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
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iii
LIST
OF
TABLES
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vi
ABBREVIATIONS
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viii
EXECUTIVE
SUMMARY
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ix
PLAIN
LANGUAGE
SUMMARY
OF
EVALUATION
OF
CHILDREN'S
RISK
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xxxiv
1.
INTRODUCTION
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1
1.1.
RISK
TO
CHILDREN
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1
1.2.
RISK
ASSESSMENT
METHODS
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6
1.3
DETERMINING
RISK
TO
CHILDREN
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14
1.4.
MAXIMUM
CONTAMINANT
LEVEL
GOAL
AND
MAXIMUM
RESIDUAL
DISINFECTANT
LEVEL
GOAL
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17
2.
ESTIMATES
OF
RISK
TO
CHILDREN
FOR
STAGE
II
DISINFECTANTS/
DISINFECTION
BYPRODUCTS
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25
2.1.
CHLORINATED
DRINKING
WATER
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25
2.1.1.
Developmental/
Reproductive
Effects
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26
2.1.2.
Systemic
Effects
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40
2.1.3.
Carcinogenicity
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40
2.2.
TRIHALOMETHANES
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42
2.2.1.
Chloroform
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42
2.2.1.1.
Developmental/
Reproductive
Effects
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42
2.2.1.2.
Systemic
Effects
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48
2.2.1.3.
Carcinogenicity
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48
2.2.1.4.
Basis
for
RfD
and
MCLG
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53
2.2.1.5.
Children's
Risk
in
Relation
to
the
MCLG
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55
2.2.2.
Brominated
Trihalomethanes
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56
2.2.2.1.
Bromodichloromethane
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64
Developmental/
Reproductive
Effects
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64
Systemic
Effects
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74
Carcinogenicity
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74
Basis
for
RfD
and
MCLG
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77
Children's
Risk
in
Relation
to
the
MCLG
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77
2.2.2.2.
Dibromochloromethane
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78
Developmental/
Reproductive
Effects
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78
iv
Systemic
Effects
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79
Carcinogenicity
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80
Basis
for
RfD
and
MCLG
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83
Children's
Risk
in
Relation
to
the
MCLG
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83
2.2.2.3.
Bromoform
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84
Developmental/
Reproductive
Effects
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84
Systemic
Effects
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86
Carcinogenicity
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87
Basis
for
RfD
and
MCLG
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89
Children's
Risk
in
Relation
to
the
MCLG
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90
2.3.
HALOACETIC
ACIDS
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90
2.3.1.
Monochloroacetic
Acid
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92
Developmental/
Reproductive
Effects
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92
Systemic
Effects
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95
Carcinogenicity
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95
Basis
for
RfD
and
MCLG
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97
Children's
Risk
in
Relation
to
the
MCLG
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97
2.3.2.
Dichloroacetic
Acid
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98
Developmental/
Reproductive
Effects
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98
Systemic
Effects
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102
Carcinogenicity
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106
Basis
for
RfD
and
MCLG
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108
Children's
Risk
in
Relation
to
the
MCLG
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109
2.3.3.
Trichloroacetic
Acid
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111
Developmental/
Reproductive
Effects
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111
Systemic
Effects
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113
Carcinogenicity
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114
Basis
for
the
RfD
and
MCLG
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117
Children's
Risk
in
Relation
to
the
MCLG
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118
2.3.4.
Monobromoacetic
Acid
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118
Developmental/
Reproductive
Effects
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118
Systemic
Effects
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120
Carcinogenicity
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120
Basis
for
RfD
and
MCLG
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121
Children's
Risk
in
Relation
to
the
MCLG
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121
2.3.5.
Bromochloroacetic
Acid
.
.
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121
Developmental/
Reproductive
Effects
.
.
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121
Systemic
Effects
.
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126
Carcinogenicity
.
.
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127
Basis
for
RfD
and
MCLG
.
.
.
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127
Children's
Risk
in
Relation
to
the
MCLG
.
.
.
.
.
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.
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127
2.3.6.
Dibromoacetic
Acid
.
.
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128
Developmental/
Reproductive
Effects
.
.
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.
128
Systemic
Effects
.
.
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.
138
v
Carcinogenicity
.
.
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.
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.
140
Basis
for
RfD
and
MCLG
.
.
.
.
.
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141
Children's
Risk
in
Relation
to
the
MCLG
.
.
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.
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142
2.4.
BROMATE
.
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.
143
2.4.1.
Developmental/
Reproductive
Effects
.
.
.
.
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.
143
2.4.2.
Systemic
Toxicity
.
.
.
.
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.
145
2.4.3.
Carcinogenicity
.
.
.
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.
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.
.
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.
.
148
2.4.4.
Basis
for
the
RfD
and
MCLG
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
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.
150
2.4.5.
Children's
Risk
in
Relation
to
the
MCLG
.
.
.
.
.
.
.
.
.
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.
.
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.
.
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.
.
.
150
2.5.
CHLORITE/
CHLORINE
DIOXIDE
.
.
.
.
.
.
.
.
.
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.
.
151
2.5.1.
Developmental/
Reproductive
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
152
2.5.2.
Systemic
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
159
2.5.3.
Carcinogenicity
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
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.
.
.
.
161
2.5.4.
Basis
for
RfD,
MCLG,
and
MRDLG
.
.
.
.
.
.
.
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.
.
.
.
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.
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.
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.
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.
.
163
2.5.5.
Children's
Risk
in
Relation
to
the
MCLG
and
MRDLG
.
.
.
.
.
.
.
.
.
.
.
163
2.6.
CHLORINE
.
.
.
.
.
.
.
.
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.
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.
.
.
164
2.6.1.
Developmental/
Reproductive
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
165
2.6.2.
Systemic
Toxicity
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
.
.
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.
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.
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.
.
167
2.6.3.
Carcinogenicity
.
.
.
.
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.
.
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.
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.
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.
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.
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.
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.
.
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.
.
.
.
.
167
2.6.4.
Basis
for
RfD
and
MRDLG
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
168
2.6.5.
Children's
Risk
Relative
to
the
MRDLG
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
168
2.7.
CHLORAMINE
.
.
.
.
.
.
.
.
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.
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.
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.
.
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.
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.
.
169
2.7.1.
Developmental/
Reproductive
Effects
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
169
2.7.2.
Systemic
Toxicity
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
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.
.
169
2.7.3.
Carcinogenicity
.
.
.
.
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.
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.
.
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.
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.
.
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
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
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.
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.
.
.
175
2.8.3.
Carcinogenicity
.
.
.
.
.
.
.
.
.
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.
.
.
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.
.
.
.
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.
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.
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.
.
.
.
.
.
.
.
.
.
.
175
2.8.4.
Basis
for
RfD
and
MCLG
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
.
.
.
.
.
177
2.8.5.
Children's
Risk
in
Relation
to
the
MCLG
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
.
.
.
.
177
3.
SUMMARY
AND
CONCLUSIONS
.
.
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.
178
4.
REFERENCES
.
.
.
.
.
.
.
.
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.
181
vi
LIST
OF
TABLES
Table
ES­
1.
Comparison
of
Toxicity
Endpoints
.
.
.
.
.
.
.
.
.
.
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.
.
.
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.
.
.
.
.
.
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
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
.
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.
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.
.
.
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
2.2.1.3).

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.

Shaw
and
colleagues
investigated
the
association
between
tap
water
consumption
and
cardiac
defects
(
Shaw
et
al.,
1990).
They
interviewed
145
mothers
of
children
with
severe
congenital
cardiac
disease
and
176
mothers
of
children
with
no
such
disease.
Subjects
were
asked
about
their
consumption
of
drinking
water
from
tap
and
bottled
sources
during
their
first
trimester
of
pregnancy.
A
relationship
was
found
between
maternal
consumption
of
4
or
more
glasses
of
tap
water
per
day
and
increases
in
cardiac
abnormality
in
the
infants
(
OR
2.0,
95%
CI
1.0­
4.0).
A
follow­
up
study
was
conducted
by
Shaw
et
al.
(
1991)
to
assess
the
relationship
between
water
chlorination
and
cardiac
defects.
In
this
study,
the
source
of
drinking
water
during
the
first
trimester
was
determined
for
each
of
the
138
study
participants.
THM
concentrations
36
corresponding
to
the
first
trimester
of
pregnancy
were
estimated
from
quarterly
utility
monitoring
records.
No
relationship
between
cardiac
abnormalities
and
chlorinated
drinking
water
was
identified,
and
average
THM
levels
during
the
first
trimester
of
pregnancy
for
cases
(
64.0
µ
g/
L)

was
less
than
for
controls
(
74.2
µ
g/
L).

Kramer
et
al.
(
1992)
conducted
a
population­
based
case­
control
study,
based
on
an
examination
of
birth
certificates
from
January
1,
1989
to
June
30,
1990
for
live
infants
born
to
white
women
in
Iowa
towns
with
1,000
to
5,000
inhabitants.
This
study
reported
an
increase
in
intrauterine
growth
retardation
with
exposure
to
chloroform
concentrations
in
drinking
water
of
greater
than
or
equal
to
10
µ
g/
L
(
OR
1.8,
95%
CI
1.1
 
2.9).
The
authors
also
found
a
nonsignificant
increase
in
intrauterine
growth
retardation
associated
with
exposure
to
drinking
water
with
BDCM
concentrations
greater
than
or
equal
to
10
µ
g/
L
compared
with
drinking
water
with
undetectable
BDCM
concentrations
(
OR1.7,
95%
CI
0.9
 
2.9).
THM
exposures,
including
chloroform
and
BDCM,
were
estimated
from
a
water
supply
survey
conducted
two
to
three
years
previously.

King
et
al.
(
2000a)
conducted
a
retrospective
cohort
study
of
50,000
deliveries
and
reported
that
exposure
to
TTHMs,
chloroform
and
BDCM
was
associated
with
an
increased
incidence
of
stillbirth.
Exposure
was
estimated
by
linking
the
mother's
residence
at
the
delivery
date
to
the
levels
of
THMs
monitored
in
public
water
supplies,
using
the
predicted
average
exposure
levels
for
the
entire
duration
of
the
pregnancy.
For
chloroform,
the
adjusted
odds
ratio
for
stillbirth
was
increased
for
exposure

100
µ
g./
L
(
OR
1.56,
95%
CI
1.04­
2.34).
Risk
doubled
for
women
exposed
to
a
BDCM
level
of
greater
than
or
equal
to
20
µ
g/
L,
when
compared
to
women
consuming
concentrations
of
less
than
5
µ
g/
L.
When
categories
of
stillbirth
(
unexplained
37
deaths
and
asphyxia­
related
deaths)
were
examined,
relative
risk
estimates
for
asphyxia­
related
deaths
increased
by
32%
for
each
10

g/
L
increase
in
concentration
of
BDCM.
As
indicated
by
King
et
al.
(
2002a),
an
in
vitro
study
(
Alston
1991)
suggested
that
chloroform
and
related
compounds
might
be
linked
to
the
asphyxia­
related
deaths
by
influencing
the
methioninehomocysteine
metabolic
pathway
and
subsequent
abruption
of
placenta.
These
data
support
the
asphyxia­
related
deaths
observed
by
King
et
al.
(
2002a),
and
indicate
that
the
observed
stillbirths
were
related
to
BDCM,
despite
their
limitation
to
one
type
of
stillbirth.
A
limitation
of
this
study
is
possible
misclassification
of
exposure
as
a
result
of
mobility
within
the
study
population.

Windham
et
al.
(
2003)
examined
menstrual
cycle
characteristics
in
relation
to
the
presence
of
brominated
trihalomethanes
in
tap
water
in
a
prospective
study
of
women
living
in
Northern
California.
Data
were
also
reported
for
TTHMs
and
for
chloroform.
The
target
population
was
married
women
of
reproductive
age
(
18­
39
years
old).
Participants
were
selected
from
among
nearly
6500
women
using
a
short
screening
interview
to
identify
women
who
were
more
likely
to
become
pregnant
(
i.
e.,
those
who
reported
a
menstrual
period
within
the
last
six
weeks,
were
not
surgically
sterilized,
did
not
use
birth
control
pills
or
intra­
uterine
devices,
and
were
noncontracepting
for
less
than
3
months).
Out
of
1092
eligible
women,
a
total
of
403
women
finished
the
study.
These
participants
collected
first
morning
urine
samples
daily
for
2­
9
menstrual
cycles
(
average
5.6
cycles)
for
measurement
of
steroid
metabolites.
The
participants
filled
out
a
daily
diary
during
the
urine
collection
phase
and
recorded
vaginal
bleeding.
These
measurements
(
diary
and
urinary
hormone
metabolites)
were
used
to
estimate
menstrual
parameters
such
as
cycle
and
phase
length.
Cycle
length
was
calculated
from
the
first
day
of
menses
to
the
day
before
onset
of
the
next
menses.
When
the
available
data
permitted,
the
cycle
was
divided
into
the
follicular
38
phase
(
first
day
of
menses
through
estimated
day
of
ovulation)
and
the
subsequent
luteal
phase.

Between
1424
and
1714
cycles
were
available
for
evaluation
of
each
parameter.
Information
on
water
consumption
and
other
variables
(
age,
race,
education,
employment,
income,
pregnancy
history,
exercise
type
and
frequency,
smoking,
alcohol
and
caffeine
consumption)
was
collected
in
a
baseline
telephone
interview
prior
to
urine
collection.
Information
on
the
number
of
showers
taken
at
home
per
week
and
their
duration
was
also
collected.
Exposure
to
trihalomethanes
was
estimated
from
quarterly
utility
monitoring
data
and
information
on
drinking
water
and
other
tap
water
use
collected
during
the
baseline
interview.

A
monotonic
decrease
in
mean
cycle
length
was
observed
with
increasing
TTHM
exposure
category.
At
TTHM
concentrations
greater
than
60
µ
g/
L,
the
adjusted
decrement
in
cycle
length
was
1.1
day
(
95%
C.
I.
­
1.8,
­
0.40)
when
compared
to
TTHM
concentrations
of
40
µ
g/
L
or
less.
The
decrease
in
follicular
phase
length
was
similar
(­
0.94
day;
95%
C.
I.
­
1.6,
­
0.24).

A
unit
decrement
in
mean
cycle
and
follicular
phase
length
of
0.18
days
per
10
µ
g/
L
increase
in
TTHM
concentration
(
95%
C.
I.
­
0.29,
­
0.07)
was
determined
when
the
cycle­
specific
TTHM
level
was
examined
as
a
continuous
variable.
Examination
of
time
spent
showering
did
not
reveal
additional
risks
with
longer
showers.
Combined
with
TTHM
concentration,
decrements
in
cycle
and
follicular
phase
length
were
seen
at
the
higher
TTHM
(>
60
µ
g/
L)
and
longer
showers
(

70
minutes)
categories
(­
1.2
and
 
1.6
days
respectively).
However,
the
confidence
intervals
were
wide
for
all
duration
categories
and
a
clear
dose
response
pattern
(
i.
e.,
shorter
lengths
at
higher
durations)
was
not
evident.

This
study
suggests
that
THM
exposure
may
have
effect
on
menstrual
cycle
length.

However,
examination
of
time
spent
showering
did
not
reveal
additional
risks
with
longer
39
showers.
This
is
counter
to
the
expected
trend,
as
elevated
blood
levels
of
THMs
have
been
documented
after
showering,
due
to
dermal
and
inhalation
exposure
in
the
shower.
However,

information
on
shower
duration
was
collected
by
interview
and
the
reported
lengths
may
not
have
accurately
reflected
actual
shower
duration.
It
would
also
be
useful
to
have
an
independent
confirmation
of
these
results
in
another
study.

The
weight­
of­
evidence
from
epidemiology
studies
suggests
that
DBPs
are
associated
with
developmental/
reproductive
effects
under
certain
exposure
conditions.
The
existing
data
are
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
menstrual
cycle
length,
fetal
growth,
fetal
viability,
and
congenital
abnormalities.
The
studies
employing
the
best
exposure
assessment
found
the
strongest
association
between
fetal
growth
and
fetal
viability
and
chlorinated
drinking
water
(
Bove
et
al.,

2002).
Evaluation
of
the
studies
is
made
more
uncertain
by
the
difficulties
of
exposure
assessment,
including
the
different
composition
of
D/
DBPs
at
different
locations,
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,
but
this
may
be
related
to
a
tendency
among
investigators
to
report
concentrations
of
THMs,
and
not
the
concentrations
of
other
DBPs.
The
available
data
suggest
that
additional
studies
on
THMs
and
other
DBPs
are
needed.

2.1.2.
Systemic
Effects
40
There
are
no
epidemiology
studies
addressing
whether
fetuses
or
children
are
more
sensitive
than
adults
to
systemic
effects
of
chlorinated
drinking
water.

2.1.3.
Carcinogenicity
The
epidemiologic
evidence
regarding
a
relationship
between
drinking
water
chlorination
and
cancer
in
adults
is
mixed.
An
association
between
chlorinated
drinking
water
and
rectal,

colon,
kidney,
and/
or
bladder
cancers
has
been
reported
based
on
a
number
of
human
studies.

The
association
is
strongest
between
DBPs
and
bladder
cancer,
but
causality
has
not
been
established.
Koivusalo
et
al.
(
1998)
and
Cantor
et
al.
(
1998)
both
found
evidence
of
increased
risk
with
increasing
exposure
duration,
but
the
increase
was
only
approximately
two­
fold
after
a
long
exposure
duration.
An
intermediate
level
of
evidence
supports
an
association
with
rectal
cancer.
Yang
et
al.
(
1998)
and
Hildesheim
et
al.
(
1998)
both
found
associations
between
chlorinated
drinking
water
exposure
and
rectal
cancer,
and
the
associations
had
a
similar
magnitude
in
both
sexes.
Hildesheim
et
al.
(
1998)
also
found
an
association
in
both
sexes
with
lifetime
average
trihalomethanes
(
THMs)
concentration.
The
consistency
of
the
dose­
response
trends,
the
consistency
between
sexes,
and
the
apparent
control
of
important
potential
confounders
in
this
study
suggest
that
the
observed
associations
between
the
exposures
and
rectal
cancer
may
be
real
(
EPA,
2001e).
Only
one
(
King
et
al.,
2000b)
of
three
key
recent
studies
(
Hildesheim
et
al.,
1998;
King
et
al.,
2000b;
Yang
et
al.,
1998)
found
an
association
with
colon
cancer,
and
a
strong
causal
association
between
DBPs
and
colon
cancer
is
considered
unlikely
(
EPA,
2001e).
The
limited
data
on
kidney
cancer
support
the
possibility
of
an
association
with
DBPs.
Yang
et
al.
(
1998)
found
fairly
high
standardized
rate
ratios
for
kidney
cancer,
and
41
Koivusalo
et
al.
(
1998)
observed
a
dose
response
by
mutagenicity
tertiles
and
by
duration
of
exposure.
Limited
human
data
(
Cantor
et
al.,
1999)
also
suggest
an
association
between
chlorinated
drinking
water
and
gastrointestinal/
urinary
tract
cancers
and
brain
cancers
(
gliomas).

However,
the
data
are
insufficient
for
a
definitive
conclusion,
and
are
insufficient
for
quantitative
cancer
risk
assessment.

Only
one
study
specifically
evaluated
the
relationship
between
childhood
cancers
and
chlorinated
drinking
water.
In
a
population­
based
case­
control
study
in
Quebec,
Infante­
Rivard
et
al.
(
2000)
examined
the
possible
association
between
childhood
acute
lymphoblastic
leukemia
(
ALL)
and
THMs.
The
authors
studied
491
cases
and
491
controls,
matched
for
age,
sex,
and
region
in
the
province;
logistic
regression
analysis
adjusted
for
maternal
age
and
level
of
schooling.
Individual
information
was
collected
on
water
source,
and
exposure
was
estimated
from
distribution
system
data
for
metals,
nitrates,
and
THMs.
No
association
with
ALL
was
found
for
prenatal
or
postnatal
exposure
to
total
THMs
or
for
specific
THMs,
and
there
was
no
evidence
of
a
dose­
response.
Based
on
these
studies,
the
data
are
inadequate
to
evaluate
the
potential
carcinogenic
effects
of
childhood
or
prenatal
exposure
to
chlorinated
drinking
water.

Overall,
the
epidemiology
data
are
inadequate
for
a
definitive
conclusion
regarding
the
carcinogenic
potential
of
chlorinated
drinking
water
for
adults
or
children.

2.2.
TRIHALOMETHANES
2.2.1.
Chloroform
42
Chloroform
is
one
of
the
best
studied
DBPs
and
has
a
very
extensive
toxicological
database.
Chloroform
and
its
metabolites
have
been
shown
to
cause
liver
and
kidney
toxicity
and
tumors
as
their
primary
adverse
effects.

2.2.1.1.
Developmental/
Reproductive
Effects
In
a
prospective
study,
Windham
et
al.
(
2003)
demonstrated
that
increasing
levels
of
TTHMs
were
associated
with
significantly
shorter
cycles
when
examined
by
quartile.
Similar
decrements
were
observed
in
follicular,
but
not
luteal,
phase
length.
However,
the
effect
was
attributed
to
exposure
to
brominated
THMs,
and
a
clear
association
with
reduced
cycle
length
was
not
observed
for
chloroform
even
at
the
highest
quartile
(

17
µ
g/
L)
(
difference
­
0.3
days;

95%
C.
I.
­
1.0,
0.40).

Two
epidemiology
studies
(
Kramer
et
al.,
1992;
King
et
al.,
2000a)
investigated
the
relationship
between
exposure
to
chloroform
in
chlorinated
drinking
water
and
developmental
effects.
These
studies
were
discussed
in
Section
2.1.,
Chlorinated
Drinking
Water.

Three
developmental
toxicity
studies
(
two
in
rats
and
one
in
rabbits)
by
the
oral
route
of
administration
(
Thompson
et
al.,
1974;
Ruddick
et
al.,
1983)
and
three
developmental
toxicity
studies
(
two
in
rats
and
one
in
mice)
by
the
inhalation
route
of
administration
(
Schwetz
et
al.,

1974;
Murray
et
al.,
1979)
were
reported.

Thompson
et
al.
(
1974)
studied
the
effects
of
chloroform
on
embryonic
and
fetal
development
of
Sprague­
Dawley
rats.
Groups
of
25
pregnant
rats
(
181
 
224
g)
were
gavaged
with
chloroform
in
corn
oil
at
total
daily
doses
of
0,
20,
50,
or
126
mg/
kg/
day
by
oral
intubation
on
days
6
 
15
of
gestation,
administered
in
two
doses/
day.
Dams
receiving
50
or
126
mg/
kg/
day
43
displayed
signs
of
maternal
toxicity
(
decreased
weight
gain,
mild
fatty
changes
in
the
liver).
There
was
no
evidence
of
maternal
toxicity
at
20
mg/
kg/
day,
although
microscopic
examinations
were
conducted
on
only
2
dams/
group.
Fetuses
were
removed
by
caesarean
section
1
or
2
days
prior
to
expected
parturition
and
examined
for
external,
skeletal
and/
or
soft
tissue
abnormalities.
There
were
no
fetal
malformations,
but
some
fetal
variations
were
noted
as
reflected
in
the
statistically
significant
increase
in
the
incidence
of
bilateral
extra
lumbar
ribs
(
p<
0.05)
at
the
high
dose;

however,
the
increase
in
affected
litters
was
not
statistically
significant.
Fetal
weight
was
also
reduced
at
the
high
dose
(
p<
0.05).
This
study
identified
a
maternal
NOAEL
of
20
mg/
kg/
day
and
a
LOAEL
of
50
mg/
kg/
day
in
rats.
For
developmental
effects,
the
NOAEL
was
50
mg/
kg/
day,

with
a
LOAEL
of
126
mg/
kg/
day.

In
the
same
study,
Thompson
et
al.
(
1974)
administered
chloroform
(
in
corn
oil)
to
Dutch­
Belted
rabbits.
In
a
preliminary
range­
finding
study,
doses
of
0,
25,
63,
100,
159,
251,
or
398
mg/
kg/
day
were
administered
to
pregnant
rabbits
on
days
6
 
18
of
gestation.
High
levels
of
maternal
death
(
60
 
100%)
were
observed
at
doses
of
100
mg/
kg/
day
and
above.
Adverse
effects
at
63
mg/
kg/
day
included
anorexia,
weight
loss,
diarrhea,
abortion,
and
one
maternal
death.
No
overt
signs
of
toxicity
other
than
mild
diarrhea
and
intermittent
anorexia
were
observed
in
dams
dosed
with
25
mg/
kg/
day.
In
the
main
study,
groups
of
15
pregnant
dams
(
1.7
 
2.2
kg)
were
dosed
by
oral
intubation
with
chloroform
at
0,
20,
35,
or
50
mg/
kg/
day
on
days
6
 
18
of
gestation.

Decreased
maternal
weight
gain
was
observed
in
dams
given
50
mg/
kg/
day.
Four
high­
dose
dams
died
from
hepatotoxicity,
but
no
evidence
of
hepatotoxicity
was
observed
in
surviving
rabbits.

Four
high­
dose
dams
aborted,
but
this
was
not
considered
to
be
a
treatment­
related
effect
as
three
control
animals
also
aborted.
Histopathology
examinations
revealed
no
evidence
of
maternal
44
toxicity
at
35
mg/
kg/
day.
Small
reductions
in
body
weights
(
7.5%
and
12%,
respectively)
were
observed
in
fetuses
from
dams
administered
20
or
50
mg/
kg/
day
(
p<
0.05),
whereas
only
a
5.5%

decrease
in
fetal
weight
was
observed
at
35
mg/
kg/
day.
At
least
some
of
the
decrease
in
fetal
weight
at
the
high
dose
may
be
attributable
to
the
larger
litter
size
at
the
high
dose
(
7.4,
vs.
6.4
in
the
controls),
although
the
mean
litter
size
at
the
mid
dose
was
only
4.5.
An
increased
incidence
of
fetuses
with
incompletely
ossified
skull
bones
(
usually
parietals)
was
observed
at
20
and
35
mg/
kg/
day
(
p<
0.05);
a
smaller
increase
at
the
high
dose
was
not
statistically
significant,
and
the
results
were
not
significant
when
the
litter
was
used
as
the
statistical
unit
of
comparison.
This
study
is
limited
by
the
high
incidence
of
abortions
and
mortality
in
the
control
group,
but
the
study
authors
did
not
consider
the
observed
effects
to
be
evidence
of
teratogenicity
or
fetotoxicity.
This
study
identified
a
maternal
NOAEL
of
35
mg/
kg/
day
and
a
maternal
LOAEL
of
50
mg/
kg/
day,

based
on
hepatotoxicity;
there
were
no
developmental
effects
related
to
chloroform
treatment.

Ruddick
et
al.
(
1983)
investigated
the
developmental
toxicity
of
chloroform
in
groups
of
15
mated
Sprague­
Dawley
rats.
Pregnant
dams
(
8
 
14
animals
per
dose
group)
were
given
0,
100,

200,
or
400
mg/
kg
chloroform
in
corn
oil
on
days
6
 
15
of
gestation.
Maternal
weight
gain
was
depressed
by
at
least
20%
at
all
dose
levels.
In
addition,
all
dose
levels
of
chloroform
produced
maternal
liver
enlargement,
decreased
hemoglobin,
and
decreased
hematocrit.
Levels
of
serum
inorganic
phosphorus
and
cholesterol
were
elevated
in
the
dams
at
the
highest
exposure
level.

Fetal
weight
was
decreased
by
about
19%
at
the
highest
dose
level.
There
were
no
fetal
malformations,
but
sternebra
aberrations
were
observed
with
a
dose­
dependent
incidence
at
200
mg/
kg/
day
and
400
mg/
kg/
day.
Interparietal
deviations
also
occurred
at
the
high
dose.
There
45
was
a
clear
increase
in
the
incidence
of
these
variations
indicating
a
potential
developmental
effect.

In
an
inhalation
study,
Schwetz
et
al.
(
1974)
exposed
pregnant
female
Sprague­
Dawley
rats
to
chloroform
at
target
concentrations
of
0,
30,
100
or
300
ppm
(
actual
concentrations
of
0,

30,
95,
or
291
ppm;
0,
146,
464
or
1,420
mg/
m3)
for
7
hours/
day
from
gestation
days
6
through
15.
Because
marked
anorexia
was
observed
in
an
earlier
experiment
in
dams
exposed
to
300
ppm
chloroform,
an
additional
control
group
was
starved;
that
is,
allowed
only
3.7
grams
of
food
per
day.
The
numbers
of
pregnant
rats
exposed
in
each
group
were
68,
22,
23,
and
3,
respectively.

The
low
percent
pregnancy
observed
at
the
high
concentration
was
not
considered
to
be
treatment­
related,
due
to
the
time
of
exposure;
however,
the
use
of
such
a
small
number
of
animals
in
the
300
ppm
group
decreased
the
statistical
sensitivity
of
any
adverse
effects
observed
in
this
group.
The
dams
were
sacrificed
on
gestation
day
21,
and
fetuses
were
removed
by
caesarian
section.
Food
consumption
and
body
weight
gain
exhibited
concentration­
related
decreases
in
all
exposure
groups.
A
significant
increase
in
relative
liver
weights
in
dams
exposed
to
100
or
300
ppm
was
observed
at
study
termination,
with
a
significant
decrease
in
absolute
liver
weight
at
300
ppm.
However,
there
was
no
effect
on
serum
glutamate­
pyruvate
transaminase
activity
at
any
concentration.
At
300
ppm,
61%
of
the
implantations
were
resorbed,
a
statistically
significant
increase.
This
high
resorption
rate
was
not
observed
in
the
"
starved"
control
group,

suggesting
that
weight
loss
cannot
account
for
the
observed
effect,
although
the
starved
control
group
was
provided
more
food
than
was
consumed
by
the
300
ppm
group.
Fetal
body
weights
were
significantly
decreased
(
40%)
at
300
ppm,
and
fetal
crown­
rump
lengths
were
slightly,
but
significantly,
decreased
(
2%)
at
30
ppm
and
significantly
decreased
(
15%)
at
300
ppm.
The
46
frequencies
of
litters
with
acaudia
or
imperforate
anus
were
significantly
increased
at
100
ppm.

Malformations
were
not
observed
at
300
ppm,
but
there
were
only
three
litters
at
this
concentration.
The
frequency
of
litters
with
delayed
ossification
was
elevated
in
all
exposure
groups.
In
addition,
there
were
statistically
significant
increases
in
wavy
ribs
at
30
ppm,
and
in
missing
ribs
and
subcutaneous
edema
at
100
ppm.
The
authors
concluded
that
chloroform
exposures
of
100
and
300
ppm
were
highly
embryotoxic
and
fetotoxic,
with
embryolethality
a
significant
effect
at
300
ppm.

Murray
et
al.
(
1979)
found
that
100
ppm
chloroform
was
teratogenic
in
CF­
1
mice
exposed
on
gestation
days
8
 
15,
but
was
fetotoxic
in
mice
exposed
on
gestation
days
1
 
7
or
6
 
15.
Groups
of
34
 
40
pregnant
females
(
as
determined
by
vaginal
plug)
were
exposed
to
0
or
100
ppm
(
0
or
490
mg/
m3)
for
7
hours/
day
on
gestation
days
1
 
7,
6
 
15,
or
8
 
15,
and
sacrificed
on
gestation
day
18.
The
ability
of
the
mice
to
maintain
pregnancy
was
significantly
decreased
in
the
groups
exposed
on
gestation
days
1
 
7
or
6
 
15,
and
there
was
a
slight
(
but
not
statistically
significant)
decrease
in
pregnancies
in
the
group
exposed
on
gestation
days
8
 
15.
Statistically
significant
decreases
in
fetal
weight
and
fetal
length
were
observed
in
the
groups
exposed
on
gestation
days
1
 
7
and
8
 
15
but
not
on
days
6
 
15.
Cleft
palate
was
observed
at
a
statistically
significant
increased
incidence
in
litters
of
mice
exposed
on
gestation
days
8
 
15,
but
not
in
the
other
groups.
Cleft
palate
was
seen
predominantly
in
fetuses
with
retarded
growth.
No
other
malformations
were
significantly
increased
in
any
group,
although
increased
incidences
of
two
skeletal
variations
were
observed.
Delayed
ossification
of
skull
bones
was
significantly
increased
in
all
exposed
groups,
and
delayed
ossification
of
sternebrae
was
significantly
increased
in
the
groups
exposed
on
gestation
days
1
 
7
and
8
 
15,
but
not
6
 
15.
The
study
authors
suggested
that
47
the
lack
of
malformations
in
the
group
exposed
on
gestation
days
6
 
15
may
have
resulted
from
the
lethality
to
the
early
embryo
obscuring
other
effects.
Maternal
toxicity
was
evident
as
increased
liver
weight
and
increased
serum
glutamate­
pyruvate
transaminase
activity.

Based
on
the
findings
of
animal
studies
discussed
above,
developmental
effects
have
been
found
after
chloroform
exposure
(
e.
g.,
pup
weight
reduction,
skeletal
variations).
These
prenatal
effects,
however,
were
typically
associated
with
exposures
causing
maternal
toxicity,
and
occurred
at
oral
doses
above
those
causing
hepatotoxicity.
The
oral
NOAEL
for
developmental
toxicity
is
in
the
range
of
35
 
50
mg/
kg/
day,
and
the
oral
LOAEL
for
hepatotoxicity
is
12.9
mg/
kg/
day
for
chloroform.

A
multigeneration
reproductive
assay
was
conducted
with
chloroform
in
CD­
1
mice
(
NTP,
1988).
This
assay
evaluated
reproductive
effects
in
two
successive
generations,
as
well
as
systemic
effects
in
the
second
generation
(
i.
e.,
F
1
animals).
In
the
first
phase,
mice
were
administered
chloroform
by
gavage
in
corn
oil
at
6.6,
16,
or
41
mg/
kg/
day,
7
days/
week
for
18
weeks.
In
the
second
phase,
the
last
litter
of
the
control
and
of
the
high­
dose
groups
were
retained.
After
weaning,
the
mice
were
administered
the
same
chloroform
dose
as
their
parents,

and
dosing
continued
through
mating
and
parturition,
when
the
study
was
terminated.
No
adverse
effects
on
fertility
or
reproduction
of
the
F
1
generation
were
observed,
although
increased
liver
weight
and
liver
lesions
(
degeneration
of
centrilobular
hepatocytes,
accompanied
by
occasional
single
cell
necrosis)
were
observed
in
all
females
exposed
to
the
single
dose
tested.

The
degeneration
was
characterized
as
minimal
in
2/
20,
mild
in
9/
20
and
moderate
in
9/
20
animals.
Thus,
a
dose
of
41
mg/
kg/
day
caused
mild
to
moderate
liver
histopathology
in
F
1
females.
No
NOAEL
can
be
identified
for
this
effect,
because
the
low­
and
mid­
dose
groups
were
48
not
evaluated
histopathologically.
However,
no
adverse
effects
on
fertility
or
reproduction
were
found.

2.2.1.2.
Systemic
Effects
Numerous
animal
studies
in
several
species
(
rats,
mice
and
dogs)
have
shown
that
liver
and
kidney
toxicity
are
primarily
target
sites
for
the
systemic
effects
of
chloroform.
Nasal
toxicity
is
also
found
in
the
rat
following
inhalation
exposure.
Organ
toxicity
(
including
liver
and
kidney
tumor
response)
following
chloroform
treatment
vary
with
the
exposure
route,
vehicle
of
administration
and
strain
of
rat
or
mouse.
The
sensitivity
to
the
organ
toxicity
induced
by
chloroform
is
associated
with
oxidative
metabolism.
These
results
were
summarized
in
two
EPA
documents
(
EPA,
1994b,
1998b).
Organ
toxicity
that
results
from
chloroform
is
considered
to
be
part
of
the
continuum
that
leads
to
tumor
development.
The
organ
toxicity
is
thus
discussed
below
in
the
context
of
the
mode
of
carcinogenic
action
for
chloroform.

2.2.1.3.
Carcinogenicity
Chloroform
has
been
found
to
cause
liver
and
kidney
tumors
in
rodents
(
discussed
in
EPA,

1994a,
1998b).
A
substantial
body
of
data
indicates
that
chloroform
is
not
a
DNA­
reactive
mutagen.
Thus,
mutagenicity
is
not
the
key
influence
of
chloroform
on
the
carcinogenic
process.

Chloroform
induces
liver
and
kidney
tumors
at
doses
that
cause
cell
injury
or
organ
toxicity.

Numerous
studies
have
shown
that
organ
toxicity
and
regenerative
proliferation
are
associated
with
tumorigenicity
of
chloroform,
and
thus
are
key
steps
in
its
carcinogenic
mode
of
action
(
EPA,
1998b;
ILSI,
1997).
To
explore
the
issue
of
whether
fetuses
or
children
are
at
increased
49
cancer
risk
compared
with
adults,
the
mode
of
carcinogenic
action
of
chloroform
in
children
versus
adults
was
examined.
To
address
this
question,
the
Agency
evaluated
the
available
data
for
children
and
adults
on
chloroform
metabolism
and
age­
related
differences
in
the
rate
of
cell
proliferation
(
EPA,
1998b).

Organ
toxicity
from
chloroform
is
dependent
on
oxidative
metabolism
primarily
by
cytochrome
P450
CYP2E1
(
as
discussed
in
EPA,
1998b).
The
oxidative
metabolism
of
chloroform
generates
highly
tissue­
reactive
metabolites.
One
metabolite
is
phosgene,
a
highly
reactive
dihalocarbonyl,
and
the
other
is
the
strong
acid
HCl.
The
very
high
reactivity
of
phosgene
prevents
it
from
entering
the
nucleus
and
forming
adducts
with
DNA.
The
chloroform
metabolites
produce
cytotoxicity
(
cell
death)
and
regenerative
hyperplasia
(
EPA,
1998b;
ILSI,

1997).
This
process
may
lead
to
tumor
development
if
sustained.
Metabolism
of
chloroform
via
a
reductive
pathway,
if
it
occurs,
could
lead
to
the
formation
of
free
radicals
and
tissue
damage,

but
the
reductive
pathway
is
absent
or
minor
under
normal
physiological
conditions
(
EPA,

2001a).

Given
that
oxidative
metabolism
is
key
to
the
carcinogenic
potential
of
chloroform,
studies
on
CYP2E1
in
fetal
and
adult
tissues
were
evaluated
(
EPA,
1998b).
Studies
in
humans
have
not
shown
consistent
results;
however,
in
those
studies
showing
expression
of
CYP2E1,
levels
in
fetuses
were
lower
than
those
in
adults
(
Boutelet­
Bochan
et
al.,
1997;
Carpenter
et
al.,
1996;

Hakkola
et
al.,
1998a;
Vieira
et
al.,
1996).
Vieira
et
al.
(
1996)
suggested
that
CYP2E1
activity
increases
rapidly
in
the
24
hours
after
birth,
and
that
activity
level
in
children
aged
1
to
10
years
is
comparable
to
that
of
adults.
50
Animal
studies
of
CYP2E1
provide
evidence
of
rapid
induction
of
this
gene
soon
after
birth
(
Song
et
al.,
1986;
Umeno
et
al.,
1988;
Schenkman
et
al.,
1989;
Ueno
and
Gonzalez,
1990).

The
study
by
Schenkman
et
al.
(
1989)
indicated
that
CYP2E1
protein
is
present
in
low
levels
in
CD
rat
neonates,
rises
to
a
peak
level
at
age
2
weeks
and
subsequently
decreases
to
adult
levels
by
puberty.
Analysis
of
protein
levels
quantified
from
western
blots
showed
a
maximum
at
2
weeks
with
decreasing
levels
at
4
and
12
weeks.
The
protein
level
at
12
weeks
was
approximately
50%
of
the
level
at
2
weeks.
The
authors
did
not
provide
a
statistical
analysis
of
this
result,
but
it
appears
from
the
error
bars
that
the
2­
week
and
12­
week
levels
(
but
not
4­
week
levels)
were
significantly
different.
Song
et
al.
(
1986)
conducted
a
similar
analysis
in
Sprague­

Dawley
rats
and
reported
a
rapid
transcriptional
induction
of
CYP2E1
(
P450)
within
1
week
following
birth
that
remained
elevated
throughout
12
weeks.
Enzyme
activity
followed
a
similar
pattern.
Ueno
and
Gonzalez
(
1990)
showed
that
extracts
from
3­
day­
old
and
12­
week­
old
rat
liver,
but
not
those
from
fetal
or
newborn
rat
liver,
were
able
to
generate
significant
CYP2E1
transcription
in
vitro.
The
ability
of
the
extract
to
result
in
transcription
of
CYP2E1
was
slightly
greater
at
12
weeks.

Taken
together,
the
animal
studies
do
not
provide
conclusive
evidence
of
an
early
period
of
increased
enzymatic
activity.
If,
however,
the
twofold
increase
in
CYP2E1
induction
in
animals
were
seen
in
humans,
its
importance
in
terms
of
chloroform
toxicity
would
depend
on
the
dose.
Under
low­
dose
conditions
(
e.
g.,
much
lower
than
the
Km)
it
is
possible
that
an
increase
in
the
level
of
enzyme
would
not
have
any
effect
on
active
metabolite
formation,
because
the
amount
of
chloroform,
and
not
CYP2E1,
would
control
the
rate
of
phosgene
production.
On
the
other
hand,
under
saturating
doses
of
chloroform,
all
the
available
enzyme
would
be
active;
thus
a
51
twofold
increase
in
CYP2E1
could
result
in
greater
activation
of
the
compound.
Although
the
animal
data
remain
unclear
regarding
the
potential
for
a
neonatal
period
of
increased
CYP2E1
activity
above
that
in
the
adult,
the
data
in
humans
show
a
rapid
induction
after
birth,
gradually
increasing
over
the
first
year
to
reach
adult
levels
during
years
1­
10
(
Vieira
et
al.,
1996).

Therefore,
although
children
may
have
the
capacity
to
metabolize
chloroform,
data
on
CYP2E1
activity
provide
no
evidence
to
suggest
that
children
have
an
increased
susceptibility
to
chloroform
toxicity
compared
with
adults
(
EPA,
1998b).
Furthermore,
the
data
from
Schenkman
et
al.
(
1989)
indicate
that
levels
of
CYP2E1
are
approximately
2­
fold
higher
in
rats
at
2
weeks
than
at
12
weeks
of
age.
The
2­
fold
magnitude
of
the
difference
in
CYP2E1
levels
is
within
the
default
UF
value
of
3
used
to
account
for
intraspecies
variability
in
toxicokinetics
(
the
CYP2E1
studies
do
not
address
the
contribution
of
toxicodynamics
to
the
total
default
UF
of
10).
Based
on
this
analysis,
differences
in
age­
dependent
metabolism
of
chloroform
by
CYP2E1
are
adequately
accounted
for
in
the
existing
UF
for
intraspecies
variability.

Thus,
the
human
data
are
inconclusive
on
age­
related
differences
in
CYP2E1
activity,
and
it
is
not
possible
to
state
whether
children
may
have
an
increased
susceptibility
to
chloroform
toxicity
and
carcinogenicity
as
compared
to
adults.
However,
based
on
animal
data,
the
degree
of
age­
dependent
differences
in
CYP2E1
expression
(
if
any)
are
likely
well
accounted
for
by
default
assumptions
about
the
magnitude
of
intraspecies
variability.

The
next
issue
to
examine
is
whether
the
developing
fetus
may
be
more
sensitive
to
the
toxicity
of
chloroform
because
of
its
greater
rate
of
cell
proliferation.
There
are
very
few
data
on
prenatal
and
postnatal
exposures
to
chloroform
and
resultant
organ
toxicity.
Liver
toxicity
was
found
in
a
multigeneration
reproductive
assay
in
CD­
1
mice
(
NTP,
1988).
The
periods
of
52
exposure
included
prenatal,
postnatal
and
adult
stages.
The
liver
toxicity
from
this
multigeneration
reproductive
study
was
compared
with
that
of
a
comparable
90­
day
study
in
adult
B6C3F1
mice
for
liver
toxicity
(
Bull
et
al.,
1986).
The
similarity
of
effects
at
comparable
doses
from
these
two
studies
suggests
that
there
is
no
increased
susceptibility
to
chloroform
that
results
from
prenatal
or
postnatal
exposures
(
EPA,
1998b).
It
should
be
noted
that
there
are
limitations
in
this
comparison;
different
strains
of
mice
were
used,
and
only
LOAELs
were
identified
in
these
two
studies
(
EPA,
1998b).

Under
the
1986
U.
S.
EPA
Guidelines
for
Carcinogen
Risk
Assessment,
chloroform
has
been
classified
as
Group
B2,
probable
human
carcinogen,
based
on
sufficient
evidence
of
carcinogenicity
in
animals
(
EPA,
1998a).
Under
the
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
of­
evidence
conclusion
is
based
on
several
lines
of
evidence.
1)
Observations
in
animals
exposed
by
both
oral
and
inhalation
pathways
indicate
that
sustained
or
repeated
cytotoxicity
with
secondary
regenerative
hyperplasia
precedes,
and
is
probably
required
for,

hepatic
and
renal
neoplasia.
2)
There
are
no
epidemiological
data
specific
to
chloroform.
The
epidemiological
data
relating
cancer
to
exposure
to
chlorinated
drinking
water
are
mixed,
and
any
observed
effects
cannot
necessarily
be
attributed
to
chloroform
amongst
multiple
other
disinfection
byproducts.
3)
Genotoxicity
data
on
chloroform
are
essentially
negative;
there
are
a
few
scattered
positive
results
that
generally
have
limitations
such
as
excessively
high
doses
or
53
confounding
factors.
Thus,
the
weight­
of­
evidence
of
the
genotoxicity
data
on
chloroform
supports
a
conclusion
that
chloroform
is
not
strongly
mutagenic,
and
that
genotoxicity
is
not
likely
to
be
the
predominant
mode
of
action
underlying
the
carcinogenic
potential
of
chloroform
(
EPA,
2001a).

Because
a
substantial
database
indicates
that
tumor
development
for
chloroform
is
secondary
to
organ
toxicity
and
regenerative
proliferation,
a
nonlinear
dose­
response
approach
for
tumorigenicity
is
viewed
as
appropriate.
Use
of
a
low
dose­
linear
approach
is
considered
overly
conservative
for
extrapolating
cancer
risk
associated
with
chloroform
exposure
(
EPA,
1998a,
b,

2001a;
ILSI,
1997).
EPA
(
2001a)
determined
a
point
of
departure
(
POD)
of
23
mg/
kg/
day,

based
on
increased
kidney
tumors
in
Osborne­
Mendel
rats
(
Jorgensen
et
al.,
1985).
The
Agency
compared
this
POD
to
the
RfD
of
0.01
mg/
kg/
day
calculated
as
described
in
the
next
paragraph,

and
concluded
that
the
margin
of
exposure
(
MOE)
of
2000
was
adequately
protective
of
public
health
for
cancer
effects.
Sustained
tissue
toxicity,
which
is
a
key
event
in
the
cancer
mode
of
action
for
chloroform,
will
not
occur
at
doses
below
the
RfD.

2.2.1.4.
Basis
for
RfD
and
MCLG
The
Agency
used
an
oral
study
in
dogs
(
Heywood
et
al.,
1979)
to
derive
the
chloroform
RfD.
The
study
identified
a
LOAEL
of
15
mg/
kg/
day
for
hepatotoxicity
(
fatty
cysts
and
an
increase
of
serum
glutamic
pyruvic
transaminase).
A
BMDL
10
(
the
95%
lower
bound
confidence
limit
on
the
dose
associated
with
a
10%
extra
risk)
of
1.2
mg/
kg/
day
was
also
calculated,
based
on
the
prevalence
of
animals
with
moderate
to
marked
fatty
cysts
in
liver
in
the
same
study.

Applying
a
factor
of
6/
7
to
account
for
exposure
6
days/
week,
the
LOAEL
was
converted
to
12.9
54
mg/
kg/
day,
and
the
BMDL
10
was
converted
to
1.0
mg/
kg/
day.
The
RfD
was
based
on
both
the
benchmark
dose
(
BMD)
and
traditional
NOAEL/
LOAEL
approach,
which
coincidentally
result
in
the
same
RfD.
An
RfD
of
0.01
mg/
kg/
day
was
calculated
using
the
BMD
approach
by
applying
an
overall
UF
of
100
(
10
for
interspecies
extrapolation
and
10
for
protection
of
sensitive
individuals)
to
the
BMDL
10
of
1.0
mg/
kg/
day.
Using
the
traditional
approach,
a
composite
UF
of
1,000
was
applied
(
100
was
used
to
account
for
inter­
and
intraspecies
differences
and
a
factor
of
10
for
use
of
a
LOAEL),
resulting
in
an
RfD
of
0.01
mg/
kg/
day.
This
RfD
corresponds
to
a
Drinking
Water
Equivalent
Level
(
DWEL)
of
0.35
mg/
L,
assuming
an
adult
tap
water
consumption
of
2
L/
day
for
a
70
kg
adult.

Because
a
nonlinear
approach
was
used
for
chloroform
carcinogenicity,
the
MCLG
is
based
on
liver
toxicity
as
the
most
sensitive
effect
for
chloroform
(
as
it
is
the
lowest
possible
LOAEL
for
any
organ)
and
as
a
precursor
response
to
a
key
step
to
its
carcinogenicity.
This
approach
is
considered
equally
protective
of
both
adults
and
children
because
the
database
on
chloroform
does
not
indicate
that
children
are
more
sensitive
than
adults
to
liver
toxicity.
The
mode
of
action
by
which
chloroform
produces
organ
toxicity
and
carcinogenicity
is
considered
to
be
the
same
for
children
and
adults.

The
MCLG
of
0.07
mg/
L
was
calculated
from
the
RfD
by
assuming
an
adult
tap
water
consumption
of
2
L
per
day
for
a
70
kg
adult,
and
by
applying
a
relative
source
contribution
(
RSC)
of
20%.
The
RSC
is
based
on
data
indicating
that
exposure
to
chloroform
by
other
routes
and
sources
of
exposure
may
potentially
contribute
a
substantial
percentage
of
the
overall
exposure
to
chloroform
(
EPA,
2001d).
Based
on
average
daily
doses
for
each
source
and
route
of
exposure
under
specific
conditions,
EPA
estimated
that
for
the
median
individual,
ingestion
of
55
total
tap
water
(
including
water­
based
drinks)
can
contribute
roughly
28%
of
the
total
dose
of
chloroform,
while
the
dose
from
showering
(
inhalation
and
dermal
exposure)
was
estimated
to
contribute
approximately
14%
of
the
total
dose
(
EPA,
2001d).
There
is,
however,
considerable
uncertainty
in
these
exposure
estimates.
For
example,
because
chloroform
is
so
volatile,
most
of
the
chloroform
would
evaporate
from
ingested
hot
liquids
such
as
coffee
or
tea,
so
the
value
of
28%
from
ingestion
of
total
tap
water
may
be
an
over­
estimate.
In
addition,
the
proportion
of
intake
from
different
sources
varies
widely
across
the
population.
Therefore,
the
default
RSC
of
20%
is
used,
a
value
that
is
consistent
with
the
percent
of
the
total
dose
from
ingestion
of
total
tap
water.
Thus,
the
MCLG
is
0.07
mg/
L:
(
0.01
mg/
kg/
day
x
70
kg
x
0.2)/
(
2/
L/
day)
=
0.07
mg/
L.

2.2.1.5.
Children's
Risk
in
Relation
to
the
MCLG
The
MCLG
derived
for
chloroform
is
considered
protective
of
both
adults
and
children,

given
that
developmental
effects
occurred
at
doses
above
those
causing
hepatotoxicity.
In
addition,
fetal
effects
were
only
seen
at
levels
at
which
maternal
toxicity
was
noted.
Also,
the
mode
of
action
data
indicates
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.
In
addition,
the
developing
fetus
would
not
be
expected
to
be
particularly
sensitive
to
a
cytotoxic
agent,
such
as
chloroform
at
low
levels,
because
cell
division
occurs
at
a
rapid
pace
and
the
cellular
repair
capacity
is
high
(
EPA,
2001a).
56
The
MCLG
is
also
considered
protective
of
carcinogenic
effects
in
children
and
fetuses.

EPA
believes
that
using
a
nonlinear
dose­
response
approach
for
setting
the
MCLG
is
protective
for
children
and
fetuses
for
the
following
reasons:
1)
the
reactive
metabolite
inside
the
cell
should
have
similar
effects
resulting
from
its
reacting
with
and
disrupting
macromolecules
in
the
cells
of
fetuses,
children,
and
adults;
2)
cell
necrosis
and
reparative
replication
are
not
likely
to
be
qualitatively
different
in
various
life
stages;
3)
cancer
risk
to
the
fetus
or
children
would
be
a
function
of
cytotoxicity,
as
in
adults,
and
protecting
fetuses
and
children
from
sufficient
levels
of
the
chemical
that
would
cause
cytotoxicity
should
protect
against
cancer
risk
(
EPA,
2001a);
4)
a
comparison
of
liver
toxicity
between
a
multigeneration
reproductive
study
and
a
comparable
90­

day
study
suggests
that
there
is
no
increased
susceptibility
to
chloroform
effects
on
the
liver
that
results
from
prenatal
or
postnatal
exposure,
indicating
that
children
are
not
more
sensitive
to
the
liver
cytotoxicity
that
is
a
precursor
to
chloroform
carcinogenicity
(
EPA,
1998b).
This
final
factor
also
supports
the
conclusion
that
children
are
not
more
sensitive
than
adults
to
the
noncancer
liver
effects
of
chloroform.

2.2.2.
Brominated
Trihalomethanes
There
is
sufficient
evidence
for
carcinogenicity
via
ingestion
of
bromoform
and
bromodichloromethane
(
BDCM)
to
consider
them
probable
human
carcinogens.
The
evidence
is
limited
for
dibromochloromethane
(
DBCM).
Based
on
the
available
data,
a
mechanism
of
action
involving
mutagenicity
was
postulated
for
the
brominated
THMs,
indicating
that
linear
low­
dose
extrapolation
for
BDCM
and
bromoform
is
appropriate.
The
proposed
mechanism
of
carcinogenicity
for
these
compounds
was
examined
to
determine
if
this
would
provide
any
reason
57
for
concern
that
children
or
fetuses
may
be
more
susceptible
to
development
of
cancer
following
exposure.
If
carcinogenicity
is
the
result
of
mutations
by
either
the
parent
compound
or
a
metabolite,
children
or
the
developing
fetus
could
be
more
susceptible
to
the
carcinogenicity
of
brominated
THMs
due
to
a
higher
rate
of
cell
proliferation
in
the
target
organs.
An
increased
risk
of
this
type
would
be
true
for
all
genotoxic
carcinogens
and
not
specific
to
brominated
THMs.

There
are
no
data
currently
available
for
brominated
THMs
to
permit
quantification
of
a
possible
increase
in
risk
to
the
developing
fetus
or
children.

A
number
of
epidemiology
studies
have
reported
an
association
between
exposure
to
THMs
and
developmental/
reproductive
effects
(
Bove
et
al.,
1995;
Dodds
et
al.,
1999;
Gallagher
et
al.,
1998;
King
et
al.,
2000a;
Klotz
and
Pyrch,
1999;
Kramer
et
al.,
1992;
Savitz
et
al.,
1995;

Waller
et
al.,
1998,
Windham
et
al.,
2003).
For
all
of
these
studies,
because
the
subjects
were
exposed
to
other
contaminants
and
disinfection
byproducts
in
the
drinking
water,
correlation
of
the
effects
directly
to
individual
brominated
THM
exposure
is
difficult.
Further
details
about
these
studies
were
presented
in
Section
2.1,
Chlorinated
Drinking
Water.

Although
the
mechanism
of
brominated
trihalomethane
toxicity
is
not
known
with
certainty,
data
indicate
that
the
adverse
effects
of
this
group
of
chemicals
are
secondary
to
metabolism.
Brominated
THMs
are
extensively
metabolized
via
oxidative
(
using
NADPH
and
oxygen)
and
reductive
(
using
NADPH
or
NADH
and
inhibited
by
oxygen)
pathways
in
humans
and
animals,
primarily
in
the
liver,
but
also
in
the
kidney
(
EPA,
2002c).
Recent
data
suggest
that
bioactivation
of
brominated
trihalomethanes
to
mutagenic
species
is
also
mediated
by
one
or
more
glutathione
S­
transferase­
mediated
conjugation
pathways.
58
Both
oxidative
and
reductive
reactions
are
mediated
by
cytochrome­
P450s.
Oxidative
metabolism
results
in
the
production
of
a
dihalocarbonyl
(
CX
2
O)
intermediate,
which
may
undergo
a
variety
of
reactions,
such
as
adduct
formation
with
cellular
nucleophiles,
hydrolysis
to
yield
carbon
dioxide
or
glutathione­
dependent
reduction
to
yield
carbon
monoxide.
Reductive
metabolism
results
in
the
production
of
free
radical
species
such
as
the
dihalomethyl
radical
(
CHX
2
.).
In
vitro
and
in
vivo
data
suggest
that
metabolism
via
the
reductive
pathway
occurs
more
readily
for
the
brominated
trihalomethanes
than
it
does
for
chloroform.
Gao
and
Pegram
(
1992)
reported
that
binding
of
reactive
intermediates
to
rat
hepatic
microsomal
lipid
and
protein
under
reductive
(
anaerobic)
conditions
was
more
than
twice
as
high
for
bromodichloromethane
as
for
chloroform.
Tomasi
et
al.
(
1985)
used
electron
spin
resonance
(
ESR)
spectroscopy
to
measure
the
production
of
a
free
radical
intermediates
(
a
product
of
the
reductive
pathway)
in
vitro
using
rat
hepatocytes
isolated
from
phenobarbital­
induced
male
Wistar
rats.
The
intensity
of
the
ESR
signal
was
greatest
for
bromoform,
followed
by
BDCM
and
then
chloroform.
The
largest
ESR
signal
was
detected
when
hepatocytes
were
incubated
under
anaerobic
conditions.

Tomasi
et
al.
(
1985)
also
used
ESR
to
evaluate
free
radical
production
in
vivo
in
rats
given
intraperitoneal
injections
of
chloroform,
BDCM,
or
bromoform.
The
intensity
of
the
ESR
signal
followed
a
ranking
similar
to
that
observed
in
in
vitro
experiments
(
bromoform
>
BDCM>

chloroform),
confirming
that
the
reductive
formation
of
free
radicals
is
greater
for
brominated
trihalomethanes
than
for
chloroform.
Together,
these
data
indicate
that
reductive
metabolism
is
a
more
important
pathway
for
metabolism
of
brominated
trihalomethanes
than
for
chloroform.
The
relative
importance
of
the
oxidative
and
reductive
pathways
for
the
brominated
THMs
in
vivo
has
not
been
determined,
but
it
is
of
note
that
oxygen
partial
pressure
in
the
kidney
and
liver
is
low
59
under
in
vivo
conditions.
Both
dihalocarbonyls
(
produced
by
the
oxidative
pathway)
and
dihalomethyl
radicals
(
produced
by
the
reductive
pathway)
are
reactive
species
and
may
form
covalent
adducts
with
a
variety
of
cellular
components
(
EPA,
2002c).

Metabolism
of
the
brominated
THMs
involves
a
number
of
enzymes,
including
CYP2E1,

CYP2B1/
2
and
CYP1A
(
EPA,
2002c).
The
roles
of
the
different
P450
isozymes
in
brominated
trihalomethane
metabolism
(
i.
e.,
oxidative
versus
reductive
pathways)
have
not
been
definitively
identified,
but
CYP2E1,
CYP2B1/
2
and
CYP1A
have
been
implicated
in
the
oxidative
pathway.

The
toxicity
of
BDCM
and
DBCM
has
been
shown
to
be
at
least
partly
related
to
bioactivation
by
the
enzyme
CYP2E1
(
e.
g.,
Thornton­
Manning
et
al.,
1994).
Thus,
a
higher
level
of
CYP2E1
activity
in
children,
as
compared
to
adults,
would
mean
that
children
could
be
at
greater
risk
for
carcinogenic
effects
from
these
compounds.

Studies
in
humans
have
not
shown
consistent
results;
however,
in
those
studies
showing
expression
of
CYP2E1,
levels
in
fetuses
were
lower
than
those
in
adults
(
Boutelet­
Bochan
et
al.,

1997;
Carpenter
et
al.,
1996;
Hakkola
et
al.,
1998a;
Vieira
et
al.,
1996).
Vieira
et
al.
(
1996)

suggested
that
CYP2E1
activity
increases
rapidly
in
the
24
hours
after
birth,
and
that
activity
level
in
children
aged
1
to
10
years
is
comparable
to
that
of
adults.

Animal
studies
of
CYP2E1
provide
evidence
of
rapid
induction
of
this
gene
soon
after
birth
(
Song
et
al.,
1986;
Umeno
et
al.,
1988;
Schenkman
et
al.,
1989;
Ueno
and
Gonzalez,
1990).

The
study
by
Schenkman
et
al.
(
1989)
indicated
that
CYP2E1
protein
is
present
in
low
levels
in
CD
rat
neonates,
rises
to
a
peak
level
at
age
2
weeks
and
subsequently
decreases
to
adult
levels
by
puberty.
Analysis
of
protein
levels
quantified
from
western
blots
showed
a
maximum
at
2
weeks
with
decreasing
levels
at
4
and
12
weeks.
The
protein
level
at
12
weeks
was
60
approximately
50%
of
the
level
at
2
weeks.
The
authors
did
not
provide
a
statistical
analysis
of
this
result,
but
it
appears
from
the
error
bars
that
the
2­
week
and
12­
week
levels
(
but
not
4­
week
levels)
were
significantly
different.
Song
et
al.
(
1986)
conducted
a
similar
analysis
in
Sprague­
Dawley
rats
and
reported
a
rapid
transcriptional
induction
of
CYP2E1
(
P450)
within
one
week
following
birth
that
remained
elevated
throughout
12
weeks.
Enzyme
activity
followed
a
similar
pattern.
Ueno
and
Gonzalez
(
1990)
showed
that
extracts
from
3­
day­
old
and
12­

weekold
rat
liver,
but
not
those
from
fetal
or
newborn
rat
liver,
were
able
to
generate
significant
CYP2E1
transcription
in
vitro.
The
ability
of
the
extract
to
result
in
transcription
of
CYP2E1
was
slightly
greater
at
12
weeks.

Taken
together,
the
animal
studies
do
not
provide
conclusive
evidence
of
an
early
period
of
increased
enzymatic
activity.
If,
however,
the
2­
fold
increase
in
CYP2E1
induction
in
animals
were
seen
in
humans,
its
importance
in
terms
of
brominated
THM
toxicity
would
depend
on
the
dose.
Under
low­
dose
conditions
(
e.
g.,
much
lower
than
the
Km)
it
is
possible
that
an
increase
in
the
level
of
enzyme
would
not
have
any
effect
on
active
metabolite
formation,
because
the
amount
of
brominated
THMs,
and
not
CYP2E1,
would
control
the
rate
of
the
enzymatic
metabolism.
On
the
other
hand,
under
saturating
doses
of
brominated
THMs,
all
the
available
enzyme
would
be
active;
thus
a
2­
fold
increase
in
CYP2E1
could
result
in
greater
activation
of
the
compound.

Although
the
animal
data
remain
unclear
regarding
the
potential
for
a
neonatal
period
of
increased
CYP2E1
activity
above
that
in
the
adult,
the
data
in
humans
show
a
rapid
induction
after
birth,

gradually
increasing
over
the
first
year
to
reach
adult
levels
during
years
1­
10
(
Vieira
et
al.,
1996).

Therefore,
although
children
may
have
the
capacity
to
metabolize
brominated
THMs,
data
on
CYP2E1
activity
provide
no
evidence
to
suggest
that
children
have
an
increased
susceptibility
to
61
brominated
THM
toxicity
compared
with
adults
(
EPA,
2002c).
Furthermore,
the
data
from
Schenkman
et
al.
(
1989)
indicate
that
levels
of
CYP2E1
are
approximately
2­
fold
higher
in
rats
at
2
weeks
versus
12
weeks
of
age.
The
2­
fold
magnitude
of
the
difference
in
CYP2E1
levels
is
within
the
default
UF
value
of
3
used
to
account
for
intraspecies
variability
in
toxicokinetics
(
the
CYP2E1
studies
do
not
address
the
contribution
of
toxicodynamics
to
the
total
default
UF
of
10).

Based
on
this
analysis,
differences
in
age­
dependent
metabolism
of
brominated
THMs
by
CYP2E1
are
adequately
accounted
for
in
the
existing
UF
for
intraspecies
variability.

Thus,
the
human
data
are
inconclusive
on
age­
related
differences
in
CYP2E1
activity
and
it
is
not
possible
to
state
whether
children
may
have
an
increased
susceptibility
to
brominated
THM
toxicity
as
compared
to
adults.
However,
based
on
animal
data,
the
degree
of
agedependent
differences
in
CYP2E1
expression
(
if
any)
are
likely
well
accounted
for
by
default
assumptions
about
the
magnitude
of
intraspecies
variability.

Information
on
age­
related
differences
in
the
activity
of
the
other
human
cytochrome
P450s
involved
in
brominated
THM
metabolism
was
reviewed
in
a
recent
summary
of
the
current
literature
(
EPA,
2000d).
For
CYP1A1,
constitutive
fetal
mRNA
levels
were
measurable
in
some
studies,
with
expression
levels
continuing
to
increase
through
gestation
(
Oesterheld,
1998).

However,
a
decline
in
CYP1A1
activity
was
reported
following
birth
(
Rendic
and
Di
Carlo,

1997).
No
data
on
age­
dependent
differences
in
inducible
levels
of
CYP1A1
were
presented
in
the
review.
For
CYP1A2,
low
fetal
mRNA
levels
and
low
enzyme
activity
in
the
neonate
were
reported.
However,
there
were
similar
levels
of
CYP1A2
enzyme
activity
in
adults
and
in
infants,

and
activity
was
greater
in
young
children
than
in
adults
(
Hakkola
et
al.,
1998a;
1998b).
Based
on
patterns
of
theophylline
clearance
as
a
measure
of
CYP1A2
activity,
CYP1A2
is
minimally
62
active
in
fetuses,
reaches
maximum
activity
during
early
childhood
and
decreases
thereafter
(
EPA,

2000d).

The
human
CYP
isoforms
CYP2A6,
CYP2D6,
and
CYP3A4
are
potential
candidates
for
metabolism
of
brominated
THMs,
based
on
overlapping
catalytic
activity
with
rodent
CYP2B1/
2
(
WHO,
2000),
but
the
identity
of
the
human
CYP
isoforms,
other
than
CYP2E1,
capable
of
metabolizing
brominated
THMs
is
not
known.
In
light
of
this
knowledge
gap,
the
toxicological
consequences
of
their
developmental
expression
patterns
are
not
clear.
However,
except
for
CYP1A2,
fetuses
and
children
generally
have
lower
expression
than
adults
of
the
CYPs
that
may
be
involved
in
brominated
THM
metabolism
(
Hakkola
et
al.,
1998a,
1998b;
Tanaka,
1998;

Oesterheld,
1998).

Higher
enzyme
activity
will
not
necessarily
result
in
higher
tissue
doses
of
metabolite,

since
at
low
doses
(
i.
e.,
much
lower
than
the
Km),
the
amount
of
metabolism
would
be
controlled
by
the
amount
of
substrate,
not
the
amount
of
enzyme.
In
addition,
if
metabolism
of
the
brominated
THMs
by
CYP1A,
CYP2B1/
2,
or
their
analogues
represents
a
low
affinity
route
of
metabolism
(
as
it
does
for
chloroform),
the
impact
of
age­
dependent
expression
of
these
enzymes
at
environmental
exposure
levels
may
be
less
important
than
for
CYP2E1.
No
data
were
identified
on
the
relative
affinity
of
the
brominated
THMs
for
the
different
CYP
isoforms.

Overall,
the
P­
450
enzyme
expression
and
activity
data
do
not
suggest
that
children
would
be
more
susceptible
than
adults
to
brominated
THM.

Only
minimal
data
are
available
regarding
the
enzymes
involved
in
the
reductive
metabolism
of
brominated
trihalomethanes.
While
experimental
evidence
indicates
that
CYP2E1
and
CYPB1/
2
catalyze
the
oxidative
pathway,
the
identities
of
the
cytochrome
P450
isoforms
that
63
catalyze
the
reductive
pathway
have
not
been
established.
In
general,
CYP2E1
protein
can
catalyze
reductive
as
well
as
oxidative
reactions
(
Lieber,
1997)
and
this
isoform
has
been
implicated
in
the
production
of
trichloromethyl
radicals
from
carbon
tetrachloride
(
see
Lieber
et
al.
1997).
However,
evidence
for
a
dual
role
of
either
CYP2E1
or
CYP2B1/
2
in
catalyzing
the
oxidative
and
reductive
pathways
for
trihalomethane
metabolism
has
been
contradictory,
perhaps
as
a
result
of
the
different
concentrations
of
chloroform
used
in
different
experiments
(
summarized
in
Testai
et
al.,
1996).
To
address
the
issue
of
concentration,
Testai
et
al.
(
1996)

studied
the
role
of
different
isoforms
in
chloroform
metabolism,
and
they
found
that
the
cytochrome
P450
isoforms
involved
in
oxidative
metabolism
of
brominated
trihalomethanes
do
not
participate
in
the
reductive
pathway.
Thus,
no
conclusion
can
be
made
on
whether
children
are
more
sensitive
than
adults
based
on
the
limited
data
on
the
reductive
pathway.

The
information
on
age­
related
differences
in
the
production
of
metabolites
via
the
glutathione
conjugation
pathway
is
also
very
limited.
Studies
in
Salmonella
typhimurium
strains
engineered
to
express
the
rat
glutathione
S­
transferase
theta
1­
1
(
GSTT1­
1)
gene
indicate
that
metabolism
of
brominated
THMs
to
mutagens
may
also
be
catalyzed
by
glutathione
S­
transferase
(
DeMarini
et
al.,
1997;
Landi
et
al.,
1999).
Data
on
the
GST
theta
genes
are
currently
quite
limited;
however,
one
study
reported
that
theta­
class
GSTs
were
expressed
in
human
adult
liver,

but
not
fetal
liver
(
Mera
et
al.,
1994).
These
results
suggest
that
the
fetus
does
not
experience
increased
risk
from
GST
theta­
mediated
mutagenicity.
The
occurrence
of
increased
risk
in
children
cannot
be
evaluated,
since
the
age
at
which
expression
of
GST
theta
begins
is
unknown.

2.2.2.1.
Bromodichloromethane
64
Developmental/
Reproductive
Effects
Developmental
and
reproductive
toxicity
data
are
available
for
bromodichloromethane
(
BDCM)
and
were
considered
in
the
derivation
of
the
MCLG.
Five
epidemiologic
studies
have
reported
a
relationship
specifically
between
developmental/
reproductive
toxicity
and
BDCM.

These
studies
were
summarized
in
detail
in
Section
2.1.1.,
but
key
conclusions
are
highlighted
here.
In
a
prospective
study,
Windham
et
al.
(
2003)
demonstrated
that
increasing
levels
of
individual
brominated
trihalomethanes
or
total
brominated
trihalomethanes
in
the
drinking
water
were
associated
with
significantly
shorter
cycles
when
examined
by
quartile.
Similar
decrements
were
observed
in
follicular,
but
not
luteal,
phase
length.
For
BDCM,
the
adjusted
decrements
were
0.74
days
(
95%
C.
I.
­
1.5,
­
0.02)
for
mean
cycle
length
and
0.8
days
(
95%
C.
I.
­
1.5,
­
0.08)

for
mean
follicular
phase
length
at
the
highest
quartile
(

16

g/
L).
The
strongest
association
for
an
individual
compound
was
observed
for
DBCM.
Menses
length
was
slightly
increased
at
the
highest
quartile
for
BDCM
exposure.

A
population
based
case­
control
study
(
Kramer
et
al.,
1992)
reported
an
increased
risk
(
not
statistically
significant)
of
intrauterine
growth
retardation
with
exposure
to
BDCM
concentrations
in
drinking
water
of
greater
than
or
equal
to
10
µ
g/
L,
compared
with
drinking
water
with
undetectable
BDCM
concentrations
(
OR
=
1.7,
95%
CI
=
0.9
 
2.9).
As
previously
discussed,
Waller
et
al.
(
1998)
reported
that
pregnant
women
exposed
to
TTHMs
in
drinking
water
at
levels
of
75

g/
L
or
higher
had
an
increased
risk
of
spontaneous
abortion.
This
study
also
reported
that
consumption
of
five
or
more
glasses
of
cold
water
with
a
BDCM
concentration
of
at
least
18
µ
g/
L
was
associated
with
an
increased
risk
of
spontaneous
abortion.
After
adjustment
for
exposure
to
other
THMs,
the
adjusted
OR
was
3.0
(
CI
=
1.4­
6.6).
65
King
et
al.
(
2000a)
reported
that
exposures
to
TTHMs,
chloroform
and
BDCM
were
associated
with
an
increased
risk
of
stillbirth.
Risk
doubled
for
women
exposed
to
a
BDCM
level
of
greater
than
or
equal
to
20

g/
L,
when
compared
to
women
consuming
concentrations
of
less
than
5

g/
L.
In
a
retrospective
cohort
study
among
49,842
residents
of
Nova
Scotia
(
Dodds
and
King,
2001),
exposure
to
BDCM
at
concentrations

20

g/
L
was
associated
with
an
increased
risk
of
neural
tube
defects
(
adjusted
relative
risk
=
2.5,
95%
CI
=
1.2
­
5.1),
but
there
was
no
evidence
of
a
dose­
response
trend.

Seven
oral
studies
in
laboratory
animals
evaluated
the
developmental
toxicity
of
BDCM
(
Bielmeier
et
al.,
2001;
CCC,
2000a,
b,
c;
Narotsky
et
al.,
1997a;
NTP,
1998a;
Ruddick
et
al.,

1983).
To
determine
the
effect
of
vehicle
on
BDCM
toxicity,
Narotsky
et
al.
(
1997a)

administered
BDCM
by
gavage
in
either
corn
oil
or
an
aqueous
vehicle
with
Emulphor
®
to
pregnant
F344
rats
(
12
 
14/
group)
at
dose
levels
of
0,
25,
50,
or
75
mg/
kg/
day
during
gestation
days
6
 
15.
Decreased
maternal
weight
gain
and
full­
litter
resorption
were
observed
at
50
and
75
mg/
kg/
day.
The
incidence
of
full­
litter
resorption
was
significantly
higher
in
the
corn
oil
vehicle
(
83%)
compared
with
the
aqueous
vehicle
(
21%)
at
the
high
dose.
Accordingly,
the
NOAEL
for
developmental
toxicity
(
full­
litter
resorptions)
was
25
mg/
kg/
day,
with
a
LOAEL
of
50
mg/
kg/
day.
The
LOAELs
for
maternal
effects
(
based
on
decreased
maternal
weight
gain
during
GD
6­
8)
in
aqueous
and
corn
oil
vehicles
were
25
and
50
mg/
kg/
day,
respectively.
No
maternal
NOAEL
for
the
aqueous
study
could
be
identified.

Bielmeier
et
al.
(
2001)
conducted
a
series
of
experiments
on
the
effect
of
BDCM
on
pregnancy
loss,
characterized
as
full
litter
resorptions,
in
female
F344
and
Sprague­
Dawley
rats.

In
one
experiment,
doses
of
0,
75,
or
100
mg/
kg/
day
were
administered
via
gavage
to
pregnant
66
rats
on
gestation
day
9
in
10%
Emulphor.
The
dose­
related
incidence
of
full
litter
resorptions
was
0%
(
0/
8),
64%
(
7/
11),
and
90%
(
9/
10),
respectively,
identifying
a
LOAEL
of
75
mg/
kg/
day.
This
effect
was
seen
in
F344
rats,
but
not
in
Sprague­
Dawley
rats.
Dosing
during
different
portions
of
pregnancy
showed
that
the
critical
period
for
induction
of
full
litter
resorption
was
limited
to
the
luteinizing
hormone
(
LH)­
dependent
phase
of
pregnancy,
suggesting
that
BDCM
may
disrupt
pregnancy
via
a
LH­
mediated
mode
of
action.
Measurement
of
serum
LH
and
progesterone
levels
indicated
that
full
litter
resorption
was
accompanied
by
a
marked
reduction
in
progesterone
concentration
without
a
corresponding
drop
in
LH
levels.
The
failure
of
BDCM
to
exert
adverse
effects
after
the
LH­
dependent
window,
the
reduction
in
serum
progesterone
level,
and
the
unchanged
serum
LH
levels
led
this
group
to
conclude
that
the
target
of
toxicity
was
the
ovary
and
that
the
mode
of
action
was
a
reduced
sensitivity
of
the
corpus
luteum
to
LH.
Although
there
are
significant
differences
between
rats
and
humans
in
the
hormonal
maintenance
of
pregnancy,

the
authors
did
consider
their
findings
possibly
relevant
to
humans.
Additional
research,
some
of
which
is
currently
ongoing,
is
needed
to
evaluate
the
relevance
to
humans
of
these
findings
in
rats.

Ruddick
et
al.
(
1983)
investigated
developmental
toxicity
in
pregnant
Sprague­
Dawley
rats
(
9
 
15/
group)
administered
BDCM
by
gavage
in
corn
oil
at
dose
levels
of
0,
50,
100,
or
200
mg/
kg/
day
from
gestation
days
6
 
15.
Maternal
weight
gain
was
significantly
depressed
in
the
high­
dose
group,
and
reduced
in
the
low­
and
mid­
dose
groups;
the
decreases
were
not
statistically
significant.
There
were
no
fetal
malformations,
but
a
dose­
dependent
increased
incidence
of
sternebra
aberrations
was
observed
in
all
dose
groups.
Although
there
was
a
clear
increase
in
the
incidence
of
these
variations,
no
statistical
analysis
was
performed
by
the
authors.

However,
a
statistical
analysis
(
using
the
Fisher
Exact
Test)
conducted
on
the
published
data
67
found
that
none
of
these
increases
differed
significantly
from
controls.
A
trend
test
showed
a
statistically
significant
dose­
related
trend
(
p=
0.03),
and
stepwise
analysis
indicated
that
the
trend
became
significant
only
when
the
high­
dose
(
200
mg/
kg/
day)
was
included
in
the
analysis.
These
findings
suggest
that
the
NOAEL
and
LOAEL
were
100
mg/
kg/
day
and
200
mg/
kg/
day,

respectively.
However,
due
to
the
small
sample
sizes,
the
statistical
power
of
the
experiment
to
detect
effects
at
lower
doses
is
limited.
The
NOAEL
and
LOAEL
for
maternal
effects
were
100
mg/
kg/
day
and
200
mg/
kg/
day,
respectively,
based
on
significantly
decreased
maternal
weight
gain
(
EPA
2002c).

NTP
(
1998a)
conducted
a
short­
term
study
screening
study
in
Sprague­
Dawley
rats
investigating
both
reproductive
and
developmental
toxicity
from
BDCM
administered
in
drinking
water.
Two
groups
of
male
rats
and
three
groups
of
female
rats
were
treated
with
BDCM
at
concentrations
of
0,
100,
700,
or
1300
mg/
L.
Based
on
measured
water
consumption,
the
authors
estimated
dose
levels
for
the
treated
males
to
be
0,
8,
41,
or
68
mg/
kg/
day,
and
for
the
treated
females
to
be
0,
14,
72,
or
116
mg/
kg/
day
(
groups
A
and
C)
and
0,
13,
54,
or
90
mg/
kg/
day
(
group
B).
The
rats
were
exposed
for
25
to
30
days,
with
the
exception
of
group
B
females,
which
were
exposed
from
gestation
day
6
to
evidence
of
littering/
birth
(
approximately
15
days).
BDCM
exposure
did
not
affect
any
reproductive
parameter
investigated
in
males
or
females,
with
the
exception
of
a
non­
dose
related
increase
in
the
number
of
live
fetuses
per
birth
at
the
14
mg/
kg/
day
dose
in
Group
C
females
and
a
slight
decrease
in
the
number
of
live
fetuses
per
birth
at
the
72
mg/
kg/
day
dose
in
Group
A
females.
On
the
basis
of
these
results,
NTP
(
1998a)
concluded
that
BDCM
was
not
a
short­
term
reproductive
or
developmental
toxicant
at
doses
up
to
approximately
68
and
116
mg/
kg/
day
in
male
and
female
rats,
respectively.
The
study
68
sensitivity
was
decreased
by
the
use
of
a
relatively
small
number
of
animals
per
group
(
5­

13/
sex/
dose)
and
the
lack
of
microscopic
examination
of
the
pups.

The
Chlorine
Chemistry
Council
sponsored
a
range­
finding
reproductive/
developmental
toxicity
study
of
BDCM
in
Sprague­
Dawley
rats
(
CCC,
2000a;
Christian
et
al.,
2001b).
BDCM
was
administered
to
parental
rats
(
P
generation,
10/
sex/
group)
in
drinking
water
at
concentrations
of
0,
50,
150,
450,
or
1350
ppm.
Exposure
began
14
days
before
cohabitation
and
continued
until
the
day
of
sacrifice.
Lactation
was
extended
for
one
week
(
LD
22­
29)
beyond
the
normal
3­

week
period
because
F
1
pup
body
weights
in
the
three
highest
dose
groups
were
significantly
reduced
on
LD
21
relative
to
control
values
(
results
are
described
below).
On
LD
29,
two
F
1
pups
per
sex
were
selected
from
each
litter
for
an
additional
week
of
postweaning
observation,

provided
ad
libitum
access
to
water
containing
the
same
concentration
of
BDCM
administered
to
their
parents
(
P
generation),
and
sacrificed
on
Day
8
postweaning.

Exposure­
dependent
reductions
in
both
absolute
(
g/
day)
and
relative
(
g/
kg
body
weightday
water
consumption
were
observed
in
all
rats
of
both
sexes
and
were
attributed
to
taste
aversion.
Treatment­
related
clinical
signs
(
e.
g.,
dehydration,
emaciation,
chromorhinorrhea)
were
observed
in
both
sexes
in
the
1350
ppm
exposure
groups
and
were
considered
to
be
generally
associated
with
reduced
water
consumption.
The
only
other
observed
effect
was
a
concentrationdependent
reduction
in
F
1
pup
body
weights
and
weight
gain
in
the
150,
450,
and
1350
ppm
exposure
groups,
both
prior
to,
and
after
weaning.
Based
on
decreased
pup
weight
and
pup
weight
gain,
the
LOAEL
for
developmental
toxicity
is
150
ppm,
and
the
corresponding
NOAEL
is
50
ppm.
Although
the
effect
of
reduced
water
consumption
on
the
decreases
in
feed
consumption,
body
weight
gain,
and
body
weight
observed
in
the
P
generation
adults
is
unclear,
69
the
LOAEL
for
parental
toxicity
is
considered
to
be
150
ppm
and
the
NOAEL
is
50
ppm.
Due
to
the
marked
changes
in
drinking
water
consumption
by
P
generation
female
rats
during
different
physiological
stages
(
pre­
mating,
mating,
gestation,
and
lactation),
it
is
not
possible
to
convert
the
administered
drinking
water
concentrations
into
a
single
biologically
meaningful
average
daily
dose;
instead,
dose
in
mg/
kg/
day
was
reported
separately
for
the
premating,
gestation,
lactation,

and
post­
weaning
phases
of
the
study.

In
the
full
study,
the
Chlorine
Chemistry
Council
(
CCC,
2000b;
summarized
in
Christian
et
al.,
2001a)
examined
the
developmental
effects
of
BDCM
in
Sprague­
Dawley
rats.
Pregnant
rats
were
given
BDCM
in
drinking
water
at
concentrations
of
0,
50,
150,
450,
or
900
mg/
L
(
0,
2.2,

18.4,
45.0,
or
82.0
mg/
kg/
day)
on
gestation
days
6­
21.
No
treatment­
related
clinical
signs
or
necropsy
results
were
observed.
Significantly
reduced
water
consumption
was
observed
in
all
dose
groups
and
was
attributed
to
taste
aversion.
Decreases
in
absolute
and
relative
feed
consumption
were
observed
in
the
three
highest
dose
groups;
associated
decreases
in
maternal
body
weight
gain
were
attributed
to
taste
aversion.
The
effect
on
maternal
body
weight
gain
was
persistent
at
the
two
highest
doses,
but
was
transient
at
the
lower
doses.
No
effects
on
early
or
late
resorptions,
fetal
body
weight,
liver
litter
size,
or
other
developmental
parameters
were
observed
from
BDCM
exposure.
There
were
no
instances
of
full
litter
resorption
and
no
dead
fetuses.
The
only
statistically
significant
changes
in
the
occurrence
of
skeletal
variations
were
reversible
delays
in
ossification
at
the
high
dose,
including
an
increased
fetal
incidence
of
wavy
ribs
and
a
decreased
number
of
ossification
sites
per
fetus
per
litter
for
the
forelimb
phalanges
and
the
hindlimb
metatarsals
and
phalanges.
The
study
authors
did
not
consider
the
increased
fetal
incidence
of
wavy
ribs
to
be
related
to
BDCM
exposure,
because
the
more
relevant
measure
of
70
litter
incidence
did
not
differ
significantly
from
the
controls
and
was
within
the
historical
range
for
this
alteration
at
the
test
facility.
The
maternal
NOAEL
and
LOAEL
were
18.4
mg/
kg/
day
and
45.0
mg/
kg/
day,
respectively,
based
on
significant
reductions
in
maternal
body
weight
and
body
weight
gains,
while
the
developmental
NOAEL
and
LOAEL
were
45.0
mg/
kg/
day
and
82.0
mg/
kg/
day,
respectively,
based
on
a
significant
decrease
in
the
number
of
ossification
sites
per
fetus
for
the
forelimb
phalanges
and
the
hindlimb
metatarsals
and
phalanges.

The
Chlorine
Chemistry
Council
also
examined
the
developmental
effects
of
BDCM
in
New
Zealand
white
rabbits
(
CCC,
2000c).
Pregnant
rabbits
were
given
BDCM
in
drinking
water
at
concentrations
of
0,
15,
150,
450,
or
900
mg/
L
(
0,
1.4,
13.4,
35.6,
or
55.3
mg/
kg/
day)
on
gestation
days
6­
29.
No
treatment­
related
clinical
signs
or
necropsy
results
were
observed.

Significantly
reduced
feed
and
water
consumption
and
body
weight
gains
were
observed
in
the
35.6
and
55.3
mg/
kg/
day
dose
groups.
No
observable
effects
on
fetal
body
weight,
live
litter
size,

and
a
number
of
other
developmental
parameters
were
noted
from
BDCM
exposure.
Statistically
significant
increases
in
the
number
of
fused
sterna
centra
were
observed
in
the
13.4
and
35.6
mg/
kg/
day
dose
groups;
however,
this
effect
was
not
dose­
related,
and
the
observed
incidences
were
within
the
historical
range
for
the
testing
facility.
The
maternal
NOAEL
and
LOAEL
identified
in
this
study
were
13.4
and
35.6
mg/
kg/
day,
respectively,
based
on
decreased
body
weight
gain,
while
the
developmental
NOAEL
was
55.3
mg/
kg/
day,
based
on
an
absence
of
statistically
significant,
dose­
related
effects
at
any
dose
tested.

Christian
et
al.
(
2002)
conduced
a
standard
two­
generation
reproductive
toxicity
study
in
which
BDCM
was
continuously
provided
to
Sprague­
Dawley
rats
in
the
drinking
water
at
concentrations
of
0,
50,
150,
or
450
ppm.
Average
daily
doses
estimated
by
the
study
authors
71
were
4.1
to
12.6,
11.6
to
40.2,
and
29.5
to
109
mg/
kg­
day,
respectively.
Exposure
of
the
parental
generation
(
P)
was
initiated
when
the
test
animals
were
approximately
43
days
of
age
and
continued
through
a
70­
day
pre­
mating
period
and
a
cohabitation
period.
P
generation
males
were
exposed
for
approximately
106
days
prior
to
sacrifice.
Exposure
of
P
generation
female
rats
continued
through
gestation
and
lactation.
F
1
generation
rats
were
exposed
to
BDCM
in
utero
and
by
consumption
of
the
dam's
drinking
water
during
the
lactation
period.
At
weaning,
F
1
rats
(
30/
sex/
concentration)
were
selected
for
a
postweaning/
premating
exposure
period,
followed
by
a
cohabitation
period,
and
the
exposure
continued
through
gestation
and
lactation.
F
1
generation
females
delivered
litters
and
the
F
2
litters
were
sacrificed
on
lactation
day
22.

Deaths
at
150
and
450
ppm
were
associated
with
reduced
water
consumption,
weight
loss
and/
or
adverse
clinical
signs
and
may
have
been
compound­
related.
Adverse
clinical
signs
occurred
at
150
and
450
ppm
were
attributed
to
reduced
water
consumption.
Body
weight
and
body
weight
gain
were
significantly
reduced
in
the
450
ppm
P
generation
males
and
females
and
150
and
450
ppm
F
1
generation
males
and
females,
and
was
associated
with
decreased
food
consumption.
Rats
at
450
ppm
also
had
decreased
absolute
and
relative
organ
weights.
Water
consumption
was
significantly
reduced
in
P
and
F
1
generation
males
and
females
at
all
concentrations
of
BDCM,
and
was
reduced
by
10­
20%
at
150
and
450
ppm.
There
were
no
gross
pathological
or
histopathological
indications
of
compound­
related
toxicity.
Most
indicators
of
reproductive
or
developmental
toxicity
examined
were
not
significantly
affected
by
BDCM
treatment.
However,
decreased
pup
body
weight
was
observed
at
weaning
and/
or
during
lactation
in
the
F
1
and
F
2
generations
in
the
150
and
450
ppm
groups.
Small,
but
statistically
significant,
delays
in
F
1
generation
sexual
maturation
occurred
at
150
(
males)
and
450
ppm
72
(
males
and
females)
as
determined
by
timing
of
vaginal
patency
or
preputial
separation,
but
no
statistically
significant
delays
were
seen
when
body
weight
at
weaning
was
included
as
a
covariate
in
the
analysis.
Delayed
estrus
was
observed
in
F
1
females
in
the
450
ppm
exposure
group.
The
study
authors
also
considered
this
effect
to
be
a
secondary
response
associated
with
reduced
pup
weights.
The
results
of
this
study
appear
to
identify
NOAEL
and
LOAEL
values
for
reproductive
effects
of
50
ppm
(
4.1
to
12.6
mg/
kg­
day)
and
150
ppm
(
11.6
to
40.2
mg/
kg­
day),
respectively,

based
on
delayed
sexual
maturation.
However,
the
study
authors
have
questioned
whether
delayed
sexual
maturation
in
F
1
males
associated
with
reduced
body
weight
should
be
treated
as
reproductive
toxicity
or
general
toxicity,
since
the
root
cause
appears
to
be
dehydration
brought
about
by
taste
aversion
to
the
compound.
The
parental
NOAEL
and
LOAEL
are
also
50
and
150
ppm,
respectively,
based
on
reduced
body
weight
and
body
weight
gain
in
F
1
males
and
females.

Klinefelter
et
al.
(
1995)
evaluated
the
effects
of
BDCM
exposure
on
male
reproduction
at
an
interim
sacrifice
as
part
of
a
two­
year
bioassay,
in
which
F344
rats
were
administered
BDCM
in
drinking
water
at
concentrations
of
0,
330
mg/
L,
or
620
mg/
L.
The
authors
estimated
the
doses
to
be
0,
22,
and
39
mg/
kg/
day.
At
52
weeks,
the
authors
conducted
an
interim
sacrifice,

which
included
an
evaluation
of
epididymal
sperm
motion
parameters
and
histopathology
of
the
testes
and
epididymides.
No
histologic
alterations
were
observed
in
any
reproductive
tissue.

Sperm
velocities
(
mean
straight­
line,
average
path,
and
curvilinear),
however,
were
significantly
decreased
at
39
mg/
kg/
day.
No
effect
on
sperm
motility
was
observed
at
22
mg/
kg/
day.
The
NOAEL
and
LOAEL
for
reproductive
effects
are
thus
22
and
39
mg/
kg/
day,
respectively.

Several
studies
on
the
developmental
toxicity
of
BDCM
gave
negative
results
at
doses
up
to
116
mg/
kg/
day
in
Sprague­
Dawley
rats
(
NTP,
1998a)
and
55.3
mg/
kg/
day
in
rabbits
(
CCC,
73
2000c)
when
administered
in
drinking
water.
However,
in
other
studies
(
CCC,
2000b;
Ruddick
et
al.,
1983),
slightly
decreased
numbers
of
ossification
sites
in
the
hindlimb
and
forelimb
were
observed
in
fetuses
of
Sprague­
Dawley
rats
administered
82
mg/
kg/
day
in
the
drinking
water
on
gestation
days
6
to
21
(
CCC,
2000b)
and
sternebral
aberrations
were
observed
in
the
offspring
of
Sprague­
Dawley
rats
administered
200
mg/
kg/
day
by
gavage
in
corn
oil
on
gestation
days
6
to
15
(
Ruddick
et
al.,
1983).
Reductions
in
mean
pup
weight
gain
and
pup
weight
were
observed
when
parental
Sprague­
Dawley
rats
were
administered
BDCM
in
the
drinking
water
at
concentrations
of
150
ppm
and
above
(
a
biologically
single
meaningful
estimate
of
intake
on
a
mg/
kg/
day
basis
could
not
be
calculated
for
this
study)
(
CCC,
2000a).
Full
litter
resorption
has
been
noted
in
F344
rats
(
Narotsky
et
al.,
1997a;
Bielmeier
et
al.,
2001),
but
not
Sprague­
Dawley
rats,
treated
with
BDCM
at
doses
of
50
to
100
mg/
kg/
day
during
gestation
days
6
to
10
or
with
a
single
dose
of
75
mg/
kg/
day
on
gestation
day
9.
Based
on
these
observations,
it
appears
that
BDCM
administered
in
the
drinking
water
can
induced
various
developmental
effects,
and
the
types
of
these
effects
are
species­
and
strain­
dependent.

Systemic
Effects
BDCM
causes
decreased
weight
gain
and
various
adverse
effects
in
the
nervous
and
immune
systems,
thyroid,
kidney,
and
liver,
but
the
predominant
systemic
effects
from
acute
and
chronic
exposure
to
BDCM
are
on
the
liver
and
kidney.
NTP
(
1987)
administered
BDCM
to
rats
by
gavage
in
corn
oil
at
doses
of
0,
50
or
100
mg/
kg/
day
for
102
weeks.
Histologic
alterations
in
the
liver
and
kidney
were
observed
at
50
mg/
kg/
day
and
higher.
In
a
similar
study
in
mice
(
NTP,

1987),
histologic
alterations
in
the
liver,
kidney
and
thyroid
of
male
mice
were
noted
at
25
74
mg/
kg/
day
and
higher
doses.
Aida
et
al.
(
1992)
administered
microencapsulated
BDCM
to
rats
in
the
diet
at
dose
levels
ranging
from
6
to
168
mg/
kg/
day
for
24
months.
At
6
mg/
kg/
day,
liver
fatty
degeneration
and
granuloma
were
observed.

Carcinogenicity
BDCM
has
been
found
to
cause
tumors
in
multiple
target
organs
of
multiple
species.

Tumors
were
observed
in
the
large
intestine
and
kidneys
of
male
and
female
rats,
kidneys
of
male
mice
and
livers
of
female
mice
when
these
rodents
were
treated
with
BDCM
in
corn
oil
via
gavage
in
a
2­
year
bioassay
(
NTP,
1987).
The
NTP
concluded
that
under
the
conditions
of
these
2­
year
gavage
studies,
clear
evidence
of
carcinogenic
activity
existed
for
male
and
female
rats
and
mice.
A
study
investigating
the
relationship
between
liver
toxicity
and
tumorigenicity
of
BDCM
concluded
that
the
shape
of
the
dose­
response
curve
was
different
for
these
two
effects
(
Melnick
et
al.,
1998).
The
authors
concluded
that
there
does
not
appear
to
be
a
causal
relationship
between
liver
toxicity
and
tumor
development
for
BDCM.

In
another
rat
study
(
George
et
al.,
2002),
BDCM
induced
heptocellular
adenomas
and
carcinomas
in
male
rats
exposed
to
the
compound
via
drinking
water
for
104
weeks.
However,
a
biphasic
dose­
response
was
observed;
the
significant
increase
in
hepatocellular
neoplasmas
only
occurred
in
the
low
dose
group
(
8/
45
vs.
2/
45
in
the
control
group)
and
not
in
the
mid­
and
highdose
groups
(
7/
48
and
4/
49,
respectively).
The
underlying
basis
for
the
biphasic
response
is
unknown,
but
this
pattern
of
response
might
be
explained
by
inhibition
of
the
hepatic
metabolism
of
BDCM
by
the
compound
itself.
The
same
authors
(
George
et
al.,
2002)
also
reported
that
BDCM
was
not
carcinogenic
to
male
mice,
but
there
is
no
evidence
that
an
adequately
high
dose
75
was
tested
in
this
study.
DeAngelo
(
2002)
reported
mixed
results
on
BDCM
induction
of
aberrant
crypt
foci
(
ACF),
putative
early
preneoplastic
lesions
of
the
colon.
The
authors
observed
increased
ACF
in
a
13­
week
drinking
water
rat
study,
but
not
in
13­
or
30­
week
mouse
studies.

The
study
authors
concluded
that
ACF
induced
by
BDCM
do
not
progress
to
neoplasia,
as
judged
by
the
absence
of
colon
neoplasms
in
the
two­
year
cancer
study
conducted
by
George
et
al.

(
2002).

BDCM
has
shown
mixed
results
in
genotoxicity
assays,
with
both
positive
(
Fujie
et
al.,

1990;
Simmon
and
Tardiff,
1978)
and
negative
results
(
Hayashi
et
al.,
1988;
NTP,
1987;
Varma
et
al.,
1988)
from
in
vitro
and
in
vivo
studies.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Although
data
are
mixed,
the
weight
of
evidence
indicates
that
BDCM
is
genotoxic.
Recent
studies
in
Salmonella
strains
engineered
to
contain
rat
GST­
theta
suggest
that
the
mutagenicity
of
the
brominated
THMs
may
be
mediated
by
glutathione
conjugation
(
DeMarini
et
al.,
1997;
Landi
et
al.,
1999).
Furthermore,
a
recent
study
in
female
B6C3F1
mice
(
Melnick,
et
al.,
1998)
suggests
that
increased
incidence
of
hepatic
tumors
occurs
at
doses
of
BDCM
that
have
no
effect
on
hepatocyte
labeling
index,
indicating
that
regenerative
hyperplasia
is
not
required
for
tumor
induction,
supporting
the
choice
of
linear
low­
dose
extrapolation
for
quantification
of
cancer
risk
associated
with
BDCM.

Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment,
BDCM
is
classified
as
Group
B2:
Probable
Human
Carcinogen
(
EPA,
2002c).
Under
the
EPA's
1999
Draft
76
Guidelines
for
Carcinogen
Risk
Assessment,
BDCM
is
likely
to
be
carcinogenic
to
humans
(
EPA,

2002c).
This
descriptor
is
considered
appropriate
when
there
are
no
or
inadequate
data
in
humans,
but
the
combined
experimental
evidence
demonstrates
the
production
or
anticipated
production
of
tumors
by
modes
of
action
assumed
to
be
relevant
to
humans.
Based
on
the
evidence
for
BDCM
genotoxicity,
and
the
lack
of
data
supporting
a
mode
of
action
for
which
nonlinear
extrapolation
to
low
doses
are
appropriate,
a
linear
extrapolation
to
low
doses
is
used
for
quantification
of
the
BDCM
cancer
risk.

Under
the
1986
cancer
guidelines,
the
Agency
calculated
a
cancer
oral
slope
factor,
based
on
renal
tumors
in
treated
male
mice,
of
6.2
×
10­
2
per
mg/
kg/
day
and
a
unit
risk
of
1.8
x
10­
6
(
µ
g/
L)­
1
(
EPA,
2002c).
Based
on
the
same
endpoint,
a
cancer
oral
slope
factor
of
8.1
x
10­
3
per
mg/
kg/
day
has
been
calculated
under
the
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,

1999),
corresponding
to
a
unit
risk
of
2.3
x
10­
7
(
µ
g/
L)­
1
(
EPA,
2002c).
Based
on
this
oral
slope
factor,
drinking
water
concentrations
of
400
µ
g/
L,
40
µ
g/
L,
and
4
µ
g/
L
are
estimated
to
be
associated
with
estimated
lifetime
risks
of
10­
4
,
10­
5,
and
10­
6.

Basis
for
RfD
and
MCLG
EPA
selected
the
chronic
study
by
Aida
et
al.
(
1992)
as
the
most
appropriate
study
for
derivation
of
the
RfD.
This
study
identified
a
LOAEL
of
6
mg/
kg/
day
based
on
liver
fatty
degeneration
and
granuloma
in
male
rats,
as
well
as
a
BMD
of
1.9
mg/
kg/
day
and
a
corresponding
BMDL
10
of
0.8
mg/
kg/
day,
based
on
the
same
endpoint.
The
LOAEL
of
6.1
mg/
kg/
day
was
used
as
the
basis
for
the
RfD.
An
UF
of
3000
was
applied
to
the
LOAEL:
a
10­
fold
UF
for
interspecies
extrapolation;
another
10­
fold
factor
for
protection
of
sensitive
human
77
subpopulations;
a
10­
fold
factor
for
extrapolation
from
a
LOAEL;
and
a
3­
fold
factor
for
database
deficiencies,
including
lack
of
a
multigeneration
reproductive
toxicity
study
and
uncertainty
related
to
possible
reproductive
or
developmental
effects
suggested
by
epidemiological
studies
(
EPA,
2002c).
These
calculations
result
in
an
RfD
of
0.002
mg/
kg/
day.
The
MCLG
for
BDCM
is
zero,
based
on
its
probable
human
carcinogenicity,
and
linear
low­
dose
extrapolation
in
light
of
the
evidence
for
a
genotoxic
mode
of
action,
coupled
with
the
absence
of
evidence
for
other
modes
of
action.

Children's
Risk
in
Relation
to
the
MCLG
The
Agency
believes
that
the
MCLG
of
zero
is
protective
for
both
the
carcinogenic
and
systemic
effects
of
BDCM
in
children
and
adults.
There
is
not
sufficient
evidence
from
studies
on
the
systemic
effects
of
BDCM,
or
on
the
metabolism
of
the
compound,
to
determine
whether
children
are
more
sensitive
to
the
toxic
effects
of
BDCM
than
are
adults.

2.2.2.2.
Dibromochloromethane
Developmental/
Reproductive
Effects
In
a
prospective
study,
Windham
et
al.
(
2003)
demonstrated
that
increasing
levels
of
individual
brominated
THMs
or
total
brominated
THMs
in
the
drinking
water
were
associated
with
significantly
shorter
cycles
when
examined
by
quartile.
Similar
decrements
were
observed
in
follicular,
but
not
luteal,
phase
length.
The
strongest
association
for
an
individual
compound
was
observed
for
DBCM
with
adjusted
decrements
of
1.2
days
(
95%
C.
I.
­
2.0,
­
0.38)
for
mean
cycle
78
length
and
1.1
days
(
95%
C.
I.
­
1.9,
­
0.25)
for
mean
follicular
phase
length
at
the
highest
quartile
(

20
µ
g/
L).

Ruddick
et
al.
(
1983)
investigated
developmental
toxicity
in
pregnant
Sprague­
Dawley
rats
(
9­
15/
group)
administered
DBCM
by
gavage
in
corn
oil
at
dose
levels
of
0,
50,
100,
or
200
mg/
kg/
day
from
gestation
days
6
to
15.
Although
maternal
toxicity
was
indicated
by
a
significant
decrease
in
body
weight
gain
at
the
highest
dose,
no
treatment­
related
developmental
toxicity
was
observed
for
DBCM.
No
statistical
analysis
was
performed
in
this
study
and
inspection
of
the
data
revealed
no
dose­
related
effects.
The
power
of
the
experiment
was
limited
by
the
small
number
of
litters
per
dose
group.
Thus,
the
developmental
NOAEL
was
200
mg/
kg/
day
and
a
LOAEL
could
not
be
identified.
Borzelleca
and
Carchman
(
1982)
conducted
a
two­
generation
reproductive
study
in
ICR
Swiss
mice.
Nine­
week­
old
mice
(
10
males
and
30
females
per
dose
group)
were
continuously
maintained
on
drinking
water
containing
0,
100,
1000,
or
4000
mg/
L
DBCM
(
0,
17,
171,
or
685
mg/
kg/
day).
Based
on
postnatal
body
weight
in
the
F
2
b
pups,
this
study
identified
a
marginal
LOAEL
of
17
mg/
kg/
day
for
DBCM,
and
a
NOAEL
could
not
be
determined.
This
LOAEL
was
considered
to
be
minimal
because
the
body
weight
decrease
was
only
observed
at
one
time
point,
the
effect
was
only
noted
in
one
of
the
two
litters
in
the
F
2
generation,
no
other
adverse
effects
were
noted
at
the
dose
level,
and
it
was
unclear
from
the
report
how
many
litters
and
pups
per
litter
were
examined
for
postnatal
body
weight.

NTP
(
1996)
conducted
a
short­
term
reproductive
toxicity
screening
study
on
Sprague­

Dawley
male
and
female
rats.
A
group
of
male
rats
and
two
groups
of
female
rats
were
treated
with
DBCM
in
drinking
water
at
concentrations
of
0,
50,
150,
or
450
mg/
L
(
10
rats/
dose/
group)

during
a
study
period
of
35
days
(
from
gestation
day
6
through
parturition).
Males
were
treated
79
on
study
days
6­
34.
Group
A
females
were
treated
on
study
days
1­
34,
and
were
mated
to
treated
males
on
study
days
13­
18.
Group
B
females
were
mated
on
study
day
1
to
untreated
males
and
treated
from
gestation
day
6
through
parturition.
Based
on
measured
water
consumption,
the
authors
estimated
dose
levels
for
the
treated
males
to
be
0,
4.2,
12.4,
or
28.2
mg/
kg/
day,
and
for
treated
females
to
be
6.3,
17.4,
or
46.0
mg/
kg/
day
(
Group
A)
and
7.1,
20.0,

or
47.8
mg/
kg/
day
(
Group
B).
The
developmental
toxicity
of
the
offspring
from
these
treated
rats
was
compared
with
those
from
the
control
group.
No
significant
reproductive/
developmental
toxicity
was
observed
at
any
dose
level.
The
NOAEL
for
reproductive/
developmental
effects
identified
in
this
study
was
28.2
mg/
kg/
day
for
males
and
47.8
mg/
kg/
day
for
females;
no
LOAEL
was
identified.
The
study
sensitivity
was
decreased
by
the
use
of
a
relatively
small
number
of
animals
per
group
and
the
lack
of
microscopic
examination
of
the
pups.

Systemic
Effects
DBCM
causes
decreased
weight
gain
and
various
adverse
effects
in
the
nervous
and
immune
systems,
kidneys,
and
liver.
The
predominant
effects
from
acute
and
chronic
exposure
to
DBCM
are
on
the
liver
and
kidney.
Tobe
et
al.
(
1982)
administered
microencapsulated
DBCM
in
the
diet
to
rats
for
24
months
and
reported
decreased
body
weight
and
changes
in
clinical
chemistry
parameters
and
gross
liver
appearance
in
males
at
49
mg/
kg/
day;
similar
effects
were
seen
in
females
at
slightly
higher
administered
doses.
Tobe
et
al.
(
1982)
identified
a
NOAEL
of
12
mg/
kg/
day,
but
this
NOAEL
is
uncertain
in
light
of
the
lack
of
histopathological
examination
and
the
results
of
Aida
et
al.
(
1992)
with
BDCM,
finding
that
adverse
liver
histopathology
occurs
at
doses
lower
than
those
observed
by
Tobe
et
al.
(
1982)
based
on
clinical
chemistry
and
gross
80
tissue
examination.
NTP
(
1985)
reported
histologic
lesions
in
the
liver
in
male
and
female
rats
at
40
mg/
kg/
day
and
lesions
in
the
liver
and
thyroid
in
female
mice
at
50
mg/
kg/
day
after
exposure
to
DBCM
via
gavage
in
oil
for
104
weeks;
no
NOAELs
were
identified.
NTP
(
1985)
evaluated
the
toxicologic
effects
of
DBCM
after
a
subchronic
(
13
week)
exposure.
Hepatic
lesions
were
identified
at
60
mg/
kg/
day,
with
a
subchronic
NOAEL
30.
Adjusting
for
exposure
5
days/
week,

the
subchronic
NOAEL
was
21.4
mg/
kg/
day,
and
the
LOAEL
was
43
mg/
kg/
day.

Carcinogenicity
Evidence
is
limited
for
DBCM
carcinogenicity
via
ingestion.
An
increase
in
tumors
occurred
in
the
livers
of
male
and
female
mice,
while
no
increase
in
tumors
occurred
in
male
and
female
rats,
when
these
rodents
were
treated
with
DBCM
via
gavage
in
corn
oil
for
2
years
(
NTP,

1985).
In
the
mouse
study
(
NTP,
1985),
a
significant
increase
in
hepatocellular
carcinomas
and
adenomas
occurred
in
female
high­
dose
group
(
19/
50
vs.
6/
50).
The
combined
incidence
of
hepatocellular
adenomas
or
carcinomas
in
high
dose
male
mice
(
27/
50
vs.
23/
50)
was
also
significant,
but
this
effect
is
considered
marginal
overall
because
it
was
only
significant
in
the
life
table
test
but
was
not
significant
(
p=
0.065)
by
the
incidental
tumor
test.
The
low­
dose
male
mouse
group
in
this
study
was
considered
to
be
unsuitable
for
data
analysis
because
an
overdose
killed
most
of
the
male
mice.
No
significant
increase
in
carcinomas
and
adenomas
was
observed
in
the
female
low­
dose
group
(
10/
50
vs.
6/
50).
Based
on
these
data,
the
NTP
concluded
that
there
was
equivocal
evidence
of
DBCM
carcinogenicity
in
male
B6C3F1
mice,
some
evidence
of
carcinogenicity
in
female
B6C3F1
mice,
and
no
evidence
of
carcinogenicity
in
male
or
female
rats.

Based
on
a
comparison
of
the
shape
of
the
dose­
response
curve
for
liver
toxicity
and
81
tumorigenicity
of
DBCM,
Melnick
et
al.
(
1998)
concluded
that
there
does
not
appear
to
be
a
causal
relationship
between
liver
toxicity
and
tumor
development
for
DBCM.
De
Angelo
(
2002)

reported
that
in
a
13­
week
drinking
water
study,
DBCM
can
induce
aberrant
crypt
foci
(
ACF),

putative
early
preneoplastic
lesions,
in
the
colons
of
male
rats.
However,
it
is
not
clear
whether
this
change
would
result
in
colon
cancer
in
exposed
rats.

DBCM
has
shown
mixed
results
in
genotoxicity
assays,
with
both
positive
(
e.
g.,
Sekihashi
et
al.,
2002;
Simmon
and
Tardiff,
1978;
Sofuni
et
al.,
1996)
and
negative
results
(
e.
g.,
Hayashi
et
al.,
1988;
NTP,
1985;
Potter
et
al.,
1996)
from
in
vitro
and
in
vivo
studies.
Synthesis
of
the
overall
weight
of
evidence
from
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,

different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
EPA
(
1994b)
has
previously
determined
that
the
weight
of
evidence
for
DBCM
mutagenicity
and
genotoxicity
is
inconclusive.
Recent
studies
in
Salmonella
strains
engineered
to
contain
rat
GST­
theta
suggest
that
the
mutagenicity
of
the
brominated
THMs
may
be
mediated
by
glutathione
conjugation
(
DeMarini
et
al.,
1997).

Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment,
DBCM
is
classified
as
Group
C:
Possible
Human
Carcinogen,
based
on
inadequate
human
data
and
limited
evidence
of
carcinogenicity
in
animals
(
EPA,
2002c).
Based
on
the
1999
Draft
Guidelines
for
Cancer
Risk
Assessment,
there
is
suggestive
evidence
of
human
carcinogenicity
of
DBCM,
but
the
data
are
not
sufficient
to
assess
human
carcinogenic
potential.
A
compound
is
described
as
having
suggestive
evidence
of
carcinogenicity
when
the
evidence
from
human
or
animal
data
is
suggestive
of
carcinogenicity,
but
is
judged
not
sufficient
for
a
conclusion
as
to
human
82
carcinogenic
potential
(
e.
g.,
a
marginal
increase
in
tumors
that
may
be
exposure­
related,
or
evidence
is
observed
only
in
a
single
study).
The
evidence
for
DBCM
carcinogenicity
is
weaker
than
the
evidence
for
the
other
brominated
THMs.
However,
the
evidence
for
DBCM
carcinogenicity
is
strengthened
by
the
evidence
for
carcinogenicity
of
BDCM
and
bromoform,

both
of
which
are
considered
likely
to
be
carcinogenic
to
humans.
BDCM,
DBCM,
and
bromoform
are
all
brominated
THMs
with
increasing
numbers
of
bromine
atoms,
suggesting
that
DBCM
is
likely
to
share
toxic
effects
common
to
both
BDCM
and
bromoform.

Under
the
1986
Guidelines
for
Carcinogen
Risk
Assessment,
the
Agency
calculated
a
cancer
oral
slope
factor
of
8.4
×
10­
2
per
mg/
kg/
day
and
a
unit
risk
of
2.4
x
10­
6
(
µ
g/
L)­
1
based
on
liver
tumors
in
female
mice
dosed
via
oral
gavage
in
corn
oil
(
EPA,
2002c).
Dose­
response
assessment
is
not
recommended
under
the
1999
guidelines
for
chemicals
for
which
the
weight
of
evidence
is
described
as
"
suggestive
evidence
of
human
carcinogenicity,
but
not
sufficient
to
assess
human
carcinogenic
potential."

Basis
for
RfD
and
MCLG
EPA
selected
the
NTP
subchronic
study
in
rats
(
NTP,
1985)
as
the
most
appropriate
basis
for
derivation
of
the
RfD
and
DWEL.
This
study
identified
a
NOAEL
of
30
mg/
kg/
day
(
duration
adjusted
to
21.4
mg/
kg/
day).
An
UF
of
1000
was
applied
to
the
subchronic
NOAEL:
a
10­
fold
factor
for
interspecies
extrapolation;
a
10­
fold
factor
for
protection
of
sensitive
human
subpopulations;
and
a
10­
fold
factor
for
subchronic
to
chronic
extrapolation
(
EPA,
2002c).
The
resulting
RfD
was
0.02
mg/
kg/
day,
corresponding
to
a
DWEL
of
0.7
mg/
L
for
a
70
kg
adult
drinking
2
L
of
water/
day.
83
A
relative
source
contribution
(
RSC)
of
80%
is
used
for
DBCM,
based
on
an
analysis
of
the
anticipated
potential
for
exposure
to
DBCM
in
disinfected
tap
water
via
ingestion,
inhalation,

and
dermal
contact,
and
because
exposure
via
other
media
(
outdoor
air,
food,
and
soil)
are
anticipated
to
be
low
(
EPA,
2002c).
There
are
some
uncertainties
in
the
80%
RSC
that
are
related
to
the
availability
of
adequate
concentration
data
for
DBCM
in
media
other
than
water.

The
MCLG
of
0.06
mg/
L
for
DBCM
is
based
on
a
noncarcinogenic
endpoint
(
the
RfD)

with
an
additional
safety
factor
of
10
to
account
for
possible
carcinogenicity.
Thus,
the
MCLG
is
calculated
as:
MCLG
=
(
0.02
mg/
kg/
day
×
70
kg
×
0.8)/(
2
L/
day
×
10)
=
0.056
mg/
L,
rounded
to
0.06
mg/
L.

Children's
Risk
in
Relation
to
the
MCLG
Developmental
and
reproductive
toxicity
data
are
available
for
DBCM
and
were
considered
in
the
derivation
of
the
RfD
for
DBCM.
The
NOAEL
in
the
two­
generation
reproductive
study
of
Borzelleca
and
Carchman
(
1982)
(
17
mg/
kg/
day)
is
comparable
to
the
duration­
adjusted
NOAEL
(
21.4
mg/
kg/
day)
that
was
the
basis
for
the
RfD.
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.
Nevertheless,
the
Agency
believes
that
the
MCLG
of
0.06
mg/
L
is
protective
of
children's
health,
because
developmental
or
reproductive
effects
have
not
been
found
to
occur
below
the
level
of
the
critical
effect
(
liver
toxicity)
used
to
derive
the
current
RfD,
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.
84
2.2.2.3.
Bromoform
Developmental/
Reproductive
Effects
In
a
prospective
study,
Windham
et
al.
(
2003)
demonstrated
that
increasing
levels
of
individual
brominated
THMs
or
total
brominated
THMs
in
the
drinking
water
were
associated
with
significantly
shorter
cycles
when
examined
by
quartile.
Similar
decrements
were
observed
in
follicular,
but
not
luteal,
phase
length.
For
bromoform,
the
adjusted
decrements
were
0.79
days
(
95%
C.
I.
­
1.4,
­
0.14)
for
mean
cycle
length
and
0.78
days
(
95%
C.
I.
­
1.4,
­
0.14)
for
mean
follicular
phase
length
at
the
highest
quartile
(

12

g/
L).
The
strongest
association
for
an
individual
compound
was
observed
for
DBCM.

Ruddick
et
al.
(
1983)
investigated
developmental
toxicity
in
pregnant
Sprague­
Dawley
rats
(
9
­
15/
group)
administered
bromoform
by
gavage
in
corn
oil
at
dose
levels
of
0,
50,
100,
or
200
mg/
kg/
day
from
gestation
days
6­
15.
No
maternal
toxicity
was
observed
at
any
dose
level;

some
fetal
skeletal
anomalies
were
observed.
Incidences
of
both
fetuses
and
litters
with
interparietal
deviations
were
increased
at
the
mid­
and
high­
dose
groups
compared
with
the
controls.
Furthermore,
incidences
of
both
fetuses
and
litters
with
sternebra
aberrations
increased
in
a
dose­
related
fashion.
No
statistical
analysis
was
performed
by
the
authors,
but
an
independent
statistical
analysis
(
EPA,
2002c,
using
the
Fisher
Exact
test)
demonstrated
that
the
increase
in
sternebral
anomalies
was
significantly
different
from
controls
at
200
mg/
kg/
day.
A
trend
test
showed
a
statistically
significant
dose­
related
trend
(
p<
0.002)
for
this
endpoint;

stepwise
analysis
indicated
that
this
trend
was
no
longer
significant
when
the
two
highest
doses
(
100
mg/
kg/
day
and
200
mg/
kg/
day)
were
removed
from
the
analysis.
These
findings
suggest
that
85
the
NOAEL
and
LOAEL
for
developmental
toxicity
from
this
study
were
50
mg/
kg/
day
and
100
mg/
kg/
day,
respectively.
The
NOAEL
for
maternal
toxicity
was
200
mg/
kg/
day,
and
a
maternal
LOAEL
could
not
be
determined.

The
prenatal
and
postnatal
effects
of
bromoform
on
fertility
and
reproduction
were
investigated
by
NTP
(
1989a)
in
Swiss
CD­
1
mice
using
a
continuous
reproductive
breeding
protocol.
Mice
were
administered
bromoform
by
gavage
in
corn
oil
at
dose
levels
of
0,
50,
100,

or
200
mg/
kg/
day
for
105
days,
including
a
seven­
day
pre­
cohabitation
phase
and
a
98­
day
cohabitation
phase.
F
1
litters
in
the
control
and
200
mg/
kg/
day
groups
were
raised
to
sexual
maturity
(
approximately
74
days)
while
receiving
the
same
treatment
as
their
parents.
At
sexual
maturity,
males
and
females
from
different
litters
within
the
same
treatment
group
were
cohabited
for
seven
days
and
then
housed
individually
until
delivery.
In
the
200
mg/
kg/
day
group,
the
body
weights
of
the
dams
at
delivery
were
consistently
less
than
the
controls.
No
effect
on
any
fertility
or
reproductive
parameter
(
numbers
of
litters
per
pair,
litter
size,
proportion
of
live
pups,
sex
ratio
of
live
pups,
and
pup
body
weight)
was
observed
in
the
P
generation.
However,
postnatal
survival
of
the
F
1
pups
in
the
200
mg/
kg/
day
group
was
significantly
less
than
in
the
control
group.
As
for
the
P
generation,
there
was
no
effect
on
any
mating,
fertility
or
reproductive
parameter
in
the
F
1
generation.
At
sacrifice,
male
and
female
F
1
mice
administered
200
mg/
kg/
day
exhibited
increased
relative
liver
weights
and
decreased
relative
kidney
weights
compared
to
controls.
Histopathological
examination
revealed
minimal
to
moderate
hepatocellular
degeneration
in
the
livers
of
the
200
mg/
kg/
day
male
and
female
mice.
Therefore,

based
on
liver
histopathology,
decreased
postnatal
survival,
and
other
signs
of
toxicity
(
e.
g.,

increased
relative
liver
and
decreased
relative
kidney
weights)
the
developmental
NOAEL
and
86
LOAEL
from
this
study
were
100
mg/
kg/
day
and
200
mg/
kg/
day,
respectively.
The
NOAEL
and
LOAEL
for
maternal
toxicity
were
also
100
mg/
kg/
day
and
200
mg/
kg/
day,
respectively,
based
on
consistently
decreased
body
weights
of
pregnant
dams
at
delivery.

Systemic
Effects
Bromoform
causes
decreased
weight
gain
and
various
adverse
effects
in
the
nervous
system,
kidney,
liver,
and
thyroid,
but
the
predominant
effects
from
acute
and
chronic
exposure
to
bromoform
are
on
the
liver.
Tobe
et
al.
(
1982)
administered
bromoform
microencapsulated
in
the
diet
of
rats
for
24
months
and
reported
gross
liver
lesions
and
changes
in
clinical
chemistry
parameters
in
male
rats
at
90
mg/
kg/
day,
but
this
study
was
not
appropriate
as
the
basis
for
the
RfD,
because
there
was
no
histopathological
analysis.
NTP
(
1989b)
conducted
a
chronic
oral
study
in
rats
and
mice
(
50/
sex/
group)
administered
bromoform
at
0,
100,
or
200
mg/
kg/
day
by
gavage
in
corn
oil.
This
study
observed
histologic
lesions
in
the
liver
(
fatty
changes
and
chronic
inflammation)
at
100
mg/
kg/
day
in
male
and
female
rats,
and
fatty
changes
in
the
liver
of
female
mice
at
the
same
dose.
Female
mice
exposed
to
200
mg/
kg/
day
also
exhibited
follicular
cell
hyperplasia
of
the
thyroid
gland.
This
study
identified
a
LOAEL
of
100
mg/
kg/
day
based
on
liver
effects
in
both
male
and
female
rats
and
in
female
mice;
no
NOAEL
was
identified.
For
male
mice,

a
NOAEL
of
100
mg/
kg/
day
was
identified.

NTP
(
1989b)
also
conducted
a
subchronic
oral
study
in
rats
and
mice.
Male
and
female
F344/
N
rats
(
10/
sex/
dose)
received
0,
12,
25,
50,
100,
or
200
mg/
kg/
day
bromoform
by
gavage
in
oil,
5
days/
week
for
13
weeks.
Male
and
female
B6C3F1
mice
received
0,
25,
50,
100,
200,
or
400
mg/
kg/
day
bromoform,
for
the
same
length
of
time
as
the
rats.
In
male
rats,
a
dose­
87
dependent
increase
in
the
frequency
of
hepatocellular
vacuolation
was
observed,
which
reached
statistical
significance
at
50
mg/
kg/
day.
This
effect
was
not
observed
in
female
rats.
In
male
mice,
a
dose­
dependent
increase
in
the
number
of
hepatocellular
vacuoles
was
seen
at
200
mg/
kg/
day.
Thus,
this
study
identified
a
NOAEL
and
LOAEL
of
25
mg/
kg/
day
and
50
mg/
kg/
day,
respectively,
based
on
hepatocellular
vacuolation
in
male
rats,
and
a
NOAEL
and
LOAEL
of
100
mg/
kg/
day
and
200
mg/
kg/
day,
respectively,
based
on
hepatocellular
vacuoles
in
male
mice.

Carcinogenicity
An
increase
in
tumors
occurred
in
the
large
intestine
in
male
and
female
rats,
while
no
statistically
significant
increase
in
the
incidence
of
tumors
was
seen
in
male
or
female
mice,
when
these
rodents
were
treated
with
bromoform
via
gavage
in
corn
oil
(
NTP,
1989b).
The
NTP
concluded
that
there
was
some
evidence
for
carcinogenic
activity
in
male
rats,
clear
evidence
in
female
rats,
and
no
evidence
in
male
or
female
mice.
De
Angelo
(
2002)
reported
that
in
a
13­

week
drinking
water
study,
bromoform
can
induce
ACF,
putative
early
preneoplastic
lesions,
in
the
colons
of
male
rats.
However,
it
is
not
clear
whether
this
change
would
result
in
colon
cancer
in
exposed
rats.
In
a
study
in
which
bromoform
was
administered
by
intraperitoneal
injection
in
mice,
the
number
of
lung
tumors
per
mouse
in
the
mid­
dose
group
was
significantly
elevated
over
controls
(
Theiss
et
al.,
1977).

Bromoform
has
shown
mixed
results
in
mutagenicity
assays,
with
both
positive
(
Simmon
and
Tardiff,
1978;
Zeiger,
1990)
and
negative
(
Hayashi
et
al.,
1988;
Ishidate
et
al.,
1982;
NTP,

1989a)
results
from
in
vitro
and
in
vivo
studies.
Synthesis
of
the
overall
weight
of
evidence
from
88
these
studies
is
complicated
by
the
use
of
a
variety
of
testing
protocols,
different
strains
of
test
organisms,
different
activating
systems,
different
dose
levels,
different
exposure
methods
(
gas
versus
liquid)
and,
in
some
cases,
problems
due
to
evaporation
of
the
test
chemical.
Although
data
are
mixed,
the
weight
of
evidence
indicates
that
bromoform
is
genotoxic.
In
addition,
recent
studies
in
Salmonella
strains
engineered
to
contain
rat
GST­
theta
suggests
that
the
mutagenicity
of
the
brominated
THMs
may
be
mediated
by
glutathione
conjugation
(
De
Marini
et
al.,
1997;

Landi
et
al.,
1999).

The
evidence
is
sufficient
to
consider
bromoform
a
Group
B2:
Probable
Human
Carcinogen
via
ingestion
under
the
EPA's
1986
Guidelines
for
Cancer
Risk
Assessment
(
EPA,

2002c).
Based
on
the
observed
rare
large
intestinal
tumors
observed
in
both
sexes
of
rats
tested
and
positive
genotoxicity
results,
bromoform
can
be
described
under
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
as
likely
to
be
carcinogenic
to
humans
(
EPA,
2002c).
Based
on
the
evidence
for
bromoform
genotoxicity,
and
the
lack
of
data
supporting
a
mode
of
action
for
which
nonlinear
extrapolation
to
low
doses
is
appropriate,
a
linear
extrapolation
to
low
doses
is
used
to
quantify
the
bromoform
cancer
risk.

The
Agency
calculated
a
cancer
oral
slope
factor
of
7.9
x
10­
3
per
mg/
kg/
day
and
a
unit
risk
of
2.3
x
10­
7
per
(
µ
g/
L)
(
EPA,
2002c),
based
on
the
incidence
of
intestinal
tumors
in
rats,
as
specified
in
the
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1986).
Calculating
a
cancer
slope
factor
based
on
the
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1999),

using
the
same
endpoint
results
in
a
value
of
4.5
x10­
3
(
mg/
kg/
day)­
1,
corresponding
to
a
unit
risk
of
1.3
x
10­
7
(
µ
g/
L)­
1
(
EPA,
2002c).
Based
on
this
calculation,
drinking
water
concentrations
of
89
approximately
800
µ
g/
L,
80
µ
g/
L
and
8
µ
g/
L
are
estimated
to
be
associated
with
lifetime
cancer
risks
of
10­
4,
10­
5
and
10­
6
,
respectively.

Basis
for
RfD
and
MCLG
EPA
derived
the
RfD
for
bromoform
based
on
hepatocellular
vacuolization
in
the
liver
of
male
rats
in
the
NTP
(
1989b)
subchronic
study.
This
study
identified
a
NOAEL
of
25
mg/
kg/
day
and
a
LOAEL
of
50
mg/
kg/
day.
After
adjustment
for
dosing
5
days/
week,
the
duration­
adjusted
NOAEL
and
LOAEL
were
17.9
and
35.7
mg/
kg/
day,
respectively.
The
duration­
adjusted
BMDL
10
was
2.6
mg/
kg/
day.
The
subchronic
study
was
preferred
to
the
chronic
study
as
the
basis
for
the
RfD.
This
is
because
there
was
much
less
uncertainty
in
the
estimate
of
the
BMDL
10
for
the
subchronic
study,
due
to
the
doses
in
the
subchronic
study
being
closer
together,
and
in
light
of
the
high
response
at
the
lowest
dose
of
the
chronic
study.
Furthermore,
a
NOAEL
was
not
identified
in
the
chronic
study,
and
the
critical
effect
was
consistent
across
both
the
subchronic
and
chronic
studies.
Thus,
the
RfD
was
based
on
a
NOAEL
of
17.9
mg/
kg/
day
and
a
composite
UF
of
1000:
10­
fold
factors
each
for
interspecies
variation
and
for
protection
of
sensitive
subpopulations;
and
a
10­
fold
factor
for
subchronic
to
chronic
extrapolation.
The
database
for
bromoform
includes
systemic
toxicity
studies
in
two
species,
a
two­
generation
study
in
mice,
and
a
developmental
toxicity
study
in
rats.
In
light
of
the
other
data
on
the
THMs,
for
which
the
critical
effect
is
systemic
toxicity,
the
single
data
gap
of
a
developmental
toxicity
study
in
a
second
species
is
insufficient
to
require
a
database
UF.
The
resulting
RfD
is
0.02
mg/
kg/
day.
90
The
MCLG
for
bromoform
is
zero,
based
on
its
probable
human
carcinogenicity,
and
linear
low­
dose
extrapolation
in
light
of
the
evidence
for
a
genotoxic
mode
of
action,
coupled
with
the
absence
of
evidence
for
other
modes
of
action.

Children's
Risk
in
Relation
to
the
MCLG
The
Agency
believes
that
the
MCLG
of
zero
is
protective
for
both
the
carcinogenic
and
systemic
effects
of
bromoform
in
children
and
adults.
There
is
no
evidence
from
studies
on
the
systemic
effects
of
bromoform,
or
on
the
metabolism
of
the
compound,
that
children
are
more
sensitive
to
the
toxic
effects
of
bromoform
than
are
adults.

2.3.
HALOACETIC
ACIDS
There
are
only
limited
data
available
for
determining
children's
risk
resulting
from
exposure
to
haloacetic
acids
(
HAAs).
The
haloacetic
acids
have
been
tested
to
varying
degrees
in
developmental
and
reproductive
toxicity
studies
in
animals,
as
described
below
for
each
HAA
covered
in
this
document.
Human
epidemiology
data
on
the
developmental
and
reproductive
toxicity
of
the
haloacetic
acids
are
lacking.
Most
of
the
human
health
data
for
halogenated
acetic
acids
are
as
components
of
complex
mixtures
of
water
disinfection
byproducts.
Although
most
studies
of
human
health
effects
following
exposure
to
water
disinfection
byproducts
have
used
total
trihalomethanes
as
the
exposure
metric,
Klotz
and
Pyrch
(
1999),
conducted
a
case­
control
study
on
the
relationship
between
neural
tube
defects
and
drinking
water
exposure
to
trihalomethanes,
haloacetonitriles
or
haloacetic
acids.
(
See
Section
2.1
for
more
study
details.)
A
statistically
significant
prevalence
odds
ratio
(
POR)
was
reported
for
the
highest
tertile
(
third)
of
91
trihalomethane
exposure;
however,
only
a
slight
non­
statistically
significant
excess
risk
(
POR
1.2:

95%
confidence
interval
0.5­
2.6)
was
found
for
cases
when
analyzed
based
on
total
haloacetic
acids
tertiles.
The
specific
haloacetic
acids
that
were
measured
as
part
of
the
total
haloacetic
acid
exposure
estimate
were
not
identified.
Based
on
the
results
of
the
study,
the
authors
concluded
that
there
was
no
clear
association
of
HAAs
with
neural
tube
defects.

Only
a
subset
of
haloacetic
acids
has
been
tested
in
standard
cancer
bioassays.
The
chlorinated
acetic
acids
have
been
tested,
but
published
reports
of
full
cancer
bioassays
are
not
yet
available
for
their
brominated
counterparts.
Cancer
studies
in
animals
are
described
separately
for
each
compound
below.

The
data
are
insufficient
to
determine
whether
there
are
age­
dependent
differences
in
the
metabolism
of
the
haloacetic
acids
that
might
lead
to
differences
in
health
risk.
The
enzymes
responsible
for
the
metabolism
of
monochloroacetic
acid
(
MCA)
and
monobromoacetic
acid
(
MBA)
have
not
been
identified.
The
metabolism
of
the
dihaloacetic
acids
has
been
more
extensively
studied.
Dichloroacetic
acid
(
DCA),
bromochloroacetic
acid
(
BCA)
and
dibromoacetic
acid
(
DBA)
all
are
metabolized
to
glyoxylic
acid
through
a
reaction
catalyzed
by
glutathione­
S­
transferase­
Zeta
(
GST­
Zeta)
(
Tong
et
al.,
1998a;
Tong
et
al.,
1998b).
Glyoxylic
acid
is
in
turn
metabolized
through
a
variety
of
competing
pathways
to
form
glycine,
glycolate,

oxalate,
or
CO
2
(
Stacpoole
et
al.,
1998).
Since
DCA
is
a
potential
metabolite
of
trichloroacetic
acid
(
TCA)
(
Bull
2000;
Lash
et
al.,
2000),
this
enzyme
pathway
may
also
be
relevant
for
evaluating
the
susceptibility
of
children
to
TCA.
Even
in
the
cases
where
relevant
metabolizing
enzymes
have
been
identified,
there
is
no
information
on
age­
dependent
changes
in
the
expression
or
activity
of
these
enzymes.
It
is
noteworthy,
however,
that
Stacpoole
et
al.
(
1998)
reported
that
92
the
toxicokinetics
of
DCA
is
similar
in
children
and
adults
given
pharmacological
doses
of
this
compound,
suggesting
that
there
are
no
significant
age­
dependent
differences
in
DCA
metabolism.

The
health
implications
of
any
differences
between
children
and
adults
in
metabolic
capacity
are
also
difficult
to
determine
for
the
haloacetic
acids;
neither
the
mode
of
action
nor
toxic
forms
are
known.

2.3.1.
Monochloroacetic
Acid
Developmental/
Reproductive
Effects
Johnson
et
al.
(
1998)
studied
the
developmental
toxicity
of
MCA
by
exposing
pregnant
Sprague­
Dawley
rats
on
gestation
days
1­
22
to
0
or
1570
mg/
L
MCA
in
drinking
water
(
reported
by
the
authors
to
be
equivalent
to
doses
of
0
or
193
mg/
kg/
day).
Dams
were
sacrificed
on
gestation
day
22,
and
implantation
sites,
resorption
sites,
fetal
placements,
fetal
weights,
placental
weights,
crown­
rump
lengths,
gross
fetal
abnormalities,
and
abnormal
abdominal
organs
were
recorded.
In
addition,
the
hearts
were
evaluated
microscopically
for
abnormalities.
There
was
no
microscopic
evaluation
of
other
internal
anomalies,
and
no
evaluation
of
skeletal
anomalies.
All
dams
survived
without
evidence
of
toxicity.
Although
the
authors
reported
that
"
weight
gain
during
pregnancy
was
not
significantly
different
for
any
group
of
dams,"
the
average
maternal
weight
gain
in
the
exposed
rats
was
markedly
reduced
(
18
g/
dam,
compared
to
122.1
g/
dam
in
the
controls).
The
magnitude
of
the
decreased
weight
gain
was
confirmed
by
the
study
authors
(
Johnson,
personal
communication),
but
it
remains
unclear
why
this
difference
was
not
considered
statistically
significant.
No
adverse
effects
on
reproductive
or
developmental
endpoints
were
reported.
There
was
no
effect
on
any
other
evaluated
endpoint
of
maternal
or
fetal
toxicity.
93
Based
on
the
marked
decrease
in
weight
gain,
the
single
dose
tested
of
193
mg/
kg/
day
is
a
maternal
LOAEL
and
a
developmental
NOAEL.
There
was
no
evidence
to
suggest
any
developmental
toxicity
resulting
from
exposure
to
MCA.
This
lack
of
effect
occurred
despite
the
apparently
large
differences
in
maternal
weight
gain
during
exposure.
Complete
fetal
examinations
were
not
conducted,
and
the
study
is
limited
by
the
small
size
of
the
exposed
group.

Smith
et
al.
(
1990
in
a
published
abstract)
reported
on
the
results
of
a
developmental
toxicity
study
in
which
pregnant
Long­
Evans
rats
were
treated
with
MCA
at
doses
of
0,
17,
35,

70,
or
140
mg/
kg/
day
by
gavage
on
gestation
days
6­
15.
MCA
treatment
significantly
reduced
maternal
weight
gain
at
140
mg/
kg/
day
but
did
not
affect
maternal
deaths
or
organ
weight
at
any
doses.
No
treatment­
related
effects
on
resorptions/
litter
or
pup
weight
were
observed.
The
mean
frequency
per
litter
of
soft
tissue
malformations
ranged
from
1.2%
in
the
control
group
to
6.4%
in
the
high­
dose
group
(
140
mg/
kg/
day);
the
change
was
not
considered
to
be
dose­
related.
The
highest
dose
of
MCA
caused
a
significantly
increased
incidence
of
malformations
of
the
cardiovascular
system,
mainly
levocardia
(
primarily
a
defect
between
the
ascending
aorta
and
right
ventricle).
There
were
no
skeletal
malformations
observed
in
this
study.
Based
on
the
malformations
of
the
cardiovascular
system,
the
LOAEL
for
developmental
toxicity
was
140
mg/
kg/
day,
and
the
next
lower
dose
of
70
mg/
kg/
day
can
be
considered
a
NOAEL.
Maternal
toxicity
was
observed
at
the
LOAEL
for
developmental
toxicity,
as
evidenced
by
a
significant
reduction
in
maternal
weight
gain.
Complete
details
of
the
study
methods
and
results
were
not
published.

MCA
has
been
evaluated
in
in
vitro
systems
for
developmental
toxicity.
Hunter
et
al.

(
1996)
treated
early
somite­
staged
conceptuses
(
3­
6
somites)
from
CD­
1
mice
to
MCA
94
concentrations
ranging
from
0
to
500
µ
M.
Statistically
significant,
concentration­
dependent
increases
in
malformations
were
seen
at
sub­
lethal
doses.
MCA
has
also
been
tested
in
developmental
screening
assays
in
non­
mammalian
systems.
The
Hydra
system
is
an
assay
that
determines
the
degree
to
which
a
test
chemical
can
perturb
embryonic
development
at
maternally
subtoxic
doses.
It
has
been
purposely
designed
to
overestimate
developmental
hazard
potential
and
its
primary
utility
is
as
a
screening
tool
for
identifying
compounds
for
in
vivo
developmental
toxicity
testing
(
Fu
et
al.,
1990).
This
assay
system
compares
the
toxicity
to
adult
Hydra
and
to
artificial
"
embryos"
made
from
disassociated
and
randomly
reaggregated
terminally
differentiated
and
pluripotent
stem
cells.
Using
this
system,
Fu
et
al.
(
1990)
studied
MCA's
developmental
toxicity.
Based
on
the
ratio
of
minimal
effective
toxic
concentrations
for
adults
and
artificial
embryos,
they
reported
that
MCA
was
more
than
8
times
more
toxic
to
in
vitro
development
than
to
the
adult
component
of
the
assay.
Ji
et
al.
(
1998)
also
studied
the
teratogenic
potential
of
MCA
using
Hydra.
The
teratogenic
potential
was
estimated
by
the
ratio
of
toxicity
to
regenerative
inhibition
(
T
50/
I
50),
which
yielded
a
ratio
of
6.16.
The
study
authors
concluded
that
MCA
had
a
high
teratogenic
potential.
The
in
vitro
studies
using
mammalian
whole­
embryo
culture
and
Hydra
provide
support
for
the
developmental
toxicity
of
MCA
observed
in
vivo
(
Smith
et
al.,
1990).

No
studies
were
located
on
the
reproductive
toxicity
of
MCA.

Systemic
Effects
Effects
in
several
short­
term
MCA
toxicity
studies
included
neurotoxicity
at
high
doses
near
the
LD
50.
In
lower­
dose
studies,
increased
nasal
discharge
and
lacrimation
were
observed
in
95
rats
beginning
at
7.5
mg/
kg/
day,
but
no
adverse
effects
were
observed
in
mice
treated
at
similar
doses
(
NTP,
1992b).
Subchronic
and
chronic
studies
suggest
that
the
primary
targets
for
MCAinduced
toxicity
include
the
heart,
spleen
and
nasal
epithelium.
In
a
13­
week
oral
gavage
study
in
rats,
decreased
heart
weight
was
observed
at
30
mg/
kg/
day
and
cardiac
lesions
progressed
in
severity
with
increasing
dose.
Liver
and
kidney
toxicity
were
only
observed
at
higher
doses
(
NTP,
1992b).
In
2­
year
oral
gavage
studies,
decreased
survival
was
observed
at
15
mg/
kg/
day
in
rats
and
decreased
survival
and
nasal
and
forestomach
hyperplasia
were
observed
in
mice
at
100
mg/
kg/
day
(
NTP,
1992b).
A
more
recent
study
appears
to
confirm
these
findings
(
DeAngelo
et
al.,
1997);
this
study
identified
a
critical
effect
of
increased
spleen
weight.

Carcinogenicity
MCA
did
not
induce
a
carcinogenic
response
in
two
chronic
rodent
bioassays
(
NTP,

1992b;
DeAngelo
et
al.,
1997).
There
was
no
evidence
for
carcinogenicity
in
the
NTP
(
1992b)

study;
however,
the
route
of
compound
administration
was
via
gavage,
only
two
doses
were
tested,
and
significant
mortality
was
observed
in
high­
dose
male
rats,
high­
dose
male
mice,
and
low­
and
high­
dose
female
rats.
The
high
mortality
rates
may
have
compromised
the
power
and
sensitivity
of
the
study
to
detect
MCA­
associated
tumor
effects.

In
the
DeAngelo
et
al.
(
1997)
drinking
water
study,
only
male
F344
rats
were
tested.

Female
F344
rats
may
be
more
sensitive
to
MCA­
associated
effects,
as
evidenced
by
some
of
the
toxicity
findings
in
the
subchronic
oral
gavage
study
(
NTP,
1992b;
Bryant
et
al.,
1992).
MCA
has
not
been
tested
for
carcinogenicity
in
a
drinking
water
assay
in
a
second
species.
96
MCA
has
yielded
mixed
results
in
genotoxicity
assays.
MCA
was
negative
in
the
S.

typhimurium
reverse
mutation
assay
(
Mortelmans
et
al.,
1986)
but
positive
in
the
mouse
lymphoma
gene
mutation
assay
in
the
absence
of
metabolic
activation
(
McGregor
et
al.,
1987).
It
was
positive
for
the
induction
of
sister
chromatid
exchanges
in
Chinese
hamster
ovary
cells
in
the
absence
of
S9
(
liver
enzyme
preparation
containing
cytochrome
P450)
activation
(
but
not
in
the
presence
of
S9)
(
Galloway
et
al.,
1987),
but
was
negative
for
sister
chromatid
exchanges
in
a
Chinese
hamster
lung
fibroblast
system
(
Sawada
et
al.,
1987).
MCA
did
not
induce
strand
breaks
in
rats
or
mice
(
Chang
et
al.,
1991).
In
a
more
recent
study
including
gene
mutation
assays
in
Escherichia
coli
and
Salmonella
TA100
and
the
newt
(
Pleurodeles
waltl
larvae)
micronucleus
test,
MCA
gave
uniformly
negative
results
(
Giller
et.
al,
1997).

Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1986),

MCA
is
best
classified
as
Group
D:
Not
Classifiable
as
to
Human
Carcinogenicity.
According
to
the
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1999),
the
data
on
MCA
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.

Basis
for
RfD
and
MCLG
The
Agency
derived
the
MCA
RfD
from
the
LOAEL
of
3.5
mg/
kg/
day
for
increased
spleen
weight
in
the
chronic
rat
drinking­
water
study
(
DeAngelo
et
al.,
1997).
A
composite
UF
of
1000
was
used:
a
10­
fold
factor
for
interspecies
extrapolation,
a
10­
fold
factor
for
human
variability,
a
3­
fold
factor
for
extrapolation
from
a
minimal
LOAEL,
and
a
3­
fold
factor
for
database
deficiencies,
including
lack
of
adequate
developmental
toxicity
studies
in
two
species
and
97
the
lack
of
a
multi­
generation
reproductive
study.
The
resulting
RfD
is
0.004
mg/
kg/
day,

corresponding
to
a
DWEL
of
0.14
mg/
L
for
a
70
kg
adult
drinking
2
L
of
water/
day.

EPA
calculated
the
MCLG
by
applying
a
relative
source
contribution
(
RSC)
of
20%
to
the
DWEL:
MCLG
=
0.14
mg/
L
x
0.2
=
0.028
mg/
L
(
rounded
to
0.03
mg/
L).
The
resulting
proposed
MCLG
for
MCA
is
0.03
mg/
L.
The
default
RSC
of
20%
was
chosen
in
accordance
with
the
exposure
decision
tree
approach
in
EPA's
Human
Health
Methodology
(
EPA,
2000f),

taking
into
account
the
likelihood
of
exposure
to
MCA
from
sources
other
than
tap
water,
such
as
ambient
air
and
food.
The
available
data
are
sufficient
to
demonstrate
that
food
and
air
are
relevant
exposure
sources
in
addition
to
drinking
water,
but
the
data
are
inadequate
to
calculate
an
RSC.

Children's
Risk
in
Relation
to
the
MCLG
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.

The
only
developmental
effects
that
have
been
reported
for
MCA
are
malformations
of
the
cardiovascular
system
reported
in
a
published
abstract
(
Smith
et
al.,
1990).
The
NOAEL
was
70
mg/
kg/
day,
and
the
LOAEL
for
developmental
toxicity
was
140
mg/
kg/
day,
where
maternal
toxicity
also
occurred
(
Smith
et
al.,
1990).
In
contrast,
Johnson
et
al.
(
1998)
reported
that
the
single
dose
tested
of
193
mg/
kg/
day
induced
maternal,
but
not
developmental
effects.

Taken
together,
these
data
suggest
that
the
developing
fetus
is
no
more
sensitive
to
MCA
than
are
adults.
The
LOAEL
for
systemic
toxicity
of
3.5
mg/
kg/
day
(
DeAngelo
et
al.,
1997)
is
20x
lower
98
than
the
NOAEL
for
developmental
effects
of
70
mg/
kg/
day.
The
MCLG
of
0.03
mg/
L
for
systemic
effects
is
adequately
protective
of
fetuses
and
children.

2.3.2.
Dichloroacetic
Acid
Developmental/
Reproductive
Effects
In
two
studies
reported
by
Smith
et
al.
(
1992),
pregnant
Long­
Evans
rats
(
approximately
20/
dose)
received
DCA
by
gavage
on
gestation
days
6­
15
at
doses
of
0,
900,
1400,
1900,
or
2400
mg/
kg/
day
(
first
study)
or
0,
14,
140,
or
400
mg/
kg/
day
(
second
study).
Dams
were
sacrificed
on
gestation
day
20,
and
both
maternal
toxicity
and
fetal
toxicity
were
assessed.
Dose­
related
increases
in
mortality
occurred
in
dams
dosed
at
1400
mg/
kg/
day
and
above,
body
weight
gain
was
significantly
reduced
at
140
mg/
kg/
day
and
above,
and
increased
liver
weight
was
observed
in
all
the
treated
groups.
Significant
implantation
loss
occurred
at
900
mg/
kg/
day
and
above,
and
the
number
of
live
fetuses
per
litter
was
reduced
at
900
mg/
kg/
day
and
above.
Fetal
weight
and
crown­
rump
length
were
significantly
lower
at
levels
of
400
mg/
kg/
day
and
above.
Dose­
related
increases
were
also
reported
for
external,
soft
tissue,
cardiovascular,
urogenital,
and
orbital
malformations
in
the
developing
fetuses
at
doses
of
140
and
above.
The
authors
identified
a
developmental
NOAEL
of
14
mg/
kg/
day
and
a
LOAEL
of
140
mg/
kg/
day.
Based
on
increased
liver
weight,
the
maternal
NOAEL
and
LOAEL
were
also
14
and
140
mg/
kg/
day,
respectively.

Epstein
et
al.
(
1992)
reported
findings
from
a
series
of
experiments
in
pregnant
Long­

Evans
rats
exposed
to
DCA
by
gavage.
There
were
three
separate,
sequential
phases
of
this
study;
in
each
phase,
the
dams
were
exposed
for
a
specific
1­
to
3­
day
period
during
gestation
and
were
sacrificed
on
gestation
day
20.
Both
maternal
and
fetal
toxicity
were
assessed,
including
99
histological
examination
of
the
fetuses.
In
all
three
phases
of
the
study,
no
treatment­
related
maternal
toxicity
was
observed
(
based
on
body
weight
and
organ
weight
data).
In
the
first
phase
of
the
study,
dams
were
exposed
to
1900
mg/
kg/
day
during
gestation
days
6­
8,
9­
11,
or
12­
15,
in
order
to
observe
the
effects
of
DCA
during
specific
periods
of
organogenesis.
A
decrease
in
average
fetal
weight
was
reported
in
the
dose
group
exposed
on
days
6
to
8,
but
no
malformations
were
reported.
In
the
groups
dosed
on
days
9­
11
and
12­
15,
the
mean
percentage
of
cardiac
malformations
per
litter
was
significantly
(
p

0.001)
increased.
In
the
second
phase
of
the
study,
pregnant
dams
were
administered
a
single
dose
of
2400
mg/
kg
on
gestation
days
10,

11,
12,
or
13.
Fetal
weights
in
each
exposure
group
were
similar
to
control
values.
Significant
(
p

0.05)
increases
in
cardiac
malformations
were
reported
in
groups
exposed
on
day
10
or
day
12.
In
the
third
phase
of
the
study,
a
single
dose
of
3500
mg/
kg
was
administered
to
dams
on
gestation
days
9,
10,
11,
12,
or
13.
This
higher
dose
level,
3500
mg/
kg,
resulted
in
a
slightly
higher
incidence
of
cardiac
defects
(
3.6%
vs.
2.9%
in
controls),
and
the
increase
was
significant
(
p

0.05)
on
day
12.
The
results
from
these
studies
suggest
that
acute
high­
dose
treatments
of
DCA
at
specific
developmental
stages
can
induce
developmental
toxicity
in
the
absence
of
maternal
toxicity.

The
developmental
toxicity
of
DCA
has
also
been
evaluated
in
in
vitro
systems.
Saillenfait
et
al.
(
1995)
exposed
explanted
embryos
from
Sprague­
Dawley
rats
for
46
hours
to
0,
1.0,
2.5,

3.5,
5.0,
7.5,
or
10
mM
DCA.
A
significant,
dose­
dependent
decrease
in
crown­
rump
length
was
seen
at
3.5
mM
and
above.
In
addition,
several
effects
(
brain
and
eye
defects,
reduced
embryonic
axis)
were
seen
at
2.5
mM.
A
similar
study
with
CD­
1
mouse
whole­
embryo
culture
exposed
to
DCA
for
24
hours
found
significant
increases
in
neural­
tube
defects
at
treatment
concentrations
of
100
5.9
mM
and
above
(
Hunter
et
al.,
1996).
Another
whole­
embryo
culture
study
was
designed
to
test
mechanisms
of
haloacetic
acid­
induced
dysmorphogenesis
(
Ward
et
al.,
2000).
The
authors
reported
that
DCA
treatment
increased
the
accumulation
of
sub­
G1
events
(
a
measure
of
cells
with
less
than
the
normal
2n
complement
of
DNA
found
during
the
G1
stage
of
the
cell
cycle).

Because
the
characteristic
breakage
of
DNA
during
apoptosis
leads
to
this
accumulation
of
sub­

G1
events,
this
measure
is
often
used
as
an
indicator
of
apoptosis.
The
results
suggested
to
the
authors
that
the
developmental
toxicity
of
DCA,
particularly
the
induction
of
embryonic
neuraltube
defects,
may
be
mechanistically
associated
with
its
ability
to
increase
apoptosis.
Bantle
et
al.

(
1999)
reported
equivocal
results
for
DCA
in
the
Frog
Embryo
Teratogenesis
Assay
­­
Xenopus
(
FETAX).
Although
these
in
vitro
studies
cannot
be
used
for
dose­
response
assessment,
they
support
the
in
vivo
observations
that
DCA
can
cause
developmental
effects.

No
single­
or
two­
generation
reproductive
toxicity
studies
have
been
conducted
with
DCA.
However,
there
are
data
indicating
that
DCA
induces
male
reproductive
tract
toxicity.

Cicmanec
et
al.
(
1991)
observed
testicular
degenerative
changes
in
beagle
dogs
exposed
to
12.5
mg/
kg/
day
DCA
in
gelatin
capsules
for
90
days.
Testicular
changes
included
syncytial
giant
cell
formation
and
degeneration
of
the
testicular
germinal
epithelium.
This
dose
was
the
study
LOAEL,
based
on
the
testicular
changes,
liver
vacuolization
and
brain
histopathology.
Toth
et
al.

(
1992)
found
decreased
absolute
weight
of
the
preputial
gland
and
epididymis
at
the
lowest
dose
tested
(
31.3
mg/
kg/
day)
in
male
Long­
Evans
rats
(
18
to
19/
group)
treated
with
gavage
doses
of
DCA
for
10
weeks,
but
the
absolute
testes
weight
was
not
affected.
Relative
liver
weights
were
also
increased
in
this
dose
group.
At
the
two
higher
doses
(
62.5
and
125
mg/
kg/
day),
significant
decreases
in
the
percentage
of
motile
sperm,
epididymal
sperm
counts
and
sperm
motion
101
parameters
were
observed.
Gross
pathologic
examination
did
not
reveal
lesions
at
any
dose
in
the
testis
or
epididymal
epithelium.
Histological
evidence
of
impaired
spermiation
was
noted
in
the
62.5
and
125
mg/
kg/
day
dose
groups
and
was
attributed
to
retention
of
late­
step
spermatids
in
the
seminiferous
tubules.
Male
fertility
was
assessed
by
mating
treated
males
with
untreated
females
on
day
70
(
final
day
of
treatment);
no
effect
on
fertility
was
seen
at
any
dose.
However,
the
study
authors
indicated
that
the
experimental
protocol
may
have
been
insensitive
for
fertility
evaluation,

because
an
overnight
mating
period
was
used
and
multiple
matings
were
possible.
Based
on
reduced
weight
in
sexual
accessory
organs
(
preputial
gland
and
epididymis),
the
LOAEL
for
this
study
was
31.3
mg/
kg/
day,
and
a
NOAEL
could
not
be
determined.

Linder
et
al.
(
1997a)
examined
the
reproductive
tract
toxicity
and
spermatotoxicity
of
DCA.
Male
Sprague­
Dawley
rats
(
8/
group)
were
given
daily
gavage
doses
of
0,
18,
54,
160,

480,
or
1440
mg/
kg
DCA
for
14
days.
The
only
evidence
of
general
toxicity
was
a
statistically
significant
decrease
in
final
body
weight
at
480
and
1440
mg/
kg/
day.
A
variety
of
male
reproductive­
tract
toxicity
parameters
were
affected.
Rats
in
the
54
mg/
kg/
day
group
exhibited
clear
histopathological
effects
on
spermiation
indicative
of
spermatotoxicity,
which
increased
in
severity
with
increasing
dose.
At
doses
of
160
mg/
kg
and
higher,
the
percentage
of
fused
cauda
sperm
was
statistically
increased
and
percent
motile
sperm
were
statistically
decreased
as
compared
with
controls.
The
characteristic
pattern
of
response
included
altered
spermiation
(
including
retention
of
mature
sperm),
atypical
formation
and
resorption
of
residual
cytoplasm,

abnormal
sperm
morphology,
and
decreased
motility,
all
of
which
increased
in
magnitude
and
severity
with
dose.
Epididymis
weight
was
decreased
at
dose
levels
of
480
and
1440
mg/
kg/
day.

The
LOAEL
for
this
study
was
54
mg/
kg/
day,
and
18
mg/
kg/
day
was
an
equivocal
NOAEL.
This
102
determination
is
based
on
unequivocal
histopathological
changes
indicative
of
spermatotoxicity
in
the
male
reproductive
tract
at
the
LOAEL.
The
susceptibility
of
the
developing
male
reproductive
tract
has
not
been
tested
for
DCA.

Systemic
Effects
Stacpoole
et
al.
(
1998)
reviewed
the
clinical
data
for
DCA.
The
sodium
salt
of
DCA
has
been
employed
experimentally
for
more
than
25
years
in
clinical
medicine
and
human
research.

Over
40
patient­
years
of
cumulative
DCA
exposure
have
been
recorded
in
more
than
50
infants
and
children
with
congenital
lactic
acidosis;
several
patients
were
treated
with
daily
oral
doses
of
approximately
25
mg/
kg
for
up
to
5
years.
Typical
doses
range
from
25
to
100
mg/
kg,
and
subsequent
clinically
adverse
findings
have
been
mainly
restricted
to
mild
neurological
effects,

such
as
a
reduction
in
anxiety
or
sedative
effects
in
both
adults
and
children
that
lasted
for
several
hours.
A
doses
of
50
to
100
mg/
kg,
there
were
a
few
cases
of
reversible
peripheral
neuropathy.

No
effects
on
reproductive
organs
or
reproductive
parameters
have
been
reported,
although
the
clinical
studies
have
not
specifically
targeted
evaluation
of
reproductive
toxicity.
Limited
pharmacokinetic
data
in
children
ranging
in
age
from
18
months
to
10
years,
who
were
treated
with
DCA
for
control
of
lactic
acidosis
due
to
malaria,
showed
kinetic
parameters
after
a
single
50
mg/
kg
infusion
similar
to
those
obtained
in
healthy
or
acidotic
adults
(
Stacpoole
et
al.,
1998).

Children
chronically
treated
with
orally­
administered
DCA
for
various
congenital
forms
of
lactic
acidosis
also
exhibited
kinetic
parameters
similar
to
those
of
adults
undergoing
the
same
treatment
regime
(
Stacpoole
et
al.,
1998).
These
human
investigations,
conducted
over
more
than
25
years
of
therapeutic
use
at
pharmacologic
doses,
suggest
that
children
are
no
more
sensitive
to
the
toxic
103
effects
of
DCA
than
adults.

As
part
of
a
series
of
neurotoxicity
studies,
Moser
et
al.
(
1999)
examined
the
neurobehavioral
toxicity
in
weanling
and
young
adults
of
two
rat
strains
(
inbred
F344
and
outbred
Long­
Evans)
exposed
to
DCA
in
drinking
water.
Treatment
for
adult
and
weanling
rats
began
at
approximately
60
and
21
days
of
age,
respectively.
Behavioral
and
neurological
changes
were
evaluated
using
a
functional
observational
battery
(
FOB)
and
assessing
motor
activity.
The
FOB
consists
of
a
neurobehavioral
screening
battery
comprised
of
a
series
of
observational
and
manipulative
tests
designed
to
assess
neurological
integrity.
Male
Long­
Evans
(
LE)
and
F344
weanling
rats
received
0,
200,
1000,
or
2000
mg/
L
in
drinking
water
(
10/
dose/
strain),
and
were
tested
at
weeks
3,
6,
9,
and
13.
Equivalent
daily
doses
were
0,
17,
88,
and
192
mg/
kg/
day
for
LE
rats
and
0,
16,
89,
and
173
mg/
kg/
day
for
F344
rats.
DCA
toxicity
was
mainly
limited
to
the
neuromuscular
system
and
to
endpoints
demonstrating
altered
neuromuscular
function;
these
included
gait
abnormalities
and
changes
in
grip
strength,
landing
foot
splay
and
righting
reaction.

The
LOAEL
was
16
mg/
kg/
day
for
weanling
F344
rats
and
17
mg/
kg/
day
for
weanling
LE
rats
after
13
weeks
of
exposure.
No
NOAELs
were
identified.
Although
the
LOAELs
for
the
two
strains
were
similar,
the
magnitude
and
severity
of
neurobehavioral
effects
were
more
pronounced
in
F344
rats.
Thus,
F344
weanling
rats
appear
to
be
more
sensitive
to
DCA­
induced
neurotoxicity
than
LE
weanling
rats.
In
the
segment
of
the
study
using
adult
rats,
the
same
experimental
design
and
protocol
was
used.
Male
LE
and
F344
adult
rats
(
10/
dose/
strain)

received
0,
250,
1250,
or
2500
mg/
L
in
drinking
water
for
8
weeks,
and
were
tested
at
weeks
2,

5,
8,
and
10
(
2
weeks
after
the
termination
of
exposure).
Equivalent
daily
doses
were
0,
23,
122,

and
220
mg/
kg/
day
for
LE
rats
and
0,
18,
91,
and
167
mg/
kg/
day
for
F344
rats.
Based
on
gait
104
impairment,
the
LOAEL
for
adult
F344
rats
was
18
mg/
kg/
day,
and
there
was
no
NOAEL.
For
adult
LE
rats,
the
LOAEL
was
122
mg/
kg/
day
and
the
NOAEL
was
23
mg/
kg/
day.
Adverse
effects
in
weanling
rats
were
considered
by
the
authors
to
be
more
severe
than
those
in
adults
similarly
tested
and
evaluated.
The
study
authors
concluded
that
F344
rats
are
more
sensitive
than
LE
rats.
They
also
noted
that
weanlings
may
be
somewhat
more
sensitive
than
adults
to
the
neurobehavioral
toxicity
of
DCA,
although
this
difference
was
seen
only
in
the
LE
rats.

The
systemic,
noncancer
effects
of
DCA
in
animals
and
humans
can
be
grouped
into
four
categories:
metabolic
alterations,
hepatic
toxicity,
reproductive/
developmental
toxicity,
and
neurotoxicity.
The
liver
is
a
major
target
organ
of
toxicity
in
DCA­
treated
mice,
rats,
dogs
and
humans
(
Cicmanec
et
al.,
1991;
DeAngelo
et
al.,
1999;
Katz
et
al.,
1981;
Parrish
et
al.,
1996;

Stacpoole
et
al.,
1998).
Frank,
dose­
dependent
toxicity
in
the
form
of
hepatic
vacuolation,

cytomegaly,
karyomegaly,
and
multi­
focal
coagulative
necrosis
has
only
been
observed
in
mice
at
drinking
water
doses
of
500
mg/
L
(
84
mg/
kg/
day)
and
higher
(
DeAngelo
et
al.,
1999).
Increased
testes
weights
have
been
observed
in
rats
dosed
with
40.2
mg/
kg/
day
DCA
in
drinking
water
for
a
lifetime
(
DeAngelo
et
al.,
1996),
although
there
was
no
clear
dose­
response.
In
a
subchronic
study
in
rats,
Katz
et
al.
(
1981)
reported
a
LOAEL
of
125
mg/
kg/
day
for
brain
lesions
(
vacuolization
of
the
myelinated
white
tracts);
no
NOAEL
was
identified.
In
a
90­
day
study
in
beagle
dogs
administered
DCA
by
capsule,
Katz
et
al.
(
1981)
found
a
LOAEL
of
50
mg/
kg/
day,

based
on
prostate
gland
atrophy
and
testicular
changes
(
degeneration
of
germinal
epithelium,

vacuolation
of
Leydig
cells,
formation
of
syncytial
giant
cells).
Other
reported
effects
at
50
mg/
kg/
day
and
higher
included
decreased
body
weight,
changes
in
hematologic
and
clinical
105
chemistry
parameters,
neurotoxicity,
and
ocular
effects
(
irreversible
lenticular
opacities).
No
NOAEL
could
be
determined
from
this
study.

In
a
subchronic
study
by
Cicmanec
et
al.
(
1991),
four­
month
old
male
and
female
beagle
dogs
(
5/
sex/
dose)
were
administered
0,
12.5,
39.5,
or
72
mg/
kg/
day
of
dichloroacetate
in
gelatin
capsules
for
90
days.
Histopathological
testicular
changes
were
reported
in
the
testes
of
males
at
all
dose
levels
(
except
for
controls),
and
included
syncytial
giant
cell
formation
and
degeneration
of
the
testicular
germinal
epithelium.
Prostate
glandular
atrophy
characterized
by
a
significant
reduction
of
glandular
alveoli
was
also
noted
in
mid­
and
high­
dose
groups.
Other
histopathology
included
hepatic
vacuolization
and
vacuolization
of
the
myelinated
white
tracts
of
the
cerebrum
and
cerebellum
at
all
dose
levels.
The
LOAEL
for
this
study
was
12.5
mg/
kg/
day,
based
on
testicular
degenerative
changes,
liver
vacuolization
and
brain
histopathology;
a
NOAEL
could
not
be
determined.

Carcinogenicity
No
epidemiological
investigations
of
the
carcinogenicity
of
DCA
in
humans
have
been
performed.
In
animals,
there
have
been
a
number
of
independent
long­
term
studies
investigating
aspects
of
the
carcinogenicity
of
DCA.
Statistically
significant
increases
in
hepatic
carcinomas
alone
and/
or
hepatic
carcinomas
plus
adenomas
were
seen
in
the
following
studies:
(
1)
all
seven
male
B6C3F1
mouse
studies
(
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);
(
2)
two
female
B6C3F1
mouse
studies
(
Pereira
and
Phelps,
1996;
Pereira,
1996),
with
the
third
female
B6C3F1
mouse
study
reporting
hyperplastic
nodules
in
the
livers
of
some
treated
106
animals,
but
no
liver
tumors
(
Bull
et
al.,
1990);
and
(
3)
male
F344
rat
studies
(
DeAngelo
et
al.,

1996;
Richmond
et
al.,
1995).
The
weight
of
evidence
clearly
indicates
that
DCA
is
hepatocarcinogenic
in
both
male
mice
and
male
rats
at
high
concentrations
of
DCA
given
in
water.
Exposure
levels
causing
increased
incidence
of
animals
with
tumors
range
from
500
to
5000
mg/
L.
However,
concentrations
as
low
as
50
mg/
L
(
8
mg/
kg/
day)
have
been
reported
to
increase
the
multiplicity
of
tumors
in
male
mice
(
DeAngelo
et
al.,
1999).
No
studies
have
been
conducted
to
investigate
the
transplacental
carcinogenicity
of
DCA
or
carcinogenicity
in
young
animals
exposed
to
DCA
from
birth.

Available
data
are
not
adequate
to
identify
any
single
mode
of
action
as
being
the
only
or
most
important
pathway
leading
to
cancer.
DCA
has
been
found
to
be
mutagenic
and
clastogenic,

but
these
responses
generally
occur
at
high
dose
levels.
There
are
differences
of
opinion
regarding
genotoxicity
of
DCA
at
lower
dose
levels.
IARC
(
1995)
and
ILSI
(
1997)
concluded
that
it
was
not
genotoxic.
DCA
was
classified
as
a
weak
mutagen
by
Moore
and
Brock
(
2000)

and
as
a
direct
acting
genotoxic
agent
by
the
National
Center
for
Environmental
Assessment
at
EPA
(
EPA,
1998e)
based
on
recently
published
studies
(
DeMarini
et
al.,
1994;

Fuscoe
et
al.,
1996;
Leavitt
et
al.,
1997;
Harrington­
Brock
et
al.,
1998).
Possible
nongenotoxic
modes
of
action
may
contribute
to
the
carcinogenic
response
but
have
not
been
conclusively
implicated
in
rodent
liver
tumorigenesis.
These
MOAs
include
cytotoxicity
and
reparative
cell
proliferation
and/
or
decreased
apoptosis,
DNA
hypomethylation,
metabolic
alterations
involving
glycogen
accumulation
in
hepatocytes,
and/
or
selective
growth
of
subpopulations
of
mutated
or
immunoreactive
cells.
107
The
EPA
reviewed
DCA
carcinogenicity
in
1994
(
EPA,
1994g),
and
this
evaluation
was
updated
in
1996
(
EPA,
1996d).
These
reviews
classified
DCA
as
a
Group
B2:
Probable
Human
Carcinogen,
in
accordance
with
the
1986
EPA
Guidelines
for
Carcinogen
Risk
Assessment.
In
1995,
IARC
concluded
that
"
DCA
is
not
classifiable
as
to
its
carcinogenicity
to
humans,"
and
placed
DCA
in
the
IARC
Group
3
category
(
IARC,
1995).
It
should
be
noted
that
at
the
time
of
the
IARC
evaluation,
there
were
no
data
demonstrating
high­
dose
DCA­
induced
hepatocarcinogenicity
in
the
rat.
In
2002,
IARC
(
2002)
changed
the
cancer
classification
to
Group
2B
(
possibly
carcinogenic
to
humans).
Under
the
1999
Carcinogen
Risk
Assessment
Guidelines
(
EPA,
1999),
DCA
is
likely
to
be
carcinogenic
in
humans,
based
on
the
weight
of
the
evidence
in
animal
bioassays.
In
particular,
the
following
characteristics
were
considered
and
found
to
increase
the
overall
weight
of
evidence
for
this
classification:
the
number
of
independent
studies
reporting
consistently
positive
results
and
at
roughly
comparable
doses,
site
concordance
for
tumor
formation
between
two
species,
consistent
observations
in
different
species
and
sexes,

and
clear
evidence
of
a
dose­
response
relationship.
There
is
no
clear
mechanistic
understanding
of
the
carcinogenic
process
and
the
shape
of
the
cancer
dose­
response
curve
at
low
doses.
This
precludes
any
consideration
of
classifying
the
tumorigenic
dose
response
for
DCA
as
nonlinear.

The
cancer
risk
from
ingestion
of
DCA
was
quantified
based
on
a
dose­
response
study
in
male
mice
(
DeAngelo
et
al.,
1999).
The
cumulative
incidence
of
hepatic
total
tumor
incidence
(
carcinoma
plus
adenoma)
in
the
test
animals
was
well­
described
by
several
different
dichotomous
models,
with
the
multistage
model
yielding
the
best
fit.
Based
on
this
model,
the
BMD
10
was
6.86
mg/
kg/
day,
and
the
BMDL
10
was
2.1
mg/
kg/
day.
In
accord
with
the
draft
guidelines
(
U.
S.
EPA,

1999),
the
BMDL
10
was
used
as
the
point­
of­
departure
(
POD)
for
quantifying
cancer
risk.
108
Because
the
mode
of
action
by
which
DCA
increases
cancer
risk
is
not
understood,
extrapolation
to
low
dose
was
performed
by
assuming
a
no­
threshold
linear
dose­
response
curve
between
the
origin
and
the
POD.
This
yields
a
cancer
slope
factor
of
0.048
per
mg/
kg/
day
and
a
unit
risk
of
1.4
x
10­
6
per
(
µ
g/
L).
Based
on
this
calculation,
drinking
water
concentrations
of
approximately
70
µ
g/
L,
7
µ
g/
L
and
0.7
µ
g/
L
are
estimated
to
be
associated
with
lifetime
cancer
risks
of
10­
4,
10­
5
and
10­
6
,
respectively.

Basis
for
RfD
and
MCLG
EPA
based
its
RfD
on
the
LOAEL
of
12.5
mg/
kg/
day
identified
for
testicular
degenerative
changes,
liver
vacuolization
and
brain
histopathology
in
the
Cicmanec
et
al.
(
1991)
study.
EPA
used
an
UF
of
3000,
for
use
of
a
LOAEL,
inter­
and
intraspecies
variability,
and
use
of
a
study
with
a
less­
than­
lifetime
duration
(
EPA,
2001c).
An
RfD
of
0.004
mg/
kg/
day
results,

corresponding
to
a
DWEL
of
0.14
mg/
L
(
rounded
to
0.1
mg/
L)
for
a
70
kg
adult
drinking
2
L
of
water
per
day.

The
MCLG
for
DCA
is
zero,
based
on
its
probable
carcinogenicity
and
the
use
of
linear
low­
dose
extrapolation.

Children's
Risk
in
Relation
to
the
MCLG
The
Agency
believes
that
the
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
are
more
sensitive
to
the
toxic
effects
of
DCA
than
are
adults.
Human
investigations,
conducted
over
more
than
25
years
of
therapeutic
use
at
109
pharmacologic
doses,
suggest
that
DCA
toxicokinetics
do
not
differ
between
children
and
adults,

and
that
children
are
not
likely
to
be
more
vulnerable
to
the
potentially
toxic
effects
of
DCA
(
Stacpoole
et
al.,
1998).
However,
these
data
are
limited
and
no
definitive
conclusions
can
be
made.

No
multi­
generation
reproductive
toxicity
animal
studies
have
been
conducted
to
assess
age­
related
differences
in
sensitivity
to
DCA.
However,
in
female
rats,
DCA
exposure
during
gestation
resulted
in
the
impairment
of
fetal
maturation
and
soft
tissue
anomalies
(
primarily
of
cardiac
origin)
indicating
that
the
developing
fetus
is
susceptible
to
DCA­
induced
toxicity.
Data
collected
by
Moser
et
al.
(
1999)
provide
limited
evidence
for
increased
susceptibility
of
rats
to
DCA­
induced
neurotoxicity
when
exposures
begin
shortly
after
weaning.

Children
with
certain
inborn
errors
of
metabolism
may
be
at
increased
risk
for
DCAinduced
toxicity.
Glycogen­
storage
disease
is
an
inherited
deficiency
or
alteration
in
any
one
of
the
enzymes
involved
in
glycogen
degradation.
The
liver
is
a
major
target
organ
of
toxicity
in
DCA­
treated
mice,
rats,
dogs,
and
humans,
and
dose­
related
increases
in
liver
size
have
been
reported
to
be
accompanied
by
an
increase
in
glycogen
deposition
in
the
liver
in
mice
and
rats
(
Kato­
Weinstein
et
al.,
1998).
Although
the
enzymatic
basis
and
functional
significance
of
this
finding
is
unclear,
glycogen
accumulation
may
be
associated
with
inhibition
of
glycogenolysis,
as
the
reported
effects
resemble
those
observed
in
glycogen­
storage
disease
VI
(
EPA,
2001c).
If
glycogen
accumulation
plays
a
role
in
the
DCA
tumorigenic
process,
children
with
this
disease
may
represent
a
group
that
is
more
sensitive
to
DCA
toxicity.
Hereditary
tyrosinemia
(
a
human
disease
involving
a
deficit
in
tyrosine
metabolism)
is
often
associated
with
the
development
of
hepatocellular
carcinoma
in
young
patients
(
Tanguay
et
al.,
1996;
LaBerge,
1986).
DCA
has
110
been
reported
to
significantly
alter
tyrosine
metabolism
as
a
consequence
of
its
inhibitory
effect
on
glutathione
S­
transferase
Zeta
(
Cornett
et
al.,
1999).
DCA­
induced
inhibition
of
tyrosine
metabolism
may
result
in
increased
levels
of
reactive
tyrosine
metabolites
that
may
adversely
affect
the
liver
and
nervous
system.
Children
with
hereditary
tyrosinemia
may
be
at
increased
risk
for
DCA­
induced
toxicity.
Oxalate
is
an
excretory
endpoint
of
DCA
metabolism.
Thus,
children
affected
with
the
metabolic
disorder,
hyperoxaluria,
may
also
be
at
an
increased
risk
to
DCA
exposure.
Adults
with
hereditary
tyrosinemia,
glycogen
storage
disease,
or
hyperoxaluria
may
also
be
at
increased
risk
for
DCA­
induced
toxicity.

2.3.3.
Trichloroacetic
Acid
Developmental/
Reproductive
Effects
Johnson
et
al.
(
1998)
studied
the
teratogenicity
of
TCA
by
exposing
pregnant
Sprague­

Dawley
rats
to
0
or
2,730
mg/
L
TCA
in
neutralized
drinking
water
on
gestation
days
1­
22.
The
authors
estimated
the
doses
to
be
equivalent
to
0
or
291
mg/
kg/
day.
Dams
were
sacrificed
on
gestation
day
22;
implantation
sites,
resorption
sites,
fetal
placements,
fetal
weights,
placental
weights,
crown­
rump
lengths,
gross
fetal
abnormalities,
and
abnormal
abdominal
organs
were
recorded.
In
addition,
the
fetal
hearts
were
removed,
dissected,
and
examined
for
abnormalities
under
microscope.
There
was
no
microscopic
evaluation
of
other
internal
anomalies,
and
no
evaluation
of
skeletal
anomalies.
Although
the
authors
reported
no
signs
of
maternal
toxicity
and
no
effect
on
maternal
weight
gain,
the
average
maternal
weight
gain
for
TCA­
exposed
animals
was
84.6
g
as
compared
with
122
g
for
control
animals,
representing
a
30%
decrease
in
maternal
body
weight
gain.
This
decrease
is
considered
toxicologically
significant.
Significant
increases
111
were
observed
in
average
resorption
sites
and
average
implantation
sites.
No
significant
differences
were
found
in
the
numbers
of
live
or
dead
fetuses,
fetal
weight,
placental
weight,

crown­
rump
length,
external
morphology,
or
gross
external
or
noncardiac
internal
congenital
abnormalities.
Cardiac
abnormalities
were
evident
in
10.5%
of
the
fetuses
in
the
TCA
group,

compared
to
2.15%
of
the
controls.
No
single
cardiac
defect
or
group
of
defects
predominated
in
the
TCA
group.
Based
on
the
increases
in
implantation
and
resorption
sites,
and
cardiac
malformations
in
the
single
TCA­
treated
group,
291
mg/
kg/
day
is
considered
a
developmental
LOAEL.
Based
on
decreased
maternal
weight
gain,
the
single
dose
of
291
mg/
kg/
day
is
also
a
maternal
LOAEL.

In
a
study
by
Smith
et
al.
(
1989),
pregnant
Long­
Evans
rats
(
20
animals/
dose)
received
TCA
at
doses
of
0,
330,
800,
1200,
or
1800
mg/
kg/
day
in
drinking
water
during
gestation
days
6­

15.
Maternal
spleen
and
kidney
weights
were
increased
significantly
in
all
dose
groups
in
a
dosedependent
manner
(
p=
0.0001);
liver
weights
of
dams
were
not
affected
by
TCA
treatment.

Postimplantation
loss
increased
at
doses
of
330
mg/
kg/
day
and
higher.
Fetal
body
weight
and
crown­
rump
length
were
significantly
(
p<
0.05)
lower
than
controls
for
all
dose
groups.

Softtissue
malformations
in
the
cardiovascular
system
were
increased
for
all
treatment
groups
in
a
dose­
dependent
manner.
Levocardia
occurred
in
0%,
32%,
71%,
71%,
and
88%
of
the
litters
in
the
0,
330,
800,
1200,
and
1800
mg/
kg/
day
groups,
respectively.
The
lowest
dose,
330
mg/
kg/
day,
was
considered
the
LOAEL
in
this
study,
based
on
the
dose­
dependent
maternal
effects
(
increased
kidney
and
spleen
weights)
and
developmental
effects
(
decreased
fetal
weight
and
crown­
rump
length
and
increased
incidences
of
levocardia
in
litters).
112
TCA
has
also
been
tested
in
a
number
of
alternative
screening
models
for
assessing
potential
developmental
toxicity.
Saillenfait
et
al.
(
1995)
exposed
explanted
embryos
from
Sprague­
Dawley
rats
on
gestational
day
10
for
46
hours
to
TCA
concentrations
up
to
6.0
mM.

TCA
induced
statistically
significant,
concentration­
related
decreases
in
the
growth
and
development
parameters
of
conceptuses.
Hunter
et
al.
(
1996)
conducted
a
24­
hour
exposure
of
CD­
1
mice
embryos
to
TCA
concentrations
up
to
5000
µ
M.
TCA
produced
abnormal
embryonic
development
at
concentrations
that
were
not
embryolethal.
In
addition,
Fort
et
al.
(
1993)

reported
that
TCA
induced
malformations
at
doses
lower
than
those
that
induced
lethality
in
the
FETAX
assay.

Conflicting
results
were
reported
by
Fu
et
al.
(
1990),
who
studied
the
developmental
toxicity
potential
of
TCA
using
a
regeneration
assay
from
reaggregated
Hydra
cells.
Based
on
similarity
in
the
minimal
effective
toxic
concentration
for
adults
and
artificial
embryos,
TCA
was
not
considered
to
interfere
with
development.
As
previously
noted,
the
Hydra
system
is
designed
to
overestimate
developmental
hazard
potential
and
is
considered
to
be
more
sensitive
to
developmental
toxicity
than
most
in
vitro
mammalian
test
systems;

Overall,
the
in
vivo
studies
indicate
that
TCA
can
cause
developmental
toxicity
at
maternally
toxic
doses.
The
in
vitro
studies
cannot
be
used
quantitatively,
but
the
results
of
Saillenfait
et
al.
(
1995)
and
Hunter
et
al.
(
1996)
support
the
findings
of
the
in
vivo
studies.
In
contrast,
the
findings
of
Fu
et
al.
(
1990)
suggest
that
TCA
would
not
be
considered
a
priority
compound
for
further
developmental
toxicity
testing
in
vivo.

One
in
vitro
study
suggested
that
TCA
might
decrease
fertilization.
The
effect
of
TCA
on
in
vitro
fertilization
was
examined
in
hybrid
C57BL6
x
DBA/
2
(
B6D2F1
)
mice
(
Cosby
and
113
Dukelow,
1992).
TCA
was
constituted
in
culture
medium
to
yield
concentrations
of
100,
250,
or
1000
ppm
(
mg/
L),
and
incubated
with
mouse
oocytes
and
sperm
for
24
hours.
The
percent
of
oocytes
fertilized
was
significantly
decreased
at
250
mg/
L
(
p<
0.025)
and
1000
mg/
L
(
p<
0.1)

compared
to
controls.

Systemic
Effects
Toxicity
studies
of
orally­
administered
TCA
have
primarily
identified
the
liver
and
kidney
as
target
organs.
Hepatic
peroxisome
proliferation
has
been
examined,
with
mice
reported
to
be
more
sensitive
to
this
effect
than
rats.
Parrish
et
al.
(
1996)
reported
that
TCA
induced
peroxisome
proliferation
in
B6C3F1
mice
exposed
for
10
weeks
to
doses
as
low
as
25
mg/
kg/
day.
Increased
liver
weight
and
significant
increases
in
hepatic
labeling
(
a
measure
of
cell
division)
have
been
observed
in
short­
term
studies
in
mice
at
doses
as
low
as
100
mg/
kg/
day
(
Dees
and
Travis,
1994).
In
a
90­
day
study
by
Mather
et
al.
(
1990)
in
Sprague­
Dawley
rats,
the
NOAEL
was
36.5
mg/
kg/
day,
and
the
LOAEL
was
355
mg/
kg/
day
for
reduced
spleen
weight,

increased
relative
kidney
weight,
and
increased
liver
weight
that
was
accompanied
by
increased
hepatic
peroxisomal
beta­
oxidation
activity.
Bull
et
al.
(
1990)
treated
groups
of
mice
with
TCA
in
their
drinking
water
at
0
or
1000
mg/
L
for
52
weeks,
at
2000
mg/
L
for
37
weeks
with
a
15­

week
recovery
period,
or
at
2000
mg/
L
for
52
weeks.
Small
increases
in
liver
size,
some
accumulation
of
lipofuscin
and
focal
necrosis
were
seen
in
all
groups.
The
LOAEL
for
hepatic
lesions
from
this
study
was
164
mg/
kg/
day.
In
rats
exposed
to
TCA
for
up
to
104
weeks
(
DeAngelo
et
al.,
1997),
the
NOAEL
for
liver
toxicity
was
32.5
mg/
kg/
day
and
the
LOAEL
was
114
364
mg/
kg/
day.
Liver
effects
included
increased
serum
levels
of
liver
enzymes
(
indicating
leakage
from
cells)
and
histopathological
evidence
of
necrosis.

Carcinogenicity
TCA
induces
liver
tumors
in
mice
but
not
in
rats.
Pereira
(
1996)
observed
an
increased
incidence
of
hepatic
adenomas
and
carcinomas
in
female
B6C3F1
mice
at
doses
of
262
mg/
kg/
day
and
higher
in
drinking
water
for
82
weeks.
In
contrast,
no
increase
in
neoplastic
liver
lesions
were
found
in
F344
rats
given
doses
up
to
364
mg/
kg/
day
for
104
weeks
(
DeAngelo
et
al.,

1997).
An
earlier
study
by
DeAngelo
and
colleagues
(
1991)
in
B6C3F1
mice
found
hyperplastic
nodules
and
hepatocellular
tumors,
both
adenomas
and
carcinomas,
primarily
in
males.
The
authors
noted
that
the
female
mice
appear
to
be
less
sensitive
than
the
male
mice
to
the
carcinogenic
potential
of
TCA.
Bull
et
al.
(
1990)
found
that
exposure
to
TCA
via
drinking
water
resulted
in
induction
of
liver
tumors
in
male
B6C3F1
mice;
female
mice,
however,
did
not
show
these
effects
after
52
weeks
of
exposure.

The
mechanism
for
TCA­
induced
mouse
hepatocarcinogenesis
has
not
been
conclusively
determined.
Mutagenicity
data
have
provided
mixed
results
(
DeMarini
et
al.,
1994;
Giller
et
al.,

1997;
Harrington­
Brock
et
al.,
1998;
reviewed
in
EPA,
1994g).
Evidence
for
DNA
stand
breaks
and
clastogenicity
is
also
mixed
(
Nelson
and
Bull,
1988;
Chang
et
al.,
1991).
A
recent
study
found
that
chromosome
damage
is
not
induced
by
TCA
in
the
absence
of
pH
changes
(
Mackay
et
al.,
1995),
but
Harrington­
Brock
et
al.
(
1998)
found
weakly
positive
evidence
of
TCA
clastogenicity
(
small
colonies)
in
mouse
lymphoma
cells
in
the
absence
of
pH
changes.
115
A
variety
of
recent
mechanistic
studies
have
observed
that
TCA
induced
or
promoted
liver
tumors
in
mice
(
Ferreira­
Gonzalez
et
al.,
1995;
Latendresse
and
Pereira,
1997;
Pereira
and
Phelps,
1996;
Stauber
and
Bull,
1997;
Tao
et
al.,
1996,
1998).
In
a
recent
review,
Moore
and
Harrington­
Brock
(
2000)
evaluated
the
weight­
of­
evidence
for
the
genotoxicity
of
trichloroethylene
and
its
metabolites,
including
TCA.
The
authors
concluded
that
it
is
unlikely
that
TCA
contributes
to
tumor
formation
through
a
mutational
mechanism.
Moreover,
laboratory
mice
have
a
high
background
rate
of
spontaneous
liver
tumor
formation,
and
chemical
compounds
are
thought
to
increase
this
incidence
via
a
nongenotoxic
mode
of
action
(
Bull,
2000).
A
variety
of
recent
mechanistic
studies
have
observed
that
TCA
induced
or
promoted
liver
tumors
in
mice
(
Ferreira­
Gonzalez
et
al.,
1995;
Latendresse
and
Pereira,
1997;
Pereira
and
Phelps,
1996;
Stauber
and
Bull,
1997;
Tao
et
al.,
1996,
1998).
However,
laboratory
mice
have
a
high
background
rate
of
spontaneous
liver
tumor
formation,
and
chemical
compounds
that
increase
this
incidence
are
thought
to
exert
their
effects
via
nongenotoxic
mode(
s)
of
action
(
Bull,
2000).

A
variety
of
other
mechanisms
have
been
suggested
as
contributing
to
TCA­
induced
liver
tumorigenesis.
Of
these,
peroxisome
proliferation
and
altered
regulation
of
cell
growth
have
been
most
well
supported.
There
is
little
evidence
for
a
role
of
oxidative
DNA
damage
(
Parrish
et
al.,
1996),
or
regenerative
hyperplasia
(
Pereira,
1996;
DeAngelo
et
al.,
1997;
Bull,
2000).

Peroxisome
proliferation
is
activated
in
both
mice
and
rats,
but
liver
tumors
are
only
induced
in
mice
(
EPA,
1994;
Pereira,
1996;
DeAngelo
et
al.,
1997;
Bull,
2000).
The
lack
of
tumorigenicity
in
F344
rats
(
DeAngelo
et
al.,
1997)
might
reflect
a
lower
affinity
of
the
peroxisome
proliferation
pathway
for
TCA,
which
would
result
in
a
smaller
peroxisome
response
in
rats
as
compared
to
mice.
Humans
have
been
reported
to
have
a
much
lower
response
to
exposure
to
peroxisomal
116
proliferators,
including
TCA,
than
either
mice
or
rats
(
Lapinskas
and
Corton,
1999;
Bentley
et
al.,

1993;
Walgren
et
al.
2000).
However,
it
is
not
yet
clear
whether
peroxisome
proliferation
is
a
key
event
in
the
development
of
TCA­
induced
mouse
hepatocarcinogenesis.
A
better
case
can
be
made
for
altered
proliferation
in
subpopulations
of
cells
having
selective
growth
advantages
(
Stauber
et
al.,
1998),
arising
for
example,
due
to
spontaneous
mutations
(
Ferreira­
Gonzalez
et
al.,
1995).

Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment,
TCA
is
best
classified
as
Group
C:
Possible
Human
Carcinogen.
In
accordance
with
EPA's
Draft
1999
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.
These
conclusions
are
based
on
the
lack
of
genotoxicity
of
TCA
in
numerous
studies,
the
induction
of
only
liver
tumors
in
one
rodent
species
(
mouse),
the
uncertainty
regarding
the
likely
mode(
s)
of
action
of
TCA­
induced
hepatocarcinogenicity,
and
the
questionable
human
relevance
of
the
finding
of
increased
liver
tumors
in
a
rodent
species
with
a
high
background
rate
of
spontaneously­
occurring
liver
tumors.
Based
on
this
same
line
of
reasoning,
the
data
are
insufficient
to
conduct
a
dose­
response
quantification
for
cancer.

Basis
for
the
RfD
and
MCLG
EPA
based
the
RfD
on
the
NOAEL
of
32.5
mg/
kg/
day
for
liver
effects
in
the
DeAngelo
et
al.
(
1997)
study.
A
composite
UF
of
1000
was
applied
to
the
NOAEL,
based
on
a
10­
fold
factor
for
extrapolation
from
an
animal
study
to
humans,
a
10­
fold
factor
to
account
for
variation
in
sensitivity
among
members
of
the
human
population,
and
a
10­
fold
factor
for
data
base
117
insufficiencies,
including
lack
of
adequate
developmental
toxicity
studies
in
two
species,
lack
of
a
multi­
generation
reproductive
study,
and
lack
of
full
histopathological
data
in
a
second
species.

The
resulting
RfD
is
0.03
mg/
kg/
day,
corresponding
to
a
DWEL
of
1.05
mg/
L
for
a
70
kg
adult
drinking
2
L
of
water
per
day.

Assuming
a
drinking
water
RSC
of
20%
of
total
exposure,
the
MCLG
=
1.05
mg/
L
x
0.2
/

10
to
account
for
possible
carcinogenicity
=
0.021
mg/
L,
rounded
to
0.02
mg/
L.
The
default
RSC
of
20%
was
chosen
in
accordance
with
the
exposure
decision
tree
approach
in
EPA's
Human
Health
Methodology
(
EPA,
2000f),
taking
into
account
the
likelihood
of
exposure
to
TCA
from
sources
other
than
tap
water,
such
as
ambient
air
and
food.
The
available
data
are
sufficient
to
demonstrate
that
food
and
air
are
relevant
exposure
sources
in
addition
to
drinking
water,
but
the
data
are
inadequate
to
calculate
an
RSC.

Children's
Risk
in
Relation
to
the
MCLG
The
MCLG
is
expected
to
be
protective
of
fetuses
and
children.
Developmental
and
systemic
toxicity
appear
to
occur
at
similar
doses
(
i.
e.
similar
LOAELs),
although
conclusions
are
limited
by
the
lack
of
developmental
NOAELs.
For
example,
LOAELs
for
developmental
toxicity
have
been
reported
as
291
mg/
kg/
day
(
Johnson
et
al,
1998)
and
330
mg/
kg/
day
(
Smith
et
al.,

1989).
The
developmental
toxicity
LOAELs
are
comparable
to
the
LOAEL
of
364
mg/
kg/
day
for
systemic
toxicity
in
the
study
used
to
derive
the
RfD
(
DeAngelo
et
al.,
1997).
The
MCLG
based
on
systemic
effects
is
also
likely
to
protect
children
because
remaining
uncertainties
regarding
the
NOAEL
for
developmental
toxicity
are
taken
into
account
through
the
application
of
UFs.
In
118
addition,
there
is
no
evidence
from
studies
on
the
systemic
effects
of
TCA
that
children
are
more
sensitive
to
the
toxic
effects
of
TCA
than
are
adults.

2.3.4.
Monobromoacetic
Acid
Developmental/
Reproductive
Effects
No
peer­
reviewed
developmental
toxicity
studies
of
MBA
are
available.
In
a
published
abstract,
Randall
et
al.
(
1991)
reported
on
the
reproductive
and
developmental
toxicity
of
MBA.

Pregnant
Long­
Evans
rats
were
given
oral
gavage
doses
of
0,
25,
50,
or
100
mg/
kg/
day
MBA
in
distilled
water
on
gestation
days
6­
15.
In
the
high­
dose
group,
maternal
weight
gain
was
reduced
and
one
dam
died.
No
effects
on
reproduction
were
observed.
Several
developmental
effects
were
noted
in
the
high­
dose
group,
including
decreased
size
of
live
fetuses
(
the
affected
measure
of
size
was
not
provided
in
the
study
summary)
and
increased
incidence
of
soft
tissue
malformations,
most
of
which
were
cardiovascular
and
craniofacial.
Based
on
the
limited
data
provided
in
the
abstract,
the
NOAEL
for
this
study
was
50
mg/
kg/
day
and
the
LOAEL
for
both
maternal
and
developmental
effects
was
100
mg/
kg/
day.

The
potential
developmental
toxicity
of
MBA
has
been
evaluated
in
whole
embryo
culture,

and
malformations
were
increased
at
sub­
lethal
doses
(
Hunter
et
al.,
1996).
These
whole­
embryo
testing
data
provide
support
for
the
developmental
toxicity
of
the
brominated
acetic
acids
observed
in
vivo.

Linder
et
al.
(
1994a)
reported
the
results
of
acute
toxicity
and
acute
spermatotoxicity
studies
of
MBA.
In
the
spermatotoxicity
study,
groups
of
eight
male
Sprague­
Dawley
rats
were
given
single
doses
of
0
or
100
mg/
kg
MBA
in
a
volume
of
5
mL/
kg
and
were
sacrificed
2
or
14
119
days
after
dosing.
The
selected
single
dose
of
100
mg/
kg
was
an
approximate
LD
01,
and
was
chosen
to
provide
a
relatively
high
dose
with
a
minimal
likelihood
of
mortality.
Measures
of
male
reproductive
toxicity
included
reproductive
organ
weights,
sperm
counts,
sperm
morphology,

sperm
motility,
and
histopathological
examination
of
the
seminiferous
tubules.
No
adverse
effects
were
observed
in
the
single­
dose
study;
therefore,
a
repeated­
dosing
protocol
experiment
was
also
conducted.
Groups
of
eight
rats
were
given
daily
doses
of
0
or
25
mg/
kg/
day
MBA
in
water
for
14
days
and
were
sacrificed
24
hours
after
the
last
dose.
MBA
did
not
induce
any
spermatotoxicity
in
this
repeated­
dosing
study.

Systemic
Effects
The
toxicity
data
for
MBA
are
very
limited.
The
oral
LD
50
for
MBA
was
reported
as
177
mg/
kg
in
male
rats
(
Linder
et
al.,
1994a).
Single­
dose
(
0
or
100
mg/
kg)
and
14­
day
studies
(
0
or
25
mg/
kg/
day)
have
been
conducted
to
assess
the
spermatotoxicity
of
MBA.
No
general
toxicity
was
observed
with
either
dosing
regimen
(
Linder
et
al.,
1994a).
No
data
were
identified
on
young
animals
to
compare
the
potential
susceptibility
of
children
and
adults
to
the
toxic
effects
of
MBA.

Carcinogenicity
No
data
were
identified
on
the
carcinogenicity
of
MBA.
The
genotoxicity
data
for
MBA
have
provided
mixed
results.
MBA
was
mutagenic
in
S.
typhimurium
(
Giller
et
al.,
1997;
Kohan
et
al.,
1998;
NTP,
2000a)
and
induced
DNA
single­
strand
breaks
in
vitro
(
Stratton
et
al.,
1981),

but
did
not
induce
a
DNA
repair
system
(
SOS
DNA
repair)
in
Escherichia
coli
that
responds
to
primary
DNA
damage,
and
did
not
cause
micronuclei
in
a
newt
larvae
system
(
Giller
et
al.,
1997).
120
Based
on
the
absence
of
human
or
animal
data
and
only
equivocal
genotoxicity
data,
MBA
is
classified
as
Group
D:
Not
Classifiable
as
to
Human
Carcinogenicity
under
the
1986
Cancer
Risk
Assessment
Guidelines.
Under
EPA's
1999
Draft
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1999),
the
data
on
MBA
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.

Basis
for
RfD
and
MCLG
The
data
on
MBA
are
inadequate
to
derive
an
RfD,
because
there
are
no
systemic
toxicity
studies
of
sufficient
duration.
Similarly,
there
are
insufficient
data
to
assess
the
carcinogenic
potential
of
MBA.
In
the
absence
of
adequate
data
on
the
noncancer
or
cancer
effects
of
MBA,

no
MCLG
is
proposed.

Children's
Risk
in
Relation
to
the
MCLG
Data
relevant
to
potential
fetal
sensitivity
are
limited
to
a
single
developmental
study
reported
in
a
published
abstract
(
Randall
et
al.,
1991).
In
rats
administered
MBA
on
GD
6­
15,

the
NOAEL
was
50
mg/
kg/
day
for
both
maternal
effects
(
decreased
maternal
weight
gain)
and
for
fetal
effects
(
decreased
live­
fetus
size
and
increased
incidence
of
soft­
tissue
malformations).
The
induction
of
developmental
effects
only
at
doses
that
also
affected
maternal
body
weight
does
not
suggest
that
the
fetus
is
more
sensitive
to
the
toxic
effects
of
MBA.
121
2.3.5.
Bromochloroacetic
Acid
Developmental/
Reproductive
Effects
No
standard
peer­
reviewed
developmental
toxicity
studies
of
BCA
are
available.
NTP
(
1998b)
reported
the
results
of
a
short­
term
reproductive
and
developmental
toxicity
screening
protocol
for
BCA.
Two
groups
of
male
rats
and
three
groups
of
female
rats
were
treated
with
BCA
at
concentrations
of
0,
60,
200,
or
600
mg/
L.
In
males,
exposed
either
on
study
days
6­
35
or
exposed
on
study
days
6­
31
followed
by
a
BrdU
treatment,
no
consistent
treatment­
related
effects
on
epididymal
sperm
measures,
spermatid
head
counts,
sperm
morphology
or
sperm
motility
were
observed
in
either
group
at
necropsy.
Group
A
females
were
treated
prior
to
mating,
during
mating
to
treated
males,
and
during
early
gestation.
Group
B
females
were
mated
with
treated
males,
and
exposed
on
gestation
days.
Group
C
females
were
treated
with
BCA
prior
to
mating
and
were
subsequently
treated
with
BrdU.
No
effects
on
indices
of
mating
or
fertility
were
affected
in
Group
A
or
C
females.
Analysis
of
the
results
after
combining
data
for
groups
A
and
C
revealed
statistically
significant
decreases
to
70%
of
controls
for
live
fetuses
per
litter
and
to
75%
of
controls
in
total
implants
per
litter
at
50
mg/
kg/
day.
For
group
B
females,

increased
(
but
not
statistically
significant)
post­
implantation
losses,
and
total
resorptions
were
observed.
No
treatment­
related
effects
were
observed
upon
soft
tissue
examination
(
heart
and
brain)
of
the
fetuses.
The
NOAEL
for
statistically
significant
decreases
in
fertility
was
the
mid
dose,
19
mg/
kg/
day
(
200
ppm
treatment
group).
The
LOAEL
for
reproductive
and
developmental
effects
(
decreased
implants
per
litter
and
live
fetuses
per
litter)
was
the
high
dose
of
50
mg/
kg/
day
(
600
ppm
treatment
group).
122
In
a
study
submitted
for
publication,
Klinefelter
et
al.
(
2002)
administered
BCA
(
dissolved
in
deionized
water
and
pH­
adjusted)
by
gavage
to
adult
male
Sprague­
Dawley
rats
(
12/
dose)
at
doses
of
0,
24,
72,
or
216
mg/
kg/
day
for
14
days.
Body
weight
was
significantly
decreased
in
the
highest
dose
group.
Testis,
epididymis,
and
seminal
vesicle
weights
were
unaffected
by
BCA
treatment,
and
there
was
no
effect
on
testis
sperm
production
or
serum
testosterone.
Although
spermatid
numbers
were
not
altered
by
BCA
exposure,
a
significant
dose­
related
decline
in
epididymal
sperm
reserves
was
observed
at
72
and
216
mg/
kg/
day,
with
the
effect
on
cauda
epididymal
sperm
being
more
severe
than
on
caput
epididymal
sperm.
Dose­
related
decreases
in
serum
luteinizing
hormone
(
LH),
follicle­
stimulating
hormone
(
FSH),
and
prolactin
were
noted
in
all
dosed
groups,
with
statistical
significance
occurring
in
the
two
highest
dose
groups.

Doserelated
decreases
were
also
observed
in
the
percentage
of
motile
and
progressively
motile
cauda
sperm,
and
in
sperm
motion
parameters
(
i.
e.,
velocity
and
linearity)
and
the
percentage
of
morphologically
normal
cauda
and
caput
sperm.
Caput
sperm
abnormalities
were
characterized
by
an
increased
number
of
sperm
with
misshapen
heads
or
tail
defects,
whereas
cauda
sperm
abnormalities
consisted
mainly
of
an
increased
number
of
isolated
heads.
Histological
evaluation
of
the
testis
showed
a
dose­
related
increase
(
statistically
significant
in
the
two
highest
dose
groups)
in
the
number
of
Step
19
spermatids
retained
in
Stage
X
and
XI
of
the
spermatogenic
cycle.
Other
findings
included
a
dose­
related
increase
in
the
number
and
size
of
atypical
residual
bodies
in
Stages
X
and
XI
(
not
quantified)
and
a
shift
in
localization
of
these
bodies,
from
basal
migration
to
luminal
release,
with
increasing
BCA
dose.
According
to
the
study
authors,
the
LOAEL
for
altered
spermiation
in
this
study
was
24
mg/
kg/
day,
and
a
NOAEL
could
not
be
determined.
123
In
a
subsequent
study
in
the
same
report
(
Klinefelter
et
al.,
2002),
adult
male
Sprague­

Dawley
rats
(
10/
dose)
were
administered
14
daily
gavage
doses
of
BCA
(
dissolved
in
deionized
water
and
pH­
adjusted)
of
0,
8,
24,
or
72
mg/
kg/
day.
End
points
evaluated
were
the
same
as
those
assessed
in
the
previous
study.
Additionally,
sperm
protein
was
extracted
and
analyzed,
and
a
fertility
assessment
was
conducted
via
in
utero
insemination
of
untreated
females
(
synchronized
by
administering
a
subcutaneous
injection
of
an
agonist
for
luteinizing
hormone
releasing
hormone,
LHRH)
with
sperm
from
treated
males.
Inseminated
females
were
sacrificed
9
days
following
treatment,
and
implanted
fetuses
and
corpora
lutea
of
pregnancy
were
counted.
No
treatment­
related
changes
in
body
weight,
testes
weight,
or
the
weight
of
the
seminal
vesicles
were
observed.
However,
in
contrast
with
the
previous
study
conducted
by
the
same
authors,

epididymal
weights
were
reduced
at
72
mg/
kg/
day,
and
there
were
no
differences
between
treated
and
control
groups
in
any
of
the
hormonal
measurements.
Sperm
motion
parameters
were
consistently
altered
by
BCA
exposure.
Although
the
percentage
of
motile
sperm
was
only
decreased
in
the
high­
dose
group
(
72
mg/
kg/
day),
progressive
sperm
motility
was
decreased
at
all
doses
tested.
Altered
sperm
morphology
was
only
observed
at
72
mg/
kg/
day;
abnormalities
in
both
cauda
and
caput
sperm
were
similar
to
those
observed
in
the
earlier
study.
The
cauda
sperm
showed
increased
incidences
of
tail
defects
and
the
caput
sperm
showed
increased
incidences
of
isolated
heads.
In
utero
insemination
of
untreated
females
with
the
cauda
epididymal
sperm
from
treated
males
showed
a
significant
reduction
in
fertility
at
all
doses,
but
there
was
no
doseresponse
Fertility
rates
in
the
8,
24,
and
72
mg/
kg/
day
groups
were
33%,
44%,
and
37%,

respectively,
as
compared
with
75%
in
control
animals.
Decreased
levels
of
the
sperm
protein
SP22
correlated
with
the
reduction
in
fertility,
supporting
the
conclusion
that
SP22
is
a
useful
124
sperm
biomarker
of
fertility.
The
LOAEL
for
this
study
was
8
mg/
kg/
day,
the
lowest
dose
tested,

and
a
NOAEL
could
not
be
determined.

The
effects
of
BCA
on
male
reproduction
have
also
been
evaluated
following
oral
gavage
dosing
in
mice.
Luft
et
al.
(
2000)
reported
in
an
abstract
on
a
study
in
which
male
C57BL/
6
mice
(
12
mice/
group)
were
administered
daily
gavage
doses
of
0,
8,
24,
72,
or
216
mg/
kg
BCA
for
14
days.
After
14
days,
five
mice/
group
were
necropsied
for
histopathological
examination
of
the
testes,
epididymis,
and
seminal
vesicles.
The
remaining
seven
males
were
used
in
a
40­
day
breeding
assay
to
evaluate
the
effects
of
treatment
on
fertility.
Coital
plug­
positive
females
(
presumably
untreated)
were
replaced
daily,
and
uteri
were
dissected
14
days
later;
the
numbers
of
implantations,
resorptions
and
fetuses
were
determined.
No
effects
on
body
weight
or
reproductive
organ
weights
were
observed
for
any
of
the
dose
groups.
Results
of
histopathological
examination
of
the
male
reproductive
tissues
were
not
reported.
BCA
treatments
with
72
or
216
mg/
kg/
day
resulted
in
adverse
reproductive
performance,
but
only
for
the
first
ten
days
after
treatment
(
data
not
shown).
Significantly
decreased
measures
of
reproductive
performance
included:
mean
number
of
litters
per
male;
percentage
of
litters
per
female
bred,
as
measured
by
the
percent
of
plug­
positive
females
that
became
pregnant;
and
total
number
of
fetuses
per
male.
There
was
no
difference
in
the
number
of
coital
plugs,
suggesting
that
treatment
did
not
result
in
behavioral
effects
on
mating.
The
number
of
fetuses
per
litter,

number
of
resorptions,
and
number
of
terata
were
not
affected,
suggesting
that,
under
the
conditions
of
this
study,
adverse
reproductive
effects
on
the
male
did
not
induce
developmental
toxicity.
This
study
appears
to
have
identified
a
NOAEL
of
24
mg/
kg/
day
and
a
LOAEL
for
125
decreased
male
fertility
of
72
mg/
kg/
day,
but
a
definitive
conclusion
is
not
possible
until
the
full
study
(
rather
than
just
an
abstract)
is
published.

Andrews
et
al.
(
1999),
in
a
published
abstract,
treated
gestational
day
9
embryos
to
BCA
concentrations
up
to
300
µ
M
BCA
for
48
hours
and
then
scored
for
dysmorphology,

developmental
score,
head
length,
somite
number,
crown­
rump
length
and
embryo
lethality.

Treatment
with
BCA
at
300
µ
M
or
greater
was
dysmorphogenic;
and
BCA
at
200
µ
M
or
greater
significantly
affected
the
developmental
score,
head
length
and
somite
number.
In
another
whole
embryo
culture
study
designed
to
test
mechanisms
of
haloacetic
acid­
induced
dysmorphogenesis,

Ward
et
al.
(
2000)
reported
that
BCA
treatment
increased
the
accumulation
of
sub­
G1
events
(
a
measure
of
cells
with
less
than
the
normal
2n
complement
of
DNA
found
during
the
G1
stage
of
the
cell
cycle).
Because
the
characteristic
breakage
of
DNA
during
apoptosis
leads
to
this
accumulation
of
sub­
G1
events,
this
measure
is
often
used
as
an
indicator
of
apoptosis.
The
apoptotic
response
was
also
confirmed
by
fluorescence
microscopy
using
an
acidophilic
dye.
The
ability
of
BCA
to
induce
an
apoptotic
response
suggested
to
the
authors
that
the
developmental
toxicity
of
BCA,
particularly
the
induction
of
embryonic
neural­
tube
defects,
is
mechanistically
associated
with
its
ability
to
increase
apoptosis.
The
whole
embryo
testing
data
provide
mechanistic
support
for
the
developmental
toxicity
of
the
brominated
acetic
acids
observed
in
vivo.
126
Systemic
Effects
Oral
toxicity
studies
have
identified
the
kidney
and
liver
as
potential
target
organs
for
BCA
toxicity.
Parrish
et
al.
(
1996)
administered
male
mice
0,
25,
125,
or
500
mg/
kg/
day
BCA
in
drinking
water
for
21
days.
Administered
doses
were
estimated
from
drinking
water
concentrations
based
on
default
water
intake
values
for
male
B6C3F1
mice
(
EPA,
1988).

Increased
liver
weight
was
induced
by
the
highest
dose
tested
(
500
mg/
kg/
day),
with
a
NOAEL
of
125
mg/
kg/
day.
NTP
(
1998b)
evaluated
target
organ
toxicity
as
part
of
a
reproductive
and
developmental
screening
assay
in
rats.
The
NOAEL
was
15
mg/
kg/
day,
and
the
LOAEL
was
39
mg/
kg/
day
for
treatment­
related
liver
histopathological
changes
(
cytoplasmic
vacuolization)
and
increased
liver
weight.
There
is
no
evidence
that
children
are
more
sensitive
to
the
toxic
effects
of
BCA
than
are
adults.
However,
these
data
are
limited
and
no
definitive
conclusions
can
be
made.

Carcinogenicity
In
a
published
abstract,
Stauber
et
al.
(
1995)
reported
that
BCA
induces
liver
tumors
in
B6C3F1
mice.
There
are
no
published
reports
of
a
full
bioassay
with
BCA.
BCA
was
mutagenic
in
S.
typhimurium
(
NTP,
2000b)
and
induced
oxidative
DNA
damage
in
the
livers
of
mice
given
treated
drinking
water
(
Parrish
et
al.,
1996).
Based
on
the
absence
of
human
or
animal
data
and
very
limited
genotoxicity
data,
BCA
is
classified
as
Group
D:
Not
Classifiable
as
to
Human
Carcinogenicity
under
the
1986
Carcinogen
Risk
Assessment
Guidelines.
Under
EPA's
Draft
1999
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1999),
the
data
on
BCA
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.
127
Basis
for
RfD
and
MCLG
The
data
on
BCA
were
insufficient
for
the
derivation
of
an
RfD,
because
there
are
no
systemic
toxicity
studies
of
sufficient
duration.
Similarly,
there
are
insufficient
data
to
assess
the
carcinogenic
potential
of
BCA.
In
the
absence
of
adequate
data
on
the
noncancer
or
cancer
effects
of
BCA,
no
MCLG
is
proposed.

Children's
Risk
in
Relation
to
the
MCLG
Data
on
the
effects
of
BCA
on
fetuses
and
children
are
very
limited.
NTP
(
1998b)

conducted
a
reproductive
and
developmental
toxicity
screening
assay.
The
NOAEL
for
decreased
live
fetuses/
litter
and
decreased
total
implants/
litter
was
19
mg/
kg/
day,
while
similar
doses
were
considered
the
NOAEL
for
general
toxicity
in
both
adult
males
(
15
mg/
kg/
day)
and
females
(
19
mg/
kg/
day).
Decreased
fertility
was
observed
when
adult
males
were
exposed
to
doses
as
low
as
8
mg/
kg/
day;
no
NOAEL
was
identified
(
Klinefelter
et
al.,
2002).
Although
this
latter
endpoint
is
a
reproductive
effect,
the
LOAEL
is
based
on
exposure
of
the
adult
male;
no
data
were
identified
on
whether
exposure
of
young
males
enhances
their
sensitivity.
Thus,
the
data
are
limited,
but
do
not
support
the
hypothesis
that
fetuses
or
children
are
more
sensitive
than
adults.

2.3.6.
Dibromoacetic
Acid
Developmental/
Reproductive
Effects
Two
recent
studies
have
evaluated
the
reproductive
and
developmental
toxicity
of
DBA
in
Sprague­
Dawley
rats.
Christian
et
al.
(
2001b)
administered
DBA
in
deionized
drinking
water
to
male
and
female
rats
(
10/
sex/
group)
at
concentrations
of
0,
125,
250,
500,
or
1000
ppm,
128
beginning
14
days
prior
to
cohabitation
and
continuing
through
gestation
and
lactation
(
63­
70
days
of
treatment).
The
average
daily
doses
(
based
on
measured
water
consumption
and
body
weights)
varied,
depending
on
the
phase
of
reproduction.
Among
the
pups,
two
male
and
two
female
weanlings
from
each
litter
were
selected
for
one
additional
week
of
observation
(
postweanling
days
1­
8,
commencing
on
LD
29);
daily
food
intake,
drinking
water
consumption,

and
body
weights
were
recorded,
and
necropsy
was
conducted
at
sacrifice.
Apparent
taste
aversion
was
associated
with
an
exposure­
dependent
reduction
in
water
consumption,
which
was
paralleled
by
a
reduction
in
food
intake
at
all
concentrations.
Decreased
body
weight
gain
was
observed
in
parental
animals
and
postweanling
pups
at
the
two
highest
exposure
levels.
Estrous
cycling
was
unaffected
in
the
female
rats.
The
only
observed
adverse
reproductive
effect
was
a
possible
reduction
in
mating
performance
in
the
1000
ppm
group,
as
evidenced
by
a
slight
but
nonsignificant
increase
in
the
number
of
days
of
cohabitation
and
a
decrease
in
the
number
of
mated
pairs
(
6/
10
in
the
1000
ppm
group
versus
9­
10/
10
in
all
other
groups).
There
were
no
effects
on
pre­
and
postimplantation
losses,
live
litter
sizes,
and
gross
external
morphology
or
sex
ratios
in
the
pups.
Although
an
exposure­
related
decrease
in
pup
body
weights
was
noted,
these
findings
were
attributed
to
decreased
water
and
food
consumption
resulting
from
the
poor
palatability
of
DBA­
treated
drinking
water.
Based
on
a
lack
of
statistically
significant,

treatmentrelated
findings,
the
parental
and
reproductive/
developmental
NOAEL
for
this
study
is
1000
ppm
(
the
highest
dose
tested),
and
a
LOAEL
could
not
be
determined.
Based
on
measured
water
intake
and
body
weights,
the
NOAEL
corresponds
to
a
dose
of
66
mg/
kg/
day
in
parental
males,

>
60
mg/
kg/
day
in
parental
females,
and
>
82
mg/
kg/
day
for
developmental
effects.
129
The
Chlorine
Chemistry
Council
(
CCC,
2001)
recently
completed
a
two­
generation
drinking
water
study
of
DBA
in
rats.
The
report
has
been
peer
reviewed
by
an
EPA
scientific
advisory
group
that
has
evaluated
and
accepted
the
results.
Male
and
female
Crl:
CD
Sprague­

Dawley
rats
(
30/
sex/
exposure
group)
were
administered
DBA
in
drinking
water
at
concentrations
of
0,
50,
250,
or
650
ppm
continuously
from
initiation
of
exposure
of
the
parental
(
P)
generation
male
and
female
rats
through
weaning
of
the
F
2
offspring.
The
high
concentration
was
chosen
based
on
a
range­
finding
study
that
found
that
650
ppm
was
the
highest
concentration
expected
to
allow
survival
of
the
F
1
offspring.
For
the
P
generation,
DBA
exposure
was
initiated
at
43
days
of
age
and
continued
from
premating
until
study
day
(
SD)
92
for
males,
and
from
premating
through
gestation
and
a
29­
day
period
of
lactation
(
LD
1­
29)
for
females.
Parental
generation
offspring
(
F
1
males
and
females)
were
exposed
in
utero
during
gestation,
and
during
lactation
(
LD
1­
29);
selected
F
1
males
and
females
(
30/
sex/
exposure
group)
were
further
exposed
during
a
postweaning
period
of
at
least
71
days,
which
continued
through
mating,
gestation,
and
lactation.

All
other
F
1
pups
were
sacrificed
on
LD
29.
All
F
1
adult
females
and
their
offspring
(
F
2
generation)
were
sacrificed
on
LD
29.

Water
consumption
was
statistically
significantly
decreased
at
all
exposure
levels,

presumably
due
to
taste
aversion,
and
food
intake
was
significantly
reduced
at
the
two
highest
exposure
concentrations.
Body
weights
and
body
weight
gains
for
high­
dose
P
males
and
females
were
significantly
reduced
during
the
premating
period
and
were
significantly
decreased
for
highdose
P
females
during
gestation
and
lactation.
F
1
male
and
female
pups
had
significantly
reduced
body
weights
at
all
exposure
levels,
sufficient
for
the
study
authors
to
delay
weaning
until
LD
29
to
ensure
pup
survival.
By
LD
29,
the
body
weights
of
pups
in
the
50
ppm
group
were
similar
to
130
control
pup
weights.
Throughout
the
postweaning/
premating
period,
F
1
males
and
females
in
the
250
and
650
ppm
groups
weighed
significantly
less
than
controls,
and
the
females
continued
to
exhibit
significant
reductions
in
body
weight
(
compared
to
controls)
during
gestation
and
lactation.
The
body
weights
of
F
2
pups
in
the
two
highest
dose
groups
were
also
reduced;

however,
these
reductions
were
not
reported
to
be
statistically
significant,
relative
to
control
group
values,
by
the
end
of
lactation.
No
treatment­
related
effects
were
reported
in
either
generation
for
estrous
cycling,
number
of
days
in
cohabitation,
duration
of
gestation,
fertility
index,
gestation
index,
number
and
sex
of
offspring
per
litter,
number
of
implantation
sites,
litter
size,
lactation
index,
percent
pup
survival,
litter
sex
ratio,
and
gross
malformations.
In
the
650
ppm
group,
preputial
separation
was
significantly
delayed
in
the
F
1
male
rats
(
50.5
days
versus
48.1
days
in
controls),
and
vaginal
patency
in
F
1
female
rats
was
also
significantly
retarded
(
36.3
days
versus
33.4
days
in
controls);
no
significant
differences
were
seen,
however,
when
the
data
were
analyzed
using
body
weight
as
a
covariate.
These
effects
were
considered
to
be
due
to
a
general
retardation
of
growth
associated
with
the
significant
reduction
in
body
weight
in
this
exposure
group
at
weaning.
In
F
2
male
and
female
pups,
anogenital
distance
did
not
differ
from
controls
on
LD
1
but
was
significantly
reduced
in
male
pups
in
the
250
and
650
ppm
by
LD
22;

these
findings
were
also
considered
to
be
associated
with
a
general
retardation
of
growth
rather
than
being
treatment­
related.

An
increased
incidence
of
malformations
of
the
male
reproductive
tract,
including
small
testes
and
small
or
absent
epididymides,
was
observed
in
four
males
in
the
F
1
group
exposed
to
650
ppm
and
was
considered
to
be
treatment­
related.
Histopathologic
examination
of
reproductive
organs
of
P
and
F
1
male
rats
in
the
250
and
650
ppm
groups
(
N
=
131
30/
group/
generation)
showed
a
consistent
and
significant
exposure­
related
increase
in
retained
Step
19
spermatids
in
Stage
IX
and
X
tubules
and
in
increased
or
abnormal
residual
bodies
in
affected
seminiferous
tubules.
Diffuse
testicular
atrophy
and
phagocytized
Step
19
nuclei
in
the
basilar
area
of
affected
seminiferous
tubules
were
also
observed,
although
at
a
lower
incidence.

Other
testicular
abnormalities
in
250
and
650
ppm
male
rats
of
both
generations
included
increased
amounts
of
exfoliated
spermatogenic
cells/
residual
bodies
in
epididymal
tubules,

atrophy,
and
hypospermia.
No
effects
were
found
on
percent
motile
sperm,
sperm
count,
sperm
density,
and
number
and
percent
of
morphologically
abnormal
sperm
for
exposed
groups.

An
increase
in
the
incidence
and
intensity
of
extramedullary
hematopoiesis
in
the
red
pulp
of
the
spleen
occurred
in
the
F
1
generation
female
rats
in
the
650
ppm
group
and
may
have
been
treatment­
related.
Decreased
cellularity
of
the
cortical
lymphoid
area
of
the
thymus
was
noted
in
P
generation
females
in
the
two
highest
dose
groups.

Based
on
testicular
histomorphology
indicative
of
abnormal
spermatogenesis
in
P
and
F
1
males,
the
parental
and
reproductive/
developmental
toxicity
LOAEL
and
NOAEL
are
250
and
50
ppm,
respectively.
Assuming
mean
daily
doses
can
be
estimated
by
the
mean
consumed
dose
from
weaning
to
termination
of
the
study,
the
LOAEL
and
NOAEL
for
the
F
1
generation
are
22.0
and
4.5
mg/
kg/
day,
respectively;
similar
doses
can
be
estimated
for
the
parental
generation.

Cummings
and
Hedge
(
1998)
studied
the
effects
of
DBA
exposure
during
early
pregnancy
in
rats.
Female
Holtzman
rats
(
8/
dose
group)
were
administered
gavage
doses
of
0,
62.5,
125,
or
250
mg/
kg/
day
on
days
1
through
8
of
pregnancy.
Administration
of
500
mg/
kg/
day
induced
moribund
behavior
and
lethality;
therefore,
dosing
was
discontinued
and
these
animals
were
not
further
evaluated
for
reproductive
endpoints.
Treated
animals
from
the
other
dose
groups
were
132
sacrificed
on
day
9
of
pregnancy,
and
the
following
endpoints
were
scored:
body
and
reproductive
organ
weights;
serum
levels
of
progesterone,
17 ­
estradiol
and
luteinizing
hormone;
number
of
implantation
sites;
number
of
resorptions;
corpora
lutea
and
pre­
implantation
loss.
The
only
response
that
was
affected
was
a
170
%
increase
in
serum
17 ­
estradiol
at
250
mg/
kg/
day.
A
second
set
of
females
were
dosed
similarly
to
the
first
set
of
females,
sacrificed
on
gestation
day
20,
and
evaluated
for
body
weight,
number
of
pups,
number
of
resorptions,
pup
weights,
weight
of
the
placentae
and
pre­
implantation
losses.
No
differences
in
any
of
these
measures
were
observed
between
treated
animals
and
controls.
The
authors
concluded
that
DBA
had
little
effect
on
female
reproduction
for
the
parameters
included
in
the
study.
They
noted
that
effects
on
ovarian
function
and
future
fertility
were
not
tested
and
such
tests
would
be
warranted
by
the
observed
increase
in
serum
17 ­
estradiol.
Based
on
this
study,
the
NOAEL
was
125
mg/
kg/
day,

based
on
increase
in
serum
17 ­
estradiol;
the
FEL
for
acute
systemic
toxicity
is
500
mg/
kg/
day.

This
study
is,
however,
limited
by
the
small
size
of
the
dose
groups.

There
has
been
considerable
interest
in
the
male
reproductive
tract
toxicity
of
DBA,
in
part
because
DCA
is
known
to
be
a
male
reproductive
toxicant.
Linder
and
colleagues
used
a
number
of
different
experimental
protocols
to
study
the
effects
of
DBA
on
spermatogenesis
and
the
resulting
consequences
for
male
fertility,
(
Linder
et
al.,
1994a;
Linder
et
al.,
1994b;
Linder
et
al.,
1995;
Linder
et
al.,
1997b).
Their
studies
show
that
DBA
is
clearly
spermatotoxic
following
high­
dose
single
gavage
exposures
or
repeated
gavage
exposures
for
longer
periods
of
time
(
up
to
79
days).
Effects
on
spermatogenesis
appear
to
be
a
sensitive
endpoint,
since
effects
are
observed
in
the
absence
of
other
toxicity.
In
general,
histopathological
evidence
for
changes
in
seminiferous
tubule
staging
was
observed
at
the
lowest
doses.
Changes
in
retention
of
Step
19
133
spermatids
were
generally
noted
as
the
earliest
effect
and
occurred
following
repeated
dosing
with
10
mg/
kg/
day,
but
not
2
mg/
kg/
day
following
31
or
79
daily
doses.
Significant
changes
in
sperm
count,
morphology
and
motility
are
generally
observed
at
higher
doses
than
those
associated
with
early
histopathological
changes.
For
example,
sperm
quality
measures
(
morphology
and
motility)

were
not
significantly
affected
until
administration
of
50
mg/
kg/
day
for
31
or
79
days
(
Linder
et
al.,
1995).
Evidence
of
spermatotoxicity
was
not
necessarily
paralleled
by
changes
in
male
fertility.
For
example,
even
at
the
high
dose
of
50
mg/
kg/
day,
which
was
clearly
spermatotoxic,

only
marginal
(
if
any)
effect
on
male
fertility
was
observed
(
Linder
et
al.,
1995).
The
results
of
the
studies
by
Linder
and
colleagues
are
supported
by
those
in
the
two­
generation
reproductive
toxicity
study
(
CCC,
2001),
in
which
altered
spermiation,
but
no
change
in
male
fertility
parameters,
was
observed.
The
most
sensitive
effect
of
long­
term
exposure
to
DBA
was
the
histopathological
effects
on
the
male
reproductive
tract,
with
a
NOAEL
of
2
mg/
kg/
day
(
Linder
et
al.,
1997b).

In
contrast
to
the
results
by
Linder
et
al.
(
1994a,
1994b),
Vetter
et
al.
(
1998)
did
not
find
evidence
of
spermatotoxicity
in
another
single­
dose
spermatotoxicity
study
with
sexually­
mature
male
Crl:
CD(
SD)
BR
rats
(
4­
5/
dose
group)
dosed
with
0,
600,
or
1200
mg/
kg
DBA
in
10
mL/
kg
deionized
water.
No
changes
in
measured
sperm
parameters
(
motility,
morphology,
and
cellmembrane
permeability)
were
reported
at
either
dose;
however,
mild
testes
histopathology
(
the
presence
of
basophilic
bodies)
was
observed
in
both
dose
groups.
In
the
high­
dose
group,
overt
toxicity
was
observed
and
included
lethargy,
irregular
gait,
decreased
feces,
ocular
discharge,

dyspnea,
and
abnormal
respiratory
sounds.
No
overt
toxicity
was
observed
in
the
low­
dose
group.
Based
on
the
clinical
findings,
1200
mg/
kg
was
considered
to
be
an
acute
frank
effects
134
level
(
FEL).
The
LOAEL
was
600
mg/
kg
for
testes
histopathology,
and
a
NOAEL
could
not
be
determined.
The
reasons
for
the
differences
between
the
findings
in
this
study
and
those
of
Linder
et
al.
(
1994)
are
not
known.

Klinefelter
et
al.
(
2000),
in
an
abstract,
reported
the
effects
of
DBA
on
pubertal
development
and
adult
reproductive
function
in
Sprague­
Dawley
rats
(
3
litters/
dose)
given
drinking
water
containing
0,
400,
600,
or
800
mg/
L
DBA
from
gestation
day
15
through
postnatal
day
98.
These
drinking
water
concentrations
were
chosen
to
result
in
doses
of
0,
50,
75,
and
100
mg/
kg/
day
(
personal
communication).
After
weaning,
male
offspring
were
exposed
to
the
same
concentrations
of
DBA
in
drinking
water.
Decreased
body
weight
in
offspring
compared
to
controls
throughout
reproductive
development
was
observed
in
the
high­
dose
animals.

Histopathology
examination
revealed
seminiferous
tubules
containing
only
Sertoli
cells
in
each
dose
group,
and
decreased
epididymis
weight
(
the
percent
decrease
was
not
specified)
in
the
600
and
800
mg/
L
dose
groups.
Fertility
of
sperm
from
treated
males
was
decreased
by
DBA
treatment.
The
number
of
implants
per
corpora
lutea
in
females
artificially
inseminated
with
sperm
from
treated
males
decreased
from
70%
for
controls
to
49%,
15%,
and
15%
for
the
400,

600,
and
800
ppm
treatment
groups,
respectively.
Levels
of
the
sperm
protein
SP22,
found
to
be
highly
correlated
with
fertility,
were
significantly
decreased
in
all
the
treatment
groups.
No
NOAEL
for
the
fertility
of
sperm
of
treated
males
was
identified.

In
a
second
recent
abstract,
Veeramachaneni
et
al.
(
2000)
exposed
male
Dutch­
belted
rabbits
(
10/
group)
to
DBA­
treated
drinking
water
from
gestation
day
15
throughout
life.
The
average
daily
doses
were
reported
as
0,
0.97,
5.05,
and
54.2
mg/
kg/
day.
The
fertility
of
sperm
from
24­
week­
old
males
was
assessed
by
artificial
insemination
of
two
6­
month­
old
rabbit
does
135
per
sample
of
sperm
from
each
male.
Conception
rates
were
significantly
decreased
(
p<
0.01)
in
does
inseminated
with
sperm
from
males
at
every
dose
group.
Of
the
53
pups
born
to
does
inseminated
with
sperm
from
the
high­
dose
males,
one
pup
had
both
cleft
palate
and
cranioschisis
and
two
had
cranioschisis.
At
25
weeks,
the
offspring
were
necropsied
and
no
difference
from
controls
in
body
weight,
anogenital
distance
or
sex
organ
weights
were
observed.
These
data
suggest
that
the
lowest
dose
tested,
0.97
mg/
kg/
day,
was
the
LOAEL
for
decreased
male
fertility
in
rabbits.

In
a
recent
reproductive
toxicity
study,
Klinefelter
et
al.
(
2001)
confirmed
a
DBA­
related
delay
in
sexual
maturity
and
effects
on
sperm
quality
in
Sprague­
Dawley
rats
exposed
to
DBA.
In
this
study,
the
rats
were
exposed
at
concentrations
of
0,
4,
40,
or
400
ppm
via
drinking
water.

Statistically
significant,
body
weight­
independent
delays
in
both
vaginal
opening
and
preputial
separation
were
observed
in
rats
exposed
to
400
ppm
DBA
in
drinking
water
(
approximately
40
mg/
kg/
day).
The
authors
stated
that
pubertal
delays
were
observed
only
when
exposure
was
continuous
from
gestation
through
adulthood,
and
not
when
exposure
only
occurred
from
weaning
through
adulthood.
A
related
study
by
the
same
authors
(
manuscript
in
preparation)

found
that
levels
of
specific
proteins
in
the
sperm
membrane
(
particularly
SP
22)
were
significantly
decreased
at
4
and
40
ppm
(
corresponding
to
0.4
and
4
mg/
kg/
day).
As
for
the
pubertal
delays,
these
changes
were
only
seen
in
animals
that
were
exposed
to
DBA
via
drinking
water
from
gestation
through
adulthood,
not
in
animals
exposed
from
weaning
through
adulthood.
The
fertility
via
artificial
(
i.
e.
in
utero)
insemination
was
also
significantly
reduced
in
animals
exposed
to
the
400
ppm
(
40
mg/
kg).
This
study
indicates
that
exposure
to
DBA
in
the
136
drinking
water
alters
sexual
maturity,
and
sperm
quality
and
fertility,
and
indicates
that
effects
may
occur
at
doses
lower
than
previously
thought.

No
peer­
reviewed
developmental
toxicity
studies
of
DBA
are
available,
but
the
developmental
toxicity
of
DBA
has
been
reported
in
two
related
abstracts.
Narotsky
et
al.
(
1996)

studied
developmental
effects
of
DBA
in
CD­
1
mice
dosed
by
gavage
with
0,
0.11,
0.23,
0.46,

0.92,
1.8,
2.8,
or
3.7
mmol/
kg/
day
(
0,
24,
50,
100,
200,
392,
610,
or
806
mg/
kg/
day)
on
gestation
days
6­
15.
Mice
were
allowed
to
deliver
and
their
litters
were
examined
on
postnatal
days
1
and
6.
Maternal
effects
were
limited
to
piloerection
and
motor
depression
at
the
highest
dose.
There
was
delayed
parturition
at
all
dose
levels,
but
it
is
not
clear
if
this
effect
is
adverse.
At
the
highest
dose
there
was
increased
prenatal
mortality,
with
only
three
of
nine
litters
viable
at
birth.

Increased
postnatal
mortality
was
also
seen
at
2.8
mmol/
kg/
day
(
610
mg/
kg/
day),
and
3.7
mmol/
kg/
day
(
806
mg/
kg/
day).
Decreased
pup
weight
was
observed
at
3.7
mmol/
kg/
day
on
postnatal
day
1.
Pup
weight
was
also
decreased
on
postnatal
day
6
at
2.8
mmol/
kg/
day.
Short,

kinked
or
absent
tails
were
observed
at
the
two
highest
doses.
Based
on
these
results,
the
authors
concluded
that
DBA
was
a
developmental
toxicant.

In
a
second
published
abstract,
DBA
was
administered
to
CD­
1
mice
by
gavage
in
distilled
water
on
gestation
days
6­
15,
at
doses
of
0,
50,
100,
or
400
mg/
kg/
day
(
Narotsky
et
al.,
1997b).

Maternal
toxicity
was
not
observed.
Litters
were
removed
by
cesarean
section
on
gestation
day
17,
and
half
of
the
fetuses
in
each
litter
were
examined
for
skeletal
defects
and
the
other
half
for
visceral
effects.
There
was
no
effect
on
prenatal
survival,
fetal
weight
or
skeletal
development.

Hydronephrosis
was
noted
at
100
and
400
mg/
kg/
day,
as
well
as
renal
agenesis
(
small
kidneys)
at
400
mg/
kg/
day.
Based
on
the
summary
data
provided
in
these
abstracts,
the
NOAEL
would
be
137
50
mg/
kg/
day
and
the
LOAEL
for
fetal
kidney
malformations
would
be
100
mg/
kg/
day.

In
addition
to
the
developmental
studies
in
animals,
the
developmental
toxicity
of
DBA
was
evaluated
in
whole
embryo
culture
studies.
Hunter
et
al.
(
1996)
reported
on
the
developmental
effects
of
DBA
in
whole
embryo
culture.
DBA
induced
malformations
in
the
embryos
at
sub­
lethal
doses.
Andrews
et
al.
(
1999)
evaluated
the
developmental
effects
of
DBA
in
rat
embryos.
Gestational
day
9
embryos
were
exposed
to
10
to
400
µ
M
DBA
for
48
hours.

Concentrations
of
200
µ
M
or
greater
DBA
resulted
in
developmental
effects.

In
another
whole­
embryo
culture
study
designed
to
test
mechanisms
of
haloacetic
acidinduced
dysmorphogenesis,
Ward
et
al.
(
2000)
reported
that
DBA
caused
only
a
limited
induction
of
sub­
G1
cell­
cycle
events
(
an
increase
in
hypodiploid
cells
or
cell
debris
resulting
from
DNA
breakage
during
apoptosis).
However,
based
on
the
more
significant
effects
of
BCA
and
DCA
treatment
on
the
induction
of
apoptosis,
the
authors
suggested
the
haloacetic
acids
might
induce
embryo
dysmorphogenesis
through
their
ability
to
increase
apoptosis.

Negative
results
were
obtained
in
a
non­
mammalian
screening
assay
regarding
the
developmental
toxicity
of
DBA.
Gardner
and
Toussant
(
1999)
evaluated
developmental
toxicity
of
DBA
in
the
frog
embryo
teratogenesis
assay
­
Xenopus
(
FETAX)
(
a
96­
hour
toxicity
test),

with
and
without
metabolic
activation.
Endpoints
evaluated
were
embryolethality
(
LC
50),

embryonic
malformations
(
EC
50),
minimum
concentration
to
inhibit
growth
(
MCIG),
and
a
teratogenicity
index
(
TI
 
the
ratio
of
the
LC
50
to
the
EC
50).
The
FETAX
assay
is
considered
to
be
a
reliable
developmental
toxicity
screening
assay;
Dawson
and
Bantle
(
1987)
have
estimated
that
its
predictive
accuracy
for
identifying
known
mammalian
or
human
developmental
toxicants
approaches
or
exceeds
85%.
Under
the
conditions
of
this
study,
DBA
did
not
exhibit
teratogenic
138
potential.
Further,
malformations
did
not
appear
to
increase
in
severity
or
prevalence
with
increasing
DBA
concentrations,
with
or
without
metabolic
activation.
Although
negative
result
was
obtained
in
one
study,
the
overall
whole­
embryo
testing
data
provide
support
for
the
developmental
toxicity
of
the
brominated
acetic
acids
observed
in
vivo.

Systemic
Effects
The
liver
toxicity,
immunotoxicity,
and
neurotoxicity
of
DBA
have
been
evaluated.

Parrish
et
al.
(
1996)
exposed
male
mice
to
DBA­
treated
drinking
water
for
21
days.
Estimated
doses
based
on
default
water­
intake
values
for
mice
were
0,
25,
125,
or
500
mg/
kg/
day.

Increased
liver
weight
was
observed
beginning
at
125
mg/
kg/
day
and
supported
by
evidence
of
oxidative
stress
at
the
same
dose;
the
NOAEL
was
25
mg/
kg/
day.
In
an
immunotoxicity
study
(
NTP,
1999),
female
mice
were
given
DBA
in
their
drinking
water
for
28
days.
The
resulting
DBA
doses
were
not
reported
by
the
study
authors,
but
based
on
body
weight
and
waterconsumption
data,
doses
were
estimated
for
each
of
four
sub­
studies
conducted.
Absolute
and
relative
liver
weights
were
increased
beginning
at
14
mg/
kg/
day
and
increased
in
a
dosedependent
fashion.
Liver
weight
changes
were
not
chosen
as
the
critical
effect
for
this
study
due
to
the
absence
of
histopathology
or
clinical
chemistry
confirming
that
the
observed
liver­
weight
increases
were
adverse.
Analogy
to
other
haloacetic
acids
suggests
that
DBA
is
likely
to
induce
adverse
liver
effects
(
histopathology
or
clinical
chemistry
changes)
at
sufficiently
high
doses,
but
none
of
these
effects
were
observed
in
this
study.

The
immunotoxicity
of
DBA
administered
in
drinking
water
has
been
evaluated
in
four
studies
in
mice
exposed
to
drinking
water
containing
0,
125,
250,
500,
1000,
or
2000
mg/
L
DBA
139
for
28
days
(
NTP,
1999).
A
number
of
different
end
points
were
assessed,
including
thymus
and
spleen
weights,
number
and
type
of
spleen
cells,
macrophage
activation,
natural
killer
(
NK)
cell
activity,
and
specific
and
general
IgM
antibody­
forming
responses.
The
most
sensitive
and
reliable
measure
was
a
decrease
in
spleen
IgM
antibody­
forming
cell
responses,
representing
a
clear
decrease
in
immune
system
function,
accompanied
by
an
increase
in
the
number
of
spleen
macrophages.
The
LOAEL
and
NOAEL
for
these
endpoints
were
approximately
70
and
38
mg/
kg/
day,
respectively.

Phillips
et
al.
(
2002,
published
abstract)
examined
the
neurobehavioral
toxicity
of
DBA
in
adolescent
(
28­
day­
old)
male
and
female
F344
rats
(
12/
sex/
dose)
given
DBA
in
drinking
water
at
concentrations
of
0,
200,
600,
or
1500
mg/
L
(
mean
doses
calculated
by
the
authors
as
0,
20,
72,

and
161
mg/
kg/
day)
for
6
months.
In
both
sexes,
body
weight
was
significantly
depressed
in
the
highest
dose
group
but
overall
health
status
was
unaltered.
A
neurobehavioral
test
battery
was
administered
to
all
animals
at
1,
2,
4,
and
6
months.
Dose­
dependent
neuromuscular
toxicity,

characterized
by
mild
gait
abnormalities,
hypotonia,
and
decreased
forelimb
and
hindlimb
grip
strength,
was
observed
in
both
sexes.
Sensorimotor
responsiveness,
as
measured
by
responses
to
a
tail
pinch
and
click,
was
reduced
at
all
doses
but
did
not
progress
with
continued
exposure
to
DBA.
Decreased
motor
activity
was
noted
in
both
sexes
in
the
high­
dose
group,
whereas
a
chest
clasping
response
was
only
observed
in
high­
dose
females.
Neuropathologic
examination
revealed
significant
myelin
fragmentation,
axonal
swelling,
and
axonal
degeneration
in
the
lateral
and
ventral
areas
of
the
spinal
cord
white
matter
in
the
high­
dose
group.
In
the
mid­
and
highdose
groups,
small
numbers
of
swollen,
eosinophilic
or
faintly
basophilic,
and
occasionally
vacuolated
neurites
were
observed
in
the
spinal
cord
gray
matter,
and
appeared
to
represent
140
axonal
degeneration.
Neuropathologic
examination
has
not
yet
been
conducted
in
the
low­
dose
group.
No
treatment­
related
neuropathology
was
noted
in
the
eyes,
peripheral
nerves,
peripheral
ganglia,
or
brain.
Based
on
neurobehavioral
abnormalities,
the
LOAEL
was
20
mg/
kg/
day,
the
lowest
dose
tested,
and
a
NOAEL
could
not
be
determined
Carcinogenicity
No
complete
reports
of
bioassays
with
DBA
have
been
published.
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.
DBA
has
also
provided
nearly
uniformly
positive
results
in
the
genotoxicity
assays
conducted.
Positive
results
have
been
reported
in
S.
typhimurium
assays
(
Giller
et
al.,
1997;
Kohan
et
al.,
1998;
NTP
2000c)
and
assays
for
DNA
damage
repair
(
Giller
et
al,
1997;
Mayer
et
al.,
1996).
DBA
has
been
shown
to
induce
oxidative
DNA
damage
(
Austin
et
al.,
1996;
Parrish
et
al.,
1996).
On
the
other
hand,
no
induction
of
micronuclei
was
reported
in
a
newt
larvae
system
(
Giller
et
al.,
1997).
The
clastogenicity
of
DBA
has
not
been
reported
in
other
assays
using
a
standard
protocol,
but
DBA
has
been
reported
to
be
co­
clastogenic
(
Sasaki
and
Kinae,
1995).
As
a
whole,
these
data
support
the
conclusion
that
DBA
is
genotoxic.

In
the
absence
of
full
cancer
bioassay
data,
DBA
is
classified
as
Group
D:
Not
Classifiable
under
the
1986
Cancer
Risk
Assessment
Guidelines.
Under
EPA's
Draft
1999
Guidelines
for
Carcinogen
Risk
Assessment,
there
is
suggestive
evidence
of
carcinogenicity
of
DBA,
but
the
evidence
is
not
sufficient
to
assess
human
carcinogenic
potential.
The
existing
weight
of
the
evidence
in
support
of
the
carcinogenicity
of
DBA
includes
its
structural
analogy
to
the
141
demonstrated
rodent
tumorigen
DCA,
strong
evidence
for
genotoxicity,
and
positive
preliminary
results
in
cancer
studies
reported
in
published
abstracts
(
So
and
Bull,
1995;
Stauber
et
al.,
1995).

Thus,
the
animal
data
raise
a
concern
for
potential
carcinogenic
effects,
but
are
not
sufficient
for
reaching
a
conclusion
with
regard
to
human
carcinogenicity.
A
2­
year
NTP
toxicity
and
carcinogenicity
study
with
DBA
is
ongoing
(
NTP,
2000a).

Basis
for
RfD
and
MCLG
The
data
on
DBA
are
insufficient
for
the
derivation
of
an
RfD,
because
there
are
no
systemic
toxicity
studies
of
sufficient
duration.
Similarly,
there
are
insufficient
data
to
assess
the
carcinogenic
potential
of
DBA.
In
the
absence
of
adequate
data
on
the
noncancer
or
cancer
effects
of
DBA,
no
MCLG
is
proposed.

Children's
Risk
in
Relation
to
the
MCLG
The
data
are
mixed
as
to
whether
children
may
be
more
sensitive
than
adults
to
the
effects
of
DBA.
Christian
et
al.
(
2001b)
found
no
effects
in
the
pups
of
rats
continuously
administered
DBA
in
drinking
water
during
premating
and
mating
(
adult
males
and
females)
and
gestation
and
lactation
(
females).
Similarly,
pups
given
DBA
in
drinking
water
for
one
week
following
weaning
did
not
exhibit
any
treatment­
related
effects.
Christian
et
al.
(
2001b)
also
analyzed
the
distribution
of
DBA
in
various
tissues.
At
high
drinking­
water
concentrations,
DBA
was
detected
in
the
placenta,
amniotic
fluid,
and
fetal
plasma,
indicating
that
it
readily
crossed
the
placenta
and
distributed
to
the
fetus,
although
it
did
not
appear
to
bioaccumulate.
In
the
two­
generation
reproductive
toxicity
study
(
CCC,
2001),
a
dose­
related
increase
in
altered
sperm
142
histomorphology
and
histopathology
of
the
male
reproductive
tract
was
observed
in
both
the
parental
and
F
1
generations,
indicating
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
F
1
generation,
and
thus
developing
males
did
not
appear
to
be
more
sensitive
than
adults.

In
contrast,
in
two
published
abstracts,
(
Narotsky
et
al.,
1996;
1997b)
DBA
was
reported
to
induce
developmental
toxicity
when
administered
by
gavage
to
CD­
1
mice.
Differences
in
the
results
between
these
studies
and
those
of
Christian
et
al.
(
2001b)
and
the
Chlorine
Chemistry
Council
(
2001)
may
have
been
due
to
differences
in
the
route
of
dosing
(
gavage
versus
drinking
water),
species
sensitivity
(
mouse
versus
rat),
dosing
regime,
or
other
study
details
not
reported
in
the
published
abstracts.
These
data
are
supported
by
recent
reports
by
Klinefelter
et
al.
(
2001;

manuscript
in
preparation),
that
reported
a
DBA­
related
delay
in
sexual
maturity
and
effects
on
sperm
quality
in
Sprague­
Dawley
rats
exposed
to
DBA
in
drinking
water.
These
effects
were
independent
of
changes
in
body
weight,
and
were
seen
when
exposure
was
from
gestation
through
puberty,
but
not
when
exposure
only
occurred
from
weaning
through
puberty.
Pubertal
delays
occurred
at
doses
as
low
as
400
ppm
DBA
in
drinking
water
(
approximately
40
mg/
kg/
day),
while
changes
in
sperm
membrane
proteins
were
observed
at
doses
as
low
as
4
ppm
(
0.4
mg/
kg/
day).
The
risk
assessment
implications
of
this
study
are
currently
being
evaluated.

Veeramachaneni
et
al.
(
2000)
reported
in
an
abstract
that
exposure
of
rabbits
in
utero
from
gestation
day
15
to
24
weeks
of
age
reduced
the
fertility
of
sperm
from
treated
males.
The
lowest
dose
tested,
0.97
mg/
kg/
day
was
the
LOAEL.
This
LOAEL
for
fertility
changes
was
10­

fold
lower
than
the
LOAEL
of
10
mg/
kg/
day
reported
in
Linder
et
al.
(
1997b)
for
altered
histopathology.
143
2.4.
BROMATE
Bromate
(
BrO
3
­)
is
formed
in
water
following
disinfection
through
ozonation
of
water
containing
bromide
ion.
In
laboratory
studies,
the
rate
and
extent
of
bromate
formation
depends
on
the
ozone
concentration
used
in
disinfection,
pH
and
contact
time.

2.4.1.
Developmental/
Reproductive
Effects
Limited
data
are
available
on
the
reproductive
or
developmental
effects
of
bromate.
No
standard
reproductive,
developmental,
or
multigeneration
studies
are
available
for
bromate.

Kurokawa
et
al.
(
1990)
reported
the
results
of
a
5­
generation
study
in
mice
and
an
8­
generation
study
in
mice,
in
which
the
animals
were
fed
bread
made
from
flour
treated
with
14­
100
ppm
potassium
bromate;
information
on
the
study
designs
was
not
presented.
No
effects
on
behavior,

weight
gain,
reproductive
performance
or
histological
abnormalities
were
reported
(
Kurokawa
et
al.,
1990).
However,
since
most
potassium
bromate
added
to
flour
is
converted
to
bromide
during
the
bread­
baking
process
(
Kurokawa
et
al.,
1986b),
it
is
unlikely
that
the
animals
in
these
multigeneration
studies
were
actually
exposed
to
bromate.

In
a
screening
study
conducted
for
NTP,
Wolf
and
Kaiser
(
1996)
evaluated
the
potential
reproductive
and
developmental
toxicity
of
sodium
bromate
in
Sprague­
Dawley
rats
following
oral
administration
in
the
drinking
water
at
concentrations
of
0.25
ppm
(
2.6
mg/
kg/
day),
80
ppm
(
9.0
mg/
kg/
day),
or
250
ppm
(
25.6
mg/
kg/
day)
over
a
35­
day
period.
(
Equivalent
bromate
ion
doses
are
2.2,
7.7,
and
22
mg
BrO
3­/
kg/
day.)
Two
groups
of
female
rats
were
treated.
Group
1
females
(
10/
group)
were
dosed
from
study
day
1
to
34
to
test
for
effects
during
conception
and
144
early
gestation.
Group
2
females
(
13/
group)
were
dosed
from
gestation
day
6
to
postnatal
day
1
to
test
for
effects
during
late
gestation
and
birth.
Male
rats
(
10/
group)
were
cohabited
with
Group
2
females
for
5
days
prior
to
dosing
(
study
days
1­
5)
and
were
then
dosed
from
study
day
6
to
day
34/
35.
Females
in
Group
2
were
allowed
to
litter
and
the
pups
were
observed
through
postnatal
day
5.
However,
there
is
no
indication
of
the
developmental
endpoints
that
were
evaluated
in
these
pups
or
if
any
effects
were
observed.
Treated
males
in
the
250
ppm
dose
group
demonstrated
a
statistically
significant
decrease
(
18%)
in
epididymal
sperm
density.
All
other
endpoints
evaluated
were
comparable
between
controls
and
treated
groups.
Female
reproductive
function
was
not
adversely
affected.
There
were
no
treatment­
related
gross
or
microscopic
changes
in
the
kidney,
liver,
spleen,
testis
or
epididymis.
These
results
indicated
that
sodium
bromate
treatment
did
not
produce
any
adverse
signs
of
general
toxicity
in
any
of
the
dose
levels
tested;
a
maximum
tolerated
dose
(
MTD)
was
not
reached.
Based
on
changes
in
sperm
density,

this
study
identified
a
NOAEL
of
80
ppm
(
7.7
mg
BrO
3
­/
kg/
day)
and
a
LOAEL
of
250
ppm
(
22
mg/
BrO
3
­/
kg/
day).

2.4.2.
Systemic
Toxicity
A
number
of
cases
of
acute
bromate
intoxication
have
been
reported
in
humans
(
both
children
and
adults)
following
accidental
or
suicidal
ingestion
of
permanent
hair­
wave
neutralizing
solutions,
which
usually
contain
either
2%
potassium
bromate
or
10%
sodium
bromate.
The
case
studies
suggest
that
there
are
no
differences
between
children
and
adults
in
either
the
target
organ
or
the
effective
dose
following
acute
oral
exposure
to
bromate.
No
epidemiological
studies
were
located
on
noncarcinogenic
or
carcinogenic
effects
of
bromate
exposure
in
humans.
145
Several
authors
report
the
effects
of
acute
oral
exposure
in
children
to
potassium
bromate
following
accidental
ingestion
of
hair
home
permanent
neutralizing
solution
(
Benson,
1951;

Parker
and
Barr,
1951;
Quick
et
al.,
1975;
Gradus
et
al.,
1984;
Warshaw
et
al.,
1985;
Lue
et
al.,

1988;
Mack,
1988;
Lichtenberg
et
al.,
1989;
Watanabe
et
al.,
1992).
The
age
of
the
children,

when
reported,
ranged
from
17
months
(
Gradus
et
al.,
1984)
to
6
years
(
Quick
et
al.,
1975).

When
estimated,
doses
ranged
from
20
mg
BrO
3­/
kg
(
Watanabe
et
al.,
1992)
to
1,000
mg
BrO
3­

/
kg
(
Lue
et
al.,
1988).
In
all
cases,
the
initial
symptoms
appeared
to
include
abdominal
pain,

vomiting
or
other
gastrointestinal
effects.
Central
nervous
system
(
CNS)
effects
such
as
sedation,

lethargy,
and
CNS
depression
appeared
to
be
early
symptoms
of
bromate
poisoning
after
doses
of
about
70
mg/
kg
or
higher
(
Parker
and
Barr,
1951;
Warshaw
et
al.,
1985;
Lue
et
al.,
1988;

Lichtenberg
et
al.,
1989).
Irreversible
deafness
is
also
an
effect
of
bromate
exposure
(
Quick
et
al.,

1975;
Gradus
et
al.,
1984);
one
review
of
bromate
ototoxicity
found
that
deafness
occurred
in
18
of
31
cases,
usually
within
4­
16
hours
of
exposure
(
Matsumoto
et
al.,
1980).

Kidney
effects
were
frequently
observed
in
children
following
acute
exposure;
although
there
is
not
a
clear
relationship
between
dose
and
the
development
of
renal
effects.
One
review
of
bromate
kidney
toxicity
found
that
renal
failure
occurred
in
26
of
31
reported
cases
(
Matsumoto
et
al.,
1980).
Anuria
persisting
for
several
days
or
longer
was
observed
following
exposure
to
20
mg/
kg
potassium
bromate
(
Quick
et
al.,
1975)
up
to
doses
of
1,000
mg
BrO
3­/
kg
(
Lue
et
al.,
1988).
In
contrast,
children
ingesting
20
mg
BrO
3­/
kg
(
Watanabe
et
al.,
1992)
and
children
ingesting
230­
460
mg
BrO
3­/
kg
(
Lichtenberg
et
al.,
1989)
did
not
demonstrate
any
renal
effects.
Histological
examination
of
renal
biopsies
from
children
with
renal
effects
indicated
interstitial
edema,
interstitial
fibrosis,
tubular
atrophy
(
Quick
et
al.,
1975),
and
epithelial
146
separation
of
the
proximal
tubules
(
Watanabe
et
al.,
1992).
Glomeruli
were
not
affected.

Although
there
are
fewer
reports
of
acute
oral
exposure
to
bromate
in
adults
(
Matsumoto
et
al.,
1980;
Kuwahara
et
al.,
1984;
Kutom
et
al.,
1990;
Hamada
et
al.,
1990),
the
symptoms
of
toxicity
appear
to
be
similar
to
those
observed
in
children.
When
reported,
the
doses
ingested
ranged
from
100­
150
mg
BrO
3­/
kg
(
Matsumoto
et
al.,
1980)
to
500
mg
K
BrO
3/
kg
(
Kuwahara
et
al.,
1984).
In
all
cases,
the
first
symptoms
to
appear
were
gastrointestinal,
including
nausea,

vomiting,
diarrhea
and
abdominal
pain.
Hearing
loss
was
reported
by
three
authors
(
Matsumoto
et
al.,
1980;
Kuwahara
et
al.,
1984;
Hamada
et
al.,
1990).
Anuria
and
renal
failure
were
also
reported
(
Kuwahara
et
al.,
1984;
Kutom
et
al.,
1990;
Hamada
et
al.,
1990).
The
amount
of
time
required
to
recover
renal
function
varied
from
7
days
(
Kutom
et
al.,
1990)
to
5
weeks
(
Hamada
et
al.,
1990),
and
in
two
cases,
renal
function
was
never
restored
(
Kuwahara
et
al.,
1984).

Several
subchronic
or
chronic
studies
in
animals
indicate
that
the
kidney
is
the
primary
target
organ
following
long­
term
oral
exposure
to
bromate.
Following
a
13­
week
exposure
of
rats
to
a
dose
of
63
mg
BrO
3­/
kg/
day
(
Kurokawa
et
al.,
1990),
the
following
non­
neoplastic
effects
were
observed:
inhibition
of
body­
weight
gain;
significant
increases
in
several
serum
parameters,
including
blood
urea
nitrogen
(
BUN);
and
droplets
of
various
sizes
and
regenerative
changes
in
the
renal
tubules.
Similar
effects
were
observed
in
chronic
studies
of
oral
bromate
exposure
(
Nakano
et
al.,
1989;
Kurokawa
et
al.,
1986b;
DeAngelo
et
al.,
1998).
The
following
non­
neoplastic
effects
have
been
reported
following
long­
term
exposure:
increased
BUN;

increased
severity
of
nephropathic
changes;
degenerative
and
necrotic
kidney
lesions
including
hyaline
casts
in
the
tubular
lumen,
hyaline
droplets,
eosinophilic
bodies,
and
brown
pigments
in
the
tubular
epithelium;
and
urothelial
hyperplasia
of
the
transitional
epithelium
of
the
renal
pelvis.
147
No
non­
neoplastic
effects
have
been
reported
in
tissues
other
than
the
kidney.

DeAngelo
et
al.
(
1998)
observed
statistically
significant
increases
in
relative
liver
weight,

absolute
and
relative
kidney
weight,
absolute
and
relative
thyroid
weight,
and
relative
spleen
weight
in
rats
treated
with
28.7
mg/
kg/
day
in
drinking
water.
Non­
neoplastic
kidney
lesions
observed
in
rats
included
a
significant
dose­
dependent
increase
in
the
incidence
of
urothelial
hyperplasia
at
doses
of
6.1
mg/
kg/
day
and
higher,
foci
of
mineralization
of
the
renal
papilla
and
eosinophilic
droplets
in
the
proximal
tubule
epithelium.
There
were
no
other
treatment­
related
non­
neoplastic
effects
observed
in
any
other
tissue
examined.
Based
on
urothelial
hyperplasia
in
male
rats,
this
study
identifies
a
NOAEL
of
1.1
mg
BrO
3­/
kg/
day
and
a
LOAEL
of
6.1
mg
BrO
3­/
kg/
day.

2.4.3.
Carcinogenicity
There
are
no
epidemiology
data
regarding
the
carcinogenic
potential
of
bromate.

DeAngelo
et
al.
(
1998)
administered
potassium
bromate
to
male
rats
and
male
mice
in
drinking
water
for
100
weeks.
Statistically
significant,
dose­
dependent
increases
in
tumors
of
the
kidney,

thyroid,
and
tunica
vaginalis
testis
(
mesotheliomas)
were
observed.
This
study
contributes
to
the
weight
of
evidence
for
the
potential
human
carcinogenicity
of
bromate
and
confirms
the
study
by
Kurokawa
et
al.
(
1986a,
b),
which
reported
renal
tumors
in
male
rats.

The
evidence
is
too
limited
to
reach
a
conclusion
about
any
mode
of
action
(
EPA,
2001b).

The
genotoxicity
of
bromate
has
been
evaluated
in
a
variety
of
in
vitro
and
in
vivo
systems,
with
consistently
positive
results.
Oxidative
stress
may
play
a
role
in
the
formation
of
kidney
tumors
induced
by
bromate.
Two
mechanisms
of
bromate­
mediated
DNA
damage
have
been
proposed:
148
direct
interaction
with
DNA
following
GSH
activation,
and
indirect
damage
via
lipid
peroxides.

The
data
suggest
that,
in
the
intact
kidney,
bromate
induces
DNA
damage
through
lipid
peroxidation
at
toxic
doses,
rather
than
via
a
direct
mechanism
(
Chipman
et
al.,
1998).
However,

the
overall
evidence
is
insufficient
to
establish
lipid
peroxidation
and
free
radical
production
as
the
key
events
responsible
for
the
induction
of
kidney
tumors,
and
the
data
are
insufficient
to
implicate
any
single
mechanism
for
the
production
of
thyroid
and
testicular
tumors
(
EPA,
2001b).

Some
evidence
suggests
that
cell
proliferation
related
to
 2u­
globulin
plays
a
role
in
enhancing
renal
carcinogenesis
by
bromate
(
Umemura
et
al.,
1993),
but
the
observation
of
kidney
tumors
in
female
rats,
and
the
observation
of
testicular
and
thyroid
tumors
in
male
rats
suggests
that
2uglobulin
is
not
the
primary
mechanism
of
bromate
carcinogenicity
(
EPA,
2001b).

Under
EPA's
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1986),
bromate
would
be
classified
as
Group
B2:
Probable
Human
Carcinogen
based
on
no
evidence
in
humans
and
adequate
evidence
of
carcinogenicity
in
male
and
female
rats
(
EPA,
2001b).
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).
Insufficient
data
are
available
to
evaluate
the
human
carcinogenic
potential
of
bromate
by
the
inhalation
route.
Although
no
epidemiological
studies
or
studies
of
long­
term
human
exposure
to
bromate
are
available,
bromate
is
carcinogenic
to
male
and
female
rats
following
exposure
in
drinking
water.
This
weight­

ofevidence
conclusion
is
based
on
sufficient
experimental
findings
that
include
the
following:
tumors
at
multiple
sites
in
rats,
tumor
responses
in
both
sexes,
and
evidence
of
mutagenicity
including
point
mutations
and
chromosomal
aberrations
in
vitro.

Cancer
risk
estimates
were
derived
from
the
DeAngelo
et
al.
(
1998)
study
by
applying
the
149
one­
stage
Weibull
model
for
the
low­
dose
linear
extrapolation.
Bromate
was
administered
to
male
F344
rats
or
B6C3F1
mice
in
drinking
water
at
concentrations
of
0,
0.02,
0.1,
0.2,
and
0.4
g/
L
or
0,
0.08,
0.4,
and
0.8
g/
L,
respectively,
for
108
weeks.
Using
a
time­
to­
tumor
model
and
a
Monte
Carlo
analysis
to
sum
across
the
three
tumor
sites,
the
upper­
bound
cancer
potency
for
bromate
ion
is
estimated
to
be
0.7
per
mg/
kg/
day.
Assuming
a
daily
water
consumption
of
2
L
for
a
70
kg
adult,
lifetime
risks
of
10­
4,
10­
5,
and
10­
6
are
associated
with
bromate
concentrations
in
water
of
5,
0.5,
and
0.05
µ
g/
L,
respectively.
This
estimate
of
cancer
risk
from
the
DeAngelo
et
al.
study
is
similar
to
the
risk
estimate
derived
from
the
Kurokawa
et
al.
(
1986a)
study
in
the
1994
proposed
rule
(
EPA,
1994a).

2.4.4.
Basis
for
the
RfD
and
MCLG
The
Agency
based
the
bromate
RfD
on
urothelia
hyperplasia
in
male
rats
in
the
DeAngelo
et
al.
(
1998)
study,
with
a
NOAEL
of
1.1
mg
BrO
3­/
kg/
day
and
a
LOAEL
of
6.1
mg
BrO
3­/
kg/
day
(
EPA,
2001b).
Kidney
effects
have
also
been
seen
in
children
and
adults
following
accidental
high­
dose
exposure,
indicating
animal
to
human
concordance
in
target.
An
UF
of
10
was
applied
for
extrapolating
from
animals
to
humans,
and
another
factor
of
10
was
used
for
variability
in
sensitivity
among
members
of
the
human
population.
A
factor
of
3
was
used
to
account
for
some
deficiencies
in
the
database.
The
bromate
database
consists
of
chronic
and
subchronic
studies
in
rats
and
mice
and
a
screening
level
reproductive/
developmental
study
in
rats.

The
database
is
missing
developmental
toxicity
studies
in
two
species
and
a
standard
multigenerational
study.
This
results
in
a
total
UF
of
300.
The
resulting
RfD
for
bromate
is
0.004
150
mg/
kg/
day.
This
RfD
corresponds
to
a
DWEL
of
0.14
mg/
L,
assuming
an
adult
tap
water
consumption
of
2
L/
day
for
a
70
kg
adult.

The
MCLG
for
bromoform
is
zero,
based
on
its
likely
human
carcinogenicity.

2.4.5.
Children's
Risk
in
Relation
to
the
MCLG
The
developmental
toxicity
of
bromate
has
not
been
adequately
evaluated,
so
no
conclusions
regarding
the
sensitivity
of
the
fetus
can
be
drawn.
Limited
evidence
suggests
that
bromate
may
be
a
male
reproductive
toxicant,
although
at
a
higher
dose
than
that
which
results
in
kidney
toxicity.
No
data
were
identified
that
describe
the
effects
of
in
utero
or
neonatal
exposure
to
bromate.
Case
reports
on
the
effects
of
acute
oral
exposure
to
bromate
suggest
that
systemic
toxicity
is
similar
in
children
and
adults.
No
data
were
identified
regarding
age­
related
differences
in
absorption,
distribution,
metabolism
or
excretion
of
bromate.
The
bromate
MCLG
is
zero,

based
on
its
likely
human
carcinogenicity.
The
mode
of
action
for
bromate
is
unclear;
however,

evidence
of
genotoxicity
led
to
use
of
a
linear
low­
dose
extrapolation.
The
Agency
believes
that,

since
there
are
no
data
suggesting
that
children
are
more
sensitive
to
bromate,
the
MCLG
of
zero
is
protective
of
both
children
and
adults.

2.5.
CHLORITE/
CHLORINE
DIOXIDE
Chlorite
and
chlorine
dioxide
are
evaluated
together
in
this
assessment
for
children's
risk.

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.
In
addition,

studies
conducted
with
chlorine
dioxide
may
be
relevant
to
characterizing
the
toxicity
of
chlorite.
151
Chlorine
dioxide
is
fairly
unstable
and
rapidly
dissociates
to
form
predominantly
chlorite
and
chloride,
and
to
a
lesser
extent,
chlorate.
There
is
a
ready
interconversion
among
these
species
in
water
(
before
administration
to
animals)
and
in
the
gut
(
after
ingestion)
(
EPA,
1994c).

Therefore,
what
exists
in
water
or
stomach
is
a
mixture
of
these
chemical
species
(
i.
e.,
chlorine
dioxide,
chlorite
and
chlorate),
and
possibly
the
products
of
their
reactions
with
the
gastrointestinal
contents.
As
a
result,
the
toxicity
data
for
chlorine
dioxide
or
chlorite
are
considered
applicable
for
assessing
the
toxicity
of
the
other.
In
the
study
descriptions
below,

concentrations
in
drinking
water
are
provided
as
reported
by
the
study
authors,
but
all
doses
are
converted
to
concentrations
of
chlorite
or
chlorine
dioxide.

2.5.1.
Developmental/
Reproductive
Effects
There
is
a
large
database
on
the
developmental
and
reproductive
toxicity
of
chlorite
and
chlorine
dioxide.
Two
epidemiologic
studies
examined
the
relationship
between
water
disinfected
with
chlorine
dioxide
and
developmental
toxicity
(
Kanitz
et
al.,
1996;
Tuthill
et
al.,
1982).
Tuthill
et
al.
(
1982)
retrospectively
compared
infant
morbidity
and
mortality
data
for
a
community
using
chlorine
dioxide
as
a
drinking­
water
disinfectant
in
the
1940s
with
a
community
using
conventional
drinking­
water
chlorination
treatment.
No
effect
was
observed
on
fetal,
neonatal,
postneonatal
or
infant
mortality;
however,
the
incidence
of
newborns
judged
premature
by
physician
assessment
was
significantly
higher
in
the
chlorine
dioxide­
treated
community.
EPA
(
1994d)
concluded
that
there
was
no
increase
in
the
proportion
of
premature
infants
when
the
age
of
the
mother
was
controlled
and
that
there
was
a
greater
postnatal
weight
loss
in
infants
from
the
exposed
community.
152
As
summarized
in
Section
2.1
on
Chlorinated
Drinking
Water,
Kanitz
et
al.
(
1996)

conducted
an
analysis
of
676
births
from
hospital
records
of
two
Italian
cities
that
used
chlorinated
surface
water,
surface
water
treated
with
chlorine
dioxide,
or
well
water
that
was
not
treated.
The
authors
found
that
the
frequency
of
newborns
with
short
body
length,
small
cranial
circumference,
and
low
birth
weight
was
increased
for
mothers
over
30
years
of
age
who
were
served
by
chlorinated
water
supplies.
The
frequency
of
newborns
with
neonatal
jaundice
was
increased
for
births
to
mothers
exposed
to
water
treated
with
chlorine
dioxide.

Six
oral
studies
have
examined
the
developmental
toxicity
of
chlorite
in
animals,
including
four
studies
in
rats
(
CMA,
1996;
Couri
et
al.,
1982;
Mobley
et
al.,
1990;
Suh
et
al.,
1983)
and
one
study
each
in
mice
(
Moore
et
al.,
1980)
and
rabbits
(
Harrington
et
al.,
1995b).
The
Chemical
Manufacturers
Association
(
CMA)
conducted
a
two­
generation
study
examining
the
developmental
neurotoxicity,
reproductive,
and
hematologic
effects
of
rats
exposed
to
sodium
chlorite
(
CMA,
1996).
Males
were
exposed
throughout
mating,
and
the
females
were
exposed
during
mating,
pregnancy,
and
lactation.
Thirty
male
and
30
female
Sprague­
Dawley
rats
received
drinking
water
containing
0,
35,
70,
or
300
mg/
L
sodium
chlorite
for
10
weeks
and
were
then
mated.
Based
on
intakes
calculated
by
the
study
authors,
doses
for
the
P
animals
were
0,
3,

5.6,
or
20,
and
0,
3.8,
7.5,
or
28.6
mg/
kg/
day
chlorite
for
males
and
females,
respectively.
For
the
F
1
animals,
doses
were
0,
2.9,
5.9,
or
22.7
mg/
kg/
day
chlorite
for
the
males
and
0,
3.8,
7.9,
or
28.6
mg/
kg/
day
chlorite
for
the
females.
At
300
mg/
L,
effects
observed
in
at
least
one
generation
and
sex
included
decreased
pup
survival
and
pup
body
weight,
decreased
absolute
and
relative
liver
weight,
delayed
sexual
development,
decreased
red
blood
cell
parameters,
and
several
neurotoxic
effects
(
e.
g.,
decreased
brain
weight
and
increased
incidence
of
abnormal
righting
153
reflex).
Effects
at
70
mg/
L
included
decreased
absolute
and
relative
liver
weights.
A
significant
decrease
in
maximum
response
to
an
auditory
startle
stimulus
was
noted
in
the
70
and
300
mg/
L
groups
on
postnatal
day
24,
but
not
on
postnatal
day
60.
The
NOAEL
in
this
study
was
35
mg/
L
(
2.9
mg/
kg/
day
chlorite)
and
the
LOAEL
was
70
mg/
L
(
5.9
mg/
kg/
day
chlorite)
based
on
lowered
auditory
startle
amplitude
and
altered
liver
weights
in
two
generations
of
rats.

Mobley
et
al.
(
1990)
exposed
groups
of
12
female
Sprague­
Dawley
rats
to
0,
20
or
40
mg/
L
chlorite
in
drinking
water
(
estimated
as
0,
3,
or
6
mg/
kg/
day)
for
10
days
prior
to
mating
with
unexposed
males,
and
during
gestation
and
lactation
until
postnatal
days
42­
53.
Significant
decreases
in
exploratory
activity
were
observed
in
the
6
mg/
kg/
day
group
on
postnatal
days
36­
39
but
not
on
days
40­
41.
In
the
3
mg/
kg/
day
group,
significant
decreases
in
exploratory
activity
were
observed
on
days
36
and
37
but
not
on
days
38­
41.
Free
T4
was
significantly
increased
in
pups
at
6
mg/
kg/
day.
Based
on
a
review
of
the
results
of
this
study
relative
to
the
findings
of
the
newer
developmental
studies
in
the
chlorite
database,
EPA
(
2000c)
concluded
that
the
NOAEL
for
neurobehavioral
effects
in
this
study
was
3
mg/
kg/
day
and
the
LOAEL
was
6
mg/
kg/
day.

Suh
et
al.
(
1983)
exposed
groups
of
six
to
nine
female
Sprague­
Dawley
rats
to
0,
1,
or
10
mg/
L
chlorite
in
drinking
water
(
estimated
as
0,
0.1,
or
1
mg/
kg/
day)
for
2.5
months
prior
to
mating
with
unexposed
males
and
during
gestational
days
0­
20.
No
significant
alterations
in
maternal
body
weight
gain,
number
of
implants,
resorptions,
dead
fetuses,
fetal
body
weights,
or
incidence
of
skeletal
abnormalities
were
observed
at
any
dose.
Crown­
rump
length
was
significantly
higher
in
the
1
mg/
kg/
day
group
compared
with
controls,
but
the
difference
was
small
and
probably
not
biologically
significant.
A
NOAEL
of
1
mg/
kg/
day
for
developmental
toxicity
was
identified
in
this
study.
154
In
another
developmental
study,
Couri
et
al.
(
1982)
exposed
groups
of
7­
13
pregnant
Sprague­
Dawley
rats
to
0%,
0.1%,
0.5%,
or
2%
sodium
chlorite
(
0%,
0.07%,
0.4%,
or
1.5%

chlorite;
estimated
as
0,
70,
440,
or
610
mg/
kg/
day
chlorite)
in
drinking
water
during
gestational
days
8­
15.
Another
group
of
four
pregnant
rats
received
daily
gavage
doses
of
200
mg/
kg/
day
sodium
chlorite
on
gestational
days
8­
15.
One
hundred
percent
mortality
was
observed
in
the
rats
receiving
this
gavage
dose,
while
no
deaths
were
observed
in
the
rats
received
sodium
chlorite
via
drinking
water.
In
the
rats
receiving
sodium
chlorite
via
drinking
water,
irregular
blood
cells,

ruptured
cells
and
hemolysis
were
observed
in
the
610
mg/
kg/
day
group.
Significant
decreases
in
crown­
rump
length
and
an
increase
in
the
number
of
resorbed
and
dead
fetuses
in
litters
delivered
on
gestational
day
22
were
observed
in
all
dose
groups.
Postnatal
growth
and
the
incidences
of
soft
tissue
and
skeletal
malformations
were
not
adversely
affected
in
this
study.
This
study
identified
a
frank
effect
level
(
FEL)
of
70
mg/
kg/
day
for
resorbed
and
dead
fetuses
and
decreases
in
crown­
rump
length.

Moore
et
al.
(
1980)
treated
groups
of
pregnant
A/
J
mice
with
0
or
100
mg/
L
sodium
chlorite
in
drinking
water
(
75
mg/
L
chlorite,
corresponding
to
approximately
22
mg/
kg/
day
chlorite)
throughout
gestation
and
lactation.
A
decrease
in
the
conception
rate
was
observed
in
the
chlorite
group.
Pup
growth
was
adversely
affected,
as
shown
by
significant
decreases
in
average
pup
weaning
weight
and
birth
to
weaning
growth
rate.
This
study
identified
a
LOAEL
of
22
mg/
kg/
day
for
developmental
effects.

Harrington
et
al.
(
1995b)
exposed
groups
of
16
New
Zealand
white
rabbits
to
sodium
chlorite
in
drinking
water
at
concentrations
of
0,
200,
600,
or
1,200
mg/
L
(
0,
10,
26
or
40
mg/
kg/
day
chlorite)
on
gestation
days
7­
20.
No
treatment­
related
effects
on
pregnancy
incidence,
155
number
of
malformations,
number
of
preimplantation
losses,
fetal
sex
ratio,
number
of
live
fetuses
or
fetal
visceral
or
structural
abnormalities
were
observed
at
any
dose.
In
the
26
and
40
mg/
kg/
day
groups,
an
increased
incidence
of
minor
skeletal
abnormalities
and
skeletal
variants
was
noted.
The
NOAEL
in
this
study
was
10
mg/
kg/
day
and
the
LOAEL
was
26
mg/
kg/
day,

based
on
decreased
fetal
weight
and
delayed
skeletal
ossification,
and
decreased
food
and
water
consumption
and
decreased
body
weight
gain
in
the
dams.

Several
studies
investigated
the
reproductive
toxicity
of
chlorite
in
drinking
water.
Carlton
and
Smith
(
1985)
and
Carlton
et
al.
(
1987)
exposed
groups
of
12
male
rats
to
sodium
chlorite
in
drinking
water
for
56
days
prior
to
mating
and
throughout
a
10­
day
mating
period;
they
also
exposed
groups
of
24
female
rats
to
the
same
sodium
chlorite
concentrations
for
14
days
prior
to
mating
and
throughout
gestation
and
lactation.
No
significant
alterations
in
body
weight
gain,

fertility
rates,
litter­
survival
rates,
median
day
of
eye
opening,
or
median
day
of
observed
vaginal
patency
were
noted.
No
alterations
were
seen
in
the
reproductive
tract
tissues,
hematologic
parameters,
testis,
epididymis,
caudal
epididymis
weights,
or
percentage
of
abnormal
sperm.
The
only
effects
noted
were
significant
decreases
in
T3
and
T4
hormone
levels.
In
follow­
up
studies,

groups
of
12
male
rats
received
sodium
chlorite
in
drinking
water
for
72­
76
days
and
a
statistically
significant
increase
in
abnormal
sperm
at
100
mg/
L
sodium
chlorite
(
70
mg/
L
chlorite,

or
approximately
7.5
mg/
kg/
day)
was
noted;
the
NOAEL
was
0.75
mg/
kg/
day
chlorite.

Reproductive
function
was
not
assessed
in
this
study.

Five
oral
studies
investigated
the
developmental
toxicity
of
chlorine
dioxide
in
rats
(
Mobley
et
al.,
1990;
Orme
et
al.,
1985;
Suh
et
al.,
1983;
Taylor
and
Pfohl,
1985;
Toth
et
al.,

1990).
Several
of
these
studies
tested
only
one
dose.
Orme
et
al.
(
1985)
exposed
groups
of
156
female
Sprague­
Dawley
rats
to
0,
2,
20
or
100
mg/
L
chlorine
dioxide
in
drinking
water
(
estimated
at
0,
1,
3
or
14
mg/
kg/
day)
for
2
weeks
prior
to
mating
and
throughout
gestation
and
lactation.

Additionally,
groups
of
5­
day­
old
Sprague­
Dawley
pups
received
gavage
doses
of
0
or
14
mg/
kg/
day
chlorine
dioxide
on
postnatal
days
5­
20.
In
the
14
mg/
kg/
day
gavage
group,
activity
was
significantly
decreased
on
postnatal
days
18­
19;
on
days
15­
17
and
20,
activity
levels
were
similar
to
controls.
In
the
14
mg/
kg/
day
drinking­
water
group,
there
was
a
significant
decrease
in
T3
and
T4
levels;
in
all
groups
there
was
a
significant
correlation
between
T4
levels
and
locomotor
activity.
T4
levels
were
not
significantly
altered
in
the
chlorine
dioxide­
exposed
dams.

A
NOAEL
of
3
mg/
kg/
day
and
a
LOAEL
of
14
mg/
kg/
day
for
neurobehavioral
effects
(
decreased
T3
and
T4
levels
and
delayed
development)
were
identified
in
this
study.

Suh
et
al.
(
1983)
exposed
groups
of
six
to
eight
female
Sprague­
Dawley
rats
to
0,
1,
10,

or
100
mg/
L
chlorine
dioxide
in
drinking
water
(
approximately
0,
0.1,
1,
or
10
mg/
kg/
day)
for
2.5
months
prior
to
mating
with
unexposed
males
and
during
gestational
days
0­
20.
There
was
a
statistically
significant
trend
for
decreasing
number
of
implants
per
litter
and
number
of
live
fetuses
per
dam.
Total
fetal
weights
and
male
fetal
weights
were
significantly
increased
in
the
10
mg/
kg/
day
group;
crown­
rump
length
was
not
significantly
affected.
Incidences
of
skeletal
abnormalities
did
not
significantly
differ
between
groups.
This
study
identified
a
NOAEL
of
1
mg/
kg/
day
and
a
LOAEL
of
10
mg/
kg/
day
for
developmental
effects.

Mobley
et
al.
(
1990)
exposed
groups
of
12
female
Sprague­
Dawley
rats
to
0
or
100
mg/
L
chlorine
dioxide
in
drinking
water
(
estimated
at
0
or
14
mg/
kg/
day)
for
10
days
prior
to
mating
with
unexposed
males
and
during
the
gestation
and
lactation
periods
(
through
postconception
day
41).
At
birth,
the
litter
weight
of
the
chlorine
dioxide­
exposed
group
was
significantly
lower
than
157
that
of
controls.
Chlorine
dioxide
exposure
also
significantly
decreased
exploratory
activity
on
postconception
days
36­
39,
but
not
on
days
40­
41.
This
study
identified
a
LOAEL
of
14
mg/
kg/
day
for
decreased
litter
weight
and
decreased
exploratory
activity.

In
another
developmental
study,
Taylor
and
Pfohl
(
1985)
exposed
groups
of
13­
16
female
Sprague­
Dawley
rats
to
0
or
100
mg/
L
chlorine
dioxide
in
drinking
water
(
estimated
at
0
or
14
mg/
kg/
day)
for
14
days
prior
to
breeding
and
throughout
gestation
and
lactation.
In
another
phase,
groups
of
male
pups
from
unexposed
dams
were
administered
0
or
14
mg/
kg/
day
chlorine
dioxide
via
gavage
from
postnatal
days
5
to
20.
No
effects
on
maternal
or
pup
body
weights
were
observed
in
the
exposed
group,
but
a
significant
decrease
in
whole­
brain
weight
was
observed
in
the
21­
day­
old
offspring
of
dams
receiving
14
mg/
kg/
day
in
drinking
water.
A
nonsignificant
decrease
in
locomotor
activity
was
observed
in
the
14
mg/
kg/
day
offspring
assessed
at
10­
20
days
of
age,
and
a
significant
decrease
in
exploratory
activity
was
observed
at
this
dose
at
60
days
of
age.
In
the
gavage­
dose
group,
significant
decreases
in
body
weight,
whole­
brain
and
forebrain
weights,
forebrain
DNA
content
and
decreases
in
home
cage
and
wheel­
running
activity
were
observed
in
the
pups.
The
LOAEL
for
neurobehavioral
effects,
decreased
brain
weight
and
cell
number
in
this
study
was
14
mg/
kg/
day,
based
on
both
drinking
water
and
gavage
dosing.

In
a
neurodevelopmental
toxicity
study
by
Toth
et
al.
(
1990),
groups
of
four
male
and
four
female
pups
per
litter
received
gavage
doses
of
0
or
14
mg/
kg/
day
chlorine
dioxide
on
postnatal
days
1­
20.
Forebrain
weights
were
significantly
lower
in
the
chlorine
dioxide­
exposed
pups
on
postnatal
days
21
and
35.
No
gross
lesions,
and
loss
of
myelin
in
the
forebrain,
cerebellum
or
brainstem
were
observed
in
the
chlorine
dioxide­
exposed
pups.
Thus,
a
LOAEL
of
14
mg/
kg/
day
for
altered
brain
development
was
identified
in
this
study.
158
Reproductive
and
developmental
toxicity
of
chlorine
dioxide
were
studied
by
Carlton
et
al.

(
1991).
Daily
gavage
doses
of
0,
2.5,
5,
or
10
mg/
kg/
day
chlorine
dioxide
in
deoionized
water
were
administered
to
groups
of
12
male
and
24
female
Long­
Evans
rats
prior
to
mating
and
throughout
the
mating
period.
The
females
continued
to
receive
the
doses
throughout
gestation
and
lactation.
No
effects
were
noted
on
mortality,
clinical
signs,
fertility
rates,
sperm
parameters,

length
of
gestation,
prenatal
deaths,
mean
litter
size
or
mean
pup
weights.
A
NOAEL
of
10
mg/
kg/
day
for
reproductive
effects
was
noted
from
this
study.

The
observed
neurodevelopmental
effects
(
CMA,
1996;
Mobley
et
al.,
1990;
Orme
et
al.,

1985;
Taylor
and
Pfohl,
1985;
Toth
et
al.,
1990)
may
have
been
related
to
decreased
T3
and
T4
levels
(
Mobley
et
al.,
1990;
Orme
et
al.,
1985).
Hypothyroidism,
defined
by
increased
serum
levels
of
TSH
and
decreased
levels
of
T3
and
T4,
has
been
associated
with
neurodevelopmental
delay
in
both
humans
and
rats.
In
humans,
congenital
hypothyroidism,
or
cretinism,
is
characterized
by
long­
term
effects
on
behavior,
locomotor
ability,
speech,
hearing,
and
cognition
(
Chan
and
Kilby,
2000).
Fetuses
and
infants
are
more
sensitive
than
adults
to
hypothyroidism,

both
because
they
are
sensitive
to
small
changes
in
thyroid
hormone
levels,
and
because
the
resulting
effect
is
more
severe.
Prompt
supplementation
of
neonates
with
thyroid
hormone
can
restore
neurodevelopmental
function.

2.5.2.
Systemic
Toxicity
Several
human
studies
examined
the
short­
term
toxicity
of
chlorite
and
chlorine
dioxide
(
Lubbers
et
al.,
1981,
1982,
1984).
In
these
studies,
groups
of
10
healthy
adult
males
drank
solutions
providing
a
dose
of
0
or
0.034
mg/
kg
chlorite
or
0.34
mg/
kg
chlorine
dioxide
(
divided
159
into
2
portions,
separated
by
4
hours);
a
second
study
involved
the
ingestion
of
distilled
water
providing
0
or
0.04
mg/
kg/
day
chlorite
or
chlorine
dioxide
for
12
weeks.
Neither
of
these
studies
found
any
physiologically
relevant
alterations
in
general
health,
vital
signs,
hematologic
parameters
or
serum
T3
or
T4
levels.
No
data
were
identified
on
chlorite
or
chlorine
dioxideinduced
systemic
toxicity
in
children.

A
number
of
animal
studies
have
examined
the
subchronic
and
chronic
toxicity
of
chlorite
and
chlorine
dioxide;
however,
the
results
are
not
consistent
and
no
firm
conclusions
can
be
made
concerning
the
systemic
toxicity
of
these
compounds.
Harrington
et
al.
(
1995a)
reported
stomach
lesions
and
alterations
in
spleen
and
adrenal
weights
in
rats
administered
sodium
chlorite
via
gavage
for
13
weeks.
Haag
(
1949)
reported
treatment­
related
renal
pathology
in
rats
exposed
to
sodium
chlorite
in
drinking
water
for
2
years,
although
the
sensitivity
of
the
study
was
limited
by
the
use
of
a
small
number
of
animals.

Daniel
et
al.
(
1990)
reported
increases
in
nasal
lesions
in
rats
exposed
to
chlorine
dioxide
in
drinking
water
for
90
days,
but
it
is
not
known
whether
the
nasal
lesions
resulted
from
inhaling
chlorine
dioxide
vapors
at
the
drinking
water
tube
or
from
off­
gassing
of
the
vapors.
Haag
(
1949)

did
not
observe
renal
pathology
in
rats
exposed
to
chlorine
dioxide
in
drinking
water
for
2
years
at
slightly
higher
molar
concentrations
than
those
that
caused
renal
pathology
in
a
similar
study
(
Haag,
1949)
with
rats
exposed
to
chlorite.

Studies
on
the
hematological
effects
of
chlorite
have
shown
mixed
results.
Moore
and
Calabrese
(
1982)
found
an
increase
in
mean
corpuscular
volume
and
osmotic
fragility
in
mice
exposed
to
sodium
chlorite
in
drinking
water
for
30
days,
while
Bercz
et
al.
(
1982)
reported
decreases
in
erythrocyte
and
hemoglobin
levels,
and
CMA
(
1996)
reported
decreased
red
blood
160
cell
parameters
in
F
1
rats
at
the
high
dose.
Two
studies
reported
decreases
in
osmotic
fragility,

blood
glutathione
levels,
and
the
activity
of
erythrocyte
glutathione
peroxidase
and
blood
catalase
in
rats
exposed
to
chlorite
in
drinking
water
for
up
to
a
year
(
Abdel­
Rahman
et
al.,
1984a;
Couri
and
Abdel­
Rahman,
1980).
However,
a
consistent
dose­
response
relationship
was
not
observed
in
these
studies.

Studies
of
hematological
effects
of
chlorine
dioxide
generally
have
been
negative.
Three
studies
(
Bercz
et
al.,
1982;
Daniel
et
al.,
1990;
Moore
and
Calabrese,
1982)
found
no
hematologic
effects
in
monkeys,
rats,
or
mice
exposed
to
chlorine
dioxide
in
drinking
water
for
4­
6
weeks,
90
days,
or
30
days,
respectively.
Abdel­
Rahman
et
al.
(
1984a)
exposed
rats
to
chlorine
dioxide
in
drinking
water
for
up
to
11
months,
and
reported
increased
osmotic
fragility
and
decreased
hematocrit
and
hemoglobin
levels,
but
there
was
no
clear
dose
response,
and
the
effects
were
not
consistent
across
interim
measurement
periods.
Two
studies
(
Abdel­
Rahman
et
al.,
1984a;
Couri
and
Abdel­
Rahman,
1980)
reported
decreases
in
erythrocyte
glutathione
levels,
increases
in
glutathione
peroxidase
activity,
and
increases
in
erythrocyte
catalase
levels
in
rats
exposed
to
chlorine
dioxide
in
drinking
water
for
up
to
one
year.
However,
a
consistent
dose­
response
relationship
was
not
observed
in
these
studies.

2.5.3.
Carcinogenicity
There
are
no
studies
on
the
carcinogenicity
of
chlorite
in
humans.
There
is
one
animal
oral
carcinogenicity
study
on
chlorite
(
Kurokawa
et
al.,
1986a).
In
this
study,
rats
and
mice
were
exposed
to
sodium
chlorite
in
drinking
water
for
80
or
85
weeks
at
doses
up
to
32
mg/
kg/
day
for
male
rats,
41
mg/
kg/
day
for
female
rats,
and
71
mg/
kg/
day
for
male
and
female
mice.
Significant
161
increases
in
liver
and
lung
tumors
were
observed
in
the
male
mice,
but
these
incidences
were
within
the
range
of
historical
controls
in
the
laboratory.
The
mouse
study
is
considered
inadequate
for
assessing
carcinogenicity
because
of
the
relatively
short
duration
and
the
high
incidence
of
early
mortality
in
the
control
male
mice.
In
rats,
the
only
effect
was
a
slight
(<
10%)

dose­
related
decrease
in
body
weight
gain;
there
were
no
exposure­
related
increases
in
tumors
in
rats.

There
have
been
no
studies
in
humans
and
no
long­
term
oral
bioassays
on
the
carcinogenicity
of
chlorine
dioxide
in
animals.
In
a
short­
term
assay
(
Miller
et
al.,
1986),
partially
hepatectomized
rats
received
a
single
dose
of
concentrated
water
(
chlorine
dioxide
concentration
not
reported)
in
2%
emulphor
followed
1
week
later
by
administration
of
500
mg/
L
phenobarbital
in
drinking
water
for
56
days.
No
significant
increases
in
the
incidence
of
 ­

glutamyltranspeptidase­
positive
foci
(
a
measure
of
preneoplastic
change)
were
observed.

Chlorite
and
chlorine
dioxide
have
shown
mixed
results
in
mutagenicity
assays.
Both
chlorite
and
chlorine
dioxide
induced
reverse
mutations
in
S.
typhimurium
(
with
S9
activation)

(
Ishidate
et
al.,
1984).
However,
in
a
study
by
Miller
et
al.(
1986),
water
samples
disinfected
with
chlorine
dioxide
did
not
induce
reverse
mutations
in
S.
typhimurium
(
with
or
without
activation).

Chlorite
increased
the
incidence
of
chromosomal
aberrations
in
Chinese
hamster
fibroblast
cells,

while
chlorine
dioxide
did
not
(
Ishidate
et
al.,
1984).
The
results
for
chlorite
and
chlorine
dioxide
from
in
vivo
assays,
such
as
the
bone­
marrow
chromosomal
aberration
and
the
sperm­
head
abnormality
assays
in
mice,
are
primarily
negative
(
Hayashi
et
al.,
1988;
Meier
et
al.,
1985).

The
available
data
do
not
provide
sufficient
evidence
to
support
conclusions
as
to
the
carcinogenic
potential
of
chlorite
or
chlorine
dioxide.
Following
the
EPA's
1986
Guidelines
for
162
Carcinogen
Risk
Assessment,
chlorite
and
chlorine
dioxide
are
best
classified
as
Group
D:
Not
Classifiable
as
to
Human
Carcinogenicity.
This
classification
is
appropriate
because
no
data
were
identified
on
human
carcinogenicity
and
there
are
only
preliminary
animal
carcinogenicity
data.

Under
the
1999
draft
guidelines
(
EPA,
1999),
the
data
for
chlorine
dioxide
and
for
chlorite
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.

2.5.4.
Basis
for
RfD,
MCLG,
and
MRDLG
EPA
based
the
RfD
for
both
chlorite
and
chlorine
dioxide
on
the
CMA
(
1996)

twogeneration
study,
which
identified
a
NOAEL
of
3
mg/
kg/
day
chlorite
based
on
lowered
auditory
startle
amplitude
and
decreased
liver
weight.
An
UF
of
100
was
applied
for
extrapolation
from
an
animal
study
to
humans
and
to
account
for
variation
in
sensitivity
among
members
of
the
human
population
(
EPA,
2000c).
The
resulting
RfD
is
0.03
mg/
kg/
day.
The
MCLG
for
chlorite
and
the
revised
MRDLG
for
chlorine
dioxide
are
both
calculated
to
be
0.8
mg/
L
for
a
70
kg
adult
drinking
2
L
of
water
per
day.
A
number
of
studies
support
both
the
conclusion
that
neurodevelopmental
toxicity
is
the
critical
effect,
and
the
identified
NOAEL.
The
principal
study
(
CMA,
1996)
is
closely
supported
by
the
study
of
Mobley
et
al.
(
1990),
who
found
a
neurobehavioral
NOAEL
of
3
mg/
kg/
day
and
a
LOAEL
of
6
mg/
kg/
day,
based
on
decreased
exploratory
activity
in
pups
exposed
to
chlorite
during
gestation
and
lactation.
Further
support
for
neurodevelopmental
toxicity
being
the
critical
effect
and
for
the
NOAEL
comes
from
Orme
et
al.
(
1985),
who
found
a
chlorine
dioxide
NOAEL
of
3
mg/
kg/
day
and
a
LOAEL
of
14
mg/
kg/
day,

based
on
neurobehavioral
effects
in
rat
pups
exposed
during
gestation
and
lactation.
163
2.5.5.
Children's
Risk
in
Relation
to
the
MCLG
and
MRDLG
The
MCLG
and
MRDLG
calculated
for
chlorite
and
chlorine
dioxide
are
considered
to
be
protective
of
sensitive
subpopulations,
including
children,
because
the
RfD
is
based
on
a
NOAEL
derived
from
a
two­
generation
reproductive
toxicity
study
that
examined
numerous
developmental,
reproductive
and
systemic
endpoints.
In
addition,
the
results
of
this
study
are
supported
by
the
results
of
four
other
developmental
studies
that
showed
similar
effects
at
similar
dose
levels.
The
UF
of
100
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.

2.6.
CHLORINE
Chlorine
forms
elemental
chlorine
(
Cl
2),
chloride
ion
(
Cl

)
,
and
hypochlorous
acid
(
HOCl)

in
pure
water.
As
pH
increases,
hypochlorous
acid
dissociates
to
hypochlorite
ion
(
OCl

)
.

Several
factors,
including
chlorine
concentration,
pH,
temperature,
exposure
to
light,
and
presence
of
catalysts
or
organic
material
affect
the
stability
of
free
chlorine
in
aqueous
solution.

Because
hypochlorite
solutions
are
more
stable
than
hypochlorous
acid,
calcium
hypochlorite
and
sodium
hypochlorite
are
often
used
as
chlorine
sources
for
disinfection
of
drinking
water
(
EPA,

1994d).
Chlorine
and
hypochlorites
are
very
reactive
and
thus
can
react
with
the
constituents
of
saliva
and
possibly
food
and
gastric
fluid
to
yield
a
variety
of
reaction
byproducts.
Thus,
the
health
effects
associated
with
administration
of
high
levels
of
chlorine
and/
or
the
hypochlorites
in
various
animal
studies
may
be
due
to
these
reaction
byproducts
and
not
the
disinfectant
itself
(
EPA,
1994d).
164
Scully
and
White
(
1991)
noted
that
reactions
of
aqueous
chlorine
with
sulfur­
containing
amino
acids
appear
to
be
so
fast
in
saliva
that
all
free
available
chlorine
is
dissipated
before
water
is
swallowed
(
EPA,
1994d).
Therefore,
there
is
very
limited
potential
for
oral
exposure
of
fetuses,
infants
and
children
to
chlorine.

2.6.1.
Developmental/
Reproductive
Effects
Animal
studies
have
demonstrated
no
evidence
of
developmental
effects
associated
with
chlorine
(
EPA,
1994d).
In
a
study
by
Carlton
and
Smith
(
1985),
developmental
landmarks
such
as
the
mean
day
of
eye
opening
and
the
average
day
of
observed
vaginal
patency
were
compared
across
groups
with
no
statistical
differences
detected.
In
this
study,
chlorine
was
administered
by
gavage
in
deionized
water
at
doses
of
1.0,
2.0,
and
5.0
mg
chlorine
per
kg/
day
to
male
and
female
Long­
Evans
rats
for
66
 
76
days.
No
statistical
differences
were
observed
between
the
control
and
dosed
groups
in
litter
survival,
litter
size
and
pup
weight.
The
NOAEL
in
this
study
is
5
mg/
kg/
day;
however,
higher
doses
were
not
tested
(
EPA,
1994h).

In
a
multigenerational
study,
rats
were
given
drinking
water
chlorinated
to
a
concentration
of
100
mg
free
chlorine/
L
(
14
mg/
kg/
day)
(
Druckrey,
1968).
The
term
"
free
chlorine"
(
free
available
chlorine,
free
residual
chlorine)
refers
to
the
concentrations
of
elemental
chlorine,

hypochlorous
acid,
and
hypochlorite
ion
that
collectively
occur
in
water.
Animals
were
mated
repeatedly
and
continued
to
drink
the
test
water
throughout
gestation
and
lactation.

Microphthalmia
of
one
or
both
eyes
was
noted
in
17
treated
progeny
but
it
was
stated
that
this
condition
has
been
known
to
occur
spontaneously
in
BDII
rats.
No
adverse
reproductive
or
developmental
effects
were
observed.
165
Meier
et
al.
(
1985)
demonstrated
that
oral
administration
of
a
sodium
hypochlorite
solution
resulted
in
dose­
related
increases
in
the
number
of
sperm­
head
abnormalities
in
male
B6C3F1
mice.
Ten
animals/
group
were
given
1
mL
of
a
free­
residual­
chlorine
solution
daily
for
5
days.
Test
solutions
were
prepared
by
bubbling
Cl
2
into
a
1
M
solution
of
NaOH
and
adjusted
to
a
pH
of
either
8.5
(
predominant
species
OCl

)
or
6.5
(
predominant
species
HOCl).
The
solutions
were
diluted
with
distilled
water
to
200
mg/
L,
100
mg/
L,
and
40
mg/
L
chlorine
equivalents
(
8.0,
4.0,
or
1.6
mg/
kg/
day,
respectively).
The
mice
were
then
sacrificed
at
1,
3,
or
5
weeks
after
the
last
dose
was
administered.
In
mice
given
OCl

,
significant
increases
in
spermhead
abnormalities
were
observed
only
at
the
3­
week
interval
at
doses
of
1.6
and
4.0
mg/
kg/
day.

These
results
were
reproduced
in
retrials
of
the
experiment.
A
similar
level
of
increase
compared
to
the
concurrent
control
value
was
seen
in
both
the
initial
and
repeat
experiments,
reaching
about
twice
the
control
values
at
4.0
mg/
kg/
day.
At
8.0
mg/
kg/
day,
no
further
increases
were
seen
above
that
seen
at
4.0
mg/
kg/
day.
A
limitation
of
this
experiment
was
the
wide
variability
in
response
seen
in
the
control
animals.
In
addition,
no
dose
of
HOCl
was
associated
with
increases
in
sperm­
head
abnormalities.
This
result
was
surprising
because
OCl

should
be
converted
to
HOCl
under
the
acidic
conditions
of
the
stomach
and
thus
similar
results
for
both
species
of
chlorine
are
expected.
In
light
of
the
wide
variability
in
the
controls,
the
lack
of
a
clear
doseresponse
and
the
inconsistent
results
between
the
two
species
of
chlorine,
no
reliable
NOAEL
or
LOAEL
could
be
determined
from
this
study.

Six
Sprague­
Dawley
rats
were
administered
0,
1,
10,
or
100
mg
HOCl/
L
in
drinking
water
for
2.5
months
prior
to
mating
(
Abdel­
Rahman
et
al.,
1982).
Animals
were
maintained
on
the
treated
water
after
pregnancy
was
confirmed
(
day
0)
and
killed
on
day
20.
Maternal
weight
at
166
time
of
death
was
not
reported.
Incidence
of
fetal
anomalies
associated
with
exposure
to
hypochlorous
acid
solutions
was
not
found
to
be
statistically
significant.
Mean
fetal
weights
from
the
10
and
100
mg/
L
groups
were
less
than
the
control,
but
this
decrease
was
not
statistically
significant.
There
was
also
no
significant
difference
between
control
and
treated
groups
in
numbers
of
resorptions.
Examination
of
general
trends
in
the
study
indicated
an
increase
(
not
significant)
in
skeletal
anomalies
in
animals
treated
with
10
mg
HOCl/
L.
Soft
tissue
anomalies
for
the
100
mg
HOCl/
L
treatment
group
were
increased
significantly
compared
with
the
control.
The
findings
of
these
experiments
were
limited
by
the
small
number
of
study
animals.
In
addition,

some
calculations
of
anomaly
percentages
were
reported
incorrectly.

2.6.2.
Systemic
Toxicity
NTP
(
1992a)
conducted
a
2­
year
bioassay
in
which
male
and
female
F344
rats
and
B6C3F1
mice
were
given
chlorine
in
distilled
drinking
water
at
levels
of
0,
70,
140,
or
275
mg/
L
(
0,
4,
8,
and
14
mg/
kg/
day).
No
effects
on
body
weight,
survival,
or
evidence
of
non­
neoplastic
lesions
were
observed
for
any
of
the
treated
groups
of
animals.
In
this
study,
dosing
began
when
rats
and
mice
were
as
young
as
7
weeks
old.

2.6.3.
Carcinogenicity
The
carcinogenicity
of
chlorine
was
tested
in
F344
rats
and
B6C3F1
mice
by
oral
exposure
to
chlorine
in
distilled
drinking
water
as
hypochlorite
at
levels
up
to
275
mg/
L
(
14
mg/
kg/
day)
over
a
2­
year
period
(
NTP,
1992a).
The
incidence
of
mononuclear
cell
leukemia
was
significantly
increased
in
mid­
dose,
but
not
high­
dose,
female
rats.
There
was
no
other
evidence
167
of
carcinogenicity
in
rats
or
mice
of
either
sex.
NTP
(
1992a)
concluded
that
there
was
equivocal
evidence
of
carcinogenic
activity
of
chlorinated
water
in
female
F344/
N
rats
based
on
an
increase
in
the
incidence
of
mononuclear
cell
leukemia.
In
addition,
NTP
concluded
that
there
was
no
evidence
of
carcinogenic
activity
of
chlorinated
water
in
male
F344/
N
rats
or
B6C3F1
mice
of
either
sex
at
the
concentrations
tested.
There
was
also
no
evidence
of
systemic
toxicity,

indicating
that
the
maximum
tolerated
dose
(
MTD)
may
not
have
been
reached,
and
making
this
study
inadequate
for
fully
assessing
the
carcinogenic
potential
of
chlorine.
Under
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1986),
EPA
has
categorized
chlorine
in
Group
D:
Not
Classifiable
as
to
Human
Carcinogenicity
(
EPA,
1994a).
Under
the
draft
guidelines
(
EPA,

1999),
the
data
on
chlorine
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.

2.6.4.
Basis
for
RfD
and
MRDLG
EPA
based
the
RfD
for
chlorine
on
the
NOAEL
of
14
mg/
kg/
day
identified
for
female
rats
in
a
2­
year
bioassay
(
NTP,
1992a).
An
UF
of
100
was
used
to
account
for
interspecies
extrapolation
and
human
variability,
resulting
in
an
RfD
of
0.1
mg/
kg/
day
(
EPA,
1994d).
This
corresponds
to
an
MRDLG
of
4
mg/
L,
based
on
a
70
kg
adult
consuming
2
liters
of
water
per
day.

2.6.5.
Children's
Risk
Relative
to
the
MRDLG
The
Agency
believes
that
the
proposed
MRDLG
of
4
mg/
L
is
protective
of
children's
health.
Since
there
are
no
reliable
data
to
suggest
chlorine­
induced
developmental
or
reproductive
toxicity,
the
current
RfD
is
based
on
a
free
standing
NOAEL
of
14
mg/
kg/
day
from
168
a
2­
year
study
in
which
chlorine
dosing
began
when
rats
and
mice
were
as
young
as
7
weeks
old.

The
UF
of
100
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.

2.7.
CHLORAMINE
Inorganic
chloramines
are
alternative
disinfectants
that
are
rapidly
formed
when
free
chlorine
is
added
to
water
containing
ammonia.
Monochloramine
is
the
principal
chloramine
formed
in
chlorinated
natural
and
wastewaters
at
neutral
pH
and
is
much
more
persistent
in
the
environment
(
EPA,
1994e).

2.7.1.
Developmental/
Reproductive
Effects
In
a
developmental
study
(
Abdel­
Rahman
et
al.,
1982),
the
authors
investigated
the
effects
of
monochloramine
administered
in
drinking
water
to
female
Sprague­
Dawley
rats.
Rats
were
administered
0,
1,
10,
or
100
mg/
L
monochloramine
daily
in
drinking
water
for
2.5
months
before
and
throughout
gestation.
On
the
20th
day
of
gestation,
animals
were
sacrificed
for
soft
tissues
and
skeletal
examination
of
the
progeny.
Monochloramine
did
not
produce
any
significant
changes
in
rat
fetuses
at
any
dose
level;
there
was
a
slight
nonsignificant
increase
in
fetal
weight
in
all
chloramine­
treated
groups
compared
with
controls.
No
reproductive
toxicity
studies
of
chloramine
were
located.
169
2.7.2.
Systemic
Toxicity
A
study
investigated
the
effects
of
monochloramine
administered
in
drinking
water
to
rats
at
concentrations
of
0,
1,
10,
or
100
mg/
L
(
doses
of
0,
0.12,
1.2,
and
12
mg/
kg/
day)
for
12
months
(
Abdel­
Rahman
et
al.,
1984b).
Treatment­
related
decreases
in
blood
glutathione
levels
were
observed
at
6
and
12
months,
but
changes
in
glutathione
levels
were
inconsistent
at
other
times.
A
significant
increase
in
red
blood
cell
fragility
was
detected
after
2
and
10
months
of
treatment
at
the
mid­
dose
level,
and
after
2
and
6
months
at
the
high­
dose
level.
At
1.2
and
12
mg/
kg/
day,
significant
changes
in
hematologic
parameters,
including
decreased
red
blood
cell
count
and
hematocrit,
were
observed
at
3
months,
but
at
the
10­
month
evaluation
the
only
affected
hematologic
parameters
were
decreased
hemoglobin
concentration
and
mean
corpuscular
hemoglobin
at
the
high
dose.
The
increased
osmotic
fragility
observed
was
not
corroborated
by
NTP
(
1992a)
and
was
not
affected
in
the
acute
experimental
series.
The
biological
significance
of
these
changes
is
uncertain,
in
light
of
inconsistent
findings
at
different
exposure
durations,
and
in
the
absence
of
food
consumption
and
water
intake
data.

NTP
(
1992a)
exposed
F344/
N
rats
to
0,
50,
100,
or
200
ppm
(
2.8,
5.3,
and
9.5
mg/
kg/
day)
chloramine
in
drinking
water.
At
the
highest
dose
tested
(
9.5
mg/
kg/
day)
there
were
statistically
significant
changes
in
body
weight,
but
mean
body
weights
were
within
10%
of
controls
until
week
97
for
females
and
week
101
for
males.
Decreases
in
liver
and
kidney
weight,

and
increases
in
the
brain­
and
kidney­
to­
body
weight
ratios
in
the
high
dose
males
were
related
to
lower
body
weights
and
were
not
considered
toxicologically
significant.
The
test
animals
consumed
a
reduced
amount
of
water,
which
was
perhaps
due
to
unpalatability,
and
NTP
did
not
consider
these
changes
in
body
weight
biologically
significant.
In
a
related
study,
NTP
(
1992a)
170
administered
chloramine
to
B6C3F1
mice
in
drinking
water
at
0,
50,
100,
or
200
ppm.
Water
consumption
was
decreased
at
the
high
dose,
but
no
treatment­
related
effects
were
observed.
The
NOAEL
identified
in
this
study
was
the
high
dose,
9.5
mg/
kg/
day.

2.7.3.
Carcinogenicity
NTP
(
1992a)
tested
chloramine
in
drinking
water
in
a
2­
year
bioassay
in
male
and
female
F344
rats
and
B6C3F1
mice.
NTP
concluded
that
there
was
no
evidence
of
carcinogenicity
in
male
rats
or
in
mice
of
either
sex
at
the
doses
tested.
In
female
rats,
mononuclear
cell
leukemia
was
marginally
increased
at
the
mid
and
high
doses.
Based
on
these
increases,
NTP
concluded
that
there
was
equivocal
evidence
of
carcinogenicity
in
female
rats.
In
the
absence
of
evidence
of
noncancer
effects,
it
is
unclear
whether
the
assay
reached
the
MTD.
Under
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1986),
the
EPA
has
categorized
monochloramine
in
Group
D:
Not
Classifiable,
because
of
inadequate
human
and
animal
evidence
(
EPA,
1994a).
Under
the
draft
guidelines
(
EPA,
1999),
the
data
on
chloramine
are
inadequate
for
an
assessment
of
human
carcinogenic
potential.

2.7.4.
Basis
for
the
RfD
and
MRDLG
EPA
based
the
chloramine
RfD
on
the
absence
of
adverse
effects
in
the
lifetime
study
in
rats
(
NTP,
1992a).
The
high
dose
of
9.5
mg/
kg/
day
was
the
NOAEL.
An
UF
of
100
was
applied
(
10
for
interspecies
differences
and
10
for
intraspecies
variability),
resulting
in
an
RfD
of
0.1
mg/
kg/
day.
An
MRDLG
of
3
mg/
L
for
chloramine
(
4
mg/
L
measured
as
total
chlorine)
was
171
derived,
based
on
the
lack
of
toxic
effects,
for
a
70
kg
adult
consuming
2
L/
day
of
water
and
assuming
a
relative
source
contribution
(
RSC)
from
drinking
water
of
80%.

2.7.5.
Children's
Risk
in
Relation
to
the
MRDLG
Only
limited
data
are
available
for
comparing
the
effects
of
chloramine
on
fetuses
and
children
with
the
effects
in
adults.
No
developmental
effects
were
observed
in
the
single
available
developmental
toxicity
study
(
Abdel­
Rahman
et
al.,
1982).
Monochloramine
caused
hematological
effects
in
rats
exposed
to
100
mg/
L
for
12
months
(
Abdel­
Rahman
et
al,
1984b),

Data
on
hematological
effects
of
chloramine
in
infants
and
young
animals
are
not
available,
but
these
groups
may
be
more
sensitive
than
adults
to
these
effects.
People
with
decreased
reducing
power
may
be
more
sensitive
to
chloramine,
because
such
deficits
make
cells,
particularly
erythrocytes,
more
vulnerable
to
the
toxic
effects
of
chemicals
such
as
monochloramine.
This
suggests
that
infants
may
be
more
sensitive
than
adults
to
hematotoxic
effects
of
chloramine.

Newborns
may
also
be
more
susceptible,
because
the
reducing
power
in
red
blood
cells
is
used
to
reduce
methemoglobin
to
hemoglobin,
and
infants
have
a
transient
deficiency
of
methemoglobin
reductase,
the
enzyme
that
reduces
methemoglobin
to
hemoglobin.
People
deficient
in
the
glucose­
6­
phosphate
dehydrogenase
(
G­
6PD)
enzyme
are
another
potential
sensitive
population,

since
they
have
a
deficit
in
reducing
power
and
in
levels
of
reduced
glutathione.
This
enzyme
deficiency
is
more
frequent
in
male
members
of
certain
ethnic
groups,
such
as
Asians,
Arabs,

Caucasians
of
Latin
ancestry,
African
Americans,
and
Africans,
indicating
that
these
ethnic
groups
may
have
a
higher
susceptibility
to
the
hematological
effects
of
chloramine.
Nevertheless,
the
Agency
believes
that
the
MRDLG
for
chloramine,
which
is
based
on
a
NOAEL
and
includes
the
172
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.

2.8.
MX
[
3­
Chloro­
4­(
dichloromethyl)­
5­
hydroxy­
2(
5H)­
furanone]

Recent
reports
have
identified
the
presence
of
3­
chloro­
4­(
dichloromethyl)­
5­
hydroxy­

2(
5H)­
furanone
(
MX)
and
MX­
related
chlorohydroxyfuranone
compounds
in
drinking
water
samples
disinfected
with
chlorine
or
other
agents,
such
as
chloramine
or
chlorine
dioxide.
MX
is
produced
as
a
by­
product
of
chlorine
disinfection
of
drinking
water
containing
humic
material
and
as
a
byproduct
of
chlorine
bleaching
of
pulpwood.
MX
was
first
identified
in
the
spent
liquors
from
kraft
pulp
bleaching
approximately
16
years
ago
(
Holmbom
et
al.,
1984).
Based
on
recent
health
assessments
completed
for
the
chlorohydroxyfuranones,
it
was
determined
that,
of
the
chlorohydroxyfuranones,
MX
is
the
most
potent
mutagen
and
has
the
more
extensive
database.

Data
regarding
possible
developmental,
reproductive
or
long­
term
chronic
health
effects
for
chlorohydroxyfuranones
other
than
MX
are
not
available.
Accordingly,
this
section
will
focus
on
MX
as
a
representative
chlorohydroxyfuranone.

2.8.1.
Developmental/
Reproductive
Effects
No
data
were
located
on
the
reproductive
toxicity
of
MX.
The
developmental
toxicity
of
MX
has
been
evaluated
in
a
study
conducted
in
Wistar
rats.
Huuskonen
et
al.
(
1997)

administered
doses
of
0,
3,
30,
or
60
mg/
kg/
day
MX
via
oral
gavage
to
pregnant
rats
on
gestation
173
days
6­
19.
The
animals
were
sacrificed
on
gestation
day
20.
Indications
of
maternal
toxicity
(
reduced
body­
weight
gain,
decreased
absolute
and
relative
kidney
weights,
and
decreased
water
consumption)
were
observed
at
the
highest
dose.
The
mean
body
weights
of
the
fetuses
were
slightly
reduced
(
4­
6%)
at
all
doses,
but
the
effect
was
not
considered
statistically
significant.
No
increases
in
gross,
visceral
or
skeletal
malformations
were
observed
in
the
fetuses
in
dose
groups
compared
to
controls,
and
fetal
mortality
was
unaffected
by
MX
exposure.
The
authors
concluded
that
MX
was
not
teratogenic
in
this
strain
of
rat.

Teramoto
et
al.
(
1998)
evaluated
the
teratogenic
properties
of
MX
using
the
micromass
in
vitro
assay.
Twelve­
day
old
rat
embryo
midbrain
(
central
nervous
system
(
CNS))
and
limb
bud
(
LB)
cells
were
exposed
to
MX
at
concentrations
of
0,
1,
2,
5
or
10
µ
g/
mL
culture
medium
for
5
days
in
the
presence
or
absence
of
S9
mix.
MX
had
no
effect
on
the
number
of
differentiated
foci
in
CNS
cells,
and
little
or
no
effect
on
LB
cells
when
the
assay
was
conducted
in
the
presence
of
S9.
Although
MX
was
rapidly
degraded
in
the
culture
medium,
treatment
in
the
absence
of
S9
mix
caused
a
significant
decrease
in
the
number
of
differentiated
foci
in
CNS
(
2,
5
and
10
µ
g/
mL)

and
LB
cells
(
5
and
10
µ
g/
mL),
indicative
of
a
teratogenic
effect.
The
IC50
(
concentration
estimated
to
result
in
50%
inhibition)
for
these
effects
was
approximately
3
µ
g/
mL.
Some
evidence
for
cytotoxicity
was
observed
in
the
absence
of
S9
fraction
in
CNS
cells
(
15%
and
50%

decreases
in
survival
at
5
and
10
µ
g/
mL,
respectively),
while
MX
was
only
weakly
cytotoxic
in
LB
cells
(
15%
decrease
in
cell
survival
at
10
µ
g/
mL).
The
authors
concluded
on
the
basis
of
these
results
that
inhibition
of
differentiation
was
not
a
simple
result
of
cytotoxicity.
The
overall
pattern
of
results
was
interpreted
by
the
authors
as
evidence
that
MX
is
a
direct­
acting
teratogen
(
i.
e.,
does
not
require
bioactivation
to
exert
teratogenic
effects)
and
that
microsomal
metabolism
174
(
S9)
minimizes
the
teratogenic
risk.
It
should
be
noted,
however,
that
the
MX
concentrations
used
in
this
in
vitro
experiment
were
high
relative
to
levels
expected
to
occur
in
vivo
following
ingestion
through
drinking
water
(
approximately
1
x
10­
6
µ
g/
mL
based
on
a
drinking
water
intake
of
2
L/
day
containing
70
ng/
L
MX,
assuming
35%
absorption
and
uniform
distribution
to
a
total
body
volume
of
40
L
for
an
adult
human)
(
Guyton,
1981).

2.8.2.
Systemic
Effects
MX
has
a
strong
local
irritant
effect
on
the
gastrointestinal
tract,
as
evidenced
by
necrosis
and
hyperplasia
of
the
duodenum
and
forestomach
in
both
acute
(
Komulainen
et
al.,

1994)
and
subchronic
(
Vaittinen
et
al.,
1995)
rat
studies.
However,
in
the
subchronic
study
(
Vaittinen
et
al.,
1995),
it
is
possible
that
the
gastrointestinal
lesions
may
have
resulted
from
bolus
dosing.
In
the
Komulainen
et
al.
(
1994)
study,
respiratory
distress,
ataxia,
cyanosis,
and
reduced
motor
activity
were
also
noted.
Indicators
of
altered
renal
and
kidney
function,
including
increased
blood­
urea
nitrogen,
creatinine,
liver
weight
and
serum­
cholesterol
levels,
were
reported
in
the
Vaittinen
et
al.
(
1995)
study,
but
were
not
accompanied
by
marked
morphological
or
histopathological
changes.

2.8.3.
Carcinogenicity
No
data
were
identified
on
the
carcinogenicity
of
MX
in
humans.
In
animals,
MX
is
a
probable
multiple­
target
organ
carcinogen,
with
tumors
observed
in
the
thyroid,
liver,
mammary
gland,
lung,
adrenal,
and
pancreas
in
a
2­
year
oral
carcinogenicity
study
in
Wistar
rats
(
Komulainen
et
al.,
1997).
The
liver
and
thyroid
glands
were
found
to
be
the
primary
target
175
organs
of
tumorigenicity.
MX
promoted
tumor
formation
in
the
glandular
stomach
of
male
Wistar
rats
when
it
was
administered
following
pretreatment
with
N­
methyl­
N­
nitro­

Nnitrosoguanidine
(
MNNG)
(
Nishikawa
et
al.,
1999).

MX
is
a
potent
mutagen
in
bacterial
assay
systems,
including
multiple
test
strains
of
S.

typhimurium
and
Escherichia
coli,
when
tested
in
the
absence
of
metabolic
activation
(
DeMarini
et
al.,
1995;
Meier
et
al.,
1987).
MX
has
also
given
consistently
positive
results
in
the
absence
of
S9
(
or
S9
status
unknown)
when
tested
in
mammalian
cells
in
vitro
for
the
induction
of
gene
mutations,
chromosome
aberrations,
DNA
adducts,
or
other
indicators
of
DNA
damage
(
Harrington­
Brock
et
al.,
1995;
Le
Curieux
et
al.,
1999;
Maki­
Paakkanen
et
al.,
1994;
Meier
et
al.,
1987,
1989).
These
results
have
been
observed,
however,
at
concentrations
significantly
higher
than
those
expected
to
occur
in
vivo.
Evidence
from
in
vivo
studies
is
equivocal,
with
negative
results
obtained
in
assays
based
on
micronucleus
formation
in
bone
marrow
and
peripheral
blood
(
Jansson,
1998;
Meier
et
al.
1987,
1996)
and
positive
results
obtained
for
micronuclei
induction
and
sister­
chromatid
exchange
in
peripheral
lymphocytes
(
Jansson
et
al.,

1995;
Maki­
Paakkanen
and
Jansson,
1995).
Additional
evidence
for
direct
DNA
damage
has
been
obtained
in
the
form
of
strand
breaks
in
multiple
tissues
and
nuclear
anomalies
in
intestinal
cells
(
Daniel
et
al.,
1991;
Sasaki
et
al.,
1997).
Overall,
the
weight
of
evidence
indicates
that
MX
is
a
direct­
acting
genotoxicant
in
mammalian
cells.

Following
the
EPA's
1986
Guidelines
for
Carcinogen
Risk
Assessment,
MX
is
best
classified
as
Group
B2:
Probable
Human
Carcinogen.
This
classification
is
appropriate
because
there
is
sufficient
evidence
in
animals
and
inadequate
evidence
in
humans.
Under
the
1999
Draft
176
Guidelines
for
Carcinogen
Risk
Assessment
(
EPA,
1999),
MX
is
likely
to
be
carcinogenic
to
humans.

EPA
selected
the
2­
year
oral­
exposure
study
conducted
in
rats
by
Komulainen
et
al.

(
1997)
for
quantitative
evaluation
of
the
carcinogenic
effects
of
MX.
Data
for
occurrence
of
thyroid
follicular­
cell
adenoma
and
adenocarcinoma,
and
liver
cholangioma
and
cholangiocarcinoma
were
modeled.
Linear
low­
dose
extrapolation
was
used
because
the
weight
of
evidence
supports
the
view
that
MX
is
a
strong
bacterial
mutagen
and
is
also
mutagenic
in
mammalian
cells
in
vitro.
Using
the
data
for
thyroid
follicular
adenomas
in
male
rats
(
as
the
most
sensitive
tumor
type)
and
linear
extrapolation
from
the
LED
10,
a
cancer
oral
slope
factor
of
3.7
(
mg/
kg/
day)­
1
was
calculated.
Based
on
this
oral
slope
factor,
drinking
water
concentrations
of
approximately
0.95

g/
L,
0.095

g/
L,
and
0.0095

g/
L
are
associated
with
lifetime
cancer
risks
of
10­
4,
10­
5,
and
10­
6,
respectively.

2.8.4.
Basis
for
RfD
and
MCLG
EPA
has
not
derived
an
RfD
for
MX.
There
are
insufficient
data
available
to
evaluate
whether
children
are
more
sensitive
to
the
toxic
effects
of
MX
than
are
adults.

2.8.5.
Children's
Risk
in
Relation
to
the
MCLG
The
Agency
is
not
proposing
an
MCLG
for
3­
chloro­
4­(
dichloromethyl)­
5­
hydroxy­

2(
5H)­
furanone
(
MX)
at
the
present
time
because
the
Agency
has
not
yet
conducted
a
full
assessment
of
the
carcinogenicity
of
MX.
177
178
3.
SUMMARY
AND
CONCLUSIONS
In
developing
the
Proposed
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
issues
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.

Table
ES­
1
and
Table
3
summarized
the
comparison
of
toxicity
endpoints
for
the
various
D/
DBPs.
As
can
be
seen
in
the
table,
bromodichloromethane
(
BDCM),
bromoform,

dichloroacetic
acid
(
DCA),
bromate,
and
3­
chloro­
4­(
dichloromethyl)­
5­
hydroxy­
2(
5H)­
furanone
(
MX)
are
considered
probable
carcinogens
for
humans
under
the
1986
cancer
guidelines
and
as
likely
to
be
carcinogenic
to
humans
under
the
revised
1999
draft
cancer
guidelines.
MCLGs
of
zero
were
selected
after
consideration
of
the
potential
carcinogenicity
of
these
chemicals,
except
179
for
MX,
for
which
the
MCLG
has
not
yet
been
determined.
The
MCLG
of
zero
would
protect
both
children
and
adults.
The
MCLG
for
chloroform
was
set
based
on
a
nonlinear
extrapolation
to
low
doses
for
the
cancer
assessment.
Liver
toxicity
was
considered
the
most
sensitive
effect
for
chloroform
and
as
a
key
event
in
carcinogenicity.
This
approach
is
considered
equally
protective
for
children
and
adults
because
the
database
does
not
indicate
that
children
are
more
sensitive
than
adults
to
liver
toxicity.
In
addition,
the
mode
of
action
by
which
chloroform
produces
organ
toxicity
and
carcinogenicity
is
considered
to
be
the
same
in
children
and
adults.

The
MCLG/
MRDLGs
for
dibromochloromethane
(
DBCM),
monochloroacetic
acid
(
MCA),
trichloroacetic
acid
(
TCA),
chlorine,
and
chloramine
were
based
on
systemic
toxicity.

The
NOAEL/
LOAELs
used
to
derive
these
numbers
are
lower
than
the
NOAEL/
LOAELs
for
developmental
effects;
therefore,
the
MCLG/
MRDLG
would
be
protective
of
developmental
effects
in
infants
and
children.
Based
on
the
available
data,
it
is
believed
that
the
MCLGs
for
these
chemicals
are
also
protective
of
systemic
effects
in
children,
in
light
of
the
absence
of
evidence
to
the
contrary,
and
because
the
UFs
used
in
the
derivation
of
these
RfDs
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.
There
is,
however,
some
uncertainty
in
this
conclusion,
since
inadequate
data
are
available
for
these
chemicals
to
directly
compare
systemic
toxicity
in
children
and
adults,
or
to
extrapolate
based
on
data
on
age­
related
differences
in
metabolic
capacity.
In
the
case
of
chlorine,
the
MRDLG
is
based
on
systemic
toxicity
because
the
NOAEL
of
5
mg/
kg/
day
based
on
developmental
effects
is
the
highest
dose
tested
and
is,
therefore,
not
a
true
NOAEL.
For
chlorine
dioxide
and
chlorite,
the
MCLG/
MRDLGs
are
calculated
based
on
data
from
developmental
studies;
hence
the
numbers
180
derived
would
be
protective
of
developmental
effects
in
both
children
and
adults.
The
chlorine
dioxide/
chlorite
MCLG/
MRDLG
is
also
considered
to
be
protective
of
systemic
effects
in
children,
because
it
is
based
on
data
from
a
two­
generation
reproduction
study
that
examined
numerous
developmental,
reproductive
and
systemic
endpoints.
For
chloramine,
the
MRDLG
was
based
on
a
NOAEL
of
9.5
mg/
kg/
day
in
a
rat
lifetime
study.
Infants
may
be
more
sensitive
than
adults
to
the
critical
effect
of
chloramine
(
hematotoxicity).
The
Agency
believes
that
the
chloramine
MRDLG,
which
includes
the
standard
UF
of
10
for
protection
of
sensitive
populations,
is
protective
of
infants,
although
there
is
some
uncertainty
in
this
conclusion.
The
data
on
monobromoacetic
acid
(
MBA),
bromochloroacetic
acid
(
BCA),
dibromoacetic
acid
(
DBA),
and
3­
chloro­
4­(
dichloromethyl)­
5­
hydroxy­
2(
5H)­
furanone
(
MX)
are
insufficient
for
the
derivation
of
an
MCLG.
It
can
be
concluded
that
the
MCLG/
MRDLGs
of
all
the
D/
DBPs
in
the
proposed
Stage
2
D/
DBP
Rule
are
protective
of
fetuses,
infants
and
children.
F
181
4.
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Abdel­
Rahman
M.
S.,
M.
R.
Bernardi
and
R.
J.
Bull.
1982.
Effect
of
chlorine
and
monochloramine
in
drinking
water
in
the
developing
rat
fetus.
J
Appl
Toxicol.
2(
3):
156­
9.
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As
cited
in
EPA
1998c)

Abdel­
Rahman,
M.
S.,
D.
Couri
and
R.
J.
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1984a.
Toxicity
of
chlorine
dioxide
in
drinking
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Am
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EPA
2000c)

Abdel­
Rahman,
M.
S.,
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D.,
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R.
J.
1984b.
Toxicity
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monochloramine
in
rat:
An
alternative
drinking
water
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825­
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K.
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Y.
Kurokawa
and
M.
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Chronic
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the
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2):
51­
68,
[
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Andrews,
J.
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J.
Schmidt,
H.
Nichols,
E.
S.
Hunter
and
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Developmental
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haloacetic
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embryo
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