HEALTH RISKS TO FETUSES, INFANTS AND CHILDREN ( PROPOSED STAGE 2 DISINFECTANT/ DISINFECTION BYPRODUCTS): A REVIEW OST ( Mail Code 4304 T) EPA 822 R 03 010 http:// www. epa. gov/ waterscience/ humanhealth/ March, 2003 iii Table of Contents Table of Contents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . iii LIST OF TABLES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vi ABBREVIATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . viii EXECUTIVE SUMMARY . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ix PLAIN LANGUAGE SUMMARY OF EVALUATION OF CHILDREN'S RISK . . . . . . xxxiv 1. INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 1.1. RISK TO CHILDREN . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 1.2. RISK ASSESSMENT METHODS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6 1.3 DETERMINING RISK TO CHILDREN . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 1.4. MAXIMUM CONTAMINANT LEVEL GOAL AND MAXIMUM RESIDUAL DISINFECTANT LEVEL GOAL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17 2. ESTIMATES OF RISK TO CHILDREN FOR STAGE II DISINFECTANTS/ DISINFECTION BYPRODUCTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25 2.1. CHLORINATED DRINKING WATER . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25 2.1.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . 26 2.1.2. Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40 2.1.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 40 2.2. TRIHALOMETHANES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 2.2.1. Chloroform . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42 <IP_ADDRESS>. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . 42 <IP_ADDRESS>. Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48 <IP_ADDRESS>. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 48 <IP_ADDRESS>. Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53 <IP_ADDRESS>. Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . 55 2.2.2. Brominated Trihalomethanes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 56 <IP_ADDRESS>. Bromodichloromethane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 64 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . 64 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . 77 <IP_ADDRESS>. Dibromochloromethane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 78 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . 78 iv Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 79 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 80 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 83 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . 83 <IP_ADDRESS>. Bromoform . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . 84 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 87 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 89 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . 90 2.3. HALOACETIC ACIDS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 90 2.3.1. Monochloroacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 92 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . 92 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 95 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . . 97 2.3.2. Dichloroacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 98 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . 98 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 102 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 108 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 109 2.3.3. Trichloroacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 111 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 114 Basis for the RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 118 2.3.4. Monobromoacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 118 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 118 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 120 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 121 2.3.5. Bromochloroacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 121 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 121 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 126 Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 127 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 127 2.3.6. Dibromoacetic Acid . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 128 Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 128 Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 138 v Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 140 Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 141 Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 142 2.4. BROMATE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 143 2.4.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 143 2.4.2. Systemic Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 145 2.4.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 148 2.4.4. Basis for the RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 150 2.4.5. Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 150 2.5. CHLORITE/ CHLORINE DIOXIDE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 151 2.5.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 152 2.5.2. Systemic Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159 2.5.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 161 2.5.4. Basis for RfD, MCLG, and MRDLG . . . . . . . . . . . . . . . . . . . . . . . . . . 163 2.5.5. Children's Risk in Relation to the MCLG and MRDLG . . . . . . . . . . . 163 2.6. CHLORINE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 164 2.6.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 165 2.6.2. Systemic Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 2.6.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 2.6.4. Basis for RfD and MRDLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 168 2.6.5. Children's Risk Relative to the MRDLG . . . . . . . . . . . . . . . . . . . . . . . 168 2.7. CHLORAMINE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 169 2.7.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 169 2.7.2. Systemic Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 169 2.7.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171 2.7.4. Basis for the RfD and MRDLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171 2.7.5. Children's Risk in Relation to the MRDLG . . . . . . . . . . . . . . . . . . . . . 172 2.8. MX [ 3 Chloro 4( dichloromethyl) 5 hydroxy 2( 5H) furanone] . . . . . . . . . . . . 173 2.8.1. Developmental/ Reproductive Effects . . . . . . . . . . . . . . . . . . . . . . . . . . 173 2.8.2. Systemic Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175 2.8.3. Carcinogenicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 175 2.8.4. Basis for RfD and MCLG . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 177 2.8.5. Children's Risk in Relation to the MCLG . . . . . . . . . . . . . . . . . . . . . . 177 3. SUMMARY AND CONCLUSIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178 4. REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 181 vi LIST OF TABLES Table ES 1. Comparison of Toxicity Endpoints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . xxix Table 1. Disinfectant Byproducts and their MCLGs Considered in this Document . . . . . 13 Table 2. Disinfectants and their MRDLGs Considered in this Document . . . . . . . . . . . . . 13 Table 3. Comparison of Toxicity Endpoints . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14 vii This document is an update of Health Risks to Fetuses, Infants and Children ( Final Stage 1 D/ DBP Rule, 1998). The new information that has been added to this document is based in large part on the contents of the Drinking Water Criteria Documents ( CDs) for Stage 2 disinfection byproducts that were available in the Office of Science and Technology as of March, 2003:

Contributors Nancy Chiu, Ph. D. U. S. EPA Office of Science and Technology, Office of Water Diana Wong, Ph. D., DABT U. S. EPA Office of Science and Technology, Office of Water Julie Du, Ph. D. U. S. EPA Office of Science and Technology, Office of Water Mary Ko Manibusan, M. P. H. U. S. EPA Office of Science and Technology, Office of Water Ambika Bathija, Ph. D. U. S. EPA Office of Science and Technology, Office of Water Joyce Donohue, Ph. D. U. S. EPA Office of Science and Technology, Office of Water Amal M. Mahfouz, Ph. D. U. S. EPA Office of Science and Technology, Office of Water Rita Schoeny, Ph. D. U. S. EPA Office of Science and Technology, Office of Water * Hend Galal Gorchev, Ph. D. World Health Organization ( WHO) Coordinator ( retired), Geneva, Switzerland Peer Reviews This document is based on Drinking Water Criteria Documents that have been peer reviewed. The epidemiology information used in this document was also peer reviewed by Pauline Mendola, Ph. D., U. S. EPA Office of Research and Development.

* Dr. Galal Gorchev helped the Health and Ecological Criteria Division's staff ( HECD) in the development of the Criteria Document on dichloroacetic acid ( DCA).

This document was prepared under the U. S. EPA contract No. 68 C 99 232 with GRAM, Inc., and Toxicology Excellence for Risk Assessment ( TERA) under the lead of Lynne Haber, and U. S. EPA contract No. 68 C 09 026 with the CADMUS Group and TERA, under the lead of Qiyu ( Jay) Zhao, respectively. viii ABBREVIATIONS BBDR biologically based dose response BCA bromochloroacetic acid BDCM bromodichloromethane BMD benchmark dose BMDL lower confidence limit on benchmark dose CI confidence interval CYP cytochrome P450 DBA dibromoacetic acid DBCM dibromochloromethane DBP disinfection byproduct DCA dichloroacetic acid D/ DBP disinfectant and disinfection byproduct DWEL drinking water equivalent level GD gestation day GLP good laboratory practice GST glutathione S transferase hCG human chorionic gonadotropin IRIS Integrated Risk Information System LH luteinizing hormone LHRH luteinizing hormone releasing hormone LOAEL lowest observed adverse effect level MBA monobromoacetic acid MCA monochloroacetic acid MCLG maximum contaminant level goal MCL maximum contaminant level MF modifying factor MRDLG maximum residual disinfectant level goal MX 3 Chloro 4( dichloromethyl) 5 hydroxy 2( 5H) furanone NOAEL no observed adverse effect level OR odds ratio pnd postnatal day POR prevalence odds ratio RfD reference dose RSC relative source contribution TCA trichloroacetic acid THM trihalomethane TTHM total THM UF uncertainty factor 1For the purposes of this document, effects on children are defined as effects on a parent's reproductive capacity, and effects on the fetus, infant, and children through sexual and physical maturity ( typically at age 18).

ix EXECUTIVE SUMMARY EPA's Office of Water is charged with ensuring that the United States population has clean water and safe drinking water. Untreated water can be contaminated with microbial agents that can cause infectious diseases. The most common result of infection by waterborne pathogens is diarrheal disease, which may have serious consequences for children, the elderly, pregnant women, and people of all ages with immune deficiencies. Water is often treated with chemical disinfectants to prevent waterborne infectious disease; this process results in formation of chemical disinfection byproducts.

EPA's drinking water assessments and regulations have historically considered sensitive subpopulations that may be at increased risk. EPA's policy, the Safe Drinking Water Act Amendments of 1996, and Executive Order 13045 call for the consideration of sensitive populations, with particular attention to fetuses, infants and children in conducting risk assessments and characterizations, and in establishing regulations and public health standards.

This document considers the potential risks to children1 from 13 disinfection byproducts ( DBPs)

and three disinfectants. Establishment of a Maximum Contaminant Level Goal ( MCLG) for a specific contaminant is based on available evidence of carcinogenicity or noncancer adverse health effects from drinking water exposure using the EPA's guidelines for risk assessment ( see 63 FR 69390 for a detailed discussion of the process for establishing MCLGs). MCLGs are nonenforceable health goals. As with MCLGs, Maximum Residual Disinfectant Level Goal ( MRDLGs) are nonenforceable health goals for drinking water disinfectants, based only on health x effects and exposure information, and are established at the level at which no known or anticipated adverse effects on the health of persons occur, and which allows an adequate margin of safety. The MCLGs and MRDLGs are used as the basis for setting the Maximum Contaminant Levels ( MCLs), which are the enforceable drinking water standards. MCLs are set as close to the MCLGs and MRDLGs as feasible, taking costs, treatment technology, and other considerations into account.

The proposed Stage 2 Disinfectant/ Disinfection Byproduct ( D/ DBP) Rule proposes new MCLGs for three DBPs: monochloroacetic acid ( MCA), trichloroacetic acid ( TCA), and chloroform. MCLGs and MRDLGs promulgated in 1998 for six DBPs and three disinfectants are also discussed in this document. Insufficient data were available to develop MCLGs for four DBPs.

This document updates the document on health risks to fetuses, infants, and children that was prepared for the final Stage 1 D/ DBP Rule ( EPA, 1998c), and includes a number of new developmental and reproductive toxicity studies in animals, new epidemiology studies of developmental/ reproductive effects, new cancer studies in animals, and the derivation of several new risk values. To evaluate the toxicity of disinfectants and disinfection byproducts ( D/ DBPs)

for fetuses, infants, and children, EPA asked the following questions in deciding that the MCLG or MRDLG for each D/ DBP is protective of children:

1. Is there information that shows that the D/ DBP causes effects in the developing fetus or harms a woman's ability to become pregnant and bear children? If it causes these effects, do these effects occur at lower doses than those that cause other types of effects? xi 2. If the D/ DBP causes cancer, are children more likely to be affected by a given dose than are adults?

3. If the D/ DBP causes a health effect other than cancer, such as an effect on the liver or kidney, are children affected at lower doses than are adults?

All available data were considered in addressing these questions, including any data directly comparing toxicity of the D/ DBP in children and adults, and any available mechanistic data. Based on an evaluation of the data for each D/ DBP, it can be concluded that the MCLGs/ MRDLGs of all the D/ DBPs in the proposed Stage 2 D/ DBP Rule are protective of fetuses, infants and children. This conclusion is based on the data summarized in this document;

data on each DBP are described in detail in the supporting criteria documents. There are uncertainties in this conclusion, however, due to incomplete information on the sensitivity of children to systemic effects, and on age related changes in the metabolism of the D/ DBPs.

Chlorinated Drinking Water. There are no reliable animal studies on the reproductive or developmental toxicity of chlorinated drinking water, nor are there animal studies addressing age related differences in the systemic response to chlorinated drinking water, or in the potential carcinogenicity of chlorinated drinking water.

Epidemiology studies suggest that DBPs are associated with developmental and reproductive effects under certain exposure conditions. The existing data are still relatively sparse,

and are insufficient for dose response analysis. There are inconsistencies among the available studies on the association between drinking water disinfection and specific effects, such as changes in the menstrual cycle, fetal growth, fetal viability, and congenital abnormalities. The xii studies with the best exposure assessment found the strongest association between changes in fetal growth and fetal viability and chlorinated drinking water exposure. Uncertainty in evaluating the effects of DBPs is enhanced by the difficulties in exposure assessment, including the different composition of D/ DBPs at different locations within a water system, and the variation in composition over time at a single location. The specific DBP( s) responsible for reproductive and/ or developmental effects is not known. The existing data suggest a relationship between THMs, particularly BDCM, and developmental effects. This may be related to a tendency among investigators to report concentrations of THMs, but not the concentrations of other DBPs. The available data suggest that additional studies on THMs and other DBPs are needed.

Chloroform. A prospective epidemiology study found no clear association between chloroform exposure and reduced menstrual cycle length. Studies in rats, rabbits and mice have reported developmental effects, including pup weight reduction, teratogenicity and embryolethality, from chloroform administration ( Murray et al., 1979; Ruddick et al., 1983;

Schwetz et al., 1974). These prenatal effects, however, were typically associated with exposures causing maternal toxicity, and occurred at oral doses above those causing hepatotoxicity. In a multigeneration reproductive assay in CD 1 mice treated with chloroform ( NTP, 1988), no adverse effects on fertility or reproduction were observed, although increased liver weight and liver lesions were observed in treated females.

Animal studies in rats, mice, and dogs have shown that the liver and kidney are the main target organs for chloroform. The current RfD for chloroform was estimated to be 0.01 mg/ kg/ day based on hepatotoxicity ( fatty cysts and an increase of serum glutamic pyruvic transaminase) observed in an oral study in dogs ( Heywood et al., 1979). xiii Chloroform has also been found to cause liver and kidney tumors in rodents ( EPA, 1994a,

1998a). Under EPA's Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999),

chloroform is likely to be carcinogenic to humans by all routes of exposure under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues ( EPA,

2001a). Chloroform is not likely to be carcinogenic to humans by any route of exposure under exposure conditions that do not cause cytotoxicity and cell regeneration. This weight

ofevidence conclusion is based on observations in animals exposed to chloroform. The carcinogenic potential of chloroform was evaluated using a nonlinear approach, in which liver toxicity was considered the most sensitive effect for chloroform and as a key event in its carcinogenicity.

The MCLG is based on an oral study in adult dogs ( Heywood et al., 1979), which was used to derive the RfD ( EPA, 2001a). This RfD corresponds to an MCLG of 0.07 mg/ L, based on a 70 kg adult consuming 2 liters of water per day ( a general assumption used in this document)

and a 20% relative source contribution ( RSC) from drinking water ( EPA, 2001d). The MCLG is considered protective of both adults and children based on several lines of reasoning. First,

developmental effects occurred at doses above those causing hepatotoxicity. Second, the weight of evidence indicates that the mode of action by which chloroform produces organ toxicity and carcinogenicity is the same for children and adults. Therefore, for both children and adults,

protection from organ toxicity would also provide protection from the carcinogenic effects of chloroform. Finally, the available data indicate that children are not uniquely sensitive to the organ toxicity caused by high doses of chloroform and there is no evidence from the available studies to suggest that children or fetuses would be qualitatively more sensitive to its effects than adults. For example, the liver toxicity observed in mice exposed prenatally, postnatally and as xiv adults in a two generation study was similar to the toxicity observed in mice exposed in a different study for a similar duration, beginning shortly after weaning. This comparison is far from perfect,

but it allows a general comparison of the chloroform doses that cause liver effects in young mice and adults. Based on these considerations, the MCLG based on systemic effects in adult dogs protects children from the reproductive, developmental, systemic, and carcinogenic effects of chloroform.

Bromodichloromethane ( BDCM). A prospective epidemiology study suggested that increasing levels of BDCM were associated with significantly shorter menstrual cycles. Because the subjects were also exposed to other contaminants and disinfection byproducts in the drinking water, establishment of a causal relationship between BDCM exposure and the observed effect is difficult.

Reproductive/ developmental assays with BDCM in animals, and epidemiology studies regarding these effects of BDCM have shown mixed results. Full litter resorptions have been reported in some rat studies ( Bielmeier et al., 2001; Narotsky et al., 1997a); the effect appears to be specific to F344 rats and not Sprague Dawley rats, and it is unclear whether the proposed mechanism of action is relevant to humans. An increase in stillbirths or spontaneous abortion has been reported in some epidemiological studies as being associated with increased BDCM levels ( King et al., 2000a; Waller et al., 1998), but the data are insufficient to show causality of BDCM.

Other studies in rats and rabbits reported mild developmental delay, possibly secondary to decreased water consumption ( CCC, 2000a, c; Ruddick et al., 1983; Christian et al., 2002), or did not report developmental or reproductive effects ( CCC, 2000b; NTP, 1998a); all tested up to maternally toxic doses. xv The liver and kidney are the main targets in rats of systemic effects of acute and chronic BDCM exposure. Effects on the liver, kidney and thyroid have been noted in male mice ( Aida et al., 1992; NTP, 1987). EPA has calculated an RfD of 0.002 mg/ kg/ day based on a LOAEL of 6.1 mg/ kg/ day in a chronic study ( Aida et al., 1992), in which the critical endpoint was liver fatty degeneration and granuloma in male rats ( EPA, 2002c). Due to the lack of a multigeneration reproductive toxicity study and uncertainty related to possible reproductive or developmental effects suggested by epidemiological studies ( EPA, 2002c), a 3 fold uncertainty factor ( UF) for database deficiencies was used in the derivation of the RfD.

Studies have shown that BDCM exposure results in tumors of the large intestine and kidneys in male and female rats, the kidneys in male mice and the liver in female mice ( NTP,

1987). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, BDCM has been classified as likely to be carcinogenic to humans, with a cancer oral slope factor of 8.1 x 10 3 per mg/ kg/ day, based on renal tumors in treated male mice ( EPA, 1998d).

The MCLG for BDCM is zero, based on its probable carcinogenicity and the use of linear low dose extrapolation in the cancer risk assessment. The MCLG of zero is protective for both the carcinogenic and systemic effects of BDCM in children and adults. In addition, there is not sufficient evidence indicating whether children are more sensitive to the toxic effects of BDCM than are adults.

Dibromochloromethane ( DBCM). Limited data are available for DBCM reproductive/ developmental toxicity. A prospective epidemiology study suggested that increasing levels of DBCM were associated with significantly shorter menstrual cycles. Because the subjects were also exposed to other contaminants and disinfection byproducts in the drinking water, xvi establishment of a causal relationship between DBCM exposure and the observed effect is difficult.

No developmental/ reproductive toxicity was observed in a rat developmental study ( Ruddick et al., 1983) or in a rat short term reproductive screening study ( NTP, 1996).

However, there was marginal evidence for decreased fetal weight in a mouse two generation reproductive study ( Borzelleca and Carchman, 1982).

DBCM has been shown to affect the nervous and immune systems, kidneys, and liver in rats ( NTP, 1985; Tobe et al., 1982). An RfD of 0.02 mg/ kg/ day was calculated based on reported histologic lesions ( fatty changes) in the liver in male rats after subchronic exposure to DBCM ( NTP, 1985).

Carcinogenicity studies have shown an increase in liver tumors in male and female mice and no increase in tumors in rats ( NTP, 1985). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, there is suggestive evidence of human carcinogenicity of DBCM,

but the data are not sufficient to assess human carcinogenic potential. Dose response assessment is not recommended under the EPA's Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999) for chemicals described by this descriptor. This descriptor is used when the evidence from human or animal data is suggestive of carcinogenicity, but further studies would be needed to determine human carcinogenic potential.

The MCLG for DBCM is 0.06 mg/ L, based on the subchronic study in rats ( NTP, 1985),

which was used to calculated the RfD of 0.02 mg/ kg/ day. Calculation of the MCLG took into account a drinking water RSC of 80% and an additional safety factor of 10 to account for the possible carcinogenicity of DBCM. EPA believes that the MCLG is protective for children's xvii health, because developmental or reproductive effects have not been found to occur below the level of the critical effect ( liver toxicity) used as the basis for the MCLG, and because development of the MCLG includes the standard UF of 10 for protection of sensitive populations.

There is, however, some uncertainty in this conclusion, because there is not sufficient evidence from studies on the systemic effects of DBCM, or on the metabolism of the compound, to determine whether children are more sensitive to the toxic effects of DBCM than are adults.

Bromoform. A prospective epidemiology study suggested that increasing levels of bromoform were associated with significantly shorter menstrual cycles. Because the subjects were also exposed to other contaminants and disinfection byproducts in the drinking water,

establishment of a causal relationship between bromoform exposure and the observed effect is difficult.

Developmental toxicities have been reported in rats and mice treated with bromoform, and these effects occurred at dose levels either below or equal to the dose that caused maternal toxicity. In a rat developmental study ( Ruddick et al., 1983), fetal skeletal anomalies were observed at a dose of 100 mg/ kg/ day, while no maternal toxicity was observed at doses as high as 200 mg/ kg/ day. In a mouse two generation reproductive study ( NTP, 1989a), developmental toxicities observed in F1 mice exposed to 200 mg/ kg/ day included hepatocellular degeneration,

decreased postnatal survival, and other signs of toxicity ( increased relative liver and decreased relative kidney weights), while the same dose also resulted in decreased body weight of pregnant dams at delivery.

Systemic effects on the liver, kidney, nervous system, and thyroid have been noted in rats and mice from bromoform exposure ( NTP, 1989b; Tobe et al., 1982). Based on hepatocellular xviii vacuolization in the liver of male rats in a subchronic study ( NTP, 1989b), the Agency derived an RfD of 0.02 mg/ kg/ day. Because the database for bromoform includes systemic toxicity studies in two species, a two generation study in mice, and a developmental toxicity study in rats, no database UF was used in developing the RfD.

Carcinogenicity bioassays have reported an increase in large intestine tumors in male and female rats, and no increase in male and female mice ( NTP, 1989b). In a study in which bromoform was administered by intraperitoneal injection in mice, the ratio of the number of lung tumors per mouse in the mid dose group was significantly elevated over controls ( Theiss et al.,

1977). Based on the observation of a rare tumor type in both sexes of rats, supported by positive genotoxicity findings, bromoform is considered likely to be carcinogenic in humans under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, and a cancer oral slope factor of 4.5 × 10

3 per mg/ kg/ day was calculated ( EPA, 2002c).

The MCLG for bromoform is zero, based on its probable carcinogenicity and the use of linear low dose extrapolation. EPA believes that the MCLG of zero is protective for both the carcinogenic and systemic effects of bromoform in children and adults. In addition, there are no studies indicating that children are more sensitive to the toxic effects of bromoform than are adults.

Monochloroacetic acid ( MCA). Developmental studies have shown mixed results. A published abstract of a developmental study in rats reported the occurrence of malformations of the cardiovascular system and maternal toxicity at the highest dose tested ( Smith et al., 1990);

another study did not find developmental toxicity at a maternally toxic dose, but this study was xix limited by a small sample size and incomplete examinations ( Johnson et al, 1998). No studies were located on the reproductive toxicity of MCA.

Subchronic and chronic studies suggest that the primary targets for MCA induced toxicity are the spleen, heart and nasal epithelium ( DeAngelo et al., 1997; NTP, 1992b). An RfD of 0.004 mg/ kg/ day was calculated based on a chronic rat drinking water study ( DeAngelo et al., 1997) in which the critical effect was increased spleen weight compared to controls. Due to the lack of adequate developmental toxicity studies in two species and the lack of a multi generation reproductive study, a 3 fold UF for database deficiencies was used in the calculation of the RfD.

MCA has not been shown to be carcinogenic in two chronic rodent bioassays ( DeAngelo et al., 1997; NTP, 1992b). However, both studies suffer from methodological limitations that limit the strength of the conclusions that can be drawn from them. Under the 1999 Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999), the data on MCA are inadequate for an assessment of human carcinogenic potential.

The proposed MCA MCLG is 0.03 mg/ L, based on the RfD of 0.004 mg/ kg/ day and assuming a RSC of 20% ( EPA, 2002a). EPA believes that this MCLG is protective for children since the levels at which systemic toxicity have been seen are much lower ( 20x) than the levels at which developmental effects have been observed. There is no evidence from studies on the systemic effects of MCA that children are more sensitive to the toxic effects of MCA than are adults. However, these data are limited. No data on potential metabolic differences between children and adults for MCA were located.

Dichloroacetic acid ( DCA). Developmental effects have been reported in two rat developmental studies. Smith et al. ( 1992) reported that DCA treatment during rat pregnancy xx resulted in developmental toxicities including implantation loss, reduced number of live fetuses,

decreased fetal body weight and crown rump length, cardiovascular, and soft tissue malformations. Epstein et al. ( 1992) reported that acute high dose treatments with DCA at specific rat developmental stages can induce developmental toxicity ( e. g., cardiac defects) in the absence of maternal toxicity.

The liver, testes, and brain are the major target organs for DCA induced systemic toxicity in DCA treated rats, mice, dogs, and humans ( Cicmanec et al., 1991; DeAngelo et al., 1999; Katz et al., 1981; Parrish et al., 1996; Stacpoole et al., 1998). The RfD of 0.004 was based on testicular degenerative changes, liver vacuolization, and brain histopathology, observed in a subchronic study in dogs ( Cicmanec et al., 1991).

Studies in male mice ( Anna et al., 1994; Bull et al., 1990; DeAngelo et al., 1991;

DeAngelo et al., 1999; Daniel et al., 1992; Ferreira Gonzalez et al., 1995; Herren Freund et al.,

1987), female mice ( Pereira and Phelps, 1996; Pereira, 1996), and rats ( DeAngelo et al., 1996)

have shown an increase in the incidence of liver tumors from DCA exposure. According to the 1999 Draft Carcinogen Risk Assessment Guidelines ( EPA, 1999), DCA is likely to be carcinogenic in humans, based on the weight of the evidence in animal bioassays and the lack of a clear mechanistic understanding of the mode of action.

The MCLG for DCA is zero, based on its probable carcinogenicity and the use of linear low dose extrapolation. The Agency believes that this MCLG of zero is protective for both the carcinogenic and systemic effects of DCA in children and adults. In addition, there is no evidence from studies on the systemic effects of DCA that children other than those with metabolic defect disorders are more sensitive to the toxic effects of DCA than are adults. However, children with xxi hereditary tyrosinemia, glycogen storage disease or hyperoxaluria may be at increased risk for DCA induced toxicity, as are adults with the same defect disorders.

Trichloroacetic acid ( TCA). Animal studies indicate that TCA can cause developmental toxicity. A developmental study in rats ( Johnson et al., 1998) reported increases in cardiac malformations; this study was limited by a small sample size and incomplete examinations. In another rat developmental study ( Smith et al., 1989), TCA treatment during gestation resulted in reduced fetal body weight and crown rump length. In addition, soft tissue malformations in the cardiovascular system were increased in a dose dependent manner. In both studies, maternal toxicity was also observed at the doses that caused developmental effects. No reproductive toxicity studies of TCA were identified.

TCA has been shown to affect the liver and kidney in rats and mice ( DeAngelo et al.,

1997; Dees and Travis, 1994; Mather et al., 1990; Parrish et al., 1996; Acharya et al. 1995)

exposed to TCA in drinking water. The Agency based the TCA RfD on increased serum levels of liver enzymes ( indicating cell damage) and histopathological evidence of necrosis in the liver in a 2 year rat study ( DeAngelo et al., 1997). A 10 fold database UF was used to account for database deficiencies, including the lack of adequate developmental toxicity studies in two species, the lack of a multi generation reproductive study, and the lack of full histopathological data in a second species. The calculated RfD is 0.03 mg/ kg/ day.

TCA has been shown to induce liver tumors in mice ( Pereira, 1996) but not in rats ( DeAngelo et al., 1997). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment,

the data on TCA provide suggestive evidence of carcinogenicity, but the data are not sufficient to assess human carcinogenic potential. Dose response assessment is not recommended under xxii the EPA's Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999) for chemicals described by this descriptor.

EPA has proposed an MCLG of 0.02 mg/ L, based on an RfD of 0.03 mg/ kg/ day, a drinking water RSC of 20%, and an additional safety factor of 10 to account for the possible carcinogenicity of TCA ( EPA, 2002a). This MCLG is expected to be protective of children, since developmental and systemic toxicity appear to occur at similar doses, and there is no evidence that children are more sensitive to the toxic effects of TCA than are adults. In addition,

uncertainties regarding the effect level for developmental toxicity have been taken into account through the application of UFs.

Monobromoacetic acid ( MBA). The toxicity data for MBA are very limited. There are no peer reviewed developmental studies, chronic toxicity, or carcinogenicity studies available. An RfD has not been established for MBA, because there are no systemic toxicity studies of sufficient duration. According to EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment ( EPA,

1999), the data on MBA are inadequate for an assessment of human carcinogenic potential ( EPA, 2002b).

An MCLG has not been established for MBA. Data relevant to potential fetal sensitivity are limited to a single developmental study reported in a published abstract in which maternal effects and fetal effects ( decreased live fetus size and increased incidence of soft tissue malformations) were noted in rats ( Randall et al., 1991).

Bromochloroacetic acid ( BCA). No long term reproductive/ developmental studies are available for BCA. A short term reproductive and developmental screening study reported significant decreases in live fetuses/ litter and total implants/ litter; however, no treatment related xxiii effects on sperm morphology or motility, mating or fertility were noted ( NTP, 1998b). In contrast, another short term drinking water study reported significant impairment in sperm motility, abnormal sperm morphology, altered spermiation, and reduced male fertility at comparable drinking water concentrations ( Klinefelter et al., 2002).

Oral toxicity studies have identified the kidney and liver as the target organs for BCA toxicity ( NTP, 1998b; Parrish et al., 1996). The data on BCA were insufficient for the derivation of an RfD, because there are no systemic toxicity studies of sufficient duration.

There are no carcinogenic bioassays available for this compound ( EPA, 2002b). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999), the data on BCA are inadequate for an assessment of human carcinogenic potential.

Due to the limited data available on the noncancer or cancer effects of BCA, EPA has not established an MCLG for this compound. Limited data suggest that decreased live fetuses/ litter and decreased total implants/ litter occurred at doses comparable to the ones that caused general toxicity in adults. The reproductive effects seen in male rats occurred at a lower dose, but no data were identified on whether exposure of young males enhances their sensitivity. These data do not support the hypothesis that fetuses or children are more sensitive than adults to the effects of BCA..

Dibromoacetic acid ( DBA). No peer reviewed developmental studies are available, but two published abstracts reported developmental effects, including increased prenatal mortality,

decreased pup weight, and hydronephrosis in gavage studies in mice ( Narotsky et al., 1996;

Narotsky et al., 1997b). DBA has been shown to be spermatotoxic following high dose single exposures or repeated longer term exposures ( CCC, 2001; Linder et al., 1994a; Linder et al., xxiv 1994b; Linder et al., 1995; Linder et al., 1997b), indicating that the male reproductive tract is a target organ in both adult and developing rats. Based on available studies that evaluated male reproductive toxicity in developing males ( Klinefelter et al., 2000, 2001; Veeramachaneni et al.,

2000), it is unclear whether the developing organism is more sensitive to these effects. However,

a two generation study found a lower incidence of males with reproductive toxicity in the F1 than the parental generation, and thus developing males did not appear to be more sensitive than adults in this study ( CCC, 2001). In a standard one generation drinking water study, Christian et al.

( 2001b) found no DBA related reproductive or developmental toxicity at drinking water concentrations high enough to cause decreased water consumption due to poor palatability. In contrast, recent results from Klinefelter et al. ( 2001) indicate that DBA causes pubertal delay independent of effects on body weight.

DBA has been shown to affect the liver and immune system in drinking water studies in mice ( NTP, 1999; Parrish et al., 1996), and to cause neurobehavioral effects in a drinking water study in rats ( Phillips et al., 2002, published abstract). The data on DBA are insufficient for the derivation of an RfD, because there are no systemic toxicity studies of sufficient duration.

No carcinogenicity bioassays are available for DBA ( EPA, 2002b). In published abstracts,

So and Bull ( 1995) reported that DBA induces aberrant crypt foci in the colon of rats, and Stauber et al. ( 1995) reported that DBA induces liver tumors in mice. Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, there is suggestive evidence of carcinogenicity of DBA, but the evidence is not sufficient to assess human carcinogenic potential. Dose response assessment is not recommended under the EPA's Draft Guidelines for Carcinogen Risk Assessment ( EPA, 1999) for chemicals for which this descriptor applies. xxv Due to the limited data available on the noncancer and cancer effects of DBA, EPA is not proposing an MCLG for the compound. The data are limited as to whether children may be more sensitive than adults. A two generation reproductive study ( CCC, 2001) reported that the male reproductive tract is a target organ in both adult and developing rats. However, the incidence of affected animals was higher in the parental than the F1 generation, and thus developing males did not appear to be more sensitive than adults. In a reproductive toxicity study ( Christian et al.,

2001b), DBA readily crossed the placenta and distributed to the fetus, but it did not appear to bioaccumulate. Developmental toxicity ( prenatal mortality, decreased pup weight, and hydronephrosis) was reported in gavage studies ( Narotsky et al., 1996; 1997b), but not in the drinking water reproductive toxicity studies ( Christian et al., 2001b; CCC, 2001). It is unclear whether this difference is due to differences in the route of dosing ( gavage versus drinking water),

species sensitivity ( mouse versus rat), or other factors. Regardless of the reason for the difference, any developmental toxicity appears to occur at doses well above those causing male reproductive toxicity. Additional reproductive studies are being conducted currently to evaluate this effect in males.

Bromate. Limited data are available on the reproductive/ developmental effects of bromate. In a screening reproductive/ developmental study in rats, a statistically significant decrease in epididymal sperm density was observed ( Wolf and Kaiser, 1996).

Acute oral poisoning of children with potassium bromate from accidental ingestion of hair home permanent neutralizing solution have shown to result in central nervous system effects such as sedation and lethargy, irreversible deafness and kidney effects ( Benson, 1951; Gradus et al.,

1984; Lichtenberg et al., 1989; Lue et al., 1988; Mack, 1988; Parker and Barr, 1951; Quick et al., xxvi 1975; Warshaw et al., 1985; Watanabe et al., 1992). Subchronic and chronic animal studies indicate that the kidney is the primary target organ for bromate toxicity following long term exposure ( DeAngelo et al., 1998; Kurokawa et al., 1986b; Kurokawa et al., 1990; Nakano et al.,

1989). EPA has derived an RfD of 0.004 mg/ kg/ day for bromate, based on urothelial hyperplasia in male rats in the DeAngelo et al. ( 1998) study ( EPA, 2001b). A 3 fold database UF was used to account for database deficiencies, including the lack of developmental toxicity studies in two species and the lack of a multi generation study.

A carcinogenicity study reported bromate exposure via drinking water to result in an increase in tumors of the kidney, thyroid and tunical vaginalis in male rats and male mice ( DeAngelo et al., 1998). Under the Draft Guidelines for Carcinogen Risk Assessment ( EPA,

1999), bromate is likely to be carcinogenic to humans by the oral route of exposure ( EPA,

2001b).

The MCLG for bromate is zero, based on its probable human carcinogenicity and the use of linear low dose extrapolation. The Agency believes that the MCLG of zero is protective for both the carcinogenic and systemic effects of bromate in children and adults. In addition, there are no data suggesting that children are more sensitive than adults to bromate,

Chlorite/ Chlorine Dioxide. Chlorite and chlorine dioxide are evaluated together in this document because it is likely that the studies conducted with chlorite, the predominant degradation product of chlorine dioxide, are relevant to characterizing the toxicity of chlorine dioxide.

A number of studies conducted with chlorite and chlorine dioxide in rats, mice, and rabbits have reported developmental/ reproductive effects, including neurobehavioral effects, xxvii increased number of resorbed and dead fetuses, and skeletal abnormalities ( CMA, 1996; Couri et al., 1982; Harrington et al., 1995b; Mobley et al., 1990; Suh et al., 1983; Moore et al., 1980;

Orme et al., 1985; Taylor and Pfohl, 1985; Toth et al., 1990). Lowered auditory startle amplitude and altered liver weights in two generations of rats have been identified as the critical effects for exposure to chlorite or chlorine dioxide ( CMA, 1996).

Subchronic and chronic animal studies reported mixed results on the systemic toxicity of chlorite or chlorine dioxide. Reported systemic effects included nasal and stomach lesions, spleen alteration, adrenal weight change, and hematological effects. However, no firm conclusions can be made regarding these observed systemic effects. An RfD for each of the two chemicals was calculated to be 0.03 mg/ kg/ day ( EPA, 2000c) based on the critical effects of lowered auditory startle amplitude and decreased liver weight observed in the two generation study in rats ( CMA,

1996).

No chronic carcinogenicity bioassays are available for chlorine dioxide. Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, the data for chlorite and chlorine dioxide human carcinogenicity are inadequate for an assessment of human carcinogenic potential.

An MCLG of 0.8 mg/ L for chlorite and an MRDLG at the same level for chlorine dioxide can be calculated from the RfD and using a drinking water RSC of 80%. This level is considered to be protective of children because it is based on data from a two generation reproduction study that examined numerous developmental, reproductive and systemic endpoints. The results of the CMA ( 1996) study are supported by the results of four other developmental studies that showed similar effects at similar dose levels. In addition, a 10 fold factor was used in the derivation of the xxviii RfD to account for human variability in response to the toxic effects of these chemicals, including the response of sensitive individuals such as children.

Chlorine. Animal studies in rats and mice have not shown evidence of reproductive or developmental effects from chlorine exposure ( Carlton and Smith, 1985; Druckrey, 1968).

A 2 year study of chlorine in drinking water reported no systemic toxicity and no carcinogenic effects in rats or mice ( NTP, 1992a). The RfD of 0.1 mg/ kg/ day for chlorine was calculated based on a free standing NOAEL identified in a 2 year bioassay in female rats ( NTP,

1992a).

Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, the data on chlorine are inadequate for an assessment of human carcinogenic potential.

An MRDLG of 4 mg/ L was calculated based on the RfD. EPA believes this level is protective of children because the current RfD is based on a free standing NOAEL in a two year study in which chlorine dosing began when rats and mice were as young as 7 weeks old. In addition, the UF used in the derivation of the RfD includes a 10 fold factor to account for human variability in response to the toxic effects of these chemicals, including the response of sensitive individuals such as children. In addition, aqueous chlorine appears to react so rapidly with sulfurcontaining amino acids in saliva that all free available chlorine is dissipated before water is swallowed ( EPA, 1994d). Therefore, the possibility of oral exposure to chlorine by fetuses,

infants and children is very limited.

Chloramine. The data on developmental effects of chloramine are limited to a rat developmental study on monochloramine in drinking water ( Abdel Rahman et al., 1982). xxix Monochloramine did not produce any significant changes in rat fetuses at any dose level. No reproductive toxicity studies of chloramine were located.

A drinking water study in rats exposed for 12 months reported hematologic effects from monochloramine exposure ( Abdel Rahman et al., 1984b); however, these effects were not reported in a chronic study ( NTP, 1992a), and the biological significance of the observed effects is unclear. In a lifetime study of chloraminated water in rats ( NTP, 1992a), there were no biologically significant effects including body weight, organ weight, or histopathology at any dose levels. The chloramine RfD of 0.1 mg/ kg/ day was calculated based on the absence of adverse effects in the lifetime study in rats ( NTP, 1992a).

No data were identified on the carcinogenic effects of chloramine in humans or animals.

Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, the data on chloramine are inadequate for an assessment of human carcinogenic potential.

An MRDLG of 3 mg/ L for chloramine ( 4 mg/ L measured as total chlorine) was derived,

based on the lack of toxic effects, assuming a drinking water RSC of 80%. Data on hematological effects of chloramine in infants and young animals are not available, but these groups may be more sensitive than adults to hematological effects, because infants have a transient deficiency of methemoglobin reductase, the enzyme that reduces methemoglobin to hemoglobin. Nevertheless, EPA believes that the MRDLG for chloramine, which is based on a free standing NOAEL in a two year study and includes the standard UF of 10 for protection of sensitive populations, is protective of infants, although there is some uncertainty in this conclusion in the absence of a quantitative analysis, in light of the possibility that newborns may have increased sensitivity to the hematological effects of chloramine. xxx MX ( 3 Chloro 4( dichloromethyl) 5 hydroxy 2( 5H) furanone). A study in rats showed no developmental toxicity at any MX dose tested, including a maternally toxic dose ( Huuskonen et al., 1997).

Subchronic rat studies have shown MX to have a strong local irritant effect on the gastrointestinal tract ( Vaittinen et al., 1995). An RfD has not been established for MX.

MX has been shown to be a carcinogen in rats, causing tumors of the thyroid, liver,

mammary gland, lung, adrenal gland, and pancreas ( Komulainen et al., 1997). Under EPA's 1999 Draft Guidelines for Carcinogen Risk Assessment, MX is likely to be carcinogenic to humans ( EPA, 2000e).

EPA has not established an MCLG for MX because the Agency has not yet conducted a full assessment of the systemic and carcinogenic effects of MX. There are insufficient data available to evaluate whether children are more sensitive to the toxic effects of MX than are adults.

Comparison of Toxicity Endpoints. Table ES 1 shows the MCLGs, MRDLGs, and the toxicity endpoints used to set the MCLGs/ MRDLGs for each of the disinfectants and disinfection byproducts evaluated in this document. xxxi Table ES   1. Comparison of Toxicity Endpoints Disinfectant/

Disinfection Byproduct Systemic Toxicity ( mg/ kg/ day) Developmental Toxicity ( mg/ kg/ day) Carcinogenicity Based on 1999 Guidelines MCLG/ MRDLGb mg/ L NOAELa LOAELa NOAEL LOAEL Chloroform None ( BMDL10 1.0) 12.9 35   50 126 likely to be carcinogenic to humans under high exposure conditionsc 0.07 ( sys tox)

Bromodichloromethane ( BDCM) None ( BMDL10 0.8) 6d 25 50 likely to be carcinogenic to humans 0 ( Ca)

Dibromochloromethane ( DBCM) None ( BMDL10 1.6) 28d 17 171 suggestive evidence of human carcinogenicity, but the data are not sufficient to assess human carcinogenic potential. 0.06 ( sys tox & Ca)

Bromoform None ( BMDL10 2.6) 36d 50 100 likely to be carcinogenic to humans 0 ( Ca)

Monochloroacetic acid ( MCA) ND 3.5 70e 140 data are inadequate for an assessment of human carcinogenic potential 0.03 ( sys tox)

Dichloroacetic acid ( DCA) ND 12.5 14 140 likely to be carcinogenic to humans 0 ( Ca)

Trichloroacetic acid ( TCA) 32.5 364 ND 291 suggestive evidence of carcinogenicity, but the data are not sufficient to assess human carcinogenic potential 0.02 ( sys tox)

Monobromoacetic acid ( MBA) ND ND 50e 100 data are inadequate for an assessment of human carcinogenic potential NDf Bromochloroacetic acid ( BCA) 15g 39 19 50 data are inadequate for an assessment of human carcinogenic potential NDf Disinfectant/

Disinfection Byproduct Systemic Toxicity ( mg/ kg/ day) Developmental Toxicity ( mg/ kg/ day) Carcinogenicity Based on 1999 Guidelines MCLG/ MRDLGb mg/ L NOAELa LOAELa NOAEL LOAEL xxxii Dibromoacetic acid ( DBA) 2 10 50 d 100 suggestive evidence of carcinogenicity, but the data are not sufficient to assess human carcinogenic potential NDf Bromate 1.1 6.1 7.7 22 likely to be carcinogenic to humans 0 ( Ca)

MX ND ND ND ND likely to be carcinogenic to humans NDf Chlorite ND ND 3 6 data are inadequate for an assessment of human carcinogenic potential 0.8 ( devel tox)

Chlorine dioxide ND ND 3 6 data are inadequate for an assessment of human carcinogenic potential 0.8 ( devel tox)

Chlorine 14 ND 5h ND data are inadequate for an assessment of human carcinogenic potential 4 ( sys tox)

Chloramine 9.5 ND 10 ND data are inadequate for an assessment of human carcinogenic potential 3 ( sys tox) i aNOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level; ND = not determined; BMDL 10 = 95% lower confidence limit on the benchmark dose for 10% extra risk.

Where appropriate, duration adjusted values shown bMCLG = Maximum contaminant level goal; ( Ca) = basis for MCLG is carcinogenic effects; ( sys tox) = basis for MCLG is systemic toxic effects; ( devel tox) = basis for MCLG is developmental effects.

cLikely to be carcinogenic to humans under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues, but not likely to be carcinogenic to humans at a dose level that do not cause these effects.

dNo NOAEL was identified for the critical effect for BDCM or DBCM; the NOAEL for bromoform was 18.

eAdverse effect levels reported in a published abstract only, and thus should be considered preliminary.

fInsufficient data for derivation of a MCLG.

gAdverse effect level for systemic effects in a short term study; no subchronic or chronic studies were available.

hHighest dose tested.

i 4mg/ L measured as total chlorine xxxiii PLAIN LANGUAGE SUMMARY OF EVALUATION OF CHILDREN'S RISK It is the duty of EPA's Office of Water to make sure that the people living in the U. S.

have clean water and safe drinking water. To make the drinking water safe to drink, disinfectants ( such as chlorine and chlorine dioxide) are added to water to kill bacteria and other organisms that can cause waterborne diseases in people of all ages. Waterborne disease can produce illnesses with symptoms such as diarrhea, nausea, and stomach cramps. Disinfectants were first added to drinking water many years ago, and their addition is today considered to be the most successful public health program in U. S. history.

The Safe Drinking Water Act authorizes EPA to set national health based standards for drinking water to protect against both naturally occurring and man made contaminants that may be found in drinking water and that may hurt people's health. U. S. EPA, states, and water systems then work together to make sure that these standards are met. When EPA sets these regulations, it considers the effects of the chemicals on many different groups within the general population, including children, pregnant women, older people, and people with serious illnesses.

The disinfectants that kill bacteria and other organisms react with naturally occurring material in the water to form other chemicals called disinfection byproducts ( DBPs), and the disinfectants and disinfection byproducts ( D/ DBPs) may present the potential for health effects. Therefore, the highest health protection requires balancing risks so that organisms that cause disease are controlled, while also having a low risk from D/ DBPs. The purpose of this document is to examine whether children may be particularly sensitive to the health effects of drinking water disinfectants and to D/ DBPs. xxxiv Studies in both animals and in people drinking chlorinated water with sufficiently high levels of DBPs have shown that DBPs may present a health risk to people drinking the water.

Some of these chemicals have been shown to cause cancer in laboratory animals, and to have effects on the developing fetus or on the liver and kidney in laboratory animals. Because it is very hard to accurately measure exposure to DBPs, there is a lot of uncertainty about how much DBP exposure may result in effects on fetuses or hurt one's ability to have children. The observed relationship is also rather weak. Some data suggest that one specific class of DBPs, the trihalomethanes, may cause the developmental effects, but it is not clear if this association is real,

or related to the methods used to measure an association. Some studies with people have suggested that drinking water containing high levels of some of the DBPs may also increase a person's risk of developing certain types of cancer. However, there are important differences in the results from different studies, and it is not certain that the risk of any cancer is higher in people drinking chlorinated water. There is very little information about whether children are more sensitive than adults to any cancer causing effects of DBPs.

EPA is using the health evaluations presented in this document for D/ DBPs to set Maximum Contaminant Level Goals ( MCLGs) for DBPs or Maximum Residual Disinfectant Level Goals ( MRDLGs) for disinfectants. MRDLGs ( and MCLGs) are nonenforceable health goals based only on health effects and exposure information, and are established at the level at which no known or anticipated adverse effects on the health of persons occur and which allows an adequate margin of safety. Therefore, MRDLGs do not reflect the improvement to health resulting from the disinfectant killing microbes and preventing disease. After EPA sets the xxxv MCLGs/ MRDLGs, it then sets the enforceable drinking water regulations for the chemicals ( Maximum Contaminant Levels  MCLs), as close to the goal levels as is practical.

This document looks at the available health data for 13 DBPs and three disinfectants. For four of the DBPs, EPA has determined that there are not enough data available on the potential health effects to set an MCLG. For the other D/ DBPs, EPA has determined that the MCLG or MRDLG for each of these chemicals is protective of children. For some chemicals, this is because the MCLG or MRDLG was set using a study that examined the effects of the chemical on the developing fetus. For other chemicals, EPA determined that the harmful health effects were not more likely to occur in children than in adults, or would not affect children at lower doses than they would affect adults. This conclusion is based in part on the standard risk assessment approach of using a protective factor to account for human variability, including potentially sensitive populations such as children, although there is some uncertainty in this conclusion,

because data on general toxic effects in young animals are not available.

This document provides a summary of the existing or proposed MCLGs and MRDLGS for each D/ DBP. New or revised MCLGs are being proposed for chloroform, monochloroacetic acid, and trichloroacetic acid. 1 1. INTRODUCTION 1.1. RISK TO CHILDREN EPA's Office of Water is charged with ensuring that the United States population has clean water and safe drinking water. This mandate covers chemical, physical, and biological pollutants. For drinking and other uses, water is often treated with chemical disinfectants to prevent waterborne infectious disease. The disinfection of public drinking water supplies to prevent waterborne disease is probably the most successful public health program in U. S. history.

The most common result of infection by waterborne pathogens is diarrheal disease. The diarrheal diseases most commonly associated with waterborne infectious agents may have a more severe outcome for children and debilitated adults; thus, the savings in lives from disinfecting water have been largely in those populations. Chemical disinfection byproducts ( DBPs) are formed from the reaction of chlorine and other disinfectants with naturally occurring organic materials in source water. These DBPs present the potential for health effects. Thus, to provide for maximum human health protection it is necessary to balance risks from water pathogens and DBPs.

EPA's drinking water assessments and regulations have historically considered sensitive subpopulations that may be at increased risk. Several initiatives in the mid to late 1990s mandated explicit consideration of fetuses, infants and children as potentially sensitive subpopulations. In 1995, EPA established an agency wide policy that calls for consistent and explicit consideration of the risk to infants and children in all risk assessments and characterizations, as well as in environmental and public health standards ( Memorandum from the Office of the Administrator,

October 20, 1995). The Safe Drinking Water Act amendments of 1996 also stipulated that in establishing maximum contaminant levels ( MCLs) the Agency shall consider " the effect of such 2 contaminants upon subgroups that comprise a meaningful portion of the general population ( such as infants, children, pregnant women, the elderly, individuals with a history of serious illness or other subpopulations) that are identifiable as being at greater risk of adverse health effects due to exposure to contaminants in drinking water than the general population." On April 21, 1997, the President Clinton signed an Executive Order ( 13045) that federal health and safety standards must include an evaluation of the potential risks to children in planned regulations. EPA's Office of Water follows the above policy and order and has historically considered risks to sensitive populations ( including fetuses, infants and children) in establishing drinking water assessments,

advisories or other guidance and standards ( Ware, 1989; EPA, 1991a).

There is a growing awareness that children, infants, and fetuses ( like other developing organisms) are not small adults, and may react to chemical exposures differently from adults.

These differences can result because the physiology of the developing body differs from that of adults; children consume more food and water and inhale more air than adults in proportion to their body mass; and the activity patterns of children, including crawling, mouthing, and outdoor play, may result in greater exposure to environmental contaminants. From the physiological perspective, the neurological, immunological, and digestive systems of children are still developing, and thus can be more sensitive to chemical damage. These differences can lead to impaired fetal or child development at doses that do not induce adverse effects in adults. In addition, children undergo substantial hormonal changes as they enter sexual maturity. As described in more detail below, chemical toxicokinetics ( absorption, distribution, metabolism, and excretion) can also differ between adults and children, particularly for children under one year of age. Thus, this document considers hazards to children that may result from exposure during 3 preconception, prenatal or postnatal development through sexual maturity. EPA must balance the potential for greater risk to children from D/ DBPs with the greater sensitivity of children to the diarrhea resulting from infection with waterborne pathogens.

Epidemiology data show that cancer risk can also differ between children and adults. For example, children were reported to be 2 fold more sensitive to the carcinogenicity of ionizing radiation than adults ( NRC, 1990). Recent experience has shown that diethylstilbestrol ( DES)

treatment during pregnancy caused some female offspring to develop clear cell vaginal carcinomas, but the same carcinomas did not appear at an increased rate in DES exposed mothers. In addition, it has been argued that developing organisms may be more sensitive than adults to the effects of strong mutagens, because there is less time for DNA damage to be repaired before the cell divides. A comparison of cancer bioassay results from 69 studies ( EPA,

1996b) indicated that combined perinatal and adult exposure slightly increases the incidence of a given type of tumor compared to the normal bioassay protocol, but it is not known if this reflects the effect of an increased length of exposure or a heightened sensitivity of the young animal to the carcinogenic effects of the chemical.

Age dependent differences in toxicokinetics can also be an important determinant of children's risk. Infants ( and children, to a lesser degree) differ from adults in a number of factors affecting absorption, distribution, metabolism, and excretion ( reviewed by Scheuplein et al.,

2002). Gastric dynamics and enzyme levels differ from adult values, but generally reach adult levels by the first year of life or earlier. Distribution can exhibit age specific differences, due to such factors as differences in total body fat and extracellular water. Age related differences in parameters affecting distribution are most apparent between adults and infants, although changes 4 continue to occur during childhood, adolescence, and adulthood. Both chemical biotransformation ( Phase I) and conjugation ( Phase II) metabolism systems are generally immature at birth, and infants may have alternative metabolic pathways. Cytochrome P450 enzymes are the most common Phase I enzymes. These enzymes appear in human fetuses during the first half of gestation, and enzyme activity levels at birth are about one third the activity in adults ( Scheuplein et al., 2002). In infants, cytochrome P450 2E1 ( CYP2E1), an isoenzyme of cytochrome P450, rapidly increases after birth to a level similar to that in young/ mature adults;

other isoenzymes of P450 have a low activity immediately after birth ( EPA, 2000d). If a chemical is converted to a less toxic form by CYP2E1, lower enzyme levels in infants would increase toxicity. If a chemical is converted to a more toxic form by CYP2E1 metabolism ( activation),

lower enzyme levels in infants would decrease toxicity. Conversely, children age 6 months up to approximately 12 years can metabolize and eliminate some chemicals more rapidly than adults ( Renwick and Lazarus, 1998; Ginsberg et al., 2002). In these cases, increased enzyme activity in young children for these chemicals would decrease toxicity if the chemical is converted to a less toxic form and increase toxicity if the chemical is converted to a more toxic form. The potential for enzyme induction ( increases in enzyme levels following exposure to certain chemicals) may also differ with age.

The expression of isoenzymes is also both species and strain dependent. In different animal species, the development of isoenzymes follows different courses, producing different patterns of age related changes in metabolizing capability. For example cytochrome P450 enzymes are virtually absent in rat fetuses, but are present in humans in the latter half of pregnancy ( Scheuplein et al., 2002). After birth, the levels of at least one cytochrome P450 5 enzyme ( CYP2E1) increase rapidly in animals to adult levels ( Song et al., 1986; Umeno et al.,

1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990). These interspecies differences need to be taken into account in extrapolating from developing laboratory animals to children, and may make this extrapolation more complex. Age related changes in enzyme inducibility can also occur.

In addition to differences in susceptibility engendered by physiological differences between children and adults, differences in exposure may also contribute to differences in risk. For example, children may incur greater exposure, because they ingest more food and fluids, and inhale more air in proportion to their body weights than do adults. For example, the mean drinking water ingestion rate for the U. S. population ( all ages) is 16 mL/ kg/ day, while the mean ingestion rate for babies younger than one year old is 46 mL/ kg/ day and for children aged 1 10 it is 19 mL/ kg/ day. Thus, for babies, the drinking water ingestion rate is estimated to be three to four times higher than for the population as a whole ( EPA, 2000b). Similarly, the breathing rate for children is higher than that for adults. A mean estimate of the breathing rate for children aged 1 2 years is 6.8 m3/ day, corresponding to a mean body weight of 12 kg ( EPA, 2000a). By contrast, the mean inhalation rate is 11.3 m3/ day for women and 15.2 m3/ day for men, and the mean adult body weight is 71.8 kg ( EPA, 1997). This means that, for a chemical that is absorbed from the respiratory tract and systemically distributed, a given concentration in air would result in a higher dose on a mg/ kg/ day basis for children. Dermal exposure would also be higher, since children have a higher surface area per unit body weight, as compared to adults. For example, the average surface area: body weight ratio for children aged 0 2 is 0.064, while the average ratio for adults is 0.028 ( EPA, 1997). This means that children would have a higher exposure than adults 6 to water contaminants from bathing, showering, or swimming. Exposure levels can also differ due to different activity patterns, the increased hand to mouth activity of infants being a commonly cited example. In summary, children may be more sensitive to the toxic effects of certain environmental contaminants than are adults, due to their developing organ systems,

different metabolic rates and processes, higher surface area per unit weight, and greater exposure potential. This document addresses the potential greater sensitivity of children, infants, and fetuses to the toxic effects of chemicals, due to the potential for toxicokinetic and toxicodynamic differences between adult and developing organisms.

1.2. RISK ASSESSMENT METHODS Risk assessment is a process by which judgments are made about an agent's potential to cause harm to humans. Risk assessment of chemicals follows the process developed by the National Academy of Sciences/ National Research Council ( NAS/ NRC). This process is based on analysis of scientific data to determine the likelihood, nature and magnitude of harm to public health associated with exposure to environmental agents ( NRC, 1983, 1994). The NAS paradigm defines four steps:

Hazard Assessment   The process of determining whether exposure to an agent can cause an increase in the incidence of a particular adverse health effect ( e. g.,

cancer, birth defect) and whether the adverse health effect is likely to occur in humans.

Dose Response Assessment   A determination of the relationship between the magnitude of an administered, applied, or internal dose and a specific biological 7 response. Response can be expressed as measured or observed incidence, percent response in groups of subjects ( or populations), or as the probability of occurrence within a population

Exposure Assessment   An identification and evaluation of the human population exposed to a toxic agent, describing its composition and size, as well as the type,

magnitude, frequency, route and duration of exposure

Risk Characterization   Summarizes and integrates the scientific findings of the hazard, dose response and exposure assessments to determine the potential human risk, and discusses uncertainties and assumptions.

To evaluate the special toxicity of disinfectants and disinfection byproducts ( D/ DBPs) for fetuses, infants, and children, the following three types of toxicity studies were evaluated:

1. Developmental and reproductive toxicity including both prenatal and postnatal exposures and effects 2. Systemic toxicity 3. Carcinogenicity These study types are evaluated in the context of EPA's guidelines for developmental toxicity risk assessment ( EPA, 1991b), reproductive toxicity risk assessment ( EPA, 1996c), and carcinogen risk assessment ( EPA, 1999).

Developmental toxicity is defined as the occurrence of adverse effects in the developing organism that may result from exposure before conception, during prenatal development, or postnatally to the time of sexual maturation. Adverse effects may be detected at any point in the life span of the organism ( i. e., in the developing organism, neonate, adolescent, or even in the 8 elderly as a late age onset disorder). There are a number of developmental abnormalities of concern: spontaneous abortions, stillbirths, premature mortality, reduced birth weight, reduced crown rump length, malformations ( overt teratogenicity, such as spinal bifida, no eyes, cleft palate, hydrocephaly, or anencephaly), delayed skeletal ossification, mental retardation, and sensory loss, as well as other adverse functional and physical effects. Developmental abnormalities from all causes are extremely common and present an enormous burden for society.

Reproductive toxicity is defined as the occurrence of biologically adverse effects on the reproductive systems of females or males that may result from exposure to environmental agents.

This may be expressed as alterations to the female or male reproductive organs, the related endocrine system, or pregnancy outcomes ( EPA, 1996c). Some examples include adverse effects on fertility, gestation, female reproductive cycle normality, the onset of puberty, or premature reproductive senescence. In males, reproductive toxicity may be manifest as adverse effects on male reproductive organ weight and morphology, sexual behavior, sperm count, sperm morphology, and sperm activity. Male reproductive effects are indirectly related to children's health risk due to the potential for male fertility problems to affect the ability of a woman to conceive and the subsequent health of the fetus and child.

In the quantification of noncarcinogenic effects of oral exposure, a reference dose ( RfD) is most often calculated. The RfD is an estimate ( with uncertainty spanning an order of magnitude)

of a daily exposure to the human population ( including sensitive subgroups) that is likely to be without an appreciable risk of deleterious health effects during a lifetime. The RfD is derived from a no observed adverse effect level ( NOAEL), or lowest observed adverse effect level ( LOAEL), identified from a chronic or subchronic study, divided by UF( s) ( UF). When the data 2All benchmark doses reported in this document reflect the lower confidence limit on the dose that would result in a 10% extra risk, i. e., a benchmark response ( BMR) of 10%, notated as the BMDL10. The benchmark doses ( BMDs) reported here reflect the central tendency estimate, while the BMDL values reflect the 95% lower confidence limit on the dose.

9 RfD

( NOAEL or LOAEL or BMDL ) Uncertainty Factor( s)

mg/ kg/ day. support it, a benchmark dose2 ( BMD) ( technically, the 95% lower confidence limit on the BMD,

termed the BMDL) may be calculated by applying an appropriate mathematical curve fitting procedure. The BMDL can then be used as a point of departure for calculation of the RfD. The RfD is calculated as follows:

Selection of the UF to be employed in the calculation of the RfD is based on professional judgment, which considers the entire database of toxicologic effects for the chemical. UFs are applied as described in the box on the next page.

The Agency is increasingly moving towards incorporating toxicokinetic and toxicodynamic data in the development of UFs, moving away from the use of default UFs. There is a continuum in the application of UFs, ranging from default, through categorical, to fully dataderived values based on chemical specific data. The International Programme on Chemical Safety ( IPCS) has developed guidelines for the use of empirical data in the development of Chemical

Specific Adjustment Factors ( CSAFs) for interspecies differences and human variability ( IPCS,

2001). In the framework adopted by the IPCS, the default UFs of 10 for interspecies differences and 10 for intraspecies differences are each divided into subfactors for toxicodynamics and toxicokinetics. When data are available, the default factors can be replaced with categorical uncertainty factors, also called Chemical Specific Adjustment Factors ( CSAFs). These guidelines 10 allow data to be used quantitatively in the development of UFs, rather than relying solely on professional judgement and the standard factors of 1, 3, and 10 noted in the box.

The potential for greater sensitivity of children, infants, and fetuses is also taken into account in several aspects of the assessment. For example, this issue is considered in the choice of UF for the adequacy of the database; part of the consideration for the database UF is whether reproductive and developmental toxicity studies are available. With the greater attention being paid to children's risk, the choice of the database UF also considers whether there is any reason for concern that developing organisms may be more sensitive to systemic toxicity caused by the chemical. In addition, the 10 fold factor for human variability, includes variability in the response of sensitive individuals such as children.

Cancer assessment includes both qualitative and quantitative components. The qualitative assessment consists of an evaluation of the weight of evidence of the carcinogenic risk from the chemical. EPA's 1986 Guidelines for Carcinogen Risk Assessment evaluated chemicals using a letter classification system ( A E) indicating the weight of evidence of carcinogenic risk ( EPA,

1986). This classification was based primarily on animal and human studies showing an increase in tumors. In 1996, EPA proposed revisions to the 1986 Guidelines ( EPA, 1996a). Further draft revisions to the Guidelines were released in 1999 ( EPA, 1999). EPA has adopted the policy that,

until final guidelines are issued, the 1999 draft revised guidelines will serve as EPA's interim guidance to EPA risk assessors conducting cancer risk assessments ( EPA, 2001g). Rather than relying exclusively on tumor findings, the new draft Guidelines ( EPA, 1999) include an expanded weight of evidence approach that emphasizes understanding mode of action ( MOA), conditions of expression of carcinogenicity ( e. g., route and magnitude of exposure) and consideration of all 11 other relevant data. A narrative with descriptors replaced the letter classifications used in 1986 guidelines. The descriptors in the draft Guidelines ( EPA, 1999) are: carcinogenic to humans;

likely to be carcinogenic to humans; suggestive evidence of carcinogenicity but not sufficient to assess human carcinogenic potential; data are inadequate for an assessment of human carcinogenic potential; and not likely to be carcinogenic to humans.

Under the 1999 Draft Guidelines, the preferred approach for dose response assessment is to use a biologically based dose response ( BBDR) model for extrapolating from animal to human doses, and for extrapolating to low doses, if the necessary data for the parameters used in such models are available. In the absence of such data, the default approach for calculating the human equivalent dose is to scale the daily applied lifetime oral dose in proportion to body weight raised to the 3/ 4 power ( BW3/ 4). In the absence of a BBDR, dose response assessment is done in two steps. First, appropriate models are fit to data in the empirical range of observation to determine a point of departure ( POD). A standard POD is the effective dose corresponding to the lower 95 percent limit on a dose associated with 10 percent increased tumor or relevant nontumor response ( LED10). In the second step, extrapolation is done below the observable range, to doses that are more characteristic of human exposures. One of several default approaches ( linear, nonlinear or both) can be used for low dose extrapolation, depending on the cancer MOA. The linear approach is used when there is an absence of sufficient tumor MOA information, the chemical has direct DNA mutagenic reactivity or other indications of DNA effects that are consistent with linearity, or the dose response relationship is expected to be linear. The nonlinear approach is used when a tumor MOA supports nonlinearity, and the chemical does not demonstrate mutagenic effects consistent with linearity, or the chemical has some indication of mutagenic activity, but it is 12 judged not to play a significant role in tumor causation. If the MOA indicates both linear and non linear dose responses, both linear and nonlinear approaches are used. The 1999 Draft Guidelines also provide additional guidance for the examination of the risk to children and other sensitive populations. Based on linear extrapolation from the POD, the cancer slope factor, in units of ( mg/ kg/ day) 1 can be calculated. The slope factor can be converted to a drinking water unit risk, in units of ( ug/ L) 1 by dividing by 70 kg, multiplying by 2 L/ day, and adjusting the units. 13 Uncertainty Factors Used in RfD Calculations Standard Uncertainty Factors ( UFs)

Use a 1, 3, or 10 fold factor when extrapolating from valid experimental results from studies using prolonged exposure to average healthy humans. This factor is intended to account for the variation in sensitivity among the members of the human population. The intermediate factor of 3 represents approximately ½ log 10 unit, ( i. e., the square root of 10).

Use an additional factor of 1, 3, or 10 when extrapolating from valid results of long term studies on experimental animals when results of studies of human exposure are not available or are inadequate. This factor is intended to account for the uncertainty in extrapolating animal data to risks for humans.

Use an additional factor of 1, 3, or 10 when extrapolating from less than chronic results on experimental animals when there are no useful long term human data. This factor is intended to account for the uncertainty in extrapolating from less than chronic NOAELs to chronic NOAELs.

Use an additional factor of 1, 3, or 10 when deriving an RfD from a LOAEL instead of a NOAEL. This factor is intended to account for the uncertainty in extrapolating from LOAELs to NOAELs.

Use an additional 3 or 10 fold factor when deriving an RfD from an " incomplete" data base. The minimum database for a high confidence RfD is: two systemic toxicity bioassays in different species, one two generation reproductive study and two developmental toxicity studies in different species. This factor is meant to account for the inability of any single type of study to consider all toxic endpoints. The intermediate factor of 3 is often used when there is a single data gap exclusive of chronic data.

The maximum composite UF for any given database is 3,000. Databases weaker than this are judged too uncertain to estimate an RfD.

Note: With each UF assignment, it is recognized that professional scientific judgment must be used. 14 1.3 DETERMINING RISK TO CHILDREN In developing the Stage 2 D/ DBP Rule, risks to sensitive subpopulations including fetuses and children were taken into account in the assessments of D/ DBPs. To determine whether fetuses and children are more sensitive than adults, the following questions were considered,

1. Is there information that shows that the D/ DBP causes effects in the developing fetus or harms a woman's ability to become pregnant and bear children? If it causes these effects, do these effects occur at lower doses than those that cause other types of effects?

2. If the D/ DBP causes a health effect other than cancer, such as an effect on the liver or kidney, are children affected at lower doses than are adults?

3. If the D/ DBP causes cancer, are children more likely to be affected by a given dose than are adults?

The ultimate goal of these questions is to determine whether the MCLG is protective of any putative special risk to children, regardless of whether the MCLG is based on developmental toxicity, systemic toxicity, or cancer effects.

The first question is addressed by directly comparing the dose response in developmental and reproductive toxicity studies with that in systemic toxicity and cancer studies. If the most sensitive endpoint for a chemical is a developmental or reproductive effect, the RfD is based on that endpoint. If adequate developmental or reproductive toxicity studies are not available, this data gap is taken into account in the choice of the database UF. 15 Ideally, to address questions 2 and 3 above, one would have data from a single study comparing the sensitivity of young animals and adult animals exposed under the same conditions and for the same duration. Such data are rarely available. In the absence of such a direct comparison, one can obtain some relevant information on relative sensitivities by comparing the NOAEL and LOAEL for effects in the first filial generation ( F 1 pups) in a two generation reproductive study with the NOAEL and LOAEL in a subchronic study. This is because the two study designs involve similar exposure durations, and the animals in the subchronic study are exposed as young adults while the F 1 pups in the reproductive study are exposed in utero through young adulthood. Therefore, if everything else were the same between studies, differences in toxicity could be attributed to age related differences in sensitivity. This approach provides a very crude comparison, since there may be differences between the subchronic and two generation studies, such as differences in dose spacing and in the strains used, complicating the comparison.

For most of the D/ DBPs, appropriate data from a two generation study were not available.

However, such a comparison was conducted for chloroform ( see Section <IP_ADDRESS>).

Since one rarely has adequate data to directly compare toxicity in adult and young animals, and the comparison described for chloroform is rather crude, questions 2 and 3 are usually addressed by considering all available data addressing the potential for age related differences in sensitivity, taking into account the factors noted in Section 1.1 that can lead to increased sensitivity of children. Age related differences in metabolism are identified where possible, and these differences are evaluated in the context of whether the parent or a metabolite is the toxic form. As noted earlier, interspecies differences in age related metabolic capabilities also need to be considered. Age related differences in metabolism was considered for the 16 trihalomethanes and the haloacetic acids, although a detailed analysis was possible only for the trihalomethanes, because neither the mode of action nor the toxic forms are known for the haloacetic acids. Where possible, potential age related toxicodynamic differences are also considered. Studies investigating toxicodynamic differences between young and adult animals are rare. However, the consideration for chloramine noted that infants may be more sensitive than adults to the hematological effects of this chemical.

An additional facet of these questions is consideration of whether differences between adults and children are so significant that systemic toxicity in children would occur at a lower dose than developmental toxicity, even though systemic toxicity in adults occurs at higher doses.

In other words, an additional question would be " does an RfD based on developmental effects as the critical effect for the available database adequately protect children from systemic effects?"

This concern can be addressed by comparing toxicity in the two generation and subchronic studies, as described above. In the absence of such data, recent assessments paying special attention to children's risk have begun to address this uncertainty using the database UF. For older assessments, this remains an uncertainty.

The third question addresses the potential for children to be more sensitive than adults to the carcinogenic effects of a chemical. To address this issue in general, one considers the chemical's mode of action and metabolic pathway to determine whether a higher cancer risk may be expected in children. Guidelines for quantitatively considering this issue are still in development. However, such quantitative consideration is not needed to determine whether MCLGs based on cancer are protective of children. MCLGs for genotoxic carcinogens are zero,

and so any MCLG for a genotoxic carcinogen would be protective of children. Chloroform is the 17 only D/ DBP that causes cancer, but has an MCLG other than 0, based on its mode of action. An analysis of the available data on chloroform metabolism and its mode of indicates that children are not expected to be at greater risk than adults from the carcinogenic effects of chloroform.

1.4. MAXIMUM CONTAMINANT LEVEL GOAL AND MAXIMUM RESIDUAL DISINFECTANT LEVEL GOAL The U. S. EPA is mandated to publish a maximum contaminant level goal ( MCLG) and promulgate a national primary drinking water regulation ( NPDWR) establishing a maximum contaminant level ( MCL) if it determines that the contaminant ( 1) may have an adverse effect on the health of persons; ( 2) is known to occur or there is a substantial likelihood that the contaminant will occur in public water systems with a frequency and at levels of public health concern; and ( 3) the Agency judges that regulation of the contaminant presents a meaningful opportunity for health risk reduction for persons served by public water systems.

As defined in the Safe Drinking Water Act amendments of 1996, the MCLG is the level of a contaminant in drinking water below which there is no known or expected risk to health.

MCLGs are nonenforceable health goals. They are set at concentration levels at which no known or anticipated adverse effects on the health of persons occur, and which include an adequate margin of safety. Establishment of an MCLG for a specific contaminant is based on the available evidence of carcinogenicity or noncancer adverse health effects from drinking water exposure using EPA's Guidelines for Risk Assessment. The Stage 1 Disinfectants/ Disinfection Byproducts ( Stage 1 DBP) Rule at 63 FR 69390 provides a detailed discussion of the process for establishing MCLGs. MCLGs can be based on an RfD for general systemic effects, on a quantitative estimate 18 for developmental or reproductive effects, or on a consideration of a cancer quantitative assessment. For carcinogenicity, when a linear low dose extrapolation is done, the MCLG is set at zero.

The maximum residual disinfectant level goal ( MRDLG) concept was introduced in the Stage 1 D/ DBP Rule to reflect the fact that these substances have beneficial disinfection properties. As with MCLGs, MRDLGs are established at the level at which no known or anticipated adverse effects on the health of persons occur and which allows an adequate margin of safety. MRDLGs are nonenforceable health goals based only on health effects and exposure information and do not reflect the benefit of the addition of the chemical for control of waterborne microbial contaminants.

The MCLGs and MRDLGs are used as the basis for setting the Maximum Contaminant Levels ( MCLs), which are the enforceable drinking water standards. MCLs are set as close to the MCLGs and MRDLGs as feasible, taking costs, treatment technology, and other considerations into account.

The MCLG or MRDLG for drinking water is calculated from the RfD for a 70 kg adult consuming 2 L of water per day and also takes into consideration the relative source contribution ( RSC) from drinking water. For risk assessment the Agency views the use of 2 L per day adult drinking water consumption as appropriate, because it represents the 84th to 90th percentile of U. S. adult drinking water consumption ( EPA, 2000b). The 90th percentile for community water consumption of young people aged 11 to 19 in the United States is 1.5 L/ day, and for children younger than 1 year old it is less than 1 L/ day, with a mean of 0.5 L/ day or less ( EPA, 2000b).

Although a conservative estimate of water intake for infants, expressed as the water intake: body 19 weight ratio, is estimated at more than 3 fold the estimate for adults. The MCLG/ MRDLG calculated from the RfD is generally assumed to be protective of children. This is for several reasons. The RfD is calculated to be protective of the human population including sensitive subgroups; UFs used in deriving the RfD include a factor for human variability in response.

Second, the RfD is protective of lifetime exposure at a certain level. For most chemicals, the severity of the effect increases with increasing exposure duration. This means that it might be safe for children to be exposed for a less than lifetime duration to a chemical at doses that are higher than the RfD ( on a body weight basis). This second consideration indicates that an important part of the consideration of children's risk is whether the effects begin to appear after only a short period of exposure, or whether lifetime exposure is required for the effects to appear. The critical effect for the RfDs for all of the D/ DBPs were based on studies of developmental effects,

or on effects in subchronic or chronic studies; none of the RfDs were based on a short term study.

The RfDs for three of the chemicals ( dibromochloromethane, bromoform, and dichloroacetic acid) were based on subchronic studies. This duration is long enough ( roughly 1/ 10 of the animal's life, corresponding to 7 years for humans) that a year of exposure at slightly more than the RfD appears unlikely to result in the critical effect. For the RfDs of some of the D/ DBPs based on developmental effects, the dose to the fetus is determined by maternal water consumption. For chlorite and chlorine dioxide, the critical effect was observed in both the F 1 and F 2 generations, and the critical exposure window was not identified.

A final consideration regarding the protectiveness of the RfD is that the RfD is defined as an estimate within " an order of magnitude." This is usually taken to mean that the " true" RfD may be a factor of three higher or lower. Because of this imprecision in the estimate of the RfD, 20 differences in dose within a factor of three are often considered to lie within the imprecision of the method. Based on all of these considerations, the MCLG calculated from the RfD is generally assumed to be protective of adverse effects over a lifetime exposure, and the Agency believes that the use of 2 L to calculate the MCLG provides sufficient protection to fetuses and children.

The RSC is derived by application of the exposure decision tree approach published in EPA's Methodology for Deriving Ambient Water Quality Criteria for the Protection of Human Health ( EPA, 2000f). The RSC is the fraction of an individual's total exposure allocated to drinking water. The Agency factors an RSC contribution into the MCLG to ensure that the contribution from drinking water does not cause the total exposure of an individual to a contaminant over a lifetime to exceed the contaminant's RfD. When data are sufficient, the Agency uses those data as the basis for the RSC from drinking water. If insufficient exposure data are available, a default assumption of a 20% drinking water contribution is used. The maximum value used for an RSC is 80%, to allow for the possibility of other sources of exposure.

The Stage 2 D/ DBP Rule proposes MCLGs for the disinfection byproducts chloroform ( a trihalomethane, or THM), and the haloacetic acids, monochloroacetic acid ( MCA) and trichloroacetic acid ( TCA). For the other D/ DBPs, the proposed Rule reflects MCLGs established in the Stage 1 D/ DBP Rule: bromodichloromethane ( BDCM), dibromochloromethane ( DBCM), bromoform, dichloroacetic acid ( DCA), bromate and chlorite. This document also addresses the health effects of some additional DBPs: monobromoacetic acid ( MBA),

bromochloroacetic acid ( BCA), dibromoacetic acid ( DBA) and 3 Chloro 4( dichloromethyl) 5

hydroxy 2( 5H) furanone ( MX); however, the data are insufficient to develop MCLGs for these 21 chemicals. The current and proposed MCLGs are shown in Table 1, and the MRDLGs for the disinfectants chlorine, chloramine and chlorine dioxide are shown in Table 2.

Table 3 summarizes the toxicity endpoints for the various D/ DBPs. BDCM, bromoform,

DCA, bromate, and MX are considered likely human carcinogens under the 1999 Draft Carcinogenic Risk Assessment Guidelines. MCLGs of zero were selected after consideration of the mode of action of these chemicals, except for MX, for which an MCLG has not yet been determined. The MCLG/ MRDLG values for chloroform, DBCM, MCA, TCA, chlorine, and chloramine were based on systemic toxicity. For chlorine dioxide and chlorite, the MCLG/ MRDLGs are calculated on data from neurodevelopmental studies. No MCLGs were derived for MX, MBA, BCA, and DBA, because the data are insufficient at this time. The analysis of the children's risk in relation to each MCLB/ MRDLG value is presented on a chemical by chemical basis in the following sections. 22 Table 1. Disinfection Byproducts and their MCLGs Considered in this Document Disinfection Byproducts Current MCLG ( mg/ L) Proposed MCLG ( mg/ L)

Chloroform 0 0.07 Bromodichloromethane ( BDCM) 0 0a Dibromochloromethane ( DBCM) 0.06 0.06a Bromoform 0 0a Monochloroacetic acid ( MCA)   b 0.03 Dichloroacetic acid ( DCA) 0 0a Trichloroacetic acid ( TCA) 0.3 0.02 Chlorite 0.8 0.8a Bromate 0 0a aNo ( new) value proposed in the Stage 2 D/ DBP rule. bNo current MCLG, based on the Stage 1 D/ DBP rule Table 2. Disinfectants and their MRDLGs Considered in this Document Disinfectants MRDLG ( mg/ L)

Chlorine 4 ( as Cl 2)

Chloramine 3 ( 4 mg/ L as Cl 2)

Chlorine dioxide 0.8 ( as ClO 2) 23 Table 3. Comparison of Toxicity Endpoints Disinfectant/

Disinfection Byproduct Systemic Toxicity ( mg/ kg/ day) Developmental Toxicity ( mg/ kg/ day) Carcinogenicity MCLG/ MRDLGb mg/ L NOAELa LOAELa NOAEL LOAEL Based on 1999 Guidelines Chloroform None ( BMDL10 1.0) 12.9 35   50 126 likely to be carcinogenic to humans under high exposure conditionsc 0.07 ( sys tox)

Bromodichloromethane ( BDCM) None ( BMDL10 0.8) 6d 25 50 likely to be carcinogenic to humans 0 ( Ca)

Dibromochloromethane ( DBCM) None ( BMDL10 1.6) 28d 17 171 suggestive evidence of human carcinogenicity, but the data are not sufficient to assess human carcinogenic potential. 0.06 ( sys tox & Ca)

Bromoform None ( BMDL10 2.6) 36d 50 100 likely to be carcinogenic to humans 0 ( Ca)

Monochloroacetic acid ( MCA) ND 3.5 70e 140 data are inadequate for an assessment of human carcinogenic potential 0.03 ( sys tox)

Dichloroacetic acid ( DCA) ND 12.5 14 140 likely to be carcinogenic to humans 0 ( Ca)

Trichloroacetic acid ( TCA) 32.5 364 ND 291 suggestive evidence of carcinogenicity, but the data are not sufficient to assess human carcinogenic potential 0.02 ( sys tox)

Monobromoacetic acid ( MBA) ND ND 50e 100 data are inadequate for an assessment of human carcinogenic potential NDf Bromochloroacetic acid ( BCA) 15g 39 19 50 data are inadequate for an assessment of human carcinogenic potential NDf Disinfectant/

Disinfection Byproduct Systemic Toxicity ( mg/ kg/ day) Developmental Toxicity ( mg/ kg/ day) Carcinogenicity MCLG/ MRDLGb mg/ L NOAELa LOAELa NOAEL LOAEL Based on 1999 Guidelines 24 Dibromoacetic acid ( DBA) 2 10 50 d 100 suggestive evidence of carcinogenicity, but the data are not sufficient to assess human carcinogenic potential NDf Bromate 1.1 6.1 7.7 22 likely to be carcinogenic to humans 0 ( Ca)

MX ND ND ND ND likely to be carcinogenic to humans NDf Chlorite ND ND 3 6 data are inadequate for an assessment of human carcinogenic potential 0.8 ( devel tox)

Chlorine dioxide ND ND 3 6 data are inadequate for an assessment of human carcinogenic potential 0.8 ( devel tox)

Chlorine 14 ND 5h ND data are inadequate for an assessment of human carcinogenic potential 4 ( sys tox)

Chloramine 9.5 ND 10 ND data are inadequate for an assessment of human carcinogenic potential 3 ( sys tox) i aNOAEL = No observed adverse effect level; LOAEL = Lowest observed adverse effect level; ND = not determined; BMDL 10 = 95% lower confidence limit on the benchmark dose for 10% extra risk.

Where appropriate, duration adjusted values shown bMCLG = Maximum contaminant level goal; ( Ca) = basis for MCLG is carcinogenic effects; ( sys tox) = basis for MCLG is systemic toxic effects; ( devel tox) = basis for MCLG is developmental effects.

cLikely to be carcinogenic to humans under high exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues, but not likely to be carcinogenic to humans at a dose level that do not cause these effects.

dNo NOAEL was identified for the critical effect for BDCM or DBCM; the NOAEL for bromoform was 18.

eAdverse effect levels reported in a published abstract only, and thus should be considered preliminary.

fInsufficient data for derivation of a MCLG.

gAdverse effect level for systemic effects in a short term study; no subchronic or chronic studies were available.

hHighest dose tested.

i 4mg/ L measured as total chlorine 25 2. ESTIMATES OF RISK TO CHILDREN FOR STAGE II DISINFECTANTS/

DISINFECTION BYPRODUCTS A variety of different types of data are used for characterizing the risks to children from D/ DBPs. This chapter begins by reviewing the relevant available epidemiology data, focusing on studies evaluating the potential developmental and reproductive effects from consumption of chlorinated drinking water. The epidemiology data on cancer risk from consumption of chlorinated drinking water are also briefly reviewed. This chapter then reviews the available data for each D/ DBP, beginning with information on metabolism when available, and addressing the D/ DBPs' developmental/ reproductive effects, systemic effects, and carcinogenic potential. The section for each D/ DBP concludes with a discussion of the derivation of the MCLG/ MRDLG for that chemical, and an evaluation of children's risk relative to the MCLG/ MRDLG.

2.1. CHLORINATED DRINKING WATER This section considers epidemiology studies on exposure to chlorinated drinking water rather than to individual D/ DBPs. There are no reliable studies on the reproductive or developmental toxicity of chlorinated drinking water in animals. While several reproductive and developmental toxicity studies have been conducted with chlorinated drinking water, they either used distilled water or chlorinated tap water ( EPA, 1994d). Since the composition of D/ DBPs in drinking water depends on source water characteristics and the chlorination procedures used,

neither of these approaches may adequately represent the composition of chlorinated drinking water. Animal studies are not available to address whether there are age related differences in the 26 systemic response to chlorinated drinking water. In addition, animal carcinogenicity studies of the D/ DBPs did not evaluate the effects of perinatal or in utero exposure.

2.1.1. Developmental/ Reproductive Effects The available epidemiologic studies on the developmental and reproductive effects from exposure to D/ DBPs in drinking water may be divided into two major groups: 1) qualitative exposure assessment studies that examine associations between the source of the water supply or disinfection method and the risk of adverse developmental effects; and 2) quantitative exposure studies that associate adverse reproductive outcomes with reported concentrations of D/ DBPs in water supplies. Developmental/ reproductive outcomes examined in both types of studies include low birth weight, intrauterine growth retardation/ small for gestational age, pre term delivery,

spontaneous abortion, stillbirth, and birth defects ( Reif et al., 2000).

Qualitative exposure studies have investigated a number of developmental/ reproductive outcomes. These studies typically do not allow the evaluation of a dose response and are limited because the effects cannot be attributed to a single contaminant or group of contaminants.

One study ( Yang et al., 2000) examined the relationship between drinking water disinfection and low birth weight, but found no statistically significant effects. The study authors conducted a study in Taiwan of the association between chlorination of drinking water and low birth weight. They compared the incidence of low birth weight in 14 cities in which chlorinated water supplied over 90 percent of the residents with the rate in 14 matched cities that used nonchlorinated water. They examined records on 18,025 births and found no association between consumption of chlorinated drinking water and low birth weight (< 2500 grams). They did 27 observe an increase in pre term deliveries (< 37 weeks) in those cities using chlorinated water ( odds ratio [ OR] 1.34, 95% confidence interval [ CI] 1.15 1.56).

Several qualitative studies compared differences in birth outcomes, including birth weight,

in populations served with untreated drinking water and water treated using differing methods of disinfection. Källen and Robert ( 2000) evaluated differences in birth outcome in Sweden based on approximately 74,300 control births, 15,400 births in areas using chlorine dioxide treated water and 24,700 births in areas served by water treated with sodium hypochlorite. Exposure data were based on published reports of drinking water disinfection practices in the municipalities at the time of the births ( 1985  1994). Two effects were significantly related to water chlorinated with sodium hypochlorite: short body length (< 43 cm, OR 1.97 95% CI 1.30 2.97) and small head circumference (< 31 cm, OR 1.46, 95% CI 1.07 1.98). In addition, increases were seen in pre term delivery (< 32 weeks gestation, OR 1.22, 95% CI 1.00 1.48) and low birth weight (< 1500 g, OR 1.11, 95% CI 0.90 1.36); the increases were not statistically significant. These effects were not observed more frequently for births in areas served by water treated with chlorine dioxide. There was no difference between the exposed and unexposed groups for neonatal death,

neonatal jaundice, congenital malformations or childhood cancer.

Kanitz et al. ( 1996) examined 548 births during 1988 1989 in Italy from women in a community exposed to filtered water disinfected with chlorine dioxide, sodium hypochlorite or both and compared them with 128 births from women in an Italian community using untreated water. Total trihalomethane levels were 8 16 ppm in the water treated with sodium hypochlorite,

and 1 3 ppb in the water treated with chlorine dioxide. Levels of chlorine dioxide in the water immediately after treatment were less than 0.3 mg/ L and chlorine residue was less than 0.4 mg/ L. 28 Infants born to women in the chlorine dioxide disinfected community had smaller cranial circumference ( OR = 2.2, 95% CI = 1.4 3.9) and smaller body length ( OR = 2.0, 95% CI = 1.2

3.3) as compared to infants born to women in the untreated water community. Statistically nonsignificant increases in low birth weight infants (

2,500 g) and preterm deliveries were reported in women living in the chlorine dioxide disinfected area. Mothers exposed to water treated with chlorine dioxide, but not to sodium hypochlorite treated water, had a significantly higher frequency of newborns with neonatal jaundice. The authors concluded that infants of women who consumed drinking water treated with chlorine dioxide or sodium hypochlorite during pregnancy were at higher risk for neonatal jaundice, cranial circumference

35 cm, and body length

49.5 cm. However, potential confounding factors, such as the lack of information on the quantity of water consumed during the pregnancy, exposure to other chemicals in the water, nutritional and smoking habits and age distribution of the women, limit the conclusions that can be drawn from this study ( EPA, 2000c).

Other qualitative studies looked for effects of chlorination on early miscarriages ( spontaneous abortions), stillbirths or birth defects. In a case control study based in the Boston metropolitan area, Aschengrau et al. ( 1989) found an increased risk of early miscarriage for surface water consumption versus ground water or mixed water consumption ( OR 2.2, 95% CI 1.3 3.6). No significant difference was found for chlorinated versus chloraminated water supplies.

Aschengrau et al. ( 1993) conducted an additional case control study of drinking water quality and the occurrence of adverse pregnancy outcomes; this included 1,039 congenital anomalies, 77 stillbirths, and 55 neonatal deaths among women who delivered infants during August 1977 through March 1980 in Massachusetts. After adjustment for confounding, the frequency of 29 stillbirths was increased ( but not statistically significant) for women exposed to chlorinated surface water ( OR 2.6, 95% CI 0.9 7.5). In addition, the risk of major malformations was also increased ( OR 1.5, 95% CI 0.7 2.1), but the increase was not statistically significant. The increase in major malformations was mainly due to a significantly increased risk of respiratory tract defects ( OR 3.2, 95% CI 1.1 9.5) and urinary tract defects ( OR 4.1, 95% CI 1.2 14.1).

Magnus et al. ( 1999) conducted an ecologic study in Norway on the relationship between the consumption of chlorinated drinking water and birth defects in the period 1993 1995. They used data from 1994 on water quality and disinfection practice. Water color was used as an indicator for natural organic matter content. Of 141,077 births in the study, 1.8% had birth defects. A significant increase in urinary tract defects ( OR 1.99, 95% CI 1.10 3.57) was seen when the high color ( high organic matter) plus chlorination group was compared with those with low color in the drinking water plus no chlorination. There were nonsignificant increases in all malformations ( OR 1.14, 95% CI 0.99 1.31) and neural tube defects ( OR 1.26, 95% CI 0.61

2.62) in the same groups. In an additional analysis of this same population reported in a published abstract ( Jaakkola et al., 1999), the risk of low birth weight, intrauterine growth retardation, or prematurity was not increased for individuals exposed to chlorinated water.

Several quantitative exposure studies examined the association between D/ DBPs and preterm delivery, low birth weight and/ or other adverse developmental outcomes. Savitz et al.

( 1995) evaluated risks of miscarriage ( n= 261), pre term delivery ( n= 413) or low birth weight ( n= 301) for total trihalomethane ( TTHM) exposure in a case control study in North Carolina.

Exposure in this study was evaluated by linking the women's dates of pregnancy to the nearest quarterly average TTHM level obtained from the water supplier. This was multiplied by the 30 average number of glasses of water consumed per day, corresponding to the fourth week of pregnancy for miscarriage cases and the 28th week of pregnancy for preterm delivery and low birth weight cases. Women in the highest sextile ( sixth) of exposures had an increased risk of miscarriages ( OR 2.8, 95% CI 1.2 6.1), but no such association was evident in the second highest sextile ( OR 0.2, 95% CI 0.0 0.5) or when the data were analyzed by exposure tertiles ( thirds).

The risk for pre term delivery was not increased, and the slight increase in risk for low birth weight in the upper tertile was not statistically significant.

Gallagher et al. ( 1998) evaluated associations between estimated TTHM concentrations in municipal water supplies and pre term delivery, low birth weight, and term low birth weight in a retrospective cohort study of 1,893 births. TTHM levels were modelled based on hydraulic characteristics of the water system and monitored TTHM levels in the distribution system.

Exposures were assigned to study participants based on the estimated exposure during the women's third trimester of pregnancy. The exposure categories for estimated TTHM concentrations were

20 µ g/ L ( referent category), 21 40 µ g/ L, 41 60 µ g/ L and  61 µ g/ L.

Women exposed to TTHM concentrations  61 µ g/ L had an increased risk for low birth weight ( OR 2.1, 95% CI 1.0 4.8) and term low birth weight ( OR 5.9, 95% CI 2.0 17.0), although the number of cases of term low birth weight was small for the exposed mothers ( n= 6).

Nuckols et al. ( 1995) conducted a pilot study in Colorado evaluating the use of geographic information system ( GIS) technology for analyzing drinking water epidemiology data.

They investigated the relationship between water disinfection in two public water systems: one using chloramination resulting in very low levels of trihalomethanes ( THMs); and the other using chlorination resulting in higher levels of THMs. Health outcome data, including birth weight and 31 gestational age for live births in two Colorado counties in 1990, was obtained from birth certificates. Exposure classification was based on the census block group of residence of the mother at the time of delivery and actual monitoring data for THMs. Preliminary analyses indicated no significant association between low birth weight and water disinfection. However,

the proportion of babies who were pre term (< 38 weeks of gestation) was significantly higher ( relative risk 0.54, 95% CI 0.33 0.88) in the chloraminated water district.

Two studies investigated early term miscarriage risk factors. The first of these is a quantitative exposure study that examined the potential association between early term miscarriage and exposure to THMs ( Waller et al., 1998). The second is a qualitative study that examined the potential association between early term miscarriage and tap water consumption ( Swan et al., 1998). Both studies used the same group of 5,144 pregnant women living in three areas of California. In the Waller et al. ( 1998) study, additional water quality information from the women's drinking water utilities were obtained so that THM levels could be determined. The Swan et al. ( 1998) study provided no quantitative measurements of THMs ( or DBPs), and thus provided no additional information on the risk from chlorination byproducts.

In the Waller et al. ( 1998) study, water utilities that served the population provided THM measurements taken during the time period when participants were pregnant. The TTHM level in a participant's home tap water was estimated by averaging water distribution system TTHM measurements taken during a participant's first 3 months of pregnancy; a similar approach was followed for estimating concentrations of individual THMs. Women who drank five or more glasses per day of cold home tap water containing at least 75 µ g/ L of TTHM were considered to be in the high TTHM group. Incidence of early term miscarriage in this group was 15.7%, 32 compared with an incidence of 9.5% among women with low TTHM exposure ( drinking fewer than five glasses per day of cold home tap water or drinking any amount of tap water containing less than 75 µ g/ L of TTHM). An adjusted odds ratio for early term miscarriage of 1.8 ( 95% CI 1.1   3.0) was determined. Only high BDCM exposure was associated with early term miscarriage ( adjusted OR = 2.0, 95% CI = 1.2   3.5). This was defined as drinking five or more glasses per day of cold home tap water containing > 18 µ g/ L BDCM. After adjustment for exposure to other THMs, the adjusted odds ratio for early term miscarriage was 3.0 ( 95% CI 1.4   6.6). The authors noted several potential limitations of these data, including misclassification of exposure, the fact that concentration levels for most subjects were based on test results from a single day, and that THM exposure from sources other than ingestion could not be fully characterized. In a follow up study, Waller et al. ( 2001) reanalyzed the exposure data from this study using several different methods. One goal of the reanalysis was to reduce exposure misclassification. The study authors reported a positive dose response relationship between spontaneous abortion rate and an exposure metric incorporating total trihalomethanes and personal ingestion. However, the authors noted that it was not possible to determine whether the reanalysis actually reduced exposure misclassification. This reanalysis did not address dose response relationships between individual trihalomethanes and occurrence of spontaneous abortion.

Additional quantitative exposure studies have investigated the association between water chlorination and birth defects, stillbirths and other adverse developmental effects. Bove et al.

( 1992) conducted a cross sectional study for four northern New Jersey counties to evaluate associations between total trihalomethane ( TTHM) concentrations in water supplies and reported birth weight, fetal deaths and birth defects. The study population consisted of 594 cases of 3This was the only confidence interval reported by the authors for this endpoint.

33 stillbirths and 669 cases of birth defects. Birthweight, fetal deaths and birth defects were evaluated by linking quarterly TTHM measurements from the water utilities with the mother's address at the birth of the child. Exposures greater than 80 µ g/ L TTHM were significantly associated with term low birth weight ( OR 1.29, 95% CI 1.08 1.5), and small for gestational age ( OR 1.14, 95% CI 1.04 1.3), although the association was weak. No association was reported for very low birth weight, stillbirth or pre term birth. Significantly elevated odds ratios were reported for birth defects, including all defects ( OR 1.53, 95% CI 1.14 2.1), central nervous system defects ( OR 2.52, 95% CI 1.40 4.5), and neural tube defects ( OR 2.98, 95% CI 1.25 7.1).

The odds ratios for oral cleft defects ( OR 1.74, 95% CI 0.88 3.4) and major cardiac defects ( OR 1.84, 95% CI 0.95 3.6) were also elevated, but the increases were not statistically significant.

This same data set was reported in a published manuscript that included analysis using an exposure level of 100 µ g/ L TTHM as the criterion to separate subjects into high and low exposure groups ( Bove et al., 1995). The association between TTHM exposure and term low birth weight ( OR 1.42, 50% CI 1.22 1.65) 3 and small for gestational age ( OR 1.50, 90% CI 1.19  1.86) was significant, and the strength of the association increased with the higher criterial exposure level. A mean decrease in term birthweight of 70.4 grams was associated with TTHM exposures greater than 100  g/ L.

Klotz and Pyrch ( 1999) followed up on the work of Bove and colleagues ( Bove et al.,

1992; 1995) and examined the potential association between neural tube defects and drinking water contaminants including trihalomethanes, haloacetonitriles and haloacetic acids. In this population based case control study, 112 births with neural tube defects reported to New Jersey's 34 Birth Defects Registry in 1993 and 1994 were matched against control births chosen randomly from across the state. Birth certificates were examined for all subjects, as were drinking water data corresponding to the mother's residence in early pregnancy. The drinking water data was obtained by analyzing tap water for haloacetic acids, haloacetonitriles, THMs, total chlorine and free chlorine, four months after the mother's due date. This sampling was intended to correspond to the critical period ( one year earlier) of neural tube development in the fetus ( i. e., taking into account seasonal variation in the concentrations of these chemicals in drinking water). In addition, TTHM exposure during the gestational period was estimated based on records from the water utility. The authors reported elevated odds ratios, generally between 1.5 and 2.1, for the association of neural tube defects with TTHMs. The only statistically significant results were seen when the analysis was isolated to those subjects with the highest THM exposures ( greater than 40 ppb) and was limited to those subjects with only neural tube defects and no other malformations ( OR 2.1, 95% CI 1.1   4.0).

Dodds and King ( 2001) conducted a retrospective cohort study of the relationship between BDCM exposure and birth defects among 49,842 residents of Nova Scotia, Canada between 1988 and 1995. Exposure was estimated from routine water monitoring samples collected from within the water distribution system. The birth defects examined had previously been reported in other epidemiological studies and included neural tube defects, cardiovascular defects, cleft defects, and chromosomal abnormalities. Exposure windows were selected to target the period before or during gestation when exposure to a potential developmental toxicant or mutagen might have the most profound effect on a particular developmental or genotoxic endpoint. Maternal age, previous births to the mother, maternal smoking, and neighborhood 35 family income were assessed as potential confounders. Exposure to BDCM at concentrations  20 µ g/ L was associated with increased risk of neural tube defects ( adjusted relative risk 2.5;

95% confidence interval 1.2 to 5.1) when adjusted for maternal age and income level. However,

there was no evidence of a concentration response trend. In addition, the study authors noted that this point estimate was " fairly unstable" as a result of the low number of cases ( n= 10) in the  20 µ g/ L exposure category. There was no apparent trend or significant association for exposure to BDCM and occurrence of cleft defects or chromosomal aberrations. A related earlier study using similar exposure measurement methods and a cohort of 93,000 women who delivered singleton births evaluated the relationship between pregnancy outcomes and TTHM exposure ( Dodds et al., 1999). This earlier study found an elevated relative risk for stillbirths in subjects with exposure to > 100 µ g/ L TTHM during pregnancy ( OR, but no significant increases in low birth weight, very low birth weight, small for gestational age, pre term delivery, or congenital anomalies.