[Federal Register Volume 59, Number 145 (Friday, July 29, 1994)]
[Unknown Section]
[Page 0]
From the Federal Register Online via the Government Publishing Office [www.gpo.gov]
[FR Doc No: 94-17651]
[[Page Unknown]]
[Federal Register: July 29, 1994]
_______________________________________________________________________
Part II
Environmental Protection Agency
_______________________________________________________________________
40 CFR Parts 141 and 142
National Primary Drinking Water Regulations; Disinfectants and
Disinfection Byproducts; Proposed Rule
ENVIRONMENTAL PROTECTION AGENCY
40 CFR Parts 141 and 142
[WH-FRL-4998-2]
Drinking Water; National Primary Drinking Water Regulations:
Disinfectants and Disinfection Byproducts
AGENCY: Environmental Protection Agency (EPA).
ACTION: Proposed rule.
-----------------------------------------------------------------------
SUMMARY: In this document, EPA is proposing maximum residual
disinfectant level goals (MRDLGs) for chlorine, chloramines, and
chlorine dioxide; maximum contaminant level goals (MCLGs) for four
trihalomethanes (chloroform, bromodichloromethane,
dibromochloromethane, and bromoform), two haloacetic acids
(dichloroacetic acid and trichloroacetic acid), chloral hydrate,
bromate, and chlorite; and National Primary Drinking Water Regulations
(NPDWRs) for three disinfectants (chlorine, chloramines, and chlorine
dioxide), two groups of organic disinfection byproducts (total
trihalomethanes (TTHMs)--a sum of the four listed above, and haloacetic
acids (HAA5)--a sum of the two listed above plus monochloroacetic acid
and mono- and dibromoacetic acids), and two inorganic disinfection
byproducts (chlorite and bromate). The NPDWRs consist of maximum
residual disinfectant levels or maximum contaminant levels or treatment
techniques for these disinfectants and their byproducts. The NPDWRs
also include proposed monitoring, reporting, and public notification
requirements for these compounds. This notice proposes the best
available technology (BAT) upon which the MRDLs and MCLs are based and
the BAT for purposes of issuing variances.
DATES: Written comments must be postmarked or hand-delivered by
December 29, 1994. Comments received after this date may not be
considered. Public hearings will be held at the addresses indicated
below under ``ADDRESSES'' on August 29 (and 30, if necessary) in
Denver, CO and on September 12 (and 13, if necessary) in Washington,
DC.
ADDRESSES: Send written comments on the proposed rule to Disinfectant/
Disinfection By-Products Comment Clerk, Drinking Water Docket (MC
4101), Environmental Protection Agency, 401 M Street, S.W., Washington,
D.C. 20460. Commenters are requested to submit three copies of their
comments and at least one copy of any references cited in their written
or oral comments. A copy of the comments and supporting documents are
available for review at the EPA, Drinking Water Docket (4101), 401 M
Street, S.W., Washington, DC 20460. For access to the docket materials,
call (202) 260-3027 between 9:00 a.m. and 3:30 p.m.
The Agency will hold public hearings on the proposal at two
different locations indicated below:
1. Denver Federal Center, 6th and Kipling Streets, Building 25, Lecture
Halls A and B (3d Street), Denver, CO 80225 on August 29 (and 30, if
necessary), 1994.
2. EPA Education Center Auditorium, 401 M Street SW., Washington, D.C.
20460, on September 12 (and 13, if necessary), 1994.
The hearings will begin at 9:30 a.m., with registration at 9:00
a.m. The Hearings will end at 4:00 p.m., unless concluded earlier.
Anyone planning to attend the public hearings (especially those who
plan to make statements) may register in advance by writing the D/DBPR
Public Hearing Officer, Office of Ground Water and Drinking Water
(4603), USEPA, 401 M Street, S.W., Washington, D.C. 20460; or by
calling Tina Mazzocchetti, (703) 931-4600. Meeting dates are tentative
and should be confirmed by calling the Safe Drinking Water Hotline
prior to making travel plans. Oral and written comments may be
submitted at the public hearing. Persons who wish to make oral
presentations are encouraged to have written copies (preferably three)
of their complete comments for inclusion in the official record.
Copies of draft health criteria, analytical methods, and regulatory
impact analysis documents are available at some Regional Offices listed
below and for a fee from the National Technical Information Service
(NTIS), U.S. Department of Commerce, 5285 Port Royal Road, Springfield,
Virginia 22161. The toll-free number is (800) 336-4700 or local at
(703) 487-4650.
FOR FURTHER INFORMATION CONTACT: General information may be obtained
from the Safe Drinking Water Hotline, telephone (800) 426-4791; Stig
Regli, Office of Ground Water and Drinking Water (4603), U.S.
Environmental Protection Agency, 401 M Street, SW., Washington, DC
20460, telephone (202) 260-7379; Tom Grubbs, Office of Ground Water and
Drinking Water (4603), U.S. Environmental Protection Agency, 401 M
Street, SW., Washington, DC 20460, telephone (202) 260-7270; or one of
the EPA Regional Office contacts listed below.
SUPPLEMENTARY INFORMATION:
EPA Regional Offices
I. Robert Mendoza, Chief, Water Supply Section, JFK Federal Bldg.,
Room 203, Boston, MA 02203, (617) 565-3610
II. Robert Williams, Chief, Water Supply Section, 26 Federal Plaza,
Room 824, New York, NY 10278, (212) 264-1800
III. Jeffrey Hass, Chief, Drinking Water Section (3WM41), 841
Chestnut Building, Philadelphia, PA 19107, (215) 597-9873
IV. Phillip Vorsatz, Chief, Water Supply Section, 345 Courtland
Street, Atlanta, GA 30365, (404) 347-2913
V. Charlene Denys, Chief, Water Supply Section, 77 W. Jackson Blvd.,
Chicago, IL 60604, (312) 353-2650
VI. F. Warren Norris, Chief, Water Supply Section, 1445 Ross Avenue,
Dallas, TX 75202, (214) 655-7155
VII. Ralph Flournoy, Chief, Water Supply Section, 726 Minnesota
Ave., Kansas City, KS 66101, (913) 234-2815
VIII. Doris Sanders, Chief, Water Supply Section, One Denver Place,
999 18th Street, Suite 500, Denver, CO 80202-2405, (303) 293-1424
IX. Bill Thurston, Chief, Water Supply Section, 75 Hawthorne Street,
San Francisco, CA 94105, (415) 744-1851
X. William Mullen, Chief, Water Supply Section, 1200 Sixth Avenue,
Seattle, WA 98101, (206) 442-1225.
Abbreviations used in this document.
AECL: Alternate enhanced coagulant level
AOC: Assimilable organic carbon
ASDWA: Association of State Drinking Water Administrators
AWWA: American Water Works Association
AWWARF: AWWA Research Foundation
BAC: Biologically active carbon
BAF: Biologically active filtration
BAT: Best Available Technology
BCAA: Bromochloroacetic acid
BDOC: Biodegradable organic carbon
BTGA: Best Technology Generally Available
CI: Confidence interval
CWS: Community Water System
DBP: Disinfection byproducts
D/DBP: Disinfectants and disinfection byproducts
D/DBPR: Disinfectants and disinfection byproducts rule
DBPP: Disinfection byproduct precursors
DBPRAM: DBP Regulatory Assessment model
DPD: N,N-diethyl-p-phenylenediamine
DWEL: Drinking Water Equivalent Level
EBCT: Empty bed contact time
EMSL: EPA Environmental Monitoring and Support Laboratory
(Cincinnati)
EPA: United States Environmental Protection Agency
ESWTR: Enhanced Surface Water Treatment Rule
FY: Fiscal year
GAC: Granular Activated Carbon
GWDR: Ground Water Disinfection Rule
GWSS: Ground Water Supply Survey
HAA5: Haloacetic acids (five)
HOBr: Hypobromous acid
IC: Ion chromotography
ICR: Information Collection Rule
IOC: Inorganic chemical
LOAEL: Lowest observed adverse effect level
LOQ: Limit of Quantitation
MCL: Maximum Contaminant Level (expressed as mg/l, 1,000 micrograms
(g) = 1 milligram (mg))
MCLG: Maximum Contaminant Level Goal
MDL: Method Detection Limit
MF: Modifying factor
mg/dl: Milligrams per deciliter
mg/l: Milligrams per liter
MGD: Million Gallons per Day
MRDL: Maximum Residual Disinfectant Level (as mg/l)
MRL: Minimum reporting level
MRDLG: Maximum Residual Disinfectant Level Goal
NCI: National Cancer Institute
ND: Not detected
NIPDWR: National Interim Primary Drinking Water Regulation
NOAEL: No observed adverse effect level
NOMS: National Organic Monitoring Survey
NORS: National Organics Reconnaissance Survey for Halogenated
Organics
NPDWR: National Primary Drinking Water Regulation
NTNCWS: Nontransient noncommunity water system
OBr: Hypobromite ion
OR: Odds ratio
PE: Performance evaluation
POE: Point-of-Entry Technologies
POU: Point-of-Use Technologies
ppb: Parts per billion
PQL: Practical Quantitation Level
PTA: Packed Tower Aeration
PWS: Public Water System
RIA: Regulatory Impact Analysis
RMCL: Recommended Maximum Contaminant Level
RNDB: Regulations Negotiation Data Base
RSC: Relative Source Contribution
SDWA: Safe Drinking Water Act, or the ``Act,'' as amended in 1986
SM: Standard Method
SMCL: Secondary Maximum Contaminant Level
SMR: Standardized mortality ratios
SOC: Synthetic Organic Chemical
SWTR: Surface Water Treatment Rule
THMFP: Trihalomethane formation potential
TOC: Total organic carbon
TTHM: Total trihalomethanes
TWG: Technologies Working Group
VOC: Volatile Synthetic Organic Chemical
WIDB: Water Industry Data Base
WS: Water Supply
Table of Contents
I. Summary of Today's Action
A. Applicability.
B. Proposed MRDLGs and MRDLs for disinfectants
C. Proposed MCLGs and MCLs for organic byproducts
D. Treatment technique for DBP precursors
E. Proposed Stage 1 MCLGs and MCLs for inorganic byproducts
F. Proposed BAT for disinfectants
G. Proposed BAT for organic byproducts
H. Proposed BAT for inorganic byproducts
I. Proposed Compliance Monitoring Requirements
J. Analytical Methods
K. Laboratory Certification Criteria
L. Variances and Exemptions
M. State Primacy, Recordkeeping, Reporting Requirements
N. System Reporting Requirements
O. D/DBP Stage 2 Rule requirements
P. Guidance
Q. Triennial Regulation Review
II. Statutory Authority
A. MCLGs, MCLs, and BAT
B. Variances and Exemptions
C. Primacy
D. Monitoring, Quality Control, and Records
E. Public Water Systems
F. Public Notification
III. Overview of Existing Interim Standard for TTHMs
IV. Overview of Preproposal Regulatory Development
A. October 1989 Strawman Rule
B. June 1991 Status Report on D/DBP rule development
C. Initiation of Regulatory Negotiation Process
V. Establishing MCLGs
A. Background
B. Proposed MRDLGs and MCLGs
1. Chlorine, hypochloriteion and hypochlorous acid
2. Chloramines
3. Epidemiology Studies of Chlorinated and Chloraminanted Water
4. Chlorine dioxide, chlorite, and chlorate
5. Chloroform
6. Bromodichloromethane
7. Dibromochloromethane
8. Bromoform
9. Dichloroacetic acid
10. Trichloroacetic acid
11. Chloral hydrate
12. Bromate
VI. Occurrence of TTHMs, HAA5, and other DBPs
A. Relationship of TTHMs, HAA5 to disinfection and source water
quality
B. Chlorination Byproducts
C. Other Disinfection Byproducts
1. Ozonation Byproducts
2. Chlorine Dioxide Byproducts
3. Chloramination Byproducts
VII. General Basis for Criteria of Proposed rule
A. Goals of regulatory negotiation
B. Concern for downside microbial risks and unknown risks from
DBPs of different technologies
C. Ecological concerns
D. Watershed protection
E. Narrowing of regulatory options through reg-neg process
VIII. Summary of the Proposed National Primary Drinking Water
Regulation for Disinfectants and Disinfection Byproducts
A. Schedule and coverage
B. Summary of DBP MCLs, BATs, and monitoring and compliance
requirments
C. Summary of disinfectant MRDLs, BATs, and Monitoring and
compliance requirements
D. Enhanced coagulation and enhanced softening requirements
E. Requirement for systems to use qualified operators
F. Basis for analytical method requirements
G. Public Notice Requirements
H. Variances and Exemptions
I. Reporting and Record Keeping requirements for PWSs
J. State Implementation Requirements
IX. Basis for Key Specific Criteria of Proposed Rule
A. 80/60 TTHM/HAA5 MCLs, enhanced coagulation requirements, and
BAT
1. basis for umbrella concept vs. individual MCLs
2. basis for level of stringency in MCLs, BAT, and concurrent
enhanced coagulation requirements
3. basis for enhanced coagulation and softening criteria
4. basis for GAC definitions
5. basis for monitoring requirements
B. Bromate MCL and BAT
C. Chlorite MCL and BAT
D. Chlorine MRDL and BAT
E. Chloramine MRDL and BAT
F. Chlorine dioxide MRDL and BAT
G. Basis for analytical method requirements
H. Basis for compliance schedule and applicability to different
groups of systems, timing with other regulations
I. Basis for qualified operator requirements and monitoring
plans
J. Basis for Stage 2 proposed MCLs
X. Laboratory Certification and Approval
A. PE-Sample Acceptance Limits for Laboratory Certification
B. Approval Criteria for Disinfectants and Other Parameters
C. Other Laboratory Performance Criteria
XI. Variances and Exemptions
A. Variances
B. Exemptions
XII. State Implementation
A. Special primacy requirements
B. State recordkeeping
C. State reporting
XIII. System Reporting and Recordkeeping Requirements
XIV. Public Notice Requirements
XV. Economic Analysis
A. Executive Order 12866
B. Predicted cost impacts on public water systems
1. Compliance treatment cost forecasts
2. Compliance treatment forecasts
3. DBP exposure estimates
4. System level cost estimates
5. Effect on household costs
6. Monitoring and State implementation costs, labor burden
estimates
C. Concepts of cost analysis
D. Benefits
XVI. Other Requirements
A. Consultation with State, Local, and Tribal Governments
B. Regulatory Flexibility Act
C. Paperwork Reduction Act
D. National Drinking Water Advisory Council and Science Advisory
Board
XVII. Request for Public Comment
XVIII. References and Public Docket
I. Summary of Today's Action
In 1992 EPA initiated a negotiated rulemaking to develop a
disinfectant/disinfection byproduct rule. The Agency decided to use the
negotiated rulemaking process because it believed that the available
occurrence, treatment, and health effects data were inadequate to
address EPA's concerns about the tradeoff between risks from
disinfectants and disinfection byproducts and microbial pathogen risk,
and wanted all stakeholders to participate in the decision-making on
setting proposed standards. The negotiators included State and local
health and regulatory agency staff and elected officials, consumer
groups, environmental groups, and representatives of public water
systems. The group met from November 1992 through June 1993.
Early in the process, the negotiators agreed that large amounts of
information necessary to understand how to optimize the use of
disinfectants to concurrently minimize microbial and disinfectant/
disinfection byproduct risk were unavailable.
Therefore, the group agreed to propose a disinfectant/disinfection
byproduct rule to extend coverage to all community water systems that
use disinfectants, reduce the current total trihalomethane (TTHM)
maximum contaminant level (MCL), regulate additional disinfection
byproducts, set limits for the use of disinfectants, and reduce the
level of compounds that may react with disinfectants to form
byproducts. These requirements were based on available information. The
group further agreed that revisions to the current Surface Water
Treatment Rule might be required at the same time to ensure that
microbial risk is not increased as byproduct rules go into effect.
Finally, the group agreed that additional information on health risk,
occurrence, treatment technologies, and analytical methods needed to be
developed in order to better understand the risk-risk tradeoff, whether
further control was needed, and how to accomplish this overall risk
reduction.
The outcome of the negotiation was three rules: a Disinfectant/
Disinfection Byproduct rule (this notice), an Enhanced Surface Water
Treatment Rule (also proposed today and appearing separately in today's
Federal Register), and an Information Collection Rule (proposed
February 10, 1994, 59 FR 6332). The Information Collection Rule will
provide information necessary to determine whether the Enhanced Surface
Water Treatment Rule needs to be promulgated and, if so, what
requirements it should set. The Information Collection Rule will also
provide information on the need for, and content of, long-term rules.
The schedule to produce these rules has also been negotiated and is
provided elsewhere in this document. A summary of today's rule follows.
A. Applicability. This action applies to all community water
systems and nontransient noncommunity water systems that add a
disinfectant during any part of the treatment process including
addition of a residual disinfectant. In addition, certain provisions
apply to transient noncommunity water systems that use chlorine
dioxide.
B. Proposed MRDLGs and MRDLs for disinfectants. EPA is proposing
the following maximum disinfectant residual level goals and maximum
residual disinfectant levels.
------------------------------------------------------------------------
Disinfectant Residual MRDLG (mg/l) MRDL (mg/l)
------------------------------------------------------------------------
(1) Chlorine................ 4 (as Cl2).......... 4.0 (as Cl2).
(2) Chloramines............. 4 (as Cl2).......... 4.0 (as Cl2).
(3) Chlorine dioxide........ 0.3 (as ClO2)....... 0.8 (as ClO2).
------------------------------------------------------------------------
C. Proposed MCLGs and MCLs for organic byproducts. EPA is proposing
the following maximum contaminant level goals and maximum contaminant
levels.
------------------------------------------------------------------------
MCLG(mg/ MCL(mg/
l) l)
------------------------------------------------------------------------
Total trihalomethanes (TTHM)....................... \1\N/A 0.080
Haloacetic acids (five) (HAA5)..................... \2\N/A .060
Chloroform......................................... 0 \1\N/A
Bromodichloromethane............................... 0 \1\N/A
Dibromochloromethane............................... 0.06 \1\N/A
Bromoform.......................................... 0 \1\N/A
Dichloroacetic acid................................ 0 \2\N/A
Trichloroacetic acid............................... 0.3 \2\N/A
Chloral hydrate.................................... 0.04 \3\N/A
------------------------------------------------------------------------
\1\Total trihalomethanes are the sum of the concentrations of
bromodichloromethane, dibromochloromethane, bromoform, and chloroform.
\2\Haloacetic acids (five) are the sum of the concentrations of mono-,
di-, and trichloroacetic acids and mono- and dibromoacetic acids.
\3\EPA did not set an MCL for chloral hydrate because the TTHM and HAA5
MCLs and the treatment technique (i.e., enhanced coagulation) for
disinfection byproduct precursor removal will control for chloral
hydrate. (See Section IX.)
D. Treatment Technique for DBP Precursors. EPA is proposing that
water systems that use surface water or ground water under the direct
influence of surface water and use conventional filtration treatment be
required to remove specified amounts of organic materials (measured as
total organic carbon) that may react with disinfectants to form
disinfection byproducts. Removal would be achieved through a treatment
technique (enhanced coagulation or enhanced softening) unless the
system met certain criteria.
E. Proposed Stage 1 MCLGs and MCLs for inorganic by-products. EPA
is proposing the following maximum contaminant level goals and maximum
contaminant levels.
------------------------------------------------------------------------
MCLG(mg/
l) MCL(mg/l)
------------------------------------------------------------------------
Chlorite.......................................... 0.08 1.0
Bromate........................................... 0 0.010
------------------------------------------------------------------------
F. Proposed BAT for disinfectants. EPA is proposing the following
best available technologies for limiting residual disinfectant
concentrations in the distribution system.
Chlorine residual--control of treatment processes to reduce
disinfectant demand and control of disinfection treatment processes to
reduce disinfectant levels
Chloramine residual--control of treatment processes to reduce
disinfectant demand and control of disinfection treatment processes to
reduce disinfectant levels
Chlorine dioxide residual--control of treatment processes to reduce
disinfectant demand and control of disinfection treatment processes to
reduce disinfectant levels.
G. Proposed BAT for organic byproducts. EPA is proposing the
following best available technologies for control of organic
disinfection byproducts in each stage of the rule.
1. Proposed Stage 1 BAT for organic by-products. Total
trihalomethanes--enhanced coagulation or GAC10, with chlorine as the
primary and residual disinfectant. Total haloacetic acids--enhanced
coagulation or GAC10, with chlorine as the primary and residual
disinfectant.
2. Proposed Stage 2 BAT for organic byproducts. Total
trihalomethanes--enhanced coagulation and GAC10, or GAC20; with
chlorine as the primary and residual disinfectant. Total haloacetic
acids--enhanced coagulation and GAC10, or GAC20; with chlorine as the
primary and residual disinfectant.
H. Proposed BAT for inorganic by-products. EPA is proposing the
following best available technologies for control of inorganic
disinfection byproducts.
Chlorite--control of treatment processes to reduce disinfectant demand
and control of disinfection treatment processes to reduce disinfectant
levels.
Bromate--control of ozone treatment process to reduce production of
bromate.
I. Proposed Compliance Monitoring Requirements. Compliance
monitoring requirements are explained in Section IX of the preamble and
were developed during the negotiated rulemaking. EPA has developed
routine and reduced monitoring schemes that address the health effects
of each disinfectant or contaminant in an individually appropriate
manner.
J. Analytical Methods. EPA is proposing to withdraw one method for
measurement of chlorine residual and to approve three new methods for
measurement of chlorine residuals. EPA is proposing to approve one new
method for measurement of trihalomethanes; two new methods for
measurement of haloacetic acids; one new method for measurement of
bromate, chlorite, and bromide; and two new methods for measurement of
total organic carbon.
K. Laboratory Certification Criteria. Consistent with other
drinking water regulations, EPA is proposing that only certified
laboratories be allowed to analyze samples for compliance with the
proposed MCLs and treatment technique requirements. For disinfectants
and other specified parameters in today's rule that the Agency believes
can be adequately measured by other than certified laboratories and for
which there is a good reason to allow analysis at other locations
(e.g., for samples which normally deteriorate before reaching a
certified laboratory, especially when taken at remote locations), EPA
is requiring that such analyses be conducted by a party acceptable to
EPA or the State.
L. Variances and Exemptions. Variances and exemptions will be
permitted.
M. State Primacy, Recordkeeping, Reporting Requirements.
Requirements for States to maintain primacy are listed in Section XII
of the preamble. In addition to routine requirements, EPA has included
special primacy requirements.
N. System Reporting Requirements. System reporting requirements
remain consistent with requirements in previous rules.
O. D/DBP Stage 2 Rule requirements. EPA is proposing a total
trihalomethane MCL of 0.040 mg/l and a haloacetic acid (five) MCL of
0.030 mg/l, to apply only to systems using surface water or ground
water under the direct influence of surface water and serving at least
10,000 persons, as part of a plan to develop new standards which
incorporates the results of additional research conducted under the
Information Collection Rule (59 FR 6332).
P. Guidance. EPA is in the process of developing guidance for both
systems and States for implementation of this rule.
Q. Triennial Regulation Review. Under the provisions of the Safe
Drinking Water Act (SDWA or the Act) (Section 1412(b)(9)), the Agency
is required to review national primary drinking water regulations at
least once every three years. As mentioned previously, today's proposed
rule revises, updates, and (when promulgated) supersedes the
regulations for total trihalomethanes, initially published in 1979.
Since that time, there have been significant changes in technology,
treatment techniques, and other regulatory controls that provide for
greater protection for health of persons. As such, in proposing today's
rule, EPA has analyzed innovations and changes in technology and
treatment techniques that have occurred since promulgation of the
initial TTHM regulations. This analysis, contained primarily in the
cost and technology document supporting this proposal, supports
amendment of the TTHM regulation for the greater protection of persons.
EPA believes that the innovations and changes in technology and
treatment techniques will result in amendments to the TTHM regulations
that are feasible within the meaning of SDWA Section 1412(b)(9).
II. Statutory Authority
Section 1412 of the Safe Drinking Water Act, as amended in 1986
(``SDWA'' or ``the Act''), requires EPA to publish Maximum Contaminant
Level Goals (MCLGs) and promulgate National Primary Drinking Water
Regulations (NPDWRs) for contaminants in drinking water which may cause
any adverse effect on the health of persons and which are known or
anticipated to occur in public water systems. Under Section 1401, the
NPDWRs are to include Maximum Contaminant Levels (MCLs) and ``criteria
and procedures to assure a supply of drinking water which dependably
complies'' with such MCLs. Under Section 1412(b)(7)(A), if it is not
economically or technically feasible to ascertain the level of a
contaminant in drinking water, EPA may require the use of a treatment
technique instead of an MCL.
Under Section 1412(b), EPA was to establish MCLGs and promulgate
NPDWRs for 83 contaminants by June 19, 1989. An additional 25
contaminants are to be regulated every 3 years. To meet this latter
requirement, EPA has developed a list of contaminants (National
Drinking Water Priority List; 53 FR 1892) including pesticides, organic
and inorganic elements or compounds, and disinfectants and disinfection
by-products (D/DBP), plus the protozoan Cryptosporidium. From this
list, EPA is to choose at least 25 contaminants for regulation every
three years. Today's regulatory proposal represents part of the first
group of 25 chemicals to be regulated. Both the general contaminants
(organics, inorganics, and pesticides), and the D/DBPs were considered
for regulation. In today's notice, EPA is proposing to regulate certain
disinfectants and disinfection byproducts; Cryptosporidium is proposed
to be regulated in a separate Notice today.
In October of 1990, EPA entered into a consent order with Citizens
Concerned about Bull Run Inc. regarding a timeframe for proposing the
first group of 25. The consent decree stipulated a June 1993 date for
proposal. That decree was subsequently amended to establish a proposal
date of May 30, 1994, for the Disinfectants/Disinfection Byproducts
Rule and a proposal date of February 28, 1995, for the other
contaminants that comprise the required group of 25.
A. MCLGs, MCLs, and BAT
Under Section 1412 of the Act, EPA is to establish MCLGs at the
level at which no known or anticipated adverse effects on the health of
persons occur and which allow an adequate margin of safety. MCLGs are
nonenforceable health goals based only on health effects and exposure
information.
MCLs are enforceable standards which the Act directs EPA to set as
close to the MCLGs as feasible. ``Feasible'' means feasible with the
use of the best technology, treatment techniques, and other means which
the Administrator finds available (taking cost into consideration)
after examination for efficacy under field conditions and not solely
under laboratory conditions (SDWA, section 1412(b)(5)). Also, the SDWA
requires the Agency to identify the best available technology (BAT)
which is feasible for meeting the MCL for each contaminant.
Also, in this proposal, EPA is introducing several new terms--
``maximum residual disinfectant level goals (MRDLGs)'' and ``maximum
residual disinfectant levels (MRDLs)''--to reflect the fact that these
substances have beneficial disinfection properties. As with MCLGs, EPA
has established MRDLGs at the level at which no known or anticipated
adverse effects on the health of persons occur and which allow 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.
MRDLs are enforceable standards, analogous to MCLs, which recognize
the benefits of adding a disinfectant to water on a continuous basis
and in addressing emergency situations such as distribution system pipe
breaks. As with MCLs, EPA has set the MRDLs as close to the MRDLGs as
feasible. The Agency has also identified the best available technology
(BAT) which is feasible for meeting the MRDL for each disinfectant.
B. Variances and Exemptions
Section 1415 authorizes the State to issue variances from NPDWRs
(the term ``State'' is used in this preamble to mean the State agency
with primary enforcement responsibility for the public water supply
system program or EPA if the State does not have primacy). The State
may issue a variance if it determines that a system cannot comply with
an MCL despite application of the best available technology (BAT).
Under Section 1415, EPA must propose and promulgate its finding of the
best available technology, treatment techniques, or other means
available for each contaminant, for purposes of section 1415 variances,
at the same time that it proposes and promulgates a maximum contaminant
level for such contaminant. EPA's finding of BAT, treatment techniques,
or other means for purposes of issuing variances may vary among
systems, depending upon the number of persons served by the system or
for other physical conditions related to engineering feasibility and
costs of complying with MCLs, as considered appropriate by EPA. The
State may not issue a variance to a system until it determines that an
unreasonable risk to health (URTH) does not exist. When a State grants
a variance, it must at the same time prescribe a schedule for (1)
compliance with the NPDWR and (2) implementation of any additional
control measures.
Under Section 1416(a), the State may exempt a public water system
from any MCL or treatment technique requirement if it finds that (1)
due to compelling factors (which may include economic factors), the
system is unable to comply, (2) the system was in operation on the
effective date of the MCL or treatment technique, or, for a newer
system, that no reasonable alternative source of drinking water is
available to that system, and (3) the exemption will not result in an
unreasonable risk to health. Under section 1416(b), at the same time it
grants an exemption, the State is to prescribe a compliance schedule
and a schedule for implementation of any required interim control
measures. The final date for compliance may not exceed three years
after the initial date of issuance unless the public water system
establishes that: (1) the system cannot meet the standard without
capital improvements which cannot be completed within the period of
such exemption; (2) the system has entered into an agreement to obtain
financial assistance for necessary improvements; or (3) the system has
entered into an enforceable agreement to become part of a regional
public water system. For systems which serve 500 or fewer service
connections and which need financial assistance to come into
compliance, the State may renew the exemption for additional two-year
periods if the system is taking all practicable steps to meet the above
requirements.
For exemptions resulting from a NPDWR promulgated after June 19,
1986, the system's final compliance date must be within 12 months of
issuance of the exemption. However, the State may extend the final
compliance date for up to three years if the public water system shows
that capital improvements to meet the MCL or treatment technique
requirement cannot be completed within the exemption period and, if the
system needs financial assistance for the improvements, it has an
agreement to obtain this assistance or the system has an enforceable
agreement to become part of a regional public water system. For systems
that have 500 or fewer service connections that need financial
assistance to comply with the MCLs, the State may renew the exemption
for additional two-year periods if the system is taking all practicable
steps to comply.
C. Primacy
As indicated above, States, territories, and Indian Tribes may
assume primary enforcement responsibility (primacy) for public water
systems under Section 1413 of the SDWA. To date, 55 States and
territories have primacy. To assume or retain primacy, States,
territories, or Indian Tribes need not adopt the MCLGs but must adopt,
among other things, NPDWRs (i.e., MCLs, monitoring, analytical, and
reporting requirements) that are no less stringent than those EPA
promulgates.
D. Monitoring, Quality Control, and Records
Under Section 1401(1)(D) of the Act, NPDWRs are to contain
``criteria and procedures to assure a supply of drinking water which
dependably complies with such maximum contaminant levels; including
quality control and testing procedures to insure compliance with such
levels * * *.''
E. Public Water Systems
Public water systems are defined in section 1401 of the Act as
those systems which provide piped water for human consumption and have
at least 15 connections or regularly serve at least 25 people. By
regulation EPA has divided public water systems into community,
nontransient noncommunity, and (transient) noncommunity water systems.
Community water systems (CWSs) serve at least 15 service connections
used by year-round residents or regularly serve at least 25 year-round
residents (40 CFR 141.2). Nontransient noncommunity water systems
(NTNCWSs) regularly serve at least 25 of the same people over six
months of the year. Schools and factories which serve water to 25 or
more of the same people for six or more months of the year are examples
of NTNCWSs. Transient noncommunity systems, by definition, are all
other public water systems. Transient noncommunity systems may include,
for example, restaurants, gas stations, campgrounds, and churches.
This rule would apply to all CWSs, all NTNCWSs, and any transient
noncommunity water systems that use chlorine dioxide as a disinfectant
or oxidant.
F. Public Notification
Section 1414(c) of the Act requires the owner or operator of a
public water system which does not comply with an applicable maximum
contaminant level or treatment technique, testing procedure, or Section
1445(a) (unregulated contaminant) monitoring requirement to give notice
to the persons served by the system. Notice must also be given if a
variance or exemption is in effect or the system fails to comply with a
compliance schedule resulting from a variance or exemption. EPA's
public notification regulations are codified at 40 CFR Section 141.32.
Those regulations were amended by EPA on October 28, 1987 (52 FR
41534).
III. Overview of Existing Interim Standard for TTHMS
In 1974, researchers in The Netherlands and the United States
clearly demonstrated that total trihalomethanes (TTHMs) are formed as a
result of drinking water chlorination (Rook, 1974; Bellar et al, 1974).
EPA subsequently conducted surveys confirming widespread occurrence of
TTHMs in chlorinated water supplies (Symons, 1975; USEPA, 1978). During
this time toxicological studies became available which supported the
contention that chloroform, one of the four trihalomethanes, is
carcinogenic in at least one strain of rat and one strain of mouse
(National Academy of Sciences, 1977).
EPA then set an interim maximum contaminant level (MCL) for the
TTHMs of 100 g/l as an annual average in November 1979 (USEPA,
1979). This standard was based on the need to balance the requirement
for continued disinfection of water to reduce exposure to pathogenic
microorganisms while simultaneously lowering exposure to animal
carcinogens like chloroform.
The interim TTHM standard only applies to systems serving at least
10,000 people that add a disinfectant (oxidant) to the drinking water
during any part of the treatment process. At their discretion, States
are allowed to extend coverage to smaller size systems. About 80
percent of the smallest systems are served by groundwater systems that
are mostly low in THM precursor content (USEPA, 1979).
The proportion of these small groundwater systems that use chlorine
is less than that of large systems; currently, less than half of these
systems disinfect. Also, the shorter hydraulic detention and chlorine
contact times in the small system distribution systems results in lower
TTHM concentrations. Therefore, drinking water systems serving less
than ten thousand people are less likely to have high concentrations of
TTHMs.
Moreover, these small systems are most likely to have greater risks
of significant microbiological contamination, especially if they reduce
or eliminate chlorination. In 1979, the majority of outbreaks
attributable to inadequate disinfection occurred in small systems.
Further, small systems have limited or no access to the financial
resources and technical expertise needed for TTHM control. Therefore,
EPA concluded that small system resources would best be spent on
maintaining and improving microbiological quality and safety. The
revised drinking water regulations now under consideration will extend
to these small systems as required by the Safe Drinking Water Act
Amendments of 1986 (P.L. 99-339, 1986). EPA will also be considering
disinfection as a treatment technique requirement and maximum
contaminant levels (MCLs) for the residual disinfectants. The impacts
these requirements will have on small systems is an important component
of the regulation development process.
Technology Basis for the Interim TTHM Standard
When an MCL is established for TTHMs or any other contaminant that
can be measured, EPA is not required to specify any particular method
for achieving that standard. Instead, the requirement for the interim
regulations was to set an MCL which could be achieved using technology
generally available in 1974. Three general control alternatives were
available:
(1) use of a disinfectant (oxidant) that does not generate (or produces
less) THMs in water;
(2) treatment to lower precursor concentrations prior to chlorination;
and
(3) treatment to remove THMs after their formation.
There are many possible choices among these broad options and in
some cases a combination of approaches might be necessary. The ultimate
choice was left up to the water supplier based on its individual
circumstances.
EPA's evaluation led to the following conclusions concerning
generally available technologies for setting the TTHM MCL:
(1) alternate oxidants like ozone, chloramines, and chlorine dioxide
are available;
(2) precursor removal strategies like changing the point of
disinfection, off-line raw water storage, and improved coagulation are
available; and,
(3) precursor removal using granular activated carbon (GAC) as a
replacement for existing filter media with a regeneration frequency of
one year is feasible as well as biologically activated carbon (ozone
plus GAC) with a regeneration frequency of every two years.
Three conditions concerning modifications of disinfection processes
were also proposed by EPA:
(1) the total quantity of chlorine dioxide added during the treatment
process should not exceed 1 mg/l;
(2) chloramines should not be utilized as a primary disinfectant; and
(3) monitoring for heterotrophic plate count bacteria (HPC) should be
conducted as determined by the State, but at least every day for a
minimum of one month prior to and six months subsequent to the
modifications.
These recommendations concerning disinfection, although useful,
were deleted from the final regulation to allow States greater
discretion. The basis for the MCL became alternate oxidants and
precursor removal.
Technology Basis for Variances
Later, in 1983, EPA promulgated regulations specifying best
technology generally available for obtaining variances (USEPA, 1983). A
variance is granted by the State when a system has installed the best
technology generally available as specified in the regulation and still
cannot meet the MCL. The best technologies generally available for
variances to the TTHM MCL are:
(1) Use chloramines as an alternate or supplemental disinfectant or
oxidant.
(2) Use chlorine dioxide as an alternate or supplemental disinfectant
or oxidant.
(3) Improve existing clarification for THM precursor reduction.
(4) Move the point of chlorination to reduce TTHM formation and, where
necessary, substituting for the use of chlorine as a pre-oxidant
chloramines, chlorine dioxide, or potassium permanganate.
(5) Use of powdered activated carbon for THM precursor or TTHM
reduction seasonally or intermittently at dosages not to exceed 10 mg/l
on an annual average basis.
EPA also identified Group II technologies, which are not
``generally available,'' but may be available to some systems:
(1) Introduction of off-line water storage for THM precursor reduction.
(2) Aeration for TTHM reduction, where geographically and
environmentally appropriate.
(3) Introduction of clarification where not currently practiced.
(4) Consideration of alternative sources of raw water.
(5) Use of ozone as an alternate or supplemental disinfectant or
oxidant.
Note that GAC and BAC are not mentioned as either Group I or Group
II technologies even though they were discussed as technologies for
standard setting purposes (USEPA, 1979). EPA concluded in its cost and
technologies document for the removal of trihalomethanes from drinking
water that (USEPA, 1981):
(1) GAC in the sand replacement mode of operation is often
inappropriate due to the short performance life and high frequency of
regeneration required to achieve substantial TTHM or THM-formation
potential reduction;
(2) the finding took into consideration costs, but primarily was made
due to the complexities of the modifications to prior unit operations,
i.e., disinfection, and in the logistics of the carbon replacement;
(3) greater operating, maintenance, and monitoring than for other
treatments; and
(4) on-site regeneration had only been demonstrated at one U.S. site.
Thus, EPA decided to defer the decision to include GAC and BAC as
best generally available technology for granting variances under the
Safe Drinking Water Act Amendments of 1974.
EPA also did not list ozonation as being ``generally available''
because:
(1) lack of experience in the U.S.;
(2) mixed results in experimental studies; and
(3) most States require a residual in the distribution system which is
not obtainable with ozone.
Thus, EPA decided to defer the decision to include ozone as best
generally available technology for granting variances under the Safe
Drinking Water Act Amendments of 1974.
Economic Impacts of the Interim Standard
Currently, there are 2,700 community water supply systems serving
at least 10,000 people required to comply with the interim TTHM
regulation. In 1988, a survey of large systems found that, on average,
the MCL of 0.10 mg/l had reduced the concentration of TTHMs by 40 to 50
percent (McGuire and Meadows, 1988). Of these, 33 were in violation of
the standard in FY88 (average=115 g/l, range=108-180
g/l). However, by FY92, only nine systems (a decrease of 73
percent) violated the requirements, for a total of 14 violations. Seven
of the nine violating systems and 12 of the 14 violations occurred in
systems serving 10,000 to 50,000 people. This indicates that even when
systems violate, they are able to return to compliance after one or two
violations of the running annual average.
In 1979, approximately 500 systems were estimated to exceed 100
g/l TTHMs. Most of these were able to come into compliance
with minor modifications of chlorination practices. A smaller portion
used alternate oxidants like chlorine dioxide and chloramines. No
system installed ozone or GAC to meet the interim TTHM regulations.
Compliance with the interim TTHM standard involved estimated capital
expenditures of between $31 million and $102 million and yearly
operating and maintenance costs of between $8 million and $29 million
for systems required to comply with the TTHM MCL (i.e., community water
systems serving a population of at least 10,000 people) (McGuire and
Meadows, 1988).
IV. Overview of Preproposal Regulatory Development
A. October 1989 Strawman Rule
1. Purpose. EPA was required to develop rules for additional
contaminants under the 1986 Amendments to the Act. In order to solicit
public comment in developing a rule, EPA released a strawman rule
(preproposal draft) in October 1989. A strawman was used because of the
complexity of the problem, the large amount of (occasionally
contradictory) information, and the ability to reorient the rule
approach based on public comment or new data. In this strawman, EPA
included a lead option of setting MCLGs and MCLs for TTHMs, haloacetic
acids, chlorine, chloramines, chlorine dioxide, chlorite, and chlorate.
The Agency also identified potential add-on compounds: chloropicrin,
cyanogen chloride, hydrogen peroxide, bromate, iodate, and
formaldehyde. Some of these compounds could also conceivably be used as
surrogate monitoring compounds for the compounds identified in Table
IV-1 below.
Additional Candidate Byproduct Compounds
------------------------------------------------------------------------
Chlorination byproducts Ozonation byproducts
------------------------------------------------------------------------
--Individual THMs: chloroform, --Aldehydes: acetaldehyde,
bromodichloromethane, hexanal, heptanal.
dibromochloromethane, bromoform.
--Individual haloacetic acids: mono-, --Organic acids.
di-, and trichloroacetic acids; mono- --Ketones.
and dibromoacetic acids. --Epoxides.
--Individual haloacetonitriles: di- --Peroxides.
and trichloroacetonitrile; --Nitrosamines.
bromochloroacetonitrile,
dibromoacetonitrile.
--Haloketones: 1,1 di- and 1,1,1-
trichloropropanone.
--Chlorophenols: 2-; 2,4-di; and --N-oxy compounds.
2,4,6-trichlorophenol. --Quinones.
--Others: chloral hydrate, N- --Bromine substituted compounds.
organochloroamines.
------------------------------------------------------------------------
In addition, the strawman provided that EPA would set treatment
technique requirements or provide guidance for control of the
following: MX, as a surrogate for mutagenicity; total oxidizing
substances, as a surrogate for organic peroxides and epoxides; and
assimilable organic carbon, as a surrogate for microbiological quality
of oxidized waters. Monitoring parameters based on the particular
disinfection process were also identified.
As BAT, EPA included precursor removal (conventional treatment
modifications, GAC of up to 30 minute duration and three months
regeneration), alternate oxidants (ozone plus chloramines, chlorine
dioxide with chlorite removal plus chloramines), and byproduct removal
(aeration, GAC adsorption, reducing agents, AOC removal). Each of the
options had problems. GAC was not universally applicable to all waters
for either precursor removal or DBP removal. Membranes were not
included as BAT because of lack of full-scale experience.
As lead options, EPA included a TTHM MCL of 25 or 50 g/l
and other MCLs based on feasibility analyses similar to those that
would be used to develop the TTHM MCL.
2. Summary of Public Comments. Several commentors expressed a
desire for EPA to look at coordination of requirements with those for
other regulations, including issues such as requirements for
maintenance of distribution system disinfectant residuals and system
optimization for multiple contaminants. Many commentors were concerned
about the lack of health data and the interpretation of existing data.
Many system operators were also concerned about the effects of
modifying their treatment processes to meet DBP MCLs. These concerns
included lowered microbiological protection, creation of conditions
that favored distribution system microbiological growth (e.g., use of
ozone would create biodegradable organics and use of chloramines would
create a nitrogen source), and creation of other environmental problems
when changing treatment (e.g., residual handling with precursor removal
and GAC regeneration). While commentors expressed concern about use of
alternate disinfectants, several offered to provide data and others
recommended epidemiological studies in systems with long histories of
alternative disinfectant use.
B. June 1991 Status Report on D/DBP Rule Development
1. Purpose and transition from Strawman Rule. EPA published a
status report on the development of D/DBPR in June 1991 that was
designed to indicate the Agency's thinking on rule criteria. The status
report indicated that EPA was considering extending coverage under the
rule to all nontransient systems (instead of just those serving at
least 10,000 people, as under the 1979 TTHM rule) and proposing a
shorter list of compounds for regulation than were included in the 1989
strawman. The 1991 list included disinfectants (chlorine, chloramines,
and chlorine dioxide), THMs, haloacetic acids, chloral hydrate,
bromate, chlorite, and chlorate). For both THMs and haloacetic acids,
three options were included: MCLs for individual compounds, a single
MCL for the total, and a combination of the two. Individual MCLs were
considered because health risks for compounds differed, in some cases
significantly. The total MCL was considered because of the precedent
established in the 1979 TTHM rule and to act as a surrogate to limit
other DBPs for which the Agency lacked adequate health effects and/or
occurrence data.
The list of compounds was shorter than that in the 1989 strawman
for several reasons. Several compounds were deleted because they did
not appear to pose significant health effects at levels present in
drinking water (e.g., haloacetonitriles, chloropicrin). Others were
deleted because the health risks were not expected to be adequately
characterized in time for rule proposal (e.g., certain aldehydes and
organic peroxides), although it was noted that these compounds might be
regulated in the future when more data became available.
2. Major issues. In the status report, EPA identified several major
issues that needed to be considered as the D/DBP rule was developed.
The first was that of trade-offs with microbial and DBP risks. The goal
was to ensure that the water remained microbiologically safe at the
level that disinfectant and DBP MCLs were set. The discussion raised
questions regarding uncertainties in defining microbial and DBP risks,
levels of risks that would be considered acceptable and at what cost,
and defining practical (implementable) criteria to demonstrate that an
achievable risk had been reached.
The second issue was the use of alternate disinfectants to limit
chlorination byproducts. The Agency recognized that while alternate
disinfection schemes (e.g., ozone and chloramines) could greatly reduce
byproducts typical of chlorination, little was known about the
byproducts of the alternate disinfectants and their associated health
risks. EPA did not want to promulgate a standard that encouraged the
shift to alternate disinfectants unless the associated risks (including
both those from byproducts and differential microbial risks from a
change in disinfectants) were adequately understood.
The third issue was integration with the Surface Water Treatment
Rule. Although the rule only mandated 3-log removal or inactivation of
Giardia and 4-log of viruses, EPA guidance recommended higher levels
for poorer quality source waters. EPA was concerned that systems would
reduce microbial protection to levels nearer to the regulatory
requirements by reducing disinfection and possibly greatly increase
microbial risks in an effort to meet DBP MCLs. The Agency wanted to
ensure adequate microbial protection while reducing risk from DBPs.
The last issue was the best available technology. The BAT defined
would determine the levels at which MCLs were set. For example,
allowing alternate disinfectants as BAT would drive the chlorination
byproduct MCLs down, but could result in increased exposure to (not
well characterized) alternate byproducts. EPA believed that it
therefore might be appropriate to define chlorine and a precursor
removal technology as BAT.
To address these issues, EPA suggested two possible regulatory
strategies. One was to define the MCL(s) based on what was possible to
achieve using the most effective DBP precursor removal strategy as BAT
(e.g., GAC or membrane filtration). While installing such precursor
removal technology might minimize health concerns, the costs would be
substantial (without finding out if other less costly technologies,
such as use of alternative disinfectants, provided similar benefits).
Also, since systems are not required to install BAT to meet MCLs, EPA
believed that many systems would attempt to meet the MCLs by lower-cost
alternative disinfectants (ozone, chloramines, chlorine dioxide). Since
health effects for alternative disinfectant byproducts are not
adequately characterized, risks may not be reduced.
The second strategy was a two-phase regulation, with the first
phase designed to address risks using lower cost options during
concurrent efforts to obtain more data on treatment alternatives and
health effects of compounds not currently adequately characterized.
This strategy would prevent major shifts into use of new treatment
technology until the full consequences of such shifts (both costs and
benefits) are better understood.
3. Suggested monitoring scenario. In its fact sheet accompanying
the status report, EPA recommended that routine TTHM and haloacetic
acid monitoring for systems serving at least 10,000 people have the
same monitoring requirements as were in the 1979 TTHM rule. Smaller
systems would have less frequent monitoring requirements, but would
have compliance based on worst-case samples. EPA included provisions
for reduced monitoring (compliance based on worst-case samples or
surrogate monitoring), waiver criteria, and requirements for
disinfectant and other DBP monitoring.
4. Summary of public comments. EPA received comments on the status
report from numerous parties. Many commentors agreed with EPA's
concerns with issues such as alternative disinfectant DBPs and
balancing microbial and DBP risks. Several commentors supported the
two-phase regulatory approach, but expressed concern about timing.
Others recommended that DBP MCLs not be set so low as to force many
systems to install expensive technology or decrease microbial
protection. Several commentors were concerned with the availability of
both analytical methods and certified laboratories for the low levels
that were being considered. One commentor recommended that EPA make it
clear that MCLs set for disinfectants should allow temporary high
levels to address distribution system microbiological problems.
Finally, many commentors supported allowing reduced monitoring wherever
possible.
C. Initiation of the Regulatory Negotiation Process
EPA became interested in pursuing a negotiated rulemaking process
for the development of the D/DBP rule, in large part, because no clear
path for addressing all the major issues identified in the June 1991
Status Report on D/DBP rule was apparent. EPA's most significant
concern was developing regulations for DBPs while also ensuring that
adequate treatment be maintained for controlling microbiological
concerns. A negotiated rule process would help people understand the
complexities of the risk-risk tradeoff issue and, hopefully, reach a
consensus on the most appropriate regulation to address concerns from
both DBPs and microorganisms.
It also appeared to EPA that the criteria for initiating a
negotiated rule under the Negotiated Rulemaking Act of 1990 for
establishing a negotiated rulemaking could be met. These include:
(1) there is a need for a rule,
(2) there are a limited number of identifiable interests that will
be significantly affected by the rule,
(3) there is a reasonable likelihood that a committee can be
convened with a balanced representation of persons who--
(A) can adequately represent the interests identified under
paragraph (2); and
(B) are willing to negotiate in good faith to reach a consensus on
the proposed rule,
(4) there is a reasonable likelihood that a committee will reach a
consensus on the proposed rule within a fixed period of time,
(5) the negotiated rulemaking procedure will not unreasonably delay
the notice of proposed rulemaking and the issuance of a final rule,
(6) the Agency has adequate resources and is willing to commit such
resources, including technical assistance, to the committee, and
(7) the Agency, to the maximum extent possible consistent with the
legal obligations of the Agency, will use the consensus of the
committee with respect to the proposed rule as the basis for the rule
proposed by the Agency for notice and comment.
In 1992 EPA hired a contractor, Resolve, which added a
subcontractor, Endispute, to assess the feasibility and usefulness of
convening a negotiated rulemaking. Resolve and Endispute conducted more
than forty interviews during the summer of 1992 with representatives of
State and local health and regulatory agencies, water suppliers,
manufacturers of equipment and supplies used in drinking water
treatment, and consumer and environmental organizations. These
interviews revealed that:
(1) The entities interested in or affected by the rulemaking were
readily identifiable and relatively few in number.
(2) The rulemaking required resolution of a limited number of
interdependent issues, about which there appeared to be a sufficiently
well-developed factual base to permit meaningful discussion. Further,
there appeared to be several ways to resolve these issues, providing a
potential basis for productive joint problem-solving.
(3) The parties expressed some common goals, along with an
unusually strong degree of good faith interest in resolving the issue
through negotiation.
(4) The Agency had adequate staff and technical resources and was
willing to commit such resources to the negotiated rulemaking.
Resolve and Endispute recommended to EPA that the negotiated
rulemaking proceed. EPA concurred with this recommendation.
However, it was also noted that reaching consensus on the proposed
rule would be a challenge. The interviews revealed that parties
differed in their perceptions about the nature and magnitude of the
risks associated with DBPs, and many expressed strong doubts about the
adequacy of available scientific and technical information. Moreover,
some parties stated that marginal improvements in disinfection
technology were all that should be done until the relative risks are
better understood, while others said that a fundamentally new approach
focusing on precursor reduction should be considered.
EPA published a notice of intent to proceed with a negotiated
rulemaking on September 15, 1992 (57 FRN 42533), proposing 17 parties
to be Negotiating Committee members. In general, comments indicated
very positive support for the negotiated rulemaking.
As part of the convening process, an organizational meeting was
held September 29-30, 1993. Participants discussed Negotiating
Committee composition and organizational protocols. Between comments
expressed at the meeting and submitted in writing, eleven additional
parties--including water suppliers not substantially represented by the
Committee's original proposed membership, and chemical and equipment
suppliers--asked to be added to the Committee. In addition,
participants discussed the need to develop accurate scientific and
technical information.
On November 13, 1992, EPA published a notice of establishment for
the Negotiating Committee (57 FRN 53866), and an 18th member was added
to the Negotiating Committee.
Based on comments received at the organizational meeting, a
Technical Workshop was organized and conducted on November 4-5, 1992.
Composed of presentations and panel discussion by 23 of the Nation's
leading experts on drinking water treatment, the workshop provided
participants with opportunities to familiarize themselves with the
technical elements in this rulemaking and to explore the range of
scientific opinions about: (1) The nature and magnitude of potential
health effects from exposure to DBPs and microbial contaminants in
drinking water, (2) available information on the cost and efficacy of
precursor removal and drinking water disinfection technologies, and (3)
EPA's efforts to model and compare chemical and microbial risks in
drinking water.
Additional presentations were given throughout the rulemaking
process, as new information became available and more questions were
raised by participants.
At the first formal negotiating session, on November 23-24, 1992,
participants formed a technologies working group (TWG) to develop
reliable and consistent information about the cost and efficacy of
drinking water treatment technologies. This approach provided a forum
for participants to arrive at a shared understanding of complex issues
in the rulemaking, setting a cooperative tone for the rest of their
discussions. The working group, which continued to meet throughout the
rulemaking, also provided a formal opportunity for input from the
chemical and equipment suppliers who had not been named to the
Committee.
In addition, three experts were hired through EPA's contract with
Resolve to provide ongoing scientific advice and technical support to
participants in the Committee and on the technologies working group,
principally for members without access to similar resources within
their own organizations.
Based on scientific data presented and discussed through the
November 23-24 meeting, participants agreed that some type of DBP Rule
was warranted.
The Committee developed and reached agreement on criteria for a
``good'' DBP Rule at the September 29-30 and November 23-24 meetings. A
good rule is one which would be flexible and affordable and would
protect public health from chemical and microbial risks. It was noted
that limiting some DBPs could encourage changes in treatment that might
increase the formation of other DBPs, or compromise protections against
microbial contaminants.
Next, Committee members and other participants were invited to
present regulatory options as a starting point for further discussion.
Sixteen options were introduced at the December 17-18 meeting, and
discussed at the meeting on January 13-14, 1993. These were merged into
three consolidated options at the January 13-14 meeting, and discussion
continued at the meeting on February 9-10. At this point, areas of
disagreement included:
(1) Whether to regulate DBPs through Maximum Contaminant Levels
(MCLs) or through a treatment technique (i.e., by exceeding DBP
``action levels,'' systems would trigger additional steps to minimize
chemical and microbial risks).
(2) Whether to minimize formation of the DBPs about which
relatively little is known by establishing a regulatory limit for their
naturally occurring organic precursors (e.g., Total Organic Carbon, or
TOC) in the water prior to the point of disinfection.
(3) Whether to provide greater protection against microbial
contaminants in drinking water, in conjunction with new DBP limits, by
developing an enhanced Surface Water Treatment Rule (ESWTR).
(4) Whether to develop a second round of DBP controls along with
the first (assuring broad improvements in drinking water quality), or
to wait until better scientific information becomes available.
Concurrently, the TWG modelled systems' potential compliance
choices under several regulatory scenarios, and presented revised
household and national compliance cost estimates at several meetings.
Using a ``strawman'' developed from the consolidated options by EPA
staff as the starting point for negotiation, the Committee worked out
an ``agreement in principle'' on the first round of DBP controls at its
February 24-25 meeting. The ``Stage 1'' agreement set MCLs for
trihalomethanes and haloacetic acids--two principal classes of
chlorination by-products--at levels the Committee deemed protective of
public health, based on current information: 80 and 60 micrograms per
liter, respectively. To limit DBP precursors, the Committee agreed to
develop a series of ``enhanced coagulation'' requirements, to vary
according to systems' influent water quality and treatment plant
configurations. Members also agreed to reconvene in several years to
develop a second stage of DBP regulations, when the results of more
health effects research and water quality monitoring are available. In
addition, members agreed that more expeditious changes to the rules may
be necessary if additional information becomes available on short- term
or acute health effects of DBPs. Members also agreed that, if data on
short-term or acute health effects warrant earlier action, a meeting
shall be convened to review the results and to develop recommendations.
A drafting group was named at the February 24-25 meeting. Assisted
by the TWG, these members drafted an ``agreement in principle'' for
presentation and discussion at the March 18-19 meeting. Using ``straw''
provisions from the facilitators, the Committee devised a regulatory
``backstop'' (i.e., Stage 2 MCLs of 0.040 mg/l for TTHMs and 0.030 mg/l
for HAA5 for surface water systems serving at least 10,000 people) at
this meeting to assure participants favoring further DBP controls that
other members would return for the ``Stage 2'' negotiation. The
Committee also agreed to recommend that EPA propose several ESWTR
options for comment, developed a collaborative process to guide the
health effects research program, and agreed to formulate short-term
water quality and technical data collection provisions within an
Information Collection Rule.
Based on the discussion to this point, one member withdrew from the
Committee at the March 18-19 meeting.
The drafting group presented regulatory language for the DBP Rule,
ESWTR, and ICR at each of the Committee's last two meetings, held May
12-13 and June 22-23, 1993. These texts provided a framework for
further discussion and resolution of remaining issues, including:
limits for residual disinfectants and individual by-products; public
notification and affordability provisions; and timing, applicability,
and conditions under which systems might qualify for exceptions from
various requirements. Committee members agreed to reserve their rights
to comment on the draft preambles.
The drafting group continued working through the summer of 1993,
and revisions to each of the rules and their preambles were mailed to
the Committee for comment on July 8, 1993, September 8, 1993, February
8, 1994, and May 12, 1994. Each member had signed the agreement by June
7, 1994.
Unless otherwise noted, EPA has adopted the recommendations of the
Negotiating Committee and its Technologies Working Group and reflects
those recommendations in the following preamble and proposed
regulations.
V. Establishing MCLGs
A. Background
1. MCLGs and MCLs Must Be Proposed and Promulgated Simultaneously
Congress revised the Safe Drinking Water Act in 1986 to require
that MCLGs and National Primary Drinking Water Regulations (NPDWRs) be
proposed simultaneously and promulgated simultaneously [SDWA section
1412 (a)(3)]. Simultaneous promulgation was intended to streamline the
development of drinking water regulations.
2. How MCLGs Are Developed
MCLGs are set at concentration levels at which no known or
anticipated adverse health effects occur, allowing for an adequate
margin of safety. Establishment of an MCLG for each specific
contaminant depends on the evidence of carcinogenicity from drinking
water exposure or an assessment for adverse noncarcinogenic health
effects.
a. MCLG Three Category Approach. EPA currently follows a three-
category approach in developing MCLGs for drinking water contaminants
(Table V-1).
Table V-1.--EPA'S Three-Category Approach for Establishing MCLGs
------------------------------------------------------------------------
Evidence of
Category carcinogenicity via MCLG approach
drinking water\1\
------------------------------------------------------------------------
I...................... Strong evidence Zero.
considering weight of
evidence,
pharmacokinetics,
potency and exposure.
II..................... Limited evidence RfD approach with
considering weight of added safety margin
evidence, of 1 to 10 or 10-5 to
pharmacokinetics, 10-6 cancer risk
potency and exposure. range.
III.................... Inadequate or no animal RfD approach.
evidence.
------------------------------------------------------------------------
\1\Considering oral exposure data such as drinking water, dietary and
gavage studies.
Each chemical is evaluated for evidence of carcinogenicity from
drinking water. For volatile contaminants, inhalation data are also
considered. EPA takes into consideration the overall weight of evidence
for carcinogenicity, pharmacokinetics, potency and exposure.
EPA's policy is to set MCLGs for Category I contaminants at zero.
The MCLG for Category II contaminants is calculated by using the
Reference dose (RfD) approach (described below) with an added margin of
safety to account for possible cancer effects. If adequate data are not
available to calculate an RfD, then the MCLG is based on a cancer risk
level of 10-5 to 10-6. MCLGs for Category III contaminants
are calculated using the RfD approach.
Category I contaminants are those for which EPA has determined that
there is strong evidence of carcinogenicity from drinking water. The
MCLG for Category I contaminants is set at zero because it is assumed,
in the absence of other data, that there is no threshold dose for
carcinogenicity. In the absence of route specific (e.g., oral) data on
the potential cancer risk from drinking water, chemicals classified as
Group A or B carcinogens (see section c below) are generally placed in
Category I.
Category II contaminants include those contaminants for which EPA
has determined that there is limited evidence of carcinogenicity from
drinking water, considering weight of evidence, pharmacokinetics,
potency, and exposure. In the absence of route specific data, chemicals
classified in Group C (see section c below) are generally placed in
Category II.
For Category II contaminants, one of two options have traditionally
been used to set the MCLG. The first option sets the MCLG based upon
noncarcinogenic endpoints of toxicity (the RfD), then applies an
additional safety factor of 1 to 10 to the MCLG to account for possible
carcinogenicity. An MCLG set by the option 1 approach is compared with
the cancer risk, if quantified. The second option is to set the MCLG
based upon a theoretical lifetime excess cancer risk level of 10-5
to 10-6 using a conservative mathematical extrapolation model. EPA
generally uses the first option; however, the second approach is used
when valid noncarcinogenic data are not available to calculate an RfD
and adequate experimental data are available to quantify the cancer
risk.
Category III contaminants include those contaminants for which
there is inadequate or no evidence of carcinogenicity from drinking
water. If there is no additional information to consider, contaminants
classified in Group D or E (see section c below) are generally placed
in Category III. For these contaminants, the MCLG is established using
the RfD approach.
b. Assessment of Noncancer Health Effects. The risk assessment for
noncancer health effects can be characterized by a Reference Dose
(RfD). The oral RfD (expressed in mg/kg/day) is an estimate, with
uncertainty spanning perhaps 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- or lowest-observed-adverse-
effect level (called a NOAEL or LOAEL, respectively) that has been
identified from a subchronic or chronic study of humans or animals. The
NOAEL or LOAEL is then divided by an uncertainty factor(s) to derive
the RfD. Although the RfD is represented as a point estimate, it is
actually a range since the RfD is a number with an inherent uncertainty
of an order of magnitude.
Uncertainty factors are used to estimate the comparable ``no-
effect'' level for a large heterogeneous human population. The use of
uncertainty factors accounts for several data gaps including intra- and
inter-species differences in response to toxicity, the small number of
animals tested compared to the size of the population, sensitive
subpopulations and the possibility of synergistic action between
chemicals (see 52 FR 25690 for further discussion on the use of
uncertainty factors).
EPA has established certain guidelines (shown below) to determine
how to apply uncertainty factors when establishing an RfD (USEPA,
1986).
Use a 1- to 10-fold factor when extrapolating from valid
experimental results from studies in average healthy humans. This
factor is intended to account for the variation in sensitivity among
the members of the human population.
Use an additional 10-fold factor 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 the case of humans.
Use an additional 10-fold factor 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 10-fold factor 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.
An additional uncertainty factor may be used according to
scientific judgment when justified.
Use professional judgment to determine another uncertainty
factor (also called a modifying factor, MF) that is greater than zero
and less than or equal to 10. The magnitude of the MF depends upon the
professional assessment of scientific uncertainties of the study and
data base not explicitly treated above, e.g., the completeness of the
overall data base and the number of species tested. The default value
for the MF is 1.
To determine the MCLG, the RfD is adjusted by the body weight of
the protected (or most sensitive) individual (usually a 70 kg adult),
average volume of water consumed daily over a lifetime (2 L/day for an
adult) and exposure to the contaminant from a drinking water source
(relative source contribution or RSC).
Generally, EPA assumes that the RSC from drinking water is 20
percent of the total exposure, unless other exposure data for the
chemical are available [see 54 FR 22069 and 56 FR 3535]. When adequate
data are available and the data indicate that drinking water exposure
contributes between 20 and 80 percent of total exposure, EPA uses the
actual percentage to determine the MCLG, as is indicated by equation
(3), below. When data indicate that contributions from drinking water
are between zero and 20 percent, or between 80 and 100 percent, EPA
utilizes a 20 percent floor and an 80 percent ceiling, respectively.
The calculations below express the derivation of the MCLG based on
noncancer health effects:
TP29JY94.000
c. Assessment of Carcinogenic Health Effects. For chemicals
suspected of being carcinogenic to humans, the assessment for non-
threshold toxicants consists of the weight of evidence of
carcinogenicity in humans, using bioassays in animals and human
epidemiological studies as well as information that provides indirect
evidence (i.e., mutagenicity and other short-term test results). The
objectives of the assessment are to determine the level or strength of
evidence that the substance is a carcinogen and to provide an
upperbound estimate of the possible risk of human exposure to the
substance in drinking water. A summary of EPA's general carcinogen
classification scheme is (USEPA, 1986):
Group A--Human carcinogen based on sufficient evidence from
epidemiological or other human studies.
Group B--Probable human carcinogen based on limited evidence of
carcinogenicity in humans (Group B1) or based on sufficient evidence in
animals with inadequate or no data in humans (Group B2).
Group C--Possible human carcinogen based on limited evidence of
carcinogenicity in animals in the absence of human data.
Group D--Not classifiable based on lack of data or inadequate
evidence of carcinogenicity from animal data.
Group E--No evidence of carcinogenicity for humans (no evidence for
carcinogenicity in at least two adequate animal tests in different
species or in both epidemiological and animal studies).
d. MRDLGs--appropriateness of a new concept? As stated in section
II.A of this preamble, EPA is proposing a new term, ``maximum residual
disinfectant level goal'' (MRDLG), in lieu of MCLGs for all
disinfectants because disinfectants are intentionally added to drinking
water as a treatment technique to kill disease-causing microorganisms.
The proposal of this concept was agreed to through the negotiated
rulemaking process.
Certain members of the Negotiating Committee were concerned that if
``MCLGs,'' which included the term ``contaminant,'' were set for
disinfectants, water treatment plant operators might be reluctant to
apply disinfectant dosages above the MCLG during short periods of time
to control for microbial risk, even though such exposure to elevated
disinfectant concentration levels would pose little or no risk. For
example, NOAELs for chlorine and chloramines are based upon animal
studies following long term exposure to high levels of the
disinfectants in drinking water. Short-term exposures at elevated
levels would not be a concern (see the following discussion on health
effects for chlorine and chloramines). During emergency situations such
as distribution system pipe breaks or significant fluctuations in
source water quality, systems will on occasion need to apply short term
disinfectant residual concentrations of chlorine or chloramines, well
above the regulatory goal, to protect from waterborne disease.
The MRDLGs are developed in the same way as MCLGs. EPA solicits
comment on the appropriateness of adopting the term ``MRDLG'' in lieu
of MCLGs for disinfectants in the final rule.
B. Proposed MRDLGs and MCLGs
The following includes a summary of the health effects information
available for each disinfectant or by-product. These summaries are
taken from more complete and comprehensive descriptions of the data
given in the cited Health Criteria Documents that have been developed
for each of these chemicals. These documents are available in the water
docket.
1. Chlorine, hypochlorite ion and hypochlorous acid
The following assessment for both chlorine and chloramines includes
a consideration of available animal data, as well as epidemiology
studies which have been conducted on chlorinated or chloraminated
drinking water. The epidemiology data are discussed in section C of
this preamble.
Chlorine (CAS # 7782-50-5) hydrolyses in water to form hypochlorite
(CAS # as sodium salt 7790-92-3) and hypochlorous acid (CAS # 7681-52-
9). Because of their oxidizing characteristic and solubility, chlorine
and hypochlorites are used in water treatment to disinfect drinking
water, sewage and wastewater, swimming pools, and other types of water
reservoirs. They are also used for general sanitation and control of
bacterial odors in the food industry.
Chlorine is a highly reactive and water soluble species. The fate,
transport, and distribution of chlorine in natural waters is not well
understood. Much of the available information comes from the addition
and oxidation reactions with inorganic and organic compounds known to
occur in aqueous solutions. Factors such as reactant concentrations,
pH, temperature, salinity and sunlight influence these reactions.
Occurrence and Human Exposure. For the purpose of setting an MRDLG,
consideration is given to chlorine levels resulting from disinfection
of drinking water. Chlorine exposures from swimming pools and hot tubs
are not evaluated in determining the MRDLG. Persons who swim frequently
or use a hot tub may have greater dermal and possibly inhalation
exposure to chlorine.
Chlorine is added to drinking water as chlorine gas (Cl2) or
as calcium or sodium hypochlorite. In drinking water, the chlorine gas
hydrolyses to hypochlorous acid and hypochlorite ion and can be
measured as the free chlorine residual. Maintenance of a chlorine
residual throughout the distribution system is important for minimizing
bacterial growth and for indicating (by the absence of a residual)
water quality problems in the distribution system. Currently, maximum
chlorine dosage is limited by taste and odor constraints and for
systems needing to comply with the total trihalomethane (TTHM) standard
regularly. Additionally, for systems using chlorination, the surface
water treatment rule (SWTR) requires a minimum residual of 0.2 mg/L
prior to the entry point to the distribution system, and the presence
of a detectable residual throughout the distribution system.
Table V-2 presents occurrence information available for chlorine in
drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' USEPA 1992a. The
table lists five surveys conducted by Federal, as well as private
agencies. Median concentrations of chlorine in drinking water appear to
range from <1 to 2 mg/L.
Table V-2.--Summary of Occurrence Data for Chlorine
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of chlorine in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (mg/L)
Survey (year)\1\ Location Sample information (No. of ----------------------------------------------------------------
samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
EPA, 1992b\2\ (1987-1991). Disinfection By-Products Finished Water:
Field Studies. At the Plant (71).............. 0.1-5.0 1.7 1.4
Distribution System (45)....... 0.0-3.2 0.7 0.5
AWWARF (1987) McGuire & National Survey........... Finished Water From: (Typical doses).
Meadow, 1988. Lakes.......................... 2.2\3\
Flowing Streams................ 2.3\3\
Ground Waters.................. 1.2\3\
Mixed-supplies................. 1.0\3\
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell 0.3-5.2 1.5 1.0 ........................
1989) Krasner et al., Nationwide. Effluent, 4 Quarters (17).
1989b.
WIDB (1989-1990).......... 228 SW Plants............. Residual Chlorine Provided to 0-3.5 0.937 0.8 ........................
215 GW Plants............. the Average Customer (systems 0-5 0.872 0.325
>50,000 people).
AWWA Disinfection Survey 283 Utilities in the U.S.. Finished Water Entering ........... 0.07-5.0 1.1 ........................
(1991). Distribution System.
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Typical dosage used by treatment plants.
SW: Surface Water.
GW: Ground Water.
AMWA: Association of Metropolitan Water Agencies.
AWWA: American Water Works Association.
AWWARF: American Water Works Association Research Foundation.
CDHS: California Department of Health Services.
EPA: Environmental Protection Agency.
WIDB: Water Industry Data Base.
Exposure to chlorine residual varies both between systems and
within systems. Chlorine residual within systems will vary based on
where customers are located within the distribution system and changes
in the system's disinfection needs over time. Using residual
concentrations from the 1989-1991 AWWA Disinfection Survey and WIDB,
exposure to chlorine due to drinking water can be estimated using a
consumption rate of 2 liters per day. Based on the estimated 25th
percentile and 75th percentile chlorine residuals in the 1991 AWWA
Disinfection Survey, exposure was determined to range from 1.5 to 3.8
mg/day and the median would be 2.2 mg/day. Using the WIDB data,
exposures to the average customer from surface and ground water sources
using chlorination, respectively, were determined to be 1.9 mg/day and
1.7 mg/day.
Little information is available concerning the occurrence of
chlorine in food and indoor air in the United States. The Food and Drug
Administration (FDA) does not analyze for chlorine in foods. However,
there are several uses of chlorine in food production, such as the
disinfection of chicken in poultry plants and the superchlorination of
water at soda and beer bottling plants (Borum, 1991). Therefore, the
possibility exists for dietary exposure to chlorine from its use in
food production. However, monitoring data are not available to
characterize adequately the extent of such potential exposures.
Additionally, preliminary discussions with FDA suggest that there are
no approved uses for chlorine in most foods consumed in the typical
diet. Similarly, EPA's Office of Air and Radiation is not currently
conducting any sampling studies for chlorine in indoor air. Data on
levels of chlorine in ambient air are forthcoming.
Considering the limited number of food groups that are believed to
contain chlorine and that no significant levels of chlorine are
expected in ambient or indoor air, it is anticipated that drinking
water is the predominant source of exposure to chlorine. Air and food
are believed to provide only small contributions, although the
magnitude and frequency of these potential exposures are issues
currently under review. EPA, therefore, is considering setting an MRDLG
for chlorine in drinking water using an RSC value of 80%, the current
exposure assessment policy ceiling. EPA requests any additional data on
known concentrations of chlorine in drinking water, food and air.
Health Effects. The health effects information for chlorine is
summarized from the draft Drinking Water Health Criteria Document for
Chlorine, Hypochlorous Acid and Hyperchlorite Ion (USEPA, 1994a). The
studies cited within this section are summarized in the draft criteria
document.
Chlorine and the 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 by-products (e.g., trihalomethanes). Thus,
the health effects associated with the administration of high levels of
chlorine and/or the hypochlorites in various animal studies may be due
to these reaction by-products and not the disinfectant itself.
Oxidizing species such as chlorine and the hypochlorites are probably
short-lived in biological systems due to both their reactivity and the
large number of organic compounds found in vivo. 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 the water is swallowed.
Oral studies with radiolabeled (i.e., 36Cl) hypochlorite and
hypochlorous acid indicate that, as measured by the radiolabel, these
compounds may be well absorbed and distributed throughout the body with
the highest levels measured in plasma and bone marrow. However,
considering the reactivity of the hypochlorites, these results may only
reflect the presence of reaction by-products (e.g., chloride). The
major route of excretion appears to be urine and then the feces.
Acute oral LD50 values for calcium and sodium hypochlorite
have been reported at 850 mg/kg in rats and 880 mg/kg in mice,
respectively. Humans have consumed hyperchlorinated water for short
periods of time at levels as high as 50 mg/L (1.4 mg/kg) with no
apparent adverse effects.
Short-term oral studies in animals have indicated decreases in
blood-glutathione levels, hemolysis and biochemical changes in liver in
rodents following a gavage dose of hypochlorite in water. No adverse
effects on reproduction (Druckery, 1968) or development were observed
in rats administered chlorine in drinking water at concentrations of
100 mg/L or less. However, Meier et al. (1985) observed an increase in
sperm-head abnormalities in mice receiving hypochlorite at 200 mg/L,
but not at 100 mg/L or less.
No systemic effects were observed in rodents following oral
exposure to chlorine as hypochlorite in distilled water at levels up to
275 mg/L over a 2 year period (NTP, 1990).
Chlorinated water has been shown to be mutagenic to bacterial
strains and mammalian cells. Investigations with rodents to determine
the potential carcinogenicity of chlorine, or chlorinated water have
been negative. In the most recent study, no apparent carcinogenic
potential was demonstrated following oral exposure to chlorine in
distilled drinking water as hypochlorite, at levels up to 275 mg/L over
a 2 year period (NTP, 1990). However, NTP observed a marginal increase
in the incidence of mononuclear cell leukemia in mid-dose female F344
rats but not in male rats or male and female mice (NTP, 1990).
Mononuclear cell leukemia has a high spontaneous rate of occurrence in
female F344 rats. The levels reported in the NTP study are within the
historical control range of incidence for the sex and strain of rat.
EPA believes that mononuclear cell leukemia can not be solely
attributed to exposures to chlorine in drinking water but rather may
reflect the high background rate of mononuclear cell leukemia in the
test species.
EPA has classified chlorine in Group D, not classifiable as to
human carcinogenicity (IRIS, 1993). This classification stems from the
findings of the NTP (1990) study indicating equivocal evidence in
female rats (increased mononuclear cell leukemia) and no evidence in
male rats or male and female mice. The International Agency for
Research on Cancer (IARC, 1991) also evaluated chlorinated drinking
water and hypochlorite for potential human carcinogenicity. IARC
determined that there was inadequate evidence for carcinogenicity of
chlorinated drinking water and hypochlorite salts in humans and
animals. (See section C for a description of these studies.) IARC
concluded that chlorinated drinking water and hypochlorite salts were
not classifiable as to their carcinogenicity to humans and thus
assigned these chemicals to IARC Group 3. This category is similar to
EPA cancer classification Group D.
Based on the previous discussion, EPA is proposing that chlorine,
hypochlorite and hypochlorous acid be placed in Category III for the
purpose of setting an MRDLG. The study selected for determining an RfD
is the previously mentioned 2 year rodent study that was conducted by
the National Toxicology Program (NTP, 1990). In this study, male and
female F344 rats and B6C3F1 mice were given chlorine in distilled
drinking water at levels of 0, 70, 140 and 275 mg/L for 2 years. Based
on body weight and water consumption values, these concentrations
correspond to doses of approximately 0, 4, 7 and 14 mg/kg/day for male
rats; 0, 4, 8, and 14 mg/kg/day for female rats; 0, 7, 14, and 24 mg/
kg/day for male mice and 0, 8, 14 and 24 mg/kg/day for female mice.
There was a dose related decrease in water consumption for both rats
and mice, presumably due to taste aversion. No effect on body weight or
survival were observed for any of the treated animals. Using a NOAEL of
14 mg/kg/d identified from female rats in the NTP (1990) study an MRDLG
of 4 mg/L, based on lack of toxicity in a chronic study is derived as
follows.
TP29JY94.001
Where 14 mg/kg/d is the NOAEL for female rats in the NTP study, and
100 is the uncertainty factor applied to account for inter and intra-
species differences in accordance with EPA guidelines when a NOAEL from
a chronic animal study is the basis for the RfD. The MRDLG is based on
a 70 kg adult consuming an average of 2 liters water per day over their
lifetime. In addition, an 80% RSC is assumed in the absence of data to
the contrary.
Public comments are requested on the following issues: 1) placing
chlorine in Category III for developing an MRDLG, 2) selection of the
study and NOAEL as the basis for the MRDLG, 3) the 80% RSC, 4) the
appropriateness of the UF of 100, 6) the cancer classification for
chlorine.
2. Chloramines
Inorganic chloramines (CAS Nos. 10599-90-3 and 10025-85-1 for mono-
and trichloramine, respectively) are formed in waters undergoing
chlorination which contain ammonia. Monochloramines, dichloramines and
trichloramines may be formed. Monochloramine is the principal
chloramine formed in chlorinated natural and wastewater at a neutral pH
and is much more persistent in the environment.
Chloramine is used as a disinfectant in drinking water to control
taste and odor problems, limit the formation of chlorinated
disinfection by-products, and maintain a residual in the distribution
system for controlling biofilm growth. At typical pHs of most drinking
waters, the predominant chloramine specie is monochloramine. For
purposes of this regulation, only monochloramine will be considered
since the other chloramines occur at much lower concentrations in
almost all drinking waters. Monochloramine has also been much more
extensively studied.
Monochloramine, the principal chloramine formed in chlorinated
natural and wastewaters at neutral pH, is relatively stable when
discharged to the environment. First-order decay rate constants of 0.03
to 0.075 hr-1 for monochloramine in the laboratory, and higher
rate constants of 0.28 to 0.31 hr-1 outdoors using chlorinated
effluents, have been reported. If discharged into receiving waters
containing bromide, monochloramine will decompose faster, probably
through the formation of NHBrCl and decomposition of the dihalamine.
The rate of monochloramine disappearance is primarily a function of pH
and salinity. For example, at pH 7 and 25 deg.C, the half-life of
monochloramine is 6 hr at 5 parts per thousand (ppt) salinity and 0.75
hr at 35 ppt salinity; at pH 8.5 and 25 deg.C, the half-life is 188 hr
at 5 ppt salinity and 25 hr at 35 ppt salinity. Monochloramine is
expected to decompose in wastewater discharges receiving waters via
chlorine transfer to organic nitrogen-containing compounds.
Occurrence and Human Exposure. Chloramine occurs in drinking water
both as a by-product and intentionally for disinfection. Chloramine is
formed during chlorination when source waters contain ammonia. It is
also used as a primary or secondary disinfectant, usually with
chloramine being generated on site by the addition of ammonia to water
following treatment by chlorination. The use of chloramines has been
shown to reduce the formation of certain by-products, notably
trihalomethanes, relative to the by-products formed with chlorination
alone. Chlorination by-product formation can be minimized when the
ammonia is added prior to or in combination with chlorine by reducing
the chlorine residual of the water being treated. In most plants,
however, ammonia is added some time after the addition of chlorine, to
allow for more effective disinfection since chlorine is a much stronger
disinfectant than chloramines.
Table V-3 presents occurrence information available for chloramine
in drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' USEPA, 1992a.
Typical dosages of chloramine used as a disinfectant in drinking water
treatment facilities range from 1.5 to 2.7 mg/L. Median concentrations
of chloramine in drinking water were found to range from 1.1 to 1.8 mg/
L.
Table V-3.--Summary of Occurrence Data for Chloramines
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of chloramine in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (mg/L)
Survey (year)\1\ Location Sample information (No. of ----------------------------------------------------------------
samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
AWWARF (1987) McGuire & National Survey........... Finished Water From: ........... ........... ........... Typical dosages:
Meadow, 1988. Lakes.......................... 1.5 mg/L
Flowing Streams................ 2.7 mg/L
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell 0.9-5.5 2.3 1.8 ........................
1989) Krasner et al., Nationwide. Effluent, 4 Quarters (13).
1989b.
EPA, 1992b\2\ (1987-1991). Disinfection By-Products At the Plant (11).............. 1.2-3.6 2.1 1.5 ........................
Field Studies. Distribution System (8)........ 0.1-3.3 1.4 1.1
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
AMWA: Association of Metropolitan Water Agencies.
AWWARF: American Water Works Association Research Foundation.
CDHS: California Department of Health Services.
EPA: Environmental Protection Agency.
Based on the residual concentrations given above, a high and low
estimate for exposure to chloramine from drinking water can be
calculated using an assumed consumption of 2 liters per day. Using the
target range of 1.5 to 2 mg/L, the exposure may range from 3 to 4 mg/
day. Some systems may deviate significantly from this range.
No information is available on the occurrence of chloramine in food
or air. Currently, the Food and Drug Administration (FDA) does not
measure for chloramine in foods since the analytical methods have not
been developed. Preliminary discussions with FDA suggest that there are
not approved uses for chloramine in foods consumed in the typical diet.
Similarly, EPA's Office of Air and Radiation is not sampling
chloramines in air (Borum, 1991).
Based on the previous discussion, EPA assumes that drinking water
is the predominant source of exposure to chloramine. Air and food
intakes are believed to provide only small contributions, although the
magnitude and frequency of these potential exposures are issues
currently under review. EPA, therefore, is proposing to establish an
MRDLG for chloramine in drinking water with an RSC value of 80%, the
current exposure assessment policy ceiling. EPA requests any additional
data on known concentrations of chloramine in drinking water, food and
air.
Health Effects. The health effects information in this section is
summarized from the draft Drinking Water Health Criteria Document for
Chloramines (USEPA, 1994b). Studies mentioned in this section are
summarized in the Criteria Document.
Short-term inhalation exposures to high levels (500 ml of 5%
household ammonia mixed with 5% hypochlorite bleach) of chloramines in
humans result in burning in the eyes and throat, dyspnea, coughing,
nausea and vomiting. Inhalation of the chloramine fumes resulted in
pneumonitis but did not result in permanent pulmonary damage.
Short-term exposures to chloramines in drinking water, in which
human subjects were administered single concentrations ranging between
1 and 24 mg/L (1, 8, 18 or 24 mg/L), have not resulted in any adverse
effects in human subjects. Following human exposure, the subject's
physical condition, urinalysis, hematology, and clinical chemistry were
evaluated. No adverse clinical effects were noted in any of the
studies.
In another study, acute hemolytic anemia, characterized by
oxidation of hemoglobin to methemoglobin and denaturation of
hemoglobin, was reported in hemodialysis patients when tap water
disinfected with chloramines was used for dialysis baths. Chloramines
were reported to produce oxidant damage to red blood cells and inhibit
the metabolic pathway used by red blood cells to prevent and repair
such damage. Many dialysis centers have installed reverse osmosis units
coupled with charcoal filtration or the addition of ascorbic acid to
prevent hemolytic anemia.
Animal studies indicate varying sensitivity and conflicting results
among different animal species. Toxic effects noted among rats are
changes in blood glutathione and methemoglobin. Both monkeys and mice
were unaffected during short-term assays with doses up to 200 mg/L
chloramines. Based on studies up to 6 weeks in length, rats appear to
be more sensitive to monochloramine than mice and monkeys.
Toxicokinetic studies of chloramines indicate that the absorption
of chloramines is rapid, peaking within 8 hours of administration. In
the rat, chloramines are metabolized to chloride ion and excreted
mostly through the urine with a small portion excreted through the
feces.
Longer-term oral studies (90 days or longer) showed decreased body
and organ weights in rodents. Some effects to the liver (weight
changes, hypertrophy, and chromatid pattern changes) appear to be
related to overall body weight changes caused by decreased water
consumption due to the unpalatibility of chloramines to the test
animals.
In addition, chloramine may induce immunotoxicity in rats in the
form of increased prostaglandin E2 synthesis, reduced antibody
synthesis, and spleen weight at levels as low as 9 to 19 mg/L
chloramines for 90 days. The significance of these findings for risk
use in risk assessment is compromised by the design flaws of the study
(i.e., animals were exposed to two antigens) and the lack of
corroboration of these findings by a follow-up study.
Two lifetime rodent studies involving oral exposures to rats and
mice via drinking water have been considered by EPA for the derivation
of the MRDLG for monochloramine. Both studies were performed by the
National Toxicology Program (NTP, 1990) and involved 70 animals/sex/
dose exposed to distilled drinking water containing 0, 50, 100 or 200
ppm chloramines.
The first NTP (1990) study was a 2-year study in mice to determine
the potential chronic toxicity or carcinogenic activity of
chloraminated drinking water. B6C3F1 mice were administered
chloramine at doses of 0, 50, 100 and 200 ppm in distilled drinking
water. These doses were calculated based on a time-weighted average to
be 0, 5.0, 8.9 and 15.9 mg/kg/day for male mice and 0, 4.9, 9.0 and
17.2 mg/kg/day for female mice. There was a dose-related decrease in
the amount of water consumed by both sexes; this decrease was noted
during the first week and continued throughout the study. Dosed male
and female mice had similar food consumption as controls except for
females in the 200 ppm dose group that exhibited slightly lower
consumption than controls.
Study results indicated that there was a dose-related decrease in
mean body weights of dosed male and female mice throughout the study.
Mean body weights of high-dose male mice were 10-22% lower than their
control group after week 37 and the body weights of high-dose female
mice were 10-35% lower after week 8. However, the survival of mice
receiving monochloramine in drinking water was not significantly
different than controls. Clinical findings observed were not attributed
to the consumption of chloraminated drinking water. Body weight loss
and systemic toxicity were not considered related to the toxicity of
chloramine, but rather due to decreased water consumption resulting
from the unpalatability of chloramines in drinking water to the test
animals. Therefore, the highest dose tested, 17.2 mg/kg/day, is
considered a NOAEL in mice.
In the second study F344/N rats were administered monochloramine
for 2 years at doses of 0, 50, 100 and 200 ppm in distilled drinking
water. These doses were calculated on the basis of a time-weighted
average to be 0, 2.1, 4.8 and 8.7 mg/kg/day for male rats and 0, 2.8,
5.3 and 9.5 mg/kg/day for female rats. There was a dose-related
decrease in the amount of water consumed by both sexes; this decrease
was noted during the first week and continued throughout the study.
Food consumption of treated rats was the same as the controls with
males consuming more. In addition, mean body weights of 200 ppm dosed
rats (both sexes) were lower than their control groups. However, mean
body weights of rats receiving monochloramine in drinking water (at all
levels) were within 10% of controls until week 97 for females and week
101 for males. Though several clinical changes were noted, no clinical
changes were attributable to chloraminated drinking water. The survival
of rats receiving chloraminated drinking water was not significantly
different than controls except that, for the 50 ppm dose groups,
survival was greater than that of controls. Therefore, EPA considers
the highest dose tested, 9.5 mg/kg/day, as the NOAEL.
Based on two bacterial assays, monochloramine appears to be weakly
mutagenic. One study examining the reproductive effects and another
which examined developmental effects of chloramines concluded that
there are no chemical-related effects due to chloramines.
The NTP evaluation, using the results of the two lifetime NTP
bioassays, concluded that chloramines exhibited equivocal evidence of
carcinogenic activity of chloraminated drinking water in female F344/N
rats. This conclusion results from an increase in mononuclear cell
leukemia. There was no evidence of carcinogenic activity in male rats
or mice of either sex. The findings do not establish a link between
chloramine exposure and carcinogenicity because of the high historical
background occurrence of this type of cancer in test animals. The
incidence of mononuclear cell leukemia in the female control groups
(16%) was substantially less than the incidence reported in untreated
historical controls (25%). Incidence of mononuclear cell leukemia in
test animals reached a high of 32% in the high dose female rats. This
study also discovered incidence of renal tubular cell neoplasms in two
high-dose male mice receiving chloraminated water. Since this type of
tumor is rarely seen in historical controls, there is some concern that
these may be treatment related. However, the overall evidence regarding
the potential carcinogenicity of chloramines in drinking water can be
described as inconclusive since no long-term study has linked any tumor
development to actual chloramination exposure. On this basis, as well
as consideration of those studies described in section C, EPA placed
chloramine in Group D: not classifiable based on inadequate evidence of
carcinogenicity.
EPA selected the lifetime study in rats (NTP, 1990) as the basis
for calculating the MRDLG for chloramines. The NOAEL for the rat (9.5
mg/kg/d) is proposed because the rat was not tested at the higher doses
where mice were tested (17.2 mg/kg/d). Rats appear to be more sensitive
considering observed changes in biochemistry. Following a Category III
approach and using the rat NOAEL of 9.5 mg/L from the NTP study, an
MRDLG of 4 mg/L (measured as total chlorine), based on lack of toxic
effects in a chronic study can be derived for the 70-kg adult consuming
2 liters of water per day applying an uncertainty factor of 100, which
is appropriate for use of a NOAEL derived from an animal study and
assuming an RSC drinking water contribution of 80 percent.
---------------------------------------------------------------------------
*Since chloramines, on a practical basis, will be measured as
total chlorine, it is necessary to present the MRDLG in terms of a
chlorine equivalent concentration. Three mg/L chloramine is
equivalent to 4 mg Cl2/liter, based on the molecular weights of
Cl2 and NH2Cl.
TP29JY94.002
EPA requests comments on the proposed MRDLG for chloramines and the
RSC of 80%, the significance of the findings of immunotoxicity for
setting the RfD instead of the NTP study, the significance of the
finding of mononuclear cell leukemia in female F344 rats, the
significance of the finding of tubular cell neoplasms in high-dose
exposed mice, and whether the adjusted MRDLG, which takes into account
the measurement of monochloramine as total chlorine, is appropriate.
TP29JY94.003
3. Epidemiology Studies of Chlorinated and Chloraminated Water
Several studies have been conducted to evaluate the association of
chlorination or chloramination with the risk of cancer, cardiovascular
disease or adverse reproductive effects in humans. A summary of some of
these studies is given below. This discussion reflects EPA's assessment
of these data and is summarized from the draft Drinking Water Criteria
Documents for Chlorine (USEPA, 1994a) and Chloramines (USEPA, 1994b),
respectively.
Introduction to Epidemiology Studies. Two distinct types of
epidemiology studies have been conducted: ecologic and analytical.
These types of studies differ markedly in what they reveal about the
association between water quality and disease. In an ecological
epidemiology study, information is available on exposure and disease
for groups of people rather than for individuals, and therefore, the
results are difficult to interpret. What is considered to be an
important or relevant group variable may not be important for or may
not pertain to individuals within that group. Theoretical and empirical
analyses have offered no consistent guidelines for the interpretation
of ecological associations, and results from these studies are
appropriate only to suggest hypotheses for further study by analytical
epidemiological methods (Piantiadosi et al., 1988; Connor and Gillings,
1974).
Analytical epidemiology studies provide an estimate of the
magnitude of risk and information which can be used to evaluate
causality. For each individual in the study, information is obtained
about disease status and exposure to various contaminants and other
characteristics. In several of the studies reported here, individual
exposures to disinfected water or specific disinfection by-products
were estimated using group exposure information. All reported
epidemiological associations from analytical studies require an
evaluation of random error (statistical significance) and potential
sources of systematic bias (misclassification, selection, observation,
and confounding biases) so that results can be interpreted properly. It
must be noted that random error or chance can never be completely ruled
out as the explanation for an observed association and that statistical
significance does not necessarily imply biological significance.
Regardless of statistical significance, it is important to consider
potential biological mechanisms. Random error does not address the
possibility of systematic error or bias. Misclassification of exposure
and disease, selection bias, and observation bias must be avoided;
confounding bias, on the other hand, can be prevented both in study
design and during analysis if information is obtained about possible
confounders. It is important to determine for each specific
epidemiology study the validity of the association observed between
exposure and disease before considering possible causality between
exposure and disease or inferring that the results apply to a larger or
target population. Systematic bias can lead to spurious associations;
in some but not all instances, the direction of the bias can be
determined. For example, a random misclassification of exposure usually
biases a study toward not observing an effect or observing a smaller
risk than may actually be present, but nonrandom misclassification can
result in either higher or lower estimates of risk depending upon the
distribution of misclassification.
In addition, because of the observational nature of epidemiology,
the interpretation of epidemiology studies requires a sufficient number
of well designed and well conducted epidemiology studies, and must
include appropriate toxicological and biological information. Judging
causality from epidemiology studies is based largely on guidelines
developed over the years, including sequence of events, strength of
association (relative risk or odds ratio), consistency of results under
different conditions of study, biological plausibility, dose or
exposure response relationship, and specificity of effect. The relative
risk represents a basic measure of an association between exposure and
disease. It is defined as the rate of disease in the exposed population
divided by the rate of disease in an unexposed population.
a. Cancer Studies. Since the early 1970's, numerous epidemiologic
studies have attempted to assess the association between cancer and the
long term consumption of water from various sources with and without
disinfection and of various chemical quality, especially chlorinated
surface waters which supply the majority of the U.S. population.
Ecological, case control, and cohort studies have been conducted. Case
control studies have included incident and decedent cases; in some
studies information about various risk factors has been collected
through interviews, but in others information was obtained primarily
from death certificates.
i. Ecological Studies. The earliest studies were analyses of group
or aggregate data available on drinking water exposures and cancer.
Usually the variables selected for analyses were readily available in
published census, vital statistics, or public records and easily
abstracted and assembled. These analyses, referred to as ecological but
also called aggregate, geographical, or correlational, were designed to
investigate cancer mortality rates, usually on a county or State level.
Areas of different water quality, source, and chlorination status were
compared to identify possible statistical associations for further
study. Drinking water exposures were most often characterized as simple
dichotomous variables which served as indicators of exposure to
differing source water quality, e.g., the drinking water source for the
county or geographic area was categorized as a surface or groundwater
source. In some instances exposure variables included estimates of the
proportion of the area's or county's population that received surface
or groundwater and whether it was chlorinated. Surface water was
assumed to be more contaminated with synthetic organics than
groundwater, but no attempt was made to estimate levels of
contaminants.
In 1974 it was discovered that when surface waters were disinfected
with chlorine, the chlorine reacted with pre-existing organic materials
in the water to create a great number of chemical by-products (Craun,
1988). The major group of disinfection by-products (by weight) was the
trihalomethanes (THMs) which included an animal carcinogen, chloroform.
Chlorinated surface water was evaluated as an exposure variable in
several of the ecological studies, and since almost all surface waters
are chlorinated, the analyses usually compared cancer mortality among
populations receiving chlorinated surface water with those receiving
unchlorinated groundwater. Chlorinated water was assumed to contain
disinfection by-products, and at higher levels than chlorinated
groundwaters. However, the quality of surface and groundwater may also
differ for other contaminants, and this was not considered. Any
observed association might be due to other water quality differences
among surface and groundwaters (e.g., organic contaminants from
nonpoint and point source discharges to surface waters from industrial,
urban, and agricultural sources before pollution control regulations).
In some ecological analyses, the investigators attempted to study
the association between cancer mortality and an estimate of group
exposure to levels of chlorinated by-products based on THM or
chloroform levels was determined from a limited number of water
samples. The exposure information used in ecological studies was
available in only broad geographic units such as census tracts or
counties (Crump and Guess, 1982; Shy, 1985). Although these exposure
variables were statistically associated with mortality rates for all
cancers combined and several site-specific cancers, the interpretation
must necessarily be cautious due to limitations of ecological studies.
In several of these studies, aggregate or group information on several
covariates, e.g., occupation, income, or population density, also was
included in the statistical analysis in an attempt to adjust for
potential confounding factors. In one study the statistical
significance of the observed associations between stomach and rectal
cancer mortality and group exposures to current THM levels disappeared
when migration patterns and ethnic data were included in the regression
model (Tuthill and Moore, 1980). A wide range of cancer sites was found
to be statistically associated with estimates of population group
exposures based on current levels of THM or chloroform including gall
bladder, esophagus, kidney, breast, liver, pancreas, prostate, stomach,
bladder, colon, and rectum. The most frequent associations observed
were for the last four sites; however, these associations were not
consistent when viewed by gender, race, and geographic region.
The ecologic design coupled with the lack of specific exposure
indicators in these studies precludes the inference of a causal
relationship (Morgenstern, 1982). A subcommittee of the National
Academy of Sciences (NAS, 1980) reviewed 12 of the ecological studies
and noted ``Results of these studies demonstrate the problems of
establishing relationships between health statistics and environmental
variables, and lend emphasis to the caution with which they should be
interpreted.'' The NAS further commented that the ecological studies in
which the current THM exposures were estimated were deemed to be more
informative than others and ``suggest that higher concentrations of THM
in drinking water may be associated with an increased frequency of
cancer of the bladder. The results do not establish causality, and the
quantitative estimates of increased or decreased risk are extremely
crude. The effects of certain potentially important confounding
factors, such as cigarette smoking, have not been determined.'' The
studies are useful, however, as an initial step for identification of
potential hazards and they indicated the need for further epidemiologic
studies or analytic studies of individuals with a specific etiologic
hypothesis.
ii. Cohort Studies. A cohort study (or follow-up) study (the study
can be called either retrospective or prospective) is one in which two
or more groups (referred to as `cohorts') of people that differ
according to the extent of exposure to a potential cause of disease are
compared with respect to incidence of the disease of interest in each
of the groups. The essential element of this study type is that
incidence rates are calculable directly for each study group (Rothman,
1986). One advantage of this study type is the ability to study
multiple disease endpoints. One disadvantage of this study type is that
a large study population is needed to detect a relatively small risk.
In addition, because of the latency period for carcinogenicity, a long
follow-up time may be required for the study.
There exists one cancer drinking water cohort study where
individual data were available for a well-defined, fairly homogeneous
area that allowed disease rates to be computed by presumed degree of
exposure to by-products of chlorination, although the population was
relatively small. Wilkins and Comstock (1981) studied the residents of
Washington County, Maryland and ascertained the source of drinking
water at home for each county resident in a private census conducted in
1963. In addition to water source, information was collected on age,
marital status, education, smoking history, number of years lived in
the household, and frequency of church attendance. Death certificates
and cancer registry information was sought for county residents whose
date of death or diagnosis occurred in the 12 year period following the
census. Sex and site-specific cancer rates were constructed for
malignant neoplasms of biliary passages and liver, kidney, and bladder.
Several additional causes of death were analyzed as well for comparison
purposes. The population was stratified into three separate exposure
subgroups: chlorinated surface water, unchlorinated deep wells, and
small municipal systems with a mixture of chlorinated and unchlorinated
water, each reflecting a different history of exposure to by-products
of chlorination. The study group which included individuals who
obtained drinking water from small municipal systems were not included
as a comparison with the other drinking water cohorts because of their
exposure to both chlorinated and unchlorinated water.
Both crude and adjusted incidence rates for liver cancer in males
and females and for cancer of the liver among males were essentially
the same for persons supplied with chlorinated surface water at home
(high THM exposure) and for persons with deep wells (low THM exposure).
The adjusted rates for bladder cancer (RR=1.6; 95% CI=0.54,6.32) and
cancer of the liver (RR=1.8; 95% CI=0.64,6.79) among females were
highest among persons using chlorinated surface water. Given the low
relative risk and broad confidence intervals, the authors indicated
that this finding could be attributed to chance (Wilkins and Comstock,
1981). Confounding bias may also influence the interpretation of a
small relative risk. EPA considers that the results of this study are
inconclusive because the results are based on small numbers of cases,
hence, the reported rates are statistically unstable and subject to
random variation.
iii. Case Control Studies. In a case control study, persons with a
given disease (the cases) and persons without the given disease (the
controls) are selected for study. The proportions of cases and controls
who have certain background characteristics or who have been exposed to
possible risk factors are then determined and compared. Exposure odds
ratios (ORs) are determined. The odds of exposure among cases is
compared with that of controls. For rare diseases, the ORs are
considered good estimates of relative risk. These studies are sometimes
called case-referent or retrospective studies. Because there are many
variations of this study design (e.g., how cases and controls are
selected, how information on exposures, risk factors, and confounding
factors are obtained, and who is interviewed), each case control study
should be evaluated individually to determine if the specific study
design parameters introduce systematic bias (Kelsy et al., 1986). As
previously noted, all epidemiology studies require careful evaluation
of systematic bias. For those studies with major bias, the results are
generally considered inconclusive. Those studies with minor bias may
still provide useful information.
Two types of studies were conducted: (1) Decedent cases without
interviewing survivors for information about residential histories and
risk factors and (2) incident cases with interviews.
Decedent Case-Control Studies. Several case-control studies were
conducted to continue to investigate the possibility that there was a
causal relationship between chlorinated drinking water, including
byproducts such as THMs, and gastrointestinal or urinary tract cancers.
Most of these case-control studies used deceased cases of the specific
cancers of interest, although some continued their investigations in a
relatively nonspecific way by using both total cancer mortality as well
as several of the site- specific cancers studied in the ecologic
studies (Crump and Guess, 1982; Shy, 1985). Controls were noncancer
deaths from the same geographic area and in all but one study matched
for several potentially confounding variables including age, race, sex,
and year of death. As in all studies of this design (i.e., death
certificate studies with no available interviews), control of
confounding factors was restricted to information that is routinely
recorded on death certificates and no information was obtained from
next-of-kin interviews. The exposure variables of interest at this time
included a comparison of surface v ground water sources, or chlorinated
v nonchlorinated ground water sources. The place of residence listed on
the death certificate was linked to public records of water source and
treatment practices in order to classify the drinking water exposure
variable for a particular case or control (Shy, 1985).
Similar to the earlier ecologic studies, the Agency considers the
results from these studies to be inconsistent in their findings. The
calculated ORs, varied by cancer site and sex, as well as in their
magnitude and statistical significance. This variability was found for
all the cancer endpoints studied including those of specific interest,
i.e., bladder, colo-rectal, and/or colon. These endpoints were found to
vary by geographic region. For example, a statistically significant
increased bladder cancer risk was observed in North Carolina for males
and females combined (OR=1.54) and New York for males (OR=2.02), but
not for females; no statistically significant risk was seen in
Louisiana, Wisconsin or Illinois. Increased colon cancer risk was
observed in Wisconsin (OR=1.35) and North Carolina (OR=1.30) for males,
but not females; no increased risk was seen in Louisiana or Illinois.
Increased rectal cancer risk was observed in North Carolina (OR=1.54)
and Louisiana (OR=1.68) for males and females combined, in Illinois for
females (OR=1.35) but not males, and in New York for males (OR=2.33)
but not females; no increased risk was seen in Wisconsin. Although
increased risk was observed for cancer of the liver and kidney
(OR=2.76), esophagus (OR=2.39), and pancreas (OR=2.23) among males in
New York, no increased risk for these cancers was seen among females in
New York, Illinois, Wisconsin or Louisiana.
Although many of the ORs were statistically significant, these
decedent case control studies with extremely limited information on
confounding factors and potential exposures to chlorinated water are of
limited usefulness in assessing whether cancer is associated with
chlorinated drinking water, or judging the causality of such as
association. Although some of the ORs were large enough to cause
concern about an exposure association, the magnitude of the OR was such
that the association could be attributed to incomplete control of
confounding factors and the ORs might represent spurious elevations
(Crump and Guess, 1982).
Although not subject to all the same limitations as ecologic
studies, decedent case-control studies are considered more limited by
some epidemiologists than others as a tool for causal inference because
of a high probability of systematic bias associated with the use of
information obtained only from the death certificate (e.g., inadequate
or no information on residential history, water exposures, and major
potential confounders). The variability seen in these five studies is
likely a combination of several factors, including available sample
size, choice of causes of death included as controls, regional
variability in true composition of the raw and treated drinking waters,
definition of exposure variables, a high probability of exposure
misclassification from imputing a lifetime exposure to a certain water
source or treatment from residence listed on the death certificate, and
uncontrolled confounding (e.g., diet and smoking).
Given the limitations of decedent case control studies without
interviews, the evidence from these studies are considered insufficient
to determine a causal association between any or all the components
which exist in the complex mixture created during the chlorination of
surface waters and any site-specific cancer. The findings provided a
stimulus for a further refined epidemiologic study using incident cases
of bladder and colon cancer and appropriate controls who could be
interviewed for residential history and numerous other covariates.
Case-Control Studies with Interviews. At the time when these more
recent studies were planned, it was still believed that THMs were the
major by-products of chlorinated drinking water that should be
investigated and studies were designed and conducted in areas where a
THM difference might be expected and somehow measurable. Exposure
assessment for individuals remained problematic in the study design.
The best available means of exposure measurement, however, was at best
a surrogate for the true exposure of interest which is the actual level
of THMs or other by-products ingested over a person's lifetime through
consumption of surface water disinfected with chlorine. Only two
studies attempted to estimate long-term exposure to THMs. Most studies
used residence at a location served by chlorinated drinking water. In
all except one of the studies, comparisons of exposure were between
chlorinated surface water and unchlorinated groundwater. As previously
discussed in the section on ecological studies, the water quality for
surface and ground water differ for many other consituents.
Because it was known that disinfection of surface water using
chloramine produced very low levels of THMs and other by-products
compared to the same water disinfected with chlorine, a study was
conducted in Massachusetts to compare the patterns of mortality in
communities which used these different disinfectants (Zierler et al.,
1986). Statewide mortality records for 1969-1983 were analyzed using
standardized mortality ratios (SMRs) and showed little variation by
community. However, mortality odds ratios (MORs) comparing bladder
cancer deaths to all other deaths were considered by the authors to
indicate a slight elevation for last residence in a chlorinated
community compared to a chloraminated community (MOR=1.7; 95% CI=1.3,
2.2). The authors noted that the results were preliminary and ``crude
descriptions of the relationship under study'' (Zierler et al., 1986).
The authors further indicated that the results may have been caused by
unidentified or uncontrolled confounding factors.
Bladder cancer deaths were investigated further using a case-
control design with proxy interviews to determine residential and
smoking histories (Zierler et al., 1988). The association of bladder
cancer was assessed for individuals with lifetime and usual exposure to
chlorinated and chloraminated water depending on the number of years of
residence at a particular water source. Residence in a community using
chlorinated drinking water was used as an index for exposure to
chlorinated by-products, while residence in a community using
chloramine for disinfection was considered an index for no exposure to
chlorinated by-products. An association was observed between bladder
cancer and both lifetime (MOR=1.6; 95% CI=1.2-2.1) and usual (MOR=1.4;
95% CI=1.1-1.8) exposure to chlorinated water. A subgroup of study
participants was noted to have lived their entire lives in an area
served with water supplied by the Massachusetts Water Resources
Authority, disinfected with either chlorine or chloramine (same water
source, different disinfectant, lifetime exposure). Within this group
the bladder cancer mortality risk was 1.6 times higher (MOR=1.6; 95%
CI=1.1, 2.4) when the water had been disinfected with chlorine compared
to chloramine (Zierler et al., 1988).
In addition to analyses using a control group which consisted of
deaths from cardiovascular disease, cerebrovascular disease, chronic
obstructive lung disease, lung cancer, and lymphatic cancer, a separate
analysis was done using only the lymphatic cancer controls. This was
considered necessary by the authors because of the possibility that
some of the other deaths among controls may also be related to the
exposure of interest. If true, then the MOR estimate would be biased
toward the hypothesis of no increased risk. When the analysis
considered only lymphatic cancer controls, the magnitude of the
association with chlorinated water increased for lifetime exposure
(MOR=2.7; 95% CI=1.7-4.3), usual exposure (MOR=2.0; 95% CI=1.4-3.0),
and lifetime exposure in the previously mentioned subgroup (MOR=3.5;
95% CI=1.8-6.7). Sources of misclassification bias that may have been
present were considered to be randomly distributed among the cases and
controls which implies that the observed MOR would be an underestimate
of risk (Zierler et al., 1988). It is also possible that
nondifferential misclassification of the variables used to control
confounding, leading to residual confounding of the summary estimates,
could have caused a systematic spurious elevation in the MORs.
The largest study to date investigating the relationship of
chlorinated water and bladder cancer incidence involved an ancillary
study to the National Cancer Institute's (NCI) 10 area study of bladder
cancer and artificial sweeteners (Cantor et al., 1985, 1987, 1990). The
original study conducted interviews with 2,982 newly diagnosed bladder
cancer cases and 5,782 population controls; lifetime information on
source and treatment of drinking water was collected and analyzed for
only a subset of the original study population (1,244 cases and 2,550
controls). Subgroup analyses of nonsmokers among participants and those
reporting beverage intake necessarily involved even smaller numbers.
Duration of exposure, measured by years of residence at a chlorinated
surface or nonchlorinated ground water source was presumed to be a
surrogate for dose of disinfectant by-products. Overall, there was no
association of duration of exposure with bladder cancer risk (Cantor et
al., 1985, 1987). In nonsmokers who never smoked, a 2-fold increased
risk was reported for those exposed for 60 or more years to chlorinated
surface water (n=46 cases, 77 controls) compared to unchlorinated
ground water (n=61 cases, 268 controls) users (OR=2.3; 95% CI=1.3,
4.2). These data were further analyzed according to beverage intake
level, type of water source and treatment (Cantor et al., 1987, 1990).
It was observed that people who reported drinking the most tap water-
based beverages from any source (>1.96 liters/day) had a bladder cancer
risk about 40% higher (OR=1.43; 95% CI=1.23-1.67, males and females
combined) than people who drank the least. The association between
water ingestion and bladder cancer risk for males was an OR=1.47; 95%
CI=1.2-1.8, and for females an OR=1.29; 95% CI=0.9-1.8.
Evaluation of bladder cancer risk by both duration of exposure and
amount of water consumed showed that the risk increased with higher
water consumption only among those who drank chlorinated surface water
for 40 or more years. Evaluation of risk by smoking status revealed
that most of the duration effect was observed in nonsmokers. Among
nonsmokers who consumed tap water in amounts above the population
median (>1.4 L/day), a risk gradient was apparent only for males.
However, a higher risk was also seen for nonsmoking females who
consumed less than the median level. The increasingly smaller numbers
of cases and controls available for these subgroup analyses produce
statistically unstable OR estimates making it difficult to evaluate the
trend results.
This is the first study of incident bladder cancer cases that
obtained and analyzed fluid consumption patterns in this way. The noted
inconsistencies in the reported data must be more thoroughly explored
and indicate a need for replication before any causal relationship can
be assumed (Devesa et al., 1990). An additional consideration is a more
refined exposure measurement; many of the disinfection by-products are
volatile. Thus exposure may occur through inhalation as well as
ingestion.
Two conflicting studies of colon cancer and presumed THM exposure
have been reported. The first one (Cragle et al., 1985) was a hospital
based case control study that included 200 incident colon cancer cases
from seven hospitals and 407 hospital controls with no history of
cancer who were diagnosed with diseases unrelated to colon cancer. It
should be noted that both colon and rectal cancer cases were included
as cases in the study. Controls were matched to cases on hospital and
admission date, as well as age, race, sex and vital status. Residential
histories were linked with water source and disinfectant information
for the 25 years prior to diagnosis. Logistic regression analysis using
qualitative data groupings for the variables of interest showed a
strong interaction of age and chlorination status (Table V-4). THM
levels were not estimated. Odds ratios computed from the regression
coefficients increased with age, and within age groups. The ORs are
higher for a longer duration of exposure.
Table V-4.--Comparison of OR's By Exposure Duration and Age
------------------------------------------------------------------------
OR (95% CI) 1-15 OR (95% CI) > 15
Age (years) years exposure years exposure
------------------------------------------------------------------------
20-29............................. 0.23
(0.11, 0.49) 0.48
(0.23, 1.01)
30-39............................. 0.36
(0.2, 0.66) 0.6
(0.33, 1.09)
40-49............................. 0.57
(0.36, 0.88) 0.75
(0.48, 1.18)
50-59............................. 0.89
(0.83, 1.12) 0.94
(0.69, 1.29)
60-69............................. 1.18
(0.94, 1.47) 1.38
(1.1, 1.72)
70-79............................. 1.47
(1.16, 1.84) 2.15
(1.7, 2.69)
80-89............................. 1.83
(1.32, 2.53) 3.36
(2.41, 4.61)
------------------------------------------------------------------------
From these data, it appears that risk is increased only in those
persons 60 years old and older with greater than 15 years of exposure
to chlorinated water and in those greater than 70 years of age,
regardless of exposure duration.
A second colon cancer study (Young et al., 1987) conducted in
Wisconsin involved 366 incident colon cancer cases, 785 controls
diagnosed with other cancers, and 654 population controls. Extensive
interviews were conducted with all participants to obtain information
on past drinking water sources, drinking water habits and a number of
potentially confounding covariates. This information was combined with
data provided by water companies to construct models to predict
historical levels of THMs to be used as both cross-sectional and
cumulative exposure variables. Simpler methods of defining exposure
were also used, (e.g., surface vs. ground, chlorinated v.
nonchlorinated), and all methods looked at the data by period specific
exposure levels. The results did not indicate any association between
THMs in Wisconsin drinking water and colon cancer risk. Odds ratios for
all exposure variables were uniformly close to 1.0 with few exceptions.
It should be noted, however, that in this study the majority of the
water supplies contained less than 20 g/l of THMs. No excess
risk was observed at these levels, given the limitations of this study
design in detecting a small risk.
The association of THM and colo-rectal cancer was studied in New
York where the THM levels were higher than those in the above Wisconsin
study (Lawrence et al., 1984). A total of 395 colon and rectal cancer
deaths among white female teachers in New York State (excluding New
York City) was compared with an equal number of deaths of teachers from
causes of death other than cancer. All deaths were ascertained using
the defined cohort of the New York State Teachers Retirement System.
Cumulative chloroform exposure was estimated by the application of a
statistical model to operational records from water systems that served
the home and work addresses of the study participants during the 20
years prior to death. The distribution of chloroform exposure was not
significantly different between cases and controls. No effect of
cumulative chloroform exposure was observed in a logistic analysis
controlling for type, population density, marital status, age, and year
of death. No excess risk was associated with exposure to a surface
water source containing THMs (OR=1.07; 90% CI=0.79, 1.43). Although the
data were not presented in the article, the authors reported that no
appreciable differences were seen when the colon and rectal cases were
analyzed separately, compared to the combined analyses reported above.
Although most all the studies reviewed here have looked at colon,
colo-rectal or bladder cancer risk, one recently published work
investigated the risk of pancreatic cancer in relation to presumed
exposure to chlorinated drinking water. Ijsselmuiden et al. (1992)
conducted a population-based case-control study in Washington County,
Maryland, using the same population data that were originally
ascertained during a private population census for an earlier cohort
study (Wilkins and Comstock, 1981). The original cohort study did not
find any association between pancreatic cancer and chlorinated drinking
water (OR=0.80, 95% CI=0.44-1.52).
This case-control study was conducted to reexamine chlorinated
drinking water as a possible independent risk factor for pancreatic
cancer in this population. It is not reported of any of the other
endpoints from the original study also were reexamined, e.g., bladder,
kidney, or liver cancer. Cases were those residents who were reported
to the County cancer registry with a first time pancreatic cancer
diagnosis during the period July, 1975 through December 1989, and who
had been included in the 1975 census (n=101). Controls were randomly
selected by computer from the 1975 census population (n=206). Drinking
water source, as obtained during the 1975 census, was the exposure
variable used. In univariate analyses, municipal water as a source of
drinking water, increasing age, and unemployment were significantly
associated with increased risk of pancreatic cancer. Multivariable
analyses that controlled for confounding variables indicated that the
use of municipal chlorinated water at home was associated with a
significant OR of 2.23 (95% CI=1.24-4.10). The OR adjusted is 2.18 (95%
CI=1.20-3.95); only age and smoking were assessed as potential
confounders.
Interpretation of these findings is hampered by several problems
regarding the assessment of exposure, including the fact that
information obtained in 1975 on type of water and other variables is an
exposure collected at one point in time and may not reflect actual
exposure patterns prior to 1975. In addition, there is no information
on the actual amounts of water consumed. Additionally, different
residential criteria were used for the cases and controls. The cases
had to still be residing in the County at the time of their cancer
diagnosis to be included in the study, but the controls may not have
been current residents. If controls emigrated out of the county
differentially on the basis of exposure, the ORs may be an over- or
underestimate of the risk depending on emigration patterns. Finally, it
can not be ruled out that the exposure variable used for this and other
studies--residence served by a particular water source--is simply a
surrogate for some other unidentified factor associated with nonrural
living. The nonspecific relationship of several different causes of
death and water source at home observed in the earlier cohort study
(Wilkins and Comstock, 1981) lends some support to this possibility.
More valid individual exposure information over a long period of time
for both specific contaminants and the use of chlorinated/unchlorinated
water are needed to assess the results of this and other analytical
epidemiology studies.
Morris et al. (1992) conducted a meta-analysis, evaluating 12
studies and pooling the relative risks from 10 epidemiological studies
of cancer and a presumed exposure to chlorinated water and its
byproducts. Meta-analysis refers to the application of quantitative
methods to combine the published results of a related body of
literature (Dickerson and Berlin, 1992). Morris et al. (1992) reported
a pooled relative risk estimate of 1.21 (95% CI, 1.09-1.34) for bladder
cancer and 1.38 (95% CI, 1.01-1.87) for rectal cancer (i.e., 9% of
bladder cancer cases and 15% of the rectal cancer cases in the U.S. or
approximately 10,000 additional cases of cancer per year could be
attributed to chlorinated water and its by-products). Pooled relative
risk estimates for ten other site specific cancers including colon,
colo-rectal and pancreas were not felt to be significantly elevated nor
were they statistically significant.
If the indications from this analysis are true such that water
chlorination could result in as many as 10,000 cases of cancer a year,
then chlorination could represent a significant cause of rectal and
bladder cancer in the U.S. However, there was disagreement among the
negotiating parties over the appropriateness of this meta-analysis.
Some believed that the use of the meta-analysis may not be appropriate
for these data. Others disagreed, expressing their view that the
analysis was statistically probative and otherwise valuable. Meta-
analysis has been used successfully to combine the results of small
clinical trials and of some epidemiology studies that have similar
experimental design and exposure conditions. Application of meta-
analysis to the water chlorination data requires careful consideration
of exposure variables and systematic bias in each of the studies.
Chlorinated drinking water is a complex mixture of many substances that
vary geographically and seasonally. There is even variability within a
geographic region. In addition, the information on exposure and
potential confounding is much more limited for the four decedent case
control studies used in the meta-analysis. Their study design is
dissimilar to the other studies included, resulting in concerns about
their inclusion in the meta-analysis. Study-specific methodological
problems, systematic bias, and problems of exposure definition and
assessment could not be corrected by this analysis (Murphy, 1993).
Thus, the overall results of the Morris et al. analysis may over- or
underestimate the risk. However, the estimate of risk in regard to
rectal cancer might be particularly affected by the inclusion of these
case control studies. It should also be noted that the results of the
Morris et al. analysis does not provide additional information to
establish causality.
The chlorinated drinking water epidemiology studies have been
reviewed extensively by EPA, the National Academy of Sciences, the
International Agency for Research on Cancer (IARC), and the
International Society for Environmental Epidemiology (ISEE). In 1987,
the National Academy of Sciences Subcommittee on Disinfectants and
Disinfectant By-Products concluded that there was a major health
concern with the chronic ingestion of low levels of disinfection
byproducts (NRC, 1987). The Subcommittee commented that some of the
epidemiology studies reported ``increased rates of bladder cancer
associated with trends of levels of certain contaminants in water
supplies. Interpretation of these studies is hampered by a lack of
control for confounding variables (e.g., age, sex, individual health,
smoking history, other exposures).'' The Subcommittee recommended that
epidemiologists continue to improve protocols and conduct studies on
drinking water and bladder cancer where exposure data can be obtained
from individuals, rather than through estimation from exposure models.
EPA and IARC, along with other individual scientists, have
interpreted the epidemiologic evidence as inadequate. IARC concluded
that ``there is inadequate evidence for carcinogenicity of chlorinated
drinking water in humans.''
The ISEE presented a full spectrum of opinion regarding the
epidemiology data (Neutra and Ostro, 1992). The ISEE reported a
``general consensus that the results of the recent EPA-sponsored
studies of cancer endpoints have strengthened the evidence for linking
bladder cancer with long term exposure to chlorinated drinking water.
The evidence for links with colon cancer are not convincing. * * * Any
risks, if real, are low when compared to the risk of infection from not
disinfecting water.''
In 1992, the International Life Sciences Institute sponsored a
conference with the Pan American Health Organization, EPA, Food and
Drug Administration, World Health Organization, and the American Water
Works Association on the safety of water disinfection. Although they do
not necessarily reflect the views of the sponsoring organizations,
conclusions prepared by the conference's editor and editorial board
(Craun et al., 1993) noted that ``Adverse human health effects may be
associated with the chemical disinfection of drinking water. However,
current scientific evidence is inadequate to conclude that water
chlorination poses a significant risk to humans. Uncertainties about
the available toxicologic evidence limit assessment of human health
risks associated with chlorine, chloramine, chlorine dioxide, and ozone
disinfection. The epidemiologic evidence for increased cancer risks of
chlorinated drinking water is equivocal.''
Some members of the reg-neg committee felt that the epidemiology
data, taken in conjunction with the results from toxicological studies,
provide an ample and sufficient basis to conclude that the usual
exposure to disinfection by-products in drinking water could result in
an increased cancer risk at levels encountered in some public water
supplies.
Because of the spectrum of conclusions concerning these data, the
Agency is pursuing additional research to reduce the uncertainties
associated with these data and better characterize the potential of
cancer risks associated with the consumption of chlorinated drinking
water.
b. Serum Lipids/Cardiovascular Disease. Laboratory studies on
animals, conducted in the early 1980's indicated a possible link
between consumption of chlorinated drinking water and elevated serum
lipid profiles which are indicators of cardiovascular disease (USEPA
1994a). The animal work was followed by a cross-sectional study in
humans (Zeighami et al., 1990) that included 1,520 adult residents,
aged 40 to 70 years, in 46 Wisconsin communities supplied with either
chlorinated or unchlorinated drinking water of varying hardness. The
study was designed to determine whether differences in calcium or
magnesium intake from water and food and chlorination of drinking water
affect serum lipids.
The communities selected for study had the following
characteristics: (1) They were small in population size (300-4,000) and
not suburbs of larger communities; (2) they had not undergone more than
20% change in population between 1970 and 1980; (3) they had been in
existence for at least 50 years; and (4) all obtained water from
groundwater sources with no major changes in water supply
characteristics since 1980 and did not artificially soften water. The
water for the communities contained total hardness of either
80 mg/l CaCO3 (soft water) or 200 mg/l
CaCO3 (hard water); 24 communities used chlorine for disinfection and
22 communities did not disinfect. Eligible residents were identified
through state driver's license tapes and contacted by telephone; an
age-sex stratified sampling technique was used to choose a single
participant from each eligible household. Only persons residing in the
community for at least the previous 10 years were included. A
questionnaire was administered to each participant to obtain data on
occupation, health history, medications, dietary history water use,
water supply and other basic demographic information. Water samples
were collected from a selected subset of homes and analyzed for
chlorine residual, pH, calcium, magnesium, lead, cadmium, and sodium.
Fasting blood specimens were collected from each participant and
analyzed for total cholesterol, triglycerides and high- and low-density
lipoprotein (HDL and LDL, respectively) subfractions.
Among females, adjusted mean total serum cholesterol levels were
statistically significantly higher in the chlorinated communities
compared to the nonchlorinated communities (249 mg/dl and 238 mg/dl,
respectively). These changes are not considered biologically
significant as they reflect background variation. Total serum
cholesterol levels were also higher for males in chlorinated
communities, on the average, but the difference was smaller and not
statistically significant (236 mg/dl vs. 232 mg/dl). LDL mean values
followed a similar pattern to that for total cholesterol, higher in
chlorinated communities for females, but not different for males.
However, for both sexes, HDL cholesterol levels are nearly identical in
chlorinated and nonchlorinated communities and there were no
significant differences found in the HDL/LDL ratios. The implications
of these findings for cardiovascular disease risk are unclear at this
time given the inconsistencies in the data. The possibility exists that
the observed association in females may have resulted from some unknown
or undetermined variable in the chlorinated communities.
The results from a second study, designed to further explore the
findings among female participants in the Wisconsin study (Zieghami et
al., 1990), were presented in 1992 (Riley et al., 1992, manuscript
submitted for publication). Participants were 2,070 white females, aged
65 to 93 years who were enrolled in the Study of Osteoporotic Fractures
(University of Pittsburgh Center) and had completed baseline
questionnaires on various demographic and lifestyle factors. Total
serum cholesterol was determined for all participants. Full lipid
profiles (total cholesterol, triglycerides, LDL, total HDL, HDL-2, HDL-
3, Apo-A-I, and Apo-B) were available from fasting blood samples for a
subset of 821 women. Interviews conducted in 1990 ascertained
residential histories and type of water source used back to 1950 and
all reported public water sources were contacted for verification of
disinfectant practices. Private water sources were presumed to be
nonchlorinated. A total of 1,896 women reported current use of public,
chlorinated water, 201 reported current use of nonchlorinated springs,
cisterns, or wells and 35 reported having mixed sources of water. Most
of the women had been living in the same home with the same water
service for at least 30 years.
Overall, there were no meaningful differences detected in any of
the measured serum lipid levels between women currently exposed to
nonchlorinated water and those exposed to chlorinated water (246 mg/dl
vs. 247 mg/dl, respectively, for total cholesterol). The data were also
stratified by age and person-years of exposure to chlorinated water at
home. There was some suggestion that women with no exposure to chlorine
had lower total cholesterol levels but this finding was inconsistent
and may represent random fluctuation since there was no trend noted
with LDL cholesterol or Apo-B, both of which are known to correlate
with total cholesterol. There was also no association between
increasing duration of exposure to chlorine and HDL cholesterol, Apo-A-
I, or triglycerides.
The only notable differences were that women with chlorinated water
reported significantly more cigarette and alcohol consumption than the
women with nonchlorinated drinking water (Riley et al., 1992). This was
evident in all age groups and across strata of duration of exposure.
This finding lends support to the possibility that the previously
reported association of chlorinated drinking water and elevated total
serum cholesterol (Zeighami et al., 1990) may have arisen due to
incomplete control of lifestyle factors which were differentially
distributed across chlorination exposure groups.
c. Reproductive/Developmental Outcomes. Several recently conducted
epidemiologic studies have examined the relationship between different
reproductive or developmental endpoints and various components of
drinking water. Kramer et al. (1992) conducted a population-based case-
control study to determine whether water supplies containing relatively
high levels of chloroform and other THMs within the state of Iowa are
associated with low birthweight, prematurity, or intrauterine growth
retardation (IUGR). Iowa birth certificate data from January, 1989
through June, 1990 served as the source of both cases and controls.
Definitions for cases and controls were as follows: the low birthweight
group included 159 live singleton infants weighing <2,500 grams and 795
randomly selected control infants weighing 2,500 grams from
the same population; the prematurity group included 342 live singleton
infants with gestational ages of <37 weeks as determined from the
mother's reported last menstrual period, and 1,710 randomly selected
control infants with gestational ages 37 weeks; IUGR
analyses included 187 IUGR infants (defined as weighing less than the
5th percentile for a particular gestational age based on California
standards for non-Hispanic whites) and 935 randomly selected controls.
Exposure status was assigned to infants according to reported maternal
residence in a given municipality at the time of birth. The assigned
THM levels came from a water survey conducted in 1987 in the state of
Iowa so the exposure information came from aggregate data. Odds ratios
were computed using multiple logistic regression to control for
measured confounders (including smoking, but not alcohol consumption).
The authors reported an increased risk for IUGR associated with
residence in communities where chloroform levels exceeded 10 ug/l
(OR=1.8; 95% CI=1.1-2.9). Prematurity was not associated with
chloroform exposure and the risk for low birthweight was only slightly
increased (OR=1.3; 95% CI=0.8-2.2).
The authors considered the results of this study to be preliminary.
Accordingly, they should be interpreted with caution. They considered
the major limitations of the study to involve assessment and
classification of individual exposure, the potential misclassification
due to residential mobility and the fluctuation of THM levels.
Aschengrau et al. (1993) conducted a case-control study in
Massachusetts to determine the relationship between community drinking
water quality and a wide range of adverse pregnancy outcomes, including
congenital anomalies, stillbirths, and neonatal deaths. The data were
obtained during a previous study of 14,130 pregnant women who delivered
infants at Brigham and Women's Hospital in Boston between 1977 and
1980. Drinking water quality information came from routine analyses of
the metal and chemical content of Massachusetts public water. An
attempt was made to link each woman in the study to the result of the
water analyses conducted in her town at the time of her pregnancy.
Information was also obtained on drinking water source and chlorination
of surface water. Drinking water samples from 155 towns were linked to
2,348 pregnant women to estimate exposure for the case-control study.
A large number of exploratory analyses were conducted with this
data set, which demonstrated both increases and decreases in risk
associated with various water quality parameters. A higher frequency of
stillbirths was correlated with chlorination and detectable lead
levels, cardiovascular defects were associated with lead levels, CNS
defects with potassium levels, and face, ear, and neck anomalies with
detectable silver levels. A decrease in neonatal deaths was associated
with detectable fluoride levels.
The authors indicated that the findings from this study, being non-
specific, must be considered as preliminary given the problems and
limitations of the exposure assessment and the lack of an a priori
study hypothesis. They indicated a need for further research
(Aschengrau et. al. 1993).
The New Jersey Department of Health recently reported the results
of a cross-sectional study and a case-control study evaluating the
association of drinking water contaminants with birth weight and
selected birth defects (Bove et al. 1992a and b). Four counties
selected for the study were included because they had the highest
levels of monitored drinking water and they were served by well defined
public water systems which used ground and surface water, or a mixture
of these sources. The exposures evaluated total volatile organic
contaminants (VOCs) as well as individual VOCs such as
trichloroethylene, tetrachloroethylene, carbon tetrachloride, benzene
and THMs. The cross sectional study base included 81,055 live single
births and 599 single fetal deaths between January, 1985 and December,
1988; 593 mothers were interviewed in the case control study. Exposure
scenarios to THMs were stratified as follows: >20-40 g/L, >40-
60 g/L, >60-80 g/L, and >80 g/L.
In the cross sectional study, ORs with exposure to THMs >80
g/L were elevated for low term birth weight (OR=1.34; 95%
CI=1.13-1.6; adjusted OR=1.29; 95% CI=1.08-1.5), small for gestational
age (OR=1.22; 95% CI=1.12-1.3; adjusted OR=1.14; 95% CI=1.04-1.3), and
prematurity (OR=1.09; 95% CI=0.99-1.2; adjusted OR=1.04; 95% CI=0.94-
1.1). Among birth defects, the ORs were elevated for all surveillance
malformations: OR=1.53; 95% CI=1.14-2.1; central nervous system defects
OR=2.6; 95% CI=1.48-4.6; neural tube defects: OR=2.98; 95% CI=1.25-7.1;
and cardiac defects: OR= 1.44; 95% CI=0.97-2.1. In the case control
study, associations were found between THMs >80 g/L and neural
tube defects (OR=4.25; 95% CI=1.02-17.7) and between THM levels >15
g/L and cardiac defects (OR=2.0; 95% CI= 0.94-4.5). The
authors note their findings should be interpreted with caution because
of possible exposure misclassification, unmeasured confounding, and
associations which could be due to chance occurrences. Although the
case control study included interviews of mothers for information about
residence and various risk factors, the authors reported a number of
limitations in the interpretation of the results from the case control
study, especially as a result of selection bias. Evaluation of
selection bias indicated that the bias led to an overestimate of the
associations with THM levels.
Some members of the Reg Neg committee viewed that these studies
indicate the possibility of a reproductive risk related to exposure to
disinfectant by-products. As a result of this concern, EPA convened a
panel of experts to review the epidemiology studies described above
(USEPA, 1993a). The panel concluded that the studies by Bove et al.
(1992a and b) were useful for hypothesis generation and identification
of a number of areas for further research. The panel further concluded
that the findings were limited by a number of issues surrounding study
design and data analysis. Some of the limitations included untested
assumptions of maternal exposure to chlorinated water, limitations in
the exposure assessment for THMs and other disinfection by-products,
possibility for exposure misclassification, confounding risk factors
and that some of these findings may have been due to chance.
d. Request for Public Comments. EPA requests comments on the
significance of the epidemiological studies with chlorine and
chloramines as indicators of risk. EPA recognizes that there are
different interpretations of these epidemiological studies and
specifically solicits comment on the rationale for EPA's
interpretations. EPA further requests comments on the studies
suggesting a reproductive risk related to disinfectant by-product
exposure.
4. Chlorine Dioxide, Chlorite and Chlorate
Chlorine dioxide is used as a disinfectant in drinking water
treatment as well as an additive with chlorine to control tastes and
odors in water treatment. It has also been used for bleaching pulp and
paper, flour and oils and for cleaning and tanning of leather. Chlorine
dioxide is a strong oxidizer that does not react with organics in the
water, as does chlorine, to produce by-products such as the
trihalomethanes. Chlorine dioxide is fairly unstable and rapidly
dissociates into chlorite, and chloride in water. Chlorate may also be
formed as a result of inefficient generation or generation of chlorine
dioxide under very high or low pH conditions. The dissociation of
chlorine dioxide into chlorite and chloride may be reversible with some
chlorite converting back to chlorine dioxide if free chlorine is
available. Chlorite ion is generally the primary product of chlorine
dioxide reduction. The distribution of chlorite, chloride and chlorate
is influenced by pH and sunlight. Chlorite, (as the sodium salt), is
used in the onsite production of chlorine dioxide and as a bleaching
agent by itself, for pulp and paper, textiles and straw. Chlorite is
also used to manufacture waxes, shellacs and varnishes. Chlorate, as
the sodium salt, was once a registered herbicide to defoliate cotton
plants during harvest, to tan leather and in the manufacture of dyes,
matches, explosives as well as chlorite.
Occurrence and Human Exposure. Based on information from the Water
Industry Data Base (WIDB), it has been estimated that for large systems
(serving greater than 10,000 people), approximately 10% of community
surface water systems serving 12.4 million people and 1% of community
ground water systems, serving 0.2 million people currently use chlorine
dioxide for disinfection in the United States. It was assumed that none
of the smaller community systems (fewer than 10,000 people) use
chlorine dioxide (WIDB, 1990).
Table V-5 presents occurrence information available for chlorine
dioxide, chlorate, and chlorite in drinking water. Descriptions of
these surveys and other data are detailed in ``Occurrence Assessment
for Disinfectants and Disinfection By- Products (Phase 6a) in Public
Drinking Water,'' (USEPA, 1992a). Typical dosages of chlorine dioxide
used as a disinfectant in drinking water treatment facilities appear to
range from 0.6 to 1.0 mg/L. For plants using chlorine dioxide, median
concentrations of chlorite and chlorate were found to be 240 and 200
g/L, respectively. However, the data base upon which these
numbers are based is very limited. A more extensive discussion of
chlorine dioxide and chlorite occurrence is described in section VI. of
this preamble.
Table V-5:--Summary of Occurrence Data For Chlorine Dioxide and Chlorite
Occurrence of Chlorine Dioxide, Chlorate, and Chlorite in Drinking Water
----------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information -----------------------------------------------------
(No. of samples) Range Mean Median Other
----------------------------------------------------------------------------------------------------------------
AWWARF (1987) Finished Water From:
McGuire &
Meadow, 1988.
Lakes................ 1.0 mg/L\3\
Flowing Streams...... 0.6 mg/L\3\
Plants Using CIO2: Positive
Detections
Chlorite at the Plant 15-740 240 110 100%
(4).
EPA, 1992b\2\ Disinfection By- Chlorate at the Plant 21-330 200 220 100%
(1987-1991). Products Field (4).
Studies
Plants Not Using
CIO2:
Chlorate at the Plant <10-660 87 16 60%
(30).
Chlorate, Distr. <10-47 18 13 75%
System (4).
----------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Typical dosage used by treatment plants.
AWWARF American Water Works Association Research Foundation.
EPA Environmental Protection Agency.
No information is available on the occurrence of chlorine dioxide,
chlorate, and chlorite in food or ambient air. Currently, the Food and
Drug Administration (FDA) does not analyze for these compounds in
foods. Preliminary discussions with FDA suggest that there are not
approved uses for chlorine dioxide in foods consumed in the typical
diet. In addition, the EPA Office of Air and Radiation does not require
monitoring for these compounds in air. However, chlorine dioxide is
used as a sanitizer for air ducts (Borum, 1991).
EPA believes that drinking water is the predominant source of
exposure for these compounds. Air and food exposures are considered to
provide only small contributions to the total chlorine dioxide,
chlorate, and chlorite exposures, although the magnitude and frequency
of these potential exposures are issues currently under review.
Therefore, EPA is considering proposing to regulate these compounds in
drinking water with an RSC value of 80 percent, the current exposure
assessment policy ceiling. EPA requests any additional data on known
concentrations of chlorine dioxide, chlorate and chlorite in drinking
water, food and air.
Health Effects. The following health effects information is
summarized from the draft Drinking Water Health Criteria Document for
Chlorine Dioxide, Chlorite and Chlorate (USEPA, 1994c). Studies cited
in this section are summarized in the draft criteria document.
As noted above, chlorine dioxide is fairly unstable and rapidly
dissociates predominantly into chlorite and chloride, and to a lesser
extent, chlorate. There is a ready interconversion among chemical
species in water (before administration to animals) and in the gut
(after ingestion). Therefore, what exists in water or the stomach is a
mixture of these chemical species and possibly their reaction products
with the gastrointestinal contents. Thus, the toxicity information on
chlorite, the predominant degradation product of chlorine dioxide, may
also be relevant to characterizing chlorine dioxide toxicity. In
addition, studies conducted with chlorine dioxide may be relevant to
characterizing the toxicity of chlorite. As a result, the toxicity data
for one compound are considered applicable for addressing toxicity data
gaps for the other.
The main health effects associated with chlorine dioxide and its
anionic by-products include oxidative damage to red blood cells,
delayed neurodevelopment and decreased thyroxine hormone levels.
Chlorine dioxide, chlorite and chlorate are well absorbed by the
gastrointestinal tract and excreted primarily in urine. Once absorbed,
36Cl-radiolabeled chlorine dioxide, chlorite and chlorate are
distributed throughout the body. Lethality data for ingested chlorine
dioxide have not been located in the available literature. A lethal
concentration for guinea pigs by inhalation was reported at 150 ppm.
Oral LD50 values for chlorite have been reported at 100 to 140 mg/
kg in rats. A more recent study indicates that the oral LD50 may
be closer to 200 mg/L. Limited data suggest an oral LD50 value
between 500 to 1500 mg/kg for chlorate in dogs.
In subchronic and chronic studies, animals given chlorine dioxide
treated water exhibited osmotic fragility of red blood cells (1 mg/kg/
d), decreased thyroxine hormone levels (14 mg/kg/d), possibly due to
altered iodine metabolism and hyperplasia of goblet cells and
inflammation of nasal tissues. The nasal lesions are not considered
related to ingested chlorine dioxide. However, it is not clear if the
nasal effects are due to off-gassing of chlorine dioxide from the
sipper tube of the animal water bottles, or from dermal contact while
the animal drinks from the sipper tube. In addition, the chlorine
dioxide treated group drank water at a pH of 4.7 which may also have
contributed to the nasal tissue inflammation. The concentration
associated with this effect (25 mg/L) is considerably greater than what
would be found in drinking water.
Studies evaluating developmental or reproductive effects have
described decreases in the number of implants and live fetuses per dam
in female rats given chlorine dioxide in drinking water before mating
and during pregnancy. Delayed neurodevelopment has been reported in rat
pups exposed perinatally to chlorine dioxide (14 mg/kg/d) or chlorite
(3 mg/kg/d) treated water. Delayed neurodevelopment was assessed by
decreased locomotor activity and decreased brain development.
Subchronic studies with chlorite administered to rats via drinking
water resulted in transient anemia, decreased red blood cell
glutathione levels and increased hydrogen peroxide formation at doses
greater than 5 mg/kg/d. Chlorite administration orally to cats at a
dose of 7 mg/kg/d produced 10 to 40 percent methemoglobin formation
within a couple of hours following dosing. Exposure to chlorite in
drinking water resulted in an increased turnover of red blood cells in
cats rather than oxidation of hemoglobin.
Oral studies with chlorate also demonstrate effects on
hematological parameters and formation of methemoglobin, but at much
higher doses than chlorite (157-256 mg/kg/d).
No clear tumorigenic activity has been observed in animals given
oral doses of chlorine dioxide, chlorite or chlorate. Chlorine dioxide
concentrates did not increase the incidence of lung tumors in mice nor
was any initiating activity observed in mouse skin or rat liver
bioassays. Lung and liver tumors were increased in mice given sodium
chlorite; however, the incidence was within the historical range for
these tumor types. Carcinogenic studies on chlorate were not located in
the available literature. Chlorate has been reported to be mutagenic in
bacterial and Drosophila tests. EPA has classified chlorine dioxide and
chlorite in Group D: not classifiable as to human carcinogenicity. This
classification is for chemicals with inadequate evidence or no data
concerning carcinogenicity in animals in the absence of human data. EPA
has not classified chlorate with respect to carcinogenicity.
There are a number of cases of poisoning in humans who used
chlorate as an herbicide. Effects observed following exposures to 11 to
3,400 mg/kg include cyanosis, renal failure, convulsions and death. The
lowest lethal dose reported in adults is approximately 200 mg/kg. It is
not clear if this is the actual dose received or if other components in
the formulation were contributors to the toxicity. In an epidemiology
study of a community where chlorine dioxide was used as the primary
drinking water disinfectant for 12 weeks, no consistent changes were
observed in the clinical parameters measured.
Three studies have been selected as the basis for the RfD and MRDLG
for chlorine dioxide. These studies identify a NOAEL of 3 mg/kg/d and a
LOAEL of approximately 10 mg/kg/d. A NOAEL of 3 mg/kg/d has been
identified in an 8 week rat study by Orme et al. (1985). In this study,
chlorine dioxide was administered to female rats via drinking water at
concentrations of 0, 2, 20 and 100 mg/L before mating, during gestation
and lactation until the pups were 21 days old. Based on body weight and
water consumption data, these concentrations correspond to doses of 1,
3 and 14 mg/kg/d. No effects were noted in dams. Pups in the high dose
group (14 mg/kg/d) exhibited decreased exploratory and locomotor
activity and a significant depression of thyroxine. These effects were
not observed at the 3 mg/kg/d dose level. In a second experiment, pups
were given 14 mg/kg/d chlorine dioxide directly by gavage during
postnatal days 5 through 20. A greater and more consistent delay in
neurobehavioral activity was observed along with a greater depression
in thyroxine. Analysis of the DNA content of cells in the cerebellum
from animals in the high dose drinking water group (14 mg/kg/day) at
postnatal day 21 and the gavage group at day 11 indicated a significant
depression (Taylor and Pfohl, 1985). Another study confirmed 14 mg/kg/d
as a LOAEL based on decreased brain cell proliferation in rats exposed
postnatally by gavage (Toth et al., 1990).
The no-effect level of 3 mg/kg/day is also supported by a monkey
study (Bercz et al., 1982), where animals were given chlorine dioxide
at concentrations of 0, 30, 100 or 200 mg/L in drinking water following
a rising dose protocol. These concentrations correspond to doses of 0,
3.5, 9.5 and 11 mg/kg/d based on animal body weight and water
consumption. Animals showed signs of dehydration at the high dose;
exposure was discontinued at that dose (11 mg/kg/d). A slight
depression of thyroxine was observed following exposure to 9.5 mg/kg/d.
No effects were seen with 3.5 mg/kg/d, which is considered the NOAEL.
MRDLG for Chlorine Dioxide. EPA is proposing an MRDLG for chlorine
dioxide based on developmental neurotoxicity following a Category III
approach. Using a NOAEL of 3 mg/kg/d and an uncertainty factor of 300,
an RfD of 0.01 mg/kg/d for chlorine dioxide is calculated. An
uncertainty factor of 300 is used to account for differences in
response to toxicity within the human population and between humans and
animals. This factor also accounts for lack of a two-generation
reproductive study. Availability of an acceptable two-generation
reproduction study would likely reduce the total uncertainty factor to
100. The Chlorine Dioxide panel of the Chemical Manufacturers
Association is conducting a two-generation reproductive study with
chlorite to address this data gap. EPA will review the results of this
study and determine if any changes to the RfD for chlorine dioxide are
warranted.
After adjusting for an adult consuming 2 L water per day, an RSC of
80 percent is applied to calculate an MRDLG of 0.3 mg/L. An RSC of 80%
is used since most chlorine dioxide exposure is likely to come from a
drinking water source.
TP29JY94.004
The Drinking Water Committee of the Science Advisory Board (SAB)
agreed with the use of the Orme et al. (1985) study as the basis for
the MRDLG and suggested that an uncertainty factor of 100 be applied
(USEPA, 1992c). They also suggested that a child's body weight of 10 kg
and water consumption of 1 L/d may be more appropriate for setting the
MRDLG than the adult parameters, given the acute nature of the toxic
effect. EPA requests comments on the SAB's suggestion.
EPA also requests comment on the appropriateness of the 300-fold
uncertainty factor, the studies selected as the basis for the RfD, and
the 80% relative source contribution.
MCLG for Chlorite. The developmental rat study by Mobley et al.
(1990) has been selected to serve as the basis for the RfD and MCLG for
chlorite. Other studies reported effects at doses higher than the
Mobley et al. study. In this study, female Sprague-Dawley rats (12/
group) were given drinking water containing 0, 20, or 40 mg/L chlorite
(0, 3, or 6 mg chlorite ion/kg/day) as the sodium salt beginning 10
days prior to breeding with untreated males until the pups were
sacrificed at 35 to 42 days postconception (a total exposure of 9
weeks). Exploratory activity was depressed in the pups treated with 3
mg/kg/day chlorite on postconception days 36-37 but not on days 38-40.
Pups from the high exposure group also exhibited depressed exploratory
behavior during days 36-39 postconception (p<0.05). Exploratory
activity was comparable among the treated and control groups on
postconception days 39-41. No significant differences in serum total
thyroxine or triiodothyronine were observed between treated and control
pups. Free thyroxine was significantly elevated in the 6 mg/kg/day
pups. A LOAEL of 3 mg/kg/day was determined in this study based on the
neurobehavioral effect (depressed exploratory behavior) in rats. This
endpoint is similar to that reported for chlorine dioxide.
EPA had considered using a study by Heffernan et al. (1979) which
described dose-related decreases in red blood cell glutathione levels
from rats orally exposed to chlorite in drinking water for up to 90
days. The decreases in glutathione were accompanied by decreases in red
blood cell concentration, hemaglobin concentration and packed red cell
volume. Taken together, these effects were considered reflective of
oxidative stress resulting from the ingested chlorite. In this study, a
NOAEL of 1 mg/kg/d and LOAEL og 5 mg/kg/d were identified.
The EPA Science Advisory Board had cautiously agreed with the
selection of the Heffernan et al. (1979) study as the basis for the
RfD, but noted that the endpoint would likely be controversial since
normal fluctuations occur with glutatione levels. Thus this effect,
alone, may not necessarily be the result of chlorite exposure. The EPA
RfD workgroup was unable to reach consensus on decreased glutathione
levels as an appropriate endpoint to base an RfD. They agreed with the
selection of the Mobley et al. (1990) study since the endpoint,
developmental neurotoxicity, represented the next critical effect and
was consistent with the toxicity observed with chlorine dioxide.
Following a Category III approach, EPA is proposing an MCLG of 0.08
for chlorite. The MCLG is based on an RfD of 0.003 determined from the
LOAEL of 3 mg/kg/day from the Mobley et al. study. This endpoint was
selected since it is similar to that reported for chlorine dioxide. An
uncertainty factor of 1,000 is used in the derivation of the RfD and
MCLG to account for use of a LOAEL from an animal study.
After adjusting for an adult consuming 2 L water per day, an RSC of
80% is applied to calculate an MCLG of 0.08 mg/L. An RSC of 80% was
used since most exposure to chlorite is likely to come from drinking
water.
TP29JY94.005
The Drinking Water Committee of the EPA Science Advisory Board
suggested that EPA consider basing the MCLG on the child body weight of
10 kg and water consumption of 1 L/day instead of the adult default
values. EPA requests comments on the SAB's suggestion along with the
study selected as the basis for the MCLG, the uncertainty factor and
the RSC of 80%.
MCLG for Chlorate. Data are considered inadequate to develop an
MCLG for chlorate at this time. A NOAEL of 0.036 mg/kg/d (the only dose
tested) was identified in the Lubbers et al. (1982) human clinical
study following a 12-week exposure to chlorate in drinking water.
NOAELs identified from animal studies are considerably higher
(approximately 78 mg/kg/d). However, doses that are lethal to humans
(200 mg/kg/d) are only 2-fold greater than this animal no-effect level.
No information is available to characterize the potential human
toxicity between the doses of 0.036, the only human NOAEL and 200 mg/
kg/d, the apparent human lethal dose. Thus, EPA considers the data base
too weak to derive a separate MCLG for chlorate at this time. The
Agency will continue to evaluate the animal data and any new
information that become available for future consideration of an MCLG
for chlorate.
EPA requests comments on the decision not to propose an MCLG for
chlorate at this time.
5. Chloroform
Chloroform [trichloromethane, CAS No. 67-66-3] is a nonflammable,
colorless liquid with a sweet odor and high vapor pressure (200 mm Hg
at 25 deg.C). It is moderately soluble in water (8 gm/L at 20 deg.C)
and soluble in organic solvents (log octanol/water partition
coefficient of 1.97). Chloroform is used primarily to manufacture
fluorocarbon-22 (chlorodifluoromethane) which in turn is used for
refrigerants and fluoropolymer synthesis. A small percentage of the
manufactured chloroform is used as an extraction solvent for various
products (e.g. resins, gums). In the past, chloroform was used in
anesthesia and medicinal preparations and as a grain fumigant
ingredient. Chloroform can be released to the environment from direct
(manufacturing) and indirect (processing/use) sources and chloroform is
a prevalent chlorination disinfection by-product. Volatilization is the
principle mechanism for removal of chloroform from lakes and rivers.
Chloroform bioconcentrates slightly in aquatic organisms and adsorbs
poorly to sediments and soil. Chloroform can be biodegraded in water
and soil (half-life of weeks to months) and ground water (half-life of
months to years), and photo-oxidized in air (half-life of months).
Occurrence and Human Exposure. The principle source of chloroform
in drinking water is the chemical interaction of chlorine with commonly
present natural humic and fulvic substances and other precursors
produced by either normal organic decomposition or by the metabolism of
aquatic biota. Because humic and fulvic material are generally found at
much higher concentrations in surface water sources than in ground
water sources, surface water systems have higher frequencies of
occurrence and higher concentrations of chloroform than ground water
systems. Several water quality factors affect the formation of
chloroform including Total Organic Carbon (TOC), pH, and temperature.
Different treatment practices can reduce the formation of chloroform.
These include the use of precursor removal technologies such as
coagulation/filtration, granular activated carbon (GAC), and membrane
filtration and the use of chlorine dioxide, chloramination, and
ozonation.
Table V-6 presents occurrence information available for chloroform
in drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' (USEPA, 1992a). The
table lists six surveys conducted by Federal and private agencies.
Median concentrations of chloroform in drinking water appear to range
from 14 to 57 g/L for surface water supplies and
0.5 g/L for ground-water supplies (many of which do
not disinfect). The lower bound median concentration for chloroform in
surface water supplies is biased to the low side because concentrations
in this survey were measured in the plant effluent; the formation of
chloroform would be expected to increase in the distribution in systems
using chlorine as their residual disinfectant.
Table V6.--Summary of Occurrence Data For Chloroform
[Occurrence of Chloroform in Drinking Water]
----------------------------------------------------------------------------------------------------------------
Concentration (g/L
Survey (year) Location Sample information -----------------------------------------------------
(No. of samples) Rage Mean Median Other
----------------------------------------------------------------------------------------------------------------
CWSS (1978) Brass 450 Water Supply Finished Water ........... ........... ........... Positive
et al., 1981. Systems (1,100):. Detections:
................ Surface Water........ ........... \3\60 ........... \4\97%
................ Ground Water......... ........... \3\<0.5 ........... \4\34%
RWS (1978-1980) >600 Rural Drinking Water from:. ........... ........... ........... Postive
Brass, 1981. Systems (>2,000 Detection:
Households)
................ Surface Water........ ........... \3\84 57 \4\82%
................ Ground Water......... ........... \3\8.9 <0.5 \4\17%
GWSS (1980-1981) 945 GW Systems: ..................... ........... ........... ........... 90th
Westrick et al. percentile:
1983.
(466 Random and Serving >10,000 (327) Max. 300 ........... 0.5 17
479 Nonrandom) Serving <10,000 (618) Max. 430 ........... 0 7.8
EPA, 1991a\2\ Unregulated Sampled at the Plant ........... 17 5 .............
(1984-1991). Contaminant (5,806).
Data Base--
Treatment
Facilities from
19 States
EPA, 1992b\2\ Disinfection By- Finished Water:...... ........... ........... ........... Positive
(1987-1991). Products Field Detections:
Studies
................ At the Plant (73).... <0.2-240 36 28 96%
................ Distribution System <0.2-340 57 42 98%
(56).
EPA/AMWA/CDHS\2\ 35 Water Samples from Max. 130 ........... \5\9.6-15 75% of Data
(1988-1989). Utilities Clearwell. was Below 33
Nationwide gL.
Krasner et al., ................ Effluent for 4 ........... ........... 14 .............
1989b. Quarters.
----------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Mean of the positives.
\4\Of systems sampled.
\5\Range of medians for individual quarters.
AMWA Association of Metropolitan Agencies.
CDHS California Department of Health Services.
CWSS Community Water Supply Survey.
GWSS Ground Water Supply Survey.
RWS Rural Water Survey.
EPA Environmental Protection Agency.
Several studies have assessed inhalation exposure to chloroform.
The major source of these data is from the USEPA's Total Exposure
Assessment Methodology (TEAM) studies, which measured chloroform
exposure to approximately 750 persons in eight geographic areas from
1980 to 1987. Personal exposure to chloroform from air was measured
over a 12-hour period (excluding showers) for individuals in three
areas. The average exposures were reported to range from 4 to 9
g/m3 in New Jersey and Baltimore, and about 0.5 to 4
g/m3 in California cities (Wallace, 1992). In the 1987
Los Angeles TEAM study, chloroform in indoor air was measured in the
living room and kitchen of private residences. Observed mean indoor
concentrations ranged from 0.9 to 1.5 g/m3 (Pellizzari et
al., 1989 and Wallace et al., 1990 in Wallace, 1992). For outdoor
levels, the 12-hour average outdoor concentrations measured in the
California and New Jersey TEAM studies ranged from 0.2 to 0.6
g/m3 and 0.1 to 1.5 g/m3, respectively
(Pellizzari et al., 1989, Wallace et al., 1990 and PEI et al., 1989 in
Wallace, 1992).
Given the limited exposure data in air, inhalation exposure can be
estimated using an inhalation rate of 20 m3/day. Resulting
estimates for average ambient air exposures range from 2 to 30
g/d and 18-30 g/d for average indoor air exposures.
However, based on the personal air monitoring data, a potentially
higher average inhalation exposure is indicated with a range of 10 to
180 g/d.
Two studies analyzed some foods for chloroform. In a pilot market
basket survey of four food groups at five sites, measured chloroform
levels were as follows: dairy composite, 17 ppb (1 of 5 sites); meat
composite, not detected; oil and fat composite, trace amounts (1 of 5
sites); beverage composite, 6 to 32 ppb (4 of 5 sites) (Entz et al.,
1982). In a study of 15 table-ready food items, chloroform was detected
in 53% of the foods tested: butter, 670 ppb; cheddar cheese, 80 ppb;
plain granola, 57 ppb; peanut butter, 29 ppb; chocolate chip cookies,
22 ppb; frozen fried chicken dinner, 29 ppb; and high meat dinner, 17
ppb (Heikes, 1987).
Limited data are available to characterize dietary exposure to
chloroform. Although some uses of chlorine have been identified in the
food production/food processing area, monitoring data are not adequate
to characterize the magnitude or frequency of exposure to chloroform.
Based on the limited number of food groups that are believed to contain
chloroform and low levels expected in ambient and indoor air, EPA
assumes that drinking water is the predominant source of chloroform
intake. The characterization of potential food and air exposures are
issues currently under review. EPA requests any additional data on
known concentrations of chloroform in drinking water, food, and air.
Health Effects. The health effects information is summarized from
the draft Drinking Water Criteria Document for Trihalomethanes (USEPA,
1994d). Studies cited in this section are summarized in the criteria
document.
Chloroform has been shown to be rapidly absorbed upon oral,
inhalation and peritoneal administration and subsequently metabolized.
The reported mean human lethal dose, from clinical observations of
overdoses, was around 630 mg/kg. The LD50 values in mice and rats
have been reported in the range of 908-1,400 mg/kg. Several reactive
metabolic intermediates (e.g. phosgene, carbene, dichloromethyl
radicals) can be produced via oxidation (major pathway) or reduction
(minor pathway) by microsomal preparations. Experimental studies
suggested that these active metabolic intermediates are responsible for
the hepatic and renal toxicity and possibly, carcinogenicity, of the
parent compound. Animal studies suggest that the extent of chloroform
metabolism varies with species and sex. The retention of chloroform in
organs after dosing was small. Due to the lipophilic nature of the
compound, the residual concentration is in tissues with higher fatty
content. In humans, the majority of the tested oral intake doses (0.1
to 1 gm) were excreted through the lungs in the form of a metabolite
CO2 or as the unchanged compound. Urinary excretion levels were
below 1%.
Mammalian bioeffects following exposure to chloroform include
effects on the central nervous system (CNS), hepatotoxicity,
nephrotoxicity, reproductive toxicity and carcinogenicity. Chloroform
caused CNS depression and affected liver and kidney function in humans
in both accidental and long term occupational exposure situations. In
experimental animals, chloroform caused changes in kidney, thyroid,
liver, and serum enzyme levels. These responses are discernible in
mammals from exposure to levels of chloroform ranging from 15 to 290
mg/kg; the intensity of response was dependent upon the dose and the
duration of the exposure. Ataxia and sedation were noted in mice
receiving a single dose of 500 mg/kg chloroform. Short-term exposure to
the low levels of chloroform typically found in air, food, and water
are not known to manifest acute toxic effects. The potential for human
effects from chronic lifetime exposure is the basis for this
regulation.
Developmental toxicity and reproductive toxicity have been
investigated in animals. One developmental study reported maternal
toxicity in rabbits administered chloroform by the oral route.
Decreased weight gain and mild fatty changes in liver were observed in
dams receiving 50 mg/kg/day (LOAEL); the maternal NOAEL was noted to be
35 mg/kg/day. There was no evidence of developmental effects.
The data from a 7.5-year oral study in dogs conducted by Heywood et
al. (1979) were used to calculate the RfD. EPA considers this study
suitable for the RfD derivation since it is a chronic study and
sensitive indices of hepatotoxicity (serum enzyme levels, liver
histology) of sufficient numbers of experimental animals were
monitored. In this study, chloroform was administered to beagle dogs
(16 per dose group) in toothpaste base gelatin capsules at dose levels
of 15 or 30 mg/kg/day 6 days/week for 7.5 years. A LOAEL of 15 mg/kg/
day was established based on the observation of hepatic fatty cysts in
treated animals at both doses. An RfD of 0.01 mg/kg/day has been
derived from this LOAEL by the application of an uncertainty factor of
1,000, in accordance with EPA guidelines.
The results of a number of assays to determine the mutagenicity
potential of chloroform are inconclusive. Studies on the in vitro
genotoxicity of chloroform reported negative results in bacteria (Ames
assays), negative results for gene mutations and chromosomal
aberrations in mammalian cells, and mixed results in yeasts. In vivo
and in vitro DNA damage tests indicate that chloroform will bind to
DNA. Gene mutation tests in Drosophila were marginal, whereas tests for
chromosomal aberrations and sperm abnormalities were mixed.
Several chronic animal studies confirmed the carcinogenicity of
chloroform. Chloroform induced hepatocellular carcinomas in mice when
administered by gavage in corn oil (NCI, 1976). Chloroform also induced
renal adenomas and adenocarcinomas in male rats regardless of the
carrier vehicle (oil or drinking water) employed (NCI, 1976; Roe et
al., 1979; Jorgenson et al., 1985).
In the study by Jorgenson et al. (1985), chloroform was
administered in drinking water to male Osborne-Mendel rats and female
B6C3F1 mice at doses of 0, 200, 400, 900 or 1,800 ppm (0, 19, 38,
81 or 160 mg/kg/day in rats and 0, 34, 65, 130 or 263 mg/kg/day in
mice) for 2 years. Chloroform increased the incidence of kidney tumors
in male rats in a dose-related manner. The combined incidence of renal
tubular cell adenomas, renal tubular cell adenocarcinomas, and
nephroblastomas in control, 200, 400, 900 and 1,800 ppm groups were 5/
301, 6/313, 7/148, 3/48, and 7/50, respectively. Jorgenson's study
reported no statistically significant increase in the incidence of
hepatocellular carcinomas in the female mice exposed to similar doses
of chloroform as reported in the 1976 NCI study.
Since hepatic changes appeared to be related to the corn oil
vehicle, the interaction of corn oil and chloroform could account for
the enhanced hepatic toxicity and thus for the difference in the NCI
and Jorgenson studies. Because the drinking water study did not
replicate hepatic tumors in female mice and the potential role of corn
oil in enhancing toxicity, the National Academy of Science Subcommittee
on the Health Effects of Disinfectants and Disinfection By-Products
(NAS, 1987) recommended that male rat kidney tumor data obtained from
Jorgenson's study be used to estimate the carcinogenic potency of
chloroform. EPA agreed with the NAS Subcommittee recommendation for
estimating risks of chloroform from drinking water exposures.
Based on all kidney tumor data in male Osborne-Mendel rats reported
by Jorgenson et al. (1985), EPA used a linearized multistage model and
derived a carcinogenic potency factor for chloroform of 6.1 x
10-3 (mg/kg/day)-1. Assuming a daily consumption of two
liters of drinking water and an average human body weight of 70 kg, the
95% upper bound limit lifetime cancer risk levels of 10-6,
10-5, and 10-4 are associated with concentrations of
chloroform in drinking water of 6, 60 and 600 g/L,
respectively.
In 1987 the Commission on Life Sciences of the National Research
Council published Drinking Water and Health (NAS, 1987). Volume 7,
Disinfectants and Disinfectant By-Products, prepared by the Subcomittee
on Disinfectants and Disinfection By-Products, discussed the available
data on chloroform, which are the same data summarized above. The
Subcommittee concluded that ``[n]oting that chloroform is the principal
THM produced by chlorination, the subcommittee found [the 100 THM]
level to be unsupportable on the basis of the risk values for
chloroform developed in this review,'' and that the level should be
reduced.
EPA has classified chloroform in Group B2, probable human
carcinogen, based on sufficient evidence of carcinogenicity in animals
and inadequate evidence in humans (IRIS, 1985). The International
Agency for Research on Cancer (IARC) has classified chloroform as a
Group 2B carcinogen, agent possibly carcinogenic to humans. (IARC,
1982).
According to EPA's three-category approach for establishing MCLGs,
chloroform is placed in Category I since there is sufficient evidence
of carcinogenicity via ingestion considering weight of evidence,
potency, pharmacokinetics, and exposure. Thus, EPA is proposing an MCLG
of zero for this contaminant. EPA requests comment on the basis for the
proposed MCLG for chloroform.
6. Bromodichloromethane
Bromodichloromethane (BDCM; CAS No. 75-27-4) is a nonflammable,
colorless liquid with a relatively high vapor pressure (50 mmHg at
20 deg.C). BDCM is moderately soluble in water (3.3 gm/L at 30 deg.C)
and soluble in organic solvents (log octanol/water partition
coefficient of 1.88). Only a small amount of BDCM is currently produced
commercially in the United States. The chemical is used as an
intermediate for organic synthesis and as a laboratory reagent. The
principle source of BDCM in drinking water is the chemical interaction
of chlorine with the commonly present organic matter and bromide ions.
Degradation of BDCM is not well studied, but probably involves
photooxidation. The estimated atmospheric half-life of BDCM is two to
three months. Volatilization is the principal mechanism for removal of
BDCM from rivers and streams (half-life of hours to weeks). Limited
studies reported that BDCM adsorbed poorly to sediments and soils. No
study of bioaccumulation of BDCM was located. Based on the data of a
few structurally similar chemicals such as chloroform, the
bioconcentration potential of BDCM in aquatic organisms is low.
Biodegradation of BDCM is limited under aerobic conditions and
extensive (completion within days) under anaerobic conditions.
Occurrence and Human Exposure. BDCM, occurs in public water systems
that chlorinate water containing humic and fulvic acids and bromine
that can enter source waters through natural and anthropogenic means.
Several water quality factors affect the formation of BDCM including
Total Organic Carbon (TOC), pH, bromide, and temperature. Different
treatment practices can reduce the formation of BDCM. These include the
use of chlorine dioxide, chloramination, and ozonation prior to
chloramination, as well as the use of precursor removal technologies
such as coagulation/filtration, granular activated carbon (GAC), and
membrane filtration.
Table V-7 presents occurrence information available for BDCM in
drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' (USEPA, 1992a). The
table lists six surveys conducted by Federal and private agencies.
Median concentrations of BDCM in drinking water appear to range from
6.6 to 15 g/L for surface water supplies and <0.5 g/L
for ground-water supplies. The lower bound median concentration for
BDCM in surface water supplies is biased to the low side because
concentrations in this survey were measured in the plant effluent; the
formation of BDCM would be expected to increase in the distribution
system when chlorine is used as the residual disinfectant.
Table V-7: Summary of Occurrence Data for Bromodichloromethane
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of Bromodichloromethane
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (Year)\1\ Location Sample Information(No. of ----------------------------------------------------------------------
Samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
CWSS (1978, Brass et 450 Systems............. Finished Water (1,100): Positive Detections:
al., 1981. Surface Water................ \3\12 6.8 94%\4\
Ground Water................. \3\5.8 <0.5 33%\4\
RWS (1978-1980) Brass, >600 Rural Systems Drinking Water from: Positive Detections:
1981. (>2,000 households). Surface Water................ ............. ........... 11 76%\4\
Ground water................. ............. ........... <0.5 13%\4\
GWSS (1980-1981) 945 GW Systems: ............. 90Percentile
Westrick et al. 1983. (466 Random and 479 Serving >10,000 (327)........ Max. 110..... ........... 0.4 9.2
Nonrandom).. Serving <10,000 (618)........ Max. 79...... ........... 0 6.1
EPA, 1991a\2\ (1984- Unregulated Contaminant Finished Water at Treatment ............. 5.6 3
1991). Data Base--Treatment Plants (4,439).
Facilities from 19
States.
EPA, 1992b\2\ (1987- Disinfection By-Products Finished Water: Positive Detections:
1989). Field Studies. At the Plant (73)............ <0.2-90...... 13 11 96%
Distribution System (56)..... <0.2-100..... 17 15 98%
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell Max. 82...... 4.1-10\5\ 6.6 75% of Data was Below 14
1989) Krasner et al., Nationwide. Effluent for 4 Quarters. g/L.
1989b.
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Mean of the positives.
\4\Of systems sampled.
\5\Range of medians for individual quarters.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
CWSS: Community Water Supply Survey.
GWSS: Ground Water Supply Survey.
RWS: Rural Water Survey.
EPA: Environmental Protection Agency.
BDCM is usually found in air at low concentrations. Based on
information obtained through a literature review, Howard (1990)
estimated the average daily intake of BDCM from air using an inhalation
rate of 20 m\3\/day. Assuming a range of 6.7 to 670 ng/m\3\, the
average exposure may be as low as 0.134 g/day or as high as
13.4 g/day.
BDCM is not a common contaminant in food. In one market study of 39
different food items, BDCM was detected in one dairy composite (1.2
ppb), butter (7 ppb), and two beverages (0.3 and 0.6 ppb). Analysis of
cola soft drinks found BDCM in three samples with reported
concentrations of 2.3 ppb, 3.4 ppb, and 3.8 ppb (Entz et al., 1982 in
Howard, 1990).
Limited data are available to characterize food and air exposures
to BDCM. Although some uses of chlorine have been identified in the
food production/food processing area, monitoring data are inadequate to
characterize the magnitude and frequency of potential BDCM exposures.
Based on the limited number of food groups that are believed to contain
BDCM and that there are not significant levels expected in ambient or
indoor air, EPA assumes that drinking water is the predominant source
of BDCM intake. EPA requests any additional data on known
concentrations of BDCM in drinking water, food, and air.
Health Effects. The health effects information in this section is
summarized from the draft Drinking Water Criteria Document for
Trihalomethanes (USEPA, 1994d). Studies mentioned below are summarized
in the criteria document.
Studies indicated that gastrointestinal absorption of BDCM is high
in animals. No studies were located regarding BDCM in humans or animals
following inhalation or dermal exposure. By analogy with the
experimental data of a structurally-related halomethane chloroform,
inhalation and dermal absorption may be high for BDCM. The reported
LD50 values in mice and rats ranged from 450 to 969 mg/kg. Under
both in vivo and in vitro conditions, several active metabolic
intermediates (e.g. dichlorocarbonyl, dichloromethyl radicals) were
produced via oxidation or reduction by microsomal preparations.
Experimental studies suggested that these active metabolic
intermediates may be responsible for hepatic and renal toxicity and
possibly, carcinogenicity of the parent compound. Animal studies
suggest that the extent of BDCM metabolism varies with species and sex.
The retention of BDCM in organs after dosing was small, even after
repeated doses. Urinary excretion levels were below 3 percent.
Mammalian bioeffects following exposure to BDCM include effects on
the central nervous system (decreased operant response),
hepatotoxicity, nephrotoxicity, reproductive toxicity, and
carcinogenicity. In experimental mice and rats, BDCM caused changes in
kidney, liver, serum enzyme levels, and decreased body weight. These
responses were discernible in rodents from exposure to levels of BDCM
that ranged from 6 to 300 mg/kg; the intensity of response was
dependent upon the dose and the duration of the exposure. Ataxia and
sedation were observed in mice receiving a single dose of 500 mg/kg
BDCM.
One study investigated developmental and reproductive toxicity of
BDCM in rodents. Ruddick et al. (1983) administered BDCM in corn oil to
groups of pregnant rats by gavage at doses of 0, 50, 100 or 200 mg/kg/
day on days 6 to 15 of gestation. At 200 mg/kg/day, BDCM significantly
(p <0.05) decreased maternal weight (25%) and increased relative kidney
weights. There were no increases in the incidence of fetotoxicity or
external/visceral malformations, but sternebral anomalies were more
prevalent at 100 and 200 mg/kg than at 50 mg/kg. The sternebral
anomalies were not considered by the authors to be evidence of a
teratogenic effect, but rather evidence of maternal toxicity.
Data from a National Toxicology Program (NTP) chronic oral study in
B6C3F1 mice (NTP, 1987) was used to calculate the RfD. BDCM in
corn oil was given to mice by gavage 5 days/week for 102 weeks. Male
mice (50/dose) were administered doses of 0, 25 or 50 mg/kg/day while
female mice (50/dose) received doses of 0, 75 or 150 mg/kg/day.
Following treatment, mortality, body weight and histopathology were
observed. Renal cytomegaly and fatty metamorphosis of the liver was
observed in male mice 25 mg/kg/day). Compound-related
follicular cell hyperplasia of the thyroid gland was observed in both
males and females. The survival rate decreased in females and decreases
in mean body weights were observed in both males and females at high
doses. Based on the observed renal, liver and thyroid effects in male
mice, a LOAEL of 25 mg/kg/day was identified. A RfD of 0.02 mg/kg/day
has been derived from the LOAEL of 25 mg/kg/day in mice by the
application of an uncertainty factor of 1,000, in accordance with EPA
guidelines for use of a LOAEL derived from a chronic animal study.
In vitro genotoxicity studies reported mixed results in bacterial
Salmonella strains and yeasts. BDCM was not mutagenic in mouse lymphoma
cells without metabolic activation, but induced mutation with
activation. An increase in frequency of sister chromatid exchange was
reported in cultured human lymphocytes, rat liver cells, and mouse bone
marrow cells (in vivo), but not in Chinese hamster ovary cells.
Overall, more studies yielded positive results and evidence of
mutagenicity for BDCM is considered adequate.
Evidence of the carcinogenicity of BDCM has been confirmed by a NTP
(1987) chronic animal study. In this study BDCM in corn oil was
administered via gavage to groups of 50 rats (Fischer 344/N) of each
sex at doses of 0, 50 or 100 mg/kg, 5 days/week, for 102 weeks (NTP,
1987). Male B6C3F1 mice (50/dose) were administered 0, 25 or 50
mg/kg by the same route while females received 0, 75 or 150 mg/kg/day.
BDCM caused statistically significant increases in kidney tumors in
male mice, the liver in female mice, and the kidney and large intestine
in male and female rats. In male mice, the combined incidence of
tubular cell adenomas or adenocarcinomas of the kidneys increased
significantly in the high-dose group (vehicle control, 1/46; low-dose,
2/49; high-dose 9/50). The combined incidences of hepatocellular
adenomas or carcinomas in vehicle control, low-dose and high-dose
female mice groups were 3/50, 18/48 and 29/50, respectively.
In rats from the NTP study, the combined incidences of tubular cell
adenomas or adenocarcinomas in vehicle control, low-dose and high-dose
groups were 0/50, 1/49 and 13/50 for males and 0/50, 1/50 and 15/50 for
females, respectively. Tumors of large intestines were significantly
increased in a dose-dependent manner in male rats, and only observed in
high-dose female rats. The combined incidences of adenocarcinomas or
adenomatous polyps were 0/50, 13/49, 45/50 for males and 0/46, 0/50,
12/47 for females, respectively. The combined tumor incidences of large
intestine and kidney were 0/50, 13/49, 46/50 for male rats and 0/46, 1/
50, 24/48 for female rats, respectively.
Using the linearized multistage model, several cancer potency
factors for BDCM were derived based on the observed cancer incidence of
various tumor types (large intestine, kidney, or combined) in mice or
rats reported in the NTP bioassay. The resulting cancer potency factors
are in the range of 4.9 x 10-3 to 6.2 x 10-2 (mg/kg/
day)-1. A potency factor of 1.3 x 10-1 (mg/kg/day)-1 was
derived from the incidence of hepatic tumors in female mice (IRIS,
1990). However, hepatic tumor data should be interpreted with caution
because studies of an analog chloroform indicated a possible role of
the corn oil vehicle in induction of these tumors. Until future studies
can provide a better understanding of the corn oil effect on hepatic
carcinogenicity, EPA considers carcinogenic risk quantification for
BDCM based on kidney or large intestine tumor data to be more
appropriate. EPA is presently conducting a cancer bioassay with BDCM in
drinking water for comparison with the NTP study. EPA will evaluate the
results of this study when available to determine if changes to the
risk assessment are warranted.
Following the Agency's Cancer Risk Assessment Guidelines (USEPA,
1986), when two or more significantly elevated tumor sites or types are
observed in the same study, the slope factor of the greatest
sensitivity preferably should be used for carcinogenic risk estimation.
Based on the potency factor of 6.2 x 10-2 (mg/kg/day)-1
derived from the kidney tumor incidence in male mice, the estimated
concentrations of BDCM in drinking water associated with excess cancer
risks of 10-4, 10-5 and 10-6 are 60, 6 and 0.6
g/L, respectively.
EPA has classified BDCM in Group B2, probable human carcinogen,
based on sufficient evidence of carcinogenicity in animals and
inadequate evidence in humans. The International Agency for Research on
Cancer (IARC) has recently classified BDCM as a Group 2B carcinogen,
agent probably carcinogenic to humans (IARC, 1991).
Following EPA's three-category approach for establishing MCLGs,
BDCM is placed in Category I since there is sufficient evidence for
carcinogenicity via ingestion considering weight of evidence, potency,
pharmacokinetics, and exposure. Thus, EPA is proposing an MCLG of zero
for this contaminant. EPA requests comments on the basis of the
proposed MCLG for BDCM and the use of tumor data of large intestine and
kidney, but not liver, in quantitative estimation of carcinogenic risk
of BDCM from oral exposure.
7. Dibromochloromethane
Dibromochloromethane (DBCM; CAS No. 124-48-1) is a nonflammable,
colorless liquid with a relatively high vapor pressure (76 mmHg at 20
deg.C). DBCM is moderately soluble in water (4 gm/l at 20 deg.C) and
soluble in organic solvents (log octanol/water partition coefficient of
2.09). Currently DBCM is not produced commercially in the United
States. The chemical has only limited uses as a laboratory agent. The
principal source of DBCM in drinking water is the chemical interaction
of chlorine with commonly present organic matter and bromide ions.
Degradation of DBCM has not been well studied, but probably involves
photooxidation. The estimated atmospheric half-life of DBCM is one to
two months. Volatilization is the principle mechanism for removal of
DBCM from rivers and streams (half-life of hours to weeks). Several
studies reported that DBCM adsorbs poorly to soil and sediments. No
experimental study was found regarding the bioconcentration of DBCM.
Based on the data of a few structurally similar chemicals, the
bioconcentration potential of DBCM in aquatic organisms is assumed to
be low. Biodegradation of DBCM is limited under aerobic conditions and
more extensive under anaerobic conditions.
Occurrence and Human Exposure. DBCM occurs in public water systems
that chlorinate water containing humic and fulvic acids and bromine
that can enter source waters through natural and anthropogenic means.
Several water quality factors can affect the formation of DBCM,
including Total Organic Carbon (TOC), pH, bromide, and temperature.
Different treatment practices can reduce the formation of DBCM in
drinking water. These include the use of precursor removal technologies
such as coagulation/filtration, granular activated carbon (GAC),
membrane filtration, and the use of chlorine dioxide, chloramination,
and ozonation.
Table V-8 presents occurrence information available for DBCM in
drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' (USEPA, 1992a). The
table lists six surveys conducted by Federal and private agencies.
Median concentrations of DBCM in drinking water appear to range from
0.6 to 3.6 g/L for surface water supplies and <0.5 g/
L for ground-water supplies. The lower bound median concentration for
DBCM in surface water supplies is biased to the low side because
concentrations in this survey were measured in the plant effluent; the
formation of DBCM would be expected to increase in the distribution in
systems using chlorine as their residual disinfectant.
Table V-8.--Summary of Occurrence Data for Dibromochloromethane
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of dibromochloromethane in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information(No. of --------------------------------------------------------------------------
samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
CWSS (1978) Brass et al., 450 Systems.......... Finished Water (1,100): Positive Detections:
1981. Surface Water............. ............. \3\5.0 1.5 67%\4\
Ground Water.............. ............. \3\6.6 <0.5 34%\4\
RWS (1978-1980) Brass, >600 Rural Systems Drinking Water from: Positive Detections:
1981. (>2,000 households). Surface Water............. ............. \3\8.5 0.8 \4\56%
Ground Water.............. ............. \3\9.9 <0.5 \4\13%
GWSS (1980-1981) Westrick 945 GW Systems: (466 90th Percentile:
et al. 1983. Random and 479 Serving >10,000 (327)..... Max. 59...... ........... 0.7 9.2
Nonran. Serving <10,000 (618)..... Max. 63...... ........... 0 5.6
dom).................
EPA, 1991a2 (1984-1991)... Unregulated Sampled at the Plant ............. 3.0 1.7 ..............................
Contaminant Data (4,439).
Base--Treatment
Facilities from 19
States.
EPA, 1992b2 (1987-1989)... Disinfection By- Finished Water: Positive Detections:
Products Field At the Plant (73)......... <0.2-41...... 4.9 2.0 92%
Studies. In Distribution System <0.2-41...... 6.6 3.4 93%
(56).
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell Max. 63...... ........... 3.6 75% of Data was Below 9.1
1989) Krasner et al., Nationwide. Effluent for 4 Quarters. 2.6-4.5\5\ g/L
1989b.
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Mean of the positives.
\4\Of systems sampled.
\5\Range of medians for individual quarters.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
CWSS: Community Water Supply Survey.
GWSS: Ground Water Supply Survey.
RWS: Rural Water Survey.
EPA: Environmental Protection Agency.
No information is available concerning the occurrence of DBCM in
food in the United States. The Food and Drug Administration (FDA) does
not analyze for DBCM in foods. However, there are several uses of
chlorine in food production, such as disinfection of chicken in poultry
plants and the superchlorination of water at soda and beer bottling
plants (Borum, 1991). Therefore, the possibility exists for dietary
exposure from the by-products of chlorination in food products.
Based on information obtained through a literature review, Howard
(1990) estimated the average daily intake of DBCM from air using an
inhalation rate of 20m\3\/day. Assuming a range of 8.25 to 425 ng/m\3\
the exposure may be as low as 0.17 g/day or as high as
8.5g/day.
Although some uses of chlorine have been identified in the food
production/food processing area, monitoring data are not available to
adequately characterize the magnitude or frequency of potential
exposure to DBCM. Additionally, preliminary discussions with FDA
suggest that there are not approved uses for chlorine in most foods
consumed in the typical diet. Based on the limited number of food
groups that are believed to contain chlorinated chemicals and that
there are not significant levels expected in ambient or indoor air, EPA
assumes that drinking water is the predominant source of DBCM intake.
Characterization of food and air exposure are issues currently under
review. EPA, therefore, is proposing to regulate DBCM in drinking water
with an RSC value at the ceiling level of 80 percent. EPA requests any
additional data on known concentration of DBCM in drinking water, food,
and air.
Health Effects. The health effects information in this section is
summarized from the Drinking Water Health Criteria Document for
Trihalomethanes (USEPA, 1994d). Studies mentioned in this section are
summarized in the criteria document.
Studies indicated that gastrointestinal absorption of DBCM is high
in animals. No studies were located regarding DBCM in humans or animals
following inhalation or dermal exposure. Based on the physical-chemical
properties of DBCM, and by analogy with the structurally-related
halomethanes such as chloroform, it is expected that the inhalation and
dermal absorption could be significant for DBCM.
The LD50 values in mice and rats range from 800 to 1,200 mg/
kg. Under both in vivo and in vitro conditions, several active
metabolic intermediates (e.g. dihalocarbonyl, bromochloromethyl
radicals) can be produced via oxidation or reduction by microsomal
preparations. Environmental studies suggest that these active metabolic
intermediates are responsible for the hepatic and renal toxicity, and
possibly carcinogenicity, of the parent compound. Animal studies
suggest that the extent of DBCM metabolism varies with species and sex.
The retention of DBCM in organs after dosing was small and relatively
higher concentrations were found in stomach, liver and kidneys. Urinary
excretion levels were below 2 percent.
Mammalian bioeffects following oral exposure to DBCM include
effects on the central nervous system (decreased operant response),
hepatotoxicity, nephrotoxicity, reproductive toxicity and possible
carcinogenicity. In experimental mice and rats, DBCM caused changes in
kidney, liver, and serum enzyme levels, and decreased body weight.
These responses are discernible in mammals from exposure to levels of
DBCM ranging from 39 to 250 mg/kg; the intensity of response was
dependent upon the dose and the duration of the exposure. Ataxia and
sedation were observed in mice receiving a single dose of 500 mg/kg
DBCM.
Developmental and reproductive toxicity of DBCM was investigated in
rodents. A multi-generation reproductive study of mice treated with )in
drinking water showed maternal toxicity (weight loss, liver
pathological changes) and fetal toxicity (decreased pup weight &
viability). The study identified a NOAEL of 17 mg/kg/day and a LOAEL of
171 mg/kg/day.
The National Toxicology Program (NTP, 1985) evaluated the
subchronic and chronic toxicity of DBCM in F344/N rats and B6C3F1
mice. In this study corn oil is used as the gavage vehicle. The chronic
data indicated that doses of 40 and 50 mg/kg/day produced
histopathological lesions in the liver of rats and mice, respectively.
However, the chronic studies did not identify a reliable NOAEL. The
subchronic study identified both a LOAEL and a NOAEL for
hepatotoxicity, and was used to calculate the RfD of 0.02 mg/kg/d.
In the NTP subchronic study, DBCM in corn oil was administered to
Fischer 344/N rats and B6C3F1 mice via gavage at dose levels of 0,
15, 30, 60, 125 or 250 mg/kg/day, 5 days a week for 13 weeks. Following
treatment, survival, body weight, clinical signs, histopathology and
gross pathology were evaluated. Final body weights of rats that
received 250 mg/kg/day were depressed 47% for males and 25% for
females. Kidney and liver toxicity was observed in male and female rats
and male mice at 250 mg/kg/day. A dose-dependent increase in hepatic
vacuolation was observed in male rats. Based on this hepatic effect,
the NOAEL and LOAEL in rats were 30 and 60 mg/kg/day, respectively.
Several studies on the mutagenicity potential of DBCM have reported
inconclusive results. Studies on the in vitro genotoxicity of DBCM
reported mixed results in bacteria Salmonella typhimurium strains and
yeasts. DBCM produced sister chromatid exchange uncultured human
lymphocytes and Chinese hamster ovary cells (without activation). An
increased frequency of sister chromatid exchange was observed in mouse
bone marrow cells from mice dosed orally, but not via the
intraperitoneal route.
The carcinogenicity of DBCM was reported in a NTP (1985) chronic
animal study. In this study DBCM in corn oil was administered via
gavage to groups of male and female F344/N rats at doses of 0, 40 or 80
mg/kg/day, 5 days/week for 104 weeks; and groups of male and female
mice at 0, 50 or 100 mg/kg/day, 5 days/week for 105 weeks.
Administration of DBCM showed a significant increase in the incidence
of hepatocellular adenomas in high-dose female mice (vehicle control,
2/50; low dose, 4/49; high dose, 11/50) and combined incidence of
hepatocellular adenomas or carcinomas (6/50, 10/49, 19/50). In high-
dose male mice, administration of DBCM showed a significant increase in
the incidence of hepatocellular carcinomas (10/50, -, 19/50); however,
the combined incidence of hepatocellular adenomas or carcinomas was
only marginally increased (23/50, -, 27/50). Administration of DBCM did
not result in increased incidence of tumors in treated rats.
Using the linearized multistage model, EPA derived a cancer potency
factor of 8.4 x 10-2 (mg/kg/day)-1 (IRIS, 1990). The
derivation was based on the tumor incidence of the hepatocellular
adenomas or carcinomas in the female mice reported in the 1985 NTP
study. Due to the possible role of the corn oil vehicle in induction of
hepatic tumors as reported in studies on chloroform, some uncertainty
exists regarding the relevance of this derived cancer potency factor to
exposure via drinking water. However, the only tumor data currently
available on DBCM are for liver tumors in mice. Until future studies
can provide additional data, EPA considers this cancer potency factor
valid for potential carcinogenic risk quantification for DBCM.
EPA has classified DBCM in Group C, possible human carcinogen,
based on the limited evidence of carcinogenicity in animals (only in
one species) and inadequate evidence of carcinogenicity in humans. The
International Agency for Research on Cancer (IARC) has classified DBCM
as a Group 3 carcinogen: agent not classifiable as to its
carcinogenicity to humans.
Using EPA's three-category approach for establishing MCLG, DBCM is
placed in Category II since there is limited evidence for
carcinogenicity via drinking water considering weight of evidence,
potency, pharmacokinetics, and exposure. As a Category II chemical, EPA
proposes to follow the first option and set the MCLG for DBCM on
noncarcinogenic endpoints (the RfD) with the application of an
additional safety factor to account for possible carcinogenicity. An
RfD of 0.02 mg/kg/day has been derived from the NOAEL of 30 mg/kg/d,
adjusted for dosing 5 days per week and divided by an uncertainty
factor of 1,000. This factor is appropriate for use of a NOAEL derived
from a subchronic animal study. EPA is proposing an MCLG of 0.06 mg/L
for DBCM based on liver toxicity and possible carcinogenicity. An
additional safety factor of 10 for possible carcinogenicity is used to
calculate the MCLG along with an assumed drinking water contribution of
80 percent of total exposure.
TP29JY94.006
EPA requests comments on the basis for the proposed MCLG for DBCM,
the RSC of 80%, and the cancer classification for DBCM.
8. Bromoform
Bromoform (tribromomethane, CAS No. 75-25-2) is a nonflammable,
colorless liquid with a sweet odor and a relatively high vapor pressure
(5.6 mmHg at 25 deg.C). Bromoform is moderately soluble in water (3.2
gm/L at 30 deg.C) and soluble in organic solvents (log octanol/water
partition coefficient of 2.38). Bromoform is not currently produced
commercially in the United States. The chemical has only limited uses
as a laboratory agent and as a fluid for mineral ore separation. The
principle source of bromoform in drinking water is the chemical
interaction of chlorine with commonly present organic matter and
bromide ion. Degradation of bromoform is not well studied, but probably
involves photooxidation. The estimated atmospheric half-life of
bromoform is one to two months. Volatilization is the principle
mechanism for removal of bromoform from rivers and streams (half-life
of hours to weeks). Studies reported that bromoform adsorbs poorly to
sediments and soils. No experimental studies were located regarding the
bioconcentration of bromoform. Based on the data from a few
structurally similar chemicals, the potential for bromoform to be
bioconcentrated by aquatic organisms is low. Biodegradation of
bromoform is limited under aerobic conditions but more extensive under
anaerobic conditions.
Occurrence and Human Exposure. Bromoform occurs in public water
systems that chlorinate water containing humic and fulvic acids and
bromine that can enter source waters through natural and anthropogenic
means. Several water quality factors affect the formation of bromoform
including Total Organic Carbon (TOC), pH, and temperature. Different
treatment practices can reduce the level of bromoform. These include
the use of chloride dioxide, chloramination, and ozonation prior to
chloramination.
Table V-9 presents occurrence information available for bromoform
in drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' (USEPA, 1992a). The
table lists six surveys conducted by Federal and private agencies.
Median concentrations of bromoform in drinking water appear to range
from <0.2 to 0.57 g/L for surface water supplies and <0.5
g/L for ground- water supplies. The lower bound median
concentration for bromoform in surface water supplies is biased to the
low side because concentrations in this survey were measured in the
plant effluent; the formation of bromoform would be expected to
increase in the distribution in systems using chlorine as their
residual disinfectant.
Table V-9.--Summary of Occurrence Data for Bromoform
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of bromoform in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information (No. of --------------------------------------------------------------------------
samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
CWSS (1978) Brass et al., 450 Systems.......... Finished Water (1,100): ............. Positive Detections:
1981. Surface Water............. \3\2.1 <1.0 13%\4\
Ground Water.............. \3\11 <0.5 26%\4\
RWS (1978-1980) Brass, >600 Rural Systems Drinking Water from: ............. Positive Detections:
1981. (>2,000 Households). Surface Water............. \3\8.7 <0.5 18%\4\
Ground Water.............. \3\12 <0.5 12%\4\
GWSS (1980-1981) Westrick 945 GW Systems: (466 .......................... ............. ........... ............. 90th Percentile:
et al. 1983. Random and 479 Serving >10,000 (327)..... Max. 68...... 0 8.3
Nonran-dom). Serving <10,000 (618)..... Max. 110..... 0 4.1
EPA, 1991a\2\ (1984-1991). Unregulated Sampled at the Plants ............. 2.5 1 ..............................
Contaminant Data (1,409).
Base--Treatment
Facilities from 19
States.
EPA, 1992b\2\ (1987-1989). Disinfection By- Finished Water: Positive Detections:
Products Field At the Plant (73)......... <0.2-6.7..... 0.7 <0.2 45%
Studies. In Distr. System (56)..... <0.2-10...... 1.0 <0.2 48%
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell Max. 72...... ........... ............. 75% of Data was Below 2.8
1989) Krasner et. al., Nationwide. Effluent for 4 Quarters. 0.33-0.88\5\
1989b. 0.57
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Mean of the positives.
\4\Of systems sampled.
\5\Range of medians for individual quarters.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
CWSS: Community Water Supply Survey.
GWSS: Ground Water Supply Survey.
RWS: Rural Water Survey.
EPA: Environmental Protection Agency.
No information is available concerning the occurrence of bromoform
in food in the United States. The Food and Drug Administration (FDA)
does not analyze for bromoform in foods. However, there are several
uses of chlorine in food production, such as disinfection of chicken in
poultry plants and the superchlorination of water at soda and beer
bottling plants (Borum, 1991). Therefore, the possibility exists for
dietary exposure from the by-products of chlorination in food products.
Bromoform is usually found in ambient air at low concentrations.
One study reported ambient air concentrations from several urban
locations across the U.S. The overall mean concentration of positive
samples was found to be 4.15 ng/m\3\ and the maximum level was 71 ng/
m\3\ (Brodzinsky and Singh, 1983 in USEPA, 1991b). Although the data
are limited for bromoform, an inhalation intake could be estimated
using the mean and maximum values from the Brodzinsky and Singh (1983)
study, indicating a possible range of 0.08 to 1.4 g/d.
Based on the limited number of food groups that are believed to
contain bromoform and that significant levels are not expected in
ambient or indoor air, EPA is assuming that drinking water is the
predominant source of bromoform intake. Characterization of food and
air exposures are issues currently under review. The EPA requests any
additional data on known concentrations of bromoform in drinking water,
food, and air.
Health Effects. The health effects information in this section is
summarized from the draft Drinking Water Health Criteria Document for
Trihalomethanes (USEPA, 1994d) and the draft Drinking Water Health
Advisory for Brominated Trihalomethanes (USEPA, 1991b). Studies
mentioned in this section are summarized in the criteria document or
health advisory.
Studies have indicated that gastrointestinal absorption of
bromoform is high in humans and animals. No studies were located
regarding bromoform in humans or animals following inhalation or dermal
exposure. Based on the physical-chemical properties of bromoform, and
by analogy with the structurally-related halomethanes such as
chloroform, it is expected that both inhalation and dermal absorption
could be significant for bromoform.
Bromoform was used as a sedative for children with whooping cough.
Based on clinical observations of accidental overdose cases, the
estimated lethal dose for a 10- to 20-kg child is about 300 mg/kg. The
clinical signs in fatal cases were central nervous system (CNS)
depression followed by respiratory failure.
The LD50 values in mice and rats have been reported in the
range of 1,147-1550 mg/kg. Under both in vivo and in vitro conditions,
several active metabolic intermediates (e.g., dibromocarbonyl,
dibromomethyl radicals) are produced via oxidation or reduction by
microsomal preparations. Experimental studies suggested that these
active metabolic intermediates are responsible for hepatic and renal
toxicity and possibly, carcinogenicity, of the parent compound. Animal
studies suggest that the extent of bromoform metabolism varies with
species and sex. The retention of bromoform in organs after dosing was
small; relatively higher concentrations were found in tissues with
higher lipophilic content. Urinary excretion levels were below 5
percent.
Mammalian bioeffects following exposure to bromoform include
effects on the central nervous system (CNS), hepatotoxicity,
nephrotoxicity, and carcinogenicity. Bromoform causes CNS depression in
humans. The reported LOAEL which results in mild sedation in humans is
54 mg/kg. In experimental mice and rats, bromoform caused changes in
kidney, liver, serum enzyme levels, decrease of body weight, and
decreased operant response. These responses are discernible in mammals
from exposure to levels of bromoform ranging from 50 to 250 mg/kg; the
intensity of response was dependent upon the dose and the duration of
the exposure. Ataxia and sedation were noted in mice receiving a single
dose of 1,000 mg/kg bromoform or 600 mg/kg for 14 days.
Few studies have investigated developmental and reproductive
toxicity of bromoform in rodents. A developmental study in rats showed
no fetal variations in a group fed with 50 mg/kg/day. An increased
incidence of minor anomalies was noted at doses of 100 and 200 mg/kg/
day. No maternal toxicity in rats was observed. One detailed
reproductive toxicity study reported no apparent effects on fertility
and reproduction when male and female rats were administered bromoform
via gavage in corn oil at doses up to 200 mg/kg/day.
EPA used subchronic data from an oral study (NTP, 1989) to
calculate the RfD. In this study, bromoform was administered to rats in
corn oil via gavage at dose levels of 0, 12, 25, 50, 100 or 200 mg/kg/
day 5 days a week for 13 weeks. Based on the observation of
hepatocellular vacuolization in treated male rats, a NOAEL of 25 mg/kg/
day was established. An RfD of 0.02 mg/kg/day has been derived from
this NOAEL by the application of an uncertainty factor of 1,000, in
accordance with EPA guidelines for use of a NOAEL from a subchronic
study.
A number of studies investigated the mutagenicity potential of
bromoform. Studies on the in vitro genotoxicity of bromoform reported
mixed results in bacterial Salmonella typhimurium strains. Bromoform
produced mutations in cultured mouse lymphoma cells and sister
chromatid exchange in human lymphocytes. Under in vivo condition
bromoform induced sister chromatid exchange, and chromosomal aberration
and micronucleus in mouse bone marrow cells. Overall, most studies
yielded positive results and evidence of mutagenicity for bromoform is
considered adequate.
The National Toxicology Program (NTP, 1989) conducted a chronic
animal study to investigate the carcinogenicity of bromoform. In this
study bromoform was administered in corn oil via gavage to F344/N rats
(50/sex/group) at doses of 0, 100 or 200 mg/kg/day, 5 days/week for 105
weeks. An evaluation of the study results showed that adenomatous
polyps or adenocarcinoma (combined) of the large intestine (colon or
rectum) were induced in three male rats (vehicle control, 0/50; low
dose, 0/50; high dose, 3/50) and in nine female rats (0/50, 1/50, 8/
50). The increase was considered to be significant since these tumors
are rare in control animals. Neoplastic lesions in the large intestine
in female rats reported in the NTP study were used to estimate the
carcinogenic potency of bromoform. EPA derived a cancer potency factor
of 7.9 x 10-3 (mg/kg/day)-1 using the linearized multistage
model (IRIS, 1990). Assuming a daily consumption of two liters of
drinking water and an average human body weight of 70 kg, the 95% upper
bound limit lifetime cancer risks of 10-6, 10-5 and 10-4
are associated with concentrations of bromoform in drinking water of 4,
40 and 400 g/L, respectively.
EPA classified bromoform in Group B2, probable human carcinogen,
based on the sufficient evidence of carcinogenicity in animals and
inadequate evidence of carcinogenicity in humans. The International
Agency for Research on Cancer (IARC) has recently classified bromoform
in Group 3: agent not classifiable as to its carcinogenicity to humans
(IARC, 1991). IARC determined that there was limited evidence of
carcinogenicity in animals, in contrast to EPA's judgment that there is
sufficient evidence in laboratory animals. EPA requests comments on the
different viewpoints between IARC and EPA regarding bromoform's
carcinogenic potential.
Using EPA's three-category approach for establishing MCLG,
bromoform is placed in Category I since there is sufficient evidence
for carcinogenicity from drinking water considering weight of evidence,
potency, pharmacokinetics, and exposure. Thus, EPA is proposing an MCLG
of zero for this contaminant. EPA requests comments on the basis for
the proposed MCLG for bromoform.
9. Dichloroacetic Acid
Chlorination of water containing organic material (humic, fulvic
acids) results in the generation of many organic compounds, including
dichloroacetic acid (DCA) (CAS. No. 79-43-6), a nonvolatile compound.
Though DCA is generally a concern due to its occurrence in
chlorinated drinking water, it is also used as a chemical intermediate,
and an ingredient in pharmaceuticals and medicine. Previously, DCA was
used experimentally to treat diabetes and hypercholesterolemia in human
patients. In addition, DCA was used as an agricultural fungicide and
topical astringent. It has also been extensively investigated for
potential therapeutic use as a hypoglycemic, hypolactemic and
hypolipidemic agent.
Occurrence and Human Exposure. DCA has been found to occur as a
disinfection by-product in public water systems that chlorinate water
containing humic and fulvic acids.
Table V-10 presents occurrence information available for DCA in
drinking water. Descriptions of these surveys and other data are
detailed in ``Occurrence Assessment for Disinfectants and Disinfection
By-Products (Phase 6a) in Public Drinking Water,'' (USEPA, 1992a).
Median concentrations of DCA in drinking water were found to range from
6.4 to 17 g/L. The lower bound median concentration for DCA in
surface water supplies is biased to the low side because concentrations
in this survey were measured in the plant effluent; the formation of
DCA would be expected to increase in the distribution system in systems
using chlorine as their residual disinfectant.
Table V-10.--Summary of Occurrence Data for Dichloroacetic Acid
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurence of dichloroacetic acid in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information (No. of -------------------------------------------------------------------
samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
EPA, 1992b\2\ (1987-1989) Disinfection By-Products Finished Water: ............. 18 16 Positive Detections:
Field Studies. At the Plant (72) <0.4-61 21 17 93%
In the Distr. System (56) <0.4-75 96%
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell <0.6-46 ........... \3\5.0-7.3 75% of Data was Below 12
1989) Krasner et al., Nationwide. Effluent for 4 Quarters. 6.4 g/L DL = 0.6
1989b. g/L
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Range of medians for individual quarters.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
EPA: Environmental Protection Agency.
Based on the above data, a range of exposure to DCA from drinking
water can be calculated using a consumption rate of 2 liters per day.
The expected median exposure from drinking water would range from 13 to
34 g/day, using these data sets.
No information is available concerning the occurrence of DCA in
food and ambient or indoor air in the United States. The Food and Drug
Administration (FDA) does not analyze for DCA in foods. However, there
are several uses of chlorine in food production, such as the
disinfection of chicken in poultry plants and the superchlorination of
water at soda and beer bottling plants. Therefore, the possibility
exists for dietary exposure from the by-products of chlorination in
food products. However, monitoring data are not available to
characterize adequately the magnitude or frequency of potential DCA
exposure from diet. Additionally, preliminary discussions with FDA
suggest that there are not approved uses for chlorine in most foods
consumed in the typical diet. Similarly, EPA's Office of Air and
Radiation is not currently sampling for DCA in air (Borum, 1991).
Little exposure to DCA from air is expected since DCA is nonvolatile.
Since only a limited number of food groups are expected to contain
chlorinated chemicals and no significant DCA levels are expected in
ambient or indoor air, EPA believes that drinking water is the
predominant source of DCA intake. Characterization of the potential
exposures from food and air are issues currently under review. EPA
requests any additional data on known concentrations of DCA in drinking
water, food, and air.
Health Effects. The health effects information in this section is
summarized from the draft Drinking Water Health Criteria Document for
Chlorinated Acetic Acids, Alcohols, Aldehydes and Ketones (USEPA,
1994e). Studies mentioned in this section are summarized in the
criteria document.
Humans treated with DCA for 6 to 7 days at 43 to 57 mg/kg/day have
experienced mild sedation, reduced blood glucose, reduced plasma
lactate, reduced plasma cholesterol levels, and reduced triglyceride
levels. At the same time, the DCA treatment depressed uric acid
excretion, resulting in elevated serum uric acid levels.
A longer term study in two young men receiving 50 mg/kg for 5 weeks
up to 16 weeks, indicated that DCA significantly reduces serum
cholesterol levels and blood glucose, and causes peripheral neuropathy
in the facial, finger, leg and foot muscles.
Estimates of acute oral LD50 values range from 2,800 to 4,500
mg/kg in rats and up to 5,500 mg/kg in mice. Short-term studies in dogs
and rats indicate an effect on intermediary metabolism, as demonstrated
by decreases in blood lactate and pyruvate. Exposures to DCA up to 3
months in dogs and rats result in a variety of adverse effects
including effects to the neurological and reproductive systems. These
effects are seen above 100 mg/kg/day in dogs and rats.
Studies on the toxicokinetics of DCA indicate that absorption is
rapid and that DCA is quickly distributed to the liver and muscles in
the rat. DCA is metabolized to glyoxylate which in turn is metabolized
to oxalate. Although there are few studies regarding the excretion of
DCA, studies in which rats, dogs and humans received intravenous
injections of DCA indicated that the half-life of DCA in human blood
plasma is much shorter than in rats or dogs. Urinary excretion of DCA
was negligible after 8 hours. Total excretion of DCA was less than 1%
of total dose.
A drinking water study by Bull et al. (1990) reported a dose-
related increase in hepatic effects in mice that received DCA at 270
mg/kg/day for 37 weeks and at 300 mg/kg/day for 52 weeks. Adverse
effects included enlarged livers, marked cytomegaly with massive
accumulation of glycogen in hepatocyte and focal necrosis. The NOAEL
for this study was 137 mg/kg/day for 52 weeks.
DeAngelo et al. (1991) conducted a drinking water study in which
mice received DCA at levels of 7.6, 77, 410, and 486 mg/kg/day for 60
or 75 weeks. While this study was intended as an assessment of
carcinogenicity, other systemic effects were measured. This study
concluded that levels at 77 mg/kg/day and above caused an extreme
increase of relative liver weights and a significant increase in
neoplasia at levels of 410 mg/kg/day and above. This study indicates a
NOAEL of 7.6 mg/kg/day for noncancer liver effects.
Based on the available data, DCA does not appear to be a potent
mutagen. Studies in bacteria have indicated that DCA did not induce
mutation or activate repair activity. Two studies have shown some
potential for mutagenicity but these results have not been
reproducible.
DCA appears to induce both reproductive and developmental toxicity.
Damage and atrophy to sexual organs has been reported in male rats and
dogs exposed to levels from 50 mg/kg/day to 2000 mg/kg/day for up 3
months. Malformation of the cardiovascular system has been observed in
rats exposed to 140 mg/kg/day DCA from day 6 to 16 of pregnancy.
A 90-day dog study was selected to determine the RfD for DCA
(Cicmanec et al., 1991). In this study, four month old beagle dogs (5/
sex/group) were administered gelatin capsules containing 0, 12.5, 39.5,
or 72 mg/kg DCA/day for 90 days. Dogs were observed for clinical signs
of toxicity; blood samples were collected for hematology and serum
chemistry analysis. Clinical signs included diarrhea and dyspnea in the
mid and high dose groups. Dyspnea was evident at 45 days and became
more severe with continued exposure leading to general depression and
decreased activity by day 90. Hindlimb paralysis was observed in 3 dogs
in the high dose group. Other effects included conjunctivitis, weight
loss, reduced food and water consumption, pneumonia, decreased liver
weights, and elevated kidney weights in the dosed animals.
Histopathology revealed toxic effects in liver, testis, and brain of
the treated dogs. A NOAEL was not identified in this study. The lowest
dose tested, 12.5 mg/kg/d, was considered a LOAEL. An uncertainty
factor of 3,000 was applied in accordance with EPA guidelines to
account for use of a LOAEL from a less-than-lifetime animal study in
which frank effects were noted as the critical effect. The resulting
RfD is 0.004 mg/kg/d.
Several studies indicate that DCA is a carcinogen in both mice and
rats exposed via drinking water lifetime studies. These studies
indicate that DCA induces liver tumors. In one study with male
B6F3F1 mice, exposure to DCA at 0.5 g/L and 3.5 g/L for 104 weeks
resulted in tumor formation in exposed animals at 75% (18/24) and 100%
(24/24) respectively. In female mice exposed for 104 weeks to DCA at
the same levels, tumor prevalence was 20% and 100%, respectively. In
male rats exposed to 0.05, 0.5 or 5 g/L DCA for 104 weeks, tumor
prevalence increased to 22% in the highest dose. No tumors were seen at
the lower doses. However, at 0.5 g/L, there was an increase in the
prevalence of proliferation of liver lesions. Some of these lesions are
likely to progress into malignant tumors.
EPA has classified DCA in Group B2: probable human carcinogen,
based on positive carcinogenic findings in two animal species exposed
to DCA in drinking water. A quantitative risk estimate has not yet been
determined for DCA.
Following a Category I approach, EPA is proposing an MCLG for DCA
of zero based on the strong evidence of carcinogenicity via drinking
water. EPA requests comments on the basis for the proposed MCLG for DCA
in drinking water and the cancer classification of Group B2.
10. Trichloroacetic Acid.
Trichloroacetic acid (TCA; CAS No. 76-03-9) is also a major by-
product of chlorinated drinking water. Chlorination of source waters
containing organic materials (humic, fulvic acids) results in the
generation of organic compounds such as TCA.
TCA is also sold as a pre-emergence herbicide. It is used in the
laboratory to precipitate proteins and as a reagent for synthetic
medicinal products. It is applied medically as a peeling agent for
damaged skin, cervical dysplasia and removal of tatoos.
Occurrence and Human Exposure. TCA occurs in public water systems
that chlorinate water containing humic and fulvic acids.
Table V-11 presents the most recent and comprehensive occurrence
information available for TCA in drinking water. Descriptions of these
surveys and other data are detailed in ``Occurrence Assessment for
Disinfectants and Disinfection By-Products (Phase 6a) in Public
Drinking Water,'' (USEPA, 1992a). Median concentrations of TCA acid in
drinking water were found to range from 5.5 to 15 g/L. The
lower bound median concentration for TCA in surface water supplies is
biased to the low side because concentrations in this survey were
measured in the plant effluent; the formation of TCA would be expected
to increase in the distribution system in systems using chlorine as
their residual disinfectant. Based on the available data sets, and
assuming a drinking water consumption rate of 2 L/day, median exposures
from drinking water would range from 11 to 30 g/day.
Table V-11.--Summary of Occurrence Data for Trichloroacetic Acid
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of trichloroacetic acid in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information (No. ------------------------------------------------------------------------
of samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
EPA, 1992b\2\ (1987-1989) Disinfection By-Products Finished Water: ............. ........... ........... Positive Detections:
Field Studies. At the Plant (72) <0.4-54 13 11 90%
Distribution System <0.4-77 15 15 91%
(56).....................
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell ............. ........... 4.0-5.8 75% of Data was Below 15.3
1989) Krasner et al., Nationwide. Effluent for 4 Quarters. 5.5 g/L DL = 0.6
1989b.
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Range of medians for individual quarters.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
EPA: Environmental Protection Agency.
No information is available concerning the occurrence of TCA in
food and ambient or indoor air in the United States. The Food and Drug
Administration (FDA) does not analyze for TCA in foods. However, there
are several uses of chlorine in food production, such as the
disinfection of chicken in poultry plants and the superchlorination of
water at soda and beer bottling plants. Therefore, the possibility
exists for dietary exposure from the by-products of chlorination in
food products. Also, TCA has limited use as a herbicide. However,
monitoring data are not available to characterize adequately the
magnitude or frequency of potential TCA exposure from diet. Similarly,
EPA's Office of Air and Radiation is not currently measuring for TCA in
air (Borum, 1991). The exposure from air for TCA is probably not a
large source since TCA is nonvolatile.
Since only a limited number of food groups are expected to contain
chlorinated chemicals and no significant TCA levels are expected in
ambient or indoor air, EPA assumes that drinking water is the
predominant source of TCA intake. Characterization of potential
exposures from food and air are issues currently under review. EPA is,
therefore, proposing to regulate TCA in drinking water with a relative
source contribution (RSC) value at the ceiling level of 80 percent. EPA
requests any additional data on known concentrations of TCA in drinking
water, food, and air.
Health Effects. The health effects information in this section is
summarized from the Drinking Water Health Criteria Document for
Chlorinated Acetic Acids, Alcohols, Aldehydes and Ketones (USEPA,
1994e). Studies mentioned in this section are summarized in the
criteria document.
Estimates of acute and LD50 values for TCA range from 3.3 to 5
g/kg in rats to 4.97 g/kg in mice. Short-term studies, up to 30 days,
in rats demonstrate few effects other than decreased weight gain after
administration of 240-312 mg/kg/day.
Few studies on toxicokinetics of TCA were located; however, a human
study and a dog study show TCA to respond pharmacokinetically similarly
to DCA. The response indicates a rapid absorption, distribution to the
liver and excretion primarily through the urine. The two studies
indicate that TCA is readily absorbed from all sections of the
intestine and that the urinary bladder may be significant in the
absorption of TCA. TCA is also a major metabolite of trichloroethylene.
Longer-term studies in animals indicate that TCA affects the liver,
kidney and spleen by altering weights, focal hepatocellular
enlargement, intracellular swelling, glycogen accumulation, focal
necrosis, an accumulation of lipofuscin, and ultimately tumor
generation in mice.
In a study by Mather et al. (1990), male rats received TCA in their
drinking water at 0, 4.1, 36.5 or 355 mg/kg/day. The high dose resulted
in spleen weight reduction and increased relative liver and kidney
weights. Hepatic peroxisomal -oxidation activity was
increased. Liver effects at the high dose included focal hepatocellular
enlargement, intracellular swelling and glycogen accumulation. The
NOAEL for this study was 36.5 mg/kg/day.
Parnell et al. (1988) exposed male rats to TCA in their drinking
water at 2.89, 29.6 or 277 mg/kg/day for up to one year. No significant
changes were detected in body weight, organ weight or histopathology
over the study duration. This study identified a NOAEL as the highest
dose tested, 277 mg/kg/day.
Bull et al. (1990) investigated the effects of TCA on liver lesions
and tumor induction in male and female B6C3F1 mice. Mice received
TCA in their drinking water at 0, 1 or 2 g/L (164 or 329 mg/kg/day) for
37 or 52 weeks. Dose-related increases in relative and absolute liver
weights were seen in females and males exposed to 1 and 2 g/L for 52
weeks. Small increases in liver cell size, accumulation of lipofuscin
and focal necrosis were also seen. A LOAEL of 164 mg/kg/day (1 g/L) was
identified.
Several studies show that TCA can produce developmental
malformations in fetal Long Evans rats, particularly in the
cardiovascular system. Teratogenic effects were observed at the lowest
dose tested, 330 mg/kg/day.
With regard to mutagenicity tests, TCA was negative in Ames
mutagenicity tests using Salmonella strain TA100, but was positive for
bone marrow chromosomal aberrations and sperm abnormalities in mice. It
also induced single-strand DNA breaks in rats and mice exposed by
gavage.
TCA has induced hepatocellular carcinomas in two tests with
B6C3F1 mice, one of 52 weeks and another of 104 weeks. In the Bull
et al. (1990) study, a dose-related increase in the incidence of
hepatoproliferative lesions was observed in male B6C3F1 mice
exposed to 1 or 2 g/L for 52 weeks. An increase in hepatocellular
carcinomas was observed in males at both dose levels. Carcinomas were
not found in females.
DeAngelo et al. (1991) administered mice and rats with TCA over
their lifetime. Male and female B6C3F1 mice were exposed to 4.5 g/
L TCA for 104 weeks. Male mice at 4.5 g/L TCA had a tumor prevalence of
86.7%. Female mice appeared to be less sensitive to TCA than males: 60%
prevalence over a 104-week exposure to 4.5 g/L. At 104 weeks, 0.5 g/L
TCA did not result in a significant increase in tumors. In a
preliminary study of 60 weeks exposure to 0.05, 0.5 and 5 g/L, no
significant additional increase in tumors was seen at 0.05 g/L, but
tumor prevalence was 37.9% and 55.2% at 0.5 and 5 g/L, respectively.
F344 male rats administered TCA over a lifetime at 0.05 to 5 g/L
did not produce a significant increase in carcinogenicity.
EPA has placed TCA in Group C: possible human carcinogen. Group C
is for those chemicals which show limited evidence of carcinogenicity
in animals in the absence of human data.
EPA is following a Category II approach for setting an MCLG for
TCA. The developmental toxicity study by Smith et al. (1989) has been
selected to serve as the basis for the RfD and MCLG. In this
developmental study, pregnant Long-Evans rats (20/dose) were
administered TCA at doses of 0, 330, 800, 1,200, or 1,800 mg/kg/d by
gavage during gestation days 6-15. Maternal body weight was
significantly reduced at doses of 800 mg/kg/d and above. Maternal
spleen and kidney weights were increased significantly in a dose-
dependent manner. Postimplantation loss was noted in the three highest
dose groups with a significant decrease in the number of live fetuses
per litter observed in the two highest dose groups. Other fetal effects
included decreased fetal weight and crown-rump length, and
malformations of the cardiovascular system, particularly the heart. The
lowest dose tested, 330 mg/kg/d, was identified as a LOAEL. A NOAEL was
not identified from this study.
An RfD of 0.1 mg/kg/day was derived using the LOAEL of 330 mg/kg/d
and an uncertainty factor of 3,000 to account for use of a LOAEL and
lack of a 2 generation reproductive study. Adjusting the RfD for a 70
kg adult drinking 2 L water per day, possible carcinogenicity and an
RSC of 80%, an MCLG of 0.3 mg/L can be determined.
TP29JY94.007
EPA requests comments on the basis for the MCLG and the cancer
classification for TCA.
11. Chloral Hydrate
Chlorination of water containing organic materials (humic, fulvic
acids) results in the generation of organic compounds such as
trichloroacetaldehyde monohydrate or chloral hydrate (CH) (CAS No. 302-
17-0).
CH is used as a hypnotic or sedative drug (i.e., knockout drops) in
humans, including neonates. CH is also used in the manufacture of DDT.
Occurrence and Human Exposure. CH has been found to occur as a
disinfection by-product in public water systems that chlorinate water
containing humic and fulvic acids.
Table V-12 presents occurrence information available for chloral
hydrate in drinking water. Descriptions of these surveys and other data
are detailed in ``Occurrence Assessment for Disinfectants and
Disinfection By-Products (Phase 6a) in Public Drinking Water,'' (USEPA,
1992a). Median concentrations of chloral hydrate in drinking water were
found to range from 2.1 to 4.4 g/L.
TABLE V-12.--Summary of Occurrence Data for Chloral Hydrate
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of chloral hydrate in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (Year)\1\ Location Sample information (No. ------------------------------------------------------------------------
of samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
EPA, 1992b\2\ (1987-1989) Disinfection By-Products Finished Water:.......... <0.2-25 5.0 2.5 Positive Detections:
Field Studies. At the Plant (67)........ <0.2-30 7.8 4.4 90%
Distribution System...... 91%
(53).....................
EPA/AMWA/CDHS\2\ (1988- 35 Water Utilities Samples from Clearwell Max. 22 ........... \3\1.7-3.0 75% of Data was below 4.1
1989) Krasner et al., Nationwide. Effluent for 4 Quarters. 2.1 g/L\4\
1989b.
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
\3\Range of medians for individual quarters.
\4\Detection limit was 0.02 g/L in the first quarter and 0.1 g/L thereafter.
AMWA: Association of Metropolitan Water Agencies.
CDHS: California Department of Health Services.
EPA: Environmental Protection Agency.
Based on the available data sets, median exposures from CH due to
drinking water would range from 3.4 to 8.8 g/day, based on the
consumption of 2 liters per day.
No information is available concerning the occurrence of CH in food
and ambient or indoor air in the United States. The Food and Drug
Administration (FDA) does not analyze for CH in foods since the
analytical methods for such an evaluation have not been developed
(Borum, 1991).
CH has been used as a sedative of hypnotic drug (see Health Effects
Section). There are several uses of chlorine in food production, such
as the disinfection of chicken in poultry plants and the
superchlorination of water at soda and beer bottling plants. Therefore,
the possibility exists for dietary exposure from the by-products of
chlorination in food products. However, monitoring data are not
available to adequately characterize the magnitude or frequency of
potential CH exposure from the diet. Similarly, EPA's Office of Air and
Radiation is not currently measuring for CH in air (Borum, 1991).
However, CH from indoor air may contribute to exposure due to the
volatilization from tap water.
Since only a limited number of food groups are expected to contain
chlorinated chemicals and no significant levels are expected in ambient
or indoor air, EPA believes that drinking water is the predominant
source of CH intake. Characterization of potential food and air
exposures are issues currently under review. EPA is therefore,
proposing to regulate CH in drinking water with an RSC value at the
ceiling level of 80 percent. EPA requests any additional data on known
concentrations of CH in drinking water, food, and air.
Health Effects. The health effects information in this section is
summarized from the draft Drinking Water Health Criteria Document for
Chlorinated Acetic Acids, Alcohols, Aldehydes and Ketones (USEPA,
1994e). Studies mentioned in this section are summarized in the
criteria document.
In its use as a sedative or hypnotic drug in humans, a history of
adverse effects related to CH exposure have been recorded. The acute
and toxic dose to humans is about 10 g (or 140 mg/kg), causing severe
respiratory depression and hypertension. Adverse reactions such as
central nervous system depression and gastrointestinal disturbances are
seen between 0.5 and 1.0 g CH. Cardiac arrhythmias are seen when
patients receive levels between 10 and 20 g (167-333 mg/kg). Chronic
use of CH may result in development of tolerance, physical dependence,
and addiction.
Estimates of acute oral LD50s in mice range from 1,265 to
1,400 mg/kg with central nervous system depression and inhibition of
respiration being the cause of death. Rats may be more sensitive than
mice with acute oral LD50 values ranging from 285 mg/kg in newborn
to 500 mg/kg in adults.
Short-term studies in mice indicate that the liver is the target of
CH toxicity with changes in liver weight as the primary effect. NOAELs
vary between 14 and 144 mg/kg/day.
Toxicokinetic studies of CH indicate that absorption is rapid and
complete in dogs and humans. CH is metabolized to trichloroacetic acid
(TCA) and trichloroethanol. CH is rapidly excreted primarily through
the urine as trichloroethanol glucuronide and more slowly as TCA.
Three 90-day studies in mice were considered by EPA to derive the
MCLG for CH. Each used the same dose levels (16 or 160 mg/kg/day) in
mice. The first study (Kallman et al., 1984) exposed groups of 12 male
mice to drinking water containing CH at concentrations of 70 and 700
mg/L for 90 days. These concentrations correspond to doses of 15.7 and
160 mg/kg/day. No treatment-related effects were observed for
mortality, body weight, physical appearance, behavior, locomotor
activity, learning in repetitive tests of coordination, response to
painful stimuli, strength, endurance or passive avoidance learning.
Both doses resulted in a decrease of about 1 deg. in mean body
temperature (p <0.05). The biological significance of this hypothermic
effect is uncertain.
In the second study, Sanders et al. (1982) supplied groups of 32
male and female CD-1 mice with CH in deionized drinking water (70 or
700 mg/L, corresponding to time-weighted average doses of approximately
16 mg/kg/day or 160 mg/kg/day, respectively). After 90 days, the liver
appeared to be the tissue most affected. Males appeared to be more
sensitive than females. In males, there was a dose-related hepatomegaly
and microsome proliferation, accompanied by small changes in serum
chemistry values for potassium, cholesterol, and glutathione. Females
did not show hepatomegaly, but did display changed hepatic microsomal
parameters. Based on hepatomegaly, this study identifies a LOAEL of 16
mg/kg/day for CH (the lowest dose tested).
In the third study, Kauffman et al. (1982) studied the effect of CH
on the immune system. Groups of 13 to 18 male and female CD-1 mice were
supplied with water containing 70 or 700 mg/L (corresponding to time-
weighted average doses of approximately 16 or 160 mg/kg/day,
respectively) for 90 days. In males, no effects were detected in either
humoral or cell-mediated immunity at either dose level. In females,
exposure to the high dose (160 mg/kg/day) resulted in decreased humoral
immune function (p <0.05), but no effects on cell-mediated immunity
were noted. Based on this study, a NOAEL of 16 mg/kg/day and a LOAEL of
160 mg/kg/day were identified.
CH is weakly mutagenic in Salmonella, yeast and molds. It has also
caused chromosomal aberration in yeast and nondisjunction of
chromosomes during spermatogenesis.
One study has observed neurobehavioral effects on mice pups from
female mice receiving CH at 205 mg/kg/day for three weeks prior to
breeding. Exposure of females continued until pups were weaned at 21
days of age. Pups from the high dose group (205 mg/kg/day) showed
impaired retention in passive avoidance learning tasks. This result can
be construed as a developmental effect of CH.
Two studies on the carcinogenicity of CH indicate that CH produces
mouse liver tumors. In the earlier study, Rijhsinghani et al. (1986),
B6C3F1 mice given a single oral dose of CH at 5 or 10 mg/kg
developed a significant increase in liver tumors after 92 weeks.
In a later study, Daniel et al. (1992), reported that male mice,
receiving 166 mg/kg/day CH for 104 weeks, showed a total liver tumor
prevalence of 71 percent (17/24). Proliferative liver lesions
recognized and tabulated in this study included hyperplastic nodules,
hepatocellular adenomas and hepatocellular carcinomas. No other studies
were located on the carcinogenicity of CH in other test species.
Based on the limited evidence of carcinogenicity in these two
studies and the extensive mutagenicity of CH, EPA has classified CH in
Group C: possible human carcinogen. The concentrations associated with
a 10-4, 10-5, and 10-6 excess cancer risk are 40
g/L, 4 g/L and 0.4 g/L, respectively.
EPA is placing CH in Category II for setting an MCLG based on liver
toxicity and limited evidence of carcinogenicity from drinking water.
EPA believes the 90-day study by Sanders et al. (1982) is most
appropriate to calculate the RfD and MCLG for CH because the liver
effects observed in this study (i.e., change to hepatic microsomal
parameters and hepatomegaly) appear to be more severe than the other
studies have indicated at similar dose levels. From the mouse LOAEL of
16 mg/kg/day and an uncertainty factor of 10,000 for use of a LOAEL
from a less than lifetime animal study, an MCLG of 0.04 mg/L is
derived.
TP29JY94.008
EPA is proposing to use an extra safety factor of 1 instead of 10
to account for possible carcinogenicity since an uncertainty factor of
10,000 has already been applied to the RfD. In addition, the proposed
MCLG equals the 10-4 excess cancer risk. EPA requests comment on
the Category II approach for setting an MCLG, the extra safety factor
of 1 instead of 10 for a Category II contaminant, and whether the
endpoint of liver weight increase and hepatomegaly is a LOAEL or NOAEL
given the lack of histopathology.
12. Bromate
Bromate (CAS #7789-38-0 as sodium salt) is a white crystal that is
very soluble in water. Bromate may be formed by the reaction of bromine
with sodium carbonate. Sodium bromate can be used with sodium bromide
to extract gold from gold ores. Bromate is also used to clean boilers
and in the oxidation of sulfur and vat dyes. It is formed in water
following disinfection via 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.
Occurrence and Human Exposure. Bromide and organobromine compounds
occur in raw waters from both natural and anthropogenic sources.
Bromide can be oxidized to bromate or hypobromous acid; however, in the
presence of excess ozone, bromate is the principal product.
Table V-13 presents occurrence information available for bromate in
drinking water. Descriptions of this data are detailed in ``Occurrence
Assessment for Disinfectants and Disinfection By-Products (Phase 6a) in
Public Drinking Water,'' (USEPA, 1992a__). Significant bromate
concentrations may occur in ozonated water with bromide. More recent
occurrence data on bromate and the influence of bromide concentration
and ozone on bromate formation is discussed in Section VI of this
preamble.
Table V-13.--Summary of Occurrence Data for Bromate
--------------------------------------------------------------------------------------------------------------------------------------------------------
Occurrence of bromate in drinking water
---------------------------------------------------------------------------------------------------------------------------------------------------------
Concentration (g/L)
Survey (year)\1\ Location Sample information (No. ------------------------------------------------------------------------
of samples) Range Mean Median Other
--------------------------------------------------------------------------------------------------------------------------------------------------------
McGuire et al., 1990\2\.. MWD Pilot Plant Studies.. Ozonation: Hydrogen Max. 60
Peroxide/Ozone. Max. 90
EPA, 1992b\2\ (1987-1991) Disinfection By-Products Finished Water, Plants <10 ........... ........... Detection
Field Studies. Not Using Ozone (33). Limit of 5 g/L
--------------------------------------------------------------------------------------------------------------------------------------------------------
\1\Dates indicate period of sample collection.
\2\May not be representative of national occurrence.
EPA: Environmental Protection Agency.
Although bromate is used as a maturing agent in malted beverages,
as a dough conditioner, and in confectionery products (Borum, 1991),
monitoring data are not available to adequately characterize the
magnitude or frequency of potential bromate exposure from the diet.
Currently, the Food and Drug Administration does not have available
data for bromate in foods, as bromate is not a part of their Total Diet
Study program. Similarly, EPA's Office of Air and Radiation is not
currently measuring for bromate in air (Borum 1991).
Since only a limited number of food groups are expected to contain
bromate and no significant bromate levels are expected in ambient or
indoor air, EPA believes that drinking water is the predominant source
of intake for bromate, and contributions from air and food would be
small. Characterization of potential exposures from food and air are
issues currently under review. EPA requests any additional data on
known concentrations of bromate in drinking water, food, and air.
Health Effects. The health effects information in this section is
summarized from the Drinking Water Health Quantification of
Toxicological Effects Document for Bromate (USEPA, 1993b). Studies
mentioned in this section are summarized in the criteria document.
The noncancer effects of ingested bromate have not been well
studied. Bromate is rapidly absorbed, in part unchanged, from the
gastrointestinal tract following ingestion. It is distributed
throughout the body, appearing in plasma and urine as bromate and in
other tissues as bromide. Following exposure to bromate, bromide
concentrations were significantly increased in kidney, pancreas,
stomach, small intestine, red blood cells and plasma. Bromate is
reduced in tissues probably by glutathione or by other sulfhydryl-
containing compounds. Excretion occurs via urine and to a lesser extent
feces.
Acute oral LD50 values range from 222 to 360 mg bromate/kg for
mice and 500 mg/kg for rats. Acute symptoms of toxicity include
decreased locomotion and ataxia, tachypnea, hypothermia, hyperemia of
the stomach mucosa, kidney damage and lung congestion. In subchronic
drinking water studies, decreased body weight gain and marked kidney
damage were observed in treated rodents. These effects were observed at
the lowest doses tested (30 mg/kg/d).
Bromate was positive in a rat bone marrow assay to determine
chromosomal aberrations. Positive findings for bromate were also
reported in a mouse micronucleus assay. Bromate has also been found to
be carcinogenic to rodents following long-term oral administration. In
these studies, an increased incidence in kidney tumors was reported for
male and female rats. Other tumors observed include thyroid follicular
cell tumor and peritoneal mesothelioma. No carcinogenic effects have
been seen in mice. Dose and time studies indicate that the minimum
exposure time to produce tumors in rats is 13 weeks.
The available data are considered insufficient to calculate an RfD.
Only one noncarcinogenic toxicity study (Nakano et al., 1989) was
located in the literature. The study failed to provide dose response
data and did not identify a NOAEL. Histopathological lesions in kidney
tubules that coincided with decreased renal function were noted in rats
exposed to 30 mg bromate/kg/d for 15 months. The available
carcinogenicity studies also do not provide sufficient information on
non-cancer related effects to determine an RfD.
In a cancer bioassay, Kurokawa et al. (1986a) supplied groups of 50
male and 50 female F344 rats (4-6 weeks old) with drinking water
containing 0, 250 or 500 mg/L (the maximum tolerated dose) of potassium
bromate (KBrO3). The high dose (500 mg/L) caused a marked
inhibition of weight gain in males, and so at week 60 this dose was
reduced to 400 mg/L. Exposure was continued through week 110. The
authors stated the average doses for low dose and high dose groups were
12.5 or 27.5 mg KBrO3/kg/day in males (equivalent to 9.6 and 21.3
mg BrO3/kg/day) and 12.5 or 25.5 mg KBrO3 in females
(equivalent to 9.6 and 21.3 mg BrO3). The incidence of renal
tumors in the three groups (control, low dose, high dose) was 6%, 60%
and 88% in males and 0%, 56% and 80% in females. The effects were
statistically significant (p <0.001) in all exposed groups. The
incidence of peritoneal mesotheliomas in males at three doses was 11%
(control), 33% (250 mg/L, p <0.05) AND 59% (500 mg/L, p <0.001). The
authors concluded that KBrO3 was carcinogenic in rats of both
sexes.
In a subsequent study, Kurokawa et al. (1986b) supplied F344 rats
with water containing KBrO3 at 0, 154, 30, 60, 125, 250 or 500 mg/
L for 104 weeks. The authors reported that these exposures resulted in
average doses of 0, 0.9, 1.7, 3.3, 7.3, 16.0 or 43.4 mg/kg/day of
KBrO3, equivalent to doses of 0, 0.7, 1.3, 2.5, 5.6, 12 or 33.4
mg/kg/day of BrO3. The incidence of renal cell tumors in these
dose groups was 0%, 0%, 4%, 21% (p <0.05), 50% (p <0.001), 95% (p
<0.001) and 95% (p <0.001). Using the linearized multistage model,
estimates of cancer risks were derived. Combining incidence of renal
adenomas and adenocarcinomas in rats, and a daily water consumption for
an adult, lifetime risks of 10-4, 105 and 106 are
associated with bromate concentrations in water at 5, 0.5 and 0.05
g/L, respectively. Equivalent concentrations in terms of
KBrO3, lifetime risks would be 7, 0.7 and 0.07 g/L,
respectively.
The International Agency for Research on Cancer placed bromate in
Group 2B, for agents that are probably carcinogenic to humans. EPA has
performed a cancer weight of evidence evaluation, and has placed
bromate in Group B2: probable human carcinogen since bromate has been
shown to produce several types of tumors in both sexes of rats
following drinking water exposures. In addition, positive mutagenicity
studies which have been reported include indications of DNA
interactions with bromate. As a result of bromate formation following
disinfection, particularly with ozone, there is a potential for
considerable exposure in drinking water. Thus, EPA is proposing an MCLG
based on carcinogenicity and a Category I approach. The resulting MCLG
is zero.
EPA is also interested in examining the mechanism of toxicity of
bromate in rats in terms of whether renal tumor formation is due to
direct action of bromate or indirectly through formation of specific
adduct in kidney DNA of rats treated with bromate.
EPA requests comment on the MCLG of zero based on carcinogenic
weight of evidence and the mechanism of action for carcinogenicity
related to DNA adduct.
VI. Occurrence of TTHMs, HAA5, and other DBPs
A. Relationship of TTHMs, HAA5 to Disinfection and Source Water Quality
1. Primary and Residual Disinfectant Use Patterns in U.S. and
Relationship to Formation of DBPs
A survey of 727 utilities nationwide was conducted for the American
Water Works Association Research Foundation (AWWARF) in 1987 to
determine the extent and cost of compliance with the 1979 maximum
contaminant level (MCL) for trihalomethanes (THMs) (McGuire et al.,
1988). The AWWARF survey reflected more than 67 percent of the
population represented by water utilities serving more than 10,000
customers. The survey found that chlorine remained the most common
disinfectant among water utilities. At the time of the survey, chlorine
was used by 85 percent of the flowing stream and lake surface water
systems and by 80 percent of the ground water systems. The median
chlorine dose for flowing stream and lake systems was 2.2-2.3 mg/L and
for ground water systems it was 1.2 mg/L. The range of chlorine doses
was 0.1 to >20 mg/L.
Chloramines were used by 25 percent of the flowing stream systems
and larger lake systems, but by only 13 percent of the smaller lake
systems. Chloramines were rarely used by ground water systems reporting
in the AWWARF THM survey. Typical chloramine doses for flowing stream
systems was 2.7 mg/L, compared with 1.5 mg/L for lake systems. In
addition, 10 percent of the flowing stream systems and 5 percent of the
lake systems reported using chlorine dioxide. The latter systems
typically served more than 25,000 customers. The typical chlorine
dioxide doses ranged from 0.6 mg/L for the flowing stream systems to
1.0 mg/L for the lake systems. No ground water systems reported using
this disinfectant. At the time of this survey, three utilities reported
using ozone.
The AWWA Disinfection Committee also performed nationwide surveys
on disinfectant use in 1978 (AWWA Disinfection Committee, 1983) and
1990 (AWWA Water Quality Division Disinfection Committee, 1992),
principally among systems serving >10,000 persons (<3 percent of the
surveyed systems served 10,000 persons or fewer). Chlorine has
historically been applied early in the water treatment process
(precoagulation) in order to utilize the benefit of chlorine as a
disinfectant and an oxidant and to control biological growths in
basins. In the 1978 survey, the vast majority (>85 percent) of those
who relied on surface waters prechlorinated (AWWA Disinfection
Committee, 1983). The 1990 survey found a significant reduction in the
frequency of chlorine addition prior to coagulation, along with an
increase in chlorine application after sedimentation (AWWA Water
Quality Division Disinfection Committee, 1992). The AWWARF THM survey
had found that 150 systems surveyed had changed the point of
disinfection to comply with the 0.10-mg/L THM MCL (McGuire et al.,
1988). However, the 1990 AWWA survey (AWWA Water Quality Division
Disinfection Committee, 1992) still found that 35 percent of the
utilities reported chlorination before coagulation or sedimentation.
The range and median chlorine doses in the 1990 AWWA survey were
similar to the AWWARF THM survey.
In the 1990 AWWA survey, disinfection modifications to reduce THMs
included (1) changes in prechlorination practices (24 percent of
respondents moved the first point of chlorination, 23 percent ceased
prechlorination, while 20 percent decreased the prechlorination dose),
(2) implementation of ammonia addition (19 percent added ammonia after
some free chlorine time, while nine percent added ammonia before
chlorination), (3) or changed preoxidant (10 percent switched to
potassium permanganate, five percent to chlorine dioxide, and 0.5
percent to ozone). A surprisingly large percentage of utilities
reported operational problems with disinfection modifications used for
THM reduction (e.g., 56 percent of utilities that implemented
postammoniation reported such problems; as well as 44, 36, and 28
percent of those who moved the first point of chlorination downstream,
ceased prechlorination, and decreased the prechlorination dose,
respectively). Neither the exact nature of the problems noted, nor
their duration, were defined in the survey. However, the Disinfection
Committee believed that many of the reported problems were probably
transitional and were alleviated after further experience.
The 1990 AWWA survey (Haas et al., 1990) found that disinfection
modifications for THM minimization differed between ground and surface
water utilities. For example, 13 percent of surface water systems
changed their preoxidation practices, while this option was rarely used
by ground water systems (which rarely preoxidize). Sixteen and 25
percent of surface and ground water utilities, respectively, reported
adding ammonia after some free chlorine contact as their modification
strategy to reduce THMs. Because 65 percent of the surveyed ground
waters had a THM formation potential (THMFP) (a worst-case measure of
the possible THM production rather than the amount actually produced in
the distribution system) of <100 g/l, most ground water
systems probably did not require modifications to meet the 1979 TTHM
rule.
AWWA established a Water Industry Data Base (WIDB) in 1990-91 (AWWA
Water Industry Data Base, 1991). The WIDB contains information from
about 500 utilities supplying water to more than 50,000 people and over
800 utilities supplying between 10,000 and 50,000 people. The utilities
in the WIDB represent a combined population of 209 million people. In
addition, a database for the Disinfectants/Disinfection By-Products (D/
DBP) negotiated regulation (``reg neg'' data base, RNDB) (JAMES M.
MONTGOMERY, CONSULTING ENGINEERS, INC., 1992) was developed for AWWA.
The RNDB comprises data on nationwide and regional DBP studies,
including data on individual THMs and haloacetic acids (HAAs), chloral
hydrate, and bromate, performed by EPA, water utilities, universities,
and engineering consultants, as well as total THM (TTHM) data from the
WIDB. The non-WIDB part of the RNDB (i.e., those studies on individual
DBP occurrence and control) includes 166 utilities serving a combined
population of about 72 million people. The majority of systems in the
non-WIDB data occurrence part of the RNDB are also in the WIDB. Thus,
the former data base represents a subset of the latter data base, in
which specialized DBP studies were conducted. In addition, many of
these studies attempted to select utilities that were representative of
source water quality, treatment plant operations, disinfectant use,
population served, and geographical locations throughout the United
States (Krasner et al., 1989). Furthermore, the RNDB includes data on
48 utilities (serving a combined population of 37 million people) which
have evaluated alternative treatments to comply with future DBP
regulations.
Figure VI-1 shows a comparison of disinfectant/oxidant uses
reported in the WIDB and the non-WIDB part of the RNDB. In general, the
current usage of disinfectants/oxidants in both data bases are
comparable, which indicates that the non-WIDB part of the RNDB is
representative of nationwide disinfectant usage. Figure VI-2 shows
disinfectants evaluated under alternative treatments in the RNDB. While
ozone is the most prevalent alternative disinfectant under
investigation in the RNDB, this data base is somewhat biased, as it
does include two AWWARF studies involving ozonation. However, Figure
VI-2 does demonstrate that ozone is an alternate disinfectant that is
being widely evaluated. While most systems currently use chlorine only,
the percentage drops when the data are population based. Figure VI-1
shows that chloramine use is higher on a population basis, probably due
to its usage by some of the larger utilities.
BILLING CODE 6560-50-P
TP29JY94.011
TP29JY94.012
BILLING CODE 6560-50-C
2. National Occurrence of TOC
The total organic carbon (TOC) level of a water is generally a good
indication of the amount of THM and other DBP precursors present in a
water (Singer et al., 1989). In the WIDB, 157 utilities provided TOC
data. For the 100 surface waters with TOC data, the range was ``not
detected'' (ND) to 30 mg/L. For these waters, the 25th, 50th, and 75th
percentiles were 2.6, 4.0, and 6.0 mg/L, respectively. For the 57
ground waters with TOC data, the range was ND to 15 mg/L. For these
waters, the 25th, 50th, and 75th percentiles were ND, 0.8, and 1.9 mg/
L, respectively. Typically, most ground waters are low in TOC. However,
there are some high-TOC ground waters, especially in the southeastern
part of the United States (EPA Region IV; see Figure VI-3 and Table VI-
1). For surface waters, the high-TOC waters also tend to be in the
southeastern part of the United States, although there are some
relatively high-TOC waters in the south central (EPA Region VI) and the
mountain (EPA Region VIII) states (see Figure VI-3 and Table VI-2).
BILLING CODE 6560-50-P
TP29JY94.013
BILLING CODE 6560-50-C
Table VI-1.--Statistics on Average Raw Groundwater Total Organic Carbon (mg/L) for Utilities in the AWWA Water
Industry Database
----------------------------------------------------------------------------------------------------------------
Number of Percentile
Number of utilities -----------------------------------------
EPA region utilities with missing Min value Max value
with data data 25th 50th 75th
----------------------------------------------------------------------------------------------------------------
1............ 2 48 ND 1.38 ............ ............ ............
2............ 3 72 ND 1.91 ............ ............ ............
3............ 3 96 ............ ............ ............ 2.74 ............
4............ 13 163 ND 15.00 ND 2.19 8.50
5............ 11 182 ND 4.00 0.63 1.45 1.92
6............ 4 82 ND 1.87 ............ ............ ............
7............ 5 53 ND 1.40 ............ ............ ............
8............ 4 46 0.71 2.00 ............ ............ ............
9............ 11 118 ND 1.00 ND ND 0.30
10........... 1 37 0.80 0.80 ............ ............ ............
ALL.......... 57 897 ND 15.00 ND 0.84 1.88
----------------------------------------------------------------------------------------------------------------
Table VI-2.--Statistics on Average Raw Surface Water Total Organic Carbon (mg/L) for Utilities in the AWWA Water
Industry Database
----------------------------------------------------------------------------------------------------------------
Number of Percentile
Number of utilities -----------------------------------------
EPA region utilities with missing Min value Max value
with data data 25th 50th 75th
----------------------------------------------------------------------------------------------------------------
1............ 6 44 3.00 9.00 3.48 3.50 4.50
2............ 10 65 2.10 20.00 2.50 4.55 5.00
3............ 20 79 ND 25.00 2.15 2.87 4.80
4............ 11 165 1.60 30.00 5.27 7.40 12.60
5............ 14 179 ND 9.17 2.70 4.50 5.90
6............ 10 76 2.00 10.00 3.90 5.50 6.90
7............ 2 56 7.00 10.00 ............ ............ ............
8............ 7 43 1.00 14.00 1.75 3.30 8.50
9............ 16 113 ND 5.90 1.95 3.25 3.88
10........... 4 34 1.25 3.30 ............ ............ ............
ALL.......... 100 854 ND 30.00 2.55 4.00 5.95
----------------------------------------------------------------------------------------------------------------
For surface waters that filter but do not soften, the median and
90th percentile TOC levels are 3.7 and 7.5 mg/L, respectively (see
Figure VI-4). However, when the data are flow-weighted (which would
represent more closely the distribution by population), the median and
90th percentile values drop to 2.7 and 5.1 mg/L, respectively (see
Figure VI-4). This is due, in part, to a number of large facilities
treating water with TOC levels <4 mg/L. When this same category of
surface waters is examined for choice of disinfectants between chlorine
and chloramines, the latter group has a higher TOC cumulative
probability than the former (see Figure VI-5). Switching from free
chlorine only to chloramination was one of the options utilized by
utilities with high-TOC waters to comply with the 0.10 mg/L TTHM MCL.
BILLING CODE 6560-50-P
TP29JY94.014
TP29JY94.015
BILLING CODE 6560-50-C
As part of a ground water supply survey (GWSS), TOC was measured at
the point of entry into the distribution system (see Figure VI-6).
Because most groundwater systems do not have precursor-removal
technology as part of their treatment, these treated-water TOC levels
provide a good indication of the range of raw-water TOC levels in
ground waters. The median and 90th percentile TOC levels of systems
without softening who chlorinate were 0.7 and 2.9 mg/L, respectively.
However, the median and 90th percentile TOC levels of systems with
softening who chlorinate were 1.7 and 6.8 mg/L, respectively.
BILLING CODE 6560-50-P
TP29JY94.016
In addition, the breakdown of treated-water TOC levels of ground
waters was examined geographically (see Table VI-3). As indicated above
(see Table VI-1), the southeastern part of the United States (i.e., EPA
Region 4) has groundwaters with a relatively higher level of TOC (see
Tables VI-3A and VI-3C). In addition, the mountain states (i.e., EPA
Region 8) also tended to have a higher distribution of TOCs in the
ground waters tested (see Tables VI-3A and VI-3C). Furthermore, these
data are also broken down into those that chlorinate and those that do
not (see Table VI-3). Typically, ground waters that are currently
undisinfected tend to be ones with lower TOC levels. Thus, promulgation
of the Ground Water Disinfection Rule will probably tend to have an
impact on waters with a lower precursor level than are currently
disinfecting.
Table VI-3A.--TOC Values for GWSS Systems That Chlorinate
----------------------------------------------------------------------------------------------------------------
Number of Percentile
EPA region utilities Max value -------------------------------------------------------
with data 25th 60th 75th 90th
----------------------------------------------------------------------------------------------------------------
1............................ 19 4.3 0.4 0.5 0.95 2.3
2............................ 59 4.6 0.3 0.5 0.9 1.5
3............................ 62 4.2 0.3 0.5 0.7 1.2
4............................ 185 14 0.8 0.7 2.1 5.3
5............................ 90 8.9 0.7 1.1 1.8 2.6
6............................ 46 8 0.6 1 1.7 2.4
7............................ 103 7.8 0.8 0.8 1.8 3.7
8............................ 26 11 0.6 1.9 3.1 6.6
9............................ 39 11 <2 0.2 0.5 1.2
10........................... 25 3.3 0.3 0.9 1.4 2.2
ALL.......................... 654 14 0.3 0.7 1.7 3.3
----------------------------------------------------------------------------------------------------------------
Table VI-3B.--TOC Values for GWSS Systems That Do Not Chlorinate
----------------------------------------------------------------------------------------------------------------
Number of Percentile
EPA region utilities Max value -------------------------------------------------------
with data 25th 50th 75th 90th
----------------------------------------------------------------------------------------------------------------
1............................ 27 3.6 0.3 0.5 0.8 1.3
2............................ 14 0.8 <2 0.3 0.5 0.6
3............................ 28 5.3 <2 0.3 0.6 1.5
4............................ 34 3.2 0.3 0.5 0.8 1.7
5............................ 41 18 0.7 1.3 2.2 3.4
6............................ 26 5.6 0.2 0.6 1.5 2.9
7............................ 18 2.9 0.6 0.9 1.1 2.2
8............................ 14 5.9 0.3 0.4 1.8 2.7
9............................ 49 3.4 <2 0.3 0.4 0.9
10........................... 40 5 0.2 0.4 0.8 2
ALL.......................... 291 18 0.3 0.5 1.1 2.2
----------------------------------------------------------------------------------------------------------------
Table VI-3C.--TOC Values for all GWSS Systems
----------------------------------------------------------------------------------------------------------------
Number of Percentile
EPA region utilities Max value -------------------------------------------------------
with data 25th 50th 75th 90th
----------------------------------------------------------------------------------------------------------------
1............................ 46 4.3 0.3 0.5 0.8 1.6
2............................ 73 4.6 0.3 0.5 0.8 1.3
3............................ 90 5.3 0.3 0.4 0.7 1.2
4............................ 219 14 0.3 0.7 1.9 4.8
5............................ 131 18 0.7 1.2 1.9 3.2
6............................ 72 8 0.4 0.9 1.7 2.4
7............................ 121 7.8 0.4 0.9 1.7 3.2
8............................ 40 11 0.3 1.8 2.7 5.9
9............................ 88 11 <2 0.2 0.5 1
10........................... 65 5 0.2 0.5 1.2 2.2
----------------------------------------------------------------------------------
ALL.......................... 945 18 0.3 0.6 1.4 2.9
----------------------------------------------------------------------------------------------------------------
3. National Occurrence of Bromide
Bromide is a concern in both chlorinated and ozonated supplies. In
chlorinated supplies, while the organic precursor level of a source
water has an impact on the amount of DBPs formed, the bromide
concentration has an impact on the speciation as well as the overall
yield (Symons et al., 1993). Typically, regardless of the organic
content in water, bromate can be formed when waters containing
sufficient levels of bromide are ozonated (Krasner et al., Jan. 1993).
In a 35-utility nationwide DBP study, bromide ranged from <0.01 to
3.0 mg/L and the median bromide level was 0.1 mg/L (Krasner et al.,
1989). Some utilities have bromide in their source water due to
saltwater intrusion (one utility had as much as 0.4 to 0.8 mg/L bromide
due to this phenomenon) (Krasner et al., 1989). However, some non-
coastal communities can have moderate-to-high levels of bromide due to
connate waters (ancient seawater that was trapped in sedimentary
deposits at the time of geological formation) or industrial and oil-
field brine discharges. The highest bromide detected in the latter
study (2.8-3.0 mg/L) (Krasner et al., 1989) was from a water in the
midsouthern part of the U.S.
Currently, a nationwide bromide survey of 70 utilities has found
bromide levels ranging from <0.005 to >3.0 mg/L (Amy et al., 1992-3).
Some waters have been sampled more than once (up to three seasonal
samples to date) in order to determine the variability in bromide
occurrence. The average raw-water bromide level per water, though,
provides an indication of the typical occurrence of bromide in each
water. Table VI-4 provides some preliminary insight into the
geographical occurrence of bromide. Ideally, more data per region are
needed; however, sufficient data are available for general trends.
Regions 6 (which includes Texas) and 9 (which includes California) have
the highest occurrence of bromide. While some California communities
have problems with saltwater intrusion, some Texas communities may have
bromide from connate waters or oil-field brines. However, most
geographical regions have at least one high-bromide water in their
area, except for the systems surveyed in the Pacific Northwest (EPA
Region 10) and the northeast (EPA Regions 1 and 2).
Table VI-4.--Statistics on Average Raw-Water Bromide (mg/L) for Utilities in the Nationwide Bromide Survey
----------------------------------------------------------------------------------------------------------------
Number of Percentile
EPA region utilities Min value Max value ---------------------------------------------------
with data 25th 50th 75th 90th
----------------------------------------------------------------------------------------------------------------
1....................... 8 0.005 0.089 0.02 0.03 0.05 0.05
2....................... 4 0.023 0.093 0.03 0.05 0.08 NA
3....................... 8 0.005 0.276 0.03 0.06 0.07 0.08
4....................... 7 0.010 0.190 0.02 0.04 0.05 0.05
5....................... 6 0.012 0.322 0.05 0.09 0.12 0.14
6....................... 7 0.014 >3.00 0.02 0.03 0.25 0.37
7....................... 6 0.042 0.206 0.06 0.08 0.09 0.10
8....................... 7 0.006 0.368 0.02 0.02 0.06 0.09
9....................... 11 0.008 0.429 0.05 0.08 0.33 0.36
10...................... 6 <0.005 0.015 <0.005 0.006 0.009 0.012
----------------------------------------------------------------------------------------------------------------
NA=Not applicable; insufficient number of utilities to determine.
Figures VI-7 and VI-8 show the cumulative probability distribution
of average raw-water bromide levels in surface and ground waters,
respectively, in the nationwide bromide survey. In surface waters in
this survey, the median, 75th, 90th, and 95th percentile bromide level
were 0.04, 0.08, 0.2, and 0.35 mg/L, respectively. In ground waters in
this survey, the median, 75th, 90th, and 95th percentile bromide level
were 0.06, 0.1, 0.25, and 0.35 mg/L, respectively. Overall, ground
waters appear to have a somewhat higher probability of bromide
occurrence than in surface waters. In the 35-utility DBP study, one
midwest utility pumped ground water into a lake to augment a low-lake
level during a drought period. Bromide rose from 0.19 to 0.68 mg/L
during this period of time.
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B. Chlorination Byproducts
1. TTHMs--Occurrence Studies
Prior to the promulgation of an MCL of 0.10 mg/L TTHMs in 1979, EPA
performed two surveys to obtain information on the occurrence of THMs
and other organic compounds: the National Organics Reconnaissance
Survey (NORS) in 1975 (Symons et al., 1975) and the National Organic
Monitoring Survey (NOMS) in 1976-77 (Brass et al., 1977 and The
National Organics Monitoring Survey, unpubl.). NORS and NOMS were
conducted primarily to determine the extent of THM occurrence in the
United States. These data were used, in part, in determining the 1979
THM regulation. Surveys in the 1980s were performed to provide data for
assessing a new MCL for THMs, as well as to develop regulations for
other DBPs.
The AWWARF THM survey used data from 1984-86, and these THM values
reflected the result of compliance with the 1979 THM regulation. Mean
TTHM values were computed for each of the utilities in the AWWARF THM
survey; these means, as well as data from the NORS and NOMS surveys,
are plotted (see Figure VI-9) on a frequency distribution curve. The
AWWARF survey's overall TTHM average was 42 g/L, which was a
40-50 percent reduction in national THM concentrations as compared to
the averages of the NORS and NOMS (all phases) results. It is important
to note that the disinfection practices of some of the utilities in the
AWWARF survey (such as the use of chloramines as a primary
disinfectant) were employed to meet the 1979 TTHM MCL, and not to meet
the requirements of the recently promulgated Surface Water Treatment
Rule (SWTR). Thus, THM and DBP levels at some utilities would most
likely be different if their current treatment practices required
modification in order to meet the new disinfection requirements of the
SWTR.
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Median TTHM concentrations in the AWWARF survey for the spring,
summer, fall, and winter seasons were 40, 44, 36, and 30 g/L,
respectively. THM levels were highest in the summer and lowest in the
winter, due primarily to the faster formation rates in warmer water
temperatures. In the 35-utility DBP study, the second highest THM
levels were in the fall (Krasner et al., 1989). For many utilities in
California and the southern United States, fall can be almost as warm
as summer. However, seasonal impacts may be due to changes in the
nature of naturally occurring organics or bromide levels as well.
Compliance with the THM regulation is based on a running annual average
to reflect these types of seasonal variations.
Because the 1979 regulation did not apply to systems that serve
<10,000 people, the running annual average TTHM distribution for small
systems is expected to be different. In the AWWARF THM survey, TTHM
data for small systems from 12 states were obtained (McGuire et al.,
1988). While the number of utilities (677) for which TTHM data were
received represents only a small percentage of the total number serving
fewer than 10,000 customers (55,449), some important observations can
be made. The range of TTHMs was from ND to 313 g/L, with a
mean of 36 g/L and a median of 18 g/L (McGuire et
al., 1988). The cumulative probability distribution differs
significantly from the NORS and NOMS data (see Figure VI-10). This lack
of agreement is probably due to many of the small systems using ground
water sources, which are generally much lower in THM precursors than
surface water sources. In addition, the overall statistics of the
AWWARF survey (for 677 cities) were markedly affected by the low TTHM
results (range of ND to 42 g/L with a mean of 2 g/L)
of the 204 systems sampled in Wisconsin. Although McGuire does not
identify a reason for low TTHMs in Wisconsin, EPA data indicate that
over 90 percent of Wisconsin systems use ground water (probably with
low precursor levels) as a primary source. Since 30 percent of the
systems in the survey were from Wisconsin, this would bias the results.
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Since the AWWARF THM survey, EPA measured DBP data in a number of
small systems. These data represent part of the non-WIDB data in the
RNDB. Figure VI-11 compares the TTHM frequency distribution for the
WIDB (large systems only) with that of the non-WIDB data on both large
and small systems. For the small systems, there is essentially a
biomodal distribution of TTHM levels: 50 percent of the small systems
have 10 g/L TTHMs, while the remaining utilities
have TTHM levels of 20 to 430 g/L. Most likely, many of the
very low THM levels are associated with treatment of low-TOC, low-
bromide ground waters. For community water, non-purchased systems
serving <10,000 people, 4562 systems treat surface water, while 17941
disinfect ground water. For systems serving >10,000 people, 1395 treat
surface water and 1117 disinfect ground water. Thus, small systems are
utilizing ground water more than surface water.
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In the WIDB (which only includes large systems), 482 utilities that
treat surface water or a mix of surface and ground waters had TTHM
median, 75th, and 90th percentile values of 43, 59, and 74 g/
L, respectively. In the WIDB, 277 utilities that treat ground water
only had TTHM median, 75th, and 90th percentile values of 13, 34, and
60 g/L, respectively. However, systems using both types of
source waters had TTHM levels in the neighborhood of 100 g/L.
Thus, while ground waters in general tend to form less THMs than
surface waters, there are some ground waters with sufficient precursor
levels to form significant amounts of THMs.
2. HAAs and Other Chlorination DBPs--Occurrence Studies
a. Discovery of Additional Chlorination By-Products. In 1985, EPA
determined chlorination DBPs at 10 operating utilities, using both
target-compound and broad-screen analyses (Stevens et al., 1989). A
total of 196 compounds that can be attributed to the chlorination
process were found in one or more of the 10 utilities' finished waters.
Approximately half of the compounds contained chlorine and many were
structurally identified; however, 128 compounds were of unknown
chemical structure. The compounds which were quantifiable represented
from 30 to 60 percent of the total organic halide (TOX) of those
supplies. That study served to significantly reduce the list of
compounds that EPA considered most significant for further work.
b. Available Data on Chlorination By-Products. Taken as an example
of subsequent survey results where quantifiable target-compound
analyses were used, Figure VI-12 shows the occurrence of DBPs in the
35-utility study (Metropolitan Water District of So. Calif et al.,
1989). The figure presents an overview of the results of four seasonal
sampling quarters combined. In addition, all sampling was performed at
treatment-plant clearwell effluents. It is important to note that these
survey results do not reflect any impacts of the SWTR under which a
substantial number of systems could be expected to modify disinfection
practice to achieve compliance. On a weight basis, THMs were the
largest class of DBPs detected in this study; the second largest
fraction was haloacetic acids (HAAs). At the time of this study,
commercial standards were only available for five of the nine
theoretical species: monochloro-, dichloro-, trichloro-, monobromo-,
and dibromoacetic acid. The data indicate that the median level of THMs
(i.e., 36 g/L) was approximately twice that of HAAs (i.e., 17
g/L). The third largest fraction was the aldehydes (i.e.,
formaldehyde and acetaldehyde). These two low-molecular-weight
aldehydes were initially discovered as by-products of ozonation, but
they also appear to be by-products of chlorination. Every target-
compound DBP was detected at some time in some utility's water during
the study; however, 2,4,6-trichlorophenol was only detected at low
levels at a few utilities during the first sampling quarter and was not
detected in subsequent samplings.
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The 35-utility DBP study assessed systems using a range of
disinfectants, a number of which used chloramines as a residual
disinfectant. In a study (in 1987-89) by EPA, primarily chlorine-only
systems were evaluated at the plant and in the distribution system
(typically a terminal location). The range of total HAAs (THAAs) (a sum
of the five aforementioned species) at the plant effluent was <1 to 86
g/L (representing 73 samples), with a median value of 28
g/L (Fair, 1992). In the distribution system (56 samples
collected), the range and median THAAs were <1-136 and 35 g/L,
respectively.
In a six-utility DBP survey in North Carolina (Grenier et al.,
1992), the sum of four measured HAAs--dibromoacetic acid was not
included in this study, as these waters are all low in bromide--ranged
from 14 to 141 g/L in the distributed waters (with utility
annual averages of 51 to 97 g/l). In this survey, HAA
concentrations consistently exceeded the concentration of TTHMs (which
ranged from 13 to 114 g/l, with utility annual averages of 34
to 72 g/l). The prevalence of the HAAs may be due, in
part, to chlorination of settled and finished waters with pH levels of
5.9 to 7.8. Chlorination at lower pH levels results in lower THM
formation but higher HAA concentrations (Stevens et al., 1989).
Recently, a commercial standard for bromochloroacetic acid (BCAA)
has become available. Studies to date suggest that the other mixed
bromochloroacetic acids may be unstable (Pourmoghaddas et al., 1992).
The RNDB includes the occurrence of BCAA for 25 utilities. The median,
75th and 90th percentile occurrence were 3, 5, and 8 g/L,
respectively. In the chlorinated distribution system of a water
containing from 0.04 to 0.31 mg/L bromide (i.e., an average- and a
high-bromide source water were being treated), BCAA was present from 6
to 17 g/L and accounted for 25 percent of the concentration of
the sum of the six measured HAA species (D/DBP Regulations Negotiation
Data Base (RNDB), 1992). Thus, most DBP studies which measured only
five of the HAA species will have some level of underestimation of
total HAAs present, although that should be a small error in low
bromide waters.
The RNDB includes HAA data, including from the 35-utility, EPA, and
North Carolina DBP studies. When a utility was sampled more than once
in time and space, a ``quasi'' running annual average value was
determined (RNDB). Figure VI-13 shows the cumulative probability
occurrence of THAAs (the four- to six-species sums) for large and small
systems. The median THAA for either population group is 30 g/
L, although the small systems have 30 percent of the utilities with
7 g/L THAAs. The difference at the low THAA levels
was probably due to treatment of low-precursor source waters in small
systems. The high end of the THAA occurrence was not significantly
different, most likely due to a lack of a HAA regulation and the fact
that pH of chlorination impacts THM and HAA formation in opposite ways.
In the RNDB, 121 of the utilities treated surface water or a ground
water/surface water mix. For those systems treating some percentage of
surface water, the median, 75th, 90th percentile, and maximum values
were 28, 50, 73, and 155 g/L, respectively. In the RNDB, 13 of
the utilities treated ground water only. For this limited ground water
data set, the median, 90th percentile, and maximum values were 4, 13,
and 37 g/L, respectively.
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