitated use of a very large continuity adjustment. For these reasons, in addition to obvious phys iological 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/m 3 . 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 histolog ical changes in 100 guinea pigs exposed continuously for 1 year to 0.08 mg/m 3 was reported by Alarie et al. (1973). Guinea pigs respond to high

concentrations of sulfuric acid by occasional laryngeal spasms th at 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 eff ects 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 acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

For comparison, a chronic REL based on the guinea pig free-standing NOAEL of 0.08 mg/m 3 in animals exposed continuous ly for one year (Alarie et al ., 1973) would be 0.8 μ g/m 3

.

VII. Data Strengths and Limitations for Development of the REL

The major strength of the study on sulfuric acid is the use of health eff ects observations from continuous long-term exposures to a primate. The major weaknesses are the lack of adequate

human health effects data and th e 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 ri sk from long-term exposures because their bodies are growing, and their developmental processes, es pecially in the lung, may well be impacted by air pollution exposures. Elevated sulfate levels (1.6 ppb or 6.6 μ g/m 3 ) have been associated with statistically significant d ecrements in FVC and FEV1 in a cohort of Canadian children (Stern et al., 1989). The chronic REL for sulfuric acid of 1 μ g/m 3 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 μ g/m3 was the lowest level of ambien t sulfate measured. The mean daily morning to afternoon peak airflow change, the use of beta-agonist medicatio n, and the number of chest symptoms versus sulfate concentration in these children extrapolated linearly down to 1

μ g/m3. Thurston et al. (1997) also examined earlie r 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 μ g/m 3 , perhaps at about 2 μ g/m 3 . 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 μ g/m 3 appears to have a relatively low margin of safety with respect to the epidemiological st udies, but these observations are consistent with the proposed

REL of 1 μ g/m 3 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, a nd Kantz RJ. 1975. Long-term exposure to sulfur dioxide, sulfuric acid mist, fly ash, and th eir mixtures. Results of studies in monkeys and guinea pigs. Arch. Environ. Health 30(5):254-262.

Amdur MO, and Chen LC. 1989. Furnace-generate d acid aerosols: Speciation and pulmonary effects. Environ. Health Perspect. 79:147-150.

Appendix D3 556 Sulfuric acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

Amdur MO, Silverman L, and Drunker P. 1952. I nhalation of sulfuric acid mist by human subjects. Arch. Ind. Hyg. Occup. Med. 6:305-313.

Bates DV, and Sitzo R. 1989. The Ontario air po llution study: Identificat ion 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 ad missions to Ontario hospitals. Environ Res 65:172-194.

CARB. 1997. California Air Resources Board. Toxic Air Contaminant Identification List

Summaries.

CARB. 2000. California Air Res ources Board. California Emissi ons 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 Am dur MO. 1991. Sulfuric acid-layered ultrafine particles potentiate ozon e-induced airway injury. J. Toxico l. Environ. Health 34:337-352.

Damokosh AI, Spengler JD, Dockery DW, Ware JH, and Speizer FE. 1993. Effects of acidic particles on respiratory sympto ms 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 respirator y illnesses in Montreal, Quebec. Am J Respir

Crit Care Med 155:568-576.

Dockery DW, Cunningham J, Damo kosh AI, Neas LM, Spengler JD, Koutrakis P, Ware JH, and Speizer FE 1993. Health effects of acid aeroso ls on North American children: respiratory symptoms. Environ Hea lth Perspect 104:500-505.

El-Saddik YM, Osman HA, and El-Gazzar RM. 1972. Exposure to sulfuric acid in manufacturing of storage batterie s. J. Occup. Med. 14:224-226.

Fenters JD, Bradof JN, Aranyi C, Ketels K, Ehrlich R, and Gar dner DE. 1979. Health effects of long-term inhalation of sulfuric acid mist-carbon particle mixtur es. Environ. Res. 19:244-257.

Fujisawa T, Iguchi K, and Uchida Y. 1986. Eff ects of exposure to sulfuric acid mist on the induction of experimental asthma in guin ea pigs. Japan J. Allergol. 35(2):137-144.

Appendix D3 557 Sulfuric acidDetermination of Noncancer Chronic Refe rence 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. Sulfur ic 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 expos ure: effect of sulfuric acid mist on function and structure. Exp. Lung Res. 14:587-605.

Gwynn RC, Burnett RT, and Thurston GD. 2000. A time-series analysis of acidic particulate matter and daily mortality and morbidity in the Buffalo, New York region. Environ Health Perspect. 108:125-133.

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Micromed ex, Inc. (Edition expires 7/31/96).

IARC. 1992. International Agency for Res earch on Cancer. IARC Monographs on the Evaluation of Carcinogenic Risk s to Humans. Vol. 54. Occupa tional Exposure to Mists and Vapours from Strong Inorganic Acids and other Industrial Chemicals. Lyon, France: IARC, pp.

41-130

Iguchi K, Fujisawa T, and Uchida Y. 1986. Eff ects of exposure to sulfuric acid mist on the induction of experimental asthma in guin ea 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

mixture of the mist and sulfur dioxide. Mie Med. J. 29(1):29-37.

Lewis TR, Moorman WJ, Ludmann WF, and Campbe ll KI. 1973. Toxicity of long-term exposure

to oxides of sulfur. Arch. Environ. Health 26:16-21.

Lippman M, Gearhart JM, Schlesinger RB. 1987. Basis for a particle size-selective TLV for sulfuric acid aerosols. Appl Ind Hyg 2:188-199.

Lippmann M, Thurston GD. 1996. Su lfate concentrations as an indicator of ambient particulate matter air pollution for health risk evaluations. J. Expo. Anal. Envi ron. Epidemiol. 6(2):123-146.

OEHHA. 2000. Office of Environmental Health Hazard Assessment. Staff report. Adequacy of California Ambient Air Quality Standards: Senate Bill No. 25 - Children's Environmental Health

Protection.

Ostro B, Lipsett M, Wiener M, and Selner JC. 1 989. A panel study of the effect of acid aerosols on asthmatics. J. Air Waste Management Assoc. 89-94.1:1-15.

Appendix D3 558 Sulfuric acidDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

Raizenne M, Neas LM, Damokosh AI, Dockery DW , Spengler JD, Koutrakis P, Ware JH, and Speizer FE. 1996. Health effects of acid aero sols on North American children: pulmonary function. Environ Health Perspect 104: 506-514.

Schlesinger RB, Chen LC, Finkelstein I, and Ze likoff JT. 1990. Comparative potency of inhaled acidic sulfates: spec iation and the role of hydrogeni on. Environ Res 52:210-224.

Silbaugh SA, Mauderly JL, Macken CA. 1981. Effects of sulfuric acid and nitrogen dioxide on airway responsiveness of the guinea pig. J. Toxicol. Environ. Health 8(1-2):31-45.

Stern B, Jones L, Raizenne M, Burnett R, Meranger JC, Franklin CA. 1989. Respiratory health

effects associated with ambient sulfates and ozone in two rural Canadian communities. Environ. Res. 49(1):20-39.

Thurston GD, Lippmann M, Scott MB, Fine JM. 1997. Summertime haze air pollution and children with asthma. Am. J. Respir. Crit. Care Med. 155(2):654-660.

U.S. EPA. 1989. U.S.Environmen tal Protection Agency. An Acid Aerosols Issue Paper: Health Effects and Aerometrics. EPA/600/8-88/00SF. OHEA/ORD. Research Triangle Park, NC: U.S.

EPA.

Utell MJ, Morrow PE, Speers DM, Darling J, Hyde RW. 1983. Airway responses to sulfate and

sulfuric acid aerosols in asthmatics: an e xposure-response relationship. Am Rev Resp Dis

128:444-450.

Appendix D3 559 Sulfuric acidDetermination of Noncancer Chronic Re ference 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/m 3 (70 ppb)

Critical effect(s) Neurotoxic effects (decr eased 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 C 7 H 8 Molecular weight 92.13 g/mol Density 0.8661 g/cm 3 @ 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/m 3 @ 25 ° C

III. Major Uses or Sources

Toluene occurs naturally as a component of crud e 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 thinne rs, lacquers, rust inhibitor,

adhesives and solvent based cleani ng agents. Toluene is also utilized in printing operations, leather tanning and chemical processes. Benzen e 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 toluen e was approximately 2.2 ppb. For 1998, annual

statewide industrial emissions of toluene from faci lities reporting under the requirements of the Air Toxics Hot Spots Act in California were es timated to be 5,176,626 pounds (CARB, 1999). Note that this estimate is for stationary sour ces, and does not include emissions from mobile

sources.

Appendix D3 560 TolueneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

IV. Effects of Human Exposures

Neurological Effects Most studies reporting adverse effects due to ch ronic toluene exposures involve either toluenecontaining solvent abuse or occupational exposure to toluene. Solvent abusers are generally exposed to higher levels of toluene than are work ers. A continuum of neurotoxic effects ranging from frank brain damage to degraded performa nce on psychometric tests which roughly track exposure levels has been observed.

Solvent abusers

Chronic toluene abuse has been shown to cause permanent changes in br ain structure (loss of grey and white matter differentiation; cerebral, cerebellar and brainstem atrophy) which correlated with brain dysfuncti on 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 Toxi c 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 indi cated the occurrence of diffuse cerebral, cerebellar, and brainstem atr ophy and loss of differentia tion between the gray and white matter throughout the CN S. 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 sugge sted that BAERs may de tect early CNS injury from toluene inhalation, even at a time wh en neurological examination and MRI scans are

normal.

Two subjects of a group of 22 hospitalized solven t 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 us ing MRI and neuropsychological evaluations. The neuropsychologi cal testing indicated that thr ee patients functioned normally, three were in a borderline range, and eight were impaired. Independent analyses of white matter changes on MRI demonstrated that the degr ee 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 wh ich protons stop rotating in phase with each other because of the local magnetic fiel ds of adjacent nuclei; OTA, 1984) , and five of six demonstrated

Appendix D3 561 TolueneDetermination of Noncancer Chronic Re ference 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 toluen e abuse. Additionally, no definite clinical eviden ce of damage to the basal ganglia and thalam i was found despite the MR imagin g finding of T2 hypointensity..

Ungar et al . (1994) developed a physical bilayered model of dipa lmitoylphosphatidylcholine (DPPC) and toluene, and subjected DPPC control and toluene-mixed bilayers to MRI. T1 (T1: “Spin-lattice” relaxation time; a time constant th at reflects the rate at which excited protons exchange energy with the surr ounding 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 lip id bilayers substantially shortene d 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 (es timated as a time-weighted average) for an average duration of 6.8 years repor ted a significantly greater incidence of sore throat, dizziness and headache than controls; the sore throat a nd headache incidence demonstrated a rough doseresponse (Yin et al. , 1987).

Orbaek and Nise (1989) examined the neurological effects of tolu ene on 30 rotogravure printers, 33-61 years of age (mean 50), employed at two Sw edish printing shops for 4-43 years (median 29) in 1985. Mean exposure levels at th e two printing shops were 43 and 157 mg/m 3 of toluene, respectively; however, before 1980 the exposure levels had exceeded 300 mg/m 3 in both shops. The authors noted that rotogravur e 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 worker s and referents was also determined (< 200 g/week or > 200 g/week). Neurological functi on in the workers and referents was evaluated using interviews and psychometric testing; the re sults from each of the two printing shops were pooled. The printers reported st atistically significantly higher occurrences of fatigue (60%), recent short-term memory problems (60%), con centration difficulties (40%), mood lability (27%), and other neurasthenic symptoms. The pr inters also scored significantly worse than referents in a number of psychometric tests, incl uding synonym, Benton revised visual retention and digit symbol tests, even afte r 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 te sts was performed in 30 female workers exposed to toluene vapors in an electroni c 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 av erage (TWA) of 88 ppm (330 mg/m 3 ) toluene while the

Appendix D3 562 TolueneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

control workers inhaled 13 ppm (49 mg/m 3 ). 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 chem icals were not addressed. In this study, 88 ppm was considered a LOAEL for central nervou s 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 re sulted in an underestima tion 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 work ers of normal hearing ability was selected from a group of 300 workers who we re apparently in go od health but were professionally exposed to toluene (12 – 14 years exposure, 97 ppm average exposure, exposure assessment not described). They were subjec ted 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 expos ed to solvents. A st atistically significant

alteration in the BAER results was noted in the toluene-exposed workers with both 11 and 90 stimuli repetitions. The author s 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. Furt hermore, 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 fo r 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 incr ease in the latencies of all the BEAP waves examined, except for P2 waves, as well as in th e 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 exposur e. The amplitude but not the latency of all the VEPs examined decreased signifi cantly with the length of exposure.

The effects of acute and chronic toluene exposur e 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 colo r vision acuity of the workers before and after an acute toluene exposure (28 – 41 minutes in duration, concentration 1115 – 1358 mg/m 3 ) was evaluated using the Farnsworth panel D-15 test, the Lanthony desa turated panel D-15 test, and the Standard Pseudoisochromatic Plates part 2. A control group of 8 unexposed workers was

also tested. Acute toluene e xposure had no effect on color visi on. 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 indi cates 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 TolueneDetermination of Noncancer Chronic Re ference 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 conc entration = 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 enlarg ement in 32 of 106 (30.2%) painters employed in an aircraft factory compared to 7% in a control group. However, th ere 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 signifi cant correlation between toluene exposure and the levels of serum enzy mes (serum aspartate am inotransferase (ASAT), alanine aminotransferase (ALAT), γ -glutamyltransferase (GGT)) considered to be indicators of hepatic damage. In another cross-sectional st udy of 289 printing workers exposed to less than 200 ppm for 8 hours/day, 8 worker s 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/gr oup) 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 wa s limited because there were no other neurohistopathological examinati ons or organ weight measurements conducted on the animals.

Forkman et al. (1991) studied the potential neurotoxicity of toluene inhalation exposure (3700 mg/m 3 (1000 ppm), 21 hours/day, 5 days/week for 4 w eeks) in male Sprague-Dawley rats. The

rats were either trained in beha vior meant to be performance test ed 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 TolueneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

activity. All tests were performed from 11 to 35 da ys after the end of the exposure. Exposure of trained rats to toluene resulted in a signifi cantly different overall test performance when compared to controls. Rats trained after toluen e exposure also had test performances different from controls, but the difference was not statistically significant.

Rats exposed to toluene concentr ations of 1000 ppm or 100 ppm, 6 h/day, 5 days/week, for 3 or 6 months, respectively, demonstrated statisti cally 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 performa nce test and decreases in spontaneous motor activity (approximately 62% of c ontrol 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 spat ial learning and memory, dopamine-mediated locomotor activity and dopamine D 2 agonist binding in rats. Spa tial learning (postexposure days 3-6) and memory (postexposure day 14) were tested using a water maze. Spontaneous and apomorphine-induced locomotor activity was evalua ted on postexposure day 17. Effects on binding parameters of the dopami ne D2 agonist S(-)[N-propyl- 3 H(N)]-propylnorapomorphine ([H]NPA) were determined using membrane prepar ations of the neostriatum of the rat brain. Toluene exposure caused a statistically significan t 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 B

B max and K D values for [ H]NPA binding. These results

indicate that subchronic toluene ex posure 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, resu lted in decreased brai n wet weights of the caudate-putamen (trend test fo r dose-response significant at p < 0.05) and subcortical limbic areas (trend test for dos e-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 recepto r activity (trend test for dose-response) as indicated by decreased IC 50 (inhibition constant) (signi ficantly less than controls ( p < 0.05) at 80 ppm), K H

(inhibition constant for high-affinity receptor sites), K L

(inhibition constant for lowaffinity receptor sites), and R H % (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 wei ghts of the whole brain, serum prolactin levels,

the K D (disassociation constant) or the B

B 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. Exposu re to xylene or styrene (80 and 40 ppm, respectively; 4 week s, 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), l eads 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 TolueneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

apomorphine-induced locomotor activity seen after toluene exposure by von Euler et al. (1993) may be related to the increased D 2 agonist activity described above ( IC 50 , K H , K L 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/m 3 ) toluene showed no treatment-related effects on histopathology of major organs, includi ng the nasal turbinates (CIIT, 1980). In this study, the nasal histopathology examination sampli ng 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/m 3 ) toluene 6.5 hours/day, 5 days/week for 103 weeks were examined for hematological and histopathological effects in addi tion to gross observations of toxicity (NTP, 1990). Significant erosion of the olfactory epithe lium was observed in male rats while degeneration of the respiratory and nasal epithelium was ob served in both sexes at 600 ppm.

Mice were exposed chronically to 0, 120, 600, or 1200 ppm (0, 452, 2261, or 4523 mg/m 3 ) 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 anim als with hyperplasia of the bronchial epithelium in the 1200 ppm exposure group.

Reproductive and Developmental Toxicity Reproductive toxicity to maternal rats was obs erved 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 sa me 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 ge station 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/m 3 ) 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 we re then exposed to the same concentrations during days 1-20 of gestati on and days 5-20 of lactati on. After weaning, the F 1 pups were exposed 80 times to the appropriate exposure leve l and then randomly mated to members of the same exposure group. The F 1 generation showed significantly decreased body weight which persisted throughout lactation. No effects were observed on histopathology. No data were presented for the F 2 generation.

Da Silva et al. (1990) exposed rats and hamsters to 0 or 800 mg/m 3 toluene for 6 hours/day on gestation days 14-20 (rats), or days 6-11 (hamster s). Exposed rats demonstrated a significant

Appendix D3 566 TolueneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

exposure-related decrease in birth weight compar ed 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 neurob ehavioral 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 neur obehavioral effects in the offspring were investigated using a test batte ry 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 activ ity 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 plat form relocation) at the age of 3.5 months indicating impaired cognitive function. The diff erence 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 TolueneDetermination of Noncancer Chronic Re ference 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 λ

a = λ h 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/m 3 ; 300 μg/m 3 )

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 m 3 /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/m 3 ; 400 μ g/m 3 )

The critical animal study (Hillefors-Berglund et al ., 1995) used to derive an REL for toluene describes adverse neurological effects in rats af ter a well characterized in halation exposure to toluene. The study results contain both a LOAE L and a NOAEL. Decreased brain (subcortical limbic area) weight and altered dopamine receptor binding compared to cont rols were noted at the NOAEL, but the changes were not statistically significant; this suggests that if a threshold for

Appendix D3 568 TolueneDetermination of Noncancer Chronic Re ference 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 dopam ine receptor binding ag rees qualitatively w ith results from similar studies (von Euler et al. , 1994). Additionally, toluene-induced neurotoxicity has been described in many studies by a variety of endpoint s in both animals and humans (ATSDR, 1999).

The adverse neurotoxic effects a ssociated with toluene exposure in the rat study by HilleforsBerglund et al . (1995), decreased brain (subc ortical limbic area) weig ht and altered dopamine receptor binding, occur in areas of the rat brain that are structurally an d 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 al so protective for other adverse endpoints, such as respiratory tract dama ge and teratogenicity.

Table 1: Reference Exposure Levels (R ELs) from Selected Neurotoxicity Studies

Study Duration Effect LOAEL LOAEL NOAEL NOAEL total REL REL (ppm) (ppm) (ppm) (ppm) UF a (ppb) (μg/m 3 ) (TWA) (TWA)

VonEuler et al . (1988) 4 weeks rat: altered brain dopamine receptor binding 80 14.3 1000 14 54

Orbaek and

Nise b (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 Leve l; 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 associat ed 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 TolueneDetermination of Noncancer Chronic Re ference 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 [subcorti cal 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 cont rast, the psychometric tests used to generate the neurotoxicity data in the human occupational exposure studie s described above tend to be less sensitive and suffer from greater measurem ent uncertainty. Additionally, the HilleforsBerglund et al . (1995) study has better exposure characte rization than the human occupational exposure studies. Nonetheless, the human studies are useful in supporting the derivation of the REL for toluene. Ordinarily, an interspecies unce rtainty 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 a ppropriate duration correc tions. Based on comparison of the data in both animals and humans, it a ppears that a REL of 271 μg/m 3 (rounded to 300 μg/m 3 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 us e of an animal study with accurate exposure

characterization and both LOAEL and NOAEL observ ations 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

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Hillefors-Berglund M, Liu Y, and von Euler G. 1995. Persistent, specific and dose-dependent effects of toluene exposure on dopamine D2 a gonist binding in the rat caudate-putamen. Toxicology 100:185-94.

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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 lear ning 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 TolueneDetermination of Noncancer Chronic Refe rence 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-meth ylbenzene; 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/m 3 (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 C 9 H 6 N 2 O 2 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/m 3 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 manufactur e 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 fo am (Howard, 1989) and coatings. The annual

statewide industrial emissions from facilities re porting 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-diis ocyanate, and 754 pounds of toluene-2,6- diisocyanate (CARB, 1999).

Appendix D3 574 Toluene diisocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

IV. Effects of Human Exposures

Diem et al. (1982) conducted a prosp ective study beginning in 19 73 of 277 male workers involved in the production of TDI. The st udy 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 m onths prior to TDI production in the plant to provide baseline pulmonary function measuremen ts. 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 le vel 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 thos e individuals who showed a normal FEV 1 to height ratio. Data were analyzed by the maximum lik elihood weighted regression approach (Diem and Liukkonen, 1988). Both FEV 1 and forced expiratory flow (25-75%) [FEF (25-75%)] among workers who never smoked were found to be signi ficantly 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 dem onstrated a significant difference in FEV 1 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 signif icant decline in FEV 1

.

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 pers onal 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 hi gh-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 asth matics) were used in evaluating the data (Diem and Liukkonen, 1988). No relationship between TDI exposure an d change in lung function was observed, although the prevalence of chronic bronchitis was significantly associated with exposure.

A longitudinal study of 780 worker s exposed to TDI in the product ion of polyurethane foam was also conducted (Bugler et al. , 1991; unpublished). Exposure le vels were established using continuous-tape personal monitori ng devices. The mean exposure level was 1.2 ± 1.1 (SD) ppb TDI among 521 workers and 0.3 ± 0.18 ppb TDI in th e 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 FEV 1 were found after regression analysis, although

FEV 1 decline was high among the control group. E xposure levels among the different groups were close, limiting the power of the study to detect changes. Appr oximately 3% of the 780 workers showed signs of TDI sens itization and, of these, over 80% were in the group exposed to 1.2 ppb.

Appendix D3 575 Toluene diisocyanateDetermination of Noncancer Chronic Refe rence 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 FEV 1 shifted in the

same direction for all data sets and the smoker/non-smoker slope difference became less meaningful with the data set combination (Ha sselblad, 1993).

Another toxicological area of con cern with exposure to TDI is th e development of sensitization, resulting in a well-documented condition known as “isocyanate asth ma” of either immediate or delayed-type onset (Moscato et al. , 1991). The level of exposure re quired to either develop or trigger a sensitization reaction is not well documen ted, however. Weaknesses of studies showing pulmonary effects of TDI exposure include use of area sampling vs. breathing-zone measurement of exposure, poor statement of cr iteria 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); decrea sed 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. Le sions 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 (s hort protocol) or to 0.02 ppm TD I (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 le vels above 0.12 ppm TD I 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 s econd 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 hypersen sitivity (measured on days 37 and 38) and all developed TDI-specific IgE antibodies, wherea s none of the animals in the 0.02 ppm TDI group showed either of these effects. Histopa thological effects in th e 1.40 ppm TDI group included interstitial inflammation, pleural thic kening, 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 report ed by Owen, 1984). Among female rats at

both dose levels and male rats at the high dose level, histopathologi cal effects observed included necrotic rhinitis, metaplasia, and inflammation of the respiratory epithelium. Female animals

Appendix D3 576 Toluene diisocyanateDetermination of Noncancer Chronic Refe rence 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 ev aluated in a two-generation stud y 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 week s, at which time the animals were randomly mated. Exposure of the females continued th rough gestation (excepting gestational day 20 through the fourth day postpartum) , and exposure of the males continued only until the delivery

of the F 1 generation. Weanlings in the F 1 generation were exposed in a manner similar to the parental (P 0 ) generation and bred afte r weaning to produce the F 2 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 P 0 generation animals included rhinitis of the epithelium in the two highest dose groups of both male and female animals. Hyperplasia of th e respiratory epithelium was also increased in the high dose groups of both sexes among P 0 animals. Among males in the F 1 generation, the incidence of rhinitis was significantly increase d at all exposure levels and the incidence of submucosal lymphoid infiltrates of the larynx and trachea was increased in the highest dose group. F 2 generation animals showed reduced pup wei ght and weight gain during the lactation period in the two highest dose groups.

Developmental toxicity of TD I was evaluated by exposing pre gnant 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 signi ficant fetal effect, a statistically significant increase in a specific skelet al 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/m 3 (1.9 ppb)

NOAEL 0.006 mg/m 3 (0.9 ppb) (non-smokers) Exposure continuity 8 h/day (10 m 3 /day occupational exposure), 5 d/wk Exposure duration 5 years Average occupational exposure 0.002 mg/m 3 for NOAEL group (0.006 x 10/20 x 5/7) Human equivalent concentration 0.002 mg/m 3 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/m 3 (0.07 μg/m 3 ; 0.01 ppb)

Appendix D3 577 Toluene diisocyanateDetermination of Noncancer Chronic Refe rence 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 foll ows. This study presented evidence of a decline in lung function, as indicated by decrements in FEV 1 , among workers involved in TDI production. Other factors suppor ting its quality include: (1) the absence of other confounding com pounds in the work environment, (2) the establishment of baseline lung f unction prior to exposure to TDI, (3) a “parallel internal comparison” of study groups for lung function, (4) an appropriate statistical an alysis which took into account interindividual variability, (5) breathing zone measurement of TDI (although co mmenced 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 TD I 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 potent ial 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 Res ources Board. California Emissi ons 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 comparativ e 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 diisocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

Hasselblad V. 1993. Analysis of studies of wo rkers exposed to toluen e diisocyanate. Report prepared for U.S. EPA, Envir onmental Criteria and Assessment O ffice, Research Triangle Park,

NC.

Howard PH. (ed) 1989. Handbook of environmen tal fate and exposure data for organic chemicals. Vol. I. Large producti on and priority pollutants. Chel sea, MI: Lewis Publishers, Inc.

HSDB. 1995. Hazardous Substances Data Bank. National Library of Medicine, Bethesda, MD (TOMES ® CD-ROM Version). Denver, CO: Micromed ex, 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. Re v. Respir. Dis. 148:671-677.

Karol MH. 1983. Concentration-dependent immunol ogic response to toluen e diisocyanate (TDI) following inhalation exposure. Toxicol. Appl. Pharmacol. 68:229-241.

Loeser E. 1983. Long-term toxicity and car cinogenicity studies with 2,4-/2,6-toluene diisocyanate (80/20) in rats a nd 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. Sens ory irritation by toluene diisocya nate 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 R un Research Center. Revised project report 50-

592.

Tyl RW, and Neeper-Bradley TL. 1989. Two-gene ration reproduction study of inhaled toluene diisocyanate in CD (Sprague-Dawley) rats. Unio n Carbide, Bushy Run Research Center. Project

No. 86-33-90704.

U.S. EPA. 1995. United States Environmenta l Protection Agency. Documentation of the reference concentration for chronic inhalation expo sure (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 diisocyanateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2000

Wong KL, Karol MH, and Alarie Y. 1985. Use of repeated CO 2 challenges to evaluate the pulmonary performance of guinea pigs exposed to toluene diisocyanate. J. Toxicol. Environ.

Health 15:137-148.

Appendix D3 580 Toluene diisocyanateDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

CHRONIC TOXICITY SUMMARY



# TRICHLOROETHYLENE

(trichloroethylene; 1,1-2-trichloroe thylene, 1,1-dichloro-2-chloroet hylene, acetylene trichloride, and ethylene trichloride ) )

CAS Registry Number: 79-01-6

I. Chronic Toxicity Summary

Inhalation reference exposure level 600 μg/m 3 (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 C 2 HCl 3 Molecular weight 131.4 Density 1.47 g/cm 3 @ 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/m 3 @ 25 ° C

III. Major Uses or Sources

Trichloroethylene was once used as an extractant in f ood processing and has been used as an anesthetic and analgesic fo r medical purposes (Waters et al. 1977). Currently, it is widely used as a solvent in the industrial degr easing of metals, with secondary solvent uses in adhesive paint and polyvinyl chloride production (U .S. EPA, 1985). Trichloroethyl ene 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 implemen ted in the manufacturing of pesticides and other chemicals (F eldman, 1979). In 1996, the latest year tabulated, the statewide

mean outdoor monitored concentration of tr ichloroethylene 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 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

IV. Effects of Human Exposure

An occupational study of trichloroethylene (T CE) vapor emissions in a pump room was conducted by Vandervort and Polnkoff (1973). Work ers 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/m 3 (32-78 ppm). Nineteen worker s (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. A bout half of the 19 exposed work ers reported that consumption

of small amounts of alcohol outsi de of work resulted in change s of skin color and severe intoxication. One worker of the 19 reported no a dverse 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 be fore 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 fr om 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 con centrations in their environments.

Nomiyama et al. (1977) examined 36 trichloroethylene workers, of which 9 males and 12 females were occupationally exposed to a consta nt concentration of trichloroethylene (TCE) and 18 males were exposed to variable concentratio ns (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 work ers. 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 hour s 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 symp toms 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 subj ects (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 tr ichlororethanol (TRI) and tric hloroacetic acid (TCA) were

determined. Trichloroethanol-blood levels we re 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 detect ed up to 12 days after the final

exposure.

Appendix D3 582 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

Okawa and Bodner (1973) studied the occupation al exposure of 24 electri cal plant workers to trichloroethylene (TCE). The plant worker grou p consisted of 22 males and 2 females ranging in age from 21-52 years old. Environmental sample s of TCE were collected over three days and yielded varying concentrations of TCE related to the task performe d in certain areas (duration of exposure unspecified). Spray boot h 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 av erage value of 39 ppm (6-82 ppm range) TCE. Although the workers wore respiratory protection during the washing procedure, the overall average of air borne TCE in this area was 48.3 ppm. In the testing area of the plant, researchers re port that the amounts of toluene and n-pr opyl 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 we re only at these elevated levels for 20-30 minutes a day. Researchers note that although othe r agents were used in the work area, TCE was the only chemical found in signi ficant 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% work ers experiencing nausea, 54.2% headache, 33.3% dizziness, 25.0% fatigue, 25% nose and throat ir ritation, 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 corre lated 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 expos ure 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 th e mouth. Liver function tests were abnormal. One of the two developed jaundice with hepato megaly. Case #3 developed a similar reaction after 5 weeks of expos ure to 216-912 mg/m 3 TCE (40-170 ppm) as did case #5 after two weeks of exposure to 370 mg/m 3 TCE (69 ppm). Case #4 involved a 39 year old man exposed to <50 mg/m 3 TCE (< 9.3 ppm) for three weeks who deve loped the characteristic rash, lesions and jaundice with slight hepatomegaly. Upon return ing to work over the next three weeks, he developed generalized erythroderm ia and facial oedema, hepatosplenomegaly and liver failure with septicemia from whic h he died 14 days later.

Stewart et al. (1974) studied the effects of subacute trichloroethylene (TCE) exposure in combination with alcohol consumption. Seven me n exposed to 200 ppm TCE ingested 1 quart of beer or 90 ml of 100-proof vodka a nd developed red blotches on thei r 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 flus hing six weeks after the last exposure.

Appendix D3 583 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

V. Effects of Animal Exposure

Kjellstrand et al. (1983) studied the effects of both inte rmittent 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 iden tical groups of airexposed controls. Liver weights increased in a non-linear fashion as the concentration level of TCE increased. All groups exhi bited statistically significant in creases in liver weights as compared to the controls. In both the 37 and th e 75 ppm groups, the increase in females was less than in males. No increase in spleen weight wa s 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 incr ease in kidney weight was seen in the male 75 ppm group and was more pronounced with increasing concen tration. Male mice in the 75 ppm group also showed st atistically significant increases in BuChE activity. In the 150 ppm group, male and female liver weight increases were statistic ally 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 incr eased 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 reha bilitation 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 slight ly 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 signifi cant increases in liver weight. Kidney weight

increased in the same manner as in the c ontinuous exposures. The authors noted that “extrapolation of the concentration-effect cu rve suggests that both li ver weight and BuChE

activities are influenced at still lower concentration.”

Briving et al. (1986) examined neurotoxicity as a resu lt 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 gr oups were used as controls. Two areas of the brain were specifically observed, the hippocampu s and the posterior part of the cerebellar vermis. These discrete brain areas were previ ously shown to be sensi tive towards chlorinated aliphatic solvents (Haglid et al. , 1981). Following exposure, gerbils were decapitated and

measurements were made of total free tissue am ino acids as well as high-affinity uptake and

release of 3 H-aminobutyric acid (GABA) and 14 C-glutamate. A significant increase in

Appendix D3 584 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

glutathione was seen in the hi ppocampus 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 hippo campus revealed no significant changes. The authors suggest that the stimulation of tran sport 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 exchange s. Also analyzed, were the cell cycle, bone marrow micronuclei in polychromatic erythroc ytes (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 signifi cant) in cell cycle. In addition, the 50 ppm group exhibited a statistically signi ficant decrease in cell cycle. Fo r all concentrations, there was an overall increase in the PCE percentage. The numb er of PCEs with micronuclei also rose with the increasing concentrations of 50 ppm and 500 pp m TCE (not statistically significant due to high control values). The research ers conclude that the resulting in crease of MN in exposed rats is indicative of aneuploidy induc tion as opposed to chromosomal breakage, and that the lack of chromosome aberrations corresponds to spindle effects such as an euploid induction. Concurrent

results of increased levels of leukocyte aneuploidy were also found by Konietzko et al. (1978) in degreasing workers occupati onally 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, gerb ils were maintained for 4 months in TCE-free

conditions in order to observe a ny restoration of neuronal function. Both of the exposed groups as well as the control gr oup consisted of six pairs of males and females. Brain samples were collected from the gerbils afte r the 4 month non-exposure period an d 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 gerb ils from each group, including the control, were used for neuropathological examination. Br ain tissue from 2 gerbils in the control group and the 320 ppm group were examined under the el ectron microscope. No difference was seen in the body and brain weights of the exposed gerb ils 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 cereb ral 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, th e 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 visu al cerebral cortex, but increased in the posterior part of the

Appendix D3 585 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

cerebellar hemisphere and the sensory-motor cort ex. However, only a slight decrease of S 100 protein was observed in the visual cerebral cortex of 320 ppm expo sed 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. Ther e was an increase in DNA in the posterior part of the cerebellar vermis in the exposed gerbil s, 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 mi croscopy 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 a nd 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 syna ptic vesicles in the granular layer. Also, the granular layer had decreased maximal nerve cell su rface area. Nerve cells were affected by the exposure as several types were reduced in size with fewer orga nelles and more lysosomes and myelin bodies. Many axons and dendrites had re duced numbers of microtubles, and there were

filament bundles observed that we re 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 durin g 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 evaluati ons, liver function tests, and re nal 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 rena l 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 P 450 activity in 9 rats exposed to 50 ppm trichloroethylene for 28 days, compared with 9 control rats.

Appendix D3 586 TrichloroethyleneDetermination of Noncancer Chronic Re ference 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/m 3 ) in the heavy assembly area

NOAEL Not observed Exposure continuity 8 hours a day (10 m 3 /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/m 3 ; 600 μg/m 3 )

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 th e time-weighted averages (TWAs) included a wide range of concentrations , the TWA of 32 ppm (170 mg/m 3 ) 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 c hosen as a LOAEL. Un certainty 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 correla ted with urinary excretion measurements.

This study was the best chronic account of the n on-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 irrita tion 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 p pb 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 interspe cies UF of 10, and an intraspe cies UF of 10 resulted in an estimated REL of 200 ppb for trichloroethylene.

Appendix D3 587 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

VII. Data Strengths and Limitations for Development of the REL

The strengths of the inhalation R EL for trichloroethylene include th e 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 observa tion of a NOAEL.

VIII. References

Bardodej Z, and Krivucova M. 1958. Determination of trichloroethyl alco hol (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 trichloroe thylene on the gerbil brain amino acids and glutathione. Neurotoxi cology 7(1):101-108.

CARB. 1999a. California Air Resources Board. T oxics Air Quality Data. Substance Chooser. Trichloroethylene. Available online at http:/www.arb.ca.gov/aqd/toxics.htm

CARB. 1999b. Air toxics emissions data colle cted 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 contaminati on 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 ch anges 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 tr ichloroethylene 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, excret ion and toxicology of trichloroethylene after inhalation: 1. Experimental exposure on rats. Arch. Toxicol 30:115-126.

Appendix D3 588 TrichloroethyleneDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

Kimmerle G, and Eben A. 1973. Metabolism, excret ion and toxicology of trichloroethylene after inhalation: 2. Experimental human e xposures. Arch. Toxicol. 30:127-138.

Kjellstrand P, Holmquist B, Alm P, Kanje M, Ro mare S, Jonsson I, Mansson L, and Bjerkemo M. 1983. Trichloroethylene: Furt her 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 genotoxic ity of trichloroethylene to rode nts. Mutat. Res. 322:87-96.

Konietzko H, Haberlandt W, Heilbronner H, Re ill G, and Weichardt H. 1978. Cytogenetische Untersuchungen an Trichlorathylen-Arb eitern. 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, Thirumoort hy T, and Goh C. Stevens-Johnson syndrome associated with occupational exposure to tric hloroethylene. Contact Dermatitis 10:270-276.

Stewart R, Hake C, and Peterson J. 1974. “D egreasers’ Flush;” Dermal response to trichloroethylene and ethanol. Ar ch. 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 evalua tion/toxicity determination. Dunham-Bush, Inc. report 72-34.

Waters E, Gerstner H, and Huff H. 1977. Trichlor oethylene. I. An overview. J. Toxicol. Environ. Health 2:271-307. [as cited by Fan, 1988]

Appendix D3 589 TrichloroethyleneDetermination of Noncancer Chronic Refe rence Exposure Levels September 2002

CHRONIC TOXICITY SUMMARY



# TRIETHYLAMINE

(diethylaminoethane; ethanamin e; N,N-diethylethanamine)

CAS Registry Number: 121-44-8

I. Chronic Toxicity Summary

Inhalation reference exposure level 200 μ g/m 3 (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 C 6 H 15 N Molecular weight 101.9 g/mol Density 0.726 g/cm 3 @ 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/m 3 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 at mospheric particulates. The annua l 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/m 3 (4.8 ppm) for 8 hours) resulted in reversible ocular effects that included corneal swelling a nd 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 TriethylamineDetermination of Noncancer Chronic Refe rence Exposure Levels September 2002

responses to TEA in volunteer su bjects, who were industrial work ers exposed to TEA during the course of their normal jobs (but had good general a nd ocular health). Four people were exposed for 4 hours to 3.0, 6.5, or 40.6 mg/m 3 triethylamine. Corneal thickness was measured by ultrasonography, clinical observati ons were recorded using ocular microscopy, and statistical analysis of the size, density and distribution of corneal endothelial ce lls was performed by automated analysis of photographs. Visual acuity and contrast sensitivity were evaluated using test charts. After exposure to 40.6 mg/m 3 there was a marked edema in the corneal epithelium, and subepithelial microcysts. Ho wever, corneal thickness increased only minimally. Vision was blurred in all subjects and visual acuity and contrast sensitivity d ecreased in three of the four. After exposure to 6.5 mg/m 3 two subjects experienced sympto ms (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/m 3 triethylamine for 4 hours.

A medical examination of 19 workers exposed at a polyurethane foam production plant to a time-weighted average c oncentration of 13 mg/m 3 (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 que stionnaire on self-reported symp toms of visual disturbances revealed repeated occurrences of temporary eye irritation and “foggy visi on.” Small quantities of dimethylethanolamine, toluene diisocyanate, and methylene diphenyl isocyanate were also present in the workplace atmosphere.

Jarvinen and Hyvärinen (1997) reported loss of vi sual acuity and contrast sensitivity in 41 foundry workers (core makers) exposed to TEA. C oncentrations of TEA we re reported to have a mean of 46 mg/m 3 and a maximum of 486 mg/m 3 , but were highly variable with numerous large excursions above a background of about 20 mg/m 3 during a two-hour period of continuous monitoring. It is therefore difficult to determin e 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/m 3 ) for 6 hours/day, 5 days/week. Groups of rats were necropsied at approximately 30, 60, and 120 da ys of exposure. The last corresponds to an elapsed time of 28 weeks. Endpoints examin ed included gross and histopathological examination of all major organs, including the lungs, nasal passage s, 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 he matological parameters were unchanged with exposure. However, all rats exposed to 247 ppm TEA manifested irrita tion. “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 fo r Virginia Chemicals (1987), necrotizing inflammation of the nasal cavity, metaplasia of the trachea, and thymic atrophy were observed

Appendix D3 591 TriethylamineDetermination of Noncancer Chronic Refe rence Exposure Levels September 2002

after exposure to 1000 ppm (4140 mg/m 3 ) 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 al l animals exhibited necrotizing inflammation in the nasal epithelium.

Rabbits (6-12 per group), exposed to 48 or 100 ppm (199 or 414 mg/m 3 ) triethylamine for 7 hours/day, 5 days/week, for 6 weeks, showed c oncentration-dependent pathology in the eyes, lungs, liver, kidney, and heart (Brieger and Hode s, 1951). The eyes showed multiple punctate erosions of the corneal epithelium, and cornea l edema at 48 ppm. Lung lesions included thickening of vascular walls; liver lesions in cluded 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 hist ologic effects among the exposed animals reported. All animals did surviv e 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 reproducti ve study in rats (10/sex/gro up) 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 d ecreased 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/m 3 ) 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 μ g/m 3 )

The U.S. EPA (1995) based its Reference Concentration (RfC) of 7 μ g/m 3 (2 ppb) for triethylamine on Lynch et al . (1990) but included a Modifying F actor (MF) of 10 for “database deficiencies” - “lack of developmental and repro ductive effects, and of appropriate data in a second species.” The criteria for use of modifyin g factors are not well specified by U.S. EPA.

Appendix D3 592 TriethylamineDetermination of Noncancer Chronic Refe rence Exposure Levels September 2002

Such modifying factors were not used by OEHH A. In addition OEHHA applied a subchronic

UF of 1 since 24 male and 24 female rats in th e 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 indicate s that the animals cl osed their eyes and buried their noses in their fur, likely to prevent the irritant e ffects 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 effect s in the eyes, lungs,

and livers of rabbits after six weeks of discon tinuous 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/m 3 TEA, which is equivalent to 4.5 mg/m 3 continuous exposure. (Other tasks were at 4-5 mg/m 3 TEA or 1.6 mg/m 3 continuous exposure.) Selectio n 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 μ g/m 3
based on human data. These workers experienced some short-term peak exposures to TEA and were also exposed to dimethylethanolamine (<0.1 mg/m 3 ), toluene diisocyanate, and methylene diphenyl isocyanate.

VII. Data Strengths and Limitations for Development of the REL

The major strengths of the triethylamine REL ar e the observation of a NOAEL in a controlled exposure experiment and finding of the same a dverse effect, eye irritation, in humans and animals. The major weaknesses are the minima l amount of adequate human health effects information, the lack of dose-re sponse 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 qu antify a differential effect of TEA in infants and children. However, it is a respiratory irritant an d thus has the potential to exacerbate asthma. In addition, other alkylamines are known to be asso ciated 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 Flor en 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. Toxi col. Appl. Pharmacol. 100:529-538.

Appendix D3 593 TriethylamineDetermination of Noncancer Chronic Refe rence 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 safe ty: 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, Bernstei n DI (eds.). 1999. Asthma in the Workplace.

2 nd ed. New York: Marcel Dekker, pp. 524-525.

Brieger H, and Hodes WA. 1951. To xic effects of exposure to vapo rs of aliphatic amines. AMA Arch. Ind. Hyg. Occup. Med. 3(3):287-291.

CARB. 2000. California Air Res ources Board. California Emissi ons 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. T oxicity 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 triethyl amine 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 triethylamin e. Occup. Environ. Med. 54(7):483-6.

Lynch DW, Moorman TR, Lewis P, Stober P, Ham lin R, and Scheuler RL. 1990. Subchronic inhalation of triethylamine vapor in Fischer-344 rats: Organ syst em toxicity. Toxicol. Ind. Health. 6(3/4):403-414.

Mosier AR, Andre CE, Viets FG Jr. 1973. Identification of aliphatic amin es volatilized from cattle f eedyard.

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 fo r chronic inhalation exposure (RfC) for triethylamine.

Appendix D3 594 TriethylamineDetermination of Noncancer Chronic Refe rence Exposure Levels September 2002

Virginia Chemicals. 1987. Pathologic findings in Fischer 344 rats expose d by inhalation to allylamine, ethylamine, diethylamine, and trieth ylamine with cover letter dated 042484. OTS #

308080. Doc # 86-870000813. Fiche # 0515251.

Appendix D3 595 TriethylamineDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

CHRONIC TOXICITY SUMMARY



# VINYL ACETATE

(1-acetoxyethylene; acetic acid, vinyl ester; ac etic 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/m 3 (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 C 4 H 6 O 2 Molecular weight 86.09 g/mol Density 0.932 g/cm 3 @ 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/m 3 @ 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 emis sions 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/m 3 ) VA. No adverse effects were noted follo wing chest x-ray, electrocardiogram, blood

Appendix D3 596 Vinyl acetateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

chemistry, and urinalysis. The control group (s ample size unspecified) co nsisted 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 ex tended period of time to 21.6 ppm (76 mg/m 3 ) VA. Upper respiratory irritati on was also experienced by a majority of 5 subjects. Odor was detected at 0.4 ppm (1.4 mg/m 3 ) 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/m 3 ) 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 majo r organs. There was no mortality resulting from these exposures. A close examination of th e 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 meta plasia/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 histio cyte accumulation in the lungs. Body weight gain of rats was significantly decreased in th e 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 ol factory 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-w ise test compared to control group

Mice also exhibited significant hi stological lesions in the respirat ory tract following exposure to 200 ppm VA or greater. The lesions include d atrophy of the olfactory epithelium and submucosal gland. At 600 pp m, hyperplasia of the trachea was observed, in addition to exfoliation/flattening of the bronchial epitheliu m and decreased body weight gain. Relative brain and kidney weights were increased in th e 600 ppm group at the en d of the study, and absolute liver, heart and kidney weights were al so 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/m 3 ) 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 pneum onitis was observed in the 1000 ppm treatment group. No adverse effects were seen in the 50 pp m treatment group. An identical study in rats was also conducted by Owen (1980b). In this stu dy, body weight gain was significantly reduced

Appendix D3 597 Vinyl acetateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

in male and female rats exposed to 1000 ppm VA. An increase in the incidence of mild histiocytic alveolitis was obs erved 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/m 3 ) VA for 6 hours/day on days 6-15 of gest ation. Significant maternal to xicity, as measured by reduced weight gain from day 10 thr ough 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 de fects 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 gest ation. 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 (espec ially delayed ossification) in fe tuses from dams exposed to 1000 ppm VA. These results indicated to the authors th at VA is not uniquely toxic to the conceptus. The NOAEL was greater than 5000 ppm via the dr inking water and 200 ppm by the inhalation

route.

Appendix D3 598 Vinyl acetateDetermination of Noncancer Chronic Refe rence 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/m 3 , 200 μ g/m 3 )

The chronic REL is the U.S. EP A 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/m 3 ). The duration-adjusted concentr ation for acetaldehyde was 16 mg/m 3

, whereas the NOAEL for histological lesions in rats by Appleman et al. (1982) was 48.75 mg/m 3
acetaldehyde. Therefore, the concentration of acetaldehyde was not cons idered 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 eq uivalent to 50 ppm continuous exposure during development. Multiplying by an RGDR of 1 and di viding 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 acet ate include the availabi lity of controlled exposure lifetime inhalation studies in multiple sp ecies at multiple exposure concentrations and with adequate histopathological analysis, and the obser vation of a NOAEL. The major area of uncertainty is the lack of adequate human exposure data.

Appendix D3 599 Vinyl acetateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

VIII. Potential for Differential Impacts on Children's Health

Since the chronic REL (0.05 ppm) is lowe r 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 litera ture 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. I nhalation toxicity of acetaldehyde in rats. I.

Acute and subacute studie s. 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, Be ems RB, Feron VJ, Cascier TC, Tyler TR, Vinegar MB, and Rickard RW. 1994. Chronic toxici ty and oncogenicity inhalation study with vinyl acetate in the rat and mouse. Fundam. Appl. Toxicol. 23:215-229.

CARB. 2000. California Air Res ources Board. California Emissi ons 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 inha lation 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: Microm edex, 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 acetat e: 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 mont h 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 acetateDetermination of Noncancer Chronic Refe rence Exposure Levels December 2001

Owen PE. 1988. Vinyl acetate: 104 week inhalation combin ed chronic toxicity and carcinogenicity study in the rat and mouse. Repor t 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 concen tration (RfC) for vinyl acetate.

Appendix D3 601 Vinyl acetateDetermination of Noncancer Chronic Re ference 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-di methylbenzene or 3-xylene), p-xylene (1,4- dimethylbenzene or 4-xylene), also noted as meth yltoluene, 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/m 3 (200 ppb) (for technical or mixed xylenes or sum of indivi dual 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 C 8 H 10 Molecular weight 106.16 g/mol Density 0.864 g/cm 3 @ 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 μ g/m 3

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 thr ee isomers, p-xylene is produced in the highest

Appendix D3 602 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

quantities in the U.S. for use in the synthesis of phthalic, isophtha lic, and terephthalic acid used in manufacture of plastics and polymer fibers in cluding mylar and dacron. In 1996, the latest year tabulated, the statewide mean outdoor mon itored 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 Califor nia 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-xyl ene - 34,573 pounds, and of m-xylene - 30,440 pounds. (Xylenes are also presen t 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 expo sures in which persons often inhaled a mixture of hydrocarbon solvents 8 hours per day, 5-6 da ys per week. These studies often have incomplete information on the airborne concentr ations 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 e xposure 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 e xposure in humans were identified.

Pharmacokinetic studies have documented the absorption of xylene in humans through inhalation, oral, and dermal rout es 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 circ ulation (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 exposur e contributed 1.8% of the total body burden.

Measurement of the rate of ab sorption through direct contact w ith the skin produced variable results ranging from 2 μg/cm 2 /min (Engstrom et al. , 1977) to 75-160 μg/cm 2 /min (Dutkiewicz and Tyras, 1968).

Xylene exposure has been associated with eff ects in a number of organ systems including the lungs, skin and eyes; neurological system; hear t 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, no se and throat (ATSDR, 1995). Chronic exposure to xylenes has been associ ated 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 na usea and headache; 200-500 ppm with dizziness, irritability, w eakness, vomiting, and slowed reac tion time; 800-10,000 ppm with

Appendix D3 603 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

lack of muscle coordination, giddiness, confusi on, ringing in the ears, and changes in sense of balance; and >10,000 ppm with loss of consci ousness (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; Savolai nen and Riihimaki 1981; Savolainen et al. , 1979; 1984; 1985).

Chronic exposure to xylenes (with other hydrocar bons) has been associated with cardiovascular and gastrointestinal effects. H eart 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 xylen e 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 b een identified in studies where exposure was to xylene only. Previous studies identifying hematological effect s included known or suspected exposure to benzene (ATSDR, 1995; ECETOC, 1986). One series of case reports iden tified lowered white cell counts in two women with chronic occupa tional 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/grou p) 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 se rious subjective or objective health responses, including neurological tests, c ognitive tests and cardiopulmonary function tests were observed. Odor was noted, but the intensit y decreased usually within the first hour of exposure. The

Appendix D3 604 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

authors concluded that p-xylene may have a weak ir ritating effect on the soft tissues starting at 100 ppm, but overall, the small sample size and hi gh 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 potentia l effects. A survey of 994 Ch inese 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 an alysis of xylene metabolites. The following day workers answered a questionnaire concerni ng 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 st udy and analysis based on completion of their

health examinations and on result s 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 dist ribution of alcohol consumption and cigarette usage. The xylene-exposed and unexposed grou ps 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, cr eatinine), and subjective symptoms (survey of symptoms occurring during work and in the previous three months).

Results of analysis of the diffusive samplers s howed 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 ex posed men than in the control population of men. Results of the survey of subjective symptoms found differe nces in symptoms occurring during work and during a similar analysis over the preceding three month period, appa rently 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 signi ficantly (p<0.01) elevated in either xyleneexposed men or women including: dimmed vision, unusual taste, di zziness, 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 floati ng sensation. Ten subjective symp toms 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 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

poor appetite, reduced grasping power, reduced musc le power in the extrem ities, and rough skin. Dose dependency appeared to exist for 3 subjecti ve symptoms noted during work: irritation in the eyes, sore throat, floating se nsation, and for one symptom occu rring in the last three months, poor appetite. V. Effects of Animal Exposure

A limited number of chronic toxici ty 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 identi fied. A spectrum of

adverse effects has been documented in shorter te rm studies which potentially could occur with chronic exposure. These studies are presented he re 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 gastrointe stinal system, the hepatic system, the renal system, the dermis, and the eye, and it has num erous neurological eff ects and developmental

effects.

Animal data are consistent with human data in documenting re spiratory effects from xylene exposure. Acute and subacute exposures in mice , rats, and guinea pigs have been associated with decreased metabolic capacity of the lungs; decrea sed 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 effect s 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 ve ry high (unspecified) levels of xylene and recorded ventricula r repolarization disturbances, atri al fibrillation, arrhythmias,

occasional cardiac arrest and cha nges 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 Toxi cology Program (NTP, 1986) and by Jenkins et al. (1970), as well as other shor ter term studies (Carpenter et al. , 1975; Wolfe, 1988), have not identified histopathological lesions of the heart.

Studies identifying adverse gastroin testinal effects, hematological effects, or musculoskeletal

effects in animals were not identified. Studi es reporting no hematological effects include Carpenter et al. (1975) (rats exposed to 810 ppm of mixe d xylenes for 10 weeks, 5 days/week, 6 hours/day and dogs exposed for 13 weeks to 810 ppm mixed xylenes, 5 da ys/week, 6 hours/day)

and Jenkins et al. (1970) (rats, guinea pigs and dogs expos ed 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 acu te exposure to high concentrations of xylene (2,000 ppm) or subacute exposure to lower concen trations (345-800 ppm) of mixed xylene or

individual isomers. These effects include increa sed cytochrome P-450 and b5 content, increased hepatic weight, increased liver to body weight ra tios, decreased hepatic glycogen, proliferation

Appendix D3 606 XylenesDetermination of Noncancer Chronic Re ference 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, do gs, and monkeys exposed to 50-2,000 ppm of xylenes. These effects incl ude 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 increas ed hyaline droplets in the kidney of Sprague-Dawley ra ts administered 150 mg/kg/day or ally of mixed xylenes. This response is consistent wi th early nephropathy.

Xylene has been found to affect the dermis and eyes of animals. Hine and Zuidema (1970) found skin erythema and edema, epidermal thicke ning, and eschar formation in response to xylene exposure. Direct instilla tion 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, a nd 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 Seppa lainen, 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 ava ilable for developmental effects is consistent with the hypothesis that xylene is fetotoxic and many of the feto toxic responses are secondary to maternal toxicity. However, the ATSDR also obs erved 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 th ese 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 incr eased 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 ai r concentrations of 12 ppm and above. The ATSDR has sugge sted that the Mirkova et al . (1983) study results may have been influenced by poor animal husbandry as indicat ed by the low conception rates and the high incidence of fetal hemorrhages seen in the co ntrols. 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 ai r or 200 ppm of xylene for 6 h/day on days 4-20 of gestation.

Appendix D3 607 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

Unlike Mirakova et al . (1983), there was no sign of maternal toxicity an d no decrease in fetal weights and no increase in soft -tissue or skeletal malformatio ns. 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 we re also noted upon te sting 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 gest ation. Maternal toxi city 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 a nd other agents, Ungvary (1985) ex posed CFY rats to either 140 ppm or 440 ppm of xylene on days 10-13 of gesta tion 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 ex posure study in the rat (CrL-CD (SD) BR strain). Rats were exposed 6 h/day during prema ting, mating, gestation an d 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 th is difference was significant only for the female fetuses. These depressed weights were, however, st ill significant on day 21 of lact ation. Other adverse effects (such as increased soft tissue and skeletal abnor malities, increased feta l resorptions) were not increased significantly at any of the test concentrations.

Ungvary et al . (1980a) tested by inhalation the individua l ortho, meta, and para isomers of xylene in the CFY rat. Pregnant rats were e xposed 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 onl y 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 pregna nt Sprague-Dawley rats by inhalati on 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 grow th 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 toxi city and no difference in the

number of live or dead fetuses. The mean birt h 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. Expos ed offspring showed a delay in the ontogeny of the air righting reflex and e xhibited impaired performance in behavioral tests for neuromotor

Appendix D3 608 XylenesDetermination of Noncancer Chronic Re ference 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 offspr ing exhibited impaired performances in the Morris water maze at 16, 28, and 55 weeks of ag e, 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 ex posure chamber for 6 h/day on days 6-12 of gestation. Shigeta et al. (1983) reported signific ant 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 dela yed 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 pe r day to 1096 ppm o-xylene. Th is exposure was a LOAEL for body weight gain in males and a NOAEL for he patic effects in male rats. Jenkins et al. (1970) exposed rats, guinea pigs, squirrel monkeys, and b eagle 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 nitr ogen, alanine aminotransferase, aspartate aminotransferase, alkaline phospha tase, and creatinine and liver function including alkaline phosphatase, tyrosine aminotransferase, and total lipid s. 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 administer ed 0, 250, or 500 mg/kg/day doses of mixed xylene in corn oil by gavage 5 days/week for 103 w eeks 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 XylenesDetermination of Noncancer Chronic Re ference 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 m 3 /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/m 3 ; 700 μg/m 3 ) for mixed xylenes or for total of individual

isomers

A number of issues are important in considering the uncertainty a ssociated with this REL. For ATSDR (1995) the animal and human toxicity data suggest that mixed xylenes and the different xylene isomers produce similar eff ects, although different isomers are not equal in potency for producing a given effect. Theref ore 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 coul d be different. The us e of a neurological endpoint for derivation of a REL is supported by th e large number of inha lation and oral studies associating neurological eff ects 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 pr incipally low incidence (<50%) of the effects.

A factor of 1 was used for subchronic uncertain ty. 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, ma ny workers would likely have been exposed for more than 8.4 years, the cut-off point for chroni c human exposure. Anothe r issue is the use of diffusive samplers in the Uchida et al. (1993) study. These samplers provide a time weighted

Appendix D3 610 XylenesDetermination of Noncancer Chronic Re ference Exposure Levels April 2000

average concentration of hydrocar bon and cannot indicate the maxi mum concentrations a worker is exposed to. It is unknown whether peak con centrations 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 esti mate 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 LOA EL UF of 3 for a mild effect, an interspecies UF of 3, and an intraspe cies 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 concentratio n, and the lack of

observation of a NOAE L in the key study.

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Appendix D3 619 Phosphine