United States
Environmental Protection
Agency
Office of Health and
Environmental Assessment
Washington DC 20460
EPA-600/8-84-004A
March 1984
External Review Draft
Research and Development
?,EPA
Health Assessment
Document for
Chloroform
Part 1 of 2
Review
Draft
(Do Not
Cite or Quote)
NOTICE
This document is a preliminary draft. It has not been formally
released by EPA and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its
technical accuracy and policy implications.
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ERRATA
The attached pages are to be substituted for the corresponding pages in
the Health Assessment Document for Chloroform (September 1985) in which certain
key individuals' names were inadvertently omitted. Most notably, that of one
of the principal authors, Or. Jean C. Parker, was not included.
In addition, several typographical errors have been corrected in the front
matter of the document.
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PREFACE
The Office of Health and Environmental Assessment has prepared this
health assessment to serve as a "source document" for EPA use. This health
assessment document was developed for use by the Office of Air Quality
Planning and Standards to support decision-making regarding possible
regulation of chloroform as a hazardous air pollutant. However, the scope of
this document has since been expanded to address multimedia aspects.
In the development of the assessment document, the scientific literature
has been inventoried, key studies have been evaluated and summary/conclusions
have been prepared in order to quantitatively identify the toxicity of
chloroform and related characteristics. Observed effect levels and other
measures of dose-response relationships are discussed, where appropriate, to
place the nature of the health responses in perspective with observed
environmental levels.
Any information regarding sources, emissions, ambient air concentrations,
and public exposure has been included only to give the reader a preliminary
indication of the potential presence of this substance in the ambient air.
While the available information is presented as accurately as possible, it is
acknowledged to be limited and dependent in many instances on assumption
rather than specific data. This information is not intended, nor should it be
used, to support any conclusions regarding risks to public health.
If a review of the health information indicates that the Agency should
consider regulatory action for this substance, a considerable effort will be
undertaken to obtain appropriate information regarding sources, emissions, and
ambient air concentrations. Such data will provide additional information for
drawing regulatory conclusions regarding the extent and significance of public
exposure to this substance.
TM
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TABLE OF CONTENTS
LIST OF TABLES x
LIST OF FIGURES xiv
AUTHORS, CONTRIBUTORS, AND REVIEWERS .... xv1
1. SUMMARY AND CONCLUSIONS 1-1
1.1. INTRODUCTION 1-1
1.2. PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYSIS 1-2
1.3. PHARMACOKINETICS 1-2
1.4. HEALTH EFFECTS OVERVIEW 1-5
1.4.1. Toxidty . . „ 1-5
1.4.2. Reproductive Effects . . 1-7
1.4.3. Mutagenicity 1-7
1.4.4. Carcinogenicity. 1-9
2. INTRODUCTION 2-1
3. BACKGROUND INFORMATION 3-1
3.1. INTRODUCTION 3-1
3.2. PHYSICAL AND CHEMICAL PROPERTIES ..... 3-2
3.3. SAMPLING AND ANALYSIS 3-4
3.3.1. Chloroform in Air 3-4
3.3.2. Chloroform in Water 3-5
3.3.3. Chloroform in Blood 3-6
3.3.4. - Chloroform in Urine 3-6
3.3.5. Chloroform in Tissue 3-6
3.4. EMISSIONS FROM PRODUCTION AND USE 3-7
3.4.1. Emissions from Production . . . 3-7
3.4.1.1. Direct Production 3-7
3.4.1.2. Indirect Production . 3-12
3.4.2. Emissions from Use. . 3-20
3.4.2.1. Emissions from Pharmaceutical
Manufacturing 3-21
3.4.2.2. Emissions from Fluorocarbon-22
Production 3-22
3.4.2.3. Emissions from Hypalon®
Manufacture 3-22
3.4.2.4. Chloroform Emissions from Grain
Fumigation 3-23
IV
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TABLE OF CONTENTS (continued)
3.4.2.5. Chloroform Losses from Loading and
Transportation 3-24
3.4.2.6. Miscellaneous Use Emissions 3-25
3.4.2.7. Summary of Chloroform Discharges
from Use 3-25
3.4.3. Summary 3-25
3.5. AMBIENT AIR CONCENTRATIONS 3-26
3.6. ATMOSPHERIC REACTIVITY .3-32
3.7. ECOLOGICAL EFFECTS/ENVIRONMENTAL PERSISTENCE 3-33
3.7.1. Ecological Effects 3-33
3.7.1.1. Terrestrial 3-33
3.7.1.2. Aquatic 3-34
3.7.2. Environmental Persistence 3-37
3.8. EXISTING CRITERIA, STANDARDS, AND GUIDELINES 3-40
3.8.1. Air 3-40
3.8.2. Water 3-42
3.8.3. Food 3-43
3.8.4. Drugs and Cosmetics 3-43
3.9. RELATIVE SOURCE CONTRIBUTIONS 3-43
3.10. REFERENCES FOR CHAPTER 3 3-44
4. DISPOSITION AND RELEVANT PHARMACOKINETICS 4-1
4.1. INTRODUCTION * 4-1
4.2. ABSORPTION 4-2
4.2.1. Dermal Absorption 4-2
4.2.2. Oral 4-3
4.2.3. Pulmonary Absorption 4-6
4.3. TISSUE DISTRIBUTION 4-12
4.4. EXCRETION 4-20
4.4.1. Pulmonary Excretion 4-20
4.4.2. Other Routes of Excretion 4-30
4.4.3. Adipose Tissue Storage 4-31
4.5. BIOTRANSFORMATION OF CHLOROFORM 4-32
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TABLE OF CONTENTS (continued)
4.5.1. Known Metabolites 4-32
4.5.2. Magnitude of Chloroform Metabolism. ....... .4-36
4.5.3. Enzymatic Pathways of Biotransformation 4-39
4.6. COVALENT BINDING TO CELLULAR MACROMOLECULES 4-45
4.6.1. Proteins and Lipids 4-45
4.6.1.1. Genetic Strain Difference 4-51
4.6.1.2. Sex Difference 4-53
4.6.1.3. Inter-species Difference 4-53
4.6.1.4. Age Difference 4-56
4.6.2. Nucleic Acids 4-56
4.6.3. Role of Phosgene . .4-58
4.6.4. Role of Glutathione 4-59
4.7. SUMMARY 4-61
4.8. REFERENCES FOR CHAPTER 4 4-65
5. TOXICITY. 5-1
5.1. EFFECTS OF ACUTE EXPOSURE TO CHLOROFORM. 5-1
5.1.1. Humans 5-1
5.1.1.1. Acute Inhalation Exposure in Humans. . . .5-1
5.1.1.2. Acute Oral Exposure in Humans 5-5
5.1.1.3. Acute Dermal and Ocular Exposure
in Humans 5-6
5.1.2. Experimental Animals 5-7
5.1.2.1. Acute Inhalation Exposure in Animals ... 5-7
5.1.2.2. Acute Oral Exposure in Animals ...... 5-8
5.1.2.3. Acute Dermal and Ocular Exposure
in Animals 5-11
5.1.2.4. Intraperitoneal and Subcutaneous
Administration in Animals 5-12
5.2. EFFECTS OF CHRONIC EXPOSURE TO CHLOROFORM 5-13
5.2.1. Humans 5-13
5.2.1.1. Chronic Inhalation Exposure in Humans . .5-13
5.2.1.2. Chronic Oral Exposure in Humans 5-15
5.2.2. Experimental Animals 5-16
5.2.2.1. Chronic Inhalation Exposure in Animals . 5-16
5.2.2.2. Chronic Oral Exposure in Animals 5-17
vi
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TABLE OF CONTENTS (continued)
5.3. INVESTIGATION OF TARGET ORGAN TOXICITY IN EXPERIMENTAL
ANIMALS 5-31
5.3.1. Hepatotoxicity 5-31
5.3.2. Nephrotoxiclty 5-39
5.4. FACTORS MODIFYING THE TOXICITY OF CHLOROFORM 5-49
5.4.1. Factors that Increase the Toxicity 5-50
5.4.2. Factors that Decrease the Toxicity 5-58
5.5. SUMMARY: CORRELATION OF EXPOSURE AND EFFECT 5-60
5.5.1. Effects of Acute Inhalation Exposure 5-60
5.5.2. Effects of Acute Oral Exposure 5-61
5.5.3. Effects of Dermal Exposure 5-62
5.5.4. Effects of Chronic Inhalation Exposure 5-63
5.5.5. Effects of Chronic Oral Exposure 5-64
5.5.6. Target Organ Toxicity 5-66
5.5.7. Factors that Modify the Toxicity of Chloroform . . .5-76
5.6. REFERENCES FOR CHAPTER 5 5-77
6. TERATOGENICITY AND REPRODUCTIVE EFFECTS 6-1
6.1. SUMMARY 6-13
6.2. REFERENCES FOR CHAPTER 6 6-14
7. MUTAGENICITY 7-1
7.1. INTRODUCTION 7-1
7.2. COVALENT BINDING TO MACROMOLECULES 7-1
7.3. MUTAGENICITY STUDIES IN BACTERIAL TEST SYSTEMS 7-4
7.4. MUTAGENICITY STUDIES IN EUCARYOTIC TEST SYSTEMS 7-10
7.5. OTHER STUDIES INDICATIVE OF DNA DAMAGE 7-16
7.6. CYTOGENETIC STUDIES 7-21
7.7. SUGGESTED ADDITIONAL TESTING 7-25
7.8. SUMMARY AND CONCLUSIONS 7-26
7.9. REFERENCES FOR CHAPTER 7 7-27
8. CARCINOGENICITY 8-1
8.1. ANIMAL STUDIES 8-1
8.1.1. Oral Administration (Gavage): Rat 8-2
8.1.1.1. National Cancer Institute (1976) 8-2
8.1.1.2. Palmer et al. (1979) 8-9
8.1.2. Oral Administration (Gavage): Mouse 8-11
vn
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TABLE OF CONTENTS (continued)
8.1.2.1. National Cancer Institute (1976) 8-11
8.1.2.2. Roe et al. (1979) 8-14
8.1.2.3. Eschenbrenner and Miller (1945) 8-18
8.1.2.4. Rudali (1967) 8-21
8.1.3. Oral Administration (Drinking Water): Rat
and Mouse 8-22
8.1.3.1. Jorgenson et al. (1985) 8-22
8.1.4. Oral Administration (Capsules): Dog 8-26
8.1.4.1. Heywood et al. (1979) 8-26
8.1.5. Intraperitoneal Administration: Mouse 8-29
8.1.5.1. Roe et al. (1968) 8-29
8.1.5.2. Theiss et al. (1977) 8-30
8.1.6. Evaluation of Chloroform Carcinogenicity
by Reuber (1979) 8-31
8.1.7. Oral Administration (Drinking Water): Mouse:
Promotion of Experimental Tumors 8-32
8.1.7.1. Capel et al. (1979) 8-32
8.2. CELL TRANSFORMATION ASSAY 8-38
8.2.1. Styles (1979) 8-38
8.3. EPIDEMIOLOGIC STUDIES .8-41
8.3.1. Young et al. (1981) 8-42
8.3.2. Hogan et al. (1979) 8-46
8.3.3. Cantor et al. (1978) 8-47
8.3.4. Gottlieb et al. (1981) 8-52
8.3.5. Alavanja et al. (1978) 8-54
8.3.6. Brenniman et al. (1978) 8-56
8.3.7. Struba et al. (1979) 8-58
8.3.8. Discussion 8-60
8.4. RISK ESTIMATES FROM ANIMAL DATA 8-63
8.4.1. Possible Mechanisms Leading to a
Carcinogenic Response for Chloroform 8-64
8.4.2. Selection of Animal Data Sets 8-67
8.4.2.1. NCI 1976 Bioassay (Mice): Liver
Tumors 8-67
8.4.2.2. NCI 1976 Bioassay (Rats): Kidney
Tumors •> .8-68
VI 11
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TABLE OF CONTENTS (continued)
8.4.2.3. Roe et al. 1979 Bioassay (Mice):
Kidney Tumors 8-69
8.4.2.4. Jorgenson et al. 1985 Bioassay
(Rats): Kidney Tumors 8-69
8.4.3. Interspecies Dose Conversion 8-71
8.4.3.1. General Considerations . ., 8-71
8.4.3.2. Calculation of Human Equivalent Doses . . 8-79
8.4.4. Choice of Risk Model 8-87
8.4.4.1. General Considerations 8-87
8.4.4.2. Mathematical Description of Low-Dose
Extrapolation Model 8-90
8.4.4.3. Adjustment for Less than Lifespan
Duration of Experiment 8-91
8.4.4.4. Additional Low-Dose Extrapolation .... 8-92
8.4.5. Unit Risk Estimates 8-93
8.4.5.1. Definition of Unit Risk 8-93
8.4.5.2. Calculation of the Slope of the
Dose-Risk Relationship for Chloroform . .8-93
8.4.5.3. Risk Associated with 1 iig/m3 of
Chloroform in Air 8-96
8.4.5.4. Risk Associated with 1 jig/L of
Chloroform in Drinking Water 8-96
8.4.5.5. Interpretation of Unit Risk Estimates . . 8-97
8.4.5.6. Reconciliation of Unit Risk Estimates
with Epidemiological Evidence 8-98
8.4.5.7. Discussion 8-98
8.5. RELATIVE CARCINOGENIC POTENCY 8-100
8.5.1. Derivation of Concept 8-100
8.5.2. Potency Index 8-100
8.6. SUMMARY 8-106
8.6.1. Qualitative 8-106
8.6.2. Quantitative 8-110
8.7. CONCLUSIONS 8-112
8.8. REFERENCES FOR CHAPTER 8 8-115
APPENDIX 8A COMPARISON AMONG DIFFERENT EXTRAPOLATION MODELS. . . 8A-1
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LIST OF TABLES
Table Page
3-1 Physical Properties of Chloroform 3-3
3-2 Chloroform Producers, Production Sites, and Capacities .... 3-9
3-3 Estimated Chloroform Discharges from Direct Sources 3-13
3-4 Ethylene Dichloride Producers, Production Sites, and
Capacities 3-15
3-5 Chloroform Discharges from Indirect Sources 3-21
3-6 Chlorodifluoromethane Producers and Production Sites ..... 3-23
3-7 Chloroform Discharges from Use 3-26
3-8 Relative Source Contribution for Chloroform 3-27
3-9 Ambient Levels of Chloroform 3-28
3-10 Acute and Chronic Effects of Chloroform on Aquatic
Organisms « •> 3-35
3-11 Values for kOH 3-38
3-12 Summary of EXAMS Models of the Fate of Chloroform 3-41
4-1 Physical Properties of Chloroform and Other
Chloromethanes 4-4
4-2 Partition Coefficients for Human Tissue at 37°C 4-4
4-3 Retention and Excretion of Chloroform by Man During and
After Inhalation Exposure to Anesthetic Concentrations .... 4-8
4-4 Chloroform Content in United Kingdom Foodstuffs and
in Human Autopsy Tissue 4-13
4-5 Concentration of Chloroform in Various Tissues of Two
Dogs After 2.5 Hours Anesthesia 4-16
4-6 Concentrations of Radioactivity (Chloroform Plus
Metabolites) in Various Tissues of the Mouse (NMRI) 4-17
4-7 Tissue Distribution of 14C-Chloroform Radioactivity
in CF/LP Mice After Oral Administration (60 mg/kg) 4-19
4-8 Pulmonary Excretion of 13CHC13 Following Oral Dose 4-25
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LIST OF TABLES (continued)
Table Page
4-9 Species Difference in the Metabolism of 14C-Chloroform . . . . 4-28
4-10 Kinetic Parameters for Chloroform After I.V. Administration
to Rats ........................... 4-29
4-11 Levels of Chloroform in Breath of Fasted Normal Healthy
Men ............................. 4-33
4-12 Cgvalent Binding of Radioactivity From 14C-Chloroform and
14C-Carbon Tetrachloride in Microsomal Incubation In Vitro. . 4-49
4-13 Mouse Strain Difference in Covalent Binding of Radioactivity
From 14C-Chloroform ..................... 4-52
4-14 In Vivo Covalent Binding of Radioactivity From 14CHC13
in Liver and Kidney of Male and Female Mice (C57BL/6) .... 4-54
4-15 In Vitro Covalent Binding of Radioactivity from
to Microsomal Protein from Liver and Kidney of Male and
Female Mice (C57BL/6) .................... 4-54
Covalent Binding of Radioactivity from *4C-Chloroform and
14C-Carbon Tetrachloride in Rat Liver Nuclear and Microson
4-16
Microsomal
Incubation In Vitro 4-58
4-17 Effect of Glutathione, Air, No or 0,0:0? Atmosphere
on the In Vitro Covalent Binding of CCT4, CHC13 and CBrCl3
to Rat Liver Microsomal Protein 4-60
4-18 Effects of 24-Hour Food Deprivation on Chloroform and
Carbon Tetrachloride In Vitro Microsomal Metabolism,
Protein, and P-450 Liver Contents of Rats 4-62
5-1 Relationship of Chloroform Concentration in Inspired
Air and Blood to Anesthesia 5-2
5-2 Dose-Response Relationships 5-6
5-3 Effects of Inhalation Exposure of Animals to Chloroform,
Five Days/Week for Six Months 5-18
5-4 Effects of Subchronic or Chronic Oral Administration of
Chloroform to Animals 5-20
5-5 Target Organ Toxicity of Chloroform 5-67
6-1 Summary of Results of the Schwetz et al. (1974) Study 6-3
6-2 Summary of Results of the Murray et al. (1979) Study 6-6
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LIST OF TABLES (continued)
Table Page
6-3 Summary of Effects of the Thompson et al. (1974) Study. . . . 6-10
7-1 Genetic Effects of Chloroform on Strain D7 of
S. Cerevlslae 7-12
7-2 Mitotic Index, Anaphase/Metaphase, and Presence of
Complete C-M1tosis 1n Grasshopper Embryos after Exposure
to CHCla Vapor 7-25
8-1 Effect of Chloroform on Kidney Epithelial Tumor Incidence
in Osborne-Mendel Rats 8-6
8-2 Effect of Chloroform on Thyroid Tumor Incidence in Female
Osborne-Mendel Rats ....... 8-7
8-3 Toothpaste Formulation for Chloroform Administration 8-9
8-4 Effects of Chloroform on Hepatocellular Carcinoma Incidence
in B6C3F1 Mice 8-13
8-5 Kidney Tumor Incidence in Male ICI Mice Treated with
Chloroform . . 8-17
8-6 Liver and Kidney Necrosis and Hepatomas in Strain A Mice
Following Repeated Oral Administration of Chloroform
in Olive Oil 8-20
8-7 Relative Tumor Incidence in Male Osborne-Mendel Rats
Treated with Chloroform in Drinking Water 8-24
8-8 Liver Tumor Incidence Rates in Female B6C3F1 Mice Treated
with Chloroform in Drinking Water 8-25
8-9 SGPT Changes in Beagle Dogs Treated with Chloroform 8-28
8-10 Effect of Oral Chloroform Ingestion on the Growth of Ehrlich
Ascite Tumors 8-35
8-11 Effect of Oral Chloroform Ingestion on Metastatic
"Tumor Takes" with 816 Melanoma 8-36
8-12 Effect of Oral Chloroform Ingestion on the Growth and
Spread of the Lewis Lung Tumor 8-37
8-13 Correlation Coefficients Between Residual Mortality
Rates in White Males and THM Levels in Drinking Water
by Region and by Percent of the County Population
Served in the United States 8-50
Xil
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LIST OF TABLES (continued)
Table Page
8-14 Correlation Coefficients Between Bladder Cancer
Mortality Rates by Sex and BTHM Levels in Drinking
Water by Region of the United States 8-50
8-15 Risk of Mortality from Cancer of the Rectum Associated
with Levels of Organics in Drinking Water 8-54
8-16 Cancer Risk Odds Ratios and 95% Confidence Intervals
(Chlorinated Versus Unchlorinated) 8-61
8-17 Incidence of Tumors in Experimental Animal Studies 8-68
8-18 Species Difference in the Metabolism of 14c-chloroform
(Oral Dose of 60 mg/kg) 8-78
8-19 Pulmonary Excretion of Chloroform Following Oral Dose. . . . 8-78
8-20 Continuous Human Equivalent Doses and Incidence of
Hepatocellular Carcinomas in Male and Female B6C3F1 Mice. . . 8-84
8-21 Continuous Human Equivalent Doses and Incidence of
Renal Tubular-Cell Adenocarcinomas in Male
Osborne-Mendel Rats 8-84
8-22 Continuous Human Equivalent Doses and Incidence of
Malignant Kidney Tumors in Male ICI Mice 8-85
8-23 Continuous Human Equivalent Doses and Incidence of
Renal Tubular-Cell Adenomas and Adenocarcinomas in Male
Osborne-Mendel Rats 8-85
8-24 Upper-Bound Estimates of Cancer Risk of 1 mg/kg/day,
Calculated by Different Models on the Basis of Different
Data Sets 8-95
8-25 Relative Carcinogenic Potencies Among 55 Chemicals Evaluated
by the Carcinogen Assessment Group as Suspect Human
Carcinogens 8-102
8A-1 Maximum Likelihood Estimate of the Parameters for Each
of the Four Extrapolation Models, Based on Different
Data Sets 8A-2
xiii
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LIST OF FIGURES
Figure Page
4-1 Rate of Rise of Alveolar (Arterial) Concentration Toward
Inspired Concentration For Five Anesthetic Agents of
Differing Ostwald Solubilities 4-9
4-2 Arteriovenous Blood Concentrations of a Patient During
Anesthesia with Chloroform 4-10
4-3 Exponential Decay of Chloroform, Carbon Tetrachloride,
Perchloroethylene and Trichloroethylene in Exhaled
Breath of a 48 Year-old Male Accidentally Exposed to
Vapors of These Solvents 4-21
4-4 Relationship Between Total 8-Hour Pulmonary Excretion of
Chloroform Following 0.5-g Oral Dose in Man and the
Deviation of Body Weight From Ideal 4-26
4-5 Blood and Adipose Tissue Concentrations of Chloroform During
and After Anesthesia in a Dog 4-32
4-6 Metabolic Pathways of Chloroform Biotransformation 4-34
4-7 Metabolic Pathways of Carbon Tetrachloride
Biotransformation 4-41
4-8 Rate of Carbon Monoxide Formation After Addition of Various
Halomethanes to Sodium Dithionite-reduced Liver Microsomal
Preparations From Phenobarbito!-treated Rats 4-46
4-9 Effect of Increasing Dosage of i.p.-Injected 14C-Chloroform
on Extent of Covalent Binding of Radioactivity in vivo to
Liver and Kidney Proteins of Male Mice 6 Hours
after Administration 4-50
4-10 Comparison of Irreversible Binding of Radioactivity from
I4C-CHC13 to Protein and Lipid of Microsomes from
Normal Rabbit, Rat, Mouse, and Human Liver Incubated
In Vitro at 37°C in 02 4-55
5-1 Probable Pathways of Metabolism of Chloroform in the
Kidney. 5-44
8-1 Survival Curves for Fisher 344 Rats in a Cardnogenicity
Bioassay on Chloroform 8-4
8-2 Negative Result in Transformation Assay of Chloroform
which was also Negative in the Ames Assay 8-40
XT V
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LIST OF FIGURES (continued)
8-3 Frequency Distribution of CHC13 Levels in 76 U.S.
Drinking Water Supplies 8-49
8-4 Effect of Increasing Dosage of i.p.-injected 14c-chloroform
on Extent of Covalent Binding of Radioactivity jn vivo
to Liver and Kidney Proteins of Male Mice 6 Hours After
Administration 8-74
8-5 Comparison of Irreversible Binding of Radioactivity
from 14C-CHC13 to Protein and Lipid of Microsomes from
Normal Rabbit, Rat, Mouse, and Human Liver Incubated
In vitro at 37°C in 02 8-75
8-6 Allometric Relationship (Y=aWn) Between Species Body
Weight (1n order: mouse, rat, squirrel monkey, and man)
and the Amount Metabolized of a Common Oral Dose of
Chloroform as Calculated from the Data of Fry et al.
(1972) and Brown et al. (1974) 8-80
8-7 The Relationship Between the Equivalent Human Exposure
Dose and Bioassay Tumor Incidence 8-86
8-8 Histogram Representing the Frequency Distribution of the
Potency Indices of 55 Suspect Carcinogens Evaluated by
the Carcinogen Assessment Group 8-101
xv
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AUTHORS, CONTRIBUTORS, AND REVIEWERS
The EPA Office of Health and Environmental Assessment (OHEA) is
responsible for the preparation of this health assessment document. The OHEA
Environmental Criteria and Assessment Office (ECAO/RTP) Research Triangle
Park, NC 27711, had overall responsibility for coordination and direction of
the document production effort (Si Duk Lee, Ph.D., Project Manager, ECAO/RTP,
919-541-4159).
AUTHORS
Larry Anderson, Ph.D.
Carcinogen Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
David Bayliss, M.S.
Carcinogen Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Chao W. Chen, Ph.D.
Carcinogen Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Joan P. Coleman, Ph.D.
Syracuse Research Corporation
Syracuse, NY
I.W.F. Davidson, Ph.D.
Bowman Gray School of Medicine
Winston-Sal em, NC
D. Anthony Gray, Ph.D.
Syracuse Research Corporation
Syracuse, NY
Si Duk Lee, Ph.D.
Environmental Criteria and Assessment Office
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Research Triangle Park, NC
Chapter 8
Chapter 8
Chapter 8
Chapter 5
Chapter 4
Chapter 3
Chapter 2
xvi
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Jean C. Parker, Ph.D. Chapters 4,8
Carcinogen Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Sheila Rosenthal, Ph.D. Chapter 7
Reproductive Effects Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Carol Sakai, Ph.D. Chapter 6
Reproductive Effects Assessment Group
Office of Health and Environmental Assessment
U.S. Environmental Protection Agency
Washington, DC
Sharon B. Wilbur, M.A. Chapter 5
Syracuse Research Corporation
Syracuse, NY
U.S. Environmental Protection Agency Peer Reviewers
Karen Blanchard
Office of Air Quality Planning and Standards
Research Triangle Park, NC
Lester D. Grant, Ph.D.
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Research Triangle Park, NC
Joseph Padgett
Office of Air Quality Planning and Standards
Research Triangle Park, NC
Jerry F. Stara, D.V.M.
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Cincinnati, OH
Environmental Criteria and Assessment Office Support Staff
F. Vandiver Bradow
Allen Hoyt
XVI 1
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The OHEA Carcinogen Assessment Group (CAG) was responsible for
preparation of the sections on carcinogenicity. Participating members of the
CAG are listed below (principal authors of present carcinogenicity materials
are designated by an asterisk [*]).
Participating Members of the Carcinogen Assessment Group
Roy E. Albert, M.D. (Chairman)
Elizabeth L. Anderson, Ph.D.
Larry D. Anderson, Ph.D.*
Steven Bayard, Ph.D.
David L. Bayliss, M.S.*
Robert P. Beliles, Ph.D.*
Chao W. Chen, Ph.D.*
Margaret M.L. Chu, Ph.D.
James C. Cogliano, Ph.D.*
Herman J. Gibb, B.S., M.P.H.
Bernard H. Haberman, D.V.M., M.S.
Charalingayya B. Hiremath, Ph.D.
Robert E. McGaughy, Ph.D.
Jean C. Parker, Ph.D.*
Charles H. Ris, M.S., P.E.
Dharm V. Singh, D.V.M., Ph.D.
Todd W. Thorslund, Sc.D.
The OHEA Reproductive Effects Assessment Group (REAG) was responsible
for preparation of the sections on mutagenicity, teratogenicity, and
reproductive effects. Participating members of the REAG are listed below
(principal authors of present sections are designated by an asterisk [*]).
Participating Members of the Reproductive Effects Assessment Group
Eric D. Clegg, Ph.D.
John R. Fowle, III, Ph.D.
David Jacobson-Kram, Ph.D.
K.S. Lavappa, Ph.D.
Sheila L. Rosenthal, Ph.D.*
Carol N. Sakai, Ph.D.*
Lawrence R. Valcovic, Ph.D.
Vicki Vaughan-Dellarco, Ph.D.
Peter E. Voytek, Ph.D. (Director)
xvi i i
-------
External Peer Reviewers
Dr. Karim Ahmed
Natural Resources Defense Fund
122 East 42nd Street
New York, NY 10168
Dr. Eula Bingham
Graduate Studies and Research
University of Cincinnati (ML-627)
Cincinnati, OH 45221
Dr. James Buss
Chemical Industry Institute of
Toxicology
Research Triangle Park, NC 27709
Dr. I.W.F. Davidson
Wake Forest University
Bowman Gray School of Medicine
Winston-Sal em, NC
Dr. Larry Fishbein
National Center for Toxicological
Research
Jefferson, AR 72079
(501) 542-4390
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, NC 27514
Dr. Marshall Johnson
Thomas Jefferson Medical College
Department of Anatomy
1020 Locust Street
Philadelphia, PA 19107
Dr. Trent Lewis
National Institute for Occupational
Safety and Health
26 Columbia Parkway
Cincinnati, OH 45226
(513) 684-8394
Dr. Richard Reitz
Oow Chemical, USA
Toxicological Research Laboratory
1803 Building
Midland, MI 48640
Dr. Bernard Schwetz
National Institute of
Environmental Health Sciences
Research Triangle Park, NC 27709
Or. James Selkirk
Oak Ridge National Laboratory
Oak Ridge, TN 37820
(615) 624-0831
Dr. Samuel Shibko
Food and Drug Administration
Division of Toxicology
200 C Street, SW
Washington, DC 20204
Dr. Robert Tardiff
1423 Trapline Court
Vienna, VA 22180
(703) 276-7700
Dr. Norman M. Trieff
University of Texas Medical Branch
Department of Pathology
Galveston, TX 77550
(409) 761-1895
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, NY 10016
(212) 340-5629
Dr. James Withey
Health and Protection Branch
Department of National Health &
Welfare
Tunney's Pasture
Ottawa, Ontario KIA 01Z Canada
Mr. Matthew Van Hook
Consultant
1133 North Harrison Street
Arlington, VA 22205
xix
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SCIENCE ADVISORY BOARD
ENVIRONMENTAL HEALTH COMMITTEE
CHLORINATED ORGANICS SUBCOMMITTEE
The content of this health assessment document on chloroform was
independently peer-reviewed in public session by the Chlorinated Organics
Subcommittee of the Environmental Health Committee of the Environmental
Protection Agency's Science Advisory Board.
ACTING CHAIRMAN. ENVIRONMENTAL HEALTH COMMITTEE
Dr. John Doull, Professor of Pharmacology and Toxicology, University of
Kansas Medical Center, Kansas City, Kansas 66207
EXECUTIVE SECRETARY. SCIENCE ADVISORY BOARD
Dr. Daniel Byrd III, Executive Secretary, Science Advisory Board, A-101 F,
U.S. Environmental Protection Agency, Washington, DC 20460
MEMBERS
Dr. Seymour Abrahamson, Professor of Zoology and Genetics, Department of
Zoology, University of Wisconsin, 500 Highland Avenue, Madison,
Wisconsin 53706
Dr. Ahmed E. Ahmed, Associate Professor of Pathology, Pharmacology, and
Toxicology, The University of Texas Medical Branch, Galveston, Texas
77550.
Dr. George T. Bryan, Professor of Human Oncology, K4/528 C.S.C Clinical
Sciences, University of Wisconsin, 500 Highland Avenue, Madison,
Wisconsin 53792.
Dr. Ronald 0. Hood, Professor and Coordinator, Cell and Developmental Biology
Section, Department of Biology, The University of Alabama, and Principal
Associate, R.D. Hood and Associates, Consulting Toxicologists, P.O. Box
1927, University, Alabama 35486.
Dr. K. Roger Hornbrook, Department of Pharmacology, P.O. Box 26901,
University of Oklahoma, Oklahoma City, Oklahoma 73190.
Dr. Thomas Starr, Chemical Industry Institute of Toxicology, P.O. Box 12137,
Research Triangle Park, North Carolina 27709.
xx
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Review
Draft
(Do Not
Cite or Quote)
EPA-600/8-84-004A
March 1984
External Review Draft
Health Assessment
Document for Chloroform
Part 1 of 2
External Review Draft
NOTICE
This document is a preliminary draft. It has not been formally released by the U.S. Environmental
Protection Agency and should not at this stage be construed to represent Agency policy. It
is being circulated for comment on its technical accuracy and policy implications.
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Research and Development
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Research Triangle Park, NC 27711
March 1984
-------
PREFACE
The Office of Health and Environmental Assessment has prepared this health
assessment to serve as a "source document" for EPA use. This health assessment
document was developed for use by the Office of Air Quality Planning and
Standards to support decision-making regarding possible regulation of chloroform
as a hazardous air pollutant.
In the development of the assessment document, the scientific literature
has been inventoried, key studies have been evaluated and summary/conclusions
have been prepared so that chemical's toxicity and related characteristics
are qualitatively identified. Observed effect levels and other measures of
dose-response relationships are discussed, where appropriate, so that the
nature of the adverse health response are placed in perspective with observed
environmental levels.
This document will be subjected to a thorough copy editing and proofing
following the revision based on the EPA's Scientific Advisory Board review
comments.
-------
TABLE OF CONTENTS
Page
LIST OF TABLES vi
LIST OF FIGURES *
1 . SUMMARY AND CONCLUSIONS 1 -1
2. INTRODUCTION 2-1
3. BACKGROUND INFORMATION 3-1
3.1 INTRODUCTION 3-1
3.2 PHYSICAL AND CHEMICAL PROPERTIES 3-2
3.3 SAMPLING AND ANALYSIS 3-4
3.3.1 Chloroform in Air 3-4
3.3.2 Chloroform in Water 3-5
3.3.3 Chloroform in Blood 3-6
3.3.4 Chloroform in Urine 3-6
3.3.5 Chloroform in Tissue 3-6
3.4 EMISSIONS FROM PRODUCTION AND USE 3-6
3.4.1 Emissions from Production 3-6
3.4.2 Emissions from Use 3-20
3.4.3 Summary of Chloroform Discharges from Use 3-26
3.5 AMBIENT AIR CONCENTRATIONS 3-26
3.6 ATMOSPHERIC REACTIVITY 3-32
3.7 ECOLOGICAL EFFECTS/ENVIRONMENTAL PERSISTENCE 3-33
3.7.1 Ecological Effects 3-33
3.7.2 Environmental Persistence 3-36
3.8 EXISTING CRITERIA, STANDARDS, AND GUIDELINES 3-39
3.8.1 Air 3-39
3.8.2 Water 3-41
3.8.3 Food 3-41
3.8 .4 Drugs and Cosmetics 3-42
3.9 RELATIVE SOURCE CONTRIBUTIONS 3-42
3.10 REFERENCES 3-43
4. DISPOSITION AND RELEVANT PHARMACOKINETICS 4-1
4.1 INTRODUCTION 4-1
111
-------
TABLE OF CONTENTS (cont.)
Page
4. 2 ABSORPTION 4-2
4.2.1 Dermal Absorption 4-2
4.2.2 Oral Absorption 4-3
4.2.3 Pulmonary Absorption 4-7
4.3 TISSUE DISTRIBUTION 4-12
4.4 EXCRETION 4-21
4.4.1 Pulmonary Excretion 4-21
4.4.2 Other Routes of Excretion 4-31
4.4.3 Adipose Tissue Storage 4-32
4.5 BIOTRANSFORMATION OF CHLOROFORM 4-34
4.5.1 Known Metabolites 4-34
4.5.2 Magnitude of Chloroform Metabolism 4-37
4.5.3 Enzymic Pathways of Biotransformation 4-40
4.6 COVALENT BINDING TO CELLULAR MACROMOLECULES 4-4 7
4.6.1 Proteins and Lipids 4-47
4.6.2 Nucleic Acids 4-58
4.6.3 Role of Phosgene 4-59
4.6.4 Role of Glutathione 4-61
4.7 SUMMARY 4-63
4.8 REFERENCES 4-67
5. TOXICITY 5-1
5.1 EFFECTS OF ACUTE EXPOSURE TO CHLOROFORM 5-1
5.1.1 Humans 5-1
5.1 .2 Experimental Animals 5-6
5.2 EFFECTS OF CHRONIC EXPOSURE TO CHLOROFORM 5-13
5.2.1 Humans 5-13
5.2.2 Experimental Animals 5-15
5.3 INVESTIGATION OF TARGET ORGAN TOXICITY IN EXPERIMENTAL
ANIMALS 5-28
5.3.1 Hepatotoxicity 5-28
5.3.2 Nephrotoxicity 5-34
-------
TABLE OF CONTENTS (cont.)
5.4 FACTORS MODIFYING THE TOXICITY OF CHLOROFORM 5-38
5.4.1 Factors that Increase Toxicity 5-39
5.4.2 Factors that Decrease Toxicity 5-44
5.5 SUMMARY; CORRELATION OF EXPOSURE AND EFFECT 5-46
5.5.1 Effects of Acute Inhalation Exposure 5-46
5.5.2 Effects of Acute Oral Exposure 5-4?
5.5.3 Effects of Dermal Exposure 5-48
5.5.4 Effects of Chronic Inhalation Exposure 5-48
5.5.5 Effects of Chronic Oral Exposure 5-50
5.5.6 Target Organ Toxicity 5-51
5.5.7 Factors that Modify the Toxicity of Chloroform 5-62
5.6 REFERENCES 5-63
6. TERATOGENICITY AND REPRODUCTIVE EFFECTS 6-1
6.1 REFERENCES 6-1 6
7. MUTAGENICITY 7-1
7.1 INTRODUCTION 7-1
7.2 COVALENT BINDING TO MACROMOLECULES 7-1
7. 3 MUTAGENICITY STUDIES IN BACTERIAL TEST SYSTEMS 7-3
7.4 MUTAGENICITY STUDIES IN EUCARYOTIC TEST SYSTEMS 7-9
7. 5 OTHER STUDIES INDICATIVE OF DNA DAMAGE 7-14
7.6 CHROMOSOME STUDIES 7-19
7.7 SUGGESTED ADDITIONAL TESTING 7-21
7.8 REFERENCES 7-23
8. CARCINOGENICITY 8-1
8.1 ANIMAL STUDIES 8-1
8.1.1 Oral Administration (Gavage): Rat 8-2
8.1.2 Oral Administration (Gavage): Mouse 8-11
8.1.3 Oral Administration (Capsules): Dog 8-21
8.1.4 Intraperitoneal Administration: Mouse 8-24
8.1.5 Evaluation of Chloroform Carcinogenicity by
Reuber (1979) 8-26
8.1.6 Oral Administration (Drinking Water): Mouse:
Promotion of Experimental Tumors 8 -26
8 .2 CELL TRANSFORMATION ASSAY 8-32
8.2.1 Styles (1979) 8-32
-------
TABLE OF CONTENTS (cont.)
Page
8.3 EPIDEMIOLOGIC STUDIES 8-34
8.3.1 Young et al. (1981) 8-37
8.3.2 Hogan et al. (1979) 8-40
8.3.3 Cantor et al. (1978) 8-42
8.3.4 Gottlieb et al. (1981) 8-46
8.3.5 Alavanja et al. (1978) 8-48
8.3.6 Brenniman et al. (1978) 8-50
8.3.7 Struba (1979) 8-51
8.3.8 Discussion 8-53
8.4 QUANTITATIVE ESTIMATION 8-56
8.4.1 Procedures for the Determination of Unit Risk 8-59
8.4.2 Unit Risk Estimates 8-70
8.4.3 Comparison of Potency with Other Compounds. 8-80
8.4.4 Summary of Quantitative Assessment 8-80
8.5 SUMMARY 8-85
8.5.1 Qualitative 8-85
8.5.2 Quantitative 8-88
8.6 CONCLUSIONS 8-89
8.7 REFERENCES 8-91
8.8 APPENDIX A: Comparison Among Various Extrapolation
Models A-l
-------
LIST OF TABLES
Table Page
3-1 Physical Properties of Chloroform 3-3
3-2 Chloroform Producers, Production Sites, and Capacities 3-8
3-3 Chloroform Discharges from Direct Sources 3-1 3
3-4 Ethylene Dichloride Producers, Production Sites and
Capacities 3-1 5
3-5 Chloroform Discharges from Indirect Sources 3-21
3-6 Chlorodifluoromethane Producers and Production Sites 3-23
3-7 Chloroform Discharges from Use 3-27
3-8 Relative Source Contribution for Chloroform 3-28
3-9 Ambient Levels of Chloroform 3-29
3-1 0 Acute and Chronic Effects of Chloroform on Aquatic
Organisms 3-34
3-11 Values for knu 3-38
Un
3-12 Summary of EXAMS Models of the Fate of Chloroform 3-40
4-1 Physical Properties of Chloroform and Other
Chloromethanes 4-4
4-2 Partition Coefficients for Human Tissue at 37°C 4-5
4-3 Retention and Excretion of Chloroform in Man During and
After Inhalation Exposure to Anesthetic Concentations 4-9
4-4 Chloroform Content in United Kingdom Foodstuffs and
in Human Autopsy Tissue 4-1 4
4-5 Concentration of Chloroform in Various Tissues of Two
Dogs After 2. 5 Hours Anesthesia 4-16
4-6 Concentration of Radioactivity (Chloroform Plus
Metabolites) in Various Tissues of the Mouse (N MRI) 4-18
4-7 Tissue Distribution of C-Chloroform Radioactivity
in CF/LP Mice After Oral Administration (60 mg/kg) 4-20
-------
LIST OF TABLES (cont.)
Table Page
4-8 Pulmonary Excretion of CHCL~ Following Oral
Dose: Percent of Dose 4-26
4-9 Species Difference in the Metabolism of C-Chloroform 4-29
4-4b Kinetic Parameters for Chloroform After I.V. Administration
to Rats 4-30
4-1 0 Levels of Chloroform in Breath of Fasted Normal
Healthy Men 4-35
4-11 Cojplent Binding of Radioactivity From C-Chloroform and
C-Carbon Tetrachloride in Microsomal Incubation
In Vitro 4-49
4-12 Mouse Strain Difference in Covalent Binding of Radioactivity
From C-Chloroform 4-53
4-13 In Vivo Covalent Binding of Radioactivity From CHC1o
in Liver and Kidney of Male and Female Mice (C57BL/6) 4-55
4-1 4 In Vitro Covalent Binding of Radioactivity from CHC1
to Microsomal Protein from Liver and Kidney of Male and
Female Mice (C57BL/6) 4-56
1 4
4-15 Covalent Binding of Radioactivity from C-Chloroform and
C-Carbon Tetrachloride in Rat Liver Nuclear and
Microsomal Incubation In Vitro 4-60
4-16 Effect of Glutathione, Air, N? or CO: 0? Atmosphere
on the In Vitro Covalent Binding of C Cl^ and C Br Cl
to Rat Liver Microsomal Protein 4-62
4-1 7 Effects of 24-Hour Food Deprivation on Chloroform and
Carbon Tetrachloride In Vitro Microsomal Metabolism,
Protein, and P-450 Liver Contents of Rats 4-64
5-1 Relationship of Chloroform Concentration in Inspired
Air and Blood to Anesthesia 5-2
5-2 Dose-Response Relationships 5-7
5-3 Effects of Inhalation Exposure of Animals to Chloroform,
Five Days/Week for Six Months 5-17
5-4 Effects of Subchronic or Chronic Oral Administration of
Chloroform to Animals 5-19
vm
-------
LIST OF TABLES (cont.)
Table
5-5 Target Organ Toxicity of Chloroform 5-52
7-1 Genetic Effects of Chloroform on Strain D7 of
S. Cerevisiae 7-10
8-1 Effect of Chloroform on Kidney Epithelial Tumor Incidence
in Osborne-Mendel Rats 8-5
8-2 Effect of Chloroform on Thyroid Tumor Incidence in Female
Osborne-Mendel Rats 8-7
Toothpaste Formulation for Chloroform Administration 8-9
Effects of Chloroform on Hepatocellular Carcinoma Incidence
in B6C3F1 Mice 8-13
8-5 Kidney Tumor Incidence in Male ICI Mice Treated with
Chloroform 8-17
8-6 Liver and Kidney Necrosis and Hepatomas in Strain A Mice
Following Repeated Oral Administration of Chloroform
in Olive Oil 8-19
8-7 SGPT Changes in Beagle Dogs Treated with Chloroform 8-23
8-8 Effect of Oral Chloroform Ingestion on the Growth of Ehrlich
Ascites Tumors 8-29
8-9 Effect of Oral Chloroform Ingestion on Metastatic Tumor Takes
with B1 6 Melanoma 8 -30
8-10 The Effect of Oral Chloroform Ingestion on the Growth and
Spread of the Lewis Lung Tumor 8 -31
8-11 Correlation Coefficients Between Residual Mortality
Rates in White Males and THM Levels in Drinking Water
by Region and by Percent of the County Population
Served in the United States 8-44
8-12 Correlation Coefficients Between Bladder Cancer
Mortality Rates by Sex and BTHM Levels in Drinking
Water by Region of the United States 8-45
8-13 Risk of Cancer of the Rectum Mortality Associated
with Level of Organics in Drinking Water 8 -48
IX
-------
LIST OF TABLES (cont.)
Table Page
8-1*4 Cancer Risk Odds Ratios and 95% Confidence Intervals
(Chlorinated Versus Unchlorinated) 8-54
8-15 Incidence of Hepatocellular Carcinomas in Female and Male
B6C3F1 Mice 8-71
8-16 Incidence of Tubular-Cell Adenocarcinomas in Male
Osborne-Mendel Rats 8-72
8-17 Incidence of Malignant Kidney Tumors in Male ICI Mice 8-72
8-18 Upper-Bound Estimates of Cancer Risk of 1 mg/kg/day,
Calculated by Different Models on the Basis of Different
Data Sets 8-75
8-19 Relative Carcinogenic Potencies Among 53 Chemicals Evaluated
by the Carcinogen Assessment Group as Suspect Human
Carcinogens 8-82
-------
LIST OF FIGURES
Figure Page
4-1 Rate of Rise of Alveolar (Arterial) Concentration Toward
Inspired Concentration For Five Anesthetic Agents of
Differing Ostwald Solubilities 4-10
4-2 Arteriovenous Blood Concentrations of a Patient During
Anesthesia with Chloroform 4-11
4-3 Exponential Decay of Chloroform, Carbon Tetrachloride,
Perchloroethylene and Trichloroethylene in Exhaled
Breath of 48 Year-old Male Accidentally Exposed to
Vapors of These Solvents 4-22
4-4 Relationship Between Total 8-Hour Pulmonary Excretion of
Chloroform Following 0.5-g Oral Dose in Man and the
Deviation of Body Weight From Ideal 4-27
4-5 Blood and Adipose Tissue Concentrations of Chloroform During
and After Anesthesia in a Dog 4-33
4-6 Metabolic Pathways of Chloroform Biotransformation.
(Identified CH Cl_ metabolites are underlines) 4-36
4-7 Metabolic Pathways of Carbon Tetrachloride
Biotransformation 4-42
4-8 Rate of Carbon Monoxide Formation After Addition of Various
Halomethanes to Sodium Dithionite-reduced Liver Microsomal
Preparations From Phenobarbitol-treated Rats 4-46
4-9 Effect of Increasing Dosage of i.p.-Injected
C-Chloroform on Extent of Covalent Binding of
Radioactivity In Vivo to Liver and Kidney Proteins of Male
Mice 6 Hours After Administration 4-51
4-1 0 Comparison of irreversible binding of radioactivity from
C-CHC1_ to protein and lipid of microsomes from
normal rabbit, rat, mouse, and human liver incubated
in vitro at 37°C in 02 4-57
8-1 Survival curves for Fisher 344 Rats in a Carcinogenicity
Bioassay on Chloroform 8-5
8-2 Negative Result in Transformation Assay of Chloroform
which was also Negative in the Ames Assay 8-35
8-3 Frequency distribution of CHC13 levels in 68 U.S.
drinking water supplies. The abscissa is linear in the
logarithm of the level 8-43
-------
LIST OF FIGURES (cont.)
8-4 Point and Upper-Bound Estimates of Four Dose-Response Models
Over Low-Dose Region on the Basis of Liver Tumor Data
for Female Mice 8-76
8-5 Point and Upper-Bound Estimates of Four Dose-Response Models
Over Low-Dose Region on the Basis of Liver Tumor Data
for Male Mice 8-77
8-6 Histogram Representing the Frequency Distribution of the
Potency Indices of 53 Suspect Carcinogens Evaluated by
the Carcinogen Assessment Group 8-81
XII
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The Office of Health and Environmental Assessment (OHEA), U.S. EPA,
is responsible for the preparation of this health assessment document. The
Environmental Criteria and Assessment Office (ECAO/RTP), OHEA, had the
overall responsibility for coordination and the document production effort.
Project Manager
Si Duk Lee, Ph.D.
Environmental Criteria and Assessment Office
U.S. EPA, Research Triangle Park, N.C. 27711
(919) 541-4159
Authors and Reviewers
The principal authors of this document are:
Larry Anderson Ph.D.
Carcinogen Assessment Group
U.S. EPA, Washington, D.C.
David Baylis, M.S.
Carcinogen Assessment Group
U.S., EPA, Washington, D.C.
Chao W.'Chen, Ph.D.
Carcinogen Assessment Group
U.S., EPA, Washington, D.C.
Carol Sakai, Ph.D.
Reproductive Effects Assessment Group
U.S., EPA, Washington, D.C.
Sheila Rosenthal, Ph.D.
Reproductive Effects Assessment Group
U.S., EPA, Washington, D.C.
I.W.F. Davidson, Ph.D.
Bowman Gray School of Medicine
Winston Salem, N.C.
D. Anthony Gray, Ph.D.
Syracuse Research Corp.
Syracuse, N.Y.
Sharon B. Wilbur, M.A.
Syracuse Research Corp.
Syracuse, N.Y.
Joan P. Coleman, Ph.D.
Syracuse Research Corp.
Syracuse, N.Y.
-------
The following individuals provided peer-review of this draft or earlier
drafts of this document:
U.S. Environmental Protection Agency
Joseph Padgett
Office of Air Quality Planning and Standards
U.S. EPA
Karen Blanchard
Office of Air Quality Planning and Standards
U.S. EPA
Jerry F. Stara, D.V.M.
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
U.S. EPA
Lester D. Grant, Ph.D.
Office of Health and Environmental Assessment
Environmental Criteria and Asssessment Office
U.S. EPA
Participating Members of the Carcinogen Assessment Group
Roy E. Albert, M.D. (Chairman)
Elizabeth L. Anderson, Ph.D.
Larry D. Anderson, Ph.D.
Steven Bayard, Ph.D.
David L. Bayliss, M.S.
Chao W. Chen, Ph.D.
Margaret M. L. Chu, Ph.D.
Bernard H. Haberman, D.V.M., M.S.
Charalingayya B. Hiremath, Ph.D.
Robert E. McGaughy, Ph.D.
Dharm W. Singh, D.V.M., Ph.D.
Todd W. Thorslund, Sc.D.
Participating Members of the Reproductive Effects Assessment Group
Peter E. Voytek, Ph.D. (Director)
John R. Fowle, III, Ph.D.
Carol Sakai, Ph.D.
Ernest Jackson, M.D.
K.S. Lavappa, Ph.D.
Sheila Rosenthal, Ph.D.
Vicki Vaughn-Dellarco, Ph.D.
XIV
-------
External Peer Reviewers
Dr. Karim Ahmed
Natural Resources Defense Fund
122 E. 42nd Street
New York, N.Y. 10168
Dr. Eula Bingham
Graduate Studies and Research
University of Cincinnati (ML-627)
Cincinnati, Ohio 45221
(513) 475-4532
Dr. James Buss
Chemical Institute of Industrial
Toxicology
Research Triangle Park, N.C. 27709
Dr. I.W.F. Davidson
Wake Forest University
Bowman Gray Medical School
Winston Salem, N.C.
Dr. Larry Fishbein
National Center for Toxicological
Research
Jefferson, Arkansas 72079
(501) 542-4390
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, N.C. 27514
Dr. Marshall Johnson
Thomas Jefferson Medical College
Department of Anatomy
1020 Locust Street
Philadelphia, Pennsylvania 19107
Dr. Trent Lewis
National Institute of Occupational
Safety and Health
26 Columbia Parkway
Cincinnati, Ohio 45226
(513) 684-8394
Dr. Richard Reitz
Dow Chemical, USA
Toxicology Research Laboratory
1803 Building
Midland, Michigan 48640
Dr. Marvin A. Schneiderman
Clement Associates, Incorporated
Arlington, Virginia 22209
(703) 276-7700
Dr. Bernard Schwetz
National Institute of Environmental
Health
Research Triangle Park, NC 27709
(919) 541-7992
Dr. James Selkirk
Oak Ridge National Laboratory
Oakridge, Tennessee 37820
(615) 624-0831
Dr. Samuel Shibko
Food and Drug Administration
Division of Toxicology
200 C Street, S.W.
Washington, D.C. 20204
Telephone:
Dr. Robert Tardiff
1423 Trapline Court
Vienna, Virginia 22180
(703) 276-7700
Dr. Norman M. Trieff
University of Texas Medical Branch
Department of Pathology, UTMB
Galveston, Texas 77550
(409) 761-1895
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, New York 10016
(212) 340-5629
Dr. James Withey
Health and Protection Branch
Department of National Health &
Welfare
Tunney's Pasture
Ottawa, Ontario
CANADA, KIA 01Z
xv
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1. SUMMARY AND CONCLUSIONS
Chloroform is a dense, colorless, volatile liquid used primarily in the
production of chlorodifluoromethane (90%) and for export (5/£). Non-consumptive
uses (5%) include use as a solvent, as a cleaning agent, and as a fumigant
ingredient. Direct United States production of chloroform in 1981 was 184
million kg, while indirect production is estimated at 13-2 million kg (=193
million kg overall). The amount of chloroform emitted to air is estimated to be
7.2 million kg, emissions to water are 2.6 million kg, and emissions to land are
0.6 million kg. Total United States emissions are estimated to be 10.4 million
kg.
Chloroform is ubiquitous in the environment, having been found in urban and
non-urban locations. The northern hemisphere background average has been deter-
—1 2
mined to be 14 ppt (10~ v/v), while the southern hemisphere has been determined
to be <3 ppt. The global average is 8 ppt. For the most part, urban ambient air
concentrations remain <1000 ppt, and rural or remote locations can be <10 ppt.
There are some notable exceptions, however, the reasons for this are not readily
apparent. The highest values reported were in Rutherford, New Jersey,
(31,000 ppt) and Niagara Falls, New York (21,611 ppt).
Hydroxyl radical oxidation is the primary atmospheric reaction of chloro-
form. Based on the rate constant for reaction with chloroform, a half-life of
21.5 weeks is expected. The principal products from this reaction are HC1 and
CO,,. It has been estimated that roughly 1/& of the tropospheric chloroform will
diffuse into the stratosphere, based on a lifetime of 0.2 to 0.3 years and a
troposphere-to-stratosphere turnover time of 30 years. An EXAMS model of chloro-
form in water confirms other data that the major removal process for chloroform
in water is evaporation.
1-1
-------
The best analytical method for detection of chloroform appears to be gas
chromatography with electron capture or electrolytic conductivity detector.
This gives a detection limit of <5 ppt.
The pharmacokinetics and metabolism of chloroform have been studied in both
humans and experimental animals. Chloroform is rapidly and extensively absorbed
through the respiratory and gastrointestinal tracts. Absorption through the
skin would make a significant contribution to body burden only in instances of
contact of the skin with liquid chloroform.
The available data suggest that, for resting human at least 2 hours of
exposure are required to reach an apparent body equilibrium with the inhaled
chloroform concentration. The percentage of the inhaled chloroform concentration
retained in the body at "equilibrium" would be -65%, and is independent of the
inhaled concentration. The magnitude of chloroform uptake into the body (dose or
body burden) is directly proportional to the concentration of chloroform in the
inspired air, the duration of exposure, and the respiratory minute volume, and
can be estimated by multiplying the percent retention by the total volume of air
breathed during exposure and by the exposure concentration.
The absorption of chloroform from the gastrointestinal tract appears to be
virtually complete, judging from recovery of unchanged chloroform and metabol-
ites in the exhaled air of humans and in the exhaled air, urine, feces, and
carcass of experimental animals. Peak blood levels occurred at =1 hour after
oral administration of chloroform in olive oil to humans or animals.
After inhalation or ingestion, highest concentrations of chloroform are
found in tissues with higher lipid contents. Results from the administration of
C-labeled chloroform to animals indicate that the distribution of radioactiv-
ity (reflecting both chloroform and its metabolites) may be affected by the route
of exposure. Oral administration appeared to result in the accumulation of a
1-2
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greater proportion of radioactivity in the liver than did inhalation exposure,
but differences in experimental protocols make this interpretation tentative.
Sex differences in the distribution of chloroform and its metabolites were found
only in mice and not in rats or squirrel monkeys. The kidneys of male mice
accumulated strikingly more radioactivity than did those of female mice. Other
than the renal accumulation of radioactivity in male mice, the tissue distribu-
tion of radioactivity after oral administration was similar in mice, rats, and
squirrel monkeys.
Chloroform has been detected in fetal liver. Chloroform would be expected
to appear in human milk, because it has been found in cow's milk, cheese, and
butter.
Chloroform is metabolized via microsomal cytochrome P-450 oxidation to a
reactive intermediate. The primary end product of chloroform metabolism is CC^?
but small amounts of the reactive intermediate bind covalently to tissue macro-
molecules or conjugate with cysteine and glutathione. Covalent binding of the
reactive intermediate to macromolecules is considered to be responsible for the
hepato- and nephrotoxicity of chloroform.
The initial product of cytochrome P-450 oxidation of cnloroform is
trichloromethanol, which spontaneously dehydrochlorinates to produce phosgene.
Phosgene is thought to be the toxic reactive intermediate produced during the
metabolism of chloroform. Phosgene reacts with water to yield C0_, with protein
to form a covalently bound product, and with cysteine and glutathione. While the
liver is the primary site for chloroform metabolism, other tissues, including the
kidney, can also metabolize chloroform to C0p.
Interspecies comparisons of the magnitude of chloroform metabolism have
been made only for the oral route. Mice, rats, and squirrel monkeys metabolized
85, 66, and "\Q%, respectively, of a 60 mg/kg body weight dose of C-chloroform
1-3
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1 4
to C0_(measured in the expired air). Most of the remainder was exhaled as
unchanged chloroform; small amounts of radioactivity (2 to 8% of the dose) were
excreted in the urine and feces. Human subjects (1 or 2/dose) who ingested 0.1,
0.5, and 1.0 g of 3C-chloroform (=1.4, 7, and 14 mg/kg body weight) metabolized
all of the low dose, 50% of the intermediate dose, and only -35% of the high dose,
judging from the excretion of unchanged chloroform and ^CO through the lungs.
This difference indi^j '.!;._:;; ./'.at the fraction of the dose metabolized is dose-
dependent.
Regardless of the route of entry into the body, chloroform is excreted
unchanged through the lungs and eliminated via metabolism, with the primary
stable metabolite, CO also being excreted through the lungs. High concentra-
tions of unchanged chloroform have been found in the bile of squirrel monkeys
after oral administration, but not in the urine or feces. The inorganic chloride
generated from chloroform metabolism is excreted via the urine.
Decay curves for the pulmonary excretion of unchanged chloroform in humans
appear to consist of three exponential components. The terminal component,
thought to correspond to elimination from adipose tissue, had a half-time of
36 hours. This long half-time, together with the relatively high levels of
chloroform found in adipose tissues of humans known to have been exposed only to
ambient levels of the chemical and of animals exposed experimentally, indicates a
potential for bioaccumulation.
The adverse health effects of exposure to chloroform include neurological,
hepatic, renal, and cardiac effects. These effects have been documented in
humans as well as in experimental animals. In addition, studies with animals
indicate that chloroform is carcinogenic and may be teratogenic.
Evidence of chloroform's effects on humans has been obtained primarily
during the use of this chemical as an inhalation anesthetic. In addition to
depression of the central nervous system, chloroform anesthesia was associated
1-4
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with cardiac arhythmias (and some cases of cardiac arrest), hepatic necrosis and
fatty degeneration, polyuria, albuminuria, and in cases of severe poisoning,
renal tubular necrosis. When used for obstetrical anesthesia, chloroform was
likely to produce respiratory depression in the infant. Experimental exposures
of humans to chloroform have focused only on subjective responses. Humans
exposed experimentally to chloroform for 20 to 30 minutes have reported dizzi-
ness, headache, giddiness, and tiredness at concentrations >1000 ppm, and light
intoxication at concentrations above 4000 ppm.
Similar symptoms occurred in workers employed in the manufacture of
lozenges containing chloroform; exposure concentrations ranged from 20 to
237 ppm, with occasional brief exposure to =1000 ppm. Additional complaints
were of gastrointestinal distress, and frequent and scalding micturition. The
only other report of adverse effects stemming from occupational exposure to
chloroform was of enlargement of the liver; this report was compromised by the
apparent lack of suitable controls.
Acute inhalation experiments with animals revealed that single exposures to
100 ppm were sufficient to produce mild hepatic effects in mice. The exposure
level that would produce mild renal effects is not known, but frank toxic effects
occurred in the kidneys of male mice exposed to 5 rag/I (1025 ppm). In subchronic
inhalation experiments, histological evidence of mild hepato- and nephrotoxicity
occurred in rats with exposures to as low as 25 ppm, 7 hours/day for 6 months.
The effects were reversible if exposure was terminated, and did not occur when
exposure was limited to 4 hours/day.
Information on the effects of acute and long-term oral exposure to chloro-
form is available primarily from experiments with animals. Human data are
mainly in the form of case reports and involve the abuse of medications contain-
ing not only chloroform, but other potentially toxic ingredients as well;
1-5
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however, a fatal dose of as little as 1/3 ounce was reported. As with inhalation
exposure, the primary effects of oral exposure were hepatic and renal damage.
Narcosis also occurred with high doses, but this effect was not usually a focus
of concern in these experiments. Subchronic and chronic toxicity experiments
with rats, mice, and dogs did not clearly establish a no-effect level of exposure
for systemic toxicity. Although a dose level of 1 7 mg/kg/day of chloroform
produced no adverse effect in four strains of mice, the lowest dosage tested,
15 mg/kg/day, elevated some clinical chemistry indices of hepatic damage in dogs
and appeared to affect a component of the reticuloendothelial system (histio-
cytes) in their livers.
No controlled studies have been performed to define dose-response thres-
holds for neurological or cardiac effects of ingested or inhaled chloroform. It
is not known whether subtle impairment of neurological or cardiac function might
occur at levels as low as or lower than those which affect the liver.
Several substances, which are of interest because of accidental or
intentional human exposure, have been shown to modify the systemic toxicity
of chloroform, usually by modifying the metabolism of chloroform to the
reactive intermediate. Examples of substances that potentiate chloroform-
induced toxicity are ethanol, PBBs, ketones and steroids. Factors that
appear to protect against toxicity include disulfiram and high carbohydrate
diets.
Chloroform appears to have teratogenic potential in laboratory animals
when inhaled. Chloroform was selectively more toxic to the fetuses than to
the dams when pregnant rats and mice were exposed to the vapor. Delayed fetal
development occurred at an exposure level (30 ppm) that produced minimal mater-
nal effects. Embryotoxic effects and low but statistically significant inci-
dences of teratogenic effects occurred at an exposure level that produced mild
to moderate maternal effects (i.e., 100 ppm). When chloroform was administered
1-6
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orally (via gavage) to pregnant rats and rabbits, however, toxic effects on
fetal development occurred only at dosage levels that produced severe toxic
effects in the dams; no teratogenic effects were observed at any dosage level
in these animals.
It has been demonstrated that chloroform can be metabolized in vivo and in
vitro to a substance(s) (presumably phosgene) that interacts with protein and
lipid. However, the sole experiment measuring interaction of metabolically
activated chloroform with DNA yielded a negative result. This result was judged
as inconclusive, because the specific activity of the CHC1 may have been too
low.
The majority of the assays for mutagenicity and genotoxicity have also
yielded negative results; however, many of these results are inconclusive
because of various inadequacies in the experimental protocols used. The major
problem is with those bacterial, sister chromatid exchange, and chromosome
aberration studies that used reconstituted exogenous activation systems (i.e.,
S-9 mix). In none of these studies was it shown that chloroform was activated or
metabolized by the activation system used. Metabolism of 2-aminoanthracene or
vinyl compounds (used as positive controls) is probably an inadequate indication
that the activation system can metabolize chloroform because these substances
are not halogenated alkanes and are therefore not metabolized like them. A
better indication that an activation system is sufficient for metabolism of
chloroform may be to show that it metabolizes CHC1_ to intermediates that bind
to macromolecules. A second problem with experimental protocols utilizing exo-
genous activation systems relates to the possiblity that any reactive metabolic
intermediates formed may react with microsoraal or membrane lipid or protein
before reaching the DNA of the test organism. A third potential problem occurs
in those in vitro protocols in which precautions were not taken to prevent escape
of volatilized CHC1-.
1-7
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Studies in which endogenous activation systems were used include those in
yeast, in Drosophila (sex-linked recessive lethal), and in mice (bone-rnarrow
micronucleus, sperm head abnormalities and host-mediated assay). The results
from several of these studies suggest that chloroform may be a weak mutagen.
In summary, with the present data, no definitive conclusions can be reached
concerning the mutagenicity of chloroform. However, there is some indication
(from the binding studies and from the mutagenicity tests that utilized endo-
genous or in vivo metabolism) that chloroform may have the potential to be a weak
mutagen. In order to substantiate this, only certain well-designed in vivo
mutagenicity studies or studies with organisms possessing endogenous activation
systems are recommended.
Chloroform in corn oil administered at maximally and one-half maximally
tolerated doses by gavage for 78 weeks produced a statistically significant
increase in the incidence of hepatocellular carcinomas in male and female
B6C3F1 mice and renal epithelial tumors (malignant and benign) in male
Osborne-Mendel rats; a carcinogenic response of female Osborne-Mendel rats to
chloroform was not apparent in this study.
A statistically significant increase in the incidence of renal tumors
(benign and malignant) was found in another study in male ICI mice treated with
chloroform in either toothpaste or arachis oil by gavage for 80 weeks; however,
treatment with a gavage dose of chloroform in toothpaste for 80 weeks did not
produce a carcinogenic response in female ICI mice and male mice of the CBA,
C57BL, and CF/1 strains. A carcinogenic response was not observed in male and
female Sprague-Dawley rats given chloroform in toothpaste by gavage for
80 weeks, but early mortality was high in control and treatment groups. Gavage
doses of chloroform in toothpaste did not show a carcinogenic effect in male and
female beagle dogs treated for 7 years; however, the treatment period was short
1-8
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in relation to the lifespan of the beagle dog. Results of preliminary
toxicity tests and the carcinogenicity studies indicate that doses of
chloroform in toothpaste given to mice, rats, and dogs in the carcinogen-
icity studies approached maximally tolerated doses. However, doses of
chloroform in toothpaste given to mice and rats were lower than those
given in corn oil.
Hepatomas were found in NLC mice given chloroform twice weekly for
an unspecified period of time and in female strain A mice given chloroform
once every 4 days for a total of 30 doses at a level which produced liver
necrosis; however, small numbers of animals were examined in pathology, the
duration of studies was below the lifetime of the animals, and no control
group of NLC mice was apparent. Although a carcinogenic effect of chloro-
form was not evident in newborn (C57 x DBA2 - Fl) mice given single or
multiple subcutaneous doses during the initial 8 days of life and observed
for their lifetimes, the dose levels used appeared well below a maximum
tolerated dose and the period of treatment after birth was quite short
compared to lifetime treatment. Chloroform was ineffective at maximally
tolerated and lower doses in a pulmonary adenoma bioassay in Strain A
mice. Although an ability of chloroform to promote growth and spread of
Tjewis lung carcinoma, Erlich ascites, and B16 melanoma cells in mice has
been shown, the mechanism by which chloroform produced this effect is
uncertain, and the relevance of this study to the evaluation of the
carcinogenic potential of chloroform is presently not clear.
There are no epidemic logic studies of cancer and chloroform per se.
There appears to be an increased risk of cancer of the bladder, rectum,
and large intestine from chlorinated drinking water and, by inference,
possibly from chloroform, the predominate contaminant.
1-9
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In conclusion, evidence that chloroform has carcinoqenic activity is
on increased incidences of hepatocellular carcinomas in mal e and female B6C3F1
mice, renal epithelial tumors in male Osborne-Mendel rats, kidney tumors in
male TCT mice, arid hepatomas in NLC and female strain A mice. Applying the
International Agency for Research on Cancer (IARC) criteria for animal studies,
this level of evidence would be sufficient for concluding that chloroform is
carcinogenic in experimental animals.
Although there is limited evidence in humans for the carcinogenicity of
chlorinated drinking water, and by inference, possibly for chloroform, the human
evidence for chloroform itself is insufficient. Considering both human and
animal evidence, the overall IARC classification would be Group 2B, meaning
that chloroform is probably carcinogenic in humans.
Four data sets that contain sufficient information are used to estimate
the carcinogenic potency of chloroform. They are: liver tumors in female mice,
Liver tumors in male mice, kidney tumors in male rats, and kidney tumors in
male mice. The carcinogenic potencies, calculated by the linearized multistage
model on the basis of these four data sets, are comparable within an order of
magnitude. The two data sets for liver tumors in female and male mice give
slightly higher potency values than the two data sets for kidney tumors in male
rats and male mice. The geometric mean, g* = 7 x 10-2/(mg/kg/day), of the
1
potencies calculated from liver tumors in male and female mice (the most sensi-
i;ve species), is taken to represent the carcinogenic potency of chloroform.
Rised on this potency, the upper-bound estimate of the cancer risk due to 1
^cj/rn3 of chloroform in air is P = 1 x 10-5. The upper-bound estimate of
the cancer risk due to 1 jug/liter in water if P = 2 x 10-6. These estimates
appear consistent with the limited epidemiclogic data available for humans.
The estimated potency of chloroform is in the fourth quartile of 53 suspect
carcinogens that have been evaluated by the Carcinogen Assessment Group.
1-10
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2. INTRODUCTION
The U.S. Environmental Protection Agency is responsible under the
authority of various laws for the identification, comprehensive assessment,
and as appropriate regulation, of environmental substances which may be of
concern to the public health. For example, under section 112 of the Clean
Air Act the EPA Administrator is directed to establish standards for any
air pollutant (other than those for which national ambient air quality
standards are applicable) which, in his judgment, "causes, or contributes
to, air pollution which may reasonably be anticipated to result in an
increase in mortality or an increase in serious irreversible, or incapacitating
reversible, illness."
Within EPA, the Office of Health and Environmental Assessment is
responsible for providing scientific assessments of health effects for
potentially hazardous air pollutants such as chloroform. These health
assessment documents form the scientific basis for subsequent agency actions,
including the Administrator's judgment as to whether regulations or standards
may be appropriate.
This Health Assessment Document for Chloroform represents a comprehensive
data base that considers all sources of chloroform in the environment, the
likelihood of human exposures and the possible consequences to man and
lower organisms from its absorption. This information is integrated into a
format that can serve as the basis for qualitative and quantitative risk
assessments, while at the same time identifying gaps in our knowledge that
limit present evaluative capabilities. Accordingly, it is expected that
this document may serve the information needs of many government agencies
and private groups that may be involved in decision making and regulatory
activities. (As with all such EPA documents, this preliminary draft is
made available to the scientific community and the general public so that
comments of interested individuals and organizations may be considered, and
the latest scientific evidence incorporated, in the final draft. This
draft will also be reviewed by the Environmental Health Advisory Committee
of EPA's Science Advisory Board at a subsequently announced public meeting).
2-1
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3. BACKGROUND INFORMATION
3.1. INTRODUCTION
This section provides background information supportive of the human health
effects data presented in subsequent sections. It is not intended to be a
comprehensive review of analytical methodology, sources, emissions, air concen-
trations, or environmental transport and fate information. In order to fulfill
this purpose and since the literature concerning chloroform is vast, only a
portion of the available literature was included. Those articles included were
chosen because of their relevance to the topic at hand and because they were
representative of the literature as a whole.
To provide the most complete overview possible, some non-peer reviewed
information has been added, from the Chloroform Materials Balance Draft Report by
Rem et al. (1982). This is an updated version of the original Level I Materials
Balance: Chloroform (Wagner et al., 1980). As described by Wagner et al. (1980):
"A Level I Materials Balance requires the lowest level of effort and
involves a survey of readily available information for constructing the
materials balance. Ordinarily, many assumptions must be made in accounting for
gaps in information; however, all are substantiated to the greatest degree
possible. Where possible, the uncertainties in numerical values are given,
otherwise they are estimated. Data gaps are identified and recommendations are
made for filling them. A Level I Materials Balance relies heavily on the EPA's
Chemical Information Division (CID) as a source of data and references involving
readily available information. Most Level I Materials Balance are completed
within a 3-6 week period; CID literature searches generally require a 2 week
period to complete. Thus, the total time required for completion of a Level I
materials balance ranges from 5-7 weeks."
3-1
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Because a greater level of effort went into the 1980 and 1982 Materials
Balance reports on chloroform than would normally be devoted to background
(non-health effects) information in an EPA health assessment document,
information in this chapter is drawn from these reports. However, because such
information has not been peer-reviewed and includes some major assumptions, it
should not be used to support any regulations or standards regarding risks to
public health.
3.2. PHYSICAL AND CHEMICAL PROPERTIES
Chloroform (CHC1 } (CAS registry number 67-66-3) is a member of a family of
halogenated saturated aliphatic compounds. Synonyms for chloroform include the
following:
Chloroforme (French)
Chloroformio (Italian)
Formyl trichloride
Methane trichloride
Methane, trichloror
Methenyl chloride
Methenyl trichloride
Methyl trichloride
NCI-C02686
Trichloromethaan (Dutch)
Trichloroform
Trichloromethane
Table 3-1 lists important physical properties. Cloroform is a colorless, clear,
dense, volatile liquid with an ethereal non-irritating odor (DeShon, 1979).
Chloroform is nonflammable; however, when hot chloroform vapors are mixed with
alcohol vapors, the mixture burns with a greenish flame. At 25°C and 1
atmosphere, a 1 ppm concentration of chloroform in air is equal to 4.88 mg/m .
Chloroform decomposes with prolonged exposure to sunlight regardless of the
presence of air (DeShon, 1979). It also decomposes in the dark in the presence
3-2
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TABLE 3-1
Physical Properties of Chloroform0
Molecular weight
Melting point (°C)
Boiling Point
Water-chloroform azeotrope (°C)
Specific gravity (25/4°C)
Vapor density (101kPa, 0°C, kg/m3)
Vapor pressure
°C kPa
-30 1.33
-20 2.61
-10 4.63
0 8.13
10 13.40
20 21.28
30 32.80
40 48.85
Solubility in water
°C
0
10
20
30
Log octanol/water partition coefficient
Conversion factors at 25 °C and 1 atm
1 ppm CHC1- in air equals 4.88 mg/nr
1 rag/m CHC1 in air equals 0.205 ppm
119.
-63.
61.
56.
1.
4.
torr
10.0
19.6
34.7
61.0
100.5
159.6
246.0
366.4
g/kg H^O
10.62
8.95
8.22
7.76
1.97
38
2
3
1
48069
36
Source: DeShon, 1979
h,
Hansch and Leo, 1979
'Calculated
3-3
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of air. The principal decomposition products include phosgene, hydrogen
chloride, chlorine, carbon dioxide, and water. Ozone causes chloroform to
decompose rapidly.
Chloroform forms a hydrate in water at 0°C (CHCl^ • 18HJD, CAS registry
number 67922-19-4); the hexagonal crystal decomposes at 1.6°C (DeShon, 1979).
Chloroform is chemically stable to water, having a hydrolysis half-life of 3100
years in neutral water at 25°C (Mabey and Mill, 1978); the half-life of
chloroform in air from hydroxyl radical reactions is 78 days (Hampson, 1980).
The hydrogen atom on chloroform can be removed in the presence of warm
alkali metal hydoxide to form a trichloromethyl anion (DeShon, 1979). This anion
can condense with carbonyl compounds. Both wet and dry chloroform will react
with aluminum, zinc, and iron.
Small amounts of ethanol are used to stabilize chloroform from oxidation
during storage (DeShon, 1979).
3-3. SAMPLING AND ANALYSIS
3.3.1. Chloroform in Air. Chloroform in air can be analyzed by a number of
methods; however, the method of Singh et al. (1980) appears to be substantially
free of artifact problems and completely quantitative. In this method, an air
sample in a stainless steel canister at 32 psig is connected to a preconcen-
tration trap consisting of a 4" x 1/16" ID stainless steel tube containing glass
beads, glass wool, or 3% SE-30 on acid washed 100/120 mesh chromosorb W. The
sampling line and trap, maintained at 90°C, are flushed with air from the
canister; then the trap is immersed in liquid 0- and air is passed through the
trap, the initial and final pressure being noted (usually between 30 and 20 psig)
on a high-precision pressure guage. The ideal gas law can be used to estimate
the volume of air passed through the trap. The contents of the trap are desorbed
onto a chromatography column by backflushing it with an inert gas while holding
-------
the trap at boiling water temperature. An Ascarite trap may be inserted before
the chromatography column to remove water. Suitable columns include 2Q% SP-2100
and 0.1? CW-1 500 on Supelcoport (100/120 mesh, 6' x 1/8" stainless steel) and 20%
DC-200 on Supelcoport (80/100 mesh, 33' x 1/8" Ni). Both columns can be operated
at 45°C with a carrier gas flow of 40 mS,/min on the former column and 25 m£/min on
the latter. An electron capture detector operating at 325°C was found to be
optimum. It should be noted that the above authors found Tenax to be unsuitable
for air analyses because of the presence of artifacts in the spectrum from
oxidation of the Tenax monomer. In addition, when Tenax is used as a sorbent,
safe sampling volumes (i.e., that volume of air which, if sampled over a variety
of circumstances, will not cause significant breakthrough) should be adhered to.
Brown and Purnell (1979) determined the safe volume for chloroform per gram Tenax
to be 9.3 I (flow rate 5-600 mS,/min; CHC1 cone. <250 mg/rn^ temp, up to 20°C) with
a safe desorption temperature of 90°C.
The detection limits of this method were not specified and are dependent on
the volume of air sampled. Analyses as low as 16 ppt have been reported using
this method (Singh et al., 1930).
3-3.2. Chloroform in Water. Chloroform in water can be analyzed by the purge-
and-trap mejhi^od (Method 502.1 ) as recommended by the Environmental Monitoring
and Support Laboratory of the U.S. EPA (1981 a). In this method, an inert gas is
bubbled through 5 ml of water at a rate of 40 m^/minute for 11 minutes, allowing
the purgable organic compounds to partition into the gas. The gas is passed
through a column containing Tenax GC at 22°C, which traps most of the organics
removed from the water. The Tenax column is then heated rapidly to 130°C and
backflushed with helium (20-60 mil/minute, 4 minutes) to desorb the trapped
organics. The effluent of the Tenax column is passed into an analytical gas
chromatography column packed with 1 £ SP-1000 on Carbopack-B (60/80 mesh, 8'x 0.1"
3-5
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I.D.) maintained at 40°C. The column is then temperature programmed starting at
45°C for 3 minutes and increasing at 8°C/minute until 220°C is reached; it is
then held there for 15 minutes or until all compounds have eluted. A halogen-
specific detector (or GC-MS) having a sensitivity of 0.10 |ig/Ji with a relative
standard deviation of <10$ must be used.
3-3-3 Chloroform in Blood. Chloroform in blood can be analyzed by using a
modified purge-and-trap method (Pellizzari et al., 1979). This method involves
diluting an aliquot of whole blood (with anticoagulant) to =50 ml with prepurged,
distilled water. The mixture is placed in a 100 mi, 3~neck round bottom flask
along with a teflon-lined magnetic stirring bar. The necks of the flask are
equipped with a helium inlet, a Tenax trap, and a thermometer. The Tenax trap is
a 10 cm x 1.5 cm i.d. glass tube containing pre-extracted (Soxhlet, methanol, 24
hrs) and conditioned (2?0°C, 30 m£/min helium flow, 20 min) 35/60 mesh Tenax
(=1.6 g, 6 cm). The sample is then heated to 50°C and purged with a helium flow
rate of 25 mJt/min for 90 min. Analysis can be performed as indicated in Section
3.3.2.
3.3.4. Chloroform in Urine. Chloroform in urine can be analyzed by using an
apparatus identical to the one described in Section 3.3.3, using 25 m& of urine
diluted to 50 mil instead of blood,.
3-3.5. Chloroform in Tissue. Chloroform in tissue can be analyzed by using an
apparatus identical to the one described in Section 3.3.3, using 5 g of tissue
diluted to 50 mS, and macerated in an ice bath instead of blood. The purge time is
reduced to 30 minutes.
3.4. EMISSIONS FROM PRODUCTION AND USE
3.4.1. Emissions from Production.
3.4.1.1. DIRECT PRODUCTION — Chloroform is produced commercially in the
United States by two methods, chloririation of methane and chlorination of methyl
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chloride produced from methanol and hydrogen chloride (Wagner et al., 1980;
DeShon, 1979). The chemistry is summarized by the following reactions:
Methane Chlorination
^CHjj + 10C12 ------- »• CH3C1 + CH2C12 + CHC13 + CCl^ + 10HC1
Methanol Hydrochlorination
Catalyst
HC1 + CH OH ----------- »• CH-C1 + H?0
J 280 - 350°C J
3CH-C1 + 6C12 --------- >• CH2C12 + CHC13 + CCl^ + 6HC1
The methanol process accounts for 7^% of capacity, while methane accounts for 26%
of capacity (SRI International, 1983).
In the chlorination of methane, natural gas is directly chlorinated in the
gas phase with chlorine at 485-510°C (Anthony, 1979). The product mixture
contains all chlorinated methanes, which are removed by scrubbing and separated
by fractional distillation.
In the second process, gaseous methanol and HC1 are combined over a hot
catalyst to form methyl chloride (Ahlstrom and Steele, 1979). The methyl
chloride is then chlorinated with chlorine to produce CH Cl?, CHC1-, and CC1.
(DeShon, 1979). The chlorination conditions for both processes can be adjusted
to optimize chloroform production.
United States production is carried out by five manufacturers at seven sites
summarized in Table 3-2.
The annual production of chloroform in the United States has risen from 35
million kg (77 million Ibs) in 1960 (DeShon, 1979) to >18M million kg (405
million Ibs) in 1981 (USITC, 1982) with few declines.
3-7
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TABLE 3-2
Chloroform Producers, Production Sites, and Capacities*
Producer
Diamond
Shamrock
Dow Chemical
Linden Chemicals
and Plastics, Inc.
Stauffer
Chemical Co.
Vulcan
Materials Co.
TOTAL
Production
Site
Belle, WV
Freeport, TX
Plaquemine, LA
Moundsville, WV
Louisville, KY
Geismar, LA
Wichita, KA
Capacity
Millions of kg
(Millions of Ibs)
18(40)
45(100)
45(100)
14(30)
34(75)
27(60)
50(110)
233(515)
Process
Methanol
Methane
Methanol
Methanol
Methanol
Methanol
Methanol and
Methane
•Source: SRI International, 1983.
3-8
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3.4.1.1.1. Chloroform Emissions from the Methane Chlorination Process —
Rem et al . (1982) reported that air emissions could come from process vents, in-
process and product storage tanks, liquid waste streams, secondary emissions
(handling and disposal of process wastes), and fugitive emissions from leaks in
process valves, pumps, compressors and pressure relief valves.
The emission factors they calculated were based on a typical methane
chlorination facility as reported by U.S. EPA (1980a) having a total chloro-
methane capacity of 2 x 10 metric tons (441 x 10 Ibs), operating continuously
(8760 hr/yr), and having a product mix of 20% CH_C1, 45? CH2C12, 25% CHClo, and
CClj.. The emission factor for the uncontrolled recycle methane inert gas
purge vent for the above plant (0.014 kg/metric ton) was calculated from an
hourly CHC1-, emission rate of 0.071 kg/hr reported by Dow Chemical Company for a
46,000 metric ton/yr facility assuming continuous (8760 hr/yr) operation (Beale,
n.d.). The uncontrolled emission factor for the distillation area emergency
inert gas vent (0.032 kg/metric ton) was calculated from an emission factor for
volatile organic compounds (VOC ) of 0.20 kg/metric ton of total chloromethane
production and composition data showing chloroform to be 4? of VOC (U.S. EPA,
1980a). In-process and product storage emissions (0.91 to 0.80 kg/metric ton,
depending on controls) were calculated from emission equations for breathing and
working losses from AP-61 (U.S. EPA, 198lb) assuming tanks to be half-full, have
95% emission controls when present, and a 12°C diurnal temperature variation
(U.S. EPA, 1980b). Rem et al . (1982) calculated the total chloroform emissions
to air to be 70.2 metric tons (155 x 10 Ibs) by multiplying the appropriate
factors by plant capacity use after including secondary emissions (0.21
kg/metric ton) and fugitive emissions (5.5 kg/hr).
Releases of chloroform to water come from scrubbers, neutralizers , and
cooling water (Rem et al . , 1982). Based on a 300 ppm CHC1- content in total
3-9
-------
wastewater discharges averaging 68 il/min and assuming that 90% would volatilize,
Rem et al. (1982) calculated a release factor of 0.023 kg/metric ton. They then
added to this the emission from indirect contact cooling water (100 ppm, 5800 1
cooling water/metric ton CHC1_, 905? evaporation) to calculate a release rate of
o
3.3 metric tons of CHC1~ (7-3 x 10J Ibs) per year to water.
No quantifiable data were available to Rem et al. (1982) regarding release
of chloroform to land from methane chlorination.
3-4.1.1.2. Chlorination Emissions from the Methanol Hydrochlorination-
Methyl Chloride Chlorination Process — Chloroform emissions to air come from
process vents, in-process and product storage tank emissions, and fugitive
emissions from leaks in valves, pumps, compressors, and pressure relief valves
(Rem et al., 1982). Rem et al. (1982) used an uncontrolled emission factor
reported by Vulcan Materials Company for process vents (Hobbs, 1978), and assumed
continuous (8,760 hr/year) operation, and controls sufficient to reduce
emissions 8Q% to obtain the emission factor for controlled process vents (0.003
kg/metric ton for controlled; 0.015 kg/metric ton for uncontrolled). Storage
emissions (0.176 kg/metric ton for controlled, 0.88 kg/metric ton for
uncontrolled) were calculated from Hobbs (1978) and from emission equations for
breathing and working losses from AP-42 (U.S. EPA, 198lb), assuming tanks to be
half-full, have 80% emission reduction controls, and a 12°C diurnal temperature
variation (U.S. EPA, 1980b). Fugitive emission factors (3-32 kg/hr for
uncontrolled; 1.08 kg/hr for controlled) for volatile organic compounds were
used (U.S. EPA, 1980c) along with a control factor of 67-5/6 based on leak
detection and repair (U.S. EPA, 1980b). Emission rates were then calculated to
be 196 metric tons (432 x 103 Ibs) based on plant capacity, capacity use (66%},
the level of emission controls used at each plant, and continuous operation
(8,760 hr/year).
3-10
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Rera et al. (1982) assumed that releases of chloroform to water from methyl
chloride chlorination resulted from contamination of cooling water and from
spent acid and spent caustic streams. For cooling water contamination, they
assumed minor spills and leaks resulted in contamination of 100 mg chloroform/Ji
cooling water, that 5,800 8, of cooling water per metric ton of chloroform
produced was consumed, and that 90/6 of the chloroform evaporated. For spent acid
and spent caustic streams they assumed that 0.04 kg of chloroform are released
for every metric ton of chloromethane produced. They further stated that this
release factor was not considered to be very reliable; however, in the absence of
better data they used this factor to approximate emissions. In addition, they
assumed that 1/3 of the chloromethane production consists of chloroform and that
90% of the released chloroform evaporates. Using these assumptions, Rem et al.
(1982) calculated that 0.070 kg of chloroform is released to water per metric ton
of chloroform produced, or 8.0 metric tons (17.6 x 10 Ibs) of chloroform were
released to water based on 1980 production levels.
Rem et al. (1982) reported that the bottoms from chloroform distillation in
the methyl chloride chlorination process are the feed for carbon tetrachloride
and perchloroethylene production, and that during their production a residue is
formed that contains chloroform. This residue is landfilled or deep-well
injected. This represents the only known release of chloroform from carbon-
tetrachloride/perchloroethylene production.
Rem et al. (1982) assumed that 1.02 kg of residue is produced/metric ton of
chloroform from methyl chloride production (see Wagner et al., 1980). They
further assumed (as did Wagner et al., 1980) that 18.4$ of the residue is
chloroform and that 25% is landfilled. This results in a release factor of 0.047
kg chloroform per metric ton of chloroform produced, or 5.4 metric tons (12 x 10^
Ibs) based on 1980 production.
3-11
-------
3.^.1.1.3- Summary of Direct Production — Direct production emits some 283
metric tons into the environment on an annual basis. Greater than 95% (266.2
metric tons) of this is emitted into the air. Direct production accounts for =3$
of all environmentally released chloroform, and -3.6% of all chloroform released
to air. Table 3-3 summarizes chloroform discharges from direct production.
3.^.1.2. INDIRECT PRODUCTION
3.4.1.2.1. Chloroform Formation During Ethylene Bichloride Production —
Ethylene dichloride (EDC) is produced by two methods, direct chlorination and
oxychlorination, and is used principally for vinyl chloride monomer (VCM)
production. A combination of the two methods is used by most VCM production
facilities in a process known as the balanced process since the HC1 from the
dehydrochlorination of EDC is used to produce more EDC from ethylene, the major
products of the overall reaction being VCM and H_0.
Direct Chlorination
Balanced Process
HC1 + CH -- CHC1
3-12
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TABLE 3-3
*
Chloroform Discharges from Direct Sources
Environment Release (metric tons/year)
Source Air Water Land Total
Methyl Chloride
Chlorination 196 8 5.4 209.4
Methane Chlorination 70.2 3.3 73.5
Total: 266.2 11.3 5.4 282.9
Source: Rem et al. (1982)
3-13
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Chloroform is formed as a byproduct during EDC manufacture. Rem et al.
(1982) estimated chloroform emissions to air from EDC production based on
emission factors developed by the EPA during field studies of domestic EDC
production facilities (see U.S. EPA, 1980d). Domestic production facilities are
listed in Table 3-4. Emission sources (e.g., process vents, fugitive emissions,
storage) for each plant, plant capacity, capacity utilization (56/6), and control
technology were all combined with the emission factors to determine the overall
chloroform emissions at the current level of control. According to their
calculations, chloroform emissions to the atmosphere are =760 metric tons/year
(1,675 x 103 Ibs/year).
Chloroform releases to water may occur during the discharge of wastewater
from the process; however, the amount of chloroform present is unknown.
Chloroform releases to land from EDC production reportedly occur when the
light ends from EDC distillation are landfilled (Rem et al., 1982). An estimated
217 metric tons (478 x 103 Ibs) were landfilled in 1980.
3.4.1.2.2. Chlorination of Drinking Water — Chloroform in drinking water
arises when humic substances or methyl ketones (e.g., acetone) in water react
with a hypochlorite anion (Stevens et al., 1976; NAS, 1978). Hypochlorite is the
principal reactant in chlorinated water above pH 5. Chloroform is produced by
the haloform reaction outlined below:
R-COCH + 30C1~ >• RCOCC1 + 30H~
R-COCC1 + OH~ >• RCOO~ + CHC1
Rem et al. (1982) used the data from the National Organics Reconnaissance
Survey (NORS) (Symons et al., 1975) and the National Organics Monitoring Survey
(NOMS) (U.S. EPA, n.d.) to estimate the concentration of chloroform in drinking
3-14
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TABLE 3-4
Ethylene Bichloride Producers, Production Sites and Capacities'
Producer
Production
Site
Capacity
millions of Kg
(millions of Ibs)
Alantic Richfield Co.
ARCO Chem. Co., div.
Borden Inc.
Borden Chem. Div.
Petrochems. Div.
Dow Chemical. U.S.A.
E.I. du Pont de Nemours and Co.,
Conoco Inc., subsid.
Conoco Chems. Co. Div.
Ethyl Corp.
Chems. Group
Formosa Plastics Corp. U.S.A.
Georgia-Pacific Corp.
Chern. Div.
The BF Goodrich Co.
BF Goodrich Chem. Group
Convent Chem. Corp., subsid,
PPG Iridust., Inc.
Chems. Group
Chera. Div.
Port Arthur, TX 204 (450)
Geismar, LA 231 (510)
Freeport, XX 726 (1600)
Oyster Creek, TX 476 (1050)
Plaquemine, LA 862 (1900)
Inc.
318
102
249
386
(700)
(225)
(550)
(850)
Lake Charles, LA 524 (1155)
Baton Rouge, LA
Pasadena, TX
Baton Rouge, LA
Point Comfort, TX
Plaquemine, LA 748 (1650)
Deer Park, TX 145 (320)
La Porte, TX 719 (1585)
Calvert City, KY 454 (1000)
Convent, LA 363 (800)
Lake Charles, LA 1225 (2700)
3-15
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TABLE 3-4 (cont.)
Producer
Production
Site
Capacity
millions of Kg
(millions of Ibs)
Shell Chem. Co.
Union Carbide Corp.
Ethylene Oxide Derivatives Div.
Vulcan Materials Co.
Vulcan Chems., div.
Deer Park, TX
Norco, LA
Taft, LA
Texas City, TX
Geisraar, LA
635
544
68'
68
159
(1400)
(1200)
(150)
(150)
(350)
TOTAL 9,206 20,295
Source: SRI International, 1983
Captive use only
3-16
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water. These surveys provided information on chloroform concentrations in 137
U.S. cities. To determine the amount of chloroform generated, the authors
multiplied the volume of water treated by each city by the chloroform concen-
tration in the drinking water. The amount of chloroform generated by each city
was then summed and divided by the volume of water generated to give a weighted
average concentration of 41 ^g/ii. This was then multiplied by the estimated
1 3
volume of water chlorinated annually (4.6 x 10 i/year) to yield the amount in
the U.S. (1,900 metric tons, 4.2x10 Ibs). Rem et al . ( 1982) stressed that this
value probably represents a minimum since NORS was conducted during the winter
(hence, chloroform levels were low) and NOMS samples were iced when taken (hence,
may be lower than if allowed full contact time). Thus, the actual value may be
higher than this estimate.
3.4.1.2.3- Chlorination of Municipal Sewage — Chlorination of municipal
sewage results in increased chloroform concentrations (NAS, 1978). Municipal
wastewater generally contains a lower concentration of chloroform precursors
(humic materials) than do ambient waters; therefore, the amount of chloroform
generated from wastewater Chlorination is smaller (wagner et al . , 1980). Rem et
al . (1982) calculated chloroform production from wastewater treatment based on
analyses of the secondary effluent from 28 municipal plants published by the EPA
(U.S. EPA, 1979a). These analyses showed that the average chloroform concen-
tration increased by 9 >>g/&, from 5 to 14 ^g/£. Rem et al . (1982) assumed that
all municipal wastewater was chlorinated and multiplied the average concen-
tration increase by the municipal wastewater flow (9.7 x 10 £/day) listed in
U.S. EPA (198lc). By this method, 320 metric tons (0.7 x 10 ibs) was calculated
to be produced annually.
3-4.1.2.4. Chlorination of Cooling Waters — Cooling water used in electric
power generating plants is treated with chlorine as a biocide to prevent fouling
3-17
-------
intake screens and condensers in both once-through and closed cycle systems (U.S.
EPA, 1980e). Bern et al. (1982) calculated chloroform production based on the
size of the average power plant (hence, the volume of water required), the type
of cooling system used (once-through or recirculating), and the fact that 65% of
all power plants chlorinate cooling water. They calculate that 72 metric tons of
o
chloroform (160 x 10 Ibs) are discharged directly into water by once-through
systems and !90 metric tons of chloroform (420 x 10 Ibs) are emitted into the
air by recirculating systems. A total of 262 metric tons of chloroform (580 x
10 Ibs) are produced annually from cooling water chlorination.
3.4.1.2.5. Chlorination in the Pulp and Paper Industry — Pulp and paper
mills emit more chloroform to the environment than any other single source.
Chloroform is produced during the bleaching of wood pulp, a process that whitens
the final paper product. Rem et al. (1982) based their estimation on information
contained in a number of documents concerned with the pulp and paper industry
(U.S. SPA, 1980f; NCASI, 1977; Metcalf and Eddy Inc., 1972; TAPPI, 1963). From
the opeBating conditions, analytical data, and production steps detailed in
these documents, Rem et al. (1982) determined the quantity of chloroform produced
for each of nine different types of mills for which monitoring data existed and
applied these valuers to all mills for which no values from sampling existed. The
authors determined that chloroform is emitted at three different stages: into the
air- during the bleaching process itself, into the air during the detention time
in wastewacer treatment plants, and into the water from treatment plant effluent.
The amount, of chloroform produced annually was calculated to be 128 metric tons
(282 x 10 Ibs) released to air during bleaching operations, 3,985 metric tons
(8.78 x 10 Ibs) released to air from wastewater detention (4,113 metric tons to
o
air), and 298 metric tons (657 x 10 Ibs) discharged into water from treatment
plants (4,411 metric tons or 9.72 x 10 Ibs total).
3-18
-------
3.4.1.2.6. Chloroform Production from Combustion of Leaded Gasoline —
Chloroforom has been reported to be a component of automobile exhaust (Harsch et
al., 1977). Its presence is reportedly the result of using ethylene dichloride
and ethylene dibromide as lead scavengers in leaded gasoline (Lowenbach and
Schlesinger Associates, 1979). Rem et al. (1982) cite other sources which state
that chloroform is not formed during the combustion of leaded fuel containing
ethylene dichloride. The Emissions Testing and Characterization Branch of EPA's
Environmental Sciences Research Laboratory measured chlorocarbon emissions from
automobiles using leaded gasoline and found no chloroform. Rem et al. (1982)
then reason that even if chloroform were present in automobile exhaust, the
decrease in the usage of leaded gasoline will decrease the amount of chloroform
produced.
The authors cite an EPA report (U.S. EPA, 1982a) which states that leaded
gasoline consumption is expected to drop by >75%, from 34 x 10 gallons to 8.3 x
109 gallons/year. Rem et al. (1982) used the estimate of Wagner et al. (1980)
that W> of the ethylene dichloride in gasoline would be converted to chloroform,
and the new lead phase-down regulations (U.S. EPA, 1982b) to calculate an annual
emission rate of 180 metric tons (397 x 103 Ibs) in 1983 and 44 metric tons (97 x
103 Ibs) by 1990.
3.4.1.2.7. Chloroform Formation During Atmospheric Trichloroethylene
Decomposition — Trichloroethylene is a major industrial solvent used
principally for vapor degreasing of fabricated metal parts (66/O (Chemical
Marketing Reporter, 1981), and the majority of each year's production is used for
replacement of evaporative loss to the environment (Rem et al., 1982). The
postulated formation of chloroform during the atmospheric decomposition of
trichloroethylene is based on laboratory experiments in which trichloroethylene,
N02, H20, and a hydrocarbon mixture were irradiated with light having the
3-19
-------
intensity and spectral distribution of the lower troposphere (U.S. EPA, 1976).
Dichloroacetylchloride, phosgene, chloroform, and HC1 were detected as products.
Rem et al. (1982) did not describe the method they used to determine the amount
of chloroform produced from this reaction; however, they estimated 780 metric
tons of chloroform are produced annually.
3.4.1.2.8. Miscellaneous Indirect Source — Wagner et al. (1980) have
listed a number of other indirect sources of chloroform that are difficult if not
impossible to quantify. These sources are: chlorination wastewaters from the
textile industry, the food processing industry, breweries, combustion of tobacco
products treated with chlorination pesticides, thermal decomposition of
plastics, biological production in red marine algae, and the reaction of
chlorinated pollutants with humic substances in natural waters.
3.4.1.2.9. Summary of Indirect Production — Chloroform is produced and
emitted into the environment from a variety of indirect sources. These sources
account for =£4£ of all chloroform air emissions, 98? of water discharges, and
36$ of all land discharges (84.6$ of all environmental releases).
The environmental discharges of chloroform from indirect sources are
summarized in Table 3-5.
3.4.2. Emissions from Use. Chloroform is used consumptively for the production
of chlorodifluoromethane or Fluorocarbon-22 (accounts for 9056 of domestic 1982
production, 60% for refrigerants, 30% for fluoropolymer) and for exports (3% in
1982) (Anonymous, 1983). Non-consumptive uses include its use as an extraction
solvent; as a solvent for penicillin, alkaloids, vitamins, flavors, lacquers,
floor polishes, artificial silk manufacture, resins, fats, greases, gums, waxes,
adhesives, oils, and rubber; as a dry cleaning agent; as an intermediate in
pesticide and dye manufacture; and as a fumigant ingredient (Rem et al., 1982;
DeShon, 1979; Merck Index, 1976). The great majority of chloroform used non-
3-20
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TABLE 3-5
o
Chloroform Discharges from Indirect Sources
Environmental Release (metric
Source
Pulp and Paper Mills
Drinking Water Chlorination
Ethylene Dichloride
Manufacture
Trichloroethylene
Photodegradation
Municipal Wastewater
Chlorination
Cooling Water Chlorination
Automobile Exhaust
TOTAL
Air
4113
0
760
780
0
190
180
6023
Water
298
1900
b
0
320
72
0
2590
Land
0
0
217
0
0
0
0
217
tons/ year)
Total
4411
1900
977
780
320
262
180
8830
aSource: Rem et al., 1982
minor releases possible
3-21
-------
consumptively is emitted into the environment since (except for expansions) the
chloroform purchased for these uses is make-up solvent used to replace that
amount not recovered from processes (Wagner et al., 1980).
3.4.2.1. EMISSIONS FROM PHARMACEUTICAL MANUFACTURING — Chloroform is used
as an extraction solvent during the manufacture of some antibiotics and ster|ijo\ds,
and during the manufacture of certain other biological and natural pharraa-
ceuticals (Rem et al., 1982). It is also used as a chemical intermediate. Based
on a Pharmaceutical Manufacturing Association (PMA) Survey, Rem et al. (1982)
reported that =1000 metric tons of chloroform are released into the environment
(no year was specified; the PMA report was dated 1978). The distribution was as
follows: 57.0? (570 metric tons) to air, 4.6? (46 metric tons) to water, and
38.4? (384 metric tons) to land.
3.4.2.2. EMISSIONS FROM FLUOROCARBON-22 PRODUCTION -- The single largest
use of chloroform is for Fluorocarbon-22 production (Chlorodifluoromethane,
CHC1F ). Fluorocarbon-22 producers and production sites are listed in Table 3-6.
Chloroform release to the environment can occur from process emissions, fugitive
emissions, and storage emissions (Rem et al., 1982). Rem et al. (1982) reported
that the first source listed does not represent a significant source of
chloroform emissions based on the design of Fluorocarbon-22 production
facilities and the process description.
Storage emission estimates were based on U.S. EPA (1980g) for chloroform
feedstock storage in fixed roof tanks. Rem et al. (1982) reported that the
Allied Chemical facility at El Segundo, California, uses a control system that
results in complete capture of chloroform vapors. Du Pont (and all others by
assumption) use refrigerated condensers that reduce the uncontrolled emission
factor of 2.5 kg/metric ton by 66?. Fugitive emissions from leaks in valves,
pumps, compressors, and relief valves was estimated to result in an uncontrolled
3-22
-------
TABLE 3-6
Chlorodifluoromethane Producers and Production Sites
Producer
Production
Site
Allied Corp.
Allied Chem. Co.
E.I. du Pont de Nemours and Co., Inc.
Petro^hems. Dept.
Freon Products Div.
Pennwalt Corp.
Chems. Group
Fluorochemicals Div.
Baton Rouge, LA
Danville, IL
El Segundo, CA
Deepwater, NJ
Louisville, KY
Calvert City, KY
Source: SRI International, 1983
3-23
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emission rate of 0.75 kg/metric ton. Total emissions to air were calculated by
Rem et al. (1982) to be 139 metric tons/year by multiplying the emission factors
by 1980 Fluorocarbon-22 production (97,500 metric tons).
No estimate was made for emissions to wastewater because of a lack of data.
Emissions to land were based on the reported practice of landfilling spent
catalyst by Allied. Rem et al. (1982) assumed a catalyst contamination level of
1056 and a total emission of 2.0 metric tons/year.
3.4.2.3- EMISSIONS FROM HYPALON* MANUFACTURE — Hypalon* is a chemically
resistant synthetic rubber made by substituting chloride and sulfonyl groups
onto polyethylene. The process involves dissolving polyethylene in chloroform
followed by reaction with chorine and sulfur dioxide. Based on a Du Pont report,
Rem et al. (1982) estimated that 54.9 metric tons of chloroform were emitted into
•
the air from Hypalon manufacture in 1980, based on an emission inventory
conducted by the Texas Air Control Board, Austin, Texas. No information was
available for water or land emissions.
3.^.2.4. CHLOROFORM EMISSIONS FROM GRAIN FUMIGATION -- Chloroform is a
registered pesticide for use on certain insects that commonly infest stored raw
bulk grains and is present in only one product (Rem et al., 1982). This product,
«
Chlorofume FC.30 Grain Fumigant (Reg. No. 5382-15), marketed by Vulcan
Materials Company, contains 12.2% chloroform, 20.4? carbon disulfide, and 7.4?
ethylene dibromide. Originally registered in 1968, the EPA issued a "Notice of
Presumption Against Continued Registration of a Pesticide Product — Chloroform
(Trichloromethane)" in 1976 because of oncogenic effects in rats and mice.
Continued study resulted in returning it to the normal registration process (U.S.
EPA, 1982c). Based on a personal communication with D. Lindsay of Vulcan
Materials, Rem et al. (1980) estimated that between 10,000 and 12,000 gallons of
chloroform/year were used in grain fumigants in the United States. Vulcan
3-24
-------
reported 1981 sales of 7,000 gallons of Chlorofume in 1981 or 5054 gallons
(19,131 &). Based on its density, 28,400 kg (28.4 metric tons) was released to
the environment (air) in this way.
3-4.2.5. CHLOROFORM LOSSES FROM LOADING AND TRANSPORTATION — Rera et al.
(1982) estimated chloroform losses from loading ships, barges, tank cars, and
tank trucks. The method was based on the degree of chloroform saturation of the
air expelled from tanks during filling, temperature, vapor pressure, control
efficiency, and filling methods as described by U.S. EPA (1979b) and Environment
Reporter (1982). The U.S. mode of transportation was taken from Sax (1981) as
follows: rail, 40.3?; barge, 47.856; and truck, 11.9?. Loading losses were
calculated to be 40.9 metric tons (90,200 Ibs).
Transit losses result from temperature and barometric pressure changes.
The losses were assumed to be the same for barges, tank trucks, and rail cars and
were estimated from the following equation:
LT= 0.1 PW
o
where LT= transit loss, Ib/week - 10J gal transported
P= true vapor pressure of transported liquid, psia
W= density of condensed vapors Ib/gal
No reference or justification for the use of the formula was presented. By this
method, and assuming 1 week transit time, Rem et al. (1982) calculated that 49-2
metric tons (0.11 x 10 Ibs) chloroform were lost to the air using 1980
production values.
3.4.2.6. MISCELLANEOUS USE EMISSIONS — Rem et al. (1982) cited the previous
materials balance (Wagner et al., 1980) in predicting the emissions from
chloroform contamination of methyl chloride, methylene chloride, and carbon
tetrachloride. Chloroform is present to some extent in these products since they
3-25
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are all made by the same process. Assuming a contamination level of 7.5 ppn,
17.5 ppm and 150 ppm for methyl chloride, methylene chloride, and carbon
tetrachloride, respectively, Rem et al. (1982) estimate that releases to air,
land and water would be 9.8, 0.6, and 0.15 metric tons, respectively.
Chloroform is also used in a variety of products (see Section 3-4.2) and as
a general solvent. Rem et al. (1982) estimate that, while these uses are
generally declining, laboratory uses in particular may account for 8.5? of
production or 14,200 metric tons of chloroform. Rem et al. (1982), however,
estimated the range of uncertainty to be + 50%.
3.4.2.7. SUMMARY OF CHLOROFORM DISCHARGES FROM USE — Chloroform
discharges from manufacturing facilities that use chloroform as a process
ingredient account for 12% of all chloroform emissions to air, 1.8/t of water
discharges, and 63/6 of all land discharges, or 12.7% of chloroform discharges to
all media. Table 3-7 summarizes chloroform discharges to all media.
3.4.3. Summary. Chloroform is produced by direct and indirect processes.
Direct production accounts for 184 million kg annually, while indirect
production accounts for =8.8 million kg annually.
Direct production of chloroform and processes associated with its use
«
(i.e., Fluorocarbon-22 production, Hypalon manufacture, loading and transit
losses, grain fumigation, pharmaceutical use) emit some 1.6 million kg to the
environment. Virtually all of the indirectly produced chloroform is emitted into
the environment; the total amount of chloroform emitted is =10.4 million kg.
This represents -5.6% of direct production. The relative source contributions
from all quantifiable sources are listed in Table 3-8.
3.5. AMBIENT AIR CONCENTRATIONS
Monitoring data for a number of U.S. and world locations are presented in
Table 3-9. For the most part, ambient concentrations remain <1000 ppt (1 ppt =
3-26
-------
TABLE 3-7
Chloroform Discharges from Use
Environmental Release (metric
Source
Pharmaceuticals
Chlorodifluoromethane
Manufacture
Loading and Transit Losses
-------
TABLE 3-8
Relative Source Contribution for Chloroform
LO
I
Environmental Release (metric
Source
Pulp and Paper Mills
Drinking water Chlorination
Pharmaceuticals
Ethylene Bichloride Manufacuture
Trichloroethylene Photodegradation
Municipal wastewater Chlorination
Cooling water Chlorination
Methyl Chloride Chlorination
Automobile exhaust
Chlorodifluoromethane Manufacture
Loading and Transit Losses
Methane Chlorination
•
Hypalon Manufacture
Grain Fumigation
Secondary Product Contamination
Laboratory Usage
TOTAL
Air
4113
0
570
760
780
0
190
196
180
139
90.1
70.2
54.9
28.4
9.8
7181.4
Jof
Total0
39
0
5.5
7.3
7.5
0
1.8
1.9
1.7
1.3
0.9
0.7
0.5
0.3
0.1
68.8
Water
298
1900
46
b
0
320
72
8
0
0
3.3
0
0.6
2647.9
Jof
Total
2.9
18
0.4
0
3.1
0.7
0.1
0
0
0.03
0
0.006
25.4
Land
0
0
384
217
0
0
0
5.4
0
2
0
0
0.2
608.6
Jof
Total
0
0
3.7
2.1
0
0
0
0.1
0
0.02
0
0
0.002
5.8
Total
4411
1900
1000
977
780
320
262
209.4
180
141
90.1
73.5
54.9
28.4
10.6
10,438°
tons/ year )
*« c
Total0
42
18
10
9.4
7.5
3.1
2.5
2.0
1.7
1.4
0.9
0.7
0.5
0.3
0.1
Source: Rem et al. (1982)
dashed lines indicate minor releases possible
c
values are rounded
not included because of uncertainty
-------
mj
I
TABLE 3-9
Ambient Levels of Chloroform
Location
Alabama
Tuscaloosa
Talladega Forest
Arizona
Phoenix
California
Stanford Hills
Point Reyes
Los Angeles
Palm Springs
Yosemite
Mill Valley
Riverside
Badger Pass
Point Arena
Point Arena
Los Angeles
Oakland
Delaware
Delaware City
Kansas
Jetmore
Maryland
Baltimore
Montana
Western Montana
Type of Site
urban
rural
urban
clean
clean marine
urban
urban-suburban
remote-high altitude
clean marine
urban- suburban
remote-high altitude
clean marine
clean marine
urban
urban
NS
remote-continental
urban
remote
Date
2/77
2/77
4-5/79
11/75
12/75
1-5/76
5/76
5/76
1/77
4-5/77
5/77
5/77
8-9/78
4/79
6-7/79
7/74
6/78
7/74
3/76
Analytical
Method
GC-ECD
GC-ECD
GC-coulometry
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-MS
Concentration (ppt, v/v)
Max.
3000
200
514.0
217
114
724
616
24
36
310
38
42
48
223.5
60.1
<10
34
<10
NR
Min.
100
NR
27.1
12
15
23
20
12
4
24
2
8
12
24.3
13.1
<10
9
<10
NR
Average
800
100
111.4
33
37
102
99
17
25
25
16
20
18
88.2
32.1
NR
16
NR
9
Holzer
Holzer
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Singh
Reference
et al
et al
et
et
et
et
et
et
et
et
et
et
et
et
et
Lillian
Singh
et
Lillian
Cronn
, 1977
, 1977
al., 1981
al.
al.
al.
al.
al.
al.
al.
al.
al.
al.
al.
al.
1979
1979
1979
1979
1979
1979
1979
1979
1979
1979
1981
1981
et al., 1975a
al.
, 1979
et al., 1975a
and Harsch, 1979
1979
Nebraska
Reese River
New Jersey
Rutherford
Newark
remote-high altitude
urban
urban
5/77
1978
1978
GC-ECD
GC-MS
GC-MS
19
31,000
7500
6
NR
NR
13
4600
3900
Singh et al., 1979
Bozzelli
and Kebbekus, 1979
Bozzelli and
Kebbekus, 1979
-------
TABLE 3-9 (cent.)
I
uo
o
Location Type of Site
Piscataway
Somerset (county)
Bridgewater
Township
Bound Brook
Patterson
Clifton
Fords
Newark
Passaic
Hoboken
Seagrlt
Seagrit
Sandy Hook
Sandy Hook
Bayonne
New York
Staten Island
New York City
New York City
White Face Mountain
White Face Mountain
Niagara Falls
Ohio
Wilmington Air
Wilmington Air
Texas
Houston
urban
urban
rural
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
urban
remote
remote
urban
Force Base
Force Base
urban
Date
1978
1978
1978
3/76
3/76
3/76
3/76
3/76
3/76
3/76
6/74
6/75
7/71
7/75
7/75
3/76
6/71
6/75
9/74
9/75
NS
7/71
7/75
6-7/77
Analytical
Method
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-MS
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-MS
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC-MS
GC-ECD
GC-ECD
GC-MS
Concentration (ppt,
Max.
2900
11,000
NR
NR
NR
NR
NR
NR
NR
NR
60
50
63
55
15,000
NR
480
150
250
350
21,611
4800
5000
11,034
Min.
NR
NR
NR
NR
NR
NR
NR
NR
NR
NR
<10
5
<10
10
<10
NR
<10
10
<10
6
215
<10
20
Trace
v/v)
Reference
Average
2200
5000
NDb
854
768
1700
3422
7582
854
427
40
35
30
25
1030
4268
160
200
9
8
NR
340
480
NR
Bozzelli and
Kebbekua, 1979
Bozzelli and
Kebbekus, 1979
Bozzelli and
Kebbekus, 1979
Pellizzari, 1977
Pellizzari, 1977
Pellizzari, 1977
Pellizzari, 1977
Pellizzari, 1977
Pellizzari, 1977
Pellizzari, 1977
Lillian et al., 1975a
Lillian et al., 1975b
Lillian et al., 1975a
Lillian et al., 1975b
Lillian et al., 1975a
Pellizzari, 1977
Lillian et al., 1975a
Lillian et al., 1975b
Lillian et al., 1975a
Lillian et al., 1975b
Pellizzari et al., 1979
Lillian et al., 1975a
Lillian et al., 1975a
Pellizzari et al., 1979
-------
TABLE 3-9 (oont.)
Location Type of Site
Washington
Pullman rural
Pullman rural
England
Liverpool/Manchester suburban
Organochlorine urban
manufacturer
Moel Faman, urban
Flintshire
Rannoch Moor, urban
Argyllshire
Rural areas rural
Ireland
Cork urban
Japan
Kobe NS
Atlantic Ocean
Northeast Atlantic
(Cape Blanc to
Lands End)
31°19'N 13°32'W to
19°51'N 05°51'W
Date
12/71 to 2/75
11/75
NS
NS
NS
NS
NS
1971
NS
7-8/72
Analytical
Method
GC-MS
GC-ECD
GC-ECD
GC-ECD
GC-ECD
GC
GC
GC-ECD
GC-ECD
GC
Concentration (jpj>tj v/v) Reference
Max. Min. Average
NR NR 20 Grlmsrud and
Rasmussen, 1975
NR NR 13 Rasmussen et al.,1977
C G
8 3 NR Pearson and McConnell,
1975
MO <10.1 NR Pearson and McConnell,
1975
0.1 <0.1 NR Pearson and McConnell,
1975
0.5 0.1 NR Murray and Riley, 1973
1.2 0.82 0.82 Murray and Riley, 1973
NR NR 26.5 Cox et al., 1976
9100 300 Okuno et al., 1971
0.96 0.11 0.35 Murray and Riley, 1973
Subject Urban Transport
Detection Limits 10 ppt
ppb by mass
NR = Not reported; NS = Not specified
ND = Not detected
-------
— 12
10 , v/v), and .some <10 ppt. There are notable exceptions, however, although
the reasons for this are not readily apparent.
Singh (1977) and Singh et al. (1979) have determined northern and southern
hemisphere background concentrations as well as a global average. The
hemispheric values are 14 ppt for the northern hemisphere and <3 ppt for the
southern hemisphere. This difference in hemispheric values suggests that the
oceans are not a significant source of chloroform, but rather, that chloroform,
for the most part, is anthropogenic. The global average concentration determined
by Singh et al. (1979) is 8 ppt.
An interesting point not presented in Table 3-9 is that chloroform concen-
trations above an inversion layer are significantly lower than concentrations
below it. In Wilmington, OH, above an inversion layer, the chloroform concen-
tration was <10 ppt, whereas below it the concentration was 120 ppt (Lillian et
al., 1975a).
3.6. ATMOSPHERIC REACTIVITY
The principal atmospheric reactant responsible for the removal of
chloroform is probably the hydroxyl radical (Atkinson et al., 1979; Graedel ,
1978; Altshuller, 1980; Singh, 1977; Crutzen and Fishman, 1977). Hydroxyl
radicals are formed in the lower atmosphere in two ways, first, by the photo-
dissociation of ozone (X <310) into 0 (1D) atoms (Atkinson et al . , 1979). These
go on to react with either water, hydrogen or methane to form hydroxyl radicals.
The second important source of hydroxyl radicals is the reaction of hydroperoxyl
radicals with nitric oxide.
Hydroxyl radical reactions probably follow the course outlined below
(Graedel, 1978):
CHC13 + HO ---------- * »CC13 + H20
cci3
3-32
-------
NO
•OCC1 --------------- v COC1 (phosgene) + Cl«
COC12 + H20 --------- >• C02 + 2 HC1
Pearson and McConnell (1975) found HC1 and C02 as the only products of chloroform
irradiation with UV (X >290nm) light. The half-life reported by these workers
(23 weeks) was of the same order of magnitude as that calculated from the
hydroxyl radical rate constant (11.5 weeks) (Singh et al., 1981).
Chloroform will not react photolytically in the trospos phere; the UV cutoff
for chloroform is 175 nm (i.e., it will not absorb light >175 nrn). Callahan et
al. (1979) calculated that roughly 1% of the tropospheric chloroform would
diffuse eventually into the stratosphere, based on a lifetime of 0.2-0.3 years
and a troposphere-to-stratosphere turnover time of 30 years.
3.7. ECOLOGICAL EFFECTS/ENVIRONMENTAL PERSISTENCE
3.7.1. Ecological Effects.
3.7.1.1. TERRESTRIAL -- Data on the terrestrial ecological effects of
chloroform are not available. Significant effects are not expected because
chloroform is quite volatile and does not accumulate in terrestrial (or aquatic)
environments, and is diluted rapidly and degraded to low concentrations in the
troposphere (NAS, 1978). Conceivably, acute effects on wildlife can occur in the
vicinity of major chloroform spills, but significant chronic effects from long-
term exposure to low ambient levels is unlikely.
3.7.1.2. AQUATIC — The toxicity of chloroform to aquatic organisms has
been reviewed by the U.S. EPA (1980h). As summarized in Table 3-10, two
freshwater fish (rainbow trout, bluegill) and one invertebrate (Daphnia magna)
species have been acutely tested under standard conditions; LC^ concentrations
ranged from 28,900 to 115,000 [ig/Jl (Bentley et al., 1975; U.S. EPA, 1978a), and
3-33
-------
TABLE 3-10
Acute and Chronic Effects of Chloroform on Aquatic Organisms'
Species
Cladoceran
Daphnia magna
Rainbow trout
Salmo gairdneri
Rainbow trout
Salmo gairdneri
OJ
i Bluegill
-(=• Lepomis macrochirus
Bluegill
Lepomis macrochirus
Pink shrimp
Penaeus duorarum
Orangespotted sunfish
Lepomis humilis
Goldfish
Carassius auratus
Duration
48 -hr
96-hr
96 -hr
96 -hr
96 -hr
96 -hr
1-hr
30 to 60 rain
Concentration
(jj.g/2,) Method
28,900 S,Ub
66,800 S,U
43,800 S,U
115,000 S,U
100,000 S,U
81,500 S,U
106,890 to NS
152,700
97,000 to NS
167,000
Effect
LC50
LC50
LC50
LC50
LC50
LC50
death
5056
anesthetized
Reference
U.S. EPA, 1978a
Bentley et al . ,
Bentley et al . ,
Bentley et al. ,
Bentley et al . ,
Bentley et al . ,
Clayberg, 1917
Gherkin and
Catchpool, 1964
1975
1975
1975
1975
1975
Threespine stickleback 90-min
Gasterosteus aculeatus
207,648
NS
anesthesia Jones, 1947a
with recovery
-------
TABLE 3-10 (cont.)
Species
Ninespine stickleback
Pungitius pungitius
Duration
NS
Concentration
(|ig/O
148,320 to
296,640
Method
NS
Effect
Avoidance
Reference
Jones, 1947b
Rainbow trout
(embryo-larval)
Salmo gairdneri
Rainbow trout
(embryo-larval)
Salmo gairdneri
27 days
27 days
2,030
1,240
F,M
F,M
LC
g/Jl H
50 mg/Jl ardness
LC
g/i, H
Birge et al., 1979
Birge et al., 1979
200 mg/i, ardness
UJ
uo
Rainbow trout 23 days
(embryo)
Salmo gairdneri
10,600 F,Me 4056
teratogenesis
Birge et al., 1979
Source: U.S. EPA, 1980
Static test, unmeasured concentration
Saltwater species
H
Corrected from vol/vol to p.g/£
eplow-through test, measured concentration
Exposures began within 20 minutes of fertilization and ended 8 days after hatching.
hr = hour; rain = minutes; NS = Not stated
-------
the trout was more sensitive than the bluegill. With stickleback, goldfish, and
oranges potted sunfish, anesthetization or death occurred after exposure to
97,000 to 207,000 \j.g/H chloroform for 30 to 90 minutes (Clayberg, 1917; Jones,
1947a; Gherkin and Catchpool, 1964). Only one test has been conducted with
chloroform and saltwater organisms; the 96~hour LC,-n for the pink shrimp was
81,500 \ig/H (Bentley et al., 1975).
Embryo-larval tests with rainbow trout at 2 levels of hardness provided 27-
day LC(-n values of 2030 and 1240 jig/S, (Birge et al., 1979). There was a 40$
incidence of teratogenesis in the embryos at hatching (23 day exposure at 10,600
Bluegills bioconcentrated radiolabeled chloroform by a factor of 6 after a
14-day exposure, and the tissue half-life was <1 day (U.S. EPA, 1978a). This
degree of bioconcentration and short biological half-life suggest that
chloroform residues would not be an environmental hazard to consumers of aquatic
life (U.S. EPA, 1980h).
3.7.2. Environmental Persistence. A number of researchers have reported the
dominance of hydroxyl radical oxidations in the fate of chloroform in the
atmosphere (see Section 3-6). Singh et al. ( 198 1 ) calculated an atmospheric
residence time for chloroform based on the NASA reviewed rate constant reported
by Hampson (1980). They reported a 116-day (16.6-week) residence time for a
hydroxyl radical concentration of 10 molecules/ cm . This compares well to the
observed 33-week lifetime of chloroform in a sunlit flask (Pearson and McConnell,
1975). This lifetime was based on experiments conducted in northwest England,
which receives less intense sunlight than most of the U.S., and may account for
its longevity.
According to recent hydroxyl radical measurements, tropospheric ambient air
£ r-i
concentrations range from =10 to 10 molecules/mi, (Atkinson et al., 1979);
3-36
-------
models of the troposphere have suggested a concentration ranging between 2 to 6 x
105 molecules/mS, (Crutzen and Fishman, 1977; Singh, 1977).
Table 3-11 summarizes the literature values for kQH> the temperature of
measurement, and the calculated lifetime based on the indicated hydroxyl radical
C "3
concentration. If a hydroxyl radical concentration of 2 x 10 molecules/cm
(typical for summer months; winter concentrations are lower) (Singh et al., 1981)
is assumed, most of the lifetimes calculated from the rate constants range
between 0.2 to 0.5 years (69 to 181 days, 122 average).
Mabey and Mill (1978) critically reviewed hydrolysis data available in the
literature. They determined that chloroform had a hydrolysis half-life of >3,000
_ij
years at pH 7 and 298 K. This is based on a base hydrolysis rate of 0.602 x 10
-7
and a OH concentration of 10 in neutral water.
Dilling et al. (1975) and Billing (1977) determined the volatilization
half-life of chloroform from water. For a 1 ppm chloroform solution stirred at
200 rpm, the time for 50% removal was 21.5 minutes (average); 90? removal was
accomplished in 71 minutes. The addition of dry granular bentonite clay,
dolomitic limestone, or peatmoss had little effect on the evaporation rate. The
rapid volatilization of chloroform was seen also by Jensen and Rosenberg (1975),
who reported that 0.1-1.0 ppn solutions of chloroform in partly open sunlit
aquaria lost 50-60$ of the chloroform in 8 days as opposed to only 5% in closed
aquaria. Pearson and McConnell (1975) suggested that the presence of chloroform
in ambient waters may be from aerial transport and washout.
An EXAMS model of the fate of chloroform in a pond, a river, and an oligo-
trophic lake and eutrophic lake revealed the dominant process in all cases to be
-9 -1
volatilization. Input parameters included hydrolysis (2.5 x 10 hr ),
octanol/water partition coefficient (91), vapor pressure (150.5 torr at 20°C),
solubility (8200 ppn), Henry's Law Constant (2.88 x 10~3), reaeration rate ratio
3-37
-------
OJ
I
TABLE 3-11
Values for k
OH
kOH
era
x 1014
molecule" sec"
6.51
10.1
16.8
11.4
6.4
7.4
10
[OH] Lifetime
K x 10~5 (years)
265 4 1.2
296
298 10 0.19
298
265 9 0.56
273
298
Reference
Singh et al . , 1979
Howard and
Evenson, 1976
Cox et al., 1976
Davis et al., 1976a
Hampson, 1980
a(4.69 + 0.71) x 10~12 exp
- (2254 + 214/RT)
Evaluated by NASA
-------
(0.583), alkoxy radical rate constant (0.7 M~1 hr~1) (RO = 1014 M), and a stream
loading of 1 g/hr. No photochemical or bacterial degradation parameters were
entered since chloroform has no UV absorbence >290 nm, and virtually no bacterial
degradation occurs with chloroform (Pearson and McConnell, 1975; Bouever et al.,
1981). Table 3-12 summarizes the EXAMS model generated fate of chloroform.
3.8. EXISTING CRITERIA, STANDARDS, AND GUIDELINES
3.8.1. Air. The Occupational Safety and Health Administration (OSHA) currently
limits occupational exposure to chloroform to a ceiling level of 50 ppm (40 CFR
1910.1000). This ceiling level is not to be exceeded in the workplace at any
time. To protect against mild central nervous system depression, irritant
effects, and fetal abnormalities (which were considered to occur at lower
exposure levels than those causing liver injury), the National Institute for
Occupational Safety and Health (NIOSH) recommended in 1974 that exposure to
chloroform be limited to 10 ppm as a Time-Weighted Average (TWA) exposure for up
to a 10-hour workday, 40 hour workweek. A ceiling level of 50 ppm was proposed
for any 10-minute period (NIOSH, 1974). NIOSH lowered the recommended criterion
to 2 ppm TWA in 1976 (NIOSH, 1977) in response to a positive NCI carcinogenesis
bioassay (NCI, 1976). NIOSH recommended that exposure to halogenated anesthetic
agents, including chloroform, be limited to 2 ppn because this is the lowest
detectable level using the recommended sampling and analysis techniques, and not
because a safe level of airborne exposure could be defined.
On the basis of recent reports of carcinogenicity and embryotoxicity, the
American Conference of Governmental Industrial Hygienists (ACGIH) currently
classifies chloroform as an Industrial Substance Suspect of Carcinogenic
Potential for Man (ACGIH, 1981). The ACGIH recommends a Threshold Limit Value
(TLV) of 10 ppm and a 15-minute Short-Term Exposure Limit (STEL) TWA of 50 ppm
for chloroform.
3-39
-------
TABLE 3-12
Summary of EXAMS Models of the Fate of Chloroform3
UJ
I
Maximum total concentrations in water
column (mg/i)
Maximum concentration in sediments
(dissolved in pore water, mg/i,)
Maximum concentration in Bios
Plankton (|ig/g)
Benthos (ng/g)
Maximum total concentration in sediment
(mg/kg, dry weight)
Total steady accumulation (kg)
% in water column
$ in sediments
Disposition
chemical transformation (?)
biotransformation (J)
volatilization (J)
Volatilization half-life
exported (J)
export half-life
Mass Flux from Volatilization (kg/hr)
Self-Purification Time
River
9.92 x 10~7
9.85 x 10~7
2.58 x 10~^
2.56 x 10~s
3.19 x 10""6
9.15 x 10"^
96.93
3-07
0.00
0.00
1.74
36 hours
98.26
0.65 hours
1.7i» x 10~5
37 hours
Pond
2.50 x 10~3
1.36 x 10~3
6.49 x 10~2
3-53 x 10"^
6.50 x 10~3
5.43 x 10~2
91.9
8.1
0.00
0.00
93.35
40 hours
6.65
566 hours
9.33 x lO"1*
31 days
Oligotrophic
Lake
1.33 x lO"4
5.82 x TO"6
3.45 x 10~3
1.51 x 10~*
2.83 x 10~5
0.33
99.95
0.05
0.00
0.00
94.98
10 days
5.02
192 days
2.28 x 10~3
65 days
Eutrophic
Lake
1.26 x 10~4
4.57 x 10"6
3.27 x 10~3
1.19 x 10"4
9.35 x 10"6
0.3
99.94
0.06
0.00
0.00
95.57
9 days
4.1(3
196 days
2.29 x 10~2
56 days
TJased on load of 1.00 g/hr
-------
Foreign industrial air standards for chloroform include (Utidjian, 1976):
Bulgaria, 10 ppm; Czechoslovakia, 10 ppn (50 ppn for brief exposures); Finland,
50 ppm; Hungary, 4 ppm (20 ppra for brief exposures); Japan, 50 ppm; Poland, 10
ppn; Rumania, 10 ppm; Yugoslavia, 50 ppm; West Germany, 10 ppm (Utidijian, 1976).
3.8.2. Water. As discussed in the Ambient Water Quality Criteria Document for
chloroform (U.S. EPA, 1980h), the EPA has proposed an amendment that would add to
the National Interim Primary Drinking Water Regulations a section on the control
of organic halogenated chemical contaminants. The proposed limit for total
trihalomethanes in drinking water, which includes chloroform as the major
constituent, is 100 ng/&. Although some estimates of cancer risk were performed,
this limit was set primarily on the basis of technological and economic
feasibility, and initially will apply only to water supplies serving >75,000
consumers. The basis and purpose of this regulation are discussed in a report
that was prepared by the Office of Drinking Water (U.S. EPA, 1978b).
The U.S. EPA (1980h) recently derived cancer-based ambient water criteria
for chloroform. Since zero level concentrations of chloroform will never be
attainable in chlorine-treated water, levels that may result in incremental
_5
increases of cancer risk over the lifetime were estimated at risks of 1 x 10 ,
r t-t
10 , and 10 . The corresponding recommended criteria, which were derived with
the tumor incidence data from the NCI bioassay with female mice (NCI, 1976), are
1.90, 0.19, and 0.019 p.g/&, respectively, if exposure is assumed to be from the
consumption of drinking water and fish and shellfish products and at 157 Hg/&,
15.7 \ig/&, and 1.57 ng/Ji,, respectively, if exposure is assumed to be from the
consumption of aquatic organisms only.
3.8.3. Food. Chloroform has been approved by the Food and Drug Administration
(FDA) as a component of articles intended for use in contact with food (i.e., an
3-41
-------
indirect food additive). The use of chloroform in the food industry is
summarized as follows:
Component of adhesives 21 CFR 175.105
Adjuvant substance required 21 CFR 177.1580
in the production of
polycarbonate resins
Chloroform also has been exempted from the requirement of tolerance when
used as a solvent in pesticide formulations that are applied to growing crops (40
CFR 180.1001), or when used as a fumigant after harvest for barley, corn, oats,
popcorn, rice, rye, sorghum (milo), or wheat (40 CFR 180.1009).
3.8.4. Drugs and Cosmetics. The positive NCI carcinogenicity bioassay of
chloroform (NCI, 1976) has prompted the FDA to restrict the use of chloroform in
drug (21 CFR 310-513) and cosmetic (21 CFR 700.18) products.
3-9. RELATIVE SOURCE CONTRIBUTIONS
The sum of all the environmental releases of chloroform from all sources
listed in Section 3-^-3 amounts to a total of 10,438 metric tons. All sources
are summarized in Table 3-8 with the percent of the total emissions. Total
emissions from all sources constitute about 5.6% of production (184,000 metric
tons). Table 3-8 does not include estimated emissions from laboratory use. Rem
et al. (1982) suggested that these are potentially large but gave no numerical
estimate.
3-42
-------
REFERENCES
ACGIH (American Conference of Governmental Industrial Hygienists). 1930.
Documentation of the Threshold Limit Values. 4th ed. Cincinnati, OH. pp.
90-91 .
ACGIH (American Conference of Governmental Industrial Hygienists). 1981.
Threshold Limit Values for Chemical Substances and Physical Agents in the
Workroom Environment with Intended Changes for 1931. Cincinnati, OH. pp. 13,
42.
Ahlstrom, R.C., Jr. and J.M. Steele. 1979. Methylchloride. In: Kirk-Othmer
Encyclopedia of Chemical Technology, 3rd ed., M. Grayson and D. Eckroth, eds.
New York: John Wiley and Sons, Inc. vol. 5. pp. 677-685.
Altshuller. 1980. Lifetimes of organic molecules in the troposphere and lower
stratosphere. In: Advances in Environmental Science and Technology, vol. 10.
Pitts and Metcalf, eds. New York: John Wiley and Sons, Inc. p. 181-219.
Anonymous. 1983. Chemical Profile. Chloroform. Chemical Marketing Reporter,
June 25, 1979, p. 54.
Anthony, T. 1979. Methylene chloride. In: Kirk-Othmer Encyclopedia of
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3-54
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4. DISPOSITION AND RELEVANT PHARMACOKINETICS
4.1. INTRODUCTION
Considering that chloroform was the major anesthetic agent in use during the
hundred years from its introduction by Simpson in 1847 (Waters, 1951; Snow, 1858;
Simpson, 1817) until after the Second World War, there is relatively little
detailed information about its pharmacokinetics and metabolism in man. This
undoubtedly is due to the fact that until recently, specific and sensitive ana-
lytical methods were unavailable for the measurement of CHCl, and its metabolites
at the concentrations in which they were likely to be present in vivo. Although
chloroform as an anesthetic agent has been replaced by drugs with less cardiac
and hepatic toxicity, it is still widely used in large bulk as an industrial
solvent, as a chemical intermediary, and as a grain fumigant. Chloroform is
present in the water supplies of many United States cities in concentrations
reaching 311 |ig/&, and also has been indentified as a contaminant of the air
(U.S. Occupational Safety and Health Administration (OSHA), 1978; National
Institute for Occupational Safety and Health (NIOSH), 1977b; Dowty et al., 1975;
Symons et al., 1975). Ordinary exposure to chloroform, therefore, includes
occupational, food, drinking water, and ambient air (NIOSH, 1977b; Dowty et al.,
1975; McConnell et al., 1975), hence, exposure can be chronic by both oral and
pulmonary routes, but at levels far below anesthetic concentrations (5000 to
10,000 ppm; 24.85 to 49.70 g/nr) . Nonetheless, chloroform has been detected in
the breath of healthy people living in non-industrial environments (Conkle
et al., 1975) and in post-mortem human tissue samples (McConnell et al., 1975).
4-1
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4.2. ABSORPTION
Chloroform is rapidly and extensively absorbed through the lungs and from
the gastrointestinal tract. Inhalation is considered the primary route of
entrance into man for occupational exposure and air pollution. Absorption after
oral ingestion is of particular interest for chloroform as a contaminating
component of drinking water and foodstuffs. Significant absorption through
intact skin occurs only with liquid contact or submersion.
4.2.1. Dermal Absorption. Absorption of chloroform through the skin from direct
liquid contact (immersion of hands or arms) is a slow process. Early studies
(Torkelson et al., 1976; Schwenkenbecher, 1904; Witte, 187*0 showed that chloro-
form does penetrate the skin and can be absorbed into the body by this route.
Tsurata (1975, 1977) has studied the percutaneous absorption of a series of
chlorinated organic solvents applied to a standard area of shaved abdominal mouse
skin for 15 minute periods. Absorption was quantitated by presence of compound
in total mouse body plus expired air, as determined by GC. For all solvents,
percutaneous absorption linearly increased with time over the short exposure
period and was directly related to water solubility. For chloroform the absorp-
'->
tion rate was 329 moles/min/cm skin, third highest of 8 solvents measured. This
investigator extrapolated this absorption rate to a calculation of the amount
absorbed into the human body as the result of 1 min immersion of both hands
p
(800 cm area). The estimated amount absorbed, 19.7 mg/min, was equated to an
inhalation exposure concentration of 2429 ppm for 1 min. Tsurata concluded that
skin absorption from liquid contact could be a significant route of entry into
the body for chloroform. More recently Jakobson et al. (1983) carried out
similar experiments with guinea pigs for 10 chlorinated organic solvents (which
4-2
-------
however did not include chloroform). Liquid contact (skin area, 3.1 cm ) was
maintained for up to 12 hours and solvent concentration in blood was monitored
during, and for some solvents after exposure. For these solvents, the blood
elimination curves following dermal exposure were non-linear, corresponding to a
kinetic model involving at least two body compartments. Furthermore
percutaneous absorption of these solvents, as reflected by blood concentration
profiles, showed three different patterns that were related to water solubility.
For solvents which were relatively hydrophilic [300 to 900 rag/100 ml water] the
blood concentration increased steadily during the entire dermal exposure,
indicating that absorption occurs faster than elimination by metabolism or
pulmonary excretion. Chloroform with a water solubility of about 750 mg/100 ml
water might be expected to be in this group.
4.2.2. Oral. The kinetics of gastrointestinal absorption of chloroform after
oral ingestion have not been specifically studied; however, transmucosal diffu-
sive passage occurs readily, as expected from its neutral and lipophilic proper-
ties (Tables 4-1 and 4-2), and as demonstrated by its biological effects produced
by peroral administration of a wide range of dosages and dosing schedules for
toxicity studies in rats, mice, guinea pigs, and dogs (Fishbein, 1979; Hill
et al., 1975; Brown et al., 1974; Kimura et al., 1971; Klaassen and Plaa, 19&7;
Miklashevskii et al., 1966; Plaa et al., 1958; Eschenbrenner and Miller, 1945),
teratolog)^ studies in rats and rabbits (Thompson et al., "974), and metabolism
studies in mice, rats, rabbits, monkeys, and man (Brown et al., 1974; Taylor
et al., 1974; Fry et al., 1972; Rubinstein and Kanics, 1964; Paul and Rubinstein,
1963). Accidental and intentional ingestion of chloroform with rapid appearance
of clinical symptoms also has been reported in man (Storms, 1973; Schroeder,
1965; Piersol et al., 1933).
4-3
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TABLE 14-1
Physical Properties of Chloroform and Other Chloromethanes*
Ostwald Solubility, 37°C
Dichloromethane
Chloroform
Carbon tetrachloride
Vapor Pressure
at 25°C, torr
400
250
100
Water/
air
7.6
4.0
0.25
Blood/
air
9.7
10.3
2.4
Olive Oil/
air
152
401
361
*Source: Sato and Nakajima, 1979
Conversion factors:
20°C; 750 mmHg 1 ppm in air =4.97 |ig/& =4.97
37°C; 760 mmHg 1 ppm in air = 4.69 Hg/fc = 4.69 rag/iir
4-4
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TABLE 4-2
Partition Coefficients for Human Tissue at 37°C
Tissue
Blood
Brain
Grey matter
White matter
Heart
Kidney
Liver
Lung
Mus cl e
Fat tissue
Coefficient
8.0
16
24
8
11
17
7 -
12
280
Relative
to
blood
2.0
3.0
1.0
1.4
2.1
0.9
1.5
35.0
Source: Steward et al., 1973
4-5
-------
Brown et al. (1974) and Taylor et al. (1974) found that l4C-chloroform in
olive oil given perorally to mice, rats, and monkeys (60 mg/kg) was essentially
completely absorbed by virtue of a 93 to 9856 recovery of radioactivity in exhaled
air, urine, and carcass (Table 4-9). Absorption was rapid, with peak blood
levels at 1 hour in mice and monkeys. In man, Fry et al. (1972) observed that
1 3
JC-chloroform (0.5 g) in olive oil swallowed in a gelatin capsule resulted in
rapid appearance of the stable isotope in exhaled breath (Table 4-8), with peak
blood levels at 1 hour.
Withey et al. (1982) have investigated the effect of dosing vehicle on the
intestinal absorption of chloroform in fasting rats (400 gm) following
intragastric intubation of equivalent doses (75 mg/kg) in about 4 ml of water or
corn oil. The postabsorptive peak blood concentration averaged 6.5 times higher
for water than corn oil (39 vs 6 |ig/ml), while the time to initial peak blood
concentration was essentially the same (5.6 vs. 6.0 min). Although the
absorption from water vehicle exhibited one blood concentration peak, the
absorption from corn oil showed two peaks in blood concentration at 6 and 40
minutes. The ratio of the areas under the blood concentration curves for 5 hours
after dosing (AUC, 5 hr) was 8.7; water, corn oil. These results suggest that
the absorption of chloroform with both vehicles is rapid; but, the rate and
extent of absorption may be diminished, and the pattern of absorption altered by
intragastric intubation of high volumes (for a rat) of corn oil vehicle. A
slower partitioning of lipophilic compounds dissolved in corn oil with mucosal
lipids can be expected in comparison with a water vehicle. Furthermore, in
contrast to aqueous absorption into the portal system and thence to the liver,
corn oil and other liquids are extensively transported via mucosal lymphatic
system which slowly drains by way of the left lymphatic thoracic duct into the
systemic circulation via the superior vena cava. While these considerations are
4-6
-------
unlikely to affect the pharmaeokinetics of chloroform in man in any practical
way, they are of importance in relation to the modes of dosing employed in long-
term carcinogenicity tests of chloroform and other lipophilic compounds.
4.2.3. Pulmonary Absorption. Chloroform has a relatively high vapor pressure
(250 torr at 25°C; Table 4-1) and a high blood/air partition coefficient (8 to
10.3 at 37°C; Table 4-2), and hence, its vapor in ambient air is a primary mode of
exposure and the lungs a principal route of entry into the body. The total
amount absorbed via the lungs (as for all vapors) is directly proportional to:
(1) the concentration of the inspired air, (2) the duration in time of exposure,
(3) the blood/air Ostwald solubility coefficient, (4) the solubility in the
various body tissues, and (5) physical activity, which increases pulmonary ven-
tilation rate and cardiac output. Hence, the basic kinetic parameters of the
pulmonary absorption of chloroform and its equilibration in the body are as valid
for low concentrations expected in the ambient environment as for the high vapor
concentrations associated with its use as an anesthetic (5000 to 10,000 ppm;
24.85 to 49.70 g/m3) (Smith et al., 1973; Morris, 1951; Waters, 1951). These
parameters have not been as well studied as they have for modern anesthetics like
halothane (Fiserova-Bergerova and Holaday, 1979) or even for other common halo-
genated hydrocarbon solvents like triohloroethylene, methylene chloride, or
methylchloroform.
The earliest attempt at controlled studies of pulmonary absorption of
chloroform in man were conducted by Lehmann and Hasegawa (1910). These investi-
gators calculated retention values for chloroform (% inspired air concentration
of chloroform retained in the body) from differences between inspired and expired
air concentrations (analyzed by alkali hydrolysis with chloride titration). As
expected, initial retention values were high, and decreased with exposure dura-
4-7
-------
tion as total body equilibrium with inspired air concentration was approached
(Table 4-3). The rate of uptake to equilibration and the final retention value
achieved is related to the solubility of chloroform in blood (blood/air partition
coefficient). Figure 4-1 illustrates for chloroform and other vapors that the
greater the Ostwald solubility coefficient for a vapor agent, the less rapid
equilibrium occurs. From the data of Lehmann and Hasegawa (Table 4-3) and recent
data of Smith et al. (1973) of blood levels during anesthesia shown in Figure
4-2, total body equilibrium with inspired chloroform concentration requires at
least >2 hours in normal man at resting ventilation rate and cardiac output. The
retention value at equilibrium suggested by the Lehmann and Hasegawa (1910) data
is ^65%, and is 67% as calculated from the data of Smith et al. (1973). The
difference, 33 to 3&%, represents body elimination of chloroform by routes other
than pulmonary (primarily by metabolism). The percent retention value is
independent of the inspired air concentration at equilibrium.
The magnitude of chloroform pulmonary uptake into the body (dose, body
burden) is directly related to the concentration of chloroform in the inspired
air and to the duration of exposure. The total amount retained in the body
during inhalation exposure can be estimated by multiplying percent retention (R)
by the volume of air inspired during the exposure period, or:
Amount uptake = (CT - C.) • V • T
JL A
where V is ventilation rate (Jl/minute), T is exposure period (minute), and C,. and
C are inspired air concentration and end alveolar air concentration,
respectively. Physical activity increases uptake by increasing the ventilation
rate, V, and the cardiac output which influences rate of distribution to the
various tissues of the body.
4-8
-------
TABLE 4-3
Retention and Excretion of Chloroform in Man
During and After Inhalation Exposure to
Anesthetic Concentrations*
Subject
Inspired air cone, (ppra)
Exposure period (min)
0 to 5
5 to 10
10 to 15
15 to 20
20 to 25
25 to 30
Postexposure (min)
4448
4920
Retention (%)
4407
74.5
72.4
68.6
67.6
NR
NR
68.4
61.6
51 .2
50.2
NR
NR
80.0
74.2
76.9
74.6
74.2
73.8
Excretion, mg/g, expired air
0 to 10
10 to 20
20 to 30
NR
NR
NR
NR
NR
NR
1.70
0.97
0.85
*Source: Lehmann and Hasegawa, 1910
NR = Not reported; min = minutes
4-9
-------
100
10 15 20
TIME, min
Figure 4-1. Rate of rise of alveolar (arterial) concen-
tration toward inspired concentration for five anesthetic
agents of differing Ostwald solubilities (blood/air par-
tition coefficients): nitrous oxide, 0.47; forane, 1.4;
halothane, 2.4; chloroform, 8; and methoxyflurane, 11.
Note rate of alveolar chloroform rise is less than that
of halothane with a smaller Ostwald coefficient and
greater than that of methoxyflurane with a larger coef-
ficient .
Source: Munson (1973)
4-10
-------
c
V
o
Z
Z
LLI
U
o
o
5
cc
o
LL
o
cc
o
I
CJ
12
10
CHLOROFORM
O VENOUS BLOOD
D ARTERIAL BLOOD
BASE
EXCESS
pH
+1.5
40.0
7.44
•2.0
36.0
7.39
-2.0
33.0
7.40
-3.0 -4.0
33.0 30.0
7.39 7.39
-5.0 -1.5
27.0 36.0
7.40 7.41
0 1.2 3 4 | 5 6
POST INDUCTION TIME, hours
Figure 4-2. Arteriovenous blood concentrations of a
patient during'anesthesia with chloroform. Note anes-
thetic blood concentration for chloroform, the decreasing
difference between arterial and venous concentrations at
2 to 3 hours, indicating whole-body equilibrium, and the
rapid fall of blood concentration with termination of
chloroform exposure.
Source: Smith et al. (1973).
'4-11
-------
During inhalation of chloroform (and in the post exposure elimination
phase), the arterial blood concentration of chloroform is directly proportional
to inspired air concentration (and end alveolar air concentration). This fixed
relationship is defined by the blood/air partition coefficient in comparison to
other solvents (Sato and Nakajima, 1979) (Table 4-1), and hence, for equivalent
ambient air exposure concentrations, the blood concentration of chloroform Is
proportionally higher. For inspired air concentration required for surgical
anesthesia (8,000 to 10,000 ppm; 39/76 to 49.70 g/m3), Smith et al. (1973)
observed a mean arterial blood chloroform concentration for 10 patients of
9.8 mg/d& with a range of 7 to 16.5 mg/dil (Figure 4-2), and Morris (1951) found
similar values for his patients. For inspired air concentrations less than
anesthetic levels, for example low vapor concentrations of 10 to 100 ppm (49.7 to
497 mg/nr), blood chloroform concentrations are lower in direct proportion.
The amount of pulmonary absorption of chloroform is also influenced by total
body weight and by the total fat content of the body (average body fat content is
8$ of body weight) (Geigy Scientific Tables, 1973). The capacity of adipose
tissue to absorb chloroform rn vivo is determined by the product of adipose
tissue weight and lipid solubility of chloroform. The lipid solubility of
chloroform is relatively high for this haloalkane (olive oil/air, 401; Tables 4-1
and 4-2), and also, the adipose tissue/blood partition coefficient is high (280
at 37°C); therefore, the uptake and storage of chloroform in adipose tissue can
be substantial, and it is increased with excess body weight and obesity.
4.3. TISSUE DISTRIBUTION
Chloroform, after pulmonary or peroral absorption, is distributed into all
body tissues. The compound crosses the placental barrier, as indicated by
4-12
-------
embryotoxicity and teratogenicity in mice, rats, and rabbits after oral and
inhalation dosing (Murray et al., 1979; Dilley et al., 1977; Schwetz et al.,
1974; Thompson et al., 1974). It has been found in fetal liver (von Oettingen,
1964). Chloroform can be expected to also appear in human colostrum and mature
breast milk, since it has been found in fresh cow's milk and in high content in
cheese and butter (Table 4-4).
As to be expected from its lipophilic nature and modest water solubility
(Table 4-1), highest concentrations are found in tissues with higher lipid
content; relative tissue concentrations are reflected by individual tissue/blood
partition coefficients. Coefficients for human tissues, given in Table 4-2,
indicate that relative tissue concentrations are expected in the order of adipose
tissue > brain > liver > kidney > blood. The absolute amounts of chloroform
found in these tissues at any given time are proportional to the body dose (i.e.,
to the concentration in the inspired air and duration of inhalation or to the
oral dose, partition coefficient, and to the tissue compartment size).
Gettler (1934) and Gettler and Blume (1931), using a modified Fujiwara
analytical method, determined the chloroform content of the brain, lungs, and
liver of nine patients who died during surgical anesthesia (presumably 5000 to
10,000 ppm; 24.85 to 49.50 g/m3 inspired air) as: brain, 120 to 182; lung, 92 to
145; and liver, 65 to 88 mg/kg tissue wet weight. Even higher values (372 to 480
mg/kg in brain tissue) were found in seven cases of death due to excessive
administration of chloroform (Gettler, 1934). The blood concentration during
surgical anesthesia has been recently determined (by GC) to range from 70 to
165 mg/JJ, in 10 patients (average, 98) by Smith et al. (1973). These tissue
concentrations are in general agreement with the tissue/blood partition
coefficients summarized from the literature by Steward et al. (1973) and given in
Table 4-2.
4-13
-------
TABLE 4-4
Chloroform Content in United Kingdom Foodstuffs
and in Human Autopsy Tissue*
Chloroform in U.K. Foodstuffs
Foodstuff
Dairy produce
Fresh milk
Cheshire cheese
English butter
Hens eggs
Meat
English beef (steak)
English beef (fat)
Pig's liver
Oils and Fats
Margarine
Olive oil (Spanish)
Cod liver oil
Vegetable cooking oil
Beverages
Canned fruit drink
Light ale
Canned orange juice
Instant coffee
Tea (packet)
Fruit and vegetables
Potatoes (S. Wales)
Potates (N.W. England)
Apples
Pears
Tomatoes
Fresh bread
Chloroform
Hg/kg
5
33
22
1.4
4
3
1
3
10
6
2
2
0.4
9
2
18
18
4
5
2
2
2
Chloroform in Human Autopsy
Age of
Subject Sex Tissue
76 F Body fat
Kidney
Liver
Brain
76 F Body fat
Kidney
Liver
Brain
82 F Body fat
Liver
48 M Body fat
Liver
65 M Body fat
Liver
75 M Body fat
Liver
66 M Body fat
74 F Body fat
Tissue
Chloroform
Hg/kg
(Wet tissue)
19
2
5
4
5
5
1
2
67
8.7
67
9.5
64
8.8
65
10.0
68
52
"Source: McConnell et al., 1975
4-14
-------
In contrast to the high tissue levels of chloroform found in response to
inspired air concentrations required for anesthesia, McConnell et al. (1975)
recently analyzed post-mortem tissue from eight persons, four males and four
females, with an age range of 48 to 82, living in the United Kingdom in ordinary
non-industrial circumstances, for chloroform and other halogenated compounds
(carbon tetrachloride, trichloroethylene, perchloroethylene, hexachloro-
butadiene). Significant tissue levels of three chlorinated hydrocarbons were
found. Chloroform level, [ig/kg wet tissue weight, were: body fat, 5 to 68
(average of 51); liver, 1 to 10 (average of 7.2); kidney, 2 to 5; and brain, 2 to
4 (Table 4-4). Presumably, these tissue levels of chloroform were derived from
air, foodstuff (Table 4-4), and drinking water contamination (OSHA, 1978; Dowty
et al., 1975; Symons et al., 1975).
There have been few controlled exposure studies in animals investigating
the distribution of chloroform in body tissues and determining dose-dependent
tissue concentrations. Chenoweth et al. (1962) determined blood and tissue con-
centrations of chloroform in two normal fasted dogs after 2.5 hours of surgical
anesthesia. Concentration of chloroform in the inhaled stream during anesthesia
was not determined, but anesthesia was judged to be satisfactory at an arterial
level of 45 to 50 mg/d£. Blood and tissue chloroform was determined by infrared
spectroscopy after tissue extraction in cold carbon disulfide and distillation.
Table 4-5 shows the relative concentration of chloroform in body tissues. The
highest concentrations were found in fat tissue, some 10-fold greater than blood,
and in adrenals (4-fold greater than blood); the concentrations in brain, liver,
and kidney were similar to blood.
Cohen and Hood (1969) used low-temperature whole-body autoradiography to
study the distribution of C-chloroform in mice. Individual mice were
administered 2.4 (i£ of C-labeled chloroform by inhalation over a 10-minute
4-15
-------
TABLE 4-5
Concentration of Chloroform in Various Tissues
of Two Dogs After ?.5 Hours Anesthesia3
Arterial blood
Brain
Adrenal (total)
Fat , omentum
Right ventricle
Skeletal muscle
Lung
Liver
Spleen
Kidney
Bile
Thyroid
Pancreas
Urine
Dog A
|ig/gm wet
275
298
1185
2820
21*4
189
147
282
237
225
209
460
296
57
Dog B
tissue weight + 5%
397
392
1305
1450
314
155
336
290
255
226
205
760
350
73
Source: Chenoweth et al., 1962
Chloroform concentration was determined by infrared spectrometry after tissue
extraction.
4-16
-------
period. The animals were sacrificed 0, 15, or 120 minutes after exposure.
Autoradiography of mice killed immediately after inhalation showed the highest
concentration of radioactivity in body fat and liver, while lesser and relatively
uniform amounts were seen in blood, brain, lung, kidney, and muscle. By 120
minutes after exposure, a considerable decrease in total radioactivity occurred,
now principally confined to liver, duodenum, and fat. A mottled appearance in
the liver suggested a segmental or localized distribution. Biopsy specimens were
taken from selected tissues in each animal and radioactivity determined by scin-
tillation counting. Table 4-6 shows the distribution of radioactivity (chloro-
form and metabolites) in these tissues, and tissue/blood concentration ratios.
Following sacrifice, after 10 minutes of exposure, most tissues approach a unit
concentration with blood. However, in both fat and liver, the concentration
exceeds unity. By 15 minutes, the ratio of radioactivity in brown fat reaches
its peak at 15 times that found in blood. The relative concentration of radio-
activity in the liver continues to increase until the termination of the experi-
ment at 120 minutes, when it reaches a final value 6.7 times in excess of that in
the blood. Kidney and lung tissues also increased in relative concentration over
the 2-hour period to a value of 1.53 and 1.43 of blood, respectively. The
increasing ratios of liver and kidney/blood radioactivity represent a continued
accumulation of metabolites within these organs. High body fat/blood concentra-
tions shows that adipose tissue represents an important storage site, prolonging
retention of chloroform in the body.
Whole-body autoradiography was also carried out by Brown et al. (1974) on
male and female Sprague-Dawley rats and squirrel monkeys given C-chloroform
perorally (60 mg/kg). Male and female rats killed 3 hours after dosage showed no
apparent sex difference in distribution of radioactivity. Radioactivity was
greatest in body fat and liver, while lesser amounts were seen in blood, brain,
4-17
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lung, kidney, and muscle. Squirrel monkeys showed a similar distribution, with
the exception that high concentrations of radioactivity were present in the bile
and increased with time. Examination of bile extract by gas-liquid chromato-
graphy showed the bile radioactivity was unchanged chloroform, indicating an
excretion of chloroform by the biliary route in the monkey.
Brown and his colleagues (Taylor et al., 1974) also investigated the tissue
distribution of chloroform in 3 strains of mice (CF/LP, CBA, and C57) by whole-
body autoradiography after oral dosing (60 mg/kg C-chloroform). In the male
mice of the three strains examined 3 hours after dosing, the greatest amounts of
radioactivity appeared in liver and kidneys, and lesser amounts in renal cortex
but not medulla. Female mice showed greatest radioactivity in liver, intestine,
and bladder, with much less radioactivity in kidney and little differentiation
between renal cortex and medulla. The same general patterns were observed 5, 7,
and 24 hours after dosing. Biopsy samples of these tissues were taken, and
radioactivity determined by scintillation counting. Table 4-7 shows the
14
distribution of C-chloroform radioactivity in male and female mice killed 5
hours after dosing. There is a 3.5-fold difference between the activity present
in the male and female kidneys of each strain. Male mice had greater activity in
the kidneys, but female mice showed relatively greater activity in the liver.
This sex difference in distribution was abolished by castration or testosterone
administration to female mice. The sex difference in tissue distribution of
chloroform and its metabolites may relate to the nephrotoxic effect of chloroform
that occurs in male mice but not in female mice (Bennet and Whigham, 1964;
Culliford and Hewitt, 1957; Hewitt, 1956; Shubik and Ritchie, 1953;
Eschenbrenner and Miller, 1945a, 1945b). The pattern of tissue distribution of
chloroform in mice also depends on mode of exposure. Cohen and Hood (19&9),
after chloroform inhalation, found highest levels in body fat (Table 4-6), while
4-19
-------
TABLE 4-7
ill
Tissue Distribution of C-Chloroform Radioactivity
in CF/LP Mice After Oral Administration (60 mg/kg)a'b
Mean DPM/100 mg
Tissue
Liver
Kidney
Brown fat
Blood
Male
18,157 (
13,759 (
1011
2910
(6)
1898)
1047)
(80)
(423)
wet weight (SEM)
Female
21,535 (
3920
1074
2906
(6)
2097)
(533)
(54)
(457)
Source: Taylor et al., 1974
Similar results were obtained for CBA and C57 strains.
4-20
-------
Brown et al. (1974) and Taylor et al. (1974) observed lower levels in fat and
highest levels in liver and kidney following oral dosing (Table 4-7). The high
liver levels of chloroform after oral administration may be due, in part, to
first passage and extraction by the liver after this route of administration, to
differences of time after exposure (2 versus 5 hours), and to metabolism and
covalent binding of metabolites to cellular macromolecules (see below).
It is worth re-emphasizing that the sex difference in tissue distribution
and binding of chloroform (and metabolites) in kidney and liver, noted by Brown
and his colleagues (Taylor et al., 1974), appeared to be peculiar to mice and
that these workers did not observe such differences in male and female rats or
squirrel monkeys (Brown et al., 1974).
4.4. EXCRETION
Elimination of chloroform from the body is perforce the sum of metabolism
and excretion of unchanged chloroform via pulmonary and other routes.
Unmetabolized chloroform is excreted almost exclusively through the lungs; how-
ever, metabolism of chloroform is extensive, with the proportion excreted
unchanged dependent on body dose. Surprisingly, considering its historical
importance, its longtime use as an industrial chemical and anesthetic agent, few
controlled experimental studies in man have been made on the kinetics of excre-
tion of chloroform.
4.4.1. Pulmonary Excretion. Figure 4-3 shows the time-course of pulmonary elim-
ination of chloroform after accidental inhalation exposure to a mixture of sol-
vents, including chloroform, carbon tetrachloride, trichloroethylene, and per-
chloroethylene. Stewart et al. (1965) determined, post-exposure, the alveolar
4-21
-------
z
o
e
•-
UJ
u
O
U
or
O
UJ
OPERCHLOROETHYLENE —
D CARBON TETRACHLORIDE —
A CHLOROFORM —
OTETRACHLOHOETHYLENE ~~
0 12 24 36 48
96 120
144 168 192 216 240 264 288 312 336 360
TIME, hr
Figure H-3. Exponential decay of chloroform, carbon tetrachloride,
perchloroethylene and trichloroethylene in exhaled breath of 48 year-old
male accidentially exposed to vapors of these solvents. The alveolar air
(end tidal) was collected in a 50-mJl glass pipette, and the sample was analyzed
by infrared spectroscopy and gas-liquid chromatography. Initial values (30 rain.
post exposure) were chloroform, 7 ppm; carbon tetrachloride, 9.5 ppm;
perchloroethylene, 11 ppm; and trichloroethylene, 4 ppm.
Source: Stewart et al. (1965).
4-22
-------
air concentrations of these solvents by infrared and gas-liquid chromatography
analysis. The kinetics of pulmonary excretion of the solvents are independent of
one another. However, all, including chloroform, demonstrate the typical kine-
tics of gaseous vapor pulmonary elimination, that has been observed
experimentally for relatively hydrophobic, volatile gaseous anesthetics and
industrial solvents (Eger, 1963; Fiserova-Bergerova et al., 1974, 1979, 1980;
Droz et al., 1977). At termination of exposure with zero concentration in
inspired air, chloroform (and the other solvents) immediately begins to be
eliminated from the body into the lungs, with blood and alveolar air concentra-
tions describing parallel exponential decay curves with three major components
(Figure 4-3). These exponential components have been related by many investi-
gators (Eger, 1963; Fiserova-Bergerova et al., 1974-, 1979, 1980; Droz et al.,
1977) to first-order kinetics of pulmonary elimination associated with desatura-
tion of physiological compartments in accordance with a blood flow-limited model
in which the rate constants are determined predominately by tissue perfusion,
volume of tissue distribution and by partition coefficients:
Tissue uptake and FT
desaturation of = F • X . exp ( - • t )
compartment,T bl/air -VT A T/bl
where F is blood flow through tissue compartment, V is volume of the compartment,
X is partition coefficient, and exp is base of natural logarithm.
0.693 V
Since three exponential components are typically observed experimentally,
three physiological tissue compartments are included in the model described by a
three-term exponential function of the form:
Uptake, Desaturation = A e ~at + B e "^ + C e ~yt
4-23
-------
where A, B, C are macrocoefficients and a, (3, Y» are hybrid constants (defined
above). These terms represent three flow-limited major body compartments (1) a
vessel-rich group of tissues (VRG) with high blood flow and high diffusion rate
constant (VRG: brain, heart, kidneys, liver and endocrine and digestive
systems), (2) lean body mass (MG: muscle and skin), and (3) adipose tissue (FG).
More recently, Fiserova-Bergerova and coworkers (1974, 1979, 1980) have
mathematically re-formulated this physiological, first-order model to accomodate
the effect of metabolism on uptake, distribution and clearance of inhaled vapor
compounds.
The half-time (t 1/2) of elimination from the physiological compartments
(VRG < MG < FG) are independent of the body dose, but are dependent on
tissue/blood partition coefficients and blood/air partition coefficients . Since
these solvent compounds have high solubility in body fat (Table 4-1), they are
eliminated slowly from fat depots with long half-time of elimination, as
illustrated by Stewart's patient (Stewart et al., 1965) in Figure 4-3. From
Figure 4-3, it can be estimated graphically that chloroform has a half-time of
elimination from the fat compartment (FG) of =36 hours, with similar long half-
times for the highly fat soluble compounds, perchloroethylene and carbon tetra-
chloride.
There is little information available for the half-times of pulmonary
elimination of chloroform from the VRG and MG. From the early data of Lehmann
and Hasegawa (1910) given in Table H-3, the half-time of pulmonary elimination
from the VRG appears to be =30 minutes. A similar estimate can be made from the
data of Smith et al. (1973) and Morris (1951) at termination of anesthesia in
man; these workers found that blood chloroform concentration rapidly fell
exponentially from 7 to 3.5 mg/d£ within 30 minutes (Figure 4-2).
4-24
-------
Pulmonary elimination of chloroform was investigated by Fry et al. (1972)
1 3
in male and female volunteers given JC-chloroform in olive oil orally (by
gelatin capsule). Chloroform was determined in expired air by GLC. Their data,
summarized in Table 4-8, show that the amount of chloroform excreted through the
lungs within 8 hours (expressed as a percentage of the dose, 0.1 to 1.0 g),
increased (0 to 65%) in proportion to the dose. Following a peak blood concen-
tration (0.5 mg/dS, for a 500 mg dose) 1 hour after oral dosage, absorption, and
distribution, the blood chloroform concentration declined exponentially with
three components: (1) a very rapid disappearance, with half-time of 14 minutes
possibly corresponding to VRG compartment kinetics, (2) a slower disappearance,
with half-time of 90 minutes corresponding to MG kinetics, and (3) a very slow
disappearance, with very long half-time from adipose tissue. This half-time was
undetermined, but chloroform was detected in blood and breath 24 hours later.
Fry and coworkers (1972) noted a linear relationship for their subjects between
pulmonary excretion of chloroform and body weight deviation from ideal, an index
of excessive leanness or excessive body fat from normal. Their data in Figure
4-4 show that for both male and female subjects given a standard oral dose of
chloroform, lean subjects eliminate via the lungs a greater percentage of the
dose, while overweight subjects eliminate less chloroform. The different slopes
of the linear relationship for men and women presumably reflect the different
proportion of adipose tissue in the two sexes. The bodies of women tend to
contain higher proportions of fat than those of men (Geigy Scientific Tables,
1973). These observations reinforce the role of adipose tissue as a storage site
for chloroform.
Brown and his coworkers (Brown et al., 1974; Taylor et al., 1974) have
demonstrated an animal species difference in the amount of pulmonary excretion of
14
C-chloroform from a standard oral dose (60 mg/kp, body weight) given in ol.ive
4-25
-------
Subjects
8 M and F
1
1
1
Pulmonary
Dose
(g)
0.5
1.0
0.25
0.10
TABLE 4-8
Excretion of 1 ^CHCl
Dose: Percent of
Mean
for 8 Hours b
40.3
64.7
12.4
nil
Following Oral
Dosea
Range
17.8 to 66.6
NA
NA
NA
-I -3
Pulmonary excretion of JCO? following 0.5 g oral dose
1 O V\
of CHC1 • Cumulative percent of dose
Time after dose (hours)
Subjects 0.5 1.75 2.5 5.5 7.5
Male (1) 2.1 24.1 35.9 49.2 50.6
Female (62.7 kg) (1) 0.5 10.7 28.3 47.5 48.5
Recalculated from the data of Fry et al., 1972
vfithin 4ft of value calculated for infinite time
NA = Not applicable
4-26
-------
-6-4-2 0 +2 +4 +6 +8
BODY-WEIGHT DEVIATION FROM CALCULATED NORMAL, kg
Figure 4-4. Relationship between total 8-hour pulmonary
excretion of chloroform following 0.5-g oral dose in man and
the deviation of body weight from ideal. The different slopes
of the linear relationship for men and women reflect the
different proportion of adipose tissue in the two sexes.
4-27
-------
oil. Mice (three strains), rats, and squirrel monkeys excrete chloroform via the
lungs (6, 20, and 79%, respectively, of the standard dose). This species
difference is primarily related to the capacity to metabolize chloroform rather
than differences in pulmonary kinetics, since, as shown in Table 4-9, the percen-
tage of the dose metabolized to CO- is inversely proportional to that of
pulmonary excretion. The mice, 48 hours after dosing, retained only 2% of
chloroform radioactivity (Table 4-9).
Withey and Collins (1980) determined the kinetics of distribution and
elimination of chloroform from blood of Wistar rats after intravenous adminis-
tration of 3, 6, 9, 12 or 15 m/kg of chloroform given in 1 ml water
intrajugularly. For all doses, the blood decay curves exhibited three components
of exponential disappearance of chloroform (a, (3> YcomP°nen^s) anc* "best" fitted
a first-order three compartment model. Table 4~'ib summarized the values obtained
for the kinetic parameters. For volatile, lipophilic compounds, for which a
major route of elimination is pulmonary, experiments utilizing dose administra-
tion via relatively large intravenous bolus injections (relative to rat total
blood volume), and which measures only blood chloroform disappearance, provide a
number of problems for data interpretation of elimination (pulmonary and
metabolism) and/or tissue distribution. In these experiments, pulmonary
elimination, which is rapid for organic solvents, occurs simultaneously with
distribution and metabolism; in contrast, experiments in which the animal is
preloaded by oral or inhalation administration, distribution is more readily
separable from pulmonary elimination. After intravenous administration (Table
4-tlb), the rate constant k (for elimination of chloroform from the central
compartment blood out of the body (principally via pulmonary excretion and/or
metabolism) was dose-dependent and consistent with a half-time of elimination of
only 3.6 min for the lowest dose and only 6.2 rain for the highest dose. Since the
4-28
-------
TABLE 4-9
14,
Species Difference in the Metabolism of ' C-Chloroform
(Oral Dose of 60 mg/kg)*
14
C-radioactivity 48 hours after dose
Species No. Mean values as percent dose
Expired 14CHC1
or metabolites
Mice
CF/LP, 19 6.1
CBA, C57
strains
Rats
S-D 6 19.7
Squirrel 6 78.7
Monkeys
Expired Urine +
14
C00 Feces Carcass Total
85.1 2.6 1.8 95.6
65.9 7.6 NR 93.2
17.6 2.0 NR 98.3
"Recalculated from the data of Brown et al., 1974
NR = Not recorded
4-29
-------
Table 4-/0. Kinetic Parameters for Chloroform After I.V. Administration to Rats
Dose* Vd
mg/kg ml
a
6
Y
k
e
kiz
k2i
kis
k3i
min"1 ± S.E.
3.0 45.07
± 0.04
6.0 53.57
± 12.61
f 9.0 64.46
o ±14.26
12.0 80.62
± 20.20
15.0 89.13
± 12.83
15.0; t, min
"2
0.72
± 0.11
0.64
± 0.13
0.64
± 0.084
0.32
± D.20
0.35
± 0.048
2.1
0.135
± 0.001
0.081
± 0.01
0.095
± 0.0001
0.060
± 0.028
0.070
± 0.006
9.9
0.0287
±0.0064
0.0158
± 0.0019
0.0189
± 0.0009
0.0074
± 0.0056
0.0134
± 0.0005
51.7
0.1907
± 0.0239
0.1874
±0.0356
0.1284
±0.0217
0.1035
± 0.0232
0.1124
± 0.0110
6.2
0.2346
± 0.0651
0.2681
± 0.0631
0.2529
± 0.0489
0.1071
± 0.0774
0.1104
± 0.0256
6.3
0.2575
± 0.0316
0.1862
± 0.0188
0.2421
± 0.0064
0.1192
± 0.0676
0.1523
± 0.0188
4.6
0.0921
± 0.0008
0.0730
± 0.0246
0.0594
± 0.0034
0.0395
± 0.0253
0.0396
± 0.0104
17.5
0.0421
± 0.0098
0.0233
± 0.0032
0.0281
± 0.0042
0.0106
± 0.0084 '
0.0193
± 0,0018
35.9
2 to 4 rats/dose . From Withey et al., 1982.
-------
half-times for distribution into other tissue compartments from the central
compartment blood have longer half-times, it is likely that, of the dose
introduced into the blood, a major portion (depending on dose) was excreted
within a few minutes by the lungs. Further indication that the dose and/or mode
of administration influenced the distribution and elimination of chloroform was
shown by the proportional increase of the apparent volume of distribution, V, (45
ml; 3 mg/kg to 89 ml; 15 mg/kg) and the decrease with dose in values for rate
constants of transfer from blood to other tissue compartments. The volume of
distribution (Vd) of chloroform was 89 ml for the highest dose or about 22% b.w.,
surprisingly low for a lipid soluble compound that is known to diffuse into all
the major organ systems (Tables 4-5 and 4-6). Adipose tissue is known to be a
major tissue compartment for chloroform [Section 4.3]; the clearance of chloro-
form from perirenal fat was found to be slow with a half-time of 106 min, and
since the rate constant, k was given as 0.0193 min~ (for 15 mg/kg dose)
indicating a half-time of 36 min, the adipose tissue appears to be a deep
compartment. These investigators believe that their kinetic data shows no
evidence in the rat of nonlinear or dose-dependent Michaelis-Menten kinetics and
they suggest that a dose of 15 mg/kg is below hepatic metabolism saturation.
4.4.2. Other Routes of Excretion. Chloroform is not eliminated in significant
amounts from the body by any route other than pulmonary. Studies of chlorinated
compounds in the urine after chloroform inhalation exposure or peroral dosage to
animals and humans have failed to detect unchanged chloroform (Brown et al.,
1974; Fry et al., 1972). Brown et al. (1974) identified chloroform in high
concentration in the bile of squirrel monkeys after oral C-chloroform dosage,
and suggested an active enterohepatic circulation in this species. For monkeys,
4-31
-------
they found only 2% of dose radioactivity in combined urine and feces collected
for 48 hours after dosing (Table 4-9), and only 8 and 3% for rats and mice,
respectively.
4.4.3. Adipose Tissue Storage. There is no definitive experimental evidence in
the literature concerning bioaccumulation after chronic or repeated daily
exposure to chloroform. However, there are practical reasons to believe that
extended residence in body fat occurs. In man, chloroform has a relatively high
fat tissue/blood partition coefficient of 35 (Table 4-2), a long half-time of
elimination from adipose tissue compartments of =36 hours (Figure 4-3), and it
has been detected in blood and breath 24 to 72 hours after a single exposure
(Stewart, 1974; Fry et al., 1972; Stewart et al., 1965.). Figure 4-5 clearly
shows the slow elimination of chloroform from the adipose tissue of dogs
following a 3-hour anesthesia (Chenoweth et al., 1962). Despite the rapid
exponential decline of blood levels of chloroform within 3 hours after
termination of anesthesia, significant levels of chloroform were still present
20 hours later.
Fry et al. (1972) has provided indirect evidence in man of the storage of
chloroform in body fat (Figure 4-4), while analysis of body fat of animals given
a single inhalation exposure or oral dosage demonstrated marked accumulation of
chloroform in this tissue (Taylor et al., 1974; Cohen and Hood, 1969) (Tables
4-5, 4-6, 4-7). Particularly pertinent are the observations of McConnell et al.
(1975), who demonstrated the occurrence of significant amounts of chloroform
(and other chlorinated hydrocarbons) in autopsy tissues (highest concentrations
in body fat) of humans exposed only to ordinary ambient air (Table 4-4), and of
Conkle et al. (1975), who analyzed the GC-MS alveolar air of eight fasting
healthy men working in a nonindustrial environment and found in three men
4-32
-------
"!;HCI?
PENTOBARBITAL
TIME.hr
Figure 4-5. Blood and adipose tissue concentrations of
chloroform during and after anesthesia in a dog. Note the high
prolonged levels of chloroform in adipose tissue (broken line)
for 20 hours even after rapid exponential fall in blood con-
centration (solid line) with termination of chloroform anes-
thesia after three hours.
Source: Chenoweth et al. (1962).
-------
significant rates of pulmonary excretion of chloroform (Table 4-10), as well as
other halocarbons (for example, methylene chloride, dichlorobenzene, methyl-
chloroform) in all eight men.
4.5. BIOTRANSFORMATION OF CHLOROFORM
4.5.1. Known Metabolites. The haloforms, and chloroform in particular, long have
been known to undergo extensive mammalian biotransformation. Zeller (1883)
clearly demonstrated an increased daily urinary inorganic chloride excretion
representing 25 to 60% of the dose in dogs given oral doses of chloroform (7 to
10 g in gelatin capsules). Eighty years later, Van Dyke et al. (1964), using
Cl-chloroform, confirmed in rats that the extra urinary inorganic chloride
originated from the metabolism of chloroform. Zeller (1883) also found, in the
urine of his dogs given chloroform, a levo-rotatory oxidative metabolite that he
suggested to be the glucuronide of trichloromethanol, a compound only recently
postulated as an intermediate of Pnc-n oxidative metabolism (Figure 4-6). Van
Dyke and coworkers (1964) also found evidence for (^metabolites (-2% dose) in
the urine of their rats given chloroform. However, other investigators (Brown
et al., 1974; Fry et al., 1972; Paul and Rubinstein, 1963; Butler, 1961) with
newer methodologies have not been able to identify lesser chloromethanes in the
urine or breath of mouse, rat, or man after chloroform exposure.
In addition to the chloride ion, it has been established from both in vivo
and in vitro studies that the major end-product metabolite of chloroform is
carbon dioxide (COp) (Brown et al., 1974; Fry et al., 1972; Rubinstein and
Kanics, 1964; Van Dyke et al., 1964; Paul and Rubinstein, 1963), with phosgene
identified as the immediate precursor metabolite from In vitro studies (Mansuy
et al., 1977; Pohl et al., 1977; Ilett et al., 1973) (Figure 4-6). C02 from
4-34
-------
TABLE 4-11
Levels of Chloroform in Breath of Fasted Normal Healthy Men*
Subject
A
B
C
D
E
F
G
H
Age
34
28
33
38
47
28
38
23
Chloroform excretion, |ig/hr
2.0
ND
ND
ND
11.0
ND
ND
0.22
*Source: Conkle et al., 1975
ND = Not detected
4-35
-------
MAJOR AEROBIC PATHWAY
H-C C13
P450, O2
NADPH
MICROSOMES
[HOCCI3]
ACCEPTOR
PROTEIN
I
CO —*
H2C - CH - CX- OH
S NH
\ /
C
it
0
2-OXOTHIAZOL1DINE-
4-CARBOXYLIC ACID
-HCI
O=C CI2
PHOSGENE
>
CYSTEINE
CONDENSATION
H2O
•»- 2 HCI + CO2
GLUTATHIONE
CONJUGATES?
MINOR ANEROBIC PATHWAY
CHCI3
ANEROBIC
NADPH
-i*~ P450 - Fe2+ : C CI2 + HCI
I +H2O
P450 - Fe2+ CO —* CO + 2 H Cl
REDUCED
MICROSOMES
Figure 4-5. Metabolic pathways of chloroform biotransformation.
(Identified CH Cl., metabolites are underlined.)
U-36
-------
chloroform metabolism is primarily excreted through the lungs, but a small per
centage (<1%) is incorporated into endogenous metabolites and excreted into the
urine as bicarbonate, urea, methionine, and other amino acids (Brown et al.,
1974). Carbon monoxide (CO) has also been identified as a very minor metabolite
of anaerobic chloroform metabolism (Figure 4-6), both from In vitro studies
(Ahmed et al., 1977; Wolf et al., 1977) and _in vivo animal studies (Anders
et al., 1978; Bellar et al., 1974).
In addition to chloroform metabolites that are excreted, phosgene and other
"reactive intermediates" of chloroform metabolism interact with and covalently
bind to tissue acceptors such as protein and lipids (Docks and Krisna, 1976;
Uehleke and Werner, 1975; Brown et al., 1974; Ilett et al., 1973; Cohen and Hood,
1969; Reynolds, 1967; Cessi et al., 1966).
The liver is the principal site of chloroform metabolism, although Paul and
Rubinstein (1963) and Butler (1961) found that rat kidney, adipose tissue, and
skeletal muscle also converted chloroform to CO (25, 0.8, and 8% of liver,
respectively) .
4.5.2. Magnitude of Chloroform Metabolism. Chloroform is metabolized to
differing extents in man and other animal species. Since chloroform and other
halogenated hydrocarbons are thought to produce pathological effects by
metabolism in target tissues to reactive intermediates that covalently bind to
macromolecules (Chapter 5), the total capacity to metabolize chloroform as well
as individual tissue sites of metabolism are important determinants of expected
interspecies differences in toxic susceptibility. This interdependence between
intensity of toxic response and metabolism, and interspecies differences in
magnitude of metabolism, are important considerations in extrapolation from
experimental animal to man (Reitz et al., 1978).
4-37
-------
Few studies have been made of the capacity of man to metabolize chloroform;
virtually no studies have been made of the phamacokinetic, endocrine, genetic,
and environmental factors modifying metabolism in man. The early studies of
Lehmann and Hasegawa (1910) on the retention of chloroform from inspired air
inhaled by three volunteers (4500 to 5000 ppm, average of 64$) (Table 4-3)
suggest that 36% of pulmonary uptake of chloroform in man is metabolized.
Similarly, a retention value of 67$, calculated from the data of Smith et al .
(1973) for patients inhaling 10,000 ppm chloroform during surgical anesthesia
(Figure 4-2), indicates 33% chloroform metabolism during inhalation exposure. A
similar estimate of the extent of chloroform metabolism during anesthesia has
been made by Feingold and Holaday (1977). These workers simulated, with a
computer, chloroform inhalation kinetics using a non-linear whole-body compart-
ment al model, and found that the percent of chloroform uptake metabolized was
30%. This rate of metabolism remained constant during 8 hours of anesthesia, and
continued for several days following termination of anesthesia, presumably from
chloroform stored during anesthesia in adipose tissue.
Fry et al. (1972) have investigated the metabolism of chloroform in man
after single oral doses. Isotopicall y labeled ^C-chloroform dissolved in
1.0 m£ olive oil/gelatin capsule was given to 12 healthy male and female volun-
teers (58 to 60 kg body weight) at doses of 0.1 to 1.0 g. For two of the
volunteers, pulmonary excretion of 13CO in expired air from the metabolism of
chloroform was serially connected over a 7.5 hour period and analyzed by mass
spectrometry. The results given in Table 4-8 show that 49 and 51$ of a 0.5 g dose
13 1 ?
was metabolized to COp. Pulmonary excretion of unchanged JC-chloroform
during a comparable period (8 hours) in a separate experiment with these two
subjects were 67 and 40$, respectively. No metabolites other than C02 (e.g.,
methylene dichloride, tetrachloroethane) were found in expired air, and chloro-
4-38
-------
form was not found in the urine. These results indicate that: (1) virtually all
of an oral chloroform dose (0.5 g) can be accounted for by pulmonary excretion of
C02 and unchanged chloroform, (2) metabolism of chloroform to CC^ is =50% of this
dose, and (3) absorption and metabolism are rapid and virtually complete within 5
hours as shown in Table 4-8, possibly because of first pass through the liver.
In this respect, the kinetics of metabolism of oral doses may differ substan-
tially from inhalation doses. Furthermore, the data of Table 4-8 suggest that
the fraction of the dose metabolized is dose-dependent. Thus, an oral dose of
0.1 g was completely metabolized (100/0, with no chloroform excreted unchanged
through the lungs; but for a 1.0 g dose, 65% was excreted and only 35% meta-
bolized. These results suggest that metabolism is rate-limited in man, since a
diminishing proportion of dose is metabolized with increasing dose (Table 4-8) .
In man, Chiou (1975) has shown that up to 3&% of an oral dose of chloroform
is metabolized in the liver, and up to about M% is excreted intact from the
lungs before the chloroform reaches the systemic circulation, an example of a
first-pass effect.
Animal experiments demonstrate a marked species difference in the metabo-
lism of chloroform. Early experiments in the 1960's by Paul and Rubinstein
(1963), Van Dyke et al . (1964), and Cohen and Hood (1969) with mice and rats
given C-chloroform indicated a minimal metabolism of chloroform (~W) occurred
in these species. More recent studies by Brown and his coworkers (Brown et al . ,
1974; Taylor et al . , 1974) have shown that mice and rats metabolize chloroform to
CO extensively (65 to 85%), and to a greater extent than non-human primates or
man. These investigators gave equivalent oral doses (60 mg/kg) of C-chloro-
14
form to mice (3 strains), rats, and squirrel monkeys, and determined C-labeled
chloroform or volatile metabolites and 1J4C02 in expired air, 1 C-radioactivity
in urine, and C-radioactivity remaining in animals at sacrifice 4 to 8 hours
4-39
-------
after dosing. Total recovery of C-chloroform radioactivity was excellent and
accounted for 93 to 98% of the administered dose. Their results are summarized
in Table 4-9. Using 1 C02 as a measure of the fraction of the chloroform dose
metabolized (intermediate chloromethane metabolites were not found in breath or
urine), mice metabolized 8555 of the dose, rats, 66%, and squirrel monkeys, 18%.
A further 2 to 8% of C-radioactivity (CO incorporated into urea, bicar-
bonate, and amino acids) were found in urine. They found no strain difference in
mice, or sex difference in mice, rats, or monkeys in capacity to metabolize
chloroform, or in tissue distribution and binding of metabolites, except for
mice, where kidney radioactivity concentration was greater in males than females
and lesser in livers of males than females (Table U-7). These findings of Brown
and coworkers of large interspecies differences for metabolism of chloroform and
marked sex differences in mice (but not other species) for tissue distribution
and covalent binding of intermediate metabolites to tissue macromolecules in
liver and kidney emphasize the difficulties and dangers of extrapolating studies
in lower animals to man (Reitz et al., 1978).
4.5.3- Enzymic Pathways of Biotransformation. It has been postulated for many
years (Docks and Krishna, 1976; Uehleke and Werner, 1975; Brown et al., 1974;
Ilett et al., 1973; Reynolds, 1967; Paul and Rubinstein, 1963) that a reactive
metabolite of CHC1- is responsible for its liver and renal toxicity in man (von
Oettingen, 1964; Conlon, 19&3) and experimental animals (Bhooshan et al., 1977;
Pohl et al., 1977; Ilett et al., 1973; Klaassen and Plaa, 1966), and possibly the
production of liver tumors in mice (Eschenbrenner and Miller, 1945a). For
example, when rats or mice are treated with C-chloroform, the extent of hepatic
necrosis parallels the amount of C-label bound irreversibly to liver protein
(Docks and Krishna, 1976; Brown et al., 1974; Ilett et al., 1973). Both necrosis
4-40
-------
and binding are potentiated by pretreatment of animals with phenobarbital, a
known inducer of liver microsomal metabolism, and inhibited by pretreatment with
the inhibitor piperonyl butoxide. Chloroform administration also decreases the
level of liver glutathione in rats Pretreated with phenobarbital, further
suggesting that a reactive metabolism is produced (Docks and Krishna, 1976; Brown
et al., 197*0. The results of in vitro studies with rat and mouse liver micro-
iti
somes support the in vivo observations by establishing that C-chloroform is
metabolized to a reactive metabolite which binds covalently to microsomal
protein (Bhooshan et al., 1977; Sipes et al., 1977; Uehleke and Werner, 1975;
Ilett et al., 1973). This metabolic process is oxygen dependent and appears to
be mediated by a cytochrome P..™ which is inducible by phenobarbitol (Sipes
et al., 1977; Uehleke and Werner, 1975; Ilett et al., 1973).
The demonstrations by Pohl et al. (1977) and Mansuy et al. (1977) of
carbonyl chloride (phosgene) formation from chloroform by rat microsomal pre-
parations suggest that phosgene may be the key causal agent for these toxic
effects. The finding of Weinhouse and collaborators (Shah et al., 1979) that
phosgene is also a reactive metabolic intermediate in the metabolism of carbon
tetrachloride emphasizes basic similarities in the metabolism and toxicities of
these two chloroalkanes. Figures 4-6 and 4-7 show for comparison the currently
proposed pathways of metabolism of chloroform and carbon tetrachloride.
Figure 4-6 indicates that the initial step in the metabolism of chloroform
involves the oxidation of the aliphatic carbon (H-C) to trichloromethanol by a
phenobarbital inducible cytochrome Ph,-0 (Sipes et al., 1977; Uehleke and Werner,
1975; Ilett et al., 1973). This metabolic step has been suggested by Mansuy
et al. (1977) and Pohl et al. (1977) as the precursor of phosgene formed by rat
microsomes in vitro from chloroform. Phosgene was confirmed as a metabolite by
reaction with cysteine to give 2-oxothiazolidine-4-carboxylic acid which was
4-'41
-------
ACCEPTOR
1
CCI
C CI3C C CI3
CHCI3
CCL4
REDUCTIVE
DECHLORINATION
ANAEROBIC
MICROSOMES
NADPH
P450 - Fe2+ • C CI3 + C\'
lP450-Fe3+.CI3COH]
ACCEPTOR
PROTEIN i
1 '
•HCI
•CCL2
H2O
• P450 - Fe2+C CI4
P450 - Fe2+ : C CI2 + CT
I H20
CO + HCOOH
LIPOPEROXIDATION
CONJUGATION
MALONALDEHYDE
2 H Cl
PHOSGENE
H,C-CH-COOH
I I
S NH
V
CYSTEINE
CONDENSATION
2-OXOTHIAZOLIDINE
4 • CARBOXYLIC ACID
Figure H-7. Metabolic pathways of carbon tetrachloride bio-
transformation. (C Cljj metabolites identified are underlined).
Source: Shah et al. (1979).
4-42
-------
identified by GC-CIMS. Trichloromethanol is highly unstable and spontaneously
dehydrochlorinates to produce phosgene (Seppelt, 1977). The electrophilic phos-
gene reacts with water to yield COp, a known metabolite of CHCl^ in vitro
(Rubinstein and Kanics, 1964; Paul and Rubinstein, 1963) and i_n vivo (Brown
et al., 1974; Fry et al., 1972), with protein to form a covalently bound product
(Pohl et al., 1977; Sipes et al., 1977; Uehleke and Werner, 1975; Brown et al.,
1974; Ilett et al., 1973), or with cysteine (Pohl et al., 1977), and possibly
with glutathione (Docks and Krishna, 1976; Brown et al., 1974). The finding that
deuterated chloroform (CDC1.J depletes glutathione in the livers of rats less
than CHC1 supports this notion (Docks and Krishna, 1976).
The postulated oxidation of the C-H bond of chloroform by P45Q to produce
trichloromethanol which spontaneously yields phosgene is further supported by
the observations of Pohl and Krishna (1978). These workers found that chloroform
metabolism to phosgene by rat liver microsomes is oxygen and NADPH dependent, and
1 R
inhibited by CO and SKF 525-A. Moreover, in the presence of cysteine and 0^
1 R
atmosphere, 0 is incorporated into the 2-oxo position of 2-oxothiazolidine-4-
carboxylic acid. Oxidative cleavage of the C-H bond appears to be the rate-
determining step, since deuterium labeled chloroform (CDC1,) is biotransformed
into phosgene slower than CHC1 • CDCl, appears also to be less hepatotoxic than
CHC1 Pohl (1980) has further characterized the metabolism of chloroform in rat
liver microsomes by measuring the covalent binding of CHC1., and C HC1-, to
microsomal protein. Chloroform does not appear to be activated by reductive
dechlorination to the radical •CHC12> because the 3H-label does not bind to
14
microsome protein as does the C-label.
Figure 4-7 summarizes current knowledge of the biotransformation of carbon
tetrachloride. The first step is a rapid reductive formation of the trichloro-
methyl ( -CCl-) radical by complexing with one or more of the PJICQ cytochromes
4-43
-------
(Shah et al . , 1979; Foyer et al . , 1978; Recknagel and Glende, 1973). This
radical undergoes several reactions in addition to binding to lipids (Villarruel
and Castro, 1975; Uehleke and Werner, 1975; Villarruel et al . , 1975; Gordis ,
1969; Reynolds, 1967) and protein (Uehleke et al . , 1977; Uehleke and Werner,
1973), although not to nucleic acids (Uehleke et al., 1977; Uehleke and Werner,
1975; Reynolds, 1967). Anaerobically, the addition of a proton and electron
yields chloroform (Glende et al . , 1976; Uehleke et al . , 1973; Fowler, 1969;
Butler, 1961), dimerization to hexachloroethane (Uehleke et al . , 1973; Fowler,
1969), or further reductive dechlorination to CO via the carbene, : CC (Wolf
et al., 1977). Aerobically, the «CC1 radical is oxidized by the P^Q system to
trichloromethanol (Cl-COH) , which is the precursor of phosgene (ClpCO) that
Weinhouse and colleagues (Shah et al . , 1979) have shown to be an intermediate in
carbon tetrachloride metabolism by rat liver homogenates. Hydrolytic
dechlorination of phosgene yields C0? (Shah et al . , 1979).
Under normal physiological conditions (i.e., aerobic conditions), a minimal
formation of chloroform might be expected to occur. Carbon tetrachloride yields
a chloroform most readily i_n vitro under anaerobic conditions and its formation
is inhibited by oxygen (Uehleke et al . , 1977; Glende et al . , 1976). Shah et al .
(1979) observed that chloroform does not compete successfully with carbon tetra-
chloride for initial binding to Ph,-n cytochrome (Sipes et al . , 1977; Recknagel
and Glende, 1973). Wolf et al. (1977) also found that binding of chloroform to
reduced cytochrome Pj.™ was very slow compared to that of carbon tetrachloride.
Anders and coworkers (Anders et al . , 1978; Ahmed et al . , 1977) have shown,
as they have for di ha lorn ethanes , that trihal omethanes , including chloroform,
also yield CO as a metabolite. Intraperitoneal administration of haloforms ( 1 to
4 mmoles/kg) to rats led to dose-dependent elevations in blood CO levels. Treat-
ment of the rats with either phenobarbital (but not 3-methyl-cholanthrene) or SKF
-------
525-A, respectively, increased or decreased metabolism to CO. The order of yield
of CO from the iodoforms was greatest for iodoform > bromoform > chloroform for
the same dose. Thus, chloroform was minimally metabolized to CO (i.e., to less
than one-tenth of that for iodoform or bromoform). Similar findings were made by
these workers with rat liver microsomes (Ahmed et al., 1977). Metabolism of the
haloforms to CO by rat liver microsomes required NADPH, could proceed anaerob-
ically but was increased 2-fold by 0~, was increased by pretreatment with pheno-
barbital and inhibited by SKF 525-A or COC12 pretreatment, and was stimulated by
glutathione or cysteine addition in both anaerobic (3-fold) or aerobic (8-fold)
conditions. These results suggested that haloforms were metabolized to CO via a
cytochrome PMJ-Q dependent system. However, chloroform was a poor substrate
compared to iodoform or bromoform, yielding <2% of its quantity of CO as formed
from equimolar concentrations of these halomethanes. Wolf et al. (1977) also
found that chloroform, to a very limited extent, was metabolized to CO by reduced
rat Pjjj-Q preparations. These workers investigated the spectral and biochemical
interactions of a series of halogenated methanes with rat liver microsomes under
anaerobic reducing conditions. Tetra- (e.g., CClj.) and trihalogens (e.g.,
CHC1 ) all formed complexes with reduced cytochrome PJ,™ with absorption peak at
460 to 465. A shift to 454 occurred with CO formation and subsequent complexing
of CO to Pjj50. CO formation required NADPH, was higher in microsomes from
phenobarbital and 3-methylcholanthrene-treated rats, and was not found at high
oxygen concentrations (<8%) . Figure 4-8 shows the relative rates of CO formation
from carbon tetrachloride and other polyhalomethanes. Chloroform, in comparison
to carbon tetrachloride, was a very poor reaction substrate, and binding of
chloroform to reduced cytochrome P^ was extremely slow compared to that of
carbon tetrachloride. Wolf et al. (1977) proposed the reduction sequence shown
in Figures 4-6 and 4-8 for the reductive dechlorination of chloroform and of
4-45
-------
120
100
o
D.
01
"o
c
2
O
t-
DC
O
LL
O
u
20 -
20
Figure U-8. Rate of carbon monoxide formation after addition of
various halomethanes to sodium dithionite-reduced liver micro-
somal preparations from phenobarbitol-treated rats. Note the low
rate of metabolism of chloroform to CO compared to carbon tetra-
ohloride.
Source: Wolf et al . (1977).
-------
carbon tetrachloride to yield CO via a carbene (CClg) intermediate. The physio-
logical importance of this pathway of metabolism appears to be more significant
for carbon tetrachloride than for chloroform.
4.6. COVALENT BINDING TO CELLULAR MACROMOLECULES
4.6.1. Proteins and Lipids. Reactive intermediates of the metabolism of chloro-
form (phosgene, carbene, *C1) and carbon tetrachloride (*CC1 phosgene, car-
bene, »C1) that irreversibly bind to cellular macromolecules (covalent binding)
are generally believed to result in an alteration of cellular integrity, which
leads to centrolobular hepatic necrosis and renal proximal tubular epithelial
damage. Chloroform, mole for mole, is generally accepted to be less hepatotoxic
than carbon tetrachloride (Brown, 1972; Klaassen and Plaa, 1969; Plaa et al.,
1958).
Chloroform in non-lethal doses produces renal damage in mice, dogs, and man;
(Bhoosan et al., 1977; Pohl et al., 1977; Ilett et al., 1973; Klaassen and Plaa,
1966, 1967; Bennet and Whigham, 1964; von Oettingen, 1964; Conlon, 1963; Plaa
et al., 1958; Culliford and Hewitt, 1957; Hewitt, 1956; Shubik and Ritchie, 1953)
whereas, in experimental animals, carbon tetrachloride does not do so (Storms,
1973; Klaassen and Plaa, 1966; Plaa and Larson, 1965; Bennet and Whigham, 1964;
Culliford and Hewitt, 1957; Hewitt, 1956; Shubik and Ritchie, 1953), although it
does in man (New et al., 1962; Guild et al., 1958). To explain these species
differences in toxicity as well as known intraspecies (Hill et al., 1975;
Deringer et al., 1953; Shubik and Ritchie, 1953) and sex differences (Taylor
et al., 1974; Ilett et al., 1973; Bennet and Whigham, 1964; Culliford and Hewitt,
1957; Hewitt, 1956; Deringer et al., 1953; Shubik and Ritchie, 1953;
Eschenbrenner and Miller, 1945b), prevailing concepts implicate (1) differences
4-47
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in the rater, of metabolism and organ system capacities for metabolism, which in
turn determines the amount of irreversible macromolecular binding and (2)
differences in the enzyme pathways for metabolism of the two haloalkanes (Figures
4-6 and 4-7).
Carbon tetrachloride, by a reductive dechlorination via complexing with
reduced P^, yields the trichloromethyl free radical (-CCl^) (Recknagel and
Glende, 1973; Slater, 1972) (Figure 4-7), which can covalently bind to lipid and
protein (Shah et al., 1979; Villarruel et al., 1975; Castro and Diaz Gomez, 1972;
Reynolds, 1967), and can also initiate peroxidation of polyenoic fatty acids
(Slater, 1972; Recknagel and Ghoshal, 1966). Chloroform does not appear to be
activated to free radicals («CC1 or »CHC1 ), but does bind covalently to liver
lipid and protein (Sipes et al., 1977; Docks and Krishna, 1976; Uehleke and
Werner, 1975; Brown et al., 1974; Ilett et al., 1973), and initiates lipid peri-
oxidation in some circumstances (Koch et al., 1974; Ilett, 1973; Brown, 1972;
Slater, 1972). Several investigators have shown that diene conjugates (products
of lipoperoxidation) are not increased .in vivo in normal rats when chloroform is
inhaled or injected intraperitoneally (Brown et al., 1974; Brown, 1972; Klaassen
and Plaa, 1969), but only when rats are pretreated with phenobarbital and the
metabolism of chloroform is greatly enhanced (Brown et al., 1974; Brown, 1972).
Studies iri vitro show diene conjugation and malonaldehyde formation (an index of
lipoperoxidation) by microsomes of phenobarbital pretreated rats were not
increased but decreased by the addition of chloroform (Brown, 1972; Klaassen and
Plaa, 1969), suggesting that with isolated microsomes, metabolism of chloroform
is too small for sufficient quantities of reactive intermediates to accumulate
and initiate lipoperoxidation. Howeve~, Rubinstein and Kanics (1964) found
chloroform to be more rapidly metabolized by rat microsomal fractions than carbon
AjM
tetrachloride (see^Table 4-11). These findings indicate that differences in
4-48
-------
TABLE 4-12.
Covalent Binding of Radioactivity From C-Chloroform and
14 i
C-Carbon Tetrachloride in Microsomal Incubation In Vitro
Incubation Microsomal
Condition Protein Lipid
nmol/mg in
1 C-CC14 N2 20.0 76.0
1J4C-CHC13 N2 5.1 4.1
02 8.5 7.0
To Added
serum albumin
60 minutes
1.4
0.9
1.7
Source: Uehleke et al., 1977
T^licrosomes from phenobarbital pretreated rabbits
4-49
-------
metabolic activation [carbon tetrachloride to produce free radicals (Figure
4-7), but chloroform primarily to phosgene (Figure 4-6)] explain the greater
potential of carbon tetrachloride for initiating lipoperoxidation. Table 4-11
shows the data of Uehleke et al. (Uehleke et al., 1977; Uehleke and Werner, 1975)
for the covalent binding of rabbit microsomes following incubation with 14C-
labeled chloroform and carbon tetrachloride. Both protein and lipid binding of
14
C-radioactivity are 4-fold and 20-fold, respectively, more extensive for
carbon tetrachloride than chloroform; lipids are labeled preferentially by
carbon tetrachloride but are not by chloroform. Furthermore, covalent binding
from chloroform metabolism occurs mainly with anaerobic conditions (a minor
metabolic pathway) and is not greatly increased with aerobic metabolism, the
major pathway for metabolism of chloroform, which is 0 dependent (Figure 4-6).
Covalent binding occurs preferentially to lipids and proteins of the endo-
plasmic reticulum proximate to P^ system for metabolism. However, consider-
able covalent binding from chloroform metabolites occurs in other cell fractions
of liver and kidney, particularly to mitochondria (Uehleke and Werner, 1975; Hill
et al., 1975). Hill et al. (1975) found when C57BL male mice were injected
interperitoneally with 0.07 raJl/kg C-chloroform in oil and sacrificed 1? hours
later, that in the liver, 50% of the radioactivity was irreversibly bound to
rnicrosome, 23% to mitochondria, 25% to cytosol, and <2% to nuclei; for kidney,
38% of radioactivity was bound to microsome, 39% to mitochondria, 22% to cytosol,
and <2% to nuclei. A similar distribution was found in male NMRI mice by Uehleke
and Werner (1975), who observed minimal binding to microsomal RNA but significant
binding to nicotine-adenine nucleotides. The data of Ilett et al. (1973), shown
in Figure 4-9, demonstrates that in C57 BL/6 mice, the amount of covalent binding
in liver and kidney microsomal fractions increases proportionally with the
chloroform dose.
4-50
-------
c
1
Q.
U>
"o
E
c
O
z
o
z
ffi
o
o
23456
CHLOROFORM DOSE, mmol/kg
Figure 4-9. Effect of increasing dosage of i.p.-injected
14
C-chloroform on extent of covalent binding of radioactivity in
vivo to liver and kidney proteins of male mice 6 hours after
administration.
Source: Ilett et al. (1973).
4-51
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4.6.1.1. GENETIC STRAIN DIFFERENCE — Hill et al. (1975) described in mice
two genetic variations in chloroform toxicity paralleling genetic differences in
covalent binding in liver and kidney. In one inbred strain (DBA/2), the male
animals were 4 times more sensitive to the lethal effects of oral doses of
chloroform (LD^ of 0.08 m£/kg) than the second strain (C57 EL/6, LD^ of
0.33 mK,/kg). Males of the F.) hybrid strain (B6D2F1/J) had an intermediate LD™
of 0.2 mjl/kg, midway between those of the two parental strains. The suscepti-
bility of DBA mice was related to a dose-dependent necrosis of the proximal
convoluted renal tubules. However, mice of all three genotypes that received
>0.17 mjl/kg chloroform exhibited both renal tubular necrosis and hepatic centro-
lobular necrosis. Males and females of the same strain exhibited similar dose
thresholds to hepatic damage, but females died of chloroform-induced hepatic
damage without developing renal lesions. This sex-related absolute difference
is dependent on androgen profile of the mice; testosterone-treated females
become sensitive to renal toxicity (Bennet and Whigham, 1964; Culliford and
Hewitt, 1957; Eschenbrenner and Miller, 1945b).
Table 4-12 shows the extent of covalent binding in liver and kidney of these
three strains after a single intraperitoneal injection of C-chloroform
(0.07 m£/kg) to the males. Kidney homogenates from DBA/2J male mice, more
sensitive to renal necrosis, contained more than 2-fold as much radioactivity as
those from resistant C57BL/6J; covalent binding in the F hybrid was interme-
diate, as expected. A significant difference was also noted in labeling of
kidney subcellular fractions. While all subcellular fractions of susceptible
male DBA mice were labeled to a greater extent than F1 or C57BL strains, the
greatest increase was in labeling of the mitochondrial fraction.
4-52
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TABLE 4-13
Mouse Strain Difference in Covalent Binding of Radioactivity
From C-Chloroform
a,b
Tissue
Liver
Kidney
Liver
Nuclei
Mitochondria
Microsoraes
Cell sap
Specific Activity
Relative to C57 BL
DBA
0.82
2.41
0.67
1.14
0.64
0.98
F1
Tissue homgenates
0.96
1.64
Subcellular fractions
0.76
1.14
0.73
1.09
C57BL
1.00
1.00
1.00
1.00
1 .00
1.00
Kidney
Nuclei
Mitochondria
Microsoraes
Cell sap
2.20
3.67
1.74
1.65
1.67
1.97
1.44
1.23
1.00
1.00
1.00
1.00
Source: Hill et al., 1975
14,
'Adult male mice of each genotype given ' C-chloroform (0.07 mJl/kg) intra-
peritoneally and sacrified 12 hours later. Genotype comparisons are given as
ratio of radioactivity to C57BL = 1.
4-53
-------
In the liver, the distribution of covalent binding was generally opposite to
that observed in the kidneys (Table 4-12), but neither liver homogenates nor
subcellular fractions showed significant strain differences.
4.6.1.2. SEX DIFFERENCE -- Kidneys of male mice are known to covalently bind
more C-chloroform radioactivity than do those of females, but females bind more
in the liver than males (Taylor et al., 1974; Ilett et al., 1973) (see Tables 4-7
and 4-13). Table 4-14 shows that pretreatment of male mice with phenobarbital
increases covalent binding in the liver but not in the kidney (Ilett et al. ,
1973). A similar observation has been made by Kluwe et al. (1978) in male mice.
They found that phenobarbital increased liver but not kidney microsomal acti-
vity; 3-methylcholanthrene, dioxin, and PCGs increased both liver and kidney
microsomal enzyme activities. From the renal and hepatic toxicity profile to
chloroform displayed by mice treated with these various inducers, these investi-
gators concluded that the chloroform metabolite(s) responsible for hepatic
damage is probably generated in the liver, and the metabolite(s) responsible for
renal damage is generated in the kidney.
4.6.1.3. INTER-SPECIES DIFFERENCE — In addition to intra-species strain
(mice) differences in covalent binding noted above, Uehleke and Werner (1975)
have also observed an apparent inter-species difference. Figure 4-10 shows the in
vitro binding of radioactivity from C-chloroform by microsomal preparations
from rat, mouse, rabbit, and man. Human and rabbit microsomes have the highest
rate of covalent binding from chloroform, with the mouse followed by the rat
considerably lower. Inter-species differences in the covalent binding rates for
carbon tetrachloride were small. These species differences in binding of chloro-
form metabolites to protein and lipid in vitro do not, however, parallel the
4-54
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TABLE 4-14
In Vivo Covalent Binding of Radioactivity From CHClo in
Liver and Kidney of Male and Female Mice (C57BL/6)a'b
Covalent Binding of C-Chloroform
nmoles/mg protein + S.E.
Liver
Kidney
Male
2.92 + 0.35
2.34 + 0.16
Female
3.66 + 0.39
0.39 + 0.02
aSource: Ilett et al., 1973
Mice were sacrificed 6 hours after intraperitoneal administration
of 3.72 nmoles/kg of CHC1.,.
4-55
-------
TABLE 4-15"
In Vitro Covalent Binding of Radioactivity from CHC1 to
3
Microsomal Protein from Liver and Kidney of Male and Female
Mice (C57BL/6)*
Male
Male
Female
Pretreatment
NA
Phenobarbital
NA
Covalent Binding of
p moles/mg protein/5
Liver
572 ± 54
1454 + 143
419 + 20
C-Chloroform
minutes + SEM
Kidney
44.6 + 4.1
41.0 + 3.2
14.6 + 2.5
"Source: Ilett et al., 1973.
NA = Not applicable
4-56
-------
0) T
cr o
cr 3
H-
W
C
T
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RADIOACTIVITY, nmol 14C from 14CHCI3 /mg
o
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CD
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-------
species differences in metabolism of chloroform in vivo, as measured by the
conversion of chloroform to CO (Table 4-9); in vivo, the mouse has the greatest
^ .1.. i i
capacity to metabolize chloroform (Q0%) , followed by rats (65%), nonhuman pri-
mates (20%), and man (30 to 50%).
4.6.1.4. AGE DIFFERENCE -- Uehleke and Werner (1975) have shown that irre-
versible protein binding of radioactivity from C-chloroform and C-carbon
tetrachloride to liver microsomes of newborn rats (18 hours old) is low compared
to that of microsomes from adult rats (32 days); however, the binding was shown
to be proportional to PH™ content of the microsomes, which was proportionally
low in microsomes from newborn rats.
4.6.2. Nucleic Acids. PJICQ systems activate chloroform and carbon tetrachloride
in vivo and in_ vitro to reactive metabolites that extensively covalently bind to
proteins and lipids, but do so only minimally to nucleic acid (Uehleke et al.,
1977; Wolf et al., 1977; Uehleke and Werner, 1975; Fowler, 1969; Reynolds, 1967),
unlike many other carcinogens that bind DNA. Reitz et al. (1980) measured DMA
alkylation in liver and kidneys of mice after an oral dose of 240 mg/kg
14 _4
C-chloroform (specific activity not given) and found values of 3 x 10 and 1 x
-4
10 mol % for liver and kidney DNA respectively. These workers judged that
chloroform has very little direct interaction with DNA when compared to known
carcinogens, as reported in the literature for dimethylnitrosamine (3.5 x 10~
p
mol % alkylation, liver DNA), dimethylhydralazine (2.6 x 10 , colon DNA) and
N-methyl-N-nitrosourea (1.5 x 10, brain DNA) but given by parenteral routes
(Pegg and Hui, 1978; Cooper et al., 1978; Kleihues and Margison, 1974). The
failure of chloroform or carbon tetrachloride reactive species to significantly
bind DNA has been ascribed to their short half-life compared to epoxides, and to
4-58
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their lack of nuclear penetration. Recently, however, Diaz Gomez and Castro
(1980) have shown that highly purified rat liver nuclear preparations are able to
anaerobically activate carbon tetrachloride, and to aerobically activate
chloroform to reactive metabolites that bind to nuclear lipids and proteins.
Their data, given in Table 4-15, show that activity in nuclear preparations is
smaller than in microsomes, but within the same order of magnitude. These
results might be relevant to the hepatocarcinogenie effects of chloroform and
carbon tetrachloride in mice and rats, since the nuclear targets (DMA, RNA,
nuclear proteins) are in the immediate neighborhood sites of activation, thus,
making unnecessary the present assumption that the highly reactive intermediates
(•CCl-j, phosgene, malonaldehyde, or carbene), produced at the endoplasmic
reticulum, must travel to the nucleus.
4.6.3- Role of Phosgene. Phosgene is a prominent intermediate of both chloroform
and carbon tetrachloride metabolisms (Figures 4-6, 4-7). It is known to be
highly reactive and toxic to cells and tissues (Pawlowsky and Frosolono, 1977),
and its two highly reactive chlorines suggest that it could act on cellular
macromolecules similar to bifunctional alkylating agents. Reynolds (1967)
14
showed that C-phosgene, given to intact rats, labeled liver protein (and lipids
to a smaller extent). The pattern of labeling was quite different from that of
C-carbon tetrachloride and more similar to C-chloroform. Moreover, ^ Cl-
ear bon tetrachloride radioactivity was also stably incorporated into liver lipid
and protein, pointing to the •CC1,, radical rather than phosgene as the reactive
form for carbon tetrachloride that labels lipid. Cessi et al. (1966) also
14
reported that C-phosgene labeled terminal amino group of polypetides in a
manner similar to in vivo protein labeling produced by carbon tetrachloride.
4-'5 9
-------
TABLE 4-16
Covalent Binding of Radioactivity from C-Chloroform and C-Carbon
Tetrachloride in Rat Liver Nuclear and Microsomal Incubation In Vitro*
Incubation
Condition
Protein
Lipid
p mol/mg + S.D.
C-CC1,
Nuclear
Microsomal
21.9 + 2.5
50.3 + 4
14? + 12
190 +11
C-CHC1,
Nuclear
Microsomal
27.0 + 3
68.0 + 9
20+3
57+8
*Source: Diaz Gomez and Castro, 1980
4-60
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4.6.4. Role of Glutathione. Ekstrom and Hogberg (1980) found that chloroform, in
freshly isolated rat liver cells, induced depletion of cellular glutathione.
Brown et al. (1974) demonstrated that exposure of rats to an atmosphere of 0.5%
chloroform for 2 hours markedly decreased glutathione (GSH) in the liver when the
animals were pretreated with phenobarbital to stimulate metabolism. GSH liver
content of untreated rats was not decreased. Phenobarbital pretreatment has been
shown to markedly potentiate toxicity of both chloroform and carbon tetra-
chloride in rats (Docks and Krishna, 1976; Cornish et al., 1973; McLean, 1970;
Scholler, 1970). However, it has not been possible to detect a decrease in the
liver glutathione levels following administration of carbon tetrachloride or
trichlorobromomethane (Docks and Krishna, 1976; Boyland and Chasseaud, 1970).
Sipes et al. (1977) have shown that the addition of GSH liver microsomes from
ill
phenobarbital pretreated rats incubated in vitro with C-labeled halocarbons,
chloroform, carbon tetrachloride, and trichloromomethane, inhibited covalent
binding =80$ for all three compounds. Their results, given in Table 4-16, also
show the effects of anaerobic and aerobic conditions on covalent binding. The
reduction in binding of chloroform by an atmosphere of N? suggests that its
bioactivation is mediated by a cytochrome P. oxidative pathway to phosgene
(Figure 4-6), while the enhanced binding of carbon tetrachloride in N~ reflects
Pj,j50 mediated reductive pathways (Figure 4-7) and formation of free radical.
These investigators suggest that in phenobarbital-treated animals, chloroform
depletes liver GSH by the formation of conjugate between the reactive
intermediate phosgene and GSH (Docks and Krishna, 1976; Brown et al., 1974). In
the case of carbon tetrachloride, they suggest that GSH addition in vitro (Table
4-16) also conjugates with the phosgene metabolite of carbon tetrachloride
produced when incubated in air, but, in addition, GSH decreases the levels of
•CC1, by reducing the free radical to chloroform. In vivo, it is postulated that
4-61
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TABLE 4-17
Effect of Glutathione, Air, N2 or CO: 02 Atmospere on the
IH Vitro Covalent Binding of C Cl^, CHC1,, and C Br Cl~ to Rat
Liver Microsomal Protein
Substrate
Incubation Conditions
CC1,
CHC1-
CBrCl.
Air
N2
C0:02(8:2)
SKF 525 A (0.5 mM), Air
Glutathione, Air
NADPH omitted, Air
ill
p moles C-bound/mg microsomal protein/minute
97+10 59+5 1456+66
310 +51 21+1 1370 + 143
18+1 20+1 853 + 62
109+5 7+7 2105 + 159
17+2 15+2 218 + 25
6+1 3+0 65+13
Source: Sipes et al., 1977
bl4 -^
C-labeled substrate is a final concentration of 1 x 10 JM incubated at 37°C
Microsomes were from phenobarbital pretreated rats.
4-62
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the oxidized glutathione may be reduced back to reduced GSH by glutatione
reductase; this would explain the lack of fall of liver GSH content with carbon
tetrachloride .in vivo (Gillette, 1972). Thus, the toxic effects of chloroform
and carbon tetrachloride may be mediated through different mechanisms of
covalent binding, and GSH may play different roles for these chlorocarbons in
preventing covalent binding of reactive intermediates of metabolism.
Chloroform and carbon tetrachloride are known to cause greater liver damage
in fasted animals than in fed animals (Diaz Gomez et al., 1975; Jaeger et al.,
1975; Krishnan and Stenger, 1966; Goldschmidt et al., 1939; Davis and Whipple,
1919). For chloroform, a decreased content of hepatic GSH from fasting has been
postulated to be responsible for the increased susceptibility of fasted mice
(Docks and Krishna, 1976; Brown et al., 1974). Nakajima and Sato (1979) have
recently offered an additional explanation. These investigators studied the
metabolism of the chlorocarbons In vitro with microsomes from livers of fasted
rats, and found that the disappearance of chloroform from incubation increased 3-
fold for a 24 hour fast, although fasting produced no significant increase in the
microsomal protein and cytochrome Pnc-n liver contents (Table 4-17). Similar
results were obtained for carbon tetrachloride. These observations suggest that
the increased toxicity of chloroform and carbon tetrachloride from food
deprivation may be due not only to decreased GSH, but also to a greater
production of reactive intermediates and covalent binding to cellular macro-
molecules .
SUMMARY
At ambient temperatures, chloroform is a volatile liquid with high lipid
solubility and appreciable solubility in water. Hence, chloroform is readily
absorbed into the body through the lungs and intestinal mucosa; the portals of
4-63
-------
TABLE 4-1$
Effects of 24-Hour Food Deprivation on Chloroform and Carbon Tetrachloride
.In Vitro Microsomal Metabolism, Protein, and P-450 Liver Contents of Rats*
Male
Female
Fed Fasted Ratio Fed
Metabolism, nmole/g/min
Chloroform
Carbon
tetrachloride
19.7 + 2.6 55.1 + 7.5 2.8
1.9 + 0.2 5.9 + 0.8 3.1
Protein content
27.7 + 3.7 23.0 + 2.7 NR
P-450, nmol/mg
0.842+0.123 0.823+0.03 NR
15.3 + 6.8
1.1 + 0.5
, mg/kg liver
22.5 + 1.5
protein
0.638 + 0.051
Fasted Ratio
39.3 + 2.5 2.6
4.5 + 0.3 4.1
23.7+1.7 NR
0.673 + 0.044 NR
*Source: Nakajiraa and Sato, 1979
NR = Not reported
-------
entry with exposure from air, water and food. Few data are available on the
pharmacokinetics of absorption and excretion of chloroform in man, particularly
at the low exposure concentrations expected in ambient air and drinking water.
However, studies show absorption from the gastrointestinal tract in man,
monkeys, rats and mice is rapid and complete, occurring by first-order passive
absorptive processes. A dose-dependent first-pass effect with pulmonary
elimination of unchanged chloroform occurs with oral ingestion in man, thus
decreasing the amount of chloroform reaching the systemic circulation. In rats,
the kinetics of peroral absorption are also influenced by the dosing vehicle; the
absorption rate is decreased for chloroform given in corn oil vehicle as compared
to an aqueous solution. Pulmonary uptake and elimination occur also by first-
order diffusion processes with three distinct components with rate constants
corresponding to tissue loading or desaturation of at least three major body
compartments. Half-times in man have been found to be approximately 14-30
minutes, 90 minutes and 24-36 hours, respectively. The longest half-time is
associated with the lipids and the adipose tissue compartment. During inhalation
exposure, at equilibrium with inspired air concentration, the blood/air
partition coefficient is about 8 at 37°C and the adipose tissue/blood partition
coefficient is 280 at 37°C. The quantity of chloroform absorbed is dependent
also on body weight and fat content of the body.
Tissue distribution of chloroform is consistent with its lipophilic nature
and modest water solubility. This chloroalkane readily crosses the blood brain
and placental barriers and distributes into breast milk. Concentrations
occurring in all major tissue organs are dose related to inspired air
concentrations or to oral dosage. Relative tissue concentrations occur in the
order of adipose tissue > brain > liver > kidney > blood.
4-65
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Elimination of chloroform from the body occurs by two major and parallel
occurring processes: 1) pulmonary elimination of unchanged chloroform by first
order kinetics, and 2) metabolism of chloroform. Chloroform is metabolized in
the liver, and to a lesser extent in the kidneys and other tissues. Metabolism
is dose-dependent and saturable, with a greater proportion of small doses being
metabolized. There are striking differences in the pharaacokinetics and
quantitative metabolism of chloroform in man as compared to other animals. For
large steady-state body burdens, 30-40/& is metabolized by man, 20% by the
nonhuman primate, > 65% by the rat, and > 85% by the mouse. Metabolism produces
phosgene and other putative reactive metabolites that covalently bind
extensively to cellular lipids and proteins, although not significantly to DNA or
other nucleic acids. The intensity of metabolite binding and organ localization
parallel the acute cellular toxicity of chloroform in liver and kidney observed
in experimental animals. Both binding and toxicity are highly dependent on
animal species and genetic strain, as well as on sex and age. An additional
variable is the tissue level of reduced glutathione which plays an important role
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and renal P450 metabolizing systems increase binding and toxicity.
4-66
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5. TOXICITY
5.1 EFFECTS OF ACUTE EXPOSURE TO CHLOROFORM
In both humans and experimental animals, characteristic effects of acute
exposure to chloroform are depression of the central nervous system and hepatic
damage. Renal and cardiac effects also occur. The systemic toxic effects of
chloroform appear to be similar regardless of whether exposure or administration
occurred by inhalation, oral, or parenteral routes. The only systemic effect
documented for dermal administration, however, is renal damage.
5.1.1 Humans
5.1.1.1 ACUTE INHALATION EXPOSURE IN HUMANS — Information on the
effects of acute inhalation exposure of chloroform on humans has been obtained
primarily during its use as an inhalation anesthetic. The relationship of the
concentration of chloroform in inspired air and blood to anesthesia is described
in Table 5-1 (Goodman and Oilman, 1980). Concentrations of chloroform used for
the induction of anesthesia were in the range of 2-3 volumes % (20,000-40,000
ppm), followed by lower maintenance levels (NIOSH, 1974; Adriani, 1970).
Chloroform inhalation has a depressive effect on the central nervous
system. Excitement due to release of inhibitions is followed by progressive
depression of the cortex, higher centers, medulla and spinal cord (Wood-Smith and
Stewart, 1964). Centers controlling temperature regulation, respiration,
vomiting, vasomotor, and vagal activity are all depressed (Adriani, 1970).
The cardiovascular system is also affected by anesthetic use of chloroform.
The myocardium is directly depressed in deeper planes of anesthesia. A blood
level sufficient to cause respiratory failure may also cause cardiac arrest. In
addition, chloroform sensitizes the autonomic tissues of the heart to
epinephrine, causing arrhythmias. It has been found that under chloroform
5-1
-------
TABLE 5-1
Relationship of Chloroform Concentration in Inspired Air and Blood to Anesthesia*
In Inhaled Air In Blood
Volumes %
*Source: Goodman and Oilman, 1980
Not sufficient for anesthesia
Light anesthesia
(after induction)
Deep anesthesia
Respiratory failure
<0.15
0.15 to 0.20
0.20 to 1.50
2.0
<2
2 to 10
10 to 20
20 to 25
5-2
-------
anesthesia regarded as normal, the heart is subject to arrythmias and extra-
systoles (Kurtz et al., 1936; Orth et al., 1951). Orth et al. (1951) found a high
incidence of ventricular arrhythmias, 20 of 52 cases investigated, and four cases
of temporary cardiac arrest. Blood pressure is lowered by chloroform as a result
of a 3-fold action: cardiac slowing due to vagal stimulation, depression of the
vasomotor center, and dilation of splanchnic blood vessels (Krantz and Carr,
1965).
Respiratory effects of chloroform inhalation include increased rate and
depth of respiration during induction and in light anesthesia, and decreased
minute volume exchange in deeper planes of anesthesia. The Hering-Breuer reflex
remains active. Bronchial smooth muscle is relaxed and secretions are increased.
Laryngeal spasms are caused by high concentrations (Adrian!, 1970).
In the gastrointestinal tract, chloroform markedly stimulates the flow of
saliva during induction and recovery, but salivation is inhibited in deeper
planes of anesthesia (Goodman and Oilman, 1980). The pharyngeal or gag reflex is
depressed. Under anoxic conditions, pharyngeal muscle spasms result in
stertorous respiration and thick mucus is excreted (Adriani, 1970). Stomach
movements are decreased or abolished as tone is reduced. Gastric secretory
activity is inhibited or abolished. Post-anesthetic dilation of the stomach
occurs in nearly all cases. Nausea and vomiting often occur during recovery from
anesthesia. The mechanism is central rather than local, but may be due in part
to irritation of the stomach by swallowed vapor (Goodman and Gilman, 1980).
Intestinal tone, motility, and secretory activity are inhibited or abolished
(Adriani, 1970).
In the urinary tract, chloroform anesthesia results in a decrease in urine
flow, possibly due to the release of antidiuretic hormone and renal vasoconstric-
tion, leading to a decrease in renal blood flow and glomerular filtration.
5-3
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Polyuria occurs after recovery (Goodman and Oilman, 1980). Chloroform
anesthesia may be followed by albuminuria and glycosuria. Post-operative urine
retention occurs frequently. Renal tubular necrosis has been found in cases of
severe poisoning (Wood-Smith and Stewart, 1964).
During obstetric use of chloroform, uterine contractions are only slightly
decreased by light anesthesia, but are markedly inhibited in deeper planes.
Chloroform rapidly crosses the placental barrier, and respiratory depression in
the infant is likely to occur (Wood-Smith and Stewart, 1964).
Chloroform anesthesia also has metabolic effects in humans. A rise in blood
glucose accompanies anesthesia. Levels may rise >2-fold and remain elevated for
several hours. The liver glycogen falls coincident with the rise in blood sugar.
This, in turn, is a result of the release of epinephrine from the adrenal medulla
during the period of excitation. There is also a decrease in glucose utilization
in the periphery (Goodman and Oilman, 1980; Krantz and Carr, 1965). Acidosis
occurs, characterized by a fall in plasma biocarbonate and phosphate.
Chloroform is acutely toxic to the liver, although in so-called delayed
chloroform poisoning, the full effects of damage done during and shortly after
administration are not seen for 24-48 hours. The glycogen content of the liver
is rapidly depleted; three-fourths in the first half-hour and less rapidly there-
after. There is centrilobular and, in severe cases, mid-zonal and massive
necrosis. Cells which survive show fatty degeneration. Symptoms include
progressive weakness, prolonged vomiting, delirium, coma, and death. They
develop from the first to the third day after exposure. Jaundice, increased
serum bilirubin, bile in the urine, reduction in liver function, increased
nitrogen excretion, lowered blood prothrombin and fibrinogen, and the appearance
of leucine, tryosine, acetone, and diacetic acid in the urine are some of the
more prominent findings. The hemorrhagic tendency is due to reduced prothrombin
5-4
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formation by the injured liver. Death usually occurs on the fourth or fifth day,
and autopsy reveals degeneration and necrosis of liver tissue, most marked around
the central veins (Goodman and Oilman, 1980; Wood-Smith and Stewart, 1964).
Hematologic effects due to acute chloroform inhalation are seen during
anesthesia. Erythrocytes are increased in number as the spleen is constricted
and red blood cells are extruded into the circulation. Leukocytes are increased
in number during the post-anesthetic period, reaching a maximum within 24 hours
and returning to normal in 48 hours. There is an increase in polymorphonuclear
cells. Platelets remain unchanged. One half-hour after exposure, there is a
decrease in clotting time. Prothrombin time is increased. Prothrombin synthesis
is impaired by liver toxicity as previously noted (Adriani, 1970).
The effects of chloroform on the eye include dilation of the pupils, with
reduced reaction to light as well as reduced intraocular pressure (Sax, 1979;
Winslow and Gerstner, 1978).
Signs of chloroform poisoning include a characteristic sweetish odor on the
breath, cold and clammy skin, and dilated pupils (Winslow and Gerstner, 1978).
Nausea and vomiting commonly occur. Ketosis, due to incomplete oxidation of
fats, as well as a rise in blood sugar, accompanies chloroform intoxication.
Initial excitation alternating with apathy is followed by prostration,
unconsciousness, and possible death due to cardiac and central nervous system
depression (Winslow and Gerstner, 1978).
The above discussion presents observations made on the effects of chloro-
form inhalation during general anesthesia. Information on the effects of experi-
mental acute inhalation exposure of chloroform in humans is limited to the work
of Lehman and Hasegawa (1910) and Lehman and Schmidt-Kehl (1936) as reviewed by
NIOSH (1974). The duration of exposure was <30 minutes and only the subjective
5-5
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responses of the subjects were measured. The dose-response relationships as
tabulated by NIOSH (1974) are presented in Table 5-2.
5.1.1.2 ACUTE ORAL EXPOSURE IN HUMANS -- Case reports of suicides
(Piersol et al., 1933; Schroeder, 1965) and of recreational abuse (Storms, 1973)
of chloroform present some information on the effects of acute imbibition. A
fatal dose of Ingested chloroform may be as little as one-third of an ounce (10
m£) (Schroeder, 1965). The initial effect is usually unconsciousness and
possibly death (within 12 hours without treatment) due to respiratory or cardiac
arrest. If the patient survives, delayed effects are observed within 48 hours
after regained consciousness. These symptoms include vomiting, anorexia,
jaundice, liver enlargement, albimuria, ketosis, ketonuria and glucosuria,
hemorrhage due to lowered blood fibrinogen and prothrombin, reduced serum
bicarbonate, increased blood sugar, coma and possible death. Upon autopsy,
extensive hepatic centrilobular necrosis is evident.
5.1.1.3 ACUTE DERMAL AND OCULAR EXPOSURE IN HUMANS — Chloroform is
absorbed through the intact skin (von Oettingen, 1964). Application of chloro-
form to the skin is followed after 3 minutes by a pungent and burning pain
reaching its maximum after 5 minutes, associated with erythema, hyperemia, and
finally vesication (Oettel, 1936). Exposure of the eye to concentrated chloro-
form vapors causes a stinging sensation. Splashing the substance into the eyes
evokes burning, pain, and redness of the conjunctival tissue. The corneal
epithelium is sometimes impaired; however, regeneration starts rapidly and leads
to full recovery within 1-3 days (Winslow and Gerstner, 1978).
5.1.2 Experimental Animals
5.1.2.1 ACUTE INHALATION EXPOSURE IN ANIMALS — Tolerance of animals to
chloroform has been summarized by Lehmann and Flury (1943) and by Sax (1979).
Similar central nervous system effects are seen in animals at approximately the
5-6
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TABLE 5-2
Dose-Response Relationships*
160 ppra (0.8 mg/fl,) for unspecified time - no odor
205 ppra (1.0 mg/il) for unspecified time - light transient odor
390 ppm (1.9 mg/&) for 30 minutes -light transient odor
920 ppm (4.5 mg/3,) for 7 minutes - stronger, lasting odor; dizziness, vertigo
after 3 minutes
680 ppm (3.3 mg/£) to 1000 ppm (5.0 mg/£) for 30 minutes - moderately strong
odor; taste
1100 ppm (5.4 mg/&) for 5 minutes - still stronger, permanent odor; dizziness,
vertigo after 2 minutes
1400 ppm (6.6 mg/£) to 1800 ppm (8.57 mg/£) for 30 minutes - stronger odor,
tiredness, salivation, giddiness, vertigo, headache, taste
3000 ppm (14.46 mg/£) for 30 minutes - all above plus pounding heart, gagging
4300 ppm (20.8 mg/£) to 5000 ppm (25 mg/S,) for 20 minutes - dizziness and
light intoxication
5100 ppm (25 mg/]i) for 20 minutes - dizziness and light intoxication
7200 ppm (35.3 mg/&) for 15 minutes - dizziness and light intoxication as
above but more pronounced
•Source: Lehman and Hasegawa (1910) and Lehman and Schmidt-Kehl (1936).
b-/
-------
same magnitude of exposure that produced these effects in humans. In mice,
exposure to 2500 ppm for 2 hours produced no obvious effects, 3100 ppm for 1 hour
produced slight narcosis, while 4000 ppm induced deep narcosis within one-half
hour. Only slight symptoms are seen at 2000-6000 ppm for longer exposures.
Fatal exposures were 4100-8200 ppm for mice, 12,300 ppm for rabbits, and
16,300-20,500 ppm for guinea pigs (duration of fatal exposures not specified).
In cats, exposure to 7200 ppm resulted in disturbance of the equilibrium after 5
minutes, light narcosis after 60 minutes, and deep narcosis after 93 minutes of
exposure. Exposure to 21,500 ppm produced disturbances in equilibrium after 5
minutes, light narcosis after 10 minutes, and deep narcosis in cats after 13
minutes of exposure.
Kylin et al. (19&3) described the effects of a single exposure of mice to
100, 200, 400, or 800 ppm of chloroform for 4 hours. The mice exposed to 100 ppm
did not develop demonstrable liver necroses, although moderate fatty infiltra-
tion of the liver was noted. In mice exposed to 200 ppm, some necrotic areas
appeared in the liver and there was an increase in serum ornithine-carbamyl
transferase. Exposure to chloroform at 400 and 800 ppm resulted in increased
hepatic necrosis and serum enzyme activity.
More recent data regarding toxic effects of acute inhalation exposure to
chloroform were presented by Wood et al. (1982), although the study was designed
primarily to investigate the role of hydrogen bonding in the anesthetic
mechanism. Groups of mice in a rotating cage were given a single exposure of
upto 3 hours of varying concentrations of chloroform or deuterated chloroform,
each concentration being held constant for about 20 minutes and then being raised
until the mice had lost their righting reflex. The concentration was then
lowered to 1/2 the ED,-n (=1500 ppm) where it remained until the mice had regained
their righting reflex. The duration of these manipulated exposures never
5-8
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exceeded 3 hours. Only 4 of 47 mice given chloroform gained the righting reflex;
indeed, some mice died or were comatose. Upon histological examination of
animals sacrificed 3-6 hours after exposure, mild hepatic centrilobular necrosis
and very mild renal tubular necrosis was observed. The animals receiving
deuterated chloroform survived for 24 hours, after which they were sacrificed.
The liver and kidney lesions in these mice were more severe, perhaps owing to the
longer survival time, which may have allowed these lesions to develop.
5.1.2.2 ACUTE ORAL EXPOSURE IN ANIMALS
Kimura et al. (1971) performed acute oral toxicity studies in newborn (5-8
g), 14-day-old (16-50 g), young adult (80-160 g), and older adult (360-470 g)
rats. The chloroform was given in undiluted form to unfasted rats. LDp... values
in mS,/kg (95% confidence limits) were reported as follows: 14-day-old, 0.3
(0.2-0.5); young adult, 0.9 (0.8-1.1), and older adult, 0.8 (0.7-0.9). The young
and older adult rats were males; the other two groups contained rats of both
sexes. When compared with 15 other solvents included in this study, the LDj-0
values for chloroform were the lowest in the two adult groups and next to lowest
in the 14-day-old rats. Only a rough approximation of the LD^ could be obtained
for the newborn rats; volumes of 0.01 m&/10 kg body weight were generally fatal.
Lower volumes could not be measured with any degree of accuracy and were not
attempted.
Torkelson et al. (1976) reported an oral LD of 2.0 g/kg (1.05-3.80) in
male rats. Animals receiving as little as 0.25 g/kg showed adverse effects.
Other recent studies of acute oral toxicity have reported LD values (with 95/t
confidence limits) of 1120 mg/kg (789-1590) in ICR male mice and 1400 mg/kg
(1120-1680) in females (Bowman et al., 1978), and 908 mg/kg (750-1082) and 1117
mg/kg (843-1514), respectively, in male and female rats (Chu et al., 1980).
5-9
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In a study that compared the toxicities of halogenated hydrocarbons, a
single oral dose of 60 mg/kg chloroform to mice had no toxic effect (Hjelle et
al., 1982).
Hill (1978) performed experiments in mice designed to study variability in
susceptibility to chloroform toxicity from single oral doses based upon genetic
sex differences. For three strains of mice, the LD values (mS,/kg) were:
DBA/2J, 0.08; B6D2F1/J, 0.20; and C57BL/6J, 0.33. The animals more sensitive to
chloroform-induced death were also found to be more susceptible to renal
toxicity. Males were found to be more sensitive to renal damage and death than
were females. This difference was related to testosterone and it was further
noted that C57BL/10J males are relatively testosterone deficient in comparison
to DBA males. The C57BL/6J strain used in the Hill (1978) study is closely
related to the C57BL/10J strain and may, therefore, also be testosterone
deficient.
In male B6C3F1 mice, severe diffuse renal necrosis occurred after a single
oral dose of 240 mg/kg and focal tubular regeneration occurred after a single
dose of 60 or 240 mg/kg. These effects were not seen after 15 mg/kg (Reitz et
al., 1980). Liver damage (hepatocellular necrosis and swelling with inflamma-
tory cell infiltration) occurred only at the highest doses.
Chu et al. (1982a) studied the effects of acute oral exposure of chloroform
on Sprague-Dawley rats. Groups consisted of 10 male and 10 female animals given
a single oral dose of 0, 546, 765, 1071, 1500, or 2100 mg/kg of chloroform in a
volume of 5 m£/kg corn oil. Clinical signs of toxicity included depression and
coma, but the authors did not specify whether these signs occurred at all dose
levels. Treated rats surviving for 14 day.3 consumed less food and had depressed
growth rates. Gross examination revealed increased liver and kidney weights at
1071 mg/kg. Upon comprehensive histological examination, only mild to moderate
5-10
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lesions, even at high doses, were observed in these organs. No changes were
noted in other organs, including brain and heart. The hepatic and renal lesions
were characterized by hepatocyte variations and occasional vesiculation of
biliary epithelial nuclei in the liver, and by bilateral focal interstitial
nephritis and fibrosis in the kidney. Changes in hematological and biochemical
parameters were also observed in the 1071 and 1500 rag/kg treated groups. Choles-
terol levels increased while lactate dehydrogenase activity and liver protein
levels decreased. In female rats, the activity of microsomal aniline hydroxylase
was induced by chloroform exposure. The numbers of lymphocytes were reduced in
both males and females, as were hemoglobin and hematocrit values. Upon longer
exposures of 5, 50, or 500 ppm chloroform in drinking water for 28 days, the only
toxic effect observed was a decreased number of neutrophils in the highest
dose-exposed rats. Examinations were performed as in the single-dose experi-
ment .
5.1.2.3 ACUTE DERMAL AND OCULAR EXPOSURE IN ANIMALS — Torkelson et al.
(1976) found that chloroform, when applied to the skin of rabbits, produced
slight to moderate irritation and delayed healing of abraded skin. When applied
to the uncovered ear of rabbits, slight hyperemia and exfoliation occurred after
one to four treatments. No greater injury was noted after 10 applications. One
to two 24-hour applications, on a cotton pad bandaged on the shaven belly of the
same rabbits, produced a slight hyperemia with moderate necrosis and a resulting
eschar formation. Healing appeared to be delayed on the site as well as on
abraded areas which were also covered for 24 hours with a cotton pad soaked in
chloroform.
Single application of either 1.0, 2.0, or 3.98 g/kg for 24 hours under an
impermeable plastic cuff held tightly around the clipped bellies of each of two
rabbits did not result in any deaths. However, extensive necrosis of the skin
5-11
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and considerable weight loss occurred at all levels. All animals were sacrificed
for study 2 weeks after exposure. All treated rabbits exhibited degenerative
changes in the kidney tubules graded in intensity with dosage levels. The livers
were not grossly affected.
In the same study (Torkelson et al., 1976), liquid chloroform, dropped into
the eyes of three rabbits, caused slight irritation of the conjunctiva which was
barely detectable 1 week after treatment. In addition, slight but definite
corneal injury occurred, as evidenced by staining with fluorescein. A purulent
exudate occurred for >2 days after treatment. Although started 30 seconds after
instilling the chloroform, thorough washing of one eye of each rabbit with a
stream of running water for 2 minutes did not significantly alter the response in
the washed eyes from that of the unwashed eyes.
5.1.2.4 INTRAPERITONEAL AND SUBCUTANEOUS ADMINISTRATION IN ANIMALS —
The toxicity of chloroform in mice after subcutaneous administration (Kutob and
Plaa, ig62b) and intraperitoneal administration (Klaassen and Plaa, 1966) has
been compared with that of other halogenated hydrocarbons (Pohl, 1979). In these
studies, the LD values for carbon tetrachloride, chloroform, and dichloro-
methane were 200, 27.5, and 76 mraol/kg after subcutaneous administration and 20,
14, and 23 mmol/kg when given intraperitoneally. When the relative hepato-
toxicity of these compounds was compared, a subcutaneous dose of 0.5 mmol/kg of
carbon tetrachloride produced approximately the same degree of liver damage as
6.2 mmol/kg of chloroform. After intraperitoneal administration, these values
dropped to 0.01 mmol/kg for carbon tetrachloride and 2.3 mmol/kg for chloroform.
Dichloromethane did not cause significant histological changes in the liver by
either route of administration. At doses that produced liver toxicity, chloro-
form caused kidney lesions which ranged from the presence of hyaline droplets,
5-12
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nuclear pycnosis, hydropic degeneration, and increased eosinophilia, to necrosis
with karyolysis and loss of epithelium of the convoluted tubules.
Ilett et al. (1973) found that intraperitoneal administration of chloroform
caused centrilobular hepatic necrosis in mice of both sexes, whereas renal
necrosis was observed only in male mice.
5.2. EFFECTS OF CHRONIC EXPOSURE TO CHLOROFORM
A characteristic effect of chronic exposure to chloroform is hepatic
damage; this effect has been documented primarily in studies with experimental
animals. As was the case for acute exposure to chloroform, hepatic damage in
chronic studies results from either inhalation or oral administration of this
chemical. Effects on the kidneys and thyroids have also been observed in some
experiments. This section will discuss both subchronic (=90 days) and chronic
exposure studies, because many of the subchronic studies were preliminary
range-finding tests for the chronic studies.
5.2.1. Humans
5.2.1.1 CHRONIC INHALATION EXPOSURE IN HUMANS — Only two chronic
inhalation studies that reported measurements of exposure concentrations, as
well as effects on human health, were found. Neither study (Challen et al.,
1958; Bomski et al., 1967) is particularly adequate or recent.
Challen et al. (1958) investigated complaints of workers (mainly women) in a
plant manufacturing lozenges that contained chloroform as a principle
ingredient. Before exhaust ventilation was installed, 9 of the 10 exposed
workers had complained of symptoms of tiredness, dull-wittedness, depression,
gastrointestinal distress, and frequent and scalding urination. Breathing-zone
monitoring during simulation of "pre-ventilation" working conditions suggested
that the emplyees had been exposed to =77-237 ppm. Discussions with management
revealed that some of these workers had occasionally been observed to behave in a
5-13
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silly manner or to stagger about during the workday. Another group of workers (N
= 10) had been exposed primarily to concentrations of 22-71 ppm. Eight of these
workers complained of less severe symptoms. Apparently, both groups of workers
had been exposed to occasional peak concentrations of =1163 ppm lasting 1.5-2
minutes. At least four workers in each group worked half-time. None of the five
controls reported symptoms similar to those reported by the exposed workers.
Eight of the higher exposure employees (77-237 ppm for 3-10 years, followed by =2
years without exposure), nine of the lower exposure employees (22-77 ppm for
10-24 months), and five unexposed anployees submitted to physical examinations,
including liver function tests (thymol turbidity, serum bilirubin, and urine
urobilinogen tests). These examinations and tests revealed no evidence of any
organic lesion, including liver damage, attributable to exposure to chloroform.
In humans, hepatic damage is the most common toxic effect of acute exposure
to chloroform, as noted previously. According to Pohl (1979), only one report of
liver abnormalities in humans after chronic exposure to chloroform has been found
in the literature and no additional reports were found in the more recent litera-
ture. In this study (Bomski et al., 1967), 17 cases of hepatomegaly were found
in a group of 68 industrial workers who were exposed to chloroform in concentra-
tions ranging from 2-205 ppm for 1-4 years. These were unknown rather than
breathing zone concentrations. Three of the 17 workers with hepatomegaly were
judged to have toxic hepatitis on the basis of elevated serum enzymes. The
frequency of viral hepatitis among the 68 chloroform-exposed workers was higher
(4.4? versus 0.38?) than the frequency among a group of inhabitants of the city,
XI8 years of age. This phenomenon also occurred in the 2 previous years. Ten
cases of splenomegaly were also diagnosed among the 68 workers. There appears to
be no comparison with incidences of these conditions in nonexposed workers.
5-14
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5.2.1.2 CHRONIC ORAL EXPOSURE IN HUMANS — There are relatively few
reports of toxic effects following chronic ingestion of chloroform (Pohl, 1979),
and more recent reports were not located in the literature. In one case, it was
estimated that a male patient ingested 1.6-2.6 g of chloroform in a cough
medicine daily for =10 years (Wallace, 1950). Blood and urine analyses, as well
as liver function tests, indicated the individual suffered from hepatitis and
nephrosis. Another report described three patients addicted to chlorodyne, a
tincture containing chloroform and morphine (Conlon, 1963). Liver biopsy showed
severe cellular damage in one of these individuals who had ingested 21 rnS, of
chloroform daily for an undetermined period of time. All three displayed
evidence of serious mental and physical deterioration, including peripheral
neuropathy. It is not possible to determine if the adverse effects were due to
chloroform, morphine, or ethanol.
More recently, the safety of a dentifrice containing 3>^% chloroform and a
mouthwash containing 0.43^ was assessed in studies lasting >_1 year (DeSalva et
al., 1975). The subjects using the dentifrice were exposed to =70 mg (0.0^7 mfi,)
of chloroform each day, whereas the groups using the mouthwash were exposed to
=178 mg (0.12 m&). The results of liver function tests and blood urea nitrogen
determinations showed no statistical differences between control and experi-
mental subjects.
Epidemiologic studies of humans exposed to chloroform in their drinking
water have focused on carcinogenic endpoints, and, hence, are discussed in the
chapter on carcinogenicity.
5.2.2 Experimental Animals
5.2.2.1 CHRONIC INHALATION EXPOSURE IN ANIMALS — Experiments with
several species of animals (Torkelson et al., 1976) give some information
regarding potential effects of long-term inhalation exposure to chloroform. The
5-15
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animals were exposed to chloroform 5 days/week for 6 months. Exposure to 25 ppm
of chloroform for up to 4 hours/day had no adverse effects in male rats as judged
by organ and body weights, and by gross and histological examination of livers
and kidneys. Exposure to 25 ppm for 7 hours/day, however, produced histo-
pathological changes in the livers and kidneys of male but not female rats.
These changes were characterized as lobular granular degeneration and focal
necrosis throughout the liver and cloudy swelling of the kidneys. The hepatic
and renal effects appeared to be reversible because rats exposed according to the
same protocol, but given a 6 week recovery period following exposure, appeared
normal by the criteria tested. Increasingly pronounced changes were observed in
the livers and kidneys of both sexes of rats exposed to 50 or 85 ppm for 7
hours/day. Hematologic indices, clinical chemistry, and urinalysis values,
tested at higher levels of exposure, were within normal limits. Each exposure
group and control group had 10-12 rats/sex except for the 25 ppm, 4 hour/day
group, which had 10 male rats and no females.
Similar experiments with guinea pigs (N = 8-12 sex/group) and rabbits (N =
2-3/sex/group) gave somewhat inconsistent results. Histopathological changes
were observed in livers and kidneys of both species at 25 ppm but not at 50 ppm in
either species, nor even at 85 ppm in guinea pigs. The results of these studies
are summarized in Table 5-3.
Other reports of effects of chronic inhalation exposure to chloroform in
experimental animals were not found in the more recent literature.
5.2.2.2 CHRONIC ORAL EXPOSURE IN ANIMALS — The data from several
studies on the effects of chronic and subchronic oral exposure to chloroform are
summarized in Table 5-4. Low levels of exposure (15-64 mg/kg/day, 6 days/week)
have been reported to increase survival in mice and rats (Roe et al., 1979;
Palmer et al., 1979) and to be associated with possible transient CNS depression
5-16
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TABLE 5-3
Effects of Inhalation Exposure of Animals to Chloroform, 5 Days/Week for 6 Months*
Species
Sex
ppm
Exposure
hours/day
Number in Group
Started Survived
Effects
rats
85
10
F 85
10
10
50 7
10
50 7
10
10
25 7
25 7
12
12
12
Excess mortality attributed to pneumonia on basis of gross
and microscopic appearance of lungs; slight depression of final
body weight; increase (p<0.05) in relative but not absolute
weights of liver, kidneys, and testes; no effect on spleen
weight; histological findings included marked centrilobular
granular degeneration of the livers and cloudy swelling of the
kidneys but no histopathological changes in testes; hematologic
values (including differential count), urinalysis values
and SGPT, SUN, and SAP values all "within normal limits"
No evidence of pneumonia; final body weights and weights of
liver and spleen unaffected, relative and absolute kidney
weights increased; histological findings included marked
centrilobular granular degeneration of the livers and cloudy
swelling of the kidneys; hematologic, urinal ysis, and
SGPT, SUN, and SAT values all "within normal limits"
Depression of final body weights (p<0.05); increases
in relative (p<0.05) but not absolute weights of kidneys,
spleen, and testes; histopathological changes in livers
and kidneys similar to those seen at 85 ppm; hematologic
values, urinalysis values, and SGPT, SUN, and SAP values
all "within normal limits"
Final body weights and weights of liver and spleen unaffected;
increase in relative kidney weight (p <0.05); histopatho-
logical changes in livers and kidneys similar to
those seen at 85 ppm but somewhat less marked; hemato-
logic values, urinal ysis values, and SGPT, SUN,
and SAP values all "within normal limits"
No effect on final body weights or weight of liver,
spleen, or testes; increased relative kidney weight
(p <0.05); lobular granular degeneration with
focal areas of necrosis throughout the liver; cloudy
swelling of renal tubular epithelium
No statistically significant effect on body
weight; increased relative but not absolute
kidney and spleen weights; no histopathologic
changes in kidneys and spleen; microscopic appearance
of livers not specified
-------
TABLE 5-3
Effects of Inhalation Exposure of Animals to Chloroform, 5 Days/Week for 6 Months* Ccont.)
Exposure Number in Group
Species
rats
Sex ppm hours/day Started
M,F 25 7 12/sex
plus 0 ppm for
6 weeks (recovery
period)
Survived Effects
8 M, "Normal" by the criteria tested at this dosage level
10 F (see 25 ppm above)
guinea
Pigs
I
CD
rabbits
dogs
25
M,F 85,50, or 25
M,F 85, 50, or 25
1,2, or 4 10, 10, 10 7,8,4, No evidence of adverse effects by the criteria
respectively tested (i.e., final body weight; weights of livers, kidneys,
spleen, testes; and probably gross and microscopic
appearance of at least the liver and kidneys)
7 8 to 12/sex/ 50 to 92% No adverse effects at 50 or 85 ppm other than
exposure level (mortality marked pneumonitis in F at 85 ppm; some histopatho-
not related logical changes in livers of both sexes and
to exposure) kidneys of M at 25 ppm (criteria tested were body weights,
organ weights, and gross and microscopic appearance
of organs)
7 2 to 3/sex/ 0 to 1 death/ No adverse effects at 50 ppm; some histopatho-
exposure level group, not logic changes in kidneys and liver and pneumonitis
related to in lungs at 25 and 85 ppm. Hematologic and
exposure clinical chemistry values within normal
limits at 85 ppm (criteria tested were same as for rats)
7 1/sex 1/sex No adverse effects in M; marked cloudy swelling of
renal tubular epithelium and increase in
capsular space in glomeruli of kidneys in F
(criteria tested were same as for rats and included
clinical chemistry and hematological studies)
•Source: Torkelson et al., 1976
M = male; F = female; SGPT = serum glutamic pyruvic transaminase; SUN = serum urea nitrogen; SAP = serum alkaline phosphatase
Controls for each species and sex included at least one unexposed and one air-exposed group, each comparable in number of animals to the exposed
groups. In statistical comparisons of organ and body weights, values for the control group (unexposed or air-exposed) closer in body weight to the
test group were used. Mortality in control groups was similar to mortality in treated groups, with the exception of excess mortality in male rats
exposed to 85 ppm for 7 hours/day or to 25 ppm for 4 hours/day. No explanation was given by Torkelson et al. (1976) for the high mortality in the
4 hours/day group. Strains of animals and age or weight at the start of the experiment were not specified. Purity of the chloroform used was 99.3%
(0.4} ethyl alcohol and <0.3% of an unknown).
M,F
25
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TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals
Species, Strain
Age/Weight at
Start sex
Rats, Sprague-Dawley M,F
weanling, 101 g M,
94 g F
No. at
Start
20/sex/dose
level
Vehicle
drinking
water
Dosage
0, 5, 50, 500,
or 2500 ppm in
drinking water
Duration
90 days, after
which 10 rats/
group were killed
Response
Increased mortality,
decreased growth rate,
and decreased food
Reference
Chu et al., 1982b
ad lib. corre-
sponding to
intakes of 0,
0.11-0.71, 1.2-
1.5, 8.9-14, or
29-55 mg/rat/day
The highest dosec
corresponds to
=291 mg/kg/day F,
310 mg/kg/day M
and 10 rats/group
observed for
additional 90 days
intake at highest dose;
increased frequency of
mild to moderate
liver and thyroid
lesions at highest
dose, including
increases in cytoplasmio
homogeneity, hepato-
cyte density, and
cytoplasmic volume,
vacuolization due to
fatty infiltration, some
vesiculation of biliary
epithelial nuclei, and
hyperplasia in livers;
and reduced follicle
and colloid density,
increased epithelial
height, some focal
collapse of follicles in
thyroid; no histopatho-
logical effects in
kidney, brain, and heart;
after the 90-day recovery,
the lesions were very
mild and similar to
those seen in controls
-------
TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
un
i
Species, Strain
Age/Weight at
Start Sex
Rats, Osborne-Mendel M
6 weeks/ 190 g
No. at
Start Vehicle Dosage Duration
30/group drinking 0, 200, 400, 600 10 rats/group
except water 900, or 1800 ppm killed at 30,
HO ad lib. in drinking water 60, and 90 days
controls ad lib. cor- of exposure
responding to
intakes3 of 0,
20, 38, 57, 81, or
160 mg/kg/day,
plus 0 ppra group
matched with
1800 ppm group
for water con-
sumption
Response Reference
Dose-related signs Jorgenson and
of depression during Rushbrook, 1980
1st week only; dose-
related reduction in
water consumption;
decreased weight gain
in 160 mg/kg group;
increased incidence of
"hepatosls" in livers
of treated rats at 30
and 60 days but not
90 days (not dose-
related) ; no other
treatment-related
effects on serum
clinical chemistry
values or urinalysis
values, or gross and
microscopic appearance
of tissues, including
kidney.
-------
TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
Species, Strain
Age/Weight at
Start
No. at
Sex Start Vehicle Dosage Duration
Response
Reference
rats, Osborne-Mendel
52 days/240 g M
175 g F
M,F 50/sex/dose corn oil,
level; 20/sex gavage
matched con-
trols; =99/
sex colony
controls
M: 0, 90, or 180
mg/kg/day;
F: 0, 100, or 200
mg/kg/day (TWA);
5 days/week
78 weeks treat-
ment plus 33 weeks
observation
i
t-o
Treated animals had NCI, 1976
dose-related decrease
in survival and weight
gain, slight decrease
in food consumption,
increased severity and
incidence of pulmonary
lesions characteristic
of pneumonia; necrosis of
hepatic parenchyma NS in
controls, 3/50 low dose M,
4/50 high dose M, 3/49 low
dose F, 11/48 high dose F;
hyperplasia of urinary
bladder epithelium 1/18
control M, 7/45 low dose
M, 1/45 high dose M, NS
for control F, 6/43 low
dose F, 2/40 high dose
F; increased splenic
hematopoiesis in 1/18
control M, 3/45 low dose
M, 6/45 high dose M; both
control and treated
animals had chronic
nephritis; low but
statistically signi-
ficant increased
incidence of renal
epithelial tumors
in treated M (see
Carcinogenic!ty
section)
-------
TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
Species, Strain
Age/Weight at
Start
Sex
No. at
Start
Vehicle
Dosage
Duration
Response
Reference
rats, Sprague-Dawley
NS
M,F
10/sex/dose
level
toothpaste,
gavage
0, 15, 30, 150,
or 110 mg/kg/day
6 days/week
13 weeks
rats, Sprague-Dawley
SPF. 180 to 240 g M
130 to 175 e F
M,F
50/sex/group toothpaste,
gavage
0 or 60 mg/kg/day
6 days/week
80 weeks exposure
plus 15 weeks
observation
At 410 mg/kg/day,
increased liver weight
with fatty change
and necrosis,
gonadal atrophy,
increased cellular
proliferation in
bone marrow; at
150 mg/kg/day,
changes less pro-
nounced but effect
(NS) on relative
liver and kidney
weights; pre-
sumably no effects at
lower dosage levels
Palmer et al., 1979
Survival of treated
animals slightly
better than that of
controls (32$ treated M,
22$ control M, 26$
treated F, 14$ control
females survived to 95
weeks); body weights of
treated rats slightly and
progressively depressed;
intercurrent respiratory
and renal disease in all
groups; minor histologi-
cal changes in livers but
no evidence of "treatment-
related toxic effect" in
livers; decrease
(p<0.01) in relative
liver weight in
treated females; no
gross or histologic
treatment-related
changes in brain;
possible effect (NS) on
incidence of severe glomerulo-
nephritis; decrease
in plasma cholines-
terase in treated females
Palmer et al., 1979
-------
TABLE 5-1. (cent.)
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
Species , Strain
Age/Weight at
Start
No. at
Sex Start Vehicle Dosage Duration
Response
Reference
mice, B6C3F1,
6 weeks/19 g
30/group
except 40
ad lib.
controls
drinking
water
0, 200, 400, 600,
900, 1800, or
2700 ppm in
drinking water
ad lib. cor-
responding to
intakes"
of =0, 32, 64,
97, U5, or 290
mg/kg/day; addi-
tional 0 ppm
group matched
with 2700 ppm
group for water
consumption
10 mice/group
killed at 30, 60,
and 90 days of
exposure
Dose-related signs
of depression during
first week only; marked
reduction in
water consumption
in higher-dose
groups during first
2 weeks; body
weight losses (p<0.05)
in 97, 145, and 290 mg/kg
groups and in matched
controls during first week;
mild hepatic centrilobular
fatty change in 64, 97,
145, and 290 mg/kg groups
at 30 days, but only in 2
highest dosage groups at
60 and 90 days; increase
in liver fat/liver
weight (p<0.05) for
290 mg/kg group at
all 3 periods;
no other treatment-
related changes in serum
enzyme levels, urinalysis
values, or gross or
microscopic appearance of
tissues including kidney
Jorgenson and
Rushbrook, 1980
mice, B6C3F1
35 days/l8g M,
17 g F
M,F
50/sex/dose
level; 20/sex
matched
controls
corn oil,
gavage
M: 0, 138, or 227
mg/kg/day;
F: 0, 238, or 447
mg/kg/day;
5 days/week
78 weeks treat-
ment plus 14 to 15
weeks observation
Survival decreased NCI, 1976
in high dose F,
unaffected in
other treated groups;
high incidences of
hepatocellular car-
cinoma in treated mice
(see Carcinogenicity sec-
tion); renal inflamma-
tion in 10/18 control M,
2/50 low dose M,
1/50 high dose M
-------
TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
Species, Strain
Age/Weight at
Start
mice, Schofield
No. at
Sex Start Vehicle Dosage Duration
M,F 10/sex toothpaste, 0, 60, 150, 6 weeks
per dosage gavage or 425 mg/kg/
day; 6 days/
week
Response Reference
At 425 mg/kg, 100? Roe et al . , 1979
mortality; at 150 mg/kg,
8/10 M died and
weight gain of F
markedly retarded; at
60 mg/kg, weight gain
of both sexes moder-
ately retarded; no
other observations men-
tioned
mice, ICI (expt. 1)
ICI (SPF) (expt. 2)
ICI, CBA, C57BL,
CF/1 (expt. 3)
<10 weeks old
\_n
i
expt. 1:
M,F
expt. 2
and 3:
M
treated
and con-
trol,
plus
some F
control
treated:
52/sex/dose
level; con-
trol: 52 to
206/sex/
strain/
vehicle plus
untreated
toothpaste in
all 3 expt.
for all
strains and
sexes plus
arachis oil
in expt. 3
for ICI,
gavage
0, 17 (expt.1),
or 60 mg/kg/day,
6 days/week
80 weeks treat-
ment; 16 to 24
weeks observa-
tion according
to numbers of
survivors
Survival generally
better in 60 mg/kg
groups than in con-
trols except for
CF/1 animals or when
chloroform given in
arachis oil; slight
retardation of weight
gain in 60 mg/kg
groups; no effect
on hematologic values
(tested in expt. 2
only); no treatment-
related adverse effect
on liver or other
tissues except in kid-
neys as follows:
60 mg/kg in tooth-
paste - increased inci-
dence of moderate
to severe renal
changes (p <0.001) in
CBA and CF/1 M,
60 mg/kg in oil-
increased incidence of
moderate to severe kid-
ney disease (p <0.05) in
1C1 M; increased incidence
of benign and malig-
nant kidney tumors
in ICI M treated with
60 mg/kg in tooth-
paste or oil (see
Carcinogenicity section)
Roe et al., 1979
-------
TABLE 5-4
Effects of Subchronic or Chronic Oral Administration of Chloroform to Animals (cont.)
Species, Strain
Age/Weight at
Start
dogs , beagle
18 to 24 weeks
7 to 8 g
No. at
Sex Start
M,F 1 /sex/dose
level for
90 and 120
mg/kg; 2/sex/
dose level
for lower
dosages
Vehicle
toothpaste
in gelatin
capsule,
orally
Dosage
30, 45, 60, 90,
or 120 mg/kg/day,
7 days /week
Duration
13 weeks for
30 and 45 mg/kg,
18 weeks for
60 mg/kg, 12 weeks
for 90 and 120
mg/kg
Response Reference
No deaths; occasional Heywood et al., 1979
vomiting; marked
weight loss in all
dogs and poor
general condition in
some at 60 mg/kg or
higher ; apparent
I
M
vn
suppression of appe-
tite initially at all
dosages and through-
out at 60 mg/kg and
higher; jaundice and
increased SAP, SCOT,
SGPT, bllirubin, and
ICD values in male at
120 mg/kg; increased
SGPT values in 1/4 and
increased SAP and SCOT
values in 2/4 at 60
mg/kg; hepatocyte
enlargement and vacuo-
1ation with fat depo-
sition at 60 mg/kg and
higher; discoloration of
liver, increased liver
weight, and slight fat
deposition in hepato-
cytes at 45 mg/kg; no
effect on any of these
clinical chemistry or
histological parameters
at 30 mg/kg
-------
TABLE 5-4
Effects of Subohronio or Chronic Oral Administration of Chloroform to Animals (cont.)
Species, Strain
Age/Weight at
Start
dogs, beagle,
8 to 24 weeks,
7 to 8 kg
No. at
Sex Start
M,F 8 /sex/ dose
level; 16/sex
vehicle
controls;
plus other
controls
Vehicle
toothpaste
in gelatin
capsule,
orally
Dosage
0, 15, or 30
mg/ kg/day;
6 days/week
Duration
7.5 years treat-
ment plus 20 to
21 weeks observa-
tion
Response
No effect on survival ,
growth, organ
weights , hematologic
or urinalysis values
(checked at intervals
thr ougho ut ) ; moderat e
Reference
Heywood et al . , 1979
I
ro
dose-related elevation
of SGPT reaching peak
in sixth year of study,
reverting to normal
levels after treatment
discontinued; other
serum enzyme indica-
tors of hepatic damage
(checked during the
latter portion of the
study) followed
pattern similar to SGPT,
but BSP retention
and ICD values were
unaffected; aggregation
of vacuolated histio-
cytea ("fatty cysts")
in livers of all groups
but cysts were larger
and more numerous in
treated dogs and
persisted after treat-
ment ended; fat depo-
sition affected more
renal glomeruli in
30 mg/kg group than
in other groups
Calculated by Jorgenson and Rushbrook (1980) from measured average body weights and water consumption.
Calculated from Jorgenson and Rushbrook's statement that the mice had actual intake levels of from 148 to 175J of the
intended levels of 20, 40, 60, 90, 180, and 270 mg/kg/day.
Calculated by Chu et al. (1982b) by multiplying the fluid intake volume by the concentration of chloroform
Growth rate data were given only for the highest dose and mg/kg/day were calculated from this information by taking average weights over the 90-day
period of exposure.
SPF = specific pathogen-free; SGPT = serum glutamic-pyruvic transaminase; SAP = serum alkaline phosphatase; BSP = bromsulphthalein;
ICD = isocitric dehydrogenase (serum)
Purity of the chloroform samples used in all these studies was generally high and is discussed in the section on Carcinogenic!ty.
M = male; F = female; TWA = time-weighted average dose for days on which chemical was administered; NS = Not specified;
-------
and mild hepatic changes in mice and rats (Jorgenson and Rushbrook, 1980; Palmer
et al., 1979), hepatic damage in dogs (Heywood et al., 1979), and renal damage in
male mice of some sensitive strains (Roe et al., 1979), and in dogs (Heywood et
al., 1979). In addition to hepatic damage, ingestion of =300 mg/kg/day of
chloroform produced thyroid lesions in rats (Chu et al., 1982b). In one study, a
decreased incidence of renal inflammation occurred in male mice treated with
chloroform at 138 or 227 mg/kg/day, 5 days/week (NCI, 1976). A similar effect
may have occurred in rats (Palmer et al., 1979), but the authors did not specify
whether the effect of chloroform was to increase or decrease the incidence of
intercurrent renal disease. The NCI (1976) report stated that hepatic necrosis,
hyperplasia of the urinary bladder epithelium, and increased splenic hemato-
poiesis in rats may have been related to chloroform treatment, but incidences in
controls for some of these effects were not reported and the data, shown in Table
5-4, are difficult to interpret. Many of the studies summarized in Table 5-4
were at least partially designed as investigations of carcinogenicity and,
hence, are also discussed in the carcinogenicity section of this document; some
experimental details are discussed more fully in that section.
From the data presented in Table 5-4, it appears that rats and mice can
tolerate higher daily intakes of chloroform when it is given in their drinking
water or in a toothpaste base (by gavage) than they can when the chemical is
administered in corn or arachis oil (by gavage). In the subchronic study of
Jorgenson and Rushbrook (1980), rats and mice appeared to adapt to low levels of
chloroform intake (up to =100 aig/kg/day); signs of depression and mild hepatic
damage that occurred initially had disappeared by 90 days of treatment. Elevated
indices of liver damage (e.g., SGPT levels) in dogs chronically exposed to
chloroform reverted to "normal" after treatment was discontinued, although
histological changes persisted. Similarly, the mild liver and thyroid lesions
5-27
-------
seen in rats exposed to high doses of chloroform via their drinking water for 90
days were no longer apparent in rats allowed to recover for an additional 90 days
(Chu et al., 1982b).
5.3 INVESTIGATION OF TARGET ORGAN TOXICITY IN EXPERIMENTAL ANIMALS
5.3-1 Hepatotoxicity. An extensive review of the early literature dealing
with chloroform-induced liver damage by von Oettigen (196M) notes studies
beginning in 1891. More recently, Groger and Grey (1979) summarized reports
describing chloroform-induced liver hepatotoxicity as follows: typical effects
of chloroform on liver cells are extensive vacuolization, disappearance of
glycogen, fatty degeneration, swelling, and necrosis, all starting in the
centrilobular areas. There is also often hemorrhaging into the parenchyma and
infiltration of polymorphonuclear cells and monocytes. Electron-microscopic
observations of liver parenchymal cells from chloroform-intoxicated rats as
carried out by Scholler (1966, 196?) revealed deposition of lipid droplets in the
cytoplasm, partial destruction of the mitochondrial matrix, proliferation of
smooth endoplasmic reticulum, and swelling of the rough endoplasmic reticulum
with detachment of ribosomes.
Kylin et al. (1963) conducted a study of the hepatotoxic effects of inhaled
trichloroethylene, tetrachloroethylene, and chloroform in mice with the objec-
tive of finding the lowest concentration of the substances producing signs of
liver damage after a single U-hour exposure period. Histological examination
showed that a concentration of 100 ppm caused moderate fatty infiltration in mice
killed 1 day after exposure. At >200 ppm, the extent of the alteration increased
with concentration and was more pronounced after 1 day than 3- Thus, judging
from the histological picture, the smallest concentrations (ppm) of the
different agents to produce more severe alterations in the exposed group than in
the controls were as follows:
5-28
-------
1 day after
exposure
3 days after
exposure
Trichloro-
ethylene
(ppm)
1600-3200
>3200
Tetrachloro-
ethylene
(ppm)
<200
200 to 400
Chloroform
(ppm)
<100
100 to 200
On this basis, the hepatotoxic effects of trichloroethylene, tetrachloro-
ethylene, and chloroform are in the approximate ratios 1:10:20. The amount of
liver fat was raised at 400 ppm of chloroform. A third indicator of liver
toxicity was an increase in serum ornithine carbamyl transferase (S-OCT)
activity at 24 hours in animals exposed to 200, 400, and 800 ppm of chloroform.
A study of the effect of oral doses of chloroform on the extent of liver
damage in white mice ("of a Swiss strain") was conducted by Jones et al. (1958).
Minimal changes characterized by midzonal fatty infiltration were observed 72
hours after the administration of 30 mg (0.02 m£)/kg. When the dose was
increased to 133 mg (0.09 m£)/kg, a massive fatty infiltration of the total liver
lobule was found. At a level of 355 mg (0.24 m£)/kg, massive fatty infiltraton
occurred along with severe central lobular necrosis. Information on the hepato-
toxicity of long-term chloroform administration has been presented in the sec-
tions on Effects of Chronic Exposure to Chloroform. Inhalation exposure of rats
to 25, 50, or 85 ppm chloroform for 7 hours/day, 5 days/week for 6 months
produced centrilobular granular degeneration and focal necrosis in their livers.
In subchronic studies, ingestion of up to =100 mg/kg/day of chloroform produced
mild, transient histological changes in the livers of rats and mice (Jorgenson
and Rushbrook, 1980), ingestion of 145 or 190 mg/kg/day produced fatty change in
the livers of mice (Jorgenson and Rushbrook, 1980), and administration of 410
mg/kg/day by gavage produced fatty change and necrosis in the livers of rats.
Dogs treated subchronically with 45 mg/kg/day by the oral route had histological
5-29
-------
evidence of slight hepatic fatty change, with increasingly severe changes noted
at dosages of 60 and 120 mg/kg/day.
In chronic oral studies, rats had minor histological change in their livers
and a decrease in relative liver weights when given 60 rag/kg/day of chloroform, 6
days/week, while the livers of mice were unaffected at this dosage. Dogs had
some evidence of liver damage (clinical chemistry parameters) and an increase in
the number and size of fatty cysts in their lifetime when administered 15 or 30
mg/kg/day orally for 6 days/week. The mechanism by which chloroform exerts its
hepatotoxic effects has been widely investigated and efforts have been made to
identify the responsible metabolite(s).
As long ago as 1928, it was suspected that the liver damage induced by
chloroform may be due not only to the chemical itself, but might be caused by a
degradation product (Lucas, 1928). The concept that chloroform is excreted
unchanged was disproved since a large number of studies (Butler, 1961; Paul and
Rubenstein, 1963; Van Dyke et al., 1964; and Reid and Krishna, 1973) indicated
that the tissue necrosis induced by chloroform is associated with the covalent
binding of toxic metabolites and alkylation of tissue proteins. Autoradiograms
have revealed that this binding occurs predominantly in the necrotic areas (Illet
et al., 1973). It was also shown by McLean (1970) that pretreatment of rats with
phenobarbital (a microsomal enzyme inducing agent) greatly enhances the
lethality of chloroform.
Brown et al. (197*4) proposed a mechanism of chloroform hepatotoxicity
implicating a free radical metabolite which can react with glutathione (GSH) (a
tripeptide which protects against hepatotoxicity), diminishing GSH levels in the
liver. According to this hypothesis, once GSH levels are depleted, further
metabolism would result in the reaction of the metabolite with microsomal
protein, and hence, necrosis. This proposal was based on observations in pheno-
5-30
-------
barbital pre-treated rats anesthetized with chloroform, that hepatotoxicity was
enhanced and GSH levels were decreased by the induction of microsomal enzymes.
Covalent binding of chloroform metabolites to microsomal proteins In vitro was
also enhanced by enzyme induction, an effect prevented by GSH. Similar findings
were reported in mice by Ilett et al. (1973), who found severe chloroform-induced
centrilobular necrosis in phenobarbital pretreated mice, but only slight centri-
lobular damage in mice exposed only to chloroform.
Thus, the hepatotoxicity of chloroform appears to depend on 1) the rate of
its biotransformation to produce reactive metabolite(s), and 2) by the amount of
GSH available to conjugate with and thus inactivate the metabolite(s).
The role of GSH in chloroform-induced hepatotoxicity was further studied by
Docks and Krishna (1976), who found that only thoses doses of chloroform that
decreased liver GSH caused liver necrosis when administered to phenobarbital
pretreated rats.
More recently, Ekstrom et al. (1982) studied the mechanism of GSH depletion
by chloroform in rats pretreated with phenobarbital. The synthesis of GSH
proceeds via two enzymatic steps, the first of which is rate limiting:
glutamate + cysteine I-Slutamyl-cysteine, dipeptide
synthetase
In the presence of glycine, the reaction continues via GSH synthetase to produce
GSH. When the soluble fraction from livers of rats sacrificed at various times
after chloroform exposure was incubated in the presence of these amino acids, it
was found that GSH synthesis was inhibited within 4-6 hours, while liver necrosis
was evident only after 6 hours. When glycine was eliminated from the initial
part of the incubation, the dipeptide accumulated, but at a lower rate in the
presence of chloroform than in its absence. Later addition of glycine resulted
in GSH synthesis at a rate similar to control values. Thus, it appears that
chloroform, or rather a reactive metabolite, inhibited GSH synthesis at the rate
5-31
-------
limiting step (i.e., the formation of dipeptide by Y-glutamyl-cysteine synthe-
tase).
The biotransformation of chloroform (as discussed in Chapter i|) depends on
the activity of the microsomal drug metabolizing enzymes. Substances that induce
these enzymes were shown to, indeed, enhance the hepatotoxicity of chloroform as
evidenced by increased serum glutamic-pyruvic transaminase (SGPT) levels and
decreased hepatic glucose-6-phosphatase activity (Lavigne and Marchand, 1974).
An inhibitor of the drug metabolizing enzymes SKF-525A, however, while
increasing the excretion of C-carbon monoxide in rats administered
14
C-labelled chloroform, failed to diminish the hepatotoxicity of chloroform,
leading these authors to conclude that factors other than metabolism may be
involved.
McMartin et al. (1981) demonstrated that altering the cytochrome P-450
concentrations in the livers of chloroform-exposed rats also altered the hepato-
toxicity, as measured by the incidence of hepatic lesions and by serum alanine
aminotransferase activities. Both fasting and phenobarbital pretreatment
increased the cytochrome P-USO content and liver damage, while cadmium produced
the opposite effect.
Theories of chloroform hepatotoxicity involve the formation of reactive
intermediates by liver enzymes. How these intermediates exert their hepatotoxic
effect has been the subject of several studies. It has been suggested by Masuda
et al. (1980) that, based on the chloroform-induced indices of hepatotoxicity of
decreased microsomal glucose-6-phosphatase activity and cytochrome P-'450 content
with increased hepatic malondialdehyde levels, the lipid peroxidation hypothesis
proposed for carbon tetrachloride may also apply to the case of chloroform.
Qualitative and mechanistic differences of hepatotoxicity between the two
chemicals were noted, however.
5-32
-------
The interactive hepatotoxicity of chloroform and carbon tetrachloride was
studied by Harris et al. (1982) who found that, while neither chemical alone
given at subthreshold dose altered SGPT activity, hepatic triglyceride content,
or hepatic calcium content, when given together, these chemicals increased the
toxic response in rats. Administration of either or both chemicals had no effect
on GSH levels or conjugated diene formation, but ethane expiration was increased
in rats given both chemicals. Diene conjugation and ethane expiration are
indices of lipid peroxidation. Histopathological changes were more severe from
the combinati^i than from either chemical alone. Although the mechanism of the
hepatotoxic interaction between chloroform and carbon tetrachloride is unclear,
the authors suggested that there might be a combined effect of phosgene formation
and lipid peroxidation initiation.
It should be noted that the prevailing theories implicate phosgene as the
major metabolite responsible for chloroform hepatotoxicity (Reynolds and Yee,
1967; Sipes et al., 1977; Mansuy et al., 1977; Pohl et al., 1977). Other
potential toxic metabolites discussed by Pohl (1979) in a review of this subject
are a trichloromethyl radical and dichlorocarbene; however, they are considered
less important than phosgene in this regard.
A study by Stevens and Anders (1981) supports the phosgene-mediated
in
mechanism. The time course of changes in SGPT levels and covalent binding of C
to proteins was examined in microsomal and soluble fractions from phenobar-
bital-pretreated rats sacrificed at various times after chloroform or
C-chloroform administration. It was found that C binding was maximal at 6
hours while indices of liver damage peaked at 18 hours after chloroform exposure.
Further experiments were performed in which diethyl maleate (a GSH depletor)
treatment caused increased C-binding to soluble and microsomal fractions and
increased SGPT levels, perhaps by inhibiting the metabolism of phosgene to carbon
5-33
-------
monoxide or stable conjugates. Cysteine, which reacts with phosgene to produce
2-oxothiazolidine-M-carboxylic acid, had a protective effect. Diethyl maleate
also diminished, but did not eliminate the deuterium isotope effect on the GSH
dependent chloroform metabolism to carbon monoxide, which would be expected if
carbon monoxide formation occurred subsequent to phosgene production. Thus, the
hepatotoxicity of chloroform can be altered by altering the various reaction
pathways of phosgene, strongly indicating that phosgene is the toxic inter-
mediate.
From the above discussion, it appears that, to be hepatotoxic, chloroform
must first be metabolized by microsomal drug metabolizing enzymes to an active
intermediate, probably phosgene, which in turn can react by various pathways,
depending on GSH levels, one of which is the covalent binding to liver proteins
resulting in necrotic lesions.
5.3.2 Nephrotoxicity
As noted by Watrous and Plaa (1972), the extensive body of research on the
hepatotoxicity of halogenated hydrocarbons has tended to overshadow the fact
that some of these agents are also nephrotoxic. Earlier reports of chloroform
nephrotoxicity include those of Heller and Smirk (1932), who found that rats
anesthetized with chloroform showed a diminished ability to excrete a water load
given prior to anesthesia, and Knocher and Mandelstam (1944), who noted that
chloroform injection produced a fatty infiltration of the kidney.
Renal necrosis produced by the oral administration of chloroform was
described by Eschenbrenner and Miller (1945a). The necrosis, observed only in
male mice, involved portions of both proximal and distal convoluted tubules. The
nuclei of the epithelial cells were often absent or fragmented and the cytoplasm
was coarsely granular and deeply eosinophilic. The glomeruli and collecting
tubules appeared normal.
5-34
-------
Sex and strain differences in the sensitivity of mice to chloroform nephro-
toxicity were further studied by Deringer et al. (1953). Exposure of strain C3H
mice to air containing =5 mg/£ of chloroform for 1, 2, or 3 hours resulted in
lesions of the kidneys of all of the males but in none of the females. In animals
dying within 1 day after exposure, epithelium of the proximal tubules and
portions of the distal tubules were generally necrotic. The lumens of those
segments of tubules were dilated. The glomeruli were relatively unaffected. The
mice dying or sacrificed at later time intervals exhibited calcification in the
necrotic area.
Similar lesions were found in males of strains C3H, C3Hf, A, and HR.
However, strains C57BL, C57BR/cd, C57L, and ST were resistant to chloroform-
induced nephrotoxicity.
Comparable results were reported by Krus and Zaleska-Rutczynski (1970).
Subcutaneous administration of chloroform to C3H/He male mice resulted in renal
tubular necrosis, with death ensuing U-9 days later. The lesions were calcified
with no evidence of regeneration. Female mice of this strain, males and females
of the C57BL/6JN and BN strains, and FI generation males of the cross of female
C3H/He with male C57BL/6JN mice survived the administration of chloroform (0.1 m?-
of 0.05 g chloroform in 1 mS, ethyl laureate). Additional studies were performed
with males and females of this F generation and the resistant BN strain, in
which animals were sacrificed at various time after chloroform administration.
All mice survived, while all female mice were resistant, showing no kidney
lesions at any time in the experiment. Renal damage was morphologically apparent
in all male mice by 12 hours, but regeneration developed by day 4 and continued
until the end of the experiment. It was concluded that although all male mice
had tubular lesions, the ones surviving had tubules that did not calcify and a
large degree of renal regeneration.
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Several investigators have studied ]the influence of testosterone on chloro-
form-induced renal damage. Eschenbrenner and Miller (1945b) performed experi-
ments in which they saw extensive necrosis of portions of the proximal and distal
renal tubules in normal male and in testosterone-treated castrated male mice
following the acute oral administration of chloroform. However, no necrosis was
found after chloroform was administered to female mice or castrated male mice.
Continuing this line of investigation, Culliford and Hewitt (1957) reported the
following results:
1) Adult male mice of two strains (CBA and WH) developed extensive necrosis
of the renal tubules after exposure to low concentrations of chloroform
vapor (7-10 mg/2, for 2 hours). Adult females showed no renal damage
after equivalent exposure.
2) Adult females became fully susceptible to necrosis after treatment with
androgens. The susceptibility of males was greatly reduced by treatment
with estrogens.
3) Castration removed the susceptibility of the males of one strain, but did
not completely remove it in another. The residual susceptibility of
castrates was abolished by adrenalectomy.
*() Male mice under 11 days old were not susceptible to necrosis even after
massive doses of androgen. Between 11 and 30 days, they were susceptible
if given androgen. Thereafter, they became spontaneously susceptible.
5) Liver damage occurred in nearly all exposed mice and was not correlated
with sex hormone status.
6) Susceptibility could be induced in gonadectomized mice by methyl testos-
terone, testosterone propionate, dehydroepiandrosterone, progesterone,
and large doses of cortisone acetate.
Hill (1978) also performed experiments demonstrating similar strain and sex
differences in chloroform-induced renal toxicity. The renal toxicity of a fixed
oral dose of chloroform to castrated male mice was increased with increasing
doses of administered testosterone. Plasma levels of testosterone in resistant
strains tended to be lower than levels In susceptible strains. Hill (1978)
conjectured that a testosterone may act by sensitizing the renal proximal convo-
luted tubules to chloroform through a testosterone receptor mechanism.
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Eschenbrenner and Miller (1945b), however, linked susceptibility to the nephro-
toxic action of chloroform to differences in kidney morphology and physiology
induced by testosterone.
Information on the nephrotoxicity of long-term chloroform administration
has been presented in the section on Effects of Chronic Exposure to Chloroform.
Inhalation of 25, 50, or 85 ppm of chloroform 7 hours/day, 5 days/week for 6
months produced cloudy swelling of the renal tubular epithelium in rats. Male
mice of certain sensitive strains had increased incidences of moderate to severe
renal disease when treated orally with chloroform at a dosage of 60 mg/kg/day, 6
days/week in a chronic study (Roe et al., 1979). Chronic oral administration of
30 mg/kg/day of chloroform, 6 days/week, to dogs produced an increase in the
numbers of renal glomeruli affected by fat deposition (Heywood et al., 1979).
The mechanism of the nephrotoxicity of chloroform has been less extensively
studied than has that of hepatotoxicity. Ilett et al. (1973) suggested that the
hepatotoxic metabolite produced in the liver is transported via the circulation
to the kidney where it exerts its nephrotoxic effects. More recent studies
(McMartin et al., 1981; Kluwe and Hook, 1981) suggest that chloroform may also be
metabolized in the kidney, but by a different mechanism.
McMartin et al. (1981) altered the concentrations of cytochrome P-450 by
fasting, by phenobarbital pretreatment, or by administration of cadmium to rats
given chloroform as a challenge. Fasting increased cytochrome P-^50 concentra-
tions in both liver and kidney, and chloroform-induced damage was enhanced in
both organs of fasted animals. Pretreatment with cadmium decreased cytochrome
P-450 in livers but not kidneys and significantly diminished liver damage due to
chloroform while having no effect on kidney damage due to chloroform exposure.
Phenobarbital pretreatment resulted in increased liver but not kidney cytochrome
P-450 and, likewise, chloroform-induced damage was enhanced in livers but not
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kidneys. Thus, pretreatments that altered hepatic cytochrorne P-450 levels had no
effect on chloroform-produced renal effects, suggesting that a metabolite of
chloroform, which is responsible for kidney damage, is produced in the kidney.
The mechanism of chloroform nephrotoxicity was also investigated by Kluwe
and Hook (1981), who found no difference between nephrotoxicity and hepato-
toxicity with respect to the effects of microsomal enzyme inhibitors and diethyl
maleate. Mice were injected intraperitoneally with chloroform either before
piperonyl butoxide or SKF-525-A administration or after diethyl maleate,
piperonyl butoxide or SKF-525A exposure. Although it inhibits microsomal
enzymes, SKF-525-A, administered either before or after chloroform administra-
tion, did not reduce the hepatotoxicity or the nephrotoxicity of chloroform.
This is consistent with a similar finding by Lavigne and Marchand (1974).
Piperonyl butoxide, when given before chloroform, protected against toxicity in
both organs, but when given after chloroform, did not. This finding indicated
that an enzymatic step in the metabolism of chloroform by both organs was
inhibited. The effect of diethyl maleate was to enhance the toxicity of chloro-
form in both organs. Thus, the mechanism of chloroform nephrotoxicity appears to
be similar to that of hepatotoxicity with respect to these substances.
5.4 FACTORS MODIFYING THE TOXICITY OF CHLOROFORM
From the preceding discussion, it is evident that the alterations of micro-
somal enzyme activity or hepatic GSH levels influence the severity of toxic
effects induced by a given amount of chloroform. It follows then that many
factors could alter chloroform toxicity by affecting these parameters or acting
through other mechanisms. These substances are of interest because they fall
into categories of accidental or intentional exposure to humans. Alcohol,
dietary components, pesticide, and steroids are some of the substances which are
discussed below.
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5.^.1 Factors that Increase the Toxicity. The effect of ethanol pretreatment
on chloroform-induced hepatotoxicity in mice was studied by Kutob and Plaa
(1962a). An intoxicating dose (5 g/kg) of ethanol was administered orally to
mice daily for 15 days initially, with a systematic shortening of the duration to
a single exposure. A challenging dose of chloroform (0.08 mil/kg) was adminis-
tered subcutaneously either 12, 15, or 24 hours after ethanol treatment. Liver
dysfunction was measured by prolongation of phenobarbital sleeping time,
bromsulphalein (BSP) retention, liver succinic dehydrogenase activity, and
histological examination. Regardless of the ethanol treatment period, pheno-
barbital sleeping time was significantly increased in mice receiving ethanol
followed by chloroform when compared with mice receiving either substance alone.
Similar findings were found for BSP retention. The iri vitro succinic dehydro-
genase activity was significantly reduced by ethanol pretreatment followed by
chloroform administration 12 or 24 hours, but not 48 hours, later, when compared
with activities from mice receiving only chloroform. Histological changes were
seen in the livers of mice given ethanol 15 hours to 4 days prior to chloroform
challenge, while mice receiving either chemical alone had morphologically nonnal
livers. It was also determined that the ethanol treatment increased liver
triglyceride content, with a maximum at 15 hours, and that ethanol pretreatment
significantly increased the concentration of chloroform in the livers with a
maximum at 12 hours after chloroform challenge. From these results, it was noted
that a single dose of ethanol was just as effective as multiple doses. A
mechanism was proposed for the ethanol enhanced chloroform-induced hepato-
toxicity in which ethanol increases liver lipid content (as evidenced by
increased triglycerides) resulting in increased concentrations of chloroform to
be metabolized in the liver.
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In support of this mechanism is the observation that oral isopropanol
pretreatment for 5 days (0.3 m£/100 g for 2 days and 0.15 mil/100 g for 3 days)
followed 12 hours later by 5 daily inhalations of chloroform (5000 ppm first day,
2500 ppm on the next four days), 2 hours/day led to severe fatty infiltration of
the liver. Chloroform alone increased the pool of triglycerides (Danni et al.,
1981).
In contrast, Sato et al. (1980) studied the mechanism by which ethanol
enhances hydrocarbon metabolism, including that of chloroform. Rats ingesting
ethanol in their drinking water for 3 weeks were sacrificed 10 hours after the
final exposure. Control rats were given isocaloric glucose solutions. Liver
microsomal enzyme systems were prepared and liver protein and cytochrome P-450
contents analyzed, the increased contents being indicative of microsomal enzyme
synthesis in response to alcohol. When chloroform was added as a substrate, its
metabolism was enhanced by 6 times, much more than could be accounted for by
enzyme induction alone. Microsomes prepared from rats that were withdrawn from
ethanol 24 hours prior to sacrifice did not show enhanced activity. In a
subsequent study (Sato et al., 1981), rats receiving a single gavage dose of 0,
2, 3, 4, or 5 g/kg ethanol were sacrificed 18 hours later. The ln_ vitro
metabolism of chloroform by rnicrosomes prepared from these rats was enhanced very
little at 2 g/kg, slightly more at 3 g/kg, and dramatically at 4 g/kg. At 5 g/kg,
however, enhanced enzyme activity was no greater than at 3 g/kg ethanol. When
ethanol was added directly to the incubation system, the metabolism of chloroform
was inhibited. Rats receiving 5 g/kg ethanol retained relatively large amounts
in the blood and liver, while those receiving 4 g/kg retained almost none. If
the ethanol remaining in the rats exerted an inhibitory effect on enzyme
activity, then microsoraal enzymes prepared from 5 g/kg ethanol-treated rats,
mixed with the soluble fraction from control rats, should show increased activity
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when compared with both microsomal and soluble fractions from ethanol-treated
rats. This was found to be the case.
Based on these results and studies on metabolism of other hydrocarbons in
vitro and in vivo, Sato et al. (1981) suggested that ethanol is both a stimulator
and an inhibitor of drug metabolizing enzymes, depending on how much ethanol
remains in the body and thus how much time has elapsed since ethanol ingestion.
Thus, when ethanol is first ingested, it acts as a competitive inhibitor of
microsomal enzyme activity, but as it disappears from the body, an optimum for
stimulation may be reached and metabolism enhanced. It was postulated that since
the metabolism of chloroform was enhanced to a much greater extent than can be
explained from enzyme induction alone, perhaps ethanol modifies the enzyme
activities by other mechanisms such as modification of membrane properties,
allosteric effects, or by displacement of substrate already bound.
Polybrominated biphenyls (PBBs) have also been found to potentiate the
toxicity of chloroform (Kluwe and Hook, 1978) . Mice were fed diets containing 0,
1, 25, or 100 ppm PBB for 14 days. One day before sacrifice, the mice were given
a single intraperitoneal injection of 0, 0.5, 2.5, 5.0, or 50 ^1/kg chloroform.
PBB enhanced the toxicity of chloroform in both the liver and the kidney as
evidenced by results of blood urea nitrogen (BUN) and serum glutamic oxaloacetic
transaminase (SCOT) determinations and by inhibition of p-aminohippuric acid
(PAH) uptake by renal slices. PBB also reduced the LDp... of chloroform in these
mice and the deaths were attributed to hepatic necrosis. Since PBBs were known
to induce the drug metabolizing enzymes, their effects on chloroform were assumed
to be due to enhanced chloroform metabolism.
Steroids appear to play a role in the potentiation of chloroform toxicity,
especially in the kidney as seen from the sex-related differences in the response
of mice (Eschenbrenner and Miller, 1945a; Deringer et al., 1953) and by experi-
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rnents involving testosterone administration to castrated male mice
(Eschenbrenner and Miller, 19l*5b; Culliford and Hewitt, 1957; Hill, 1978)
discussed previously. Clemens et al. (1979) further studied this phenomenon in
castrated male and intact female mice. Dose-dependent testosterone sensitiza-
tion of renal tubules to a fixed dose of chloroform was observed in castrated
males with the response ranging from kidney dysfunction to death at high doses.
The androgenic progestin, medroxyprogesterone acetate, enhanced chloroform-
induced kidney damage in both castrated males and intact females. Progesterone
or hydrocortisone potentiated chloroform toxicity in DBA/2J castrated male mice,
but not in the C57BL/J6 strain males nor in any of the females. The mechanism by
which the androgens exerted their potentiation may have been mediated through
strain specific androgen receptors of the proximal convoluted tubular cells. The
mechanism for the potentiating action of the other steriods was less clear.
The potentiation of chloroform toxicity by ketones and ketogenic substances
has been studied extensively in recent years. Hewitt et al. (1979) and Cianflone
et al. (1980) found that while pretreatment of mice with the insecticide, kepone
(a ketone), enhanced the liver damage caused by chloroform exposure, the struc-
turally related mirex (a non-ketone) did not. Other ketones were compared for
their ability to enhance the hepatotoxic and nephrotoxic action of chloroform in
rats with the following results: methyl n-butyl ketone and 2, 5-hexanedione were
the most potent enhancers, followed by acetone, followed by n-hexane (a ketogenic
chemical) (Hewitt et al., 1980). Jernigan and Harbison (1982) studied the
potentiation by 2,5-hexanedione of chloroform hepatotoxicity in mice with
specific reference to sex differences. Female mice were more susceptible to the
dose-dependent enhancement of chloroform hepatotoxicity as well as to the
dose-related increase in hepatic microsomal enzymes. Pretreatment of male mica
with 2>5-hexanedione potentiated the toxicity of deuterated chloroform but to a
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lesser extent than chloroform; however, this deuterium isotope effect did not
occur in female mice in_ vivo. Phenobarbital pretreatment did elicit a deuterium
isotope effect in female mice iri vivo, suggesting that 2,5-hexanedione pretreat-
ment altered chloroform metabolism by a different mechanism than did pheno-
barbital. Jernigan and Harbison (1982) speculated that perhaps female mice have
greater microsomal enzyme activities, different membrane properties, or perhaps
produce a different reactive metabolite of chloroform than do males.
The mechanism of ketone potentiation of chloroform-induced hepato- and
nephrotoxicity was also investigated by Branchflower and Pohl (1981) using
methyl n-butyl ketone (MBK). Male rats were pretreated with MBK followed by
chloroform administration. The metabolism of chloroform by liver and kidney
microsomal enzymes and the toxicity to these organs were examined. Control
experiments were conducted in which rats were either not pretreated, not given
chloroform, or given deuterated chloroform (CDC1 ) instead of chloroform. MBK
increased cytochrome P-450 levels and NADPH-dependent cytochrome reductase
activity in liver microsomes, while having no effect on renal levels of these
microsomal components. MBK pretreatment doubled the rate of metabolism of
chloroform to diglutathionyl dithiocarbonate (GSCOSG) in microsomal prepara-
tions, and more GSCOSG was excreted into the bile of the pretreated animals when
compared with rats receiving only chloroform. The amount of GSCOSG in bile was
less in MBK-CDC1,. treated animals. GSH levels were significantly decreased by
MBK treatment and this decrease was enhanced following chloroform exposure, and
to a lesser extent, following CDC1,, exposure. Rats pretreated with MBK followed
by chloroform had greatly elevated levels of SGPT associated with liver necrosis
and significantly greater BUN levels associated with renal cortical tubule
lesions over the control groups. A mechanism was proposed whereby MBK, by
increasing cytochrome P-450 levels, enhanced the metabolism of chloroform to
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phosgene. Furthermore, according to the hypothesis, the phosgene was converted
to GSCOSG through GSH, levels of which were diminished by MBK, because the more
phosgene formed, the more GSH was depleted in the reaction. The results with
CDC1.J indicated that C-H bond was involved in the mechanism. Although MBK also
potentiated chloroform toxicity to the kidney, a different mechanism may have
been involved since renal cytochrome P-450 and renal GSH levels were not
affected.
5.4.2 Factors that Decrease the Toxicity. As discussed above, the experiments
of Sato et al. (1980, 1981) indicate that ethanol is both a stimulator and an
inhibitor of microsomal enzymes, and hence, of chloroform metabolism and
toxicity, depending upon the length of time after ingestion.
Disulfiram and its metabolites have also been studied with respect to their
protective effects of chloroform'-induced hepatotoxicity (Scholler, 1970; Masuda
and Nakayama, 1982). Disulfiram is used in treating chronic alcoholism and is
metabolized to carbon disulfide and diethyldithiocarbamate (a herbicide) (IARC,
1976). Disulfiram, a known inhibitor of the microsomal drug metabolizing
enzymes, given to rats prior to chloroform anesthesia completely prevented the
elevated SGPT activity and liver necrosis observed in rats administered chloro-
form alone (Scholler et al., 1970). More recently, Masuda and Nakayama (1982)
studied the effects of diethyldithiocarbamate and carbon disulfide pretreatment
in mice challenged with chloroform. Both substances had a protective effect as
measured by SGPT activity, liver calcium content, and centrilobular necrosis.
Diethyldithiocarbamate and carbon disulfide decreased the activity of the drug
metabolizing enzymes rn vivo and iii vitro, but only in the presence of NADPH,
indicating that these substances must first be metabolized before exerting their
inhibitory effect on chloroform metabolism. Gopinath and Ford (1975) also found
that dithiocarbamate and carbon disulfide protected against chloroform hepato-
-------
toxicity in rats, and the effect was presumed to be due to suppression of the
drug metabolizing enzymes.
Dietary components can also alter the toxicity of chloroform. It is a
widely held opinion that low protein content of the diet decreases microsomal
enzyme activities, while a high protein diet increases the activities (McLean and
McLean, 1969; Nakajima et al., 1982). If this is the case, a low protein diet
should protect against chloroform hepatotoxicity by inhibiting the enzymes
responsible for chloroform metabolism. It was found, however, that protein
depletion did not alter the toxicity of chloroform in rats given a single oral
dose (McLean and McLean, 1969; McLean, 1970). If pretreated with phenobarbital
or NDDT to induce microsomal enzymes, rats maintained on a standard diet were no
more susceptible to chloroform-induced liver damage that were pretreated,
protein-depleted rats (McLean, 1970).
More recently, Nakajima et al. (1982) studied the individual effects of
protein, fat, and carbohydrate on the metabolism of chloroform in relation to its
toxicity in male rats. Test diets were varied with respect to carbohydrate,
protein, or fat while maintaining isocaloric contents. Microsomal enzymes were
prepared and chloroform was added as a substrate. The following results were
obtained: decreased food intake increased liver microsomal enzyme activities;
decreased sucrose content in the diet increased the metabolic rate; varying the
protein and fat content, while holding the sucrose content constant, had no
effect on the metabolic rate; a carbohydrate-free diet, which contained high
protein and high fat, accelerated the rate of chloroform metabolism almost as
much as 1 day of food deprivation. The authors concluded that it is a high
carbohydrate content, rather than a low protein content, which is responsible for
the decreased microsomal enzyme metabolism of chloroform and, hence, its
toxicity.
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5.5 SUMMARY; CORRELATION OF EXPOSURE AND EFFECT
The purpose of this section is to delineate dose-response relationships for
the systemic toxicity of chloroform.
5.5.1 Effect of Acute Inhalation Exposure. The adverse effects on humans of
inhaling high concentrations of chlorof^opi have been well documented in the
course of its use as an anesthetic. Studies that define the threshold region of
exposure for such effects in humans are, however, sparse at best. Experiments
involving subchronic exposure of several species of animals give some informa-
tion on toxicity thresholds for renal and hepatic effects, but little for CNS and
none for cardiovascular effects.
The only experimental studies conducted with humans (Lehmann and Hasegawa,
1910; Lehmann and Schmidt-Kehl, 1936) involved relatively short exposures and
subjective responses. The results of these studies indicate that the odor of
chloroform can be perceived at about 200 ppm. Subjective CNS effects (dizziness,
vertigo) apparently did not occur at 390 ppm during a 30-minute exposure but were
perceived at about 900 ppm after 2-3 minutes of exposure. Subjects exposed to
1400 ppm for 30 minutes experienced tiredness and headache in addition to the
above CNS symptoms. The threshold for "light intoxication" was about 4300 ppm
(20 minutes). An exposure duration of 30 minutes or less is insufficient to
achieve pulmonary steady state (or total body equilibrium, Section 4.2.3).
Hence, longer exposures at these concentrations would be expected to cause more
severe effects.
Chloroform concentrations used for the induction of anesthesia ranged from
about 20,000-40,000 ppm (NIOSH, 1974; Adriani, 1970) and for the maintenance of
anesthesia ranged from 1500 ppm (light anesthesia) to 15,000 ppm (deep
anesthesia) (Goodman and Oilman, 1980). Continued exposure to 20,000 ppm could
result in respiratory failure, direct depression of the myocardium, and death
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(Section 5.1.1). Levels of exposure sufficient to produce anesthesia have also
caused cardiac arrythmias and extrasystoles (Kurtz et al., 1936; Orth et al.,
1951) and hepatic necrosis and fatty degeneration (Goodman and Oilman, 1980;
Wood-Smith and Stewart, 1964).
Data from acute animal exposures tend to show similar CNS effects at roughly
the same levels of exposure that produced these effects in humans (Lehmann and
Flury, 19^3). In addition, some data on the threshold for hepatic effects has
been obtained for mice. Kylin et al. (1963), in experiments with female mice of
an unspecified strain, found that single, 4-hour exposures to chloroform
produced mild hepatic effects (increased incidence of moderate fatty infiltra-
tion) at 100 ppm. At 200 ppm, in addition to fatty infiltration, hepatic
necrosis and increased serum ornithine carbamyl transferase activity occurred.
(An elevation in serum levels of this enzyme indicates liver damage according to
Divincenzo and Krasavage, 1974.) Further increases in fatty infiltration,
necrosis, and serum enzyme activity were observed at 400 and 800 ppm. These
effects appeared to be reversible because the extent of change was less severe 3
days after exposure than it was 1 day after exposure.
Damage to the kidneys of male mice of sensitive strains (e.g., C3H) has
occurred at exposure levels as low as 5 mg/£ (1025 ppm) for 1 hour (Deringer et
al., 1953). The damage consisted of necrosis of the epithelium of the proximal
tubules.
5.5.2 Effects of Acute Oral Exposure. Dose-response data for acute oral expo-
sure of humans to chloroform is limited to case reports. A fatal dose of as
little as 1/3 ounce (10 m£) was reported (Schroeder, 1965).
A variety of dose-response data is available for acute oral administration
of chloroform to animals. Single doses that were sufficient to adversely affect
kidney function (measured as excessive loss of glucose and/or protein in the
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urine in male mice) ranged from 89-149 ing/kg in sensitive and relatively Insensi-
tive strains (Hill, 1978). At single oral doses of 1071 mg/kg, but not 756 or 546
ing/kg chloroform, increases in organ weights and mild to moderate lesions were
observed in the livers and kidneys of Sprague-Dawley rats (Chu et al., 1982a).
In male B6C3F1 mice, renal necrosis occurred after 20 mg/kg and focal tubular
regeneration occurred after 60 or 240 mg/kg but not after 15 mg/kg (Reitz et al.,
1980). A low observed adverse-effect level (LOAEL) for hepatic effects in mice
can be identified from the study of Jones et al., 1958, in which 30 mg/kg caused
midzonal fatty infiltration. Doses in the range of 133-355 mg/kg (Jones et al.,
1958; Reitz et al., 1980) represent a PEL (Frank-Effect-Level) for hepatic damage
(including centrilobular necrosis) in mice. According to Torkelson et al.
(1976), rats given "as little as" 250 mg/kg chloroform "showed adverse effects"
on liver and kidney as determined by gross pathological examination. Reported
oral LD values for mice ranged from 119-1400 mg/kg, depending on sex, strain,
and age (Kimura et al., 1971; Hill, 1978; Bowman et al., 1978). For rats, LD™
values of 908-2000 mg/kg have been reported (Chu et al., 1980; Torkelson et al.,
1976). The lethal dose studies included both 24-hour and delayed deaths.
5.5.3 Effects of Dermal Exposure. Chloroform is irritating to the skin. It
has been reported to cause degenerative changes in the renal tubules of rabbits
exposed dermally to high doses under extreme conditions (1-3-98 g/kg body weight
for 24 hours under an impermeable plastic cuff) (Torkelson et al., 1976). In
humans, toxicity from dermal exposure is probably not important in comparison
with other routes.
5.5.4 Effects of Chronic Inhalation Exposure. Limited information on the
effects of long-term intermittent exposure of humans or animals to chloroform is
available. A study involving a small number of workers (Challen et al., 1958)
o
indicates that long-term exposure to 20-71 ppm (98-346 mg/m ) for a 4-8 hour
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workday, with occasional brief excursions to =1163 ppm, may represent a LOAEL for
symptoms of CNS toxicity. No evidence of liver damage or other organic lesion
was detected by physical examination and clinical chemistry tests. A single
report linking liver enlargement and viral hepatitis to occupational exposure to
10-200 ppm chloroform (Bomski et al., 1967) is flawed by the apparent lack of
suitable controls. The available data do not define a NOAEL (no-observed-
adverse-effect-level) or NOEL (no-observed-effect- level) for humans.
Experiments with several species of animals (Torkelson et al., 1976) give
some information regarding the threshold region for hepatic and renal effects of
inhalation exposure to chloroform (Table 5-3). The animals were exposed to
chloroform 5 days/week for 6 months. Exposure to 25 ppm of chloroform for up to 4
hours/day had no adverse effects in male rats as judged by organ and body weights
and probably the gross and microscopic appearance of at least the liver and
kidneys, although the authors were not explicit about the latter. Exposure to 25
ppm for 7 hours/day, however, produced histopathologic changes in the livers and
kidneys of male but not female rats. These changes were characterized as lobular
granular degeneration and focal necrosis throughout the liver and cloudy
swelling of the kidneys. The hepatic and renal effects appeared to be reversible
because rats exposed in the same way, but given a 6-week "recovery period" after
the exposure period, appeared normal by the criteria tested. Increasingly
pronounced changes were observed in the livers and kidneys of both sexes of rats
exposed to 50 or 85 ppm. Hematologic, clinical chemistry, and urinalysis values,
tested at the two higher levels of exposure, were "within normal limits."
Similar experiments with guinea pigs and rabbits gave somewhat inconsistent
results. Histopathological lesions were observed in liver and kidneys of both
species at 25 ppm but not at 50 ppm in either species and not in guinea pigs even
at 85 ppm (Torkelson et al., 1976).
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The experiments of Torkelson et al. (1976) indicate that subchronic expo-
sure to 25 pprn (123 mg/m^), U hours/day, 5 days/week represents a NOAEL and
exposure to ?5 ppm, 7 hours/day, 5 days/week respresents a LOAEL for rats.
Guinea pigs and rabbits may be slightly less sensitive.
5.5.5 Effects of Chronic Oral Exposure. Little dose-response data for oral
exposure of humans to chloroform appear to be available (Chapter 5). A single
controlled study has been performed. In this study, subjects were exposed to 70
or 178 mg of chloroform/day (-1 or 2.5 mg/kg/day assuming 70 kg body weight) for
at least 1 year (DeSalva et al., 1975). Neither liver function tests nor blood
urea nitrogen determinations (a measure of kidney function) revealed statis-
tically significant differences between exposed and control subjects. Case
reports involving abuse of medicines containing chloroform (Wallace, 1950;
Conlon, 1963) are not adequate for risk assessment because of the small numbers
of patients, exposure to other agents, and imprecise estimates of intake.
Subchronic and chronic toxicity experiments with rats, mice, and dogs, when
considered together (Table 5-4), do not clearly establish a NOAEL or NOEL.
Although no adverse effects were observed in four strains of mice given 17
mg/kg/day of chloroform, 6 days/week for 2 years (Roe et al., 1979), at the
lowest dose level tested (i.e., 15 mg/kg/day, 6 days/week for 7.5 years) in dogs,
chloroform treatment was associated with an elevation in SGPT in some but not all
of the other tested clinical chemistry indicies of hepatic damage (Heywood et
al., 1979). The livers of dogs treated with chloroform at this dosage level had
larger and more numberous "fatty cysts" than were found in controls. These fatty
cysts consisted of aggregations of vacuolated histiocytes. No effect on
survival, growth, organ weights, gross and histological appearance of other
organs, or hematologic or urinalysis values was observed at this dosage level.
Hence 15 mg/kg/day (6 days/week) represents a LOAEL for dogs for effects on the
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liver. Chronic oral administration of 60 mg/kg/day of chloroform (6 days/week)
was associated with slight hepatic changes in rats (Palmer et al., 1979) and with
increased incidences of moderate to severe renal disease in male mice of sensi-
tive strains (Roe et al., 1979).
None of the three species tested in long-term experiments appeared to be
markedly more sensitive to the toxicity of chloroform than any other; dogs may
have been slightly more sensitive. There were considerable differences among
strains of mice in the sensitivity of the males to chloroform nephrotoxicity, as
had also been observed in acute toxicity experiments.
As would be expected, dosages that produced little or no histologic or
clinical chemistry evidence of toxicity when given subchronically (15 and 30
mg/kg/day; rats, dogs) resulted in greater evidence of toxicity when given for
longer periods of time (Palmer et al., 1979; Heywood et al., 1979). The response
to chloroform in the long-term studies may have been modified by the presence or
absence of intercurrent respiratory and renal disease, but no consistent pattern
is obvious from an inspection of the data in Table 5-1*.
5.5.6 Target Organ Toxicity. Target organs characteristic of the acute
toxicity of chloroform are the central nervous system, liver, kidney, and heart.
For chronic exposure to chloroform, characteristic target organs are the liver
and kidney, and possibly the central nervous system. Some dose-response data are
available for the toxicity of chloroform to the liver, kidney, and central
nervous system; these data are summarized in Table 5-5 by target organ. The
studies from which these data are drawn are discussed more fully elsewhere in
Chapter 5, but a comparison on the basis of endpoinb (target organ) was also
considered to be useful.
Manifestations of liver damage include centrilobular necrosis, vacuoliza-
tion, disappearance of glycogen, fatty degeneration and swelling (Groger and
5-51
-------
Table 5-5
Target Organ Toxicity of Chloroform
Target
Organ
Route and
Type of
Exposure
Species
Dose or
Exposure
Effect on
Target Organ
Reference
liver inhalation, acute
(surgical anesthesia)
human
liver inhalation, acute
liver inhalation, acute
liver inhalation, chronic
liver inhalation, chronic
liver inhalation, chronic
mice
mice
rats
rats
rats
induction =
20,000-40,000 ppm x
a few minutes, plus
maintenance = 1500-
15,000 ppm x variable
duration
100 ppm x 4 hours,
single exposure
200 ppm x 4 hours,
single exposure
25 ppm, 4 hours/day,
5 days/week x
6 months
25 ppm, 7 hours/day,
5 days/week x
6 months
50 or 85 ppm, 7 hours/
day, 5 days/week x
6 months
necrosis, fatty degen-
eration in some
patients
fatty infiltration
necrosis, fatty infil-
tration, increase in
SOCT
no effect
lobular granular
degeneration, focal
necrosis
marked centrilobular
granular degeneration
NIOSH, 1974;
Goodman and
Gilman,
1980;
Wood-Smith
and Stewart,
1964
Kylin et al.,
1963
Kylin et al.,
1963
Torkelson
et al., 1976
Torkelson
et al., 1976
Torkelson
et al., 1976
-------
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
£
oo
Target
Organ
liver
liver
liver
liver
Route and
Type of
Exposure
oral, acute
oral, acute
oral, acute
oral, acute
Dose or
Species Exposure
mice 30 mg/kg bw;
single dose
mice 133 mg/kg, single
dose
mice 355 mg/kg, single
dose
mice 60 mg/kg, single
Effect on
Target Organ Reference
fatty infiltration Jones et al . ,
1958
massive fatty infiltra- Jones et al.,
tion and severe necrosis 1958
massive fatty infiltra- Jones et al.,
tion and severe necrosis 1958
no effect Reitz et al.,
liver oral, acute
mice
liver oral (drinking water) rats
subchronic
liver oral (drinking water) mice
subchronic
dose
240 mg/kg, single
dose
20, 38, 57, 81, or
160 mg/kg/day x
90 days
64 or 97 mg/kg/day x
90 days
hepatocellular necrosis
and swelling; inflamma-
tion
transient hepatosis
(at 30 and 60, but
not at 90 days)
transient centrilobular
fatty change (at 30
and 60 but not at
90 days)
1980
Reitz et al.,
1980
Jorgenson
and
Rushbrook,
1980
Jorgenson
and
Rushbrook,
1980
-------
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
ui
-Cr
Target
Organ
liver
liver
liver
liver
liver
liver
liver
liver
Route and
Type of
Exposure Species
oral (drinking water) mice
subchronic
oral (gavage) rats
subchronic
oral (capsule), dogs
subchronic
ora] (capsule), dogs
subchronic
oral (capsule), dogs
subchronic
oral (capsule), dogs
subchronic
oral (gavage), rats
chronic
oral (gavage), mice
chronic
Dose or
Exposure
145 or 190 mg/kg/day x
90 days
410 mg/kg/day, 6 days/
week x 13 weeks
30 mg/kg/day, 7 days/
week x 13 weeks
45 mg/kg/day, 7 days/
week x 13 weeks
60 mg/kg/day, 7 days/
week x 18 weeks
120 mg/kg/day, 7 days/
week x 12 weeks
60 mg/kg/day, 6 days/
week x 80 weeks
60 mg/kg/day, 6 days/
week x 80 weeks
Effect on
Target Organ
fatty change
fatty change and
necrosis
no effect
slight fatty change
fatty degeneration,
increase in SCOT and
SGPT
fatty degeneration,
jaundice, increase
in SCOT, SGPT, bili-
rubin
minor histological
changes and decrease
in relative liver weight
no effect
Reference
Jorgenson
and
Rushbrook,
1980
Palmer et al . ,
1979
Heywood
et al., 1979
Heywood
et al., 1979
Heywood
et al., 1979
Heywood
et al., 1979
Palmer et al . ,
1977
Roe et al . ,
1979
-------
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
liver
Route and
Type of
Exposure
oral (capsule),
Species
dogs
Dose or
Exposure
15 or 30 mg/kg/day,
Effect on
Target Organ
increases in SGPT and
Reference
Heywood et al . ,
chronic
liver oral, chronic
liver dermal, acute
kidney inhalation, acute
kidney inhalation, chronic
kidney oral, acute
humans
rabbits
mice, males
of sensitive
strains
rats
mice, males
of sensitive
strains
6 days/week x 7.5
years
2.5 mg/kg/day for
>_1 year
3.98 g/kg x 24 hours
under plastic cuff,
single exposure
5000 ppm, 1 hour,
single exposure
25, 50, or 85 ppm,
7 hours/day, 5 days/
week x 6 months
89 mg/kg, single dose
other serum indicators
of hepatic damage,
increase in size and
number of fatty cysts
(vacuolated histiocytes)
no effect
no macroscopic
pathologic changes
necrosis and calcifi-
fication of tubular
epithelium
cloudy swelling of
tubular epithelium
loss of glucose or
protein in urine
1979
De Salva
et al., 1975
Torkelson
et al., 1976
Deringer
et al., 1953
Torkelson
et al., 1976
Hill, 1978
-------
ON
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
kidney
kidney
kidney
kidney
kidney
kidney
kidney
Route and
Type of
Exposure
oral, acute
oral, acute
oral, acute
oral, acute
oral (drinking water)
subchronic
oral (drinking water)
subchronic
oral (drinking water)
subchronic
Dose or
Species Exposure
mice, males 149 mg/kg, single dose
of sensitive
strains
mice, male 15 mg/kg, single dose
mice, male 60 mg/kg, single dose
mice, male 240 mg/kg, single dose
rats 160 mg/kg/day x
90 days
rats =300 mg/kg/day x
90 days
mice 290 mg/kg/day x
90 days
Effect on
Target Organ
loss of glucose or
protein in urine
no effect
focal tubular
epithelial regenera-
tion
severe diffuse cortical
necrosis, focal tubular
epithelial regeneration
no effect
no effect
no effect
Reference
Hill, 1978
Reitz et al. ,
1980
Reitz et al . ,
1980
Reitz et al. ,
1980
Jorgenson
and
Rushbrook,
1980
Chu et al . ,
1980b
Jorgenson
and
Rushbrook,
1980
-------
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
kidney
kidney
v* kidney
kidney
kidney
kidney
Route and
Type of
Exposure
oral (capsule),
subchronic
oral, chronic
oral (gavage),
chronic
oral (gavage),
chronic
oral (gavage),
chronic
oral (gavage)
chronic,
Species
dogs
humans
rats
mice
mice , males
of sensitive
strains
mice , males
of sensitive
Dose or
Exposure
120 mg/kg/day,
7 days /week x
12 weeks
2.5 mg/kg/day,
7 days /week, for
>1 year
200 mg/kg/day,
5 days /week x
78 weeks
138 or 227 mg/kg/day,
5 days /week x
78 weeks
17 mg/kg/day,
6 days /week x
80 weeks
60 mg/kg/day,
6 days /week x
Effect on
Target Organ
no effect
no effect on BUN
no effect
decreased incidence
of renal disease
no effect
increased incidence
of moderate to severe
Reference
Jorgenson
and
Rushbrook,
1980
De Salva
et al., 1975
NCI, 1976
NCI, 1976
Roe et al . ,
1979
Roe et al . ,
1979
strains
80 weeks
renal disease
-------
CNS
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
kidney
kidney
kidney
CP,
vil
OO
kidney
central
nervous
system
(CNS)
Route and
Type of
Exposure
oral (gavage)
chronic
oral (capsule) ,
chronic
oral (capsule),
chronic
dermal, acute
inhalation, acute
Species
mice, males
of insensi-
tive strains,
females
dog
dog
rabbits
humans
Dose or
Exposure
60 mg/kg/day,
6 days /week x
80 weeks
15 mg/kg/day
6 days/week x
7.5 years
30 mg/kg/day,
6 days /week x
7.5 years
1.0, 2.0, and
3.98 g/kg x 24 hours
under plastic cuff,
single exposure
900-1400 ppm for
>30 minutes, single
exposure
Effect on
Target Organ
no effect
no effect
increase in fat
deposition in glomeruli
degenerative changes
in tubules
dizziness, tiredness,
headache
Reference
Roe et al. ,
1979
Heywood et al . ,
1979
Heywood
et al., 1979
Torkelson
et al., 1976
Lehman n
and
Hasegawa, 1910;
Lehmann and
inhalation, acute
humans
M300-5100 ppm x
20 minutes, single
exposure
dizziness, light
intoxication
Schmidt-Kehl,
1936
Lehmann and
Hasegawa, 1910;
Lehmann and
Schmidt-Kehl,
^936
-------
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
ui
Target
Organ
CNS
Route and
Type of
Exposure
inhalation, acute
Species
humans
Dose or
Exposure
1500-2000 ppm, si
Effect on
Target Organ
ngle maintenance of light
Reference
Goodman and
CNS inhalation, acute
CNS inhalation, acute
CNS inhalation, acute
CNS inhalation, acute
CNS inhalation, acute
CNS inhalation, acute
CNS inhalation, acute
humans
humans
mice
mice
mice
cats
cats
exposure
15,000 ppm, single
exposure
20,000-40,000 ppm x
a few minutes, single
exposure
2500 ppm x 12 hours,
single exposure
3100 ppm x 1 hour,
single exposure
4100 ppm x 0.5 hours,
single exposure
7200 or 21,500 ppm x
5 minutes, single
exposure
7200 ppm x 60 minutes
single exposure
anesthesia (after
induction)
maintenance of heavy
anesthesia (after
induction)
Gillman, 1980
Goodman and
Gillman, 1980
induction of anesthesia NIOSH, 197H
Adriani, 1970
no obvious effects
light narcosis
deep narcosis
disturbance of equilib-
rium
light narcosis
Lehmann and
Flury, 1943
Lehmann and
Flury, 1943
Lehmann and
Flury, 1943
Lehmann and
Flury, 1943
Lehmann and
Flury, 1943
-------
CNS
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
CNS
CNS
CNS
CNS
Route and
Type of
Exposure
inhalation, acute
inhalation, acute
inhalation, acute
inhalation, chronic
Species
cats
cats
cats
humans
Dose or
Exposure
7200 ppm x 93 minutes
single exposure
21,500 ppm x 10 minutes
single exposure
21,500 ppm x 13 minutes
single exposure
20-71 ppm (with excur-
Effect on
Target Organ
deep narcosis
light narcosis
deep narcosis
tiredness
Reference
Lehmann and
Flury, 19^3
Lehmann and
Flury, 19^3
Lehmann and
Flury, 1943
Challen
inhalation, chronic
humans
sions to 1163 ppm
lasting 1.5-2 minutes)
for 4-8 hours/day,
5 days/week
77 to 237 ppm (with
excursions to =1163
lasting 1.5-2 minutes)
for 4-8 hours/day,
5 days/week
et al., 1958
tiredness, depression,
occasional silliness
or staggering during
the workday
Challen
et al., 1958
-------
01
Table 5-5
Target Organ Toxicity of Chloroform (cont.)
Target
Organ
CNS
CNS
CNS
CNS
CNS
Route and
Type of
Exposure Species
oral, acute rats
oral (drinking water), rats
subchronic
oral (drinking water), mice
subchronic
oral (drinking water) rats
subchronic
oral (gavage), rats
subchronic
Dose or
Exposure
350 mg/kg
single dose
20-160 mg/kg/day x
90 days
32-290 mg/kg/day x
90 days
=300 mg/kg/day x
90 days
60 rag/kg/day
6 days /week x
80 weeks
Effect on
Target Organ
minimum narcotic dose
(MND )
dose-related signs
of depression during
1st week only
dose-related signs of
depression during 1st
week only
no histopathologic
changes in brain
no effect on gross
or histologic appear-
ance of brain
Reference
Jones et al . ,
1958
Jorgenson
and
Rushbrook,
1980
Jorgenson
and
Rushbrook ,
1980
Chu et al . ,
1982b
Palmer et al . ,
1977
-------
Grey, 1979). In the kidney, chloroform exposure produces necrosis of the
proximal and distal convoluted tubules (Eschenbrenner and Miller, 1945a). The
mechanism by which chloroform produces these effects has been extensively
studied in experimental animals. From the studies summarized in Section 5.3.1
(Brown et al., 1974; Ilett et al., 1973; Docks and Krishna, 1976; Ekstrbm et al.,
1981; Lavigne and Marchand, 1974; McMartin et al., 1981; Masuda et al., 1980;
Harris et al., 1982; Stevens and Anders, 1981), it appears that chloroform is
first metabolized in the target organ by microsomal drug metabolizing enzymes to
a reactive intermediate, probably phosgene, which in turn can react by various
pathways, depending on glutathione levels, one of which is the covalent binding
to liver proteins resulting in necrotic lesions. A similar mechanism may or may
not occur in the kidney (Kluwe and Hook, 1981).
5.5.7 Factors that Modify the Toxicity of Chloroform. Several substances
alter the toxicity of chloroform, most probably by modifying the metabolism of
chloroform to a reactive intermediate (see Section 5.4). These substances are of
interest because humans may be accidentally or intentionally exposed to them.
Factors that potentiate the toxic effects induced by exposure to chloroform
include ethanol (Kutob and Plaa, 1962; Sato et al., 1980, 1981), polybrominated
biphenyls (Kluwe and Hook, 1978), steroids (Clemens et al., 1979), and ketones
(Hewitt et al., 1979; Jernigan and Harbison, 1982; Branchflower and Pohl, 1981).
Disulfiram and its metabolites (Scholler et al., 1970; Masuda and Nakayama, 1982;
Gopinath and Ford, 1975) and high carbohydrate diets (Nakajima et al., 1982),
appear to protect against chloroform toxicity.
5-62
-------
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Lehmann, K.B. and Hasewaga. 1910. [Studies of the absorption of chlorinated
hydrocarbons in animals and humans.] Arch. Hyg. 72: 327-342. (Ger.) (Cited in
NIOSH, 1974.)
5-69
-------
Lehmann, K.B. and L. Schmidt-Kenl. 1936. [The thirteen most important chlor-
inated aliphatic hydrocarbons from the standpoint of industrial hygiene.] Arch.
Hyg. 116: 131-200. (Ger.) (Cited in NIOSH, 1974.)
Lucas, G.H.W. 1928. A study of the fate and toxicity of bromine and chlorine
containing anesthetics. Jour. Pharmacol. Exp. Ther. 3**: 223.
Mansuy, D., P. Beaune, T. Cresteil, M. Lange and J.P. Lerous. 1977. Evidence
for phosgene formation during liver microsomal oxidation of chloroform.
Biochem. Biophys. Res. Commun. 79: 513-517.
Masuda, Y. and N. Nakayama. 1982. Protective effect of diethyldithiocarbamate
and carbon disulfide against liver injury induced by various hepatotoxic agents.
Biochem. Pharmacol. 31: 2713-2725.
Masuda, Y., I. Yano and T. Murano. 1980. Comparative studies on the hepatotoxic
actions of chloroform and related halomethanes in normal and phenobarbital-
pretreated animals. Jour. Pharmacobio. Dyn. 3: 53-6M.
McLean, A.E.M. and E.K. McLean. 1969. Diet and toxicity. Brit. Med. Bull.
25: 278-281.
McLean, A.E.M. 1970. The effect of protein deficiency and microsomal enzyme
induction by DDT and phenobarbitone on the acute toxicity of chloroform and a
pyrrolizidiue alkaloid, retrorsine. Brit. Jour. Exp. Pathol. 51: 317-321.
5-70
-------
McMartin, N.D., J.A. O'Connor, Jr., L.S. Kaminsky. 1981. Effects of differen-
tial changes in rat hepatic and renal cytochrome P-450 concentrations on hepato-
toxicity and nephrotoxicity of chloroform. Res. Commun. Chem. Pathol.
Pharmacol. 31: 99-110.
Nakajima, T., Y. Koyama and A. Sato. 1982. Dietary modification of metabolism
and toxicity of chemical substances with specific reference to carbohydrate.
Biochem. Pharmacol. 31: 1005-1011.
NCI (National Cancer Institute). 1976. Report on Carcinogenesis Bioassay of
Chloroform. Available from National Technical Information Service, Springfield,
Virginia. (NTIS PB-26H-018).
NIOSH (National Institute for Occupational Safety and Health). 1974. Criteria
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National Technical Information Service, Springfield, Virginia. (NTIS
P8-246-695).
Oettel, H. 1936. Arch. Exp. Pathol. Pharmacol. 183: 655.
Orth, O.S., R.R. Liebenow and R.T. Capps. 1951. III-The effect of chloroform on
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Palmer, A.K., A.E. Street, F.J.C. Roe, A.N. Worden and N.J. Van Abbe'. 1979.
Safety evaluation of toothpaste containing chloroform. II. Long-term studies in
rats. Jour. Environm. Pathol. Toxicol. 2: 821-833-
5-71
-------
Paul, B.B. and D. Rubinstein. 1963- Metabolism of carbon tetrachloride and
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Piersol, G.M., A.J. lumen and L.S. Kau. 1933- Fatal poisoning following the
ingestion of chloroform. Med. Clin. North Am. 17: 587-601.
Pohl, L.R., B. Bhooshan, N.F. Whittaker and G. Krishna. 1977. Phosgene: A
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Reitz, R.H., J.F. Quast, W.T. Stott, P.G. Watanabe and P.J. Gehring. 1980.
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1M
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5-72
-------
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5-73
-------
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5-74
-------
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5-75
-------
United States
Environmental Protection
Agency
Office of Health and
Environmental Assessment
Washington DC 20460
EPA-600/8-84-004A
March 1984
External Review Draft
&EPA
Research and Development
Health Assessment
Document for
Chloroform
Part 2 of 2
Review
Draft
(Do Not
Cite or Quote)
NOTICE
This document is a preliminary draft. It has not been formally
released by EPA and should not at this stage be construed to
represent Agency policy. It is being circulated for comment on its
technical accuracy and policy implications.
-------
Review
Draft
(Do Not
Cite or Quote)
EPA-600/8-84-004A
March 1984
External Review Draft
Health Assessment
Document for Chloroform
Part 2 of 2
External Review Draft
NOTICE
This document is a preliminary draft. It has not been formally released by the U.S. Environmental
Protection Agency and should not at this stage be construed to represent Agency policy. It
is being circulated for comment on its technical accuracy and policy implications.
U.S. ENVIRONMENTAL PROTECTION AGENCY
Office of Research and Development
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
Research Triangle Park, NC 27711
March 1984
-------
PREFACE
The Office of Health and Environmental Assessment has prepared this health
assessment to serve as a "source document" for EPA use. This health assessment
document was developed for use by the Office of Air Quality Planning and
Standards to support decision-making regarding possible regulation of chloroform
as a hazardous air pollutant.
In the development of the assessment document, the scientific literature
has been inventoried, key studies have been evaluated and summary/conclusions
have been prepared so that chemical's toxicity and related characteristics
are qualitatively identified. Observed effect levels and other measures of
dose-response relationships are discussed, where appropriate, so that the
nature of the adverse health response are placed in perspective with observed
environmental levels.
This document will be subjected to a thorough copy editing and proofinq
following the revision based on the EPA's Scientific Advisory Board review
comments.
-------
TABLE OF CONTENTS
LIST OF TABLES vi
LIST OF FIGURES x
1 . SUMMARY AND CONCLUSIONS 1 -1
2. INTRODUCTION 2-1
3. BACKGROUND INFORMATION 3-1
3.1 INTRODUCTION 3-1
3.2 PHYSICAL AND CHEMICAL PROPERTIES 3-2
3.3 SAMPLING AND ANALYSIS 3-4
3.3.1 Chloroform in Air 3-4
3.3.2 Chloroform in Water 3-5
3.3.3 Chloroform in Blood 3-6
3.3.4 Chloroform in Urine 3-6
3.3.5 Chloroform in Tissue 3-6
3.4 EMISSIONS FROM PRODUCTION AND USE 3-6
3.4.1 Emissions from Production 3-6
3.4.2 Emissions from Use 3-20
3.4.3 Summary of Chloroform Discharges from Use 3-26
3.5 AMBIENT AIR CONCENTRATIONS 3-26
3.6 ATMOSPHERIC REACTIVITY 3-32
3. 7 ECOLOGICAL EFFECTS/ENVIRONMENTAL PERSISTENCE 3-33
3.7.1 Ecological Effects 3-33
3.7.2 Environmental Persistence 3-36
3.8 EXISTING CRITERIA, STANDARDS, AND GUIDELINES 3-39
3.8.1 Air 3-39
3.8.2 Water 3-41
3.8.3 Food 3-41
3.8 .4 Drugs and Cosmetics 3-42
3.9 RELATIVE SOURCE CONTRIBUTIONS 3-42
3.10 REFERENCES 3-43
4. DISPOSITION AND RELEVANT PHARMACOKINETICS 4-1
4.1 INTRODUCTION 4-1
-------
TABLE OF CONTENTS (cont.)
5.
4.2
4.3
4.4
4.5
4.6
4.7
4.8
ABSORPTION ,
4.2.1 Dermal Absorption
4.2.2 Oral Absorption ,
4.2.3 Pulmonary Absorption ,
TISSUE DISTRIBUTION
EXCRETION
4.4.1 Pulmonary Excretion
4.4.2 Other Routes of Excretion
4.4.3 Adipose Tissue Storage
BIOTRANSFORMATION OF CHLOROFORM
4.5.1 Known Metabolites
4.5.2 Magnitude of Chloroform Metabolism
4.5-3 Enzymic Pathways of Biotransformation
COVALENT BINDING TO CELLULAR MACROMOLECULES
4.6.1 Proteins and Lipids
4.6.2 Nucleic Acids
4.6.3 Role of Phosgene
4.6.4 Role of Glutathione
SUMMARY
REFERENCES
TOXICITY
5.1
EFFECTS OF ACUTE EXPOSURE TO CHLOROFORM
5.1 .1 Humans
5.1 .2 Experimental Animals
Page
4-2
4-2
4-3
4-7
4-12
4-21
4-21
4-31
4-32
4-34
4-34
4-37
4-40
4-47
4-47
4-53
4-59
4-61
4-63
4-67
5-1
5-1
5-1
5-6
5.2 EFFECTS OF CHRONIC EXPOSURE TO CHLOROFORM 5-13
5.2.1 Humans 5-13
5.2.2 Experimental Animals 5-15
5.3 INVESTIGATION OF TARGET ORGAN TOXICITY IN EXPERIMENTAL
ANIMALS 5-28
5.3.1 Hepatotoxicity 5-28
5.3.2 Nephrotoxicity 5-34
IV
-------
TABLE OF CONTENTS (cont.)
Page
5.4 FACTORS MODIFYING THE TOXICITY OF CHLOROFORM 5-38
5.4.1 Factors that Increase Toxicity 5-39
5.4.2 Factors that Decrease Toxicity 5-44
5.5 SUMMARY; CORRELATION OF EXPOSURE AND EFFECT 5-46
5.5.1 Effects of Acute Inhalation Exposure 5-46
5.5.2 Effects of Acute Oral Exposure 5-47
5.5-3 Effects of Dermal Exposure 5-48
5.5.4 Effects of Chronic Inhalation Exposure 5-48
5.5.5 Effects of Chronic Oral Exposure 5-50
5.5.6 Target Organ Toxicity 5-51
5.5.7 Factors that Modify the Toxicity of Chloroform 5-62
5.6 REFERENCES 5-63
6 . TERATOGENICITY AND REPRODUCTIVE EFFECTS 6-1
6.1 REFERENCES 6-16
7. MUTAGENICITY 7-1
7.1 INTRODUCTION 7-1
7.2 COVALENT BINDING TO MACROMOLECULES 7-1
7. 3 MUTAGENICITY STUDIES IN BACTERIAL TEST SYSTEMS 7-3
7.4 MUTAGENICITY STUDIES IN EUCARYOTIC TEST SYSTEMS 7-9
7.5 OTHER STUDIES INDICATIVE OF DNA DAMAGE 7-14
7.6 CHROMOSOME STUDIES 7-19
7.7 SUGGESTED ADDITIONAL TESTING 7-21
7.8 REFERENCES 7-23
8. CARCINOGENICITY 8-1
8.1 ANIMAL STUDIES 8-1
8.1.1 Oral Administration (Gavage): Rat 8-2
8.1.2 Oral Administration (Gavage): Mouse 8-11
8.1.3 Oral Administration (Capsules): Dog 8-21
8.1.4 Intraperitoneal Administration: Mouse 8-24
8.1.5 Evaluation of Chloroform Carcinogenicity by
Reuber (1979) 8-26
8.1.6 Oral Administration (Drinking Water): Mouse:
Promotion of Experimental Tumors 8-26
8 .2 CELL TRANSFORMATION ASSAY 8-32
8.2.1 Styles (1979) 8-32
-------
TABLE OF CONTENTS (cont.)
8 . 3 EPIDEMIOLOGIC STUDIES 8-34
8.3.1 Young et al. (1981) 8-37
8.3-2 Hogan et al. (1979) 8-40
8.3.3 Cantor et al. (1978) 8-42
8.3.4 Gottlieb et al. (1981) 8-46
8.3.5 Alavanja et al. (1978) 8-48
8.3.6 Brenniman et al. (1978) 8-50
8.3.7 Struba (1979) 8-51
8.3-8 Discussion 8-53
8.4 QUANTITATIVE ESTIMATION 8-56
8.4.1 Procedures for the Determination of Unit Risk 8-59
8 .4.2 Unit Risk Estimates 8-70
8.4.3 Comparison of Potency with Other Compounds 8-80
8.4.4 Summary of Quantitative Assessment 8-80
8.5 SUMMARY 8-85
8.5.1 Qualitative 8-85
8.5.2 Quantitative 8-88
8.6 CONCLUSIONS 8-89
8.7 REFERENCES 8-91
8.8 APPENDIX A: Comparison Among Various Extrapolation
Models A-l
VI
-------
LIST OF TABLES
Table
3-1
3-2
3-3
3-4
3-5
3-6
3-7
3-8
3-9
3-10
3-11
3-12
4-1
4-2
4-3
14-4
4-5
4-6
4-7
Physical Properties of Chloroform
Chloroform Producers, Production Sites, and Capacities
Chloroform Discharges from Direct Sources
Ethylene Dichloride Producers, Production Sites and
Capacities
Chloroform Discharges from Indirect Sources
Chlorodifluoromethane Producers and Production Sites
Chloroform Discharges from Use
Relative Source Contribution for Chloroform
Ambient Levels of Chloroform
Acute and Chronic Effects of Chloroform on Aquatic
Organisms
Values for knu
Un
Summary of EXAMS Models of the Fate of Chloroform
Physical Properties of Chloroform and Other
Chloromethanes
Partition Coefficients for Human Tissue at 37°C
Retention and Excretion of Chloroform in Man During and
After Inhalation Exposure to Anesthetic Concentations
Chloroform Content in United Kingdom Foodstuffs and
in Human Autopsy Tissue
Concentration of Chloroform in Various Tissues of Two
Dogs After 2 . 5 Hours Anesthesia
Concentration of Radioactivity (Chloroform Plus
Metabolites) in Various Tissues of the Mouse (N MRI)
1 4
Tissue Distribution of C-Chloroform Radioactivity
in CF/LP Mice After Oral Administration (60 mg/kg)
Pajje
3-3
3^
3-1 3
3-15
3-21
3-23
3-27
3-28
3-29
3-34
3-38
3-40
4-4
4-5
4-9
4-14
14-16
4-18
4-20
vn
-------
LIST OF TABLES (eont.)
Table Page
1 3
4-8 Pulmonary Excretion of CHCL_ Following Oral
Dose: Percent of Dose 4-26
4-9 Species Difference in the Metabolism of C-Chloroform 4-29
4-4b Kinetic Parameters for Chloroform After I.V. Administration
to Rats 4-30
4-10 Levels of Chloroform in Breath of Fasted Normal
Healthy Men 4-35
1 ii
4-11 Covalent Binding of Radioactivity From C-Chloroform and
C-Carbon Tetrachloride in Microsomal Incubation
In Vitro 4-49
4-12 Mouse Strain Difference in Covalent Binding of Radioactivity
From C-Chloroform 4-53
1 4
4-13 In Vivo Covalent Binding of Radioactivity From CHC1
in Liver and Kidney of Male and Female Mice (C57BL/6) 4-55
1 4
4-14 In Vitro Covalent Binding of Radioactivity from CHC1_
to Microsomal Protein from Liver and Kidney of Male and
Female Mice (C57BL/6) 4-56
1 4
4-15 Coyalent Binding of Radioactivity from C-Chloroform and
C-Carbon Tetrachloride in Rat Liver Nuclear and
Microsomal Incubation I_n Vitro 4-60
4-16 Effect of Glutathione, Air, N or CO: 0- Atmosphere
on the In Vitro Covalent Binding of C Cl^ and C Br Cl
to Rat Liver Microsomal Protein 4-62
4-17 Effects of 24-Hour Food Deprivation on Chloroform and
Carbon Tetrachloride In Vitro Microsomal Metabolism,
Protein, and P-450 Liver Contents of Rats 4-64
5-1 Relationship of Chloroform Concentration in Inspired
Air and Blood to Anesthesia 5-2
5-2 Dose-Response Relationships 5-7
5-3 Effects of Inhalation Exposure of Animals to Chloroform,
Five Days/Week for Six Months 5-17
5-4 Effects of Subchronic or Chronic Oral Administration of
Chloroform to Animals 5-19
vm
-------
LIST OF TABLES (cont.)
Table Page
5-5 Target Organ Toxicity of Chloroform ............................ 5-52
7-1 Genetic Effects of Chloroform on Strain D7 of
S. Cerevisiae ................................................ 7-1 o
3-1 Effect of Chloroform on Kidney Epithelial Tumor Incidence
in Osborne-Mendel Rats ....................................... 8-5
8-2 Effect of Chloroform on Thyroid Tumor Incidence in Female
Osborne-Mendel Rats .......................................... 8-7
Toothpaste Formulation for Chloroform Administration ........... 8-9
Effects of Chloroform on Hepatocellular Carcinoma Incidence
in B6C3F1 Mice ............................................... 8-13
8-5 Kidney Tumor Incidence in Male ICI Mice Treated with
Chloroform [[[ 8-17
8-6 Liver and Kidney Necrosis and Hepatomas in Strain A Mice
Following Repeated Oral Administration of Chloroform
in Olive Oil ................................................. 8-19
8-7 SGPT Changes in Beagle Dogs Treated with Chloroform ............ 8-23
8-8 Effect of Oral Chloroform Ingestion on the Growth of Ehrlich
Ascites Tumors ............................................... 8 -29
8-9 Effect of Oral Chloroform Ingestion on Metastatic Tumor Takes
with B1 6 Melanoma ............................................ 8 -30
8-10 The Effect of Oral Chloroform Ingestion on the Growth and
Spread of the Lewis Lung Tumor ............................... 8-31
8-11 Correlation Coefficients Between Residual Mortality
Rates in White Males and THM Levels in Drinking Water
by Region and by Percent of the County Population
Served in the United States .................................. 8 -4^4
8-12 Correlation Coefficients Between Bladder Cancer
Mortality Rates by Sex and BTHM Levels in Drinking
Water by Region of the United States
8-13 Risk of Cancer of the Rectum Mortality Associated
-------
LIST OF TABLES (cont.)
Table Pat:e
8-14
8-15
8-16
8-17
8-18
Cancer Risk Odds Ratios and 95? Confidence Intervals
(Chlorinated Versus Unchlorinated)
Incidence of Hepatocellular Carcinomas in Female and Male
B6C3F1 Mice
Incidence of Tubular-Cell Adenocarcinomas in Male
Incidence of Malignant Kidney Tumors in Male ICI Mice
Upper-Bound Estimates of Cancer Risk of 1 mg/kg/day.
8-54
8-71
8-72
8-72
Calculated by Different Models on the Basis of Different
Data Sets 8-75
8-19 Relative Carcinogenic Potencies Among 53 Chemicals Evaluated
by the Carcinogen Assessment Group as Suspect Human
Carcinogens 8-82
-------
LIST OF FIGURES
1-1 Rate of Rise of Alveolar (Arterial) Concentration Toward
Inspired Concentration For Five Anesthetic Agents of
Differing Ostwald Solubilities 1-10
1-2 Arteriovenous Blood Concentrations of a Patient During
Anesthesia with Chloroform 1-11
1-3 Exponential Decay of Chloroform, Carbon Tetrachloride,
Perchloroethylene and Trichloroethylene in Exhaled
Breath of 18 Year-old Male Accidentally Exposed to
Vapors of These Solvents 1-22
1-1 Relationship Between Total 8-Hour Pulmonary Excretion of
Chloroform Following 0.5-g Oral Dose in Man and the
Deviation of Body Weight From Ideal 1-27
1-5 Blood and Adipose Tissue Concentrations of Chloroform During
and After Anesthesia in a Dog 1-33
1-6 Metabolic Pathways of Chloroform Biotransformation.
(Identified CH Cl_ metabolites are underlines) 1-36
1-7 Metabolic Pathways of Carbon Tetrachloride
Biotransformation 1-12
1-8 Rate of Carbon Monoxide Formation After Addition of Various
Halomethanes to Sodium Dithionite-reduced Liver Microsomal
Preparations From Phenobarbitol-treated Rats 1-16
1-9 Effect of Increasing Dosage of i.p.-Injected
C-Chloroform on Extent of Covalent Binding of
Radioactivity In Vivo to Liver and Kidney Proteins of Male
Mice 6 Hours After Administration 1-51
1-10 Comparison of irreversible binding of radioactivity from
C-CHC1- to protein and lipid of microsomes from
normal rabbit, rat, mouse, and human liver incubated
in vitro at 37°C in 0_ 1-57
fL
8-1 Survival curves for Fisher 311 Rats in a Carcinogenicity
Bioassay on Chloroform 8-5
8-2 Negative Result in Transformation Assay of Chloroform
which was also Negative in the Ames Assay 8-35
8-3 Frequency distribution of CHC1 levels in 68 U.S.
drinking water supplies. The abscissa is linear in the
logarithm of the level 8-13
XI
-------
LIST OF FIGURES (cont.)
Page
8-4 Point and Upper-Bound Estimates of Four Dose-Response Models
Over Low-Dose Region on the Basis of Liver Tumor Data
for Female Mice 8-76
8-5 Point and Upper-Bound Estimates of Four Dose-Response Models
Over Low-Dose Region on the Basis of Liver Tumor Data
for Male Mice 8-77
8-6 Histogram Representing the Frequency Distribution of the
Potency Indices of 53 Suspect Carcinogens Evaluated by
the Carcinogen Assessment Group.... 8-81
XII
-------
The Office of Health and Environmental Assessment (OHEA), U.S. EPA,
is responsible for the preparation of this health assessment document. The
Environmental Criteria and Assessment Office (ECAO/RTP), OHEA, had the
overall responsibility for coordination and the document production effort.
Project Manager
Si Duk Lee, Ph.D.
Environmental Criteria and Assessment Office
U.S. EPA, Research Triangle Park, N.C. 27711
(919) 541-4159
Authors and Reviewers
The principal authors of this document are:
Larry Anderson Ph.D.
Carcinogen Assessment Group
U.S. EPA, Washington, D.C.
David Baylis, M.S.
Carcinogen Assessment Group
U.S., EPA, Washington, D.C.
Chao W. Chen, Ph.D.
Carcinogen Assessment Group
U.S., EPA, Washington, D.C.
Carol Sakai, Ph.D.
Reproductive Effects Assessment Group
U.S., EPA, Washington, D.C.
Sheila Rosenthal, Ph.D.
Reproductive Effects Assessment Group
U.S., EPA, Washington, D.C.
I.W.F. Davidson, Ph.D.
Bowman Gray School of Medicine
Winston Salem, N.C.
D. Anthony Gray, Ph.D.
Syracuse Research Corp.
Syracuse, N.Y.
Sharon B. Wilbur, M.A.
Syracuse Research Corp.
Syracuse, N.Y.
Joan P. Coleman, Ph.D.
Syracuse Research Corp.
Syracuse, N.Y.
xm
-------
The following individuals provided peer-review of this draft or earlier
drafts of this document:
U.S. Environmental Protection Agency
Joseph Padgett
Office of Air Quality Planning and Standards
U.S. EPA
Karen Blanchard
Office of Air Quality Planning and Standards
U.S. EPA
Jerry F. Stara, D.V.M.
Office of Health and Environmental Assessment
Environmental Criteria and Assessment Office
U.S. EPA
Lester D. Grant, Ph.D.
Office of Health and Environmental Assessment
Environmental Criteria and Asssessment Office
U.S. EPA
Participating Members of the Carcinogen Assessment Group
Roy E. Albert, M.D. (Chairman)
Elizabeth L. Anderson, Ph.D.
Larry D. Anderson, Ph.D.
Steven Bayard, Ph.D.
David L. Bayliss, M.S.
Chao W. Chen, Ph.D.
Margaret M. L. Chu, Ph.D.
Bernard H. Haberman, D.V.M., M.S.
Charalingayya B. Hiremath, Ph.D.
Robert E. McGaughy, Ph.D.
Dharm W. Singh, D.V.M., Ph.D.
Todd W. Thorslund, Sc.D.
Participating Members of the Reproductive Effects Assessment Group
Peter E. Voytek, Ph.D. (Director)
John R. Fowle, III, Ph.D.
Carol Sakai, Ph.D.
Ernest Jackson, M.D.
K.S. Lavappa, Ph.D.
Sheila Rosenthal, Ph.D.
Vicki Vaughn-Dellarco, Ph.D.
xiv
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External Peer Reviewers
Dr. Karim Ahmed
Natural Resources Defense Fund
122 E. 42nd Street
New York, N.Y. 10168
Dr. Eula Bingham
Graduate Studies and Research
University of Cincinnati (ML-627)
Cincinnati, Ohio 45221
(513) 475-4532
Dr. James Buss
Chemical Institute of Industrial
Toxicology
Research Triangle Park, N.C. 27709
Dr. I.W.F. Davidson
Wake Forest University
Bowman Gray Medical School
Winston Salem, N.C.
Dr. Larry Fishbein
National Center for Toxicological
Research
Jefferson, Arkansas 72079
(501) 542-4390
Dr. Derek Hodgson
Department of Chemistry
University of North Carolina
Chapel Hill, N.C. 27514
Dr. Marshall Johnson
Thomas Jefferson Medical College
Department of Anatomy
1020 Locust Street
Philadelphia, Pennsylvania 19107
Dr. Trent Lewis
National Institute of Occupational
Safety and Health
26 Columbia Parkway
Cincinnati, Ohio 45226
(513) 684-8394
Dr. Richard Reitz
Dow Chemical, USA
Toxicology Research Laboratory
1803 Building
Midland, Michigan 48640
Dr. Marvin A. Schneiderman
Clement Associates, Incorporated
Arlington, Virginia 22209
(703) 276-7700
Dr. Bernard Schwetz
National Institute of Environmental
Health
Research Triangle Park, NC 27709
(919) 541-7992
Dr. James Selkirk
Oak Ridge National Laboratory
Oakridge, Tennessee 37820
(615) 624-0831
Dr. Samuel Shibko
Food and Drug Administration
Division of Toxicology
200 C Street, S.W.
Washington, D.C. 20204
Telephone:
Dr. Robert Tardiff
1423 Trapline Court
Vienna, Virginia 22180
(703) 276-7700
Dr. Norman M. Trieff
University of Texas Medical Branch
Department of Pathology, UTMB
Galveston, Texas 77550
(409) 761-1895
Dr. Benjamin Van Duuren
Institute of Environmental Medicine
New York University Medical Center
New York, New York 10016
(212) 340-5629
Dr. James Withey
Health and Protection Branch
Department of National Health &
Welfare
Tunney's Pasture
Ottawa, Ontario
CANADA, KIA 01Z
xv
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6. TERATOGENICITY AND REPRODUCTIVE EFFECTS
Schwetz et al. (1974) evaluated the effects of reagent grade chloroform
(lot no. 9649, Burdick and Jackson Laboratories, purity not reported) on the
maternal and fetal well-being of Sprague-Dawley rats. Twenty female rats were
exposed, by inhalation (7 hr/day), to 30, 100, and 300 ppm chloroform on days
6 through 15 of gestation. The authors analyzed the results statistically
using Fisher's Exact Probability test, analysis of variance, Dunnett's test,
or Tukey's test to compare the frequency of anomalies, resorptions, maternal
and fetal weights, body lengths, liver weights, or serum glutamic pyruvic
transaminase (SGPT) activity in the exposed versus the control groups. The
level of significance was chosen at p < 0.05, and the litter was used as the
experimental unit.
When the animals were exposed to the highest dose of chloroform (300 ppm),
there was a significant increase in the number of resorptions and a decrease in
the conception rate (Schwartz et al., 1974). At the lower doses (30 and 100
ppm), no alterations in resorption rate, fetal body weight, conception rate,
number of implantations, or average litter size were observed. Fetal crown-
rump length was significantly decreased at 30 and 300 ppm, but not at the 100
ppm level. At 100 ppm, an increase in the incidence of acaudia (absence of
tail), short tail, imperforate anus, subcutaneous edema, missing ribs, and
delayed ossification of sternebrae were observed. At 300 ppm, subcutaneous
edema and abnormalities of the skull and sternum were observed, but the
incidence of these was not statistically significant. The authors pointed out
that the small numbers of survivors in the 300 ppm group (4+7 versus the
control 10+4 live fetuses/litter) may have prevented adequate statistical
evaluation of this effect.
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In this study (Schwetz et al., 1974), chloroform also produced maternal
toxicity, such as a decrease in the rate of maternal weight gain at all dose
levels and a decrease in food consumption during pregnancy at the 100 and 300
ppm level. Other maternal effects observed were changes in liver weight gain
during pregnancy (no change in absolute liver weight gain at 30 ppm, but an
increase at 100 ppm, possibly due to the concomitant anorexia at this dose).
No significant changes in SGPT activity were observed in groups exposed to
chloroform at the 100 and 300 ppm levels (the only two doses evaluated).
Since the developmental effects observed in the 300 ppm group were associated
with anorexic effects in the mother, a starvation control group was included
in this study. This starvation control group was restricted to food consumption
comparable to the 300 ppm chloroform group. Animals on a starvation diet
(allowed 3.7 g/day of food on days 6 through 15 compared with control animals
whose food consumption average 19-25 g/day on days 6 through 15) had a signifi-
cant decrease in the absolute weight of the liver and an increase in the relative
weight of the liver. The effects of 300 ppm chloroform on the increase in the
relative weight of the liver were much greater than starvation alone. Addition-
ally, and perhaps most importantly, exposure to 300 ppm chloroform resulted in
a dramatic decrease in the number of animals pregnant at sacrifice (15% pregnant
versus 88% in air control), a decrease in the number of live fetuses per litter
(4 versus 10 live fetuses/litter), and an increased percentage of resorptions
(100% vs. 57%). Examination of the uteri indicated that the conceptus had
been completely resorbed very soon after implantation. These effects appeared
to result primarily from chloroform exposure, and not maternal toxic influence,
since anorexia and liver weight changes associated with starvation were not
accompanied by embryotoxic and teratogenic effects.
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Murray et al. (1979) evaluated the effects of chloroform (spectral grade,
Mai 1inckrodt, lot CSZ, code 4434, purity not reported) administered by inhalation
(7 hr/day) to CF-1 mice. Thirty-five animals per group were exposed on gesta-
tion days 1 through 7 and 6 through 15; forty animals per group were exposed
on gestation days 8 through 15. Only one dose, 100 ppm, was tested during
three different time periods (days 1 through 7, 6 through 15, 8 through 15 of
gestation) in CF-1 mice. The varying exposure periods were designed to evaluate
the effects of chloroform in very early pregnancy, organogenesis, and somewhat
later in pregnancy. Special sodium sulfide staining of the uteri was used to
detect very early pregnancies.
The authors (Murray et al., 1979) analyzed the results statistically
using the Fisher's Exact Probability test to evaluate pregnancy incidence; the
modified Willcoxan test for fetal outcomes; the Mann-Whitney signed rank test
for SGPT activity; and one-way analysis of variance for fetal body weights and
body measurements, maternal body weights, liver weights, food consumption, and
number of implantations and resorptions. The level of significance was chosen
at p < 0.05.
Murray et al. (1979) reported that 100 ppm chloroform resulted in a decrease
in the total number of pregnancies when the animals were exposed on days 1-7 or
6-15 of gestation but not on days 8-15 (see Table 6-1 for summary of data). In
the pregnant animals, however, there was no significant effect on the average
number of implantation sites. In animals exposed on days 1-7 of gestation, but
not in those exposed on days 6-15 or 8-15, there were significant increases in
resorptions per litter. This effect was accounted for by the loss of two
entire litters. Mean fetal body weight and crown-rump length were significantly
decreased in the groups exposed on days 1-7 and 8-15, but not in those exposed
on days 6-15. Maternal toxicity (slight decrease in body weight gain during
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TABLE 6-1. SUMMARY OF EFFECTS
(Murray et al., 1979)
Chemical: Chloroform, spectral grade, Lot CSZ, Code 4434, Mallinckrodt, Inc.
Animal: CF-1 mice, 35 animals in groups exposed on days 1-7 and 6-15;
40 animals in groups exposed on days 8-15
Route of exposure: Inhalation, 100 ppm (one dose only)
Duration of exposure: 7 hr/day, days 6-15 of gestation
Pregnant
Days (Implantation sites)
Additional pregnancies
(special stain)
(number of animals)
Total pregnancies
(implantation sites
and special stain)
Exposed Control Exposed
1-7 11/34 (32%) 22/35 (63%) 4
6-15 13/35 (37%) 29/34 (85%) 2
8-15 18/40 (45%) 25/40 (62%) 6
Control Exposed Control
4 15/34 (44%) 26/35 (74%)
2 15/35 (43%) 31/34 (91%)
1 24/40 (60%) 26/40 (65%)
(continued on the following page)
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TABLE 6-1. (continued)
Days
1-7
6-15
8-15
Fetal effects
resorption
fetal body weight
and crown-rump
length
*
delayed skeletal
ossification
fetal body weight
and crown-rump
length
cleft palate
delayed skeletal
ossification
Maternal effects
body weight gain
body weight gain
liver weight
SGPT (only one
dose)
body weight gain
1iver wei ght
*Not significantly different.
6-5
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pregnancy) was seen in groups exposed on days 6-15, with a more severe decrease
in groups exposed on days 1-7 and 8-15. Less food and water1 were consumed in
all experimental groups as compared to controls. Absolute and relative weight
of the liver were increased in groups exposed on days 6-15 and 8-15, hut not
in those exposed on days 1-7. Serum glutamic pyruvic transaminase (SGPT)
activity was increased in mice exposed on days 6-15, which was the only time
period evaluated for this measurement.
A summary of this study (Murray et al., 1979) is presented in Table 6-1.
One major result was that chloroform, under the conditions of this experiment,
caused a decrease in pregnancy. The authors concluded that chloroform affected
the stages either prior to, or in, early implantation. However, because of the
small numbers of animals and the lack of dose-response evaluation (only one
dose was tested, 100 ppm), this conclusion must be considered tentative until
future studies confirm this observation. Other results of this study (Murray
et al., 1979) indicated that the incidence of cleft palates increased in pups
exposed in utero on days 8-15 of gestation, but not on days 1-7 or 6-15. The
authors (Murray et al., 1979) suggested three possibilities to explain this last
result. The first was that earlier exposure on days 6-15 prevented susceptible
concept! from implanting. The second possibility was that the number of litters
available (11 in the group exposed on days 6-15) was insufficient to detect
this effect. The third was that the teratogenic effect (cleft palate) did not
occur in concepti exposed on days 1-7 since they were exposed before organo-
genesis. The number of offspring coming to term was consistently less in all
exposure groups than in the controls (days 1-7, 9 litters versus 22 in control;
days 6-15, 11 litters versus 29 in control; days 8-15, 18 litters versus 24 in
control). Therefore it was not clear whether chloroform produced teratogenicity
separate from embryotoxicity. Since the pups with cleft palate in groups
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were also retarded in growth, it was suggested that the ability of chloroform
to cause malformations was indirect and not direct. However, this was not
experimentally determined, and it is not known if the tendency for lowered
fetal weight and/or delayed skeletal growth is correlated with a higher
incidence of malformations.
In conclusion, this study (Murray et al., 1979) indicated that chloroform
administration (100 ppm) by inhalation (7 hr/day) produced teratogenic and
embryotoxic effects and interfered with pregnancy in addition to causing
maternally toxic effects (changes in liver weight and decreases in weight gain
during pregnancy). Exposure in the early stages of pregnancy appeared to
produce a decreased incidence of conception, but the results of this study did
not conclusively determine which days of pregnancy were most susceptible to the
effects of chloroform. To answer this question, it would be necessary to use
a greater number of doses and larger numbers of animals.
Thompson, Warner, and Robinson (1974) investigated the effect of chloroform
administered orally, using Sprague-Dawley rats and Dutch-Belted rabbits. The
rats were intubated with chloroform (Mai 1inckrodt, Batch ZJL dissolved in corn
oil, purity not reported) twice a day in divided doses of 20 to 501 mg/kg/day.
The rabbits were intubated once a day in doses of 20 to 398 mg/kg/day. Each
study was divided into two parts, a range-finding portion designed to establish
the proper dose range (six rats were administered 79, 126, 300, 316, and 501
mg/kg/day of chloroform; five rabbits were administered 63, 100, 159, 251, and
398 mg/kg/day), and a teratology study, using greater numbers of animals and
three doses (25 rats were administered 20, 50, and 126 mg/kg/day). The rats
were exposed to chloroform on days 6 through 15 of gestation, while the rabbits
were exposed on days 6 through 18 of gestation.
6-7
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Statistical evaluations of maternal body weight gains, food consumption,
implantations, corpus luteum, resorptions, litter size, and fetal weights were
made by an analysis of variance and Dunnett's Two-Tailed Multiple Range test.
Sex ratios and frequency of anomalies among the fetal population and among
litters were analyzed by the chi-square test. In all analyses, the level of
significance chosen was p < 0.05.
The data for the range-finding portion of this study (Thompson et al.,
1974) was not presented; however, the authors reported that rats treated orally
with greater than 126 mg/kg/day chloroform had signs of maternal toxicity, such
as a decrease in food consumption, acute toxic nephrosis, hepatitis, and gastric
erosion. Fetal development was adversely affected in rats receiving 316 and
501 mg/kg/day with a decrease in fetal viability, litter size, and fetal weight
and an increase in the number of resorptions. Only two rats survived when
given 501/mg/kg/day; one was not pregnant, the other had complete early
resorptions. In the rat teratology study, animals receiving 50 and 126 mg/kg/day
displayed signs of maternal toxicity (lowered body weight gain, lowered food
consumption, fatty changes in the liver). No overt toxic effects were observed
in animals given 20 mg/kg/day, and no malformation was noted at any dose level.
In the range-finding portion of the study by Thompson et al. (1974) using
rabbits, severe acute hepatitis and nephrosis were observed in animals given
63 mg/kg/day and higher. No overt signs of toxicity were observed at the
25 mg/kg/day level. In the two surviving dams given 100 mg/kg/day, one had four
resorption sites with no viable concepti, while the other was not pregnant. No
other embryotoxic or teratogenic effect was observed. In the teratology study
of rabbibs, maternal toxicity (depressed weight gain) was observed at the 50
mg/kg/day level. In the fetus, mean body weight was depressed at the 20 and
50 mg/kg/day levels, but no abnormalities were observed. Incomplete skeletal
6-8
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ossification at the 25 or 35 mg/kg/day level was observed amongst fetuses, but
not amongst litters when analyzed statistically.
In this study (Thompson et al., 1974), adverse effects resulting from
chloroform exposure, such as skeletal deformities and deficiencies in pregnancy
maintenance that were observed by Schwetz et al. (1974) and Murray et al. (1979),
were not observed except at doses acutely toxic to the dam. The authors
suggested that the greater maternal toxicity observed in this study compared
to that of others (Schwetz et al., 1974; Murray et al., 1979) could be explained
by the route or duration of exposure. In the study by Thompson et al. (1974),
rats and rabbits were exposed orally to chloroform or twice a day, while Schwetz
et al. (1974) and Murray et al. (1979) administered chloroform by inhalation
7 hr/day. The lack of information on the pharmacokinetic interaction of chloro-
form, however, prevents an evaluation of the role of exposure in producing
adverse reproductive outcome.
Burkhalter and Ralster (1979) evaluated the potential of chloroform to
adversely affect behavior in developing ICR mice. This study was designed as a
preliminary screening study, with the parental generation of male and female
mice exposed 21 days prior to mating, during mating, and for an additional 21
days. The offspring were exposed starting on day seven until day 21 after
birth. Only one oral dose of chloroform, 31.1 mg/kg/day, was administered to
five control and five experimental animals. Each litter was reduced to eight
pups, and three pups were randomly selected for behavorial teratogenic testing.
The chloroform (Mai 1inckrodt, nanograde purity) was administered by gavage and
delivered in a solution of 1 part polyoxyethylated vegetable oil, Emulphor
(EL-620, GAP Corp., New York), and 8 parts saline. A variety of behavorial
responses were evaluated which included: righting reflex, forelimb placing
response, forepaw grasp, rooting reflex, cliff drop aversion, auditory startled
6-9
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response, bar-holding ability, eye opening, motor performance, and learning
ability. The scoring for these responses was based upon predesignated criteria.
These criteria established behavioral ability by measuring both objective
standards (time to complete test) and subjective standards ("weak" or "complete"
grasp of paws).
Analysis of variance was used to statistically evaluate tests of passive
avoidance. Screen test latencies were analyzed by t-test. The data from the
neurobehavioral developmental scale were analyzed using a Mann-Whitney U test.
The level of statistical significance was chosen at p < 0.05.
Burkhalter and Ralster (1979) reported that the sizes of litters were
similar for both the control and experimental groups; however, fetal body weight
gain of pups during the 14 days of exposure (days 7-21 following birth) was
decreased. Forelimb placement response was reduced in the exposed group on
day 5 and 7 of birth, but not on day 9. The significance of this reduction is
not known, although the recovery on day 9 suggested that the effect was
reversible. The other behavorial responses were not significantly different
in the exposed groups. Burkhalter and Balster (1979) concluded that 31.1
mg/kg/day of chloroform produced no significant adverse behavorial effects in
pups exposed both in utero and after birth (days 7-21). However, since this
study was designed as a preliminary screening study, using one dose and small
numbers of animals, there was no attempt to evaluate the full range of dose-
related effects. In future studies, it would be desirable to evaluate at
least three doses, including doses high enough to produce some maternal toxicity,
in addition to using a larger sample size.
In an abstract, Dilley et al. (1977) reported the effects of exposure to
chloroform on pregnant rats (strain not reported, number of animals not reported),
The animals were exposed by inhalation (20 +_ 1.2 gm3/day) on days 7-14 of
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gestation. Two lower concentrations were administered in the study, but the
doses were not reported in the abstract. Dilly reported that chloroform
increased fetal mortality and decreased fetal weight gain; however, there were
no malformations.
In another abstract, Ruddick et al. (1980) exposed 15 Sprague-Dawley rats
to 100, 200, and 400 mg/kg/day of chloroform by gavage on days 6-15 of gestation.
Chloroform (purity not reported) was reported to cause maternal toxicity
(changes in weight gain, biochemical and hematological parameters, and liver or
kidney changes). Chloroform also produced adverse effects in fetal development
(type not specified), but the authors attributed these effects primarily to
maternal toxicity and not directly to chloroform exposure. Without the
presentation of this data, however it is not possible to fully evaluate
these results.
6.1 SUMMARY
In summary, the results of four articles and two abstracts indicated that
under the conditions of the experiments, chloroform has the potential for
causing adverse effects in pregnancy maintenance, delays in fetal development
and production of terata in laboratory animals. The adverse effects on the
conceptus were observed in association with maternal toxicity. The type and
severity of effects appeared to be specific to the conceptus, affecting the
fetus to a much greater degree than the mother. Therefore, it was concluded
that chloroform has the potential for causing embryotoxicity and teratogenicity
(Schwetz et al., 1974; Murray et al., 1979). The results of other studies
indicated that chloroform has no significant effect on neonatal behavior
(Burkhalter and Balster, 1979) and does not cause adverse fetal effects except
at maternally toxic levels (Thompson et al., 1974). The two abstracts did not
contain enough detail for critical scientific review (Oil ley et al., 1977;
6-11
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Ruddick et al., 1980).
Studies administering chloroform by inhalation for 7 hr/day (Schwetz et
al., 1974; Murray et al., 1974) reported more severe outcomes than other studies
which administered chloroform by intubation once or twice a day (Thompson et
al., 1974; Burkhalter and Balster, 1979). However, since the pharmocokinetic
relationship associated with route or duration of exposure have not been studied,
it is not possible to evaluate the importance of the route of exposure in
causing adverse reproductive outcome. To evaluate more fully the influence of
these factors, additional investigations would have to be conducted.
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6.2 REFERENCES
Burkhalter, J., and R.L. Balster. 1979. Behavioral teratology evaluation of
chloroform in mice. Neurobehavioral Toxicol. 1:199-205.
Dilley, J.V., N. Chernoff, D. Kay, N. Winslow, and E.W. Newell. 1977. Inhalation
teratology studies of five chemicals in rats. Toxicol. Appl. Pharmacol.
41:196.
Murray, F.A., R.A. Schwetz, J.G. McBride, and R.E. Staples. 1979. Toxicity of
inhaled chloroform in pregnant mice and their offspring. Toxicol. Appl.
Pharmacol. 50:151-522.
Ruddick, J.A., D.C. Villenouve, I. Chu, and V.E. Balli. 1980. Teratogenicity
assessment of four halomethanes. Teratology 21.-66A.
Schwetz, B.A., B.K.J. Leong, and P.J. Gehring. 1974. Embryo- and fetotoxicity
of inhaled chloroform in rats. Toxicol. Appl. Pharmacol. 28:442-451.
Thompson, D.O., S.O. Warner, and V.B. Robinson. 1974. Teratology studies on orally
administered chloroform in the rat and rabbit. Toxicol. Appl. Pharmacol. 29:
348-357.
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7. MUTAGENICITY
7.1 TNTROniJCTION
The nutagenic potential of chloroform (CHC^) has been assessed by
evaluation of the results from five jn yjtro bacterial studies, one
host-mediated assay using Sajnonella as the indicator organism, one yeast
study, one Drosophila sex-linked recessive lethal test, one in vitro mammalian
cell mutagenicity assay, two sperm head abnormality tests, three chromosome
aberration studies, and six DNA damage studies (sister chromatid exchange and
unscheduled DNA synthesis). These reports are discussed below. Also, several
assays from a recently published screening study, in which 42 chemicals were
tested in various short-term protocols, are briefly discussed. The majority
of the above studies were negative. Information relating to the binding of
metabolically activated CHC13 to cellular macromolecules is presented before
the sections assessing the genetic damage caused by CHC13. This was done to
set the stage for the discussion of the negative mutagenicity studies
described below and to support the suggestion that CHC13 may be a weak
mutagen. In addition, suggestions for further testing are presented.
7.?. CnVALENT RINDING TO MACROMOLECULES
As mentioned in Chapter 4, the primary reactive metabolite of CHC13 is
phosgene, COCl^. Phosgene is a crosslinking agent and may covalently bind
to and crosslink macromolecules. Thus, the toxicity and carcinogenicity of
CHC13 nay be related to its metabolism to phosgene. The DNA binding
potential and carcinogenicity of phosgene are currently under investigation in
Or. R.L. Van Duuren's laboratory at New York University Medical Center.
Preliminary evidence indicates that phosgene binds to DNA (Dr. Sipra Ranerjee,
New York University Medical Center, personal communication). The binding
potential of metabolically activated CHC13 has been assessed in several
7-1
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studies. Some of these studies have been described in Chapter 4, section 4.6.
Additional studies and studies not adequately covered in Chapter 4 for the
purposes of this chapter on mutagenicity are described below.
Diaz Gomez and Castro (1980a) assessed the CHC13 activation potential
of purified rat liver nuclei by measuring covalent binding of nuclear-
activated CHC13 to nuclear protein and lipid. The results were compared to
results obtained from similar incubation mixtures containing microsomes
instead of purified nuclei. The incubation mixtures containing either nuclei
(1.3 mg protein/ml) or microsomes (1.56 mg protein/ml) were incubated for 30
min in 6.4 nM 14CHC13 (5.4 Ci/mol) and an NADPH generating system. The
authors observed that the extent of binding to protein in the nuclear
preparation was approximately 40% of that observed for microsomes (nuclei, 27
pmol/mg; microsomes, 68 pmol/mg). Rinding to nuclear lipid was approximately
35% of that observed for microsomes (nuclei, 20 pmol/mg; microsomes, 57
pmol/mg). Thus, isolated nuclei were less efficient than microsomes in
metabolizing CHC13, but the results were within the same order of magnitude.
This study suggests that metabolism of CHC13 to a reactive
intermediate(s) can occur in nuclear membranes, as may be the case with other
xenobiotics (Weisburger and Williams, 1982). Tt should be mentioned, however,
that the nuclear preparations were contaminated with trace amounts of
endoplasmic reticulum, which may have been sufficient to result in at least
part of the nuclear activation observed.
In a subsequent study, niaz Gomez and Castro (1980b) exposed rat or mouse
liver HNA or RNA to l^CHCl 3 in_ vjj/o or JJT_ vitro without finding any
significant binding of 14C to the nucleic acids. However, the specific
activity of the l^CHC^ was only 5.4 Ci/mol, which may have been too low
to allow for observation of binding to DNA (Brookes and Lawley, 1971),
7-2
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especially with a background count of 160 dpm.
Reitz et_ aj_. (1980) published results of DNA alkylation studies in livers
and kidneys of male mice that were exposed orally to CHC13 at 240 mg/kg. A
very small amount (1-3 x 10~6 alkylations/deoxynucleotide) of alkylation was
observed. However, these results cannot be interpreted because the specific
activity of the l^CHC^ used was not specified and because the
experimental procedures used were not described.
In summary, binding of metabolically activated CHC13 to liver
microsomal and nuclear protein and lipid has been observed. The only studies
that attempted to measure binding to nucleic acids were inconclusive.
7.3 MUTAGENICITY STUDIES IN BACTERIAL TEST SYSTEMS
Uehleke et_ a_L (1977) tested CHd3 for mutagenicity in suspension assays
with _S. typhimurium strains TA1535 and TA1538. No nutagenic activity was
observed. About 6-9 x 10^ bacteria were incubated for 60 min under N£ in
tightly closed test tubes with 5 mM CHC13 and microsomes (5 mg protein) plus
an NADPH generating system. The mutation frequencies (his+ colony forming
units/10^ his" colony forming units) were less than 10 for both strains
and the spontaneous mutation frequencies were 3.9 +_ 3.7 for strain TA1535 and
4.4^ 3.5 for strain TA1538. At this concentration of CHC13, survival of
the bacteria was at least 90%. Additional higher concentrations should have
been tested, because the mutagenic range can occur at higher toxicities.
Dimethylnitrosamine (50mM), cyclophosphamide (0.5mM), 3-methylcholanthrene
(0.1 mM), and benzo[ajpyrene (0.1 mM) were the positive controls used in this
study. Information was not provided on the survival of the bacteria at these
concentrations of the positive control chemicals. Although these chemicals
were mutagenic in the presence of the S9 activation system, they may not be
appropriate as controls for CHC13 because they are not halogenated alkanes
7-3
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and therefore are not metabolized like them.
Studies demonstrating that rnetabolically activated CHC13 hinds to
protein and lipid in the presence of rabbit nicrosomes were also described in
the paper by Uehleke et_ aj_. (1977) and are mentioned in Chapter 4 of this
document. However, it is not clear from the description provided in this
report by Uehleke et_ aj_. (1977) whether rat, mouse, or rabbit microsomes were
used in the mutagenicity studies. If mouse or rat microsomes rather than
rabbit microsomes were used for the mutagenicity experiments, it cannot be
assumed that CHC13 was sufficiently activated, since activation sufficient
for binding of l^CHClg ^0 macronolecules was shown in this paper only with
rabbit microsomes. Another deficiency in this study is that the Ames strains
TA98 and TA10D were not used. These strains contain an R factor plasmid that
increases the sensitivity of the tester strains to certain mutagens.
The mutagenicity of CHC13 was also tested in a study designed to
evaluate the mutagenic potential of chemicals identified in drinking water
(Simmon et_ a_K , 1977). No mutagenic activity was detected with CHC13. The
authors tested 71 of the 300 chemicals that had been identified in public
water supplies. CHC13 was tested at 10% by volume (1.24 M) in a suspension
assay with Salmonella strains TA1S3S, TA1537, TA1538, TA9R, and TA100 and S9
mix prepared from Aroclor 1254-treated rats. This concentration of CHC13
exceeds its solubility. Mutagenic activity was not observed. However,
information on toxicity was not provided.
CHC13 was also tested in this study in a desiccator to assess
mutagenicity due to vapor exposure (Simmon et_ aj_., 1977). Agar plates were
placed uncovered in a desiccator above a glass petri dish containing the
CHC13. The desiccator contained a magnetic stirrer which acted as a fan to
aid in evaporation of the measured amount of CHC13 and to maintain an even
7-4
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distribution of the vapors. Plates were exposed to the vapors for 7-10 hr and
then removed from the desiccators, covered, and incubated approximately 40 hr
before scoring. As in the suspension assay described above, mutagenic
activity was not observed and no information on toxicity was provided.
This study by Simmon et_ aj_. (1977), although lacking some specific
details of the CHC13 assay, clearly identifies other trihalomethanes
(CHBT3, CHBT2C1, CHBrCl?) as mutagens in the vapor assay in desiccators.
Methyl bromide, methyl chloride, methyl iodide, and methylene chloride were
also found to be mutagenic in the desiccator assay. However, these seven
halogenated compounds did not require metabolic activation to exhibit
mutagenic activity. It may be that CHC13 itself is not mutagenic and the
rat liver S9 was not sufficient to metabolize CHC13 to a potential mutagenic
reactive intermediate (? phosgene), even though the demonstration of
mutagenicity of three of the chemicals tested [bis(2-chloroisopropyl)ether,
vinyl chloride, and vinylidene chloride] required or was enchanced by this S9
mix. Because CHC13 is a liver carcinogen in the mouse and not in the rat
(NCI bioassay, 1976), mouse liver microsomes may be more appropriate than rat
liver microsomes as a component of an activation system for CHC13
mutagenesis. It may also be that a reactive intermediate was formed, but it
was too reactive or short-lived to be detected in a test system that uses
exogenous metabolic activation.
Kirkland et_ aj[. (1981) studied the mutagenicity of CHC13 in Escherichia
coli strains WP2p and WP2uvrA"p, using reversion to tryptophan prototrophy
as the endpoint. The bacteria were treated with CHC13 in plate
incorporation and preincubation tests both with and without rat liver
microsomes (plus cofactors) prepared from Aroclor 1254-induced rats. The
concentration of protein in the microsomal suspension was not given. CHC13
7-5
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was added at 10,000, 1,000, 100, 10, 1, or 0.1 ug/plate. Negative results
were obtained in both tests. However, there was no indication that
volatilization of CHC13 was prevented in the preincubation test. Also, it
appears from the description of the protocol for the plate incorporation test
that the procedure used to prevent loss of CHC13 was inadequate. CHC13
was added to suspensions of the bacteria in molten agar, and each mixture was
rapidly mixed on a Whirlimixer and poured onto agar plates. The plates were
then incubated in gas-tight containers. Excessive evaporation of CHC13 may
have occurred during the mixing of the molten agar/bacteria/CHCl3
suspension. 2-Aminoanthracene was used as a positive control requiring
netabolic activation and N-methyl-N1-nitro-N-nitrosoguanidine was used as the
positive control not requiring activation. These chemicals are not volatile
and are therefore inadequate positive controls for CHC13. Also, it cannot
be assumed that a microsomal activation system that metabolizes
2-aminoanthracene is sufficient to metabolize CHC13.
Gocke eit_ a_U (1981) assessed the mutagenicity of 31 chemicals (including
CHC13) used as ingredients in European cosmetics. Three test systems were
used: the Salmonel1 a/mi crosome test, the sex-linked recessive lethal test in
Orosophija, and the micronucleus test for chromosome aberrations in mice. The
latter two tests will be discussed in the following sections.
For the Sajmonel1 a/microsome assays at least five doses of each compound
were tested, usually up to 3.6 mg/plate for nontoxic and soluble compounds.
Salmonella strains TA1535, TA100, TA1538, TA98, and TA1537 were used with and
without activation by S9 mix prepared from Aroclor-pretreated rats. Because
this was a screening study in which 31 chemicals were tested, details of the
assay protocols were not given. Three halogenated aliphatic hydrocarbons were
tested (1,1,1-trichloroethane, dichloromethane, and CHC^). Because of the
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volatility of these compounds, the bacteria were exposed in airtight
desiccators for 8 hr. The composition and purity of these chemicals were not
specified.
The first two substances exhibited mutagenic activity with and without
metabolic activation. CHC13 was inactive. However, as discussed above in
the evaluation of the Simmon et^ aj_. (1977) paper, CHC13 may require
metabolic activation to be mutagenic, and the rat liver S9 preparation may not
have been sufficient. Also, if a reactive metabolite was formed, it may have
been too reactive to be detected under conditions of exogenous activation.
Phosgene, the primary reactive metabolite of CHC13, is very reactive and
unstable (Kirk-Othmer, 1971).
In a recently published screening study of 42 chemicals (de Serres and
Ashby, 1981), CHC13 was evaluated in 38 jn vivo and jn vitro short-term
tests designed to assess potential genotoxicity. Results from bacterial
assays carried out in 18 laboratories using Salmonejla (Ames reversion test)
or 1E_. coli (forward mutation test) were essentially negative. However, these
results are inconclusive, because nowhere was it mentioned in the protocols
that excessive volatilization and escape of CHC13 from the culture dishes
was prevented. In addition, the problems with external activation systems
mentioned above also apply to the bacterial assays carried out in this
screening study.
Agustin and Lim-Sylianco (1978) investigated the mutagenicity of CHC13
in a host-mediated assay using Salmonella strains TA1535 and TA1537 as the
indicator organisms that were injected into male and female mice. The authors
found that male mice metabolized the CHC13 to a mutagen active in strain
TA1537. However, they reported only the ratios of mutation frequencies for
treated vs. control animals and gave no indication of the actual colony counts
7-7
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observed. The mutation frequency (tested/control) for strain TA1537 in male
mice was 36.75 and that for female mice, 2.30. The mutation frequencies for
strain TA1535 were 0.61 and 0.12, respectively. Details of the procedures
used (i.e., doses of CHC13, numbers of bacteria injected and recovered,
route of exposure, and time of exposure before the animals were killed) were
not presented. Thus, although there is suggestive evidence of a positive
response, definitive conclusions cannot be reached because of inadequacies in
the way the data were reported and because details of the procedures used were
not provided.
Agustin and Lim-Sylianco (1978) also studied the mutagenicity of
ether-extracted urine concentrates from 10 male mice in a bacterial spot test,
using strain TA1537. The mice were exposed to CHC13 at 700 mg/kg. Urine
concentrates from CHCl3-treated mice yielded 302 revertant colonies and a
zone of inhibition of 29 mm, whereas urine from control animals yielded 10
revertant colonies with no zone of inhibition. Details of the ether
extraction procedure were not provided, but the likelihood of a false positive
result due to the presence of histidine in the extracted urine is unlikely,
because the urine concentrate from the control animals yielded only 10
colonies and was presumably subjected to the same extraction procedure.
In summary, the results from the above bacterial studies are inconclusive,
because false negative results could have been obtained due to a number of
factors, including:
1. The activation systems used may have been inadequate for metabolism of
CHC13.
2. Phosgene, the primary reactive metabolite of CHC13, is unstable and
highly reactive. Because exogenous activation systems were used in
many of these studies, any phosgene generated (assuming an adequate
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activation system) may have been scavenged by microsomal protein or
lipid before reaching the DNA.
3. Adequate exposure to CHC13 may not have occurred if appropriate
precautions were not taken to prevent the evaporation of CHC13.
It is, of course, also conceivable that the negative results reflect the
possibility that CHC13 is not a mutagen.
The positive result in the host-mediated study utilizing jn_ vivo
metabolism of CHC13 could not be evaluated because the details of the
procedures used and the appropriate data were not reported. This result may
be suggestive of a positive response and indicates the need for additional
testing. The results from the urine spot test are suggestive of a positive
response.
7.4 MUTAGENICITY STUDIES IN EUCARYOTIC TEST SYSTEMS
Callen et_ a_K (1980) carried out a study on the mutagenicity of CHC13 in
the D7 strain of Saccharomyces cerevjsi ae, which contains an endogenous
cytochrome P-450 dependent monooxygenase metabolic activation system. By
using this strain of yeast, Callen and his coworkers eliminated the need for
the exogenous type of metabolic activation system used in the above bacterial
studies. Three different genetic endpoints can be examined with this system:
gene conversion at the trp5 locus, mitotic crossing over at the adej? locus,
and gene reversion at the i1v1 locus. The effect of CHC13 on these
endpoints was measured by exposing cells in suspension to 2.5, 5.0, and 6.3 g
of CHC13 per liter of buffer (21 mM, 41 mM, and 54 mM, respectively). The
purity of the CHC13 sample (from J.T. Baker) was not provided. Escape of
volatilized CHC13 is not expected to have occurred to any significant
extent, because the incubations were carried out in screw-capped glass tubes.
Results of the Callen et al. study are presented in Table 7-1. A 1-hr
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TABLE 7-1. GENETIC EFFECTS OF CHLOROFORM ON STRAIN D7 OF S. CEREVISIAE
—I
I
o
Concentration, mM
Survi val
Total colonies
% of control
trp5 locus (gene conversion)
Total convertants
Convertants/105 survivors
ade2 locus (nitotic crossing over)
Total twin spots
Mitotic cross-overs/10^ survivors
Total genetically altered colonies
Total genetically altered colonies/
lf)3 survivors
ilvl locus (gene reversion)
Total revertants
Revertants/10" surviviors
0
1423
100
246
1.7
1
1.6
6
1.0
61
4.3
21
1302
91
274
2.1
1
1.7
11
1.9
46
3.5
41
982
69
450
4.6
2
4.1
43
8.9
81
8.2
54
84
6
278
33.1
4
44.8
47
52.7
50
60.0
-------
treatnent of cells with 54 mM CHC13 resulted in an increased convertant to
survivor ratio with a marginal increase in the observed number of convertant
colonies. Similar results were obtained for mitotic crossing over and gene
reversion. Toxicity at this concentration was high (6% survival).
At the lower concentrations of CHC13 (21 mM and 41 mM), a small
dose-related increase (1.2-fold and 2.7-fold, respectively) in gene
convertants was observed. In addition, a 9-fold increase in the frequency of
genetically altered colonies, which are due to gene conversion and mitotic
crossing over, was observed at 41 mM CHC13. For gene reversion, a 2-fold
increase was observed. Toxicity was low at these levels. These results are
suggestive of a positive response, but additional studies are needed before it
can be conclusively stated that CC14 causes genetic effects in yeast.
Sturrock (1977) tested the mutagenic effects of CHC13 at the
8-azaguanine locus in Chinese hamster lung fibroblast cells (V-79 cells) in
culture. The cells were grown to a monolayer and exposed for 24 hr to an
atmosphere containing 1 to 2.5% CHC13. Cells were then plated onto media
with or without 8-azguanine. After incubation, the plates were examined for
mutations and survival. No significant increase in the frequency of mutants
was observed in treated cultures as compared with untreated controls.
However, the xenobiotic biotransformation capability of the cells used in this
study is unknown.
fiocke et_ a_K (1981) evaluated the mutagenicity of CHC13 by carrying out
tests in ProsopMl a to detect sex-linked recessive lethal mutations. The
flies were exposed by the adult feeding method to 25 mM CHd3. Three
successive broods (3-3-4 days) of flies were examined for sex-linked recessive
lethal mutations. Over 40DO chromosomes per brood were tested. In two of the
broods, small increases in mutations were observed. Results (sex-linked
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recessive lethal s/chromosomes) were as follows: Brood 1, 20/4616 (0.430/,);
Brood ?., 13/4349 (0.29%). Controls were 0.27% and 0.14%, respectively. These
increases were not significant. To obtain significant increases more
chromosomes would have to he tested.
Because sperm head abnormalities are thought by Hyrobek and Rruce (1978)
to arise from mutations in the genes that code for spermatogenesis, it is
possible that assays for sperm head abnormalities may be used as an indication
of the mutagenic potential of chemicals. In a screening study in which 54
chemicals were tested for induction of sperm head abnormalities, Topham (1980)
reported that CHC13 did not induce sperm head abnormalities when injected in
mice. Groups of five male mice received five daily intraperitoneal injections
of corn oil alone (5 ml/kg/day) or CHC13 in corn oil at 0.025, 0.05, 0.075,
0.1, and 0.25 mg/kg/day. Topam (1980) reported that the highest of these
doses (0.25 mg/kg/day) was lethal. However, in Chapter 5 of this document, it
is stated that the LO^n, of CHC13 for ICR male mice is 789-1590 mg/kq. The
reason for the discrepancy in reported topic response to CHC13 is not clear.
Five weeks after the last dose, caudal sperm smears were examined for head
abnormalities. The raw data for the experiments with CHC13 were not
presented in the paper. Topham stated that CHC13 induced sporadic small
increases in abnormal sperm heads at a high dose, but this result could not be
repeated.
Another sperm head abnormality study was carried out by Land et al.
(1981). This study was designed to determine whether certain anesthetics
affect mouse sperm morphology. Groups of five male mice (11 weeks old) were
exposed by inhalation to CHC13 at 0.04 and 0.08% (vol/vol) for 4 hr/day for
5 days in glass exposure chambers. Control mice (N = 15) were exposed to
compressed air under similar conditions. Twenty-eight days after the first
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exposure, the nine survivors from each exposure level were killed and the
caudal spern were examined for head abnormalities. The results were reported
as % abnormal sperm (+ SEM) and were as follows: control, 1.42 (n.08); 0.08%
CHC13, 3.48 (0.66); and 0.04% CHC13, 2.76 (0.31). The authors concluded
that exposure of mice to CHC13 resulted in a significant increase in sperm
head abnormalities compared to the control (P < 0.01). However, significance
was calculated by the t-test and by the F test. Use of these tests may not be
appropriate in this case because of the non-homogeneity of the variance in
CHCl3-treated and control groups (Dr. Chao Chen, Carcinogen Assessment
Group, U.S. EPA, personal communication). This study is suggestive of a
positive response, but a more appropriate statistical analysis of the data is
needed. However, the data necessary to carry out such a statistical analysis
were not provided.
Testing of the mutagenic potential of CHC13 in eucaryotic systems was
carried out in the same screening study edited by de Serres and Ashby (1981)
that was discussed in the previous section of this chapter for bacterial
assays. Seven yeast assays, two jn vitro mammalian ONA damage assays
(unscheduled DNA synthesis and sister chromatid exchange), and three
whole-animal tests (Drosophila sex-linked recessive lethal, mouse bone marrow
micronucleus, and mouse sperm abnormality) were described for CHC^. The
ONA damage studies will be discussed in the next section of this chapter.
The seven yeast assays involved both forward and reverse point nutations,
mitotic crossing over, mitotic gene conversion, and induction of aneuploidy in
mitotic cells. The latter three assays test for ONA damage. A positive
result was obtained only in the forward mutation assay utilizing
S c h i20s a c c haromycej potnbe as the test organism. In the reverse mutation
assay, CHC13 was tested only with stationary cells in the presence of rat S9
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mix. Exposure was for 24 hr. Growing cells are none sensitive to the
mutagenic effects of several chemicals than are stationary cells, possibly
because log-phase yeast cells contain an endogenous cytochrome P-450
metabolizing system (Callen et_ aj_., 1980). The ONA damage studies in yeast
yielded negative results. The negative results in the mitotic qene conversion
assay, which was carried out in strain 07, are in conflict with the weakly
positive results reported for CHC13 by Callen et_ a_L (1980) as described
above.
The three whole-animal tests on CHC13 (Drosophila sex-linked recessive
lethal, mouse bone marrow micronucl eus, and mouse sperm abnormality) described
in the de Serres and Ashby (1981) report yielded negative results. However,
de Serres and Ashby, in their overview of the performance of the assay systems
used in this study, state that the whole-animal tests had low sensitivity and
that a negative result has "very little significance." It was recommended
that CHC13 be tested further in irj \nvo short-term tests.
In summary, the results from the eucaryotic test systems suggest that
CHC13 may be a weak mutagen. Results indicative of a positive response were
obtained only in studies using test organisms possessing endogenous activation
systems (i.e., yeast and mice). More studies, particularly with organisms
possessing endogenous activation, are needed before a definite conclusion on
the mutagenicity of CHC13 can be reached. Suggested studies are described
later in this chapter.
7.5 OTHER STUDIES INDICATIVE OF ONA DAMAGE
Two types of DNA damage studies, sister chromatid exchange (SCE) and
unscheduled DNA damage (HDS), are described in this section.
Sister chromatid exchange is thought to involve DNA breakage. For this
reason, assays for SCE have been used as an indicator of primary DNA damage.
7-14
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White et_ aK (1979) studied the induction of SCEs by CHC13 and other
anesthetics. Information about the purity of these compounds was not given.
Exponentially growing Chinese hamster ovary cells were exposed to the gases JJT_
vitro with and without S9 mix (10% by volume) prepared from Aroclor-induced
rat livers. Exposure was at 0.71% (vol/vol) CHC13 (88 mM) for 1 hr in
closed screw-capped culture flasks. The cells were then incubated for 24 hr
in medium containing 10 uM 5-bromo-2-deoxyuridine. SCEs per chromosome were
0.544 _+ 0.018 for CHCl3-exposed cells and 0.536 _+ 0.018 for controls. The
short exposure time and low concentration raise serious questions concerning
the conduct of this assay. Also, even though the rat liver S9 was sufficient
for activation of vinyl-containing compounds to derivatives (presumably
epoxides) that induced SCE, it may not have been adequate for activation of
CHC13 and the other haloalkanes that tested negative.
Another SCE study was carried out by Kirkland et_ aj_. (1981) using human
lymphocytes. The cells were treated with CHC13 at 25, 50, 75, 100, 200, and
400 ug/ml (0.2, 0.4, 0.6, 0.8, 1.6, and 3.3 mM, respectively) for 2 hr in the
presence of S9 mix from Aroclor-induced rats. Metaphase spreads from
approximately 100 cells per treatment were examined. Acetone was the negative
control, and a positive control was not included in the assay because the same
donor's lymphocytes had previously shown a dose-related increase in SCE after
treatment with benzo[a]pyrene in the presence or absence of S9 mix. A small
increase in SCE occurred at 50 ug of CHC13 per ml, but no dose-response
trend was observed.
There are several problems with this study. First, since a positive
dose-related increase in SCE after treatment of lymphocytes with
benzo[a]pyrene was not dependent on S9 mix, this positive control is
inadequate for substances that are likely to require metabolic activation. In
7-15
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addition, this control was not concurrent and is therefore not an appropriate
positive control. Second, there is no indication that volatilization and
escape of CHC13 was prevented. Third, the maximal dose was only 3.3 mM.
Fourth, infornation about the toxicity of CHC13 for the lymphocytes was not
provided.
Two jrj vitj-Q mammalian DNA damage studies on CHC13 (tins and SCE) were
described in the volume edited by de Serres and Ashby (1981). The SCE assay
(Chapter 51) utilized an exogenous activation system and yielded negative
results. The Chinese hanster ovary cells were exposed to CHOI 3 in the
presence of rat liver S9 mix for only 1 hr. This length of time may be
insufficient, particularly since a positive response for 2-acetylaninofluorene
was obtained in the presence of S9 after 2 hr of exposure and not after a 1-hr
exposure. Also, as mentioned above, exogenous activation may be inappropriate
when testing CHC13 for genotoxic activity. Thus, the negative results
obtained in this study are inconclusive. In addition, as in the bacterial
tests described above, there was no indication that precautions were taken to
prevent evaporation and loss of CHC13 from the culture flasks.
Unscheduled DNA synthesis (UDS), measured as repair of chemically induced
DNA damage, is an additional method of testing for genetic damage. Mirsalis
et_ aj_. (198?) measured IJDS in primary rat hepatocyte cultures following rn_
vivo treatment of adult male Fischer-344 rats (175-275 g) with CHC13 at 40
and 400 mg/kg by gavage. Control rats received corn oil (the vehicle for
CHC^) by gavage. Several additional chemicals were also tested in this
study. At 2 or 12 hr after treatment, the livers were perfused in sjtu and
hepatocytes were isolated. Approximate"!/ 6 x 105 viable cells were seeded
in 35-mm culture dishes containing coverslips and allowed to attach to the
coverslips for about 90 min. After the coverslip cultures were washed, they
7-16
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were incubated in a medium containing in uCi [^H]thymidine (40-50 Ci/mmol)
per ml for 4 hr. The cultures were washed again and incubated in medium
containing 0.25 mM cold thymidine for 14-16 hr. The extent of LIDS was
assessed by autoradiography. Net grains/nucleus were calculated as the silver
grains over the nucleus minus the highest grain count of three adjacent
nuclear-sized areas over the cytoplasm.
Cells from negative control animals (given vehicle only) ranged from -3.0
to -5.1 net grains/nucleus. Several chemicals tested positive in this assay
(<_ 5 net grains/nucleus was considered positive), including methyl
methanesulfonate, dimethylnitrosamine, 2-acetylaminofluorene, benzidine, and
others. CHC13 at 40 and 400 mg/kg yielded a negative response (-2.7 to -4.4
net grains/nucleus). However, rats are not susceptible to CHCl3-induced
hepatocarcinogenesis (NCI bioassay, 1976). The negative response observed in
this study is consistent with this fact. Renzo[a]pyrene and
7,12-dimethylbenzCaJanthracene, carcinogens that, like CHC13, are not rat
liver carcinogens, also tested negative in this assay. These chemicals tested
positive in the jn_ vitro rat hepatocyte (IDS assay (Williams e>t_ a]_., 1981).
This discrepancy suggests that the jrj^ vitro test may be more sensitive than
the in vivo assay. CHC13 has not been tested in the jn vitro rat hepatocyte
Uns assay. The mouse is susceptible to CHCl3-induced liver tumors (NCI
bioassay, 1976) and therefore may be a more appropriate test animal for the j_n_
vjvQ DOS assay.
Also, it is uncertain whether the method of assaying for UHS used in this
study (subtraction of cytoplasmic grain counts from nuclear grain counts) will
allow for detection of a weak response. In a recent article discussing the
validity of the autoradiographic procedure for detecting UPS in rat
hepatocytes, Lonati-Gal ligani et_ aU (1983) describe some potential problems
7-17
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with this method. First, they found that it is difficult to obtain hepatocyte
preparations of reproducible quality. Preparations can differ in their
metabolic capabilities. In order to avoid false negative results with
potential weak UDS inducers due to poor hepatocyte preparations, they suggest
that test chemicals should be studied in conjunction with a potent
UnS-inducing analog and that negative results should be accepted only in tests
in which the analog is strongly positive. No known positive analog of CHC13
was tested in the study of Mirsalis et_ aj_. (1982). Second, the cytoplasmic
layer covering the nucleus is thinner than the cytoplasmic area next to the
nucleus. Therefore, a variable overcorrection is probably applied, as
witnessed by the usually higher cytoplasmic than nuclear counts observed in
control cells (e.g., see above results for cells from control animals). This
effect would tend to obscure a weakly positive UDS response. Lonati-Galligani
et_ £l_. (1983) suggest that an alternative endpoint be determined. Instead of
subtracting cytoplasmic grain counts from nuclear grain counts, the grains
over the nucleus and over a cytoplasmic area should be scored and
dose-response curves plotted separately. Both dose-response curves should be
considered before a decision is reached on whether exposure to a certain
chemical results in DOS.
Results of an additional DNA repair study were published by Reitz et al.
(1980). Mice were exposed orally to CHC13 at 240 mg/kg and repair in the
livers was assayed. Negative results were obtained. However, these results
cannot be interpreted because no information on the methodology used to assay
for HNA repair was given. In order to interpret these results one would have
to know the length of time the mice were exposed to CHC13. Some compounds
require a longer exposure than others (e.g., 2-acetylaminofluorine in Mirsalis
et_ aj_., 1982). Also, the opportunity for false negative results, as discussed
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above, exists in this study as well.
The uns assay discussed in Chapter 48 of the de Serres and Ashby (1981)
volume was carried out with Hela cells, which do not contain a P-450
activation system. An exogenous rat liver S9 activation system was employed.
Although CHC13 tested positive in this assay in the absence of the
activation system, the discussion of this assay in the de Serres and Ashby
book suggests that this result is misleading because of an inadequacy in the
statistical method employed. In the presence of rat liver S9, CHC13 tested
negative. However, as already discussed, exogenous activation is probably
inappropriate when testing compounds such as CHC13 for their ability to
cause genetic damage.
In summary, because of various deficiencies in the above studies, the
determination of the DNA-damaging potential of CHC13 requires additional
studies. Indication of the DNA-damaging potential of CHC13 was suggested by
the small increases in conversion and mitotic crossing over observed in yeast
(Callen et_ a_K, 1980), as described in the previous section.
7.6 CHROMOSOME STUDIES
Kirkland et_ aj_. (1981) studied the ability of CHC13 to induce
chromosome breakage in cultured human lymphocytes. The cells from one donor
were treated with CHC13 at 50, 100, 200, and 400 ug/ml for 2 hr in the
presence of an S9 activation system derived from Aroclor 1254-induced rats.
The positive control compound, benzo[a_]pyrene, in a separate experiment with
the same donor's lymphocytes induced chromosome breakage with or without S9
treatment. The response of this donor's lymphocyte chromosomes to CHC13 was
a random variation around the control value. The highest breakage level was
at 200 ug/ml with 8 breaks/100 cells compared with 5.5 breaks/100 cells in the
control. However, this difference was not significant according to the
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chi-square test. The same four problems discussed in the previous section for
the SCE study carried out by Kirkland et_ aK (1981) apply to their study on
chromosome breakage.
According to Schmid (1976), the bone marrow micronucleus test can be used
to detect clastogens and spindle poisons. Micronuclei are small elements that
contain either pieces of chromosomal fragments originating from clastogenic
events or whole chromosomes resulting from malfunction of the spindle
aparatus. Gocke et_ aj_. (1981) used this test to study chromosome aberrations
in mice exposed to CHC13. The animals were treated with CHC13 at 0 and 24
hr, and bone-marrow smears were prepared at 30 hr. The purity of the CMC!3
(purchased from Merck) was not provided. Four mice (two males and two
females) were used for each of three doses and one control. The animals were
given two intraperitoneal injections of CHC13, each at 238, 476, and 952
mg/kg (2, 4, and 8 mmol/kg, respectively). The authors state that the assay
was performed according to Schmid (1976); thus, it can be assumed that the
doses chosen included the highest tolerable dose. Slides were coded, and 1000
polychromatic erythrocytes were scored per mouse.
The results were as follows (dose, % micronucleated polychromatic
erythrocytes): 0 mg/kg, 1.2%; 2 x 238 mg/kg, 2.2%, 2 x 476 mg/kg, 2.6%; 2 x
952 mg/kg, 2.2%. Thus, a dose-related increase was not observed. Three
halogenated alkanes were tested (dichloromethane, 1,1,1-trichloroethane, and
CHC13) and all yielded negative results. Of 30 chemicals tested, only two
(pyrogallol and hydroquinone) yielded positive results in the micronucleus
test. Positive controls were not included in the assay, but the positive
results for pyrogallol and hydroquinone indicate that the assay system was
working.
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The micronucleus test was also used by Agustin and Lim-Sylianco (1978) to
study the clastogenic potential of CHC13. The authors tested seven
concentrations of CHC13 up to 900 mg/kg in the mouse. The number of mice
used and their sex was not specified. The CHC13 was purchased from
Mallinckrodt and was redistilled before use. Information on purity of the
CHC13 was not provided. For each slide, 1000 polychromatic erythrocytes
were scored. The authors reported that CHC13 was clastogenic. Results were
as follows (dose in mg/kg, number of micronucleated polychromatic erythrocytes
per 1000 polychromatic erythrocytes _+ SE): 0, 4 +_ 1; 100, 3 +_ 1; 200, 5 _+ 1;
400, 5 4. 1; 600, 9 _+ 2; 700, 17 _+ 4; 800, 9 ± 2; 900, 10 _+ 2. The authors
stated that these data indicate that CHC13 must be metabolized to a
clastogenic substance, because a straight-line dose-response relationship was
not observed. However, the data provided by the authors are not sufficient to
support this interpretation.
In the same paper, Agustin and Lim-Sylianco demonstrated that vitamin E
administered 1 hr after CHd3 reduced the number of micronucleated cells
observed at 700 mg of CHCl3/kg ( 17 +_ 4) to the control level (4 _+ 1). The
significance of this result is not clear.
This study by Agustin and Lim-Sylianco (1976) is difficult to interpret
because details of the experimental procedures necessary to permit an
evaluation of the results were not provided (e.g., number and sex of the
animals and positive and negative controls). This study suggests that CHC13
may affect chromosomes. However, corroborative studies are needed to confirm
or refute this suggested response.
In summary, based on the results of the three chromosome aberration
studies described above, there are suggestive but not conclusive data that
CHC13 is clastogenic. Negative results were reported by Kirkland et al.
7-21
-------
(1981) and by Gocke et_ aj_. (1981), while Agustin and Lim-Sylianco (1976)
reported a positive result. More studies are needed before it can be
conclusively determined whether or not CHC13 is clastogenic.
7.7 SUGGESTED ADDITIONAL TESTING
More studies are needed on the covalent binding of ^CHCl3 to DNA.
Such studies should be done with 14CHCl3 at a higher specific activity
[about 40 Ci/mol (Brookes and Lawley, 1971)] than used by Diaz Gomez and
Castro (1980b).
Additional studies on the ability of CHC13 to cause DNA damage are
needed. Examples of such tests include measurement of unscheduled DNA
synthesis jrj vi^/o in mice and in both rat and mouse hepatocytes after in vitro
exposure (Williams, 1981), measuring endpoints suggested by Lonati-Gal1igani
et_a_l_. (1983).
Additional studies are needed to corroborate and extend or refute the
Cal len et_ aj_. (1980) study in yeast, which possesses an endogenous activation
system and is capable of assaying for point mutations, mitotic crossing over,
and gene conversion.
Further testing for the ability of CHC13 to cause chromosome aberrations
is needed, particularly in jn vivo systems.
In addition to the study by Gocke et^ aj_. (1981), further testing of the
ability of CHC13 to cause sex-linked recessive lethal mutations in
Drosophjla is needed. However, in order to detect a weak response, a larger
number of chromosomes than analyzed by Gocke et_ aJN (1981) should be scored.
7.8 SUMMARY AND CONCLUSIONS
It has been demonstrated that chloroform (CHC13) can be metabolized j_n_
vivo and in vitro to a substance(s) (presumably phosgene) that interacts with
protein and lipid. However, the only experiment measuring interaction of
7-22
-------
metaholically activated CHC13 with DNA, in which adequate information was
given concerning the experimental procedures used, yielded a negative result
(Diaz Gomez and Castro, 1980h). This result was judged as inconclusive
because the specific activity of the 14CHCl3 may have been too low.
The majority of the assays for mutagenicity and genotoxicity have also
yielded negative results. However, many of these results are inconclusive
because of various inadequacies in the experimental protocols used. The major
problem is with those bacterial, sister chromatid exchange, and chromosome
aberration studies that used reconstituted exogenous activation systems (i.e.,
S9 mix). In none of these studies was it shown that CHC13 was activated or
metabolized by the activation system used. Metabolism of 2-aminoanthracene or
vinyl compounds (used as positive controls) is probably an inadequate
indication that the activation system can metabolize CHC13, because these
substances are not halogenated alkanes and are therefore not metabolized like
them. A better indication that the activation system is sufficient for
metabolism of CHC13 may be to show that it metabolizes l^CHC^ to
intermediates that bind to macromolecules. A second problem with experimental
protocols utilizing exogenous activation systems relates to the possibility
that any reactive metabolic intermediates formed may react with microsomal or
membrane lipid or protein before reaching the DNA of the test organism. A
third potential problem occurs in those jn yjtro protocols in which
precautions were not taken to prevent escape of volatilized CHC13.
Studies in which endogenous or in yi^p activation systems were used
include those reported by Callen et_ aj_. (1980) in yeast, Gocke et_ aj_. (1981)
in Drosophija (sex-linked recessive lethal test) and mice (bone-marrow
micronucleus test), Topham (1980) and Land et_ a]_. (1981) in mice (sperm head
abnormalities), and Agustin and Lim-Sylianco (1978) in mice (bone marrow
7-23
-------
micronucleus test and host-mediated assay). The results from several of these
studies suggest that CHC13 may be a weak mutagen.
Tn summary, with the present data, no definitive conclusions can be
reached concerning the mutagenicity of CHC13. However, there is some
indication (from the binding studies and from the mutagenicity tests that
utilized endogenous or ui vivo metabolism) that CHC13 may have the potential
to be a weak mutagen. In order to substantiate this, only certain
wel1-designed in vivo mutagenicity studies or studies with organisms
possessing endogenous eucaryotic P-450 activation systems are recommended.
7-24
-------
7.9 REFERENCES
Agustin, J.S., and C.Y. Lim-Sylianco. 1978. Mutagenic and clastogenic
effects of chloroform. Bull. Phil. Biochem. Soc. 1:17-23.
Brookes, P., and P.O. Lawley. 1971. Effects on DNA: Chemical methods.
In: A. Hollaender (ed.) Chemical Mutagens. Vol 1, pp. 121-144. New
York: Plenum Press.
Callen, n.F. C.R. Wolf, and R.M. Phil pot. 1980. Cytochrome P-450 mediated
genetic activity and cytotoxicity of seven halogenated aliphatic
hydrocarbons in Saccharomyces cereyjsjae. Mutat. Res. 77:55-63.
de Serres, F.J., and J. Ashby (eds.) 1981. Evaluation of Short-Term Tests
for Carcinogens. Progress in Mutation Research, Vol I. Elsevier/North
Hoi land.
Diaz Gomez, M.I., and J.A. Castro. 1980a. Nuclear activation of carbon
tetrachloride and chloroform. Res. Commun. Chem. Pathol. Pharmacol.
27:191-194.
Diaz Gomez, M.I. and J.A. Castro. 1980b. Covalent binding of chloroform
metabolites to nuclear proteins-no evidence for binding to nucleic acids.
Cancer Lett. 9:213-218.
Gocke, E., M.-T. King, K. Eckhardt, and n. Wild. 1981. Mutagenicity of
cosmetics ingredients licensed by the European communities. Mutat. Res.
90:91-109.
Kirkland, O.J., K.L. Smith, and N.J. Van Abbe. 1981. Failure of chloroform
to induce chromosome damage or sister-chromatid exchanges in cultured
human lymphocytes and failure to induce reversion in Esc her ichi a cojj .
Fd. Cosnet. Toxicol. 19:651-656.
Kirk-Othmer Encyclopedia of Chemical Technology, Second Edition, Supplement
volume. 1971. pp. 674-683. Interscience Publishers.
Land, P.C., E.L. Owen, and H.W. Linde. 1981. Morphologic changes in mouse
spermatozoa after exposure to inhalational anesthetics during early
spermatogenesis. Anesthesiol. 54:53-56.
Lonati-Galligani, M., P.H.M Lohman, and F. Berends. 1983. The validity of
the autoradiographic method for detecting DNA repair synthesis in rat
hepatocytes in primary culture. Mutat. Res. 113:145-160.
Mirsalis, J.C., C.K. Tyson, and B.E. Butterworth. 1982. Detection of
genotoxic carcinogens in the jn yiy°, ~ j.n. vitro hepatocyte DNA repair
assay. Environ. Mutagen. 4:553-562.
National Cancer Institute (NCI). 1976. Report on Carcinogenesis Bioassays of
Chloroform. National Technical Information Service, Springfield, Virginia
(NTIS PB-264-018).
7-25
-------
Schmid, W. 1976. The micronucleus test for cytogenetic analysis. In: A.
Hollaender (ed.). Chemical Mutagens. Vol. 4, pp. 31-53. New York:
Plenum Press.
Simmon, V.F., K. Kauhanen, and R.G. Tardiff. 1977. Mutagenic activity of
chemicals identified in drinking water. In: 0. Scott, B.A. Bridges, and
F.H. Sobels (eds.), Progress in Genetic Toxicology, pp. 249-258. New
York: Elsevier/North Holland Biornedical Press.
Sturrock, J. 1977. Lack of mutagenic effect of halothane or chloroform on
cultured cells using the azaguanine test system. Br. J. Anaesth.
49:207-210.
Topham, J.C. 1980. Do induced sperm-head abnormalities in mice specifically
identify mammalian mutagens rather than carcinogens? Mutat. Res.
74:379-387.
Uehleke, H. T. Werner, H. Greim, and M. Kramer. 1977. Metabolic activation of
haloalkanes and tests in vjtro for mutagenicity. Xenobiotica. 7:393-400.
Weisburger, J.H. and G.M. Williams. 1982. Metabolism of chemical carcinogens.
In: F.F. Becker (ed.). Cancer, a Comprehensive Treatise, 2nd Edition.
vol. 1, pp. 241-333. New York: Plenum Press.
Reitz, R.H., J.F. Quast, W.T. Stott, P.G, Watanabe, and P.J. Gehring. 1980.
Pharmacokinetics and macromolecular effects of chloroform in rats and
mice: Implications for carcinogenic risk estimation. In: R.L. Jolley,
W.A. Brungs, and R.B. Cumming (eds.). Water chlorination: Environmental
Impact and Health Effects, Vol. 3, pp. 983-992.
White, A.E., S. Takehisa, E.I. Eger, S. Wolff, and W.C. Stevens. 1979.
Sister chromatid exchanges induced by inhaled anesthetics. Anesthesiol.
50:426-430.
Williams, G.M. 1981. Liver culture indicators for the detection of chemical
carcinogens. In: Short-Term Tests for Chemical Carcinogens, H.F. Stich and
R.H.C. San (eds.), pp. 275-289. New York: Springer Verlag.
Wyrobek, A. and W.R. Bruce. 1978. The induction of sperm-shape abnormalities
in mice and humans. In: A. Hollaender (ed.). Chemical Mutagens.
Vol. 5, pp. 257-285. New York: Plenum Press.
7-26
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8. CARCINOGENICITY
8.1. ANIMAL STUDIES
The carcinogenicity of chloroform has been evaluated in mice, rats, and
dogs. Evidence for carcinogenic activity by chloroform includes induction of
renal epithelial tumors in male Osborne-Mendel rats (National Cancer Institute
[NCI] 1976), hepatocellular carcinomas in male and female B6C3F1 mice (NCI
1976), kidney tumors in male ICI mice (Roe et al. 1979), and hepatomas in
female strain A mice (Eschenbrenner and Miller 1945) and NIC mice (Rudali
1967). Cape! et al. (1979) demonstrated an ability of chloroform to promote
growth and metastasis of murine tumors. Chloroform was not shown to be
carcinogenic in (C57 x DBA2 Fl) mice (Roe et al. 1968), female Osborne-Mendel
rats (NCI 1976), female ICI mice, and male mice of the CBA, C57BL, and CF/1
strains (Roe et al. 1979), male and female Sprague-Dawley rats (Palmer et al.
1979), and male and female beagle dogs (Heywood et al. 1979). Chloroform was
negative in a pulmonary tumor induction bioassay in male strain A/St mice
(Theiss et al. 1977). Chloroform in liquid solution did not induce trans-
formation of baby Syrian hamster kidney (BHK-21/C1 13) cells in vitro. Under
the conditions of the carcinogenicity bioassays showing carcinogenic activity
for chloroform specifically in kidney and liver of mice and rats, the conclu-
sion can be made, by applying the IARC classification approach for carcinogens,
that there is sufficient evidence for the carcinogenicity of chloroform in
experimental animals.
8-1
-------
8.1.1. Ural Administration (Gavage): Rat
8.1.1.1. NATIONAL CANCER INSTITUTE (1976) -- A carcinogenesis bioassay
on chloroform in Osborne-Mendel rats was reported by the NCI (1976). The
chloroform product (Aldrich Chemical Company, Milwaukee, Wisconsin) was shown
to be 98% pure chloroform and 2% ethyl alcohol (stabilizer) by gas-liquid
chromatography, flame ionization detection, and infrared spectrometry at the
carcinogenesis bioassay laboratory. Chloroform solutions in corn oil were
prepared fresh each week and stored under refrigeration.
Fifty animals of each sex were assigned to each of two dose groups.
Treated animals were compared with matched vehicle-control groups (20 males and
20 females) and with vehicle colony control groups (99 males and 98 females) that
included the matched control group and three other controls groups put on study
within 3 months of the matched control group. Matched control and treated
animals were housed in the same room, and colony controls were housed in two
different rooms.
Doses selected for the main study in rats were estimated as those maximally
and one-half maximally tolerated based on survival, body weights, clinical signs,
and necropsy examinations in a preliminary toxicity test in which chloroform was
given by gavage for fa weeks with a subsequent observation period of 2 weeks
without treatment. The chronic study began when rats were 52 days old and
ended with sacrifice of survivors at 111 weeks. Chloroform was administered in
corn oil by gavage 5 days each week during the initial 78 weeks. Doses of 90
and 180 mgAg/day were administered to male rats throughout the chronic study;
however, since initial doses of 125 and 250 mg/kg/day were reduced to 90 and 180
mgAy/day at 22 weeks, doses given to female rats were expressed as time-weighted
averages of 100 and 200 mg/kg/day.
8-2
-------
Decedents and survivors were necropsied, and tissues and organs were ex-
amined microscopically. Body weights and food consumption were monitored weekly
for the first 10 weeks and monthly thereafter. Animals were observed twice daily.
In matched control and both dose groups, at least 50% of the male arid
female rats survived as long as 85 and 75 weeks, respectively. Seven matched
control, 24 low-dose, and 14 high-dose males and 15 matched control, 22 low-dose,
and 14 high-dose females survived until the end of the study. Only one control
male rat died before 90 weeks; the increase in death rate of control males after
90 weeks was, according to the NCI (1976) report, "probably due to respiratory
and renal conditions." Overall survival was less in treated animals than in
controls (Figure 8-1).
Decreased body weight gain was evident in both sexes of rats in both
treatment groups. Initial mean body weights for all groups were about 175 g
for females and 250 g for males. By 50 weeks, mean body weights were approxi-
mately 400 g in control, 350 g in low-dose, and 330 g in high-dose females;
by 100 weeks, mean body weights were about 375 g in all groups of females. In
males, mean body weights were about 640 g in the control group, 550 g in the
low-dose group, and 500 g in the high-dose group by 50 weeks; by 100 weeks,
mean body weights were approximately 500 g in all groups. Food consumption
was reported as slightly lower in treated animals, but data were not provided.
Appearance and behavior among groups were generally similar, but hunching,
urine stains on the lower abdomen, redness of eyelids, and wheezing were noted
in treated animals early in the study.
A statistically significant (P < 0.05) increase in renal epithelial
tumors of tubular cell origin was found in treated male rats (Table 8-1). The
epithelial tumors were described as follows: Of 13 tumors in high-dose males,
10 were carcinomas and three were adenomas; two carcinomas and two adenomas
8-3
-------
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-------
TABLE 8-1. EFFECT OF CHLOROFORM ON KIDNEY EPITHELIAL TUMOR INCIDENCE
IN OSBORNE-MENDEL RATS
(NCI 1976)
Male
Female
Treatment3
Controls'3
Colony Matched
Dose (mgAg/day)c
90 180
Controlsb
Colony Matched
Dose (mgAg/day)c
100 200
CO
1
cr
Kidney tumor 0/99(0%)
incidence0'
P value"
Time to first
tumor (weeks) —
0/19(0%) 4/50(8%)e
0.266
102
12/50(24%)f
0.0141
80
0/98(0%) 0/20(0%) 0/49(0%) 2/48(4%)9
0.495
102
Survival at
terminal
sacrifice
(111 weeks)
26%
37%
48%
28%
51%
75%
45%
29%
aChloroform in corn oil administered by gavage 5 times per week for 78 weeks.
^Colony controls consist of four vehicle-control groups, including matched controls, given corn oil.
cDoses are time-weighted averages.
dAnimals with tumor/animals examined.
eTwo with tubular cell adenocarcinoma and two with tubular cell adenoma.
^Ten with tubular cell adenocarcinoma and two with tubular cell adenoma.
90ne with tubular cell adenocarcinoma and one with squamous cell carcinoma in the renal pelvis.
^Fisher's Exact Test, compared with matched controls.
iFor adenocarcinomas alone, P value is 0.03.
-------
comprised the tumors found in four low-dose males; one renal epithelial car-
cinoma and one squamous cell carcinoma from renal pelvic transitional epithelium
were noted in two high-dose females. One low-dose male had both a malignant
mixed tumor and a tubular cell adenoma in the left kidney, and a high-dose
male had a tubular cell carcinoma and a tubular cell adenoma in the right
kidney. Renal epithelial carcinomas were large and poorly circumscribed, and
they infiltrated surrounding normal tissue. Renal epithelial adenomas were
circumscribed and well-differentiated. Additional kidney tumors included
malignant mixed tumors in two low-dose and two colony control males and
hamartomas in one low-dose male, one high-dose male, and one colony control
male.
Although a statistically significant (P < 0.05) increase in thyroid tumors
in both treatment groups of female rats as compared with colony controls was
reported, the toxicologic significance of this finding appears questionable in
that C-cell tumors and follicular cell tumors, which have different embryonic
origins and different physiologic functions, are combined in the incidences
described in Table 8-2; the majority of tumors were adenomas; the spontaneous
incidence of thyroid tumors in Oshorne-Mendel females is variable as stated,
without presentation of historical data, in the NCI (1976) bioassay report;
and the increased incidence of thyroid tumors in treated females is not
significant (P > 0.05) when compared with data for matched controls.
No significant (P < 0.05) differences for other tumor types among groups
were apparent. Four rats were lost (missing or autolyzed) for pathology.
Non-neoplastic lesions described as treatment-related include necrosis of
liver parenchyma, epithelial hyperplasia in the urinary bladder, and
hernatopoiesis in spleen. Inflammatory pulmonary lesions characteristic of
8-6
-------
TABLE 8-2. EFFECT OF CHLOROFORM ON THYROID TUMOR INCIDENCE
IN FEMALE OS60RNE-MENDEL RATS
(NCI 1976)
Dosea>b
(mg/kg/day)
0 (matched)"
o° 0 (colony)h
100
200
Fol licular cell
tumors0
Incidence^
1/19(5%)
1/98(1%)
2/49(4%)
6/49(12%)
C-cell
tumorsd
Incidence
0/19(0%)
0/98(0%)
6/49(12%)
4/49(8%)
Incidence
1/19(5%)
1/98(1%)
8/49(16%)
10/49(20%)
Total tumors6
Time to first tumors
P value9 (weeks)
110
110
0.216 73
0. 121 49
aChloroform in corn oil administered by gavage 5 times per week.
^Time-weighted average doses.
cAdenomas except for carcinoma in one low-dose and two high-dose animals.
^Adenomas except for carcinoma in one high-dose animal.
eSee text.
fAnimals with tumors/animals examined.
SFisher's Exact Test, compared with matched controls.
nColony controls consist of four vehicle-control groups, including matched controls, given corn oil.
-------
pneumonia were found in all groups, but the severity and incidence of these
lesions were stated (data not reported) to have been greater in treatment groups.
Under the conditions of this bioassay, chloroform treatment significantly
(P < 0.05) increased the incidence of renal epithelial tumors in male Osborne-
Mendel rats. Although the number of matched vehicle controls was low, the
use of pooled colony controls gives additional support for treatment-related
effects. Moreover, historical control incidence of renal epithelial tumors in
Osborne-Mendel rats was reported as rare.
Lower survival rates and body weights in rats than in matched controls
provide evidence that the chloroform doses used were toxic to the rats used in
this study, and a more precise estimate of dose-response perhaps could have
been obtained if additional lower doses had been given, and if constant doses
rather than time-weighted averages had been used. Treated animals were housed
in the same room as rats treated with other volatile compounds (1,1,2,2-tetra-
chloroethane, 3-chloropropene, ethylene dibromide, carbon tetrachloride);
however, since controls were in the same room as treated animals and oral
chloroform doses probably would have been much higher than ambient levels of
other volatiles, the likelihood that the other volatile compounds were responsible
for the observed results is considered to be low. Additionally, these other
volatile compounds did not induce kidney tumors in Osborne-Mendel rats (NCI
1976; Weisburger 1977). It should be noted that ambient levels of volatiles in
the animal quarters were not measured.
8.1.1.2. PALMER ET AL. (1979) -- Palmer et al. (1979) reported carcin-
ogenicity studies on chloroform in Sprague-Dawley rats. Chloroform was prepared
in toothpaste, as described in Table 8-3 herein for the Roe et al. (1979) study,
and administered by gavage. Dose levels of 15, 75, and 165 mg CHCl3A9/day were
-------
selected for the carcinogenicity study based on results of a preliminary range-
finding study showing the lowest toxic dose, indicated by liver and kidney
changes, as 150 mg/kg/day.
TABLE 8-3. TOOTHPASTE FORMULATION FOR CHLOROFORM ADMINISTRATION
(Roe et al. 1979)
Ingredient Percentage w/w
Chloroform 3.51
Peppermint oila 0.25
Eucalyptoia 0.50
Glycerol 39.35
Carragheen gum 0.45
Precipitated calcium carbonate 48.53
Sodium lauryl sulphate 1.16
Sodium saccharin 0.03
White mineral oil 1.10
Water 5.12
Total 100.00
aEssential oil flavor components.
An initial carcinogenicity study was done in which 25 rats of each sex per
group received one of the selected doses in toothpaste containing essential
oils (flavor components), indicated in Table 8-3, 6 days per week. A concurrent
control group of 75 males and 75 females was administered toothpaste without
chloroform and essential oils. A second carcinogenicity study was done in which
8-9
-------
50 male and 50 female specific pathoyen-free (SPF) Sprague-Dawley rats were
dosed with 60 mg CHCl3/kg/day in toothpaste with essential oils b days per week,
and 50 control rats of each sex were given toothpaste without chloroform but
with essential oils.
Body weights were measured weekly, and food consumption was recorded.
Body weights were initially 180 to 240 g for males and 130 to 175 g for females.
Blood and urine analysis were performed in the first study, and serum and
erythrocyte cholinesterase activities were monitored in the second study.
All animals were necropsied, and tissues and organs were examined histopatho-
logically. Adrenals, kidneys, liver, lungs, and spleen were weighed.
Chloroform was not carcinogenic in these studies. Significant (P < O.Ob)
body weight loss in high-dose males in the first study (data not reported) and
maximal body weight gain of approximately 370 g in control males, 330 g in
treated males, 220 g in control females, and 180 g in treated females in the
second study suggest an effect from chloroform treatment, Other than a 40%
reduction of plasma cholinesterase levels and slight decreases in serum
glutamic-pyruvic transaminase and serum alkaline phosphatase in treated females,
additional toxic effects from chloroform treatment were not evident.
Low survival, attributed to respiratory disease, was apparent in both
studies. The initial study was terminated at 52 weeks; 50% of the animals
in all groups had died by 52 weeks in the second study, which was ended at 95
weeks. Except for 48 control females in the initial study, no more than 18
animals were alive in each group at the conclusion of either study. Although
carcinogenic activity for chloroform was not observed, these studies on
Sprague-Dawley rats are weakened by the high early mortality in control and
treated animals.
8-10
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8.1.2. Oral Administration (Gavage): Mouse
8.1.2.1. NATIONAL CANCER INSTITUTE (1976) -- The carcinogenicity of
chloroform in B6C3F1 mice was evaluated by the NCI (1976). The chloroform
product and chloroform solutions in corn oil were those used in the NCI (1976)
carci nogenesis bioassay in rats discussed herein.
Each of two dose groups was composed of 50 males and 50 females. Treated
mice were compared with matched vehicle control groups (20 males and 20 females)
put on study 1 week earlier and with vehicle colony control groups (77 males and
80 females), which included the matched control group and three other control
groups put on study within 3 months of the matched control group. All control
mice were housed in the same room with treated mice.
Maximally and one-half maximally tolerated doses for the main study in
mice were estimated from a preliminary toxicity test done as described for the
NCI (1976) rat study. Mice were started in the chronic study at 35 days of
age, and the study was concluded with sacrifice of survivors at 92-93 weeks.
Chloroform in corn oil was administered by gavage 5 days per week during the
first 78 weeks. Initial dose levels of 200 and 100 mg/kg/day for males and
400 and 200 mg/kg/day for females were raised to 300 and 150 mg/kg/day for males
and 500 and 250 mg/kg/day for females at 18 weeks. Thus, doses expressed as time-
weighted averages for the entire study were 138 and 277 mg/kg/day for males and
238 and 477 mg/kg/day for females.
Survival in mice was similar among groups except for high-dose females.
At least 50% of the animals in each group survived as long as 85 weeks. Ten
matched control, 33 low-dose, and 30 high-dose males, and 15 matched control,
34 low-dose, and 9 high-dose females survived for the duration of the study.
All but two deaths in high-dose females occurred after 70 weeks.
8-11
-------
Body weight gain among groups was comparable. Male and female mice in-
itially weighed about 18 and 15 g, respectively. Mean body weights at 50 weeks
were approximately 35 g in males and 28 g in females, and these levels were
generally sustained throughout the remainder of the study. Food consumption
was stated to have been equivalent among groups. Appearance and behavior among
groups were similar except for bloating and abdominal distension noted in
treated animals beginning after 42 weeks of treatment.
Statistically significant (P < 0.05) increases in hepatocellular carcinomas
in both sexes in both treatment groups of mice were observed (Table 8-4).
Various histopathologic types of hepatocellular carcinomas were observed.
Hepatocellular carcinoma metastasized to the lung in two low-dose males and
two high-dose females, and to the kidney in one high-dose male. Twenty animals
were reported as missing or autolyzed, and therefore were not included in the
pathology report.
Non-neoplastic lesions in mice attributed to treatment include nodular
hyperplasia of the liver in 10 low-dose males, six low-dose females, and one
high-dose female; and liver necrosis in one low-dose male, four low-dose females,
and one high-dose female. Nine high-dose females with hepatocellular carcinoma
had cardiac atrial thrombosis. Kidney inflammation was diagnosed in 10 matched
control males, two low-dose males, and one high-dose male.
Under the conditions of this bioassay, chloroform treatment significantly
(P < 0.05) increased the incidence of hepatocellular carcinoma in male and
female B6C3F1 mice. Although the number of matched vehicle controls was low,
the use of pooled colony controls gives additional support for treatment-related
effects. Moreover, historical control incidence of hepatocellular carcinomas
in B6C3F1 mice was reported as 5-10% in males and 1% in females.
8-12
-------
TABLE 8-4. EFFECTS OF CHLOROFORM ON HEPATOCELLULAR CARCINOMA INCIDENCE IN B6C3F1 MICE
(NCI 1976)
Treatment3
Male
Controls5
Colony Matched
Dose (mg/kg/day)c
138 277
Female
Controls'5
Colony Matched
Dose (mg/kg/day)c
238 477
00
Hepatocellular
carcinoma
i ncidence^
P value8
Time to first
tumor (weeks)
5/77(8%) 1/18(6%) 18/50(36%) 44/45(98%)
0.011 3.13xlO-13
1/80(1%) 0/20(0%)
36/45(80%) 39/41(95%)
4xlO-10 3.7xlO-14
72
72
80
54
90
66
aChloroform in corn oil administered by gavage 5 times per week.
bColony controls consist of four vehicle-control groups, including matched controls, given corn oil
cDoses a^e time-weighted averages.
dAnima1s with tumor/animals examined.
9Fishea's Exact Test, compared with matched controls.
67
Survival at
terminal
sacrifice
(92 weeks)
48%
50%
65%
65%
81%
75%
75%
20%
-------
A more precise estimate of dose-response perhaps could have been obtained
if additional lower doses had been used and if constant doses rather than time-
weighted averages had been used. Treated animals were housed in the same room
as animals treated with other volatile compounds*; however, since 1) controls
were in the same room as treated animals, 2) oral chloroform doses probably
would have been much higher than ambient levels of other volatiles, 3) the
cages had filters to limit the amount of chemical released into the ambient
air, 4) the total room air was exchanged 10 to 15 times per hour, and 5) dosing
was done in another room under a large hood, the likelihood that the other vol-
atile compounds were responsible for the observed results is considered to be
low. It should be noted that ambient levels of volatiles in the animal quarters
were not measured.
8.1.2.2. ROE ET AL. (1979) -- Roe et al . (1979) studied the carcinogenicity
of chloroform in toothpaste in four strains of mice (C57BL, CBA, CF/1, and ICI).
The toothpaste formulation used is presented in Table 8-3. The chloroform
product was described as British Pharmacopoeia grade which was not contaminated
with other haloalkanes or phosgene. Toothpaste was prepared fresh each month.
Chloroform in arachis oil was also tested in ICI mice.
Dose levels for the carcinogenici ty studies were selected based on results
of a 6-week preliminary range-finding study in male and female Schofield mice
which indicated moderate weight gain reduction at the lowest toxic dose of
60 mg CHCl3/kg/day. Three different carci nogenicity studies were conducted in
*l,l,2,2-tetrachloroethane, 3-chloropropene, chloropicrin,
1,1-dichloroethane, trichloroethylene, sulfolene, iodoform, ethylene dichloride,
methyl chloroform, 1,1,2-trichloroethane, tetrachloroethylene, hexachloroethane,
carbon disulfide, trichlorofluoromethane, carbon tetrachloride, ethylene
dibromide, dibromochloropropane.
8-14
-------
which mice, initially no more than 10 weeks old, were given chloroform by
gavage 6 days per week for 80 weeks followed by observation for 13 to 24 weeks.
In one study, 52 male and 52 female ICI mice per dose group were given 17 or
60 mg/kg/day of chloroform in toothpaste and compared with 100 ICI mice of each
sex concurrently given toothpaste without chloroform, peppermint oil, or
eucalyptol. A second study, confined to male ICI mice, included 52 untreated
mice, 260 mice given toothpaste alone without chloroform, eucalyptol, or
peppermint oil, and groups of 52 mice each given, in toothpaste, 60 mg
CHCl3/kg/day, eucalyptol up to 32 mg/kg/day, or peppermint oil up to 16 mg/kg/day;
treatment with chloroform, eucalyptol, or peppermint oil was done in the
absence of the other two compounds. In the third study, groups of 52 male mice
of each of the C57CL, CBA, CF/1, and ICI strains were given 60 mg CHCl3/kg/day
in toothpaste and compared with concurrent vehicle-control groups of 52 mice
each, and with 100 untreated ICI mice. Fifty-two ICI male mice given 60 mg
CHCl3/kg/day in arachis oil, and concurrent control mice given arachis oil
alone, were also evaluated in the third study.
Body weights were recorded in each study, and food consumption was
estimated in the second and third studies. In each study, the animals were
necropsied, and tumors and lesions as well as routine tissues and organs were
examined histopathologically. Adrenals, kidneys, livers, lungs, and spleens
were wei ghed.
Although the authors stated (data were not reported) that body weight gain
was poorer in each treatment group than in controls in the third study on four
mouse strains, differences in survival, body weights, and food consumption
between control and treatment groups were not statistically (P < 0.05) signifi-
cant, either as shown with data or as stated by the authors without data.
Median survival was ^> 73 weeks for all groups in the two studies on ICI mice
8-15
-------
alone; by terminal sacrifice in the study on four strains (survival patterns
were not reported), 52 to 79% of the C57BL and CBA mice and 12 to 31% of the
CF/1 and ICI mice were alive. Liver and kidney weights were slightly lower
(data not reported) in male ICI mice given chloroform in toothpaste. Incidences
of tumors and lesions between control and chloroform-treated animals were not
significantly (P < 0.05) different except for: 1) increased kidney tumor
incidences in treated male ICI mice, as shown in Table 8-5, and 2) a significantly
(P < 0.001, chi-square test) higher incidence of moderate to severe kidney
"changes" in treated CBA and CF/1 males than in corresponding controls and of
moderate to severe kidney disease (P < 0.05, chi-square test) in ICI males
given CHC13 in arachis oil than in arachis oil controls, as described by the
authors without presentation of data. Results in Table 8-5 indicate more
effective induction of kidney tumors by chloroform in arachis oil than by
chloroform in toothpaste. Kidney tumors were not found in C57BL, CBA, and
female ICI mice, and malignant kidney tumors were diagnosed in two control
and one treated CF/1 mice. Malignant kidney tumors were identified as hyper-
nephromas, and benign kidney tumors were characterized as cortical adenomas.
Eucalyptol and peppermint oil were not toxic in male ICI mice in these studies.
Results of the studies by Roe et al. (1979) show the ability of chloroform
to induce kidney tumors in male ICI mice. The stronger induction of kidney tu-
mors by chloroform in arachis oil compared with chloroform in toothpaste may re-
flect an effect of dosing vehicle on chloroform absorption, since Moore et al.
(1982) demonstrated greater severity of acute toxicity and regenerative changes
in kidneys of male CFLP mice given single gavage doses of 60 mg CHCl3/kg when
corn oil rather than toothpaste was the dosing vehicle. Kidney pathology was
noted in treated animals in the study with four strains of mice; however,
although poorer body weight gain reported for treated mice in each of the
8-16
-------
TABLE 8-5. KIDNEY TUMOR INCIDENCE IN MALE ICI MICE
TREATED WITH CHLOROFORM
(adapted from Roe et al. 1979)
Dose group
Numbers of mice
examined histo-
logically
Number of mice with kidney tumors
Benign Malignant Total
First study
Vehicle-control3
17 mg CHCl3Ag/dayb
60 mg CHCl3/kg/dayb
Second study
Untreated control
Vehicle-control3
60 mg CHCl3/kg/dayc
Third study
Untreated control
Vehicle-control^
60 mg CHCl3/kg/daye
Vehicle-control^
60 mg CHCl3/kg/day9
72
37
38
45
237
49
83
49
47
50
48
0
0.
0
1
2
1
3
0
0
31
0
0.
21
0
0
3
&
0
0
8J
1
6
9j
0
1
5
0
12J
toothpaste base vehicle without chloroform, eucalyptol, and peppermint oi
bChloroform given in toothpaste base with eucalyptol and peppermint oil.
cChloroform given in toothpaste base without eucalyptol and peppermint oil,
toothpaste base vehicle without chloroform.
eChloroform given in toothpaste base.
fArachis oil.
9Chloroform given in arachis oil.
"Statistically significant versus vehicle-control (P < 0.05).
""Statistically significant versus vehicle-control (P < 0.01).
^Statistically significant versus vehicle-control (P < 0.001).
8-17
-------
strains would suggest that a maximun tolerated dose was being approached, the
observation that survival, body weights, and other pathology between control
and treated mice in each of the four strains were not significantly (P < 0.05)
different also suggests that higher doses could have been tested to more strongly
challenge the mice for carcinogenicity. Since mice were as old as 10 weeks at
the start of the studies, it is evident that treatment could have been started
when the mice were younger to cover a greater portion of their lifespan during
growth. A fuller evaluation of chloroform carcinogenicity could have been made
if female mice had also been included in each study.
8.1.2.3. ESCHENBRENNER AND MILLER (1945) — An early study on hepatoma
induction by chloroform in mice was described by Eschenbrenner and Miller
(1945). Strain A mice, initially 3 months old, with a historical spontaneous
hepatoma rate of < 1% at 16 months of age were selected for treatment.
"Chemically pure" chloroform was used, but chemical analysis was not indicated.
Dose groups of five males and five females each were treated with doses of
2.4, 1.2, 0.6, 0.3, or 0.15 g/kg of chloroform in olive oil by gavage. Controls
received olive oil alone.
In the study of hepatoma induction, mice were dosed every 4 days for a
total of 30 doses. When 8 months old, mice were examined for hepatomas at one
month after the last dose; however, these animals were given an additional dose
of chloroform 24 hours before necropsy. Tissues and organs were examined
histopathologically. Liver necrosis also was examined in mice given a single
gavage treatment of one of the indicated doses of chloroform (one male and two
females per group) 24 hours before removal of liver for microscopic evaluation.
Incidences of liver and kidney necrosis and hepatomas are shown in Table 8-6.
Liver necrosis was noted in both sexes in the three highest-dose groups. Males in
8-18
-------
all treatment groups developed kidney necrosis, whereas kidney necrosis was not
apparent in females. No males in the three highest-dose groups and no females
in the highest-dose group survived the study. All deaths occurred by 48 hours
after the second administration of chloroform. All surviving females dosed
with 0.6 or 1.2 g CHCl3/kg had hepatomas.
TABLE 8-6. LIVER AND KIDNEY NECROSIS AND HEPATOMAS IN STRAIN A MICE
FOLLOWING REPEATED ORAL ADMINISTRATION OF CHLOROFORM IN OLIVE OIL
(adapted from Eschenbrenner and Miller 1945)
Observation
Liver necrosis
Kidney necrosis
Deaths3
Hepatomas in
survi ving
animals receiv-
ing 30 dosesa
Sex
F
M
F
M
F
M
F
M
Dose (g/kg)
2.4 1.2 0.6
+ + +
+ + +
000
+ + +
5/5 1/5 2/5
5/5 5/5 5/5
4/4 3/3
— — — — *" —
0.3
0
0
0
+
0/5
2/5
0/5
0/3
0.15
0
0
0
+
0/5
0/5
0/5
0/5
Control
0
0
0
0
0/5
0/5
0/5
0/5
aNumerator is positive occurrences.Denominator is animals observed.
In the experiment on the ability of a single dose of chloroform to produce
tissue necrosis, there was sharp distinction between normal and necrotic cells
in liver. Doses of 2.4 and 1.2 g/kg produced extensive necrosis in all liver
lobules, and the 0.6 gAg dose produced necrosis in some lobes. Mice given
30 doses of chloroform in the hepatoma study had moderate liver cirrhosis and
8-19
-------
necrosis, however, animals given 3U doses that did not result in necrosis had
livers that appeared normal. Necrosis was not found in hepatoma cells, and
hepatomas contained cords of enlarged liver-like cells which formed disorganized
anastamosing columns. The hepatomas did not appear invasive, and metastasis was
not found.
Renal necrosis in males was localized in the areas of the proximal and
distal tubules. Glomeruli and collecting tubules appeared normal. The severity
of renal necrosis was less with lower doses. The different kidney responses
by males and females to chloroform treatment may be due to the unique lining
of the Bowman's capsules with flat and cuboidal epithelium in females and males,
respectively (an anatomic sexual dimorphism in mice). Although few animals
were available for pathologic examination, the Eschenbrenner and Miller
study (1945) indicates that hepatomas in female mice were induced at chloroform
doses that also produced liver necrosis. Early mortality precluded all animals
given chloroform doses that produced liver necrosis from developing heoatomas,
Hepatomas were not induced by non-necrotizing doses of chloroform; however, a
lifetime study perhaps could have given a stronger indication of the carcinogenic
potential of chloroform at these lower doses. The observation of kidney necrosis
in males without tumor formation and lack of necrosis in hepatomas suggests
that liver in strain A mice was uniquely sensitive to tumor induction at
necrotizing doses, or that there might have been additional factors in liver
tumor formation besides necrosis. Furthermore, since a dose of chloroform was
given 1 day before sacrifice—a factor which in itself could have been responsible
for producing necrosis, as supported by liver necrosis found in mice which died
after one or two treatments with chloroform—it is not clear what the extent of
necrosis was during the last month of observation, when the animals were un-
treated.
8-20
-------
8.1.2,4. RUDALI (1967) -- Rudali (1967) reported a carcinogenicity study
on chloroform in NLC mice. Details such as age and sex of the mice were not
given. The mice received twice-weekly doses of 0.1 ml of a 40% solution of
chloroform in oil by force-feeding for an unspecified treatment period.
Twenty-four animals were initially on study, but only five "sound mice" were
evidently given a pathologic examination. An average survival period of 297
days was reported, but it is not clear if this period applied to the total
group of 24 or to the smaller group of five. The observation period for the
study was not mentioned. The use of a control group was not indicated, nor was
a chemical analysis of the chloroform sample provided.
Three of the five mice examined in pathology were diagnosed with hepatomas
and hepatic lesions; however, details of the pathologic observations were not
reported. Although the study by Rudali (1967) gives evidence for carcinogenic
activity by chloroform in NLC mice, it is weakened by a lack of experimental
details, the absence of a control group, and the small number of animals examined
in pathology.
8.1.3. Oral Administration (Capsules): Dog
8.1.3.1. HEYWOOD ET AL. (1979) -- The carcinogenicity of chloroform in
toothpaste was evaluated in beagle dogs by Heywood et al . (1979). The tooth-
paste formulation used was that previously described in Table 8-3 herein except
for reduced amounts of carragheen gum and glycerol. Chloroform in toothpaste
was transferred from a syringe to gelatin capsules immediately before dosing.
Doses were selected from results of a preliminary range finding-study in
which one or two dogs of each sex per group were given oral chloroform doses 7
days per week for 13 (30 and 45 mg/kg/day), 18 (60 mg/kg/day), or 12 (90 and
8-21
-------
120 my/kg/clay) weeks. Because 45 mg/kg/day, which produced pathologic changes in
the liver, was the lowest toxic dose, dose levels of 0, 15, and 30 mg CHCl3/kg/day
were chosen for the carcinogenicity study.
In the carcinogenicity study, chloroform was given orally in capsules 6 days
per week for over 7 years. Eight males and eight females were assigned to each
treatment group and to an untreated control group, and 16 dogs of each sex
composed a vehicle-control group. The dogs were initially 18 to 24 weeks old.
All of the dogs were clinically examined before treatment, and had been receiving
medication annually for common diseases. Dogs were fed 200 g of diet twice
daily until week 300, when obese dogs received reduced daily rations of 300 g.
Body weights, food consumption, and water intake were estimated during the
study. Hematology, serum biochemistry, and urinalysis were included in the
evaluation of chloroform toxicity. Treatment was stopped at 376 weeks, and
survivors were sacrificed for macroscopic examination at 395 to 399 weeks.
Major organs were weighed. Tumors, lesions, and routine tissues and organs
were evaluated microscopically. Liver and kidney specimens from control and
high-dose dogs were also examined by electron microscopy.
Survival, body weights, food and water consumption, and appearance of the
eyes were unaffected by chloroform treatment. Mean body weights increased from
7 to 8 kg initially to a maximum of 14 to 15 kg; however, reduction of diet
portions for obese dogs complicated the body weight results. Results of blood
and urine analyses were unremarkable except for dose-related increases in SGPT
levels (Table 8-7), which could reflect liver pathology.
No treatment-rel ated carcinogenic effects were found in necropsy and
microscopic examination of tissues and organs. Non-neoplastic diagnoses showed
that fatty cysts in the livers of all groups were larger and more numerous in
treated dogs.
8-22
-------
TABLE 8-7. SGPia CHANGES IN BEAGLE DOGS TREATED WITH CHLOROFORM
(adapted from Heywood et al. 1979)
CO
ro
CJ
Group mean SGPT (MU/ml )
Treatment
(mg CHCl3/kg/day)
30 mg
15 mg
Vehicle-control
Untreated
Pretreatment
6 26
24 34b 58C
22 29 33
22 29 30
24 30 30
Treatment stage (weeks)
52 104 156
52C 64C 76C
32 45 46d
29 40 30
27 37 29
208
91C
55d
40
30
260 312 372
147C 128C 102C
95C 89C 66
33 47 51
32 50 50
Post-treatment (weeks)
14
105d
53
56
53
19
111
48
128
56
aberum giutamic-pyruvic transaminase.
^Comparison with untreated group; P < 0.05.
cComparison with untreated group; P < 0.01.
^Comparison with untreated group; P < 0.001.
-------
The study by Heywood et al. (1979) did not show a carcinogenic effect of
chloroform in toothpaste given to beagle dogs. Range-finding tests and SGPT
and liver fatty cyst diagnoses in the carcinogenicity study suggest that a
maximally tolerated dose was approached in the carcinogenic!ty study. It is
not certain if 7 years was long enough for carci nogenicity testing with respect
to the lifespan of the beagle dog (13 to 14 years), but by 7 years spontaneous
tumor formation was becoming evident.
8.1.4. Intraperitoneal Administration: Mouse
8.1.4.1. ROE ET AL. (1968) -- Roe et al . (1968) investigated the carcino-
genicity of chloroform in newborn (C57 X DBA£ Fl) mice. ChToroform was subcu-
taneously injected into one group of mice as a single dose of 200 ug when the
animals were less than 24 hours old, and into another group of mice as eight
daily doses of 200 ug, each beginning when the animals were 1 day old. Control
groups were given the dosing vehicle, a*rachis oil, alone. Survivors were
sacrificed for necropsy at 77 to 80 weeks.
No carcinogenic effect of chloroform was found. However, since the study
was reported as an abstract, experimental details were not provided. ChToroform
doses were rather low, and the use of newborn mice given one or a few doses of
chloroform is not equivalent to lifetime treatment of animals given doses as
high as those maximally tolerated. Additionally, there may be differences in
chloroform metabolism between newborn and adult (C57 x DBA? Fl) mice. Hence,
it is concluded that the study by Roe et al. (1968) does not present sufficient
evidence for an absence of carcinogenicity in chloroform.
8-24
-------
8.1.4.2. THEISS ET AL. (1977) -- The carcinogenicity of chloroform was
evaluated by Theiss et al. (1977) by means of the pulmonary tumor induction
bioassay in strain A mice.
Test animals were male strain A/St mice initially 6 to 8 weeks old. Pre-
liminary toxicity tests were performed for selection of maximum tolerated doses;
in these tests, mice received six intraperitoneal injections of chloroform for
2 weeks and were observed for another 4 weeks. Results of the preliminary test
were not reported. In the bioassay, chloroform doses in tricaprylin were 80
and 200 mg/kg, administered 3 times weekly for a total of 24 intraperitoneal
injections; and 400 mg/kg, which was injected only twice. Fifty control mice
were given tricaprylin alone. Each treatment group contained 20 animals. Mice
were sacrificed 24 hours after the last dose, and lungs were removed for counting
and examining surface adenomas microscopically. The chloroform product was
reagent grade (Aldrich Chemical Company), but its chemical composition was not
reported. A positive control group of 20 mice was given one injection of 1 g/kg
of urethan in saline, and compared with 50 controls given saline alone.
Chloroform treatment did not produce a pulmonary adenoma response in this
study. The average number of lung tumors per mouse was 0 to 0.39 in each
group, except for the positive controls, which had an average of 19.6 lung tumors.
At least 90% of the mice in each group survived, except for the mice given 400
mg CHCl3/kg, where there was 45% survival. However, since this type of bioassay
is basically a screen for carcinogens, a negative result does not necessarily
indicate a lack of carcinogenic potential. Evidence for the carcinogenic
activity of chloroform is available in other studies described in this document,
and, according to the authors, there is evidence for carcinogenic activity in
other compounds, e.g., 2-chloroethyl ether and hexachlorocyclohexane in liver,
which also tested negative in the Theiss et al. (1977) study. Carcinogenic
8-25
-------
effects of chloroform have been shown in the liver and kidney, whereas the lung
was apparently not a target organ in the Theiss et al. (1977) study and in
other studies,
8.1.5. Evaluation of Chloroform Carcinogenicity by Reuber (1979). Reuber
(1979) evaluated the carcinogenicity of chloroform based on his review of slides
in the NCI (1976) bioassay, and his review of data in other carcinogenicity
studies described in this document. Reuber concurred with reported findings of
rat kidney tumors and mouse hepatocel lular carcinomas in the NCI (1976) study,
mouse hepatomas in the Eschenbrenner and Miller (1945) and Rudali (1967) studies,
and mouse kidney tumors in the Roe et al . (1979) study. However, Reuber also
concluded that there were treatment-related neoplasms in the NCI study in addition
to those reported. In rats, Reuber concluded that chloroform treatment induced
liver tumors (hepatocellular carcinomas and neoplastic nodules) and cholangio-
fibromas and cholangiocarcinomas in addition to kidney tumors. Besides hepato-
cellular carcinomas, malignant lymphoma was also concluded by Reuber to have
been induced by chloroform treatment in mice. Reuber also noted that treated
rats and mice did not exhibit liver cirrhosis, that treated rats with thyroid
tumors generally did not have liver or kidney tumors, and that liver necrosis
was apparent only in high-dose female mice. The differences in histopathologic
interpretation of tissue specimens in the NCI bioassay between the Reuber study
and the NCI report, outside of a difference of opinion between pathologists,
are not clear.
8.1.6. Oral Administration (Drinking Water): Mouse: Promotion of Experimental
Tumors
8-26
-------
8.1.6.1. CAPEL ET AL. (1979) — The effect of chloroform on the growth of
murine tumors was assessed by Capel et al. (1979). Redistilled analar chloroform
was used, but chemical analysis of the product was not indicated. Test animals
were male C57CL/105cSn/01a and male Theiller-Original (TO) mice, 20 to 22 g body
weight. A cage of 20 mice drank 80 to 100 ml of water each day; hence, chloroform
was added to yield doses of 0.15 or 15 mg CHCl3Ag/day for two dose groups, with
each mouse drinking 4 ml water per day. Fresh chloroform solution was given
daily and was protected from light.
In one experiment, the authors stated that 100 TO mice in each dose group
were divided into three "approximately equal" subgroups. One subgroup
(pretreated) was treated with chloroform for 14 days before and after inoculation
of Ehrlich ascites tumor cells. Another subgroup (post-treated) was given
chloroform only after inoculation of tumor cells. The third subgroup, also
inoculated with tumor cells, served as untreated controls. Tumor cells had
been maintained in the peritoneal cavity of male TO mice by weekly passage of
10^ cells. Peritoneal fluid was collected 7 days after inoculation of
cells and diluted with buffered saline. All mice in the three subgroups were
given intraperitoneal injections of 0.1 ml diluent (10^ cells). At the end
of exponential growth at 10 days following inoculation of cells, animals were
sacrificed for removal of peritoneal fluid. The peritoneal cavity was washed
with heparinized buffered saline. Fluid and washings were combined and diluted
with buffered saline. Cells were disrupted by sonication for estimation of
DNA levels per ml cell suspension as a measure of total cell content.
A second experiment was done in which 100 C57BL mice in each dose group
were subdivided into three subgroups (pretreated, post-treated, control) each
of which was treated with chloroform according to the protocol in the first
8-27
-------
experiment. Each mouse received a subcutaneous injection of 10^ B16 melanoma
cells suspended in 0.1 ml buffered saline. Inoculum was obtained from a C57BL
mouse which had received a transplant of syngeneic 816 melanoma cells maintained
by intramuscular passage every 14 days. Animals were sacrificed at 21 days
after inoculation, and spleen, mesenteric lymph nodes, and lungs were examined
for metastases.
In the third experiment, Lewis lung tumor cells were maintained by
serial intramuscular transplantation in C57BL mice. A group of 100 mice
was divided into three approximately equal subgroups (pretreated, post-treated,
control) to investigate the effect of 15 mg CHCl3Ag/day on tumor growth and
spread according to the protocol used in the first experiment. Each mouse
received intramuscular injections of 2 x 10^ cells suspended in 0.1 ml buffered
saline into a thigh. Animals were killed 14 days after administration of
tumor cells, and both the tumor-bearing and the normal thighs were skinned and
severed at the knee and hip. Tumor weight was estimated as the difference
between the weights of the thighs. Pulmonary tumor foci were also counted.
For estimation of the effect of 0.15 mg CHCl3Ag/day in the third experi-
ment, 100 mice were divided into subgroups of 20 animals each and were pretreated
with chloroform before (for 8, 6, 4, or 2 weeks) and after injection of the
Lewis cells. Mice were sacrificed at 16 days after inoculation of tumor cells,
and tumor weights and numbers of lung foci were determined. In these animals,
homogenates of primary tumors were prepared for B-glucuronidase estimation
and protein content.
The results of these experiments are summarized in Tables 8-8, 8-9, and
8-10. Body weights and survival were not affected by chloroform treatment.
Ehrlich ascites tumor cells, as equated with DNA content, were significantly
(P < 0.05) increased in high-dose mice, and slightly though not significantly
8-28
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TABLE 8-8. EFFECT OF ORAL CHLOROFORM INGESTION
(Cape! et al.
)N THE GROWTH OF EHRLICH ASCITES TUMORS
1979)
CO
I
Dose
0. 15 mg /kg /day
15 mg/kg/day
Treatment
group
Control
Post-treated
Pretreated
Control
Post-treated
Pretreated
Number of
animals per
group
33
33
33
43
37
30
Average body
weight (g)a
38.3 + 3.7
39.4 + 2.9
37.9 _+ 3.2
39.4 + 3.4
37.5 + 3.0
37.0 + 3.9
Tumor DNA
(ug/ml )a
661 + 222
724 + 254
770 _+ 283
637 + 221
1143 + 324
827 + 245
Si gnif icant
NS
NS
P < 0.001
P < 0.001
mean + S.D.
NS = Not significant; P > 0.05.
-------
TABLE 8-9. EFFECT OF ORAL CHLOROFORM INGESTION ON METASTATIC "TUMOR TAKES" WITH B15 MELANOMA
(Capel et al. 1979)
CO
CO
o
Dose
0. 15 mg /kg /day
15 mg/kg/day
Treatment
Control
Post-treated
Pretreated
Control
Post-treated
Pre-treated
Number of
animals per group
26
31
28
30
32
32
Animals
Spleen
15
35
36
15
31
31
with B16 melanoma invasion in organs (%)
Mesenteri c
lymph nodes
13
10
29
12
25
32
Lung
(a)a
12
10
18
6
19
20
(b)b
3
5
10
4
6
20
aNumbers in column (a) refer to the percentage of animals with tumor foci on the lungs.
^Numbers in column (b) refer to the average number of lung metastases.
-------
TABLE 8-10. EFFECT OF ORAL CHLOROFORM INGESTION ON THE GROWTH AND SPREAD OF THE LEWIS LUNG TUMORa
(Capel et al. 1979)
Number of Average Tumor
animals per body weight weight
Dose Treatment group
0.15 mg/ 8d
kg /day 6
4
CO o
1 L.
- 0
(control )
15 mg/ Control
kg /day Post-treated
Pretreated
20
20
20
20
20
33
33
33
(g)
30.6 +
30.6 +
29.0 +
29.0 +
29.3 +_
23.6 +
24.5 T
24.4 +_
— — : — f — R
3.8
3.8
3.6
3.5
2.7
1.3
2.3
2.2
(g)
3.5 + 0.81
3.3 + 0.72
3.3 + 0.54
3.2 + 0.72
3.1 +_ 0.11
1.6 + 0.31
1.7 + 0.51
1.8 + 0.12
Lung
metastases
165 + 56
170 + 41
154 + 39
147 + 44
142 _+ 34
44 + 26
57 + 19
61 + 19
B-glucuronidase
Si gni ficance
NS
NS
NS
NS
P < 0. 05
P < 0.01
acti vityb
0.33 +
0.27 +
0.38 +
0.49 +
0.58 +
0.56
0.79
0.073
0.070
0.094
Protei
n
contentc
78.2
66.8
60.1
60.3
50.8
+
+
+
T
T
4.2
2.7
5.1
4.7
6.2
1 » ^- -J 'Jl I V «J \~ f\ ^S \ *_ .J ,J ^_ V4 lull V^ Lrlll, \\t\f \A I I J ' *J % U/ •
^Expressed as mole product/ng protein/fain.
cMi11igrams of protein after extraction mg/g wet weight,
^Duration of treatment (weeks).
NS = not significant, P > 0.05.
-------
increased in low-dose animals. Invasion by B16 melanoma cells, especially in
the spleen, was augmented by both doses, and the numbers of lung foci were
also greater in both treatment groups. Metastasis of Lewis cells was increased
only by treatment with 15 mg CHCl3/kg« There was no change in B-glucuronidase
levels based on tumor protein content in the low-dose group. The increased
tumor protein levels appear to reflect tumor growth which was not evident by
weighing.
The study by Capel et al. (1979) shows an ability of chloroform to enhance
the growth of three types of murine tumors in mice. A dose of 15 mg CHC13A9
was effective in each experiment, whereas a dose of 0.15 mg CHCl3Ag was effective
only in the test with 1316 melanoma cells. Although this study does not evaluate
the ability of chloroform to induce primary tumors, it does give evidence for a
promoting effect of chloroform on the growth and spread of experimental tumors
at low doses. However, the mechanism by which chloroform enhanced tumor growth
in the study by Capel et al. (1979) is not certain, and the relevance of this
study to the evaluation of the carcinogenic potential of chloroform is not
clear.
8.2. CELL TRANSFORMATION ASSAY
8.2.1. Styles (1979). Styles (1979) reported an investigation on chloroform
in a cell transformation system with BHK cells, using growth in semi-solid agar
as an endpoint, as part of a larger study (Purchase et al. 1978) done to screen
chemicals for carcinogenic potential. The BHK-agar transformation assay tech-
nique used has been described by Styles (1977) and Purchase et al. (1978).
In the study reported by Styles (1979), baby Syrian hamster kidney (BHK-21/C1 13)
cells were exposed to five different doses of test substance in vitro in serum-
8-32
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free liquid tissue culture medium in the presence of rat liver microsomal
fraction and cofactors (S-9 mix; Ames et al. 1975). The liver microsomal
fraction was obtained from Sprague-Dawley rats induced with Arochlor 1254.
Cells were grown and maintained in Dulbecco's modification of Eagle's
medium in an atmosphere of 20% COg in air. Cells were maintained at 37°C until
confluent, and then were trypsinized and resuspended in fresh growth medium.
Resuspended cells were grown until 90% confluent for transformation assays or
100% confluent for stock. Only cells with .normal morphology were used for
assays. To minimize spontaneous transformation frequency, cells were obtained
at low passage, grown to 90% confluency, and frozen in liquid nitrogen. Cells
were thawed at 37°C in growth medium for further use.
Test compounds were dissolved in DMSO or water as appropriate. Each dose
was tested in replicate assays. Cells incubated until 90% confluency were
trypsinized and resuspended in Medium 199 at a concentration of 10^ cells/foil.
Resuspended cells (10^) were incubated with test chemical and S-9 mix at 37°C
for 4 hours. After treatment, cells were centrifuged and resuspended in growth
medium containing 0.3% agar. Survival after treatment was estimated by incubating
1,000 cells at 37°C for 6 to 8 days before counting colonies. Transformation
was evaluated by counting colonies after cells were plated and incubated for 21
days at 37°C. The dose-response for transformation was compared with that for
survival. Styles (1977) accepted a fivefold increase in transformation frequency
above control values at the 1X50 as a positive result. The spontaneous trans-
formation frequency of BHK cells (72 experiments) in this study was 50 _+ 16 per
10° survivors. The suitability of the soft agar medium for colony growth was
checked by assays with polyoma-transformed BHK-21/C1 13 cells or Hela cells.
Cell transformation results were negative with exposure to chloroform so-
lution in DMSO added to culture medium in a dose range that included levels at
8-33
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which toxicity was observed (Figure 8-2). Although chloroform doses high
enough to produce toxicity did not induce transformation, exposure of cells to
chloroform as a vapor could have provided a comparison of the transformation
potential of chloroform as a vapor and chloroform in liquid solution.
The study by Purchase et al. (1978), which was done on 120 chemicals of
various classes, showed that the BHK-agar transformation assay system was about
90% accurate in discriminating between compounds with demonstrated carcinogenic
or noncarcinogenic activity, and was in approximately 83% agreement with the
results of assays done by the authors with S. typhimurium (TA 1535, TA 1538,
TA 98, TA 100). Styles (1979) indicated, without presenting numerical data,
that the results obtained in Salmonella assays on chloroform in liquid solution
were similar to the findings of the transformation assays. Purchase et al.
(1978) also observed that metabolically activated agents transformed BHK cells
more strongly in the presence of S-9 mix, thus suggesting that BHK cells have
limited intrinsic metabolic capability.
8.3 EPIDEMIOtOGIC STUDIES
In the last decade there has appeared in the literature a host of epi-
demiologic and statistical studies of cancer and exposure to the constituents
of drinking water, of which chloroform is one (Harris 1974, Page et al. 1976,
Tdrone and Gart 1975, Buncher 1975, Vasilenko and Magno 1975, De Rouen and Diem
1975, McCabe 1975, Kruse 1977, Alavanja et al. 1978, Rafferty 1979, Kuzma et al.
1977, Harris et al. 1977, Salg 1977, Mah et al. 1977, Brenniman et al. 1978,
Tuthill et al. 1979, Wilkins 1978). These studies have been subjected to
several critical reviews (Wilkins et al. 1979, U.S. Environmental Protection
8-34
-------
g 100
DC
D
50
0
1100
c/)
tr
§ 900
>
DC
2 700
QC
uu
Q_
500
DC
O
DC
h-
300
100 -
0
CHCU
0.25 2.5 25 250
CONCENTRATION (//I/ml)
ALSO AMES-VE
2500
Figure 8-2. Negative result in transformation assay of chloroform which
was also negative in the Ames assay. (Styles 1979)
8-35
-------
Agency 1979, National Academy of Sciences 1978) and have been discussed in some
detail. Some very general relationships have been noted by the reviewers. Of
particular importance is the appearance of some consistency in the finding of
cancer of the large intestine, rectum, and bladder associated with the consti-
tuents of drinking water.
It must be emphasized that none of the studies discussed in this section
implicates chloroform directly as the sole or dominant constituent of drinking
water responsible for the excess of cancer at these sites. Over 300 volatile
organic contaminants have been identified in drinking water, and many of these
have been identified as carcinogens (Wilkins et al . 1979).
However, chloroform at a peak concentration of 266 ug/1 has been shown to
exceed peak concentrations of other detected carcinogens by levels 37 times
higher than those of the next highest carcinogen, vinyl chloride (Wilkins et al .
1979).
Chloroform measurements appear to range largely between 1 and 112
ug/1, according to a survey of 76 drinking water supplies (Cantor et al . 1978).
Although a direct association cannot be made, the possibility still exists
that since chloroform is apparently the predominant component in chlorinated
drinking water, it could be a contributing factor in the etiology of the cancer
associated with the consumption of drinking water.
Almost all of the above-referenced studies were ecological correlation
investigations, and only a few utilized case-control methods. The studies
varied by sample size, cancer sites considered, control variables, and the types
of endpoints used as indicators. Among the problems posed by the data in these
studies are the following: 1) a lack of data measuring the quantity of chlorine
and chloroform in drinking water; 2) the limited nature of recently acquired
data on the quality and quantity of organics in drinking water; 3) the limited
8-36
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amount of information given regarding personal consumption of drinking water;
4) the long latency periods associated with most cancers (current cancer rates
reflect exposures received decades earlier); and 5) the demographic effects of
migration, which adds another dimension of difficulty to the quantification of
personal consumption of drinking water over time.
Since publication of the three reviews referred to above, several additional
studies of cancer and exposure to trihalomethanes have been published. The
following pages discuss each of these studies in detail.
8.3.1. Young et al. (1981). Young et al. (1981) conducted a case-control study
in which cancer deaths in 8,029 white females were matched with non-cancer deaths
in some 8,029 white females for county of residence, year of death, and age
recorded on death certificates in 28 counties in the State of Wisconsin from
1972 through 1977. Information about the chlorine content of the drinking
water of the 16,058 cases and controls was derived from mail-back questionnaires
recently submitted to the superintendents of 202 waterworks encompassing the
counties sampled. The questions pertained to prechlorination and postchlorina-
tion dosages used over the past 20 years (average daily dose in ppm). For 14%
of the sample who were not served by a waterworks, decedents were assigned
chlorine dosages of zero. The assignment was on the basis of water supplied to
decedent's usual place of residence.
Odds ratios were calculated from a logistic regression model. This model
provided estimates of the relative risk of site-specific cancer deaths for
exposure of the previous 20 years to high, medium, and low chlorine doses, as
compared with no chlorination. Urbanicity, marital status, and site-specific
high-risk occupation were controlled in the model. Only colon cancer showed a
significant (P < 0.05) association with chlorine intake in all three dosage
8-37
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categories. However, no gradient of increasing risk with increasing dosage was
apparent. For the high, medium, and low dosage categories, the odds ratios
were 1.51, 1.53, and 1.53, respectively. All were significant at P < 0.05. In
those counties where the drinking water supplies were exposed to rural runoff,
the odds ratios for colon cancer increased to 3.43, 3.68, and 2.94 for high,
medium, and low average daily chlorine doses when controlled for water source
depth and purification. These were statistically significant at the P = 0.025
level. Colon cancer mortality was not related to chlorination in counties not
exposed to rural runoff. This finding is consistent with the hypothesis that
trihalomethanes are formed through the action of chlorine on organic substances
in drinking water.
Nonsignificant risks were evident at the remaining sites, i.e., esophagus,
stomach, rectum, liver, pancreas, kidney, bladder, lung, brain, and breast. The
average daily chlorine dose categories were designated by the authors as follows:
none (less than 0.01 ppm), low (0.01-0.99 ppm), medium (1.00-1.70 ppm), and high
(1.71-7.00 ppm).
The authors made a number of assumptions regarding exposure of subjects
and controls to chloroform. They assumed that chlorine in drinking water would
represent a good surrogate for exposure of cases and controls to chloroform,
reasoning that trihalomethanes such as chloroform are believed to result from
the reaction of chlorine with naturally occurring organics in water. Although
drinking water at the tap was not analyzed for chloroform or other trihalomethanes,
the authors assumed that the measured levels of chlorine at the respective water-
works would correlate well with presumed exposure to chloroform in drinking water.
Such implicit assumptions appear questionable for several reasons. First,
the latent period for the development of several, if not most, of the cancer sites
8-38
-------
is most probably greater than 20 years. This is longer than the period covered
by the exposure data on chlorination of water supplies used by the authors.
Second, migration within and around the 28-county area could have masked
any real risk that was related to exposure. A diagnosis of colon cancer, which
has a 5-year survival rate of better than 46%, could have induced victims to
migrate to urban areas (where chlorine levels were higher) in order to obtain
better medical care, thus leading to a false positive association.
Third, the amount of chloroform that is formed from the addition of
chlorine is a function of several important variables: the quantity of organics
in the water supply, treatment practices, and chlorine dosages. The quantity of
organics in the water supply is, in turn, determined by the nature of the water
supply source. Surface water (rivers and streams) receives large quantities of
organics from land runoff, whereas groundwater contains little or no organic
material; hence, the likelihood of chloroform formation from the addition of
chlorine to a groundwater supply is minimal.
Fourth, liquids intake rates and amounts vary considerably from person to
person. It is clear that most people satisfy their liquid requirements through
a variety of drinks besides tap water, (e.g., milk, orange juice, coffee,
soda). It is conceivable that many may drink little water because of these
competing sources of liquid refreshment. Therefore, it is probable that many
persons who were ranked as having been exposed to chloroform may in fact have
had little exposure to it. The resulting miselassification of cases and controls
by exposure category would tend to mask any gradient of increasing risk with
exposure if one existed.
Another possibly confounding variable not controlled for in this study 1s
the dietary intake of meat and foods low in fiber content (Reddy et al. 1980),
both of which have been hypothesized as being related to colon or rectal cancer.
8-39
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The dietary intake of such foods, however, is not known to be correlated with
the quantity of chlorine in drinking water, although the possiblity of a spurious
correlation cannot be ruled out. In more urbanized counties where chlorine
levels are higher, residents may consume a diet of more meat and less fiber.
In summary, a definite association of chlorine or chloroform in drinking
water with an increased risk of colon cancer should not be made for the reasons
stated.
8.3.2. Hogan et al. (1979). Hogan et al. (1979) conducted an ecological study
of site-specific cancer rates based on National Cancer Institute (NCI) cancer
mortality data by county for the years between 1950 and 1969 (Mason and McKay
1974) and on chloroform levels in finished drinking water, as determined by the
U.S. Environmental Protection Agency (EPA) in two separate surveys (U.S. EPA
1975). The first survey, known as the National Organics Reconnaissance Survey
(NORS), consisted of samples from 80 water treatment facilities across the
country. The second survey covered 83 utilities in the states of Illinois,
Indiana, Michigan, Minnesota, Ohio, and Wisconsin. Linear multiple regression
analyses were done for each set of data separately. The dependent variable was
county site-specific cancer mortality. Weighted and unweighted regression
coefficients were determined for a number of independent variables selected by
the author based on a study by Hoover et al. (1976). A variety of demographic
characteristics related to cancer mortality were used in addition to the variable
"chloroform levels" as determined from the NORS and regional surveys to explain
cancer mortality. These characteristics were: county population density, per-
cent of urbanization per county, percent of nonwhite people, percent of foreign-
born, county median family income, educational level, percent of workforce
employed in manufacturing, chloroform level in drinking water samples, and
8-40
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county population. According to the authors, the weighting was based on the
inverse of the square root of the population of the race-sex county stratum,
and was done chiefly to improve the precision of the regression estimates.
Significant positive statistical correlations were found between chloroform
levels in treated drinking water and cancer mortality specific for bladder,
rectum, and large intestine in the "weighted" regression for white females. On
the other hand, only stomach cancer appeared to be positively correlated sig-
nificantly with chloroform levels in white males. Without weighting, cancers
of the bladder, rectum, thyroid, and breasts were significantly correlated with
chloroform levels in white females. In white males, cancers of the pancreas
and rectum were significantly correlated with chloroform without weighting.
Only estimated regression coefficients were provided with their corresponding P
values. The study contained no information on actual levels of chloroform
observed in drinking water. Nonwhites were not considered because of the small
sizes of the populations from which rates were derived.
Ecological studies such as this one are necessarily weak because their infor-
mation is based on aggregate rather than individual data. The evidence for an
association is indirect and definite conclusions cannot be drawn, although
hypotheses may be formulated. It is not certain whether a multiple linear
regression technique is the proper method for analyzing such data, since the
assumption of linearity implied in its selection may not be warranted. Also,
since the model contains no interaction terms, it is implicit that the chosen
control variables are independent of each other, and such an assumption may also
be unwarranted. Furthermore, as was mentioned in the Young et al. (1981)
study, these data are weakened because it was assumed that the subjects were
actually exposed to the levels of chlorine (or chloroform) indicated. Another
limitation is that since the chloroform data were collected in 1975, the more
8-41
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relevant exposure data (assuming a general cancer latency of 1U to 30 years)
should be those of 1920 to 1959, given that the site-specific cancer mortality
data covered the period 1950-1969.
8.3.3. Cantor et al. (1978). Cantor et al . (1978), in an ecological study of
cancer mortality and halomethanes in drinking water, used age-standardized cancer
mortality rates by site and sex in whites for the years 1968-1971, but only in
the 923 U.S. counties that were more than 50% urban in 1970. This study was
similar to the Hogan et al. (1979) study with respect to its design; i.e., a
weighted linear regression model was used with sex- and site-specific cancer
rates as the dependent variable. The weight was directly proportional to the
square root of the counties' person-years at risk and thus inversely proportional
to the standard deviation of the estimated mortality rate. Chloroform (CHC13),
bromochloromethane (BTHM), and total trihalomethane (THM) levels were obtained
from the two EPA surveys (U.S. EPA 1975) used in the Hogan et al . study.
Demographic variables used in the regression model on a county-wide basis were:
percent of urbanization (1970); median school years completed by persons over age
25; population size (ratio of 1970 to 1950 population); percentage of the work
force in manufacturing; and percentage of foreign-born. Although a predicted,
age-adjusted, site-specific cancer rate was calculated for each county based on
this regression technique, only the data for 76 counties, where more than half
of the population of the counties was served by a sampled water supply, were
actually used in this correlation analysis of THM levels with residual mortality
rates. Figure 8-3 gives a frequency distribution of the chloroform levels in these
76 U.S. drinking water supplies. The three indicators, chloroform, bromochloro-
methane, and total trihalomethane, were highly correlated with one another.
Positive nonsignificant correlations with THM levels were evident with
8-42
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25
20 -
u
z
UJ
cc
DC
D
U
O
O
u.
O
o
LU
CC
15 -
10 —
0
—
1
0.001
, 1
I
0.01 0.1 1.0
1
100
5 —
MICROMOLES CHCI3/LITER
Figure 8-3. Frequency distribution of CHC13 levels in 76 U.S. drinking
water supplies. The abscissa is linear in the logarithm of
the level. (Cantor et al. 1978)
8-43
-------
respect to several forms of cancer, including lymphoma and kidney cancer
in males (Table 8-11). But according to the authors, bladder cancer mortality
rates gave the strongest and most consistent association with THM exposure
after controlling for differences in social class, ethnic group, urbanicity,
region, and extent of county industrialization (Table 8-12). However, the
association appeared to be greatest with respect to BTHM and not chloroform.
The corresponding correlations for chloroform were positive but nonsignificant.
The authors report that although other sites appeared to be positively correlated
with THM levels, the inconsistencies "outweigh the consistencies," thus casting
doubt on the reliability of these correlation coefficents; i.e., the direction
anc} strength of the correlations bear little relationship to the percent of
population served by treated drinking water and/or by region.
TABLE 8-11. CORRELATION COEFFICIENTS BETWEEN RESIDUAL MORTALITY RATES IN
WHITE MALES AND THM LEVELS IN DRINKING WATER BY REGION AND BY
PERCENT OF THE COUNTY POPULATION SERVED IN THE UNITED STATES
(Cantor et al. 1978)
Site of THM
cancer Indicator
Kidney CHC13
Lymphoma BTHM
(non-
Hodgkins )
Kidney CHC1 3
Lymphoma BTHM
(non-
Hodgkins)
Correlation
coefficients
for regions
• North South Mountain Pacific
0.11 -0.
(0.54)a (o.
0.06 0.
(0.74) (0.
Correlati on
the percent
50-64%
-0.16
(0.44)
-0.33
(0.11)
11 0
73) (0
08 0
79) (0
coefficients
of the popul
65-84%
-0.11
(0.60)
-0.19
(0.36)
.66
.11)
.05
.92)
for counties
ation served
85-100%
0.42
(0.04)
0.36
(0.08)
of the U.S.
All regions
0.14
(0.33)
0.06
(0.70)
in which
was:
50-100%
0.07
(0.55)
-0.08
(0.81)
aP value for two-tai1edt-test is snown in parentheses.
8-44
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TABLE 8-12. CORRELATION COEFFICIENTS BETWEEN BLADDER CANCER MORTALITY RATES BY
SEX AND BTHM LEVELS IN DRINKING WATER BY REGION OF THE UNITED STATES
(Cantor et al. 1978)
Correlation coefficients by region
Bladder cancer
Number
Male white
Female white
North
31
0.529
(0.002)
0.30
(0.11)
South
13
0.04
(0.90)
0.20
(0.51)
Mountain
7
-0.02
(0.96)
0.63
(0.13)
Total
51
0.30
(0.03)
0.33
(0.02)
aP value for two-tailed t-test is shown in parentheses. Counties with at
least 65% of their populations served by one water supply were included in
this analysis.
The authors noted an association of kidney cancer with chloroform exposure
that was restricted to males, but was significant only in counties where at least
85% of the public was served by treated drinking water. In counties where less
than 85% was served by treated drinking water, the correlation coefficients
were actually negative. Combining all counties with greater than 50% served by
treated drinking water, the correlation coefficent was nonsignificant and close
to zero. One interesting observation was that without controlling for ethnicity,
the authors found a "fairly strong" association of THM levels with colon cancer
and lung cancer rates in both sexes, and even a dose-response relationship
between these tumor sites and the proportion of the population exposed. However,
when ethnicity was added to the regression model, these relationships disappeared.
Again, this is a descriptive study from which hypotheses can be
formulated only for future in-depth study. It cannot be concluded that even the
significant positive correlations in the study indicate any evidence of real
associations. As the authors point out, potential sources of error (i.e.,
8-45
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control of confounders such as cigarette smoking and diet) are particularly
difficult since no direct information is available on the individuals studied.
The main problem with such studies, as mentioned earlier, is that the data are
aggregate rather than individual. Such data frequently include large numbers
of individuals who never received the exposure in question. Associations
derived from such data may be misleading and are often unreliable.
8.3.4. Gottlieb et al. (1981). Gottlieb et al. (1981) completed a case-control
study of the relationship between Mississippi River drinking water and the risk
of rectum and colon cancer. The study was based on mortality data gathered from
20 parishes in southern Louisiana. Rectal and colon cancer deaths (692 and 1,167,
respectively) from 1969 to 1975 were matched one-to-one to non-cancer deaths by
age at death, year of death, sex, and race, within the same parish group. A
parish group consisted of similar parishes with respect to industrial and urban-
rural characteristics and were defined so that each parish included nearly
equal populations using groundwater and surface water supply sources, based on
information from the 1970 census.
Three different estimators of exposure were used. The first, "source!ife,"
is defined as follows: "mostly surface" (birth and death in a surface-water-using
parish); "some surface" (some known surface water use at birth or death); "pos-
sible surface" (death in a groundwater parish but had either unknown or out-of-
state birthplace); and "least surface" (birth and death in a groundwater-using
parish). Length of residence was also considered, if known and for more than
10 years. The second index used was chlorine level (none, low [less than 1.09
ppm], or high [greater than 1.08 ppm]). The third index was the level of
organics in the drinking water (low [less than 68 ppm] and high [greater than
or equal to 68 ppm]). Sourcelife could be determined for 99.2% of the entire
8-46
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group of 3,718 cases and controls, but 51% had no data for length of residence
or had lengths of residence of under 10 years. For those with lengths of
residence of less than 10 years, water sources during the possible carcinogenic
period were unknown. Chlorine values were available for some 78.9% of the
3,718 sources, while organics levels were available for only 50.1% of the sources.
The analyses using the latter two variables were equivocal, possibly due to the
lack of information on these parameters.
Colon cancer was found not to be related significantly to any water variable,
although the number of colon cancer cases in this study (1,167) was greater than
the number of rectum cancer cases (692). The authors hint that the earlier
correlation found in ecological studies could have resulted from confounding
with urban lifestyles. Rectal cancer, on the other hand, was found to be
significantly elevated with respect to surface or Mississippi River water
consumption. Based on sourcelife, the odds ratio for rectal cancer for those
who were born and died using groundwater sources was 2.07 (95% confidence interval
[C.I.] 1.49-2.88) based on a multidimensional contingency table analysis.
Chlorination was significantly associated with rectal cancer, and for those who
used river water, the risk decreased as the distance from the mouth of the
river increased. The odds ratio for cancer of the rectum at a location below
New Orleans versus one above the city was 1.82 (95% C.I. 1.01-3.26). The authors
noted that both sexes were at increased risk. With respect to controlling for
the effect of chlorination where adequate numbers existed, the surface water versus
groundwater effect on rectal cancer was of only borderline significance (P =
0.05), implying a chlorine effect.
With respect to levels of organics, information was available for over 48%
of the rectal cancer group and their controls. The odds ratio calculated based
on these data was nonsignificant (Table 8-13), but was probably subject to some
8-47
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TABLE 8-13. RISK OF MORTALITY FROM CANCER OF THE RECTUM ASSOCIATED
WITH LEVELS OF ORGANICS IN DRINKING WATER
(Gottlieb et al. 1981)
High (J> 68 ppm)
Low (< 68 ppm)
Total
Odds ratio
Cases
110
232
342
Controls
97
220
317
1.08
bias with respect to availability of exposure data as a function of date of death.
With respect to colon cancer, the authors felt that since they had grouped
the parishes according to industry and urban characteristics (matching was done
within the parish group), they successfully eliminated urban lifestyle as a
confounder in their evaluation of colon cancer and drinking water.
The results of this study suggest that cancer of the rectum is linked to
the consumption of surface water, and since chlorination appears to be an effect
modifier altering the risk ratio to only borderline significance, it would
seem that chlorination does contribute to the risk of rectal cancer.
8.3.5. Alavanja et al. (1978). Alavanja et al. (1978) reported on a case-
control study of 3,446 gastrointestinal and urinary tract cancer deaths
(1,595 females and 1,851 males) occurring during a 3-year period from 1/1/68
to 12/31/70 in seven counties of New York State. Some 3,444 individually
matched noncancer deaths were also selected. Independent variables were:
residence in an urban or rural area, residence in an area served by chlorinated
or nonchlorinated water, residence in an area served by surface water or ground-
8-48
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water, and occupation. Cases were taken from computer tapes of New York State
death certificates, and were individually matched with an equal number of
non-cancer deaths for the same year. Matching variables were age, race, sex,
foreign- versus United States-born, and county of usual residence. If potentially
confounding variables could not be controlled via the matching process, the
cases and controls were stratified by these confounding variables. The data
were analyzed by the chi-square test. A statistically significant excess of
gastrointestinal and urinary tract cancer mortality occurred among women in the
urban county of Erie (odds ratio [OR] = 2.08), with nonsignificant excesses
in Schenectady County (OR = 2.98) and Allegany County (OR = 4.13). Likewise,
among men a statistically significant excess of gastrointestinal and urinary
tract cancer mortality occurred in Erie County (OR = 2.15) and Rensselaer
County (OR = 1.98), and a nonsignificant excess occurred in Schenectady County
(OR = 1.96) and Allegany County (OR = 2.85). Although the study encompassed a
seven-county area, almost two-thirds of the deaths occurred in Erie County.
The combined overall odds of dying from gastrointestinal and urinary tract
cancer for all seven counties combined (including Erie), were only 1.79 based
on 3,446 cases, whereas in Erie County alone they were 3.15 based on 2,177 cases.
The authors concluded that males and females residing in the chlorinated water
areas of the counties noted above were at a greater risk of gastrointestinal
and urinary tract cancer mortality not due to age, race, ethnic distribution,
urbanicity, occupation, inorganic carcinogens (Cd, As, Be, Pb, Ni, N03), or
surface/groundwater difference. No environmental data are provided, however,
to characterize quantities of chlorine (or chloroform) exposure. "Inadequate
water quality data" prevented the authors from making a "definitive claim that
the process of chlorination is directly or indirectly responsible for the
8-49
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greater risk of gastrointestinal and urinary tract cancer mortality" in
chlorinated cancer areas. No description is given of how residence was
classified into chlorinated versus nonchlorinated water areas or surface water
versus groundwater areas through the use of water distribution maps, a practice
which can result in misclassification on the basis of exposure. Again, because
of the lack of individual dosage data on chloroform exposure and the low signifi-
cance of the risks described, this study can only be regarded as suggestive for
gastrointestinal and urinary tract cancer mortality.
8.3.6. Brenniman et al. (1978). Brenniman et al. (1978) attempted to confirm
the findings of Alavanja et al. (1978) in a case-control study of gastrointestinal
and urinary tract cancer mortality among whites in 70 Illinois communities using
both chlorinated and nonchlorinated groundwater. The authors limited the study to
groundwater because of the possible introduction of confounding effects due to
agricultural runoff and industrial sewage in surface water. The 3,208 cases and
43,666 controls used were those of Illinois deaths occurring between 1973 and
1976. Controls were selected from a pool of non-cancer deaths after the elimina-
tion of certain special types of deaths, such as perinatal deaths.
Chlorinated groundwater communities were matched with nonchlorinated
groundwater communities that were similar with respect to urbanicity and
Standard Hetropolitan Statistical Area (SMSA) description. To ensure a minimum
follow-up period, water supplies were categorized as chlorinated or nonchlorinated
according to a "1963 inventory of municipal water facilities." Additionally,
questionnaires were sent to water treatment plants in the communities to verify
the 1963 data. The beginning dates for chlorination were obtained for many of
the plants. Based on an EPA survey, it was found that 14 chlorinated groundwater
8-50
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supply sources in Illinois had chloroform concentrations ranging from less than
1 ug/1 to 50 ug/1, with a mean concentration of 10.8 ug/1.
In females, statistically significant increased relative risks of cancer
of the large intestine and rectum (OR = 1.19, P £ 0.05), as well as total
digestive tract cancer (excluding liver) (OR = 1.15, P £0.05), were found for
chlorinated versus nonchlori nated Illinois groundwater supplies. With respect
to total gastrointestinal and urinary tract cancer, the risk was significantly
increased in females living within standard metropolitan statistical areas (OR
= 1.28, P <_ 0.025) and within urban areas (OR = 1.24, P £ 0.025) between
chlorinated and nonchlorinated groundwater communities.
Where evidence was available concerning a history of chlorination, the
authors noted that the relative risk of total gastrointestinal and urinary
tract cancer tended to increase with time from initial chlorination, although
the change was small. The greatest increase occurred in urban nonstandard
metropolitan areas (OR = 1.14 if chlorinated since 1963 and nonsignificant, but
OR = 1.28 if chlorinated since 1953 and significant, P £0.025). Although
several significant findings were observed in this study, it is surprising that
the authors so readily dismissed the results of their own study on the basis that
confounding factors such as diet, smoking, and occupation were not controlled.
These authors felt that the findings were tenuous and did not confirm the findings
of Alavanja et al. (1978) either in strength or in consistency. They state that
"chlorination of groundwater does not seem to be a major factor in the the
etiology of site-specific gastrointestinal and urinary tract cancers."
8.3.7. Struba (1979). Struba (1979), as part of his Ph.D. thesis, completed a
case-control study of mortality in North Carolina on individuals who died at
age 45 or under during the period 1975-1978. The cancer sites studied were the
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rectum, colon, and urinary bladder. Between 700 to 1,500 cases per site were
matched with controls by age, race, sex, and geoeconomic region (coastal,
piedmont, or mountain). Non-cancer deaths were excluded if cancer was listed
as a contributory or underlying cause of death. For colon and rectal cancer,
certain precancerous colonic disorders were excluded (ulcerative colitis,
familial polyposis, and adenomatous polyposis). Water data were classified by
source, treatment, and previous use. "Source" was defined as ground, surface
uncontaminated, or uncontaminated by upstream pollution. "Treatment" was
defined as none, prechlorinated, post-chlorinated, or both. "Previous use"
included the following 15 categories of upstream pollution for contaminated
water only:
(1) Tobacco manufacturing
(2) Textile manufacturing
(3) Textile bleaching and dyeing
(4) Furniture manufacturing
(5) Pulp and paper mills
(6) Chemical industries
(7) Petroleum refining
(8) Rubber and plastics manufacturing
(9) Leather tanning and finishing
(10) Abrasives, asbestos, minerals
(11) Primary metals industries
(12) Electroplating
(13) Electric power generation
(14) Urban areas > 50,000
(15) Out-of-state upstream discharges
The author found small but significant odds ratios (1.3 to 2.0) for all
three sites in rural areas, as well as significant odds ratios for each of the
water quality variables in many stratified or combined analyses. Odds ratios for
urban areas (population over 10,000) were generally not significant. Urbanization
was shown to be an effect modifier for colon cancer and a likely confounder for
rectal and bladder cancers. The author considered socioeconomic status to be a
likely confounder for cancer of the rectum and bladder. Multivariate analyses
showed no evidence that occupation acted as a confounder for bladder cancer in
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this study. To estimate migration effects, cases and controls were stratified
by place of birth and death (birth and death in the same county; birth and
death in North Carolina; and death in North Carolina, birth unspecified) and
substratified by region, age, race, sex, and urbanization. Odds ratios for
treatment (chlorinated and nonchlorinated) were computed for all of the strata.
For all three cancer sites, the group with least migratory influence had the
highest odds ratio, thus lending support to the author's supposition that an
increasing migratory effect is associated with a decreasing risk of cancer of
all three sites.
Additionally, Struba found an increasing gradient of risk from the coastal
regions of North Carolina to the mountains, a finding that he maintains is
consistent with a stronger contrast between surface water and water from deep
wells than between surface water and water from shallow wells, which are known
to be susceptible to contamination by surface water seepage into groundwater
aquifers. However, the author notes that this difference could be due to
differences in water treatment practices or confounding by uncontrolled factors
such as dietary habits or lifestyles.
8.3.8. Discussion. These later ecological and case-control studies of chlorine
exposure and cancer risk from water supplies consistently support the finding
of an increased risk of bladder, colon, and rectal cancer from exposure to
chlorinated water. This association is at best weak, although significant, as
evidenced by odds risk ratios that range up to 3.6 in the Young et al. (1981)
study, but generally fall between 1.1 and 2.0 (see Table 8-14) in the remaining
case-control studies. The risk ratios derived in these studies could be
explained by the confounding effects of uncontrolled influences such as smoking,
diet, air pollution, occupation, and lifestyle. However, the consistency of
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TABLE 8-14. CANCER RISK ODDS RATIOS AND 95% CONFIDENCE INTERVALS
(CHLORINATED VERSUS UNCHLORINATED)
(Crump and Guess 1982)
Site
Rectum
Alavanja
et al.
1978a
1.93
(1.32, 2.83)
Brenniman
et al.
1978b
1.26 (crude)
(0.98, 1.61)
Young
et al.
1981°
1.39 high
(0.67, 2.86)
Gottlieb
et al.
1981d>e
1.41
(1.07, 1.87)
Struba
1979d,e
1.53
(1.24, 1.
89)
1.22 (ad-
justed)
1.16 medium
(0.58, 2.32)
1.13 low
(0.61, 2.08)
Colon 1.61
(1.28, 2.03)
1.08 (crude)
(0.96, 1.22)
1.1 (ad-
justed
1.51 high 1.05
(1.06, 2.14) (0.95, 1.18)
1. 53 medium
(1.08, 2.00)
1.30
(1.13, 1.50)
1.53 low
(1.11, 2.11)
Bladder
1.69
(1.11, 2.56)
1.04 (crude) 1.04 high
1.07
1.54
0.98 (ad- 1.03 medium
justed) (0.42, 2.54)
1.06 low
(0.60,
3.09)
Calculated for both sexes and all races combined. Confidence intervals
were not stated in Alavanja et al. (1978). Crump (1979) calculated them by
applying the method of Fleiss (1979) to data in Alavanja et al. (1978).
^Calculated for Caucasians of both sexes. Adjusted values were adjusted for
age, sex, urban/rural, and SMSA/nonSMSA. Confidence intervals were not stated
in the original report. Crump (1979) calculated them by applying the method
of Fleiss (1979) to data on total cases and total controls supplied by Dr.
Brenniman in personal communication.
Calculated for white females and for high, medium, and low
chlorine doses compared with no chlorination. Odds ratios
intervals computed by logistic regression, controlling for urbanization,
marital status, and site-specific occupation.
Calculated for both sexes and all races combined.
eStruba and Gottlieb et al. also computed odds ratios for surface water versus
groundwater as follows. Struba: rectum 1.55 (1.26, 1.91); colon 1.27
(1.10, 1.46); bladder 1.48 (1.22, 1.80). Gottlieb et al.: rectum 1.51 (1.21,
1.90); colon 0.95 (0.88, 1.03).
average daily
and confidence
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the finding across several independent and diverse study groups supports the
finding of a definite risk. Of course, all of the case-control studies use
residence data and cause-of-death information from death certificates, and
thus are not strictly incidence studies. Bias can creep in from several
sources: differential survivorship rates due to proximity to better medical
care and treatment facilities, higher socioeconomic status, and the possibility
of migration of newly diagnosed cancer patients to major medical care centers
where chlorination is used to a greater extent. Underestimates of risk can
result from failure to control for migration before to diagnosis, misclassifica-
tion of cause of death, and use of chlorination as a surrogate variable in
place of more direct measurements of chloroform, especially if the chlorinated
source contains few organic contaminants. Hence, the association is weak but
significant with regard to the three cancer types and exposure to chlorinated
drinking water. Since exposure to chlorine in water is not the same as exposure
to chloroform, the most that can be said is that there is a suggestion of an
increased risk of cancer of these three sites from exposure to chloroform. If
this risk truly exists, it may be due to an intermediate in the natural synthesis
of chloroform (communication with Dr. Kenneth P. Cantor, NCI).
In summary, it appears that there may be a suggestion of an increased risk
of certain forms of cancer (bladder, large intestine, and rectum) due to the
presence of tri'halomethanes in drinking water. Beyond this, little more can be
said. The significant excess risk of colon cancer from chlorine in drinking
water does not constitute evidence of an association of colon/rectal cancer with
chloroform. The statistically significant positive correlation of bladder can-
cer and BTHM levels in drinking water is not readily attributable to chloroform.
The evidence of a significant association of kidney cancer with chloroform
exposure in drinking water is even more questionable, since it was based on
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the findings of only one study, which was confined to males residing in counties
where more than 85% of the population was served by treated drinking water. A
statistically positive correlation was seen only in males residing in counties
with over 85% treated drinking water. No association was observed in females
in these same counties, and the correlations were actually negative for both
males and females in counties with less than 85% treated drinking water.
It appears that these case-control and ecological studies in humans suggest
a weak association of certain forms of cancer with trihalomethanes and chloroform
in drinking water, but further epidemiologic research should be performed to
confirm these findings.
8.4. QUANTITATIVE ESTIMATION
This section evaluates the unit risk for chloroform in air and water and
the potency of chloroform relative to other carcinogens that the Carcinogen
Assessment Group (CAG) has evaluated. The unit risk is defined as the lifetime
cancer risk to humans from daily exposure to a concentration of 1 ug/1 in water
by ingestion or daily exposure to 1 ug/m-^ in air by inhalation. If the unit
risk is calculated from a model that is linear at low doses, then the unit risk
could be used as the slope for calculating risk at low doses.
The unit risk estimate for chloroform represents an extrapolation below the
dose range of experimental data. There is currently no solid scientific basis
for any mathematical extrapolation model that relates exposure to cancer risk
at extremely low concentrations, including the unit concentration given above.
For practical reasons, such low levels of risk cannot be measured directly, either
by animal experiments or by epidemiologic studies. Low-dose extrapolation must
therefore be based on current understanding of the mechanisms of carcinogenesis.
8-56
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At the present time, the dominant view of the carcinogenic process involves the
concept that most cancer-causing agents also cause irreversible damage to DNA.
This position is based in part on the fact that a very large proportion of agents
that cause cancer are also mutagenic. There is reason to expect that the quantal
response, which is characteristic of mutagenesis, is associated with a linear non-
threshold dose-response relationship. Indeed, there is substantial evidence from
mutagenicity studies with both ionizing radiation and a wide variety of chemicals
that this type of dose-response model is the appropriate one to use. This is
particularly true at the lower end of the dose-response curve; at higher doses,
there can be an upward curvature, probably reflecting the effects of multistage
processes on the mutagenic response. The linear non-threshold dose-response
relationship is also consistent with the relatively few epidemiologic studies
of cancer responses to specific agents that contain enough information to make
the evaluation possible (e.g., radiation-induced breast and thyroid cancer, skin
cancer induced by arsenic in drinking water, liver cancer induced by aflatoxin
in the diet). Some supporting evidence also exists from animal experiments (e.g.,
the initiation stage of the two-stage carcinogenesis model in rat liver and mouse
skin). Linearity is also supported when the mode of action of the carcinogen
in question is similar to that of the background cancer production in the exposed
population.
Because its scientific basis, although limited, is the best of any of the
current mathematical extrapolation models, the linear non-threshold model has
been adopted as the primary basis for risk extrapolation to low levels of the
dose-response relationship. The risk estimates made with such a model should
be regarded as conservative, representing the plausible upper limits for the
risk; i.e., the true risk is not likely to be higher than the estimate, but it
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could be lower. For several reasons, the unit risk estimate based on animal
bioassays is only an approximate indication of the risk in populations exposed
to known carcinogen concentrations. First, there are important species differences
in uptake, metabolism, and organ distribution of carcinogens, as well as species
differences in target site susceptibility. Second, the concept of equivalent
doses for humans compared with animals on a mg/surface area basis is virtually
without experimental verification as to carcinogenic response. Finally, human
populations are variable with respect to genetic constitution and diet, living
environment, activity patterns, and other cultural factors.
The unit risk estimate can give a rough indication of the relative potency
of a given agent compared with other carcinogens. The comparative potency of
different agents is more reliable when the comparison is based on studies in
the same test species, strain, and sex, and by the same route of exposure.
The quantitative aspect of the carcinogen risk assessment is included here
because it may be of use in the regulatory decision-making process (setting
regulatory priorities, evaluating the adequacy of technology-based controls, etc).
However, it should be recognized that the estimation of cancer risks to humans
at low levels of exposure is uncertain. At best, the linear extrapolation
model used here provides a rough but plausible estimate of the upper limit of
risk; i.e., it is not likely that the true risk would be much more than the
estimated risk, but it could very well be considerably lower. The risk estimates
presented in subsequent sections should not be regarded as an accurate representa-
tion of the true cancer risks even when the exposures are accurately defined.
The estimates presented may be factored into regulatory decisions to the extent
that the concept of upper risk limits is found to be useful.
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8.4.1. Procedures for the Determination of Unit Risk.
8.4.1.1. LOW-DOSE EXTRAPOLATION MODEL -- The mathematical formulation
chosen to describe the linear nonthreshold dose-response relationship at low
doses is the linearized multistage model. This model employs enough arbitrary
constants to be able to fit almost any monotonical ly increasing dose-response
data, and it incorporates a procedure for estimating the largest possible
linear slope (in the 95% confidence limit sense) at low extrapolated doses that
is consistent with the data at all dose levels of the experiment.
Let P(d) represent the lifetime risk (probability) of cancer at dose d. The
multistage model has the form
P(d) = 1 - exp [-q0 + qjd + qxd2 + ... + qkdk)]
where
qi _> 0, i = 0, 1, 2, ..., k
Equivalently,
Pt(d) = 1 - exp [qjd + q2d2 + ... + qkdk)]
where
Pt(d) = P(d) - P(0)
1 - P(0)
is the extra risk over background rate at dose d.
The point estimate of the coefficients q-j, i = 0, 1, 2, ..., k, and con-
sequently, the extra risk function, Pt(d), at any given dose d, is calculated
by maximizing the likelihood function of the data.
The point estimate and the 95% upper confidence limit of the extra risk,
Pt(d), are calculated by using the computer program GLOBAL79, developed by
Crump and Watson (1979). At low doses, upper 95% confidence limits on the
8-59
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extra risk and lower 95% confidence limits on the dose producing a given risk
are determined from a 95% upper confidence limit, q*, on parameter q . Whenever
qi > 0, at low doses the extra risk P-t(d) has approximately the form Pt(d) =
q* x d. Therefore, q* x d is a 95% upper confidence limit on the extra risk
and R/q* is a 95% lower confidence limit on the dose, producing an extra risk
of K. Let Lg be the maximum value of the log-likelihood function. The upper-
limit q* is calculated by increasing q^ to a value q* such that when the log-
likelihood is remaximized subject to this fixed value q* for the linear
coefficient, the resulting maximum value of the log-likelihood LI satisfies the
equation
2 (L0 - L!) = 2.70554
where 2.70554 is the cumulative 90% point of the chi-square distribution with one
degree of freedom, which corresponds to a 95% upper limit (one-sided). This
approach of computing the upper confidence limit for the extra risk P^(d) is an
improvement on the Crump et al. (1977) model. The upper confidence limit for
the extra risk calculated at low doses is always linear. This is conceptually
consistent with the linear non-threshold concept discussed earlier. The slope,
q*, is taken as an upper bound of the potency of the chemical in inducing cancer
at low doses. [In the section calculating the risk estimates, Pt(d) will be
abbreviated as P.]
In fitting the dose-response model, the number of terms in the polynomial
is chosen equal to (h-1), where h is the number of dose groups in the experiment,
including the control group.
Whenever the multistage model does not fit the data sufficiently well, data
at the highest dose are deleted and the model is refit to the rest of the data.
This is continued until an acceptable fit to the data is obtained. To determine
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whether a fit is acceptable, the chi -square statistic
X =
is calculated where N^ is the number of animals in the i-t dose group, X^ is
the number of animals in the ith dose group with a tumor response, P^ is the
probability of a response in the itn dose group estimated by fitting the multi-
stage model to the data, and h is the number of remaining groups. The fit is
determined to be unacceptable whenever X^ is larger than the cumulative 99%
point of the chi-square distribution with f degrees of freedom, where f equals
the number of dose groups minus the number of non-zero multistage coefficients.
8.4.1.2. SELECTION OF DATA — For some chemicals, several studies in
different animal species, strains, and sexes, each run at several doses and
different routes of exposure, are available. A choice must be made as to which
of the data sets from several studies to use in the model. It may also be
appropriate to correct for metabolism differences between species and for
absorption factors via different routes of administration. The procedures used
in evaluating these data are consistent with the approach of making a maximum-
likely risk estimate. They are as follows:
1. The tumor incidence data are separated according to organ sites or tumor
types. The set of data (i.e., dose and tumor incidence) used in the model is the
set where the incidence is statistically significantly higher than the control for
at least one test dose level and/or where the tumor incidence rate shows a statis-
tically significant trend with respect to dose level. The data set that gives
the highest estimate of the lifetime carcinogenic risk, q*, is selected in most
cases. However, efforts are made to exclude data sets that produce spuriously
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high risk estimates because of a small number of animals. That is, if two
sets of data show a similar dose-response relationship and one has a very small
sample size, the set of data having the larger sample size is selected for
calculating the carcinogenic potency.
2. If there are two or more data sets of comparable size that are identical
with respect to species, strain, sex, and tumor sites, the geometric mean of q*,
estimated from each of these data sets, is used for risk assessment. The geo-
metric mean of numbers Aj, A2, ..., Am is defined as
x A2 x ... x AJ
1/fo
3. If two or more significant tumor sites are observed in the same study,
and if the data are available, the number of animals with at least one of the
specific tumor sites under consideration is used as incidence data in the model.
8.4.1.3. CALCULATION OF HUMAN EQUIVALENT DOSAGES -- Following the sugges-
tion of Mantel and Schneiderman (1975), it is assumed that mg/surface area/day
is an equivalent dose between species. Since, to a close approximation, the
surface area is proportional to the two-thirds power of the weight, es would be
the case for a perfect sphere, the exposure in mg/day per two-thirds power of
the weight is also considered to be equivalent exposure. In an animal experiment,
this equivalent dose is computed in the following manner.
Let
Le = duration of experiment
le = duration of exposure
m = average dose per day in mg during administration of the agent (i.e.,
during le), and
W = average weight of the experimental animal
8-62
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Then, the lifetime exposure is
I e x m
Le x W
2/3
8.4.1.3.1. Oral -- Often exposures are not given in units of nig/day, and
it becomes necessary to convert the given exposures into mg/day. Similarly, in
drinking water studies, exposure is expressed as ppm in the water. For example,
in most feeding studies exposure is given in terms of ppm in the diet. In these
cases, the exposure in mg/day is
m = ppm x F x r
where ppm is parts per million of the carcinogenic agent in the diet or water, F
is the weight of the food or water consumed per day in kg, and r is the absorption
fraction. In the absence of any data to the contrary, r is assumed to be equal
to one. For a uniform diet, the weight of the food consumed is proportional to
the calories required, which in turn is proportional to the surface area, or two-
thirds power of the weight. Water demands are also assumed to be proportional to
the surface area, so that
m « ppm x
or
m a ppm.
o /o
rW2/3
As a result, ppm in the diet or water is often assumed to be an equivalent expo
sure between species. However, this is not justified for the present study,
since the ratio of calories to food weight is very different in the diet of man
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as compared with laboratory animals, primarily due to differences in the moisture
content of the foods eaten. For the same reason, the amount of drinking water
required by each species also differs. It is therefore necessary to use an
empirically derived factor, f = F/W, which is the fraction of an organism's body
weight that is consumed per day as food, expressed as follows:
Fraction of body
weight consumed as
Species W ^food ^water
Man
Rats
Mice
70
0.35
0.03
0.028
0.05
0.13
0.029
0.078
0.17
Thus, when the exposure is given as a certain dietary or water concentration in
ppm, the exposure in mg/W2/3 is
* F =
x f x W = ppm x f x w1/3
_
rW2/3 w2/3 w2/3
When exposure is given in terms of mg /kg/day = m/Wr = s, the conversion is
simply
JD _ = s x W1/3
rW2/3
8.4.1.3.2. Inhalation -- When exposure is via inhalation, the calculation
of dose can be considered for two cases where 1) the carcinogenic agent is
either a completely water-soluble gas or an aerosol and is absorbed proportion-
ally to the amount of air breathed in, and 2) the carcinogen is a poorly water-
soluble gas which reaches an equilibrium between the air breathed and the body
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compartments. After equilibrium is reached, the rate of absorption of these
agents is expected to be proportional to the metabolic rate, which in turn is
proportional to the rate of oxygen consumption, which in turn is a function of
surface area.
8.4.1.3.2.1. Case 1. Agents that are in the form of particulate matter
or virtually completely absorbed gases, such as sulfur dioxide, can reasonably
be expected to be absorbed proportionally to the breathing rate. In this case
the exposure in mg/day may be expressed as
m = I x v x r
where I = inhalation rate per day in m3, v = mg/m3 of the agent in air, and
r = the absorption fraction.
The inhalation rates, I, for various species can be calculated from the
observations of the Federation of American Societies for Experimental Biology
(FASEB 1974) that 25-g mice breathe 34.5 1/day and 113-g rats breathe 105
1/day. For mice and rats of other weights, W (in kilograms), the surface area
proportionality can be used to find breathing rates in m3/day as follows:
For mice, I = 0.0345 (W/0.025)2/3 m3/day
For rats, I = 0.105 (W/0.113)2/3 rn3/day
For humans, the value of 30 m3/day* is adopted as a standard breathing rate
(International Commission on Radiological Protection 1977). The equivalent
exposure in mg/W2/3 for these agents can be derived from the air intake data in a
*From "Recommendation of the International Commission on Radiological
Protection." page 9. The average breathing rate is 10^ cm3 per 8-hour workday
and 2 x 10' cm3 in 24 hours.
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way analogous to the food intake data. The empirical factors for the air intake
per kg per day, i = I/W, based on the previously stated relationships, are
tabulated as follows:
Species W i = I/Id
Man
Rats
Mice
70
0.35
0.03
0.29
0.64
1.3
Therefore, for particulates or completely absorbed gases, the equivalent exposure
in mg/W^/3 -js
d = m = Ivr = iUvr = -jwl/3vr
w2/3 w2/3 w2/3
In the absence of experimental information or a sound theoretical argument
to the contrary, the fraction absorbed, r, is assumed to be the same for all
species.
8.4.1.3.2.2. Case 2. The dose in mg/day of partially soluble vapors is
proportional to the Q^ consumption, which in turn is proportional to W*-/3 and
is also proportional to the solubility of the gas in body fluids, which can be
expressed as an absorption coefficient, r, for the gas. Therefore, expressing
the 02 consumption as 02 = k W2/3, where k is a constant independent of species,
it follows that
m = kW2/3xvxr
or
d = m = kvr
W2/3
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As with Case 1, in the absence of experimental information or a sound theoretical
argument to the contrary, the absorption fraction, r, is assumed to be the same
for all species. Therefore, for these substances a certain concentration in ppm
or ug^n3 in experimental animals is equivalent to the same concentration in
humans. This is supported by the observation that the minimum alveolar concen-
tration necessary to produce a given "stage" of anesthesia is similar in man
and animals (Dripps et al. 1977). When the animals are exposed via the oral
route and human exposure is via inhalation or vice versa, the assumption is
made, unless there is pharmacokinetic evidence to the contrary, that absorption
is equal by either exposure route.
8.4.1.4. CALCULATION OF THE UNIT RISK FROM ANIMAL STUDIES — The risk
associated with d mgAg^/3/day is obtained from GLOBAL79 and, for most cases of
interest to risk assessment, can be adequately approximated by P(d) = 1 - exp
(-q*d). A "unit risk" in units X is simply the risk corresponding to an exposure
of X = 1. This value is estimated by finding the number of mg/kg2/3/day that
corresponds to one unit of X, and substituting this value into the above
relationship. Thus, for example, if X is in units of ug/m3 in the air, then for
case (1), d = 0.29 x 701/3 x 10~3 mgAg2/3/day, and for case (2), d = 1, when
ug/fa3 is the unit used to compute parameters in animal experiments.
If exposures are given in terms of ppm in air, the following calculation
may be used:
1 ppm = 1.2 x mo1ecu1ar weight (gas) mg/m3
molecular weight (air)
Note that an equivalent method of calculating unit risk would be to use mg/kg
for the animal exposures, and then to increase the j^*1 polynomial coefficient by
8-67
-------
an amount
(wh/wa)J/3 j = l, 2, .... k,
and to use mgAg equivalents for the unit risk values.
8.4.1.4.1. Adjustments for Less Than Lifespan Duration of Experiment --
If the duration of experiment Le is less than the natural lifespan of the test
animal L, the slope q*, or more generally the exponent g(d), is increased by
1 o
multiplying a factor (L/Le) . We assume that if the average dose d is continued,
the age-specific rate of cancer will continue to increase as a constant function
of the background rate. The age-specific rates for humans increase at least by
the third power of the age and often by a considerably higher power, as demon-
strated by Doll (1971). Thus, it is expected that the cumulative tumor rate
would increase by at least the third power of age. Using this fact, it is assumed
that the slope q*, or more generally the exponent g(d), would also increase by
at least the third power of age. As a result, if the slope q* [or g(d)] is
calculated at age Le, it is expected 1 that if the experiment had been continued
for the full lifespan L at the given average exposure, the slope q* [or g(d)]
O
would have been increased by at least (L/Le) .
This adjustment is conceptually consistent with the proportional hazard
model proposed by Cox (1972) and the time-to-tumor model considered by Daffer et
al. (1980), where the probability of cancer by age t and at dose d is given by
P(d,t) = 1 - exp [-f(t) x g(d)]
8.4.1.5. MODEL FOR ESTIMATION OF UNIT RISK BASED ON HUMAN DATA — If human
epidemiologic studies and sufficiently valid exposure information are available
8-68
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for the compound, they are always used in some way. If they show a carcinogenic
effect, the data are analyzed to give an estimate of the linear dependence of
cancer rates on lifetime average dose, which is equivalent to the factor B^.
If they show no carcinogenic effect when positive animal evidence is available,
then it is assumed that a risk does exist, but is smaller than could have been
observed in the epidemiologic study, and an upper limit to the cancer incidence
is calculated assuming hypothetically that the true incidence is below the level
of detection in the cohort studied, which is determined largely by the cohort
size. Whenever possible, human data are used in preference to animal bioassay
data.
Very little information exists that can be used to extrapolate from high-
exposure occupational studies to exposures at low environmental levels. However,
if a number of simplifying assumptions are made, it is possible to construct a
crude dose-response model whose parameters can be estimated using vital
statistics, epidemiologic studies, and estimates of worker exposures.
In human studies, the response is measured in terms of the relative risk
of the exposed cohort of individuals as compared with the control group. The
mathematical model employed for the low-dose extrapolation assumes that for low
exposures the lifetime probability of death from cancer, PQ, may be represented
by the linear equation
P0 = A + BHx
where A is the lifetime probability in the absence of the agent, and x is the
average lifetime exposure to environmental levels in units such as ppm. The
factor BH is the increased probability of cancer associated with each unit
increase of x, the agent in air.
8-69
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If it is assumed that R, the relative risk of cancer for exposed workers
as compared with the general population, is independent of length of exposure
or age at exposure and depends only on average lifetime exposure, it follows
that
R = JL = A + BH (XT = x?)
P0 A + BH x 1
or
RP0 = A + BH (xi + x2)
where xj = lifetime average daily exposure to the agent for the general popula-
tion, X2 = lifetime average daily exposure to the agent in the occupational
setting, and PQ = lifetime probability of dying of cancer with no or negligible
exposure.
Substituting PQ = A + BH xj and rearranging gives
BH = P0 (R - 1)A2
To use this model, estimates of R and X2 must be obtained from epiderr.i ologi c
studies. The value PQ is derived by means of the life table methodology from
the age- and cause-specific death rates for the general population found in
U.S. vital statistics tables.
8.4.2. Unit Risk Estimates.
8.4.2.1. DATA AVAILABLE FOR POTENCY CALCULATION — Evidences of carcino-
genic activity of chloroform from lifetime treatment studies in laboratory
animals include: 1) significantly (P < 0.05) increased incidences of hepato-
8-70
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cellular carcinomas in female and male B6C3F1 mice (Table 8-15) and kidney
tumors in male Osborne-Mendel rats (Table 8-16) in an NCI (1976) bioassay, where
animals were given chloroform in corn oil by gavage; and 2) kidney tumors in
male ICI mice given chloroform in arachis oil by gavage (Roe et al. 1979, Table
8-17). These data sets are used to estimate the carcinogenic potency of chloro-
form using the linearized multistage model and the dose conversion procedure
as described previously. For comparison, three other low-dose extrapolation
models, the probit, Weibull, and one-hit, are also used to calculate the carcino-
genic potency of chloroform.
TABLE 8-15. INCIDENCE OF HEPATOCELLULAR CARCINOMAS IN FEMALE AND MALE B6C3F1 MICE
(NCI 1976)
Human (animal)
dose (mg/kg/day)a Incidence rate
Female 0 0/20 (0%)
11.59 (238) 36/45 (80%)
23.23 (477) 39/41 (95%)
Male 0 1/18 (6%)
6.72 (138) 18/50 (36%)
13.49 (277) 44/45 (98%)
aHurnan equivalent dose is calculated by d x (5/7) x (78/90) x (0.034/70)1/3
= 4.87 x 10-2 x d, where d is the animal dose given 5 days per week for 78
weeks (out of a lifespan of 90 weeks). The body weights are assumed to be
34 g for mice and 70 kg for humans.
3-71
-------
TABLE 8-16. INCIDENCE OF TUBULAR-CELL ADENOCARCINOMAS IN MALE OSBORNE-MENDEL RATS
(NCI 1976)
Human (animal)
dose (mg/kg/day)a Incidence rate
0 0/19
8.62 (90) 2/50
17.24 (180) 10/50
aHuman equivalent dose is calculated by d x (78/104) x (5/7) x (0.4/70)1/3
= 9.58 x 10~2 x d, where d is the animal dose given 5 days per week for 78
weeks (out of a lifespan of 104 weeks). The body weights are assumed to be
400 g for rats and 70 kg for humans.
TABLE 8-17. INCIDENCE OF MALIGNANT KIDNEY TUMORS IN MALE ICI MICE
(Roe et al. 1979)
Human (animal)
dose (mg/kg/day)a Incidence rates
0 0/50
3.79 (60) 9/48
dHuman equivalent dose is calculated by d x (6/7) x (80/90) x (0.04/70)1/3
= 6.32 x 10~2 x d, where d is the animal dose given 6 days per week for 80
weeks (out of a lifespan of 90 weeks). The body weights are assumed to be 400
g for mice and 70 kg for humans.
8-72
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8.4.2.2. CHOICE OF LOW-DOSE EXTRAPOLATION MODELS — In addition to the
multistage model currently used by the CAG for low-dose extrapolation, three
more models, the probit, the Weibull, and the one-hit models, are used for
comparison (Appendix A). These models cover almost the entire spectrum of risk
estimates that could be generated from existing mathematical extrapolation
models. Generally statistical in character, these models are not derived from
biological arguments, except for the multistage model, which has been used to
support the somatic mutation hypothesis of carcinogenesis (Armitage and Doll
1954, Whittemore 1978, Whittemore and Keller 1978). The main difference among
these models is the rate at which the response function P(d) approaches zero or
P(0) as dose d decreases. For instance, the probit model would usually predict
a smaller risk at low doses than the multistage model because of the difference
of the decreasing rate in the low-dose region. However, it should be noted
that one could always artificially give the multistage model the same (or even
greater) rate of decrease as the probit model by making some dose transformation
or by assuming that some of the parameters in the multistage model are zero.
This, of course, would not be reasonable if the carcinogenic process for the
agent were not known a priori. Although the multistage model appears to be
the most reasonable or at least the most general model for low-dose extrapola-
tion, the point estimate generated from this model is of limited value because
the shape of the dose-response curve beyond the experimental exposure levels
remains in question. Furthermore, point estimates at low doses extrapolated
beyond the experimental doses could be extremely unstable and could differ
drastically, depending on the amount of lowest experimental dose. Since upper-
bound estimates from the multistage model at low doses are relatively more
stable than point estimates, it is suggested that the upper-bound estimate for
the risk (or the lower-bound estimate for the dose) be used in evaluating the
8-73
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carcinogenic potency of a suspect carcinogen. The upper-bound estimate can be
taken as a plausible estimate if the true dose-response curve is actually
linear at low doses. The upper-bound estimate means that the risks are not
likely to be higher, but could be lower if the compound has a concave upward
dose-response curve or a threshold at low doses. Another reason one can, at
best, obtain an upper-bound estimate of the risk when animal data are used is
that the estimated risk is only a probability conditional on the assumption
that an animal carcinogen is also a human carcinogen. Therefore, in reality,
the actual risk could range from a value near zero to an upper-bound estimate.
8.4.2.3. CALCULATION OF THE CARCINOGENIC POTENCY OF CHLOROFORM ~ Using
the incidence data in Tables 8-15 to 8-17 and the corresponding human equivalent
doses, the maximum likelihood estimates of the parameters were calculated for
each of the four models referred to above (see Table A-l in Appendix A). These
models can be used to calculate either point estimates of risk at a given dose,
or the virtually safe dose for a given level of risk. The upper-bound estimates
of the risk at 1 rng/kg/day, calculated from each of these models on the basis
of different data sets, are presented in Table 8-18. From this table, it is
observed that the multistage model predicts a comparable risk on the basis of
four different data sets, while the probit and Weibull models are very unstable
and predict a wide range of risk depending on which data base is used for the
risk calculation. Figures 8-4 and 8-5 present the point and upper-bound estimates
of risk over the low-dose range, calculated from the four models. The carcino-
genic potency of chloroform is represented by the geometric mean of the risk
estimates calculated from the linearized multistage model, on the basis of liver
tumor data for female and male mice. Although the risk calculated from the data
for female mice is greater than that calculated from the data for male mice, the
8-74
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TABLE 8-18. UPPER-BOUND ESTIMATES OF CANCER RISK OF 1 mg/kg/day, CALCULATED BY DIFFERENT MODELS
ON THE BASIS OF DIFFERENT DATA SETS3
CO
I
en
Data base
Liver tumors in female
mice (NCI 1976)
Liver tumors in male mice
(NCI 1976)
Kidney tumors in male rats
(NCI 1976)
Kidney tumors in male mice
(Roe et al. 1979)
Multistage
1.6 x 10-1
3.0 x 10-2
1.3 x 10-2
9.0 x lO-2
Probit Weibull One-hit
1.5 x lO'1 4. 2 x 10-1 1.6 x 10-1
7.3 x ID'12 2.4 x 1Q-3 1.5 x 1Q-1
3.0 x ID'5 1.3 x 10-3 1.6 x 1Q-2
NA NA 9.0 x ID'2
aUpper-bound estimates are calculated by the one-sided 95% confidence limit.
NA = not applicable. Models are not applicable because there is only one dosed group.
-------
0.58
Dose in mg/kg/day
Multistage/One-hit. —•
Probit:
Weibull:
M.L.E. (Maximum likelihood estimate)
upper-bound estimate
M.L.E.
upper-bound estimate
—• upper-bound estimate
Figure 8-4. Point and upper-bound estimates of four dose-response models over
low-dose region on the basis of liver tumor data for female mice.
(NCI 1976)
8-76
-------
0.25
0
c
o
Q.
0
cc.
•| 0.125
0
O)
o
c
o
CD
CJ
n
1 X'
X
/
/ X
xx /
X /
/ /
- // -
x /
X / ^XD
X
X / _DX
-•^ X X0-
x / ^-°
// ^" ^
'^<^^^--
0
Multistage: —
1
Dose in mg/kg/day
M.L.E.
n— — D— — G— upper-bound estimate
0 0 ^ upper-bound estimate
Probit: (Risk is too small to show in the graph)
Weibull: — -- — --- M.L.E.
^r— -*^— — &— — upper-bound estimate
Figure 8-5. Point and upper-bound estimates of four dose-response models over
low-dose region on the basis of liver tumor data for male mice.
(NCI 1976)
8-77
-------
estimates from both data sets are combined because the data for males includes
an observation at a lower dose, and the response at this dose does not appear
to be inconsistent with the female data, if the linear dose-response relation-
ship is assumed.
Thus, the risk at 1 mg/kg/day is
P = (1.6 x 10-1 x 3.0 x 10-2)l/2 = 7 x iQ-2
This number differs little from the geometric mean of the q* (upper-bound
1
of the linear parameter) calculated from the two data sets, and thus is used
herein as the slope for calculating risk at low doses.
8.4.2.4. RISK ASSOCIATED WITH 1 ug/to3 OF CHLOROFORM ,IN AIR -- A paucity
of information presently exists on the retention of inhaled chloroform. The
only available estimate of pulmonary absorption of chloroform is a study by
Lehmann and Hasegawa (1910), who estimated an approximate average of 64.1%
(67.6%, 50.2%, 74.6%) retention of a mean chloroform exposure level of 4,592
ppm breathed by three humans for 20 minutes. The relationship is not certain
between these early data on short exposures to high chloroform levels and data
on pulmonary absorption of chloroform by humans for longer periods at lower
exposure levels. In the absence of additional data, absorption rates of 65%
by inhalation and 100% orally are assumed for the purpose of a unit risk estimate.
Under this assumption, 1 ug/ni3 of chloroform in the air would result in an
absorbed effective dose of 1.7 x 10-4 mg/kg/day or
d - 0.65 x (10-3 mg/m3 x 20 m3/day)/70 kg = 1.7 x 10-4 mg /kg/day
8-78
-------
Therefore, the risk, P, associated with 1 ug/m3 of chloroform in air is
P = 7 x ID'2 x 1.7 x 10-4 = 1 x 10'5
8.4.2.5. RISK ASSOCIATED WITH 1 ug/LITER OF CHLOROFORM IN DRINKING WATER
For drinking water exposure, it is assumed that 100% of the chloroform in
drinking water can be absorbed, and that water intake is 2 1/day. Under these
assumptions, the daily dose from consumption of water containing 1 ug/1
(1 ppb) of chloroform is calculated as follows:
d = 1 ug/1 x 2 I/day x 10~3 mg/ug x 1/70 kg = 2.9 x 10~5 mg/kg/day
Therefore, the risk associated with 1 ug/1 of chloroform in water is
P = 7 x lO'2 x 2.9 x 10-5 = 2 x 10'6
This estimate appears consistent with available epidemiologic data such as
the odd ratios for bladder cancer, which were estimated to range from 1.04 to
1.69 (Table 8-14). According to a survey of 76 water supply systems in the
United States, the chloroform measurements ranged from 1 ug/1 to 112 ug/1.
A rough estimate of the cancer risk on the basis of these statistics ranges
from:
B = (1.04 - 1) x 7 x 10-4/112 = 3 x 10-7/(ug/1 )
to
B = (1.69 - 1) x 7 x 10-4/1 = 5 x 10-4/(Ug/l )
where 7 x ID'4 is the estimated background bladder cancer mortality rate in the
United States.
8-79
-------
8.4.3. Comparison of Potency with Other Compounds. One of the uses of the
quantitative potency estimate is to compare the relative potencies of carcinogens.
Figure 8-6 is a histogram representing the frequency distribution of potency
indices for 53 suspect carcinogens evaluated by the CAG. The actual data
summarized by the histogram are presented in Table 8-19. The potency index is
derived from q*, the 95% upper bound of the linear component in the multistage
model, and is expressed in terms of (mMol /kg/day)"1. Where no human data
were available, animal oral studies were used in preference to animal inhalation
studies, since oral studies constitute the majority of animal studies.
Based on the available data concerning liver tumors in female and male
mice (NCI 1976), the potency index for chloroform has been calculated as 8 x
10°. • This figure is derived by multiplying the slope q* = 7 x 10~2 mgAg/day
and the molecular weight of chloroform, 119.4. This places the potency index
for chloroform in the fourth quartile of the 53 suspect carcinogens evaluated
by the CAG.
The ranking of relative potency indices is subject to the uncertainties
involved in comparing estimates of potency for different chemicals based on
varying routes of exposure in different species, by means of data from studies
whose quality varies widely. All of the indices presented here are based on
estimates of low-dose risk, using linear extrapolation from the observational
range. These indices may not be appropriate for the comparison of potencies if
linearity does not exist at the low-dose range, or if comparison is to be made
at the high-dose range. If the latter is the case, then an index other than
the one calculated above may be more appropriate.
8.4.4. Summary of Quantitative Assessment. Four data sets that contain
sufficient information are used to estimate the carcinogenic potency of
8-80
-------
>
J2>
-------
TABLE 8-19. RELATIVE CARCINOGENIC POTENCIES AMONG 53 CHEMICALS EVALUATED
BY THE CARCINOGEN ASSESSMENT GROUP AS SUSPECT HUMAN CARCINOGENS1.2'3
Compounds
Acryloni tri le
Aflatoxin B^
Aldrin
Allyl Chloride
Arsenic
B[aJP
Benzene
Benzi dine
Beryl 1 ium
Cadmium
Carbon tetrachloride
Chlordane
Chlorinated ethanes
1 ,2-dichloroethane
hexachloroethane
1,1,2,2-tetrachloroethane
1 ,1 ,1-trichloroethane
1 ,1 ,2-trichloroethane
Chloroform
Chromium
DDT
Dichlorobenzi dine
1 , 1-dichloroethy lene
Di el drin
SI ope
(mg/kg/day)"1
0.24(W)
2924
11.4
1.19x10-2
lb(H)
11.5
5.2xlO-2(W)
234(W)
4.86
6.65(W)
1.30x10-1
1.61
6.9x10-2
1.42x10-2
0.20
1.6xlO--3
5.73x10-2
7x10-2
41 (W)
8.42
1.69
1.47x10-1(1)
30.4
Molecular
weight
53.1
312.3
369.4
76.5
149.8
252.3
78
184.2
9
112.4
153.8
409.8
98.9
236.7
167.9
133.4
133.4
119.4
100
354.5
253.1
97
380.9
Potency
i ndex
1x10+1
9xlO+5
4x10+3
9x10-!
2x10+3
3x10+3
4x100
4x10+4
4x10+1
7x10+2
2x10+1
7x10+2
7x10°
3x100
3x10+1
2x10-1
8x100
8x10°
4x10+3
3x10+3
4x10+2
1x10+1
1x10+4
Order of
magnitude
(logic
index )
+ 1
+6
+4
0
+3
+3
+1
+5
+2
+3
+1
+3
+ 1
0
-1
+1
+1
+4
+3
+3
+1
+4
continued on the following page]
8-8Z
-------
TABLE 8-19. (continued)
Compounds
Dini trotol uene
Diphenylhydrazine
Epichlorohydrin
Bis(2-chloroethyl )ether
Bis(chloromethyl )ether
Ethylene dibromide (EDB)
Ethylene oxide
Heptachlor
Hexachlorobenzene
Hexachlorobutadiene
Hexachlorocyclohexane
technical grade
alpha isomer
beta isomer
gamma isomer
Methylene chloride
Nickel
Nitrosamines
Dimethylnitrosamine
Diethylnitrosamine
Dibutylni trosamine
N-ni trosopyrrol i di ne
N-ni troso-N-ethy 1 urea
N-ni troso-N-methyl urea
N-nitroso-diphenylamine
PCBs
Slope
(mg/kg/day)"1
0.31
0.77
9.9x10-3
1.14
9300(1)
8.51
0.63(1)
3.37
1.67
7.75xlO-2
4.75
11.12
1.84
1.33
6.3xlO-4
1.15(W)
25.9(not by q*)
43.5(not by q*)
5.43 1
2.13
32.9
302.6
4.92xlO-3
4.34
Molecular
weight
182
180
92.5
143
115
187.9
44.0
373.3
284.4
261
290.9
290.9
290.9
290.9
84.9
58.7
74.1
102.1
158.2
100.2
117.1
103.1
198
324
Potency
index
6x10+1
1x10+2
9x10-1
2x10+2
1x10+6
2x10+3
3x10+1
1x10+3
5x10+2
2x10+1
1x10+3
3x10+3
5x10+2
4x10+2
5x10-2
7x10+1
2x10+3
4x10+3
9x10+2
2x10+2
4x10+3
3x10+4
IxlQO
1x10+3
Order of
magnitude
(login
index)
+2
+2
0
+2
+6
+3
+1
+3
+3
+1
+3
+3
+3
+3
-1
+2
+3
+4
+3
+2
+4
+4
0
+3
(continued on the following page)
8-83
-------
TABLE 8-19. (continued)
Compounds
Phenols
2,4,6-trichlorophenol
Tetrachlorodioxin
Tetrach loroethylene
Toxaphene
Trichloroethylene
Vinyl chloride
Remarks:
1. Animal slopes are 95%
Slope
(mg/kg/day)"1
1.99xlO-2
4.25xl05
3.5xlO-2
1.13
1.9x10-2
1.75x10-2(1)
upper-limit slop
Molecular
weight
197.4
322
165.8
414
131.4
62.5
es based on the
Potency
index
4x10°
lxlO+8
6x10°
5x10+2
2.5x100
1x10°
linearized multis
Order of
magnitude
(logio
index)
+1
+8
+1
+3
0
0
tage model .
They are calculated based on animal oral studies, except for those indicated by
I (animal inhalation), W (human occupational exposure), and H (human drinking water
exposure). Human slopes are point estimates based on the linear non-tnreshold
model.
2. The potency index is a rounded-off slope in (mMol/kg/day)"1 and is calculated by
multiplying the slopes in (mg/kg/day)"l by the molecular weight of the compound.
3. Not all the carcinogenic potencies presented in this table represent the same
degree of certainty. All are subject to change as new evidence becomes available.
8-84
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chloroform. They are liver tumors in female mice (NCI 1976), liver tumors in
male mice (NCI 1976), kidney tumors in male rats (NCI 1976), and kidney tumors
in male mice (Roe et al. 1979). The unit risks at 1 mg/kg/day, calculated by
the linearized multistage model on the basis of these four data sets, are
comparable. The geometric mean, q* = 7 x 10-2/(mg/kg/day), of the potencies
calculated from liver tumors in male and female mice, is taken to represent
the carcinogenic potency of chloroform. The upper-bound estimate of the cancer
risk due to 1 ug/m3 of chloroform in air is P = 1 x 10-5. The upper-bound
estimate of the cancer risk due to 1 ug/1 in water is P = 2 x 10-6. The
carcinogenic potency of chloroform is in the fourth quartile among the 53 suspect
carcinogens evaluated by the CAG.
The unit risks given above are calculated under the assumption that mg per
unit of body surface area is equivalent between mice and humans. If the dose
in mg/kg/day is assumed to be equivalent, then these unit risks would be reduced
approximately by a factor of 12.
8.5. SUMMARY
8.5.1. Qualitative. Chloroform in corn oil administered at estimated maximally
and one-half maximally tolerated doses by gavage for 78 weeks produced a
statistically significant increase in the incidence of hepatocellular carcinomas
in male and female B6C3F1 mice and renal epithelial tumors (malignant and
benign) in male Osborne-Mendel rats; a carcinogenic response of female Osborne-
Mendel rats to chloroform was not apparent in this study. Use of more than
two doses in these studies might have given a more precise estimate of dose-
response.
8-85
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A statistically significant increase in the incidence of renal tumors
(benign and malignant) was found in another study in male ICI mice treated with
chloroform in either toothpaste or arachis oil by gavage for 80 weeks; however,
treatment with a gavage dose of chloroform in toothpaste for 80 weeks did not
produce a carcinogenic response in female ICI mice and male mice of the CBA,
C57BL, and CF/1 strains. Induction of malignant kidney tumors in male ICI mice
was greater when chloroform was administered, at the same dose, in arachis oil
instead of toothpaste. A carcinogenic response was not observed in male and
female Sprague-Dawley rats given chloroform in toothpaste by gavage for 80
weeks, but early mortality was high in control and treatment groups. Gavage
doses of chloroform in toothpaste did not show a carcinogenic effect in male
and female beagle dogs treated for over 7 years. The results of preliminary
toxicity tests and the carcinogenicity studies suggest that doses of chloroform
in toothpaste given to mice, rats, and dogs in the carcinogenicity studies
approached those maximally tolerated. However, chloroform doses given to mice
and rats in toothpaste or arachis oil were lower than those above given in corn
oi 1.
Hepatomas were found in NIC mice given chloroform in oil by force-feeding
twice weekly for an unspecified period of time, and in female strain A mice given
chloroform in olive oil by gavage once every 4 days for a total of 30 doses at
a level which produced liver necrosis; however, small numbers of animals were
examined for pathology, the duration of these studies was either uncertain or
appeared to be below the lifetime of the animals, and no control group of NIC
mice was apparent. Although a carcinogenic effect of chloroform was not evident
in newborn (C57 x DBA2 - Fl) mice given single or multiple subcutaneous doses
during the initial 8 days of life and observed for their lifetimes, the dose
levels used appeared well below a maximum tolerated dose and the period of
8-86
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treatment after birth was quite short compared to lifetime treatment. Chloroform
was ineffective at maximally tolerated and lower doses in a pulmonary adenoma
bioassay in strain A mice; however, other chemicals that have shown carcinogenic
activity in other tests were also ineffective in this pulmonary adenoma bioassay.
Although an ability of chloroform to promote the growth and spread of Lewis lung
carcinoma, Erlich ascites, and B16 melanoma cells in mice has been shown, the
mechanism by which chloroform produced this effect is uncertain and the relevance
of this study to the evaluation of the carcinogenic potential of chloroform is
presently not clear. Chloroform in liquid solution did not induce transforma-
tion of baby Syrian hamster kidney (BHK - 21/C1 13) cells in vitro at doses
high enough to produce toxicity; additional testing of chloroform as a vapor
could have provided a comparison of cell transformation potential between
chloroform as a vapor and chloroform in a liquid solution.
An additional carcinogenicity study on female B6C3F1 mice and male Osborne-
Mendel rats given chloroform in drinking water over a wide range of dose levels
is in progress at the Stanford Research Institute.
There are no epidemiologic cancer studies dealing with chloroform per se.
Chlorinated drinking water can contain significant amounts of chloroform. There
appears to be a weak but statistically significant risk of cancer of the bladder,
large intestine, and rectum with the presence of chlorine in drinking water.
The odd ratios calculated in the latter ecological and case-control studies
range up to a high of 3.68 for cancer of the colon in the Young et al. study,
but most fell between 1.1 and 2.0. The risk ratios derived in each study
could be explained by the confounding effects of several factors; i.e., smoking,
diet, air pollution, occupation, or lifestyle. However, the consistent finding
of a statistically significant excess of cancer across several independent
and diverse study populations supports the finding of a definite risk. Bias
8-87
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can creep into these studies from differential surrounding rates due to
proximity to better medical care and treatment facilities, higher socioeconomic
status, and the possibility of migration of cancer patients to medical care
facilities in areas where chlorination is used to a greater extent. Under-
estimates of risk may result from failure to control for migration effects
prior to diagnosis, misclassification of cause of death, and use of chlorination
as a surrogate variable for chloroform, especially if few organic contaminants
are in the water. Exposure to chlorinated drinking water will not necessarily
result in exposure to chloroform if organic contaminants are not present.
Many contaminants found in drinking water other than chloroform are carcinogenic,
but they generally are found in much smaller quantities as compared with
chloroform levels found in water sources containing large quantities of
organics. The presence of these other substances, some carcinogenic, makes it
impossible to incriminate chloroform directly as the cause of the excess cancer
at the three sites. Hence, there appears to be an increased risk of cancer of
the bladder, rectum, and large intestine from chlorinated water and, by inference,
from chloroform.
8.5.2. Quantitative. Four data sets that contain sufficient information are
used to estimate the carcinogenic potency of chloroform. They are liver tumors
in female mice (NCI 1976), liver tumors in male mice (NCI 1976), kidney tumors
in male rats (NCI 1976), and kidney tumors in male mice (Roe et al. 1979).
The unit risks at 1 mg/kg/day, calculated by the linearized multistage model
on the basis of these four data sets, are comparable. The geometric mean, q*
= 7 x 10"2/(mgAg/day), of the potencies calculated from liver tumors in
male and female mice, is taken to represent the carcinogenic potency of
chloroform. The upper-bound estimate of the cancer risk due to 1 ug/m3 of
8-88
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chloroform in air is P = 1 x 10"5. The upper-bound estimate of the cancer risk
due to 1 ug/liter in water is P = 2 x 10'6. This estimate appears consistent
with the limited epidemiologic data available for humans.
8.6. CONCLUSIONS
Evidence that chloroform has carcinogenic activity is based on increased
incidences of hepatocellular carcinomas in male and female B6C3F1 mice, renal
epithelial tumors in male Osborne-Mendel rats, kidney tumors in male ICI mice,
and hepatomas in NLC and female strain A mice. As concluded elsewhere in this
document, no definitive conclusions can be reached concerning the mutagenicity
of chloroform based on present evidence. The toxicity of chloroform in liver
and kidney, as noted in this document, is considered to occur through covalent
binding of a reactive metabolic intermediate, possibly phosgene, with cellular
macromolecules; the evidence discussed in the metabolism section herein
indicates that reactive metabolites of chloroform can react extensively with
proteins and lipids, but minimally with nucleic acids. Applying the Interna-
tional Agency for Research on Cancer (IARC) criteria for animal studies, the
level of evidence for carcinogenicity would be sufficient for concluding that
chloroform is carcinogenic in experimental animals.
There are no epidemiologic studies of cancer and chloroform per se. There
appears to be an increased risk of cancer of the bladder, rectum, and large
intestine from chlorinated drinking water and, by inference, from chloroform.
Applying the IARC criteria to assess the carcinogenicity of chloroform in
humans, there is limited evidence for the carcinogenicity of chlorinated
drinking water in humans, and inadequate evidence for the carcinogenicity of
chloroform in humans.
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The overall IARC classification for chloroform is 2B, which by definition
designates chloroform as probably carcinogenic to humans. In the IARC scheme,
Group 2 chemicals are divided into higher (Group A) and lower (Group B) degrees
of evidence depending on whether evidence for their carcinogenicity in humans
is concluded to be limited (Group A) or inadequate (Group B).
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APPENDIX A
COMPARISON AMONG DIFFERENT EXTRAPOLATION MODELS
Four models used for low-dose extrapolation, assuming the independent
background, are:
Multistage: P(d) = 1 - exp [-(q^d + ... + qkdk)]
where q^ are non-negative parameters
A + Bln(d)
Probit: P(d) = / f(x) dx
— 00
where f(.) is the standard normal probability density function
Weibull: P(d) = 1 - exp [-bdk]
where b and k are non-negative parameters
One-hit: P(d) = 1 - exp [-bd]
where b is a non-negative parameter.
The maximum likelihood estimates (MLE) of the parameters in the multistage
and one-hit models are calculated by means of the program GL08AL82, which was
developed by Howe and Crump (1982). The MLE estimates of the parameters in the
probit and Weibull models are calculated by means of the program RISK81, which
was developed by Kovar and Krewski (1981).
Table A-l presents the MLE of parameters in each of the four models that
are applicable to a data set.
A-l
-------
TABLE A-l. MAXIMUM LIKELIHOOD ESTIMATE OF THE PARAMETERS FOR EACH OF THE FOUR EXTRAPOLATION MODELS,
BASED ON DIFFERENT DATA BASES
Data base
Liver tumors in female
mice (NCI 1976)
Liver tumors in male mice
(NCI 1976)
3»
I
rsi
Kidney tumors in male rats
(NCI 1976)
Kidney tumors in male mice
(Roe et al. 1979)
Multistage
qj = 1.35 x
q2 = 0
(q* = 1.7 x
qi = o
q2 = 1.34 x
(q* = 3.0 x
qi = 0
q2 = 7.07 x
(q* = 1.3 x
Probit
10'1 A = -2.03
B = 1.17
ID'1)3
9 A = -7.15
10"^ B = 3.51
ID'2)
. A = -4.58
ID'4 B = 1.31
10-2)
q1 = 5.5 x 10"2 NA
(q* = 9. 1 x 10-2)
Weibull One-hit
b = 1.68 x 10"1 b = 1.35 x 10'1
k = 0.92
b = 7.95 x 10~4 b = 1.21 x 10'1
k = 3.25
b = 2.08 x 10'4 1.56 x 10"2
k = 2.45
NA 5. 5 x 10~2
a q* is the 95% upper-bound confidence limit of the linear parameter in the multistage model,
NA = not applicable. The models are not applicable since there is only one dose group.
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