REVIEW DRAFT
DRINKING WATER CRITERIA DOCUMENT
FOR
BETA AND GAMMA EMITTING RADIONUCLIDES
Prepared by
Clement International Corporation
1201 Gaines Street
Ruston, Louisiana 71270
and
Wade Miller Associates
1911 North Myer Drive
Arlington, Virginia 22209
July 1991
Prepared for
Drinking Water Standards Division
Office of Ground Water and Drinking Water
and Office of Radiation Programs
U.S. Environmental Protection Agency
Washington, DC 20460
e*»
Printed on Recycled Paper
U-3
CD
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FOREWORD
Section 1412 (b)(3)(A) of the Safe Drinking Water Act, as amended in
1986, requires the Administrator of the Environmental Protection Agency to
publish maximum contaminant level goals (MCLGs) and promulgate National
Primary Drinking Water Regulations for each contaminant, which, in the
judgment of the Administrator, may have an adverse effect on public health and
which is known or anticipated to occur in public water systems. The MCLG is
nonenforceable and is set at a level at which no known or anticipated adverse
health effects in humans occur and which allows for an adequate margin of
safety. Factors considered in setting the MCLG include health effects data
and sources of exposure other than drinking water.
This document provides the health effects basis to be considered in
establishing the MCLG. To achieve this objective, data on pharmacokinetics,
human exposure, acute and chronic toxicity to animals and humans, epidemiology
and mechanisms of toxicity are evaluated. Specific emphasis is placed on
literature data providing dose-response information. Thus, while the
literature search and evaluation performed in support of this document has
been comprehensive, only the reports considered most pertinent in the
derivation of the MCLG are cited in the document. The comprehensive
literature data base in support of this document includes information
published up to 1991; however, more recent data may have been added during the
review process.
When adequate health effects data exist, Health Advisory values for less
than lifetime exposures (1-day, 10-day and longer-term, -10% of an
individual's lifetime) are included in this document. These values are not
used in setting the MCLG, but serve as informal guidance to municipalities and
other organizations when emergency spills or contamination situations occur.
James R. Elder
Director
Office of Ground Water and Drinking Water
DRAFT
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INTRODUCTION
The Environmental Protection Agency (EPA) classifies all radionuclides
as Group A carcinogens based on their property of emitting ionizing radiation
(EPA 1991). A Group A carcinogen is one in which there is sufficient evidence
from epidemiological studies to support a causal association between exposure
to the agent(s) and cancer. The studies that provide this evidence indicate
that, depending on radiation dose and the pattern of exposure, ionizing
radiation can induce cancer in nearly any tissue or organ in the body (ATSDR
1990a,b,c). Radiation-induced cancers in humans are found to occur in the
hemopoietic system, the lung, thyroid, liver, bone, skin, and other tissues.
Man may be exposed to radionuclides by several routes, one of which is
through ingestion of drinking water containing these radionuclides. Several
Drinking Water Criteria Documents for radionuclides have already been
completed, including those on naturally occurring radionuclides such as
radium, radon, and uranium. These three radionuclides decay (transform)
primarily, by alpha emission. This document estimates the risks of ingesting
drinking water containing radionuclides which decay by beta particles or gamma
rays. Beta emitters decay by the emission of either a negative or positive
electron. If the electron originates in the nuclides during the
transformation of a neutron into a proton (or the reverse reaction), the decay
is referred to as a beta decay. A gamma ray is an energetic packet of
electromagnetic radiation (referred to as a "photon") produced in the nucleus
during a change in the energy level of a nucleon (i-e., a proton or a
neutron). Three beta or gamma-emitting radionuclides are discussed in depth
because drinking water, either surface or groundwater, are routinely monitored
for their presence. These include strontium-90, tritium, and iodine-131. For
these three radionuclides occurrence in drinking water, chemical and physical
properties, toxicokinetics, and health effects are discussed. In addition,
discussions on mechanisms of toxicity and quantification of cancer effects are
presented.
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TABLE OF CONTENTS
Page
FOREWORD i
INTRODUCTION ii
I. SUMMARY 1-1
II. NATIONAL OCCURRENCE OF BETA OR GAMMA EMITTING
ACTIVITY IN DRINKING WATER II-l
A. INTRODUCTION II-l
B. CURRENT STANDARDS AND COMPLIANCE MONITORING REQUIREMENTS . . 11-2
C. MAN-MADE RADIONUCLIDE OCCURRENCE DATA 11-5
Environmental Radiation Data Reports II-5
National Inorganics and Radionuclides Survey (NIRS) .... 11-10
D. CONCLUSIONS 11-11
III. PROFILES OF REPRESENTATIVE BETA OR GAMMA EMITTING RADIONUCLIDES . III-l
A. INTRODUCTION I II-l
B. STRONTIUM 111-2
Physical and Chemical Properties III-2
Toxicokinetics III-6
Health Effects in Animals III-8
Health Effects in Humans 111-15
C. IODINE 111-17
Physical and Chemical Properties 111-17
Toxicokinetics 111-20
Health Effects in Animals ... 111-23
Health Effects in Humans 111-27
D. TRITIUM 111-31
Physical and Chemical Properties 111-31
Toxicokinetics 111-34
Health Effects in Animals 111-36
Health Effects in Humans 111-40
E. BIOACCUMULATION AND RETENTION 111-41
Hydrogen Model 111-43
Helium Model 111-43
Lithium Model 111-43
Beryllium Model 111-44
Boron Model 111-44
Carbon Model 111-44
Nitrogen Model 111-45
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Oxygen Model 111-46
Fluorine Model 111-46
Neon Model 111-46
Sodium Model 111-47
Magnesium Model 111-47
Aluminum Model 111-47
Silicon Model 111-48
Phosphorus Model 111-48
Sulfur Model 111-49
Chlorine Model 111-49
Argon Model 111-50
Potassium Model 111-50
Calcium Model 111-51
Scandium Model 111-51
Titanium Model 111-52
Vanadium Model 111-52
Chromium Model 111-52
Manganese Model 111-53
Iron Model 111-53
Cobalt Model 111-54
Nickel Model 111-54
Copper Model 111-55
Zinc 111-55
Gallium Model 111-56
Germanium Model 111-56
Arsenic Model 111-57
Selenium Model 111-57
Bromine Model 111-58
Krypton Model 111-58
Rubidium Model 111-59
Strontium Model 111-59
Yttrium Model 111-60
Zirconium Model 111-60
Niobium Model 111-61
Molybdenum Model 111-62
Technetium Model 111-63
Ruthenium Model 111-63
Rhodium Model 111-64
Palladium Model 111-64
Silver Model 111-65
Cadmium Model 111-65
Indium Model 111-66
Tin Model 111-66
Antimony 111-67
Tellurium Model 111-67
Iodine Model 111-68
Xenon Model 111-68
Cesium Model 111-69
Barium Model 111-69
Lanthanide (Rare Earth) Models 111-70
Dysprosium 111-72
Holmium 111-72
Erbium 111-73
Thulium • 111-73
DRAFT iv
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Ytterbium 111-74
Lutetium 111-74
Hafnium Model 111-75
Tantalum Model 111-75
Tungsten Model 111-76
Rhenium Model 111-76
Osmium Model 111-77
Iridium Model 111-77
Platinum Model 111-78
Gold Model 111-78
Mercury Model 111-79
Thallium Model 111-79
Lead Model 111-80
Bismuth Model '. 111-80
Polonium Model 111-81
Astatine Model 111-81
Radon Model 111-82
Francium Model 111-82
Radium Model 111-83
Actinium Model 111-83
Thorium Model 111-84
Protactinium Model 111-84
Uranium Model 111-85
Neptunium Model 111-85
Plutonium Model 111-86
Americium Model . . 111-87
Curium Model 111-87
IV. MECHANISM OF TOXICITY IV-1
A. IONIZING ENERGY DEPOSITION IV-1
B. RADIONUCLIDE DOSIMETRY IV-3
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS V-l
A. NONCARCINOGENIC EFFECTS V-l
Method for Quantification of Noncarcinogenic Effects .... V-l
Quantification of Noncarcinogenic Effects V-4
B. CARCINOGENIC EFFECT V-4
Method for Quantification of Carcinogenic Effects V-4
Quantification of Cancer Risk for Chemicals V-5
Quantification of Carcinogenic Effects V-6
VI. UNCERTAINTY ANALYSIS VI-1
A. UNCERTAINTY IN ASSESSMENT OF NONCARCINOGENIC EFFECTS
OF BETA AND GAMMA EMITTING RADIONUCLIDES VI-1
B. UNCERTAINTY IN ASSESSMENT OF CARCINOGENIC EFFECTS
OF BETA AND GAMMA EMITTING RADIONUCLIDES VI-1
Uncertainty in Parameters Used in the Metabolic Model .... VI-1
Uncertainty in Distribution of Isotope . . . .' VI-2
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Uncertainty in Dosimetric Calculations VI-3
Uncertainty in Risk Coefficients VI-3
Uncertainty in Other Factors Influencing Risk vi-4
Conclusion VI-4
VII. REFERENCES VI-5
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LIST OF TABLES
Page
II-l Summary of Recent ERAMS Data on Beta Emitting
Man-Made Radionuclides in Drinking Water II-7
III-l Physical and Chemical Properties of Strontium III-4
III-2 Physical Properties of Iodine 111-18
III-3 Atomic Properties of Hydrogen, Deuterium, and Tritium .... 111-32
III-4 Physical Properties of Hydrogen, Deuterium, and Tritium . . . 111-32
V-l 50-Year Committed Absorbed Dose per Unit Intake
(millirad/pCi) from Beta and Gamma Emitters in
Drinking Water V-8
V-2 Organ-Specific Lifetime Cancer Risks Used in the RADRISK
Model from High-LET and Low-LET Irradiation V-ll
V-3 Concentration of Beta and Gamma Emitters In Drinking
Water to Yield a Specific Risk of Cancer and Death
for Lifetime Consumption V-14
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LIST OF FIGURES
Page
III-l Half-Lives and Decay Schemes of Strontium-89 and -90 III-5
III-2 Half-Lives and Decay Schemes of Iodine-129 and -131 111-19
III-3 Half-Lives and Decay Schemes of Tritium 111-33
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I. SUMMARY
This document addresses potential risks of cancer from ingestion of man-
made radionuclides that decay by emitting beta particles or gamma rays. Data
on distribution of these radionuclides in the environment are limited.
However, some data on drinking water supplies monitored for strontium-90,
tritium, and iodine-131, as well as gross gamma activity, are available.
In drinking water, average concentrations of the beta emitting
radionuclides of concern are as follows: <0.2 pCi/L (strontium-90), 100 to
300 pCi/L (tritium), and <0.1 pCi/L (iodine-131). None of the drinking water
samples monitored for gamma activity had measurable levels. However, the
detection limit for gamma activity was not reported. These average
concentrations were obtained from the Environmental Radiation Ambient
Monitoring System (ERAMS).
The potential radiation dose to tissues from these radionuclides in
drinking water is dependent on the amount of the substance that may be
absorbed through the gastrointestinal tract and distribution of the absorbed
amount to potential target tissues. The beta emitting radionuclides vary
widely in their potential for absorption through the gut. For example, about
15 to 36 percent of strontium-90 is absorbed, while both iodine-131 and
tritium are very well absorbed, with approximately 95 percent absorption from
the gut.
Once absorbed into the body, each of these radionuclides distributes
according to a different pattern. Strontium distributes preferentially to the
skeleton followed by lesser amounts distributed to soft tissues; while iodine-
131 distributes almost entirely to the thyroid gland. Tritium in the form of
tritiated water behaves like ordinary water and distributes throughout bodily
fluids and, thus, uniformly throughout the body.
Data on humans exposed to these radionuclides are limited. In an
in vitro test using human blood cells exposed to strontium, chromosomal
aberrations, including dicentrics, were induced. No data on effects of
tritium exposure in humans were located. Several studies on effects of
DRAFT 1-1
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iodine-131 exposure in humans were considered in this document. However, most
of these studies were on individuals with preexisting thyroid problems and the
iodine was used as a diagnostic tool or used to destroy abnormal thyroid
tissue. In a study of residents of the Marshall Islands who were exposed to
radioactive iodine from fallout, hypothyroidism, thyroid nodules, and thyroid
cancer were reported. The incidence of thyroid nodules was found to be
statistically significantly related to total radiation dose.
In laboratory animals treated orally with strontium over many months,
effects on blood and blood-forming tissues were most often reported. These
effects included decreased counts of numerous types of blood cells, bone
marrow hypoplasia or hyperplasia, and other forms of marrow proliferative
disease. In addition, immune function was impaired in treated rats. Dogs and
swine administered strontium orally developed bone tumors, tumors in the
mouth, and leukemia.
In animal bioassays, all effects resulting from exposure to iodine-131
were seen in the thyroid gland or adjacent tissues. Oral doses to sheep (5 to
15 mCi) have resulted in damage to or complete destruction of the thyroid
gland. Oral exposure of sheep, dogs, and rats has induced thyroid adenomas
and/or carcinomas in all three species.
Long-term oral exposure of rats to tritium resulted in numerous
hematological effects, including various changes in white blood cells and
decreased erythrocyte, leukocyte, and reticulocyte counts. Pregnant mice and
rats receiving tritium had decreased reproductive capacity such as reduced
litter size and offspring body weights. Direct empirical information on
possible carcinogenic effects of tritium is limited. Male mice given tritium
in drinking water sired offspring having intestinal adenocarcinomas, a tumor
which had not been previously observed in this particular strain of mice.
There is limited evidence that exposure of rats in utero increased the
incidence of ovarian tumors.
Levels of these radionuclides in drinking water estimated to result in a
risk of 1 cancer death in 10,000 persons exposed have been calculated using
the RADRISK model. For the three radionuclides specifically discussed in this
DRAFT 1-2
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document, the levels are as follows: 62 pCi/L (strontium-90), 516 pCi/L
(iodine-131), and 5640 pCi/L (tritium). Uncertainty in these estimates
results from uncertainty in the gastrointestinal absorption factors used, the
kinetic behavior predicted by the model, and the risk factors derived by the
model.
Because noncancer effects have only been observed in animal bioassays
using test doses that exceed any potential doses from ingesting drinking water
containing these radionuclides, criteria for noncancer effects were not
developed.
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II. NATIONAL OCCURRENCE OF BETA OR GAMMA EMITTING
ACTIVITY IN DRINKING WATER
A. INTRODUCTION
This report provides current monitoring data on the presence of man-made
radionuclides in drinking water. The main purpose of the information
presented in this report is to assist EPA in evaluating the potential impacts
of alternative monitoring requirements and Maximum Contaminant Levels (MCLs)
under consideration for a forthcoming proposal of national primary drinking
water regulations on radionuclides under the Safe Drinking Water Act (SDWA).
Two other documents addressing the occurrence of man-made radionuclides
have previously been prepared for EPA's Office of Drinking Water. A document
was prepared by Peeters (1985) for the EPA Office of Drinking Water that
focused primarily on identifying the important sources that could potentially
introduce man-made radionuclides into public drinking water supplies. Four
major sources were discussed:
• DOE nuclear sites
t Commercial nuclear power plants
• Institutional sources (e.g., research facilities, hospitals,
universities)
• Industrial sources (e.g., pharmaceutical companies, commercial
analytical laboratories)
Atmospheric fallout, another frequently noted potential source of man-
made radionuclides, was also discussed briefly by Peeters. Atmospheric
fallout is not currently considered a major source because of the moratorium
begun in 1958 suspending atmospheric weapons testing worldwide.
Peeters (1985) provided an assessment of the potential for various
specific man-made radionuclides to be released from each of the four major
sources. That assessment included an estimate of the number of DOE facilities
that are potential sources and a listing of specific man-made radionuclides
that may be released from each facility. It was concluded in the report,
DRAFT II-l
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however, that it is not possible to make any generalizations concerning the
type of man-made radionuclides that may result as drinking water contaminants
from any of the four major sources because of the widely varying nature of
activities occurring at specific facilities in each category.
The second report prepared for the Office of Drinking Water focused on
available drinking water monitoring data on man-made radionuclides (Ellis et
al. 1986). As indicated in that report, the amount of actual monitoring data
available on man-made radionuclides in drinking water are very limited. The
primary source for such data is the series of "Environmental Radiation Data"
(ERD) quarterly reports prepared by EPA using data from the Environmental
Radiation Ambient Monitoring System (ERAMS). In addition to presenting the
available occurrence data on man-made radionuclides, Ellis et al. (1986)
provided some estimates of the potential radiation dose associated with those
substances in drinking water.
This document is primarily an update of the occurrence data summary
provided in the Ellis et al. (1986) report. Specifically, this report
presents a summary of the ERAMS data that have become available since the
Ellis et al. report was prepared. In addition, data from the National
Inorganics and Radionuclides Survey (NIRS) are presented; these data were not
available when the Peeters and Ellis reports were prepared.
B. CURRENT STANDARDS AND COMPLIANCE MONITORING REQUIREMENTS
To provide a context for evaluating the man-made radionuclide occurrence
information, it is useful to summarize the existing drinking water standards
and compliance monitoring requirements.
The category of substances referred to as man-made radionuclides
comprises more than 800 elements of which approximately 200 are considered to
be potential drinking water contaminants (Lowry and Lowry 1988). Man-made
radionuclides have been regulated under SDWA interim primary regulations since
1976 (40 CFR 141.16). Unlike most drinking water regulations in which a
drinking water concentration of each contaminant regulated is specified as the
MCL, the regulation for man-made radionuclides is a categorical standard that
DRAFT II-2
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seeks to limit the combined health risk of all man-made radionuclides that may
be present in a given water supply.
Under current regulations, the MCLs for man-made radionuclides are
defined in terms of "beta particle and photon radioactivity from man-made
radionuclides in community water systems." Specifically, they include all
radionuclides emitting beta particles and/or photons (except uranium-235 and
uranium-238) that are listed in the 1963 National Bureau of Standards (NBS)
Handbook 69 entitled, "Maximum Permissible Body Burden and Maximum Permissible
Concentration of Radionuclides in Air or Water for Occupational Exposures."
The MCL states that the average annual concentration of beta particle and
photon radioactivity from man-made radionuclides in drinking water is not to
produce an annual dose equivalent to the total body or any organ greater than
4 millirem (mrem) per year.1 The regulation specifies the drinking water
levels assumed to produce the 4 mrem dose for only two specific man-made
radionuclides, namely tritium and strontium-90; for all other man-made
radionuclides, the 4 mrem equivalent level is to be calculated from data on
dose to critical organs provided in the NBS Handbook.
Although the MCLs for man-made radionuclides apply to all public
community water supplies, only certain categories of supplies were
specifically required to monitor them. The monitoring requirements that
accompany the MCL for man-made radionuclides (40 CFR 141.26(b)) are depicted
in Figures II-l and II-2. Only systems using surface water sources and
serving more than 100,000 people are specifically designated as having to
monitor man-made radionuclides under the Federal standards. Those systems
must monitor gross beta activity, tritium and strontium-90, with more specific
monitoring to identify and quantify other man-made radionuclides "triggered"
if gross beta activity exceeds 50 pCi/L (see Figure II-l).
essentials of the radiochemistry of man-made radionuclides, the
practical utility of using gross beta as a screen or surrogate measure, and
the relationships between activity levels (measured in pCi/L) and dose levels
(measured in millirems per year) are discussed in detail in the 1986 ANPRM as
well as in other Agency documents, and are not repeated here.
DRAFT II-3
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In addition to the large surface water systems, the states were to have
identified other community water supplies "utilizing waters contaminated by
effluents from nuclear facilities." For those systems, gross beta, tritium,
strontium-90, and iodine-131 were to be measured, again with additional
monitoring requirements applying if certain levels of gross beta activity are
exceeded {see Figure II-2).
It should be noted that the man-made radionuclides interim regulations
were directed mainly at surface water supplies. Ground water has generally
been considered less likely to be vulnerable to man-made radionuclides because
it is not directly accessible to fallout. Long-lived isotopes, such as
strontium-90 and cesium-137, are tightly bound by soil, and short-lived
isotopes decay before the water is used (Lowry and Lowry 1988). However,
notwithstanding the implied emphasis on surface water systems, the interim
regulations do explicitly note that states have the discretion to require
monitoring for man-made radionuclides in supplies using only ground water
sources.
In 1986, EPA published an Advance Notice of Proposed Rulemaking (ANPRM)
addressing radionuclides (EPA 1986). Some changes to the man-made radionuclide
regulations were discussed in the ANPRM and public comment was solicited on
them. The notable changes discussed for man-made radionuclides included the
elimination of the reference to the specific NBS Handbook list of man-made
radionuclides as being too limiting, suggesting instead the more generalized
definition be used to allow for the potential development of new elements.
Also, it was indicated that three additional substances may be included as
man-made radionuclides. These are potassium-40 (a beta emitter that is
naturally occurring, but not part of the uranium or thorium decay series), and
plutonium-239 and americium-241 (both man-made radionuclides that are alpha
rather than beta emitters).
No changes were made in the use of the 4 mrem equivalent level as the
MCL for man-made radionuclides, although more recent data on the activity
levels equivalent to 4 mrem for specific substances were presented. Potential
changes to the monitoring requirements for man-made radionuclides were also
not discussed in the ANPRM.
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C. MAN-MADE RADIONUCLIDE OCCURRENCE DATA
As noted in the Introduction, the focus of this report is on the most
recent occurrence information on man-made radionuclides in drinking water
available from the ERAMS and NIRS. One other potential source of drinking
water occurrence data that merits mention is the compliance monitoring data
required to be collected under the existing standards. However, except for
showing that there have been no reports of gross beta activity levels
exceeding 50 pCi/L for surface water systems serving more than 100,000, the
available compliance monitoring data do not provide any specific information
on the levels of man-made radionuclides present in public water supplies.2
In general, available data on the occurrence of man-made radionuclides
in drinking water are very limited. As noted previously, the most important
source of relevant information is the ERAMS program, the data from which are
presented in the ERD reports published by EPA.
In addition to the ERAMS, there are recent data available from the
National Inorganics and Radionuclides Survey (NIRS). Although, as discussed
below, man-made radionuclides were not specifically included as analytes in
NIRS, gross beta activity levels were measured, and these data are presented
here.
Environmental Radiation Data Reports
The drinking water program of the ERAMS monitors ambient radiation
levels in drinking water grab samples from 78 sites that are either major
population centers or selected nuclear facility environs. These sites all
2There are approximately 275 surface water systems serving 100,000 or
more people required to perform compliance monitoring under the interim
regulations. An attempt was made to determine how many other water supplies
have been designated by the states as being affected by nuclear facilities.
No information on this is available at EPA. Based on a limited number of
contacts made at the state level, it appears that very few, if any, such water
supplies have been so designated.
DRAFT 11-5
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involve surface water sources. Data for a few other sites are also
occasionally reported as well.
The relevant man-made radionuclide analyses performed at the ERAMS
drinking water sites are:
• tritium on a quarterly basis
• gross beta, strontium-90 and gamma radiation as annual composites
• iodine-131 on one quarterly sample per year for each station
At the time that the Ellis et al. (1986) report was prepared, ERD
reports providing information on the occurrence of man-made radionuclides in
drinking water were only available through March 1985. Published ERD reports
are now available that cover the period through March 1989. Table II-l
presents a summary of the ERAMS data on "beta-emitting" man-made radionuclide
levels found in drinking water as reported in the ERD reports covering the
period from mid-1986 through March 1989. The following sections summarize the
ERAMS data by parameter measured.3
It should be noted that the data presented here for gross beta, tritium,
strontium-90 and iodine-131 taken from the more recent ERAMS data do not
differ markedly from the values reported in the earlier ERD reports and
summarized in Ellis et al. (1986).
GROSS BETA
Although there is currently no MCL per ie for gross beta, gross beta
activity measurements are used in the compliance monitoring scheme as a
3Note that some ERAMS data presented in these tables are shown as
negative values. As discussed in the ERD reports, negative values are a
result of the procedure for measuring radionuclides in which a previously
determined background level is subtracted from a measured sample value that
happens to fall below the background level. Though having no real physical
meaning, these negative values when taken together with all other observations
can be used to describe the overall distribution of activity levels. Also,
these values allow for better evaluations of trends in the data and facilitate
estimates of biases in the nuclide analytical methods.
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Table II I Summary "I Recent KItAMS Data on llcla Emitting Man-Made Uadiomiclides in Drinking Water"
KRAMS Site
AK: Pairbanks
Al.: Dolhan
Al.: Montgomery
Al.: Muscle Shoals
Al.: Scoltsboro
Alt: I.illle Itock
CA: Uerkelcy
CA: l.os Angeles
CO: Denver
CO: I'lalleville
CT: llarllord
DC: Washington
DI-: Dover
DE: Wilmington
PL: Miami
PL: Tampa
CiA: llaxley
CiA: Savannah
III: Honolulu
IA: Cedar Rapids
ID: Hoise
ID: Idaho Palls
II.: Morris
Gross Belab
Min Max Avg
2.2 3.0 2.7
1.6 3.0 2.2
1.6 1.7 1.6
0.7 2.7 1.5
2.1 2.K 23
O.I 1.9 I.I
0.7 1.5 1.1
0.0 6.4 3.8
1.2 2.0 1.7
6.5 7.6 6.9
0.5 1.5 1.1
2.0 3.0 2.6
3.5 5.4 4.3
0.0 2.9 1.5
0.9 2.1 1.5
1.8 4.8 3.2
1.8 1.8 1.8
1.2 2.7 1.7
1.2 2.1 1.6
2.8 3.3 3.0
0.4 1.7 1.0
2.6 4.7 3.9
16.1 17.2 16.6
Tritium0
Min Max Avg
0 300 180
100 300 150
100 300 150
2(K) 400 270
200 400 240
100 200 170
100 300 170
100 200 150
100 300 170
100 400 240
100 300 160
100 200 180
100 200 130
100 400 190
100 300 140
100 200 150
100 2500 1120
100 200 170
100 400 210
100 300 140
100 300 160
100 200 150
lodinc-131d
Min Max Avg
O.I 0.0 -0.1-
-0.2 0.1 -0.0
0.0 0.2 0.1
0.1 0.0 -0.1
0.2 0.3 0.1
O.I 0.1 0.1
-0.5 O.I -0.2
-0.1 0.1 0.0
-0.1 -0.1 -0.2
0.1 O.I 0.0
0.0 0.3 0.1
O.I 0.2 0.2
-0.3 0.1 -O.I
0.2 0.0 -O.I
-0.1 0.0 -0.1
2.5 0.0 -1.3
0.0 0.2 0.1
-0.3 0.0 -0.2
-0.2 0.1 0.0
-0.1 0.2 0.0
-0.4 0.3 -0.0
-0.1 0.2 0.1
Slrontium-90e
Min Max Avg
-0.2 0.2 0.0
-0.1 0.1 0.0
-0.2 0.3 O.I
-0.1 0.3 0.1
0.0 0.2 0.1
0.0 0.2 O.I
-0.2 0.2 0.0
-0.4 0.2 -0.0
-0.3 0.2 0.0
0.1 0.3 0.2
-0.1 O.I 0.0
O.I O.I 0.1
0.0 0.1 0.1
0.3 0.3 0.3
0.0 0.1 0.0
0.0 0.3 0.2
-0.1 -O.I -0.1
0.0 0.6 0.2
-0.1 0.2 0.0
-0.3 O.I -O.I
0.0 0.3 O.I
-0.3 0.3 0.0
-0.2 0.3 0.1
-------
Table ll-l. (Continued)
CRAMS Site
IL: W. Chicago
K.S: Topeka
I.A: Nc\v Orleans
MA: 1 jiwrcnce
MA: Uowe
Ml): liallimore
MO: Conowingo
ME: Augusta
Ml: Detroit
Ml: Grand Rapids
MN: Minneapolis
MN: Ked Wing
MO: Jefferson City
MS: Jackson
MS: I'ort Gibson
MT: Helena
NC: Charlotte
NC: Wilmington
ND: liismnrck
NT,: Lincoln
Nil: Concord
N.I: Trenton
N.I: Warelown
Gross Helab
Min Max Avg
14.9 22.9 17.8
5.5 7.5 6.4
2.4 3.7 3.0
1.0 2.1 1.4
0.3 0.7 0.5
1.2 3.4 2.2
1.2 2.6 1.9
O.I 1.1 0.6
1.7 1.9 1.8
2.3 2.5 2.4
1.8 2.8 2.3
5.3 7.4 6.2
3.7 7.1 5.4
2.1 4.2 3.1
3.6 7.0 5.1
2.5 4.2 3.2
0.6 1.9 1.4
2.2 2.<> 2.4
2.4 4.7 3.3
10.4 11.7 11. 1
0.2 0.9 0.6
0.6 1.3 1.0
0.1 1.5 0.9
Triliumc
Min Max Avg
100 300 150
100 400 180
100 300 220
100 300 160
100 200 150
100 400 190
100 500 250
100 400 ' 210
200 400 260
200 500 250
200 300 220
100 200 150
100 200 150
100 200 170
100 300 150
100 300 240
300 1000 640
100 300 180
100 300 190
100 300 200
100 300 160
100 300 160
1(K) 200 120
Iodine-I31d
Min Max Avg
0.0 0.0 0.0
0.0 0.1 0.1
0.4 0.0 -0.2
0.0 0.3 0.2
0. 1 0.2 0.2
0.1 0.3 0.2
0.0 O.I 0.1
-0.2 0.1 -0.0
-0.2 0.1 -0.1
O.I 0.2 0.1
0.0 0.2 0.1
-O.I 0.1 0.0
0.3 0.3 0.3
0.0 0.3 0.1
0.1 0.2 0.0
-0.2 0.1 0.0
0.4 0.1 -0.1
-O.I 0.0 -0.1
0.0 0.2 0.1
0.1 0.2 0.1
-0.4 0.2 -0.1
O.I 0.3 0.2
0.0 0.1 0.0
Slronlium-90e
Min Max Avg
-0.2 O.I 0.0
-0.3 0.0 -0.2
0. 1 0.2 0. 1
-0.2 O.I -0.0
-0.5 O.I -O.I
-0.2 0.2 -0.0
-O.I 0.2 0.1
-0.1 0.1 0.0
0.5 0.8 0.7
-0.1 0.6 0.4
-0.4 0.1 -0.1
-0.6 0.4 -0.1
0.1 0.1 O.I
0.3 0.5 0.4
-0.2 O.I -0.1
-0.2 0.4 O.I
-0.5 0.2 -0. 1
0.0 0.1 O.I
-O.I 0.1 0.0
O.I 0.3 0.2
-0.3 O.I -O.I
-0.3 0.4 O.I
-0.2 0.1 0.0
-------
Table ll-l. (Continued)
liKAMS Site
NM: S;int;i Te
NV: l,as Vegas
NY: Albany
NY: New York Cily
NY: Niagara Palls
NY: Syracuse
Oil: Cincinnati
Ol 1: Columbus
Oil: liasi Liverpool
Oil: I'ainesville
Oil: Toledo
OK: Oklahoma Cily
OR: Portland
1'A: Columbia
I'A: 1 larrisburg
I'A: I'illsburgli
1'R: Alteon
I'K: Coro/al
I'K: Cristobal
Kl: Providence
SC: llarnwell
SC: llnrnwell
S(.': Columbia
Gross »elab
Min Max Avg
1.9 7.6 4.';
3.0 6.7 5.4
0.1 1.3 0.8
0.6 1 . 1 (I.1;
I.I 2.5 1.6
1.6 2.4 2.0
2.2 3.1 2.5
3.4 18.2 8.4
1.3 3.1 2.3
1.6 4.1 2.7
1.7 2.8 2.2
2.1 2.7 2.3
0.6 0.9 0.7
1.2 2.9 2.2
0.6 1.2 0.9
1.3 3.4 2.1
0.1 0.8 0.6
1.0 1.5 1.2
0.4 1.5 1.1
2.2 2.2 2.2
2.0 2.0 2.0
Tritium'
Min Max Avg
100 200 120
100 200 180
100 300 160
100 300 170
100 300 180
KM) 400 210
100 300 200
100 300 190
KM) 200 150
200 300 210
100 300 210
100 200 180
100 200 150
100 300 180
100 300 180
100 300 170
100 200 130
400 400 400
100 100 100
100 200 140
100 200 150
200 600 370
lodine-131d
Min Max Avg
0.2 0.0 -0.1
-0.3 0.0 -0.1
-0.2 0.2 0.0
-O.I 0.2 0.0
0.0 0.2 0.1
00 O.I O.I
-0.2 0.4 0.2
0.0 0.1 0.0
0.0 0.4 0.2
-0.6 0.1 -0.2
0.0 0.1 0.1
-0.3 0.2 -0.1
0.0 0.3 0.1
O.I 0.3 0.2
-0.2 0.1 -0.0
-0.1 0.1 0.0
-O.I 0.3 O.I
-O.I -0.1 -0.1
-0.3 O.I -0.1
0.0 0.1 O.I
0.2 0.2 0.2
-0.2 0.1 -0.0
Stronlium-90c
Min Max Avg
-0.2 0.0 -O.I
0.1 0.6 0.4
-0.4 0.3 -0.0
0.1 0.3 0.2
O.I 0.7 0.4
0.2 0.6 0.4
0.0 0.3 0.1
-O.I O.I 0.0
-0.7 0.4 0.0
-0.2 0.9 0.3
0.1 0.6 0.3
-0.2 0.2 0.1
-0.3 0.2 0.0
0.1 0.7 0.3
0.0 0.2 0.1
0.0 0.5 0.2
-0.8 0.2 -0.2
-0.3 0.3 O.I
-1.3 0.0 -0.4
0.3 0.3 0.3
-0.4 0.3 -O.I
-------
Table II-1. (Continued)
ERAMS Site
SC: llarlsville
SC: Jcnkinsvillc
SC: Seneca
TN: Chattanooga
TN: Knoxvillc
'1'X: Austin
VA: Doswcll
VA: l.ynthburg
VA: Virginia Reach
VI: St. Thomas
WA. Kichland
WA: Seattle
Wl: Genoa City
Wl: Madison
WV: Wheeling
Gross Hciab
Min Max Avg
1.2 1.5 1.3
4.3 6.1 5.3
-1.2 1.2 0.3
2.0 3.9 2.8
1.5 2.7 1.9
2.8 4.4 3.5
4.3 6.1 5.0
0.6 1.5 1.0
2.6 3.2 3.0
O.I 0.7 0.3
1.3 1.9 1.6
0.5 1.2 0.8
0.8 1.7 1.2
1.0 1.2 1.1
Tritium0
Min Max Avg
100 200 140
100 300 190
100 300 200
100 500 300
100 300 180
100 400 190
100 300 190
100 300 170
100 200 130
100 300 150
100 600 250
100 200 130
100 200 130
100 200 150
100 100 100
lodine-131d
Min Max Avg
-0.2 0.1 -0.0
-0.1 0.0 -0.0
-0.2 0.2 0.0
0.0 0.4 0.2
•0.4 0.2 -0.0
-0.3 0.1 -O.I
-0.1 0.1 -0.0
-0.4 O.I -0.1
-O.I 0.4 0.1
-O.I -0.1 -0.1
-0.1 0.4 0.1
O.I 0.3 0.2
-0.3 0.0 -0.2
0.0 0.2 O.I
Slrontium-90e
Min Max Avg
-0.5 0.1 -O.I
-0.7 0.1 -0.2
-0.6 0.5 0.1
0.1 0.4 0.2
0.1 0.4 0.2
-0.7 0.2 -0.2
-0.3 O.I -O.I
-0.3 0.2 0.0
0.2 0.6 0.4
-0.6 0.2 -0.1
-0.4 0.2 -0.1
-0.4 0.4 0.0
-0.1 0.0 -0.0
-0.3 0.4 0.0
a All values in pCi/L; see Footnote 3 in text regarding negative values.
h Gross beta activity data are for annual composites 1985-1987 from EKD Reports 47, 51 and 54.
c Tritium activity data are for quarterly samples from ERD Reports 47 (July-Sept. 1986) through 57 (Jan.-March 1989).
d lodine-131 activity data are for one quarterly sample per year from ERD Reports 48 (1985), 53 (1987) and 57 (1988).
c Stronlium-90 activity data are for annual composites 1985-1987 from ERD Reports 47, 51 and 54.
-------
screening device to determine whether a water supply should conduct more
detailed monitoring for specific man-made radionuclides. There are two gross
beta "trigger values" in effect: 50 and 15 pCi/L. Exceedance of the 50 pCi/L
level results in the requirement to conduct more comprehensive analyses to
identify the specific man-made radionuclides present and to determine the
annual dose from all such substances found for comparison with the 4 mrem/year
MCL. The second screening value of 15 pCi/L is included only in the
monitoring scheme for those water supplies specifically designated as having
source water potentially contaminated by effluents from nuclear facilities.
Exceedance of the 15 pCi/L level in that scheme requires that additional
monitoring for strontium-89 and cesium-134 be conducted.
The gross beta activity levels shown in Table II-l reflect the averages
of annual composites taken at the indicated sites for 1985 through 1987. In
almost all cases, data were provided for all 3 years, although in a few cases
results were provided for only 1 or 2 years. As indicated by these data, the
average gross beta levels at these sites ranged from 0.3 to 17.8 pCi/L, with
average values generally falling below 3 pCi/L.
There were no instances in which the gross beta levels during this
period exceeded 50 pCi/L at any of these sites; however, there were a few
instances in which gross beta activity exceeded the 15 pCi/L value, most
notably at the Morris and West Chicago, Illinois sites where the gross beta
activities for the 3-year period ranged from 14.9 to 22.9 pCi/L. No
information was provided in the ERD reports on these sites to explain the
relatively high gross beta levels. However, a contact with the Illinois
Department of Nuclear Safety revealed that those relatively high gross beta
activity levels are apparently due to naturally occurring rather than man-made
radionuclides; however, the specific nature of the naturally occurring sources
was not specified.
DRAFT 11-8
-------
TRITIUM4
The current drinking water regulations specify a tritium level of 20,000
pCi/L as the amount assumed to produce a total body/organ dose of 4 mrem/year.
It should be noted that the Proposed Rule Making (PRM) shows a substantially
higher level of 60,000 pCi/L to be the 4 mrem/year equivalent.
The available ERAMS data show that tritium levels are very low by
comparison to the 4 mrem/year equivalent levels. As shown in Table II-l, the
range of the quarterly tritium levels reported between 1985 and 1987 was 0 to
2,500 pCi/L, with average values generally falling in the 100 to 300 pCi/L
range.
STRONTIUM-90
The current drinking water regulations specify a strontium-90 level of 8
pCi/L as the amount assumed to produce a total body/organ dose of 4 mrem/year;
the ANPRM shows levels of 50 and 700 pCi/L as the 4 mrem/year equivalent. As
shown in Table II-l, annual composite samples for strontium-90 for 1985 to
1987 do not approach any of those values, ranging from -1.3 to 0.9 pCi/L, with
typical average values falling below 0.2 pCi/L.
IODINE-131
Although not having a 4 mrem per year equivalent level specified in the
current drinking water regulations as do tritium and strontium-90, the
compliance monitoring scheme indicates that an iodine-131 level of 3 pCi/L is
the MCL compliance level (presumably derived from the NBS Handbook); the ANPRM
indicates that 700 pCi/L is the 4 mrem/year equivalent. As shown in Table
II-l, the annual composite measurements for iodine-131 are far below these
levels, with a range of -2.5 to 0.4 pCi/L, and typical average values falling
below 0.1 pCi/L (including many that are negative values).
4Note that, although tritium is a beta emitter, tritium levels are not
reflected in the gross beta analyses because the calibration procedure for
gross beta analyses precludes detecting the low energy beta emissions from
tritium.
II-9
-------
GAMMA RADIATION
The interim drinking water standards include "photon radioactivity" in
the definition of the MCL for man-made radionuclides. Gamma radiation, which
provides the measure of photon radioactivity, is included as a parameter in
the annual composites at ERAMS sites. For 1985 to 1987, all sites were
reported as having nondetectable (ND) levels for gamma radiation; however, no
information was given in the ERD reports on the detection limit for gamma
radiation in drinking water samples.
National Inorganics and Radionuclides Survey (NIRS)
As noted in the Introduction, the NIRS results were not available at the
time Ellis et al. (1986) was prepared. NIRS was designed to provide data to
support the development of drinking water regulations for a wide range of
inorganic and radionuclide contaminants. With regard to radionuclides,
however, the survey focused primarily on certain naturally-occurring
radionuclides (radon, radium-226, radium-228, and uranium). In addition, the
NIRS sample sites (approximately 1,000) were exclusively ground water supplies
which, as noted previously, are generally not considered to be at high risk of
contamination from man-made radionuclides. Although there were no analyses
conducted for specific man-made radionuclides in NIRS, gross beta activity
analyses were performed. The results of those analyses are summarized below.
There were 990 sites in NIRS for which gross beta analyses were
reported. Of these, 440 had levels above the minimum reporting level of 2.3
pCi/L. There were nine sites (0.9%) reporting gross beta activity levels
above 50 pCi/L (maximum of 94 pCi/L), and 102 sites (10.3%) with levels above
15 pCi/L. The mean gross beta activity for all sample sites (assuming a value
equal to one-half the minimum reporting level for the "non-detects") was
approximately 6 pCi/L.
Longtin (1988a) conducted a "beta activity balance" analysis to
determine the contribution of radium-228, the only beta emitter specifically
measured in NIRS, to the total gross beta levels observed. Finding that the
bulk of the gross beta activity was not accounted for by radium-228, Longtin
DRAFT 11-10
-------
(1988b) subsequently expanded the activity balance analysis to take into
consideration potential contributions of potassium-40 and rubidium-87, two
important naturally-occurring, non-series beta emitters. Based on measured
potassium levels in the NIRS samples, and using the known relative isotopic
abundance of potassium-40, Longtin concluded that potassium-40 would generally
contribute more to the gross beta activity in the NIRS samples than radium-
228. Based on crustal abundance considerations, Longtin concluded that
rubidium-87 would contribute only about an order of magnitude less activity
than potassium-40. Even with those two sources considered along with radium-
228, however, there remained a considerable "deficit" of beta activity in the
NIRS samples (generally ranging from 50 to 90 percent of the gross beta
activity) that could not be accounted for by these specific substances.
Longtin concluded that the unaccounted for beta activity was most probably due
to beta emitting decay products of the uranium and thorium series and/or
errors in the gross beta analyses. No suggestion of contributions by man-made
radionuclides was made by Longtin.
D. CONCLUSIONS
The absence of any violations of the current interim drinking water
regulations reported in the compliance monitoring data, together with the
consistently low levels of gross beta and specific man-made radionuclide
levels reported for the ERAMS sites, strongly suggest that there is little or
no occurrence of man-made radionuclides in surface water based drinking water
supplies at levels of concern relative to current MCL values. The limited
data for ground water supplies are not sufficient for drawing conclusions
about man-made radionuclide occurrence, although other considerations suggest
that contamination is unlikely.
It also appears that, with the current monitoring schemes, very few if
any public water supplies using surface water would find sufficiently high
levels of gross beta activity to require more specific monitoring. The NIRS
data, however, suggest that current screening levels for gross beta activity
of 50 pCi/L could, if applied to all ground water systems, result in about
1 percent of them being required to conduct further monitoring; about
DRAFT 11-11
-------
10 percent of ground water systems may exceed a gross beta activity screen of
15 pCi/L.
It must be recognized, however, that the available data upon which these
conclusions are based are extremely limited.
DRAFT 11-12
-------
III. PROFILES OF REPRESENTATIVE BETA OR GAMMA EMITTING RADIONUCLIDES
A. INTRODUCTION
In the following sections profiles on strontium, iodine, and tritium are
provided. For each radionuclide, the profile includes a discussion of
physical and chemical properties, toxicokinetics, health effects in animals,
and health effects in humans. Discussions of the toxicokinetics include
absorption, distribution, bioaccumulation and retention, and excretion.
Health effects sections are organized in the following fashion: first by
route of exposure (oral, inhalation, parenteral), toxicological endpoint, and
species. For all of these radionuclides, data on toxicokinetics and health
effects following oral exposure are available. These data are certainly more
applicable than data from other routes of exposure for estimating risks from
ingestion of radionuclides in drinking water. However, since health effects
by other routes are generally similar, information on toxicokinetics and
health effects from exposure by other routes is also presented.
The relevance of information on other routes of exposure to estimates of
potential health concerns from the radionuclides in drinking water is
dependent upon whether the pattern of distribution observed after exposure by
that route is similar to that observed after oral exposure. For example,
following oral exposure to strontium, distribution of the absorbed amount in
laboratory animals was primarily to the skeleton followed by the liver and
other soft tissues. Data on distribution kinetics of strontium following
parenteral exposure were not located; however, at long times following single
injections of high doses of strontium, hematological effects (indicating
damage to bone marrow) and other skeletal effects (such as fractures, bone
marrow hypoplasia, and bone tumors) have been observed. These effects are
similar to those observed following high oral doses of strontium.
When compared to an oral dose of the same radiological activity,
intravenous exposure delivers a much higher dose of strontium to these target
tissues resulting in effects much more severe than should be expected
following oral exposure. But this example illustrates that the main target
tissues are the same for the two routes of exposure, and, by. use of data on
DRAFT III-l
-------
absorption of strontium from the gastrointestinal tract, relative
distribution, dose equivalent, and cancer risks to these tissues may be
predicted.
Use of data on toxicokinetics and health effects following inhalation
exposure is more complex due to the fact that distribution of beta/gamma
emitting radionuclides from the lung is largely dependent on the isotopic form
and solubility of the particular form administered. However, depending upon
the particular isotope, distribution to other tissues than the lung, such as
the liver and skeleton have been reported, as well as cancer in each of these
tissues.
Therefore, even for inhalation and injection routes, data on health
effects may be relevant when data on distribution, when compared to
distribution following oral exposure, are similar.
B. STRONTIUM
Physical and Chemical Properties
Chemical Properties - Elemental
Strontium is an alkaline earth element usually found as celestite or
strontianite. The metal may be prepared by reducing strontium oxide or by
electrolysis of the fused chloride mixed with potassium chloride. Three
allotropic forms of strontium are known to exist with transition points at
235°C and 540°C. Freshly cut strontium has a silvery appearance that quickly
changes to a yellowish color upon oxidation. Finely divided strontium is
pyrophoric (Weast 1981).
Strontium only forms divalent compounds, and the solubility of its
compounds are usually greater than those of barium (Carson et al. 1986).
Strontium chemistry resembles that of calcium; however, strontium has greater
base-forming characteristics than calcium but less than those of barium
(Hampel 1968). Because strontium is a reactive metal, exposure to air results
in an oxide coating which quickly covers the metal. Two strontium oxides
DRAFT III-2
-------
known to exist are strontium oxide and the peroxide, strontium dioxide.
Strontium reduces halides and oxides of many metals at elevated temperatures,
thereby producing the corresponding metal (Hampel 1968). Strontium is an
active reducing agent that reacts violently with water to liberate hydrogen
and form strontium hydroxide. Strontium reacted with acids forms hydrogen and
the strontium salt of the acid. Strontium will burn when heated in air,
oxygen, chlorine, bromine gas, and sulfur, thus producing a bright crimson
color. Strontium forms the nitride only in nitrogen at temperatures greater
than 380°C (Hampel 1968).
Strontium compounds are analogous to the corresponding calcium
compounds. Consequently, the strontium compounds of sulfide, chloride,
bromide, iodide, nitrate, etc. are soluble, whereas the carbonate, fluoride,
sulfate, oxalate and phosphate compounds are not. The only major difference
occurring between the solubility of strontium compounds and calcium compounds
is that strontium hydroxide is very soluble in hot water (100°C) and calcium
hydroxide is not (Hampel 1968).
Physical Properties
Strontium (atomic number 38, atomic weight 87.62) is an alkaline earth
element located between calcium and barium in Group IIA of the Periodic Table
and its physical and chemical properties resemble both elements. Strontium is
a hard, silver-white metal softer than calcium, ductile and malleable, and
capable of being formed into wire. Strontium has a melting and boiling point
of 769°C and 1384°C, respectively, and its specific gravity is 2.54 (Weast
1981). Some chemical and physical properties of elemental strontium are
listed in Table III-l, adapted from Faure and Powell (1972).
DRAFT 111-3
-------
Table III-l. Physical and Chemical Properties of Strontium
Properties
Atomic number 38
Atomic weight (based on 12C) 87.62
Ionic radium (A) (Pauling) 1.13
Radius ratio (O2" = 1.40 A) 0.81
Electronegativity (Pauling) KO
Strontium occurs naturally as a mixture of four stable isotopes:
strontium-84 (0.56 percent), strontium-86 (9.86 percent), strontium-87
(7.02 percent), and strontium-88 (82.56 percent). In addition, 12 unstable
radioactive isotopes are known to exist, the most important being strontium-90
(half-life of 28 years) (Weast 1981) because of its presence in radioactive
fallout (Hampel 1968), and strontium-89 (half-life of 50 days). Strontium-89
and strontium-90 emit beta energies of 1.5 MeV and 0.61 MeV, respectively
(Hampel 1968). Rubidium-89 decays to strontium-89 which subsequently decays
to the stable isotope of yttrium-89. Strontium-90, a decay product of
rubidium-90, decays to yttrium-90 which in turn decays to stable zirconium-90
(Brucer 1979). Half-lives and decay schemes of strontium-89 and -90 may be
seen in Figure III-l, adapted from Brucer (1979).
Summary
Strontium is a divalent alkaline earth element (Group IIA of the
Periodic Table) with physical and chemical properties similar to those of
calcium and barium. It is a silver-white metal, ductile and malleable, and
readily oxidizes in air. Strontium is an active reducing agent and reduces
halides and oxides of many metals at elevated temperatures, consequently
producing the corresponding metal. Strontium compounds are analogous to the
corresponding calcium compounds and include the soluble sulfide, chloride,
bromide, iodide, and nitrate compounds and the insoluble carbonate, fluoride,
sulfate, oxalate, and phosphate compounds.
Strontium occurs naturally as a mixture of four stable isotopes:
strontium-84, strontium-86, strontium-87, and strontium-88. •Twelve
DRAFT III-4
-------
Figure 111-1. Half-Lives and Decay Schemes
of Strontium-89 and -90
Sr -89 (a)
50.5 d (b)
B" = 1.489
Y-89
(Stable)
Sr-90
29.12 y
B~ = .546
Y -90 m -
3.19 h
B- = .480
* Y-90
2.67 d
B'= .2779
Zr-90
(Stable)
a = half-life
b = maximum B' decay energy in MeV
m = metastable
Brucer 1979
ICRP1983
DRAFT
III-5
-------
radioactive isotopes of strontium are known to exist, the most important being
strontium-90 because of its presence in radioactive fallout and the fact that
it has the longest half-life (28 years) of the artificially produced strontium
isotopes. Strontium-89 is another important radioactive isotope and has a
half-life of 50 days. The beta energies emitted by strontium-90 (decays to
xenon-90) and strontium-89 (decays to xenon-89) are 0.61 and 1.5 MeV,
respectively.
Toxicokinetics
Absorption
In this analysis strontium, unless otherwise noted, refers to strontium-
89 or -90. Human gastrointestinal absorption values ranging from 15 to 36
percent have been suggested (Stara et al. 1971, Spencer et al. 1972a). ICRP
(1979) lists 30 percent as the accepted value. Strontium inhaled as strontium
chloride is rapidly absorbed into the bloodstream (McClellan et al. 1972).
Net strontium absorption in young adults is similar to that found in older
persons (Spencer et al. 1972b). Gastrointestinal absorption has been
determined to be 11 percent for dogs, 26 percent for cats, 16 percent for
monkeys, and 15 percent for rats (Stara et al. 1971). Spencer et al. (1972b)
noted that calcium alone and in conjunction with phosphorous decreased
strontium absorption in rats but, for reasons unknown, no such effect occurred
in humans.
Distribution
Of the absorbed amount, initial fractional uptake to bone in humans is
reported to be 27 percent, while the reported uptake to other tissue is 73
percent (ICRP 1979). Final strontium distribution in adults is primarily to
the skeleton (uniformly to within 10 percent) due to the fact that strontium
can substitute metabolically for calcium, though no time frame is given for
this or for the ICRP data (Stara et al. 1971, Kulp et al. 1959). Strontium
distribution in the skeleton is relatively uniform throughout gestation and in
young children (Kawamura et al. 1986, Kulp et al. 1959). Since strontium is
DRAFT III-6
-------
known to be a bone-seeker, 98 percent of the associated risk of cancer is from
the irradiation of the skeleton and bone marrow (Thorne and Vennart 1976).
A study of strontium chloride in 452 beagle dogs exposed orally
beginning in utero (by oral exposure of the dam) and continuing to age 540
days found strontium distribution to be dependent on local bone turnover and
the beagle's age (Momeni et al. 1976). Deposition was primarily to newly
forming bone surfaces or other areas of active bone formation (Stara et al.
1971). Distribution was relatively uniform to all areas of the skeleton at
the end of the dosing period but, over time, became variable due to the uneven
rate of bone turnover (Pool et al. 1973). Age of the subject was a factor
because younger organisms have more areas of active bone formation than mature
individuals. During the feeding period strontium uptake to bone, plasma,
gastrointestinal tract, and to soft tissues was uniform. Thirty days after
the end of the feeding period in these beagles 78.5 percent of skeletal burden
of strontium was deposited in cortical (compact) bone, with a retention half-
life of 10.3 years, while the remainder was deposited in trabecular (spongy)
bone, with a retention half-life of 0.45 years (Momeni et al. 1976). The
gastrointestinal tract, plasma, and soft tissues retained a low level of
radioactivity (10 percent) at one year post-exposure.
Strontium inhaled as strontium chloride was distributed primarily to the
skeleton. By 8 hours after exposure, 85 percent of the absorbed strontium was
found in the skeleton, while only 2.5 percent remained in the lungs (Gillett
et al. 1987, McClellan et al. 1972). In beagles aged 1.5 years exposed via
single injection, 27 percent of the injected dose was retained after 30 days;
of this amount 46 percent had a retention half-life of 0.45 years (because of
its deposition in spongy bone) while the remainder (54 percent) had a
retention half-life of 7.6 years (because of its deposition in cortical bone)
(Momeni et al. 1976). The rate of clearance from the skeleton showed no
relation to dose (Parks 1991).
Excretion
Although little data are available regarding human excretion of
strontium, one relevant study was located. Warren and Spencer (1978) exposed
DRAFT III-7
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ten humans to 3.3 pCi strontium/day via the diet, and reported that about 84
percent of the intake was excreted via the feces, while 14 percent was
excreted in the urine.
Bioaccumulation and Excretion
See Section E. In this section toxicokinetic models for beta or gamma
emitting radionuclides are used to predict the potential bioaccumulation and
retention for these radionuclides. Some of the metabolic models are based on
extensive human and animal data, some on primarily animal data and some on
similarity of elemental characteristics. The model for strontium may be found
on page 111-59.
Summary
Human absorption of strontium was reported to be between 15 and 36
percent, though 30 percent was the accepted value. In humans, initial
distribution to bone was 27 percent of the absorbed amount, with uniform
distribution to all areas of the skeleton, while final distribution was
primarily to the skeleton. The distribution varied within the skeleton due to
the uneven rate of bone turnover. Excretion of strontium was mainly via the
feces (84 percent), with 14 percent excreted in the urine.
Health Effects in Animals
Short-Term Exposure
Short-term exposure to strontium has been associated with hematopoietic
changes, bone marrow hypoplasia, and liver lesions. A study of 44 cats
exposed by gavage to 25, 50, or 100 (iCi strontium/day for a month reported
dose-related mortality, hemorrhaging, depression of platelet counts, and bone
marrow lesions, and nondose-related bone marrow hypoplasia and depression of
lymphocyte and neutrophil counts (Nelson et al. 1972). Hemorrhaging was also
reported in pigs, dogs, mice, and cattle exposed to strontium (Ward and Wright
1972). In a group of seven adult monkeys exposed orally to a single dose of
500 or 1,000 jiCi strontium, one monkey exposed to 1,000 nCi.died at 4 months
DRAFT 111-8
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post-exposure with pancytopenia, i.e., a decrease in all cellular components
of blood, while others developed tumors and leukemia (discussed in the section
on carcinogenicity) (Mays and Lloyd 1972).
Few inhalation studies were located in which the soluble form of
strontium was used. In one such study, 71 beagle dogs were exposed to a
single inhalation of strontium chloride resulting in initial body burdens of
2.5 to 250 pCi strontium/kg (Snipes et al . 1981). Atrophy of the liver and
nonneoplastic liver lesions were reported (Snipes et al . 1981). A related
study exposed dogs to 0.999 to 118.8 pCi strontium chloride/kg, and reported
the development of dose-dependent pancytopenia (Gillett et al . 1987).
Beagles intravenously injected with 32.4 to 105.3 nCi strontium/kg
developed depressed platelet, lymphocyte, and neutrophil counts, and anemia by
2 to 5 weeks after injection (Gillett et al . 1987). Deaths due to bone marrow
hypoplasia were reported in the 64.8 and 105.3 jiCi treatment groups. In a
further study, Taylor et al . (1966) determined that, in contrast to the
behavior of other radionucl ides, a single injection of 100 ^Ci strontium/kg
seldom produced fractures in beagle dogs at intervals up to 3,400 days post-
exposure. Finkel et al . (1972), however, found that a series of daily
injections of 5.8 jiCi of strontium given over 90 days to beagle dogs beginning
at 4 to 8 days postpartum resulted in many bone fractures.
The injection of strontium in mice caused prolonged hypoplasia of bone
marrow and impaired marrow hematopoiesis, resulting in a decrease in
lymphocyte counts (Ito et al . 1976). A high rate of cell death was noted in
lymphocyte populations, as would be expected based on the radiosensitivity of
these cells (Stevenson et al . 1982). Further, Monig et al . (1980) noted that
a single injection of strontium resulted in the disturbance of iron
incorporation into the peripheral red blood cells of mice. The incorporation
of iron reflects the level of red blood cell production. Red blood cell
production decreased initially but later increased to slightly above control
values.
DRAFT III-9
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Lonqer-Term Exposure
Longer-term exposure to strontium has been associated with hematopoietic
changes, bone marrow hypoplasia, excess growth of bone marrow, and a decrease
in immune system function. Shubik et al. (1978) exposed rats to drinking
water containing various concentrations of strontium or strontium chloride in
an effort to study immune system response. Rats exposed to 10,000 times the
average annual permissible concentration (AAPC = 1.2xlO"8 Ci/1) of strontium
showed the inhibition of nonspecific protection factors (including humoral
factors such as lysozyme activity and cellular factors such as phagocytic
activity of neutrophils), while 1,000 times the AAPC resulted in increased
production of anti-tissue autoantibodies (which attack the organism's own
tissue). These changes may affect the ability of the rat's immune system to
function properly, possibly allowing the development of malignant neoplasms
and infectious complications (Shubik et al. 1978). Studies using beagle dogs
orally exposed to strontium beginning j_n utero. via exposure of the dam,
continuing in offspring to age 19 months at 0.081 to 118.8 jiCi/day have
revealed the development of dose-related myeloproliferative disease (Gillett
et al. 1987). A gradual and persistent decrease in leukocyte count developed
due to a decrease in neutrophils during the first 1.5 years of exposure. In
animals in the high-dose group, the maximum depression of neutrophils was 47
percent (Gillett et al. 1987).
An ingestion study performed over the lifetime of an unknown number of
swine, at 0.999 to 2,970 yCi strontium/day, resulted in the development of
bone marrow hypoplasia and myeloproliferative disease (Gillett et al. 1987).
Ragan et al. (1972) exposed more than 700 swine to 1 to 3,100 jiCi
strontium/day over three generations and noted the development of a dose-
related decrease in platelet, neutrophil, and lymphocyte counts, in addition
to the development of myeloproliferative disease.
Reproductive Effects
No effects on mortality, reproductive fitness, or survival have been
noted at oral or injected doses up to and including those subacutely lethal to
the mother (Goldman and Bustad 1972, Clarke et al. 1972). Goldman and Bustad
DRAFT 111-10
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(1972) indicated that placenta! and mammary barriers reduce strontium
concentration in the offspring, relative to the maternal concentration.
Following maternal ingestion the placenta discriminates against the movement
of strontium across the placenta, relative to calcium reducing the maternal-
fetal ratio by a factor of 5 (Delia Rosa et al. 1972).
Mutaoenicity
Ito et al. (1976) studied mice exposed via injection to strontium and
detected chromosomal aberrations in lymph nodes and bone marrow. Chinese
hamsters injected with 1 nCi of strontium also incurred an increased frequency
of chromosomal aberrations of bone marrow (Volf 1972).
Carcinoqenicity
In addition to the effects already discussed, exposure to strontium
resulted in tumor development and leukemia in animals. Wright et al. (1972)
studied the effects of strontium ingested by mice for a period of 180 days,
beginning at conception or weaning. Mice were exposed to 20 or 60 yCi
strontium chloride/g of dietary calcium/day. Hematopoietic neoplasms, mainly
lymphatic leukemia, and bone tumors, primarily of the appendicular skeleton,
were the major causes of death. Leukemia appeared early, while bone tumors
occurred after a latent period, the length of which was not reported.
Pool et al. (1973) studied 373 dogs exposed via ingestion to 0.03 to 36
yCi strontium/day from in utero to 1.5 years of age. At the high-dose, normal
bone formation was accompanied by a high rate of osteocyte cell death and
cortical focal necrosis, to the extent that the damage was irreversible. At
lower doses (12.5 ^Ci/day) repair attempts were more "successful" and injury
developed more slowly, which may account for the longer latency period noted
at lower doses (Pool et al. 1973). Dose-dependent bone tumors, osteosarcomas
(98 percent of which were malignant), occurred first in high- dose animals and
later in lower dose groups, though the relative times to tumor were not
reported. Pool et al. (1973) noted that this study and others in mice and
swine suggested that attempts at repairing initial radiation injury may be a
prerequisite to radiogenic tumor induction.
DRAFT III-ll
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Parks et al . (1984) exposed 387 dogs via ingestion to 0.02 to 36
strontium/day for 1.5 years (21 days post-conception to 540 days) noting a
dose-related temporal effect on the development of squamous cell carcinoma
(SCC). Dogs at higher doses developed SCC at an earlier age. The highest
dose group was the exception presumably because animals died earlier with bone
tumors, primarily osteosarcomas, and leukemia. In addition to significantly
higher numbers of animals with SCC, the percentages of carcinomas per location
in dosed groups differed from the percentages common in the controls. In
controls, tumors were about evenly divided between tumors located near the
molars and pre-molars and tumors located toward the front of the mouth, 43 and
58 percent, respectively. In dosed animals, 79 percent of the tumors were
located near the molars and pre-molars and 21 percent were located toward the
front of the mouth (Parks et a.l . 1984). Another chronic ingestion study was
performed using 384 dogs exposed beginning in utero and continuing for 19
months to daily amounts of 0.081 to 118.8 jiCi strontium/day. Gillett et al .
(1989) reported that, of the 46 primary bone tumors developed by 41 dogs, 50
percent of the tumors were located in the appendicular skeleton. In addition,
85 percent of the total number of tumors were osteoblastic osteosarcomas (bone
producing tumors) .
The exposure of female swine via ingestion to 1 to 3,100
strontium/day resulted in the development of leukemia, but few bone tumors
(Clarke et al . 1972). Ragan et al . (1972) exposed swine orally to 1 to 3,100
nCi strontium and also noted a low incidence of bone tumors. In contrast to
other studies where the primary effect was bone tumors, primary toxic effects
in swine were leukemogenic (causing leukemia) in bone marrow and
lymphoreticular systems (Clarke et al . 1972). Monkeys exposed via ingestion
to 500 or 1,000 jiCi strontium developed bone tumors (osteosarcomas and
chondrosarcomas, cartilaginous tumors) and leukemia (Mays and Lloyd 1972).
Evidence compiled from studies using monkeys, swine, rats, mice, and dogs
indicated that strontium enhances the occurrence of the most common leukemia
in a species rather than eliciting a distinct type of leukemia. In addition,
species with extremely low incidences of spontaneous leukemia appear more
resistant to the induction of leukemia by radiation (Clarke et al . 1972).
DRAFT 111-12
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Following inhalation exposure to strontium chloride beagle dogs
developed bone tumors, including hemangiosarcomas (tumors associated with
blood vessels), osteosarcomas, and squamous cell carcinomas, and leukemia
(Benjamin et al . 1976, Gillett et al . 1989). More specifically, Benjamin et
al . (1974) noted that many of the 72 dogs exposed via inhalation to 2.5 to 250
\iC\ strontium chloride/kg developed mainly hemangiosarcomas and osteosarcomas.
The incidence of hemangiosarcomas was initially greater but the number of new
cases that developed declined over time. Snipes et al . (1981) found that
inhalation of strontium chloride also resulted in local cancers, including
cancer of the lung, heart, and tracheobronchial lymph nodes. Most tumors
found in the tissues of these dogs were hemangiosarcomas.
Dougherty et al . (1972) found that beagles injected with 64 to 98
strontium/kg developed bone tumors, including hemangiosarcomas, osteosarcomas,
and squamous cell carcinomas, after an average latency period of 1,243 days.
Nilsson (1972) studied the effects of strontium in rodents. The injection of
0.2 to 1.6 \id strontium/g in mice resulted in leukemia and bone tumors.
Incidence of leukemia was highest in 0.2 to 0.4 yCi strontium/g dose groups,
while osteoblastic and fibroblastic (originating in connective tissues)
osteosarcomas increased with dose to 1.6 nCi strontium/g. A clear relation
was seen between dose, tumor incidence, and time-to-tumor. Those receiving
the high dose had more tumors (219) in a shorter period (an average of 267
days), while those receiving the low dose had fewer tumors (8) and those
tumors had a longer latency period (an average of 485 days) (Nilsson 1972).
Ito et al . (1976) also noted the development of leukemia following strontium
injection in mice. Bierke and Nilsson (1990) administered a single strontium
injection to mice and reported the development of osteosarcomas, squamous cell
carcinomas, malignant bone marrow lymphomas, and liver tumors which progressed
to mal ignancy.
Summary
Short-term exposure to strontium via ingestion, inhalation, and
injection resulted in hematopoietic changes in cats, monkeys, dogs and swine.
Among the most common changes were the depression of lymphocyte and neutrophil
counts. In addition, cats, dogs, and mice exposed to strontium exhibited bone
DRAFT 111-13
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marrow hypoplasia and, in many cases, myeloproliferative disease. Researchers
using dogs exposed via inhalation reported liver lesions and atrophy of lobes
of the liver.
Longer-term exposure resulted in similar effects including the
depression of lymphocyte and neutrophil counts and bone marrow hypoplasia and
myeloproliferative disease in dogs and swine exposed orally. Rats exposed
orally exhibited a decrease in indicators of immune response, including
changes in factors of nonspecific protection and autoantibodies, possibly
indicative of reduced immune capabilities.
No adverse reproductive effects were reported at levels below those
causing adverse effects in the dams. Mutagenic effects were seen in mice
exposed orally and in mice and hamsters exposed via injection. All three
groups exhibited an increase in the frequency of chromosomal aberrations in
bone marrow.
Carcinogenic effects of radioactive strontium in animals are well
documented. Oral exposure to strontium resulted in leukemia in mice, dogs,
monkeys, rats, and swine. Mice reportedly developed dose-dependent
appendicular bone tumors, while dogs were reported to develop dose-dependent
squamous cell carcinomas (SCC) and appendicular osteosarcomas. The SCC
induced by strontium occurred in different percentages per location relative
to spontaneously occurring SCC in controls. Exposure via the inhalation of
strontium chloride resulted in leukemia, bone tumors, and local tumors in
beagle dogs. Injections of strontium led to bone marrow tumors, liver tumors,
and leukemia in mice, as well as bone tumors in dogs. Bone carcinomas in mice
occurred with a definite dose-relationship between tumor incidence and time-
to-tumor. As with the ingestion study, higher doses resulted in more tumors
in a shorter period of time, while lower doses led to fewer tumors after a
longer latency period.
DRAFT 111-14
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Health Effects in Humans
Clinical Case Studies
Few case studies were located in which humans were exposed to strontium.
Human blood was irradiated (in vitro) using an external source of strontium to
produce doses of between 13.8 and 276 rads. Lymphocyte aberrations, total
aberrations and the number of dicentrics (chromosomes with two centromeres)
per cell, increased with dose to 220 rads (Vulpis and Scarpa 1986). The high
dose produced fewer total aberrations and dicentrics, perhaps because more
cells were killed at this dose.
Epidemioloqical Studies
Stannard (1973) conducted a study of 103 luminous dial painters exposed
to a compound containing strontium and radium-226 ratio unreported. The
incidence of chromosomal aberrations detected in the peripheral blood and bone
marrow of these individuals was twice that seen in the control group, which
reportedly consisted of healthy volunteers investigated at the same time and
in the same manner. Body burden estimates of strontium in the exposed group
ranged from 0.1 to 1.9 jiCi (based on whole body counting and on strontium
excretion). Pain in long bones and tendency toward subcutaneous hematomas was
noted in nearly half the subjects. Evaluation of the hematology of the
subjects identified cases of leukopenia, neutropenia, and thrombocytopenia
(Volf 1972). While these cases involve individuals also exposed to radium,
preventing any definitive correlations with the toxicity of strontium, these
effects correlate well with bioassay data derived from animals exposed to
strontium (Ito et al. 1976).
High-Risk Populations
While there are currently no data available on the risk to young
children exposed to strontium, animal bioassay data suggested that this group
may be a high risk population for the development of bone tumors and of blood
disorders such as leukemia. Evidence in mice, dogs, and pigs indicated that
the young may be up to four times as sensitive as adults to the induction of
111-15
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bone tumors and may be more sensitive to the induction of blood disorders
(Mays and Lloyd 1972).
In addition to possible increased sensitivity to strontium exposure, the
human fetus may be exposed to maternal concentrations of strontium via the
placenta. Borisov (1972) studied skeletal ashes of human fetuses dying at
various stages of development in utero. Based on average strontium blood
levels of women in Moscow and concentrations of strontium in the skeletal
ashes, Borisov determined that the placenta was not a strong barrier to the
movement of strontium. In contrast, Warren (1972), using bone samples from
children whose mothers were exposed to nuclear fallout, determined that the
placenta provided protection to the fetus by impeding the movement of
strontium. This conclusion is supported by Kawamura et al. (1986) who
determined the observed ratio of strontium in bone to strontium in diet for
adults to be 0.12 and the observed ratio of strontium in fetal bone to
strontium in maternal diet to be 0.055.
Summary
Clinical case studies in vitro have revealed an increase in chromosomal
aberrations in human lymphocytes following irradiation with strontium.
Epidemiological studies of workers exposed to both strontium and radium-226
reported increased incidences of chromosomal aberrations in blood and bone
marrow of these individuals. In addition, decreased leukocyte, neutrophil,
and platelet counts were noted. Young children may be a high risk population
due their to increased sensitivity to the induction of bone tumors and blood
disorders following strontium exposure. Studies suggested that the placenta!
membrane hindered the movement of strontium, reducing the concentration in the
fetus, relative to the maternal concentration, and thereby providing some
protection to the fetus.
DRAFT 111-16
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C. IODINE
Physical and Chemical Properties
Chemical Properties - Elemental
Iodine (I) is a member of the halogen family (Group VIIA of the Periodic
Table), and is a bluish-black, lustrous, crystalline solid that volatilizes at
ordinary temperatures to a blue-violet gas (Weast 1981, Hawley 1981). It
exhibits some metallic-like properties and is a semiconductor of electricity
(Mills 1968). Although iodine is nonmetallic, it dissolves in many organic
solvents including alcohol, chloroform, ether, carbon tetrachloride, glycerol,
carbon disulfide, and alkaline iodide solutions; however, it is only slightly
soluble in water and the least soluble of the common halogens (Weast 1981,
Hawley 1981). In solutions of chlorides, bromides, and other salts, iodine is
somewhat more soluble. No hydrates of iodine are known to exist (Mills 1968).
Iodine is the least reactive of the halogens but will form compounds with all
the elements except the noble gases, sulfur, and selenium. Unlike the
corresponding carbonyl, nitrosyl, or sulfuryl chlorides, iodine does not form
compounds with carbon monoxide, nitrous oxide, or sulfur dioxide. Iodine does
not react with carbon, nitrogen, or oxygen unless high temperatures and a
platinum catalyst are used (Mills 1968).
Physical Properties
Elemental iodine has an atomic weight of 126.9, atomic number of 53, and
a melting and boiling point of 113.5°C and 184.35°C, respectively. The
density of the gas is 11.27 g/L and its specific gravity (solid) is 4.93 at
20°C. Iodine can exist in valences of 1, 3, 5, and 7. Physical properties of
iodine are listed in Table III-2, adapted from Mills (1968). Iodine-127 is
the only stable isotope of iodine found in nature, although 23 isotopes are
known to exist. Iodine-131 is produced from the radioactive decay of
tellurium-131 and decays to a stable isotope of xenon, xenon-131. Iodine-131
has a radiological half-life of 8.04 days and emits beta energy of 0.807 MeV
(Brucer 1979). Iodine-129, a decay product of tellurium, is a radioactive
isotope with a half-life of 1.57xl07 years, emits beta energy of 0.150 MeV,
DRAFT 111-17
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and decays to the stable xenon-129 (Brucer 1979). Half-lives and decay
schemes of iodine-129 and -131 may be seen in Figure III-2, adapted from
Lederer and Shirley (1978).
Table III-2. Physical Properties of Iodine
Atomic number 53
Atomic weight 126.9044
Solid Iodine
Color Bluish-black
Melting point, °C 113.6
Density, g/cc, 20°C 4.93
60°C 4.886
Crystal structure Orthorhombic
Vapor pressure, mm Hg, 25°C 0.31
113.6°C 90.5
Liquid Iodine
Color Bluish-black
Boiling point, °C 185
Critical temperature, °C 553
Critical pressure, atm 116
Density, g/cc, 120°C 3.960
180°C 3.736
Gaseous Iodine
Color Violet
Density, g/1, 185°C, 1 atm 6.75
Summary
Iodine is a non-metallic, crystalline solid that volatilizes to a gas at
ordinary temperatures. Although iodine is non-metallic, it will dissolve in
many organic solvents. It is the least reactive of the halogens but will form
compounds with most of the elements. Only one stable isotope for iodine is
found in nature, iodine-127; however, 23 radioactive isotopes are known to
exist. Two important radioactive isotopes are iodine-131 and iodine-127 with
half-lives of 8 days and 1.57xl07 years, respectively. Iodine-131 (beta
energy 0.807 MeV) decays to xenon-131 and iodine-129 (beta energy 0.150 MeV)
decays to xenon-129.
DRAFT 111-18
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Figure 111-2. Half-Lives and Decay Schemes
of lodine-129and -131
-129
(a)
1.6x107y(b)
R-= .150
Xe-129
(Stable)
1-131
8.04 d
(T = .807
Xe-131
(Stable)
a = half-life
b = maximum B" decay energy in MeV
ICRP1983
DRAFT
111-19
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Toxicokinetics
Absorption
Iodine is readily absorbed following oral, inhalation, or dermal
exposure. Following oral iodine exposure, the majority was absorbed from the
stomach and the upper portion of the small intestine. The oral rate of
absorption was approximately 5 percent of the administered dose per minute.
The USNRC (1975) has reported the absorption of iodine from the intestine at
95 percent. Following inhalation exposure, between 50 and 90 percent of
deposited methyl iodine was taken up into the blood. Deposition of iodine
particles in the lung was dependent on the physical properties of the
particulates (Johnson 1982).
Distribution
Iodine is distributed to the thyroid almost immediately following an
oral or parenteral administration. A maximum of 10 to 50 percent of the
administered oral or parenteral dose was found in the thyroid within 1 to 2
days (Stara et al. 1971). Following oral administration of iodine in women
with normal thyroid glands, uptake by the thyroid at 4 and 24 hours ranged
from 3 to 13 percent and 7 to 26 percent of the administered dose,
respectively (Schober and Hunt 1976). Uptake to the thyroid following the
oral exposure of men with normal thyroid glands was 17.8 percent, slightly
less than that in the female (Ghahremani et al. 1971). In addition to the
thyroid, iodine was also concentrated in the kidneys, gastric glands, mammary
glands, and salivary glands of humans (Maier and Bihl 1987). Twenty-four hour
iodine uptake into breast tissue was statistically significantly higher in
abnormal tissue, tissue with carcinoma or dysplasia (12.5 percent), compared
to histologically normal tissue (6.9 percent) (Eskin et al. 1974). The
greatest tissue concentrations of iodine (those which exceeded concentrations
in the blood) in animals were found in the thyroid followed by the liver,
ovary, kidney, adrenal, pituitary, lung, lacrimal gland, heart, pancreas,
spleen, thymus, and brain (Stara et al. 1971).
DRAFT 111-20
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The amount of beta energy distributed to the thyroid gland is dependent
on the gland size, geometry, and the beta particle energy. It has been
estimated that 70 percent of the total beta energy from absorbed iodine is
distributed to the thyroid of mice and rats, 90 percent to the thyroid of
human infants, and greater than 95 percent to the thyroid of adult humans (NAS
1990). In workers exposed to iodine, a biological half-life in the thyroid of
48 to 75 days has been reported (Raghavendran et al. 1978).
The uptake of orally administered radioactive iodine by the thyroid is
dependent on the amount of stable iodine in the diet. Large amounts of stable
iodine in the diet may block the uptake of radioactive iodine by the thyroid.
During the 1950's, human studies determined that, in individuals with normal
thyroid function, 24 hour uptake to the thyroid was 28.6 percent, while
comparable studies conducted in the late 1960's reported the 24 hour uptake to
be 15.4 percent (Pittman et al. 1970). This decrease in thyroid uptake seems
to be due to increased dietary intake of iodine in the more recent decade
(Pittman et al. 1970, Oddie et al. 1970). In rats fed a diet with a low
stable iodine level of 2 ng/day, the level of radioactive iodine in the
thyroid peaked at greater than 50 percent of intake in 6 hours. The effective
half-life was 2.5 days. When rats were fed a high stable iodine diet (15
ng/day), the maximum uptake of radioactive iodine, occurring at 24 hours, was
9 percent of the administered amount. The effective half-life was 4 days.
Similar uptake by the thyroid has been observed in sheep and chicks (Stara et
al. 1971).
Inorganic iodine crossed the placental membrane and distributed from
maternal to fetal blood in humans (Johnson 1982). In the fetus, the thyroid
began to accumulate iodine at about 90 days of age and this continued
throughout gestation. The average absolute iodine dose to the fetal thyroid
in hyperthyroid patients was predicted to be highest when administration of
iodine occurred during the sixth month of gestation (Stabin et al. 1991). The
concentration of iodine in fetal blood at term was approximately 75 percent of
that in maternal blood. With organic iodine, little if any placental transfer
has been observed (Johnson 1982). No information was located on the transfer
of radioiodine across the placenta.
DRAFT IH-21
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Excretion
Iodine was excreted primarily in the urine of animals and man with a
urine:feces ratio of 27:1 in man. In rats, fecal excretion was also an
important route (Stara et al. 1971). In ten humans administered iodine to
destroy normal thyroid tissue in preparation for treatment of thyroid
carcinoma, 46 to 83 percent of the administered dose was excreted within
4 days (Andrews et al. 1954). Small amounts of iodine have also been shown to
be excreted in humans via perspiration and exhalation (Nishizawa et al. 1980).
Iodine is eliminated exponentially with an effective half-life dependent on
species, thyroid function, and diet (Stara et al. 1971).
Bioaccumulation and Retention
See Section E. In this section toxicokinetic models for beta or gamma
emitting radionuclides are used to predict the potential bioaccumulation and
retention for these radionuclides. Some of the metabolic models are based on
extensive human and animal data, some on primarily animal data and some on
similarity of elemental characteristics. The model for iodine may be found on
page 111-68.
Summary
Iodine is readily absorbed following all routes of exposure. The oral
rate of absorption was approximately 5 percent of the administered dose per
minute and the absorption of iodine from the intestine was 95 percent of the
administered dose (USNRC 1975). Distribution of iodine is primarily to the
thyroid. An estimated 90 to 95 percent of the absorbed dose is distributed to
the thyroid of humans. The distribution of iodine to the thyroid may be
influenced by the amount of stable iodine in the diet. Iodine is excreted
primarily in the urine of both animals and humans. In humans, the urine to
feces ratio is 27:1. The excretion of iodine is dependent on thyroid function
and diet.
DRAFT 111-22
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Health Effects in Animals
The following discussion of health effects in animals is based on
experiments using iodine-131.
Short-term Exposure
Oral exposure to iodine in animals has resulted in damage to the thyroid
and effects on the hematological system. Exposure to high parenteral doses
(50 mCi) resulted in complete destruction of the thyroid gland and a lack of
epithelial regeneration. With doses of 25 to 40 yCi, focal regenerative
hyperplasia progressed to benign nodule formation (Lindsay and Chaikoff 1964).
Exposure to lower doses resulted in damage to thyroid and parathyroid tissue,
skeletal abnormalities, and hematological effects.
In adult sheep administered 1.25 mCi iodine by the oral route, no
damage was observed 60 days after administration (Goldberg et al. 1950).
Exposure to 5 mCi resulted in severe damage to the thyroid, while 15 mCi
completely destroyed the thyroid gland (Garner 1963). Adult sheep
administered a single dose of 15 mCi iodine exhibited a transient depression
in lymphocytes and leukocytes at 15 days after exposure which returned to
normal by 50 days (Garner 1963).
No information was located regarding the health effects of iodine in
animals following short-term inhalation exposure.
In male Long-Evans rats administered a single intraperitoneal injection
of 18 jiCi iodine, no evidence of damage to any tissue was reported. In
animals which received 300 or 525 ^Ci, damage to the thyroid occurred in five
phases: degeneration and necrosis of epithelial cells, vascular degeneration
and thrombosis, inflammatory changes, fibrotic changes, and epithelial
regeneration. Similar effects occurred in animals receiving 875 ^Ci with the
exception that epithelial regeneration did not occur. Effects were seen as
early-as 12 hours after exposure to the two highest doses. In animals
receiving 525 or 875 pCi, fibrosis occurred in the peripheral portions of the
parathyroid glands and slight glomerular congestion and slight cloudy swelling
DRAFT 111-23
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was observed in the kidneys up to one week after exposure (Goldberg et al.
1950).
Juvenile beagles were administered intravenous injections of 0, 0.1,
0.3, 0.6, or 1 mCi iodine/kg. In dogs receiving 1 mCi/kg, abnormal skeletal
development was observed within 3 to 4 months and death occurred in two of the
four dogs within 12 to 15 months following exposure. An absence of thyroid
tissue was seen in the remaining two dogs at autopsy (30 months after
exposure). Skeletal growth retardation, hypertrophy of the adrenal medulla,
lymph-node hyperplasia, and neoplastic thyroid tissue were observed in dogs
administered 0.3 or 0.6 mCi/kg. These dogs also had roughened and dry coats
and excessive folding of the skin. No effects were observed in animals
receiving 0.1 mCi/kg (Andersen 1958).
In mice administered 3 to 50 mCi iodine/kg by subcutaneous injection,
complete destruction of the thyroid occurred within 3 days at the high dose.
Exposure to levels as low as 3 to 4 mCi/kg resulted in 90 percent destruction
of the thyroid by 120 days. The thyroid tissue which survived was epithelial
tissue in the isthmus and cranial apex. By 120 days after injection of 3 to
4 mCi/kg, a total loss of the parathyroids occurred (Gorbman 1947).
Longer-term Exposure
In sheep fed 1800 ^Ci for 2 to 4 months or 240 ^Ci for 6 to 10 months,
signs of hypothyroidism were reported. These included lethargy, clumsy
movements, bloating, ulceration of the oral mucosa and tongue, decreased milk
production, and dry skin and fleece (Garner 1963). In sheep fed 5 jiCi/day, a
depression in lymphocytes and leukocytes was reported (Garner 1963).
No information was located on the health effects of iodine in animals
following longer-term exposure via other routes.
Reproductive/Developmental Effects
Limited information was available on the reproductive/developmental
effects of iodine in animals. In lambs exposed j_n utero by. feeding the dams
DRAFT 111-24
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135 i»Ci to 1800 nCi iodine, hypothyroidism was observed at birth. Ewes
exposed to 1800 jiCi or 240 jiCi during pregnancy gave birth to either dead
lambs or lambs which did not survive for more than a few days (Garner 1963).
In sheep receiving daily quantities of 1.5 or 5 yCi iodine from conception
(either j_n utero or in feed), thyroid adenomas were observed after 4 to 6
years (Lindsay and Chaikoff 1964). Thyroid changes, including early fibrosis,
compensatory hyperplasia with adenoma formation, and colloid goiters, have
been observed in the offspring of mouse dams administered iodine injections
during pregnancy (Lindsay and Chaikoff 1964). No adverse effects on fertility
have been reported in hypothyroid sheep or cows (Garner 1963).
Mutagenicity
No information on the mutagenic effects of iodine was located in the
reviewed literature.
Carcinogenicity
Both benign and malignant thyroid neoplasms have been reported following
single or repeated exposure to iodine. In sheep fed daily quantities of 1.5
to 135 nCi beginning at 15 months of age, thyroid adenomas were observed at 4
to 8 years of age (Lindsay and Chaikoff 1964). In a separate study in sheep
fed 5 ^Ci per day for life, both thyroid adenomas and fibrosarcomas were
reported (Lindsay and Chaikoff 1964). Three years after feeding adult sheep
45 nCi per day for 12 months or 135 ^Ci for 8 months, thyroid adenomas were
reported (Garner 1963).
No information was located regarding the carcinogenicity of iodine in
animals following inhalation exposure.
Juvenile beagles were administered single intravenous injections of 0,
0.1, 0.3, 0.6, or 1 mCi iodine/kg. Thyroid adenomas were observed at
sacrifice (30 months after exposure) in five out of eight dogs which received
0.3 or 0.6 mCi iodine/kg and in one of two dogs receiving 1 mCi/kg (Andersen
1958).
DRAFT 111-25
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Levels of iodine as low as 1 nCi have been reported to result in thyroid
tumors in Long-Evans rats (Lindsay and Chaikoff 1964). The highest incidence
of thyroid neoplasms reported has been in Long-Evans rats that received either
a single dose of 25 nCi or four doses of 10 ^Ci at weekly intervals (Lindsay
and Chaikoff 1964). In male rats which received a single intraperitoneal
injection of 25 (iCi iodine or four injections of 10 jiCi at monthly intervals,
a statistically significant increase in follicular adenomas, papillary
adenomas and carcinomas, and follicular carcinomas of the thyroid were
observed (Potter et al. 1963). Following three intraperitoneal injections of
10 nCi at monthly intervals in female rats, benign thyroid nodules or adenomas
were observed in 39 percent of the animals compared to 3 percent in controls.
The incidence of carcinomas in females was 7 percent compared to no carcinomas
in controls. The incidence of carcinomas in females was lower than in males
(21 percent) which were treated with lower doses (Lindsay et al. 1963). In
female Long-Evans rats administered 0.48 to 5.4 ^Ci iodine by intraperitoneal
injection, a dose-related increase in thyroid carcinomas was observed. The
authors suggest that carcinoma induction was dependent on total dose and not
dose rate (Lee et al. 1982).
Summary
Exposure to iodine by either the oral or parenteral route has been shown
to result in damage to the thyroid gland. Depending on the dose administered,
effects on the thyroid ranged from epithelial degeneration, fibrotic changes,
and cancer, to complete obliteration of the gland. Other effects which have
been reported include skeletal abnormalities, a decrease in lymphocytes and
leukocytes, damage to the parathyroid, and kidney effects. Following longer-
term oral exposure, ulceration of the oral mucosa and tongue were also
reported. Developmental effects which have been reported following exposure
of mouse and sheep dams included changes in the thyroid gland. No adverse
effects on reproduction were reported. Both benign and malignant thyroid
neoplasms have been reported following iodine exposure. Repeated oral
exposure to iodine in sheep has been reported to result in thyroid adenomas
and fibrosarcomas as early as 3 years after exposure. Levels of iodine as low
as 1 jiCi administered by injection have been reported to result in tumors in
rats.
DRAFT 111-26
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Health Effects in Humans
Clinical Case Studies
Tenderness and swelling of the salivary glands, as well as a decrease in
salivary flow, have been reported in humans following oral doses of iodine
high enough to destroy thyroid tissue (Maier and Bihl 1987). Oral
administration of 20 to 70 mCi iodine in patients with carcinoma of the
thyroid gland resulted in a "moderate" increase in red blood cell membrane
permeability against hemoglobin in peripheral blood at 4 days after
administration. The authors attributed this effect to beta radiation of the
peripheral blood resulting in direct damage to the formed elements (Geszti et
al. 1973). Chromosomal damage (type not specified) was reported in one
patient following administration of two 100 mCi doses of iodine (Gestzi et al.
1973).
Oral administration of 350 mCi iodine in a thirteen year old boy with
thyroid cancer has been reported to result in testicular damage, including a
lack of sperm in the semen (Ahmed and Shalet 1985). In addition, persistently
elevated basal follicle stimulating hormone level was reported. Four years
after treatment there appeared to be no recovery of sperm production in the
boy (Ahmed and Shalet 1985). The possibility of a reversal of this condition
was undetermined.
In patients administered greater than 100 mCi of iodine-labeled
monoclonal antibodies by intraperitoneal injection, for the treatment of
ovarian cancer, reversible bone marrow suppression was reported. Neutropenia
and thrombocytopenia were reported 3 to 5 weeks following injection with the
lowest blood counts observed from days 30 to 40 (Stewart et al. 1989).
Epidemiological Studies
In patients with Graves' Disease, a form of thyrotoxicosis, who were
treated with iodine, thyroid nodules were found in eight of 256 patients 5 to
14 years following iodine therapy. Histological examination revealed typical
radiation effects including nodules consisting of follicular atrophy,
DRAFT 111-27
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perifollicular fibrosis, and mild chronic inflammatory infiltration. Nodules
did not occur in patients who developed hypothyroidism following iodine
therapy for thyrotoxicosis which indicated that the gland was so severely
damaged that it was incapable of regeneration and nodule formation. The
glands of these patients resembled those of rats receiving 400 pCi (Lindsay
and Chaikoff 1964).
In a prospective study of 1,005 patients treated with iodine for
thyrotoxicosis, the risk of thyroid cancer was 9 fold greater compared to
patients who were treated surgically for this disease, but was not
statistically significant when compared with data from the Connecticut Cancer
Registry. In 21,714 patients treated with iodine compared to 11,732 patients
treated surgically, the incidence of thyroid cancer was not significantly
increased. In 4,557 iodine treated patients, the thyroid cancer incidence was
not increased compared with data from the Swedish Cancer Registry (NAS 1990).
In a follow-up study of 10,133 patients who received diagnostic doses of
iodine, no evidence of an increased risk of thyroid cancer was observed. A
separate study was conducted of 35,074 patients who survived 5 years or more
after receiving a mean diagnostic dose of 0.05 mCi iodine. The mean follow-up
period was 20 years and the mean age at exposure was 44 years. Fifty thyroid
cancers were observed in the exposed group compared to an expected 39.37
cases. Of the 50 observed cases, ten were of a type not shown to be
associated with radiation exposure, six were from subjects 50 to 74 years of
age at time of exposure, and 15 cancers occurred after 5 to 9 years of
exposure suggesting that they were present at the time of exposure.
Therefore, the authors concluded that the data do not support an increased
risk of thyroid cancer from diagnostic doses of iodine (NAS 1990).
Residents of the Marshall Islands were exposed to radioactive iodine
from fallout from the BRAVO test bomb. Doses of ingested radionucl ides were
calculated from the iodine content of pooled urine samples taken 15 days after
the first exposure. Doses were calculated to range from 0.3 to 15 Grays (30
to 1500 rads). The prevalence of hypothyroidism, thyroid nodules, and thyroid
cancer appeared to increase with dose. General radiation sickness, including
nausea and hair loss, were reported within 2 weeks following exposure. When
DRAFT 111-28
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additional studies of the Islanders were undertaken, a statistically
significant dose-related increase in thyroid nodules was observed. A
significant dependence of nodule prevalence on distance from the test site,
age at exposure, sex, and latitude was observed when using logistic regression
analysis (MAS 1990).
In school-age children who ingested milk contaminated with iodine from
atmospheric fallout resulting from atomic bomb tests, a suggestive 20 to 30
percent greater incidence in all thyroid abnormalities compared to unexposed
controls was reported. The cumulative radiation doses to the thyroid were
estimated to average as high as 1 Gray (100 rads) (NAS 1990).
No increase in reproductive problems, i.e., infertility, miscarriage,
prematurity, or birth defects, were reported in 33 patients receiving a mean
dose of 196 mCi as children or adolescents (Sarkar et al. 1976). However,
further examination of this and other data prompted Handelsman et al. (1980)
to note infertility in some men treated after puberty. Toxicity of the testes
was seen in 3 of the 6 men in this category, including two men included in the
Sarkar et al. (1976) report (Handelsman et al. 1980, Ahmed and Shalet 1985).
Total iodine dose administered (100 to 400 mCi) correlated positively with
serum follicle-stimulating hormone concentration and negatively with sperm
density (Handelsman et al. 1980). More detailed follow-up studies of
similarly treated individuals would be required to determine the reversibility
of this effect.
111-29
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High-Risk Populations
While little data were located concerning probable high-risk populations
following exposure to iodine, infants and children may be included in this
category. Various incidents involving the treatment of pregnant women with
iodine have resulted in infants with depressed or absent thyroid function
(Stabin et al. 1991). These effects were likely the result of iodine
accumulation in the fetal thyroid, which has been shown to begin at 90 days of
age. In addition, iodine uptake to the thyroid in newborn infants is much
greater than that in adults, 62 percent compared to 16.9 to 26 percent
(depending on sex) (Fisher et al. 1962, Schober and Hunt 1976). The decline
of thyroid absorption to adult values may occur within 5 days or require as
long as a year.
Studies concerning other radionuclides, specifically strontium, have
shown children to be more sensitive to the induction of cancer following
exposure (Mays and Lloyd 1972). The data suggest that children exposed to
radiation during the first five years of life have a considerably increased.
risk of cancer, compared to those exposed later (NAS 1990). NAS (1990)
concludes that the risk of radiation-induced thyroid cancer in children is '
twice as great as that in adults. Iodine radiosensitivity in children has not
yet been fully explored. Pending further relevant studies, children should be
considered a high-risk population for iodine exposure.
Summary
The effects of iodine exposure have been studied in patients receiving
large therapeutic doses, patients receiving smaller doses of iodine for
diagnostic purposes, and those exposed to environmental iodine fallout (NAS
1990). Injury to the human thyroid gland, including cancer, has occurred
following irradiation for thyrotoxicosis or thyroid carcinoma. Effects
observed in the human thyroid gland at long times after administration of
large doses include chronic inflammation, interlobular and perifollicular
fibrosis, and foil icular atrophy (Lindsay and Chaikoff 1964). Other effects
observed in humans include changes in blood cell membrane permeability and
chromosomal damage. Effects observed in humans differ from.those in
DRAFT 111-30
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laboratory animals in that the sequence of effects appears to occur more
rapidly in animals. In addition, the size of the organs in small animals in
relation to the path of the beta radiation is such that changes are more
commonly produced in adjoining structures, such as the parathyroids and
trachea (Andrews et al. 1954). Due to data concerning adverse effects in
infants exposed to iodine and the possible increased sensitivity of children
to iodine exposure, infants and children should be considered a high-risk
population for iodine-induced effects.
D. TRITIUM
Physical and Chemical Properties
Chemical Properties - Elemental
Tritium, a radioactive isotope of hydrogen, may be formed naturally by
interactions of cosmic rays with gases in the upper atmosphere, or may be
prepared by the bombardment of lithium with low energy neutrons in nuclear
reactors, or by nuclear bombardment of deuterium with other hydrogen species.
Tritium is also present in effluents from nuclear reactors and weapons (Hawley
1981, Hobbs and McClellan 1986).
With respect to chemical reactions, tritium reacts similarly to ordinary
hydrogen. However, hydrogen isotopes are easily distinguished from one
another because of the relatively large differences in mass. The reaction of
tritium to form tritiated water (HTO) is favored at room temperature, and most
tritium present in the environment is in the form of tritiated water (Jacobs
1982).
Physical Properties
Tritium (mass number 3, isotopic weight 3.017) is the heaviest of the
three hydrogen isotopes and has one electron outside the nucleus and two
neutrons and one proton inside the nucleus. Tables III-3 and III-4, adapted
from Greenwood and Earnshaw (1984), list atomic and physical properties,
respectively, for hydrogen, deuterium, and tritium. The binding energy
111-31
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between the three elementary nuclear particles in tritium is relatively low.
Consequently, the nucleus is not stable and will decay by emission of an
electron (beta particle) with energy up to 18.6 kev and a neutrino, where one
neutron transforms into a proton. The nuclide gains an additional charge that
can hold a second electron in orbit and thus undergo beta decay to a new
chemical identity, primarily helium 3 (stable). In this process an average of
approximately 11 ev of excitation may be added to the helium ion (Feinendegen
1967). The half-life is 12.4 years and the decay constant, X, equals 0.0565
per year, or 1.791xlO"9 per second. The half-life and decay scheme for
tritium may be seen in Figure III-3, adapted from Lederer and Shirley (1978).
Table III-3. Atomic Properties of Hydrogen, Deuterium, and Tritium
Property
Relative atomic mass
Radioactive stability
Hydrogen
1.007825
Stable
Deuterium
2.014102
Stable
Tritium
3.016049
IT tv, 12.35 ya
18.6 keV;
5.7 keV.
Table III-4. Physical Properties of Hydrogen, Deuterium, and Tritium
Property3
Melting Point °C
Boiling Point °C
Heat of fusion/kJ mol"1
Heat of vaporization/kj mol"1
Critical temperature/K
Critical pressure/atmb
Hydrogen,
-259
-253
0.117
0.904
33.19
12.98
Deuterium,
-254
-249
0.197
1.226
38.35
16.43
Tritium,
-252
-248
0.250
1.393
40.6 (calc)
18.1 (calc)
Data refer to H2 of normal isotopic composition (i.e., containing 0.0156
atom percent of deuterium, predominantly as HD). All data refer to the
mixture of ortho- and para-forms that are in equilibrium at room
temperature.
1 atm = 101.325 kN m2 = 101.325 kPa.
DRAFT
111-32
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Figure 111-3. Half-Life and Decay Scheme
of Tritium
H -3 (a)
12.33y(b)
3-=0.0186
Helium
a = half life
b = maximum 3'decay energy in MeV
Lederer and Shirley 1978
111-33
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Summary
Tritium is a radioactive isotope of hydrogen that has a half-life of
12.4 years, emits beta radiation up to 18.6 kev, and decays to helium-3.
Tritium reactions are similar to those involving hydrogen, and most of the
tritium present in the environment is in the form of tritiated water. Tritium
is often used as tritium oxide (HTO, tritiated water), tritiated thymidine,
and other forms to label atoms or reactions for the purpose of following
mechanism pathways or to identify and analyze products. Tritium atoms can
replace hydrogen atoms in many compounds, rendering the new molecule
radioactive and providing a means of monitoring the presence and concentration
of labeled compounds with beta-particle detecting devices (Andrews 1968).
Toxicokinetics
Absorption
In this analysis all references to tritium refer to tritium oxide (HTO),
usually administered as tritiated water. Tritium is easily absorbed from the
gastrointestinal tract, skin, and lungs into the bloodstream (Hobbs and
McClellan 1986). Gastrointestinal absorption is about 95 percent (Killough
and Rohmer 1978). Pinson and Langham (1957) conducted an experiment in which
fasted adult males ingested 100 to 1000 milliliters (mL) of tritiated water.
Absorption through the gastrointestinal tract began after 2 to 9 minutes and
was complete in 40 to 45 minutes (Pinson and Langham 1957). Within the range
of 100 to 1000 ml, the volume of tritiated water transferred from the
intestine to the blood was linear with time and proportional to the water
volume ingested (Pinson and Langham 1957). In humans 98 to 99 percent of
inhaled tritium was exchanged and absorbed through the respiratory tract, the
remainder being exhaled (Moskalev 1968).
Distribution
Though limited definitive data are available, the distribution of
tritium has been described as follows. An equilibrium distribution of tritium
in the bodily fluids of humans was established 2 to 4 hours.after ingestion or
DRAFT 111-34
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inhalation of tritium. In addition to its presence in body fluids, 0.5 to 4
percent of the tritium activity may be incorporated into organic molecules of
the body (Balonov et al. 1984). Subsequent to intake, tritium content in the
urine, saliva, sweat, feces, and exhaled air was the same, indicating that
tritiated water migrates through biological barriers at the same rate as
ordinary water (Moskalev 1968). According to ICRP (1979) tritium distribution
to soft tissues is uniform. However, since some tissues contain more water
than others, the dose administered to tissue via the associated body water
would be expected to vary accordingly (Balonov et al. 1984). Moskalev (1968)
reported that in humans occupationally exposed to tritium (average specific
activity in these individuals was 23 jiCi/L body water), the specific activity
detected in certain organic components, specifically in fat and skin, was
higher than the specific activity detected in the body water, though no
explanation was offered as to why these components would contain a higher
specific activity.
Excretion
Data concerning the excretion of tritium are limited. Since tritium is
incorporated into body water, most excretion is presumably via the urine,
though some is expired via respiration. The rate of tritium elimination
increased slightly with age. The biological half-life of tritium for persons
20 to 29, 30 to 39, 40 to 49, and 50 to 59 years of age has been shown to be
10.5, 9.5, 9.0, and 8.2 days, respectively (Moskalev 1968). Other factors
influencing the elimination of tritium are the amount of water consumed and
the rate of water turnover.
Bioaccumulation and Retention
See Section E. In this section toxicokinetic models for beta or gamma
emitting radionuclides are used to predict the potential bioaccumulation and
retention for these radionuclides. Some of the metabolic models are based on
extensive human and animal data, some on primarily animal data and some on
similarity of elemental characteristics. The model for tritium (hydrogen) may
be found on page 111-43.
DRAFT 111-35
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Summary
Tritium is rapidly absorbed into the body water, regardless of route of
administration (Pinson and Langham 1957). Humans absorbed 98 to 99 percent of
inhaled tritium. Tritium distribution in humans was uniform in the body
water. Distribution in tissues, though thought to be relatively uniform, may
vary depending on the amount of body water contained in the tissue. The rate
of tritium excretion has been shown to increase with age. Excretion may also
vary with rate of water turnover and amount of water consumed.
Health Effects in Animals
Short-term Exposure
Data currently available concerning health effects in animals exposed to
tritium via ingestion or inhalation are limited. A single intraperitoneal
injection of 1400 jiCi tritium/g body weight resulted in death in all treated
mice (Zhuravlev 1968). Reported LD50/30s (the dose which was lethal to 50
percent of the animals in 30 days) for injection in mice ranged from 250 to
1000 tiCi/g body weight; however, the range of 250 to 350 yCi/g was generally
agreed upon (Zhuravlev 1968). The LD50/30 for rats exposed via all routes is.
about 1,070 pCi/g body weight, confirming that the toxicity of tritium for
rats is not dependent upon the route of exposure, since tritium is rapidly
absorbed and distributed following exposure by any route (Zhuravlev 1968).
Rats exposed via injection to 1,200 or 2,400 jiCi tritium/g body weight
(doses in excess of the LD50) exhibited an immediate decrease in erythrocyte
count. However, in rats dosed with 300 or 600 pCi tritium/g there was an
initial increase in erythrocyte count before the decrease began 3 to 5 days
after injection (Zhuravlev 1968). Rats injected with 300 to 2,400 ^Ci
tritium/g body weight exhibited a decrease in absolute leukocyte numbers on
the first day after exposure. Leukocyte count on days 7 to 10 was observed to
be 400 cells per cubic millimeter (mm3), compared with 16,000 cells/mm3 before
injection. By the end of a month leukocyte values in the surviving animals
were between 5,500 and 9,000 cells/mm3. Dogs exposed via injection to 300
body weight also exhibited decreases in leukocyte, platelet,
DRAFT 111-36
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erythrocyte, and lymphocyte counts. Zhuravlev (1968) reported that the
ability of the liver to synthesize hippuric acid was inhibited in rats exposed
to 40 pCi tritium/g body weight and dogs exposed to 300 nCi tritium/g body
weight. The production of hippuric acid declined to 50 and 55 percent of the
initial output for rats and dogs, respectively. A gradual decrease in the
synthesis of hippuric acid was seen when smaller doses were administered over
a period of time. Shal'nova (1968) demonstrated that tritium administered in
single injections of 80 to 300, 10 to 300, and 150 to 300 nCi/g body weight,
in mice, rats, and dogs, respectively, resulted in damage to immunological
function. All three species exhibited disruption of phagocytic activity of
neutrophils and all began to produce autoantibodies designed to attack the
organism's tissue (Shal'nova 1968).
Longer-term Exposure
Repeated oral doses of 5, 10, and 40 jiCi tritium/g body weight
administered to rats produced a gradual increase in toxicity (Zhuravlev 1968).
In rats, administration of 40 jiCi tritium/g resulted in changes in the white
blood cells, including an increased number of lysed cells, vacuolization of
the protoplasm, and fragmentation of the nuclei of the neutrophils.
Additional effects observed were a decrease in the number of erythrocytes,
leukocytes, and reticulocytes, and a less distinct decrease in the ability of
the liver to synthesize needed compounds. At doses of 10 nCi tritium/g, a
gradual inhibition of hematogenesis (specifically of erythrocytes and
leukocytes) was observed. No changes were seen in rats administered 5 nCi
tritium/g for 2 months.
Reproductive/Developmental Effects
Female germ cells in some mammals have been shown to be extremely
sensitive to radiation, such as that emitted by tritium (Dobson and Kwan
1977). Pregnant mice were given tritiated drinking water containing less than
0.1 to more than 10 nCi/mL from conception to 14 days postpartum (Dobson
et al. 1986). Results indicated that the tritium-induced decrease of
reproductive capacity was dose-dependent but not directly proportional to
decreased oocyte number (Dobson et al. 1986). No decrease in reproductive
DRAFT 111-37
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capacity was seen unless early oocyte loss exceeded 50 percent. Fecundity of
exposed females tended to be normal during early reproductive life but failed
prematurely as oocytes diminished (Dobson et al . 1986).
Rats were exposed to 0.01, 0.10, 1.0, or 10.0 (iCi tritium/mL drinking
water from conception of the F1 generation through delivery of the F2
generation (offspring of the F,) (Laskey et al . 1973). A 30 percent reduction
in the weight of testes of the F, male and decreased litter size and increased
resorptions in the dam were observed following exposure to 10 fiCi tritium/mL
(Laskey et al. 1973). A similar study, using only intrauterine exposures,
produced statistically significant reductions in litter size and offspring
body weight. The dose levels used in this second study (though not reported)
were 10 to 100 times greater than those used in the lifetime study (Laskey
et al. 1973).
Cahill and Yuile (1970) exposed pregnant rats to 1 to 100 nCi tritium/mL
of body water5 throughout pregnancy and reported statistically significant
litter size reduction and increased resorptions. Cahill et al . (1975) exposed
pregnant rats to 1, 10, 50, or 100 jiCi tritium/mL of body water to determine
late effects of tritium exposure. No late effects were reported at the two
lower doses. All rats exposed in utero to the two highest doses were sterile.
Cahill and Yuile (1970) exposed pregnant rats to 1 to 100 nCi tritium/mL of
body water throughout pregnancy and reported microencephaly, sterility, and
stunting, all of which were statistically significant.
In addition to reproductive effects, tritium also affected the
offspring. When parents were exposed to tritiated drinking water, neonatal
effects included the reduction of the brain to body weight ratio and decreased
body weight at doses as low as 0.1 and 1.0 pCi/mL, respectively (Laskey et al .
1973). Cahill et al . (1975) reported that all rats exposed in utero to 100
tritium/mL had reduced mean life spans, when compared to controls.
Females were injected with tritium with a high specific activity so
that when the tritium was diluted by the body water, the desired
concentrations were obtained. Following the initial injection, tritiated
water was given ad libitum to maintain the desired concentration.
DRAFT 111-38
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Mutaqenicitv
Injection of rats with a single dose of 75 or 100 jiCi tritium/g body
weight resulted in a statistically significant increase in chromosomal
aberrations in bone marrow cells within 24 hours (Andreuta and Racoveanu
1974). A gradual decrease in aberrations was seen 3 to 7 days after
administration. In rats administered daily injections of 30 yCi/g body weight
for 18 days, an increase in the incidence of bone marrow cell aberrations was
reported (Andreuta and Racoveanu 1974).
Carcinoqenicity
In addition to noncarcinogenic effects, animals exposed to tritium via
drinking water developed tumors. Following multigeneration exposure of male
mice to 10 (iCi tritium/mL of drinking water, a heritable, multiple intestinal
adenocarcinoma (HMIA) was observed in the offspring of five successive
generations. This tumor type, not previously seen in the C57 Black/6M strain,
was detected in 44 percent of the males and 78 percent of the females in the
five generations (Mewissen 1983).
In a study performed by Cahill and Yuile (1970), pregnant rats received
1 to 100 jjCi tritium/mL of body water. Within 270 days, five tumors (one in
the thymus gland, two ovarian, and two mammary) had arisen in an unreported
number of animals exposed in utero to 50 to 100 jiCi tritium/mL) compared to
one benign facial adenoma in a control. Cahill et al . (1975) exposed pregnant
rats to 1, 10, 50, or 100 (iCi tritium/mL body water from conception to birth
of offspring. Females exposed iji utero to 50 or 100 jiCi tritium/mL had an
increased incidence of ovarian tumors, compared to controls (Cahill et al .
1975). The controls had no ovarian tumors, while rats receiving 50 and 100
i tritium/mL developed five and two tumors, respectively.
Summary
Short-term exposure of rats and dogs to tritium has been reported to
cause hematological changes, such as the depression of erythrocyte and
leukocyte counts in both species and decreased platelet and lymphocyte counts
DRAFT 111-39
-------
in dogs. In addition, the ability of the liver to synthesize hippuric acid
was reduced, indicating some compromise in hepatic function. Rats, mice, and
dogs reportedly developed damage to immunological function, such as the
disruption of phagocytic activity of neutrophils. Longer-term exposure has
been shown to produce changes in white blood cells and to inhibit
hematogenesis in rats.
Reproductive studies in mice exposed to tritium j_n utero have detected
decreased reproductive capacity and oocyte death. Studies in rats have shown
reduced testes weight, brain to body weight ratio, overall body weight, and
litter size. Mutagenicity studies reveal a significant increase in
chromosomal aberrations of bone marrow in exposed rats. Carcinogenic effects
include the observation of a heritable, multiple intestinal adenocarcinoma in
the offspring of exposed mice, as well as the development of mammary and
ovarian tumors in exposed rats.
Health Effects in Humans
C1inical Case Studies
There are currently no data available from clinical case studies
regarding health effects resulting from human exposure to tritium.
Epidemiological Studies
There are currently no data available from clinical case studies
regarding health effects resulting from human exposure to tritium.
High-Risk Populations
Female germ cells in some mammals have been shown to be extremely
sensitive to radiation, such as that emitted by tritium (Dobson and Kwan
1977). However, germ cell radiosensitivity in unborn females has not been
studied extensively (Dobson et al. 1986). For this reason females exposed in
utero may be a high-risk population for tritium induced damage to germ cells.
DRAFT 111-40
-------
Summary
There are currently no data available from clinical case studies or
epidemiological studies regarding human health effects which may be associated
with exposure to tritium. Females exposed in utero may be a high risk
population.
E. BIOACCUMULATION AND RETENTION
The toxicokinetic models for beta and gamma radionuclide uptake and
retention are derived from a number of sources: EPA (1977), Adams et al .
(1978), Bernard and Snyder (1975), ICRP (1973, 1979) and Sullivan et al .
(1981). Some of the metabolic models are based on extensive human and animal
data, some on primarily animal data and some on similarity of elemental
characteristics. These radionucl ides include naturally occurring forms such
as: uranium, radium and radon, which are described in other criteria
documents; tritium, carbon-14, etc. and some man-made radionucl ides such as
iodine-131 and cesium-137.
In general, retention is defined as the amount of a substance remaining
in a tissue or organ at some time after the intake of the substance. The
retention function R(t) specifies the fraction of a substance remaining in a
tissue or organ at some time, t, after introduction. The retention function
is frequently defined as a sum of terms of the form F x exp(-XBt) where each
fraction, F, of a substance entering an organ or tissue has a corresponding
biological removal rate coefficient, XB, or equivalently a biological half-
life TB. TB and AB are related by the equation XB = (In 2)/TB.
For a radionuclide with a radiological decay constant, XR, the retention
function is modified by replacing each value of A.B with the quantity
Retention periods can be long for certain elements in some organs.
Under those conditions, the organ burden of a long-lived radionuclide may
include a significant fraction of all previous uptake of the radionuclide at
that site. For continuous intake, this process of bioaccumulation can provide
DRAFT 111-41
-------
an organ burden which increases significantly over a persons's entire
lifetime. The bioaccumulation is proportional to the convolution of intake
rate and retention.
The numerical values for F and TB are usually determined in studies
where there is a known intake of a radionuclide. The whole body or specific
organs or tissues are radioassayed to determine the amount of activity present
at various times after intake. These data are used to develop a retention
curve from which the numerical values of parameters can be obtained.
Metabolic studies to determine retention parameters are usually done in
laboratory animals; however, for some radionuclides there are metabolic data
from both humans and animals.
The EPA models differ from ICRP models in the treatment of radioactive
daughters produced in the body (Dunning et al. 1984). ICRP assumes all
radioactive progeny produced in the body follow the metabolic behavior of the
parent nuclide (ICRP 1979). EPA permits each daughter, as it is formed, to
assume the metabolic properties of that element. For example, calculating the
ionizing energy deposition in tissues associated with intake of alpha emitting
radionuclides involves metabolic models for all elements from curium to lead
and samarium, so that all progeny can be evaluated and energy deposition
within tissues summed to calculate the absorbed radiation dose.
Toxicokinetic models for all elements may be used since almost every
element has at least one beta or gamma emitting radioisotope. Elements from
hydrogen to curium are listed here.
DRAFT 111-42
-------
Hydrogen Model
The toxicokinetic model for hydrogen is taken from Killough and Rohmer
(1978).
1. Absorption from th.e intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
(a) Assume a daily intake of 350 g of 1H.
(b) Also assume the 3H/1H ratio in an organ is equivalent to that in
the daily intake.
(c) If the fractional weight of hydrogen in an organ is FH, then the
activity concentration of H in the organ (pCi per gram) is
(FH/350)(daily intake of UC in pCi).
Helium Model
There is no toxicokinetic model for helium in Sullivan et al. (1981).
Isotope half-lives are too short for internal dosimetry.
Lithium Model
There is no toxicokinetic model for lithium in Sullivan et al. (1981).
Isotope half-lives are too short for internal dosimetry.
DRAFT 111-43
-------
Beryl 1ium Model
The toxicokinetic model for beryllium is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f,) is 0.002.
2. Fractional tissue uptake (fl) is: bone, 0.32; liver, 0.10; kidneys,
0.03; spleen, 0.002; other tissues, 0.548.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
Spleen
Other tissues
1.0
1.0
1.0
1.0
1.0
450
270
120
540
180
Boron Model
There is no toxicokinetic model for boron in Sullivan et al. (1981).
Isotope half-lives are too short for internal dosimetry.
Carbon Model
For carbon-11 and carbon-15 in the case of carbon as C02, the following
is used:
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f^) is: bone, 0.008; other tissues, 0.992.
3. Retention parameters:
F TB
(Fraction of (Biological hal/-
Organ organ burden) 1 ife in days)'
Bone
Other tissues
*Data are not available at this time.
DRAFT 111-44
-------
For carbon-14, the specific activity model of Killough and Rohmer (1978)
is used:
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
(a) Assume a daily intake of 300 g of carbon.
(b) Also assume the UC/C ratio in an organ is equivalent to that in
the daily intake.
(c) If the fractional weight of carbon in an organ is Fc, then the
activity concentration of C in the organ (pCi per gram) is
(Fc/300)(daily intake of UC in pCi).
Nitrogen Model
The toxicokinetic model for nitrogen is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
All tissues 1.0 90
DRAFT 111-45
-------
Oxygen Model
The toxicokinetic model for oxygen is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
f TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues M) 14
Fluorine Model
There is no toxicokinetic model for fluorine in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for fluorine is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: bone, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone 1.0 »
Neon Model
There is no toxicokinetic model for neon in Sullivan et al. (1981).
DRAFT 111-46
-------
Sodium Model
The toxicokinetic model for sodium is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: bone, 0.30; other tissues, 0.70.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Or gan organ burden) 1 ife in days)
Bone
Other tissues
1.0
1.0
15
15
Magnesium Model
There is no toxicokinetic model for magnesium in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for magnesium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 0.50.
2. Fractional tissue uptake (f'2) is: bone, 0.4; other tissues, 0.4;
excretion, 0.2.
3. Retention parameters:
F . T8
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Other tissues
1.0
1.0
100
100
Aluminum Model
There is no toxicokinetic model for aluminum in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for aluminum is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: bone, 0.30; other tissues, 0.70.
3. Retention parameters:
DRAFT 111-47
-------
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Other tissues
1.0
1.0
100
100
Silicon Model
There is no toxicokinetic model for silicon in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for silicon is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues
0.40
0.60
5
100
Phosphorus Model
The toxicokinetic model for phosphorus is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.80.
2. Fractional tissue uptake (f'2) is: bone, 0.30; other tissues, 0.55;
excretion, 0.15.
3. Retention parameters:
F T8
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Other tissues
1.0
0.27
0.73
00
2
19
DRAFT 111-48
-------
Sulfur Model
The toxicokinetic model for sulfur is taken from Adams et al . (1978)
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 0.20; excretion,
0.80.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues
0.75
0.25
20
2000
Chlorine Model
There is no toxicokinetic model for chlorine in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for chlorine is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 1.0 10
DRAFT 111-49
-------
Argon Model
The toxicokinetic model for argon is taken from Bernard and Snyder
(1975).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'z) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues
0.885
0.092
0.021
1 . 5x10"^*
S.OxlO'4
5.3xlO'3
6.2xlO'4
5.7xlO'3
0.029
0.19
Potassium Model
The toxicokinetic model for potassium is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (fj) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 1.0 30
DRAFT 111-50
-------
Calcium Model
There is no toxicokinetic model for calcium in Sullivan et al . (1981).
The current RADRISK toxicokinetic model for calcium is taken from Adams et al
(1978).
1. Absorption from the intestine (f,) is 0.30.
2. Fractional tissue uptake (fl) is: cortical bone, 0.33; trabecular
bone, 0.32; other tissues, 0.35.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Cortical bone
Trabecular bone
Other tissues
0.30
0.15
0.55
0.34
0.16
0.50
0.57
0.29
0.14
10
1100
10500
10
700
2900
1.5
45
100
Scandium Model
The toxicokinetic model for scandium is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f^ is 0.0001.
2. Fractional tissue uptake (f'2) is: bone, 0.20; liver, 0.15; kidneys,
0.02; other tissues, 0.63.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Kidneys
Other tissues
1.0
1.0
1.0
1.0
33
36
75
30
DRAFT 111-51
-------
Titanium Model
There is no toxicokinetic model for titanium in Sullivan et al. (1981)
The ICRP (1981) toxicokinetic model for titanium is listed below.
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: All tissues, 1.0.
3. Retention parameters:
f TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues . T.O 600
Vanadium Model
There is no toxicokinetic model for vanadium in Sullivan et al. (1981).
The current RADRISK toxicokinetic model for vanadium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'z) is: bone, 0.25; other tissues, 0.05;
excretion, 0.70.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Other tissues
1.0
1.0
10,000'
10,000
Chromium Model
The toxicokinetic model for chromium is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: bone, 0.05; other tissues, 0.65;
excretion, 0.30.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife .in days)
DRAFT 111-52
-------
Bone 1.0 1000
Other tissues 0.62 6
0.38 80
Manganese Model
The toxicokinetic model for manganese is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: bone, 0.35; liver, 0.25; other
tissues, 0.40.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Other tissues
1.0
0.4
0.6
0.5
0.5
40
4
40
4
40
Iron Model
The toxicokinetic model for iron is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: liver, 0.08; spleen, 0.013; other
tissues, 0.907.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Liver
Spleen
Other tissues
1.0
1.0
1.0
2000
2000
2000
DRAFT 111-53
-------
Cobalt Model
The toxicokinetic model for cobalt is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (f'2) is: kidneys, 0.05; other tissues,
0.45; excretion, 0.50.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Kidneys
Other tissues
0.6
0.2
0.2
0.6
0.2
0.2
6
60
800
60
60
800
Nickel Model
The toxicokinetic model for nickel is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (fl) is: kidneys, 0.97; other tissues,
0.03.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Kidneys 1.0 1
Other tissues 1.0 10,000
DRAFT 111-54
-------
Copper Model
There is no toxicokinetic model for copper in Sullivan et al . (1981).
The current RADRISK toxicokinetic model for copper is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 0.5.
2. Fractional tissue uptake (f'2) is: liver, 0.1; brain, 0.1; pancreas,
0.006; other tissues, 0.794.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Liver
Brain
Pancreas
Other tissues
1.0
1.0
1.0
1.0
40
40
40
40
Zinc
The toxicokinetic model for zinc is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.50.
2. Fractional tissue uptake (f'2) is: bone, 0.20; other tissues, 0.80.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone 1.0 400
Other tissues 0.3 20
0.7 400
DRAFT 111-55
-------
Gallium Model
The toxicokinetic model for gallium is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f'2) is: bone, 0.30; liver, 0.25; spleen,
0.01; other tissues, 0.44.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Spleen
Other tissues
1.0
1.0
1.0
0.5
0.5
40
5
40
5
40
Germanium Model
There is no toxicokinetic model for germanium in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for germanium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f^ is 1.0.
2. Fractional tissue uptake (f'2) is: kidneys, 0.5; other tissues, 0.5.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Kidneys 1.0 0.2
Other tissues 1.0 1
DRAFT 111-56
-------
Arsenic Model
The toxicokinetic model for arsenic is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f,) is 0.03.
2. Fractional tissue uptake (f'2) is: liver, 0.03; kidneys, 0.01; other
tissues, 0.96.
3. Retention parameters:
(Fraction of (Biological half-
Organ organ burden) life in days)
Liver
Kidneys
Other tissues
1.0
1.0
1.0
550
550
280
Selenium Model
The toxicokinetic model for selenium is taken from Adams et al . (1978).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: bone, 0.10; liver, 0.20; kidneys,
0.10; other tissues, 0.60.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Kidneys
Other tissues
0.4
0.3
0.3
0.4
0.3
0.3
0.4
0.3
0.3
0.4
0.3
0.3
1
10
70
1
10
70
1
10
70
1
10
70
DRAFT 111-57
-------
Bromine Model
There is no toxicokinetic model for bromine in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for bromine is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 1_10 10
Krypton Model
The toxicokinetic model for krypton is taken from Bernard and Snyder
(1975).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 0.89 8.8xlO"5
0.09 l.OxlO'3
0.02 9.5xlO'3
2.9xlO"3 0.049
1.5xlO~3 0.321
DRAFT IH-58
-------
Rubidium Model
The toxicokinetic model for rubidium is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'z) is: bone, 0.25; other tissues, 0.75.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Other tissues
1.0
1.0
40
40
Strontium Model
The toxicokinetic model for strontium is taken from Adams et al. (1978)
and ICRP Publication 30 (1979).
1. Absorption from the intestine (f,) is 0.30.
2. Fractional tissue uptake (f'2) is: bone, 0.27; other tissues, 0.73.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Other tissues
0.393
0.0496
0.186
0.168
0.203
0.8
0.15
0.041
0.003
5
170
1100
2500
8800
1.8
30
200
. 1600
DRAFT 111-59
-------
Yttrium Model
The toxicokinetic model for yttrium is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.0001.
2. Fractional tissue uptake (f'2) is: bone, 0.50; liver, 0.15; other
tissues, 0.10; excretion, 0.25.
3. Retention parameters:
F. TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Other tissues
1.0
1.0
1.0
00
oo
30
Zirconium Model
The toxicokinetic model for zirconium is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.002.
2. Fractional tissue uptake (f'2) is: bone, 0.50; other tissues, 0.50.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone 1.0 8000
Other tissues 1.0 7
DRAFT 111-60
-------
Niobium Model
The toxicokinetic model for niobium is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: bone, 0.71; kidneys, 0.018;
spleen, 0.01; testes, 0.002; other tissues, 0.26.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Kidneys
Spleen
Testes
Other tissues
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
0.5
6
200
6
200
6
200
6
200
6
200
DRAFT 111-61
-------
Molybdenum Model
The toxicokinetic model for molybdenum is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.80.
2. Fractional tissue uptake (f'2) is: bone, 0.15; liver, 0.30; kidneys,
0.05; other tissues, 0.50.
3. Retention parameters:
F \
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
Other tissues
0.1
0.9
0.1
0.9
0.1
0.9
0.1
0.9
1
50
1
50
1
50
1
50
DRAFT 111-62
-------
Technetium Model
The toxicokinetic model for technetium is taken from Adams et al .
(1978).
1. Absorption from the intestine (f.,) is 0.80.
2. Fractional tissue uptake (f'2) is: liver, 0.08; kidneys, 0.01;
thyroid, 0.02; other tissues, 0.89.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Liver
Kidneys
Thyroid
Other tissues
0.76
0.20
0.04
0.76
0.20
0.04
0.76
0.20
0.04
0.76
0.20
0.04
1.6
3.7
22
1.6
3.7
22
1.6
3.7
22
1.6
3.7
22
Ruthenium Model
The toxicokinetic model for ruthenium is taken from Adams et al. (1978)
.1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (f'2) is: all tissues, 0.85; excretion,
0.15.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues
0.41
0.35
0.24
8
35
1000
DRAFT 111-63
-------
Rhodium Model
The toxicokinetic model for rhodium is taken from Adams et al. (1978)
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (f'2) is: all tissues, 0.85; excretion,
0.15.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
All tissues
0.41
0.35
0.24
8
35
1000
Palladium Model
There is no toxicokinetic model for palladium in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for palladium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 5xlO"3.
2. Fractional tissue uptake (f'2) is: bone, 0.07; liver, 0.45; kidneys,
0.15; other tissues, 0.03; excretion, 0.30.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Testes
Other tissues
1.0
1.0
1.0
1.0
15
15
15
15
DRAFT 111-64
-------
Silver Model
The toxicokinetic model for silver is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (f'2) is: liver, 0.80; other tissues, 0.20.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Liver
Other tissues
0.10
0.90
0.10
0.90
3.5
50
3.5
50
Cadmium Model
There is no toxicokinetic model for cadmium in Sullivan et al. (1981).
The current RADRISK toxicokinetic model for cadmium is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (fj is 0.05.
2. Fractional tissue uptake (f'2) is: liver, 0.30; kidneys, 0.30; other
tissues, 0.40.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Liver 1.0 9131
Kidneys 1.0 9131
Other tissues 1.0 9131
DRAFT 111-65
-------
Indium Model
The toxicokinetic model for indium is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.02.
2. Fractional tissue uptake (f^) is: bone, 0.30; liver, 0.20; kidneys,
0.07; spleen, 0.01; other tissues, 0.42.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
Spleen
Other tissues
1.0
1.0
1.0
1.0
1.0
00
00
00
00
00
Tin Model
The toxicokinetic model for tin is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (f'2) is: bone, 0.50; other tissues, 0.50.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone 1.0 50
Other tissues 1.0 50
DRAFT 111-66
-------
Antimony
The toxicokinetic model for antimony is taken from Adams et al. (1978),
1. Absorption from the intestine (f,) is 0.20.
2. Fractional tissue uptake (f'2) is: bone, 0.14; other tissues, 0.56;
excretion, 0.30.
3. Retention parameters:
(Fraction of (Biological half-
Organ organ burden) life in days)
Liver
Other tissues
1.0
1.0
20
20
Tellurium Model
The toxicokinetic model for tellurium is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.20.
2. Fractional tissue uptake (f'2) is: bone, 0.25; other tissues, 0.25;
excretion, 0.50.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone 1.0 5000
Other tissues 1.0 20
111-67
-------
Iodine Model
The toxicokinetic model for iodine is taken from the U.S. Nuclear
Regulatory Commission (1975).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: thyroid, 0.30; other tissues,
0.70.
3. Retention parameters:
F. TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Thyroid
Other tissues
0.05
0.95
0.996
-0.0725
0.0765
11.3
117
0.243
11.3
117
Xenon Model
The toxicokinetic model for xenon is taken from Bernard and Snyder
(1975).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.-
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 0.87 1.78xlO"4
0.088 2.1xlO'3
0.037 0.019
0.0051 0.097
0.0028 0.642
DRAFT 111-68
-------
Cesium Model
The toxicokinetic model for cesium is taken from ICRP Publication 30
(1979).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
All tissues 0.10 2
0.90 110
Barium Model
The toxicokinetic model for barium is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: bone, 0.60; other tissues, 0.30;
excretion, 0.10.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Other tissues
0.83
0.10
0.08
0.0385
0.0235
0.73
0.13
0.1
0.017
0.2
3.5
390
1400
5500
0.8
18
130
1000
DRAFT IH-69
-------
Lanthanide (Rare Earth) Models
The lanthanides are a group of 15 elements which have many similar
characteristics. They are modeled in 3 groups.
(A) Lanthanum Model
Cerium Model
Praseodymium Model
The toxicokinetic model for lanthanum, cerium and praseodymium is taken
from ICRP Publication 30 (1979).
1. Absorption from the intestine (f,) is 0.0003.
2. Fractional tissue uptake (f'z) is: bone, 0.20; liver, 0.60; spleen,
0.05; other tissues, 0.15.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone 1.0 3500
Liver 1.0 3500
Spleen 1.0 3500
Other tissues 1.0 3500
DRAFT 111-70
-------
(B) Neodvmium Model
Promethium Model
Samarium Model
Europium Model
Gadolinium Model
Terbium Model
The toxicokinetic model for neodymium, promethium, samarium, europium,
gadolinium and terbium is taken from Adams et. al. (1978).
1. Absorption from the intestine (f,) is 0.0001.
2. Fractional tissue uptake (f'2) is: bone, 0.45; liver, 0.45;
excretion, 0.10.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
1.0
1.0
3500
3500
(C) Dysprosium Model
Holmium Model
Erbium Model
Thulium Model
Ytterbium Model
Lutetium Model
There is no toxicokinetic model for dysprosium, holmium, erbium,
thulium, ytterbium or lutetium in Sullivan et al. (1981).
DRAFT 111-71
-------
Dysprosium
The current RADRISK toxicokinetic model for dysprosium is taken from
ICRP Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO"4.
2. Fractional tissue uptake (f'2) is: bone, 0.60; liver, 0.10; kidneys,
0.02; excretion, 0.28.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Kidneys
1.0
1.0
1.0
3500
3500
10
Holmium
The current RADRISK toxicokinetic model for holmium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO"4.
2. Fractional tissue uptake (f'z) is: bone, 0.40; liver, 0.40; pancreas/
0.05; excretion, 0.15.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ _ organ burden) _ 1 ife in days)
Bone 1.0 3500
Liver 1.0 3500
Pancreas 1.0 . 3500
DRAFT 111-72
-------
Erbium
The current RADRISK toxicokinetic model for erbium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO"4.
2. Fractional tissue uptake (f'2) is: bone, 0.60; liver, 0.05; other
tissues, 0.10; excretion, 0.25.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Other tissues
1.0
1.0
1.0
3500
3500
3500
Thulium
The current RAORISK toxicokinetic model for thulium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO~4.
2. Fractional tissue uptake (f'2) is: bone, 0.65; liver, 0.04; other
tissues, 0.10; excretion, 0.21.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone 1.0 3500
Liver 1.0 3500
Other tissues 1.0 3500
DRAFT 111-73
-------
Ytterbium
The current RADRISK toxicokinetic model for ytterbium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO"4.
2. Fractional tissue uptake (f^) is: bone, 0.5; liver, 0.03; kidneys,
0.02; spleen, 0.005; excretion, 0.445.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
Spleen
1.0
1.0
1.0
1.0
3500
3500
10
3500
Lutetium
The current RADRISK toxicokinetic model for lutetium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 3xlO"A.
2. Fractional tissue uptake (f'2) is: bone, 0.6; liver, 0.02; kidneys,
0.005; excretion, 0.375.
3. Retention parameters:
F T8
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
1.0
1.0
1.0
3500
3500
10
DRAFT 111-74
-------
Hafnium Model
The toxicokinetic model for hafnium is taken from ICRP Publication 2
(1959).
1. Absorption from the intestine (f,) is 0.0001.
2. Fractional tissue uptake (f^) is: bone, 0.15; liver, 0.45; kidneys,
0.02; spleen, 0.13; other tissues, 0.25.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Tantal
Bone
Liver
Kidneys
Spleen
Other tissues
urn Model
1.0
1.0
1.0
1.0
1.0
600
625
563
350
563
There is no toxicokinetic model for tantalum in Sullivan et al. (1981),
The current RADRISK toxicokinetic model for tantalum is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f'2) is: bone, 0.30; kidneys, 0.06; other
tissues, 0.64.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Kidneys
Other tissues
1.0
0.50
0.50
0.50
0.50
100
4
100
4
100
DRAFT 111-75
-------
Tungsten Model
The tungsten (wolfram) model is taken from ICRP Publication 2 (1959),
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: bone, 0.07; liver, 0.06; other
tissues, 0.87.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Other tissues
1.0
1.0
1.0
9
4
1
Rhenium Model
There is no toxicokinetic model for rhenium in Sullivan et al . (1981).
The current RADRISK toxicokinetic model for rhenium is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 0.80.
2. Fractional tissue uptake (f'z) is: thyroid, 0.04; stomach wall, 0.10;
liver, 0.03; other tissues, 0.83.
3. Retention parameters:
(Fraction of (Biological half-
TB
Organ
Thyroid
Stomach wall
Liver
Other tissues
organ burden)
1.0
0.75
0.20
0.05
0.75
0.20
0.05
0.75
0.20
0.05
1 ife in days)
0.5
1.6
3.7
22
1.6
3.7
22
1.6
3.7
22
DRAFT 111-76
-------
Osmium Model
There is no toxicokinetic model for osmium in Sullivan et al (1981)
The current RADRISK toxicokinetic model for osmium is taken from ICRP
Publication 30 (1980).
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: liver, 0.20; kidneys, 0.04;
spleen, 0.02; other tissues, 0.54; excretion, 0.20.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Liver
Kidneys
Spleen
Other tissues
0.2
0.8
0.2
0.8
0.2
0.8
0.2
0.8
8
200
8
200
8
200
8
200
Iridium Model
The iridium toxicokinetic model is taken from Furchner et al. (1971)
1. Absorption from the intestine (f,) is 0.01.
2. Fractional tissue uptake (f'2) is: liver, 0.20; kidneys, 0.04;
spleen, 0.02; other tissues, 0.54; excretion, 0.20.
3. Retention parameters:
F T8
(Fraction of (Biological half-
Organ organ burden) life in days)
Liver
Kidneys
Spleen
Other tissues
0.20
0.80
0.20
0.80
0.20
0.80
0.20
0.80
8
200
8
200
8
200
8
200
DRAFT 111-77
-------
Platinum Model
There is no toxicokinetic model for platinum in Sullivan et al. (1981)
The current RADRISK toxicokinetic model for platinum is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (fj is 0.01.
2. Fractional tissue uptake (fl) is: kidneys, 0.1; liver, 0.1;
spleen, 0.01; adrenals, 0.001; other tissues, 0.589; excretion,
0.2.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Kidneys
Liver
Spleen
Adrenals
Other tissues
0.95
0.05
0.95
0.05
0.95
0.05
0.95
0.05
0.95
0.05
8
200
8
200
8
200
8
200
8
200
Gold Model
There is no toxicokinetic model for gold in Sullivan et al. (1981). The
current RADRISK Toxicokinetic model for gold is taken from ICRP Publication 30
(1980).
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
All tissues 1.0 3
DRAFT 111-78
-------
Mercury Model
The mercury toxicokinetic model is taken from Adams et al . (1978).
1. Absorption from the intestine (f,) is 0.02.
2. Fractional tissue uptake (f'2) is: kidneys, 0.08; other tissues,
0.92.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ _ organ burden) life in days)
Kidneys 0.95 40
0.05 10,000
Other tissues 0.95 40
0.05 10,000
Thallium Model
Thallium metabolism has been studied in humans and animals. The model
is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.95.
2. Fractional tissue distribution (f'2) is: kidneys, 0.05 and other
tissues 0.95.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Kidneys 1.0 7
Other tissues 1.0 7
DRAFT 111-79
-------
Lead Model
The lead model used is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.20.
2. Fractional tissue distribution (f'2) is: bone 0.55; liver, 0.25;
kidney, 0.02; other tissues 0.18.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Kidneys
Other tissues
0.60
0.15
0.25
0.80
0.18
0.02
0.80
0.18
0.02
0.80
0.18
0.02
12
180
12000
12
180
12000
12
180
12000
12
180
12000
Bismuth Model
The bismuth model is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.05.
2. Fractional tissue uptake (fl) is: kidney, 0.40; other tissues,
0.60.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Kidneys
Other tissues
0.60
0.40
0.60
0.40
0.60
5
0.60
5
DRAFT 111-80
-------
Polonium Model
The polonium model is taken from ICRP 30 (1979).
1. Absorption from the intestine (f,) is 0.10.
2. Fractional tissue uptake (f^) is: liver, 0.10; kidney, 0.10;
spleen, 0.10; other tissues, 0.70.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Kidneys
Spleen
Other tissues
1.0
1.0
1.0
1.0
1.0
50
50
50
50
50
Astatine Model
The current RADRISK toxicokinetic model for astatine is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
All tissues 1.0 10
DRAFT 111-81
-------
Radon Model
The radon model used by EPA is based on human and animal studies using
radon and is generally supported by human and animal studies with inert gases
(Bernard and Snyder 1975).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters:
Organ
Bone
Other tissues
F
(Fraction of
organ burden)
0.70
0.30
0.874
0.0913
0.0198
0.00863
0.00612
TB
(Biological hal
life in days)
2.65xlO"4
00
2.65xlO"4
0.0031
0.0288
0.146
0.963
f-
Francium Model
The current RADRISK toxicokinetic model for francium is taken from ICRP
Publication 30 (1981).
1. Absorption from the intestine (f,) is 1.0.
2. Fractional tissue uptake (f'2) is: all tissues, 1.0.
3. Retention parameters: Due to the short half-life of
francium isotopes, francium is considered to transform in the
organ in which it is deposited.
DRAFT 111-82
-------
Radium Model
The radium model is based on human and animal studies of radium. The
radium model is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.20.
2. Fractional tissue uptake (f'2) is: bone, 0.46; other tissues, 0.54.
3.
4.
Bone
Other
Radon retention in bone is 0.
Retention parameters:
F
(Fraction of
Organ organ burden)
0.525
0.435
0.022
0.00875
0.013
tissues 0.16
0.54
0.11
0.046
0.009
30.
TB
(Biological half-
life in days)
0.023
3.6
1300
3500
9600
0.05
1.0
35
200
1400
Actinium Model
The actinium model is taken from Adams et al. (1978).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue distribution (f'2) is: bone 0.20; liver, 0.60;
spleen, 0.05; other tissues, 0.15.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Spleen
Other tissues
1.0
1.0
1.0
1.0
3500
3500
3500
3500
DRAFT 111-83
-------
Thorium Model
Thorium uptake and retention have been studied in humans and animals,
The thorium model is taken from ICRP 30 (ICRP 1979).
Absorption from the intestine (f,) is 2xlO'4.
1.
2.
Fractional tissue distribution (fl) is: bone 0.70; liver, 0.04;
other tissues, 0.16; excreta, 0.10.
3. Retention parameters:
Organ
(Fraction of
organ burden)
(Biological half-
life in days)
Bone
Liver
Other tissues
1.0
1.0
1.0
8000
700
700
Protactinium Model
The protactinium model is based on the models for transuranium elements
developed by EPA for "Proposed Guidance for Transuranium Elements in the
Environment" (EPA 1977).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f'?) is: bone, 0.45; liver, 0.45;
testes, 0.00035; ovaries, 0.00011; other tissues, 0.07; excretion,
0.03.
3. Retention parameters:
Organ
(Fraction of
organ burden)
(Biological half-
life in days)
Bone
Liver
Testes
Ovaries
Other tissues
1.0
1.0
1.0
1.0
1.0
36525
14610
14610
14610
36525
DRAFT
111-84
-------
Uranium Model
The toxicokinetic model for uranium is based on human and animal studies
of uranium administered intravenously or orally (ICRP 1979). It is taken from
the EPA models (Sullivan et al. 1981, Dunning et al. 1984, Begovich et al.
1981).
Absorption from the intestine (f,) is 0.05 to 0.20.
1.
2.
Fractional tissue uptake (f'2) is: bone, 0.22; kidneys, 0.12; other
tissues, 0.12; excretion, 0.54.
3. Retention parameters:
Organ
(Fraction of
organ burden)
(Biological half-
life in days)
Bone
Kidneys
Other tissues
0.9
0.1
0.996
0.004
0.996
0.004
20
5000
6
1500
6
1500
Neptunium Model
The neptunium model is based primarily on animal studies and taken from
the EPA model for transuranium elements (EPA 1977).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f'2) is: bone, 0.45; liver, 0.45;
testes, 0.00035; ovaries, 0.00011; other tissues, 0.07; excretion,
0.03.
3. Retention parameters:
Organ
(Fraction of
organ burden)
(Biological half-
life in days)
Bone
Liver
Testes
Ovaries
Other tissues
1.0
1.0
1.0
1.0
1.0
36525
14610
14610
14610
36525
DRAFT
111-85
-------
Plutonium Model
The plutonium model is based on extensive animal studies supported by
some studies of accidental exposure in humans. It is taken from the EPA model
for transuranium elements (EPA 1977).
1. Absorption from the intestine (f,) is 0.001; absorption from the
intestine (f,) for plutonium-239, plutonium-240 and plutonium-242
oxides only, is 0.0001.
2. Fractional tissue uptake (f'2) is: bone, 0.45; liver, 0.45;
testes, 0.00035; ovaries, 0.00011; other tissues, 0.07; excretion,
0.03.
3. Retention parameters:
Organ
Bone
Liver
Testes
Ovaries
Other tissues
(Fraction of
organ burden)
1.0
1.0
1.0
1.0
1.0
(Biological hal
1 ife in days)
36525
14610
14610
14610
36525
f-
DRAFT
111-86
-------
Americium Model
The americium model is based primarily on animal studies with some
support from studies of accidental exposure of humans. It is taken from the
EPA model for transuranium elements (EPA 1977).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f^) is: bone, 0.45; liver, 0.45;
testes, 0.00035; ovaries, 0.00011; other tissues, 0.07; excretion,
0.03.
3. Retention parameters:
TB
(Fraction of (Biological half-
Organ organ burden) 1 ife in days)
Bone
Liver
Testes
Ovaries
Other tissues
1.0
1.0
1.0
1.0
1.0
36525
14610
14610
14610
36525
Curium Model
The curium model is based primarily on animal studies and taken from the
EPA model for transuranium elements (EPA 1977).
1. Absorption from the intestine (f,) is 0.001.
2. Fractional tissue uptake (f^) is: bone, 0.45; liver, 0.45;
testes, 0.00035; ovaries, 0.00011; other tissues, 0.07; excretion,
0.03.
3. Retention parameters:
F TB
(Fraction of (Biological half-
Organ organ burden) life in days)
Bone
Liver
Testes
Ovaries
Other tissues
1.0
1.0
1.0
1.0
1.0
36525
14610
14610
14610
36525
DRAFT 111-87
-------
All of the models listed previously are used in the an overall computer
program, RADRISK, to calculate the concentration of a parent radionuclide and
all subsequent daughters in body tissues (Dunning et al . 1980, Sullivan et al .
1981).
The activity in an organ at any time is calculated as:
) Aiik + cllk (*; I]:! Bfj ZZ Ajr + Pik) 1 = 1 ...... , L
1k
where:
Ajlk = rate of change of activity of isotope i in compartment 1
organ k
Ljk = number of exponential terms in the retention function for
isotope i in organ k
Br = branching ratio of nuclide to nuclide i
X* = rate coefficient (time"1) for radiological decay of isotope
i
A.jllc = rate coefficient (time"1) for biological removal of isotope
i , from compartment 1 of organ k
CHk = fractional coefficient for nuclide i in the 1th compartment
of organ k
pjk = inflow rate of isotope i into organ k
The function is evaluated for a constant inflow rate per year, 1
pCi/year in most cases.
Retention is calculated in a similar manner (Sullivan et al . 1981):
_
where:
R-k(t) = retention of isotope i at time t after intake, in organ k.
Concentration of activity or integrated activity in an organ or tissue
can be calculated using these computer programs. Computed values are stored
DRAFT 111-88
-------
in a file and later passed to another program which calculates dose and dose
rate from the activity concentrations. Dosimetric estimates calculated in
this way are used with age-specific risk coefficients to calculate the risk
from a specified radionuclide intake. Further information on these
calculations can be found in Section IV and V.
111-89
-------
IV. MECHANISM OF TOXICITY
A. IONIZING ENERGY DEPOSITION
The potential for ionizing radiation to cause tissue damage and cancer
was established by 15 years after the discovery of x-rays (EPA 1989, Upton
1975). Radiogenic effects in humans and animals following exposure to
ionizing radiation from x-rays and gamma rays and several specific
radionuclides such as radium, radon, thorium, and plutonium have been studied
extensively (NAS 1972, 1988; UNSCEAR 1986, 1988).
The interaction of ionizing radiation with matter involves a number of
rapid radiochemical reactions. Absorption of energy from alpha and beta
particles induces ions and excited radicals in matter, while absorption of
energy from x- and gamma radiations by atoms results in the ejection of
electrons from the atoms. These ejected high speed electrons then produce
ionizations in the same manner as the alpha and beta particles.
Although the exact mechanisms of action of ionizing radiation are not
known with certainty, radiation injury is considered to be related to the
production of these ions and free radicals within the cells (Hobbs and
McClellan 1986). One mechanism by which radiogenic injury may occur is by the
interaction of radiation with cellular aqueous solutions resulting in the
formation of superoxide (02~) and hydrogen peroxide, and free radicals, such
as the hydroperoxyl radical (H02-) or the hydroxyl radical (OH"). These
reactive intermediates may then react with biologically important molecules in
the cell. Further chemical interactions can lead to bond scission, chain
scission, cross-linking, development of adducts, single and double strand
breaks, etc., in DNA and in nearby organic molecules. Radiation effects have
been known to occur with proteins, nucleic acids, lipids, and
carbohydrates.
The intracellular changes caused by radiation are numerous, varied and
complex and can be similar to those seen with other types of cellular injury.
The response of individual cells to radiation is variable and may depend on
such things as the cycle of the cell and the oxygenation status of the cell
DRAFT IV-1
-------
(Hobbs and McClellan 1986). The response of cells to radiation has been
extensively reviewed (Rubin and Casarett 1968, ICRP 1984, Mettler and Mosley
1985). Ionizing radiation delivered at a sufficiently high dose and dose rate
may result in cell death, which, if extensive, may result in the death of the
organism. At nonlethal doses and dose rate, cellular damage may undergo
either partial or total repair; however, unrepaired alterations may be
expressed as mutations or tumors at later times.
For the purposes of radiation protection, radiogenic effects may be
classified as stochastic or nonstochastic (ICRP 1984). Stochastic effects are
those for which the probability of occurrence of effect, and not its severity,
varies as a function of dose in the absence of a threshold (Hobbs and
McClellan 1986). The major stochastic effects are cancer and heritable
effects (mutations).
While the potential for carcinogenesis and mutagenesis has not been
investigated for every radionuclide, the results of studies that have been
performed are consistent with expected effects of ionizing radiation. When
exposed to levels of ionizing radiation exceeding those in the environment,
mammalian cells in culture have been transformed; chromosomal aberrations have
been observed in cultured peripheral lymphocytes; even activation, by frame
shift due to a single base deletion in an oncogene, has been reported (EPA
1989). Induction of enzymes of unscheduled DNA repair, perhaps signaling
error-prone repair, has also been noted (Tuschl et al. 1980, 1983; Olivieri et
al. 1984).
Some epidemiological studies have shown increased cancer mortality in
persons exposed to x-rays, gamma rays and various internal alpha emitters,
such as radon, radium, and thorium (EPA 1989). Additional support comes from
the studies of mice, rats, hamsters, guinea pigs, cats, dogs, sheep, cattle,
pigs and monkeys that have demonstrated increased incidence of cancer
following exposure to a source of ionizing radiation (EPA 1989).
DRAFT IV-2
-------
B. RADIONUCLIDE DOSIMETRY
Since ionizing radiation is known to induce cancer in many organs and
tissues, risk estimates for internally deposited radionuclides can be
calculated from the distribution of the radionuclides in tissues and the
ionizing radiation dose associated with that distribution. The toxicokinetic
(metabolic) models used to estimate distribution have been listed in Section
III, Part E. As explained there, the radionuclide concentration in various
organs is calculated for a lifetime intake of 1 pCi/year. For all organs or
tissues where there is no organ-specific model, the activity in the organ
compartment "other" is considered to be uniformly distributed in all body
tissues not specified.
The absorbed radiation dose rate is calculated as follows:
D5(X;t) = £k=1 Dj(X-Yk;t)
where:
D,-(X;t) = Absorbed dose rate to organ X at time t due to radionuclide
i in source organs Y,, Y2 YM.
D,(x-Yk;t) = S, (X-Yk) Aik(t)
and
A,-k(t) = The activity at time t of radionuclide i in source organ Yk
S,(X-Yk) = The S factor
S,(X-Yk) = c L, f. Em*m(X-Yk)
c = A constant that depends on units of dose, energy and time being
used
and there is a summation across all events m of:
fm = Intensity of decay event (number per disintegration)
Em = Average energy of decay event in MeV
'UX-Yk) = Specific absorbed fraction, i.e., the fraction of emitted
energy from source organ Yk absorbed by target organ X per
gram of X
DRAFT IV-3
-------
The S factor is similar to the SEE factor used by the ICRP in ICRP
Publication 30. Both S factor and SEE are dosimetric factors to account for
the fraction of energy absorbed in a target organ from emission of radiation
from a source in another organ, for example, irradiation of a kidney by a
radionuclide deposited in the bone. However, the SEE factor includes a
quality factor for the type of radiation emitted during the transformation
(disintegration) (EPA 1989, Dunning et al . 1984).
Summation across all organs and tissues is used to estimate the absorbed
dose rate in target organs due to one unit of activity of radionuclide i
distributed in source organs Y1 ---- Yk.
The expression for this dose rate is:
) - ZkL AJk(t) Sim(X-Yk).
Alpha and beta particles are not usually energetic enough to produce
significant cross-irradiation terms, so m(X-Y) = 0. The absorbed fraction
for alpha and beta particles [4>m (X-X)j is assumed to be the inverse of the
mass of organ X (EPA 1989). Absorbed fractions for beta particles in skeletal
tissues are taken from ICRP Publication 30 (1980). In the case of alpha
particles special calculations for active bone marrow and endosteal cells in
bone are based on the method of Thorne (1977).
A simplified model of the dose rate estimate would be (EPA 1989):
D,.(t) = elf, (f;)E [l-e-x'M
mXE
where:
D(t) = organ absorbed dose rate (rad/day)
c = proportionality constant ([51.2 x 10"6 g rad] Ci~1 Mev"1 d"1)
I = radionuclide intake rate (Ci/day)
f, = fraction of ingested activity transferred to the blood
DRAFT IV-4
-------
f'2 = fraction of blood activity transferred to the target organ
E = energy absorbed by the target organ for each radioactive
transformation
m = target organ mass (grams)
AE = elimination constant (day"1) ln2/TE or 0.693/TE
and
TR TB
TE = ^-4-
TB = biological removal half-time
TR = radiological decay half-time
Such a simplified estimate does not include cross-irradiation terms from
radionuclides deposited in other nearby organs. These cross-irradiation terms
are included in the detailed calculation of radiation dose (S factors)
performed by the RADRISK program employed by EPA for calculating radiation
dose.
There are some differences between the EPA model and the ICRP models as
noted earlier. The EPA calculations carry forward high-LET and low-LET dose
rates as separate entities, whereas the ICRP combines high- and low-LET dose
rate through the use of a quality factor for high-LET. The ICRP calculations
treat radioactive progeny in the body as following the metabolic behavior of
the parent nuclide, whereas the EPA model assumes each of the progeny has the
metabolic behavior of the elements specific to the progeny. The assumption
can have significant impact on the dose estimate (Dunning et al. 1884).
EPA elected to assign the metabolic behavior of progeny on the basis of
the elemental form of the progeny. This is clearly appropriate for soft
tissues where the elemental form of the progeny will govern their metabolic
behavior in the aqueous phase in tissues. In skeletal tissues, however, the
progeny may be born and trapped in the crystal lattice of the hydroxyapatite.
They would, in this situation, continue to have the metabolic characteristics
more like the parent radionuclide (Neuman and Neuman 1958, McLean and Budy
IV-5
-------
1964). That portion of the progeny born in skeletal tissues outside the
crystal lattice would be expected to behave as their elemental characteristics
direct them. When radium transforms to radon-222 within a crystal, 70 percent
of the radon-222 escapes from the crystal (McLean and Budy 1964); very little
of other radon isotopes born in the bone crystal can escape before decaying
radiologically.
In estimating radiation risk, EPA generally uses the risk coefficients
derived from exposure to low-LET radiation for all tissues except bone, liver,
and inhaled radon and radon daughters. Bone and liver risk coefficients are
derived from high-LET, radium and Thorotrast studies (MAS 1980). Radon and
radon daughter risk coefficients are derived from studies of underground
miners (NAS 1988). The dosimetry for external x-rays and gamma rays is well
established and has minimal uncertainty due to the nature of the radiation.
The risk coefficients times the calculated radiation absorbed dose delivered
by internally deposited radionuclides yield the risk. The radiation absorbed
dose for internally deposited radionuclides, as noted above, is based on
pharmacokinetic models and radiation dosimetric calculations.
The pharmacokinetic models assume uniform distribution of radionuclides
in organ or tissues. While many studies have shown uneven distribution of
radionuclides in tissues, there are not sufficient data on any radionuclide to
establish a detailed pharmacokinetic model using non-uniform tissue or organ
nuclide distribution. To the extent that the pharmacokinetic model cannot
describe the microscopic distribution of deposited radionuclide within organs,
the dose calculated and the risk estimated will be imprecise.
Epidemiologic studies of ingested or injected radionuclides usually
involve activity concentrations far in excess of expected environmental
exposures. A prime example of this situation can be found in epidemiological
studies of persons exposed to radium-226 and radium-228 as dial painters or
those injected with radium-226.
The pharmacokinetic model estimated for radium, but taken from studies
of persons injected with radium, differs appreciably from a model estimate
based on environmental exposure to radium (Stehney and Lucas 1956). In
DRAFT IV-6
-------
addition, gross pathology, including peritrabecular fibrosis and
osteoradionecrosis with osteoporosis and osteosaphyrosis (softening and
fragility of bone), has been noted in many species, including humans,
following elevated doses of radiation to bone (Sharpe 1976, Lloyd and Henning
1983, Ackerman and Spjut 1962, Sikov and Mahlum 1976, Wilson et al. 1976, Jee
et al. 1969, Taylor and Bensted 1969, Clarke 1962). The peritrabecular
fibrosis may occur within about 150 days post-exposure, following an abortive
attempt by osteoblasts to lay down bone (Jee et al. 1969). At the same time,
the microvasculature of bone and bone marrow are disrupted. Jee (1971)
reported a 50 percent destruction of marrow vasculature within days of
injection of 10 ^Ci radium-226/kg. The extensive pathology reported by Jee
(1971) following radiation exposure of bone demonstrates that metabolism and
physiological responses in bone must be considered abnormal.
The circumstances influencing pharmacokinetics, physiology and pathology
must be carefully evaluated for epidemiological studies of internally
deposited radionuclides. In particular, it is questionable whether bone with
abnormal physiology and metabolism is suitable for estimating pharmacokinetics
and whether epidemiology of persons with such internal pathology can be
compared to environmental situations. The results of epidemiological studies
where atypical pathological conditions are prevalent must be regarded with
extreme caution. Because data are available on health effects following
exposure to lower doses of penetrating radiation (x- or gamma rays), risks at
environmental exposure levels are, in some cases, better estimated using risk
coefficients derived from penetrating radiation and pharmacokinetics models
than by applying data from epidemiological studies of persons exposed to
extremely high activity concentrations of radionuclides.
For many radionuclides there is no direct evidence for carcinogenesis in
humans. Often, for a specific radionucl ide, there is little evidence in
animal studies. Nevertheless, it must be presumed that such internally
deposited radionuclides can cause cancer since they deposit ionization energy
in tissues which is no different in character than that which has been shown
to be carcinogenic. In such cases, it is prudent to base risk estimates on
dosimetry models and risk coefficients derived from other types of exposures
to ionizing radiation.
DRAFT IV-7
-------
Risk estimates for internally deposited radionuclides should be based on
positive epidemiological studies where those are available, but care must be
taken in extrapolating to low doses; in particular, tissue damage associated
with high dose levels may distort the results. Where direct epidemiological
data are lacking, risk estimates can be calculated using the dosimetric
method. Where epidemiological studies exist but show no statistical excess, a
careful analysis should be conducted which considers the sensitivity and
possible shortcomings of the epidemiologic studies as well as uncertainties in
the dosimetric analysis. However, even in the absence of direct
epidemiological data, use of dosimetry models and risk coefficients is
reasonable since risk coefficients are based broadly on the fact that ionizing
radiation is a known carcinogen.
DRAFT IV-8
-------
V. QUANTIFICATION OF TOXICOLOGICAL EFFECTS
This introductory section summarizes the general approach used to
evaluate the hazard of contaminants in drinking water. This text appears in
all criteria documents to provide information relevant to the basic issue of
how toxicological effects are quantified for chemical toxicity assessment.
The quantification of toxicological effects of a chemical consists of
separate assessments of noncarcinogenic and carcinogenic health effects.
Chemicals which do not produce carcinogenic effects are believed to have a
threshold dose below which no adverse, noncarcinogenic health effects occur,
while carcinogens are assumed to act without a threshold.
A. NONCARCINOGENIC EFFECTS
Method for Quantification of Noncarcinoqenic Effects
In the quantification of noncarcinogenic effects, a Reference Dose (RfD)
(formerly called the Acceptable Daily Intake (ADI)) is calculated. The RfD is
an estimate of a daily exposure to the human population that is likely to be
without appreciable risk of deleterious health effects, even if exposure
occurs over a lifetime. The RfD is derived from a No-Observed-Adverse-Effect
Level (NOAEL), or Lowest-Observed-Adverse-Effect Level (LOAEL), identified
from a subchronic or chronic study, and divided by an uncertainty factor(s).
The RfD is calculated as follows:
D,n (NOAEL or LOAEL) , . . .
RfD = —- — = mg/kg bw/day
Uncertainty Factor(s)
Selection of the uncertainty factor to be employed in the calculation of
the RfD is based on professional judgment, while considering the entire data
base of toxicological effects for the chemical. In order to ensure that
uncertainty factors are selected and applied in a consistent manner, the
Office of Water (OW) employs a modification to the guidelines proposed by the
National Academy of Sciences (NAS 1977, 1980) as follows:
DRAFT V-l
-------
• An uncertainty factor of 10 is generally used when good chronic,
or subchronic human exposure data identifying a NOAEL are
available and are supported by good chronic, or subchronic
toxicity data in other species.
• An uncertainty factor of 100 is generally used when good chronic
toxicity data identifying a NOAEL are available for one or more
animal species (and human data are not available), or when good
chronic, or subchronic toxicity data identifying a LOAEL in humans
are available.
• An uncertainty factor of 1,000 is generally used when limited or
incomplete chronic, or subchronic toxicity data are available, or
when good chronic, or subchronic toxicity data identify a LOAEL,
but not a NOAEL for one or more animal species are available.
The uncertainty factor used for a specific risk assessment is based
principally on scientific judgment rather than scientific fact and accounts
for possible intra- and interspecies differences. Additional considerations
not incorporated in the NAS/OW guidelines for selection of an uncertainty
factor include the use of a less than lifetime study for deriving a RfD, the
significance of the adverse health effect and the counterbalancing of
beneficial effects.
From the RfD, a Drinking Water Equivalent (DWEL) can be calculated. The
DWEL represents a medium specific (i.e., drinking water) 1ifetime exposure at
which adverse, noncarcinogenic health effects are not anticipated to occur.
The DWEL assumes 100% exposure from drinking water. The DWEL provides the
noncarcinogenic health effects basis for establishing a drinking water
standard. For ingestion data, the DWEL is derived as follows:
DWEL , (RfD x (Body Weight in kg) = mg/|_
Drinking Water Volume in L/day
DRAFT V-2
-------
where:
Body weight = assumed to be 70 kg for adult
Drinking water volume = assumed to be 2 liters per day for an adult
In addition to the RfD and the DUEL, Health Advisories (HAs) for
exposures of shorter duration (One-day, Ten-day and Longer-term) are
determined. The HA values are used as informal guidance to municipalities and
other organizations when emergency spills or contamination situations occur.
The HAs are calculated using a similar equation to the RfD and DWEL; however,
the NOAELs or LOAELs are identified from acute or subchronic studies. The HAs
are derived as follows:
HA _ (NOAEL or LOAEL) x (bw) _
(UF) x (L/day)
Using the above equation, the following drinking water HAs are developed
for noncarcinogenic effects:
1. One-day HA for a 10-kg child ingesting 1 L water per day.
2. Ten-day HA for a 10-kg child ingesting 1 L water per day.
3. Longer-term HA for a 10-kg child ingesting 1 L water per day.
,4. Longer-term HA for a 70-kg adult ingesting 2 L water per day.
The One-day HA calculated for a 10-kg child assumes a single acute
exposure to the chemical and is generally derived from a study of less than
7 days duration. The Ten-day HA assumes a limited exposure period of 1 to
2 weeks and is generally derived from a study of less than 30-days duration.
A Longer-term HA is derived for both the 10-kg child and a 70-kg adult and
assumes an exposure period of approximately 7 years (or 10% of an individual's
lifetime). A Longer-term HA is generally derived from a study of subchronic
duration (exposure for 10 percent of animal's lifetime).
DRAFT V-3
-------
Quantification of Noncarcinoqenic Effects
There is little evidence for noncarcinogenic effects of exposure at
environmental levels to the beta or gamma emitting radionuclides of interest
in this document (see section III). Numerous noncarcinogenic effects such as
hematopoietic effects, atrophy of various tissues, and reproductive effects
occur in laboratory animals only at doses exceeding the potential dose that
might result from drinking water containing these radionuclides. Therefore,
EPA considers it inappropriate to derive a Reference Dose for these beta or
gamma emitting radionuclides in drinking water.
B. CARCINOGENIC EFFECTS
Method for Quantification of Carcinogenic Effects
The EPA categorizes the carcinogenic potential of a chemical, based on
the overall weight-of-evidence, according to the following scheme:
• Group A: Human Carcinogen. Sufficient evidence exists from
epidemiology studies to support a causal association between
exposure to the chemical and human cancer.
t Group B: Probable Human Carcinogen. Sufficient evidence of
carcinogenicity in animals with limited (Group Bl) or inadequate
(Group B2) evidence in humans.
• Group C: Possible Human Carcinogen. Limited evidence of
carcinogenicity in animals in the absence of human data.
• Group D: Not classified as to Human Carcinogenicitv. Inadequate
human and animal evidence of carcinogenicity or for which no data
are available.
• Group E: Evidence of Noncarcinogenicitv for Humans. No evidence
of carcinogenicity in at least two adequate animal tests in
DRAFT V-4
-------
different species or in both adequate epidemiologic and animal
studies.
If toxicological evidence leads to the classification of the contaminant
as a known, probable or possible human carcinogen, mathematical models are
used to calculate the estimated excess cancer risk associated with the
ingestion of the contaminant in drinking water.
Quantification of Cancer Risk for Chemicals
For chemicals, the data used in these estimates usually come from
lifetime exposure studies in animals. In order to predict the risk for humans
from animal data, animal doses must be converted to equivalent human doses.
This conversion includes correction for noncontinuous exposure, less than
lifetime studies and for differences in size. The factor that compensates for
the size difference is the cube root of the ratio of the animal and human body
weights. It is assumed that the average adult human body weight is 70 kg and
that the average water consumption of an adult human is 2 liters of water per
day.
For contaminants with a carcinogenic potential, chemical levels are
correlated with a carcinogenic risk estimate by employing a cancer potency
(unit risk) value together with the assumption for lifetime exposure via
ingestion of water. The cancer unit risk for chemicals is usually derived
from a linearized multistage model with a 95 percent upper confidence limit
providing a low dose estimate; that is, the true risk to humans, while not
identifiable, is not likely to exceed the upper limit estimate and, in fact,
may be lower. Excess cancer risk estimates may also be calculated using other
models such as the one-hit, Weibull, logit and probit. There is little basis
in the current understanding of the biological mechanisms involved in
chemically-caused cancer to suggest that any one of these models is able to
predict risk more accurately than any others. Because each model is based
upon differing assumptions, the estimates which were derived for each model
can differ by several orders of magnitude.
DRAFT V-5
-------
The scientific data base used to calculate and support the setting of
cancer risk rate levels for chemicals has'an inherent uncertainty due to the
systematic and random errors in scientific measurement. In most cases, only
studies using experimental animals have been performed. Thus, there is
uncertainty when the data are extrapolated to humans. When developing cancer
risk rate levels, several other areas of uncertainty exist, such as the
incomplete knowledge concerning the health effects of contaminants in drinking
water, the impact of the experimental animal's age, sex and species, the
nature of the target organ system(s) examined and the actual rate of exposure
of the internal targets in experimental animals or humans. Dose response data
usually are available only for high levels of exposure, not for the lower
levels of exposure closer to where a standard may be set. When there is
exposure to more than one contaminant, additional uncertainty results from a
lack of information about possible synergistic or antagonistic effects.
For radionuclides, human epidemiologic data rather than animal
experiments form the basis of the cancer risk rate levels, which are derived
as best estimates rather than as 95% upper confidence limits of the unit risk
from a linearized multistage model. The true risks to humans may be higher or
lower than the predicted risks, but the overall uncertainty is probably less
than an order of magnitude. Because human data are used, the individual sites
of cancer are predicted, as well as the total risk. Therefore, projections
can be made both of cancer incidence and of cancer fatality, which are related
for a given organ to site-specific survival data, which ranges from 90%
survival (10% mortality) for thyroid cancer to 0% survival (100% mortality)
for liver cancer (EPA 1989).
Quantification of Carcinogenic Effects
Organ Doses and Risks from Ingestion of Beta and Photon Emitters in Drinking
Water Based on the RADRISK Model
The risk calculations in RADRISK are based on annual dose rates. For
this purpose, the dose rates at specific ages are computed for an annual unit
intake of a radionuclide (Sullivan et al. 1981, Dunning et al. 1980, EPA
1989). The calculated dose rate in a given year for continuous intake is
DRAFT V-6
-------
numerically equal to the integrated lifetime dose for the cumulative intake
until that year (ICRP 1971). For illustration, Table V-l lists the 50th year
and the 70th year doses (integrated doses) for intake of beta and gamma
emitters. (A 50 year period is standard for occupational exposure and 70
years is recommended by ICRP for environmental exposures.)
As noted earlier, the RADRISK model differs from ICRP models in that no
quality factors are used. Instead, separate dose calculation files are
maintained for each organ, for high-LET and for low-LET radiations.
The beta and gamma emitters are a diverse group of radionuclides
including examples from all elements. Due to the large number of beta and
gamma emitters only selected examples are given. Isotopes selected are:
tritium, a ubiquitous radionuclide arising from both man-made and natural
sources; cesium-137, a nuclide in both reactor effluents and weapons fallout;
strontium-90, another nuclide in reactor effluents and fallout; iodine-131, a
nuclide present in reactor effluents and fallout but also used in medical
practice and some laboratory tests; and lead-210, one of the long-lived
daughters of radon-222.
DRAFT V-7
-------
Table V-l
50-year Committed Absorbed Dose per Unit Intake
(millirad/pCi) from Beta and Gamma Emitters in Drinking Water
Orqan
Pulmonary
Lung
Stomach
Intestine
Kidneys
Liver
Breast
Pancreas
Red Marrow
Endosteum
Thyroid
All other
LET
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Lead-210
Isotope
Tritium-3*
2.400xlO"6(2.471xlO~6)** 8.360xlO"8
2.388xlO'5(2.475xlO'5)
1.380xlO"6(1.416xlO~6)
2.377xlO"5(2.463xlO'5)
9.948xlO~6(9.982xlO"6)
2.377xlO'5(2.464xlO~5)
5.008xlO'5(5.156xlO'5)
5. 192xlO'4(5. 382xlO~4)
1.081xlO'4(1.113xlO'4)
1.119xlO'3(1.160xlO~3)
2.389xlO'6(2.461xlO'6)
2.388xlO'5(2.475xlO'5)
2.397xlO'6(2.468xlO'6)
2.388xlO'5(2.475xlO'5)
5.155xlO'4(5.366xlO~4)
2.597xKT4(2.708xlO"4)
1.036xlO"3(1.079xlO'3)
3.962xlO"3(4.133xlO"3)
2.359xlO~6(2.429xlO'6)
2.388xlO'5(2.475xlO"5)
2.397xlO'6(2.468xlO'6)
2.388xlO'5(2.475xlO"5)
l.OSOxlO'7
1.187xlO"7
8.560xlO"8
8.280xlO"8
8.300xlO"8
8.060xlO'8
8.260xlO'8
6.560xlO'8
8.280xlO'8
8.300xlO'8
Iodine-131*
3.670xlO'7
1.091xlO'6
1.098xlO'6
1.745xlO"7
1.843xlO"7
4.468xlO"7
2.135xlO"7
3.506xlO"7
2.877xlO'7
1.670xlO"3
2.135xlO~7
* Committed absorbed dose in year 50 and year 70 are the same.
** Committed absorbed dose in the 70th year, otherwise doses in the 50th year
and 70th year are the same.
DRAFT
V-8
-------
Table V-l (continued)
50-year Committed Absorbed Dose per Unit Intake
(millirad/pCi) from Beta and Gamma Emitters in Drinking Water
Oroan
Pulmonary Lung
Stomach
Intestine
Kidneys
Liver
Breast
Pancreas
Red Marrow
Endosteum
Thyroid
All other
LET
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Low
Isotope
Strontium-90
1.776xlO"7
9.653xlO"7
4.793xlO'5
1.776xlO"7
1.904xlO'7
1.776xlO'7
1.776xlO"7
2.167xlO'5(2.170xlO'5)**
4.794xlO~5(4.925xlO'5)**
1.776xlO'7
1.776xlO"7
Cesium-137*
4.438xlO"5
3.568xlO"5
2.928xlO'5
5.055xlO"5
4.786xlO'5
4.799xlO'5
4.440xlO"5
4.365xlO'5
3.052xlO"5
5.080xlO'5
4.440xlO'5
* Committed absorbed dose in year 50 and year 70 are the same.
** Committed absorbed dose in the 70th year, otherwise doses in the 50th year
and 70th year are the same.
DRAFT
-------
Isotopes in the natural radionuclide chains heavier than thallium, if
they start as beta and gamma emitters, often have an alpha emitter as one of
their progeny. Lead-210 is an example of such a situation in that lead-210
and its daughter, bismuth-210, are essentially beta and gamma emitters; while
polonium-210, the daughter of bismuth-210, is an alpha emitter. Lead-210,
therefore, has both low- and high-LET emissions associated with its decay.
RADRISK contains a bookkeeping section of the program that sums
intakes, calculates dose rates (see Table V-l), and uses the calculated dose
rates with risk coefficients (see Table V-2) to calculate lifetime risk by
organ for high- and low-LET radiations for continuous intake of 1 pCi/year
(EPA 1989). Since risk coefficients were based on cancer incidence/mortality
from the low-LET radiation, a relative biological effectiveness (RBE) of 8
should be used to estimate risk from high-LET radiation. That is, the risk
per low-LET radiation listed in Table V-2 should be multiplied by a factor of
8 to obtain a high-LET risk coefficient (EPA 1989).
Irradiation of bone marrow by low-LET radiation leads to leukemia in
humans and animals. Leukemia has also been reported in patients receiving
thorium oxide and has been associated with enriched uranium exposure in
animals. However, radium dial painters with huge radium deposits in bone have
not developed leukemias, but have developed bone sarcomas, sinus and mastoid
carcinomas. Nor have ankylosing spondylitic patients treated with radium-224
developed an appreciable number of leukemias.
In the case of internally distributed radioisotopes, the dose to
"sensitive" tissues is calculated from pharmacokinetic data and physics
dosimetry principles. Compared to external x- or gamma radiation which gives
a fairly uniform dose, internal emitters, particularly alpha emitters, give an
irregular, rather uncertain dose distribution. In bone, for example, target
DRAFT V-10
-------
Table V-2
Organ-Specific Lifetime Cancer Risks
Used in the RADRISK Model
from High-LET and Low-LET Irradiation
Organ
Pulmonary Lung
Stomach wall
Intestine
Kidneys
Liver
Breast
Pancreas
Red marrow
Bone surface
Thyroid
Esophagus
Lymphoma
All other tissue
LET
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Low
High
Fatality
7.0xlO'5
5.7xlO/4
4.6xlO'5
3.7xlO'4
2.3xlCT5
l.SxlO'4
l.SxlO'5
1.4xlO~4
B.OxlO'5
4.0xlO'4
5.5xl(T5
4.4xlO'4
3.5xlO'5
2.8xlO'4
4.5xl(T5
3.6xlO'4
2.5xlO'6
2.0xlO"5
6.4xlO'6
B.lxlO'5
9.1xlO"6
7.3xlO'5
1.4xlO"5
l.lxlO'4
1.9xlO'5
1.6xlO'5
Risk/Rad
Incidence
7.5xlO"5
6.0xlO'4
6.0xlO'5
4.8xlO"4
4.3xlO"5
3.4xlO"4
4.3xlO"5
3.4xlO'4
S.OxlO'5
4.0xlO"4
1.4xlO'4
l.lxlO'3
3.8xlO"5
3.1xlO"4
4.5xlO"5
3.6xlO"4
2.5xlO"6
2.0xlO"5
6.4xlO"5
S.lxlO'4
9-lxlO"6
7.3xlO'5
1.9xlO"5
l.SxlO'4
3.4xlO'5
2.7xlO"4
Adapted from EPA (1989) Tables 6-6 and 6-7.
DRAFT
V-ll
-------
cells for osteosarcoma induction are considered to be, and the dose estimated
for, cells up to 10 jim from bone surfaces.
For leukemia induction, the possible non-uniform distribution of the
target cells in bone marrow is neglected. Harley and Pasternack (1976)
calculated alpha dose rate for radium-226 and its daughters in uniformly
labelled bone in the tissues outside the mineral bone. Noting Lloyd's report
(1970) of a 5 jim thick layer of osteoid between mineralized bone surfaces and
cellular layers, Harley and Pasternack (1976) adopted a distance of 10 jim from
mineral bone as appropriate for calculating endosteal dose. On the other
hand, Scott (1967) indicated cell nuclei at a distance of 10 to 15 \im from
mineralized bone surface in rats, and James and Taylor (1971) calculated dose
out to a distance of 20 ^m from bone surface. ICRP (1968) recommended
calculating endosteal dose rate over a distance 5 to 10 \im from the trabecular
surface.
Mays and Tueller (1964) calculated doses for radium-226 out to 70 ^m
from bone surface and stated that the average dose rate for soft tissues from
0 to 10 urn was used because: (a) osteosarcoma induction appeared to occur in
cells adjoining bone surfaces and (b) dose-rates were maximum at bone
surfaces. They also reported Owen's (1963) suggestion that cells located over
1 cell-layer from bone surfaces were at risk. Vaugan (1972) had shown a 5 to
20 \im thick cell layer in rabbits and Lloyd and Henning (1983) demonstrated a
single cell layer 0.6 to ^8.1 \im thick in human bone.
Bone exemplifies the dosimetry problems associated with alpha emitting
isotopes. The short range of the alpha particle in tissue and the high rate
of deposition of energy make the distance from source to target cells a very
important consideration. The situation described for bone also applies to a
greater or lesser extent, for all tissues with a non-uniform distribution of
radioisotope. This complicates direct application of a risk coefficient
derived from gamma exposure with a dose calculation using a pharmacokinetic
model. It also increases the uncertainty in the risk assessment.
EPA does not list all beta and gamma emitters in the list of
radionuclides in drinking water. Radioisotopes with very short half-lives,
DRAFT V-12
-------
which are not members of decay chains, are not considered. They will not
exist long enough to be a hazard in drinking water. Those radionuclides that
the Agency deems important are tabulated in Table V-3.
DRAFT V-13
-------
Table V-3
Concentration of Beta and Gamma Emitters
In Drinking Water to Yield a Specific Risk
of Cancer and Death from Cancer for Lifetime Consumption
Ch = Concentration in water for 4 mrem/y Committed EDE
Cm = Concentration in water for mortality risk = IxlO"4
Ci = Concentration in water for incidence risk = IxlO"4
NUCLIDE
H-3
BE-7
N-13
C-ll
C-14
C-15
0-15
F-18
NA-22
NA-24
SI-31
P-32
P-33
S-35
CL-36
CL-38
K-40
K-42
CA-45
CA-47
SC-46
SC-47
SC-48
V-48
CR-51
MN-52
MN-54
MN-56
FE-55
FE-59
CO-57
CO-58
CO-58M*
CO-60
NI-59
NI-63
NI-65
CU-64
ZN-65
Ch(pCi/L)
6.09E+04
4.35E+04
1.52E+05
9.92E+04
3.20E+03
6.69E+06
4.95E+05
3.95E+04
4.66E+02
3.35E+03
1.02E+04
6.41E+02
1.87E+03
1.29E+04
1.85E+03
2.12E+04
3.02E+02
3.90E+03
1.73E+03
8.46E+02
8.63E+02
2.44E+03
7.66E+02
6.44E+02
3.80E+04
7.33E+02
2.01E+03
5.64E+03
9.25E+03
8.44E+02
4.87E+03
1.59E+03
6.49E+04
2.18E+02
2.70E+04
9.91E+03
8.81E+03
1.19E+04
3.96E+02
Cm(pCi/L)
5.64E+04
1.10E+05
7.69E+04
5.24E+04
3.28E+03
3.12E+06
2.36E+05
2.60E+04
4.39E+02
2.80E+03
1.34E+04
7.33E+02
4.32E+03
1.42E+04
1.73E+03
1.14E+04
2.89E+02
3.23E+03
2.33E+03
1.55E+03
2.09E+03
5.41E+03
1.74E+03
1.54E+03
7.76E+04
1.44E+03
2.67E+03
7.22E+03
1.04E+04
1.12E+03
5.09E+03
1.98E+03
9.96E+04
2.00E+02
3.58E+04
1.39E+04
1.13E+04
1.84E+04
3.73E+02
Ci(pCi/L)
3.55E+04
6.76E+04
5.81E+04
3.94E+04
2.14E+03
2.38E+06
1.80E+05
1.97E+04
2.86E+02
1.87E+03
8.83E+03
5.55E+02
3.46E+03
8.75E+03
1.08E+03
8.56E+03
1.81E+02
2.17E+03
1.96E+03
9.80E+02
1.23E+03
3.02E+03
1.03E+03
9.15E+02 .
4.60E+04
9.08E+02
1.82E+03
4.83E+03
7.05E+03
6.97E+02
3.34E+03
1.25E+03
6.01E+04
1.30E+02
2.22E+04
8.50E+03
7.51E+03
1.14E+04
2.27E+02
DRAFT
V-14
-------
Table V-3 (continued)
NUCLIDE
ZN-69
ZN-69M
GA-67
GA-72
GE-71
AS-73
AS-74
AS-76
AS-77
SE-75
BR-82
RB-82
RB-86
RB-87
RB-88
RB-89
SR-82
SR-85
SR-85M
SR-89
SR-90
SR-91
SR-92
Y-90
Y-91
Y-91M
Y-92
Y-93
ZR-93
ZR-95
ZR-97
NB-93M
NB-94
NB-95
NB-95M
NB-97
NB-97M
MO-99
TC-95
TC-95M
TC-96
Ch(pCi/L)
6.31E+04
4.22E+03
7.02E+03
1.19E+03
4.36E+05
7.85E+03
1.41E+03
1.06E+03
4.33E+03
5.74E+02
3.15E+03
4.36E+05
4.85E+02
5.01E+02
2.91E+04
5.27E+04
2.41E+02
2.83E+03
2.37E+05
5.99E+02
4.20E+01
2.16E+03
3.10E+03
5.10E+02
5.76E+02
1.32E+05
2.87E+03
1.20E+03
5.09E+03
1.46E+03
6.50E+02
1.05E+04
7.07E+02
2.15E+03
2.39E+03
2.35E+04
1.37E+06
1.83E+03
6.97E+04
3.12E+03
2.05E+03
Cm(pCi/L)
4.76E+04
7.53E+03
1.63E+04
2.44E+03
4.35E+05
1.12E+04
2.07E+03
1.83E+03
7.55E+03
4.64E+02
2.76E+03
2.06E+05
4.74E+02
7.66E+02
1.49E+04
2.75E+04
4.11E+02
3.98E+03
2.30E+05
9.99E+02
6.24E+01
3.70E+03
5.67E+03
1.11E+03
1.27E+03
1.23E+05
4.27E+03
2.33E+03
1.47E+04
3.44E+03
1.40E+03
2.33E+04
1.48E+03
5.08E+03
5.27E+03
2.20E+04
1.09E+06
1.66E+03
5.82E+04
2.79E+03
1.87E+03
Ci(pCi/L)
3.48E+04
4.47E+03
9.44E+03
1.46E+03
2.71E+05
6.75E+03
1.26E+03
1.08E+03
4.41E+03
3.06E+02
1.73E+03
1.57E+05
3.29E+02
5.65E+02
1.13E+04
2.07E+04
2.59E+02
2.54E+03
1.62E+05
6.46E+02
5.90E+01
2.27E+03
3.39E+03
6.13E+02
6.98E+02
8.68E+04
2.73E+03
1.36E+03
1.12E+04
1.99E+03
7.98E+02
1.30E+04
9.06E+02
3.02E+03
2.93E+03
1.56E+04
7.87E+05
1.14E+03
3.43E+04
1.09E+03
1.06E+03
V-15
-------
Table V-3 (continued)
NUCLIDE
TC-96M
TC-97
TC-97M
TC-99
TC-99M
RU-97
RU-103
RU-105
RU-106
RH-103M
RH-105
RH-105M
RH-106
PD-100
PD-101
PD-103
PD-107
PD-109
AG-105
AG-108
AG-108M
AG-109M
AG-110
AG-110M
AG-111
CD-109
CD-115
CD-115M
IN-113M
IN-114
IN-114M
IN-115
IN-115M
SN-113
SN-121
SN-121M
SN-125
SN-126
SB-122
Ch(pCi/L)
1.76E+05
3.25E+04
4.45E+03
3.79E+03
8.96E+04
7.96E+03
1.81E+03
4.99E+03
2.03E+02
4.71E+05
3.72E+03
5.51E+06
1.24E+06
1.30E+03
1.34E+04
6.94E+03
3.66E+04
2.12E+03
2.70E+03
6.26E+05
7.23E+02
1.67E+07
1.84E+06
5.12E+02
1.08E+03
2.27E+02
9.58E+02
3.39E+02
5.24E+04
9.76E+05
3.23E+02
3.51E+01
1.64E+04
1.74E+03
6.06E+03
2.26E+03
4.46E+02
2.93E+02
8.10E+02
Cm(pCi/L)
1.46E+05
2.29E+04
3.13E+03
2.61E+03
8.22E+04
1.91E+04
3.78E+03
8.36E+03
3.55E+02
3.87E+05
7.94E+03
4.34E+06
5.78E+05
3.36E+03
2.71E+04
1.60E+04
7.83E+04
4.28E+03
3.69E+03
2.98E+05
7.27E+02
7.77E+06
8.52E+05
5.46E+02
2.03E+03
2.39E+02
2.00E+03
4.66E+02
5.71E+04
4.57E+05
5.48E+02
6.21E+01
2.63E+04
3.89E+03
1.28E+04
5.48E+03
9.78E-t-02
5.89E+02
1.71E+03
Ci(pCi/L)
8.47E+04
1.28E+04
1.70E+03
1.44E+03
3.83E+04
1.14E+04
2.18E+03
5.20E+03
2.04E+02
2.81E+05
4.47E+03
3.10E+06
4.42E+05
1.97E+03
1.66E+04
8.78E+03
4.32E+04
2.45E+03
2.64E+03
2.28E+05
5.51E+02
5.94E+06
6.51E+05
4.08E+02
1.21E+03
1.25E+02
1.10E+03
2.46E+02
3.96E+04
3.50E+05
3.34E+02
4.81E+01
1.66E+04
2.25E+03
7.18E+03
3.56E+03
5.43E+02
3.48E+02
9.68E+02
DRAFT
V-16
-------
Table V-3 (continued)
NUCLIDE
SB-124
SB-125
SB-126
SB-126M
SB-127
SB-129
TE-125M
TE-127
TE-127M
TE-129
TE-129M
TE-131
TE-131M
TE-132
1-122
1-123
1-125
1-126
1-129
1-130
1-131
1-132
1-133
1-134
1-135
CS-131
CS-134
CS-134M
CS-135
CS-136
CS-137
CS-138
BA-131
BA-133
BA-133M
BA-137M
BA-139
BA-140
LA-140
CE-141
CE-143
CE-144
Ch(pCi/L)
5.63E+02
1.94E+03
5.44E+02
5.85E+04
8.18E+02
3.09E+03
1.49E+03
7.92E+03
6.63E+02
2.72E+04
5.24E+02
2.68E+04
9.71E+02
5.80E+02
2.11E+05
1.07E+04
1.51E+02
8.10E+01
2.10E+01
1.19E+03
1.08E+02
8.19E+03
5.49E+02
2.14E+04
2.34E+03
2.28E+04
8.13E+01
1.01E+05
7.94E+02
5.18E+02
1.19E+02
2.56E+04
2.95E+03
1.52E+03
2.62E+03
2.15E+06
1.38E+04
5.82E+02
6.52E+02
1.89E+03
1.21E+03
2.61E+02
Cm(pCi/L)
1.14E+03
3.86E+03
1.17E+03
3.58E+04
1.75E+03
5.32E+03
3.14E+03
1.44E+04
1.13E+03
2.39E+04
9.55E+02
2.37E+04
1.91E+03
1.25E+03
1.02E+05
1.59E+04
7.27E+02
3.87E+02
1.03E+02
1.76E+03
5.16E+02
7.98E+03
8.72E+02
1.43E+04
3.30E+03
2.15E+04
7.64E+01
6.79E+04
7.82E+02
4.76E+02
1.14E+02
1.39E+04
6.87E+03
2.47E+03
5.59E+03
1.08E+06
1.33E+04
1.29E+03
1.44E+03
4.22E+03
2.62E+03
5.68E+02
Ci(pCi/L)
6.77E+02
2.33E+03
6.98E+02
2.67E+04
9.97E+02
3.28E+03
2.27E+03
8.52E+03
8.72E+02
1.72E+04
6.14E+02
6.81E+03
5.13E+02
6.60E+02
7.75E+04
1.90E+03
7.46E+01
4.02E+01
1.04E+01
2.13E+02
5.37E+01
1.86E+03
9.36E+01
6.78E+03
4.57E+02
1.37E+04
4.66E+01
4.77E+04
4.90E+02
2.88E+02
7.01E+01
1.04E+04
4.11E+03
1.64E+03
3.15E+03
8.14E+05
9.37E+03
7.30E+02
8.32E+02
2.34E+03
1.47E+03
3.15E+02
DRAFT
V-17
-------
Table V-3 (continued)
NUCLIDE
PR-142
PR-143
PR-144
PR-144M
ND-147
ND-149
PM-147
PM-148
PM-148M
PM-149
SM-151
SM-153
EU-152
EU-154
EU-155
EU-156
GD-153
GD-159
TB-158
TB-160
DY-165
DY-166
HO-166
ER-169
ER-171
TM-170
TM-171
YB-169
YB-175
LU-177
HF-181
TA-182
W-181
W-185
W-187
RE-183
RE-186
RE-187
RE-188
OS-185
OS-191
OS-191M
Ch(pCi/L)
1.04E+03
1.17E+03
4.70E+04
1.12E+05
1.25E+03
1.17E+04
5.24E+03
5.05E+02
5.75E+02
1.38E+03
1.41E+04
1.83E+03
8.41E+02
5.73E+02
3.59E+03
6.00E+02
4.68E+03
2.76E+03
1.25E+03
8.15E+02
1.51E+04
8.30E+02
9.81E+02
3.64E+03
3.80E+03
1.03E+03
1.27E+04
1.83E+03
3.11E+03
2.55E+03
1.17E+03
8.42E+02
1.90E+04
3.44E+03
2.66E+03
5.04E+03
1.88E+03
5.82E+05
1.79E+03
2.46E+03
2.38E+03
1.43E+04
Cm(pCi/L)
2.17E+03
2.58E+03
2.69E+04
7.02E+04
2.80E+03
1.51E+04
1.10E+04
1.12E+03
1.37E+03
2.98E+03
2.82E+04
3.99E+03
1.33E+03
9.57E+02
6.60E+03
1.37E+03
1.09E+04
5.75E+03
2.58E+03
1.89E+03
1.89E+04
1.89E+03
2.08E+03
8.01E+03
7.05E+03
2.27E+03
2.81E+04
4.28E+03
6.86E+03
5.65E+03
2.71E+03
1.96E+03
3.57E+04
6.85E+03
5.31E+03
6.55E+03
1.50E+03
4.00E+05
1.58E+03
4.96E+03
5.15E+03
2.91E+04
Ci(pCi/L)
1.23E+03
1.41E+03
2.03E+04
5.26E+04
1.55E+03
l.OOE+04
6.32E+03
6.23E+02
7.92E+02
1.65E+03
1.75E+04
2.22E+03
9.23E+02
6.46E+02
4.31E+03
7.73E+02
6.30E+03
3.26E+03
1.62E+03
1.08E+03
1.27E+04
1.03E+03
1.16E+03
4.40E+03
4.24E+03
1.25E+03
1.57E+04
2.45E+03
3.80E+03
3.12E+03 .
1.54E+03
1.13E+03
2.18E+04
3.80E+03
3.09E+03
2.42E+03
7.01E+02
2.21E+05
6.32E+02
3.08E+03
2.89E+03
1.66E+04
DRAFT
V-18
-------
Table V-3 (continued)
NUCLIDE
OS-193
IR-190
IR-192
IR-194
PT-191
PT-193
PT-193M
PT-197
PT-197M
AU-196
AU-198
HG-197
HG-203
TL-202
TL-204
TL-207
TL-208
TL-209
PB-203
PB-209
PB-210
PB-211
PB-212
PB-214
BI-206
BI-207
BI-210
BI-214
FR-223
AC-228
TH-231
TH-234
PA-233
PA-234
PA-234M
NP-236
NP-238
NP-239
NP-240
NP-240M
PU-243
AM-242
Ch(pCi/L)
1.69E+03
1.01E+03
9.57E+02
1.04E+03
3.81E+03
4.61E+04
3.02E+03
3.40E+03
1.75E+04
3.66E+03
1.31E+03
5.76E+03
2.39E+03
3.84E+03
1.68E+03
4.00E+05
2.83E+05
3.58E+05
5.06E+03
2.53E+04
1.01E+00
1.28E+04
1.23E+02
1.18E+04
6.56E+02
1.01E+03
8.53E+02
1.89E+04
3.41E+03
3.27E+03
4.07E+03
4.01E+02
1.51E+03
2.56E+03
9.30E+05
5.96E+03
1.39E+03
1.68E+03
2.31E+04
1.74E+05
1.64E+04
3.83E+03
Cm(pCi/L)
3.59E+03
2.37E+03
1.97E+03
2.15E+03
8.74E+03
9.94E+04
6.59E+03
7.03E+03
2.38E+04
8.38E+03
2.82E+03
1.26E+04
5.07E+03
3.53E+03
1.65E+03
1.93E+05
1.40E+05
1.89E+05
9.64E+03
3.47E+04
3.55E+00
1.44E+04
3.83E+02
1.52E+04
1.50E+03
2.27E+03
1.94E+03
1.92E+04
1.20E+04
6.10E+03
8.67E+03
8.87E+02
3.42E+03
4.61E+03
4.36E+05
1.36E+04
3.18E+03
3.71E+03
2.11E+04
8.93E+04
2.73E+04
8.66E+03
Ci(pCi/L)
2.01E+03
1.41E+03
1.15E+03
1.22E+03
5.11E+03
5.44E+04
3.64E+03
3.98E+03
1.55E+04
5.05E+03
1.59E+03
7.15E+03
2.85E+03
2.15E+03
9.39E+02
1.47E+05
1.06E+05
1.42E+05
6.06E+03
2.27E+04
2.99E+00
1.06E+04
2.70E+02
1.09E+04
8.89E+02
1.31E+03
1.01E+03
1.42E+04
9.11E+03
3.81E+03
4.87E+03
4.87E+02
1.91E+03
2.84E+03
3.33E+05
8.32E+03
1.82E+03
2.07E+03
1.51E+04
6.78E+04
1.69E+04
5.34E+03
* M = metastable.
DRAFT
V-19
-------
VI. UNCERTAINTY ANALYSIS
The types of health effects expected following exposure to radiation in
general have been established based on human data following high level
exposure and supporting animal studies. Much of the uncertainty in estimating
risk from exposure to radiation arises in evaluation of the relatively small
doses to target organs that result from environmental exposures. The
remainder of this section briefly describes the assumptions and parameters
contributing uncertainty to the quantification of the toxicological effects of
beta and gamma emitting radionucl ides found in drinking water.
A. UNCERTAINTY IN ASSESSMENT OF NONCARCINOGENIC EFFECTS OF BETA AND GAMMA
EMITTING RADIONUCLIDES
It is assumed that noncancer health effects would not be of concern at
levels of exposure from beta and gamma emitting radionuclides in drinking
water. Noncancer health effects observed in animals and humans occur at high
levels of exposure and do not appear to be more sensitive indicators of
adverse health effects than cancer. Therefore, consideration of
carcinogenicity should be sufficiently protective against other health
effects.
B. UNCERTAINTY IN ASSESSMENT OF CARCINOGENIC EFFECTS OF BETA AND GAMMA
EMITTING RADIONUCLIDES
Assessment of the carcinogenicity of beta and gamma emitting
radionuclides is uncertain in several areas. These areas of uncertainty
include: parameters used in the metabolic model, including absorption,
distribution and dosimetry; the risk coefficients used for calculating
lifetime risk; and factors influencing both.
Uncertainty in Parameters Used in the Metabolic Model
One issue is the proper choice of the gastrointestinal absorption factor
(f,) for these radionuclides. Because the number of radionuclides considered
in this analysis is large, and the data used to derive f1 factors are highly
variable, the degree of confidence for the f, factors chosen also varies. In
assessing the uncertainty relating to f,, one should keep in mind that it is
the dose to target organs (or cells) that is of interest and not the f1 per
DRAFT VI-1
-------
se. The organ doses depend also on rates of radioisotope uptake from blood
(f'2) and of loss from the organ. In principle, one can partly bypass the
problem of determining these metabolic parameters and assess the dose to an
organ or tissue for a given intake by simply measuring the organ or tissue
burden under steady state conditions. However, an important limitation to
this approach is the difficulty in estimating past intake.
Estimates of body burden and the distribution of beta emitters among
organs vary widely (Dunning et al. 1984, Sullivan et al. 1981). Uncertainty
in organ burden influences estimates of total body burden and selection of
metabolic model parameters. For a specific radionuclide, the calculated body
burden varies with the fraction absorbed, the respective fraction transferred
to organs, and the retention of the radioisotope in these organs. If intake
and body burden are specified, then varying any metabolic parameter requires
that some other parameter(s) change in compensation to maintain the constant
body burden. Because of their relatively short range, the dosimetry for beta
particles depends on the precise localization of the radionuclide within the
body; in contrast, the dose from gamma rays is less sensitive to the detailed
distribution. Consequently, dosimetry for the latter is generally more
certain than for the former.
Uncertainty in Distribution of Isotope
Few of the beta emitting radionuclides have been well characterized with
respect to distribution in the human body. Distribution within the skeleton
has been reported in some studies (Stara et al. 1971, Pool et al. 1973).
Distribution between and within soft tissues has also been reported (Schober
and Hunt 1976, Moskalev 1968). Uncertainty concerning the distribution within
organs will influence dose and risk calculations.
Estimating the degree of uncertainty for modeled dose estimates for each
of the radionuclides is not only difficult, but is also somewhat arbitrary.
No model has been verified in man for any long-term exposure scenario, and the
data selected to establish the parameters used in the model may not be
representative of the population being evaluated (EPA 1989). Nevertheless,
EPA has characterized the potential magnitude of dosimetric.uncertainties for
DRAFT VI-2
-------
the four groups of radionuclides: 1) essential elements and their analogs,
2) inert gases, 3) well-studied toxic metals, and 4) other elements. The
section on Bioaccumulation and Retention (Section III.E) present models of
radionuclides for all four of these groups. EPA (1989) states that for the
essential elements, uncertainty in dose is a factor of two or less in major
critical organs, while the uncertainty for analogs is perhaps three or less in
the critical organs. Because the available information on kinetics of the
well-studied toxic metals varies widely among these compounds, the uncertainty
in dose for this class ranges from about 2 to 5, depending on the specific
radionuclide. The Group 4 radionuclides include those for which kinetics
information is limited and, therefore, EPA states that dose estimates may be
in error by at least an order of magnitude. EPA considers these estimates to
represent the Agency's best judgment but also states that this quantification
should be considered rough estimates.
Uncertainty in Dosimetric Calculations
While calculations of average dose deposited in a tissue can be made,
these may not accurately reflect the dose to target cells. This is
particularly true for some beta emitting nuclides due to the short range of
beta particles in tissue. For example, it is well known that radioactive
iodine distributes primarily to the thyroid. However, the degree to which
specific tissues of the thyroid are irradiated is not well known.
The pharmacokinetic models used assume uniform distribution of
radionuclides in organs or tissues. While many studies would indicate that
distribution is often non-uniform, there are not sufficient data on any
radionuclide to establish a detailed pharmacokinetic model using a non-uniform
distribution. This limitation of the models results in some degree of
uncertainty.
Uncertainty in Risk Coefficients
EPA (1989) has derived risk coefficients for low-LET radiation based on
data derived from atomic bomb survivors (NAS 1980). New data and analyses on
irradiated populations have recently become available (UNSCEAR 1988, NAS
DRAFT VI-3
-------
1990); however, the resulting central estimates of risk from uniform, whole
body irradiation are in reasonable agreement with that based on EPA's current
model (to within a factor of 2). For whole body irradiation, the uncertainty
in the risk estimate is thought to be about a factor of 3 or less, but larger
errors in the quantification of risk at very low doses and dose rates cannot
be excluded. In addition, the estimate of risk to specific irradiated target
tissues is often more uncertain than the estimate for whole body irradiation.
Hence, in cases where the radiation is concentrated in specific organs, the
uncertainty in risk may be larger.
Uncertainty in Other Factors Influencing Risk
When considering a specific individual rather than population (average)
risks, variability in physiological parameters must be considered. These will
include such factors as age, genetic makeup, gender, and diet. For example,
uptake of radioactive iodine into the thyroid is dependent on the amount of
stable iodine in an individual's diet (Pittman et al. 1970).
Conclusion
There are many uncertainties inherent is estimating the risk from
ingestion of beta and gamma emitting radionuclides in drinking water. These
include uncertainties in internal dosimetry and in the appropriate risk
factors for specific tissues. Overall, EPA regards its risk estimate as a
reasonable central estimate but emphasizes that the actual risk could be at
least a factor of 3 higher or lower.
DRAFT VI-4
-------
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