-------
The fractions absorbed from the stomach and large intestine are usually
considered negligible compared with f]_, the fraction from the small
intestine. The latter varies considerably depending on the radionuclide
and on the material to which it is attached. The variability in f^
probably represents the greatest uncertainty associated with the Gl tract
model for most radionuclides and will be further discussed in the next
section.
(2) Uncertainties Due to Parameter Variability
Parameters employed in internal dosimetric models are often
quantified by values that represent "best estimates" or "average" values
from parameter distributions and ignore the recognized variability among
individuals. As a result, there are limitations inherent in taking a
deterministic approach in applying "Reference Man" parameters to assess
the dose to individuals in the general population. When assessing the
uncertainties associated with using EPA dosimetric models, the
variability of such parameters as radionuclide intake rate, organ mass,
blood transfer factor, organ uptake rate, and biological half-times of
the ingested radionuclides must be considered. These parameters vary
among individuals in the general population primarily because of age and
sex differences. Other factors that contribute are biological,
environmental, and geographical differences.
In order to fully assess the uncertainty in model predictions of the
doses to organs due to parameter variability (assuming that the model
structure is correct), a parametric uncertainty analysis must be done.
This process involves taking frequency distributions of values for each
model parameter to produce a frequency distribution of model predictions
of the doses.
Parameter values used in radiological assessments are generally
taken from the literature. However, a wide range of reported values is
expected for some parameters, implying a large uncertainty in the
estimated dose.
As a numerical illustration, we will consider the uncertainties in
parameters associated with age variations for the simple case of chronic
Ingestion of a single radionuclide. The model used to define the
absorbed dose rate to a target organ X due to radioactivity located in
source organ Y^ can be expressed as:
D (X «• Y ;t) = c I £l
where:
«•
D (X «- Y ;t) - absorbed dose rate (rad/day);
E [l-exp(-\ik)]/mVXik
(24)
= radionuclide intake rate (Ci/day) ;
6-28
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El
f2
m
Mk
E
= fraction of ingested activity transferred to the blood;
= fraction of blood activity transferred to the organ;
= target organ mass (g);
= elimination constant (day"1);
= energy absorbed by the target organ for each radioactive
transformation;
= proportionality constant (51.2 x 106g rad Ci"1 MeV~1d~1.
For simplicity, we will consider the case where t is very large
compared to the biological half-life of the incorporated radionuclide,
so that the term in the bracket is approximately 1:
D (X «- Y ;t) = c I
f2 E/mV\ik
(25)
in addition, it is assumed that- the parameters remain constant
throughout the period of investigation.
Equation 25 is a simplified form of the actual equations used by EPA
to estimate the absorbed dose rates to target organs resulting from the
ingestion of radioactive material. It represents the absorbed dose rate
to a target organ from particulate radiation due to radioactivity that is
uniformly distributed in that organ (i.e., 4>(X«- Yk) = <1>(X<- X.)).
For this analysis, we will consider the chronic intake of iodine-131
assuming that it behaves metabolically the same as stable iodine. It is
further assumed that iodine is rapidly and almost completely absorbed into
the bloodstream following inhalation or ingestion. From the blood, iodine
enters the extracellular fluid and quickly becomes concentrated in the
salivary, gastric, and thyroid glands. It is rapidly secreted from the
salivary and gastric glands, but is retained in the thyroid for relatively
long periods.
The intake and metabolism of iodine have been reviewed extensively in
the literature. There are two principal parameter data bases to be used in
this analysis. The first is found in an article, published by Dunning and
Schwarz (Du81), in which the authors compiled and evaluated the variability
in three of the principal biological parameters contained in Equation 25: m,
Tl/2- and f2- Tne second is taken from a paper by Bryant (Br69),
which provides age-specific values for most of the model parameters. These
data will serve as a means of illustrating how parameter variabilities in the
above model can affect absorbed dose rate estimates to members of the general
population.
(a) Intake Rate, I
The amount of radioactive material taken into the body over a specified
period of time, by ingestion or inhalation, is expected to be proportional to
the rate of intake of food, water, or air by individuals, and it would depend
6-29
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on such factors as age, sex, diet, and geographical location. Therefore,
understanding the patterns of food intake for individuals in the population
is important in assessing the possible range of intake rates for
radionuclides.
Recent EPA studies were done to assess the daily intake rates of food
and water for individuals in the general population. These studies showed
that age and sex played an important role (Ne84). Age significantly affects
food intake rates for all of the major food classes and, with one exception,
subclasses. The relationships between food intake and age are, in most
cases, similar to growth curves; there is a rapid increase in intake at an
early stage of physical development, then a plateau is reached in adulthood,
followed by an occasional decrease after age 60.
When sex differences were significant, males, without exception,
consumed more than females. The study also showed that relative consumption
rates for children and adults depend on the type of food consumed. The amount
of radioactivity taken into the body per unit intake of food, air, and water
depends on its relative density (amount of radioactivity contained in the
material per unit volume). The most likely pathway to organs in the body for
the ingestion of radioactive iodine comes from drinking milk. According to
the above study, the daily intake rate for milk for children (under 1 yr) to
that of an adult (25 to 29 yr) for males, varied by a factor of 2. Thus, if
milk contains radioactive iodine, the absorbed dose rate to the thyroid due
to the milk intake rate alone would also vary by a factor of 2. The intake
rates for milk used in this analysis are 0.71/day and 0.51/day for the child
and adult, respectively.
(b) Transfer Fraction, f^
While uncertainty in fj is not an important consideration for iodine,
it can be very significant for other elements. Experimental studies suggest
that the fi value for some radionuclides may be orders of magnitude higher
in newborns than in adult mammals, with the largest relative changes with age
occurring for those nuclides with small adult fj values (Le83). For some
radionuclides there appears to be a rapid decrease in the fx -value in the
first year of life. This can be related to the change in diet during this
time period, which could affect both the removal rate from the small
intestine to the upper large intestine and the absorption rate from the small
intestine to the bloodstream. Studies have indicated that the wall of the
small intestine is a selective tissue and that absorption of nutrients is to
a large extent controlled by the body's needs (Le83). in particular, the
fraction of calcium or iron absorbed depends on the body's needs for these
elements, so the fj value for these elements and for related elements such
as strontium, radium, and barium (in the case of calcium) and plutonium (in
the case of iron) may change as the need for calcium or iron changes during
various stages of life.
In the case of some essential elements such as potassium and chemically
similar radioelements such as rubidium .and cesium, however, absorption into
the bloodstream is nearly complete at all ages, so that changes with age and
possible homeostatic adaptations in absorption are not discernible. The
fraction of a radioelement that is transferred to the blood depends on its
6-30
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chemical form, and wide ranges of values are found in the literature for
individuals who ingest the material under different working conditions. For
example, f^ values for uranium were found to range from 0.005 to 0.05 for
industrial workers, but a higher value of 0.2 is indicated by dietary data
from persons not occupational ly exposed (ICRP79). EPA has used the 0.2 value
for uranium ingest ion by the general population. It appears that all iodine
entering the small intestine is absorbed into the blood, and hence the fj_
value is taken as 1 for all ages, which is the value we will use for this
analysis.
(c) Organ Masses, m
To a large extent, the variability in organ masses among individuals in
the general population is related to age. For most of the target organs
listed in Table 6-2, the mass increases during childhood and continues to
increase until adulthood, at which time the net growth of the organ ceases;
there may be a gradual .decrease in mass (for some organs) in later years.
The associated uncertainty in estimating the dose to the thyroid of
exposed individuals can be estimated by considering how the mass varies with
age, as well as among individuals of the same age. Based on data reviewed by
Dunning and Schwarz (Du81), the mass of an adult thyroid ranges from 2 to 62
g. As a result, the absorbed dose rate to the thyroid would vary by a factor
of 31 just among adults. In comparing estimates for children and adults,
children in the age group from .5 to 2 yr were found to have a mean thyroid
mass of 2.1 g, while the adult group had a mean mass of 18.3 g. Based on
these values, the absorbed dose rate to the thyroid of the average child, and
adult would differ by about a factor of 9. For this analysis, we have used
the same values employed by the ICRP (20 g for the adult thyroid mass and 1.8
g for that of a 6-month-old child), which are also consistent with the
recommendation of Bryant (Br69).
(d) Organ Uptake Fraction, fg.
The fraction of a radionuclide taken up from the blood in an organ is
strongly correlated with the size of the organ, its metabolic activity, and
the amount of material ingested. Iodine introduced into the bloodstream is
rapidly deposited in the thyroid, usually reaching a peak within 24 hours.
The uptake of iodine-131 by the thyroid is similar to that of stable iodine
in the diet (Doll), and can be influenced by sex and dietary differences.
There can be considerable variation among populations.
Dunning and Schwarz (Du81) found a mean £•% value of O-4"7
newborns, 0.39 for infants, 0.47 for adolescents, and 0.19 for adults. Other
data (Pi69, Be70, Wo77) would suggest that a value of 0.10 to 0.20 may better
represent adult populations in the United states. For purposes of this
analysis, we have used £2 values of .35 and .30 for a child and adult,
respectively.
6-31
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(e) Biological Half-Life.
Data suggest that there is a strong correlation between biological
half-lives of radionuclides in organs in the body and the age of the
individual. Children are expected to exhibit smaller values of Two
and greater uptakes (Ro58), and this relationship appears to be independent
of the type of radionuclide ingested (Br85). For iodine-131, a range of
21 to 200 days for adults was observed and a similarly wide range would be
expected for other age groups (Du81). Rosenberg (Ro58) found a significant
correlation between the biological half-life and the age of the individual,
and an inverse relationship between uptake and age in subjects from 22 to 50
yr of age. Dunning and Schwarz (Du81) concluded that for adults the observed
range was from 21 to 372 days, implying for adults about an 18-fold variation
in absorbed dose rate, other factors being held constant. For children in
the age group from .5 to 2 yr, the range was 4 to 39 days, which would affect
the absorbed dose rate estimate by about a factor of 10.
In light of the possible inverse relation between the biological
half-life and the f2 value, we will, for the purposes of this analysis,
use biological half-lives of 24 and 129 days, respectively, for children and
adults, based on the paper by Bryant (Br69).
(f) Effective Energy per Disintegration, E
The effective energy per disintegration of a radionuclide within an
organ is dependent upon the decay energy of the radionuclide and the
effective radius of the organ containing the radionuclide (ICRP59). It is
expected, therefore, that E is an age-dependent parameter which could vary as
the size of the organ changes. While very little work has been done in
determining E values for the radionuclides found in low-level radioactive
waste products, some information has been published for iodine-131 and
cesium-137. Considering the differences between the child and the adult
thyroid, Bryant (Br69) derives E values of 0.18 MeV/dis for the child and
0.19 MeV/disintegration for the adult. The above values correspond to a
6-month-old child with a mass of 1.8 g and an f2 value of 0.35. The
corresponding E value for the adult was calculated for a 20 g thyroid with an
£2 value of 0.3.
(3) Differences in Child and Adult Doses Associated with
Age-Dependent Changes in Model Parameters
To examine the uncertainties in thyroid dose associated with changes in
model parameters with age, values for child and adult parameters were chosen
as discussed above and are listed in Tables 6-5 and 6-6.
Using Equation 25, the absorbed dose rate to the thyroid of a child
Dc, can be compared to that of an adult Da, by the following:
Dc/Da (0.7X1X0.35X0.18X20X24)
(0.5X1X0.30)<0.19X1.8X139) "
2.96
(26)
6-32
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Table 6-5. Model parameters for iodine metabolism in the
thyroid of a child (Age 0.5 to 2 yr)
Parameters
I
£1
fl
E
m
m
Tl/2
C = iodine
Values
0.7/day x C@
1
0.35
0.18 Me v/ 1 r ans format ion
1.8 g
24 days
concentration in milk = 1 v>Ci/L
Reference
Br69
ICRP59
Br69
Br69
Br69
Br69
Table 6-6. Model parameters for iodine metabolism in the
thyroid of "an adult (Age > 18 yr)
Parameters
I
fl
f1
2
E
m
Tl/2
3
C = iodine
Values
@
0.5/day x C
1
0.30
0.19 Mev/transforraation
20 g
139 days
concentration in milk = 1 yd
Reference
Br69
ICRP59
Br69
Br69
Br69
Br69
/L
6-33
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_u fc c ! °n these Parameters, therefore, the analysis indicates
that, for a given concentration of 1-131 in milk, the estimated absorbed
dose rate to the thyroid of a 6-month-old child would be a factor of
approximately 3 times that to the adult thyroid, in other words, use of
adult parameters would underestimate the thyroid dose to the child bv
almost a factor of 3. This difference is expected to change with age
with other radionuclides, however. .. 9
Depending on the type of radionuclide ingested, the age and
element dependency in the metabolic and physiological processes
determines how the dose to target organs varies with age. For example,
strontium tends to follow the calcium pathways in the body and deposits
to a large extent in the skeleton. In fact, the fraction of ingested
strontium eventually reaching the skeleton at a given age depends largely
on the skeletal needs for calcium at that age, although the body is able
to discriminate somewhat against strontium in favor of calcium after the
first few weeks of life.
in summary, it is difficult to make generalizations concerning
the uncertainty involved in making calculations. More work is necessary
to properly characterize the effect of age and individual dependent
morphological and metabolic changes on dose.
6.3.2 Special Radionuclides
The following paragraphs briefly summarize some of the special
considerations for particular elements and radionuclides.
(A) Tritium and Carbon-14
Most radionuclides are nuclides of elements found only in
trace quantities in the body, others like tritium (hydrogen-3) or
carbon-14
must be treated differently since they are long-lived nuclides of
elements that are ubiquitous in tissue. An intake of tritium is'assumed
to be completely absorbed and will rapidly mix with the water content of
tne body (Ki78a).
The estimates for inhalation include consideration of
absorption through the skin. Organ dose estimates are based on the
steady-state specific-activity model described by Killough et al. (Ki78a).
Carbon-14 is assumed to be inhaled as CO2 or ingested in a
biologically bound form, inhaled carbon-14 is assumed to be diluted bv
stable carbon from ingestion (Ki78b). This approach allows separate
consideration of the ingestion and inhalation pathways. The
specific-activity model used for organ dose estimates is also that of
Killough et al. (Ki78a). Short-lived carbon radionuclides (e g ,
carbon-11 or carbon-15) are treated as trace elements and the organ doses
are calculated accordingly. uw&es,
6-34
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(B) Noble Gases
The retention in the lung of noble gases uses the approach described
by Dunning et al. (Du79). The inhaled gas is assumed to remain in the
lungs until it is lost by radiological decay or respiratory exchange.
Translocation of the noble gas.to systemic organs is not considered, but
is included for any decay products produced in the lungs. The inhalation
of the short-lived decay products of radon is assessed using a potential
alpha energy exposure model (see Chapter 7) rather than by calculating
the doses to lung tissues from these radionuclides.
(C) Transuranics
The metabolic models for transuranic elements (Po, Np, Pu, Am, and
Cm) are consistent with those used for the EPA transuranic guidance (EPA
77). Basically, a GI tract to blood absorption factor of 10 3 is used
for the short-lived nuclides of plutonium (plutonium-239, -240,
and -242), while a value of 10~4 is used for other transuranics. For
soluble forms of uranium, a GI tract to blood absorption factor of 0.2 is
used in accordance with the high levels of absorption observed for
low-level environmental exposures by Hursh and Spoor (Hu73 and Sp73).
6.3.3 External Dose Models
This section is concerned with the calculation of dose rates for
external exposure to photons from radionuclides dispersed in the
environment. Two exposure models are discussed: (1) immersion in
contaminated air and (2) irradiation from material deposited on the
ground surface. The immersion source is considered to be a uniform
semi-infinite radionuclide concentration in air, while the ground surface
irradiation source is viewed as a uniform radionuclide concentration on
an infinite plane. In both exposure modes, the dose rates to organs are
calculated from the dose rate in air. For low-level waste assessments,
ground surface irradiation is, almost without exception, more significant
than air immersion.
Dose rates are calculated as the product of a dose rate factor which
is specific for each radionuclide, tissue, and exposure mode and the
corresponding air or surface concentration. The dose rate factors used
in the low-level waste modeling assessments were calculated with the DOSE
FACTOR code (Ko81). Note that the dose rate factors for each
radionuclide do not include any contribution for decay products. For
example, the ground surface dose factors for cesium-137 are all zero,
since no photons are emitted in its decay. To assess surface deposition
of cesium-137, one must first calculate the ingrowth of its decay
product, metastable barium-137, which is a photon emitter.
(A) immersion
For immersion exposure to the photons from radionuclides in air, EPA
assumes that an individual is standing at the base of a semi-infinite
cloud of uniform radionuclide concentration. We first calculate the dose
rate factor (the dose rate for a unit concentration) in air for a source
6-35
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of photons with energy E^. At all points in an infinite uniform
source, conservation of energy considerations requires that the rates of
absorbed and emitted energy per unit mass be equal. The absorbed energy
rate per unit mass at the boundary of a semi-infinite cloud is just half
that value. Hence
(27)
where:
a
DRFy » the immersion dose rate per unit air concentration (rad m3/ci s);
Ey =* emitted photon energy (MeV);
k « units conversion factor
= 1.602 10~13 (J/MeV) x 3.7 1010 (1/ci s) x 103 (g/kg) x 102 (rad kg/J)
- 5.93 102 (g rad/MeV Ci s); and
P - density of air (g/m3).
The above equation presumes that for each nuclide transformation, one
photon with energy E^ is emitted. The dose rate factor for a nuclide is
obtained by adding together the contributions from each photon associated
with the transformation process for that radionuclide.
(B) Ground Surface Irradiation
In the case of air immersion, the radiation field was the same
throughout the source region. This allows the dose rate factor to be
calculated on the basis of energy conservation without having to explicitly
consider the scattering processes taking place. For ground surface
irradiation, the radiation field depends on the height of the receptor above
the surface, and the dose rate factor calculation is more complicated. The
radiation flux per unit solid angle is strongly dependent on the angle of
incidence, it increases from the value for photons incident from
immediately below the receptor to a maximum close to the horizon.
attenuation and buildup due to scattering must be considered to calculate
the dose rate factor. Secondary scattering provides a distribution of
photon energies at the receptor, which increases the radiation flux above
that calculated on the basis of attenuation. Trabey (Tr66) has provided a
useful and reasonably accurate expression to approximate this buildup:
(28)
where Ben is the buildup factor (i.e., the quotient of the total energy
flux and that calculated for attenuation) only for energy in air; pa
is the attenuation coefficient at the energy of the released photon
(ra"1); r is the distance between the photon source and the receptor;
6-36
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and the Berger buildup coefficients Ca and Da are dependent on energy
and the scattering medium. The buildup factor is dimensionless and
always has a value greater than unity. The resulting expression for the
dose rate factor at a height z (m) above a uniform plane is
\
(29)
is the mass energy-absorption coefficient (m2/g)
where (ven/p)a
for air at photon energy E« (MeV) ; E! is the first order
exponential integral function, i.e.,
E (X) = I
X
exp(-u)
u
du
(30)
a a are the buildup coefficients in air at energy E~; and
k=5.93 102 (g rad/MeV Ci s) as for the immersion calculation.
Ca and Da
As for immersion, the dose rate factor for a nuclide combines the
contribution from each photon energy released in the transformation
process.
(C) Organ Doses
The dose rate factors in the preceding two sections are for the
absorbed dose in air. For a radiological assessment, the absorbed doses
in specific tissues and organs are needed. For this purpose, Kerr and
Eckerman (Ke80, KeSOa) have calculated organ dose factors for immersion
in contaminated air. Their calculations are based on Monte carlo
simulations of the absorbed dose in each tissue or organ for the spectrum
of scattered photons in air resulting from a uniform concentration of
monoenergetic photon sources. Kocher (Ko81) has used these data to
calculate values of the ratio of the organ dose factor to the air dose
factor, Gk(Ey), for 24 organs and tissues at 15 values of Ey
ranging from 0.01 to 10.0 MeV.
The resulting organ-specific dose rate factor for immersion is
DRFk(E )
(31)
For a specific nuclide, the dose rate factor is obtained by taking the
sum of the contributions from each photon energy associated with the
radionuclide decay.
Ideally, a separate set of Gk(Ey) values would be used for the
angular and spectral distributions of incident photons from a uniform
plane source. Since these data are not available, Kocher (Ko81) has used
the same set of Gk(Ey) values for calculating organ dose rate
factors for both types of exposure.
6-37
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(D)
Uncertainty Considerations in External Dose Rate Factor;
in computing the immersion dose rate factor in air, the factor of
1/2 in Equation (27), which accounts for the semi-infinite geometry of
the source region, does not provide a rigorously correct representation
of the air ground interface. However, Dillman (Di74) has concluded that
this result is within the accuracy of available calculations. The
radiation field between the feet and the head of a person standing on
contaminated ground is not uniform, but for source photon energies
greater than about 10 keV, the variation about the value at 1 meter
becomes minimal. A more significant source of error is the assumption of
a uniform concentration. Kocher (Ko81) has shown that sources would have
to be approximately uniform over distances of as much as a few hundred
meters from the receptor for the dose rate factors to be accurate for
either ground surface or immersion exposures. Penetration of deposited
materials into the ground surface, surface roughness, terrain
irregularities, as well as the shielding provided by buildings to their
inhabitants, all serve to reduce doses.
The effect of using the same factors to relate organ doses to the
dose in air for ground surface as for immersion photon sources has not
been studied. The assumptions that the radiation field for the ground
surface source is isotropic and has the same energy distribution as for
immersion clearly do not hold true, but more precise estimates of these
distributions are not likely to change the organ dose rate factors
substantially.
Kocher (Ko81) has noted that the idealized photon dose rate factors
are "likely to be used quite extensively even for exposure conditions for
which they are not strictly applicable... because more realistic
estimates are considerably more difficult and expensive [to make]."
6'4 Distribution of Doses in the General Population
Although the use of extreme parameter values in a sensitivity
analysis indicates that large uncertainties in calculated doses are
possible, this uncertainty is not usually reflected in the general
population. There are several reasons for this: the parameter values
chosen are intended to be typical of an individual in the population; it
is improbable that the "worst case" parameters would be. chosen for all
terras in the equation; and not all of the terms are mutually independent,
e.g., an increased intake may be offset by more rapid excretion.
This smaller range of uncertainty in real populations is
demonstrated by studies performed on various human and animal
populations, it should be noted that there is always some variability in
observed doses that results primarily from differences in the
characteristics of individuals. The usual way of specifying the dose, or
activity, variability in an organ is in terms of the deviation from the
average, or mean, value.
6-38
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REFERENCES
Be70 Bernard, J.D., McDonald, R.A. .and.. Nesmith J.A,, New Normal
Ranges for the-Radioiodine Uptake"study, J. Nucl. Med.,
II:(7):449-451, 1970. ,
Br52 Bruckner, H. Die Anatomie der Lufttrohre beim lebenden Menchen,
A. Anat., Entwicklungsgeschichte, 116:276, 1952 [cited inLi69].
Br69 Bryant, P.M., Data for Assessments Concerning Controlled and
Accidental Releases of 131I and 137Cs to Atmosphere, Health
Phys., 17U):51-57, 1969,
Di74 Dillman, L.T., Absorbed Gamma Dose Rate for Immersion in a
Semi-infinite Radioactive Cloud, Health Phys., 27(6):571, 1974.
Du79 Dunning, D.E. Jr., Bernard, S.R., Walsh, P.J., Klllough, G.G.
and Pleasant, J.C., Estimates of Internal Dose Equivalent to 22
Target Organs for Radionuclides Occurring in Routine Releases
from Nuclear Fuel-Cycle Facilities, Vol. II, Report No.
ORNL/NUREG/TM-190/V2, NUREG/CR-0150 Vol. 2, Oak Ridge National
Laboratory, Tennessee, 1979.
Du81 Dunning, D.E. and Schwort, G. Variability of Human Thyroid
Characteristics and Estimates of Dose from Ingested i<3*I.-
Health Phys., 40(5):661-675, 1981.
Ec85 Eckerman, K.F.. Absorbed Fraction Data for Radiosensitive
Tissues of the Skeleton, Part 1, Beta Emitters in Trabecular
Bone, (in preparation).
Fi35 Findeisen, W., Uber das Absetzen Kleiner in der Luft
suspendierten Teilchen in der Menschlichen Lunge bei der Atmung,
Pflugers Arch, f d ges. Physiol., 236, 367, 1935.
FRC67 Federal Radiation Council, Guidance for the Control of Radiation
Hazards in Uranium Mining, FRC Report No. 8, Revised. U.S.
Government Printing Office, Washington, D.C., 1987.
Ho75 Holden, W.S. and Marshal, R., Variations in Bronchial Movement,
Clin. Radiol., 26:439-454^ 1975.
Hu72 Hughes, J.M.B., Hoppin, F.G., Jr. and Mead, J. Effect of Lung
Inflation on Bronchial Length and Diameter in Excised Lungs,
J. Appl. Physiol., 32:25-35, 1972.
Hu73 Hursh, J.B. and Spoor, N.L., Data on Man, Chapter 4 in Uranium,
Plutonium and the Transplutonic Elements, Springer, New York,
1973.
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ICRP66
ICRP80
ICRP81
ICRP84
ICRU80
Ke78a
Ke78b
Ke80
KoSla
KoSlb
ICRP Task Group on Lung Dynamics, Depositions and Retention
Models for Internal Dosimetry of the Human Respiratory Tract
Health Phys., 12(2):173-207, 1966. ,
International Commission on Radiological Protection, Limits for
Intakes of Radionuclides by Workers, ICRP Publication 30, Part
2, Annals of the ICRP, Vol. 4 (3/4), Pergamon Press, Oxford,
1980.
International Commission on Radiological Protection, Limits for
Intakes of Radionuclides by Workers, ICRP Publication 30, Part
3, Annals of the ICRP, Vol. 6 (2/3), Pergamon Press, Oxford,
1981.
International Commission on Radiological Protection, A
Compilation of the Major concepts and Quantities in use by ICRP,
ICRP Publication No. 42, Pergamon Press, Oxford, (1984)
International Commission on Radiation Units and Measurements,
ICRU Report No 33, Washington, D.C., 1980.
Killough, G.C., Dunning, D.E Jr., Bernard, s.R. and Pleasant,
J.C., Estimates of internal Dose Equivalent to 22 Target Organs
for Radionuclides Occurring in Routine Releases from Nuclear
Fuel Cycle Facilities, Vol. 1, Report No. ORNL/NUREG/TM-190, Oak
Ridge National Laboratory, Tennessee, June 1978.
Killough, G.c. and Rohwer, P.S., A New Look at the Dosimetry of
i4C Released to the Atmosphere as Carbon Dioxide, Health
Phys., 34(2):141, 1978.
Kerr, G.D. and Eckerman, K.F., Oak Ridge National Laboratory,
private communication; see also Abstract P/192 presented at the
Annual Meeting of the Health Physics Society, Seattle,
Washington, July 20-25, 1980, and the discussion section in
ref. 16.
Kocher, D.C. and Eckerman, K.F., Electron Dose-Rate Conversion
Factors for External Exposure of the Skin, Health Phys.,
40(1):67, 1981.
Kocher, D.C., Dose-Rate Conversion Factors for External Exposure
to Photon and Electron Radiation from Radionuclides Occurring in
Routine Releases from Nuclear Fuel-Cycle Facilities, Health
Phys., 38(4):543-621, 1981.
6-40
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LeSOa
NCRP71
ORNL85
R058
Sp73
SU81
Th77
Th78
We63
Kerr. G.D. , A Review of Organ Doses from Isotropic Fields of
Y-Rays, Health Phys., 39(1):3, 1980.
National Council on Radiation Protection and Measurements, Basic
Radiation protection criteria, NCRP Report No. 39, National
Council on Radiation Protection and Measurements, Washington,
D.C., 1971.
Oak Ridge National Laboratory Report of Current Work of the
Metabolism and Dosimetry Research Group, ORNL/TM-9690 , Oak
Ridge, Tennessee, 1985.
Rosenberg, G., Biologic Half-life of "li in the Thyroid of
Healthy Males, J. Clin. Endocrinol. Metab. , 18., 516-521, 1958.
spoor, N.L. and Hursh, J.B., Protection Criteria, Chapter 5 in
Uranium, Plutonium and the Transplutonic Elements, Springer, New
York, 1973.
Sullivan, R.E., Nelson, N.S., Ellett. W.H. , Dunning, D.E. Jr.,
Leggett, R.W., Yalcintas, M.G. and Eckerman, K.F., Estimates ot
Health Risk from Exposure to Radioactive Pollutants, Report No.
ORNL/TM-7745, Oak Ridge National Laboratory, Oak Ridge,
Tennessee, 1981.
Thome, M.D., Aspects of the Dosimetry of Alpha-Emitting
Radionuclides in Bone with Particular Emphasis on -"°Ra and
239Pu, Phys. Med. Biol., 22:36-46, 1977.
Thurlbeck, W.M. Miscellany, 287-315 in The Lung: structure
Function and Disease, Thurlbeck, W.M. and Abell, M.R., editors,
The Williams and Wilkins Co., Baltimore, Maryland, 1978.
Weibel, E.R. Morphometry of the Human Lung, Springer-Verlag,
Berlin, 1963.
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Chapter 7: ESTIMATING THE RISK OF HEALTH EFFECTS RESULTING
FROM EXPOSURE TO LOW LEVELS OF IONIZING RADIATION
7.1 Introduction
This chapter describes how EPA estimates the risk of fatal cancer,
serious genetic effects, and other detrimental health effects caused by
exposure to low levels of ionizing radiation.
Ionizing radiation refers to radiation that strips electrons from
atoms in a medium through which it passes. The highly reactive electrons
and ions created by this process in a living cell can produce, through a
series of chemical reactions,, permanent changes (mutations) in the cell's
genetic material, the DNA. These may result in cell death or in an
abnormally functioning cell. A mutation in a -germ cell (sperm or ovum)
may be transmitted to an offspring and be expressed as a genetic defect
in that offspring or in an individual of a subsequent generation: such a
defect is commonly referred to as a genetic effect. There is also strong
evidence that the induction of a mutation by ionizing radiation in a
non-germ (somatic) cell can serve as a step in the development of a
cancer. Finally, mutational or other events, including possible cell
killing, produced by ionizing radiation in rapidly growing and
differentiating tissues of an embryo or fetus, can give rise to birth
defects: these are referred to as teratological effects. At acute doses
above about 25.rads, radiation induces other deleterious effects in man;
however, for the low doses and dose rates of interest in this document
only those three kinds of effects referred to above are thought to be of
significance.
Most important from the standpoint of the total societal risk from
exposures to low-level ionizing radiation are the risks of cancer and
genetic mutations. Consistent with our current understanding of their
origins in terms of DNA damage, these are believed to be stochastic
effects; i.e., the probability (risk) of these effects increases with the
absorbed dose of radiation, but the severity of the effects is
independent of'dose. For neither induction of cancer nor genetic ^
effects, moreover, is there any convincing evidence for a "threshold,
i.e., some dose level below which the risk is zero. Hence, so far as we
know, any dose of ionizing radiation, no matter how small, might give
rise to a cancer or to a genetic effect in future generations.
Conversely, there is no way to be certain that a given dose of radiation,
no matter how large, has caused an observed cancer or will cause one in
the future.
In summary, knowledge of the radiation dose absorbed by an
individual allows us to estimate the probability that the dose will
result in a cancer or a genetic effect (or somewhat more precisely - to
estimate the number of excess cancers and genetic effects resulting from
the same dose to a large group of similar individuals).
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Beginning nearly with the discovery of x rays in 1895 but especially
since World War II, there has been an enormous amount of research
conducted into the biological effects of ionizing radiation. This
research continues at the level of the molecule, the cell, the tissue,
the whole laboratory animal, and man. There are two fundamental aspects
to most of this work:
1.
2.
Estimating the radiation dose to a target (cell, tissue, etc.).
This^aspect (dosimetry), which may involve consideration of
physiological, metabolic, and other factors, is discussed more
fully in Chapter 6.
Measuring the number of effects of a given type associated with
a certain dose (or exposure).
For the purpose of assessing the risk to man from exposures to
ionizing radiation, the most important information comes from human
epidemiological studies in which the number of health effects observed in
an irradiated population is compared to that in an unirradiated control
population. The human epidemiological data regarding radiation-induced
cancer are extensive. As a result, the risk can be estimated to within
an order of magnitude with a high degree of confidence. Perhaps for only
one other carcinogen - tobacco smoke - are we in a better position with
regard to the reliability of risk estimates.
Nevertheless, there are serious gaps in the human data on radiation
risks. No clear-cut evidence of excess genetic effects has been found in
irradiated human populations, for example. Likewise, no statistically
significant excess of cancers has been demonstrated below about 10 rads,
the dose range of interest from the standpoint of environmental
exposures. _Since the epidemiological data are incomplete in many
respects, risk assessors must rely on mathematical models to estimate the
risk from exposures to low-level ionizing radiation. The choice of
models, of necessity, involves subjective judgments, but should be based
on all^relevant sources of data collected by both laboratory scientists
and epidemiologists. Thus, radiation risk assessment is a process that
continues to evolve as new scientific information becomes available.
The EPA's estimates of cancer and genetic risks in this BID are
based largely on the results of a.National Academy of Sciences (NAS)
study as given in the BEIR-3 report (NAS80). The study assessed
radiation risks at low exposure levels. As phrased by the President of
the Academy, "We believe that the report will be helpful to the EPA and
other agencies as they reassess radiation protection standards. It
provides the scientific bases upon which standards may be decided after
nonscientific social values have been taken into account."
In this discussion, we outline the various assumptions made in
calculating radiation risks based on the 1980 NAS report, and compare
these risk estimates with those prepared by other scientific groups, such
7-2
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as the 1972 NAS BEIR Committee (NAS72). the United Nations Scientific
Committee on the Effects of Atomic Radiation (UNSCEAR77, 82), and the
ICRP (ICRP77). We recognize that information on radiation risks is
incomplete and do' not argue that any of the estimates derived by the 1980
NAS BEIR Committee on the basis of alternative assumptions is highly
accurate. Rather, we discuss some of the deficiencies in the_available
data base and point out possible sources of bias in current risk
estimates. Nevertheless, we believe the risk estimates made by EPA are
reasonable in light of current evidence.
In the sections below, we first consider (Sections 7.2-7.2.8) the
cancer risk resulting from whole-body exposure to low-LET (see Chapter 6)
radiation, i.e., sparsely ionizing radiation like the energetic electrons
produced by x rays or gamma rays. Environmental contamination by
radioactive materials also leads to the ingestion or inhalation of the
material and subsequent concentration of the radioactivity in selected
body organs. Therefore, the cancer risk resulting from low-LET
irradiation of specific organs is examined next (Sections 7.2.9-7.2.11;.
Organ doses can also result from high-LET radiation, such as that
associated with alpha particles. The estimation of cancer risks for
situations where high-LET radiation is distributed more or less uniformly
within a body organ is the third situation considered (Section 7.3).
Because densely ionizing alpha particles have a very short range in
tissue, there are exposure situations where the dose distribution to
particular organs is extremely nonuniform. An example is the case of
inhaled radon progeny, polonium-218, lead-214, and polonium-214. For
these radionuclides we base our cancer risk estimates on the amount ot
radon progeny inhaled rather than the estimated dose, which is highly
nonuniform and cannot be well quantified. Therefore, risk estimates, of
radon exposure are examined separately (Section 7.4). In Section 7.5, we
review the causes of uncertainty in the cancer risk estimates and the
magnitude of this uncertainty so that both the public and EPA decision
makers have a proper understanding of the degree of confidence to place
in them. In Section 7.6, we review and quantify the risk of deleterious
genetic effects from radiation and the effects of exposure in utero on
the developing fetus. Finally, in Section 7.7, we calculate cancer and
genetic risks from background radiation using the models described in
this chapter.
7.2 Cancer Risk Estimates for Low-LET Radiations
Most of the observations of radiation-induced carcinogenesis in
humans are on groups exposed to low-LET radiations. These groups include
the Japanese A-bomb survivors and medical patients treated with
diagnostic or therapeutic radiation, most notably for ankylosing
spondylitis in England from 1935 to 1954 (Sm78). Comprehensive reviews
of these and other data on the carcinogenic effects of human exposures
are available (UNSCEAR77, NAS80).
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The most important source of epidemiological data on radiogenic
cancer is the population of Japanese A-bomb survivors. The A-bomb
survivors have been studied for more than 38 years and most of them the
Life Span Study Sample, have been followed since 1950 in a carefully
planned and monitored epidemiological .survey (Ka82, Wa83). They are the
largest group that has been studied, and they provide the most detailed
information on the response pattern for organs by age and sex over a wide
range of doses of low-LET radiation. Unfortunately, the doses received
by various individuals in the Life Span Study Sample are not yet known
accurately. The 1980 BEIR Committee's analysis of the A-bomb survivor
data collected up to 1974 was prepared before bias in the dose estimates
for the A-bomb survivors (the tentative 1965 dose estimates, T65) became
widely recognized (Lo81). It is now clear that the T65 dose equivalents
to organs tended, on average, to be overestimated (Bo82, RERF83 84) so
that the BEIR Committee's estimates of the risk per unit dose are likely
Co be too low. A detailed reevaluation of current risk estimates is
indicated when the A-bomb survivor data have been reanalyzed on the basis
of new and better estimates of the dose to individual survivors. These
estimates should become available during 1988.
Uncertainties in radiation risk estimates do not result just from
the uncertainties about the Japanese and other epidemiological studies.
As discussed below, risk projections based on these studies require
certain assumptions (e.g., with regard to low dose extrapolation). The
degree of uncertainty associated with these assumptions is probably
greater than the uncertainty of the estimated risk per unit dose among
the A-bomb survivors or other sources of risk estimates for radiogenic
cancer in humans.
7-2.1 Assumptions Needed to Make Risk Estimates
A number of assumptions must be made on how to extrapolate
observations made at high doses to estimate effects from low doses and
low dose rates. Excess cancers have been observed, for the most part,
only following doses of ionizing radiation that are relatively high when
compared to those likely to occur as a result of the combination of
background radiation and environmental contamination from controllable
sources of radiation. Therefore, a dose response model must be chosen to
allow extrapolation from the number of radiogenic cancers observed at
high doses to the number of cancers at low doses resulting from all
causes including background radiation.
The range of extrapolation is not the same for all kinds of cancer
because it depends upon the radiosensitivity of a particular tissue. For
example, the most probable radiogenic cancer for women is breast cancer.
As described below, the incidence of radiogenic breast cancer does not
seem to diminish when the dose is protracted over a long period of time.
For example, the number of excess cancers per unit dose among Japanese
women, who received acute doses, is about the same per unit dose as women
exposed to small periodic doses of x rays over many years. If this is
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actually the case, background radiation is as carcinogenic per unit dose
for breast tissue as the acute exposures from A-bomb gamma radiation.
Moreover, the female A-bomb survivors show an excess of breast cancer at
doses below 20 rads which is linearly proportional to that observed at
several hundred rads (To84). [Evidence of a nonlinear dose response
relationship for induction of. breast cancer has been obtained in a study
of Canadian fluoroscopy patients, but only at doses above about 500 rads
(Ho84).l Women in their 40's, the youngest age group in which breast
cancer is common, have received about 4 rads of whole-body low-LET
background radiation and usually some additional dose incurred for
diagnostic medical purposes. Therefore, for this cancer, the difference
between the lowest dose at which radiogenic cancers are observed, less
than 20 rads, and the dose resulting from background radiation is less
than a factor of 5, not several orders of magnitude as is sometimes
claimed. Based on data from irradiated tinea capitis patients, induction
of thyroid cancer also seems to be linear with doses down to 10 rads or
lower (NCRP85). However, for most other cancers, a statistically
significant excess has not been observed at doses below 50 rads of
low-LET radiation. Therefore, the range of dose extrapolation is often
large.
7.2.2 Dose Response Functions
The 1980 NAS report (NAS80) examined only three dose response
functions in detail: (1) Linear, in .which the number of effects (risk)
is directly.proportional to dose at all doses; (2) linear-quadratic, in
which risk is very nearly proportional to dose at very low doses and
proportional to the square of the dose at high doses; and (3) a quadratic
dose response function, where the risk varies as the square of the dose
at all dose -levels.
We believe the first two of these functions are compatible with most
of the data on human cancer. Information which became available only
after the BEIR-3 report was published indicates that a quadratic response
function-is inconsistent with the observed excess risk of solid cancers
at Nagasaki, where the estimated gamma-ray doses are not seriously
confounded by an assumed neutron dose component. The chance that a
quadratic response function underlies the excess cancer observed in the
Nagasaki incidence data has been reported as only 1 in 10,000 (Wa83;.
Although a quadratic response function is not incompatible with the Lite
Span Study Sample data on leukemia incidence at Nagasaki, Beebe and
others (Be78, E177) have pointed out how unrepresentative these data are
of the total observed dose response for leukemia in that city. There is
no evidence that a quadratic response function provides a better fit to
the observed leukemia excess among all A-bomb survivors in the Lite bpan
Study Sample than a simple linear.model (NAS80). Based on these
considerations, we do not believe a quadratic response can be used in a
serious effort to estimate cancer risks due to ionizing radiation.
7-5
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The 1980 NAS BEIR Committee considered only the Japanese mortality
data in their analysis of possible dose response functions (NAS80).
Based on the T65 dose estimates, this Committee concluded that the excess
mortality from solid cancers and leukemia among the A-bomb survivors is
compatible with either a linear or linear-quadratic dose response to the
low-LET radiation component and a linear response to the high-LET neutron
component (NAS80). Although the.1980 BEIR report indicated risk
estimates for low-LET radiation based on a linear-quadratic response were
preferred" by most of the scientists who prepared that report, opinion
was not unanimous, and we believe the subsequent reassessment of the
A-bomb dose seriously weakens the Committee's conclusion. The
Committee's analysis of dose response functions was based on the
assumption that most of. the observed excess leukemia and solid cancers
among survivors in Hiroshima resulted from neutrons (see Tables V-13
A-7, Equations V-10, V-ll in NAS80). Current evidence, however, is
conclusive that neutrons were only a minor component of the dose among
all but a few survivors in both Hiroshima and Nagasaki (Bo82,
RERF83,84). Therefore, it is likely that most of the response attributed
to neutrons was caused by the gamma dose, not the dose from neutrons.
This point is discussed further in Section 7.3.
The revised dosimetry will involve changes in individual absorbed
doses that vary with distance from the explosion in each of the two
cities and with shielding characteristics. As a consequence, though it
seems clear that there will generally be a higher response per unit dose,
there will also be an unpredictable change in the shape of the dose
response exhibited by the data. Reanalysis of the Japanese experience
after completion of the dose reassessment may then provide more
definitive information on the dose response of the A-bomb survivors;
nevertheless, it is unlikely to produce a consensus on the dose response
at environmental levels, i.e., about 100 mrad/yr. This is because at low
enough doses there will always be sampling variations in the observed
risks so that observations are compatible, in a statistical sense, with a
variety of dose response functions. In the absence of empirical evidence
or a strong theoretical basis, a choice between dose response functions
must- be based on other considerations.
Although there is evidence for a nonlinear response to low-LET
radiations in some, but not all, studies of animal radiocarcinogenesis
(see below), we are not aware of any data on human cancers that are
incompatible with the linear model. In such a case, it may be preferable
to adopt the simplest hypothesis that adequately models the observed
radiation effect. Moreover, EPA believes that risk estimates, for the
purpose of assessing radiation impacts on public health, should be based
on scientifically credible risk models that are unlikely to understate
the risk. The linear model fulfills this criterion. Given the current
bias in the doses assigned to A-bomb survivors (see Section 7.5.1 below),
such an approach seems reasonable as well as prudent. Therefore, EPA has
primarily used the BEIR-3 linear dose response model for estimating the
risk of radiogenic cancer due to low-LET radiations.
7-6
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For low-LET radiations, the BEIR-3 Committee preferred the
linear-quadratic dose response model. In this model, the risk from
an acute dose, D, of low-LET radiation is assumed to be of the form
«D + D2. The BEIR-3 Committee assumed that the linear and
quadratic terms were equal at 116 rads, leading to a linear coefficient
a which was a factor of 2.5 times lower than the coefficient obtained
from the linear model (NAS80). At low doses the quadratic term becomes
negligible; at chronic low-dose rates it is ignored, for reasons
discussed below. For environmental exposures, therefore, risk estimates
based on the BEIR-3 linear-quadratic dose response model are only
40 percent of those based on the BEIR-3 linear model.
A theoretical basis for the linear-quadratic dose response model has
been put forth by Keilerer and Rossi (Ke72). In this theory of dual
radiation action," events leading to "Lesions" (i.e., permanent changes)
in cellular DNA require the formation of interacting pairs of
"sublesions." The interacting pairs can be produced by a singLe
traversing particle, or track, or by two separate tracks, giving .rise,
respectively, to a linear and quadratic term in the dose response
relationship. According to the theory, a sublesion may be repaired
before it can interact to form a lesion, the probability of such repair
increasing with time. Consequently, as dose rate is reduced the
formation of lesions from sublesions caused by separate tracks becomes
less important,"and the magnitude of the V2 term diminishes. Hence,
the theory predicts that at sufficiently low doses or dose rates the
response .should be a Linear function of dose. Moreover, the constant of
proportionality is the same in both cases, i.e., a.
Results of many animal experiments are qualitatively consistent
with the theory: low-LET radiation often seems to have a reduced
effectiveness per unit dose at Low dose .rates (NCRP80); however, it is
usually not possible from the data to verify that the dose response curve
has the Linear-quadratic form. Another success of the .dual action theory
has been in explaining observed differences between the effects ot
Low-LET and high-LET radiations. In this view, the densely ionizing
nature of the Latter results in a much greater production of interacting
pairs of subiesions by single tracks, leading in turn to higher relative
biological effectiveness at low doses and a linear dose response
relationship for high-LET radiation (except for possible cell-kiLLing
effects).
7-7
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The dual action theory has nevertheless been challenged on
experimental grounds, and observed variations in response with dose dose
rate (see below), and LET can also be explained in terms of a theory
involving only single lesions and a "saturable" repair mechanism that
decreases in effectiveness at high dose rates on the microscopic scale
(Go82). One property of such a theory is that the effectiveness of
repair, and'therefore the shape of the dose response curve, can in
principle vary substantially with cell type and species. Hence, results
obtained on laboratory animals would not necessarily be entirely
applicable to people.
Finally, some mention should be made of "supralinear models" in
which the risk coefficient dec-reases with increasing dose (downward
bending, or convex, dose response curve). Such models imply that the
risk at low doses would actually be greater than predicted by linear
interpolation from higher doses.
The evidence from radiation biology investigations, at the cellular
as well as the whole animal level, indicates that the dose response curve
for induction of mutations or cancer by low-LET radiation is either
linear or concave upward for doses to mammalian systems below about
2jO rads (NCRP80). Somewhere above this point the dose response curve
often begins to bend over: this is commonly attributed to
cell-killing." Analysis of the A-bomb survivor data, upon which most of
our risk estimates depend, is dominated by individuals receiving about
250 rads or less. Consequently, the cell-killing phenomenon should not
produce a substantial underestimate of the risk at low doses.
_ Noting that human beings, in contrast to pure strains of laboratory
animals, may be highly heterogeneous with respect to radiation
sensitivity, Baum (Ba73) proposed an alternative mechanism by which a
convex dose response relationship could arise. He pointed out that
sensitive subgroups may exist in the population who are at very high risk
from radiation. The result could be a steep upward slope in the response
at low doses, predominantly reflecting the elevated risk to members of
these subgroups, but a decreasing slope at higher doses as the risk to
these'highly sensitive individuals approaches unity.
Based on current evidence, however, it seems unlikely that the
effect postulated by Baum would lead to substantial overestimation of the
risk at low doses. While there may indeed be small subgroups at vejy —
high risk, it is difficult to reconcile the A-bomb survivor data: with a
strongly convex dose response relationship. For example, if most of the
leukemias found among the cohort receiving about 200 rads or more in fact
arose from subgroups whose risk saturated below 200 rads, then many more
leukemias ought to have occurred in lower dose cohorts than were actually
observed (Ro78). The U.S. population, it could be argued, may be more
heterogeneous with respect to radiation sensitivity than the Japanese.
The risk of radiation-induced breast cancer appears, however, to be
similar in the two populations, so it is difficult to see how the size of
.7-8
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the hypothetical sensitive'group could be large enough in the former to
alter the conclusion reached above. The linear dose-response
relationship seen for radiogenic breast cancer in several populations •
(NIH85) further argues against Baum's hypothesis.
7.2.3 The Possible Effects of Dose Rate on Radiocarcinogenesis
The BEIR-3 Committee limited its risk estimates to a.minimum dose
rate of 1 rad per year and stated that it "does not know if-dose rates of
gamma rays and x rays of about 100 mrad/yr are detrimental to man. At
dose rates comparable to the background everyone receives from
naturally-occurring radioactive materials, a considerable body of
scientific opinion holds that the effects of radiation are reduced
compared to high dose rates. NCRP Committee 40 has suggested that
carcinogenic effects of low-LET radiations may be a factor of from 2 to
10 times less per unit dose for small doses and dose rates than have been
observed at high doses and dose rates (NCRP80).
The low dose and low dose rate effectiveness factors estimated by
NCRP Committee 40 are based on their analysis of a large body of plant
and animal data that showed reduced effects at low doses for a number of
biological endpoints, including radiogenic cancer in animals, chiefly
rodents. However, no data for cancer in humans confirm these findings as
yet; indeed, a few human studies seem to contradict them. Highly
fractionated small doses to human breast tissue are apparently as
carcinogenic as large acute doses (NAS80, La80). Furthermore, small
acute (less than 10 rads) doses to the thyroid have been found to be as
effective per rad as much larger doses in initiating thyroid cancer
(UNSCEAR77, NAS80). Moreover, the increased breast cancer resulting from
.chronic, low dose, occupational, gamma ray exposures among British dial
painters is comparable to, or larger than, that expected on the basis ot
acute, high dose exposures (Ba81).
While none of these examples is persuasive by itself, collectively
they indicate that it may not be prudent to assume that all kinds of
cancers are reduced at low dose rates and/or low doses. However, it may
be overly conservative* to estimate the.risk of all cancers on the basis
of the linearity observed for breast and thyroid cancer. The ICRP and
UNSCEAR have used a dose rate effectiveness factor (DREF) of about 2.5 to
estimate the risks from occupational (ICRP77) and environmental exposures
(UNSCEAR77). That choice of a DREF is fully consistent with and
equivalent to the reduction of risk at low doses obtained by substituting
the BEIR-3 linear-quadratic response model for their linear model (see
above). Therefore, use of both a DREF and a linear-quadratic model for
risk estimation in the low-dose region is inappropriate (NCRP80).
*In carrying out risk assessments, one is often forced to choose
among alternative assumptions, norie of which can be definitively
shown to be more accurate than the others. A conservative choice
in this connection, is one leading to higher estimates of risk.
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7.2.4 Risk Projection Models
None of the exposed populations have been observed long enough to
assess the full effects of their exposures if, as currently thought, most
radiogenic cancers occur throughout an exposed person's lifetime
(NAS80). Therefore, another major choice that must be made in assessing
tile lifetime cancer risk due, to radiation is to select a risk projection
model to estimate the risk for a longer period of time than currently
available observational data will allow.
To estimate the risk of radiation exposure that is beyond the years
of observation, either a relative risk or an absolute risk projection
model (or suitable variations) may be used. These models are described
at length in Chapter 4 of the 1980 NAS report (NAS80). The relative risk
projection model projects the currently observed percentage increase in
annual cancer risk per unit dose into future years, i.e., the increase is
proportional to the underlying (baseline) risk. An absolute risk model
projects the average annual number of excess cancers per unit dose into
future years at risk, independent of the baseline risk.
_ Because the underlying risk of most types of cancer increases
rapidly with age, the relative risk model predicts a larger probability
or excess cancer toward the end of a person's lifetime. In contrast, the
absolute risk model predicts a constant incidence of excess cancer across
time. Therefore, given the incomplete data we have now, less than
lifetime follow-up, a relative risk model projects a somewhat greater
total lifetime cancer risk than that estimated using an absolute risk
model.
Neither the NAS BEIR Committee'nor other scientific groups (e.g.,
UNSCEAR) have concluded which projection model is the appropriate choice
tor most radiogenic cancers. However, recent evidence favors the
relative risk projection model for most solid cancers. As pointed out by
the 1980 NAS BEIR Committee:
"If the relative-risk model applies, then the age of the exposed
groups, both at the time of exposure and as they move through
life, .becomes very important. There is now considerable
evidence in nearly all the adult human populations studied that
persons irradiated at higher ages have, in general, a greater
excess risk of cancer than those irradiated at lower ages, or at
least^they develop cancer sooner. Furthermore, if they are
irradiated at a particular age, the excess risk tends to rise
7-10
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pari passu (at equal pace) with the risk of the population at
iSrTe. IS other words, the relative-risk model with respect to
cancer susceptibility at least as a function of age, evidently
applies to some kinds of cancer that have been observed to
result from radiation exposure." INAS80, p.33;
This observation is confirmed by the Ninth A-bomb Survivor Life Span
Study, published two years after the 1980 Academy report, This latest
report indicates that, for solid cancers, relative risks have continued
to remain constant in recent years, while absolute risks have increased
substantially (Ka82). Smith and Doll (Sm78) reached similar conclusions
on the trend in excess cancer with time among the irradiated spondylitic
patients. •
Although we believe considerable weight should be given to the
relative risk model for most solid cancers (see below), the model does
not necessarily give an accurate projection of lifetime risk. The mix ot
tumor types varies with age so that the relative frequency of some common
radiogenic tumors, such as thyroid cancer, decreases for older ages.
Land has pointed out that this may result in overestimates of the
lifetime risk when they are based on a projection model using relative
risks (La83). While this may turn out to be true for estimates ot cancer
incidence that include cancers less likely to be fatal, e.g., thyroid, it
may not be too important in estimating the lifetime risk of fatal
cancers, since the incidence of most of the common fatal cancers, e.g.,
breast and lung cancers, increases with age.
Leukemia and bone cancer are exceptions to the general validity of a
lifetime expression period for radiogenic cancers. Most, if not ail, ot
the leukemia risk has apparently already been expressed in both the
A-bomb survivors and the spondylitics (Ka82, Sm78>. Similarly, bone
sarcoma from acute exposure appears to have a limited express ion Pe^°d
UAS80, Ma83). For these diseases, the BEIR-3 Committee believed that an
absolute risk projection model with^a limited expression period is
adequate for estimating lifetime risk ^NAS80).
Note that, unlike the NAS BEIR-1 report (NAS72), the BEIR-3
Committee's relative and absolute risk models are age dependent; that is,
the risk coefficient changes, depending on the age of the exposed
persons. Observational data on how cancer risk resulting ^om.^iafon
changes with age are sparse, particularly so in the case of childhood
exposures. Nevertheless, the explicit consideration of the variation in
radiosensitivity with age at exposure is a significant improvement in
methodology. It is important to differentiate between age sensitivity at
exposure and the age dependence of cancer expression. In general, people
seem to be most sensitive to radiation when they are young. In contrast
most radiogenic cancers seem to occur late in life, much like cancers
resulting from other causes. In this chapter we present lifetime cancer
risk estimates for a lifetime exposure of equal annual doses. H°wever,
it is important to note that the calculated lifetime risk of developing a
7-11
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fatal cancer from a single year of exposure varies with the age of the
recipient at the time of exposure.
7'2"5 Effect of Various Assumptions on the Numerical Risk Estimates
Differences between risk estimates made by using various
combinations of the assumptions described above were examined in the 1980
NAS report. Table 7-1 below, taken from Table V-25 (NAS80), shows the
range of cancer fatalities that are induced by a single 10-rad dose as
estimated using linear, linear-quadratic, and quadratic dose response
functions and two projection models, relative and absolute risk (NAS80).
_ As illustrated in Table 7-1, estimating the cancer risk for a given
projection model on the basis of a quadratic as compared to a linear dose
response reduces the estimated risk of fatal cancer by a factor of about
18. Between the more credible linear and linear-quadratic response
functions, the difference is less, a factor of about 2.2. For a given
dose response model, results obtained with the two projection models for
solid cancers, differ by about a factor of 3.
Even though the 1980 NAS analysis estimated lower risks for a
linear-quadratic response at 10 rads, it should not be concluded that
this response function always provides smaller risk estimates. In
contrast to the 1980 NAS analysis where the proportion of risk due to the
dose-squared term (e.g., C3 in Footnote c of Table 7-1) was constrained
to positive values, the linear-quadratic function that agrees best with
Nagasaki cancer incidence data has a negative coefficient for the
dose-squared term (Wa83). Although this negative coefficient is small
and indeed may not be significant, the computational result is a larger
linear terra, which leads to higher risk estimates at low doses than would
be estimated using a simple linear model (Wa83).
Differences in the estimated cancer risk introduced by the choice of
the risk projection model are also appreciable. As pointed out above
the 1980 NAS analysis indicates that relative lifetime risk estimates
exceed absolute risk estimates by about a factor of 3 (see Table 7-1).
However, relative risk estimates are quite sensitive to how the risk
resulting from exposure during childhood persists throughout life. This
question is addressed in the next section, where we compare risk
estimates made by the 1972 and 1980 NAS BEIR Committees with those of the
ICRP and UNSCEAR.
7>2'6 Comparison of Cancer Risk Estimates for Low-LET Radiation
A number of estimates of the risk of fatal cancer following lifetime
exposure are compared in Table 7-2. The BEIR-1 and BEIR-3 values were
calculated for this table using risk model data from NAS72 and NAS80. .
The BEIR-3 values in this table differ slightly from those in NAS80 and
Table 7-1 because of some minor calculational corrections including
revised age-specific mortality data. The risk estimates in this table
7-12
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Table 7-1.
Range of cancer fatalities induced by a single 10-rad,
low-LET radiation exposure to the general population
(Average value per rad per million persons exposed)
Dose response
functions
Lifetime risk projection model
Relative3 Absolute
Linear'3
Linear Quadratic0
Quadratic^
501
226
28
167
77
10
a Relative risk projection for all solid cancers except
leukemia and bone cancer fatalities, which are projected
by means of the absolute risk model (NAS80).
b Response R varies as a constant times the dose, i.e.:
R=C D.
C =
(L,L)
(LQ.LQ)
See text for model notation.
d R=C4D2 (Q,Q)
Source: NAS80, Table V-25.
7-13
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Table 7-2. A comparison of estimates of the risk of fatal
cancer from low-LET radiation
Source of
estimate
BEIR-1
BEIR-1
BEIR-3
BEIR-3
BEIR-3
BEIR-3
UNSCEAR
UNSCEAR
CLM
ICRP
Reference
NAS72
NAS80
NAS80
NAS80
NAS80
NAS80
UNSCEAR? 7
UNSCEAR77
Ch83
ICRP77
Fatalities per
*
10 person-rad
118
622
168
395
71
163
.200-300
75-175
100-400
125
Projection model
Absolute3
Relative3
Absolute3'13
Relativea> b>c
Absolute3 »d
Relative3 >c>d
None6 - high dose
(j>100 rad)
None6 - low dose,
dose rate
None - UNSCEAR77
without A-bomb data
None - Occupational
low dose, low dose rate
a Lifetime projection for constant dose rate calculated for
1970 U.S. general population lifetable and mortality rates;
see text.
Linear model (L-L for leukemia and bone, L-L for all other
sites; notation explained in text).
c Leukemia and bone are calculated with absolute risk model.
d Linear-Quadratic model (LQ-L for leukemia and bone, LQ-L
for all other sites; notation explained in text).
e Taken from paragraphs 317 and 318 in UNSCEAR77.
7-14
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are based on different assumptions regarding the extrapolation to low
doses and dose rates; they also differ considerably because of other
assumptions. In contrast with absolute risk estimates, which have
increased since the 1972 NAS BEIR-1 Committee report (NAS72),.the 1980
NAS BEIR-3 Committee's estimates of the relative risk, as shown in
Table 7-2, have decreased relative to those in the BEIR-1 report. This
illustrates the sensitivity of risk projections to changes in modeling
assumptions. For the NAS80 report, the relative risk coefficient
determined for ages 10 to 19 was substituted for the considerably higher
relative risk coefficient that would be calculated for those exposed
during childhood, ages 0 to 9. In addition, the relative risk
coefficients used in the BEIR-3 analysis are based on matching excess
cancer for a 30-year follow-up of Japanese A-bomb survivors with 1970
U.S. lifetime and cancer mortality rates. In the 1972 NAS report this
excess was compared to cancer mortality in Japan.
By comparing the three relative risk estimates from Table 7-2, it is
apparent that the relative risk estimates are fairly sensitive to the
assumptions made as to what extent the observed high relative risk of
cancer from childhood exposure continues throughout adult life. The Life
Span Study (Ka82) indicates that the high-risk adult cancer caused by
childhood exposures is continuing, although perhaps not to the extent
predicted by the NAS BEIR-1 Committee in 1972.
The major reason the risk estimates in Table 7-2 differ^is
because of the underlying assumption in each set of risk estimates. The
NAS BEIR estimates are for lifetime exposure and lifetime expression of
induced cancers (NAS72, NAS80). Neither the age distribution of the
population at risk nor the projection models (if any) have been specified
by either the UNSCEAR (UNSCEAR77) or the ICRP (ICRP77). UNSCEAR
apparently presumes the same age distributions as had occurred in the
epidemiological studies they cited, mainly the A-bomb survivors, and a
40-year period of cancer expression. The ICRP risk estimates are. for
adult workers, presumably exposed between ages 18 and 65, and a similar
expression period. These are essentially age-independent absolute risk
models with less than lifetime expression of induced cancer mortality.
For these reasons, risks estimated by ICRP and UNSCEAR are expected to be
smaller than those made on the basis of a lifetime relative risk model in
the BEIR-3 report.
The next to the last entry in Table 7-2 (Ch83) is of interest
because it specifically excludes the A-bomb survivor data based on T65
dose estimates. The authors reanalyzed the information on radiogenic
cancer in UNSCEAR77 so as to exclude all data based on the Japanese
experience. Their estimate of fatalities ranges from 100 to 440 per
10" person-rad based on data from exposure at high doses and dose
rates. As indicated in Table 7-2, this is somewhat greater but
comparable to the UNSCEAR estimate, which includes the A-bomb survivor
data., The upper bound estimate for the number of fatalities given in
Ch83 is 400 per 106 person-rem, which is nearly identical to the value
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EPA has used in this report for a linear dose response model—395
fatalities per 10^ person-rad (see below).
7.2.7 EPA Assumptions About Cancer Risks
Resulting from Low-LET Radiations
The EPA estimates of radiation risks, presented in Section 7.2.8
below, are based on a presumed linear dose response function. We
believe, however, that the linear-quadratic model is also credible.
Using the BEIR-3 linear-quadratic model is equivalent to using a dose
rate effectiveness factor of 2.5; thus, at low doses, it would project
2.5 times lower risk than the linear model.
Except for leukemia and bone cancer, where we use a 25-year
expression period for radiogenic cancer, we use a lifetime expression
period, as was done in the NAS report (NAS80). Because the most recent
Life Span Study Report (Ka82) indicates absolute risks for solid cancers
are continuing to increase 33 years after exposure, the 1980 NAS
Committee choice of a lifetime expression period appears to be well
founded. We do not believe limiting cancer expression to 40 years (as
has been done by the ICRP and UNSCEAR) is compatible with the continuing
increase in solid cancers that has occurred among irradiated populations
(Ka82).
To project the number of fatalities resulting from leukemia and bone
cancer, EPA uses an absolute risk model, a minimum induction period of
2 years, and a 25-year expression period. To estimate the number of
fatalities resulting from other cancers, EPA has used the arithmetic
average of absolute and relative risk projection models (EPA84). For
these cancers, we assume a 10-year minimum induction period and
expression of radiation-induced cancer for the balance of an exposed
person's lifetime after the minimum induction period.
7'2.8 Methodology for Assessing the Risk of Radiogenic Cancer
EPA uses a life table analysis to estimate the number of fatal
radiogenic cancers in an exposed population of 100,000 persons. This
analysis considers not only death due to radiogenic cancer, but also the
probabilities of other competing causes of death which are, of course,
much larger and vary considerably with age (Bu81, Co78). Basically, it
calculates for ages 0 to 110 the risk of death due to all causes by
applying the 1970 mortality data from the National Center for Health
Statistics (NCHS75) to a cohort of 100,000 persons. Additional
information on the details of the life table analysis is provided in
Appendix B. It should be noted that a life table analysis is required to
use the age-dependent risk coefficients in the BEIR-3 report. For
relative risk estimates, we have used age-specific cancer mortality data
also provided by NCHS (NCHS73). The EPA computer program we use for the
life table analysis was furnished to the NAS BEIR-3 Committee by EPA and
7-16
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used by the Committee to p.repare its risk estimates. Therefore, the
population base and calculations should be essentially the same in both
the NAS and EPA analyses.
We have considered both absolute and relative risk models to project
the observed risks of most solid radiogenic cancers beyond the period of
current observation. As indicated in Table 7-2, the range of estimated
fatal cancers resulting from the choice of a particular projection model
and its internal assumptions is about a factor of 3. Although the
relative risk model has only been tested in some detail for lung and
breast cancer (La78), based on current evidence, it appears to be the
better projection model for solid cancers. We have, therefore, adopted
it for our risk estimates in this report. Previously, we have used an
average of the risks calculated by the absolute and relative risk
projection models (EPA84).
To estimate the cancer risk from low-LET, whole-body, lifetime
exposure,' we use the relative risk projections (the BEIR-3 XFL model) for
solid cancers and the absolute risk projection for leukemia and bone
cancer (the BEIR-3 L-L model). ' Since the expression period for leukemia
and bone cancer is less than the follow-up period, the same risk values
would be calculated for these cancers using either projection method.
For a dose to the whole body, this procedure yields about 400 fatalities
per million person-rad: For the BEIR-3 linear-quadratic model, which is
equivalent to applying a DREF of about 2*5 to the linear model, a low-LET
whole-body dose yields an estimated lifetime risk of about 160 fatalities
per million person-rad.
BEIR-3 also presented estimates of excess soft tissue cancer
incidence for specific sites, as a function of age at exposure, in their
Table V-14. By summing the site-specific risks, they then arrived at an
estimate for the whole-body risk of cancer incidence (other than leukemia
and bone cancer) as given in Table V-30. Finally, by using the weighted
incidence/mortality ratios given in Table V-15 of the same report
(NAS80),,the results in Table V-30 can be expressed in terms of mortality
to yield (for lifetime exposure) a risk estimate of about 242 and 776
cancer fatalities per 106 person-rad, depending on whether an absolute
or a relative risk projection model, respectively, is used to estimate
lifetime risk. These values are about 1.6 and 2.1 times their counterparts
for the BEIR-3 ~C=L model and 3.9 and 9.1 times the LQ-L values.
These models all presume a uniform dose to all tissues at risk in
the body. In practice, such uniform whole-body exposures seldom occur,
particularly for ingested or inhaled radioactivity. The next section
describes how we apportion this risk estimate for whole-body exposure
when considering the risks following the exposure of specific organs.
7.2.9 Organ Risks
For most sources of environmental contamination, inhalation and
ingestion of radioactivity are more common than external exposure. In
7-17
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many cases, depending on the chemical and physical characteristics of the
radioactive material, inhalation and ingestion result in a nonuniform
distribution of radioactive materials within the body so that some organ
systems receive much higher doses than others. For example, since iodine
isotopes concentrate preferentially in the thyroid gland, the dose to
this organ can be orders of magnitude larger than the average dose to the
body.
Fatal Cancer at Specific Sites
To determine the probability that fatal cancer occurs at a
particular site, we have performed life table analyses for each cancer
type using the information on cancer incidence and mortality in NAS80. '
For cancer other than leukemia and bone cancer, we have used NAS80
lable V-14 Uge Weighted Cancer Incidence by Site Excluding Leukemia and
Bone Cancer) and NAS80 Table V-15*, which lists the BEIR Committee's
estimates_of the ratio of cancer fatality to cancer incidence for these
various sites, to calculate a set of site-specific mortality risk
coefficients. The excess mortality for the T^TTmodel was presumed to be
distributed similarly. The proportions of leukemia and fatal bone cancer
caused by low-LET radiation were estimated using the results of the
models given in Table V-17** of NAS80. Normalized results, which give
the proportion of fatal radiogenic cancers resulting from uniform
whole-body irradiation, by cancer site, are listed in Table 7-3. These
proportions were calculated for the average of the absolute and relative
risk projections as in EPA84. • Since it was not practicable to reanalyze
all the scenarios considered for this report, we have adjusted the
original risk estimate by the factor 395/280 to approximate the effects
of using a relative risk projection model for solid cancers. As noted
above, these proportions are assumed to be the same for the BEIR-3
linear-quadratic dose response model.
Information on the distribution of fatal, radiogenic cancers by
organ is not precise. One reason is that the data in NAS80 (and
Table 7-3) are based on whole-body exposures, and it is possible that the
incidence of radiogenic cancer varies depending on the number of exposed
organs. Except for breast and thyroid cancer, very little information is
available on radiogenic cancer resulting from exposure of only one region
in the body.
* The mortality to incidence ratio for thyroid (male: 0.18,
female: 0.20) in NAS80 is.high compared to other references, e.g.,
NCRP80 Uses a mortality to incidence ratio for thyroid of 0.1 for
both males and females.
** The low-LET risk rate coefficient for bone has been changed to
0.125xlO~b sarcoraa/yr per person-rad to be consistent with an alpha
particle RBE of 8.
7-18
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Table 7-3. Proportion of the total risk of fatal
radiogenic cancer among different sites
as given in EPA 84e
Site
Proportion of
total risk
Fatalities per
10" person-rad"
Lung
Breast3
Red bone marrow'5
Thyroid
Bone surface
Liver
Stomach
Intestines
Pancreas
Kidneys and urinary tract
Other0
0.207
0.130
0.150
0.099
0.009 .
0.085
0.084
0.039
0.059
0.025
0.113
58.2
36.4
•42.1
27.7
2.4 •
23.9
23.6
10.9 '
16.4
7.0
31.8
Total
280.4
a Average tor both sexes.
b Leukemia.
c Total risk for all other organs, including the esophagus,
lymphatic system, pharynx, larynx, salivary gland, and brain.
d Lifetime exposure and cancer expression. Total risk for
all sites calculated using the L-L absolute risk model for
leukemia and bone and the L-L model for the total of the
remaining sites averaged for absolute and relative risk
projections (EPA84). The risks for these remaining sites
have been apportioned as for the site specific absolute risk :
model using NAS80 Tables V-14 and V-15.,
e In projecting cancer deaths for this proposed rulemaking, the organ
risks above have been scaled up by a factor of 395/280 (see text).
7-19
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Another reason is that most epidemiological studies use mortality
data from death certificates, which often provide questionable
information on the site of the primary cancer. Moreover, when the
existing data are subdivided into specific cancer sites, the number of
cases becomes small, and sampling variability is increased. The net
result of these factors is that numerical estimates of the total cancer
risk are more reliable than those for most single sites.
The 1977 UNSCEAR Committee's estimated risks (UNSCEAR77) to
different organs are shown in Table 7-4. For all of the organs, except
the breast, a high and low estimate was made. This range varies by a
factor of 2 or more for most organs (Table 7-4). Other site-specific
estimates show a similar degree of uncertainty (Ka82), and it is clear
that any system for allocating the risk of fatal cancer on an
organ-specific basis is inexact. Table 7-5 compares proportional risks
by the NAS BEIR-3 Committee, UNSCEAR, and the ICRP. ICRP Report 26
provides organ-specific weights for assessing combined genetic and cancer
risks from occupational exposure (ICRP77). In Table 7-5, we have
renormalized ICRP risks so that they pertain to cancer alone.
Considering that the cancer risk for a particular site is usually
uncertain by a factor of 2 or more, as indicated by the range of UNSCEAR
estimates in Table 7-4, we would not expect perfect agreement in
apportionment of total body risks. Table 7-5, however, does indicate
reasonable agreement among the three sets of estimates considered here.
The differences between the proportions of the total risk of fatal
cancer shown in Table 7-5 are, for the most part, small in comparison to
their uncertainty. We have used the BEIR-3 organ risks in preference to
those made by other groups such as UNSCEAR or the ICRP for several
reasons. BEIR estimates of organ risk are based on a projection of
lifetime risk using age-specific risk coefficients, rather than just
observations to date. Moreover, the 1980 BEIR Committee considered
cancer incidence data as well as mortality data. This gives added
confidence that the diagnostic basis for their estimates is correct.
And, finally, because we apply these proportional organ risk estimates to
the NAS80 cancer risk estimates for whole-body exposures, we believe it
is consistent to use a single set of related risk estimates. The way we
have used NAS80 to estimate mortality resulting from cancer at a
particular site is outlined in the next section.
7-2.10 'Thyroid Cancer from Iodine-131 and Iodine-129
Iodine-131 has been reported to be only one-tenth as effective as
x rays or gamma rays, in inducing thyroid cancer (NAS72, NCRP77, NCRP85).
On this basis, EPA has employed a thyroid cancer risk coefficient for
internal exposures to iodine-131 and iodine-129 which is one-tenth that
used for gamma rays or beta radiations from other radionuclides.
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Table 7-4. UNSCEAR estimates of cancer risks (fatalities per
person-rad) at specified sites
Range
Site
Average
a Average for both sexes.
" Leukemia.
c Includes esophagus and lymphatic tissues.
Source: Adapted from UNSCEAR77.
Proportion of
total risk
Lung
Breasta
Red bone marrow'3
Thyroid
Bone
Liver
Stomach
Intestines
Pancreas
Kidneys and urinary
tract
Other0
25-50
25
15-25
5-15
2-5
10-15
10-15
14-23
2-5
2-5
4-10
37.5
25.0
20.0
10.0
3.5
12.5
12.5
17.5
3.5
3.5
7.0
. 245 •
.164
.131
.066
.023
.082
.082
.115
.023
.023
.046
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Table 7-5. Comparison of proportion* of the total risk of
radiogenic cancer fatalities by body organ
Site/Source
Lung
Breast
Red Marrow
Thyroid
Bone Surface
Remainder
Liver
Stomach
Intestine
Pancreas
Urinary
Other6
EPA84a>b
0.21
0.13
0.16
0.10
0.01
—
0.08
0.08
0.04
0.06
0.02
0.11
UNSCEAR77
0.25
0.16
0.13
0.07
0.02
__
0.08
0.08
0.12
0.02
0.02
0.05
ICRP77C
0.16 .
0.20
0.16
0.04
0.04
0.40d
— _
— —
* Values rounded to 2 decimal places.
a Lifetime exposure and cancer expression. UNSCEAR and ICRP estimates
use different age distributions and periods of expression.
° EPA Radionuclides Background Information Document; EPA 520/1-84-022-1
(EPA84). Also see Table 7-3 and text.
c Normalized for risk of fatal cancer (see text).
d Five additional target organs which have the highest doses are
assigned 0.08 each for a total of 0.4.
e Other includes esophagus, lymphatic system, pharynx, larynx, salivary
gland, and brain.
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7.2.11 Cancer Risks for a Constant Intake Rate
The fatal cancer risks shown in the tables of this chapter presume a
lifetime exposure at a constant dose rate. Even for a dosimetric model
with age invariant parameters, dose rates vary with time for a constant
intake rate. This variation reflects the time dependent activity levels
associated with the retention of the radionuclide in the organs and
tissues. The ingrowth of radioactive decay products can also contribute
further to the time dependence of dose rates.
Traditionally, risk estimates for chronic intake of a radionuclide
have been determined using a dose commitment model to calculate dose
rates following a fixed period (e.g., a 70-year average lifespan). For
the purpose of estimating risk, these dose rates are considered to be
invariant over the individual's lifetime. This approach is overly
conservative for estimating risk for many long-lived radionuclides.
Therefore, EPA estimates risks for constant radionuclide intakes by first
determining dose rates to each radiosensitive organ or tissue as a
function of time. Then these dose rates and the risk models of this
chapter.are used to calculate lifetime risk based on 1970 life table
data. The resulting risks are consistent with both the dosimetric and
risk models, and the arbitrary choice of a dose commitment period is
avoided.
7.3 Fatal Cancer Risk Resulting from High-LET Radiations
In this section we explain how EPA estimates the risk of fatal
cancer resulting from exposure to high-LET radiations. Unlike exposures
to x rays and gamma rays where the resultant charged particle flux
results in linear energy transfers (LET) of the order of 0.2 to 2 keV per
Dm in tissue, 5-MeV alpha particles result in energy deposition at a
track average rate of more than 100 keV per Dm. High-LET radiations
have a larger biological effect per unit dose (rad) than low-LET
radiations. How much greater depends on the particular biological
endpoint. being considered. For cell killing and other readily observed
endpoints, the relative biological effectiveness (RBE) of high-LET alpha
radiations is often 10 or more times greater than low-LET radiations.
The RBE may also depend on the dose level; for example, if linear and
linear-quadratic dose response functions are appropriate for high- and
low-LET irradiations, respectively, then the RBE will decrease with
increasing dose.
7.3.1 Quality Factors and RBE for Alpha Particles
For purposes of calculating dose equivalent, each^type of
biologically important ionizing radiation has been assigned a quality
factor, Q, to account for its relative efficiency in producing biological
damage. Unlike an RBE value, which is for a specific tissue and
well-defined endpoint, a quality factor is based on an average overall
assessment by radiation protection experts of potential harm of a given
7-23
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radiation relative to x- or gamma radiation. In 1977, the ICRP assigned
a quality factor of 20 to alpha particle irradiation from radionuclides
(ICRP77). However, the appropriateness of this numerical factor for
estimating fatal radiogenic cancers is still unclear —- particularly fo:r
individual sites.
The dose equivalent, in rem, is the dose, in rad, times the
appropriate quality factor for a specified kind of radiation. For the
case of internally deposited alpha-particle emitters, the dose equivalent
from a one-rad dose is 20 rem. It should be noted that prior to ICRP
Report 26 (ICRP79), the quality factor assigned, to alpha particle
irradiation was 10. That is, the biological.effect from a given dose of
alpha particles was estimated to be 10 times that from an acute dose of
low-LET x rays or gamma rays of the same magnitude in rad. The ICRP
decision to increase this quality factor to 20 followed from their
decision to estimate the risk of low-LET radiations, in occupational
situations, on the assumption that biological effects were reduced at low
dose rates. There is general agreement that dose rate effects do not
occur for high-LET (alpha) radiations. Implicit in ICRP's risk estimates
for lox<7 dose/dose rate gamma radiation is a dose rate reduction factor of
about 2.5. The EPA (linear) risk model for low-LET radiation does not
involve such a DREF; therefore, in order to avoid an artifactual
inflation in our high-LET risk estimates, we have assumed an RBE of 8
(20/2.5) for calculating the. risks from alpha particles (see
Section 7.3.3).
In 1980 the ICRP published the task group report "Biological Effects
of Inhaled Radionuclides," which compared the results of animal
experiments on radiocarcinogenesis following the inhalation of
alpha-particle and beta-particle emitters (ICRP80). The task group
concluded that: "...the experimental animal data tend to support the
decision by the ICRP to change the recommended' quality factor from 10 to
20 for alpha radiation."
7.3.2 Dose Response Function
In the case of high-LET radiation, a linear dose response is
commonly observed in both human and animal studies. This response is not
reduced at low dose rates (NCRP80). Some data on human lung cancer
indicate that the carcinogenic response per unit dose of alpha radiation
is higher at low doses than higher ones (Ar81, Ho81, Wh83); in addition,
some studies with animals show the same response (Ch81, U182). We agree
with the NAS BEIR-3 Committee that: "For high-LET radiation, such as
from internally deposited alpha-emitting radionuclides, the linear
hypothesis is less likely to lead to overestimates of the risk and may,
in fact, lead to underestimates" (NAS80). However, at low doses,
departures from linearity are small compared to the uncertainty in the
human epidemiological data, and we believe a linear response provides an
adequate model for evaluating risks in the general environment.
7-24
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A possible exception to a linear response is provided by the data
for bone sarcoma (but not sinus carcinoma) among U.S. dial painters who
ingested alpha-emitting radium-226 (NAS80). These data are consistent
with a dose-squared response (Ro78). Consequently, the NAS BEIR-3
Committee estimated bone cancer risk on the basis of both linear and
quadratic dose response functions. However, as pointed out in NAS80, the
number of U.S. dial painters at risk who received less than 1,000 rads
was so small that the absence of excess bone cancer at low doses is not
statistically significant. Therefore, the consistency of these data with
a quadratic (or threshold) response is not remarkable and, perhaps, not
relevant to evaluating risks at low doses. In contrast to the dial
painter data, the incidence of bone cancer following short-lived
radium-224 irradiation, observed in spondylitics by Mays and Spiess
(Ma83, NAS80), in a larger sample at much lower doses, is consistent with
a linear response. Therefore, for high-LET radiations EPA has used a
linear response function to evaluate the risk of bone cancer.
Closely related to the choice of a dose response function is what
effect the rate at which a dose of high-LET radiation is delivered has on
its carcinogenic potential. This is an active area of current research.
There is good empirical evidence, from both human and animal studies,
that repeated exposures to radium-224 alpha particles is 5 times more
effective in inducing bone sarcomas than a single exposure which delivers
the same dose (Ma83, NAS80). The 1980 NAS BEIR Committee took this into
account in its estimates of bone cancer fatalities, which EPA is using.
We do not know to what extent, if any, a similar enhancement of
carcinogenicity may occur for other cancers resulting from internally
deposited alpha-particle emitters.
7.3.3 Assumptions Made by EPA for Evaluating
the Dose from Alpha-Particle Emitters
We have evaluated the risk to specific body organs by applying an
RBE of 8 for alpha radiations to the risk estimates for low dose rate
low-LET radiations as described above. For some organs, this factor may
be too large. Several authors have noted that estimates of.leukemia
based on an RBE of 20 for bone marrow alpha irradiation (relative to a
low dose rate low-LET risk model which includes a DREF of 2.5)
overpredicts the observed incidence of leukemia in persons receiving
thorptrast (thorium oxides) (Mo79) and in the U.S. radium dial painters
(Sp83). Nevertheless, in view of the paucity of applicable human data
and the uncertainties discussed above, the ICRP quality factor provides a
reasonable and prudent way of evaluating the risk due to alpha emitters
deposited within body organs.
All EPA risk estimates for high-LET radiations are based on a linear
dose response function. For bone cancer and leukemia we use the absolute
risk projection model described in the previous section.
cancers we use relative risk projections.
For other
7-25
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Table 7-6 indicates the Agency's estimates of the risk of fatal
cancer due to a uniform organ dose in various organs from internally
deposited alpha-particle emitters. It was prepared by multiplying the
average risk based on the linear model for a uniformly distributed
whole-body dose of low-LET radiation by an RBE of 8 and then apportioning
this risk by organ, as indicated in Table 7-6. These estimates are for
lifetime doses at a constant dose rate. This procedure was not followed
for bone cancer. As outlined above, the risk estimate for this cancer in
the BEIR-3 report is based directly on data for high-LET (alpha)
radiation. ;
Some readers may note that the risk estimate in Table 7-6, 19 bone
cancer fatalities per 106 person-rad, is less than the 27 fatalities
listed in Table A-27 of (NAS80) for alpha particles. This is because the
analysis in Appendix A of NAS80 (but not Chapter V of that report)
assumes that in addition to a 2-year minimum induction period, 27 years
are available for cancer expression. This is usually not the case for
doses received beyond about age 50. Hence, the-estimated lifetime risk
is smaller when it is based on a life table analysis that considers
lifetime exposure in conjunction with competing causes of death.
In the next section, we describe how we estimate the risk due to
inhalation of alpha-emitting radon progeny, a situation where the organ
dose is highly nonuniform.
7.4 Estimating the Risk Resulting from Lifetime
Population Exposures from Radon-222 Progeny
The Agency estimates of the risk of lung cancer due to inhaled radon
progeny do not utilize the dosimetric approach, outlined above, but
rather are basedton what is sometimes called an epidemiological approach,
that is, on the excess human lung cancer in groups known to have been
exposed to radon progeny.
When radon-222, a radioactive noble gas, decays, a number of short
half-life radionuclides, principally polonium-218, lead-214, bismuth-214,
and polonium-214, are formed, some of which attach to inhalable dust
particles in air. When inhaled, the radon progeny are deposited on the
surfaces of the larger bronchi of the lung. Since two of these
radionuclides decay by alpha-particle emission, the bronchial epithelium
is irradiated by high-LET radiation. There is a wealth of data
indicating that a range of exposures to the bronchial epithelium of
underground miners causes an increase in bronchial lung cancer, both in
smoking and in nonsmoking miners. Two recent reviews on the underground
miner experience are of particular interest. The 1980 NAS BEIR-3 Report
(NAS80) contains a review of the epidemiological studies on these
miners. A lengthy report, "Risk Estimates for the Health Effects of
Alpha Radiation" by D. C. Thomas and K. C. McNeil for the Atomic Energy
Control Board (AECB) of Canada (Th82), reanalyzes many of these
epidemiological studies in a consistent fashion, so that the modeling
assumptions are similar for all of the data sets.
7-26
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Table 7-6. Estimated number of cancer fatalities from a lifetime
exposure to internally deposited alpha-particle
emitters as given in EPA 84*
Site
Proportional risk'
Fatalities per
10 person-rad
Lung
Breastc
Red bone marrowd
Thyroid
Bone surface6
Lxver
Stomach
Intestine
Pancreas
Kidneys and urinary tract
Other
0.207
0.130
0.150
0.099
0.009
0.085
0.084
0.039
0.059
0.028
0.113
466
291
337
222
19
191
189
87
131
56
254
Total
2243
a Proportion of whole-body risk from Table 7-3.
b Rounded to two figures.
c Average for both sexes.
d Leukemia.
e Bone surface (endosteum) as defined in ICRP-30 (ICRP79).
f As in the case of low-LET radiation, the organ risks above have been
scaled up by a factor of 395/280 for this proposed rulemaking.
7-27
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Although considerable progress has been made in modeling the
deposition of radon daughters in the lung (Ha82, Ja80, Ja81), it is not
yet possible to adequately characterize the bronchial dose delivered by
alpha radiation from inhaled radon-222 progeny. This is in part due to
the uncertainty concerning the kinds of cells in which bronchial cancer
is initiated (Mc78) and the depth of these cells in the bronchial
epithelium. Current estimates of the dose actually causing radiogenic
cancer due to inhaled radon-222 progeny are based on average doses which '
may or may not be relevant.
Even if accurate estimates of the dose delivered to the target cells
in the bronchial epithelium could be obtained for both low-LET and alpha
radiations, they would probably not be adopted as the basis for
estimating the risk to the public from airborne radon daughters. To do
so would mean extrapolating risk estimates derived from observations on
populations (particularly, the A-bomb survivors) receiving acute doses of
low-LET radiation over the whole lung to the case of chronic, nonuniform
lung doses from high-LET alpha irradiation. It would appear that more
reliable estimates of the risk can be derived on the basis of observed
cancers following occupational exposure to radon progeny, i.e., through
the epidemiological approach. Dosimetric considerations may nevertheless
be helpful in refining the risk estimates for the general population. In
particular, they were used, as discussed below, in formulating our
age-specific risk estimates for members of the general public through the
use of an "exposure equivalent."
7.4.1 Characterizing Exposures to the General
Population vis-a-vis Underground Miners
Exposures to radon progeny under working conditions are commonly
reported in a special unit called the working level (WL). One working
level is any combination of short half-life radon-222 progeny having 1.3
x 1Q5 MeV per liter of potential alpha energy (FRC67). This unit was
developed because the concentration of specific radon progeny depends on
ventilation rates and other factors. A working level month (WLM) is the
unit used to characterize a miner's exposure to one working level of
radon progeny for a working month of about 170 h. Because the results of
epidemiological studies are expressed in units of WL and WLM, we outline
below -how they can be interpreted for members of the general population
exposed to radon progeny.
For a given concentration of radon progeny, the amount of potential
alpha energy inhaled in a month by a member of the general population is
more than that received in a miner's working month. These individuals
are exposed longer, up to 24 h/da, 7 da/wk. However, the average amount
of air inhaled per minute (minute volume) by a member of the general
population is less than the amount for a working miner when such
activities as sleeping and resting are taken into account. To compare
the radon progeny exposure of a working miner to a member of the general
population, we have calculated the amount of potential alpha energy each
inhales per year.
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We have assumed that (averaged over a work day) a miner inhales
30 L/min. This average corresponds to about 4 h of light activity and
4 h of moderately heavy work per day (ICRP75). -We recognize that the new
ICRP radon model assumes a 20-L/min volume for miners, which corresponds
to 8 h of light activity per day (ICRP81). Although this may be
appropriate for nuclear workers, studies of the metabolic rate,of working
miners clearly show that they are not engaged only in light activity
(Sp56, ICRP75, NASA73). Therefore, we have chosen 30 L as a more
realistic estimate of their average minute ..volume. A working miner with
this minute volume inhales 3.6E+03 m^ in a working year of 2,000 h
(ICRP79). One working level of radon-222 progeny is 2.08E-05
joules/m^. Therefore, in a working year the potential alpha energy
inhaled by a miner exposed to one working level is 7.5E-02 joules.
For adult males and females in the general population, we follow the
ICRP Task Group on Reference Man (ICRP75) in assuming an inhaled air
volume of 23 m-^/da for males and 21 m-^/da for adult females. We use
the average of these two values, 22 m^/da, for an adult member of the
general public. This average volume results in 1.67E-01 joules/yr of
inhaled potential alpha energy from a continuous exposure to 1 WL of
radon-222 progeny for 365.25 da. Although it may be technically
inappropriate to quantify the amount of potential alpha particle energy
inhaled by a member of the general population in WLM, this corresponds to
about the same inhaled potential alpha energy as a 27 WLM exposure would
to a miner. Hence, a one WL concentration of radon progeny provides an
adult a 27 WLM annual exposure equivalent (see Table 7-7). For indoor
exposure, we assume an occupancy factor of 0.75, so that an indoor
exposure to 1 WL results in an annual exposure equivalent to 20 WLM,
(EPA78) in terms of the amount of potential alpha energy actually inhaled.
Children have a smaller bronchial area than adults, which more than
offsets their lower minute volume, so that the bronchial deposition and
expected dose, for a given concentration of radon progeny, is greater.
This problem has been addressed by Hofmann and Steinhausler (Ho77).
Their results indicate that doses received during childhood are about
50 percent greater than adult doses for a given air concentration of
radon daughters. We have used the information in (Ho77) to prepare
Table 7-7, which lists the age-dependent exposure equivalents we have
used in the risk assessments described below. (The assumptions on minute
volume, etc., for miners and the general population described above are
the same as those used in the preparation of EPA79,82,83a,b.) The
results in Table 7-7 have been rounded to two .significant figures. The
larger effective exposure to children relative to adults increases the
estimated mortality due to lifetime exposure from birth by about
20 percent.
7.4.2 The EPA Model
Since 1978, the Agency has based risk estimates due to inhaled
radon-222 progeny on a linear dose response function, a relative risk
projection model, and a minimum induction period of 10 years. The life
7-29
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table analysis described in Appendix E is used to project this risk over
a full life span. Lifetime risks were initially projected on the
assumption that an effective exposure of 1 WLM increases the age-specific
risk of lung cancer' by 3 percent over the age-specific rate in the U.S.
population as a whole (EPA79).
The initial EPA model for calculating radon risks has been described
in detail (EPA79, E179). In reviewing this model in terms of the more
recent information described below, we have found that our major
assumptions, linear response and relative risk projection, have been
affirmed. Data on the A-bomb survivors clearly indicate that for low-LET
1 radiation their absolute risk of radiogenic lung cancer has continued to
increase while their relative risk has remained reasonably constant
(Ka82). The UNSCEAR, ICRP, and 1980 NAS Committee have continued to use
a linear dose response to estimate the risk of lung cancer due to inhaled
radon progeny. Thomas and McNeill's analysis (Th82) indicates that the
use of linearity is not unduly conservative and may, in fact,
underestimate the risk at low doses. As noted above, the 1980 NAS BEIR
Committee reached a similar conclusion.
A major limitation of the EPA model is the uncertainty in the choice
of relative risk coefficient, the percent increase per WLM. This value
is based on the excess mortality due to lung cancer among exposed miners
of various ages, many of whom smoked. Therefore, it is an average value
for a mixed population of smokers and nonsmokers. Smoking was more
prevalent among some of the groups of miners studied than it is among the
U.S. general population today; this may inflate the risk estimate, as
discussed below.
Radford and Renard (Ra84) reported on the results of a long-term
study of Swedish iron miners who were exposed to radon progeny. This
study is unique in that most of the miners were exposed to less than
100 WLM, and the risks to smokers and nonsmokers were considered
separately. The absolute risk of the two groups was similar, 20
fatalities per 10^ person WLM year for smokers compared to 16 for
nonsmokers. While absolute risks were comparable for the smoking and
nonsmoking miners, relative risks were not. Nonsmokers have a much lower
baseline incidence of lung cancer mortality than smokers. As a result,
the relative risk coefficient for nonsmoking miners was about 4 times
larger than for smoking miners. In each case, the risk' was calculated
relative to baseline rates in nonsmokers and smokers, respectively.
Although occupational exposures to pollutants other than radon-222
progeny are probably not important factors in the observed lung cancer
risk for underground miners (E179, Th82, Mu83, Ra84), the use of
occupational risk data to estimate the risk of a general population is
far from optimal, as it: provides no information on the effect of radon
progeny exposures to"children and women. While we have continued to
assume that the risk per unit dose during childhood is no more effective
than that occurring to adults, this assumption may not be correct. The
A-bomb survivor data indicate that, in general, the risk from childhood
7-30
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Table 7-7. Annual exposure equivalent (WLM) by age for members of
the general public continuously exposed to radon
progeny at 1 WL (2.08 x 10~5 joules per cubic meter)
Age
(yr)
Exposure equivalent
(WLM)
0-2
3-5
6-11
12-15
16-19
20-22
23 or more
Lifetime Average
35
43
49
43
38
32
27
31.4
7-31
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exposure to low-LET radiation is greater than from exposure of adults and
continues for at least 33 years, the time over which A-bomb survivors
have been observed (Ka82). There are not, as yet, specific data for lung
cancer (Ka82). Another limitation of the underground miner data is the
absence of women in the studied populations. The A-bomb survivor data
xndicate women are as sensitive as men to radiogenic lung cancer from
low-LET radiation even though, on1 the whole, they smoke less (Pr83).
These data are not conclusive, however.
7.4.3 Comparison o'f Risk Estimates
Several estimates of the risk due to radon progeny have been
published since the EPA model was developed. One of particular interest
was expounded by the BEIR Committee in NAS80. The BEIR-3 Committee
formulated an age-dependent absolute risk model with increasing risk for
older age groups. The Committee estimates of the risk per WLM for
various ages are listed on page 325 in NAS80 and its estimated minimum
induction period for lung cancer following exposure on page 327. We have
used these data, summarized in Table 7-8, to calculate the lifetime risk
of lung cancer mortality from lifetime exposure to persons in the general
population by means of the same life table analysis used to calculate
other EPA risk estimates.
It should be noted that the zero risk shown in Table 7-8 for those
under 35 years of age at exposure does not mean no harm occurs,, but . ••'
rather that it is not expressed until the person is at least 35 years
old, i.e., only after the minimum induction period. The sequence of
increasing risk with age shown in Table 7-8 is not unlike the increase in
lung cancer with age observed in unexposed populations, so that the
pattern of excess risk over time is similar to that found using a
relative risk projection model.
Recently, Thomas and McNeil conducted .a thorough analysis of lung
cancer among uranium and other hard rock miners for the AECB of Canada
(Th82). These investigators tested a number of risk models on all of the
epidemiological studies that contained enough data to define a
dose-response function. They concluded that for males a 2.3 percent
increase in lung cancer per WLM and a relative risk projection model were
more consistent with the excess lung cancer incidence observed in
underground miner groups than other models they tested. This is the only
analysis we are aware of that treated each data set in a consistent
fashion and utilized modern epidemiological techniques, such as
controlling, to the extent possible, for age at exposure and duration of
follow-up.
The initial EPA risk estimates for lifetime exposure to a general
population, along with AECB, NAS, UNSCEAR, ICRP, and NCRP estimates of
the risk of lung cancer resulting from inhaled radon progeny, are listed
in Table 7-9. The AECB estimate for lifetime exposure to Canadian males
is 830 fatalities per million person-WLM (Th82). In Table 7-9 this
estimate has been adjusted for the U.S. 1970 male and female population.
7-32
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There is good agreement between the EPA, NAS80 (BEIR-3), and the
AECB estimates shown in Table 7-9. Each of these estimates is based on
lifetime exposure and lifetime expression of the incurred risk. In
contrast, the ICRP and UNSCEAR risk estimates in Table 7-9 do not
explicitly include these factors.
The ICRP estimates are for occupational exposure to working adults.
The larger ICRP estimate is based on their epidemiological approach, that
is, the exposure to miners in WLM and the risk per WLM observed in
epidemiological studies of underground miners. The ICRP epidemiological
approach assumes an average expression period of 30 years for lung
cancer. Children, who have a much longer average expression period, are
excluded from this estimate. The ICRP has not explicitly projected the
risk to miners beyond the years of observation even though most of the
miners on whom its estimates are based are still alive and continuing to
die of lung cancer.
The smaller of the two ICRP estimates listed in Table 7-9 is based
on this dosimetric approach. The ICRP assumes that the risk per'rad for
lung tissue is 0.12 of the risk of cancer and genetic damage following
whole-body exposure (ICRP77). For the case of exposure to radon^progeny,
the ICRP divided this factor of 0.12 into two equal parts. A weighting
factor of 0.06 was. used to assess the risk from the high dose to
bronchial tissue, where radiogenic lung cancer is observed in exposed
underground miners. The other half of the lung weighting factor, another
0.06 of the total body risk, was used to assess the risk to the pulmonary
region which receives a comparatively small dose from radon-222 progeny
and where human lung cancer is seldom, if ever, observed.
The UNSCEAR estimate is for a general population and assumes an
expression time of 40 years. Like the ICRP, UNSCEAR did not make use of
an explicit projection of risk of fatal lung cancer over a full lifetime.
The last entry in Table 7-9, the NCRP risk estimate based on an
analysis by Harley and Pasternack (Ha82), is of particular interest
since, like those of EPA and AECB, it is based on a life table analysis
of the lifetime risk due to lifetime exposure. This estimate utilizes an
absolute risk projection model with a relatively low risk coefficient, 10
cases per 106 person-WLM per year at risk, the smallest of those listed
by the NAS BEIR-3 Committee: cf. Table 7-8. Moreover, they have assumed
that the risk of lung cancer following irradiation decreases
exponentially with a 20-year half-life, so that exposures occurring early
in life have very little risk. The NCRP assumption of a 20-year
half-life for radiation injury reduces the estimated lifetime risk by
about a factor of 2.5. Without this assumption the NCRP risk estimate
would be the same as the midpoint of the UNSCEAR estimate, about 325
fatalities per million person-WLM. Note that if lung cancer risk from
low-LET radiation decreased over time with a 20-year half-life, the
excess lung cancer observed in Japanese A-bomb survivors should have
decreased during the period they have been followed, 1950-1978. During
this period the absolute lung cancer risk in every age cohort has
markedly increased (Ka82).
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Table 7-8. Age-dependent risk coefficients and minimum induction
period for lung cancer due to inhaling Radon-222
progeny (NAS80)
Age
(yr)
0-14
15-34
35-49
50-65
65 or more
Excess
(cases per 10^
WLM person-years)
0
0
10
20
50
Minimum induction period
(years)
25
15-20
10
10
10
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Table 7-9. Risk estimate for exposures to radon progeny
Organization
E.PAa
NAS BEIR-3a
AECBC
ICRP
UNSCEAR
NCRPd
Fatalities per
106 person-WLM
760 (460)b
730 (440)b
600 (300)b
150-450
200-450
130
Exposure period
Lifetime
Lifetime
Lifetime
Working Lifetime
Lifetime
Lifetime
Expression
period
Lifetime
Lifetime
Lifetime
30 years
40 years
Lifetime
a The number of fatalities per 106 person-WLM listed for EPA and
NAS80 in this table differs from figures we have previously published
(e.g., EPA83b) because we have now included, correctly we believe, the
increased potential alpha energy exposure during childhood in the
denominator of this ratio. Our risk estimates for various sources of
radon in the environment have not changed, because all were calculated
via a life table analysis yielding deaths per 100,000 exposed, not
deaths per 106 person-WLM.
b EPA and AECB based their estimates of risk for the general population
on an exposure equivalent, corrected for breathing rate (and other
factors). For comparison purposes, the values in parentheses express
the risk in more customary units, in which a continuous annual exposure
.to 1 WL corresponds to 51.6 WLM.
c Adjusted for U.S. General Population, see text.
d NCRP84: Table 10.2; assumes risk, diminishes exponentially with a
20-year halftime.
Sources: EPA83b; NAS80; Th82; ICRP81;
UNSCEAR77; NCRP84; USRPC80.
7-35
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Good agreement exists among the EPA, NAS (BEIR-3), and the AECB
estimates listed in Table 7-9. Each of these estimates is based on
lifetime exposure and lifetime expression of the incurred risk.
Conversely, the three lower risk estimates shown in Table 7-9 either do
not explicitly include these conditions or they include other modifying
factors. Nevertheless, Table 7-9 indicates a divergence, by a factor of
about 6, in risk estimates for exposure to radon-222 progeny. Thus, the
use of a single risk coefficient may not be appropriate, as it could give
the impression that the risk is known more precisely than is warranted by
available information. The EPA, BEIR-3, and AECB estimates may be
slightly high because they represent relative risks based on adult males,
many of whom smoked. The actual risk may be smaller for a population
that includes adult females, children, and nonsmokers. The UNSCEAR and
ICRP estimates are probably low because they represent absolute risk
estimates that do not completely take into account the duration of the
exposure and/or the duration of the risk during a lifetime. The NCRP
estimate is likely to be very low, as a low risk coefficient was used in
an absolute risk model, and it was assumed that the risk decreases
exponentially after the exposure.
7.4.4 Selection of Risk Coefficients
To estimate the range of reasonable risks from exposure to radon-222
progeny for use in the Background Information Document for Underground
Uranium Mines (EPA85), EPA averaged the estimates of BEIR-3, the EPA
model, and the AECB to establish an upper bound of the range. The lower
bound of the range was established by averaging the UNSCEAR and ICRP
estimates. The Agency chose not to include the NCRP estimate in its
determination of the lower bound because this estimate is believed to be
outside the lower bound. Therefore, the EPA chose relative risk
coefficients of 1.2 to 2.8 percent per WLM exposure equivalent (300 to
700 fatalities per million person-WLM exposure equivalent) as estimates
of the possible range of effects from inhaling radon-222 progeny for a
full lifetime. Although these risk estimates do not encompass the full
range of uncertainty, they seemed to illustrate the breadth of much of
current scientific opinion.
The lower limit of the range of 1985 EPA relative risk coefficients,
1.2 percent per effective WLM, is similar to that derived by the Ad Hoc
Working Group to Develop Radioepidemiological Tables, which also used 1.2
percent per WLM (NIH85). However, some other estimates based only on
U.S. and Czech miner data averaged 1 percent per WLM (Ja85) or 1.1
percent per WLM (St85). On the other hand, three studies, two on miners
(Ra84, Ho86) and one on residential exposure (Ed83, 84), indicate a
relative risk coefficient greater than 3 percent per WLM, perhaps as
large as 3.6 percent.
The EPA has, therefore, increased the upper limit of its estimated
range of relative risk coefficients. To estimate the risk due to
radon-222 progeny, the EPA now uses the range of relative risk
coefficients of 1 to 4 percent per WLM. [See EPA86 for a more detailed
7-36
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discussion.] Based on L980 vital statistics, this yields, for members of
the general public, a range of lifetime risks from 380 to 1,520 fatal
cases per 106 WLM (expressed in exposure equivalents). In standard
exposure units, uncorrected for breathing rate and age, this corresponds
to 230 to 920 cases per 106 WLM. Coincidentally, the geometric mean
estimate obtained in this way, 4.6E-04/WLM in .standard units of exposure,
is numerically the same as that obtained using a 3 percent relative risk
coefficient and 1970 vital statistics (see Table 7-9).
7.5 Uncertainties in Risk Estimates for Radiogenic Cancer
As pointed out in the Introduction of this chapter, numerical
estimates of risks due to radiation are not precise. A numerical^
evaluation of radiogenic cancer risks depends, both on epidemiological
observations and on a number of ad hoc assumptions which are largely .
external to the observed data. These assumptions include such factors as
the expected duration of risk expression and variations in
radiosensitivity as a function of age and demographic characteristics. A'
major assumption is the shape and slope of the dose response curve,
particularly at low doses, i.e., below 1 rad, where there is insufficient
epidemiological data to directly base risk estimates. In 1971, the BEIR
Committee based its estimates of cancer risk on the assumption that
effects at low doses are directly proportional to those observed at high
doses, the so called linear-nonthreshold hypothesis. As described above
in Section 7.2, the BEIR-3 Committee considered three dose response
models and indicated a preference for the linear-quadratic model. The
risk coefficients the BEIR-3 Committee derived for its linear-quadratic
model, and to a lesser extent its linear model, are subject to
•considerable uncertainty primarily because of two factors:
(1) systematic errors in the estimated doses of the individual A-bomb
survivors, and (2) statistical uncertainty because of the small number of
cancers observed at various dose levels.
7.5.1 The BEIR-3 Analysis of .the A-bomb Survivor Data
For its analysis of the A-bomb survivor data, the BEIR-3 Committee
expanded the equations for low-LET radiations to include a linear dose
response function for neutrons:
F(Dg, DQ> =
F(Dg, Dn) -
FCD Dn);-
g
n
(7-1)
(7-2)
(7-3)
where Dg is the gamma dose and Dn is that part of the dose due to
high-LET radiations from neutron interactions. Note that Equation 7-1
7-37
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and Equation 7-3 each have two linear terms: one for neutrons and one
for gamma radiation. In analyzing approximately linear data in terms of
these equations, the decision as to how much of the observed linearity
should be assigned to the neutron or the gamma component is crucial. As
discussed below, the BEIR-3 Committee attributed much of the observed
radiogenic cancer to a linear response from neutron doses that did not
occur.
The BEIR-3 Committee's general plan was to examine the dose response
for leukemia and for solid cancer separately to find statistically valid
estimates of the coefficients G! C4 and K! K3 by means of
regression analyses. The T65 neutron and gamma doses to individual
survivors are highly correlated since both are strongly decreasing
functions of distance. This makes accurate determination of the
coefficients in Equation 7-3 by means of a regression analysis extremely
difficult. In addition, there is considerable sampling variation in the
A-bomb survivor data due to small sample size, which exacerbates the
regression problem (He83). Because of these and other problems,
agreement between the observed response and that predicted by any of the
dose response functions examined by the BEIR-3 Committee provides little
basis for a choice between models.
The Committee analyzed the A-bomb survivor data in two separate
sets: first, leukemia; second, all cancers excluding leukemia (solid
cancers).^ Its treatment of these two cases was not equivalent. The
Committee's analysis of leukemia considered the Nagasaki and Hiroshima
data separately. The Committee's regression analysis of the leukemia
mortality data provided stable values for all of the coefficients in
Equation 7-3, and hence for the neutron RBE and the ratio of linear to
dose-squared terms for leukemia induction by gamma rays, as a function of
dose.
Estimating the linear-quadratic response coefficients for solid
cancers proved to be less straightforward, however, and it was
7-38
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decided that the observations on solid cancers were "not strong enough to
provide stable estimates of low 'dose, Ibw-LET cancer risk when analyzed
in this fashion" (NAS80,p. 186).
As outlined in the BEIR-3 Report, the Committee decided to use a
constrained regression analysis, carrying over some of the parameters for
Equation 7-3 found in its analysis of leukemia deaths to the regression
analysis of the dose response for solid cancers. Specifically, both the
neutron RBE at low dose (the ratio of the coefficient K3 to 63) and
the ratio of GS to €4, as estimated from the leukemia data, were
assumed to apply to the induction of fatal solid cancers. These
estimates became the basis for the "preferred" linear-quadratic
(LQ-L) risk estimates for solid cancers presented in BEIR-3 (NAS80):p. 187,
7.5.2 Uncertainty of the Dose Response Models
Due to Bias in the A-bomb Dosimetry
A careful state-of-the-art evaluation of the dose to A-bomb
survivors was carried out by investigators from Oak Ridge National
Laboratory in the early 1960's (Au67, Au77). The results of these
studies resulted in a "T65" dose being assigned to the dose (kerma) in
free air at the location of each survivor for both gamma rays and
neutrons. A major conclusion of the ORNL study was that the mix of gamma
ray and neutron radiations was quite different in the two cities where
A-bombing occurred. These results indicated that at Hiroshima the
neutron dose was more important than the gamma dose when the greater
biological efficiency of the high-LET radiations produced by neutrons was
taken into account. Conversely, the neutron dose at Nagasaki was shown
to be negligible compared to the gamma dose for that range of doses where
there were significant numbers of survivors. Therefore, the 1980 BEIR
Committee evaluated the cancer risks to the survivors at Hiroshima on the
assumption that the combined effects of gamma rays and particularly
neutrons caused the observed cancer response.
Since the BEIR-3 report was published, it has become evident that
the organ doses due to neutrons at Hiroshima were overestimated by about
an order of magnitude, at distances where most of the irradiated persons
survived bomb blast and yet received significant doses (1,000-1,500 m).
7-39
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In fact, the neutron doses at Hiroshima are quite comparable to those
previously assigned, at similar distances, to Nagasaki survivors (KeSla,
KeSlb, RERF83, RERF84). Moreover, there are now grounds to believe the',
T65 estimates of gamma-ray' doses in both cities are also incorrect
(RERF83, RERF84).
At the time of this writing, a major effort is underway to reassess
the dosimetry in both cities (RERF83, RERF84). Preliminary indications
are that gamma-ray doses in air will decrease in Nagasaki, but only
slightly. In Hiroshima there may be substantial increases in the
gamma-ray kerma beyond about 1500 m, but only small increases closer to
the^hypocenter, where most of the collective dose was received. In
addition, recalculation of shielding from structures and body tissue is
expected to decrease the average gamma-ray organ doses somewhat. The net
effect of these changes in gamma-ray doses is still unclear, but they are
unlikely to result in more than a 50 percent change in risk estimates for
gamma irradiation. More important, it seems, is the anticipated effect
of revised estimates of the neutron dose to the Hiroshima survivors.
Given the information discussed above, it is possible to see at
least qualitatively, how the large bias in the estimated T65 neutron dose
to the Japanese survivors affects the 1980 BEIR Committee's estimates of
the risk coefficients for leukemia. The Committee's age-adjusted risk
coefficients for leukemia are listed in Table V-8. For the linear fit
the neutron RBE (K^/CX) was 11.3, while for the linear-quadratic case '
the neutron RBE (K3/C3) was 27.8. Tables A-ll and V-13 provide the
estimates of neutron and gamma doses to the bone marrow of Hiroshima
survivors that were used by the,Committee. Substituting these doses in
its risk equations (Table V-8) indicates that, for either model, almost
50 percent of the total leukemia deaths were ascribed to the neutron dose
component then thought to be present at Hiroshima. (At first sight, it
might appear that a substantially larger fraction of leukemias would be
attributed to neutrons in the LQ-L model, in view of its higher neutron
RBE. However, when one takes into account the D| term in Equation 7-3,
it turns out that the contribution of gamma rays is about the same as in
the L-L model.)
^In a similar way, the conversion factors given in Table V-13 for
obtaining tissue dose from kerma can be used to derive the fraction of
all solid tumors attributed to neutrons. Here again almost half the
cancers were attributed to the neutron component. Since, as noted above,
the neutron dose was overestimated, almost all of the excess leukemias
and solid tumors will probably now have to be attributed to gamma rays.
There is no simple way of adjusting the 1980 BEIR risk estimates to
account for the risk they attributed to neutrons. Adjustment of neutron
doses alone is clearly inappropriate, since there is good reason to
believe that T65 estimates of the dose due to gamma rays are also subject
to considerable change. Moreover, not all of the individuals in a given
T65 dose category will, necessarily, remain grouped together after new
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estimates of neutron and gamma doses are obtained. Both the numerator
and denominator in the ratio of observed to expected cases are subject to
change and indeed could change in opposite directions, a fact not
considered in some preliminary analyses (St81). Nevertheless, it. is
reasonable to conclude that bias in the estimated neutron doses at
Hiroshima has led to considerable uncertainty in the BEIR-3 risk
estimates and probably to a systematic underestimation of the risk due to
low-LET radiations. In addition, random errors in dose estimates will
contribute to the uncertainty in risk coefficients. As discussed by
Gilbert (Gi84), these errors also tend to bias risk estimates downward.
In light of these biases arising from errors in dosimetry, we believe
that estimates based on the more conservative linear dose response should
be given considerable weight vis a vis those made using the BEIR-3
linear-quadratic models.
In conclusion, the overall effect of the revised dosimetric
calculations will probably be' to increase the estimated risk per unit
dose of low-LET radiation in the A-bomb survivor population. The
magnitude of the increase is unknown, but will probably not be more than
about a factor of 2.
From the standpoint of estimating risks from low-level, low-LET
radiation, however, the most important result of the new dosimetric
calculations may be in helping to determine which models best describe
the data on human radiation carcinogenesis. After all, the greatest
uncertainties in radiation risk estimation generally reflect model
uncertainties, not uncertainties in the magnitude of risk coefficients.
7.5.3 Sampling Variation
Besides the systematic bias in the BEIR-3 risk estimates for low-LET
radiation outlined above, the precision of the estimated linear and
linear-quadratic risk coefficients in the BEIR-3 report is limited by
statistical fluctuations due to sample size. The uncertainty bounds ( + 1
SD) attached to the gamma-ray risk coefficient in the BEIR-3 linear model
are about +25 percent, for either leukemia (Table V-8) or for all other
cancers (Table V-ll). For the latter groups of cancers, however, the
neutron RBE was constrained to the value obtained from analysis of the
leukemia data. If this constraint is removed, the uncertainty in the
estimate increases to +150 percent (Table V~9). This increase reflects
the large uncertainty associated with the neutron contribution in the
analysis and the strong correlation between neutron and gamma-ray doses.
Following the dosimetry reassessment, neutron doses will decrease
markedly, but will remain correlated with gamma-ray doses. It is alio
likely that the estimated risk per unit dose will still turn out to be
significantly different between Hiroshima and Nagasaki. Attributing this
difference to the much smaller neutron fluxes may imply a biologically
implausible value for the neutron RBE (<0 or >100, for example). -
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^An alternative approach would then be to impose a constraint: in
particular, it might be assumed that the neutron contribution to the
excess cancer is negligible, the apparent difference between the two
cities being due to some residual systematic errors in dosimetry, or to-
other unknown causes. If the fit is constrained in this way, the
standard deviation in the linear coefficient obtained from the combined
data may be reduced back down to approximately that.obtained from the
constrained analysis based on the T65 dosimetry, i.e., to about
+25 percent; there could, however, be a residual uncertainty relating to
any unexplained differences between the two cities.
With the linear-quadratic model, there is the additional uncertainty
over the relative magnitudes of the linear and quadratic coefficients.
If the quadratic term is constrained to be non-negative, then the linear
model estimate provides an upper bound on the magnitude of the risk at
low doses. On the other hand, a pure quadratic model (linear coefficient
equal to zero) based on the T65 dosimetry is consistent with the A-bomb
survivor data on leukemia as well as on solid tumors.
7.5.4 Low Dose Extrapolation
As discussed above, the A-bomb'survivor data on leukemia and all
solid tumors, when analyzed in terms of the linear-quadratic model, are
consistent with a very small, possibly zero, linear coefficient and thus .
a risk at low doses/dose rates, which is much smaller than .what would be
predicted from the linear model. A reasonable lower bound on the risk
coefficient at low doses and dose rates can, however, be derived from
other considerations. ,
Results from animal and cellular studies often show decreasing
effects (e.g., cancers, mutations, or transformations) per rad of low-LET
radiation at low doses and dose rates. Based on a review of .this
literature, the National Council on Radiation Protection (NCRP80) has
concluded that "linear interpolation from high doses (150 to 300 rads)
and dose rates (>5 rads min l) may overestimate the effects of either
low doses (0-20 rads or less) or of any dose delivered at dose rates of
5 rad y~J- or less by a factor of 2 to 10." Judged solely from
laboratory experiments, therefore, about a factor of 10 reduction from
the linear prediction would seem to constitute a plausible lower limit on
the effectiveness of low-LET radiation under chronic low dose
conditions. Epideraiological evidence, however, would seem to argue
against such a large DREF for human cancer induction.
Data on the A-bomb survivors and patients irradiated for medical
reasons indicate that excess breast cancer incidence is proportional to
dose and independent of dose fractionation (NAS80, NIH85). The evidence
regarding thyroid cancer induction is less firm, but the data would again
suggest a linear dependence on dose (NAS80, NIH85). The only other
cancer for which there are human data "good enough" to provide any test
of dose response models is leukemia. An analysis of the A-bomb survivor
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data based on T65 dosimetry suggests a quadratic component; however,, the
best estimate of the linear coefficient obtained from the linear
quadratic fit to the data is only about a factor of 2.5 less than the
coefficient derived from the linear model.
A lower bound estimate of risk might be constructed by assuming that
a linear dose-response function holds for breast cancer induction, but
that for low dose rates the pure linear model overpredicts other cancers
by a factor of 10 (DREF = 10). Using a linear model for all cancers, it
was estimated (see Table 7-3) that about 17 percent of all fatal cancers
resulting from a uniform whole-body dose to the general population are
breast cancers. Thus, under the assumption above, the lower bound
estimate is 23 percent of the linear estimate (1 x 14 percent +
0.1 x 86 percent). This would still seem to be an extreme lower bound
estimate of the risk, especially in light of the evidence on thyroid
cancer and leukemia referred to above. We believe a reasonable lower
bound on the effectiveness of low-LET radiation in causing fatal cancers
at low doses and dose rates is about 30 percent of that computed by
linear extrapolation from high acute doses (equivalent to DREF=3.3).
7.5.5 Other Uncertainties Arising from Model Selection
In addition to a dose response model, a "transportation model" is
needed to apply the risks from an observed irradiated group to another
population having different demographic characteristics. A typical
example is the application of the Japanese data for A-bomb survivors to
Western people. Seymour Jablon (Director of the Medical Follow-up Agency
of the National Research Council, NAS) has called this the
"transportation problem," a helpful designation because it is often
confused with the risk projection problem described below. However,
there is more than a geographic aspect to the "transportation problem."
Risk estimates for one sex must sometimes be based on data for the
other. In transporting risk estimates from one group to another, one may
have to consider habits influencing health status, such as differences
between smokers and nonsmokers, as described in Section 7.4 for the case
of risk estimates for radon progeny.
The BEIR-3 Committee addressed this problem in its 1980 report and
concluded, based largely on the breast cancer evidence, that the
appropriate way to transport the Japanese risk to the U.S. population was
to assume that the absolute risk over a given observation period was
transferrable but that relative risk was not. Therefore, the Committee
calculated what the relative risk would be if the same number of excess
cancer deaths were observed in a U;S. population having the same age
characteristics as the A-bomb survivors. A constant absolute risk model,
as postulated by the Committee, would imply that, whatever the factors
are which cause Japanese and U.S. baseline cancer rates to differ, they
have no effect on the incidence of radiation-induced cancers; i.e., the
effects of radiation and these factors are purely additive.
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An alternative approach to solving the "transportation problem" is
that of the 1972 NAS BEIR-1 Committee. This Committee assumed relative
risks would be the same in the United States and Japan and transferred
the observed percentage increase directly to the U.S. population. Since
the U.S. and Japanese baseline rates differ drastically with respect to
mortality from specific cancers, this approach implies some large
differences in the predicted number of specific cancers resulting from a
given dose of radiation in the two countries. The most important
differences relate to cancers of the breast, lung, and stomach. Baseline
rates of breast and lung cancers are higher in the U.S. by factors of
about 4 and 2, respectively, while the risk of stomach cancer is about
8 times higher in Japan (Gi85). As noted above, it now appears that the
absolute risk should be transported for breast cancer. Evidence is
lacking regarding the other diseases, however. If lung cancer risk were
to be transported with a relative risk model, retaining the absolute
model for other cancers, the estimated risk from a whole-body exposure
would increase by about 20 percent; on the other hand, applying the
relative risk model to stomach cancer alone would lower the whole-body
risk by about 8 percent. Based on these considerations, including the
tendency for changes in specific cancers to cancel one another, we
believe that using the absolute risk "transportation model" is unlikely
to cause errors of more than +20 percent in the total risk estimate.
Thus, in the case of uniform whole-body doses, the amount of uncertainty
introduced by transporting cancer risks observed in Japan to the U.S.
population appears to be small compared to other sources of uncertainty
in this risk assessment.
The last of the models needed to estimate risk is a risk projection
model. As outlined in Section 7.2, such models are used to project what
future risks will be as an exposed population ages. For leukemia and
bone cancer, where the expression time is not for a full lifetime but
rather 25 years, absolute and relative risk projection models yield the
same number of radiogenic cancers, but would distribute them somewhat
differently by time after exposure, and hence by age. For solid cancers,
other than bone, the BEIR-3 Committee assumed that radiogenic cancers
would occur throughout the estimated lifetime. This makes the choice of
projection model more critical, because the relative risk projection
yields estimated risks about three times larger than those obtained with
an absolute risk projection, as shown in Table 7-2. Recent follow-up of
the^A-bomb survivor population strongly suggests that the relative risk
projection model better describes the variation in risk of solid tumors
over time (NIH85). However, there may be some cancers, apart from
leukemia and bone cancers, for which the absolute risk projection model
is a better approximation to reality. For other cancers, the relative
risk may have been roughly constant for the current period of follow-up,
but may eventually decrease over time. Thus, while the relative risk
model^was used in this report for calculating a "best estimate" of the
lifetime risk of solid tumors, it may overestimate the risk by as much as
a factor of 2.
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Similarly, there is as yet insufficient information on
radiosensitivity as a function of the age at exposure. The age-dependent
risk coefficients we have used are those presented in the BEIR-3 report.
As yet, there is little information on the ultimate effects of exposure
during childhood. As the "A-bomb survivors' population ages, more
information will become available on the cancer mortality of persons
irradiated when they were young. Table 7-2 indicates that the more
conservative BEIR-1 assumption for the effect of childhood exposures
would increase BEIR-3 risk estimates by about 40 percent. This is
probably an upper bound. A lower bound can be estimated by assuming that
the relative risk coefficient for those irradiated between ages 0-19 is
actually only as large as that calculated for the next higher age
category (20-34). This assumption leads to about a 20 percent decrease
in the lifetime risk as compared to the BEIR-3 calculation. Therefore,
the lack of precise information concerning the dependence of risk on age
at exposure does not appear to be a major source of uncertainty in
estimates of risk caused by either lifetime exposure or by a single
exposure to the general population. Similarly, the BEIR-3 Committee did
not include rn utero exposures when calculating population risks for
radiogenic cancer because they felt the estimate of the effect of
in utero radiation is uncertain. We have deferred to their judgment in
this regard. The BEIR-1 report did include in utero cancer risk. These
had little effect, 1 to 10 percent, on the lifetime risk of cancer from
lifetime exposure. An effect this small is not significant relative to
other sources of uncertainty in the risk assessment.
7.5.6 Summary
We can only semi-quantitatively estimate the overall uncertainty in
the risk per rad for low-LET radiations. We expect that more
quantitative estimates of the uncertainty will be possible only after the
A-bomb dose reassessment is completed and the A-bomb survivor, data are
reanalyzed on the basis of the new dose estimates. It should be noted,
however, that even if all systematic bias is removed from the new dose
estimates, there will still be considerable random error in the -dose
estimate for each survivor. This random error biases the estimated slope
of the dose response curve so that it is smaller than the true dose
response (Da75, Gi84, Ma59). The amount of bias introduced^depends on
the size of the random errors in the dose estimates and their
distribution, which are unknown quantities at this stage of the dose
reassessment.
Table 7-10 summarizes the various sources of uncertainty, as
discussed above. The numerical entries represent multiplicative factors
by which our estimates might have to be adjusted due to each source. To
fully assess the magnitude of the combined uncertainty from all these
sources, one must first characterize the underlying distribution of
uncertainty relating to each source. This is beyond the scope of this
report. However, a rough estimate of the overall uncertainty can be
derived employing the general approach outlined in the Report of the
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Table 7-10. Uncertainties in fatal cancer risk estimates
Source of uncertainty
Factor for limit
lower
upper
Use of linear model to extrapolate from
acute high dose to chronic low dose
exposures 0.3 1.0
Slope of dose response resulting from
sampling variation 0.5b (.6)° 1.5b(1.6)c
Use of T65 dosimetry 1.2 2.0
Use of lifetime relative risk
projection model 0.5 1.0
Use of absolute risk "transportation
model"3 0.8 1.2
Influence of age at exposure 0.8 1.4
Overall uncertaintyd 0.23 1.6
a For the total of all cancers resulting from a uniform whole-body
exposure to low-LET radiation; uncertainties relating to specific
cancers may be considerably larger.
Estimated 95% confidence limits based on a normal distribution.
c Estimated 95% confidence limits based on a lognormal distribution
having the same mean and variance as the normal distribution defined
by b.
° Combined uncertainty estimate from sources listed above; 95%
confidence interval assuming the uncertainties from each source
are independent and lognormally distributed (see text).
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Ad Hoc Working Group to Develop Radioepictemioiogical Tables (NIH85) . [It
might be noted that the sources of uncertainty listed here differ
somewhat from those in the NIH Report. To a large extent this reflects a
difference in focus: here, it is on estimating the total number of
cancers from whole-body exposures to a population; there, it is on
estimating the probability that a particular cancer was caused by a given
exposure. In addition, we have tried to incorporate sources of
uncertainty (viz., those relating to sampling variation and choice of
transportation model) not included by the Working Group in its
calculation of combined uncertainty.]
As in the NIH Report, the uncertainties due to each source are
assumed to be independent and lognormally distributed, the geometric mean
of the distribution being set equal to the geometric mean of the upper
and lower bounds. The respective upper and lower bounds are further
assumed to be commensurate with one another; in particular, all are taken
to be 95 percent confidence interval limits. The combined standard
deviation or confidence interval can then be readily calculated.
Denoting the geometric standard deviation of each source i
(i = 1, 2, ...k) by S^ , the geometric standard deviation of the
combined distribution is given by
ln2S =
. ..ln2Sk.
This procedure is highly arbitrary since there is generally no
objective information on the actual underlying distribution of
uncertainty for each source, the lognormal distribution being adopted, in
large part, for calculational ease. Even the choice of upper and lower
bounds involves a largely subjective judgment in most cases. One partial
exception is the uncertainty relating to sampling variation (second entry
in Table 7-10). This uncertainty is directly derived from a linear
regression analysis of the A-bomb survivor data, the upper and lower
bounds reflecting +2 standard deviations about the best estimate of the
risk coefficient. Given the properties of the data, these bounds should
indeed represent approximate 95 percent confidence limits; however, the
underlying distribution of uncertainty is expected to be normal rather
than lognormal. To better reflect this fact, while retaining the
calculational simplicity of the lognormal assumption, a lognormal
distribution having the same arithmetic mean and standard deviation as
the normal distribution of uncertainty was constructed for this source
and used for the purpose of computing the combined uncertainty. As seen
in Table 7-10, the lognormal construct had upper and lower bounds shifted
upward slightly with respect to the normal distribution, the geometric
mean of the former (0.97) falling very close to the arithmetic raean/of
the latter (1.0).
Scaled to our estimate of the average risk from whole-body low-LET
radiation, the upper and lower confidence limits, calculated as described
above, span the range from 0.23 to 1.6. This corresponds to a range of
91 to 630 fatal cancers/106 person-rad. The geometric mean of the
range is about 240 fatal cancers/106 person-rad, suggesting that our
estimate of 395 fatal cancers /106 person-rad may be biased high
slightly. However^ our estimate falls well within the range of
uncertainty, and we believe it represents a prudent and reasonable
choice for the purposes of radiation protection.
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Finally, it should be noted that the analysis above pertains to
whole-body, low-LET radiation exposures. The uncertainties in risk to
specific organs may be considerably larger. This is particularly
important for internal emitters that concentrate in certain organs.
Often the dose estimates for these radionuclides are more uncertain as
well.
7.6 Other Radiation-Induced Health Effects
The earliest report of radiation-induced health effects was in 1896
(Mo67), and it dealt with acute effects in skin generally caused by very
large x-ray exposures. Within the six-year period following, 170
radiation-related skin damage cases had been reported. Such injury, like
many other acute effects, is the result of exposure to hundreds or
thousands of rads. Under normal situations, environmental exposure does
not cause such large doses, so possible acute effects will not need to be
considered in assessing the risk to the general population from
non-accidental radionuclide emissions. -
Radiation-induced carcinogenesis was the first delayed health effect
described: the first case was reported in 1902 (Vo02), and 94 cases of
skin cancer and 5 of leukemia were reported by 1911 (Up75).
Radiation-induced genetic changes were noted soon afterward. In 1927,
H.J. Muiier described x-ray-induced mutations in animals (in the insect,
Drosophila) and in 1928, L.J. Stadier reported a similar finding in
plants (Ki62). At about the same time, radiation effects on the
developing embryo were observed. Case reports in 1929 showed a high rate
of microcephaly (small head size) and central nervous system disturbance
and one case of skeletal defects in children irradiated jin utero
\.UNSCEAR69). These effects, at unrecorded but high exposures and at
generally unrecorded gestational ages, appeared to produce central
nervous system and eye defects similar to those reported in rats as early
as 1922 (Ru50).
For purposes of assessing the risks of environmental exposure to
radionuclide emissions, the genetic effects and in utero developmental
effects are the only health hazards other than cancer that are addressed
in this Background Information Document (BID).
7.6.1 Types of Genetic Harm and Duration of Expression
Genetic harm or the genetic effects of radiation exposure are those
effects induced in the germ cells (eggs or sperm) of exposed individuals,
which are transmitted to and expressed only in their progeny and future-
generations.
Of the possible consequences of radiation exposure, the genetic
risk is more subtle than the somatic risk, since it does not affect the
persons exposed, but relates only to subsequent progeny. Hence, the time
scales for expression of the risk are very different* Somatic effects
are expressed over a period on the order of a lifetime, while about 30
subsequent generations (about 1,000 yr) are needed for near complete
expression of genetic effects. Genetic risk is incurred by fertile
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people when radiation damages the nucleus of the cells which become their
eggs or sperm. The damage, in the form of a mutation or a chromosome
aberration, is transmitted to, and may be expressed in, a child conceived
after the radiation exposure or in subsequent generations. However, the
damage may be expressed only after many generations or, alternatively, it
may never be expressed because of failure to reproduce or failure of the
chance to reproduce.
EPA treats genetic risk as independent of somatic risk even though
somatic risk may be caused by mutations in somatic cells because, whereas
somatic risk is expressed in the person exposed, genetic risk is
expressed only in progeny and, in general, over many subsequent
generations. Moreover, the types of damage incurred often differ in kind
from cancer and cancer death. Historically, research on genetic effects
and development of risk estimates have proceeded independently of the
research on carcinogenesis. Neither the dose response models nor the
risk estimates of genetic harm are derived from data on studies of
carcinogenesis.
Although genetic effects may vary greatly in severity, the genetic
risks considered by the Agency evaluating the hazard of radiation
exposure include only those "disorders and traits that cause a serious
handicap at some time during lifetime" (NAS80). Genetic risk may result
from one of several types of damage that ionizing radiation can cause in
the DNA within gonidial cells or eggs and sperm. The types of damage
usually considered are: dominant and recessive mutations in autosomal
chromosomes, mutations in sex-linked (x-linked) chromosomes, chromosome
aberrations (physical rearrangement or removal of part of the genetic
message on the chromosome or abnormal numbers of chromosomes), and
irregularly inherited disorders (genetic conditions with complex causes,
constitutional and degenerative diseases, etc.).
Estimates of the genetic risk per generation are conventionally
based on a 30-yr reproductive generation. That is, the median parental
age for production of children is age 30 (one-half the children are
produced by persons less than age 30, the other half by persons over age
30). Thus, the radiation dose accumulated up to age 30 is used to
estimate the genetic risks. Using this accumulated dose and the number
of live births in the population along with the estimated .genetic risk
per unit dose, it is possible to estimate the total number of genetic
effects per year, those in the first generation and the total across all
time. Most genetic risk analyses have provided such data. EPA
assessment of risks of genetic effects includes both first generation
estimates and total genetic burden estimates.
(A) Direct and Indirect Methods of Obtaining
Risk Coefficients for Genetic Effects
Genetic effects, as noted above, may occur in the offspring of the
exposed individuals or they may be spread across all succeeding
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generations. Two methods have been used to estimate the frequency of
mutations in the offspring of exposed persons, direct and indirect. In
either case, the starting point is data from animal studies, not data
obtained from studies of human populations. This is required since the
human evidence available is inadequate to provide statistically valid
estimates of the dose response relationship for radiation-induced
mutations in humans. .
For a direct estimate, the starting point is the frequency of a
mutation per unit exposure in some, experimental animal study. The 1982
UNSCEAR (UNSCEAR82) report gave an example of the direct method for
estimating induction of balanced reciprocal translocations (a type of
chromosomal aberration) in males per rad of low-level, low-LET radiation.
1) Rate of induction in rhesus
monkey spermatogonia: cytogenetie
data [x rays delivered at
>/>30 rad/min]
2) Rate of induction that relates to
recoverable translocations in the
FI (1st filial generation) progeny
[divide (1) by 4] [based on mouse
data, UNSCEAR 1977]
3) Rate after low dose rate x rays:
based on mouse cytogenetic
observations [divide (2) by 2]
Induction rate/rad
8.6E-05
2.15E-05
1.075E-05
4) Rate after chronic gamma-irradiation:
based on mouse cytogenetic
observations [divide (2) by 10] 2.2E-06
*•• . ~ '
fD Expected rate of unbalanced products:
[multiply (3) and (4) by 2]
for (3): 2.15E-05
for (4): - 4.3E-06
6) Expected frequency of congenitally
malformed children in the F^, assuming
that about 6% of unbalanced products :
[item (5) above] contribute to this
for low dose rate x rays: 1.3E-06
for chronic gamma radiation: i/*3E-07
For humans, UNSCEAR (UNSCEAR 82) estimates that as a consequence of
induced balanced reciprocal translocations in exposed fathers, an
estimated 0.3 to 1.3 congenitally malformed children would occur in each
10° live births for every rad of paternal low-level radiation exposure.
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A complete direct estimate of genetic effects would include
estimates derived in a manner similar to, that shown above, for each type
of genetic damage. These direct estimates can be used to calculate the
risk'of genetic effects.in the first generation (F^) children of
exposed parents.
The indirect (or doubling dose) method of estimating genetic risk
also uses animal data but in a different way. The 1980 BEIR-3 report
(NAS80) demonstrates how such estimates are obtained.
Induction rate/rad
1) Average radiation-induced mutation
per gene for both sexes in mice
[based on 12 locus data in male
mice adjusted to a chronic gamma
radiation estimate]: induction
rate per rad, observed in the F^
generation
2) Estimated human spontaneous
mutation rate per gene
3) Relative mutation risk in humans
[divide (1) by (2)]
2.5E-08
5E-07
5E-0.6
0.005 to 0.05
4) Doubling dose: the exposure needed
to double the human mutation rate 200 to 20 rads
The doubling dose can then be used to estimate the equilibrium
genetic effects or the genetic burden in all future generations caused by
the exposure of parents. Since the genetic component of congenital
defects occurring in the population can be estimated by epidemiological
surveys, and this component is considered to be maintained at an
equilibrium level by mutations, a doubling dose of ionizing radiation
would double these genetic effects. Dividing the number of the various
genetic effects in 10^ live-births by the doubling dose yields the
estimate of genetic effects per rad. For example:
1) Autosomal dominant and x-linked
diseases, current incidence
2) Estimated doubling dose
3) Estimate of induced autosomal
dominant and x-linked diseases
10,000 per 106
live births
20 to 200 rads
50 to 500 per 106
live births per rad
of parental exposure.
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A doubling dose estimate assumes that the total population of both
sexes is equally irradiated, as occurs from background radiation, and
that the population exposed is large enough so that all genetic damage
can be expressed in future offspring. Although it is basically an
estimate of the total genetic burden across all future generations, it
can also provide an estimate of effects that occur in the first,
generation. Usually a fraction of the total genetic burden for each type
of damage is assigned to the first generation using population genetics
data as a basis to determine the fraction. For example, the BEIR-3
committee geneticists estimated that one-sixth of the total genetic
burden of x-linked mutations would be expressed in the first generation,
five-sixths across all subsequent generations. EPA assessment of risks
of genetic effects includes both first generation estimates and total
genetic burden estimates.
7.6.2 Estimates of Genetic Harm Resulting from Low-LET Radiations
One of the first estimates of genetic risk was made in 1956 by the
NAS Committee on the Biological Effects of Atomic Radiation (BEAR
Committee). Based on Drosophila (fruit fly) data and other
considerations, the BEAR Genetics Committee estimated that 10 roentgens
(10 R*) per generation continued indefinitely would lead to about 5,000
new instances of "tangible inherited defects" per 106 births, and about
one-tenth of them would occur in the first generation after the
irradiation began (NAS72). The UNSCEAR addressed genetic risk in their
1958, 1962, and 1966 reports (UNSCEAR58, 62, 66). During this period,
they estimated one rad of low-LET radiation would cause a 1 to 10 percent
increase in the spontaneous incidence of genetic effects.
In 1972, both the NAS BEIR Committee (NAS72) and UNSCEAR (UNSCEAR72)
reexam-ined the question of genetic risks. Although there were no
definitive human data, additional information was available on the
genetic effects of radiation on mammals and insects. In 1977, UNSCEAR
reevaluated the 1972 genetics estimates (UNSCEAR77). These new estimates
used recent information on the current incidence of various genetic
conditions, along with additional data on radiation exposure of mice and
marmosets and other considerations.
In 1980, an ICRP Task Group [ICRPTG] summarized recommendations that
formed the basis for the genetic risk estimates published in ICRP Report
26 (Of80). These risk estimates are based on data similar to those used
by the BEIR and UNSCEAR Committees, but with slightly different
assumptions and effect categories (Table 7-11).
R is the symbol for roentgen, a unit of measurement of x-radiation
exposure, equivalent to an absorbed dose in soft tissue of
approximately 0.9 rad.
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Table 7-11. ICRP Task Group estimate of number of cases of serious
genetic ill health in liveborn from parents irradiated
with 10^ person-rem in a population of constant size3
(Assumed doubling dose = 100 rad) [low level radiation
exposure]
Category of
genetic effect
First generation
Equilibrium
Unbalanced translocations:
risk of malformed liveborn
Trisomics and XO
Simple dominants and sex-
linked mutations
Dominants of incomplete
penetrance and multifactorial
disease maintained by mutation '•
Multifactorial disease not
maintained by mutation
Recessive disease
Total
23
30
20
16
89
30
30
100
160
320
a This is equivalent to effects per 10& liveborn following
an average parental population exposure of 1 rem per 30-yr
generation, as used by BEIR and UNSCEAR.
Source: Of80.
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The 1980 NAS BEIR Committee revised genetic risk estimates (NAS80).
The revision considered much of the same material that was in BEIR-1
(NAS72), the newer material considered by UNSCEAR in 1977 (UNSCEAR77),
and some additional data. Estimates for the first generation are about a
factor of 2 smaller than those reported in the BEIR-1 report. For all
generations, the new estimates are essentially the same (Table 7-12).
The most recent genetic risk estimate, in the 1982 UNSCEAR Report
(UNSCEAR82), includes some new data on cells in culture and the results
of genetic experiments using primates rather than rodents (Table 7-13).
Although all of the reports described above used somewhat different
sources of information, there is reasonable agreement in the estimates.
However, all these estimates have a considerable margin of error, both
inherent in the original observations and in the extrapolations from
experimental species to man. Some of the committee reports assessing the
situation have attempted to indicate the range of uncertainty; others
have simply used a central estimate. The same uncertainties exist for
the latter (central estimates) as for the former (see Table 7-14). Most
of the difference is caused by the newer information used in each
report. Note that all of these estimates are based on the extrapolation
of animal data to humans. Groups differ in their interpretation of how
genetic experiments in animals might be expressed in humans. While there
are no comparable human data at present, information on hereditary
defects among the children of A-bomb survivors provides a degree of
confidence that the animal data do not lead to underestimates of the
genetic risk following exposure to humans. (See "Observations on Human
Populations," which follows.)
It should be noted that the genetic risk estimates summarized in
Table 7-14 are for low-LET, low-dose, and low-dose-rate irradiation.
Much of the data was obtained from high dose rate studies, and most
authors have used a sex-averaged factor of 0.3 to correct for the change
from high-dose rate, low-LET to low-dose rate, low-LET exposure (NAS72,
80, UNSCEAR72,77). However, factors of 0.5 to 0.1 have also been used in
estimates of specific types of genetic damage (UNSCEAR72,77,82).
(A) " Beta Particles
Studies with the beta-particle-emitting isotopes carbon-14 and
tritium yielded RBEs of 1.0 and 0.7 to about 2.0, respectively, in
comparison to high-dose rate, high-dose exposure to x rays (UNSCEAR82).
At the present time, the RBE for genetic endpoints due to beta particles
is taken as one (UNSCEAR77,82).
7.6.3 Estimates of Genetic Harm from High-LET Radiations
Although genetic risk estimates are made for low-LET radiation, some
radioactive elements, deposited in the ovary or testis, can irradiate the
germ cells with alpha particles. The relative biological effectiveness
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Table 7-12. BEIR-3 estimates of genetic effects of an average
population exposure of 1 rem per 30-yr generation
[chronic x-ray or gamma radiation exposure]
Type of genetic
disorder
Current incidence
per lO^ liveborn
Effects per 10^ liveborn
per rem per generation
Autosomal dominant
and x-1 inked
Irregularly inherited
Recessive
Chromosomal aberrations
Total
10,000
90,000
1,100
6,000
107,100
First generation*
5-65
(not estimated)
Very few
Fewer than 10
5-75
Equilibrium**
40-200
20-900
Very slow
increase
Increases
only
slightly
60-1100
* First generation effects estimates are reduced from acute fractionated
exposure estimates by a factor of 3 for dose rate effects and 1.9 for
fractionation effects (NAS80, p. 117).
** Equilibrium effects estimates are based on low dose r,ate
studies in mice (NAS80, pp. 109-110).
Source: NAS80.
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Table 7-13. UNSCEAR 1982 estimated effect of 1 rad per generation of
low-dose or low-dose rate, low-LET radiation on a
population of 10^ liveborn according to the doubling
dose method (Assumed doubling dose = 100 rad)
[low level, low-LET radiation]
Disease classification
Autosomal dominant and
x-linked diseases
Recessive diseases
increase
Chromosomal diseases
Structural
Numerical
Congenital anomalies,
anomalies expressed later,
constitutional and
degenerative diseases
Total
Current Effect of 1 rad
incidence per generation
First generation Equilibrium
10,000 15 100
2,500 Slight Slow
' 400 2.4 4
3,000 Probably :
very small
90,000 4.5 " 45
105,900 22 149
Source: (UNSCEAR82).
7-56
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Table 7-14. Sunsnary of genetic risk estimates per 10° liveborn
for an average population exposure of 1 rad of low-dose
or low-dose rate, low-LET radiation in a 30-yr
generation
Source
Serious hereditary effects
First generation
Equilibrium
(all generations)
BEAR, 1956 (NAS72)
BEIR-I, 1972 (NAS72)
UNSCEAR, 1972 (UNSCEAR72)
UNSCEAR, 1977 (UNSCEAR77)
ICRP, 1980 (Of80)
BEIR-3, 1980 (NAS80)
UNSCEAR, 1982 (UNSCEAR82)
49a (12-200)
9a (6-15)
63
89 "
L9a (5-75)
22
500
300a (60-1500)
300
185
320
257a (60-1100)
149
Numbers in parentheses are the range of estimates.
a Geometric Mean. The geometric mean of two numbers is the square root
of their product; in general, it is the Nth root of the product of
N numbers.
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(RBE) of high-LET radiation, such as alpha particles, is defined as the
ratio of the dose (rad) of low-LET radiation to the dose of high-LET
radiation producing the same specific patho-physiological endpoint.
Studies of the RBE for alpha-emitting elements in germinal tissue
have been carried out only with plutonium-239. Studies comparing
cytogenetic endpoints after chronic low-dose-rate gamma radiation
exposure, or incorporation of plutonium-239 in the mouse testis, have
yielded RBEs of 23 to 50 for the type of genetic injury (reciprocal
dranslocations) that might be transmitted to liveborn offspring (NAS80,
UNSCEAR77,82). However, an RBE of 4 for plutonium-239 compared to
chronic gamma radiation was reported for specific locus mutations
observed in neonate mice (NAS80). Neutron RBE, determined from
cytogenetic studies in mice, also ranges from about 4 to 50 (UNSCEAR82,
Gr83a, Ga82). Most reports use an RBE of 20 to convert risk estimates
for low-dose rate, low-LET radiation to risk estimates for high-LET
radiation.
7.6.4 Uncertainty in Estimates of Radiogenetic Harm
Chromosomal damage -and mutations have been demonstrated in cells in
culture, in plants, in insects, and in mammals (UNSCEAR72,77,82).
Chromosome studies in peripheral blood lymphocytes of persons exposed to
radiation have shown a dose-related increase in chromosome aberrations
(structural damage to chromosome) (UNSCEAR82). In a study of nuclear
dockyard workers exposed to external x-radiation at rates of less than
5 rad/yr, Evans et al. (Ev79) found a significant increase in the
incidence of chromosome aberrations in peripheral lymphocytes. The
increase appeared to have ,a linear dependence on cumulative dose. In a
study of people working and living in a high natural background area
where there was both external gamma-radiation and internal
alpha-radiation, Pohl-Ruling et al. (Po78) reported a complex dose
response curve. For mainly gamma-radiation exposure (less than
10 percent alpha-radiation), they reported that chromosome aberrations
increased linearly from 100 to 200 mrad/yr, plateaued from 300 mrad to
2 rad/yr, and then increased linearly again for doses above 2 rad/yr.
Although chromosomal damage in peripheral blood lymphocytes cannot
be used for predicting genetic risk in progeny of exposed persons, it is
believed .by some to be a direct expression of the damage, analogous to
that induced in germ cells, resulting from the radiation exposure. It is
at least evidence that chromosome damage can occur in vivo in humans.
Since human data are so sparse, they can be used only to develop
upper bounds of some classes of genetic risks following radiation
exposure. Most numerical risk estimates are based on extrapolations from
animal data. As genetic studies proceeded, emphasis shifted from
Drosophila (fruit flies) to mammalian species in attempts to find an
experimental system that would reasonably project what might happen in
humans.
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For example, Van Buul (Va80) reported the slope (b) of the linear
regression, Y = a + bD, for induction of reciprocal translocations in
spermatogonia (one of the stages of sperm development) in various species
as follows:
b x E+04 + sd x E+04
Rhesus monkey
Mouse
Rabbit
Guinea Pig
Marmoset
Human
0.86 + 0.04
1.29 + 0.02 to
2.90 + 0.34
1.48 + 0.13
0.91 + 0.10
7.44 + 0.95
3.40 +0.72
These data indicate that animal-based estimates for this type of genetic
effect would be within a factor of 4 of the true human value. In this
case, most of the animal results would underestimate the risk in humans.
However, when risk estimates such as this are used in direct
estimation of risk for the first generation, the total uncertainty in the
estimate becomes indeterminate. Even if studies have been made in a
species that can predict the-dose response and risk coefficient for a
specific, radiation-induced genetic damage, there is no certainty that it
predicts the response for all genetic damage of that type. In addition,
as shown in the example from the 1982 UNSCEAR report (UNSCEAR82) in
Section 7.6.1, additional assumptions based on observations, usually in
other animal species, are used to adjust the risk coefficient to what is
expected for humans. The uncertainty in these extrapolations has not
been quantified.
A rough estimate of the uncertainty can be obtained by comparing
direct estimates of risk for the first generation with doubling-dose
estimates in the 1977 UNSCEAR report (UNSCEAR77). The estimates differ
by a factor between 2 and 6, with the direct estimate usually smaller
than the doubling-dose estimate.
A basic assumption in the doubling-dose method of estimation is that
there is a proportionality between radiation-induced and spontaneous
mutation rates. Some of the uncertainty was removed in the 1982 UNSCEAR
report with the observation that in two-test systems (fruit flies and
bacteria), there is a proportionality between spontaneous and induced
mutation rates at a number of individual gene sites. There is still some
question as to whether the sites that have been examined are repre-
sentative of all sites and all gene loci dr not. The doubling-dose
estimate dose, however,, seems better supported than the direct estimate.
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While there is still some uncertainty as to which hereditary
conditions would be doubled by a doubling dose, future studies on genetic
conditions and diseases can only increase the total number of such
conditions. Every report, from the 1972 BEIR and UNSCEAR reports to the
most recent, has listed an increased number of conditions and diseases
which have a genetic component and hence may be increased by exposure to
ionizing radiations.
(A) Observations on Human Populations
As noted earlier, the genetic risk estimates are based on
interpretation of animal experiments as applied to data on
naturally-occurring hereditary diseases and defects in man. A study of
the,birth cohort consisting of children of the Japanese A-bomb survivors
was initiated in mid-1946. In a detailed monograph, Neel and Schull
(Ne56) outlined the background of this first study and made a detailed
analysis of the findings to January 1954 when the study terminated. The
study was designed to determine: (1) if during the first year of life,
any differences could be observed in children born to exposed parents
when compared to children born to suitable control parents, and (2) if
differences existed, how should they be interpreted (Ne56). At the time
the study started, there were data on spontaneous and radiation-induced
mutation in Drosophila, but little was known concerning spontaneous
mutation rates in mammals and less on the effects of radiation on
mammals. The authors concluded that, based on the human data, it was
improbable that human genes were so sensitive that exposures as low as
3 R, or even 10 R, would double the mutation rate.
While this first study addressed a number of endpoints, including
sex ratio, malformations, perinatal data and anthropometric data,
subsequent studies have addressed other endpoints. The most recent
reports on this birth cohort of 70,082 persons have reported data on six
endpoints. Frequency of stillbirths, major congenital defects, prenatal
death, and frequency of death prior to age 17 have been examined in the
entire cohort. Frequency of cytogenetic aberrations (sex chromosome
aneuploidy) and frequency of biochemical variants (a variant enzyme or
protein electrophoresis pattern) have been measured on large subsets of
this cohort.
There are small but statistically insignificant differences between
the number of effects in the children of the proximally and distally
exposed with respect to these various indicators. These differences are
in the direction of the hypothesis that mutations were produced by the
parental exposure. Taking these differences then as the point of
departure for an estimate of the human doubling dose, an estimated
doubling dose for low-LET radiation at high doses and dose rates for
human genetic effects of about 156 rem (Sc81) or 250 rem (Sa82) was
obtained as an unweighted average. When each individual estimate was
weighted by the inverse of its variance, an average of 139 rem was found
(Sc84). Because of the assumptions necessary for these calculations, as
7-60
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well as the inherent statistical errors, the errors associated with these
estimates are rather large. As a result, a reasonable lower bound to the
human estimate overlaps much of the range based on extrapolation from
mouse data.
As noted above, animal studies indicate that chronic, exposures to
low-LET radiation would be less hazardous than acute exposures by a
factor of about 3 (NAS72, 80). If applicable to the Japanese A-bomb
survivors, this would increase the estimated' doubling doses cited above
to 468 rem, 750 rem, and 417 rem, respectively. These recent reports
thus suggest the minimum doubling dose for humans may be 4 to 7 times
higher than those in Table 7-14 (based on animal data). It would be
premature to estimate the exact magnitude since these reports are based
on the T65 dosimetry in Japan (see Section 7.2), which is being revised.
The EPA is using the geometric mean of the BEIR-3 range of doubling
doses, about 110 rads. Although the best estimate of the minimum
doubling dose derived from human data is 4 to 7 times greater than the
EPA estimate, the 95 percent lower confidence limit averages about
70 rem. Therefore, EPA believes the estimate of doubling of about
100 rads is on the conservative side; however, it is compatible with both
human and mouse data and should not be changed at this time. However,
the EPA estimates of genetic risks will be reviewed and revised, if
necessary, when the dosimetry of A-bomb survivors is revised.
(B) Ranges of Estimates Provided by Various Models
EPA has continued to follow the recommendations of the 1980 BEIR-3
and earlier committees and uses a linear nonthreshold model for
estimating genetic effects, although, as pointed out by the 1982 UNSCEAR
committee, a number of models other than linear (Y '= c + aD) have been
proposed: e.g., linear-quadratic (Y = C + bD + eD2), quadratic (Y = k
+ fD2), or even a power function (Y = K + gD^1)* .
; Some data on specific genetic endpoints obtained with acute low-LET
exposures are well described by a linear-quadratic function. Moreover,
in some of these cases, it has been found that a reduction in dose rate
(or fractionation of dose) produced a reduction in the quadratic term
seen at high doses with little or no effect on the linear component.
Such observations can be qualitatively explained, as previously discussed
in reference to somatic effects (Section 7.2.2), in terms of the dual
radiation action theory of Kellerer and Rossi (Ke72), as well as
alternative theories, e.g., one involving enzyme saturation (Go80, Ru58).
* Y is yield of genetic effects; D is radiation dose; c, C, k, and K are
spontaneous incidence constants for genetic effects; and a; b, e, f,
g, and h are the rate constants for radiation-induced genetic effects.
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The linear model adopted by BEIR-3 and EPA incorporates a factor of
3 reduction in extrapolating results obtained with high acute exposures
to low dose rates. For this reason, the predictions obtained with the
model in the low dose region, are roughly consistent with what one would
obtain with a linear-quadratic model based on the same data.
Most of the arguments for a nonlinear dose response have been based
on target theory (Le62) or microdosimetry site theory (Ke72, NAS80).
However, other theories based on biology [e.g., enzyme
induction-saturation (Go80,82), repair-misrepair (To80)] could also
provide models that fit the observed data. There is still much
disagreement on which dose response model is appropriate for estimating
genetic effects in humans. Until there is consensus, EPA will continue
to use the linear nonthreshold model.
Even though genetic risk estimates made by different committees
based on the linear non-threshold model vary, the agreement is reasonably
good. While the authors of the reports used different animal models,
interpreted them in different ways, and had different estimates of the
level of human genetic conditions in the population, the range of risk
coefficients is about an order of magnitude (see Table 7-14). For the
most recent, more comparable estimates, the range is a factor of 2 to 4
(see ICRP, BEIR-3, and UNSCEAR 1982 in Table 7-14).
7.6.5 The EPA Genetic Risk Estimate
There is no compelling evidence for preferring any one set of the
genetic risk estimates listed in Table 7-14. EPA has used the estimates
from BEIR-3 (NAS80). These "indirect" estimates are calculated using the
tnormal prevalence of genetic defects and the dose that is considered to
double this risk. The NAS estimates that EPA uses are based on a
"doubling dose" range with a lower bound of 50 rem and an upper bound of
250 rera. We prefer these risk estimates to those made by the ICRP task
group (Of80), which used "direct" estimates for some types of genetic
damage with doubling-dose estimates for others. We also prefer the
BEIR-3 risk estimates to the "direct" estimates of UNSCEAR 1982, which
tabulates genetic risk separately by the direct method and by the
doubling-dose method.
Our reasons are as follows: mutation rates for all gene loci
affected by ionizing radiation are not known nor have loci associated
with "serious" genetic conditions been identified. Therefore, the risk
estimated by the direct method, at this time, is incomplete, does not
include the same types of damage estimated by doubling doses, and was not
considered further. Moreover, the BEIR-3 genetic risk estimates provide
a better estimate of uncertainty than the UNSCEAR 1982 and ICRPTG
estimates because the BEIR-3 Committee assigned a range of uncertainty
for multifactorial diseases (>5 percent to <50 percent) that reflects
the uncertainty in the numbers better than the other estimates (5 percent
and 10 percent, respectively).
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In developing the average mutation rate for the two sexes used in
the calculation of the relative mutation risk, the BEIR-3 Committee
postulated that the induced mutation rate in females was about 40 percent
of that in males (NAS80). Recent studies by Dobson et al. suggest that
the assumption was invalid and that human oocytes should have a risk
equivalent to that of human spermatogonia. This would increase the risk
estimate obtained from doubling-dose methods by a factor of 1.43 (.DoBJa,
Do83b, Do84a, Do84b).
We recognize, however, that the use of the doubling-dose concept
does assume that radiation-induced genetic damage is in some way
proportional to "spontaneous" damage. As noted earlier, the recent
evidence obtained in insects (Drosophila) and bacteria (E^'coli) supports
the hypothesis that, with the exception of "hot spots" for mutation, the
radiation-induced mutation rate is proportional to the spontaneous rate
(UNSCEAR82). No proof that this is also true in mammals is available yet.
The BEIR-3 estimates for low-LET radiations give a considerable
range. To express the range as a single estimate, the geometric mean of
the range is used, a method first recommended by UNSCEAR (UNSCEAR58) for
purposes of calculating genetic risk. The factor of 3 increase in risk
for high-dose rate, low-LET radiation, noted earlier, is also used.
The question of RBE for high-LET radiation is more difficult. As
noted above, estimated RBEs for plutonium-239 alphas versus chronic gamma
radiation for reciprocal translocations as determined by cytogenetic
analyses are between 23 and 50 (NAS80, UNSCEAR82). However, the observed
RBE for single locus mutations in developing offspring of male mice given
plutonium-239 compared to those given chronic gamma irradiation is 4
(NAS80). The average of RBEs for reciprocal translocations and for
specific locus mutations is 20. Since reported neutron RBEs are similar
to those listed above for plutonium-239 alpha radiation, we use an RBE of
20 to estimate genetic risks for all high-LET radiations. Tins is
consistent with the RBE for high-LET particles recommended for estimated
genetic risks associated with space flight (Gr83b).
Genetic risk estimates used by EPA for high- and low-LET radiations
are listed in Table 7-15. As noted above, EPA uses.the dose received
before age 30 in assessing genetic risks.
The EPA estimates in Table 7-15 are limited, like all other human
genetic risk estimates, by the lack of confirming evidence of genetic
effects in humans. These estimates depend on a presumed resemblance ot
radiation effects in animals to those in humans. The magnitude of the
possible error is indeterminable. The largest source of data, the
Japanese A-bomb survivors, appears, at best, to provide only an estimate
of the minimum doubling dose for calculating the maximum genetic risk in
man. However, doubling-dose estimates are also uncertain since the
number of human disorders having a recognized genetic component is
constantly increasing, and the type of genetic damage implicated in a
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Table 7-15. Estimated frequency of genetic disorders in a birth
cohort due to exposure of the parents to 1 rad per
generation
Serious heritable disorders
(Cases per 10 liveborn)
Radiation
Low Dose Rate,
LOW-LET
High Dose Rate,
LOW-LET
High-LET
First generation.
low high
20 30
60 90
400 600
All generations
lowa high5
260 370
780 1110
5200 7400
a Female sensitivity to induction of genetic effects is
40 percent as great as that of males.
Female sensitivity to induction of genetic effects is
equal to that of males.
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specific disorder may change. The combined uncertainties in
doubling-dose estimates and the magnitude of genetic contributions to
various disorders probably introduce an overall uncertainty of about an
order of magnitude in the risk estimates. Moreover, the BEIR Committee
in deriving its estimate has assumed that almost all of the risk was due
to recessive mutations which would eventually be eliminated. To what
extent this occurs will depend on medical practices in the future. It is
possible, as our knowledge of medicine improves, that recessive ,
hereditary defects will be carried on for many more generations than
assumed by the BEIR Committee.
The relative risk of high-LET radiation compared to low-dose-rate,
low-LET radiation (RBE) is also uncertain. The data are sparse, and
different studies often used different endpoints. In addition, the
microscopic dosimetry, i.e., the actual absorbed dose in the cells at
risk, is poorly known. However, the RBE estimate used by EPA should be
within a factor of 5 of the true RBE for high-LET radiation.
7.6.6 Effects of Multigeneration Exposure
As noted earlier, while the somatic effects, i.e., cancer, occur in
persons exposed to ionizing radiation, the genetic effects occur in
progeny, perhaps generations later. The number of effects appearing in
the first generation is based on direct estimates of the mutations
induced by irradiation and should not change appreciably regardless of
the background or -"spontaneous" mutation rate in the exposed population.
The estimate for total genetic effects, or the equilibrium estimate, is
based on the doubling-dose concept. For these estimates, the background
mutation rate is important: it is the background rate that is being
"doubled."
If there is long-lived environmental contamination, such that 30
generations or more are exposed UlOOO years), the background mutation
rate will change and come into equilibrium with the new level of
radiation background. There will be an accumulation of new radiation-
induced mutations until the background mutation rate has reached
equilibrium with this continued insult.
While predicting 1,000 yr in the future is chancy, at best, if it is
assumed that there are no medical advances, and no changes in man or his
environment, then an estimate can 'be made. In Table 7-15, it is
estimated that exposure to 1 rad per generation of low-dose-rate low-LET
radiation will induce 260 cases of serious heritable disorders per 10°
live births in all generations. This is for a background mutation rate
leading to 29,120 cases of serious heritable disorders per 10° live
births. The "all generations" estimate in Table 7-15 is equal to the
"equilibrium" estimate in Table 7-12. The "all generations" estimate is
used for exposures to a single generation; the same number is employed as
the "equilibrium" estimate for multigeneration exposures (see NAS80,
p. 126, note 16). Thus, the risk estimate can be reexpressed as an
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estimate of the effects expected for a given change in the level of
background radiation (Table 7-16). Since these calculations are based
both on the background level mutations and the doubling dose, changes in
either must be reflected in new calculations.
7.6.7 Uncertainties in Risk Estimates for
Radiogenic Genetic Effects
As noted throughout the preceding sections, there are sources of
uncertainty in the genetic risk estimates. The overall uncertainty can
be addressed only in a semi-quantitative manner. The identified sources
of uncertainty are listed in Table 7-17. Uncertainties listed in
Table 7-17 are likely to be independent of each other and therefore
unlikely to be correlated in -sign. Although the root mean square sum of
the numerical uncertainties suggests the true risk could be a factor of 4
higher or lower [x/* by a factor of 4], it is unlikely in light of the
Japanese A-bomb survivor data that the upper bound is correct.
7.6.8 Teratogenic Effects ,
Although human teratogenesis (congenital abnormalities or defects)
associated with x-ray exposure has a long history, the early literature
deals mostly with case reports. Stettner reported a case in 1921 (St21)
and Murphy and Goldstein (Mu29, Gol29) studied a series of pregnancies in
which 18 of the children born to 76 irradiated mothers had microcephaly
(reduced head circumference). However, the irradiation exposures were
high.
In 1930, Murphy exposed rats to x rays at doses of 200 R to 1600 R.
Thirty-four of 120 exposed females had litters, and five of the litters
had animals with developmental defects (Mu30). He felt that this study
confirmed his clinical observations and earlier reports of animal
studies. Although there were additional studies of radiation-induced
mammalian teratogenesis before 1950, the majority of the studies were
done after that time (see Ru53 for a review), perhaps reflecting concerns
about radiation hazards caused by the explosion of nuclear weapons in
1945 (Ja70).
Much of the work done after World War II was done using mice (Ru50,
Ru54, Ru56) and rats (Wi54, Hi54). Early studies, at relatively high
radiation exposures, 25 R and above, established some dose response
relationships. More important, they established the timetable of
sensitivity of the developing rodent embryo and fetus to radiation
effects (Ru54, Hi53, Se69, Hi66).
Rugh, in his review of radiation teratogenesis (Ru70), listed the
reported mammalian anomalies and the exposure causing them. The lowest
reported exposure was 12.5 R for structural defects and 1 R for
functional defects. 'He also suggested human exposure between ovulation
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Table 7-16.
Increase in background level of genetic
effects after 30 generations or more
Increase in
background
radiation
(mrad/y)
Increase in serious heritable
disorders per
10 live births
Low-dose-rate,
low-LET radiation
High-LET
radiation
0.1
1.0
10.0
0.8
8
80
16
160
1600
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Table 7-17. A list of the causes of uncertainty in the
genetic risk estimates
Source of uncertainty
Degree of uncertainty
in risk estimates
Selection of species to use in
developing a direct estimate
Selection of species and loci to
use in developing a doubling-dose
estimate
Use of - division by a factor of 3 -
to convert acute, high-dose low-LET
estimates to chronic low-LET estimates
Sensitivity of oogonia compared to
spermatogonia as described in BEIR-3
Background rate selected for use
with a doubling dose
Selection of RBE for high-LET
radiation compared to an RBE of 20
Underestimate of the doubling dose
required in man
x/*, a factor of 4
-100% to
+indeterminate
x/*, a factor of 3
0.5b>d -1.0c»d
+/-, indeterminate
X/T, a factor of 5
*, a factor of 3e
a The risk estimate cannot go below zero, -100%, but it may not be
possible to determine the upper bound; indeterminate.
Assumes no radiation-induced mutations from oocytes.
c Assumes equal radiation-induced mutations from oocytes and
spermatocytes.
d In reference to the high estimate in Table 7-15.
e If the most recent analysis of the Japanese A-bomb survivors is
correct, the lower bound for an estimate of the doubling dose in
man is at least 3 times greater than the average doubling dose
in the mouse.
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and about 7 weeks gestational age could lead to structural defects, and
exposures from about 6 weeks gestational age until birth could lead to
functional defects. In a later review (Ru71), Rugh suggested structural
defects in the skeleton might be induced as late as the 10th week of
gestation and functional defects as early as the 4th week. It should be
noted that the gestation period in mice is much shorter than that in
humans and that weeks of gestation referred to above are in^terms of
equivalent stages of mouse-human development. However, estimates of
equivalent gestational age are not very accurate.
Rugh (Ru71) suggested there may be no threshold for
radiation-induced congenital effects in the early human fetus. In the
case of human microcephaly (small head size) and mental retardation, at
least, there are.some data to support this theory (Ot83, 84). However,
for most teratogenic effects, the dose response at low doses is not
known. In 1978, Michel and Fritz-Niggli (Mi78) reported induction of a
significant increase in growth retardation, eye and nervous system
abnormalities, and post-implantation losses in mice exposed to 1 R. The
increase was still greater if there was concurrent exposure to
radiosensitizing chemicals such as iodoacetimide or tetracycline (Mi78).
In other reports of animal studies it appeared as if teratologic
effects, other than perhaps growth retardation, had a threshold for
induction of effects (Ru54, Ru53, Wi54). However, Ohzu (Oh65) showed
that doses as low as 5 -R to preimplantation mouse embryos caused
increased resorption of implanted embryos and structural abnormalities in
survivors. Then in 1970, Jacobsen (Ja70) reported a study in which mice
were exposed to 5, 20, or 100 R on the 8th day of pregnancy. He
concluded that the dose response function for induction of skeletal
effects was linear, or nearly linear, with no observable threshold.
appears consistent with a report by Russell (Ru57), which suggested a
threshold for some effects whereas others appeared to be linearly
proportional to dose.
One of the problems with the teratologic studies in animals is the
difficulty of determining how dose response data should be interpreted.
Russell (Ru54) pointed out some aspects of the problem:. (1) although
radiation is absorbed throughout the embryo, it causes selective damage
that is consistently dependent on the stage of embryonic development at
the time of irradiation, and (2) the damaged parts respond, in a
consistent manner, within a narrow time range. However, while low-dose
irradiation at a certain stage of development produces changes only in
those tissues and systems which are most sensitive at that time, higher
doses may induce additional abnormalities in components which are most
sensitive at other stages of development, and may further modify
expression of the changes induced in parts of the embryo at maximum
sensitivity during the time of irradiation. In the first case, damage
may be to primordial cells themselves, while in the second, the damage
may lead indirectly to the same or different endpoints.
This
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The human embryo/fetus starts as a single, fertilized egg and
divides and differentiates to produce the normal infant at term. (The
embryonic period, when organs develop, is the period from conception
through 7 weeks gestational age. The fetal period, a time of in utero
growth, is the period from 8 weeks gestational age to birth.) The
different organ and tissue primordia develop independently and at
different rates. However, they are in contact through chemical induction
or evocation (Ar54). These chemical messages between cells are important
in bringing about orderly development and the correct timing and fitting
together of parts of organs or organisms. While radiation can disrupt
this pattern, interpretation of the response may be difficult. Since the
cells in the embryo/fetus differentiate, divide, and proliferate at
different times during gestation and at different rates, gestational
times when cells of specific organs or tissues reach maximum sensitivity
to radiation are different. Each embryo/fetus has a different
timetable. In fact, each half (left/right) of an embryo/fetus may have a
slightly different timetable.
In addition, there is a continuum of variation from the hypothetical
normal to the extreme deviant which is obviously recognizable. There is
no logical place to draw a line of separation between normal and
abnormal. The distinction between minor variations of normal and frank
malformation, therefore, is an arbitrary one, and each investigator must
establish his or her own criteria and apply them to spontaneous and
induced abnormalities alike (HWC73).
The limitations of the human data available make the use of animals
in both descriptive and experimental studies inevitable. However, this
gives rise to speculation about the possible relevance of such studies to
•man. There are species differences in development attributable partly to
the differing complexity of the adult organs, but especially to
differences in growth rates and timing of birth in relation to the
developmental events. For example, the histological structure of the
brain is, in general, surprisingly similar, both in composition and in
function, from one mammalian species to another; and the sequence of
events is also similar (Ref.). However, the processes of brain
development that occur from conception to about the second year of life
in man are qualitatively similar to those seen in the rat during the
first six weeks after conception (Do79, 81).
For example, a major landmark, the transition from the principal
phase of multiplication of the neuronal precursors to that of glial
multiplication, occurs shortly before midgestation in man, but at about
the time of birth in the rat (Do73). In this respect, then, the rat is
much less neurologically mature at birth than the newborn human infant.
Many other species are more mature at birth; the spectrum ranges from the
late-maturing mouse and rat to the early-maturing guinea pig, with
non-human primates much closer to the guinea pig than to man (Do79, 81).
As a consequence, it is unreasonable to compare a newborn rat's brain,
which has not begun to myelinate (Do79, 81), with that of a newborn
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human, which has (Do79, 81), or with that of a newborn guinea pig in
which myelination has been completed (Do79, 81).
Nevertheless, in the study of teratogenic effects of prenatal
exposure to ionizing radiation,• in which the timing of the exposure in
relation to the program of developmental events dictates the consequences
of that insult, it is only necessary to apply the experimental exposure
at the appropriate stage (rather than at a similar age) of embryonic or
fetal development in any species to produce similar results in all (Do79,
81). The duration of exposure must, however, match the different time
scales in the different species. Unless these elementary rules of
cross-species adjustments are followed, extrapolation of even qualitative
estimates of effects will be of dubious relevance and worth.
Because of the problems in interpretation listed above, a pragmatic
approach to evaluation of studies is useful. The dose response should be
given as the simplest function that fits the data (often linear or linear
with a threshold). No attempt should be made to develop complex dose
response models unless the evidence is unequivocal.
(A) Teratologic Effects: Mental Retardation in Humans
The first report of congenital abnormalities in children exposed i.n
utero to radiation from atomic bombs was that of Plummer (P152). Twelve
children with microcephaly, of which 10 also had mental,retardation, had
been identified in Hiroshima in a small set of the in utero exposed
survivors. They were found as part of a program started in 1950 to study
children exposed in the first trimester of gestation. However, not all
of the in utero exposed survivors were examined. In 1955, the program
was expanded to include all survivors exposed in utero.
Studies initiated during the program have shown radiation-related
(1) growth retardation; (2) increased microcephaly; (3) increased
mortality, especially infant mortality; (4) temporary suppression of
antibody production against influenza; and (5) increased frequency of
chromosomal aberrations in peripheral lymphocytes (Ka73).
Although there have been a number of studies of Japanese A-bomb
survivors, including one showing a dose- and gestational age-related
increase in postnatal mortality (Ka73), only the incidences of
microcephaly And mental retardation have been investigated to any great
extent. In t/he most recent report, Otake and Schull (Ot83, 84) showed
•that mental retardation was particularly associated with exposure between
8 and 15/weeks of gestation (10 to 17 weeks of gestation if counted from
the las€ menstrual period). They further found the data suggested
littley if any, non-linearity and were consistent with a linear
dose-response relationsh/p for induction of mental retardation that
yielded a probability of occurrence of severe mental retardation of
4.16^0.4 cases per 1,000 live births per rad of exposure (Ot84). A child
was classified as severely mentally retarded if he or she was "unable to
7-71
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perform simple calculations, to make simple conversation, to care for
himself or herself, or if he or she was completely unmanageable or had
been institutionalized" (Ot83, 84). There was, however, no evidence of
an effect in those exposed at 0 to 7 weeks of gestation (Ot83). Exposure
at 16 weeks or more of gestation was about a factor of 4 less effective,
with only a weak relationship between exposure and risk, and with few
cases below 50 rads exposure (Ot84).
Mental retardation can be classified as mild (IQ 50-70), moderate
(IQ 35-49), severe (IQ 20-34), and profound (IQ<20) (WH075). However,
some investigators use only mild mental retardation (IQ 50-70) and severe
mental retardation (IQ<50) as classes (Gu77b, HaSla, St84). Mental
retardation is not usually diagnosed at birth but at some later time,
often at school age. Since the mental retardation may have been caused
before or during gestation, at the time of birth, or at some time after
birth, that fraction caused before or during gestation must be estimated.
In like manner, since mental retardation caused before birth may be due
to genetic conditions, infections, physiologic conditions, etc., the
fraction related to unknown causes during gestation must be estimated.
This is the fraction that might possibly be related to radiation exposure.
Estimates of the risk of mental retardation for a rad of
embryo/fetus exposure in the U.S. population can be derived using the
absolute risk calculated by Otake and Schull for the Japanese survivors
(Ot84). Otake and Schull (Ot84) gave an estimate for one case entitled,
"The Relationship of Mental Retardation to Absorbed Fetal Exposure in the
'Sensitive' Period When All 'Controls' Are Combined." This estimate of
frequency of mental retardation, 0.416 per 100 rads, could be directly
applicable to a U.S. population. In this case, the risk estimate would
be about:
Four cases of severe mental retardation per
1,000 live births per rad of exposure
during the 8th and 15th week of gestation.
Data on mental retardation in school age populations in developed
countries suggest a prevalence of 2.8 cases/1,000 (Uppsala County,
Sweden) to 7.4 cases/1,000 (Amsterdam, Holland) of severe mental
retardation, with a mean of about 4.3 _+ 1.3 cases/1,000 (St84). Where
data are available for males and females separately, the male rate is
about 30 percent higher than the female rate (St84). Historically, the
prevalence of mild mental retardation has been 6 to 10 times greater than
that of severe mental retardation. However, in recent Swedish studies,
the Crates of prevalence of mild and severe mental retardation have been
similar (St84). This was suggested to be due to a decline in the
"cultural-familial syndrome." That is, improved nutrition, decline in
infection and diseases of childhood, increased social and intellectual
stimulation, etc. , combined to reduce the proportion of nonorganic mental
retardation and, therefore, the prevalence of mild mental retardation
(St84).
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In studies of the causes of mental retardation, 23 to 42 percent of
the cases have no identified cause (Gu77a, HaSlb, St84). It is this
(idiopathic) portion of the mental retardation that may be susceptible to
increase from radiation exposure of the embryo/fetus and should be used
as the "background" incidence for comparison with radiation-induced
effects. In that case, the prevalence of idiopathic mental retardation
would be 0.6 to 3.1 cases per 1,000 of severe mental retardation and
perhaps an equal number of cases of mild mental retardation. This
estimate may be biased low because mental retardation induced during
gestation is often associated with a high childhood death rate (St84).
If this is generally true for idiopathic causes of mental retardation, it
would cause an underestimation of the risk.
The risk of increased mental retardation per rad of embryo/fetus
exposure during the 8- to 15-week gestational period estimated as 4 x
10~3 cases per live birth, compares with an earlier UNSCEAR (UNSCEAR77)
estimate of 1 x 10~3 excess cases of mental retardation per rad per
live birth. The UNSCEAR estimate, however, did not consider gestational
age at the time of exposure. The Otake and Schull report (Ot84) did
address gestational age and estimated a higher risk, but with what
appears to be a narrower window of maximum susceptibility*
If the estimate is applicable, the low-LET background radiation
(about 15 mrads) delivered during the 8- to 15-week gestational
age-sensitive period could induce a risk of 6 x 10~5 cases of severe
mental retardation per live birth. This can be compared to an estimate
of a spontaneous occurrence of 0.6 x 10~3 to 3.1 x 10~3 cases of
idiopathic severe mental retardation per live birth.
(B) Teratologic Effects: Microcephaly in Humans
Plummer (P152) reported microcephaly associated with mental
retardation in Japanese A-bomb survivors exposed in utero. Wood (Wo65,
66) reported both were increased. The diagnosis of reduced head
circumference was based on "normal distribution" statistical theory
(Wo66); i.e., in a population, the probability of having a given head
circumference is expected to be normally distributed around the mean head
circumference for that population.
For example, in a population of live born children, 2.275 percent
will have a head circumference 2 standard deviations or more smaller than
the mean, 0.621 percent will have a head circumference 2.5 standard
deviations or more smaller than the mean, and 0.135 percent will have a
head circumference 3 standard deviations or more smaller than the mean
(statistical estimates based on a normal distribution).
For most of the studies of the Japanese A-bomb survivors exposed in
utero, if the head circumference was two or more standard deviations
smaller than the mean for 'the appropriate controls in the unexposed •
population, the case was classified as having reduced head circumference
7-73
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even if the data had not been adjusted for differences in stature (Ta67, '
Mi72, Wo65). While a definitive relationship between reduced head
circumference and mental retardation has not- been established, there is
evidence that they are related.
For example, there is evidence in a nonselected group of 9,379
children that mental retardation may be estimated using incidence of
microcephaly, even though head circumference in the absence of other
supporting data, e.g., height or proportion, is an uncertain indicator of
mental retardation.. Based on this study of 9,379 children, Nelson and
Deutschberger (Ne70) concluded that about half of the children with a
head circumference 2.5 standard deviations or more smaller than average
had IQs of 79 or lower. Since 0.67 percent of those studied were in this
size group, the observed number is about what would be-expected based on
a normal distribution of head size in a population (0.62 percent). The
estimated incidence of mental retardation per live birth in a population
would then be:
(6.7 cases of microcephaly per 1,000 live births) x
case of mental retardation
case of microcephaly
0.5
or about 3.Ceases of mental retardation per 1,000 live births. This
might be divided roughly into 1,7 cases each of ,mild and severe mental
retardation. " "•'./•
Studies of the Japanese survivors also show a relationship between
reduced head size and mental retardation, but all these studies are based
on subsets of the total in utero population. The fraction of mentally
retarded with reduced head circumference has been reported as 50 percent
(RERF78) to 70 percent (Wo66). While the fraction of those selected for
reduced head circumference who had mental retardation has been reported
as 11 percent (Wo66) to 22 percent (Mi72). Thus, while the relationship
appears present, the quantitative relationship is not known.
The majority of the cases of reduced head size are observed in those
exposed in the first trimester of gestation, particularly the 6th or 7th
Co 15th weeks of gestation (Mi59, Wo66, Mi72, Wo65, Ta67). Most
recently, it has been shown that reduction in head circumference was a
linear function of dose (Is84). However, the authors noted that the
analysis was based on T65 dosimetry, and the data should be reanalyzed
after completion of the dosimetry reassessment currently in progress.
These findings of reduction in head circumference, with a window of
effect in the same time period of gestation as mental retardation, help
support the observations on mental retardation. Although the exact dose
response functions are still uncertain, data on both types of effects '
have so far been consistent with a linear, no-threshold dose response.
7-74
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(C) Teratologic Effects: Other
Japanese A-bomb, survivors exposed in utero also showed a number of
structural abnormalities and, particularly in those who were . ,
microcephalic, retarded growth (Wo65). No estimate has been made of the
radiation-related incidence or dose-response relationships for these
abnormalities. However, UNSCEAR (UNSCEAR77) made a very tentative
estimate based on animal studies that the increased incidence of
structural abnormalities in animals may be 5E-03 cases per R per live
born, but stated that projection to humans was unwarranted. In any
event, the available human data cannpt show whether the risk estimates
derived from high-dose animal data overestimates the risk in humans.
It should be noted that all of the above estimates.are based on
high-dose-rate, low-LET exposure. In 1977, UNSCEAR also investigated the
dose rate question and stated:
"In conclusion, the majority of the data available for most
species indicate a decrease of the cellular and malformature
effects by lowering the dose rate or by fractionating the dose.
However, deviations from this trend have been well documented in
a few instances and are not inconsistent with the knowledge
about mechanisms of the teratogenic effects. It is therefore
impossible to assume that dose rate and fractional:ion factors
have the same influence on all teratological effects."
(UNSCEAR77).
From this analysis, EPA has concluded that there is risk of-
4E-03 cases of mental retardation per live birth per rad of low-LET
.radiation delivered between weeks 8 and 15 of gestation with no threshold
identified at this time.
No attempt can be made now to estimate total teratogenic effects.
However, it should be noted that the 1977. UNSCEAR estimate from animals
was 5E-03 cases of structural abnormalities per R per live birth (about
the same number per rad of low-LET radiation). This estimate must be
viewed as a minimum one since it is based, to a large extent, on
observation of grossly visible malformations. Differences in criteria
for identifying malformations, have compounded the problem, and questions
of threshold and species differences have made risk projection to humans
unwarranted.
7.6.9 Nonstochastic Effects
Nonstochastic effects, those effects that increase in severity with
increasing dose and have a threshold, have been reviewed in the 1982
UNSCEAR report (UNSCEAR82). In general, acute doses of 10 rads low-LET
radiation and higher are required to induce these effects. It is
possible that some, of the observed effects of in utero exposure are
nonstochastic, e.g., the risk of embryonic loss, estimated to be 10"
7-75
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per R (UNSCEAR77), following radiation exposure soon after
fertilization. However, there are no data to address the question.
Usually, nonstochastic effects are not expected at environmental levels
of radiation exposure.
7.7 Radiation Risk - A Perspective'
^ To provide a perspective on the risk of fatal radiogenic cancers and
the hereditary damage due to radiation, we have calculated the risk from
background radiation to the U.S. population using the risk coefficients
presented in this chapter and the computer codes described in Appendix E.
The risk resulting from background radiation is a useful perspective for
the risks caused by emissions of radionuclides. Unlike cigarette
smoking, auto accidents, and other measures of common risks, the risks
resulting from background radiation are neither voluntary nor the result
of self-induced damage. The risk caused by background radiation is
largely unavoidable; therefore, it is a good benchmark for judging the
estimated risks from radionuclide emissions. Moreover, to the degree
that the estimated risk of radionuclides is biased, the same bias is
present in the risk estimates for background radiation.
The radiation dose equivalent rate from low-LET background radiation
has three major components: cosmic radiation, which averages to about
28 mrad/yr in the U.S.; terrestrial sources, such as radium in soil,
which contributes an average of 26 mrad/yr (NCRP75); and the low-LET dose
resulting from internal emitters. The last differs among organs, to some
extent, but for soft tissues it is about 24 mrad/yr (MCRP75). Other,
minor radiation sources such as fallout from nuclear weapons tests,
naturally-occurring radioactive materials in buildings, and consumer
products, contribute about another 10 mrad for a total low-LET whole-body
dose of about 90 mrad/yr. The lung and bone receive somewhat larger
doses, not included in the 90 mrad/yr estimate, due to high-LET
radiations; see below. Although extremes do occur, the distribution of
this background annual dose to the U.S. population is relatively narrow.
A population weighted analysis indicates that 80 percent of the U.S.
population would receive annual doses that are between 75 mrad/yr and
115 mrad/yr (EPA81).
As outlined in Section 7.2, the BEIR-3 linear, relative risk models
yield, for lifetime exposure to low-LET radiation, an average lifetime
risk of fatal radiogenic cancer of 395 per 106 person-rad. Note that
this average is for a group having the age- and sex-specific mortality
rates of the 1970 U.S. population. We can use this risk estimate to
calculate the average lifetime risk due to low-LET background radiation
as follows. The average duration of exposure in this group is 70.7 yr,
and at 9E-02 rad/yr, the average lifetime dose is 6.36 rads. The risk of
fatal cancer per person in this group is:
395 fatalities
6
10 person-rad
x 6.36 rem = 2.5 x 10
-3
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or about 0.25 percent of all deaths. The vital statistics we use in our
radiation risk analyses indicate that the probability of dying from
cancer in the United States from all causes is about 0.16, i.e..,
16 percent. Thus, the 0.25 percent result for the BEIR-3 linear dose
response model indicates that about 1.6 percent of all U.S. cancer is due
to low-LET background radiation. The BEIR-3 linear-quadratic model
indicates 'that about 0.1 percent of all deaths are due to low-LET
background radiation or about 0.6 percent of all cancer deaths.
Table 7-6 indicates a risk of 466 fatalities per 10b organ rad for
alpha emitters in lung tissue. UNSCEAR estimated in "normal" areas the
absorbed dose from alpha emitters, other than radon decay products, in
the lungs would be about 0.51 mrad (UNSCEAR77). The individual lifetime
cancer risk from this exposure is:
-4 • • •
395 460 fatalities 5.1 x 10 rad ' " -5
x —; x — x 70.7 yr = 2.4 x 10 .
280 10 organ rad
yr
This is about 1/100 of the risk due to low-LET background radiation
calculated by means of the BEIR-3 linear model. .
The 1982 UNSCEAR report indicates that the average annual dose to
the endosteal surfaces of bone due to naturally occurring, high-LET alpha
radiation is about 6 mrad/yr or, based on a quality factor of 20,
120 mrem/yr (UNSCEAR82). Table 7-6 indicates that the individual
lifetime risk of fatal bone cancer due to this portion of the naturally
occurring radiation background is:
280
19 cases
10 person-rad
0.006 rad
year
x 70.7 years = 1.1 x 10
—5
The exposure due to naturally occurring background radon-222 progeny
in the indoor environment is not well known. The 1982 UNSCEAR report
lists for the U.S. an indoor concentration of about 0.004 working levels
(15 Bq/m3) (UNSCEAR82). This estimate is not based on a national
survey and is known to be exceeded by as much as a factor of 10 or more
in some houses. However, as pointed out in UNSCEAR82, the national
collective exposure may not be too dependent on exceptions to the mean
concentration. The UNSCEAR estimate for the U.S. now appears low (Ne86);
the average residential exposure is probably 0.2-0.3 WLM/yr (in standard
exposure units).
Assuming 0.25 WLM/yr is a reasonable estimate for indoor exposure to
radon-222 progeny in the U.S., the mean lifetime exposure, indoors, is
about 18 WLM. Based on the geometric me'an lifetime risk coefficient from
Section 7.4.4, 460 cases/10^ WLM, a lifetime risk of 0.83 percent is
estimated. For comparison, roughly 5 percent of all deaths in 1980 were
7-77
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due to lung cancer. Based on these assumptions, therefore, about one out
of six lung cancer deaths may be attributable to background radon
exposure. This would correspond to about 4 percent of all cancer deaths
and 0.8 percent of all deaths. We note that this is comparable to the
1 percent cancer fatality 'incidence estimated above for low-LET
background radiation. The reader is cautioned, however, that this risk
estimate'only applies to the U.S. population taken as a whole, i.e.,.men
and women, smokers and nonsmokers. While we believe it is a reasonable
estimate for the U.S. 1980 population in which the vast majority of the
lung cancer mortality occurred in male smokers, we do not believe this
risk estimate can be applied indiscriminately to women or nonsmokers. As
noted in Section 7.4, the risk to these groups may not be comparable.
The spontaneous incidence of serious congenital and genetic
abnormalities has been estimated to be about 105,000 per 10^ live
births, about 10.5 percent of live births (NAS80, UNSCEAR82). The
low-LET background radiation dose of about 90 mrad/year in soft tissue
results in a genetically significant dose of 2.7 rads during the 30-year
reproductive generation. Since this dose would have occurred in a large
number of generations, the genetic effects of the radiation exposure are
thought to be an equilibrium level of expression. Since genetic risk
estimates vary by a factor of 20 or more, EPA uses a log mean of this
range to obtain an average value for estimating genetic risk. Based on
this average value, the background radiation causes 700 to 1,000 genetic
effects per 10^ live births, depending on whether or not the oocyte is
as sensitive to radiation as the spermatogonia (see Section 7.6). This
result indicates that about 0.67 to 0.95 percent of the current
spontaneous incidence of serious congenital and genetic abnormalities may
be due to the low-LET background radiation.
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REFERENCES
Ar81
Ar54
Au67
Au77
Ba73
Ba81
Be78
Bo82
Bu81
Ch81
Archer, V.E., Health Concerns in Uranium Mining and Milling,
J. Occup. Med., 23, 502-505, 1981.
Arey, L.B., Developmental Anatomy, 6th ed., W.B. Saunders,
Philadelphia, 1954.
Auxier, J.A., Cheka, J.S., Haywood, F.F., Jones, T.D. and
J.H. Thorngate, Free-Field Radiation Dose Distributions from
the Hiroshima and Nagasaki Bombings, Health Phys.
12(3):425-429, 1967.
Auxier, J.A., Ichiban - Radiation Dosimetry for the Survivors
of the Bombings of Hiroshima and Nagasaki, TID 27080,
Technical Information Center, Energy Research and Development
Administration, National Technical Information Service,
Springfield, Virginia, 1977.
Baum, J.W., Population Heterogeneity Hypothesis on Radiation
Induced Cancer, Health Phys., ^5(1):97-104, 1973.
Baverstock, K.F., Papworth, D. and J. Vennart, Risks of
Radiation at Low Dose Rates, Lancet, 430-433, Feb. 21, 1981.
Beebe, G.W., Kato, H. and C.E. Land, Studies of the Mortality
of A-bomb Survivors, 6: Mortality and Radiation Dose,
1950-74, Rad. Res., 75, 138-201 (RERF TR 1-77, Life Study
Report 8), 1978.
Bond, V.P. and J.W. Thiessen, Reevaluations of Dosimetric
Factors, Hiroshima and Nagasaki, DOE Symposium Series 55,
CONF-810928, Technical Information Center, U.S. Department of
Energy, Washington, D.C., 1982.
Bunger, B., Cook, J.R. and M.K. Barrick, Life Table
Methodology for Evaluating Radiation Risk: An Application
Based on Occupational Exposure, Health Phys., 40(4):439-455.
Chameaud, J., Perraud, R., Chretien, J., Masse, R. and J.
Lafuma, Contribution of Animal Experimentation to the
Interpretation of Human Epidemiological Data, in: Proc. Int.
Conf. on Hazards in Mining: Control, Measurement, and
Medical Aspects, October 4-9, 1981, Golden, Colorado, pp.
228-235, edited by Manual Gomez, Society of Mining Engineers,
New York, 1981.
7-79
-------
Ch83
Co78
Da75
Do73
Do79
Do81
Do83a
Do83b
t -r
Do84a
Bo84b
EL77
Charles, M.E. , Lindop, P.J. and A.J. Mill, A Pragmatic
Evaluation of the Repercussions for Radiological Protection
of the Recent Revisions in Japanese A-bomb Dosimetry, IAEA
SM-266/52, Proceedings, International Symposium on the
Biological Effects of Low-Level Radiation with Special Regard
to Stochastic and Non-stochastic Effects, Venice, IAEA,
Vienna, April 11-15, 1983.
Cook, J.R., Bunger, B.M. and M.K. Barrick, A Computer Code
for Cohort Analysis of Increased Risks of Death (CAIRD), ORP
Technical Report 520/4-78-012, U.S. Environmental Protection
Agency, Washington, B.C., 1978.
Davies, R.B. and B. Hulton, The Effects of Errors in the
Independent Variables in a Linear.Regression, Biometrika,
£2:383-391, 1975.
Bobbing, J. and J. Sands, Quantitative Growth and Bevelopment
of the Human Brain. Arch. Bis. Child., 48:757-767 (1973).
Bobbing, J. and J. Sands, Comparative Aspects of the Brain
Growth Spurt, Early Human Bev. , 3_: 109-126 (1979).
Bobbing, J. The later development of the brain and its
vulnerability, pp. 744-758, in: Scientific Foundations of
Pediatrics, 2nd edition, J.A. Bavis and J. Bobbing, editors,
William Heinemann Medical Books Ltd., London, 1981.
Bobson, R.L. and J.S. Felton, Female Germ Cell Loss from
Radiation and Chemical Exposures, Amer. J. Ind. Med., 4,
175-190, 1983.
Bobson, R.L., Straume, J., Felton, J.S. and T.C. Kwan,
Mechanism of Radiation and Chemical Oocyte Killing in Mice
and Possible Implications for Genetic Risk Estimation
[abstract], Environ. Mutagen. , _5, 498-499, 1983.
Bobson, R.L. and T. Straume, Mutagenesis in Primordial Mouse
Oocytes Could Be Masked by Cell Killing: Monte Carlo
Analysis, Environ. Mutagen. £,393, (1984) [Abstract].
Bobson, R.L., Kwan T.C. and T. Straume, Tritium Effects on
Germ Cells and Fertility, pp. 285-298, in Radiation
Protection European Seminar on the Risks from Tritium
Exposure, EUR9065en, Commission of the European Communities,
1984.
Ellett, W.H. and A.C.B. Richardson, Estimates of the Cancer
Risk Bue to Nuclear Electric Power Generation, pp. 511-527,
in Origins of Human Cancer, Book A., H. H. Hiatt et al.,
eds., Cold Spring Harbor Laboratory, 1977.
7-80
-------
E179
EPA78
EPA79
EPA81
EPA82
EPA83a
EPA83b
EPA84
Ev79
FRC67
ELlett W.H. and N.S. Nelson, Environmental Hazards from Radon
Daughter Radiation, pp. 114-148, in: Conference/Workshop on
Lung Cancer Epidemiology and Industrial Applications of
Sputum Cytology, Colorado School of Mines Press, Golden,
Colorado, 1979.
U.S. Environmental Protection Agency, Response to Comments:
Guidance on Dose Limits for Persons Exposed to Transuranium
Elements in the General Environment, EPA Report 520/4-78-010,
Office of Radiation Programs, Washington, D.C., 1978.
U.S. Environmental Protection Agency, Indoor Radiation
Exposure Due to Radium-226 in Florida Phosphate Lands, EPA
Report 520/4-78-013, Office of Radiation Programs,
Washington, D.C., revised printing, July 1979.
U.S. Environmental Protection Agency, Population Exposure to
External Natural Radiation Background in the United States,
Technical Note ORP/SEPD-80-12, Office of Radiation Programs,
Washington, D.C., 1981.
U.S. Environmental Protection Agency, Final Environmental
Impact Statement for Remedial Action Standards for Inactive
Uranium Processing Sites (40 CFR 192), Volume I, EPA Report
520/4-82-013-1, Office of Radiation Programs, Washington,
D.C., 1982.
U.S. Environmental Protection Agency, Draft Background
Information Document, Proposed Standards for Radionuclides,
EPA Report 520/1-83-001, Office of Radiation Programs,
Washington, B.C.,.1983.
U.S. Environmental Protection Agency, Final Environmental
Impact Statement for Standards for the Control of Byproduct
Materials from Uranium Ore Processing (40 CFR 192), Volume I,
EPA Report 520/1-83-008-1, Office of Radiation Programs,
Washington, D.C., 1983.
U.S. Environmental Protection Agency, Radionuclides
Background Information Document for Final Rules, Volume I,
EPA Report 520/1-84-022-1, Washington, D.C., October 1984.
Evans, H.J., Buckton, K.E., Hamilton, G.E., et al.,
Radiation-induced Chromosome Aberrations in Nuclear Dockyard
Workers, Nature, 277, 531-534, 1979.
Federal Radiation Council, Radiation Guidance for Federal
Agencies, Memorandum for the President, July 21, 1967, Fed.
Reg., 3_2, 1183-84, August .1, 1967.
7-81
-------
Ga82 Garriott, M.L. and D. Grahn, Neutron and Gamma-Ray Effects
Measured by the Micronucleus Test, Mut. Res. Let., 105,
157-162, 1982,
Gi84 Gilbert, E.S., Some Effects of Random Dose Measurements
Errors on Analyses of Atomic Bomb Survivor Data, Rad. Res.,
.98, 591-605, 1984.
Gi85 Gilbert, E.S., Late Somatic Effects, in: Health Effects Model
for Nuclear Power Plant Accident Consequence Analysis by J.S.
Evans, D.W. Cooper, and D.W. Moeller, NUREG/CR-4214, U.S.
Nuclear Regulatory Commission, 1985. ..
Go29 Goldstein, L. and D.P. Murphy, Etiology of Ill-health of
Children Born After Maternal Pelvic.Irradiation: II,
Defective Children Bprn After Post Conception Pelvic
Irradiation, Amer. J. Roentgenol. Rad* Ther., 22, 322-331,
1929.
Go80 Goodhead, D.T., Models of Radiation Interaction and
Mutagenesis, pp. 231-247, in Radiation Biology in Cancer
Research, R.E, Meyn and H. R. Withers, eds., Raven, New York,
1980.
Go82 Goodhead, D.T., An Assessment of the Role of Microdosimetry
in Radiobiology, Rad. Res., j?l_, 45-76, 1982.
Gr83a Grahn, D., et al., Interpretation of Cytogenetic Damage
Induced in the Germ Line of Male Mice Exposed for Over 1 Year
to 239pu Alpha Particles,.Fission Neutrons, or 6°Co Gamma
Rays, Rad. Res., 9j>, 566-583, 1983.
Gr83b Grahn, D., Genetic Risks Associated with Radiation Exposures
During Space Flight, Adv. Space Res., _3(8), 161-170, 1983.
Gu77a Gustavson, K.H, Hagberg, B. , Hagberg, G. ,and K. Sars, Severe
. Mental Retardation in a Swedish County, I, Epidemiology,
Gestational Age, Birth Weight and Associated CNS Handicaps in
Children Born 1959-70, Acta Paediatr. Scand., ^6, 373-379,
1977. .
Gu77b Gustavson, K.-H., Hagberg, B.:, Hagberg, G. and K. Sars,
Severe Mental. Retardation in a Swedish County, II. Etiologic
and Pathogenetic Aspects of Children Born 1959-70,
Neuropadiatrie, J3:293-304 (1977).
HaSla Hagberg, B., Hagberg, G., Lewerth, A. and U. Lindberg, Mild
Mental Retardation in Swedish School Children, 1. Prevalence,
Acta Paediatr. Scand., 70, 441-444, 1981.
7-82
-------
HaSlb Hagberg, B., Hagberg, G. , Lewerth, A. and U. Lindberg, Mild
Mental Retardation in Swedish School Children, II. Etiologic
and Pathogenetic Aspects, Acta Paediatr. Scand., 70:445-452,
1981.
Ha82 Harley, N.H. and B.S. Pasternak, Environmental Radon Daughter
Alpha Dose Factors in a Five-Lobed Human Lung,.Health Phys.,
42, 789-799, 1982.
He83 Herbert, D.E., Model or Metaphor? More Comments on the BEIR
III Report, pp. -357-390, in Epidemiology Applied to Health
Phys., CONF—830101, DE-83014383, NTIS, Springfield,
Virginia, 1983.
Hi53 Hicks, S.P., Developmental Malformations Produced by
Radiation, A Timetable of Their Development, Amer. J.
Roentgenol. Radiat. Thera., 69, 272-293, 1953.
Hi54 Hicks, S.P., The Effects of Ionizing Radiation, Certain
Hormones, and Radiomimetic Drugs on the Developing Nervous
System, J. Cell. Comp. Physiol., 43 (Suppl. 1), 151-178, 1954.
Hi66 Hicks, S.P. and C.J. D'Amato, Effects of Ionizing Radiations
on Mammalian Development, Adv. Teratol., _l, 195-266, 1966.
Ho77 Hofmann, W. and F. Steinhausler, Dose Calculations for
Infants and Youths Due to the Inhalation of Radon and Its
Decay Products in the Normal Environment, in: Proceedings of
the 4th International Congress of the International Radiation
Protection Association, Paris, _2, 497-500, 1977..
Ho81 Hornung, R. W. and S. Samuels, Survivorship Models for Lung
Cancer Mortality in Uranium Miners - Is Cumulative Dose an
Appropriate Measure of Exposure?, in: Proc. Int. Conf. on
Hazards in Mining: Control, Measurement, and Medical
Aspects, October 4-9, 1981, Golden, Colorado, 363-368, edited
by Manuel Gomez, Society of Mining Engineers, New York, 1981.
Ho84 Howe, G.R., Epidemiology of Radiogenic Breast Cancer, in:
Radiation Carcinogenesis: Epidemiology and Biological
Significance, 119-129, edited by J.D. Boice, Jr. and J.F.
Fraumeni, Jr., Raven Press, New York, 1984.
Ho86 Howe, G.R. , Nair, R.C. , Newcotab, H.B., Miller, A.B. and J.D.
Abbatt,' Lung Cancer Mortality (1950-1980) in Relation to
Radon Daughter Exposure in a Cohort of Workers at the
Eldorado Beaver Lodge Uranium Mine, JNCI, 77, 357-362, 1986.
HWC73 Health and Welfare Canada, The Testing of Chemicals for
Carcinogenicity, Mutagenicity'and Teratogenicity, Health
Protection Branch, HWC, Ottawa, 1973.
7-83
-------
ICRP75
SCRP77
ICRP79
ICRP80
ICRP81
Is84
Ja80
Ja70
Ja81
Ka73
International Commission on Radiological Protection,
Committee II on Permissible Dose for Internal Radiation, Task
Group on Reference Man, ICRP Publ. 23, Pergamon Press, 1975.
International Commission on Radiological Protection,
Recommendations, of the International Commission on
Radiological Protection, ICRP Publ. 26, Ann. ICRP, 1, (1),
Pergamon Press, 1977.
International Commission on Radiological Protection, Limits
for Intakes of Radionuclides by Workers, ICRP Publication 30,
Part 1, Ann. ICRP, 2 (3/4), Pergamon Press, New York, 1979.
International Commission on Radiological Protection, Effects
of Inhaled Radionuclides, ICRP Publication 31, Pergamon
Press, 1980.
International Commission on Radiological Protection, Limits
for Intakes of Radionuclides by Workers, ICRP Publication 32,
Part 3, Ann. ICRP, 6 (2/3), Pergamon Press, 1981.
Ishimaru, T. , Nakashima, E. and S. Kawamoto, Relationship of
Height, Body Weight, Head Circumference and Chest
Circumference to Gamma and Neutron Doses.Among In Utero
Exposed Children, Hiroshima and Nagasaki. Technical Report
RERF TR 19-84, Radiation Effects Research Foundation,
Hirsohima, 1984.
Jacobi, W. and K. Eisfeld, Dose to Tissue and Effective.Dose
Equivalent by Inhalation of Radon-222 and Radon-220 and Their
Short-Lived Daughters, GFS Report S-626, Gesellschaft fuer
Strahlen 'und Unweltforschung mbH, Munich, 1980.
Jacobsen, L., Radiation Induced Fetal Damage, Adv. Teratol.,
4, 95-124, 1970.
James, A. C. et al., Respiratory Tract Dosimetry of Radon and
Thoron Daughters: The State-of-the-Art and Implications for
Epidemiology and Radiobiology, in: Proc. Int. Conf. on
Hazards in Mining: Control, Measurement, and Medical
Aspects, October 4-9, 1981, Golden, Colorado, 42-54, edited
by Manuel Gomez, Society of Mining Engineers, New York, 1981.
Kato, H., Late Effects in Children Exposed to the Atomic Bomb
While In Utero, Technical Report 18-73, Atomic Bomb Casualty
Commission, Hiroshima, 1973.
7-84
-------
Ka82
Ke72
KeSla
KeSlb
Ki62
La78
La80
La83
Le62
Lo81
Ma 5 9
Kato, H. and W.J. Schull, Studies of the Mortality of A-bomb
Survivors, 7. Mortality, 1950-1978: Part I, Cancer
Mortality, Rad. Research 90, 395-432, 1982, (Also published
by the Radiation Effect Research Foundation as: RERF TR
12-80, Life Span Study Report 9, Part 1.)
Kellerer, A.M. and H.M. Rossi, The Theory of Dual Radiation
Action, Curr. Topics Rad., Res. Quart., J5, 85-158, 1972.
Kerr, G.D., Review of Dosimetry for the Atomic Bomb
Survivors, in: Proceedings of the Fourth Symposium on
Neutron Dosimetry, Gessellschaft fur Strahlen- und
Umweltforschung, Munich-Neuherberg, Federal Republic of
Germany, June 1-5, l_, 501, Office for Official Publications
of the European Communities, Luxemburg, 1981.
Kerr, G.D., Findings of a Recent ORNL Review of Dosimetry for
the Japanese Atomic Bomb Survivors, ORNL/TM-8078, Oak Ridge
National Laboratory, Oak Ridge, Tennessee, 1981.
King, R.C., Genetics, Oxford University Press, New York, 1962.
Land, C.E. and J.E. Norman, Latent Periods of Radiogenic
Cancers Occurring Among Japanese A-bomb Survivors, in: Late
Biological Effects of Ionizing Radiation, I_, 29-47, IAEA,
Vienna, 1978.
Land, C.E., Boice, J.D., Shore, R.E., Norman, J.E. and M.
Tokunaga, et al., Breast Cancer Risk from Low-Dose Exposures
to Ionizing Radiation: Results of Parallel Analysis of Three
Exposed Populations of Women, J. Nat(l. Cane. Inst., 65,
353-376, 1980.
Land, C.E. and D.A. Pierce, Some Statistical Considerations
Related to the Estimation of Cancer Risk Following Exposure
to Ionizing Radiation, pp. 67-89, in Epidemiology Applied to
Health Phys., CONF-830101, DE83014383, NTIS, Springfield,
Virginia, 1983.
Lea, D.E., Actions of Radiations on Living Cells, 2nd
edition, Cambridge University Press, 1962.
Loewe, W.E. and E. Mendelsohn, Revised Dose Estimates at
Hiroshima and Nagasaki, Health Phys., 41, 663-666, 1981.
Mandansky, A., The Fitting of Straight Lines When Both
Variables Are Subject to Error, J. Amer. Statis. Assoc., 54,
173-205, 1959.
7-85
-------
Ma83
Me 7 8
Mi78
Mi59
Mi72
Mo67
Mo 79
Mu29
Mu30
Mu83
NAS72
Mays, C»W. and H. Spiess, Epidemiological Studies in German
Patients Injected with Ra-224, pp. 159-266, in: Epidemiology
Applied to Health Physics, CONF-830101,, DE-83014383, NTIS,
Springfield, Virginia, 1983.
McDowell, E.M., McLaughlin, J.S., Merenyi, D.K., Kieffer,
R.F., Harris, C.C. and B.F. Trump, The Respiratory Epithelium
V. Histogenesis of Lung Carcinomas in Humans, J. Natl. Cancer
Inst., 61, 587-606, 1978.
Michel C. and H. Fritz-Niggli, Radiation-Induced
Developmental Anomalies in Mammalian Embryos by Low Doses and
Interaction with Drugs, Stress and Genetic Factors,
pp. 399-408, in: Late Biological Effects of Ionizing
Radiation, Vol. II, IAEA, Vienna, 1978.
Miller, R.W., Delayed Effects Occurring Within the First
Decade After Exposure of Young Individuals to the Hiroshima
Atomic Bomb, Technical Report 32-59, Atomic Bomb Casualty
Commission, Hiroshima, 1959.
Miller, R.W. and W.J. Blot, Small Head Size Following In
Utero Exposure to Atomic-Radiation, Hiroshima and Nagasaki,
Technical Report 35-72, Atomic Bomb Casualty Commission,
Hiroshima, 1972.
Morgan, K.Z. and J.E. Turner, Principles of Radiation
Protection, John Wiley and Sons, Inc., New York, 1967.
Mole, R.H., Carcinogenesis by Thorotrast and Other Sources of
Irradiation, Especially Other Alpha-Emitters, Environ. Res.,
1.8, 192-215, 1979.
Murphy, D.P., The Outcome of 625 Pregnancies in Women Subject
to Pelvic Radium or Roentgen Irradiation, Amer. J. Obstet.
Gyn. , JJ3, 179-187, 1929.
Murphy, D.P., and M. DeRenyi, Postconception Pelvic
Irradiation of the Albino Rat (Mus Norvegieus): Its Effects
Upon the Offspring, Surg. Gynecol. Obstet., 50, 861-863, 1930.
Muller, J., Wheeler, W.C., Gentleman, J.F., Suranyi, G. and
R.A. Kusiak, Study of Mortality of Ontario Miners, 1955-1977,
Part I, Ontario Ministry of Labor, Ontario, May 1983.
National Academy of Sciences - National Research Council, The
Effects on Populations of Exposures to Low Levels of Ionizing
Radiation, Report of the Committee on the Biological Effects
of Ionizing Radiations (BEIR Report), Washington, D.C., 1972.
7-86
-------
NAS80 National Academy of Sciences - National Research Council, The
Effects on Populations of Exposure to Low Levels of Ionizing
Radiation, Committee on .the Biplogical Effects of Ionizing
Radiation, Washington, D.C., 1980.
NASA73 National Aeronautics and Space Administration,
Bioastronautics Data Book, NASASP-3006, 2nd Edition, edited
by J. R. Parker and V. R. West, Washington, B.C., 1973.
NCHS73 National Center for Health Statistics, Public Use Tape, Vital
Statistics - Mortality Cause of Death Summary - 1970,
PB80-133333, Washington, D.C., 1973.
NCHS75 National Center for Health Statistics, U.S. Decennial Life
Tables for 1969-71, 1(1), DREW Publication No. (HRA) 75-1150,
U.S. Public Health Service, Rockville, Maryland, 1975.
NCRP75 National Council on Radiation Protection and Measurement,
Natural Background Radiation in the United States, NCRP
Report No. 45, Washington, D.C., 1975.
NCRP77 National Council on Radiation Protection and Measurements,
Protection of the Thyroid Gland in the Event of Releases of
Radioiodine, Report No. 55, Washington, D.C., 1977.
NCRP80 National Council on Radiation Protection and Measurements,
Influence of Dose and Its Distribution in Time on
Dose-Response Relationships for Low-LET Radiation, NCRP
Report No. 64, Washington, D.C., 1980.
NCRP84 National Council on Radiation Protection and Measurements,
Evaluation of Occupational and Environmental Exposures to
Radon and Recommendations, NCRP Report No. 78, Washington,
D.C., 1984.
NCRP85 National Council on Radiation Protection and Measurements,
Induction of Thyroid Cancer by Ionizing Radiation, NCRP
Report No. 80, Washington, D.C., 1985.
Ne56 Neel, J.V. and W.J. Schull, The Effect of Exposure to the
Atomic Bombs on Pregnancy Termination in Hiroshima and
Nagasaki, National Academy of Sciences, Publ. 461,
Washington, D.C., 1956.
Ne70 Nelson, K.B. and J. Deutschberger, Head Size at One Year as a
Predictor of Four-Year I.Q. , Develop. Med. Child Neurol. , 1^2,
487-495, 1970.
Ne86 Nero, A.V., Schwehr, M.B., Nazaroff, W.W. and K.L. Revzan,
Distribution of Airborne Radon-222 Concentrations in U.S.
Homes, Science, 234, 992-997, 1986.
7-87
-------
QRNL84 Oak Ridge National Laboratory, Age Dependent Estimation of
Radiation Dose, [in press], 1984.
Of80 Oftedal, P. and A.G. Searle, An Overall Genetic Risk
Assessment for Radiological Protection Purposes, J. Med.
Genetics, _L7, 15-20, 1980.
Oh65 Ohzu, E., Effects of Low-Dose X-Irradiation on Early Mouse
Embryos, Rad. Res. 26, 107-113, 1965.
Ot83 Otake, M., and W.H. Schull, Mental Retardation in Children
Exposed In Utero to the Atomic Bombs: A Reassessment,
Technical Report RERFTR 1-83, Radiation Effects Research
Foundation, Hiroshima, 1983.
Ot84 Otake, M., and W.J. Schull,' In Utero Exposure to A-bomb
Radiation and Mental Retardation: A Reassessment, Brit. J.
Radiol. , .5_7, 409-414, 1984.
P152 Plummer, G.W., Anomalies Occurring in Children Exposed In
Utero to the Atomic Bomb in Hiroshima, Pediat., 10, 687-692,
1952.
P°78 Pohl-Ruling, J., Fischer, P. and E. Pohl, The Low-Level Shape
of Dose Response for Chromosome Aberration, pp. 315-326, in:
Late Biological Effects of Ionizing Radiation, Volume II,
International Atomic Energy Agency, Vienna, 1978.
Pr83 Prentice, R.L. , Yoshimoto, Y., and M.'W. Mason, Relationship
of Cigarette Smoking and Radiation Exposure to Cancer
Mortality in Hiroshima and Nagasaki, J. Nat. Cancer Inst.,
70, 611-622, 1983.
Ra84 Radford, E.P., and K.G. St. Cl. Renard, Lung Cancer in
Swedish Iron Miners Exposed to Low Doses of Radon Daughters,
N. Engl. J. Med., 310, 1485-1494, 1984.
RERF78 Radiation Effects Research Foundation. Radiation Effects
Research Foundation, 1 April 1975 - 31 March 1978. RERF
Report 75-78, Hiroshima, 1978.
RERF83 Radiation Effects Research Foundation, Reassessment of Atomic
Bomb Radiation Dosimetry in Hiroshima and Nagasaki, Proc. of
the U.S.-Japan Joint Workshop, Nagasaki, Japan, Feb. 16-17,
1982, Radiation Effects Research Foundation, Hiroshima, 730,
Japan, 1983.
RERF84 Radiation Effects Research Foundation, Second U.S.-Japan
Joint Workshop for Reassessment of Atomic Bomb Radiation
Dosimetry in Hiroshima and Nagasaki, Radiation Effects
Research Foundation, Hiroshima, 730, Japan, 1984.
7-88
-------
Ro78
Ro78
Ru50
Ru53
Ru54
Ru56
Ru57
Ru58
Ru70
Ru71
Sa82
Sc81
Sc84
Rossi, H.H. and C.W. Mays, Leukemia Risk from Neutrons,
Health Phys., 3_4, 353-360. 1978.
Rowland, R.E., Stehney, A.F. and H.F. Lucas, Dose Response
Relationships for Female Radium Dial Workers, Rad. Res., 76,
368-383, 1978.
Russell, L.B., X-ray Induced Developmental Abnormalities ,in
the Mouse and Their Use in the Analysis of Embryological
Patterns, I. External and Gross Visceral Changes, J. Exper.
Zool., 114, 545-602, 1950.
Rugh, R., Vertebrate Radiobiology: Embryology, Ann. Rev.
Nucl. Sci., 3_, 271-302, 1953.
Russell, L.B. and W.L. Russell, An Analysis of the Changing
Radiation Response of the Developing Mouse Embryo, J. Cell.
Comp. Physiol., 43 (Suppl. 1), 103-149, 1954.
Russell, L.B., X-Ray Induced Developmental Abnormalities in
the Mouse and Their Use in the Analysis of Embryological
Patterns, II. Abnormalities of the Veretebral Column and
Thorax, J. Exper. Zool., 131, 329-390, 1956.
Russell, L.B., Effects of Low Doses of X-rays on Embryonic -
Development in the Mouse, Proc. Soc. Exptl. Biol. Med., 95,
174-178, 1957.
Russell, W.L., Russell, L.B. and E,M. Kelly, Radiation Dose
Rate and Mutation Frequency, Science, 128:1546-1550, 1958.
Rugh, R., The Effects of Ionizing Radiation on the Developing
Embryo and Fetus, Seminar Paper No. 007, Bureau of
Radiological Health Seminar Program, U.S. Public Health
Service, Washington, D.C., 1970.
Rugh, R., X-ray Induced Teratogenesis in the Mouse and Its
Possible Significance to Man, Radiol., 99, 433-443, 1971.
Satoh, C. et al., Genetic Effects of Atomic Bombs, in: Human
Genetics, Part A: The Unfolding Genome, A. R. Liss, Inc.,
New York, 267-276, 1982.
Schull, W.J., Otake, M. and J.V. Neel, Genetic Effects of the
Atomic Bombs: A Reappraisal, Science, 213, 1220-1227, 1981.
Schull, W.J. and J.K. Bailey, Critical Assessment of Genetic
Effects of Ionizing Radiation on Pre- and Postnatal
Development, pp. 325-398, in: Issues and Reviews in
Teratology, Volume 2, H. Kal'ter, editor. Plenum Press, New
York, 1984.
7-89
-------
Se69
Sra78
Sp56
Sp83
St21
St81
St84
Ta67
Th82
To80
To84
Senyszyn, J.J. and R. Rugh, Hydrocephaly Following Fetal
X-Irradiation, Radiol. , 93^, 625-634, 1969.
Smith, P.G. and R. Doll, Radiation-Induced Cancers in1
Patients with Ankylosing Spondylitis Following a Single
Course of X-ray Treatment, in: Proc. of the IAEA Symposium,
Late Biological Effects of Ionizing Radiation, 1,, 205-214,
IAEA, Vienna, March 1978. ~
Spector, W.S., editor, Handbook of Biological Data,
Table 314, Energy Cost, Work: Man, W. B. Sanders Co.,
Philadelphia, 1956.
Spiers, F.W., Lucas, H.F., Rundo, J. and G.A. Anast, Leukemia
Incidence in the U.S. Dial Workers, in: Conference Proc. on
Radiobiology of Radium .and; J:he Actinides in Man,
October 11-16, 1981, He,a 1th Phys. , 44(Suppl. l):65-72, 1983.
Stettner, E., Ein weiterer Fall einer Schadingung einer
raenschichen Frucht durch Roentgen Bestrahlung., Jb.
Kinderheilk. Phys. Erzieh., 95, 43-51, 1921.
Straume, T. and R. L. Dobson, Implications of New Hiroshima
and Nagasaki Dose Estimates: Cancer Risks and Neutron RBE
Health Phys., 4_U4) : 666-671, 1981.
Stein, Z.A. and M.W. Susser, The Epidemiology of Mental
Retardation, in: Epidemiology of Pediatric Neurology, B.
Schoenberg, editor, Marcel Dekker, Inc., New York, [in
press], 1984.
Tabuchi, A., Hirai, T., Nakagawa, S., Shimada, K. and J.
Fugito, Clinical Findings on In Utero Exposed Microcephalic
Children, Technical Report 28^-67, Atomic Bomb Casualty
Commission, Hiroshima, 1967.
Thomas, D.C. and K.G. McNeill, Risk Estimates for the Health
Effects of Alpha Radiation, Report INFO-0081. Atomic Energy
Control Board, Ottawa, 1982.
Tobias, C. A. et al., The Repair-Misrepair Model,
pp. 195-230, in: R. E. Meyn and H. R. Withers, eds. , Raven,
New York, 1980.
Tokunaga, M., Land, C.E., Yamamoto, T., Asano, M., Takioka,
S., Ezaki, E. and I. Nishimari, Incidences of Female Breast
Cancer Among Atomic Bomb Survivors, Hiroshima and Nagasaki,
1950-1980, RERF TR 15-84, Radiation Effects Research
Foundation, Hiroshima, 1984.
7-90
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U182 Ullrich, R.L., Lung Tumor Induction in Mice: Neutron RBE at
Low Doses, NTIS-DE 82009642, National Technical Information
Service, Springfield, Virginia, 1982.
UNSCEAR58 United Nations, Report of the United Nations Scientific
Committee on the Effects,of Atomic Radiation, Official
.Records: Thirteenth Session, Supplement No. 17 (A/3838),
United Nations, New York, 1958.
UNSCEAR62 United Nations, Report of the United Nations Scientific
Committee on the Effects of Atomic Radiation, Official
Records: Seventeenth Session, Supplement No. 16 (A/5216),
United Nations, New York, 1962.
UNSCEAR66 United Nations, Report of the United Nations Scientific
Committee on the Effects of Atomic Radiation, Official
Records: Twenty-First Session, Supplement No. 14 (A/6314),
United Nations, New York, 1966.
UNSCEAR69 United Nations, Report of the United Nations Scientific
Committee on the Effects of Atomic Radiation, Supplement
No. 13 (A/7613), United Nations, New York, 1969.
UNSCEAR72 United Nations Scientific Committee on the Effects of.Atomic
Radiation, Ionizing Radiation: Levels and Effects,
Volume II: Effects, Report to the General Assembly. Sales
No. E. 72. IX.18., United Nations, New'York, 1972.
UNSCEAR77 United Nations Scientific Committee on the Effects of Atomic
Radiation, Sources and Effects of Ionizing Radiation, Report
to the General Assembly, with Annexes, Sales No. E.77 IX.1.,
United Nations, New York, 1977.
UNSCEAR82 United Nations Scientific Committee on the Effects of Atomic
Radiation, Ionizing Radiation: Sources and Biological
Effects, 1982 Report to the General Assembly, Sales No. E.82.
IX.8, United Nations, New York, 1982.
Up75. Upton, A.C., Physical Carcinogenesis: Radiation—History and
Sources., pp. 387-403, in: Cancer 1, F.F. Becker, editor.
Plenum Press, New York, 1975.
USRPC80 U.S. Radiation Policy Council, Report of the Task Force on
Radon in Structures, USRPC-80-002, Washington, B.C., 1980.
Va80 Van Buul, P.P.W., Dose-response Relationship for X-ray
Induced Reciprocal Translocations in Stem Cell Spermatogonia
of the Rhesus Monkey (Macaca mulatta), Mutat. Res., 73,
363-375, 1980. (Cited in UNSCEAR82.)
7-91
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Vo02
Wa83
Wh83
WHO 7 5
Wi54
Wo 65
Wo 66
Von Frieben, A., Demonstration lines cancroids des rechten
Handruckens das sich nach lang dauernder Einwirkung von
Rontgenstrahlen entwickelt hatte. Fortschr. Geb.
Rontgenstr., 6^:106 (1902) cited in Up75.
Wakabayashi, T., Kato, H., Ikeda, T. and W.J. Schull, Studies
of the Mortality of A-bomb Survivors, Report 7, Part III,
Incidence of Cancer in 1959-78 Based on the Tumor Registry
Nagasaki, Radiat. Res., 93_, 112-142, 1983.
Whittemore, A.S. and A. McMillan, A Lung Cancer Mortality
Among U.S. Uranium Miners: A Reappraisal, Technical Report
No. 68, SIAM Inst. Math. Soc., Stanford University, Stanford
1983.
World Health Organization, International Statistical
Classification of Diseases, Injuries, and Causes of Death,
9th Revision, Geneva, 1975.
Wilson, J.G., Differentiation and the Reaction of Rat Embryos
to Radiation, J. Cell. Comp. Physiol., 43 (Suppl. 1). 11-37,
1954.
Wood, J.W., Johnson, K.G. and Y. Omari, In Utero Exposure to
the Hiroshima Atomic Bomb: Follow-up at Twenty Years,
Technical Report 9-65, Atomic Bomb Casualty Commission,
Hiroshima, 1965.
Wood, J.W., Johnson, K.G., Omari, Y., Kawamoto, S. and R.J.
Keehn, Mental Retardat-ion in Children Exposed In Utero to the
Atomic Bomb—Hiroshima and Nagasaki, Technical Report 10-66,
Atomic Bomb Casualty Commission, Hiroshima, 1966.
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Chapter 8: METHODOLOGY FOR THE ASSESSMENT OF HEALTH IMPACT
8.1 Introduction
A health impact analysis or risk assessment is required when
developing an EPA standard. It is the primary technical submission from
Agency staff to Agency management in the decision-making process. In the
case of the standard for the disposal of LLW, its purpose is to answer
the following important questions: (1) What are the potential health
impacts of LLW disposal? (2) What are the possible technical means for
limiting these impacts? and (3) Using various technical means, what
reduction in health impact (benefit) can be achieved? In addition, it is
informative to estimate the health impact resulting from LLW disposal as
it has been conducted in the past, and as it might be conducted! with and
without an EPA standard.
Because the standard must be generally applicable, it will apply
over the entire U.S. and not be specific to certain disposal methods or
management alternatives. Therefore, the health impact assessment
compares combinations of as many as 10 different disposal methods,
3 general hydrogeologic and climatic settings, 24 waste streams!, 4 waste
forms, and a variety of other variables. The assessment must also
evaluate the potential health impacts from all important pathways, using
the Agency's radiation risk methodology. Because of these complexities,
the only feasible means of estimating the health impacts was through the
use of a computer model. To meet these needs, the PRESTO-EPA computer
code was developed jointly by EPA and Oak Ridge National Laboratory
(EPA83). This model, which was completed in 1983, was expanded by EPA
and Rogers and Associates Engineering Company into a family of codes used
to estimate health irapaqt from the disposal of LLW under a variety of
disposal alternatives (Ro84a). TheSe codes are described in more detail
in later sections.
8.2 Health Impact Assessment Modeling; PRESTO-EPA
The evaluation of the potential health impact, consisting of
cumulative population health effects and maximum annual dose to the
critical population group, from shallow-land disposal of LLW is performed
with the computer code PRESTO-EPA (Prediction of Radiation Effects from
Shallow Trench Operations - Environmental Protection Agency). The
PRESTO-EPA code models the transport of radionuclides through
hydrogeologic and atmospheric pathways to the eventual ingestion and
inhalation by or direct exposure of humans. The results from the
environmental transport portion of the model are used in the assessment
of cumulative population health effects (consisting of fatal cancers and
serious genetic effects to a local population for a period of 1,000 years
and to a regional basin population for 10,000 years) or to estimate the
maximum annual dose to a critical population group located close to the
disposal site.
8-1
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Environmental transport of radionuclides away from the LLW disposal
site occurs through the hydrologic pathway, including infiltration,
overflow, surface runoff, and ground-water flow, and through the
atmospheric pathway. Figure 8-1 presents a schematic representation of
these pathways. Radiation dose to humans is estimated for internal
exposure from inhalation and ingestion and for direct exposure from
contaminated air and soil. For a schematic representation of the food
chain and direct exposure pathways, see Figure 8-2. Exposures can be
calculated for onsite intruders who may grow crops with roots into the
waste or build houses over the waste.
It is important to point out that the PRESTO-EPA model was developed
for a specific purpose, to estimate health impacts from various'disposal
methods, as an aid in developing a generally applicable LLW standard.
The model is a relatively simple, generic model, consisting of a
one-dimensional ground-water transport submodel and other submodels of
the compartment type. The model is not designed to be site specific, and
the results should not be interpreted as pertaining to any particular
disposal facility. This analytical approach is used since it is adequate
for comparison of disposal methods; avoids potential errors in the
numerical approach; and is consistent with the quality of the input data
currently available.
A generic model is required because EPA will not be establishing
site-specific standards but generally applicable standards, which will be
implemented on a site-specific basis by me and DOE. The potential
impact of disposing of a wide variety of wastes by different disposal
methods under sharply different hydrogeologic and climatic conditions
dictates use of a broad-based model. For example, three separate
hydrogeological environments are modeled which encompass the majority of
environmental conditions in the U.S. that would be feasible for LLW
disposal.
The model is modular in design to allow substitution or modification
of submodels when analyzing different disposal systems and is
sufficiently flexible to analyze disposal systems within a wide range of
hydrogeologic and climatic conditions. A disposal system includes the
wastes, the disposal method, the site geology, hydrology, and climate,
and all the applicable exposure pathways. In this way, the performance
of the system can or will change if changes are made to any of its parts.
The use of a simple, one-dimensional, generic model has certain
limitations. The model was developed for EPA's comparison of disposal
alternatives and, therefore, was not designed, and should not be used,
for making site-specific estimates of health impact, in simplifying the
model, certain processes had to be omitted. The model assumes that all
inputs, such as population size, demographics, and food consumption,
remain constant over the modeling period. Radioactive daughter in-growth
is not included in the PRESTO-EPA model, although activity due to the '
daughters can be accounted fo.r by direct input of the daughter source
8-2
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PRECIPITATION
Ul
ATMOSPHERIC TRANSPORT
oo
I
U)
RESUSPENSION
DEPOSITION
TRENCH CONTAINING
LOW-LEVEL WASTE
TRANSPORT TO AQUIFER
AT RETARDED VELOCITIES
STREAM
AQUIFER
...MK«.a
-------
CONTAMINATED AIR
DEPOSITION ON CROPS
, DUE TO PLUME DEPLETION
DEPOSITION ON CROP SURFACES DUE TO IRRIGATION
CONTAMINATED V~/
STREAM \y
UPTAKE BY PLANTS
DUE TO
IRRIGATION
a) Crop contamination
00
I
DRINKING
STREAM
WATER
CONTAMINATED^/
STREAM \/
DRINKING
WELL
WATER
CONTAMINATED
CROPS
b) Livestock contamination
CONTAMINATED AIR
CONTAMINATED \~f
STREAM A/
CONTAMINATED AIR
DRINKING
STREAM
WATER
CONTAMINATED
CROPS
CONTAMINATED
LIVESTOCK
CONTAMiriAiED
WELL
DEPOSITION ON
GROUND DUE
TO PLUME
DEPLETION
DIRECT
EXPOSURE
FROM AIR
DIRECT EXPOSURE
FROM GROUND
CONTAMINATED
STREAM
c) Internal exposure to man
DEPOSITION ON GROUND
FROM IRRIGATION
d) External exposure to man
CONTAMINATED
WELL
Figure 8-2. Food Chain and Direct Exposure Pathways
Used in PRESTO-EPA
-------
term. The cumulative health effects beyond 1,000 years are estimated
using conversion factors that are based on the analysis over the first
1,000 years. While these limitations exist, the PRESTO-EPA model is
appropriate for its intended use.
Because PRESTO-EPA was developed specifically for this
standard-setting effort and is a new code, a program of code improvement
and verification was conducted. This program includes:
• Quality assurance audits of all codes;
• Improvements and corrections based on extensive test and
production runs;
• Peer review (PR84);
• Sensitivity testing and analysis;
• Improvements and corrections based on review and use by others,
including national laboratories and universities;
• Review by EPA's Science Advisory Board (SAB85);
• Review and comparison of the geological transport portion of the
code with results obtained by the U.S. Geological Survey (USGS)
simulated for existing LLW disposal sites;
• intercomparison with similar codes, such as those of the NRC; and
• comparison of PRESTO-EPA results with data from actual LLW
disposal sites.
The program to verify the PRESTO-EPA codes is described in more
detail elsewhere (Me85). The sensitivity analysis portion of the program
is described separately (Ba86a). A discussion of the sensitivity
analysis is included in Chapter 11 and a discussion of model
uncertainties in Chapter 12.
The original PRESTO-EPA code was developed to estimate cumulative
population health effects from shallow-land burial of regulated LLW.
Additional information was required, however, for the standard
development, including: maximum annual doses to a critical population
group located close by the disposal site, cumulative population health
effects and maximum annual doses from deep disposal options, and
cumulative population health effects and maximum annual doses from less
restrictive disposal of BRC wastes. In order to provide these estimates,
the original PRESTO-EPA code was expanded into a family of codes that
includes:
8-5
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PRESTO-EPA-CPG
PRESTO-EPA-BRC
PATHRAE-EPA
PRESTO-EPA-POP Estimates cumulative population health effects to local
and regional basin populations from land disposal of LLW
by shallow methods; long-term analyses are modeled
(generally 10,000 years);
PRBSTO-EPA-DEEP Estimates cumulative population health effects to local
and regional basin populations from land disposal of LLW
by deep methods;
Estimates maximum annual dose to a critical population
group from land disposal of LLW by shallow or deep
methods; dose in maximum year is determined;
Estimates cumulative population health effects to local
and regional basin populations from disposal of BRC
wastes by sanitary landfill, municipal dump, and
incineration methods; and
Estimates annual pathway doses to a critical population
group from less restrictive disposal of BRC wastes by
sanitary landfill, municipal dump, and incineration
methods.
These codes and how the Agency uses them have been described in
detail (Hu83a, Ga84, Ro84b). Information on obtaining complete
documentation and-users' manuals for the PRESTO-EPA family of codes
(EPA87a through EPA87g, Me81, Me84) is available from the Agency.
8>3 Health Impact Assessment Methodology r Overview
In this section the basic PRESTO-EPA code is discussed in a general
manner. The specific codes that make up the PRESTO-EPA family are
discussed in more detail in Sections 8.4 and 8.5.
While being a relatively simple computer model, the PRESTO-EPA code
incorporates a number of different submodels that are used to analyze
various pathways and scenarios and to determine health impact. Some
members of the PRESTO-EPA family of codes use a unit response approach
(see Section 8.3.6) and a conversion factor for long-term analyses In
order to reduce computer time (see Section 8.3.5). Numerous input
parameters are required for the PRESTO-EPA codes. These input parameters
and their values are listed in Appendix c.
8-3.1 Infiltration/Leaching
At many sites, water is the most important medium for transport of
radionuclides away from the trench. Whether the transport pathway is
predominantly through the aquifer or by overland flow, the parameter that
generally drives the system is the amount of water entering the trench
via Infiltration through the .trench cap.
8-6
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infiltration is computed using a method by C.Y.. Hung (Hu83b). This
method simulates the infiltration of rainwater through a trench cover by
modeling three separate flow systems: subsurface, overland, and
atmospheric. Normal infiltration rates, calculated by the model, occur
on the intact portions of the trench cap. On the failed portions, the
infiltration also includes all the surface runoff that is diverted into
the trench from the area of the trench cap up-slope from the failure
area. Failure of the trench cap is through erosion or other processes.
Erosion is determined by the model based on input parameters. However,
in most cases, an actual trench cap will fail from such processes as
subsidence, gully formation, or mechanical disturbance. To model these
cases, the failure of the trench cap is based on assumed failure
percentages occurring in user-specified years. The assumed values used
are listed in Appendix C.
Water infiltrating the trench will be contaminated through contact
with the waste in the trench. The amount of contamination is determined
by the selected leaching methodology. Options include the use of
specific distribution coefficients (Kd) for'trash and absorbing waste
forms, or a specified annual release fraction for activated metal and
solidified waste forms. The amount of leaching can be modified by
options such as solidification of the waste, use of high integrity
containers (HIC), or active site maintenance. A listing of the options
that were used for various scenarios is included in Appendix C*
Contaminated water may leave the trench by draining through the
trench bottom or by overflowing. The model estimates the radionuclide
activities and the amount of water leaving the trench on an annual basis.
8.3.2 Transport/Uptake Pathways
The PRESTO-EPA model estimates health impact from the hydrologic,
atmospheric, food chain, and direct exposure pathways (Ba85). These
pathways are described in the following sections. The hydrologic and
atmospheric environmental transport pathways in PRESTO-EPA are shown in
Figures 8-3 and 8-4. • ~
(A) Hvdrologic Pathways
in the ground-water flow model (Hu81), material that leaches from
the trench is transported vertically to the underground aquifer and
horizontally through the aquifer to a well. Transport velocities are
calculated using saturated or unsaturated flow models, depending on
sub-trench hydrological conditions. Radionuclides are transported at
velocities lower than the characteristic flow velocity of the water in
the aquifer. This "retardation" is due to interaction of the
radionuclides with the solid materials in the aquifer.
8-7
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•SEEPAGE-
-OVERFLOW-
SOIL
SURFACE
TRENCH
LEACHING
LEACHING
VERTICAL
SOIL COLUMN
AQUIFER
GROUND-WATER TRANSPORT
SURFACE
WATER BODY
WELL
IRRIGATION
SOIL
IRRIGATION
PLANT UPTAKE
__;
CROPS
DRINKING
INGESTION
A
LIVESTOCK
AND MILK
INGESTION
DRINKING
HUMANS
Figure 8-3. Hydrologic Environmental Transport
Pathways in PRESTO-EPA
8-8
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SURFACE
CONTAMINATION
EXPOSED WASTE
FROM EROSION
SUSPENSION
i ;
DIRECT
EXPOSURE
AIR
INHALATION
HUMANS
INHALATION
DEPOSITION
INGESTION
INGESTION
CROPS
INGESTION
LIVESTOCK,
DAIRY
Figure 8-4.
Atmospheric Environmental Transport
Pathways in PRESTO-EPA
8-9
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™»n a cer*ain Period. the contaminated aquifer water reaches a
well. The activity reaching the well is diluted by the annual volume of
water available at the well and used in the food chain and direct
exposure pathways. In calculating cumulative population health effects
the concentrations of radionuclides in water are averaged over the length
of the simulation period and used in determining exposure to humans, in
calculating the maximum annual dose, the concentration in the maximum
year is used.
Depending upon precipitation, infiltration rate, and hydrogeological
characteristics, contaminated trench water may overflow from the trench
onto the soil surface. When this occurs, radionuclides are added to the
surface inventory of radionuclides from any initial operational
spillage. The radionuclides on the ground surface are made up of two
components, dissolved and adsorbed. The component adsorbed to soil is a
source terra for resuspension and subsequent atmospheric transport. The
dissolved component may enter deep soil layers or nearby surface streams
via overland flow. The radionuclides entering the deep soil layers will
be modeled in the hydrologic pathway. The radionuclides entering the
stream are divided by the annual stream flow and used in the food chain
and direct exposure pathways.
(B) atmospheric Pathways
in determining radionuclide concentrations during atmospheric
transport, PRESTO-EPA includes a simple algorithm suitable for those
sites where the population is concentrated into a single, small
community. Because of the need to model more complex population
distributions, EPA has used an optional externally computed ratio of the
air concentration to source strength (chi/Q). The externally computed
ratio is population weighted and can be used in cases of complex
population distribution (see Section 8.3.8).
The atmospheric dispersion equation involves characterization of the
source strength and calculation of the atmospheric concentration of
radionuclides at the receptor location. The atmospheric emission rate
(source strength) is made up of wind-driven suspension and mechanical
suspension of radionuclides sorbed to the surface soil. NO gaseous
emissions, such as methane, carbon dioxide, or water vapor, are included
in the source term. Radionuclides at the disposal site are deposited on
the soil surface by spillage of wastes during operation and through
overflow from the trench after site closure.
The atmospheric concentration of radionuclides at the receptor
location is calculated using a Gaussian-plume atmospheric dispersion
model (S168). Dispersion parameters are the standard deviation of the
plume concentration in the horizontal and vertical directions. The
radionuclides are transported at a height-independent wind speed to the
receptor location. Plume depletion, effective plume height, and stable
air layers at high altitudes are taken into account. Neutral atmospheric
8-10
-------
stability is generally assumed. The model calculates a radionuclide
concentration in air, averaged over a 22.5 degree sector for use in the
inhalation, ingestion, and direct exposure pathways.
(C) Food Chain *"d Direct Exposure Pathways
Radionuclides in water, either well or stream, may impact humans
through both internal and external radiological exposure. The internal
exposure occurs from drinking contaminated water and from ingesting milk,
beef, and crops contaminated through irrigation or livestock watering.
External exposure results directly from exposure to crops and soil that
have been irrigated with contaminated well or stream water.
Radionuclides in air may also expose humans through both internal and
external pathways. Direct external exposure may result from immersion in
a plume of contaminated air or by exposure to soil surfaces contaminated
by plume deposition. Internal exposures may result from inhalation of
contaminated air or ingestion of food products contaminated by plume
deposition.
To calculate the impact to humans from ingestion, inhalation, and
direct exposure from radionuclides in air and water, the computer code
DARTAB (see Section 8.3.4) is used as a subroutine in PRESTO-EPA.
Radionuclide input data to DARTAB from the air and water pathways must be
in four specific formats:
(1) average concentrations in air (Ci/m3);
(2) average concentrations on ground surface (Ci/m2);
(3) collective average inhalation rates (person-pci/yr); and
(4)
collective average ingestion rates (person-pci/yr)
The mean concentrations of radionuclides in air are calculated for
the period of simulation or for the year of maximum annual dose. Mean
concentrations of radionuclides in air are used to calculate the direct
exposure to humans from immersion in contaminated air.
The concentration of each radionuclide on the ground surface is due
to both atmospheric deposition and irrigation. Atmospheric deposition
results from plume depletion through the atmospheric transport pathway.
Deposition from well and stream water used for irrigation is simulated in
the irrigation subroutine. The concentration of radionuclides on the
ground surface is used to calculate the direct external exposure to
humans from contaminated ground.
The collective exposure from inhalation is calculated by multiplying
the size of the population of interest by the average individual
inhalation rate and by the average concentration of radionuclides in the
air The collective exposure from inhalation is used to calculate organ
8-11
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doses and effective whole-body dose equivalent from breathinq
radionuclide-contaminated air.
The collective ingestion rate includes intake of drinking water,
beef, milk, and crops. Except for drinking water, all of these media may
be contaminated by either atmospheric deposition or by irrigation from a
well and/or the stream. Humans may ingest water directly from either the
well or the stream. The water may also be ingested by cattle or used to
jjrlgate crops. Crops are contaminated through atmospheric deposition
directly onto the plants, and/or by growing in soil contaminated through
atmospheric deposition or irrigation with contaminated water. Once
contaminated, the crops can be ingested by humans or by animals. Humans
ingest beef and milk that have become contaminated from cattle eating and
drinking contaminated forage and/or water. The annual individual
ingestion rate is multiplied by the size of the population and the
average concentration of radionuclides ingested to calculate the annual
collective ingestion rate, in calculating population health effects, the
mean annual collective ingestion rate is assumed to be constant for the
population over the period of simulation, in contrast, for the maximum
annual dose, the average concentration for the year iri which the peak
concentration occurs is used. The annual collective ingestion rate is
used to calculate the internal organ doses and effective whole-body dose
equivalent to humans, through ingestion of contaminated water, milk,
beef, and crops.
8.3.3 Intruder Scenarios
The PRESTO-EPA code can be used to estimate exposures to an
intruder, it is assumed that the intruder would enter the site after
institutional control has ended. The intruder scenarios allow for onsite
farming with crop roots growing directly into the waste or for building
homes with basements dug into the waste.
If the intruder farms the disposal site, it will cause mechanical
disturbance of the trench cap, as well as the possibility of crop roots
growing into the waste. The mechanical disturbance is taken into account
£hrough the atmospheric transport pathway. The crop roots growing into
the waste are considered in the food chain pathway.
Intru , Protection a9ainst inadvertent
intrusion into the disposal areas. EPA feels that this exposure pathway
is probabilistic in nature and that safeguards against inadvertent
intrusion should be carried out on a site-specific basis. For these
reasons, EPA has not included intrusion scenarios in its health impact
assessments .
8-3.4 Health Impact Assessment
The PRESTO-EPA computer code estimates radionuclide concentrations
in ground and surface water, concentrations in the air, rates of
8-12
-------
deposition on the ground, concentrations on the ground, and the amounts
of radionuclides taken into the body via inhalation of air and ingestion
of water, meat, milk, and fresh produce. The amounts of radionuclides
that are inhaled are calculated from these air concentrations and a
knowledge of how much air is inhaled by an average person. The amounts
of radionuclides ingested in the water, meat, milk, and fresh produce
that people consume are estimated using daily food intake values based on
data from the 19T7-1978 USDA nationwide food consumption survey (Ne84).
The subprogram DARTAB combines the information on the amounts of
radionuclides that are ingested or inhaled (as provided by the pathway
analysis) with radiation health risk data for a unit quantity of each
radionuclide (Be81). The health risk data are calculated by the code
RADRISK (SU81).
The RADRISK code first computes dose rates to organs resulting from
a unit quantity of a radionuclide that is ingested or inhaled. These
dose rates are then used in a subroutine adaptation of the program CAIRO
to estimate the risk of fatal cancers in an exposed cohort of 3.00,000
persons (Bu81). All persons in the cohort are assumed to be born at the
same time and to be at risk of dying from competing causes (including
natural background radiation). Estimates of potential health risk due to
exposure to a known quantity of approximately 500 different radionuclides
are tabulated and stored until needed. These risks are summarized in
terms of the probability of premature death for a member of the cohort
due to a unit quantity of each radionuclide that is ingested or inhaled.
This information is then combined with the unit response data to
determine the cumulative population health effects, which include fatal
cancers and serious genetic effects, and the maximum annual doses to the
CPG (see Section 8.3.6).
8.3.5 Regional Analysis - Use of Health Effect Conversion Factor (HECP)
Two population groups are used to estimate cumulative population
health effects — a local population and a regional basin population.
The cumulative population health effects to the local population are
determined directly by the PRESTO-EPA-POP, PRESTO-EPA-DEEP, and
PRESTO-EPA-BRC models, through a number of detailed pathway analyses
using iterative yearly updates. This impact is to a relatively small,
nearby community for the first 1,000 years after site closure. After
1,000 years, the local community is assumed to become part of the larger
regional basin community. This assumption is made to reduce the amount
of computer time, but will not affect the model results, as the local
population provides only a small percentage of the cumulative population
health effects in comparison to the much larger regional basin population.
Because of the small size of the local population, not all
contaminated water is used. All excess contaminated water, either ground
water or surface stream, flows past the local community to enter a river
8-13
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and then the downstream basin. Upon reaching the downstream basin, this
residual activity is assessed as part of the regional basin analysis.
After the 1,000-year local analysis period has ended and the local
population is considered to be a part of a larger regional basin
community, all activity that leaves the disposal site, in either ground
or surface water, travels directly to a regional basin river and enters
the downstream basin, where the activity is assessed as part of the
regional basin analysis. The residual activity from the first 1,000
years and the regional basin activity from the remaining 9,000 years are
added together. This total activity, in curies, is termed the regional
basin activity (see Figure 8-5) and is used to assess cumulative
population health effects as part of the regional basin analysis.
Fatal cancers and serious genetic effects to the regional basin
population are estimated by multiplying the regional basin activity
F^orf SUhe 5aSin fiV6r by nuc"-
-------
YEARS 1-1000
oo
I
i-1
Ul
BASIN RESIDUAL
RADIOACTIVITY:
REGIONAL BASIN
HEALTH EFFECTS:
B(Ci) = (A - W) + (S - Y)
HE = B x HECF
YEARS 1,001-10,000
BASIN RESIDUAL
RADIOACTIVITY:
REGIONAL BASIN
HEALTH EFFECTS:
B(Ci) = D + S
HE = Bx HECF
Figure 8-5. Regional Basin Health Effects Pathways
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* The HECF values include health effects caused by the ingestion of
contaminated fish. This pathway is not considered for the local
population exposure, since the consumption of fish from a small,
local stream would, in general, be minimal; and
» Separate HECF values are calculated for each hydrogeologic
setting, based on water usage factors for that region.
(A) Calculation of HECF Values
The HECF values, which are nuclide specific, are made up of a
terrestrial pathway portion (HECFti) and a portion based upon an
aquatic pathway (HECFfi). separate HECF values are determined for
fatal cancers and for genetic effects. The following discussion,
however, is applicable to either.
(1) Terrestrial Pathway;
The nuclide-specific health effects conversion factors for the
terrestrial pathway (HECFti) are calculated in two steps using the
results from PRESTO-EPA analyses for the local population. In the first
step, the health effects to the local population resulting from the
withdrawal of a unit curie of a specific nuclide from the local well or
stream are determined. The second step involves the calculation of the
fraction of activity discharged into the basin which will be withdrawn
from the regional stream. The HECFti is determined by multiplying the
fraction of activity withdrawn, by the health effect per curie conversion
factor:
HECFti
withdrawn) x f ,
where ;
HECFti
= the terrestrial HECF for radionuclide i;
withdrawn = the health effects to the local population per unit
curie of radionuclide i withdrawn from the local
well or stream; and
f = the fraction of activity withdrawn from the basin
river per unit activity released to the regional
basin.
The fraction of activity withdrawn by the regional basin population
is based on the local population water usage and a standard ratio of
river flow to population. The per capita water use (drinking water,
cattle watering, and irrigation) of the local population is taken from
appropriate PRESTO-EPA model inputs. Assuming that the per capita
regional basin water use will be uniform, and using a standard ratio of
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population to river flow for the regional basin, the fraction of regional
basin activity that will be withdrawn by the regional basin communities
can be calculated:
f = (U/P) T (Q/P)
where:
f = fraction of activity withdrawn from the basin river per unit
activity released to the regional basin;
(U/P) = per capita water usage (m-Vperson-yr); and
(Q/p) = ratio of river flow to population, 3,000 m3/person-yr (EPA85a).
Thus, when the local per capita water consumption is divided by a
standard ratio of river flow to population, the fraction of the regional
basin radioactivity that will be withdrawn by the regional basin community
is determined.
This calculation assumes two points. The first is that local water
usage is comparable to regional basin water usage. The second point is
that the global ratio of river flow to population is applicable to the
regional basin. As noted in the High-Level Radioactive Waste Background
Information Document (EPA85a), studies show that while regional basin
population and river flow vary widely, the ratio of the river flow to the
population remains relatively constant and that the value we have used is
within the range of similar values associated with various river basins in
the U.S.
It is again noted that the fraction of nuclides that are not taken up
by the regional basin community is assumed to enter the ocean, which acts
as a nuclide sink. Health effects from the activity released to the ocean
are assumed to be negligible. The percentage of activity in the basin
river that reaches the ocean varies from almost zero in the arid region to
about 95 percent in the humid regions. The reason that so much activity
reaches the ocean in the humid regions is that, in general, very little
surf.ace water is used. This assumption is tested in the sensitivity
analysis program and is discussed in Chapter 11.
(2) Fish Pathway:
One pathway that was viewed as negligible for the local population but
is considered in calculating regional basin health effects, is that of
contaminated fish ingestion. A separate HECFfi for fish is determined,
based on the following equation:
HECFfi = (P/Q) x Bfi x Uf x (D/C)i
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where:
Uf
health effects conversion factor from consumption of fish, per
curie of radionuclide i released to the regional basin;
(P/Q) - river-flow-to-population ratio (person-yr/3,000 m3) (EPA85a);
- fish bioaccumulation factor (Ci/kg-fish per Ci/1 of
radionuclide i in water) (NAS71);
= annual fish consumption rate (6.9 kg/person-yr) (Ru80); and
s conversion factor for health effects per curie of nuclicte i
ingested, obtained from PRESTO-EPA-POP calculations.
(B) Basin Health Effects Conversion Factor
The HECFfi is added to the HECFti calculated earlier to
determine the nuclide-specific regional basin KECP^ values:
= HECFti •
(C) Combined Health Effects
The total nuclide-specific activity reaching the regional basin over
10,000 years (q^i) is multiplied by the nuclide-specific HECF values
(HECFi), and summed over all nuclides, to determine the regional basin
health effects:
Regional Basin Health Effects = Z q , x HECF .
. bi i
These are added to the local health effects, already calculated
directly by PRESTO-EPA, to estimate the total health effects over 10,000
years. Note that the above discussion relates to the calculation of
either cancer deaths or serious genetic effects. These are calculated
separately, using separate HECFi values, and then combined to estimate
the cumulative population health effects.
Sensitivity analysis was performed on the input parameters
associated with the health effects conversion factors. This analysis, as
well as a general discussion on the calculation of the HECF values, is
discussed in Appendix G. Also listed in this appendix are the
nuclide-specific fish bioaccumulation factors.
8.3.6 Use of Unit Response
In order to evaluate 3 generic hydrogeologic disposal sites, 10
disposal methods, 24 waste streams, and 4 or 5 waste forms, a large
number of computer runs would have to be performed. Therefore, a unit
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response methodology was devised to reduce the number of computer runs. .
This methodology was used with the PRESTO-EPA-POP, PRESTO-EPA-DEEP, and
the PRESTO-EPA-BRC codes. The unit response methodology is not practical
for use with the PRESTO-EPA-CPG or PATHRAE-EPA codes, as an actual source
term is necessary in order to determine the year in which the maximum
annual dose would occur.
The unit response approach has been used in many applications and
has been proved to be a valid approach. The application of the unit
response approach to estimate cumulative population health effects was
evaluated by comparison with a direct assessment method. The results of
the evaluation showed that for the purpose of cumulative population
health effects estimation, both approaches give comparable results.
The data for a unit response analysis are altered slightly from the
data for a full facility analysis. In order to model a variety of waste
streams with a single PRESTO-EPA run, a single curie of each radionuclide;
is assumed to be present in the waste inventory. The volume of the waste
disposal trench is reduced to a single cubic meter of waste. The results
of this unit response analysis are used as inputs to an accounting model
program.
The accounting model adjusts the PRESTO-EPA unit-curie results in
proportion to the number of curies of a particular nuclide, per cubic
meter of a given waste stream. This results in a tabulation of health
effects arising from the disposal of one cubic meter of waste from a
particular waste stream. The accounting model then multiplies the impact
resulting from one cubic meter of a particular waste by the number of
cubic meters of that waste stream for the particular scenario being
modeled. By adding together the impact from all the waste streams
located at a site, the total impact from the waste disposed of at a
disposal site can be estimated. This calculation is also performed by
the accounting model. The accounting model is described in greater
detail in a separate report (EPA87g).
8.3.7 Time Periods Analyzed
in estimating health impact from the shallow-land disposal of LLW,
two main time periods are analyzed: (1) 1,000 years for impact to the
local population and to the critical population group, and (2) 10,000
years for impact to a regional basin population.
The PRESTO-EPA-CPG and PATHRAE-EPA codes are used in estimating
maximum annual dose to the critical population group. These two codes
model a 1,000-year period, although the important output is the annual
dose in the maximum year. Because the critical population group is
located close to the disposal site, maximum annual doses for the mobile
radionuclides generally occur soon after the assumed institutional
control period has ended and almost always before 1,000 years. In some
arid scenarios and with more restrictive disposal methods, annual doses
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continue to rise slowly after 1,000 years, although the maximum is
usually reached long after 1,000 years and is always small.
The PRESTO-EPA-POP and PRESTO-EPA-BRC codes estimate cumulative
population health effects to both a local population (for a 1,000-year
period) and a regional basin population (for 10,000 years). After 1,000
years, the local population is included within the larger regional basin
population. A modeling period of 10,000 years is necessary since several
radionuclides are hazardous for this period or longer and, some disposal
methods require long time periods before radionuclides reach the
population. While there is a great deal of uncertainty in many
parameters, especially when long time spans are used, these uncertainties
are present in each of the disposal methods to the same degree. Thus,
when comparing methods, the uncertainty becomes less important. This is
discussed in more detail in chapter 12.
The cumulative population health effect assessments also analyze two
shorter time frames (100 years and 500 years). Both local and regional
basin health effects are estimated. The results from these analyses are
useful in learning more about the different disposal alternatives and how
they compare in the earlier periods after disposal.
The PRESTO-EPA-DEEP code estimates cumulative population health
effects to a local and regional basin population from the deep disposal
of LLW. Because of the deep disposal methodology., very little
radionuclide release occurs before 1,000 years. Therefore,
PRESTO-EPA-DEEP estimates both local and regional basin health effects
for 10,000 years.
8.3.8 Modeling Inputs
The PRESTO-EPA codes require a large number of input parameters.
While some of the inputs are constant over all conditions, most vary
depending upon the site characteristics, waste stream, waste form,
disposal method, or radionuclide. The input parameters and values used
for the various PRESTO-EPA codes are listed in Appendix C. Discussed
below are the conditions upon which the inputs are dependent.
(A) Use of Generic Sites
Three sets of generic hydrogeologic disposal site characteristics
are modeled. These are termed "generic" sites because the data used in
the model are not specific to ,any actual disposal site or to any exact
geographic location. To generate health impact data that are useful in
the standard-setting process, the sites are typical of large areas of the
United States. The three generic settings chosen are: (1) a humid site
with low permeability; (2) a humid site with moderate permeability; and
(3) an arid site with moderate permeability. Three sets of standard
hydrogeologic input data are used to characterize the three different
generic sites. Data for water usage patterns, population distribution,
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and farming activities are representative of the general region of the
U.S. where the site is located. More specific data requirements, such
as precipitation patterns, temperature variations, and ground-water
movement, require the use of actual site data. The detailed
characteristics of the sites used as a source of these data are described
more fully in Appendix D. All of the site-dependent input parameters are
listed in Appendix C.
(B) Disposal Methods
Several of the input parameters to the PRESTO-EPA code serve to
define the waste disposal method. In particular, trench width and depth,
cap or cover thickness, porosity, and permeability are required. These
parameters are obviously dependent on site characteristics and design.
To compare the effectiveness of ten disposal methods at three generic
hydrogeologic sites, a standard conceptual design was prepared for each
disposal method. These conceptual designs use average dimensions and
specifications and are held constant from one site to another. This
generalization is consistent with the use of generic rather than actual
disposal sites. The combination of the unit response analysis
methodology (see Section 8.3.6), combined with generic disposal sites and
standardized disposal methods, results in well-defined data sets for each
site which can be easily modified to reflect different disposal methods.
Detailed conceptual designs for each disposal method are presented in
Chapter 4. Input values dependent upon the disposal method are listed in
Appendix c.
(C) Waste Forms
Although LLW may take a variety of forms, from disposable gloves and
scrap paper to large steel parts, five general waste forms are used to
simplify modeling. These waste forms influence some of the physical
characteristics of the trench material, such as porosity, permeability,
and density. The waste form also determines the rate at which the
radionuclides are released from the waste material, such as leaching rate
and release fraction. The five waste forms are: trash (TR), absorbing
materials (AW), solidified waste (SW), activated metal (AM), and
incinerated waste which is then solidified (IS). The waste form
dependent input values are listed in Appendix C.
(D) Waste Streams
The NRC identified 36 separate waste streams in supporting its
regulation for LLW disposal facilities (NRC81). A more recent NRC
document describes 148 waste streams, with greater emphasis on higher
activity wastes and nonroutine sources (NRC86). For EPA's analysis,
similar waste streams are combined to form a total of 24 waste streams.
This results in fewer scenarios to be modeled without sacrificing
accuracy. In addition to the 24 LLW streams, separate streams were
identified for NARM wastes and BRC.
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(1) NARM
To provide a basis upon which to consider the regulation of NARM
wastes, EPA commissioned a study of NARM waste streams (EPA85b). special
emphasis is placed oh higher activity wastes and those exhibiting
characteristics analogous to LLW regulated under the AEA. The NARM waste
streams included in EPA's LLW radiological source term are based on .this
study, and include radium sources and radium-contaminated ion exchange
resins (Ba86b).
(2) BRC
To characterize waste streams appropriate for the BRC analysis, a
select group of LLW waste streams was constructed. These are designated
as "surrogate" BRC wastes, i.e., waste streams meant to represent the
kinds of wastes that may be deregulated by other regulatory agencies
under various EPA-designated BRC levels. In general, the surrogate BRC
waste streams are lower activity LLW waste streams or, where enough
information is available, substrearas of previously defined LLW waste
streams. Various wastes are represented by these surrogate BRC wastes,
including those from nuclear fuel-cycle, institutional, and industrial
LLW generators.
A complete description of the waste streams is found in Chapter 3.
The input parameters that are dependent upon the waste streams are listed
in Appendix c.
(E) Radionuclides
Each of the waste streams'modeled contains various radionuclides. A
total of 40 radionuclides, including 5 NARM radionuclides, are modeled.
All of the nuclides and nuclide-dependent parameters are included as
input in the PRESTO-EPA models, in those models that use the unit
response methodology (POP, BRC, and DEEP), one curie of each nuclide is
assumed. For the models not utilizing the unit response methodology (CPG
and PATHRAE), the actual source terra for each radionuclide is included
for each separate computer run.
The complex physiochemical interaction between the radionuclides and
the solid geologic media has been grouped into a single factor, the
distribution coefficient (Kd). Separate Kd values are used for soil,
the mixture of soil and waste material in the trench, the subtrench soil,
and the aquifer. The Kd values for each radionuclide are listed in
Appendix C.
In addition to the K
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(F) Atmospheric Parameters (RAPE Program)
The PRESTO-EPA codes include a Gaussian plume model for atmospheric
transport. However, the codes only allow one wind direction, speed,
stability, distance, and population to characterize a community for each
scenario modeled. In order to facilitate the assessment of hecilth
impacts for multiple communities, a utility program called RADK
(Radioactive Atmospheric Dispersion and Exposure) has been designed
(Ro84b). The RADE code performs standard Gaussian plume atmospheric
dispersion calculations using subroutines from PRESTO-EPA. RADE produces
results that are suitable for use as input parameters in the PRESTO-EPA
codes, which allow for atmospheric modeling of multiple communities. The
atmospheric input parameters, including those which come from the RADE
program, are listed in Appendix C.
8.4 Health Impact Assessment - Regulated Disposal of LLW
The EPA LLW health impact assessment consists of two broad types of
analyses: regulated and unregulated (BRC) disposal.
General characteristics of the health impact assessment for
regulated disposal are discussed in this section. The second area, that
of unregulated (BRC) disposal, is discussed in Section 8.5.
8.4.1 PRESTO-EPA-POP
The original PRESTO-EPA model was used as a basis for a family of
PRESTO-EPA codes. The basic model is used to estimate cumulative
population health effects and is called PRESTO-EPA-POP.
This model calculates the cumulative population health effects,
resulting from the regulated disposal of LLW, to a local population and
to the population in a regional basin in which the disposal site is
located. The health impacts to the local and regional basin populations
are both analyzed for a period of 1,000 years. The regional basin
analysis, however, is extended for an additional 9,000 years. The values
used for the PRESTO-EPA-POP input parameters are listed in Appendix c.
Some characteristics specific to PRESTO-EPA-POP are discussed in the
following sections.
Because of the many waste disposal scenarios considered in the
PRESTO-EPA-POP analysis, the modeling was simplified. By utilizing a
unit response approach, the numbers of PRESTO-EPA-POP analyses were
reduced significantly. The unit response approach, which is based on the
disposal of one curie of each radionuclide, is described in section 8.3.6.
A radionuclide-specific health effect conversion factor (HECF) is
used to determine health effects to the regional basin population over
10,000 years. The HECF is described in more detail in Section 8.3.5.
8-23
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The starting point for the health impact assessment is assumed to be
immediately after the closure of the disposal site. The radionuclide
inventories of the waste in the trench are reduced to account for the
radioactive decay during the operational period (assumed to be 20
years). The waste is assumed to be containerized, with the leaching of
radionuclides from the waste beginning after the container fails. The
length of the container integrity is a user-specified parameter and is
listed in Appendix C.
Because in-growth of radiological decay products is not calculated
by the model to maintain simplicity (daughter product in-growth is
considered by RADRISK for internal exposures), cases where the major dose
contribution is from external exposure to a short-lived progeny in
equilibrium with a parent radionuclide present in the trench inventory,
the progeny is included directly in the trench inventory. This is the
case for the Cs-137 daughter, Ba-137ra, which is included in the source
term with the same activity and radiological characteristics as the
parent.
Operational spillage is defined as the radionuclides spilled from
waste packages and remaining on the ground surface at the close of
disposal operations. These radionuclides would subsequently be
transported either by the atmospheric pathway to the local population or
by the surface-water pathway to the local stream and basin river.
Each member of the population is assumed to eat the same quantities
of food, all grown on the same fields, and to obtain his or her drinking
and crop irrigation water from the same sources (a certain percentage of
which is assumed to be contaminated). This assumption simplifies the
calculations and is appropriate because of the large uncertainties in
predicting individual mobility, population demography, agricultural
practices, and geologic and hydrologic changes that might occur during
the analysis period. As input parameters, the user may specify the
fraction of the drinking and irrigation water that is supplied by
contaminated sources, as outlined in Appendix c.
Cumulative population health effects are calculated by
characterizing the population center for each site with a single,
geographically central location and the total population size, in
calculating the cumulative population health effects, the population age
distribution and size are held constant over the assessment period.
(A) General Characteristics of the PRESTO-EPA-POP Analyses
The results from the PRESTO-EPA-POP analyses are cumulative
population health effects to both a local and a regional basin
population. The specific estimates from our various computer runs are
detailed in Chapters 9 and 11. in the following sections, general
results for the PRESTO-EPA-POP code are discussed. Typical results from
the PRESTO-EPA-POP analyses show that, in general, the local population
8-24
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health effects do not dominate in any of the three regional hydrogeologic
and climatic scenarios. This is due to the limited amount of water and
food, contaminated by nuclides, that can be ingested and expose the
relatively small local population. The majority of the cumulative
population health effects are incurred by the regional basin population.
The cumulative health effects to the regional basin population, and the
pathways by which they occur, vary considerably over the three
hydrogeologic and climatic regions (see Figure 8-6). The results are
more easily reviewed by first separating them into the three general
settings: humid permeable, humid impermeable, and arid permeable.
At the site characterized by relatively permeable soil and high
rainfall, most of the mobile radionuclides leach out of the trench and
into the aquifer during the initial 1,000-year period. The majority of
the cumulative population health effects are incurred by the regional
basin population through the ground-water pathway during the first 1,000
years. ,
At the site characterized by high rainfall and relatively low soil
permeability, the trenches fill with water after the trench cap has been
assumed to have failed. The activity will be leached from the wastes and
will escape from the trench through overflow (bathtub effect) in a
relatively short period of time. The regional basin population receives
the majority of the cumulative population health effects through the
surface-water pathway during the first 1,000 years.
At the site characterized by relatively permeable soil but low
rainfall, most activity does not reach either the local or the regional
basin populations until relatively late in the modeling period. The
local population incurs some minor impact in the first few years due to
wind blown (atmospheric) transport of nuclides spilled onto surface soils
during site operations. This health impact, while quite small, can be
the only health impact during the first 1,000 years. This is due to the
long travel time required for contamination to reach the aquifer and then
travel to the local and regional basin populations by ground water. The
overall cumulative population health effects, which are always very
small, are dominated by health effects from activity reaching the
regional basin population through the ground-water pathway after 1,000
years.
8.4.2 PRESTO-EPA-CPG
The maximum annual doses to groups of individuals located close to
the disposal site are estimated using the PRESTO-EPA-CPG code. These
groups are assumed to have certain characteristics and to be associated
with environmental pathways where they are likely to receive a greater
exposure than the average population. These individuals are called the
critical population group (CPG).
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INFILTRATING WATER
RAIN & SNOW
WELL | EXPOSURE PATHWAYS
..^v I.., -,.....
AQUIFER
a) PERMEABLE MEDIUM WITH A HUMID CLIMATE
INFILTRATING WATER I' I H ' '
RELEASE TO
GROUND WATE
MINOR
RAIN & SNOW
MAJOR RELEASE BY
DIRECT OVERFLOW
MAJOR
EXPOSURE
STREAM
b) IMPERMEABLE MEDIUM WITH HUMID CLIMATE
\II7
EXPOSURE SMALL BUT
OVER LONG TIME
I' ' LITTLE INFILTRATING WATER
VERY SLOW RELEASE
TO GROUND WATER
i . -.• •.
I ,_•' g'"'^-
AQUIFER
C) PERMEABLE MEDIUM WITH AN ARID CLIMATE
Figure 8-6. Environmental Pathways at a Shallow LLW Disposal
Facility in the Three General Hydrogeologic and
Climatic Settings.
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In assessing the impact of waste disposal alternatives and
site-specific characteristics on CPG dose, a methodology was employed
which relied on the established PRESTO-EPA approach, with necessary and
appropriate modifications. The values used for PRESTO-EPA-CPG input
parameters are listed in Appendix C.
The transport of radionuclides was evaluated for the same pathways
pictured in Figure 8-1, although, as shown in Figure 8-7, the location of
the population of interest is changed.
Maximum annual dose rates were determined using the DARTAB
subroutine in PRESTO-EPA-CPG. The RADRISK data file, used in DARTAB,
contains the organ-specific dose factors for each radionuclide in the
inventory; effective whole-body dose equivalent rates are generated in
DARTAB using standard EPA organ-weighting factors. See Chapters 6 and 1
and Section 8.3.4 for more detailed information on DARTAB/RADRISK.
(A) Differences Between PRESTO-EPA-CPG and PRESTO-EPA-POP Methodologies
The calculation of maximum annual CPG dose relies upon the
PRESTO-EPA-POP approach with appropriate modifications. Table 8-1
summarizes the major differences between the two models. The first
difference is that the radionuclide-specific activity used in the CPG
calculations is the best estimate of the total activity and volume of
waste to be disposed of at a facility (assumed to be generally 250,000
m3). The radionuclide-specific inventories, by waste volume, are
listed in Chapter 3. This is in contrast to the normalized unit volume
and unit curie inventory used in PRESTO-EPA-POP, in which the cumulative
population health effects from a fully-loaded disposal facility are
calculated in a separate utility program. This aspect of the
PRESTO-EPA-POP code is discussed in section 8.3.6.
Another difference lies in the treatment of waste leaching, which
varies according to waste form. The total source term for the CPG
analysis contains various waste streams which are grouped into one of
five waste forms: (1) absorbing materials, (2) trash, (3) solidified,
(4) activated metals, and (5) incinerated-solidified.
The leaching of absorbed wastes and a fraction of the trash are
estimated using only a distribution factor (Kd). Leaching from the
other waste forms is characterized by a radionuclide-independent annual
leach fraction for each waste form, followed by K
-------
(MAXIMUM ANNUAL DOSE
TO CRITICAL POPULATION GROUP)
PRESTO-EPA-CPG
PRECIPITATION
CUMULATIVE POPULATION HEALTH EFFECTS
, Jfllt10 ^OCAL AND REG'ONAL BASIN POPULATION
(FATAL CANCERS AND SERIOUS GENETIC EFFECTS)
PRESTO-EPA-POP
1. MAXIMUM ANNUAL
DOSE TO CPG
2. YEAR OF OCCURRENCE
CUMULATIVE POPULATION
HEALTH EFFECTS ASSESSMENT
FOR LOCAL USE POINT UP
TO 1,000 YEARS
CUMULATIVE POPULATION HEALTH
„ ^_ EFFECTS ASSESSMENT FOR REGIONAL
^•F«H»M^ BASIN UPTO:(1) 1,000 YEARS
10,000 YEARS
Figure 8-7. Differences in Health Impacts Estimated and Locations
and Populations Evaluated for PRESTO-EPA-POP and
PRESTO-EPA-CPG
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Table 8-1. Main differences between PRESTO-EPA-POP and PRESTO-EPA-CPG
Characteristic
PRESTO-EPA-POP
PRESTO-EPA-CPG
Population Analyzed
Local and Regional Basin
Population
Critical Population
Group
Impact Analyzed
Cumulative Population
Health Effects
Maximum Annual
Whole-Body Dose
Modeling Period
(years)
Source Term
Site Size
Waste Form/Leaching
1,000 local population,
10,000 combined local
and regional population
1 curie of each nuclide
(unit response approach)
1 a? volume
(unit response approach)
Separate runs performed
for each waste form with
appropriate leaching
(unit response approach)
1,000 modeled, with
maximum year determined
Best estimate of actual
20-year disposal
activities
250,000 m3
Waste forms (five types)
combined with a two-
step leaching process
performed
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amount (curies) of each radionuclide leached in the facility water from
these wastes is added to the absorbing material inventory present at the
beginning of that year. This new absorbing material inventory is
partitioned between the contacting water and the absorbing phase
according to the distribution coefficients (Kd).
In PRESTO-EPA-CPG, the focus of attention for each disposal
alternative at each regional site is on identifying the year when the
dose to the CPG is maximum. Therefore, in PRESTO-EPA-CPG, at the end of
the simulation when the year of maximum equivalent whole-body dose rate
has been identified, radionuclide uptake information for that year is
input to the DARTAB subroutine. This subroutine then calculates organ,
pathway, and radionuclide whole-body dose equivalent rates for the
critical population group for the maximum year.
Lifetime risk to a member of the CPG is estimated using the maximum
annual dose rate, assuming it remains constant over the individual's
lifetime (average of 71 years). This calculation is done outside of the
PRESTO-EPA model and is based on a standard dose to risk conversion
factor. The association of a definite level of individual risk with a
maximum year's whole-body dose equivalent is tentative. Individuals in
the CPG may experience several years at or near the maximum exposure.
Where the half-life of the radionuclide is long and the maximum CPG
exposure occurs relatively late after disposal site closure, the exposure
may continue at close to the maximum level for many years. If the
radionuclldes causing the exposure have short half-lives and the maximum
CPG exposure occurs soon after closure, an individual may be exposed to
the maximum rate for only a few years. This leads to a possible
overestimate of the risk. In most cases, however, the maximum dose
remains near the maximum level for many years.
Another uncertainty is the duration of individual residence time at
the CPG location. A wide range of individual risk estimates is possible
depending upon the length of residence. In order to be conservative,
however, we assume that the critical population group will remain in the
same location throughout its lifetime.
(B) General Characteristics of the PRESTO-EPA-CPG Analyses
The results from PRESTO-EPA-CPG are maximum annual whole-body doses
to a critical population group located close to the disposal site. The
results of the analysis allow EPA to determine which disposal
alternatives would meet various levels of the standard under certain
generic circumstances. The specific estimates from our various computer
runs are detailed in chapters 9 and 11. in the following sections,
general results for the PRESTO-EPA-CPG code are discussed.
Typical results show that, in general, the impacts to the critical
population group and the pathways by which they occur vary considerably
over the three hydrogeologic and climatic regions (see Figure 8-4).
8-30
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Therefore, the analysis results are reported for each of the three
general settings. At the humid permeable site, the maximum dose occurs
through the ground-water pathway. The important nuclides are those with
high mobility (low Kd values), such as H-3, C-14, and 1-129. They
reach the critical population group within 1,000 years when combined with
relatively high ground-water velocities.
At the humid impermeable site, the maximum dose occurs within about
100 years of failure of the trench cap (assumed to occur in year 100) via
trench overflow directly to the surface-water pathway. The important
nuclides are those that are relatively mobile and have longer
half-lives. An example is 1-129, which reaches the critical population
group soon after the trench cap fails. It leaves the trench via overflow
and is transported directly to the local stream by surface water, thus
bypassing the greater retardation it would have if it had moved through
the ground. Nuclides with shorter half-lives, such as H-3, cause few
high doses due to their decay during the period the trench cover remains
intact.
At the arid permeable site, the maximum dose can occur in the first
year after closure because of the atmospheric transport of less mobile
nuclides, such as Co-60, Cs-137, and Ba-137m, spilled onto the surface
soil during site operations. This dose is very small (much less than one
mrera), however, since only a fraction of the total activity brought onto
the site is assumed to have been spilled during operations and even less
reaches the downwind population after dilution and dispersion by
atmospheric transport. A greater dose may occur through the ground-water
pathway, either late in the modeling period or even after 1,000 years.
The later doses can be significantly larger, although still very small
(much less than one mrem), and are dominated by mobile nuclides with
relatively long half-lives, such as C-14 and 1-129.
8.4.3 PRESTO-EPA-DEEP
The PRESTO-EPA-DEEP code estimates cumulative population health
effects from LLW disposal by deep-well injection, hydrofracture, and deep
geologic disposal. The code estimates, for up to 10,000 years following
the end of LLW disposal operations, local and regional basin health
effects. The maximum annual dose to a critical population group is
calculated by the PRESTO-EPA-CPG code, using all of the modifications and
assumptions assumed in this section, and the full source term as required
for the CPG analysis. The values used for the PRESTO-EPA-DEEP input
parameters are listed in Appendix C.
Water, principally from deep aquifers, is the primary transport
medium for radioactivity from LLW stored in deep facilities. Water
moving upward through the deep facility may ultimately enter a shallow
aquifer. Radionuclides that enter the aquifer may eventually reach
irrigation or drinking wells or surface streams and be consumed. The
consumption of radionuclides is through the food chain pathway, modeled
in the same manner as that in PRESTO-EPA-POP.
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The deep disposal scenarios implemented in PRESTO-EPA-DEEP consider
only the naturally occurring pathways such as natural ground-water and
surface-water flows and, in some scenarios, atmospheric air transport.
Intrusion scenarios, such as accidental drilling, geological faulting,
and the failure of the access shaft sealing, have a probabilistic nature
and are not considered.
In general, the environmental transport pathways are the same for
PRESTO-EPA-POP and PRESTO-EPA-DEEP. The major modifications found in
PRESTO-EPA-DEEP include extension of the period for local population
analysis, modification of the ground-water transport submodel, and the
bypassing of some submodels that are not applicable. These bypasses
include the infiltration submodel in the case of all the deeper
geological disposal alternatives and the air transport submodels in the
case of deep disposal in a mined cavity.
PRESTO-EPA-DEEP considers the vertical movement of ground water from
a lower confined aquifer through the waste facility and surrounding
strata to an upper aquifer, as shown in Figure 8-8. Water in the upper
aquifer moves horizontally to a receptor location where the water and
radioactive contaminants are introduced into the food chain pathway in
the same manner as that in PRESTO-EPA-POP.
In addition to the major changes made in the ground-water transport
pathway, minor modifications had to be made to certain portions of the
model to simulate the deep disposal scenarios. One change was to
increase the simulation time frame for the local population analysis from
1,000 to 10,000 years. This change was necessary because of the long
time periods required for radionuclides to travel from the deeply, located
facilities to the local population. The PRESTO-EPA-DEEP model is
discussed in detail in the PRESTO-EPA-DEEP User's Manual (EPA87c).
8.5 Risk Assessment - Unregulated Disposal (BRC)
The methods used to estimate the health impacts resulting from
disposal of certain very low activity LLW by less restrictive practices
than those currently used are presented in this section. A number of
surrogate waste streams and several types of facilities that would
ultimately receive them for disposal are identified. The migration of
radionuclides from these facilities through various pathways and the
resulting human exposures are calculated. From this information,
cumulative population health effects and maximum annual doses are
calculated. The results of this analysis are made compatible with the
results of the LLW analysis so that comparisons of disposal methods can
be made for LLW and BRC waste.
The PRESTO-EPA-BRC health impact assessment model (EPA87e), a
modified version of the PRESTO-EPA code, is the primary analytical tool
used in the cumulative population health effects assessment. In
addition, an accounting model (see Section 8.3.6) is used to facilitate
8-32
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GROUND SURFACE
WELL
SURFACE
STREAM
oo
U)
U)
GROUND-WATER FLOW IN AQUIFER AT RETARDED VELOCITIES
AQUICLUDE
WASTE REPOSITORY
LOWER AQUIFER
Figure 8-8. Ground-Water Model for Deep Disposal Scenarios
-------
the use of the unit response analysis methodology. The PATHRAE-EPA code
(EPA87f) is the primary tool used to estimate maximum annual doses to the
CPG.
8.5.1 PRESTO-EPA-BRC
The PRESTO-EPA-BRC code, which is used to estimate cumulative
population health effects from unregulated disposal of LLW, is a modified
version of the PRESTO-EPA model, in addition to the standard pathways,
the code includes the capability to determine health ef-fects from onsite
operations and airborne radioactivity released during incineration of the
waste. The number of onsite workers during the active operation of the
site is defined, and the maximum exposure level to which they are
subjected is estimated. For scenarios involving incineration of the
waste, a second set of air pathway parameters is required to account for
exposures to onsite and offsite individuals during the period of
incineration. These parameters are the fractions of time that the wind
blows toward the population of interest when incineration is considered
and the exposure per unit source strength. These parameters are
determined in the same manner as were the general air pathway parameters,
but apply only during the operational period of the incinerator, in
addition, a fractional volatility factor must be specified for each
radionuclide to facilitate calculation of the quantity of radioactivity
being emitted by the incinerator. The input values associated with these
parameters are listed in Appendix C. For more detailed information on
the PRESTO-EPA-BRC code, see the User's Manual (EPA87e).
A number of scenarios were developed for the BRC health impact
assessments, including a variety of deregulated disposal alternatives,
i.e., sanitary landfills, municipal dumps, onsite landfills, and
incineration methods situated in rural, suburban, and urban settings.
For more detailed information on the disposal methods, see chapters 4, 5,
and 10.
In evaluating whether some types of LLW might potentially be BRC
wastes, a number of surrogate LLW wastes were analyzed, as they have been
generated at nuclear power reactors, uranium fuel fabrication and uranium
process facilities, and industrial, medical, and educational facilities,
as well as by consumers, since the BRC analysis assumes that the waste
will receive no special treatment or packaging, the "as is" waste form is
used. For more detailed information on waste streams and waste form, see
Chapter 3.
The hydrogeologic settings used were comparable to those analyzed in
the LLW scenarios, although demographic characteristics were modified to
model more urban settings (see chapter 4). The input parameters and
parameter values for the PRESTO-EPA-BRC code are listed in Appendix c.
.The radlonuclides used in the BRC analysis were comparable to those used
in the LLW analysis.
8-34
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The results of the PRESTO-EPA-BRC analysis are cumulative population
health effects, consisting of fatal cancers and serious genetic effects,
to a local and a regional basin population. These results are similar to
those from the PRESTO-EPA-POP analysis. The results of the
PRESTO-EPA-BRC analysis are presented in Chapter 10.
8.5.2 PRESTO-EPA-BRC Pathways
The evaluation of cumulative population health effects from disposed
BRC waste involves exposure during the operating period of the disposal
facility For example, workers at a sanitary landfill are not radiation
workers and their doses must be considered in evaluating the cumulative
population health effects. In addition, the general public has access to
many facilities appropriate for BRC waste disposal. Since the primary
PRESTO-EPA computer code for assessing population health effects did not
consider exposure mechanisms possible with BRC waste disposal, it was
necessary to develop a new computer program to accomplish this function.
The PRESTO-EPA code was modified by additional exposure pathways as
discussed below. The resulting computer code, PRESTO-EPA-BRC, more
completely assesses the cumulative population health effects that could
result from unregulated or less restrictive disposal of BRC waste. The
other pathways are the same as those used with the PRESTO-EPA-POP model.
The model assumes that the radionuclide inventory is the amount of
activity found in the facility at the end of the disposal operation. The
wastes disposed of in the facility are assumed to be a homogeneous
mixture of radionuclides and waste materials.
The exposure pathways for the BRC scenarios are listed in
Table 8-2. Other combinations of pathways may be specified by changing
the input parameters. We consider these to be the maximum exposures
involving the actual BRC waste disposal activities, either during
operations or after site closure.
The major modifications to the PRESTO-EPA code that were required
for BRC analysis, involve adding the ability to calculate and accumulate
exposures during the disposal facility operations (Ro84a). since
PRESTO-EPA is oriented to post-closure events, all impacts from BRC
operations are summed over the 20 years of operations and assumed to
occur in the first year after site closure in PRESTO-EPA-BRC. Major
modifications incorporated into PRESTO-EPA-BRC include the following
supplemental pathways in addition to the regular PRESTO-EPA pathways:
• Worker and site visitor dust inhalation during operations;
• population dust inhalation from mechanical disturbances during
operations;
• Population inhalation of incinerator releases during operations;
8-35
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Table 8-2. Radiological exposure pathways for PRESTO-EPA-BRC scenarios
Scenario
For population
Normal
Farming
Eroded Trench
Ingests offsite water
Ingests offsite foods
Inhales downwind air
Ingests onsite foods
Ingests offsite water
Inhales suspended material at centroid
Direct exposure plume
8-36
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• worker and site visitor gamma exposure during operations; and
• Changes in the onsite farming option; includes human intrusion
onsite (construction of a house, i.e., residential use)
immediately after closure.
These additional exposure pathways and changes in PRE STO-EPA are
described in the following sections. For our modeling we assume that
during the first year following closure the maximum health impact will
occur .
(A) Dust Exposures
When calculating health effects to non-radiation workers at typical
as
.
site will be exposed to varying levels of potentially contaminated
atmospheric dust and direct gamma radiation, depending on their locations
and times spent at each location.
Dust exposures to the population occur through the following four
mechanisms :
(1) worker and Site Visitor Dust Exposure
worker and site visitor dust exposures are assumed to be constant
from year to year during operations. Thus, the cumulative dust exposure
over the operational life of the facility can be determined by
multiplying the annual exposure rate by the operational life of the
facility in PRESTO-EPA-BRC , this cumulative exposure is calculated and
included'with the other post-closure exposures that do not occur during
the 20 years of BRC waste disposal operations. This cumulative exposure
is then considered in the first year after operations cease, which is the
first year of the PRESTO-EPA-BRC simulation.
(2) Population Dust Exposure from Mechanical Disturbances
The cumulative population dust exposure resulting from mechanical
disturbances during operations is also calculated for 20 years and
considered with the other events during the first year of post-closure
during the PRESTO-EPA-BRC simulation.
(3) Population Exposures from Incinerator Releases
The cumulative population exposure resulting from incinerator
releases during operations is also calculated for 20 ^«^tj£
with the other events in the first year of post-closure simulation
8-37
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Population Exposure from Natural Resuspension
The population exposure resulting from natural resuspension is
calculated during operations in the identical manner as in PRESTO-EPA
after operations. As with other dust exposures during operations, this
exposure is accumulated over the facility's operational life and is added
(B)
Worker and Site Visitor Gamma Exposure
r^r-m^ Calculatin9 exposures resulting from direct gamma radiation bv
PRESTO-EPA-BRC, the maximum expected gamma radiation level is used for
calculating gamma doses to workers and visitors. Depending on their
locations, the workers and visitors are then exposed to some gamma
radiation depending on time spent at various locations near the waste
trenches The equivalent full-time, full-exposure population is used to
represent population exposure to workers and visitors from gamma
radiation. The cumulative gamma exposure to both workers and site
^vJ"!?r^ f a^° accumulated ov^ the 20-year facility lifetime and
maximized in the first year after closure.
(C) Qnsite Farming
in the post-operation period, i.e., the period after the site is
closed and returned to unrestricted use (considered to be the first year
after closure), for the farming or reclamation scenario, the onsite
farmer may grow and eat his or her own vegetables, beef, and milk
hn? ^fV^f irrigated ^ *ne Potentially contaminated onsite water,
but drink offsite public supply water equal in concentration to the
™n^STari° The farmer also "iay inhale suspended,
contaminated soil from the residual operational spillage. Population
dose and risk calculations under the farming scenario may assume that the
£nPrp S grown °n tne site are ingested by the general population or
by the farmer and his family. (Appendix c presents food product
parameters.) *
of vJ;RfS?°~EPA-BRC W11J- calculate exposures resulting from consumption
of vegetation grown onsite. PRESTO-EPA-BRC modeling also calculates
population dust exposures from post-closure activities such as onsite
tarraing.
8.5.3 PATHRAE-EPA
The PATHRAE-EPA code, which was originally developed by Rogers and
Associates Engineering Company (Ro84a), is used to assess the exposures
to the CPG from the unregulated disposal of BRC waste, while this code
is not directly based on PRESTO-EPA, it was modified extensively to
incorporate the analytical concepts used for the PRESTO-EPA family of
codes. The PATHRAE-EPA code most closely resembles PRESTO-EPA-CPG, and
8-38
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studies have been done to compare the output. The results of the studies
show the output of the two codes to be comparable under the circumstances
for which they are being used in the LLW standard development (Sh86).
The computer ,code PATHRAE-EPA is designed to assess the maximum
annual CPG dose for each exposure pathway resulting from the disposal of
LLW. Maximum annual doses are calculated to workers during disposal
operations, to offsite personnel after site closure, and to reclaimers
and inadvertent intruders after site closure. Dose conversion
calculations are performed to give annual doses. The PATHRAE-EPA code is
described in greater detail in the PATHRAE-EPA User's Manual (ISPA87f).
The main advantages of the PATHRAE-EPA model are its ease of
operation and simplicity of presentation, although with the simplicity
comes some sacrifice in the accuracy of the dose assessment and the loss
of the ability to assess combined effects from each pathway. It is felt
that this loss is not significant, however. Site performance and
facility designs for LLW disposal can be readily investigated, with
relatively few parameters needed to define the problem. Some important
parameter values are obtained from the results of PRESTO-EPA
calculations. Results are annual doses, by radionuclide, as a function
of time for each pathway, as well as total annual dose rates with time.
This permits quick focusing on key pathways and parameters.
For conservatism, the entire radionuclide inventory is used as the
source term for each pathway calculation, and depletion of the inventory
via migration through other pathways is ignored. This, of course,
provides conservative estimates while saving computer time.
The PATHRAE-EPA methodology considers both offsite and onsite
pathways through which man can be exposed to radioactive waste. The
onsite pathways include the ingestion of food grown onsite, direct gamma
exposures to workers and intruders, and the inhalation of radioactive
dust by workers and intruders. Offsite pathways are the same as for the
PRESTO-EPA codes.
The PATHRAE-EPA analysis produces a set of annual doses to an
individual as a function of time, nuclide, and pathway. Radionuclide
concentrations in river water and well water are also given for times up
to 10,000 years.
8.5.4 PATHRAE-EPA Pathway Analysis
In considering the pathways through which man can be exposed to BRC
waste, both onsite and offsite, two major assumptions were made:
1. Waste materials in the trench are assumed to be a homogeneous
mixture of radionuclides and other waste materials.
2. Radionuclides are transported vertically from the trench bottom
to the aquifer and then horizontally through the aquifer.
8-39
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(A) The onsite Worker Pre-Closure Pathways
Pre-closure exposures to onsite workers occur through two pathways:
(1) direct exposure to gamma radiation from the buried waste and
(2) internal exposures from radioactive dust inhaled during operations.
These two pathways are calculated in the last year of operation for
the 20-year accumulation of decayed radioactive waste and occur for all
three hydrogeologic/climate settings. This is based on the assumption
that the last year would provide the maximum exposure for the 20-year
accumulation of waste.
(B) The Post-Closure Onsite Resident Pathways
Post-closure exposures to onsite residents occur through two
pathways - ingestion of food grown onsite and biointrusion.
In the food grown onsite pathway radionuclides are brought to the
surface by construction activities or burrowing animals, both of which
disturb trench cover to a depth of 3 meters. The food plants grown in
onsite gardens are then assumed to absorb radionuclides from the
disturbed ground.
In the biointrusion pathway, the roots of food plants grown in
onsite gardens are assumed to penetrate into the undisturbed waste
(greater than 6 meters), and the plants are later consumed by humans.
These two pathways are calculated in the first year after closure
for the 20-year accumulation of decayed radioactive waste and occur for
all three hydrogeologic/cliraatic settings. This is based on the
assumption that the first year following closure, and the site's return
to unrestricted use, will provide the maximum exposure based on 20 years
of accumulated BRC waste.
(C) Post-Closure offsite Resident Pathways
Post-closure exposures to offsite residents occur mainly through
water pathways: ground-water to the river, facility overflow (bathtub
effect), surface erosion, and ground-water to the well.
In the surface water or ground-water to the river pathway, the
contaminated leachate from the waste trenches migrates through the ground
water to a major aquifer that supplies a nearby river used for
irrigation, livestock, or domestic purposes.
In the facility overflow or commonly called the bathtub effect
pathway, the disposal trenches accumulate water because of trench cap
failure and eventually overflow to nearby surface streams.
8-40
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During surface erosion, the cover and subsequently the waste itself
are eroded. The radionuclides then may reach nearby surface waters by
overflow.
The ground-water to well pathway is based on a nearby well used for
irrigation, livestock watering, or domestic purposes, and is contaminated
with leachate through ground-water migration from the waste trenches.
The river water and facility overflow pathways occur only for the
humid impermeable hydrogeologic/climatic setting; the surface erosion
pathway occurs only for humid climatic settings. The well water pathway
occurs at all settings. The ground-water to river, ground-water to well,
erosion, and overflow pathways all occur at from 100 to thousands of
years after closure.
(D) Pre-Closure Offsite Pathways
Pre-closure exposures to offsite residents occur through two
pathways: spillage and atmospheric inhalation.
The spillage pathway occurs during placement in the trench and the
spilled material mixes with surface water and discharges to a nearby-
stream.
For the atmospheric inhalation pathway, dust resuspension, a trench
fire, or a waste incinerator may be a source of contaminated gas and
particulate matter in which the radioactive plume migrates offsite before
affecting people.
For these two pathways the spillage and atmospheric inhalation occur
in the last year of operation. The spillage pathway occurs only for the
humid impermeable hydrogeologic/climate setting. The atmospheric
inhalation pathway occurs at all settings.
8.5.5 Additional BRC Analyses ,
In addition to the PRESTO-EPA-BRC and PATHRAE-EPA runs, other
analyses are required to evaluate BRC waste disposal scenarios. These
analyses include the evaluation of transportation exposures and exposures
caused by flooding of the disposal site.
A special analysis was made to evaluate direct radiation exposures
to workers who would collect and transport BRC wastes from the generator
to the disposal facility. For this analysis, additional short-lived
nuclides (half-lives of less than 1 year) were included, as they might
provide additional direct radiation doses. The results of this analysis
are described in Chapter 10.
Preliminary analyses were performed to evaluate the risks from
disposal site flooding (La84). The exposures from these scenarios were
8-41
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found to be minimal or much less than other scenarios because of the
effect of dilution from flooding. These analyses were not included in
our final methodology or results.
8-42
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Ba85
Ba86a
Ba86b
Be81
Bu81
BPA83
EPA85a
EPA85b
REFERENCES
Bandrowski, M.S. and C.Y. Hung, Environmental Transport Pathways
of the EPA Model (PRESTO-EPA) Used to Determine Health Impact
from Low-Level Radioactive Waste Disposal, Environmental
Radiation '85: Proceedings of the Eighteenth Midyear Topical
Symposium of the Health Physics Society, January 6-10, 1985,
Colorado Springs, Colorado, pp. 477-484, Compiled by Proceedings
Committee of the Central Rocky Mountain Chapter of the Health
Physics Society, 1404 Bridger Street, Laramie, Wyoming, 82070,
1985.
Bandrowski, M.S., Hung, C.Y. and o.L. Meyer, Sensitivity Analyses
of EPA's Codes for Assessing Potential Health Risks from Disposal
of Low-Level Wastes: Proceedings of 7th Annual Participants'
information Meeting on DOE Low-Level Waste Management Program,
Las,Vegas, Nevada, September 10-13, 1985, CONF-8509121, Las
Vegas, Nevada, 1986.
Bandrowski, M.S. and J.M. Gruhlke, Inclusion of NARM in the LLW
Standard, Proceedings of 8th Annual Participants' Information
Meeting on DOE Low-Level Waste Management Program, Denver,
Colorado, 1986.
Begovich, C.L., Eckerman K.F., Schlatter B.C. and S.Y. Ohr,
DARTAB- A Program to Combine Airborne Radionuclide Environmental
Exposure Data with Dosimetric and Health Effects Data to Generate
Tabulations of Predicted Impacts, Oak Ridge National Laboratory
Report ORNL-5692, Oak Ridge, Tennessee, 1981.
Bunger, B.M., Cook, J.R. and M.K. Barrick, Life Table Methodology
for Evaluating Radiation Risk: An Application Based on
Occupational Exposures, Health Physics, 40.:439-455, 1981.
U.S. Environmental Protection Agency, PRESTO-EPA: A Low-Level
Radioactive Waste Environmental Transport and Risk Assessment
Code - Methodology and User's Manual, Prepared under Contract No.
W-7405-eng-26, Interagency Agreement No. EPA-D—89-F-000-60, U.S.
Environmental Protection Agency, Washington, D.C., April 1983.
U.S. Environmental Protection Agency, High-Level and Transuranic
Radioactive Wastes - Background information Document for Final
Rule, EPA 520/1-85-023, Washington, D.C., August 1985.
U S Environmental Protection Agency, Radiation Exposures and
Health Risks Associated with Alternative Methods of Land Disposal
of Natural and Accelerator-Produced Radioactive Materials (NARM),
performed by PEI Associates, Inc., and Rogers and Associates
Engineering Corp. under EPA Contract 68-02-3878, October 1985.
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EPA86 U.S. Environmental Protection Agency, Environmental Pathways
Models for Estimating Population Health Effects from Disposal of
High-Level Radioactive Waste in Geologic Repositories, EPA
520/5-85-026, Washington, D.C., May 1986.
EPA87a U.S. Environmental Protection Agency, in press, PRESTO-EPA-POP•
A Low-Level Radioactive Waste Environmental Transport and Risk'
Assessment Code, Volume I, Methodology Manual, RAE-8706-1, Rogers
and Associates Engineering Corporation, salt Lake city, Utah,
1987.
U.S. Environmental Protection Agency, in press, PRESTO-EPA-POP:
A Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Code, Volume II, User's Manual, RAE-8706-2, Rogers and
Associates Engineering Corporation, Salt Lake city, Utah, 1987.
U.S. Environmental Protection Agency, in press, PRESTO-EPA-DEEP•
A Low-Level Waste Environmental Transport and Risk Assessment
Code, Documentation and User's Manual, RAE-8706-3, Rogers and
Associates Engineering Corporation, salt Lake city, Utah, 1987.
U.S. Environmental Protection Agency, in press, PRESTO-EPA-CPG:
A Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Code, Documentation and User's Manual, RAE-8706-4,
Rogers and Associates Engineering Corporation, Salt Lake City,
Utah, 1987..
U.S. Environmental Protection Agency, in press, PRESTO-EPA-BRC-
A Low-Level Radioactive Waste Environmental Transport and Risk"
Assessment Code, Documentation and User's Manual, RAE-8706-5,
Rogers and Associates Engineering Corporation, Salt Lake City,
Utah, 1987.
U.S. Environmental Protection Agency, in press, PATHRAE-EPA: A
Performance Assessment Code for the Land Disposal of Radioactive
Waste, Documentation and User's Manual, RAE-8706-6, Rogers and
Associates Engineering Corporation, Salt Lake City, Utah, 1987.
EPA87g U.S. Environmental Protection Agency, in press, Accounting Model
for PRESTO-EPA-POP, PRESTO-EPA-DEEP, and PRESTO-EPA-BRC Codes,
RAE-8706-7, Rogers and Associates Engineering Corporation, Salt
Lake City, Utah, 1987.
EPA87b
EPA87c
EPA87d
EPA87e
EPA87f
Ga84
Galpin, F.L. and G.L. Meyer, Overview of EPA's Low-Level
Radioactive Waste Standards Development Program, 1984:
Proceedings of 6th Annual Participants' Information Meeting on
DOE Low-Level Waste Management Program, Denver, Colorado,
September 11-13, 1984, CONF-8409115, Idaho Falls, Idaho, 1984.
8-44
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HU81
Hu83a
Hu83b
La84
Me81
Me84
Me85
NAS71
Hung, C.Y., An Optimum Model to Predict Radionuclide Transport in
an Aquifer for the Application to Health Effects Evaluation, in:
Proc. Modeling and Low-Level Waste Management: An Interagency
Workshop held December 1-4, 1980, Denver. (C.A. Little and L.E.
Stratton, compilers), Department of Energy Report ORO-821, pp.
65-80, Oak Ridge, Tennessee, 1981.
Hung, C.Y., Meyer, G.L. and V.C. Rogers, Use of PRESTO-EPA Model
in Assessing Health Effects from Land Disposal of LLW to Support
EPA's Environmental standards: U.S. Department of Energy,
Proceedings of 5th Annual Participants' Information Meeting on
DOE Low-Level Waste Management Program, Denver, Colorado, August
30, 1983, CONF-8308106, Idaho Falls, Idaho, 1983.
Hung, C.Y., A Model to Simulate Infiltration of Rainwater Through
the Cover of a Radioactive Waste Trench Under Saturated and
Unsaturated Conditions, in: Role of the Unsaturated Zone in
Radioactive and Hazardous Waste Disposal, Edited by J.W. Mercer,
et al., pp. 27-48, Ann Arbor Science, Ann Arbor, 1983.
Lachajczyk, T., et al.. Final Report: Radiation Exposures and
Health Risks Resulting from Less Restrictive Disposal
Alternatives for Very Low-Level Radioactive Wastes, Performed by
Envirodyne Engineers, Inc., for EPA, under contract No.
68-02-3178, Work Assignment 20, U.S. Environmental Protection
Agency, Washington, D.C.» 1984.
Meyer, G.L. and C.Y. Hung, An Overview of EPA's Health Risk
Assessment Model for the Shallow Land Disposal of LLW,
Proceedings of an Interagency Workshop on Modeling and Low-Level
Waste Management, Denver, Colorado, December 1-4, 1980, ORO-821,
Oak Ridge National Laboratories, Oak Ridge, Tennessee, 1981.
Meyer, G.L., Modifications and Improvements Made to PRI5STO-EPA
Family of LLW Risk Assessment Codes Based on Recommendations of
Peer Review, February 1984, U.S. Environmental Protection Agency,
letter dated July 13, 1984, to members of PRESTO-EPA Peer Review,
February 7-8, Airlie, Virginia: Washington, D.C., 1984.
Meyer, G.L., Galpin, F.L. and J.M. Gruhlke, Overview of EPA's
Low-Level Radioactive Waste Standards Development Program, 1985:
Proceedings of 7th Annual Participants' Information Meeting on
DOE Low-Level Waste Management Program, Las Vegas, Nevada,
September 10-13, 1985, CONF-8509121, Las Vegas, Nevada, 1985.
National Research Council/National Academy of Sciences,
Radioactivity in the Marine Environment, Washington, D.C.
1971,
8-45
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Ne84
NRC81
NRC82
NRC86
PR84
Ro84a
Ro84b
RU80
SAB85
Sh86
Nelson, C.B. and You-Yen Yang, An Estimation of the Daily Food
Intake Based on Data from the 1977-1978 USDA Nationwide Food
Consumption Survey, U.S. Environmental Protection Agency EPA
520/1-84-015, Washington, D.C., May 1984.
U.S. Nuclear Regulatory Commission, Draft Environmental Impact
Statement on 10 CFR Part 61, Licensing Requirements for Land
Disposal of Radioactive Waste, NUREG-0782, Washington, D.C .
September 1981. _»..*,...
U.S. Nuclear Regulatory Commission, Final Environmental Impact
Statement on 10 CFR 61, Licensing Requirements for Land Disposal
of Radioactive Wastes, NUREG-0945. November 1982.
U.S. Nuclear Regulatory Commission, Update of Part 61 impacts •
Analysis Methodology, NUREG/CR-4370, January 1986.
The Peer Review of PRESTO-EPA (Release 2.4), A Draft Summary,
February 7 and 8, 1984, Airlie House, Virginia, U.S.
Environmental Protection Agency, Washington, D.C.
!
Rogers, v.C., Hung, c.Y., Cuny, P.A. and F. Parraga, An Update on
Status of EPA's PRESTO Methodology for Estimating Risks from
Disposal of LLW and BRC Wastes, U.S. Department of Energy,
Proceedings of 6th Annual Participants' Information Meeting on
DOE Low-Level Waste Management Program, Denver, coloradq,
September 11-13, 1984, CONF-8409115, Idaho Falls, Idaho, 1984.
Rogers, v.C., Klein, R.B. and R.D. Baird, Radioactive Atmospheric
Dispersion and Exposure - The RADE 3 Air-Pathway Unit Response
Code with Plume Rise Effects, Rogers and Associates Engineering
Corporation - Technical Information Memorandum TIM-51-7, Rogers
and Associates Engineering Corporation, Salt Lake City, Utah,
June 29, 1984.
Rupp, E.M., Miller, F.L., and C.F. Baes III, some Results of
Recent Surveys of Fish and Shellfish Consumption by Age and
Region of U.S. Residents, Health Physics, 39(2), 1980.
Science Advisory Board, U.S. Environmental Protection Agency
Report on the March 1985 Draft Background Information Document
for Proposed Low-Level Radioactive Waste Standards,
SAB-RAC-85-002, Washington, D.C., November 1985.
shuraan, R. and v.c. Rogers, A Comparison of PATHRAE and
PRESTO-CPG Simulation Results. Rogers and Associates Engineering
Corporation - Technical Information Memorandum TIM-8469/11-1?,
Rogers and Associates Engineering Corporation, Salt Lake citv,
Utah, January 31, 1986.
8-46
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S168 slade, D.H. ed., Meteorology and Atomic Energy (1968), U.S.
Atomic Energy commission Report TID-24190, Washington, D.C., 1968.
Sm82 Smith, J.M., Fowler, T.W. and A.S. Goldiri, Environmental Pathway
Models for Estimating Population Risks from Disposal of
High-Level Radioactive Waste in Geologic Repositories, EPA
52ti/5-80-002, 1982.
Su81 Sullivan, R.E.. et al., Estimates of Health Risk from Exposure to
Radioactive Pollutants, ORNL/TM-^7745, Oak Ridge National
Laboratory, Oak Ridge, Tennessee, November 1981.
8-47
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Chapter 9: ESTIMATED DOSES AND HEALTH EFFECTS FROM, THE
REGULATED DISPOSAL OF LLW
9.1 Introduction
Previous chapters provided detailed descriptions of the key elements
needed to perform a risk assessment of LLW disposal. LLW has been
described in terms of volumes and concentrations representing numerous
waste streams (Chapter 3). Disposal methods have been identified and
described (Chapter 4), along with major categories of hydrogeological and
climatic 'settings (Chapter 5). The calculational models describing
environmental transport and radiation dosimetry have also been defined
(Chapters 6, 7, and 8).
This chapter presents, the rationale for the selection and results of
the base case assessments of health impact performed for the regulated
disposal of LLW. Comparing the potential impacts from LLW disposal under
a broad range of disposal alternatives and regional conditions is an
important element supporting the development of EPA's generally
applicable LLW standards, as these standards will apply to LLW facilities
throughout the United States.
The only practical method of reducing the hazard of the land
disposal of LLW is to isolate it from people and the environment until
the radioactivity has decayed to very low levels. This assessment
projects the capability of the disposal system for isolating the
radioactivity in LLW from human populations. These results reflect the
undisturbed performance of an engineered LLW disposal system without
disruption by human intrusion or unlikely natural events. Such external
disruptions may best be handled on a site-specific basis.
This chapter compares the undisturbed performance of engineered
disposal systems located in various hydrogeologic settings in terms of
two critical radiological effects: (1) the maximum annual critical
population group (CPG) dose (provided in terms of a committed effective
whole-body dose equivalent), and (2) cumulative population health effects
(in terms of fatal cancers or serious genetic effects). The results
given in this chapter can be described as the base case analysis, because
the selection of disposal methods and waste forms is intended to
illustrate a step-wise progression in technological sophistication rather
than all possible combinations of disposal methods and waste forms.
Chapter 11, sensitivity analysis, presents a complete listing of the
results obtained for all combinations of disposal methods and waste forms
investigated.
9.2 Input Data and Rationale for Base Case Analysis
The purpose of this chapter is to compare the health impacts from
the undisturbed performance of engineered disposal systems covering a
9-1
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wide range o£ technological sophistication. In order to evaluate health
impacts, certain critical input data must be specified for each
analysis. These factors include the source term, the hydrogeologic
setting, the engineered disposal method and associated waste forms, and
the radiological risk assessment model. These factors are described in
Chapters 3, 4, 5, and 8. The'following sections discuss the pertinent
data, assumptions, and rationale related to each factor that is used to
construct the base case analysis.
9.2.1 The Low-Level Radioactive Waste Source Term
The overall LLW source term is described in chapter 3. Twenty-six
waste streams are defined based on similarities in origin and their
general physical, chemical, and radiological characteristics. Table 3-2
indicates 24 waste streams are presently regulated under AEA authority.
As explained in Chapter 3, the EPA source term is based in large part
upon the NRG characterization of commercial LLW (NRC86, Gr86) regulated
under the AEA. In addition, EPA includes two high radionuclide-
concentration source streams containing NARM wastes (PEI85).
The 26 waste streams included in EPA's overall LLW source term and
their projected 20-year volumes for 1985-2004 are shown in Table 9-1.
The NRG classification of these waste streams under 10 CFR 61 and the
physical-chemical form assigned to each waste stream are also indicated
in the table. Since the EPA analysis of regulated LLW disposal is
principally directed at estimating and comparing impacts over the long
term, the EPA LLW source term considers only longer-lived (i.e.,
half-life of more than one year) radionuclides. In addition to long
half-life, certain radionuclides are included if they exhibit high
radiotoxicity (e.g., 1-129, Np-237) and/or are present in significant
amounts in LLW (e.g., Cs-134). Table 9-2 lists 34 radionuclides
considered in the EPA analysis. The inventories shown in this table
reflect the total projected activities in commercial LLW and NARM from
1985 to 2004. The total volume associated with the commercial LLW
inventory is about 3E+06 m3 over the same time period.
In order to derive the source term used for the base case analysis,
it is necessary to modify the overall LLW source term described above.
To illustrate implementation of all phases of the EPA LLW standard,
namely, a generally applicable radiation protection standard with
provisions for inclusion of high specific activity NARM wastes and
implementation of a BRC level, an appropriate LLW source term is
constructed for the base case analysis. To illustrate implementation of
BRC,» seven waste streams are excluded from consideration as a regulated
LLW. These waste streams are as follows: N-SSTRASH, N-SSWASTE,
F-COTRASH, F-NCTRASH, U-PROCESS, F-PROCESS, and I-LQSCNVL.
9-2
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Table 9^-1. Overall LLW source term: Commercial LLW and NARM
volumes by waste stream, 1985 - 2004
(PHB85) '
Waste stream3
L-IXRESIN
L-CONCLIQ
L-FSLUDGE
P-FCARTRG
L-DECONRS
L-NFRCOHP
F-PROCESS
U-PROCESS
L-COTRASH
L-NCTRASH
F-COTRASH
F-NCTRASH'
I-COTRASH
N-LOTRASH
N-SSTRASH
N-SSWASTE
I-LQSCNVL
I-ABSLIQO
I-BIOWAST
N-LOWASTE
N-ISOPROD
N-SOURCES
N-TRITIUH
N-TARGETS
R-RAIXRSN
R-RASOURC
10 CFR 61
Class
B
A
B
A
C
A
A
A
A
A
A
A
A
A
' A
A
A
A
A
A
C
C
B
B
C
C
Waste
form5
AW
AW
AW
TR
AW
AH
AW
AW
TR
TR
TR
TR
TR
TR
TR
TR
AW
AW
AW
AW
TR
AH
TR
AM
AW
AM
Volume
(m3)
9.91E+04
3.31E+05
1.31E+05
1.28E+04
2.24E+03
6.45E+04
5.95E+04
2.14E+04
5.98E+05
4.78E+05
1.79E+05
3.17E+04
2.82E+05
1.01E+05
3.59E+05
6.34E^4
1.50E+04
1.11E+04
7.52E+03
6.03E+04
9.97E+03
5.82E-f02
6.94E+03
2.23E+02
6.60E+03
4.45E-01
Total Volume
2.93E+06
aSee Table 3-2 for description.
^As-generated waste form:
AW Absorbing Waste
AM Activated Metal
TR Trash
9-3
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Table 9-2. Estimated total activity of major radionuclides
in commercial LLW and NARH, 1985 - 2004
Nuclide
H-3
C-14
Fe-55
Ni-59
Co-60
Ni-63
Sr-90
Nb-94
Tc-99
Ru-106
Sb-125
1-129
Cs-134
Cs-135
Cs-137
Ba-137m
Eu-154
Activity*
1.80E+06
5.88E+03
3.99E+06
2.66E+03
3.35E+06
3.61E+05
7.33E+05
2.70E+01
2.49E+01
4.00E+03
5.68E+03
6.95E+01
6.62E+05
2.47E+01
9.66E+05
9.66E+05
5.70E+02
Nuclide
Po-210
Pb-210
Bi-214
Pb-214
Ra-226
U-234
U-235
Np-237
U-238
Pu-238
Pu-239
Pu-241
Am-241
Pu-242
Am-243
Cm-243
Cm-244
Activity*
6.82E+02
6.82E+02
6.82E-h02
6.82E+02
7.41E-1-02
8.21E-h01
2.85E+00
1.55E-04
3.41E+01
1.13E+03
4.12E-»02
1 . 76E+04
1 .68E+03
8.55E-01
2.53E+01
2.56E-»01
3.33E4-02
Total Activity: 1.29E+07 Ci.
*Activity is in Ci.
9-4
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These particular waste streams have low concentrations of radionuclides.
Table 9-3 lists the remaining waste streams and estimates the volume of
each for a disposal site capacity of 2.5E+05 m3. This reference site
capacity is derived by consideration of the total LLW volumes presented
in Table 9-1 divided by an assumed number of 10 to 12 disposal sites
formed under the LLRWPAA of 1985. While the number of State compacts may
vary, present indications are that there will be quite a few such
compacts. Exclusion of the seven "designated" BRC waste streams reduces
the 20-year volume of regulated LLW from about 3E+06 m3 to about
2.2E+06 m3 (Table 9-1). Rather than reduce site capacity accordingly,
the base case analysis presumes that the 2.5E+05 m3 site is filled to
capacity with the remaining waste streams shown in Table 9-3. Chapter 11
provides additional analyses of the implications of including or
excluding certain categories of LLW (e.g., BRC, WARM). In general,
excluding BRC wastes from the site inventory while simultaneously
allowing the corresponding volume to be filled with the remaining higher
activity LLW results in slightly higher CPG doses and health effects.
9.2.2
Hvdroqeologic/Climatic/Demoqraphic Conditions
The hydrogeologic and climatic conditions at a site can directly
affect and change the importance of pathways and impacts of releases from
a LLW disposal facility. Because the LLW standards will be applied to
LLW facilities throughout the United States, they must be applicable
under a wide range of conditions. Therefore, we have conducted all of
our base case health impact assessments under three very different
regional hydrogeologic/climatic scenarios. These three scenarios were
also used for our sensitivity analyses of the health impact assessment
codes.
The three hydrogeologic/climatic scenarios used are for sites in
humid permeable, humid impermeable, and arid permeable regions.
Figure 8-6 illustrates the environmental pathways for the transport of
water (and radionuclides) at these three hydrogeologic/climatic
settings. Realistic site data, which are typical for these scenarios,
were obtained from USGS studies at the LLW disposal facilities at
Barnwell, West Valley, and Beatty. General population distributions and
water usage patterns typical of these three climate settings were also
used. These scenarios span a wide range of conditions under which a
disposal facility would normally be sited in the continental United
states, chapter 5 provides more detail concerning these
hydrogeologic/climatic/demographic settings.
9.2.3 Disposal Methods
Nine different disposal methods are defined for the overall analysis
of the land-based disposal of LLW. Conceptual designs for these disposal
methods are developed in sufficient detail to estimate disposal costs and
health impacts. The disposal methods chosen for analysis represent a
wide range of technological sophistication and would need little or no
9-5
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Table 9-3. Input data for EPA's base case analysis: commercial
LLW and NARK disposed of in a regulated disposal
facility, 1985-2004
Waste stream3
L-IXRESIN
l-CONCLIQ
L-FSLUOGE
P-FCARTRG
L-DECONRS
L-NFRCOMP
L-COTRASH
L-NCTRASH
I-COTRASH
N-LOTRASH
I-ABSLIQD
I-BIOWAST
N-LOWASTE
N-ISOPROD
N-SOURCES
N-TRITIUM
N-TARGETS
R-RAIXRSN
R-RASOURC
10 CFR 61
Class
B
A
B
A
C
A
A
A
A
A
A
A
A
C
C
B
B
(NARM)
(NARH)
Waste
fornr1
AW
AW ;
AW
TR
AW
AM
TR
TR
TR
TR
AW
AW
AW
TR
AH
TR
AM
AW
AM
Total Site Volume
Volume
(m3)
1 . 13E+04
3.75E+04
1.48E+04
. U>46E+03
2.54E+02
7.32E+Q3
6.78E+04
5.43E+04
3,20E+04
1.15E+04
1.26E+03
8.54E+02
6.85E+03
1 . 13E+Q3 - •
6.61E+01
7.88E+02
2.54E+01
7.49E-»J02
5.05E-02
2.5E+05
aSee Table 3-2 for description.
bAs-generated waste form:
AW Absorbing Waste
AM Activated Metal
TR Trash
9-6
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engineering development. A special tenth disposal option is also
presented to illustrate current disposal practice according to the
requirements of the 10 CFR 61 disposal technology. These regulations
actually imply the use of a combination of two of the nine basic disposal
methods listed in Table 9-4, as described below.
Disposal alternatives chosen to represent the results of the base
case analysis include those that can handle all 26 LLW streams, represent
a wide range of technological sophistication, and typify to some degree
past, present, and potential disposal techniques. They include:
1. Regulated Sanitary Landfill;
2. Shallow Land Disposal;
3. improved Shallow Land Disposal;
4. Current Disposal Practice: 10 CFR 61;
5. Intermediate Depth Disposal; and
6. Concrete Canister Disposal.
A regulated SLF is the simplest land disposal technology analyzed.
This would essentially be a landfill operating in accordance with EPA
regulations, 40 CFR 241 to 257, but with the additional features required
for a facility authorized to receive, handle, and dispose of radioactive
materials. SLD represents LLW disposal as practiced between 1963-1980 in
the United States. ISO incorporates all of the requirements of 10 CFR
61, but places all classes of LLW in narrower, deeper trenches than those
used by SLD. Current disposal practice according to 10 CFR 61 disposal
technology is actually a combination of the SLD and ISO methods. Class A
and Class B wastes are disposed of in separate disposal trenches typical
of SLD, while Qlass C wastes are disposed of in trenches typical of ISD.
IDD meets the requirements of 10 CFR 61 disposal technology, but places
the waste even deeper than ISD (i.e., 15 meters vs 8 meters below ground
level). Finally, the CC method is analyzed. This method has been
designed and engineered within the last few years and emphasizes concrete
as a barrier to: limit, radioactivity releases to the environment. All LLW
received at the site is re-packaged into standardized hexagonal concrete
containers, or canisters. Void spaces are filled with grout, creating a
solid hexagonal container.
Four other disposal methods are analyzed but not included in the
base case results. They are: ,
1. Hydrofracture;
2. Deep-Well Injection;
3. Deep Geological Disposal; and
4. Earth-Mounded concrete Bunker.
The results for these disposal methods are listed in chapter 11. The
first three methods are deemed appropriate only for selected waste
streams. The EMCB technique is used in France. More details on all of
these disposal methods are presented in Chapter 4.
9-7
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Table 9-4. Input data for EPA's LLW risk assessments:
Disposal options and waste form
Disposal option
Pi sposal
Acronym
Regulated Sanitary Landfill SLF
Shallow Land Disposal (as practiced from SLD
1963-1982)
Improved Shallow Land Disposal ISO
Current Disposal Practice 10 CFR 61
(Combination of SLD and ISO)
Intermediate Depth Disposal IDD
Hydrofracture HF
Deep-Well Injection DWI
Deep Geologic Disposal DGD
Concrete Canister CC
Earth-Mounded Concrete Bunker EMCB
Pretreatment
Waste form option Acronym
Packaged As Generated AG
Solidified S
Incinerated, Then Solidified i/s
Packaged in a High Integrity
Contai ner HIC
9-8
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In addition to the various disposal methods, the form of the waste
being disposed of may also vary for a given disposal method. Table 3-7
illustrates the basic waste forms analyzed and the 10 CFR 61 disposal
technology waste class for each of the 26 waste streams. The base case
analysis relates two categories of waste form to the base case disposal
methods shown in Table 9-5. The "as-generated" waste form designation
actually encompasses any one of three basic waste forms (trash, activated
metal, and absorbing waste) for a given waste stream, reflecting minimal
treatment and packaging. Use of the 10 CFR 61 disposal technology waste
classes in Table 9-5 simplifies the description of matching waste streams
to disposal methods as well. For example, according to Table 9-5, all
Class A waste streams are in the "as-generated" waste form for a
regulated SLF. Referring back to Table 3-7, the Class A waste streams
are identified, along with the basic "as-generated" waste form for each
Class A waste stream. Appendix B presents the NRC's waste classification
system as set forth in 10 CFR 61.55. Chapter 11 (Sensitivity Analysis)
presents the results for other combinations of waste forms and disposal
methods.
9.2.4 Health Impact Assessment Codes
The original PRESTO-EPA model was developed jointly by EPA and Oak
Ridge National Laboratory for use in the LLW standard development effort
(EPA83). This model, which was completed in 1983, was expanded by EPA
and Rogers and Associates Engineering Company into a family of health
impact assessment codes in order to estimate such impacts from the
disposal of regulated LLW under a wider range of conditions. The
following derivatives of PRESTO-EPA have been developed:
PRESTO-EPA-POP
Cumulative health effects to local and regional
populations from land disposal of LLW by shallow
methods; long-term analyses are modeled (10,000
years).
PRESTO-EPA-CPG Maximum annual committed effective whole-body dose
equivalent to a critical population group (CPG) from
land disposal of LLW by shallow or deep methods;
dose in maximum year is estimated. Maximum annual
CPG dose can be converted to an annual or lifetime
risk using appropriate conversion factors.
Chapter 8 provides detailed descriptions of the above health impact
assessment codes.
9.3 Summary of Base Case Analysis
As indicated above, the purpose of the base case analysis is to
estimate and compare the health impacts from the undisturbed performance
of engineered LLW disposal systems. The disposal systems examined
include disposal methods requiring little or no further engineering
9-9
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Table 9-5. Base case analyses of LLW disposal
10 CFR 61
waste class
SLF
Waste forms assigned to disposal methods
SLD
SLD
SLD ISO
(10 CFR 61)*
ISO
IOD
CC
Class A AG AG AG AG
Class B AG AG S S
Class C, NARH AG AG S
AG
S
S
AG
S
S
Notes:
1. Abbreviations Used:
SLF = Sanitary Landfill
SLD = Shallow Land Disposal
ISO = Improved Shallow Land Disposal
IDD = Intermediate Depth Disposal
CC = Concrete Canister Disposal
AG = "As-Generated" Waste Form
S = Solidified
2. See Table 3-7 for the relationship between "waste class" and specific waste
streams.
*10 CFR 61 disposal technology incorporates practices from both SLD and ISO.
9-10
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development, methods applicable for all types of LLW, and methods
illustrative of a wide range of technological sophistication.' They are
SLF, SLD, ISO, 10 CFR 61, IDD, and CC. The health impact assessments
presume a disposal site with a capacity of 2.5E-H-05 m3 of LLW.
The base case analysis source term is derived from the overall LLW
source term presented in Chapter 3. Modifications include the
elimination of seven lower activity LLW streams from those presently
regulated under AEA authority to reflect implementation of a BRC
criterion, and the inclusion of two high concentration NARM waste streams
to reflect inclusion of these NARM wastes as regulated wastes. Table 9-3
indicates the volume of each waste stream contributing to the
2.5E+05 m3 disposal capacity of the model site.
Having defined the waste streams associated with the model site and
the disposal methods designed to accept such wastes, it is also necessary
to define the waste form associated with each disposal method. Table 9-5
relates the waste form (according to the 10 CFR 61 waste classification)
to the disposal method. (Note that Tables 9-1 and 9-3 provide a listing
of the LLW streams and their corresponding 10 CFR 61 waste class.)
Selection of waste form was first made for the 10 CFR 61 disposal method
using the waste form requirements therein. Disposal methods that are
less sophisticated generally are assigned the simpler waste forms.
Likewise, more sophisticated disposal methods are assigned waste forms
comparable to, or more engineered than, those employed by 10 CFR 61
disposal technology requirements.
The base case health impact assessments are then carried out for the
matrix of disposal method/waste form combinations shown in Table 9-5.
Each disposal method/waste form combination is evaluated for all three
settings. Thus, each disposal method/waste form/hydrogeologic setting
combination is evaluated in terms of maximum annual CPG dose and
cumulative population health effects using the appropriate health impact
assessment computer code (i.e., PRESTO-EPA-CPG or PRESTO-EPA-POP). For
example, Table 9-5 indicates that the base case source term, Table 9-3,
is to be evaluated in the "as-generated" waste form for all classes of
LLW (as defined by 10 CFR 61) disposed of in a regulated SLF. By varying
the hydrogeologic input data, maximum annual CPG doses are calculated for
a regulated SLF located in each of the three different hydrogeologic
settings. Similarly, cumulative population health effects are also
evaluated for an SLF located in each of the three hydrogeologic
settings. Corresponding health impact assessments are then carried out
for the remaining disposal method/waste form combinations in Table 9-5.
The results of this base case analysis are described in the following
sections.
9.4 Results of Base Case Health Impact Analyses
The purpose of the EPA assessment of LLW disposal is to compare
potential risks from LLW disposal methods. In so doing, EPA has taken
9-11
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great care to use input data as realistic as possible and health impact
assessment computer codes that are "state-of-the-art." None of the
predicted population health effects or CPG doses described below should
be taken to be predicted impacts from any existing or future site.
Site-specific predictions would require site-specific assessment code(s)
and site-specific engineering, hydrogeologic/climatic, and waste stream
data. The results presented below reflect the undisturbed performance of
natural and engineered barriers.
9-4.1 Health Effects to the General Population
Figure 9-1 summarizes the potential health effects incurred by a
general population from disposing of 2.5E+05 m3 of regulated LLW by six
different disposal methods in three different hydrogeologic/climatic
settings. These disposal methods include SLF, SLD, 10 CFR 61, ISD, IDD,
and disposal using CC. As shown in Figure 9-1, these disposal methods
incorporate more sophisticated waste forms as the sophistication of the
disposal method increases. Two different waste form cases are considered
for SLD. Current disposal practice under 10 CFR 61 is modeled as a
combination of the SLD and ISD methods.
In the humid permeable setting, estimated total health effects
(fatal cancers and-genetic effects) ranged from 7.1 for SLF, the least
stringent disposal method, to 1.9 for CC disposal, the most highly
engineered method, with 4.4 health effects for 10 CFR 61 disposal. These
health effects were incurred primarily by the regional basin population
via the ground-water pathway, and usually occurred within the first
500 years.
In the humid impermeable setting, estimated total health effects
ranged from 47 for SLF to 0.3 for CC disposal, with 2.5 health effects
for 10 CFR 61 disposal. Again, the majority of the health effects were
incurred by the regional basin population. However, the major release
pathway shifted from the ground-water pathway to surface water by direct
overflow onto the land surface because of the "bathtub" effect (Me76).
In the direct overflow case, because of the impermeable disposal medium,
both mobile and less mobile radionuclides were released more quickly than
would have been expected by the ground-water pathway.
In an arid permeable setting, estimated total health effects ranged
from 4.4 for SLF to 0.4 health effects for CC disposal, with 2.6 health
effects for the 10 CFR 61 disposal technology. As in the case of the
humid permeable settings, essentially all of the health effects were
incurred by the regional basin population via the ground-water pathway.
In this case, however, they occurred during the second thousand years
rather than the first 500 years because of the much lower rainfall and
therefore the much smaller flux of water entering the trench to leach the
wastes. This, in turn, caused a much smaller flux of water to leave the
trench. A much thicker unsaturated zone between the trench bottom and
the aquifer also provides additional delay. Once the contaminant reached
the aquifer, however, it moved at the same rate as the ground water, less
any retardation from exchange with geologic media.
9-12
-------
50-
(0
5 40-
Ul
o
o
o
o
3 fl-
oe
UJ
o
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1-
O 20-
UJ * w
u.
u.
Ul
X
2 10-
X
0-
DISPOSAL CLASS A
OPTION: 9.LASS ?
WASTE sSsr"
FORM: CLASS C, NARM
47
i:
"* t
'• •:
t
\
^ ,
\
1
:|
s* v"
"^
HUMID IMPERMEABLE
F/3 HUMID PERMEABLE
l^^j
[""] ARID PERMEABLE
^~~^
10 CFR 61
7 1 ??*
• • • lj?S?!5
V/ I "'.v//r~~\ 2.Syy2_B K/^[2.6 rjni 3 .8
// I •• '%f yyi I '^ 1 **~ryj>i i 1 • 2 _-^f * 1 flry/ii .3 <\ « i . 9 « ^
y/ I % v i: (//l 1 // 1 WmfsA 1 88& vl 1 psBiyyi"! i .. Kyj- — i
SLF SLD SLD SLD ISD IDD CC
SLF SLD SLD SLD ISD IDD CC
S' " S' D SLD !SD ISD IDD CC
AG AG AG AG AG AG s
AG AG 8 8 ss s
AG AG 8 88 8 S
•REFER TO TABLE 9-3 FOR DEFINITION OF ACRONYMS.
Figure 9-1.
Comparison of Population Health Effects over 10,000
Years by Disposal Options for a Reference Disposal
Facility Containing 250,000 m 3 of Regulated LLW
-------
Both disposal method and hydrogeologic setting influence the
magnitude of predicted population health effects. Analysis of Figure 9-1
shows that for simpler disposal methods such as SLF, the humid
impermeable setting results in the greatest number of health effects.
This method provides for minimal containment of all radionuclides that
are able to reach numerous surface water pathways in a relatively short
time because of trench overflow, or the "bathtub effect." However, with
enhanced waste form technology, such as solidification, radionuclide
containment is improved. When both waste form and disposal method are
improved, such as for 10 CFR 61 disposal technology or more advanced
methods, radionuclide containment at the humid impermeable site is much
improved. In such cases, the humid permeable setting results in a
greater number of health effects.
It is also of interest to examine the distribution of cumulative
population health effects over time for the whole United states.
Nationwide impacts may be calculated by weighting the results of the
three hydrogeologic settings by the percentage of LLW expected to be
disposed of in each setting. Considering only the 10 CFR 61 waste
technology disposal option, the following illustrates the proportion of
nationwide health effects projected to occur over the indicated time
periods:
Time period (vrs)
0 -
101 -
501 -
1,001 -
100
500
1,000
10,000
Total:
0 - 10,000
% of total
health effects
5.7
37.5
9.8
47.0
100.
For less sophisticated disposal methods, total health effects are
greater, and a much larger fraction of total health effects occurs in the
first 500 years. For more advanced disposal methods and waste forms, the
bulk of total health effects is slightly less and occurs later, in the
If001- to 10,000-year time frame. Such advanced technology provides for
greater containment, allowing decay of short-lived radionuclides and very
slow release of the remaining long-lived radionuclides.
Table 9-6 shows the critical radionuclides and their relative
contribution to total health effects for the 3 different hydrogeologic
settings using 10 CFR 61 disposal technology and waste form
requirements. Carbon-14 is the dominant radionuclide at all the
settings. Only at the humid impermeable setting does any other
radionuclide make a substantial contribution, namely, Am-241. The humid
impermeable setting allows radionuclides to reach surface water pathways
(due to the "bathtub effect"). Thus, radionuclides that would normally
be retarded in their movement through the soil reach the surface via
9-14
-------
Table'9-6. Critical radionuclides at a model LLW site
Hydrogeologic
setti ng
Humid
Permeable
Humid
Impermeable
Arid
Nuclide
1-129
C-14
Other
1-129
C-14
Anv-241
Other
C-14
Other
Percent of
CPG
(exposure)
93%
7%
781
22%
100%
total impact
Population
(health effects)*
90%
10%
70%
20%
10%
95%
5%
Note: Model site assumes 250,000 m3 of regulated LLW using 10 CFR 61
disposal technology methods.
*Approximate values.
9-15
-------
overflow and become more readily available to pathways affecting human
populations.
9.4.2
Exposure of Critical Population Groups
Figure 9-2 summarizes the estimated maximum annual effective
whole-body doses to a CPG living within a few tens of meters of a
standard reference disposal facility containing 2.5E+05 m3 of regulated
LLW. Estimated exposures are calculated for the same 6 disposal methods
and waste form combinations as for the preceding health effects
assessments. Analyses are terminated at 1,000 years for all of the
standard CPG assessments, but in some sensitivity runs, analyses are
extended to 10,000 years to look for possible significant exposures
beyond 1,000 years. This is discussed in more detail in Chapter 11.
In the humid permeable setting, estimated maximum CPG doses range
from 62 mrem/yr for SLF to 1.3 mrem/yr for CC disposal, with 9.2 mrem/yr
for 10 CFR 61 disposal. These maximum doses occur at 60 years for SLF,
780 years for 10 CFR 61 disposal, and at 1,000 years for CC disposal. As
shown in Table 9-6, c-14 and 1-129 are the principal radionuclides
contributing to CPG dose for 10 CFR 61 disposal technology. Less
sophisticated disposal methods allow short-lived radionuclides to be
released sooner, along with long-lived radionuclides, resulting in larger
maximum CPG doses at earlier times. For example, use of SLD method
without solidification of any wastes results in a maximum CPG dose of
about 35 mrem/yr at 30 years. In this case, the peak dose is caused
primarily by H-3 and C-14. over time, other radionuclides are released,
but they result in annual CPG doses lower by about a factor of 4. These
long-terra CPG doses are attributed to the slow release of long-lived
radionuclides, particularly 1-129. chapter 11 provides a graphical
representation of CPG dose over the first 1,000 years for SLD in all
3 hydrogeologic settings. The benefits of solidification are apparent by
comparing the two SLD cases depicted in Figure 9-2. on the other hand,
disposal methods more sophisticated than the 10 CFR 61 disposal method
modeled do not appreciably reduce maximum CPG dose, except for the CC
disposal.
In the humid impermeable setting, estimated maximum CPG doses are
all less than 1 rarera/yr, ranging from 0.8 to 0.001 mrem/yr for SLF and
CC disposal, respectively, with 0.03 mrem/yr for 10 CFR 61 disposal.
Exposures are due to releases to land surface and subsequently to surface
waters, again due to the "bathtub" effect caused by the low permeability
disposal medium. The year of maximum exposure occurs at 24 years for
SLF, 190 years for 10 CFR 61 disposal technology, and 250 years for
CC disposal. These 190- to 250-year periods for 10 CFR 61 and CC
disposal include 100 years of institutional control, during which it is
assumed that the trench covers are maintained intact, with the balance of
the time for the caps to fail and the trenches to fill and overflow. No
active maintenance on the cover is assumed for SLF. The critical
radionuclides for the humid.impermeable setting are C-14 and 1-129.
Though the peak CPG dose is short-lived and low (i.e., less than
9-16
-------
I
I-1
~J
g
§
Ul
(0
o
Q
O
OL
O
UJ
Q.
60-1
50-
40-
30-
20-
10-
0.5
7?
Y/
\
'//
I
35
0.4 0.1
9.1
0 0 0.0 'ViO.O
l||| HUMID IMPERMEABLE
V/\ HUMID PERMEABLE
r~| ARID PERMEABLE
10 CFR 61
[.2
^ 5.1 5.0
^ ^ ^ 1-3
o.o/J/o.o o.o y^o.o o.oi//|o.o O.OTT/IO.O
DISPOSAL
OPTION:
CLASS A
CLASS B
CLASS C, NARM
CLASS A
WASTE CLASS B
FORM: CLASS C, NARM
SLF
SLF
SLF
AG
AG
AG
SLD
SLD
SLD
AG
AG
AG
SLD
SLD
SLD
AG
8
8 '
SLD
SLD
ISO
AG
8
8
ISD
iSD
ISD
AG
8
8
IDD
• V%fl^
IUU
IDD •
AG
8
8
cc
s* r*
\* \s
CC
8
s
s
FIGURE 9-2.
Comparison of Effective Whole-Body Dose to Critical
Population Group by Disposal Options for a Reference
Disposal Facility Containing 250,000 m3 of Regulated LLW
-------
0.1 mrem/yr), the expected chronic CPG dose over the remaining time
period is expected to be relatively constant. For SLD, the chronic CPG
dose is expected to remain at about 10 percent of the peak CPG dose for
many hundreds of years.
The estimated maximum exposures were all less than 1 mrem/yr for the
arid permeable setting. They ranged from 0.4 mrem/yr for SLF to almost
zero for cc disposal (calculations terminated at 1,000 years) with
0.007 mrem/yr for 10 CFR 61 disposal technology, in the arid setting,
the principal radionuclide contributing to CPG dose is c-14, as shown in
Table 9-6. Peak doses occur much later in the arid settings. For
example, for SLD without waste solidification, the peak dose is very low
(less than 0.1 mrem/yr) and occurs at about 950 years. More
sophisticated disposal methods result in maximum CPG doses via ground
water occurring even later.
As indicated for population health effects estimates, these results
should not be considered as applying to any specific facility, either
present or future. They are only applicable to generic comparisons of
methods. Also, the absolute values have considerable uncertainty
associated with them even within the context of a generic analysis, and a
factor of 2 variability in the scale would not be unreasonable, chapter
12 discusses uncertainty in more detail.
It is interesting to observe the relationship between maximum CPG
dose and estimated health effects for the disposal methods and waste
forms analyzed for the base case (see Figures 9-1 and 9-2). For those
methods where waste is not solidified (i.e., the as-generated waste
form), the maximum CPG dose occurs in a different hydrogeologic setting
than that corresponding to the maximum health effects. Typically, the
maximum CPG dose occurs in the humid permeable setting, whereas maximum
health effects occur in the humid impermeable setting. For these less
sophisticated methods, greater amounts of short-lived radionuclides are
released. For the humid permeable setting, the bulk of such a release is
permitted to flow into the ground water and reach a well, creating a
relatively high concentration in ground water (and well water), resulting
in a large CPG dose. However, the population using such a well is
relatively small compared to the overall population .at risk, and thus
total health effects also remain relatively small, in the humid
impermeable setting, however, virtually no radionuclides migrate via the
ground water to well pathway because of the low permeability soil.
Instead, many more radionuclides are forced to the surface via trench
overflow, dispersing among numerous surface water and airborne pathways.
Dilution afforded by these numerous surface pathways ensures a lower CPG
dose, but the short-lived and wider variety of radionuclides forced to
the surface result in the exposure of much larger populations, thus
producing greater health effects.
For disposal methods where some or all of the waste is solidified,
the maximum CPG dose occurs in the same hydrogeologic setting as for
maximum health effects (humid permeable). As the sophistication of the
disposal method increases, however, the reduction in maximum CPG dose and
health effects is generally modest.
9-18
-------
Note that the results provided above for CPG exposures and general
population health effects represent the base case analysis of the impacts
from LLW disposal. Many additional analyses have been performed to
investigate the effects of varying key input or assumptions. These
special analyses are presented in Chapter 11 and examine other disposal
methods, site size, special waste treatment options, the effects of
regional compact waste volumes and characteristics, and many other
factors that may have a bearing on the base case analysis presented here.
9-19
-------
REFERENCES
RPA83
Gr86
Me76
NRC86
PEI85
PHB85
U.S. Environmental Protection Agency, PRESTO-EPA: A Low-Level
Radioactive Waste Environmental Transport and Risk Assessment
Code - Methodology and User's Manual, Prepared under Contract
No. W-7405-eng-26, Interagency Agreement No.
EPA-D—89-F-000-60, U. S. Environmental Protection Agency,
Washington, DC, April 1983.
Gruhlke, J.M., EPA Source Term for Low-Level Radioactive Waste
Risk Assessment, Office of Radiation Programs, U.S.
Environmental Protection Agency, Washington, DC 20460, 1986
(draft).
Meyer, G.L., Recent Experience with the Land Burial of Solid
Low-Level Radioactive Wastes: Management of Radioactive Wastes
from the Nuclear Fuel Cycle - Volume II, IAEA-SM-207/64,
pp. 383-394, International Atomic Energy Agency, Vienna.
Austria, 1976.
U.S. Nuclear Regulatory Commission, Update of Part 61 Impacts
Analysis Methodology, NUREG/CR-4370, Washington, DC,
January 1986.
PEI Associates Inc., Radiation Exposures and Health Risks
Associated with Alternative Methods of Land Disposal of Natural
and Accelerator-Produced Radioactive Materials (NARM), (Draft)
performed under Contract No. 68-02-3878 for the u. S.
Environmental Protection Agency, October 1985.
Putnam, Hayes, & Bartlett, Projected Waste Volume by State and
Compact, 1985-2004, Data transmitted from Charles Queenan,
Putnam, Hayes, & Bartlett to James M. Gruhlke, office of
Radiation Programs, U.S. Environmental Protection Aqencv.
August 1, 1986.
9-20
-------
Chapter 10: THE ESTIMATED HEALTH IMPACT ASSESSMENT OF DISPOSAL OF BRC WASTES
10.L Introduction
The health impact assessment associated with the nonregulated disposal of
those wastes considered "Below Regulatory Concern" (BRC) is a major factor in
developing proposed disposal criteria for use by other regulatory agencies.
The risk assessments carried out are intended to be generic in nature. To
begin to determine the health impacts resulting from BRC waste disposal, we
developed several scenarios (described in Chapter 4 for ^deling based on
surrogate waste streams (described in Chapter 3), disposal methods (described
in Chapter 4), and various hydrogeologic/climatic settings (described in
Chapter 5).
The candidate or surrogate waste streams were chosen to represent .a
postulated .set of BRC waste types. These surrogate waste types originate
from a variety of waste generators (power reactors, institutional, industrial
facilities, etc.). For the sake of analysis, these surrogate wastes are
declared BRC and therefore qualify for less restrictive disposal P^ctices.
In order to scope the range of cumulative impacts from many S""OS^ B*Lg
waste types, numerous realistic scenarios or combinations of BRC waste types
and disposal methods were constructed (see Chapter 4, Section 4.4). For each
scenario, the radionuclide source term was defined, as well as the numerous
parameters necessary to define the potential pathways of human exposure to
radiation. . .
These data servers the input to the PRESTO-EPA-BRC and PATHRAE-EPA
methodologies developed specifically to calculate the health impacts
consisting of cumulative population health effects and maximum CPG risk (see
Chapter 8).
10.1.1 Wastes.- ... . .• . . .
To determine if it was feasible to allow some types of LLW to be BRC
wastes a number of waste streams were identified that had very low
radio-activity, were reasonably well characterized, and had P0^^^
volumes to provide a cost savings. We chose 18 waste streams (see Table 3 10)
generated at nuclear power reactors, uranium fuel fabrication and uranium
process facilities, and industrial, medical, and educational facilities, as
well as by consumers (see Chapter 3). To make our work comparable with that
of others, EPA's BRC waste sources and volumes are based on waste
characterizations done by NRC, the AIF, and others (NRCSlb, NRC86, NRC84
Oz84, AIF78). Two of the surrogate consumer waste streams, smok^f 6^°"
and timepieces, were chosen because they presently are not £S«JjgJ. and they
help to provide a comparison and perspective for our analysis (NRC80).
Ano'ther s^ial deregulated waste stream (BIOMED) modeled after the upper
limits of the NRC biomedical rule (NRC81. , «** *lso ch°"« !°Ln 3 3 3(A)(3)]
incineration scenarios for further comparison [Chapter 3, Section 3.3.3UUUJJ
10-1
-------
10.1.2 Disposal Methods
Once a group of surrogate BRC waste streams was selected, it was
necessary to determine reasonable disposal methods for such wastes
Consideration of the numerous generators represented by the surrogate BRC
wasce types indicated the very real possibility that a given waste disposal
site might receive BRC wastes from more than one generator. To account for
this, several realistic scenarios were constructed involving the disposal
of various BRC waste streams from different generators. The choice of both
surrogate BRC waste streams and disposal scenarios was made for the purposes
of the assessment. EPA is not implying that these are the only streams or
the only disposal scenarios available, but rather that they are the most
likely, considering the types of generators and their locations (see
Chapter 4, Section 4.4). Each scenario combined a generic BRC waste disposal
method with selected groups of surrogate BRC waste types. Generic BRC waste
disposal methods included a variety of options, i.e., sanitary landfills
municipal dumps, onsite landfills, and incineration methods situated in rural
suburban, and urban demographic settings. - Chapter 4 discussed the selection '
ot these methods and their varying parameters. Table 10-1 shows the major
characteristics of the disposal methods and associated demographic settings.
10.1.3 Hydrogeologic/CIimatic Settings
In developing the scenarios to model the disposal methods, three
nydrogeologic/climatic settings were used that we believe cover the expected
range of values for parameters affecting radionuclide retention and site
performance anywhere in the United States. The settings include: (a) an arid
zone site with permeable disposal medium (water infiltrating through the waste
trench into the ground and radionuclides moving very slowly to ground water)-
tb; a humid zone site with permeable disposal medium (water infiltrating ' /
through the waste trench into the ground and radionuclides moving more rapidly
to ground water); and (c) a humid zone site with impermeable disposal medium
(.water infiltrating into the waste trench and radionuclides potentially
overflowing to surface waters rather than moving to the ground water).
1°'2 Selection of Health Impact Assessments
^ Health impacts were estimated for (a) the general population and (b) the
critical population group (CPG) (including both .onsite arid offsite workers).
This impact consists of cumulative population health effects and a maximum CPG
whole body effective annual dose equivalent. The assessment of impacts from
the disposal of BRC wastes involves the simulation of the transport of
radionuclides through geological, atmospheric, and ecological systems, and the
evaluation of human organ doses and fatal cancer risks after ingestion and
inhalation of radionuclides.
1°'3 Cumulative Population Health Effects Assessment
Cumulative population hea.lth effects were estimated as both fatal cancers
and serious genetic effects for both the local and regional basin general
populations over 100, 500, 1,000, and 10,000 years.
10-2
-------
1.
2.
3.
4.
5.
6.
Table 10-1. Major characteristics of BRC waste disposal methods
MD - Municipal Dump (pop. served = 60,000)
capacity = 2.1 million m3
size = 35 ha
SF - Suburban Sanitary Landfill (pop. served = 175,000)
capacity = 6.0 million m3
size = 100 ha •
UF - Urban Sanitary Landfill (pop. served = 1,000,000)
34. 7 million m3
size = 576 ha
capacity
LURO - Large University/Medical Center with Onsite Landfill
and Onsite Incineration (pop. served =175,000)
capacity = 0.17 million m3
incinerator at disposal site
size = 2.8 ha
SI - Suburban Sanitary Landfill with Onsite Incineration
(pop. served = 175,000)
capacity = 1.0 million m3
incinerator at disposal site [aggregate VRF - 6.0]
size = 16 ha
UI - Urban Sanitary Landfill with Onsite Incineration
(pop. served = 1,000,000)
capacity = 5.78 million m3
incinerator at disposal site [aggregate VRF =6.01]
size = 96 ha
10-3
-------
Fifteen scenarios were defined and used to assess the consequences to the
general population of deregulated or less restrictive disposal (i.e. BRC) of
some radioactive wastes. The scenarios consist of combinations of radioactive
wastes, disposal methods and diverse demographic, climatic, and hydrogeologic
settings. The, health effects' to the general population resulting from
disposal of radioactive waste streams without regard to radioactivity have
been estimated using the PRESTO-EPA-BRC computer model (Ro84, EPA87a) as
described in Chapter 8. '
The cumulative population doses and resulting health effects are
separated into projections for a local population during the first 1,000 years
10 n£n yS1S' ^f pr°Jections for a regional basin population during the entire
10,000-year analytical period. After the first 1,000 years, the local
population is assumed to become part of the larger regional basin population.
The magnitude of local versus regional basin health effects is highly variable
and either may predominate, depending on the waste streams and radionuclides
present, local and regional water uses, and site-specific hydrogeology,
climate, and demographics. (See Chapter 8, Section 8.3.5 for more detail on
local and regional analysis.)
effects °°CUr thro"8h various pathways, including ingestion,
immersion, and surface (gamma radiation) exposure. In most cases
ingestion of radionuclides by their presence in contaminated ground and
surface waters, either directly or through ingestion of contaminated food
appears to be the predominant pathway and accounts for greater than 90 percent
of the total projected health effects. Ground surface exposures can
predominate in certain situations. For example, in the case of the arid
southwest climate with permeable soil, it is hypothesized that lack of
rainfall infiltrating the trenches reduces pollutant transport rates and the
importance of the ground- and surface-water pathways, and enhances the air
Unhalation) and ground surface exposure pathways. The air immersion pathway
is responsible for less than one percent of predicted health effects in all
cases.
t _ Health effects include the effect of onsite exposures of workers and
visitors to the surrogate BRC waste streams through surface (gamma) exposure
and inhalation. These effects are only a small fraction of total predicted
effects and are probably due to the relatively brief operational period of the
disposal site (20 years), compared to the lengthy period used in the
post-closure analysis (10,000 years).
L0'4 Maximum Annual Dose Estimates from BRC Wastes to a Critical PC
»Group (CPG) ~—— _ _
ilat
Eleven of the 15 localized disposal scenarios described in Sections 4.4.1
through 4.4.11 have been used to assess the consequences of less restrictive
disposal (i.e., BRC) of some radioactive wastes in terms of a maximum annual
dose to a Critical Population Group. The four reference disposal scenarios
(Chapter 4, Sections 4.4.12 th.rough 4.4.15) are not relevant to this analysis
tor regulatory considerations and were only used for comparison purposes. The
maximum whole-body dose equivalent rates to a CPG, located at or near the BRC
(EPA87b ' Sre eStimated using the PATHRAE-EPA computer modeling program'
10-4
-------
The individual exposures were calculated as maximum annual radiation dose
and year of occurrence over 10,000 years for the CPG. For those otisite
individuals living close to the disposal site, the major pathway is via water
from a well or stream a few tens of ' meters from the site boundary. Other
individual exposures were also estimated for incinerator disposa operations
garbage collectors who might collect the wastes, onsite workers during routine
disposal operations, reclaimers, and off site personnel from other exposure
pathways besides water. In the case of BRC waste disposal, the onaite worker
is considered because exposure of these personnel cannot bacons trued as
occupational (i.e., radiation workers). For the PATHRAE-EPA model , it was
assumed that the only major significant human exposure pathways available are
those listed in Table 10-2. In its review (SAB85), the SAB commented that in
general believes, the [BRC] scenarios discussed. .. to be sufficient.
A special CPG pathway analysis was made to evaluate direct radiation
expotureTto transportation workers who would collect and transport BRC wastes
from generators to the various less restrictive disposal facilities. The
primary exposure pathway to the transportation workers will be gamma exposures
(PEISS! RoL). The methodology used was based upon NRC's de .minims
methodology (0.84). Figure 10,1 shows the exposure pathways evaluated for our
analyses.
10.5 Health Effects Results to the General Population
Two different assessments were made concerning health effects to the
general population. The first was based on the localized scenarios (described
in Chapter 4 Section 4.4) where EPA felt these to be realistic cases
involving the disposal of various Arrogate BRC waste streams The localized
scenarios also provide results concerning the disposal of multiple BRC waste
streams at one site.
The second assessment was done to examine a total BRC waste Disposal
m .
one site and the resulting national health effects calculated and summed (see
also EIA, Chapter 7).
10.5.1 Population Health Effects by Scenario
The general population health effects analyses for the BRC wastes using
the UlocaUzed scenarios (the 4 reference scenarios were not included for
regulatory considerations) were based on the 16 surrogate waste streams from
nuclear fuel-cycle, industrial, and institutional sources. Two of he
i? —^ •«•«•
The
site hydrogeologic/climatic settings are discussed in Chapter 5.
10-5
-------
Table 10-2. PATHRAE-EPA CPG pathways considered by which exposure may
reach humans from the less restrictive disposal of BRC wastes
1. Ground-water migration with discharge to a river.
2. Ground-water migration with discharge to a well.
3. Surface erosion of the cover material.
4. Spillage of the waste.
5. Saturation of the waste and surface-water contamination by the
bathtub effect.
6. Food grown on land.
7. Biointrusion by plant roots.
8. Direct gamma exposure.
9. Atmospheric inhalation of radioactive airborne contaminants from dust
resuspension, incinerator, or trench fire. :
10. Inhalation of radioactive dust stirred up by workers.
10-6
-------
PATHWAYS OR
SCENARIOS
o
-J
ONSITE
— EXTERNAL GAMMA FROM BURIAL OPERATIONS
— DUST INHALATION
— RECLAIMER FOOD
OFFSITE
— BIOINTRUSION
L— INHALATION OF RADIOACTIVE GAS
— EXTERNAL GAMMA FROM TRANSPORTATION
— GROUND-WATER MIGRATION TO A WELL
— GROUND-WATER MIGRATION TO A RIVER
— EROSION AND SURFACE WATER CONTAMINATION
— TRENCH OVERFLOW AND SURFACE WATER CONTAMINATION
— ATMOSPHERIC DISPERSION
Figure 10-1. Pathways Included In the EPA Analysis
-------
^ combination of surrogate waste streams and the disposal site to which
they would be shipped is based in part on regional considerations and actual
situations currently known to exist. The volumes of the surrogate waste
streams are based on probable routine quantities generated over 20 years. The
various scenario waste volumes do not represent the quantity of that waste
stream generated on a national basis.
The two surrogate consumer waste streams as described in Chapter 3 (smoke
.detectors and timepieces) were chosen because they presently are not regulated
and provide a reference comparison and perspective on our analysis. Another
r^in^of TSte stream> raodeled after the upper limits of the NRC biomedical
rule (.NRCSla), was aiso selected for a comparison and perspective on our risk
analysis. These are incineration scenarios where, in one case, 100 percent of
the waste is incinerated and in the other case 50 percent of the waste is
incinerated.
The estimated excess health effects (total cancers plus serious genetic
m ufCtf« ;r°m the surr°Sate BRC waste disposal over 10,000 years are listed in
Table 10-3 for the 11 BRC scenarios and 4 reference scenarios.
10.5.2 Population Health Effects on a Total Nationwide Basis
A population health effects assessment was performed on each of the
surrogate BRC waste streams for the 20-year total U.S. inventory of surrogate
BRC waste (the BRC wastes are all assumed to be Class A wastes as defined by
the NRC). A comparison was made between BRC disposal and the same waste being
disposed of in a regulated LLW disposal facility (referred to as SLD — see
Chapter 4). Table 10-4 lists the excess population health effects over 10 000
years from the nationwide disposal of the surrogate commercial and DOE BRC
waste streams for a 20-year accumulation of waste.
The methodology behind the compilation of the estimated nationwide
population Health effects for the BRC waste streams is presented in detail in
the EIS Volume 2 — Economic Impact Assessment (Chapters 3 and 7). Briefly
the incremental or excess health effects are calculated from the difference'
between the health effects from the current regulated disposal practice (SLD)
and the health effects for each unregulated BRC waste stream. The health
effects are determined for each of the disposal options across the three
hydrogeologic/climatic regions, and the total health effects are added
together for all three regions. To take into consideration the five
unregulated disposal options, a weighted average is used.
10*6 Results of the Maximum CPG Dose Assessments
The BRC waste disposal scenarios discussed in Section 10.5 (and described
in Chapter 4, Section 4.4) were also used for the CPG exposures. The results
of the CPG exposure assessments are in terms of millirem per year. The life
span currently used is 70.76 years and the annual risk from a 1-mrem low-LET
exposure (gamma and beta) is 3.95E-07. Therefore, the CPG can be converted
from millirem per year to lifetime risk by using the factor 2.8E-05.
10-8
-------
Table 10-3. Excess population health effects over 10,000 years from BRC waste disposal for
various scenarios, disposal sites, and hydrogeologic/ciimatic settings
I.
2.
3.
4.
5.
6.
7.
8.
9..
. 10.
11.
' *12.
*13.
*14:
*15.
.- ! Disposal scenario :
Description '
3-Unit pressurized water power reactor complex -
municipal dump . x:.. .,......„
2-Unit boiling water power reactor complex - municipal dump
University and medical center complex - urban sanitary
landfill
Metro area with fuel-cycle facility - suburban sanitary
landfill '
Metro area with fuel-cycle facility - suburban sanitary
landfill with incineration
2-Unit power reactor, institutional, and industrial -
municipal dump '
Uranium hexafluoride facility - municipal dump
Uranium foundry - municipal dump
Large university /medical center; volatilization of 90% H-3
and 75% C-14; onsite landfill with onsite incineration
Large metropolitan area with consumer wastes - suburban
sanitary landfill with incineration
Large metropolitan area with consumer wastes - urban
sanitary landfill with incineration
Consumer product wastes - suburban sanitary landfill
Consumer product wastes - urban sanitary landfill
Large university/medical center; 100% volatilization of
H-3 and C-14; onsite landfill with onsite incineration
Large university/medical center; 50% volatilization of H-3
and C-14; onsite landfill with onsite incineration
Hydrogeo logic /climatic setting
Humid
permeable
9E-02
1.3E-01
3E+01
1E+01
2.6E+00
1E+00
2.3E-04
7. IE-OS
1.5E+00
2.1E+00
1.1E+01
1E-02
1.4E-01
2.8E-01
2.5E+01
Humid
impermeable
6.6E-02
1.5E-01
2.9E-01
2.4E-01
7.3E-02
9E-02
2.2E-03
6.8E-04
4.5E-02
8.8E-02
2.1E-01
2.5E-03
3.4E-02
7.5E-03
6.2E-01
Arid
permeable
1.7E-02
2.8E-02
2.1E+01
5.9E-02
2.6E-02
1.6E-02
5.6E-04
1.6E-04
2.3E-02
3.7E-02
7.5E+00
1.9E-05
2.9E-04
4.8E-03
3.7E-01
indicates those reference scenarios where the waste streams are already deregulated.
NOTE: Analysis is based on 20 years of waste accumulation.
10-9
-------
Table 10-4.
Excess populatipn health effects over IQ.QQ0 years from
unregulated disposal of cenpereial piu$ 00S BRC waste
streams versi* regulated LLW disposal on a nationwide total
Surrogate
BRC waste stream
Excess population health effects
p
p
L
B
I
. I
I
I
N
N
N
N
F
U
• F
F
- COTRASH
- CONDRSN
- WASTOIL
- COTRASH
- COTRASH
- BIOWAST
- ABSLIQD
- LQSCNVL
- SSTRASH
- SSWASTE
- LOTRASH , •
- LOWASTE
- PROCESS
- PROCESS
- COTRASH
- NCTRASH ' -
3.8
6.0009
0*0001
c - ' - . • I ' ,
••• •• : 2.6
•••-. , • -.-371*4 ':...;:.
18.7
22.4
. ; 0.93
6f0011
0.0037
32.8
10*7
, 0.0035
6.0011
0.0006
6,0001
C - SMOKDET*
C - TIMEPCS*
Total 46^3411
'• ,' ': . i. i
7.3
*Waste streams already deregulated.
NOTE: Analysis is based on a 20-ye«r accumulation of wa^te. The reason there
are so many significant nun»|»fr? i» thai; each waste stream is considered
on a separate basis. ' ;
10-10
-------
Appendix F provides tables for all 15 localized scenarios listing the
maximum CPG exposure, the radionuclide providing the major exposure, and the
year in which the maximum CPG exposure occurs for each of the 10 major
pathways (Table 10-2) at each of the 3 hydrogeologic/climatic settings. The
transportation pathway- is not included in these analyses.'
•• • , '. • '. i>i''' ••; , '. " '.') ••*''. . ".,.'• j' • • • . . •'
;.;• Table 10-5 list's'each of the BRC' disposal scenarios, indicating the
predominant pathway,, the associated max imum,.,C,PG exposure, and the disposal
setting. Table 10-6 lists the four scenarios analyzed for reference and
comparison purposes containing the Consumer and the BIOMED wastes, showing the
pathway through which the maximum CPG exposure is delivered and the disposal
setting. Neither of these analyses include the transportation pathway.
10.6.1 Results of the Tra-nsportation CPG Dose Assessment
Based upon the NRC methodology (Oz84) and values for parameters
determined by EPA (Ro86, PEI85), external gamma doses to transportation
workers handling BRC waste streams were estimated, the primary exposure
pathway to the transportation workers is gamma exposure. Additional
short-lived nuclides (half-lives ranging from 2 days to 1 year) were included
to ensure consideration of any important nuclides exhibiting gamma exposure
during transportation (Table"3-12 lists these additional nuclides). The
estimated average annual radiation exposures for the transportation worker
exposed to BRC wastes and assuming a 30-day storage time prior to transporting
the wastes are presented in Table 10-7.
The transport analyses are centered around the BRC disposal scenarios,
except that the transport scenarios assume a single transport worker hauls all
the waste volume (based:on volumes in Chapter 4, Section 4.4) associated with
a group of waste streams from' the generator's site to the landfill or dump.
As shown in Table 10-7,2all the surrogate BRC waste streams are covered by the
analyses, although not all BRC disposal scenarios are included. It was
assumed that these transport scenarios would represent;the maximum for each
group of wastes (i.e., reactor, institutional, foundry, etc.) where a
combination .of was te. groups pc.cur, the, dfita may, be extrapolated to obtain any
other transport or disposal scenario desired.
Vi -. . ' - •'•'' < '' . • "
10.7 Discussion of the Health Impacts from BRC Waste Disposal
^•; -..i.v,.f4': ."ft; .,.-•?.-«: '.... -,i^. .*>•••:•-:•.:•••.•'!•••.•:'.."•-• '•':'••.'•
• •'•••; This:vseetion''will examine'"and-discuss tne health impacts from the
disposal of BRC waste. These impacts consist of cumulative population health
effects and maximum CPG doses.
10.7.1 Cumulative Population Health Effects
The predicted population health effects for the localized scenarios.
ranged from extremely small fractions, 0.00007, to about 30 excess health
effects over 10,000 years for 20 years of waste accumulation, where excess
health effects is defined as both fatal cancers and serious genetic effects.
The genetic effects range from 5 to 24 percent of the total health effects.
10-11*
-------
Table 10-5.
The maximum CPG annual doses of BRC waste disposal by scenario,
setting, and pathway for 20 years of accumulated waste
Scenario
(disposal
setting)
1 PWR-MD
2 BWR-MD
3 LUMC-UF
4 MAFC-SF
5 MAFC-SI
6 PWRHU-MD
7 UHX-MD
8 UF-MD
9 LUR03-ON
10 LMACW-SI
11 LMACW-UI
Humid
Impermeable
Dose
(mrem)
12
1.1
11
2.7
0.18
0.89"
5.4
8.8
0.13
0.039
0.54
21
4.4
Pathway
Gamma
Bioint.
Gamma
Bioint.
Gamma
Gamma
Gamma
Gamma
Dust
Dust
Well
Gamma
Gamma
Humid
• Permeable
Dose
(mrem)
12
11
1.6
0.18
1.1
5.4
1.5
8.8
0.13
0.039
2.4
21
1.1
4.4 .
Pathway
Gamma
Gamma
Bioint.
Gamma ,
Well
Gamma
Well
Gel nuns
Dust
Dust
Well
Gamma
. Well
Gamma
Arid
Permeable
Dose
(mrem)
12
11
2.2
0.18
0.89
5.4
8.8
0.13
0.039
0.16
21
4.4
Pathway
Gamma
Gamma
Bioint.
Gamma
Gamma
Gamma
Gamma
Dust
Dust
Gamma
Gamma
Gamma
Notes: The transportation pathway was not considered in this analysis.
Key for Disposal Settings (see also Chapter 4, Section 4.4)
MD = Municipal Dump
SF s Suburban Sanitary Landfill
UF « Urban Sanitary Landfill
SI * Suburban Sanitary Landfill with Incineration
UI 3 Urban Sanitary Landfill with Incineration
ON = Onsite Disposal with Incineration
Key for Pathways (see also Table 10-2)
Bioint. s Biointrusion by plant roots to onsite resident
Dust s Onsite worker dust inhalation
Gamma = Direct gamma- radiation to onsite worker
Well = Exposure to offsit'e resident from drinking contaminated
well water
10-12
-------
Table 10-6. The maximum CPG annual doses from already deregulated
waste streams for 20 years of accumulated waste for
4 specific reference scenarios
Humid
Scenario
(disposal
setting)
'Humid
Permeable
Dose Pathway
(mrem)
Dose
(mrem)
Pathway
Arid
Permeable
Dose Pathway
(mrem)
12 CW-SF
13 CW-UF
14 LUROl-ON*
15 LUR02-ON**
0.0018 ' Dust
0.0017 Dust
0.00031 Atmos.
9.1 Well
0.043 Well
0.017 Well
0.00076 Atmos.
40 Well
0.0018 Dust
0.0017 Dust
0.00032 Atmos.
0.0022 Well
* 100% incineration of wastes.
** 50% incineration of wastes. .
Notes: The transportation pathway was not considered in these scenarios.
Key for Disposal Settings^ , .
See Notes in Table 10-5 and Chapter 4, Section 4.4.
Key for Pathways
See Notes in Table 10-5 and Table 10-2.
Atmos. - Exposure to offsite residents from atmospheric inhalation of
radioactive airborne contaminants.
10-13
-------
Table 10-7. Transportation worker exposures to BRC
wastes with a 30-day storage time (Ro86)
Scenarios*
and waste streams
(1)** PWR
(P-COTRASH, P-CONDRSN,
L-WASTOIL)
(2)** BWR
(B-COTRASH, L-WASTOIL)
(3)** Institutional
(I-COTRASH, I-ABSLIQD,
I-BIOWAST, I-LIQSCVL)
(7)** Uranium Hexafluoride
(U-PROCESS)
(8)** Foundry
(N-SSTRASH, N-SSWASTE)
ruel Cycle Wastes
(V— PROPR5?! 17— rWTDACU
Dose
(mrem/yr)
160
87
10
3.5
1.7
1.5
410
69
9. 1
8.0
11 -
0.93
0.0014
0.00025
0.0026
Major
nuclide(s)
Cs-134
Co-60
Co-58
Cs-137
1-131
Cs-136
Cs-134
Co-60
Cs-137
Co-58
Co-60
Cs-137
U-235
U-235
U-235
Half-life
2.05 yr
5.26 vr
~* * *~\j y L
71.3 da
30 yr
8.05 da
13.7 da.
2.05 yr
5.26 yr
30 yr
71.3 da
5.26 yr
30 yr
7.1E+08 yr
7. 1E+08 yr
' 7.1E+08 yr
F-NCTRASH)
Industrial Wastes 1.8
(N-LOTRASH, N-LOWASTE)
(12)** Consumer Wastes*** 0.0001
(C-TIMEPCS, C-SMOKDET)
Co-60
Am-241
5.26 yr
458 yr
disposal scenarios
* Scenarios do not necessarily reflect the same BRC
as listed in Chapter 4, Section 4.4.
** These scenarios are the same as the BRC waste scenarios listed in
Chapter 4, Section 4.4.
***Indicates a reference scenario where the waste streams are already
deregulated.
10-14
-------
(A) Health Effects Versus Demographic Setting
As expected from a cumulative population health effects analysis, the
demographics will play an important part in how the health effects are
distributed, i.e., rural populations will generally incur the least number of
health effects, suburban populations the next largest number, and the urban
population the greatest number of health effects. As shown in Table 10-1, the
populations served for each of the disposal method settings are:
rural
suburban and onsite
urban
60,000
175,000
1,000,000
(B) Health Effects Versus Disposal Method Using Incineration
The analysis showed that where the waste was disposed of in conjunction'
with incineration, the number of health effects was usually reduced by a
factor of two.
The reason behind the reduction in health effects is that the
incineration transforms the majority of the radionuclides from the water
ingestion pathway to the air inhalation pathway. In the air pathway, the
radionuciides are diluted considerably and the body response from inhalation
is generally much less than when a material is ingested.
(C) Health Effects Versus Hydrogeologic/Climatic Setting
The health effects resuLts for the three hydrogeolpgic/climatic settings
— humid permeable, humid impermeable, and arid permeable — showed several
trends.
o In the arid and humid permeable scenarios, health effects to the local
populations dominated in the first 1,000 years. In these regions the
limited amount of ground-water dilution is the major factor, along
with the larger defined populations. In the humid impermeable ,
scenarios, regional basin populations dominated in the first 1,000
years. In this region the larger dilution becomes a factor, along
with the much larger regional basin population.
o The number of estimated excess health effects was greatest in the
humid permeable region, next greatest in the arid, permeable region,
and least in the humid impermeable region. These comparisons are in
relation to one another and, as such, in the humid impermeable region,
the greater ground-water dilution becomes a factor in the amount of
health effects observed. In the arid region, there is limited
ground-water dilution; thus, more health effects are observed.
o The majority (greater than 95 percent) of the health effects were
incurred in the first 1,000 years in all the hydrogeologic/climatic
settings. This is because it is assumed that water infiltration
through the trench cap will be greater; thus, increasing the movement
of radionuclides through the ground will be greater for landfills than
for regulated LLW sites.
10-15
-------
(D) Health Effects Versus Waste Stream
On a national basis, the surrogate BRC waste streams causing the most
health effects for the cumulative population analysis are as follows (in :
descending order): I-COTRASH, N-LOTRASH, I-BIOWAST, and I-ABSLIQD. This can
also be seen in Table 10-4.
(E) Health Effects Versus Radionuclide
The dominant radionuclide causing the most population health effects in
the four surrogate BRC waste streams mentioned above is carbon-14.
10.7.2 Critical Population Group (CPG) Exposures
The individual exposures were calculated for 10 pathways (shown in
Table 10-2 and Figure 10-1) as a maximum annual radiation dose (effective
whole-body dose equivalent) and year of occurrence over 10,000 years for the
CPG. For the overall time span, the maximum individual in any given year may
be one of three persons: onsite worker, onsite resident, or offsite resident.
In the case of BRC waste disposal, the onsite worker is employed at a
BRC waste disposal facility and is not regulated for radiological protection.
The onsite visitor is also considered in this context (see Chapter 4,
Section 4.3). Both the worker and visitor are considered members of the
general public. (The offsite transportation worker is discussed in Section
10.8.) The onsite resident is any member of the general public building a
house and living on the BRC waste disposal site after closure and growing
crops for human consumption. The offsite resident is any member of the
general public who lives away from the BRC waste disposal site, but is
subjected to the various pathways capable of exposing radionuclides to the
human population (see Chapter 8, Section 8.5.4).
There are three time periods involved in the CPG analysis. In all cases
we are assuming that the disposal site has a full 20-year inventory of BRC
waste, with radioactive decay taken into consideration. The first time period
is 0 year or the last year before closure or (pre-closure). In the 0 year,
the maximum individual is either the onsite worker involved with direct gamma
and dust inhalation or an offsite resident exposed via the atmospheric
inhalation or spillage pathway.
Second, the year 1 is considered to be the first year of the post-closure
phase with the maximum exposure from the full site inventory. In the year 1,
the onsite resident is the individual most likely to be exposed via the food
pathways. Finally, there are the variable years, i.e., greater than the first
year in which the offsite resident is most likely to be exposed via the major
water pathways.
10-16
-------
Three of the 10 pathways examined for CPG exposures — the ground water
to the river, the spillage of waste on the surface and subsequent discharge to
surface waters, and the saturation of waste in the trench with overflow to
surface, waters (bathtub effect) — are applicable only to the humid
impermeable hydrogeologic/clima.tic setting and not to the other two settings.
This is because this setting deals only with surface water flow, while the
other two settings deal with ground-water migration to water sources.
In the erosion pathway, the arid permeable setting has no results, mainly
because it is estimated that erosion will not uncover the was'te for over
13,000 years. This is due to the minimal rainfall for the arid setting.
For the urban demographic disposal settings, it was assumed that there
would be no food grown onsite after site closure.
Appendix F shows the detailed exposures per scenario per pathway, the
critical radionuclide, and the year of maximum CPG dose.
A general overview of the CPG doses indicated the following:
o the maximum annual dose was less than 4 mrem (1.6E-06 annual risk or a
lifetime risk of 1.1E-04) in roughly one-half of the principal
localized scenarios with presently regulated wastes (see Table 10-5);
o the dominant radionuclides were cobalt-60 through direct exposure of
workers, cesium-137 through biointrusion, and carbqn-14 through well
water usage;
o the maximum annual dose occurs within the first year in most scenarios
for non-ground-water pathways ; and
o in all regions, the dominant pathways providing the maximum annual
doses that exceed 4 mrem were external gamma radiation, biointrusion,
and ground water.
(A) Exposure Versus Hydrogeologic/Climatic Setting
The maximum annual doses for a given set of waste streams generally show
little variability between disposal method, demographic, or hydrogeologic/
climatic settings. The only exceptions are the ground water-to-well and
erosion pathways for the hydrogeologic/climatic settings. In the ground
water-to-well pathway, the maximum annual doses were greatest in the humid
permeable, next greatest in the humid impermeable, and least in the arid
permeable.
(B) Exposure Versus Specific Humid Impermeable Site Pathways
Three CPG pathways apply only to the humid impermeable hydrogeologic/
climatic setting and affect only offsite residents. They are the ground
water-to-river, spillage, and .the bathtub effect (saturation of the waste and
surface water contamination by trench overflow) pathways.
10-17
-------
In all cases, the spillage pathway occurs in the Last year of site
operation or year 0; the bathtub effect .occurs in the year 100 after site
closure, and the ground-water migration with discharge to a river occurs •
beyond 2000 years in most scenarios. In some cases, it occurs beyond tens of
thousands to hundreds of thousands of years. Table 10-8 presents the CPG
results for each of the 11 localized scenarios for the 'three humid impermeable
pathways only. (See Table 10-5 for acronyms of'disposal scenarios.)
(C) Exposure to the Onsite Worker or Visitor , •
For the onsite worker there are'two, pathways: direct gamma exposure and
inhalation of radioactive dust. These onsite worker pathways are independent
of the hydrogeologic/climatic setting. Where cobalt-60 is present in'the
waste, it becomes the dominant radionuclide for that pathway, and the maximum
exposure takes place during operations (in the year 0). In two scenarios,
UHX-MD and UF-MD, there is .only uranium in the wastes; and although there is
direct gamma exposure to the onsite worker (less than l.OE-08 mrem/yr),
the maximum gamma exposure is not to a worker and does not take place until
erosion removes the trench cover, in the humid areas beyond 3,000 years and
beyond 10,000 years in the arid areas. Table 10-9 shows these data.
The high doses from the direct gamma exposure is the result of Co-60 in
the waste streams. The dust inhalation always occurs to the onsite worker
during operations (in the year 0). >*'
(D) Exposure to Offsite Residents ; :
For the offsite residents there are six pathways. Three of the pathways
only occur in the humid impermeable setting and were discussed previously in
Section 10.7.2 (B). The remaining three pathways''areI: (l) ground-water
migration to a well; (2) surface erosion and deposition to a nearby water
source; and (3) inhalation from atmospheric contamination from dust
resuspension, incineration, or trench fire. The atmospheric pathway exposure
occurs during site operations (in the year 0), while the erosion occurs in the
humid areas beyond 3,000' years and in the arid areas beyond 10,000 years.
•In most cases, releases through the ground water-to-well pathway occur
beyond several hundred years. The longest time is in the humid impermeable
(greater than 2,000 years), next longest (greater than 200 years) in the arid,.
and least (greater than 16 years)- in-the- humid-.permeable hydrogeologic/— •--•
climatic setting.
Table 10-10 presents exposure data versus the scenario for the ground
water-to-well pathway. In all scenarios, the humid permeable hydrogeologic/
climatic setting has the highest maximum annual exposure; the next highest is
in the humid impermeable setting, with the arid setting having the least
exposure.
Table 10-11 presents exposure data versus scenarios for the erosion
pathway at- the humid permeable and humid impermeable sites. Table 10-12
presents the doses delivered through the atmospheric inhalation pathway to all
three hydrogeologic/climatic settings.
10-18
-------
tab16 16-8. CP6 exposures for humid impermeable settings affecting
offsite resident*
Scenario
i
2
3
4
5
6
7
8
9
10
11
~
Maximum
Ground Water
to River
7.5E-0?
1.7E-06
7.3E-07
1.4E-06
8.0E-07
5.0E-07
9.0E-09
4.2E-09
1*1E^06
7, lE-67
5.5E-07
exposure, ,iftretn/year
Pathway
Spillage
2.1E-03
3.5E-03
5.4E-04
5.2E-04
2.8E-04
1.6E-03
t 5.9E-05
1.8E-05
. 1.4E-05
6.9E-04
9.0E-04
Bathtub
2.1E-04
4.7E-04
6.3E-04
5.7E-04
1 . 5E-04
2.5E-04
1.5E-05
4.7E-06
8.9E-05
2.4E-04
3.5E-04
-------
Table 10-9. CPG exposures for direct'gamma and dust' inhalation
pathways to onsite workers and visitors
Maximum exposure,
Pathway
Scenario Direct Gamma
1 1.2E+01
2 1.1E+01
3 1.8E-01
4 8.9E-01
5 5.4E+00
6 8.8E+00
7 2.4E-02
8 4.9E-03
9 1.6E-01
10 2.1E+01
11 4.4E+00
mrem/year
Dust Inhalation
3.2E-02
1.1E-02
1.6E-04
5.3E-02
2.1E-01
2.1E-02
1.3E-01
3.9E-02
7. 6E-03
3.9E-02
1.3E-02
10-20
-------
Table 10-10. CPG exposures for the ground water-to-well pathway
to offsite residents
Scenario
1
2
3
4
5
6
• 7
8
9
10
11
Humid
Impermeable
Site
3.1E-02
7.1E-02
1.5E-03
1.8E-02
6.6E-02
2.0E-02
3.7E-04
1.7E-04
5.4E-01
6.1E-02
7.5E-03
Maximum exposure, mrem/year
Ground water-to-well pathway
Humid
Permeable
Site
1.5E-01
3.3E-01
1.2E-01
1.1E+00
1 . 5E+00
6.7E-01
L.2E-03
5.7E-04
2.4E+00
1.1E+00
4.9E-01
Arid
Permeable
Site
7.3E-04
1 . 7E-03
6.4E-05
1.4E-04
9.4E-05
4.9E-04
4.7E-05
2.2E-05
1.3E-04
7.3E-04
3.0E-04
10-21
-------
Table 10-11. CPG exposures for erosion pathway
to offsite residents
Maximum
exposure, mrem/year
Erosion j>athway
Humid
Scenario Impermeable
Site
1 2.8E-06
2 1.8E-06
3 1.1E-06
4 4. 1E-06
5 3.4E-06
6 2.0E-06
7 2.7E-06
8 "8.5E-07
9 6.4E-08
10 2.1E-06
11 2.5E-06
Humid
Permeable
Site
3.3E-03
1.9E-03
1.7E-03
5.6E-03
4.5E-03
2 . 5E-03
3.6E-03
1.1E-03
l.OE-04
2.5E-03
3.1E-03
10-22
-------
Table 10-12. CPG exposures for atmospheric inhalation pathway
to offsite residents
Humid
Scenario Impermeable
Site
1 2.1E-06
2 " 1.8E-06
3 1.1E-09
4 . 8.0E-06
5 3.6E-02
6 1.2E-06
7 6.7E-06
8 1.1E-05
9 1.3E-03
10 1.8E-03
11 5.8E-04
Maximum exposure, mrem/year
Atmospheric inhalation pathway
Humid
Permeable
Site
4.4E-06
3.8E-06
2.3E-09
1.7E-05
6.0E-02
2.6E-06
1.5E-05
2.4E-05
3.2E-03
3.0E-03
9.4E-04
Arid
Permeable
Site
5.0E-06
4.3E-06
2.6E-09
2.0E-05
3.7E-02
2.9E-06
1.6E-05
2.7E-05
1.3E-03
1.8E-03
6.0E-04
10-23
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(E) Exposure to Onsite Resident
After the disposal facility is closed, it is assumed that the land area
will be available for certain functions. There are two pathways assumed for
an onsite resident, the biointrusion by plant roots to the undisturbed waste
giving a dose to people from eating the plants and a reclamation pathway,
where the land is disturbed by excavating for a basement and the waste is
brought to the surface and mixed with the surface soil and food grown within
this mixed soil/waste. The maximum exposure from these two pathways occurs in
the first year after closure. Only the rural and suburban disposal settings
are assumed to have food grown on the land. Cesium-137 is the dominant
radionuclide in those scenarios handling several waste streams and
uranium-234/238 in the two uranium disposal scenarios, UHX-MD and UF-MD.
Tables 10-13 and 10-14 present maximum annual doses versus scenarios for
these two onsite resident pathways.
(F) CPG Dominant BRC Waste Streams
The surrogate BRC waste streams causing the highest CPG exposures are as
follows (in descending order): P-COTRASH, B-COTRASH, I-COTRASH, N-LOTRASH,
and I-ABSLIQD.
(G) CPG Dominant BRC Waste Stream Radionuclides
The radionuclides dominating the CPG analyses for the various pathways
involved in the BRC waste disposal scenarios are listed in Table 10-15.
10.8 Discussion of Transportation CPG Results
The primary exposure pathway to the transportation workers will be gamma
exposures. The study (see Section 10.4) showed that the major exposure to the
transportation worker appears to occur from the cobalt-60/58 and the
cesiura-134/137 radionuclides (see Table 10-7). These radionuclides primarily
occur in the following waste streams: P-COTRASH, B-COTRASH, I-COTRASH,
I-ABSLIQD, and N-LOTRASH.
10.9 Discussion of. CPG Versus Population Results
Table 10-16 presents a combination of the cumulative population health
effects versus a range of BRC standards using CPG exposures and the surrogate
waste streams considered. .
As shown in Table 10-16 and Sections 10.7.2 (F) and (G), certain
radionuclides and specific waste streams are the major contributors causing
the excess health effects and CPG exposure. It is therefore possible to vary
both the population health effects and the CPG doses by selecting appropriate
waste streams with and without specific radionuclides to be declared BRC. The
risks of any alternative BRC scenario can be examined by constructing a
scenario which eliminates those waste streams contributing the highest dose.
Also restricting any of the waste streams will influence the amount of BRC
waste as a percent of the total volume.
10-24
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Table 10-13. CPG exposures for biointrusion pathway to onsite residents
Humid
Scenario Impermeable
Site
1 1.1E+01
2 2.7E+00
3 N/A
4 7.6E-02
5 3.0E-01
6 8.1E-01
7 4.2E-04
8 1.3E-04
9 N/A
10 6.8E-01
11 N/A
Maximum exposure , mrem/year
Biointrusion pathway
Humid
Permeable
Site
6.4E-01
• 1 . 6E+00
N/A
4.6E-02
1.8E-01
4.8E-01
3.8E-04
1.2E-04
N/A
4.1E-01
N/A
Arid
Permeable
Site
9.0E-01
2.2E+00
N/A
6.4E-02
2.6E-01
6.8E-01
4.4E-04
1.4E-04
N/A
5.7E-01
N/A
Note: NA - Not applicable.
10-25
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Table 10-14. CPG exposure for Che food grown onsite pathway
Humid
Scenario Impermeable
Site
1 3.2E-01
2 8.1E-01
3 N/A
4 2.3E-02
5 9.0E-02
6 2.4E-01
7 1.2E-04
8 3.8E-05
9 N/A .
10 2.1E-01
11 N/A
Maximum exposure, mrem/year
Food grown onsite pathway
Humid
Permeable
Site
1.9E-01
4.8E-01
N/A
1.4E-02
5.5E-02
1.5E-01
1.1E-04
3.5E-05
N/A
1.2E-01
N/A
Arid
Permeable
Site
2.7E-01
6.7E-01
N/A
1.9E-02
7.7E-02
2.0E-01
1.3E-04
4.1E-05
N/A
1.7E-01
N/A
Note: N/A - Not applicable.
10-26
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Table 1.0-15. Dominant radionuclides for the CPG pathways
Pathway
Dominant
radionuclides
Ground Water-to-River
Ground Water-to-Well
Spillage
Erosion
Bathtub
Food Grown on Site
Biointrusion
Direct Gamma
Dust Inhalation
Atmospheric
C-14
1-129
U-234/238
C-14
1-129
U-234/238
Cs-137
Co-60
U-234/238
Pu-239
U-234/238,
C-14
1-129
C-14
1-129
U-234/238
Cs-137
U-234/238
Cs-137
U-234/238
Co-60
U-235
Am-241
U-234/238
Cs-60
Am-241
U-234/238
H-3
Co-60
10-27
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I
NJ
CO
Table 10-16. Excess health effects over 10,000 years nationwide
from disposal of 20 years of accumulated DOE and
commercial wastes versus regulated LLW disposal
Alternative
BRC levels Lifetime
(maximum CPG dose) risk
mrem/yr
BRC waste percent
of total volume
Additional health effects
versus current practice*
BRC surrogate waste stream rejected^
15 4.2E-04
4 1.1E-04
1 2.8E-05
0.4 I.IE-05
O.I 2.8E-06
43%
34%
30%
28%
25%
457
85
30
20
1
P-COTRASH and B-COTRASH
I-COTRASH and above
• ~ N-LOTRASH, I-ABSLIQD, and above1
N-LOWASTE and above
F-PROCESS, U-PROCESS, I-BIOWAST,
P-CONDRSN, and above
Under current practice, commercial LLW is treated as Class A and disposed of at a SLD- consumer wastes are
unregu ated. DOE waste is disposed of in the as-generated form in an SLD site.
Table 10-5 presents a complete listing of the BRC surrogate waste streams used in the assessment.
-------
10.10 Discussion of the Reference Scenarios
Four localized scenarios using presently deregulated waste streams were
chosen as a reference for.comparison and give a perspective for our health
impacts analysis. The deregulated waste streams used were consumer wastes and
those wastes associated with the NRC biomedical rule. Sections 3.3.3, 4.4,
and 10.6 describe and discuss the waste streams and the four scenarios.
10.10.1 Cumulative Population Health Effects
Table 10-3 shows the four BRC reference scenarios and their excess
population health effects over 10,000 years. As shown in the table, the
deregulated consumer wastes presented less than 0.1 health effects over the
10,000 years for urban, suburban, and all hydrogeologic/climatic settings.
For the deregulated BIOMED waste stream containing only institutional wastes
with just carbon-14 and tritium (H-3) and disposed of in a suburban setting,
the 100 percent volatilization of the two radionuclides for this incineration
scenario (LURO-1) causes less than 0.3 health effect over 10,000 years for all
hydrogeologic/climatic settings. For the same incineration scenario (LURO-2),
except that there is only 50 percent volatilization of the two radionuclides,
the humid permeable hydrogeologic/climatic setting indicates the possibility
of greater than 20 health effects over 10,000 years, while the other two
hydrogeologic/climatic settings cause less than one health effect over 10,000
years.
As mentioned in Section 10.7.1 (B), the use of incineration can have the
effect of reducing population health effects in relation to direct burial. In
this case, the 50 percent volatilization incineration scenario allows enough
of the wastes' radionuclides (C-14 and H-3) to be disposed of by burial, thus
'allowing a larger source to reach the population through the water pathway in
the humid permeable setting. During the 100 percent volatilization
incineration scenario, the source terms are diluted and the inhalation pathway
provides a much lower body response, as compared to the water ingestion
pathway.
10.10.2 CPG Exposures
Table 10-6 shows the BRC reference scenarios and their maximum CPG
doses. As shown, releases from the deregulated consumer wastes presented less
than 0.02 mrem/yr for the urban and suburban areas, and for all the
hydrogeologic/climatic regions. For the deregulated BIOMED institutional
wastes and the 100 percent volatilization incineration scenario (LURO-1), the
CPG doses are less than 0.0008 for all three hydrogeologic/climatic regions.
However, the 50 percent volatilization incineration scenario (LURO-2) does
show higher doses in the ground water-to-well pathway (at year 16 after
closure for humid permeable and at 2,320 years for humid impermeable), where
carbon-14 is the critical radionuclide.
As indicated in the previous section, the total incineration of the
wastes provides for lower health impacts due to the fact that: the inhalation
pathway affords less of a risk to the whole body than does the ingestion
pathway.
10-29
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REFERENCES
Atomic Industrial Forum, De Minimus Concentrations of Radionuclides in
Solid Wastes, AIF/NESP-016, Prepared by Nuclear Safety Associates,
Washington, D.C., April 1978.
U.S. Environmental Protection Agency, in press, PRESTO-EPA-BRC: A
Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Cose, Documentation and User's Manual, RAE-8706-5, Rogers
and Associates Engineering Corporation, Salt Lake City, Utah, 1987.
U.S. Environmental Protection Agency, in press, PATHRAE-EPA: A
Performance Assessment Code for the Land Disposal of Radioactive
Wastes, Documentation and User's Manual, RAE-8706-6, Rogers and
Associates Engineering Corporation, Salt Lake City, Utah, 1987.
NRC80 U.S. Nuclear Regulatory Commission, Environmental Assessment of
Consumer Products Containing Radioactive Materials, NUREG/CR-1775,
Washington, D.C., October 1980.
AIF78
EPA87a
EPA87b
NRCSla
NRCSlb
NRC84
NRC86
Oz84
PEI85
Ro84
U.S. Nuclear Regulatory Commission, Biomedical Waste Disposal, Final
Rule, 10 CFR Part 20, Federal Register, 46(47):16230-16234, March 11,
1981. —
U.S. Nuclear Regulatory Commission, Data Base for Radioactive Waste
Management, 3 Volumes, NUREG/CR-1759, November 1981,
U.S. Nuclear Regulatory Commission, Edison Electric Institute and
Utility Nuclear Waste Management Group; Filing of Petition for
Rulemaking, Federal Register, 49(183):36653-36655, September 19, 1984.
U.S. Nuclear Regulatory Commission, Update of Part 61 Impacts Analysis
Methodology, 2 Volumes, NUREG/CR-4370, Washington, D.C. January 1986.
Oztunali, O.I. and G.W. Roles, De Minimis Waste Impacts Analyses
Methodology, NUREG/CR 3585, prepared by Dames & Moore for the U.S.
Nuclear Regulatory Commission, Washington, D.C., February 1984.
PEI Associates, Inc., CPG Dose Rates for Transportation Workers and
Onsite Workers Exposed to BRC Waste Streams, EPA Contract No.
68-02-3878, Work Assignment No. 18, Cincinnati, Ohio, October 1985.
Rogers, V.C. et al., An Update on Status of EPA's PRESTO Methodology
for Estimating Risks from Disposal of LLW and BRC Wastes, Proceedings
of the 6th Annual Participants' Information Meeting of DOE Low-Level
Waste Management Program, CONF-8409115, December 1984.
10-30
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Ro86
SABS 5
Rogers, V.C. et al., Gamma Doses to Maximally Exposed Workers from
Transportation of BRC Waste Streams, TIM-8621-2, Rogers and Associates
Engineering Corp., Salt Lake City, Utah, July 28, 1986.
Science Advisory Board, U.S. Environmental Protection Agency,
Report on the March 1985 Draft Background Information Document for
Proposed Low-Level Radioactive Waste Standards, SAB-RAC-85-002,
Washington, D.C., 1985
10-31
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-------
Chapter 11: SENSITIVITY ANALYSIS OF THE PRESTO-EPA MODELS
11.1 Introduction
This chapter describes the program of sensitivity analysis conducted
on the PRESTO-EPA models. The analysis methodology is described and the
results summarized. The rationale for the analyses, as well as their
limitations, is presented.
11.1.1 Background
In developing the LLW Standard, it was necessary for EPA to assess the
impacts from the disposal of LLW using a variety of disposal methods, site
locations, and other variables. These assessments were performed using the
PRESTO-EPA computer models.
The PRESTO-EPA model was developed jointly by EPA and Oak Ridge
National Laboratory (EPA83). The model, completed in 1983, was expanded by
EPA and Rogers and Associates Engineering Company into a family of health
impact assessment codes (Ro85). These codes are described in more detail in
Chapter 8.
Because PRESTO-EPA was developed specifically for the LLW standard-
setting effort and is a new code, a program of code improvement and
verification was conducted. This program included: quality assurance ^
audits'of all codes, extensive test runs, peer review, and review by EPA s
Science Advisory Board. Another important aspect of this program,
sensitivity analysis, is discussed in this chapter.
11.1.2 Description of Sensitivity Analysis Program
Sensitivity analysis can be defined as changing the values of specified
input parameters, either individually or as a group, in order to assess the
change in the model output. The output from the test runs is compared to
the output from standard runs, where all the input parameters remain con-
stant. In this way the results can be quantified and a relative measure of
the model's sensitivity to changes in various input parameters is determined,
In our sensitivity analysis program we conducted two broad types
of analyses. The first, called single parameter sensitivity analysis,
consisted of varying only a single parameter (in some cases, a few
parameters) at a time. Examples of this type of analysis would be
increasing or decreasing the aquifer flow rate or the permeability of
the trench cap. The single parameter sensitivity analysis is summarized
in this chapter and the analysis results and conclusion given. A more
detailed discussion of the analysis and results is contained in a separate
EPA technical report (EPA88).
11-1
-------
The second type, which we called scenario sensitivity analysis,
consisted of varying a group of input parameters associated with a specific
scenario variable, in order to modify the scenario associated with one of
the base case analyses. Examples of this type of analysis would be changing
the waste form, the size of the site, or the disposal methods. We termed as
scenario sensitivity analyses any assessments we performed other than the
base case analyses outlined in Chapters 9 and 10.
H.1.3 Rationale for Conducting Sensitivity Analyses • .
Since the PRESTO-EPA models were new,codes, it was important to test
them as extensively as possible. The single parameter sensitivity analyses
were an important part of this test program, as they allowed us to identify
the most sensitive input parameters. The identification of the sensitive
input^parameters prior to the final production runs allowed for more
efficient use of limited resources in better characterizing those parameters
which would most affect model output. Also, sensitive input parameters were
flagged for more thorough review when checking input lists for accuracy
prior to production runs.
In addition to identifying the sensitive parameters, the program of
single parameter sensitivity analyses allowed us to:
o
o
Reconfirm code logic and reliability;
Test the effects of parameters .with wide ranges of values or large
degrees of uncertainty; and
o Evaluate controversial input parameter values.
We also carefully reviewed the results of each of the sensitivity tests,
which provided us with a great deal of knowledge about how the codes
responded to changes in individual input parameter values and how various
aspects of the output were affected.
The scenario sensitivity analyses allowed us to analyze the results of
scenarios different from those,chosen as our standard base-case analyses for
the LLW standard-setting effort (discussed in Chapters 9 and 10). In this
way we were able to test how scenario assumptions that were made about the
base cases affected the output results.
The scenario sensitivity analyses differed from the single parameter
analyses in that, in general, a group of input parameters related to a
specific scenario variable was varied. .The purpose of the analyses was
not necessarily to determine what would happen when certain input para-
meter values were changed, but to see what would happen when the scenario
variables were changed. In choosing a set of standard scenarios to analyze
for our standard-setting effort, it was necessary to make certain assump-
tions about the scenarios, such as the volume of waste disposed of, the
form of the waste, or the disposal methods that would be used. The scenario
analyses allowed us to determine how sensitive the results were to these
assumptions.
11-2
-------
11.1.4 Limitations of the Sensitivity Analyses
During the testing and verification of the PRESTO-EPA code, thfe
question of the uncertainty of the results was raised. A Monte Carlo _
analysis technique was suggested as'a method for quantifying the uncertainty
in the risk assessments. In this technique, the PRESTO-EPA deterministic
risk assessment models would be combined with random input parameter
sampling and statistical analysis submodels to form a probabilistic risk
assessment model. A Monte Carlo analysis of a large, complex nsk^
assessment model such as PRESTO-EPA would require a large number of computer
calculations. This was not feasible from the standpoint of either the funds
or time available, especially considering that the main purpose of our
generic LLW standard setting- analyses was to conduct a relative comparison
of various control methods, for the purpose of setting a standard, rather
than obtaining absolute values from site-specific disposal situations.
A combination of. single parameter and scenario sensitivity analyses was
selected as a less rigorous but acceptable method of determining some of the
uncertainties associated with the PRESTO-EPA results. In addition,_this
type of analysis is a very useful method for determining the sensitivity
associated with the various PRESTO-EPA input parameters and disposal
scenarios, since sensitivity analysis provides the relative sensitivity
of model results due to changes in the parameters tested. It also helps
in re-verifying the PRESTO-EPA code and in discerning differences in over
all disposal system performance due to changing scenario assumptions, such
as the use of a buffer zone or high-integrity containers. -
The type of sensitivity analysis that we performed, however,
has shortcomings and limitations in that the relative importance of
each parameter could be affected by the values'of the other - parameters.
We addressed this problem to a certain degree by performing a large number
of sensitivity runs under many different scenarios. This helped to identity
parameters and scenario variables that were sensitive under various assump-
tions, but did not eliminate the overall problem of having determined the
sensitivity based upon some assumed set of input parameters. The problem
of choosing an assumed ;set of input parameters was dealt with by using
"standard input data sets," which were the values used for our base case :
analyses. In this way we were able to determine which parameters and
scenario variables were most sensitive under the base case scenarios and
which were the most important, since the results of these scenarios were
what would be used as a basis to develop the LLW Standard. If the base
case scenarios were changed, the sensitivity of certain input parameters
might change, but characterizing this was felt to be beyond, the scope of
this.study and of lesser importance.
Another limitation of our analyses was that not every input parameter
or scenario was tested. Because of the large number of input parameters
and possible scenarios, it was impractical, if not impossible, to test each
one. For these analyses, we tested those parameters and scenarios which we
felt would be most sensitive, based on the'extensive test runs and code
11-3
-------
review we had performed prior to the sensitivity analysis program, as well
as on good engineering judgment. In addition, we tested parameters and
scenarios if we felt the results to be of particular interest or if there
was some uncertainty or controversy over the value or assumptions we were
using for the base case.
Finally, because of the inherent limitations of the single parameter
and scenario sensitivity analyses, no direct measure of model uncertainty
could be made. In order t.o provide some direct measure of the uncertainty
of the model output, an analysis of uncertainty was conducted. This
analysis and the results are described in Chapter 12.
11*2 Single Parameter Sensitivity Analysis
The single parameter sensitivity analyses, as conducted by EPA
consisted of systematically varying the values of specific input parameters
to quantify their effects on code output. To determine the relative
sensitivity of each of the parameters tested, we developed a quantitative
sensitivity index, as well as an associated qualitative rating of
sensitivity. For these analyses, we chose specific standard data sets,
with input parameter values equivalent to those used in the base case runs
described in Chapters 9 and 10, and known output against which to compare
the output from the sensitivity runs. Table 11-1 outlines the important
features of our "standard" input data sets.
The majority of the parameters tested were related to infiltration
nuclide retention and release, transport, and exposure mechanisms. The
health risk factors used in the codes were not tested. The health risk
factors are calculated by the EPA RADRISK code and are used as inputs to
the program DARTAB. The DARTAB code, which is used as a subroutine by
PRESTO-EPA, combines radionuclide uptakes with the RADRISK health risk
factors to determine health impacts. The DARTAB portion of the PRESTO-EPA
code was not included in the sensitivity analyses, as it received extensive
review during development of the AIRDOS-EPA code, for which it was
originally developed (Be81/ Mo79). We did, however, perform sensitivity
analyses on the health effect conversion factors (HECF), which are used to
assess long-term health effects in the regional basin population, as
described in Chapter 8.
The various PRESTO-EPA codes, while basically similar in design and
function, have differences that are important to understanding the analysis
results. The_PRESTO-EPA-POP code is used to estimate the cumulative health
effects, consisting of fatal cancers and serious genetic effects, to both
local and regional basin populations over 10,000 years. Local population
health effects are calculated through a number of detailed pathway analyses
using iterative yearly updates for a period of 1,000 years. Health effects
tor the regional basin population are calculated using a HECF for both an
initial 1,000-year period and for an additional 9,000 years (during which
the local population is included within the regional basin) for a total of
10,000 years. The local and regional basin populations are assumed to live
at distances from the disposal site similar to what one might encounter
today in those geographic regions. Because the PRESTO-EPA-POP code was the
basis for the other codes, it was tested extensively, with a total of 54
test runs performed on 30 input parameters (see Table 11-2).
11-4
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Table 11-1. Important features of "standard" data sets
used in the single parameter sensitivity analyses
Characteristics
Sites Evaluated
Disposal Method
Waste Type
Site Capacity
Modeling Period
Population Analyzed
Impact Analyzed
PRESTO-EPA-POP
Humid Permeable
Arid Permeable
Humid Impermeable
Conventional
Shallow Disposal
Absorbed Waste
250,000 m3*
10,000 years
Local and Regional
Populations
\jumu i. at ive
Health Effects
PRESTO-EPA-CPG
Humid Permeable
Arid Permeable
Humid Impermeable
Conventional
Shallow Disposal
Trash Waste
Absorbed Waste
Solidified Waste
Activated Metal
Incinerated/
Solidified Waste
250,000 m3
1,000 years
Critical Population
Group
Maximum Annual
Whole-Body Dose
PRESTO-EPA-DEEP
PRESTO-EPA-BRC
Humid Permeable
Arid Permeable
Humid Impermeable
Deep Geological
Hydrofracture
Deep Well Injection
Absorbed Waste
Variable
10,000 years ;
Local and Regional
Populations
Cumulative
Health Effects
Humid Permeable
Arid Permeable
**
Urban Sanitary Landfill
Municipal Dump
Urban Sanitory Landfill
w/Incineration
Absorbed Waste
Variable
10,000 years
Local and Regional
Populations
Cumulative
Health Effects
* individual nuclide activity is used based on 250,000 m3 site, but since only 10 of the 40 nuclides
are evaluated, the actual source term used is less than 250,000 m .
**The humid impermeable site was not evaluated in the BRC sensitivity analysis, since the results for the
parameters tested would be the same as for the other sites tested.
-------
Table 11-2. Summary of input parameters analyzed and tests performed
Code
PRESTO-EPA-POP
PRESTO-EPA-CPG
PRESTO-EPA-DEEP
PRESTO-EPA-BRC
Total Input
Parameters
146
151
154
151
Input Parameters
Evaluated
30
56
20
12
Sensitivity
Tests Performed
54*
121
41
22
*An additional 10 test runs were performed using the PRESTO-EPA-POP
code in testing the health effect conversion factor (HECF).
11-6
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The PRESTO-EPA-CPG code is used to estimate the maximum annual dose and
the year in which the maximum dose occurs for a critical population group.
This group is assumed to live adjacent'to the disposal site and obtain its
water from a well or stream located at_the boundary fence. Only the first
1,000 years are evaluated and impacts to the regional basin population are
not determined, as discussed in Chapter 8. Figure 11-1- illustrates the
differences between the population groups and their locations,, as modeled
by these two codes. The PRESTO-EPA-CPG code is quite different from the
PRESTO-EPA-POP code, in the impacts that are assessed and in how the source
term is modeled. Therefore, this code was also tested extensively, with a
total of 121 test runs performed on 56 input parameters, (see Table 11-2).
The PRESTO-EPA-DEEP code is used to estimate the cumulative population
health effects to both local and regional basin populations for 10,000 years
from deep disposal options. As with the PRESTO-EPA-POP code, regional basin
health effects are determined using a conversion factor. The results of the
PRESTO-EPA-DEEP code are not used as extensively in our standard development
effort as are those from the other codes; however, since the pathways are
different, the code was tested fairly extensively. A total of 41 test runs
were performed on 20 input parameters (see Table 11-2).
The PRESTO-EPA-BRC code is used to estimate the cumulative population
health effects to both local and regional basin populations for 10,000
years, in the same manner as the PRESTO-EPA-POP code. The major difference
between the two codes is that PRESTO-EPA-BRC also determines health effects
to onsite workers from unregulated disposal of BRC wastes through
incineration, dust inhalation, and direct exposure pathways. Because the
two codes are so similar otherwise, the major portions of the PRESTO-EPA-BRC
code that were tested were those having to do with the pathways for the
onsite workers. Therefore, only 22 test runs were performed on 12 input
parameters (see Table 11-2).
The PATHRAE-EPA code, which estimates the annual pathway doses to a
critical population group and to onsite workers, is not based on the
PRESTO-EPA codes, although it is compatible with these codes (Sh86).
Because PATHRAE-EPA is a different code, it was tested in a separate study
by Rogers and Associates Engineering Company (Sh87a). The results of this
analysis are summarized in section 11.2.2 (G).
The codes described briefly above are discussed in detail in Chapter 8.
The number of input parameters associated with each code, how many were
tested, and the number of test runs performed, are outlined in Table 11-2.
In addition to the evaluations of the codes, a set of tests was
performed to evaluate the HECF, The PRESTO-EPA-POP code was used for this
purpose, although the results apply equally to the other codes using the
HECF. A total of 10 tests were conducted on the HECF.
11-7
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(MAXIMUM ANNUAL DOSE
TO CRITICAL POPULATION QROUP)
PRESTO-EPA-CPG
PRECIPITATION
CUMULATIVE POPULATION HEALTH
EFFECTS (FATAL CANCERS AND
SERIOUS GENETIC EFFECTS)
PRESTO-EPA-POP
MAXIMUM ANNUAL DOSE TO CPQ
YEAR OF OCCURRENCE
^^^^m^>^ f ^mim
CUMULATIVE POPULATION
HEALTH EFFECTS ASSESSMENT
FOR LOCAL USE POINT UP TO
1,000 YEARS
»3l
CUMULATIVE POPULATION
HEALTH EFFECTS ASSESSMENT
FOR REGIONAL BASIN
UP TO: J1) 1,000 YEARS
(2) 10,000 YEARS
Figure 11-1. Differences in the Health Impacts Estimated and the
Locations and Populations Evaluated for the
PRESTO-EPA-POP and PRESTO-EPA-CPG Analyses
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11.2.1 Methodology of Single Parameter Sensitivity Analysis
In conducting the single parameter sensitivity 'analyses, a number of
separate steps were required. Before performing the actual sensitivity runs
the "standard" scenarios had to be chosen, the input parameters to be tested
had to be identified, and input parameter test values had to be determined.
After these initial steps were completed, the sensitivity runs were
performed and the results processed so that they could be more easily
analyzed. The analysis of the results consisted of comparisons of the test
run to standard run results; calculation of test run and standard run output
and input ratios; and the determination of quantitative sensitivity indices
and a qualitative sensitivity rating for each parameter tested. Each of
these steps is discussed briefly in the following sections, with a more
detailed discussion contained in an EPA technical report on the single
parameter sensitivity analyses (EPA88).
(A) Choosing Standard Scenarios
The standard data sets and their output serve as a comparison against
the test data sets and output. In conducting the sensitivity analyses we
typically ran the code using a standard data set and recorded the results.
We then changed one input parameter value and ran the code again using this
test data set, again recording the results. Finally, we compared the
results from the standard data set to the results from the test data set to
see how much of a change in output had occurred due to the input change.
This process, therefore, required a set of standard data sets.
In choosing our standard data sets for the sensitivity analyses, it
was decided to use the data sets associated with our "base case" scenarios
described in Chapters 9 and 10, since these data sets were used extensively
in performing the LLW standard setting effort. Although there were 7 base
cases analyzed for regulated disposal and 15 for BRC disposal, because of
the large number of sensitivity analyses we would be performing we could not
test all of the base case scenarios in detail. Instead, we elected to more
carefully analyze only a few scenarios to keep the analysis from becoming
unwieldy. The scenarios that were used are summarized in Table 11-1 and are
discussed in more detail in the EPA technical report on sensitivity analysis
(EPA88).
(B) Choosing Input Parameters to Test
The PRESTO-EPA codes have approximately 150 input parameters associated
with each of them. Because of the large number of input parameters, it was
not practical to test every one. We had, however, performed a large number
of test and production runs prior to our sensitivity analysis program and
were able to choose, based on the results of these runs and good engineering
judgment, those input parameters which we felt were most important to
analyze. We chose parameters which we felt would be most sensitive, which
had a large degree of uncertainty pr a wide range of possible values
associated with them, or which were controversial for some reason.
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Using these criteria, we chose 30 PRESTO-EPA-POP, 56 PRESTO-EPA-CPG
20 PRESTO-EPA-DEEP, and 12 PRESTO-EPA-BRC input parameters to analyze. In
addition, we tested a number of variables associated with the HECF. A
listing of those parameters which were tested is included in a separate EPA
technical report (EPA88).
(C) Determining Test Input Values
Once the input parameters to be analyzed were chosen, we had to
determine what test values to use for the input parameter. We generally
chose values at either one or both ends of a reasonable range around the .
standard value or, if the parameter was very uncertain, one or two orders
of magnitude above or below the standard value. In all cases we tried to
select test values that were realistic and would give us meaningful results.
Based on these criteria, 54 separate PRESTO-EPA-POP test runs were
performed. A total of 121 were done for PRESTO-EPA-CPG, 41 for PRESTO-EPA-
DEEP, and 22 for PRESTO-EPA-BRC. Ten test runs were performed on the HECF.
Each of the sensitivity tests performed, along with the test and standard
values used are listed in a separate EPA technical report (EPA88).
(D) Processing of Output
In order to analyze the results of the test runs, a convenient measure
of output was required so that comparisons could be made to the standard
runs. Both the standard and test run output was processed to result in
a single measure of impact, i.e., maximum annual whole-body dose for
PRESTO-EPA-CPG and 10,000 year cumulative population health effects for
PRESTO-EPA-POP, PRESTO-EPA-DEEP, and PRESTO-EPA-BRC. A more detailed •
discussion of how the output was processed is contained in the EPA technical
report (EPA88). How these measures of impact are used is discussed in the
next section.
(E) Comparing Test to Standard Runs
Most sensitivity tests consisted of changing only one specific input
parameter value, although in some .cases it was more efficient or made more
sense to change a set of input values as a group. Once the input parameter
value(s) was changed from the standard to the test value, the code was run
and the output results recorded.
The output results from the test runs were compared to the output
results from the standard runs. This was done not only to measure the
sensitivity of the input parameter, but also to help re-evaluate the code
logic and reliability and gain some knowledge of how changing a certain
input parameter would affect the model output. The output was evaluated in
two ways: using ratios of standard values to test values and using summaries
of the test output, such as the test result summary form developed for most
of the PRESTO-EPA test runs.
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: The use of ratios in evaluating the sensitivity tests provided a
general measure of an input parameter's sensitivity and of the
reasonableness of the results. Ratios .were calculated for the input
values by simply dividing the input parameter test value by the input
paramete'r standard value. The output ratios were based on a convenient
measure of health impact, the total number of cancers over 10,000 years
in the case of PRESTO-EPA-POP, PRESTO-EPA-DEEP, and PRESTO-EPA-BRC, and
maximum annual dose to the CPG in the case of PRESTO-EPA-CPG. .The output
ratios were calculated by dividing the test run impact by the standard
run impact for each sensitivity test. The sensitivity run could then be
evaluated by comparing the input and output ratios.
•A results summary was also completed for each sensitivity test.
This summary contains a description of the input parameters, how they
vary from the standard, a listing of how the output varies from the
standard, and the input and output ratios. Input and output ratios and
resuts summaries for each test are included in a separate EPA technical
report (EPA88).
(F) Determination of Sensitivity
In order to determine which were the most sensitive input
parameters, input and output ratios were used to determine a quantita-
tive sensitivity index for each input parameter tested. The quantita-
tive sensitivity index was then used to assign each input parameter a
qualitative sensitivity rating of none, low, medium, or high. A listing
of input parameters with sensitivity ratings of medium or high is
contained in Table 11-3.
A more detailed discussion of the quantitative sensitivity index and
the qualitative sensitivily rating, including a listing of the
sensitivity index and qualitative rating associated with each input
parameter tested, is included in a separate EPA technical report .(EPA88).
11.2.2 Results and Discussion of Single Parameter Sensitivity Analyses
In performing the sensitivity analyses and completing the test
results summary forms, basic information was learned about the codes and
how they respond to changes in input values. This is discussed below
under the separate model headings. Results from tests on specific
parameters also provided useful information for later production runs,
and our final runs incorporated many modifications based upon information
learned during the sensitivity analysis program. The major emphasis of
the sensitivity analysis, however, was to identify those parameters which
were most sensitive.
The identification of the most sensitive parameters was felt to be
particularly important since small differences in the value used for an
input parameter might affect the output results to a large degree. By
identifying the sensitive input parameters, limited resources (both time
and money) could be used to characterize those parameters which would
11-11
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Table 11-3. PRESTO-EPA input parameters identified as
exhibiting relatively medium or high sensitivity
under the conditions of this sensitivity analysis
PRESTO-EPA-POP
o
o
Percent of Trench Cap Failure
Release Fraction for Solidified Waste
PRESTO-EPA-CPG
o
o
o
o
o
o
o
o
o
o
o
o
o
Percent of Trench Cap Failure
Trench Cover Permeability and Porosity
Waste Release Fraction and Distribution Coefficients
Waste Container Related Parameters
Trench and Sub-Trench Porosity and Residual Saturation
Distance from Trench to Well and Trench to Aquifer
Aquifer Porosity and Thickness •
Ground and Surface Water Velocity
Mobile Nuclide Source Term
Spillage Fraction for Arid Sites
Atmospheric Pathway Parameters for Arid Sites
Duration of Institutional Control (Active Site Maintenance)
Amount of Water .Uptake by Humans
PRESTO-EPA-DEEP
o
o
o
o
Vertical Water and Groundwater Velocity
Density of the Confining Stratum
Distribution Coefficient (Kd) for the Vertical Zone
Waste Release Fraction
PRESTO-EPA-BRC
o
o
Volatilization Factor for Incinerated Radionuclides
Fraction of Surface Spillage for Arid Sites
HEALTH EFFECT CONVERSION FACTOR
o *Fish Bioaccumulation Factors
o Fish Consumption Rates
o Human Water Consumption Rates
o River-flow-to-population Ratio
11-12
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have the greatest effect on the results. Once a parameter was identified as
sensitive, it could be flagged for more detailed consideration or at least
noted as a parameter that should be reviewed carefully when performing later
production runs.
Each parameter tested was given a qualitative sensitivity rating ranging
from none to high. In general, these ratings were based on the input and
output ratios. If an input ratio was large (i.e.-, a relatively large
difference between the input values used in two runs) and the output ratio
relatively small, the qualitative rating would be low or none. If the input
and output ratios were approximately equal, the sensitivity rating was
labeled as medium. A qualitative sensitivity rating of high meant that the
output ratio was relatively high compared to the input ratio. A listing ot
the quantitative sensitivity ratings for each of the parameters tested is
given in a separate EPA technical report (EPA88). Those parameters
identified as having "medium" or "high" sensitivity are summarized in
Table 11-3.
In reviewing Table 11-3, one notices that the number of sensitive
parameters varies depending upon the code. The reason for this is based
on the differences between the codes and on what health impact they are
evaluating. The following sections discuss each of the codes and how their
characteristics affect the sensitivity analysis results. The results from
the analysis on the HECF is also discussed.
(A) PRESTO-EPA-POP
In order to understand the sensitivity analysis results, it is necessary
to have a basic idea of the results of the base case analyses, as discussed
in Chapters 9 and 10. The general results from the PRESTO-EPA-POP base case
analyses show that the local population health effects do .not dominate in
any of the three regional hydrogeologic and climatic scenarios. This is due
to the limited amount of contaminated nuclides which the relatively small
local population can take in. The majority of health effects are incurred
by the much larger regional basin population. The health effects to the
regional basin population, and the pathways by which they occur, vary
considerably over the three hydrogeologic and climatic regions. The general
trends in the results are more easily seen by reviewing separately the three
general settings: humid permeable, humid impermeable, and arid permeable.
At the site characterized by relatively permeable soil and high
rainfall, most of the mobile radionuclides leach out of the trench and
into the aquifer during the initial 1,000-year period. The majority of the
total health effects are incurred by the regional basin population through
the groundwater pathway during the first 1,000 years.
At the site characterized by high rainfall and soil with relatively
low permeability, the trenches fill with water after a portion of the
trench cap has failed, and much of the activity will be leached from the
waste and will escape from the trench th'rough overflow (bathtub effect) in
a relatively short period of time. Because of transport through the surface
11-13
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water pathway, less mobile nuclides can reach the populations more quickly
than they would have through the groundwater pathway. The regional basin
population receives the majority of the total health effects through the
surface water pathway during the first 1,000 years.
At the site characterized by relatively permeable soil but low rainfall
most activity does not reach either the local or regional basin populations'
until relatively late in the modeling period. The local population incurs
a(few health effects in the first few years after site closure due to
windblown (atmospheric) transport of nuclides spilled onto surface soils
during site operations. These health effects, while quite small, are the
only health effects until late in the modeling period. This is due to the
long travel time required for contamination to reach the aquifer and then
travel to the local and regional basin populations by groundwater. The
overall impact is dominated by health effects from activity reaching the
basin population through the groundwater pathway between 500 and 10,000
years, depending upon the disposal method.
_ Based on the above general results, we can .identify those parameters
which are most sensitive in causing changes in the PRESTO-EPA-POP output.
First of all, there are a number of input parameters that affect the health
effects to the local population during the first 1,000 years. However
since the health effects to the local population are only a small portion
of the total health effects, these changes will not usually affect overall
results significantly. Second, since regional basin health effects are
incurred only from the surface water and groundwater pathways, only input
parameters that ultimately affect these pathways will cause significant
changes in the total health effects.
At the humid permeable site, the health effects will be most sensitive
to parameters that change the release and transport of nuclides from the
trench to the groundwater system, such as the integrity of the trench cap.
At a humid impermeable site, the health effects will be most sensitive
to parameters that increase or decrease the rate of transport of nuclides
from the trench into the surface water system, such as the integrity of the
trench cap and the solidified waste release fraction.
At the arid permeable site, the very short-term health effects will be
most sensitive to parameters that change the amount of surface spillage or
downwind nuclide concentrations. The overall health effects, however, will
be most sensitive to parameters that alter groundwater concentrations, the
same^as for the humid permeable site. Because of the long travel time
required for groundwater transport of even the mobile nuclides at the arid
permeable site, radiological decay of the longer-lived nuclides becomes more
important. Parameters that modify the time required for these nuclides to
reach the local and regional basin populations such as trench cap failure
and, to a lesser degree, trench-to-aquifer distances, aquifer flow rates,
or distances to the population, could lead to significant changes in total
health effects.
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In summary, the PRESTO-EPA-POP code exhibits relatively low sensitivity
to changes in input parameter values for the impacts that were assessed--
cumulative population health effects. This is because the impact that is
assessed, long-term cumulative health effects, is buffered by the long time
period analyzed and the cumulative nature of the output. Changing an input
parameter may have a short-term effect, but over the long term, the results
will tend to change very little. Input parameters that were found to be
relatively sensitive were the parameters which affected infiltration through
the trench cap and leaching out of the trench. In addition, other input
parameters were found to be sensitive when evaluating local population
health effects or short-term impacts.
(B) PRESTO-EPA-CPG
In a similar manner to the PRESTO-EPA-POP analysis, the sensitivity
analysis results for PRESTO-EPA-CPG are more easily understood if the
general base case analysis results are first reviewed. Unlike the
PRESTO-EPA-POP code, the health impact that is assessed by the
PRESTO-EPA-CPG code is not cumulative population health effects, but the
maximum annual dose to a critical, population group located close to the
disposal site. The peak doses to the CPG and the pathways by which they
occur vary considerably over the three hydrogeologic and climatic regions.
Therefore, the base case results for PRESTO-EPA-CPG are also broken down
into the three general settings.
At the humid permeable site, the maximum dose rate occurs relatively
quickly from the groundwater pathway. The important nuclides are those with
high mobility (low Kd values), such as H-3, C-14, and 1-129. They can
•reach the CPG very quickly when combined with permeable soil characteristics
and relatively high groundwater velocities.
At the humid impermeable site, the maximum dose rate occurs soon after
failure of the trench cap (assumed to occur in year 100 for our "standard"
scenario) via-trench overflow directly to the surface water pathway. The
important nuclides are those that are relatively mobile and have longer
half-lives. An example is 1-129, which reaches the CPG soon after the
trench cap fails. It leaves the trench via overflow and is transported
directly to the local stream by surface water, thus bypassing the greater
retardation its higher K^ might afford.if it had moved through
groundwater. Nuclides with shorter half-lives, such as H-3, will not
contribute high doses due to their decay during the period the trench
cover remains intact.
At the arid permeable site, an initial peak in the CPG dose rate
occurs in the first year after site closure due to atmospheric transport
of less-mobile, high-dose nuclides, such as Co-60 and Cs-137, spilled onto
the surface soil during site operations. This peak is relatively small,
however, since only a fraction of the total activity brought onto the site
is assumed to have been spilled during operations and even less reaches the
downwind population after dilution and dispersion by atmospheric transport.
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A much greater peak can occur through the groundwater pathway, although not
until much later in the analysis and even after the 1,000 year modeling
period for many scenarios. This later -peak, if it occurs, would be
significantly larger and would be dominated by mobile nuclides with
relatively long half-lives, such as C-14 and 1-129.
Based on the above general results, we can identify those parameters
which are most sensitive in causing the largest changes'in PRESTO-EPA-CPG
results. In general, the parameters found to be sensitive in the PRESTO-
EPA-POP code were also sensitive parameters in the PRESTO-EPA-CPG code.
At the humid permeable site, the maximum dose rate to the CPG is
most sensitive to parameters that have an effect on: the amount of
water infiltrating into the trench, such as the percentage of trench
cap failure and the trench cover permeability and porosity; the rate
at which radionuclide contaminated leachate leaves the waste matrix and
then the trench, such as waste container related parameters, duration
of institutional control, and nuclide specific release fractions and
distribution coefficients; and radionuclide transit time in groundwater,
such as the distance from the trench to the aquifer and the well.
At the humid impermeable site, the maximum dose rate to the CPG is most
sensitive to parameters that affect the release to the surface water system
and transit time of mobile and relatively long-lived nuclides, such as the
percentage of trench cap failure, waste container related parameters, and
the nuclide specific release fractions.
At the arid permeable site, the maximum dose to the CPG is most
sensitive to parameters that modify groundwater transport characteristics,
such as increasing the amount of trench cap failure, decreasing the
trench-to-aquifer distance, or increasing the aquifer flow rate. In
addition, the spillage fraction and atmospheric pathway parameters are
very sensitive for the scenarios where short-term, atmospheric pathway
doses dominate.
In summary, the PRESTO-EPA-CPG code exhibits greater relative
sensitivity to changes in input parameter values than does the
PRESTO-EPA-POP code. This is because the impact that is assessed, maximum
annual dose to the CPG, is sensitive to small changes due to the model's
assessing peak doses over short time periods to individuals close to the
disposal site. Because the model is evaluating maximum doses relatively
soon after disposal, sensitive parameters are those that affect leaching and
transport of highly mobile, short-lived nuclides, such as H-3. In addition,
the maximum dose will be very sensitive to the source term and release of
the mobile radionuclides. In general, in a similar manner to the PRESTO-
EPA-POP code, the most sensitive parameters will be those affecting
infiltration through the trench cap, leaching out of the trench, and
transport to the CPG.
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(D) PRESTO-EPA-DEEP
The PRESTO-EPA-DEEP model is based ,on and is very similar to the
PRESTO-EPA-POP code. The impact that is assessed, cumulative population
health effects, is the same and, in general, the sensitivity results are
similar. The major exceptions are that the PRESTO-EPA-DEEP code models a
pathway of vertical water movement from a lower aquifer, through the waste,
to an upper aquifer. This pathway is very significant, so input parameters
associated with this pathway, such as vertical water velocity, density of
confining stratum, and distribution coefficients (Kd) for the vertical
zone were found to be sensitive. In addition, since deep disposal assumes
solidified waste, the assumed waste release fraction is a very sensitive
input parameter. Finally, unlike the shallow disposal options, infiltration
through the trench cap is not applicable. Otherwise, input parameter
sensitivity is the same as for the PRESTO-EPA-POP code.
(E) PRESTO-EPA-BRC
The PRESTO-EPA-BRC model is also based on and very similar to the
PRESTO-EPA-POP code. The major difference is that exposure to on-site
workers and visitors from direct gamma exposure and dust inhalation is
included, as well as an incineration pathway to the general public.
Because the number of on-site workers and visitors is small compared, to
the total number of persons affected over 10,000 years, parameters affecting
this exposure pathway are not sensitive in changing overall impact. When
assessing exposures from the incineration pathway, however, H-3 and C-14
become large contributors and the assumed volatilization fraction for these
nuclides is found to be very sensitive. Also, because of the importance of
the atmospheric pathway for the arid sites, the fraction of surface spillage
is sensitive at these sites. Otherwise, the sensitivity of the input
parameters are similar to those found for the PRESTO-EPA-POL3 model.
(F) Health Effect Conversion Factor (HECF)
The health effect conversion factor is used to determine cumulative
population health effects to the regional basin popultion in the
PRESTO-EPA-POP, PRESTO-EPA-DEEP, and PRESTO-EPA-BRC models. In many
cases, the health effects to the regional basin population dominate.
Therefore, it was felt to be important to test this parameter in some
detail. This section summarizes the sensitivity analysis and results on
the HECF. For a more detailed discussion, see the separate EPA technical
report on sensitivity analysis (EPA88).
A number of assumptions associated with the calculation of the HECF
values were tested, including: the assumed fraction of water useage, the
amount of food and water consumption, the river-flow-to-population-ratio,
and fish consumption and bioaccumulation rates. How these values are used
in the calculation of the HECF is discussed in some detail in Chapter 8.
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The results of the analysis on the HECF show that the sensitivity varies
depending upon the nuclide, as the HECF values are nuclide dependent. Since
the majority of population health effec.ts are attributable to C-14 and
1-129, the sensitivity results for these nuclides are the most important.
A few parameters, however, such as the river-flow-to-population ratio', will
affect the HECF independently of specific nuclides.
Because PRESTO-EPA output values are used to determine the HECF values,
changing PRESTO-EPA input parameters can affect the values of the HECF.
The basic parameters being determined from the PRESTO-EPA output is the
number of health effects per unit of activity removed from a well or stream
and the amount of water used by the local community. Therefore, the input
parameters which will most affect the HECF values are those related to •
water_usage. The input parameter UWAT, which is the assumed consumption
of drinking water, is the most sensitive of the PRESTO-EPA input parameters
in regard to the HECF values. Even this parameter, however, does not cause
a large change in the HECF values. In general, changing PRESTO-EPA input
parameter values will not affect the HECF values greatly.
The calculation of the HECF values also requires some parameters which
are independent of PRESTO-EPA. These inlcude the river-flow-to-population
ratio and, for the fish pathway, fish .bioaccumlation factors and fish
consumption rates. These parameters, which affect the HECF values directly,
are generally more sensitive to changes than are those which are used
indirectly through the PRESTO-EPA code. In fact, the HECF values are very
sensitive to the river-flow-to-population ratio, the fish consumption rate,
and the fish bioaccumulation factor, with changes in these parameters • •
causing proportional changes in the HECF values for many nuclides. •• "• •;
The river-flow-to-population ratio (3000 m3/person-yr) is used to
calculate both the health effects from water usage and the health effects
from fish consumption (see Chapter 8). The value used will, therefore, - >•''
affect all components of the HECF calcuation and will affect them directly.
The fish consumption rate and bioaccumulation factors will affect only the
fish component of the HECF calculation. The importance of the fish pathway
in the HECF values varies by nuclide, but for the most important nuclide for-'
cumulative health effects, which is C-14, the fish pathway contributes over
95% of the total health effects. This shows that the HECF values for the
most important nuclides and, therefore, the population health effects in'
general, will be very sensitive to the values used for the river-flow-
to-population ratio, the annual consumption rate for fish, and the fish
bioaccumulation factor for C-14.
In qummary, the HECF values are calculated using PRESTO-EPA related
parameters_and direct- input parameters. The PRESTO-EPA related parameters
will not, in general, affect HECF values greatly, although of the PRESTO-EPA
input parameters, the most sensitive will be the human water consumption
rate. The direct input parameters, river-flow-to-population ratio, fish
consumption rate, and fish bioaccumulation factors (C-14 especially) are
very sensitive in affecting the HECF values and, therefore, the.cumulative
population health effects. It should be noted, however, that the HECF is
not used in calculating CPG dose, so these conclusions are not applicable.to
the PRESTO-EPA-CPG model or to CPG doses.
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(G) PATHRAE-EPA
The PATHRAE-EPA code was analyzed separately (Sh87a). The results
of this analysis showed that sensitivity was comparable to that of the
PRESTO-EPA-CPG code with which PATHRAE-EPA is very similar. The sensitivity
rankings for the PATHRAE-EPA input parameters tested indicate that a number
of the parameters exhibit moderate to high sensitivity. However, the
Overall importance of the majority of these parameters was judged small
owing to the fact that the dose projections of the affected exposure
pathways were negligible to all parameter values tested. The parameters
judged to be significant pertained to hydrogeologic characteristics at the
humid permeable site and disposal facility characteristics at all three
sites, including facility area, waste and cover thickness, and operational
period.
11.2.3 Summary and Conclusions of Single
Parameter Sensitivity Analysis
.Single parameter sensitivity analyses were performed on each of the
PRESTO-EPA codes. The testing concentrated on the PRESTO-EPA-POP and
PRESTO-EPA-CPG codes, as these were the main codes used in the LLW analysis
and were the basis for the other codes. Because of the importance of the
HECF values in calculating cumulative population health effects, parameters
which affected the HECF were also analyzed. The PRESTO-EPA-DEEP and
PRESTO-EPA-BRC codes were evaluated to a lesser degree. The PATHRAE-EPA
code .was evaluated in a separate analysis (Sh87a).
In conducting the sensitivity analysis program, over 100 input
parameters were evaluated and over 200 separate sensitivity tests
performed. The input parameters tested were related to waste form and
composition, hydrogeologic conditions at the disposal site, and engineering
barriers. PRESTO-EPA parameters related to the calculation of the HECF
values were evaluated, as well as parameters independent of PRESTO-EPA which
were used directly in calculating the HECF values. All three
hydrogeologic/climatic sites were included in the analyses.
The main conclusions from the s-ingle parameter sensitivity analyses were:
o Single parameter sensitivity analysis allows for the identification
of those parameters which have the greatest impact on model results;
in addition, it is useful for checking that the code performs in a
logical and consistent manner.
o The identification of sensitive input parameters allows for more
efficient use of limited resources in better characterizing those
parameters which would most affect model output. In addition,
sensitive input parameters can be flagged for more thorough review
when checking input lists for accuracy prior to production runs.
Al&o, identification of the most sensitive input parameters allows
for.the evaluation of some degree of the uncertainty in model output,
based upon knowledge of the uncertainty associated with the most
sensitive input parameters.
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In general, the PRESTO-EPA-CPG and PATHRAE-EPA codes showed much
greater sensitivity than the PRESTO-EPA-POP, PRESTO-EPA-DEEP, and
PRESTO-EPA-BRC codes to an equivalent change in PRESTO-EPA input
values. This is because the PRESTO-EPA-CPG and PATHRAE-EPA codes
estimate the maximum annual dose to a nearby population group,
whereas the other codes evaluate long-term cumulative population
health effects to local and regional basin populations, with intakes
and exposures averaged over the entire period of interest.
Based on quantitative measures of sensitivity, a qualitative
sensitivity rating was given to each parameter tested and the most
sensitive input parameters associated with the various codes were
identified. These parameters are listed in Table 11-3. The other
parameters tested were found to have either low or no sensitivity.
The sensitivity results showed, in general, that model results were
most sensitive to PRESTO-EPA input parameters which affected the
infiltration through the trench cover and the leaching out of the
trench. In addition, the parameters associated with the PRESTO-EPA-
CPG groundwater movement and uptake, the PRESTO-EPA-DEEP vertical
water movement, and the PRESTO-EPA-BRC volitalization factors were
very sensitive.
The results of the single parameter sensitivity analysis seem to
imply that for the PRESTO-EPA model, the most effort should
be placed in better characterizing input parameter values and trans-
port processes having to do with infiltration through the trench cap
and leaching out of the trench. Furthermore, this implication might
be carried on to the actual disposal sites, as suggesting that
reducing infiltration .through the trench cap and leaching out of the .
trench would be an effective means of reducing health impact from the
disposal of LLW.
Based on the analysis of the HECF calculations, it was determined
that PRESTO-EPA related parameters will not, in general, affect
HECF values greatly. The input parameters which are used directly
to calculate the HECF values, river-flow-to-population ratio, annual
fish consumption rate, and river-to-fish bioaccumulation factors
(especially for C-14), are very sensitive in affecting the HECF
values and, therefore, cumulative population health effects.
In determining cumulative population health effects, these input
parameter values should be evaluated very carefully. It should be
noted, however, that HECF values are not used in calculating the
maximum annual. CPG dose.
Although the population health effects analyses were carried out
to 10,000 years., the majority of the impacts in most scenarios
occur before year 1,000. Therefore, a program of single parameter
sensitivity analyses can. be used to evaluate changes to the "disposal
system," which can reduce impacts in the first several hundred
years. These evaluations, along with the results from the
PRESTO-EPA-CPG analyses, may be useful in deciding among several
management or disposal alternatives.
11-20
-------
11.3 Scenario Sensitivity Analyses
In developing the LLW Standard, it 'was necessary to assess the health
impacts that would result from the disposal of LLW under various assumed
scenarios. These scenarios, which reflect a broad range of disposal
options, include assumptions on the disposal sites, disposal methods, waste
form, waste volume, and regional waste mix. Because there was such a large
number of possible scenario combinations that could have been assessed, we •
felt that it was necessary to choose a limited number for our basic
analyses. These basic analyses, which are called the base case scenarios,
are discussed in Chapters 9 and 10.
Additional analyses were also performed where the basic sicenarios were
changed to see how different assumptions would affect the results. These
tests were conducted as part of our sensitivity analysis program as scenario
sensitivity analyses,and are described in this section.
11.3.1 Methodology of Scenario Sensitivity Analyses '.
The two basic codes that we tested (by varying input parameters
associated with a particular scenario variable we wanted to evaluate the
sensitivity of) were PRESTO-EPA-POP, which is used to estimate long-term
population health effects, and PRESTO-EPA-CPG, which is used to estimate
maximum annual doses to a nearby CPG. Since the other codes are based on
and are very similar to these two, scenario sensitivity analyses were not
performed on the other PRESTO-EPA codes. ;
As part of the basic economic and cost-benefit analysis, a total of 93
runs were performed with the PRESTO-EPA-POP and PRESTO-EPA-CPG codes. These
runs were evenly distributed over the three standard site locations. Among
these runs were the 21 base case scenarios (seven for each site location)
described in Chapter 9. The additional runs were assessed as part of the
scenario sensitivity analyses. Tables 11-4, 11-5, and 11-6 list all of the
93 runs performed, by site location, with the base case scenarios
identified. These tables list each run by a scenario number and'include
information on the disposal method, waste form, and waste volume. Acronyms
were used and are described in a key. Also included are the PRESTO-EPA-CPG
and PRESTO-EPA-POP results, in terms of maximum annual dose and long-term
population health effects, respectively. Additional runs, not listed in
Tables 11-4, 11-5, and 11-6, were performed to test specific scenario
assumptions. These scenario sensitivity analyses and their results will
be described in later sections.
Certain scenario variables, such as site location, waste form,
regional waste mix, or site size, can affect the output results. We
felt it was important to determine how sensitive the results from the base
case scenarios were to changes to the variables associated with the assumed
scenarios. In the following sections, the analyses associated with certain
variables and the results of those analyses are described.
11-21
-------
table 11-4. Listing of FRES10-EPA tuns perfumed for a luaid penneable site,
with associated modmm annual dose and emulative population health effects
Disposal method
SCENARIO
HMBER
!•
4»
7»
10*
13*
16
1%
23
26
29
32
35
38*
41
44
47
50
53
56
59
62
65
68
71
74
77
80
95
98
101
103
Class
A
SLO
SLD
SLF
IS!)
IDD
EM
OC
DWI
SLD
SLD
SID
SID
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
Class
B
SLD
SLD
SLF
ISO
ICO
O)
OC
Oil
SLD
SLD
SLD
SLD
SLD
ISO
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLO
SLD
SLD
SLD
SLD
SLD
SLD
Class
C.D.N
SLD
ISO
SLF
ISO
IDD
O)
OC
GUI
SLD
ISO
ISO
ISO
SLO
ISD
ISO
ISD
ISD
ISD
ISD
ISD
ISD
SLD
ISD
SLD
SLD
ISD
LSI)
SLD
ISD
ISD
ISD
Haste forms
Class
A
AS IS
AS IS
AS IS
AS IS
AS IS
GK
SOL
AS IS
INCIN/SOL
INCIN/SOL
AS IS
HIC
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
Class
B
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
AS IS
INCIN/SOL
INCIN/SOL
HIC
HIC
SOL
AS IS
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
AS IS
SOL
AS IS
AS IS
SOL
SOL
AS IS
SOL
SOL
SOL
As gen
Class volute
C.D.N (1000 m3)
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
AS IS
INCIN/SOL
INCIN/SOL
SOL
HIC
SOL
AS IS
AS IS
SOL
SOL
SOL
SOL
SQL
SOL
AS IS
SOL
AS IS
AS IS
SOL
SOL
AS IS
SOL
SOL
SOL
250
250
250
250
250
250
250
6.72
250
250
250
250
250
250
250-
373
250
170
366
250
170
249
249
500
100
500
100
250
250
590
250
CFG topulation
dose in ttalth
peak year • Effects
(nrem/yr) (cancer deatlia)* Coments
35
9.2
62
5.1
5.0
2.0
1.3
7.3
13
13
82
40
9.1
44
35
8.6
7.0
5.7
8.6
7.0
5.7
35
9.1
48
22
13
5.6
37
9.7
15
9.2
4.7
3.9
6.2
3.4
3.4
2.0
1.7
0.036
3.1
3.1
4.6
4.5
3.9
4.6
4.7
4.4
2.9
2.0
4.3
3.0
2.0
4.7
3.9
9.4
1.9
7.8
1.6
2.8
2.4
5.6
3.9
Base Case Run
Base Case Run
Base Case Run
Base Case Run
Base Case Run
QtCB Disposal
Base Case Run
Deep Disposal
Incineration
Incineration
Use of HIC
Use of HIC
Base Case Run
Waste as is
Waste as is
Larger Waste Voluae
LLW + NABMonly
Smaller Waste Volune
Larger Waste Vfolume
LLW Only
Smaller Waste Volume
UH-fiRC+NAiW-Class D
UH-BRCHWRM-Oass D
Larger Waste Volume
Smaller Waste Volune
Larger Waste Vblune
Smaller Waste Vblurae
Regional Compact
Regional Compact
Regional Compact
10,000 Year CFG Run
*Unulative topulation Health Effects Over 10,000 years (cancer deaths only).
-------
Table 11-5. listing of PRESIO-EPA runs performed for a humid imperrreable site,
with associated maxiuun annual dose and emulative population health effects
Disposal
SCENARIO Class
UMBER A
2»
5,
8*
11*
14*
17
20*
22
27
30
33
36
39*
42
45
48
51
54
57
fin
uu
63
66
69
72
75
78
81
96
99
102
104
• Base
SID
SID
SLF
ISO
HDD
EM
OC
HF
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
case scenarios
method
Class Class
B C,D,N
SID
SID
SLF
ISD
IDD
OS
OC
HF
SID
SID
SID
SID
SID
ISD
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
SID
3LD
SID
SID
SID
SID
SID
(see
SID
ISD
S1F
ISD
IDD
CB
CC
HF
SID
ISD
ISD
ISD
SID
ISD
ISD
ISD
ISD
ISD
ISD
ISD
ISD
SID
ISD
SID
SID
ISD
ISD
SID
ISD
ISD
ISD
Chapter 9,
Waste forms
Class
A
AS IS
AS IS
AS IS
AS IS
AS IS
GR
SOL
SOL
Class Class
B C.D.N (1
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
SOL
INCDl/SOL nCK/SOL
INCIN/SOL INCIN/SOL
AS IS
me
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
HIC
me
SOL
AS IS
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
AS IS
SOL
AS IS
AS IS
fWI
OUU
SOL
AS IS
SOL
SOL
SOL
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
SOL
INCIN/SOL
INCIN/SOL
SOL
HIC
SOL
AS IS
AS IS
SOL
SOL
SOL
sa
SOL
SOL
AS IS
SOL
AS IS
AS IS
SOL
SOL
AS IS
SOL
SOL
SOL
As gen
volume
.000m3)
250
250
250,
250
250
250
250
15.2
250
250
250
250
250
250
250
373
250
170
366
250
170
249
249
500
100
(uvi
100
250
250
470
250
CFG
dose in
peak year
(mran/yr)
0.13
0.03
0.77
0.012
9.3E-03
1.9E-03
1.4E-03
1.7E-04
0.03
0.025
0.12
0.13
0.03
0.048
0.12
0.034
0.023
0.015
0.033
0.023
0.015
0.12
0.030
0.25
0.05
0.060
0.012
0.15
0.031
0.058
0.032
Population
Health
Effects
(cancer deatt
8.1
2.2
39
1.1
0.85
0.64
0.26
0.007
6.4
3.4
3.8
6.1
2.2
3.9
4.6
2.5
1.7
1.1
2.5
1.7
1.2
3.6
2.2
16
3.2
4=5
0.9
12
3.8
7.1
2.2
a)* Contents
Base Case Rim
Base Case Bun
Base Case Run
Base Case Run
Base Case Run
EMCB Disposal
Base Case Run
Deep Disposal
Incineration
Incineration
Use of HIC
Use of HIC
Base Case Run
Waste as is
Waste as is
Larger Waste Volume
LIH+NMM only
Smaller Waste Volume
Larger Waste Volume
UHonly
Smaller Waste Volume
UH-BROfNMM-Class D
UW-BRC+NARM-Class D
Larger Waste Volume
Smaller Waste Vblune
Larger Waste Volume
Smaller Waste Volume
Regional Compact
Regional Compact
Regional Compact
10,000 year CFG Run
Figures 9-2 and 9-3).
*Qjnulative Population Health Effects oyer 10,000 years (cancer deaths only).
-------
Table
lasting of PRESTO-EPA runs performed for an arid pemeable site,
with associated maxima annual dose and emulative population health effects
Disposal method
SCENARIO
UMBER
3*
6*
9*
12*
15»
18
21*
24
25
28
31
34
37
40s
43
46
49
52
55
58
61
64
67
70
73
76 -
79
82
97
100
105
Class
A
SID
SID
SLF
ISO
mo
Bi
cc
DGD
DGD
SID
SID
SLD
SID
SLD
SIC
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
Class
B
SLD
SLD
SLF
ISO
HID
CB
CC
DGD
DGD
SLD
SLD
SLD
SLD
SLD
ISD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLD
SLfl
SLD
SLD
SLD
Class
C.D.N
SLD
ISD
SLF
ISD
no
CB
CC
DGD
DGD
SLD
ISD
ISD
ISD
SLD
ISD
ISD
ISD
ISD
ISD
ISD
ISD
ISD
SLD
ISD
SLD
SLD
ISD
ISD
SLD
ISD
ISD
Haste fonns
Class
A
AS IS
AS IS
AS, IS
AS IS
AS IS
GR
SOL
SOL
SOL
mON/SOL
DBDJ/SOL
AS IS
me
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
AS IS
Class Class
B C.D.N
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
nem/soL
DKIN/SOL
me
me
SOL
AS IS
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
AS IS
SOL
AS IS
AS IS
SOL
SOL
AS IS
SOL
SOL
AS IS
SOL
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
INCDJ/SOL
INCIN/SOL
SOL
HIC
SOL
AS IS
AS IS
SOL
SOL
SOL
SOL
SOL
SOL
AS IS
SOL
AS IS
AS IS
SOL
SOL
AS IS "
SOL
SOL
As gen
volune
(1000 m3)
•250
250
. 250
250
250
250
250
118.5
250
250
250
250
250
250
250
250
373
250
170
366
250,
170
249
249
500
100
500
100
250
250
250
CFG
dose in
peak year
(mren/yr)
2.2E-03
9.2E-04
0.41
4.4E-05
3.7E-05
3.1E-06
0
0
0
3.0E-04
3.0E-04
3.9E-03
1.9E-03
9.2E-04
1.4E-03
2.2E-03
8.4E-04
6.8E-04
5.5E-04
8.4E-04
6.8E-04
5.5E-04
2.2E-03
9.2E-04
3.1E-03
1.3E-03
1.3E-03
5.6E-Q4
1.5E-03
l.OE-03
1.7E-02
Population
Health
Effects
(cancer deaths)* Contents
3.6
2.3
3.9
1.7
1.2
6.69
0.31
0.047
0.023
1.4
1.4
3.5
3.5
2.3
3.3
3.6
2.6
1.7
1.2
2.5
1.7
1.2
3.5
2.3
7.1
1-4
4.6
0.92
2.5
2.1
2.3
Base Case Run
Base Case Run
Base Case Run
Base Case Run
Base Case Run
EMCB Disposal
Base Case Run
Deep Disposal
Deep Disposal
Incineration
Incineration
Use of HIC
Use of HIC
Base Case Run
Waste as is
Waste as is
Larger Waste Volume
LUWMRMonly
Smaller Waste Volume
Larger Waste Volune
LlWonly
Smaller Waste Volune
UH-BKC*NHM-Class D
LLW-BRGrtlARM-Caass D
Larger Waste Volune
Smaller Waste Volune
larger Waste Volume
Smaller Waste Volune
Regional Compact
Regional Compact
10,000 Year CFG run
• Base case scenarios (see Chapter 9, Figures 9-2 and 9-3).
*Cunulative Itopulation Health Effects over 10,000 years (cancer deaths only).
-------
Key to Tables 11-4. 11-5, and 11-6
Disposal Method
SLF
SLD
ISD
IDD
EM
CB -
CC
DWI
HF
DGD
Waste Form
As Is
SOL
INCIN/SOL
HIC
GR
Waste Class
A
B
C
D -
N
sanitary landfill
conventional shallow land disposal**
improved shallow land disposal**
intermediate depth disposal
earth-mounded disposal***
concrete bunker disposal***
concrete canister disposal
deep well injection disposal
hydrofracture disposal
deep geologic disposal
no treatment, disposed of as generated
solidified
incinerated and then the ash is solidified
placed in a high-integrity container
supercompacted and grouted
NRC definition (10 CFR 61)
NRG definition (10 CFR 61)
NRC definition (10 CFR 61)
greater than Class C AEA waste (An EPA designation
for the purpose of
this analysis)
naturally occurring and accelerator-produced waste
(NARM)
base case scenarios (see Chapter 9)
Cumulative population health effect .values are for cancers only and do
not include the serious genetic effects. The addition of those effects
will, in general, increase the total health estimates by about
** SLD-ISD - combination is equivalent to 10 CFR 61 disposal
(see Chapter 4)
***EMCB - combination is earth-mounded concrete bunker disposal
(see Chapter 4)
11-25
-------
(A) Site Location
As noted in Chapter 5, three generic locations (humid permeable,
humid impermeable, and arid permeable) are used to reflect the range of
disposal sites in the United States. Since the disposal site is
considered a critical protection factor, all base case assessments were
made for each of the three locations. In addition, with the exception of
the three deep disposal methods and the eight regional compact scenarios,
the scenarios listed in Tables 11-4, 11-5, and 11-6 are spread evenly
over the three sites. The deep disposal methods not assessed for all
three sites are hydrofracture, deep well injection, and deep geological
disposal. These methods are costly and not necessarily suitable for all
wastes or locations. The regional compact scenarios preclude certain
'site characteristics, i.e., some regions have no arid sites.
The humid impermeable location is evaluated using 31 scenarios,
including one scenario for the hydrofracture deep disposal method and
three scenarios to evaluate a northeast compact waste mix. The humid
permeable location is evaluated using 31 scenarios, including one
scenario for the deep well injection deep disposal method and three
scenarios to evaluate a southeast compact waste mix. The arid permeable
location is Devaluated using 31 scenarios, including two scenarios for
deep geological disposal and two scenarios to evaluate,a Rocky Mountain
compact waste mix.
In the scenarios described above, it was assumed,that the humid
permeable site was located in the southeast and the humid impermeable in
the northeast. A set of scenario sensitivity analyses were performed to
evaluate a humid permeable site located in the northeast and a humid
impermeable site located in the southeast. In addition, scenario
sensitivity analyses were performed to evaluate an arid site which is-
impermeable. The results of these analyses are discussed in a later
section.
(B) Disposal Methods
As discussed in Chapter 4, nine disposal methods were analyzed (SLF,
SLD, ISO, IDD, EMCB, CC, DWI, HF, and DGD (see Key to Tables 11-4, 11-5,
and 11-6 for explanation of these acronyms). In the scenario analyses
there is at least one assessment for each disposal method* Each disposal
method includes all of the waste that can be suitably disposed of by that
method. Some of the assessments, however, involve combinations of
different disposal methods. For example, the disposal method (Scenarios
4, 5, and 6) that most closely resembles existing NRG rules requires
shallow land disposal (SLD) "for Class A and B wastes and improved shallow
land disposal (ISO) for Class C waste.
In most of the scenarios, Class A waste and Class B waste are
disposed of by the SLD method, with Class C waste and NARM waste disposed
of by the ISO method. This is. consistent with existing NRG rules (10 CFR
Part 61) and reflects the policy that greater isolation should be
required for the more hazardous wastes.
11-26
-------
The sanitary landfill (SLF), earth-mounded concrete bunker (EMCB),
intermediate depth disposal (IDD), and .concrete canister (CC) methods
have three assessments each, one for each site. The deep geological
disposal (DGD) method has two assessments, both for the arid permeable
location, one for a normal volume and one for a smaller volume. The,
hydrofracture (HF) and deep well injection (DWI) methods have one
assessment each, for humid impermeable and humid permeable sites,
respectively.
(C) Waste Forms
The form the waste is in when placed in the disposal site is a
variable which was analyzed. As part of the scenario sensitivity
analyses, waste forms which were analyzed included, "as is", solidified,
placed within high-integrity containers, and special. The "as is" form
means that the waste is left as generated or unsolidified and includes
trash, adsorbing waste, and activated metal. The special form includes
wastes that are suitable for disposal by the hydrofracture, deep well
injection, or deep geological disposal methods.
*
Waste that is placed in high integrity containers is done so in the
"as is" form., For six scenarios, all of the waste is incinerated and
then the ash is solidified.
The waste form parallels the waste disposal method in most cases,
conforming to the requirements of 10 CFR Part 61 that the more hazardous
waste should receive greater isolation from the environment. Thus, most
scenarios place the Class A waste (the least hazardous) in the "as is"
form, while Class B, Class C, and NARM waste are generally in the
solidified form. -
(D) Waste Mixes (Regional Compacts)
The waste mix reflects the relative volume of each waste stream
which makes up the total waste volume that is disposed of at a site.
The United States average waste mix is used in the base case analyses
and the majority of the other scenarios. There are, however, eight
scenarios which reflect the waste mixes in three geographical regions:
Southeast, Northeast, and Rocky Mountain. There are three.Southeast
waste mix scenarios, three Northeast,, and two Rocky Mountain. Waste
forms and disposal methods are adjusted in six scenarios. Waste volumes •
are varied in two scenarios. -
These scenarios are used to estimate the potential effects that
the Low-Level Waste Policy Act will have on LLW disposal. In particular,
waste volumes can be expected to vary significantly by region. For more
detail on the differences between the U.S. average and the regional waste
mixes, see Appendix A of the EIA.
(E) Waste Groups
For this analysis, wastes were grouped into five broad classes. The
classes represent: AEA-regulated low-level waste (LLW class), the waste
-------
that would be considered a candidate for less restrictive disposal
(BRC class), LLW not covered under the AEA (NARM class), wastes whose
activities are greater than the NRC's Class C (designated by EPA as
Class D for the purpose of this study), and special waste. NARM wastes
are represented by the R-RASOURC and R-RAIXRSN waste streams. Class D
wastes are represented by a select set of Class C wastes (N-SOURCES,
R-RASOURC, and L-DECONRS) for scenarios 65-70 only (See Tables 11-4,
11-5, 11-6). These particular waste streams are, on average, Class C
wastes. However, they contain one or more substreams whose activity
concentrations are greater than the NRC's Class C specifications in
10 CFR Part 61. Special wastes include those suitable for disposal by
hydrofracturing, deep well injection, or deep geological methods.
These waste groups are described in more detail in Chapter 3.
The base case analyses, as well as the majority of the other
analyses use the "LLW-BRC+NARM" waste group. Additional scenarios that
were analyzed include: LLW+NARM, LLW only, LLW-BRC+NARM-Class D, and
special.
By comparing scenarios with and without the different groups, it is
possible to determine the impacts that any one group contributes to the
total. For example, the impact from disposal of NARM wastes can be
obtained by subtracting the LLW impact from the LLW+NARM impact.
(F) Waste Volumes
A 250,000 cubic meter site is used for the base case scenarios.
Other waste volumes are analyzed to assess the effect of varying the
waste volume on disposal site impact, to address the lower volume special
wastes, to address the larger volumes for two waste compacts, and to
address the larger and smaller volumes when NARM, BRC, and Class D wastes
are removed from the total. In addition, in order to evaluate the
consequences of volume reduction, analyses were performed where the total
activity disposed of was kept constant, while the site size was reduced.
(G) Time Horizon
The analysis period used for the base case and most of the other
CPG assessments is 1,000 years, although 10,000 years was used for
three scenarios. The additional three runs were performed in order
to identify any peak CPG doses that might occur after 1,000 years.
In order to provide additional information on how the CPG doses vary
over time, the annual CPG dose is plotted over the 1,000 year analysis
period. This information is provided from PRESTO-EPA-CPG analyses using
conventional shallow land disposal technology and 10 CFR Part 61 disposal
technology. Each of the three hydrogeological sites is evaluated.
Cumulative population health effects, estimated using the
PRESTO-EPA-POP model, are assessed for 10,000 years. In order to
evaluate how these health effects accumulate over time, the model
estimates cumulative health effects for shorter time periods, including
100, 500, and 1,000 years. This information is provided for a site using
10 CFR Part 61 disposal technology in all three hydrogeologic environments,
11-28
-------
(H) Radionuclides
Although specific scenarios were riot developed to analyze the
sensitivity of various radionuclides, -the unit response methodology
employed for our analyses lends itself to such an assessment. There-
fore, we evaluated the sensitivity of the disposal site impact to the
various radionuclides included within the disposal site.
(I) Additional Analyses
A number of additional analyses were performed to answer specific
questions about the scenarios. These included tests to evaluate the
importance of a buffer zone on CPG dose by varying the distance from the
trench to the well. Because global health effects were not assessed in
the PRESTO-EPA analyses, a separate analysis was performed to evaluate
the health effects to the global population associated with releases into
the ocean. The- results of these analyses are discussed in following
sections.
Because of the importance of the volatility fraction on incineration
doses, a set of analyses were performed to test this variable. These
analyses are discussed, however, in Chapter 10, which relates to the BRC
incineration scenarios.
11.3.2 Results and Discussion of Scenario Sensitivity Analyses
The results from the scenario assessments, including those listed in
Tables 11-4, 11-5, and 11-6, can be organized into various groupings to
evaluate the sensitivity of certain scenario assumptions. Test runs are
described according, to their scenario number, if applicable, and results
are generally characterized by both maximum annual dose to the CPG and
long-term population health effects. The results of the_scenario
sensitivity analyses are described in the following sections.
(A) Sensitivity to Disposal Site
The sensitivity of the results to the hydrogeology and climate of
the disposal site can be observed by reviewing Tables 11-4, 11-5, and 11-6.
The CPG dose estimates indicate significant differences in impact among, the
three-sites. Maximum CPG dose comparisons for the different site locations
are all evaluated assuming a well or stream located 100 meters from the edge
of the waste disposal area (different buffer zone distances are evaluated as
well, see Section 11.3.2(1)). The differences are best observed by com-
paring the scenarios for conventional shallow disposal (1, 2, 3), the 10 CFR
Part 61 combination of conventional shallow and improved shallow land
disposal specified by current NRG rules (4, 5, 6), and regulated sanitary
landfill (7, 8, 9). The maximum annual doses at the humid impermeable
location are approximately 100 times lower than at the humid permeable
location. The maximum annual doses at the arid permeable location are
approximately 10,000 times lower than at the humid permeable site for the
SLD and 10 CFR Part 61 disposal technology scenarios and approximately _
100 times lower than at the humid permeable 'site for the sanitary landfill
scenario. This pattern generally holds for all comparable scenarios for the
CPG assessments, and is due to both hydrogeology and climate.
11-29
-------
At the humid permeable site, leachate travels v.ia the groundwater
pathway to a well, while at the humid impermeable site the major pathway
is via surface water flow to a river. 'Maximum doses at the humid
impermeable site are less than at the humid permeable due to the dilution
that a river provides compared to consumption of contaminated groundwaters.
Maximum doses are much less at the -arid site due to the much smaller amount
of water infiltrating the trench and the extremely long travel times in the
arid environment.
The long-term, cumulative population health effects estimates do not
demonstrate as specific a pattern regarding site location as do the maximum
CPG doses. For SLF disposal (scenarios 7, 8, 9) and SLD disposal (1, 2, 3),
the humid impermeable site has the greatest number of health effects, while
for 10 CFR Part 61 disposal technology (4, 5, 6) and IDD disposal (13, 14,
15) the humid permeable site has the greatest number.
The reason for these results has to do with both the hydrogeology
of the site and the form in which the waste is disposed. With the less
sophisticated disposal methods, such as SLF and SLD, the waste is assumed
to be disposed of in the as-generated form. This method provides minimal
containment for the nuclides, so that when trenches overflow at the humid
impermeable site due to the "bathtub effect," larger releases of mobile
and non-mobile nuclides occur than would at the permeable site through
groundwater transport processes. With 10 CFR Part 61 disposal technology
and other greater confinement type disposal, the waste is assumed to be
solidified or in high-integrity containers and trench covers do not fail
as readily. Therefore, overflow does not take place and transport through
impermeable soil leads to lower health effects than at the site with
permeable soil.
While site characteristics and disposal technology causes the
relationship between the maximum number of health effects for a particular
scenario to vary between the two humid sites, long travel times cause the
arid site to have consistently lower health effects estimates. However, in
contrast to the maximum CPG doses, health effects estimates vary by a much
smaller degree over the three sites, generally within a range of about 10
for the various scenarios.
Because of concerns related to the scenario assumption that a humid
permeable site would be located in the Southeast and a humid impermeable
site in the Northeast, an additional set of analyses were conducted to
evaluate the hydrogeological differences between the humid permeable and
humid impermeable site, separate from regional climatic and lifestyle
differences. Also evaluated was ah arid site with impermeable hydro-
geology. All analyses were conducted using 10 CFR .Part 61 disposal
technology and are summarized in Table 11-7. The results are discussed
in the following paragraphs and in more detail in reference Sh87b. When
reviewing the results, it should be understood that there are many complex,
interrelated input parameters associated with the hydrogeology of the sites
and that these are limited analyses, with results that are informative but
not conclusive.
11-30
-------
Table 1L-7. Summary of Varying Site Characteristics
Using 10 CFR Part 61 Disposal Technology
Scenario-'-
Southeast (HP) Site2
.Standard HP Scenario
Soil Characteristics from HI Site-3
Water usage from HI Site ^
Soil arid Water Usage from HI Site
Northeast (HI) Site5
Standard HI Scenario
Soil Characteristics from HP Site6
Southwest (AP)Site7
Standard AP Site
Soil Characteristics from HI Site-3
Maximum CPG
Dose (mrem/yr)
9.2
8.6
3.6
12.0
3.0E-2
2.0E-6
9.2E-4
1.7E-5
Year of
max. Dose
777
390
777
196
191
1
1000
2
1. For a more complete description of scenarios and analysis, see Sh87b.
2. Humid Permeable (HP) Site is assumed to be located in Southeast.
3. For the following soil characteristic input parameters, values were used
from the standard humid impermeable (HI) site: trench, trench_cap, and
sub-trench permeability; trench, sub-trench, and aquifer porosity; density
of waste material; local soil infiltration rate; fraction of residual
saturation; bulk density of soil; runoff fraction; porosity in gravity and
pellicular zone; upward diffusivity; upward hydraulic conductivity; and
pellicular and gravity infiltration capacity.
4. Water usage characteristics (i.e., fractional usage of stream and well
water) values taken from humid impermeable (HI) site input parameters.
5. Humid Impermeable (HI) site is assumed to be located in Northeast.
6. Soil characteristic input parameter values (as listed in 3) of the
standard humid permeable (HP) site were used.
7. Arid permeable (AP) site assumes an arid site located in the Southwest.
11-31
-------
With standard Southeast site, humid permeable, scenario input
parameters, the maximum dose (9.2 mrem/yr) is due to ground water
transport to a well. When soil characteristic parameters (see Table 11-7)
from the humid impermeable site are used, the maximum dose goes down but
occurs earlier. This seems somewhat contradictory, but is due to competing
effects. The impermeable soil characteristics cause water to pool in the
trench,^leading to increased concentrations of nuclides in the leachate'due
to the increased contact time. Even though the leachate has greater
nuclide concentrations, the impermeable soil causes a slower rate of travel
to the well. With the normal scenario, C-14 reaches the well early in the
analysis, but with a small peak, while 1-129 takes longer to arrive at the
well, but with the larger peak dose. In the test scenario, the greater
concentration of the leachate causes the peaks to increase, but the slower
rate of travel causes them to occur later. Therefore, the dose due to C-14
goes up due to its greater concentration in the leachate, but the dose from
1-129 no longer shows up due to its arrival at the well after 1,000 years.
In other words, doses from individual nuclides go up, but occur later. The
nuclide contributing the maximum dose changes from 1-129 in year 777 'to
C-14 in year 390. '
When humid impermeable (HI) site water usage characteristics (i.e.,
fractional usage of well water decreases and stream water increases) are
used the dose goes down due to the greater dilution afforded by the use of
a surface water stream over an aquifer. However, when HI water usage is
combined with HI soil characteristics, doses go up and occur earlier due to
trench overflow and surface water usage combined with some groundwater flow
and eventual discharge to the stream of mobile nuclides. It should be
noted, however, that all doses are very similar for these scenarios.
^ When evaluating the Northeast HI site, doses are greatly reduced when
humid permeable soil characteristics (see Table 11-7) are assumed. This is
because the permeable soil does not allow trench overflow to occur, but
groundwater travel times are still too slow to allow contamination to reach
the well within 1,000 years. Therefore, the maximum dose occurs in year
one, due to atmospheric transport of surface spillage.
Maximum doses are reduced at the arid permeable site when the HI
soil characteristics (see Table 11-7) are used. This is due to ground-
water transport being sufficiently slowed so that nuclides will not reach
the well until after 1,000 years. This causes the max-imum dose to occur in
year two, due to atmospheric transport of surface spillage.
(B) Sensitivity to Disposal Method
The health impact from the disposal of LLW varies depending upon the
disposal method used. In Figures 9-2 and 9-3 the sensitivity of health
impact to a number of different disposal methods are shown and the results
for these methods are discussed in Chapter 9. The analysis in Chapter 9 is
based on specific base case disposal methods which do not include all of
the disposal methods that were assessed. In our broader analysis, a total
of nine disposal methods were considered, including SLF, SLD, ISD, IDD
EMCB, CC, HF, DWI, and DGD. The HF and DWI methods only apply to certain
waste streams, however.
11-32
-------
The results of the analysis of the nine disposal methods are displayed
in Table 11-8, grouped by site location. These results show the sensitivity
of using the various disposal methods at a single site. Three deep disposal
assessments (22, 23, 24) consider smaller volumes and limited waste streams,
as noted. Because of these differences, these three scenarios cannot be
directly compared with the others.
The maximum annual CPG dose and long-term health effects are greatest
from the sanitary landfill method at all sites. The concrete canister
method leads to the lowest CPG dose and health effects at all sites,
although the more rigorous disposal methods offer essentially an equivalent
level of protection. .
The difference in maximum annual CPG doses among the disposal methods
varies depending on location. The range between the least protective and
the most protective is about 50 times more protective for the humid
permeable location, about 500 times more protective for the humid
impermeable location, and at least five orders of magnitude for the arid
permeable location.
The difference in long-term health effects among the disposal methods
also varies depending on location, but to a lesser degree. The range
between the least protective and the most protective methods is about four
for the humid permeable location, about 150 times more protective for the
humid impermeable location, and about 200 times more protective for the arid
permeable location.
The above discussion is based on disposal methods where it is assumed
that as the methods change the waste form changes as well. For example,
when waste is disposed of under SLF methodology all waste is assumed to be
as generated, while for IDD disposal it is assumed that only class A waste
is disposed of as generated and Class B and C are solidified. To gain an
idea of how CPG dose and population health effects vary when only the
disposal method is changed, a review of scenarios 8, 2, 45 and 42 from the
humid impermeable site are shown in Table 11-9, where all waste is di.sposed
of as generated.
From these scenarios it can be seen that as the disposal methods become
more rigorous, while the waste form remains the same, maximum CPG dose and
population health effects are reduced. This is not as true, however, once
the waste is in a solidified form, as can be seen in scenarios 11 and 14,
where ISO disposal is compared to IDD disposal, both with Class B and C
wastes solidified. Under those conditions the reduction in health impact is
much less.
11-33
-------
Table 11-8. Summary of maximum annual CPG doses and long-term health
effects from nine disposal methods at different site locations
Scenario
Disposal
me thod
CPG dose in
peak ,year
(mrem/yr)
Health effects over
10,000 years
(fatal cancers only)
Humid Impermeable Location
8
2
11
14
17
20
22*
SLF
SLD
ISD
IDD
EMCB
CC
HF
Humid Permeable Location
T SLF
1 SLD
10 ISD
13 IDD
16 EMCB
19 CC
23* DWI
Arid Permeable Location
9 SLF
3 SLD
12 ISD
15 IDD
18 EMCB
21 CC
25 DGD
24* DGD
0,77
0.13
1.2E-02
9.3E-03
9E-03
4E-03
1.7E-04
62
35
5.1
5.0.
2.0
1.3
7.3
4.1E-01
2.2E-03
4.4E-05
3.7E-05
3.1E-06
0
0
39
8.1
1.1
0.85
0.64
0.26
0.007
6.2
4.7
3.4
3.4
2.0
1.7
0.036
3.9
3.6
1.7
1.2
0.69
0.31
0.023
0.047
* 22 includes L-CONCLIQ, I-ABSLIQD, L-DECONRS, L-FSLUDGE, L-IXRESIN, and
RAIXRSN; 23 includes L-CONCLIQ, I-ABSLIQD, and L-DECONRS; 24 includes
L-IXRESIN and Class C waste.
11-34
-------
Table 11-9. Summary of maximum annual CPG dose and long-term
health effects when disposal methods vary but the
waste form remains constant
D
Scenario-'-
Waste "As Is"3
8
2
45
42
isposal Method^
A
SLF
SLD
SLD
SLD
B
SLF
SLD
SLD
ISD
C
SLF
SLD
ISD
ISD
Health Effects
CPG Dose in Peak Over 10,000 years
Year (mrem/yr) (fatal cancers only)
0.77
0.13
0.12
0.04
39
8.1
4.6
3.9
Waste Solidified^
11
14
ISD
1DD
ISD
IDD
ISD
IDD
0.012
0.009
1.1
0.9
1. All waste is disposed of at the humid impermeable site.
2. Class A, B, and C waste are disposed of as listed. See Key to
Tables 11-4, 11-5, and 11-6, for a description of wa:ste classes
and disposal methods.
3. All waste disposed of in the "as generated" form.
4. Class A waste disposed of "as generated," while class B arid C are
disposed of in a solidified waste form.
11-35
-------
(C) Sensitivity to Waste Forms
T . 1Theifensitivity °f the results to the waste forms analyzed is presented in
Tables 11-10 and 11-11, grouped by site characteristics. Table 11-10 lists
tests in which the SLD disposal method is used (past practice shallow land
disposal). Table 11-11 lists tests in which the disposal method is according
fern? * PSrt 61 dlsP°sal technology, which require shallow land disposal
tSLD; for Class A and B wastes and improved shallow disposal (ISD) for Class C
wastes. Any NARM wastes are assigned the same disposal method as Class C
wastes.
Regarding the sensitivity of the output to the waste form, a major result
is that solidification appears to provide four or five times the protection
that unsolidified waste forms provide. A secondary finding is that high-
integrity waste canisters provide no additional population health effects
protection when they are filled with wastes in the "as is" form, due to the
long time periods analyzed. In addition, when evaluating CPG doses, maximum
doses actually may go up about 10 percent since releases occur quickly once
containers fail causing a larger peak dose. Incineration of the waste prior
to solidification causes little additional reduction in health impact.
(D) Sensitivity to Regional Waste Mixes
The impact from three different waste mixes is compared to the base case
scenarios (1, 2, 3, and 4, 5, 6) in Table 11-12. There appears to be little
significant difference in impact from regional waste mixes. The maximum range
of CPG doses is less than 50 percent, with the greatest being for the
comparison of scenarios 3 and 97. The maximum range for long-term health
effects is about 70 percent for three of the scenario comparisons. This is
due to the fact that the differences between the U.S. average waste mix and
the various compact s waste mix are not significant for the nuclides and waste
streams causing the majority of the health impacts.
_Two additional assessments were made for the Northeast and Southeast
regional compact areas. The volumes of waste are 470,000 m^ for the
Northeast scenario (102) and 590,000 m3 for the Southeast scenario (101).
The impacts from these scenarios are compared to risks from scenarios 78 and
77, which use a 500,000 mj volume of the U.S. average waste mix. The
maximum CPG doses are essentially the same. The long-term health effects are
within a factor of 1.6.
(E)
Sensitivity to Waste Groups
The sensitivity of the results to the various combinations of wastes
is presented in Table 11-13. The wastes are grouped into LLW-BRC+NARM
LLW+NARM,_LLW only, and LLW-BRC+NARM-CLASS D. The maximum annual CPG doses
fall within 30 percent of each other at any location, as do the long-term
health effects.
11-36
-------
Table 11-10. Suranary of long-terra health effects and maxinum annual
CFG doses from selected waste forms at different site
locations using the SII> disposal method
Waste form
Scenario Class A Class B Class C mm
ttmid
2
27
39
Humid
1
26
38
Arid
3
28
40
Impermeable Location
ASIS ASIS ASIS ASIS
— Incinerated/Solidified —
AS IS SOL SOL SOL
Permeable Location
AS IS AS IS AS IS , AS IS
— Incinerated/Solidified —
AS IS SOL SOL SOL
Permeable Location
ASIS AS IS AS IS ASIS
— Incinerated/Solidified —
AS IS SOL SOL SOL
CPG dose in
peak year
(mrem/yr)
0.13
0.03
0.03
35
13
9.1
2.2E-03
3.0E-04
9.2E-04
Health effects
over 10,000 years
(fatal cancers only)
8.1
6.4
2.2
4.7
3.1
3.9 ' ,
3.6
1.4
2.3
11-37
-------
Table 11-11. Sunmary of long-term health effects and maxinum annual
CPG doses from selected waste forms at different locations
using 10 GBR Part 61 disposal technology
Scena
Hunid
5
30
33
36
45
ttmid
4
29
32
35
44
Waste form
rio Class A Class B Class C
Impermeable Location
AS IS SOL SOL
— Incinerated/Solidified —
AS IS HIC SOL
KC HIC HIC
AS IS AS IS AS IS
Permeable Location
AS IS SOL SOL
— Incinerated/Solidified —
AS IS HIC SOL
( HIC HIC HIC
AS IS AS IS AS IS
NAR4
SOL
SOL
HIC
AS IS
SOL
SOL
HIC
AS IS
CPG dose in
" peak year
(mrem/yr)
0.03
0.025
0.12
0.13
0.12
9.2
13
82
40
35
Health effects
over 10,000 years
(fatal cancers only)
•
2.2
3.4
3.8 '
6.1
4.6
3.9
• 3.1
4.6
4.5
4.7
Arid Permeable Location
6
31
34
37
46
AS IS SOL SOL
— Incinerated/Solidified
AS IS HIC SOL
HIC HIC HIC
AS IS AS IS AS IS
SOL
—
SOL
HIC
AS IS
9.2E-04
3.0E-04
3.9E-03
1.9E-03
2.2E-03
2.3
1.4
3.5
3.5
3.6
11-38
-------
Table 11-12. Suiraary of long-term health effects and maxinun annual
CPG doses from various waste mixes at different site
locations .
Scenario
Waste
mix
Disposal
method
CPG dose
in peak 'year
(mrem/yr)
tealth effects
over 10,000 years
(fatal cancers only)
ftjnid Impermeable Location
2
96
5
99
U.S. average
N.E. compact
U.S. average
N.E. compact
SID
SID
SID/ISD
SLD/ISD
Humid Permeable Location
1
95
4
98
U.S. average
S.E. compact
U.S. average
S.E. compact
SID
SID
SID/ISD
SID/ISD
Arid Permeable Location
3
97
6
100
SID
U.S. average
Rocky Mtn. compact SID
U.S. average SLD/ISD
Rocky Mtn. compact SLD/ISD
0.13
0.15
0.03
0.03
35
37
9.2
9.7
2.2E-03
1.5E-03
9.2E-04
l.OE-03
8.1
12
2.2
3.8
4.7
2.8
3.9
2.4
3.6
2.5
2.3
2.1
11-39
-------
Table 11-13.
Summary of long-term health effects and maximum
annual CPG doses from four waste groups at different
site locations using 10 CFR Part 61 disposal technology
Scenario
Waste
group
Humid Impermeable Location
5 LLW-BRC+NARM
51 LLW+NARM
60 LLW
69 LLW-BRC+NARM
-CLASS D
Humid Permeable Location
4
50
59
68
LLW-BRC+NARM
LLW+NARM
LLW
LLW-BRC+NARM
-CLASS D
Arid Permeable Location
CPG dose in
peak year
(rarem/yr)
3.0E-02
2.3E-02
2.3E-02
3.0E-02
9,2
7.0
7.0
9.1
Health effects
over 10,000 years
(fatal cancers only)
2.2
1-7
1.7
2.2
3.9
2.9
3.0
3.9
6
52
61
70
LLW-BRC+NARM
LLW+NARM
LLW
LLW-BRC+NARM
-CLASS D
9.2E-04
6.8E-04
6.8E-04
9.2E-04
2.3
1. 7
1.7
2.3
11-40
-------
An additional test of waste groups compares scenarios 1, 2, 3 against
scenarios 65, 66, 67, from which Class D wastes are removed. In these
scenarios all wastes are disposed of by the SLD method. This comparison
indicates that removal of'the Glass D waste, because it is solidified, has
very little effect on estimated impact. This is due to the fact that the
Class D waste, which is always solidified, results in little release to the
environment, thus very little health impact.
Removing the BRC component from the wastes (4,5,6) causes the impacts to
increase. This is because the size of the disposal site is kept constant.
Therefore, when lower activity BRC wastes are removed, they are replaced with
higher activity LLW, which causes impacts to increase. The reason that
removing NARM or Class D components from the wastes causes little change_in
'impact is due to the disposal method that was analyzed (SLD/ISD). In this
method the NARM and Class D wastes are solidified. This causes the activity
to be released very slowly, thus not reaching the population until long after
the time period analyzed has been exceeded. Therefore, since these nuclides
do not reach the population, removing them from the waste has no effect on
impacts.
It is worth noting that the NARM wastes included in this analysis are not
regulated at the Federal level. Many of these NARM wastes are presently being
stored. For the purpose of our analyses, however, we assume that when these
NARM wastes are disposed of, they are disposed of in ISO facilities. The
total U.S. health effects from their disposal, using this assumption, is less
than one health effect. This is from 20 years of waste, evaluated for 10,000
years. If these wastes were disposed of in a sanitary landfill, the health
effects would increase to 71. -The difference in health effects is due to the
solidification of the waste when using the ISD method and the very low
releases which result.
(F) Sensitivity to Waste Volumes > .
The sensitivity of the results to various waste volumes at a site is
presented in Table 11-14 for two different disposal methods. The long-term
health effects vary directly with the volume of waste at all locations, i.e.,
when the volume of waste at a site is doubled, the number of fatal cancers
doubles. The maximum CPG dose for the humid impermeable site also varies
directly with the volume of waste. However, the dose estimates for the other
two locations do not demonstrate a similar relationship. For the humid
permeable site, the CPG dose increases 30 to 40 percent when the volume is
doubled, while for the arid permeable site, the CPG dose increases 40 to 50
percent when the volume is doubled, as explained below.
-Additional comparisons of waste volume effects can be made by evaluating
variations in volumes caused by changes in the waste groups (LLW, BRC, NARM
combinations). These volume tests are presented in Table 11-15 using 10 CFR
Part 61 disposal technology (SLD/ISD). The same observation applies to these
tests as to the tests made specifically to observe the effect of volume
changes. The long-term health effects vary directly with waste volume at
all sites and for all waste group combinations. The maximum annual CPG dose
varies directly with volume for the humid impermeable site, but varies less
than directly with volume at the two permeable sites, as explained below.
11-41
-------
Table 11-14.
Siranary of long-term healtfi effectjs and
annual CFG doses from different volumes of
Waste
volume
Scenario (1000 m3)
Disposal
method
CPG dose in
peak year
(mrern/yr)
. , Health, effi-xsts
over 10,000 years
(fatal cancers only)
Humid linpermeable Location
75
2
72
81
5
78
100
250
500
100
250
500
SID
SID
SID
SID/ISD
SID/ISD
SID/ISD
0-05
0.13
0.25
0,01
0.03
0.06 .'
3.2
"** *V '
8.1
\>9 f.
16
0.9
2,2
**V 4*
.'••,,'. 4'5 '' -. '.!
Huraid Permeable Location • < ? . ' ; '•" • 'r r •'•.:'••
74
1
71
80
4
77
Arid Permeable
76
3
73
82
6
79
100
250
500
100
250
500
Location
100
250
500
100
250
500
SID
SID
SID
SID/ISD
SID/ISD
SID/ISD
SID
SID
SID
SID/ISD
SID/ISD ,
SLD/ISD
22
35
48
5.6 , ,
9.2
' ' 13 •••'
1.3E-03
2.2E-03
3.1E-Q3
5.6E-04
9.2E-O4
1.3E-03
,J.. ,g • • ' ••- :;.."
+r* *
4.7
9.4 ;
1.6
, - *»Xf
3.9
> > *^™ if •
1.4
3,6 '' .' '•'
74
0.92
2.3
'"" 4,6
11-42
-------
Table.,, 1,1-15. Summary of. long-term health effects and maximum
annual CPG doses from different volumes of waste
using 10 CFR Part 61 disposal technplogy
t '*
Scenai
Humid
48
51
54
57
60
63
Humid
47
50
53
56
59-
62
Arid
49
52
55
58
61
64
. " , , .Waste -
Volume
-io •'"'• ' (1000 m3) ' " •"
Impermeable Location
373
250
•;. 170
'' 366
250
170
Permeable Location
373
250
170
366
250 '
170
Permeable Location
; 373 '
'" " 250
170
• ' :', , .366
.•••., V 250
""170
Waste
group
LLW+NARM
LLW+NARM
LLW+NARM
LLW
LLW
LLW
-' -- *
LLW+NARM
LLW+NARM
LLW+NARM
LLW ,
LLW '
LLW- ""-*
LLW*NARM
LLW+NARM
LLW+NARM
LLW
LLW ,
LLW " ;
PPG dose in
peak year
(mrem/yr)
3.4E-02
2, 3E-02
1.5E-02
3.3E-02
2.3E-02
1.5E-02
8.6
7.0
.' ' . '; 5'7'
8.6 ..',
7,.0 '
• , •- 5.7 - -
• 8i'4E-04
6.8E-04
1 5.5E-04
8,.4E-04 .
6.8E-04
5.5E-04
Health Effects,/
over 10,000 years
(fatal cancers .only)
2.5
1.7
1.1
2.5
1.7
1.2
'/'' j- ; ; ».:• .•:
4.4
2.9
2.0
, 4,3
3.0
. 2.0
: •. ' •. 2.6
' •••:•'., 1.7
1.2
. , ' 2.5
1.7
1.2
• , , „„,'„•. ^.\ V; ... " - "' I ' ' - . '
11-43
-------
The sensitivity of maximum annual CPG doses to site size is shown in
Figure 11-2. The changes in CPG dose from disposing of 100,000, 170,000,
250,000, 373,000, and 500,000 m3 of the U.S. average mix of LLW by the
10 CFR Part 61 disposal technology under three different hydrogeologic/-
climatic settings are shown. Based on this figure and Table 11-15, it can
be seen that with a linear increase in inventory: (1) the long-term health
effects increase linearly and (2) the CPG dose increases (a) linearly with
the inventory in a humid, .low-permeability (overflow) setting and (b) at a
less than linear rate for humid and arid permeable (contaminant movement to
ground water) settings. This less than linear increase for the permeable
sites is due to areal dilution of ground water caused by the larger site
area required for the source term. This areal dilution does not occur at
the impermeable site, since the surface water pathway predominates, for
which the site area is less important in calculating water concentrations.
It is worth noting, however, that the CPG dose did not exceed 15 mrem/yr for
a 500,000 m3 capacity site in a humid permeable setting and was always
less than 0.1 mrem/yr for the other settings.
An additional set of analyses relating to site volume were performed
to evaluate a scenario where the waste was volume reduced, such as by
compaction, but the total activity remained the same. This would result
in a smaller site size, but with the same radionuclides and activity.
This analysis was performed at humid permeable and humid impermeable
sites, using 10 CFR Part 61 disposal technology (consisting of SLD for
Class A and B waste and ISD for Class C waste). Two separate sets of
analyses were performed; the first with the well located the same distance
from the center of the disposal site (which results in a larger buffer zone,
since the site area will be smaller) and the second with the well moved
closer to the site (resulting in the same buffer zone distance as in the
base case analyses). The results are shown in Table 11-16.
Peak dose projections for the humid permeable disposal site increase as
the disposal site area decreases. The increase in dose is due to the fact
that, as the site area is decreased, the concentration of the site leachate
increases since the volume of water into which the contaminants released
from the site are diluted declines. The PRESTO-EPA-CPG code calculates an
aquifer dilution volume based on the disposal site width, the aquifer
thickness, and the aquifer dispersivity, as described in the PRESTO-EPA
Methodology Manual (EPA87). As the site area is reduced, the groundwater
flow rate drops and, in light of constant rate of radionuclide releases from
the waste, contaminant concentrations rise. Radionuclide releases are
constant despite the fact that contaminant concentrations in the trench are
higher. While the higher concentrations result in higher leachate
concentrations, the volume of water draining from the trench is reduced.
These effects balance one another such that total curie releases remain
unchanged.
11-44
-------
10.0
1.0—=
iu
o 0.1
o
Q.
o
0.01
Q.
?=
5—
~9—
2-
"91
2—
MH^MMOM^Mn
5.6
,_ ;
«
s ?
*
-
"*
I
100
13
8.6
5.7 7'°
ui
-------
Table Ll-16. Results summary of reducing waste volume
with total activity constant
Scenario^-
CPG Dose in Peak Year (mrem/yr)
Humid Permeable Site
Standard Scenario
Volume Decreased by 50%, larger buffer zone
Volume Decreased by 75%, larger buffer zone
Volume Decreased by 50%, same buffer zone
Volume Decreased by 75%, same buffer zone
Humid Impermeable Site
Standard Scenario
Volume Decreased by 50%, larger buffer zone
Volume Decreased by 75%, larger buffer zone
Volume Decreased by 50%, same buffer zone
Volume Decreased by 75%, same buffer zone
9.2
12
15
13
17
0.03
0.03
0.03
0.03
0.03
1.
All scenarios are for disposal in a site using 10 CFR Part 61 disposal
technology. Volume is reduced as listed, with buffer zone distances
either increasing or staying the same, as discussed in the text.
11-46
-------
The elevajzed doses seen for the smaller sites are greater still
when the distance from the site boundary to the well (buffer zone) is
held constant. With a decline in the distance to the well (from site
center) the degree of dispersion seen in the aquifer prior to arrival
at the well is minimized, thereby reducing the overall aquifer dilution
volume. This reduction results in higher aquifer nuclide concentrations
and, consequently, higher doses from the use of the contaminated water.
In contrast to the humid permeable site, reductions in disposal
site area has no affect on peak dose projections at the humid impermeable
site. The peak dose at this site arises from the overflow of the waste
trenches and the subsequent transport of released contaminants to a
stream. While the concentrations of nuclides in the overflow water are
higher due to waste volume reduction, the volume of water exiting the
trenches is reduced as it is based on trench surface area. The end
result is that total curie releases remain constant.
These releases are diluted in the same amount of water as the
dilution volume of the receiving stream in either case and, therefore,
are unaffected by the site area. Consequently, contaminant concentra-
tions are unchanged from the base case simulations and peak doses remain
constant, regardless of the size of the buffer zone.
From these results it can be seen that at permeable sites, if volume
reduction takes place such that the outcome is a smaller site with the
same total activity, peak CPG doses will increase. The increase will be
even greater if the size of the buffer zone, (distance from trench
boundary to well) remains constant, resulting in the well moving closer
to the site center. It should be noted, however, that even with a 50%
decrease in site size, the peak CPG dose is still not above.15 mrem/yr
and that it can be reduced to 12 mrem/yr by using the smaller site size
as a means of increasing the .size of the buffer zone.
(G) Sensitivity to the Time Horizon
The maximum annual CPG doses for a 1,000- and a 10,000-year
time horizon are shown in Table 11-17, where 10 CFR Part 61 disposal
technology is analyzed at all three hydrogeologic/climatic sites.
The peak CPG doses for the two humid locations are the same for both
periods. This finding reflects the fact that the movement of the
high-dose, mobile radionuclides to the CPG at humid locations occurs
relatively quickly, so that extending the analysis past 1,000 years has
no effect on peak CPG dose
At the arid site, however, extending the analysis period to
10,000 years causes an increase in the peak dose from about 0.0009 to
about 0.02 mrem/yr. This increase is due to the nature of the arid site,
where radionuclide transport through the groundwater pathway is very slow
for even the more mobile nuclides. In the 1,000 year analysis,. C-14
reaches the GPG" and causes the peak dose, with other less mobile nuclides
remaining in transit. Extending the analysis period to 10,000 years
allows some of the less mobile nuclides, such as 1-129, to reach the CPG,
causing a greater peak dose. It should be noted, however, that^all
doses, whether from C-14 or 1-129, are very small at the arid site.
11-47
-------
Table 11-17 Sunuary of maximum annual CPG doses from IIW
using different time periods and 10 CER. Part 61
disposal technology
r '
Scenario
Humid Impermeable location
5
104
Humid Permeable Location
4
103
Arid Permeable Location
6
105
Period
Analyzed
(yr)
1,000
10,000
1,000
10,000
1,000
10,000
CPG dose in
peak year
(mrem/yr)
3.0E-02
3.2E-02
9.2
9.2
9.2E-04
1.7E-02
11-48
-------
In Figures 11-3, 11-4, and 11-5 the maximum annual CPG dose is plotted
over time (with time zero corresponding to site closure) for conventional
shallow land disposal and 10 CFR Part 61 disposal technology. For
conventional shallow land disposal, all waste is in the as-generated waste
form, while for 10 CFR Part 61 disposal technology, much of the waste is
solidified (Class B and C). As can be seen from the three figures, the
combination of more rigorous disposal practices and the solidified waste
form for the 10 CFR Part 61 disposal technology leads to similar behavior
of CPG dose versus time as for conventional disposal for the radionuclide
releases, but, in general, significantly smaller dose rates.
Figure 11-3, shows the relationship of dose rate over time at the humid
permeable site. The annual dose rate rises quickly from the very mobile
radionuclides H-3 and C-14. As this dose diminishes, the Tc-99 and then the
1-129 reach the receptor. The maximum annual dose rate is from 1-129,
around year 700. After about 900 years the dose rates start, to level off.
Less mobile radionuclides will continue to reach the CPG after 900 years,
but the doses will remain below that of 1-129..
Figure 11-4 shows the relationship of dose rate over time at the humid
impermeable site. There is very little dose until the trench cover fails in
year 100. Failure of the trench cover allows a quick release of many
nuclides and a large peak in dose rate soon after, due to overflow of the
trench and transport via surface water. The peak dose is mainly due to
1-129 in about year 200. The peak drops quickly, leveling off after about
year 300.
The arid permeable site is depicted in Figure 11-5. The small dose.
rate in the first few years is due to windblown transport of nuclides-
spilled during operation. The most mobile nuclide, C-14, reaches the CPG
about year 900. As discussed earlier, extending the analysis past 1,000
years results in a larger peak dose from 1-129, although this peak dose is
still very small. In general, the plot of CPG dose versus time for the arid
site will look similar to that of the humid permeable site, except that the
doses will be much lower and the time at which the peaks occur will be
shifted to the right.
Cumulative population health effects are assessed for 10,000 years, but
occur at different rates over that period. This can be seen in Table 11-18,
where health effects are broken down by time period. Looking at the U.S.
totals, as opposed to the three specific sites, and assuming 10 CFR Part 61
disposal technology, we see that 43 percent of the population health effects
occur in the first 500 years. An additional 10 percent occur over the
following 500 years, with 47 percent occurring during the last 9000 years of
the analysis.
It should be noted that these results are estimates of the total
potential U.S. health effects, weighted for the waste in each of the three
hydrogeologic regions. These results would be less for each specific region
and would vary depending on the region.
11-49
-------
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«-T
£
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Ul
CO
o
° 10
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(
n
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j
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a
a
a
a
< Q
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100
•niae /..^«..
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r
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1 j i • .
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a
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V-.' ***., "'• j
.. ..... ».,. .1 1 1 '
;
i CONVENTIONAL
. , SHALLOW; DISPOSAL
10 CI"R Part 61
•• DISPOSAL
.TECHNOLOGY ,
1000 •
.-_;,%,- •• •,-,-. t ••-", i.
/•^ -in,' ,ri", ' , • _
Figure 11-3. CPG Dose Versus Time for the Humid Permoable
Hydrogeologlc Site
''" '"'' ' ' !
-------
• 0.08 .
2002 -
Oflfl .
<
1
/'•:',,
'- b
1 2C
1
a ...,'•, • •
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*********
10 10
(
1
.
00
CONVENTIONAL
SHALLOW DISPOSAL
10 CFR Part 61
> DISPOSAL
TECHNOLOGY
TIME (years)
Flour* 11-4. CPQ Dos* Vtrsus Time for the Humid Impermeable
HydrogeoIoflSc Site
U-5L-
-------
u.uua-
*•«••
W •
>,
!
£. 0.002
Ul
H
2
w
g 0.001
(3
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-
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ODD
a Dlf
,a
a
a ••••<
a •
^
-
0.000
200
800
400 600
TIME (years)
Figure 11-5. CPG Dose Versus Time for the Arid Permeable
Hydrogeologlc Site
H? CONVENTIONAL
SHALLOW DISPOSAL
10 CFR Part 61
» DISPOSAL
TECHNOLOGY
1000
11-52
-------
Table 11-18. U.S. total health effects over time, from disposal of
20 year U.S. waste volume using 10 CFR Part 61 disposal
technology
Time Period (yr)
0 - 100
101 - 500
501 - 1,000
1,001 - 10,000
Total (0 - 10,000)
Cumulative Health Effects1
1.6
10.6
2.8
13.3 -
28.3
% of Total
5.7
37.5
9.8
47.0
100
1. Health effects are for fatal cancers and serious genetic effects.
11-53
-------
If the assessment that was described above was done on a regional basis,
the population health effects for the humid sites would be very similar to
the results shown. About 50% of the health effects occur before 1,000 years
and the other half after 1,000 years. For less restrictive disposal
methods, such as conventional shallow, or with less waste solidified, the
proportion of health effects occuring before 1,000 years would increase.
For the arid sites, a greater percentage of the health effects occur in
later years due to the slower rate of groundwater transport. In general,
numbers of health effects occuring over the hydrogeologic/climatic regions
or between various disposal methods are fairly constant, but with the health
effects occuring later at the arid sites and with more restrictive disposal
methods or solidification of the waste.
In assessing long-term, cumulative population health effects over
10,000 years, a question is raised as to what.health effects could
potentially occur due to radionuclides which do not reach the population
within 10,000 years. There is some concern, of course, that any analysis
over 10,000 years contains so much inherent uncertainty that estimates at
extremely long time periods are ludicrous. Some rough estimate may be
instructive, however.
In Table 11-19, the radionuclides which reach the population
(breakthrough) after 10,000 years at the humid permeable site are listed,
along with their half-life and the percentage of their original activity
which would remain at the time of breakthrough. As can be seen from this
table, many of these less-mobile nuclides will have decayed
completely prior to breakthrough. For a few, however, a significant
percentage of their original activity remains.
For the nuclides with a significant fraction of their original inventory
remaining at time of breakthrough, a rough estimate of potential health
effects can be made as outlined in Table 11-20. The percent of inventory
remaining (i.e., not yet decayed) at breakthrough can be multiplied by the
original inventory to determine the inventory remaining at breakthrough.
This inventory is very conservative, however, as much of it could still be
in_the trench or in transit at the time some reaches the population. Using
this conservative inventory, however, an estimate of potential health •
effects can be made by multiplying the inventory by the radionuclide
specific HECF values for the humid permeable site. The HECF values, which
are described in detail in Chapter 8, provides an estimate of the health
effects to the population from releases of activity to the basin, assuming
the same water useage patterns for the basin population as for the local
population.
11-54
-------
Table 11-19. Radionuclides which reach the population after
10,000 years and the percent of their original
activity remaining at time of breakthrough
Nuclide
Fe-55
Ni-59
Ni-65
Sr-90
Nb-94
Ru-106"
Cs-134
Cs-135
Cs-137
Ba-137m
Eu-154
Po-210
Pb-210
Bi-214
Pb-214
Ra-226
U-234
U-235
U-238
Pu-238
Pu-239
Pu-241
Am-241
Pu-242
Am-243
Cm-243
Cm-244
Half-Life
(yr)
2.7E+0
8.0E+4
1.1E+2
2.9E+1
2 . OE+4
l.OE+0
2.1E+0
2.3E+6
3.0E+1
3.0E+1
8.5E+0
3.8E-1
2.0E+1
8.6E-1
5.1E-5
1.6E+3
2.5E+5
7.0E+8
4.5E+9
8.8E-H
2.4E+4
1.3E+1
4.6E+2
3.8E+5
7.4E+3
3.2E+1
1.8E+1
Breakthrough Time
to Population*
(yr)
7.6E+5
1.9E+4
1.9E+4
1.4E+4
4.5E+4
2.8E+4
1.1E+5
1.1E+5
1.1E+5
1.1E+5
5.1E+5
2.8E+4
2.8E+4
2.8E+4
2.8E+4
2.8E+4
9 . 6E+4
9.6E+4
9.6E+4
4.5E+5
4.5E+5
. 4.5E+5
l.OE+7
4.5E+5
l.OE+7
4.2E+5
4.2E+5
% of Original
Inventory Remaining
at Time of Breakthrough
0
85
0
0
21
0
0
97
0
0
0
0
0
0
0
0
77
100
100
0
0
0
0
44
0
0
0
*Breakthrough time is the time at which nuclides reach the population and is based
on the humid permeable site.
11-55
-------
Table 11-20.
i
ui
CT\
Potential population health effects from
nuclides which reach the population after
10,000 years
Nuclide
Ni-59
Nb-94
Cs-135
U-234
U-235
U-238
Pu-242
% of Original In-
ventory Remaining at
Time of Breakthrough1
85
21
97
77
100
100
44
Original
Inventory (Ci)2
3.0E+2
3.1
2.8
3.8
7.5E-2
7.2E-1
9.7E-2
"HECF3
2.0E-5
5.2E-2
5.9E-3
3.1E-4
3.5E-4
3.0E-4
6.2E-3
Total Potential Health Effects
Potential
Health Effects4
5.1E-3
3.4E-2
1.6E-2
9.0E-4
2.6E-5
2.2E-4
2.6E-4
6.2E-2
1. Breakthrough time is the time at which nuclides reach the population and is based on the humid
permeable site.
2. Original inventory is the total activity for each nuclide used in the base case runs.
3. HECF values are the health effect conversion factors from the humid permeable site used to
calculate health effects from radionuclides released to the regional basin.
4. Potential health effects are calculated by multiplying the amount of inventory remaining at time of
breakthrough by the appropriate HECF value.
-------
Using the methodology described above, it can be seen that potential
health effects from radionuclides which reach the population after 10,000
years would be very low. Using conservative assumptions, total potential
health effects from these nuclides are much less than one health effect.
The results of this analysis would not vary significantly for the other
hydrogeologic sites. An additional point can be raised as to what the
contribution would be from radionuclides which reach the well prior to
10,000 years, but are not completely consumed in the regional basin.
This is evaluated in Section (I), as part of the global analysis.
(H) Sensitivity to Radionuclides
An additional consideration regarding health impacts from LLW disposal
is identifying the radionuclides that contribute most to the health impacts
under different disposal situations. Since a unit curie and unit volume
methodology was used to calculate population health effects (see Chapter 8),
the results from these separate analyses can be used to illustrate which
radionuclides could be most sensitive under various disposal scenarios. The
analyses and results are described in the following sections for population
health effects with a separate discussion of CPG peak doses.
The "unit curie" methodology, wherein the population health effects
from disposing of one curie of each radionuclide of interest by a specific
combination of disposal- methods, waste forms, and hydrogeologic/climatic
conditions is modeled (see chapter 8), is used to identify potentially
important radionuclides. The relative importance of radionuclides change,
depending on hydrogeologic/climatic setting, the form of the waste, and the
critical release, transport, and exposure pathways.
Where ground water pathways are important, such as at the humid
permeable and arid permeable sites, the majority of the population health
effects are contributed by long-lived, mobile radionuclides with high risk
factors, such as C-14, Tc-99, 1-129, and Np-237 . Less mobile radionuclides,
such as Nb-94, Cs-135, and 'Cm-243, become important in cases where the
trenches overflow, such as at the humid impermeable site. When the trenches
overflow, these radionuclides are discharged directly onto the land surface
and subsequently into surface waters, after retardation by the soils.
For ground surface exposure pathways, such as in the first years after site
closure when atmospheric transport of nuclides spilled during operation are
important, gamma-emitting radionuclides such as Co-60, Cs-134, and Cs-137
can dominate. Figure 11-6 shows the relative importance, on a per curie
basis, of selected radionuclides in each hydrogeologic region as derived
from a "unit curie" analysis.
The contribution to health impact of individual radionuclides
depends on the concentration of the nuclide in the waste. The "unit volume"
analysis (see Chapter 8), on a waste stream by waste stream basis, is useful
for identifying potentially important waste streams and radionuclides. In
the "unit volume" calculation, one cubic meter of waste is loaded with the
scaled concentration of activity for each radionuclide in that waste stream.
The population health effects are then calculated based on the unit volume
being disposed of by specific disposal methods, waste forms, and hydro-
geologic and climatic combinations.
11-57
-------
0.5
0.0
1.0-
ill
O
cc
111
o
<
UJ
cc
0.5-
0.0-
1.0-
0.5-
0.0-
J-XC*?'
ARID PERMEABLE
SITE
HUMID PERMEABLE SITE
HUMID IMPERMEABLE SITE
- «
«
Figure 11-6. Relative Impact from the Disposal of a Unit Curie of
Various Radionuclides by Hydrogeoiogic Setting
11-58
-------
Figure 11-7 identifies the important radionuclides in each waste stream
for three different hydrogeologic/climatic settings, based on the "unit
volume" analysis. This allows identification of important radionuclides on
a waste stream basis; for example, how specific radionuclides can shift in
importance, depending on the hydrogeologic/climatic setting. As can be
seen, in the permeable sites, where groundwater transport is the major
pathway, C-14 is the most significant nuclide due to its long half-life,
mobility, and high risk factor. Note however, that this is based on a
unit-volume approach, with one cubic meter of each waste stream. The
results can change, with a full inventory. At the impermeable site, the
significant radionuclides vary due to the predominance of the overflow and
surface water pathway, which allows less mobiles nuclides, such as Am-241 to
dominate certain waste streams.
If Figure 11-6 is compared with Figure 11-7, it can be seen how
radionuclides which were important on a per curie basis are no longer
important on a waste stream basis. Neptunium-237 is a good example. In
the "unit curie" analysis, Np-237 is identified as potentially the most
critical radionuclide because of its high mobility in water pathways, its
high toxicity, and its long half-life. However, since only negligible
concentrations of Np-237 are in commercial LLW, Np-237 is usually not
important, as can be seen from its absence in Figure 11-7.
Finally, we can evaluate a fully loaded, 250,000 m3 site to determine
which nuclides contribute the greatest number of population health effects
for the base case 10' CFR Part 61 disposal technology scenario. Table 11-21
shows the radionuclides that contribute the greatest percentage of the
population health effects at a site using 10 CFR Part 61 disposal technology
at the three site locations. As can be seen from Table 11-21, C-14 is the
critical radionuclide, contributing 90 percent of the population health
effects at the humid permeable site and 95 percent at the arid permeable
site. At the humid impermeable site, with its overflow pathway, C-14
contributes only 70 percent of the population health effects, with Am-241
contributing most of the remainder. The importance of C-14 is due to its
high mobility, long half-life, and, to a lesser degree, its relatively large
source term and high dose conversion factor.
The unit curie and unit volume methodology, which was used for
calculating population health effects, was not used in calculating peak
doses to the CPG. The impact that is determined for the CFG, maximum
annual dose rate, is not amenable to that type of a methodology, as nuclide
concentration over time is required. Instead, analyses are done directly
with a full source term as described in Chapter 8.
The nuclides which are most critical to the CPG peak dose at each
of the three sites, using a complete source term and the 10 CFR Part 61
disposal technology, are shown in Table 11-21. At the humid sites, 1-129 ,
is the major nuclide, while at the arid site, C-14 dominates. These results
can be most easily explained by reviewing Figures 11-3, 11-4, and 11^-5.
11-59
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RELATIVE IMPACT (%)
i
| WASTE STREAMS
L-IXRE8IN
L-CONCLIQ
L-FSLUDQE
P-FCARTRQ
L-DECONRS
L-NFRCOMP
L-COTRASH
L-NCTRASH
I-COTRASH
N-LOTRASH
I-ABSLIQD
I-BIOWA3TE
N-LOWASTE
N-ISOPROD
N-SOURCES
N-TRITIUM
N-TARQETS
R-RAIXRSN
R-RASOURC
c
LU
ID IMPERMEABLE S
S
X
IO 4k 09 CD C
' ? ? ? ? ]
c-14" :%;;"J
AM141 |
ff:,v'l
nm!i?i! „ 1
Wlfff
C-14:
AMI41J
AM*V J
C-14
>
i C
LU
H
(/)
MEABLE
C-14 |5
•;w ;
C-14
C-14
AMI41 ' ]
C-14 ]
C-141"11'"""''"1 '
M-*
'9-a»° 1
fp-tio ;•• ; J
X
> O 0
OS O)
&^Wm$m
«)
;isj
O M 4k o OJ C
O O OOO PC
J T 1 1 T
C-14 J
b-u
'i-i'i
c-u
C-14
01214
8!?K
Figure 11-7. Dominant Radionuclides When Unit Volumes of Various
Waste Streams are Disposed of in Three Different
Hydrogeologic Regions
-------
Table 11-21. Critical radionuclldes at model LLW site
Scenario
Critical
Radionuclides
Percent of total health impact
caused by critical radionuclides
Peak CPG Population
Dose Health Effects2
Humid
Permeable
Humid
Impermeable
Arid
Permeable
1-129
C-14
Other
1-129
C-14
Am-241
Other
C-14
Other
93%
7%
78%
22%
100%
90%
10%
70%
20%
10%
95%
5%
1. Model site assumes 250,000 m3 of U.S. average waste mix,
using 10 CFR Part 61 disposal technology.
2. Approximate values.
11-61
-------
At the humid permeable site (Figure 11-3), where over 90 percent of
the peak CPG dose is due to 1-129, one sees that C-14 contributes a small
peak early in the analysis (about year 200), but that at the time of the
main peak (about year 800), 1-129 dominates. This is contrasted with the
humid impermeable site (Figure 11-4), where 1-129 contributes 78 percent of
the peak CPG dose and C-14 contributes 22 percent. The reason that both
nuclides contribute toward the peak CPG dose is due to the overflow pathway
that^predominates at the humid impermeable site, which allows both of the
nuclide peaks to occur at about the same time. While both peaks occur at
approximately the same time (about year 200), the contribution from 1-129 is
higher.
At the arid permeable (Figure 11-5), the critical radionuclide is
C-14. As discussed earlier in this chapter, this is due to the long travel
times at the arid site, which causes 1-129 to reach the CPG after the 1,000
year analysis is over. If the analysis was extended to 10,000 years, the
percentages would be similar to those of the humid permeable site, although
the peak CPG dose would still be very small.
(I) Additional Analyses
Additional analyses were performed to answer specific questions related
to the scenarios about effects of a buffer zone, the assessment of global
health effects, and the- importance of the volatility fraction on
incineration doses. Because the incineration pathway relates to the BRC
scenarios, the sensitivity analyses of volatility fraction are discussed in
Chapter 10. The buffer zone and global analyses are discussed below.
A set of analyses were performed to see what the effect of reducing the
size of the buffer zone (the distance from the trench boundry to the nearest
well) around a disposal site would have on peak CPG doses. These analyses
were performed at a humid permeable and an arid permeable site using 10 CFR
Part 61 disposal technology. The results are shown in Table 11-22.
The results show that for both sites the CPG peak doses will increase
as the buffer zone is reduced. This effect is mainly due to a decrease in
the Amount of dispersion that takes place as the nuclides move through the
aquifer to the well. The dispersion is reduced since a smaller buffer zone
results in the nearest well being closer to the disposal site. In addition,
because the well is closer, nuclides reach the well sooner, which results in
somewhat less decay. This effect is minor, however, since the peak dose is
due to long-lived nuclides occuring in later years. As can be seen above,
the year of peak dose does not vary at the arid permeable site. This is due
to the peak not having been reached by the end of the 1000 year analysis,
although the later peak will remain very small.
11-62
-------
Table 11-22.
Results of varying size of buffer
zone on peak CPG dose
Scenario1
Peak CPG Dose
(mrem/yr)
Humid Permeable Site
Standard Scenario^
Buffer Zone Decreased by 50%
Buffer Zone Decreased by 75%
No Buffer Zone
Arid Permeable Site
9.2
9.5
9.7
9.9
Year of Peak
CPG Dose
777
754
742
730
Buffer Zone Increased
Standard Scenario2
Buffer Zone Decreased
Buffer Zone Decreased
No Buffer Zone
by
by
by
100%
50%
75%
8.
9.
9.
9.
9.
6E-4
2E-4
6E-4
8E-4
9E-4
1000
1000
1000
1000
1000
1. All scenarios assume disposal using 10 CFR Part 61 disposal technology.
2. Standard scenario assumes a 100 meter buffer zone, which is
the distance from the trench boundry to the nearest well.
11-63
-------
In calculating population health effects, the HECF is used to estimate
health effects to the regional basin population. The HECF methodology
assumes that water useage patterns for the basin population are the same
as the local population. As discussed in Chapter 8, this leads to radio-
nuclides reaching the downstream basin prior to 10,000 years, but since the
contaminated water is not completely consumed by the downstream population,
radionuclies leave the regional basin and enter the ocean. The nuclides
reaching the ocean are ignored in the analysis, with no health effects
resulting from their potential consumption. This was considered reasonable,
since it was felt that global population health effects due to eventual
consumption of radionuclides entering the ocean would be minimal. To
further evaluate this assumption, a rough global assessment was performed.
The analysis was performed by calculating nuclide-specific, health
effect conversion factors for the consumption of ocean fish and seafood,
based on nuclide transfer factors (water-to-fish and water-to-seafood),'
annual average consumption rates for fish and seafood, and world population,
as outlined below. These were compared to the health effect conversion
factors for fish consumption from the basin river (the source of the
majority of the cumulative basin health effects), the calculation of which
is described in Chapter 8.
The^health effects conversion factor for ocean fish and seafood
consumption is given by the following equation:
HEF.|
(P/R)
i [BfiUf + BsiUs]/XRi
Where:
HEF,-
P
R
Bfi
B .
si
Uf
U
cancer deaths from ocean fish and seafood consumption
per curie of nuclide i entering the ocean
world population consuming the ocean fish and seafood
volume of upper 75 m of ocean (L)
nuclide transfer factor, water-to-fish (Pc*/kS £ist>.)
pCi/L water
nuclide transfer factor, water-to-seafood (PCj-/kg seafood
pCi/L water
annual ocean fish consumption rate per person (kg)
Annual seafood consumption rate per person (kg)
health risk per curie of nuclide i ingested by the population
effective fraction of nuclide removal rate from top 75m of
ocean (yr~l).
11-64
-------
The value of P used is 10 billion people worldwide. R has a value
of 2.7E+19 litres. The parameters Uf and Us equals 6.9 and 1.0
kg/person-yr, respectively (NRC77). The values of Bfi and Bsi are
given in columns 2 and 3 of Table 11-23 for each nuclide. These values are
obtained from NRC77. The conversion factor (H/O^ is calculated from the
DARTAB subroutine of PRESTO-EPA. The effective nuclide removal rate is
around 0.2 yr-1 for long-lived nuclides, based upon comparisons-of the
present calculations to ratios given in EPA82.
The resulting ocean fish and seafood health effect conversion factors
(HEF£> for the ocean scenario were divided by the river fish health
effects factors for the river scenario, as listed in Ro87, to obtain the
ratios given in column 4 of Table 11-23. These ratios demonstrate that
health effects over 10,000 years from nuclides entering the oceans and
ingested in ocean fish and seafood are all less than one percent of the
health effects of nuclides entering the rivers and ingested in river fish,
except for americium, which is less than four percent of the. corresponding
health effects from the river scenario. Based on the results of this
analysis, it appears reasonable to ignore the contribution of global health
effects in our analysis of population health effects.
11.3.3 Summary and Conclusions of Scenario Sensitivity Analyses
' Based on the results of the scenario sensitivity analyses, a number^
of conclusions can be made on the sensitivity of the model output to various
scenario assumptions. These conclusions are presented in the following
section. Also presented is a summary of the scenario sensitivity analysis
results, in the order in which they were presented in the previous sections.
It should be noted, however, that as the models are generic and not site
specific, the results are as well.
The sensitivity of the results to the choice of disposal site is very
pronounced when determining maximum annual doses to a nearby CPG. With
other things equal, an impermeable site provides more protection than a
permeable one, while arid regions provide more protection than one which
is humid. In determining long-term population health effects, the results
are much .less sensitive to site location with long-term, population health
effects fairly constant over all sites. This is due to the, fact that over
long time periods, the cumulative amount of radionuclides that reach the
population does not vary a great deal.
The sensitivity of the results to the type of disposal method used is
very similar to the results shown for the site location. Placing waste into
a more stringent disposal system which provides greater isolation is similar
to disposing of waste in less permeable or less humid sites. Again, the
effects are more sensitive when assessing the CPG dose than when assessing
long-term, population health effects.
Concerning the waste form, the results are very sensitive to
solidification. Both CPG dose and cumulative population health effects
are reduced.by solidification, due to a reduction in the rate at which
radionuclides leave the disposal trench. A secondary point is that
high-integrity containers affect the cumulative population health effects
11-65
-------
Table 11-23.
Comparison of ocean health effects to
river health effects
Nuclide
H-3
C-14
Mn-54
Fe-55
Ni-59
Co-60
Ni-63
Sr-90
Nb-94
Tc-99
Ru-106
Sb-125
1-129
Cs-134
Cs-135
Cs-137
Ce-144
Eu-154
Ra-226
U-234
U-235
Np-237
U-238
Pu-238
Pu-239
Pu-241
Ara-241
Pu-242
Am-243
Cm-243
Cm-244
Bf
(L/kg)
9.0E-1
4.5E+3
l.OE+2
l.OE+2
l.OE+2
2.0E+1
l.OE+2
1.1E+1
3.0E+4
4.3E+1
l.OE+1
1.0
3.3E+1
1.3#+3
1.3E+3
1.3E+3
2.5E+1
2.5E+1
5.0E+1
l.OE+1
l.OE+1
5.0E+2
l.OE+1
8.0
8.0
8.0
8.1E+1
8.0
8.1E+1
2.5E+1
2.5E+1
Bs
(L/kg)
9.3E-1
3.5E+3
7.2E+1
6.7E+2
2.5E+2
2.0E+2
2.5E+2
1.1E+2
l.OE+2
2.1E+2
3.3E+3
l.OE-1
1.7E+2
8.1E+2
8.1E+2
8.1E+2
2.5E+1
2.5E+1
l.OE+2
l.OE+1
l.OE+1
5.0E+2
l.OE+1
5.3
5.3
5.3
l.OE+3
5.3
103
l.OE+3
1 . OE+3
Ratio of
Ocean to River
Health Effects
1.3E-5
1.2E-5
1.2E-4
2.2E-4
9.8E-3
2.7E-4
1.5E-4
2.7E-4
5.1E-3
8.7E-3
5.2E-3
1.1E-4
8.9E-3
1 . 2E-4
4.1E-3
8.1E-4
1.3E-4
1.3E-4
4.7E-3
1.3E-3
1.3E-3
4.1E-3
1.3E-3
1 . OE-3
1.9E-3
9.8E-4
1.7E-2
1.9E-3
3.3E-3
6. OE-3
6. OE-3
11-66
-------
very little, since the containers are assumed to fail relatively quickly.
In addition, high integrity containers can actually cause an increase in
maximum CPG dose of about 10 percent, due to their failure and release of
radionuclides over a shorter time period, causing a larger peak in CPG dose*
Separating the waste into waste mixes indicative of various regional
compacts has little effect on impacts from disposal. The waste mix used
does not vary greatly over the regional compacts for the critical nuclides.
The waste groups used in the analyses have little effect on the results,
since the major impacts are due to general LLW, which was not varied, and
higher activity wastes generally being solidified, such that their inclusion
or removal does not affect the results significantly.
Increasing or decreasing the site volume (waste volume and activity)
causes a linear increase or decrease in long-term health effects and in CPG
dose at impermeable sites. Permeable sites show a less than linear effect,
due to areal dilution of larger source terms. In general, site size or
waste volume affects health impact in a linear manner.
The sensitivity of the results to the time period used in the analysis
is generally not important, as long as a minimum period of about 1,000 years
is used for the CPG analyses. Increasing the time period in the CPG
analyses at humid sites has no effect, as the CPG peak dose occurs prior
to 1,000 years. Since the peak doses can occur as late as 800 years,
however, decreasing the time period to much less than 1,000 years could
cause the peak dose to be missed. For arid sites and some humid scenarios,
the dose rate continues to rise after 1,000 years, due to the long travel
time required for radionuclides to reach the CPG or due to greater isolation
disposal methods. In these cases, however, the peak annual dose to the CPG
is always small. For the long-term, population health effects, nuclides
which had not reached the regional basin in 10,000 years, such-as nuclides
of uranium or plutonium, would require extremely long time periods to do so
and would not contribute significantly to the total health effects. In
general, it was felt that continuing the analysis period for greater than
10,000 years incorporated too much uncertainty to be useful.
' Our analyses show that the inventories of various radionuclides are
sensitive, depending upon the site location and the predominant pathways.
In general, the mobile nuclides with longer half-lives, such as C-14 and
1-129,- predominate, although in cases,of atmospheric or direct exposure
pathways the critical radionuclides may change.
Additional analyses on the size of the buffer zone show this variable
to be unimportant in affecting CPG dose or population health effects.
Analyses of global health effects show them to be minor in comparison to the
river basin health effects.
11-67
-------
Based on the scenario sensitivity analyses, the following general
conclusions can be made (the conclusions of the single parameter sensitivity
analyses are in section 11.2.3.):
o The assumptions concerning the scenario variables, when analyzing
^ the impacts from LLW disposal, cause a much greater change in the
output when assessing the maximum dose to the CPG than when
assessing the long-term, population health effects. This is due to
the fact that CPG dose is based on only the peak release, whereas
long-term health effects are based on cumulative releases.
o The health impacts from LLW disposal are very sensitive to methods
that provide greater isolation, such as disposal in less permeable
or less humid sites, disposal using more rigorous disposal
technologies, or solidifying the waste prior to disposal.
o In terms of waste form or waste treatment, the most sensitive method
for reducing health impacts is solidification of the waste.
o The health impacts from LLW disposal are dominated by the
longer lived, mobile radionuclides with high dose conversion
factors, such as C-14 and 1-129, although in cases of atmospheric or
direct exposure pathways the critical radionuclides may change.
o In analyzing the health impacts from the disposal of LLW, an
analysis period of 1,000 years is usually sufficient to assess any
significant peak dose. For long-term, population health effects,
10,000 years is long enough to assess the health impacts from all
but the most immobile radionuclides, which would require an
extremely long analysis period to assess and are minor contributors
to total health effects.
o In determining population health effects, the contribution
from nuclides released to the ocean and taken up through
the consumption of ocean fish or seafood will be minor.
11-68
-------
REFERENCES
Be81 Begovich C.L. , Eckerman K.F., Schlatter E.G., Ohr S.Y., and R.O.
Chester, DARTAB: A Program to Combine Airborne Radionuclide
Environmental Exposure Data with Dosimetric and Health Effects Data
to Generate Tabulations of Predicted Impacts, ORNL/5692, Oak Ridge
National Laboratory, Oak Ridge, Tenn., August 1981.
EPA82 J.M. Smith et al., "Environmental Pathway Models for Estimating
Population Health Effects from Disposal of High-Level Radioactive
Waste in Geologic Repositories," EPA 520/5-80-002, U.S.
Environmental Protection Agency, Montgomery, Alabama, December 1982.
EPA83 U.S. Environmental Protection Agency, PRESTO EPA: A Low-Level
Radioactive Waste Environmental Transport and Risk Assessment Code -
Methodology and User's Manual, Prepared under Contract No.
W-7405-eng-26, Interagency Agreement No. EPA-D-89-F-000-60, U.S.
Environmental Protection Agency, Washington, D.C., April 1983.
EPA87 U.S. Environmental Protection Agency, Low-Level and NARM Radioactive
Wastes, Model Documentation PRESTO-EPA-POP, Volume 1, Methodology
Manual, EPA 520/1-87-024-1, U.S. Environmental Protection Agency,
Washington, D.C., December, 1987.
EPA88 U.S. Environmental Protection Agency, Sensitivity Analysis Results
of PRESTO-EPA Low-Level Waste Risk Assessment Models, U.S.
Environmental Protection Agency, Washington, D.C., 1988.
Mo79 Moore R.E., Baes C.F. Ill, McDowell-Boyer L.M., Watson A.P., Hoffman
F.O., Pleasant J.C., and C.W. Miller, AIRDOS-EPA: A Computerized
Methodology for Estimating Environmental Concentrations and Dose to
Man from Airborne Releases of Radionuclides, EPA 520/1-79-009
(Reprint of ORNL-5532), U.S. Environmental Protection Agency, Office
of Radiation Programs, Washington.,.-B-.C., December 1979.
NRC77 U.S. Nuclear Regulatory Commission, "Calculation of Annual Doses to
Man from Routine releases of Reactor Effluents for the Purpose of
Evaluating Compliance with 10 CFR Part 50, Appendix I," Regulatory
Guide 1.109, Washington, D.C., October 1977.
Ro85 Rogers V.C., Hung C.Y., Cuny P.A., and F. Parraga, An Update on
the Status of EPA1s PRESTO Methodology for Estimating Risks from
Disposal of LLW and BRC Wastes, U.S. Department of Energy,
' Proceedings of 6th Annual Participants' Information Meeting on DOE
Low-Level Waste Management Program, Denver, Colorado, September
11-13, 1984, CONF-8409115, Idaho Falls, Idaho, 1984.
Ro87 V.C. Rogers, "Health Effects Conversion Factors for Fish
Consumption," TIM-47/2-3R2, Rogers and Associates Engineering
Corporation, Salt Lake City, Utah, May 2, 1984.
11-69
-------
ShS6 Shuraan R. , and V.C. Rogers, A Comparison of PATHRAE and PRESTO-CPG
Simulation Results, Rogers and Associates Engineering Corporation -
Technical Information Memorandum, TIM-8469/11-12, Rogers and
Associates Engineering Corporation, Salt Lake City, Utah, January
31, 1986.
Sh87a Shuman, R., and V.C. Rogers, PATHRAE-EPA Model Sensitivity Studies,
RAE8706/4-2, Rogers and Associates Engineering Corporation, Salt
Lake City, Utah, August 21, 1987.
Sh87b Shuman, R., and V.C. Rogers, PRESTO-EPA-CPG Sensitivity Studies,
Rogers and Associates Engineering Corporation-Technical Information
Memorandum, TIM-8706/4-3, Rogers and Associates Engineering
Corporation, Salt Lake City, Utah, November 13, 1987.
11-70
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Chapter 12:
UNCERTAINTY OF CUMULATIVE POPULATION HEALTH
EFFECTS AND MAXIMUM CPG DOSE ANALYSES
12.1 Introduction
Deterministic models were used for the assessments of maximum CPG
dose and cumulative population health effects conducted in support of the
proposed LLW standards. This chapter presents the results of uncertainty
analyses, specifically the upper bound of the maximum CPG dose and
cumulative population health effects assessments described in Chapter 9,
and discusses the methodology used in determining the uncertainty.
Because of the complexity of the assessment, uncertainty analysis is
divided into several components. Every effort is made to quantify the
upper bound of the uncertainty associated with each component. However,
the quantification of uncertainty is limited to certain components
because the analyses include extensive field and laboratory data as well
as actual uncertainties.
Several uncertainties are not included in this analysis because of
time and budget constraints. These include the uncertainties inherited
from the approximation of the basic equations used in the PRESTO-EPA
model and those resulting from variations of site location input
parameters.
12.1.1 Uncertainty Analysis
Ideally, the dose and health effects assessments should be performed
using a probabilistic model that expresses the results in the form of
random variables (value versus its probability density function). The
results of the analysis could then fully inform readers of the absolute
value of the dose or health effects and their correspondent, probability
of occurrence.
Since this type of analysis would require large numbers of
calculations, from a practical standpoint this probabilistic model should
not be applied to complex assessment models such as the PRE!STO-EPA models.
An alternative method is to use a deterministic model with sets of
deterministic input parameters to calculate the most probable results of
the risk assessment, and to use a separate model to estimate the probable
variation of the assessed values (known as "standard deviation"). The
estimated probable variation of the assessed values is commonly known as
the uncertainty of the results; this analysis therefore intends to
evaluate the uncertainty of the results of risk assessments used to
support the LLW standards.
12.1.2 Complexity of the Uncertainty Analysis
Maximum CPG dose and cumulative population health effects
assessments are extremely complex analyses involving the analysis of many
12-1
-------
raultidisciplinary areas of expertise, including the processes of
atmospheric, hydrogeological, and biological transports and the
bioeffects resulting from radionuclide exposure.
Since the dose and health effects analyses supporting the LLW
standards are generic, i.e., nonsite specific, the discussion of the
uncertainty of the results obtained from the assessments is far more
complex than if these same assessments had been performed for a specific
site. This is because a site-specific analysis would deal with
relatively well-defined site parameter values, while a nonsite-specific
analysis would be concerned with the site parameter values having both
variation of parameter values for a specific site .and variations of
several disposal sites within a region. Therefore, the uncertainty
caused by the site parameter variations was not included in this
analysis, because the results of the risk assessment used for supporting
the LLW standards did not intend to cover the entire spectrum of possible
variations of disposal sites. When the uncertainty due to the variation
of sites is considered, the results of the uncertainty analysis are
expected to be dominated overwhelmingly by this uncertainty.
12.1.3 components of Overall Uncertainties
Because of the complexity of the processes in the risk assessment,
the discussion of the overall uncertainties may be simplified by dividing
the uncertainties into several components.
For the purpose of this discussion, the overall uncertainties are
divided into five components: (1) source term radionuclide
concentration, (2) radionuclide geosphere transport, (3) radionuclide
food chain transport, (4) human organ dosimetry, and (5) health effects
conversion factors. The interactions of the parameters between //
components are presented in Figures 12-1 and 12-2 for the cumulative '
health effects analysis and maximum CPG dose analysis, respectively.
12.1.4 Significance of Sensitivity Analysis
A sensitivity analysis, as discussed in Chapter 11, quantifies the
sensitivity of model outputs to a change in a specific parameter or group
of input parameters under a set of presumed input parameter values, if
the input parameter values are altered, a change in the sensitivity of
that particular parameter may subsequently follow. Furthermore, the
sensitivity analysis results do not provide information on the
probability of occurrence associated with the particular input parameter
value and the value of the analyzed results. Therefore, if a parameter
is found to be very sensitive within a certain range and insensitive in
another range it does not necessarily mean that the parameter is
important, since the importance of that parameter should be determined by
both sensitivity and its associated probability of occurrence. Using an
extreme case as an example, if the probability of occurrence for the
above case is found to be zero within the range value of the parameter
12-2
-------
RADIONUCLIDE SOURCE TERMS
• Concentration
RADIONUCLIDE TRANSPORT IN GEOSPHERE
• Release from Trench
• Ground-Water Transport
• Cumulative Discharge to River Basin
RADIONUCLIDE TRANSPORT
• Drinking Water
• Crop Irrigation
• Cattle Feed
IN FOOD CHAIN
HUMAN ORGAN DOSIMETRY
• Organ Dose Conversion Factors
CUMULATIVE HEALTH EFFECTS
• Health Effects Conversion Factors
Figure 12-1. Major Components of Uncertainty Analysis:
Cumulative Population Health Effects Analysis
12-3
-------
RADIONUCLIDE SOURCE TERMS
• Concentration
RADIONUCLIDE TRANSPORT
• Release from Trench
• Ground-Water Transport
• Maximum Concentration
IN GEOSPHERE:
RADIONUCLIDE TRANSPORT IN FOOD CHAIN
• Drinking Water
• Irrigation
• Cattle Feed
HUMAN ORGAN DOSIMETRY
° Organ Dose Conversion Factors
MAXIMUM CPG DOSE
• Body Dose Conversion Factors
Figure 12-2. Major Components of Uncertainty Analysis?
Maximum CPG Dose Analysis
12-4
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which is found to be sensitive, then the parameter should be judged as
unimportant. Even if a sensitivity analysis provides some information
concerning the uncertainty of the overall analysis, it can never be used
as a substitute for the uncertainty analysis.
On the other hand, the sensitivity analysis has its. own merits; it
can be performed with much less effort and cost when the deterministic
model is already established. Because the quantification of the overall
uncertainty of the results obtained from the PRESTO-EPA model is not
possible at this stage, the results of the sensitivity analysis, as
discussed in Chapter 11, play an important role by providing some
information regarding the uncertainty of the assessment result.
12.1.5 Approach to the Uncertainty analysis
As discussed in Chapters 8 and 11, PRESTO-EPA is a dynamic model
simulating the radionuclide transport in the geosphere and food chain
pathways, with the use of the organ doses and health effects conversion
factors to calculate organ doses and health effects to be incurred from
the disposal of LL¥. The organ dose and the health effects conversion
factors are the input data for PRESTO-EPA models and are obtained from
RADRISK, another human organ dosimetry and health effects model.
All of these components of calculation are conducted in series, and
the parameters transferred from one component of calculation to another
are clearly defined and evaluated in PRESTO-EPA. The uncertainty of each
component of calculation can be evaluated separately as a part of the
overall uncertainty of the analysis. Thus, the discussion of the
uncertainty is divided into the.five components displayed in Figures 12-1
and 12-2.
After determining the uncertainty for each component, the
uncertainty of the overall dose and health effects assessments can be
estimated. Although a quantitative analysis of the uncertainty for each
of the components is desirable, such an analysis is limited to the
radionuclide transport in the geosphere only, because analyses for the
other components are limited by the time and human resources available.
in order to quantify the uncertainty of the analysis for the
radionuclide transport in the geosphere, a simplified model from
PRESTO-EPA is developed. The model simplifications include:
® Using a quasi-steady-state approximation;
@ Converting a numerical model to an analytical model; and
® Considering the humid permeable hydrogeological settings only.
in a quasi^steady-state simplification, the model assumes that the
trench cap failure will reach its maximum at the time when the analysis
is started, and that the same level of, cap failure is maintained for the
entire period of analysis. This approximation is essential for
12-5
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converting the numerical approach to an analytical approach, which
greatly reduces the complexity of simulation. This approximation will
not introduce serious errors into the results of the risk assessment, and
is adequate for the purposes of an uncertainty analysis. The analysis is
conducted for a humid permeable site only, because the maximum CPG dose
from disposal activities was found to be greatest at this site among the
three sites investigated. In any case, it is speculated that the same
degree of uncertainty can also be expected for the other two sites.-
12.2 Uncertainty Due to Radionuclide Source Term
In order to perform a radiological risk assessment supporting
a generally applicable environmental radiation protection standard
for the land disposal of LLW, it was necessary to develop a radio-
logical source term representative of LLW to be disposed of in the
foreseeable future. The radiological source term consists of best
estimates of the radionuclide concentrations and projected volumes of the
various categories of LLW. EPA has relied on the best information
available to construct the source term used in its radiological risk
assessment of LLW disposal and this source term is presented in Chapter 3.
Considering the enormous amount of detail associated with the data
base used to construct the EPA source term, it is not possible to perform
a rigorous mathematical evaluation of the uncertainty in the EPA LLW
source term within the constraints of available resources. However, one
can evaluate such an uncertainty in a more qualitative manner by taking
into consideration the limitations associated with the data base
supporting the LLW source term, the results of the risk analysis
(Chapter 9), and the results of the sensitivity analysis (Chapter 11).
12.2.1 Origin of the EPA Source Term for LLW
Extremely diverse radioactive wastes fall under the definition of
LLW. EPA has derived its radiological source term for LLW from the vast
LLW data base developed by the NRC in conjunction with the development
of NRC technical requirements for near-surface LLW disposal facilities,
10 CFR Part 61 (NRC81, NRC82a,b, NRC86). The NRC data base defines
numerous waste categories, or "waste streams," each of which consists of
a consolidation of groups of wastes having common sources and similar
physical, chemical, and radiological characteristics. Earlier NRC
assessments (NRC81, NRC82a) for its draft and final EIS examined 36
and 37 waste streams, respectively. Table 3-2 compares the waste
streams defined by NRC for its 10 CFR Part 61 rulemaking and the
condensed listing of these wastes used in the EPA analysis.
The more compact EPA set of waste streams was achieved either by
combining waste volumes from large and small facilities into one waste
stream or by combining similar wastes from similar generators into one
waste stream. This latter simplification was achieved by weighting
12-6
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radionuclide concentrations from each contributing waste stream by its
proportional contribution of volume to the overall waste stream volume.
For example, the EPA waste stream LWR Ion Exchange Resins volume-weighted
the radionuclide concentrations of two NRC waste streams, PWR Ion
Exchange Resins and BWR Ion Exchange Resins.
NRC's "Update of Part 61 Impacts Analysis Methodology" report
(NRC86) further revised, updated, and supplemented NRC's LLW source term
characterization. This "updated" report provided more detail on the
higher specific activity waste streams and nonroutine or unusual
low-level waste streams. Updated information on routine LLW streams was
used to provide revised radionuclide concentrations for those affected
EPA waste streams as shown in Chapter 3.
Although they are discussed in Chapter 3, nonroutine or unusual
low-level waste streams were not included in the EPA source terra because
there is much uncertainty over the timing and characteristics of such
wastes. Moreover, such wastes do not comprise a significant fraction of
projected LLW volumes or activities. In summary, the EPA radiological
source term for LLW is based upon the most complete and recent
information available for routine sources of commercial LLW. DOE has
indicated that its LLW are similar to commercial LLW, but a breakdown
comparable to that provided by the NRC for commercial wastes is not
available (DOE86).
EPA has made one minor addition to the basic LLW source term for
commercial wastes. The EPA source term includes two Naturally Occurring
and Accelerator-produced Radioactive Materials (NARM) waste streams:
radium sources and radium-contaminated water treatment ion exchange
resins. These two waste streams comprise less than one percent of the
commercial LLW volume and an even smaller percentage of the total
activity. They have been included to reflect EPA's intention to regulate
high specific activity, low volume NARM wastes under the TSCA.
12.2.2 Uncertainties Associated with Data Bases
Since EPA has relied heavily upon the NRC characterization of LLW
(NRC81, NRC82, NRC86), uncertainties introduced into the NRC source term
for. LLW would also apply to the EPA source term. The major sources of
information supporting the development of the NRC and EPA
characterizations of LLW include:
• Computer-assisted calculations;
• Surveys of waste generators;
• Disposal site records; and
• Radiochemical analysis.
The following discussion will focus on the limitations of each of these
information sources as they relate to the characterization of LLW.
12-7
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Computer-assisted calculations are typically used to estimate
the radionuclide composition and quantities generated by "burn-up"
of nuclear fuels. Such models are based on numerous parametric values
for a given power reactor design. As such, they are reasonably
well-suited to identifying important radionuclides that are produced in
the nuclear fuel and activated in the surrounding structural materials.
Since virtually every reactor design is different, computer-assisted
calculations performed for one or several plant designs would introduce
some error when attempting to project LLW characteristics over all
reactors. Use of one or more of the remaining information sources in
conjunction with such computer calculations could reduce the uncertainty
associated with any LLW predictions.
Past surveys of waste generators or disposal facility site records
may have certain limitations. In practice, radionuclide distributions
listed on such records frequently were calculated by applying
predetermined radionuclide distributions to the total gross
radioactivities obtained during screening measurements made at the time
of shipment. Such measurements were probably conservative in terras of
the total radioactivity measured since less sophisticated measurement
techniques have been applied in the past, and because the radioactivity
contribution of short-lived isotopes was included in the total activity
reading. When predetermined radionuclide distributions are used, changes
in actual radionuclide concentrations on a day-to-day basis may have been
missed as well.
The sensitivities (minimum detection limits) of the analytical
procedures for the various radionuclides are not identical, especially
with respect to "hard-to-measure" radionuclides (e.g., C-14, 1-129).
Such radionuclides are more likely to be "scaled" from previous
specialized measurements, using a "scaling factor" associated with a more
abundant, easy-to-measure nuclide. This scaling factor would then be
applied on a routine basis as a calculational tool, with the possibility
that day-to-day variations in the actual radionuclide concentration would
be overlooked.
In order to minimize the uncertainties associated with these
information sources, NRC has updated its LLW source term for those LLW
streams contributing large volumes,, such as power reactor LLW, or those
waste streams possessing relatively high specific activities, such as
certain industrial LLW (NRC86). More recent LLW shipment records, more
detailed surveys of certain waste generators, and, in general, more
up-to-date waste volume generation rates all contributed to the revised
NRC characterization of commercial LLW. These improvements were
subsequently incorporated into the EPA LLW source term.
12.2.3 Estimated Uncertainty
As the discussion above indicates, numerous information sources have
contributed to the development of the EPA characterization of LLW. For
12-8
-------
each of the four categories of information sources cited above, numerous
individual data bases have contributed to the characterization of LLW.
To characterize the uncertainty associated with each individual source of
data would be a monumental task well beyond the scope of this effort.
A more manageable approach is to develop a qualitative
characterization of the uncertainty associated with the EPA LLW source
term. The results of the health impacts assessment, discussed in
Chapter 9, indicated that the two most important radionuclides,
considering both CPG dose and population health effects, are C-14 and
1-129. Other radionuclides contribute virtually nothing to the CPG dose
and on the order of 20 percent or less to population health effects.
Therefore, the remaining discussion of uncertainty in the EPA LLW source
term will focus on C-14 and 1-129.
Characterization of source terms for C-14 in LLW has received
particular attention over the last few years. Such attention is
well-deserved considering that C-14 is found in many categories of LLW
and has a long half-life (5,700 years), high mobility via water pathways,
and a relatively high dose conversion factor. The NRG has continually
updated its characterization of C-14 in LLW (NRC81, NRC82, NRC86). The
most recent update greatly improved the characterization of C-14
occurring in many higher activity waste streams. A recent EPA study of
C-14 in LLW reviewed the various source terms (Gr86).
Each of the three major categories of LLW streams (nuclear fuel
cycle, institutional, and industrial) was reviewed by comparing the EPA
source term to independent estimates of LLW containing C-14,. In all
three cases, the EPA source term was found to be a "reasonable
representation" of the C-14 in all three categories of LLW (Gr86). A
recent NRC document reports on the results of the analyses of hundreds of
process and waste samples from power reactors in an effort to establish
useful correlation factors between "easy" and "difficult" to measure
radionuclides (C-185). Analysis of the data suggested that an empirical
scaling factor for C-14 with Co-60 would be most useful. The scaling
factors derived separately for BWRs and PWRs showed statistical
uncertainties in the range of 27 to 45 percent. This implies that for
any given reactor design, the concentration of C-14 in a given waste may
vary up to plus or minus 50 percent (rounded). For institutional wastes
containing C-14, average concentrations of C-14 in the EPA source term
were compared with reported average concentrations. The average C-14
concentration in institutional wastes was reported by EPA as
5.1 E-3 Ci/ra3 for 1982. At the same time, the CRCPD reported an
average concentration of 1.3 E-2 Ci/m3 for 1982. For the year 1983,
the National Institutes of Health cited an average concentration of
2.4 E-3 Ci/m3 for C-14 in its LLW (Gr86).
For institutional wastes, these estimates represent a variation of
about a factor of 2 from the EPA source term. More limited data are
available for industrial sources of C-14, although manufacturers of
12-9
-------
chemical compounds labeled with C-14 (and H-3), the major source of C-14,
appear to be reasonably well characterized by detailed generator surveys
(NRC86, Ke85). Considering the above evaluations, a qualitative
uncertainty of a factor of 2 up or down is assigned to the EPA source
term characterization of the c-14 concentrations in LLW.
Like C-14, 1-129 possesses a long half-life (17 million years), high
mobility via water pathways, and a relatively high dose conversion
factor. 1-129 occurs in fewer LLW streams, however. Other than power
reactor waste streams, 1-129 is found only in wastes from industrial
radioisotope manufacturers. The contribution of 1-129 from industrial
radioisotope manufacturing waste is very small because of the extremely
low 1-129 concentrations and relatively small volumes of such wastes (see
Chapter 3). Therefore, this discussion will concentrate on 1-129 in
power reactor LLW. The NEC and EPA characterization of 1-129 in LLW
relies upon previous work that attempted to derive a "scaling factor" for
1-129, as related to the measured concentration of Cs-137 in the same
sample. Since there were so few samples in which both were measured,
statistical averaging was not possible (NRC81). A more recent
investigation also attempted to develop scaling factors for 1-129 in LLW
(C185). In this study 1-129 was compared with Cs-137, since both are
fission products, have similar transport properties in reactor systems,
and release mechanisms from reactor fuel. In this case, however, only a
small percentage of BWR samples (16 out of 191) and PWR samples (22 out
of 259) contained both 1-129 and Cs-137. Most of the time, 1-129 was at
or near its detectable limit. Thus, the scaling factors used to estimate
the 1-129 activities in the waste contain large uncertainties, typically
ranging from 50 to 90 percent. These results suggest that 1-129
concentrations in LLW are highly variable. On the other hand, since
1-129 was detected so seldom, it is likely that 1-129 does not occur very
often in LLW at levels comparable with detectable limits. Based on the
measured data, therefore, a factor of 2 up or down may be reasonably
assigned to the uncertainty of the 1-129 concentration in LLW. However,
it is felt that the EPA characterization of 1-129 in LLW is probably very
conservative (i.e., too high).
In summary, the results of EPA's risk analyses (see Chapter 9) have
identified two radionuclides, C-14 and 1-129, as predominant in producing
exposures to the critical population group (CPG) and causing total
population health effects. Information concerning the occurrence of
these two radionuclides in LLW indicates that the concentration of each,
as used in the EPA source term, has an uncertainty of approximately a
factor of 2 up or down; that is, the likely concentration of each
radioriuclide may be as large as twice the EPA concentration or as small
as one-half the EPA concentration. Data with respect to 1-129 indicate
that the EPA source term is probably conservative, however.
12.3 Uncertainty Due to Radionuclide Geosphere Transport for Cumulative
Health Effects Analysis
Since the output requirements for the two analyses, assessment of
maximum CPG dose and assessment of the cumulative population health
12-10
-------
effects, are different, the uncertainties of these analyses are treated
separately. This section discusses the uncertainty for the; cumulative
population health effects analysis due to geosphere transport. The key
difference in the twp uncertainty analyses for the radionuclide transport
in the geosphere is that the cumulative health effects analysis requires
an analysis- of the cumulative radionuclides being discharged into the
regional river basin, while the maximum CPG dose analysis requires the
analysis of the maximum annual average concentration of radionuclides at
the nearest accessible environment.
12.3.1 Method of Analysis
In order to evaluate the cumulative radionuclide release, a
leaching-release model simplified from the PRESTO-EPA model is employed.
After imposing the simplifications discussed in Section 12.1.5, the
unsteady-state leaching-release model used in PRESTO-EPA wan converted
into a steady-state solute transport system model for which an analytical
solution is obtainable. The rate of the radionuclides being discharged
into the regional river basin is established using Hung's ground-water
transport model, which is the same transport model used in PRESTO-EPA
(Hu86).
For the humid permeable site, the total health effects incurred from
LLW disposal are dominated by the effects of the residual radionuclides
being discharged into a regional river basin for 10,000 years of
analysis. The health effects incurred from the local community (made up
of a few farmhouses) for 10,000 years of analysis are therefore combined
into the downstream river basin effects. Based on this simplification,
the cumulative radionuclides being discharged into the regional river
basin can be obtained by integrating the discharged activity over the
time frame of 10,000 years. The result is expressed by:
QT = " ^R X. A/e VR H^W
- Exp^-U, + E/eVR) • 10,000 +
(12-1)
where:
QT
the cumulative radionuclide being discharged into the regional
river basin;'
Hung's ground-water transport correction factor;
the initial radionuclide inventory;
the equivalent rate of infiltration through the trench cap (a
constant);
12-11
-------
e •» the porosity of the waste material;
V ~ the volume of waste;
R = the radionuclide retardation factor;
X. = the radionuclide decay constant; and
t^ = the sum of radionuclide transit time through the host soil
(between waste trench and aquifer) and aquifer (between disposal
, site and the ground-water discharging point).
This equation brings together the major parameters that will
significantly affect the cumulative radionuclide release and,
subsequently, the total health effects. Equation 12-1 also shows that
the cumulative radionuclide release can be expressed as a simple
mathematical function of major parameters, and therefore the uncertainty
of the OTftulative radionuclide release can be calculated by the
analytical method proposed by Hung (Hu87), instead of using a time-
consuraing Monte Carlo or other simulation method. Rung's method
calculates the joint probability density distribution for two successive
random variables based on Equation 12-1. The overall joint probability
density distribution is the uncertainty of the cumulative amount of
radionuclide being discharged into the regional river basin.
12.3.2 Postulated Probability Density Distribution of Parameters
Because limited data are available for analyzing the probability
density distribution of the major parameters,appearing in Equation 12-1,
values for the arbitrary distribution of probability density distribution
are estimated for this analysis on the basis of engineering judgment.
The input parameters selected for the analysis are the radionuclide
distribution coefficients for the trench material, the host soil, and the
aquifer; the degree of trench cap failure; the distance from trench
bottom to aquifer; the distance from the disposal site to the regional
river basin; the ground-water velocity in the aquifer; and the
percolation velocity in the host soil. The assumed probability density
distribution for each parameter is normalized and presented in
Figures 12-3 through 12-7.
Probability density distributions have not been assigned to some of
the Input parameters because (1) the probable standard deviation is so
small that it can be considered a deterministic variable, or (2) the
probability density distribution is programmed to be calculated from
other random Input parameters.
12,3.3 Results of Uncertainty Analysis
A computer program was developed based on the methodology described
in Section 12.3.1 for the uncertainty analysis of the cumulative
radionuclide being discharged into the regional river basin. The
12-12
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12-15
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12-16
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computer program was implemented on a personal computer AT or a
compatible. By using the estimated probability density distribution for
the input parameters described in Section 12.3.2, the uncertainty o£ the
cumulative radionuclide being discharged to the regional river basin is
analyzed.
The normalized results of the uncertainty analysis are presented in
Figures 12-3 through 12-7. respectively, for H-3, C-14, Tc-99, 1-129, and
Np-237 (£or a humid permeable site only). The analysis is conducted
solely for these radionuclides because the other relatively immobile
radionuclides are retained either in the waste trench or in the aquifer
at the end of analysis and did not contribute to the total health effects.
The results indicate that the uncertainty or the standard deviation
of the cumulative activities being discharged into the regional river
basin for each radionuclide is practically zero except for H-3. This is
due to the fact that when the radionuclide transit time through the
geosphere is prolonged because of a higher distribution coefficient,
there would be additional loss from radioactive decay, and vice versa.
Since the half-lives of these radionuclides, other than H-3, are
relatively long, the change in the radioactive decay loss due to the
change in the transit time is negligibly small and thus the uncertainty
of the results is small also. On the other hand, H-3 has a relatively
short half-life, so that there is a significant effect on the cumulative
activity of H-3, which can be transported to the regional river basin due
to the change in radionuclide transit time.
The above results are expected to remain unchanged even if there are
slight changes in the probability density distributions of input
parameters.
12.3.4 Summary
Since C-14 is the critical radionuclide contributing a major portion
of the health effects for all three generic sites analyzed (see
Chapters 9 and 10), and since the uncertainty of the analysis for the
cumulative activity of c-14 being discharged into the regional river
basin is near zero, one may logically conclude that the uncertainty for
the cumulative health effects assessment due to geosphere transport is
near zero.
The results of analyses for Tc-99, 1-129, and Np-237 indicated that
they have characteristics that are similar to those of C-14 (Figures 12-5
through 12-7); that is, their uncertainties may also be considered to be
zero. Figure 12-3 showed that the uncertainty for H-3 is analyzed to be
approximately 65 percent.
The uncertainty of the cumulative population health effects analysis
for the humid impermeable site was not analyzed, because the total
population health effects assessments for this region were much smaller
12-18
-------
than that for a humid permeable region and thus will not play as
important a role as that for the humid permeable region. Furthermore,
the radionuclide release pathway—trench overflow pathway—analyzed for a
humid impermeable region could possibly be avoided in future designs by
using an improved engineering disposal method.
12.4 Uncertainty Due to Geosphere Transport for Maximum Dose analysis
As discussed in Section 12.3, the maximum annual average
concentration of radionuclides being released to the nearest accessible
environment is the output parameter generated from the geosphere
transport analysis for the maximum CPG dose analysis. This information
is transmitted to the food chain calculation (see Figures 12-1
and 12-2). The risk assessment conducted in support of the development
of EPA's LLW standards assumed that a farmhouse well is located right on
the fence line and that the well continues to operate at the same
location even after institutional control is lifted. The fence line is
assumed to be 100 m from the edge of the trench area.
12.4.1 Method of Analysis
The basic equation used to calculate the maximum annual average
concentration of radionuclides in the accessible environment combines the
simplified leaching-release model and the ground-water transport model
used in PRESTO-EPA-CPG. The same simplifications presented in
Section 12.3 are also used for the leaching-release model.
Because the critical radionuclides that contributed over 90 percent
of the maximum CPG dose were long half-life radionuclides (3C-129
or C-14), the maximum concentration of a radionuclide is found to occur
at the time when the contribution of the radionuclide from the far end of
the trench area reaches the well. The time required to reach its maximum
concentration is known as the time of arrival. Knowing the time of
arrival, the maximum concentration of a specific radionuclide is obtained
by integrating the contributions from each subdivided segment over the
entire trench area at the time of arrival. The analytical solution for
this integration is (HU87):
max
A Waq'Ht
and
V
V,
(12-2)
12-19
-------
aq
where:
C s the maximum concentration of radionuclide at the well;
max
A - the total surface area of the disposal trenches;
T - the thickness of the aquifer;
the porosity of the aquifer;
R - the retardation factor for the specific radionuclide;
d = the distance between the trench bottom and the top of the aquifer;
V - the interstitial ground-water velocity;
L s the length of the disposal site without including a buffer zone
and measured in the ground-water flow direction;
X = the distance from the near edge of the disposal site (excluding.
the buffer zone) to the well;
5 ~ the leaching rate correction factor;
E s the equivalent rate of infiltration;
V - the volume of waste material (including the backfill material); and
subscripts, h, v, and w designate aquifer, host soil, and waste material,
respectively.
Equation 12-2 demonstrates that the parameters appearing in the
equation will contribute significantly to the maximum annual average
radionuclide concentration and are in a simple mathematical relationship
with the model output. It should be noted that those parameters which do
not have any significant effect on the radionuclide concentration are
discarded from Equation 12-2. Equation 12-2 implies that the uncertainty
of the maximum annual average concentration of a radionuclide at the well
can be calculated by the analytical method proposed by Hung (Hu87).
12.4.2 Estimated Probability Density Distribution of Input Paramaters
For the same reasons as were stated in Section 12.3.2, for this
analysis the distributions of probability density for each random input
parameter are estimated through engineering judgment. The predominant
random input parameters selected for the analysis are: (1) the
distribution coefficients for trench material, host soil, and the
aquifer; (2) the degree of trench cap failure; (3) the distance from the
trench bottom to the top of the aquifer; (4) the length of the disposal
site in the direction of ground-water flow; (5) the ground-water velocity
in the host soil; and (6) the ground-water velocity in the aquifer.
12-20
-------
Their distributions are shown in Figures 12-8 and 12-9. The remainder of
the input parameters appearing in Equation 12-2 are considered to be
deterministic numbers, and thus the same numbers as are used in
PRESTO-EPA-CPG for the humid permeable site are assigned.
12.4.3 Results of Uncertainty Analysis
A separate computer program was also developed for the uncertainty
analysis of the maximum annual average radionuclide concentiration based
on Equation 12-2. Using the probability density distribution of the
input parameters discussed previously, the uncertainty of the maximum
concentration in the well is analyzed. The results of the uncertainty
analysis for radionuclides C-14 and 1-129 are presented in Figures 12-8
and 12-9 in a form of a normalized probability density distribution.
Analyses are conducted only for C-14 and 1-129 because they acted as the
predominant radionuclides, i.e., those that contribute the major portion
of the maximum CPG dose, for all scenarios analyzed for the humid
permeable site. One should notice that the contributions of C-14 and
1-129 to the maximum body dose would occur in different time frames
because of the difference in retardation factors.
The results indicated that the uncertainty or the standard deviation
of the maximum annual average concentration for C-14 and 1-129 in the
well (occurring at different time frames) are approximately 8 percent and
26 percent of the mean values, respectively. Compared to the standard
deviation for the input parameters, which have values ranging from 12 to
25 percent, the output parameters are considered to be converging from
the uncertainties of input parameters for C-14 and diverging very slowly
from the uncertainties of input parameters for 1-129.
The analysis also indicated that the uncertainty of the maximum
radionuclide concentration does not increase in proportion to the
increase in the uncertainty of the input parameters. This is due to the
fact that when the distribution coefficient increases, the initial
release rate will decrease and the time of arrival of the peak
concentration at the well will increase. These two effects, known as
primary effects, tend to minimize the peak concentration at the well. In
addition, when the rate of the initial leaching rate decreases, the rate
of the radionuclide inventory depletion rate will decrease as well. This
secondary effect will tend to increase the subsequent radionuclide-
leaching rate and to slow down the decrease of the leaching rate. This
secondary effect is also known as buffer action.
in the same manner, when the distribution coefficient decreases, the
initial radionuclide release rate will increase and the time of peak
concentration arrival will decrease. These two primary effects will tend
to maximize the peak concentration at the well. Conversely, when the
rate of the initial leaching rate increases, the rate of the radionuclide
depletion rate also increases, which will tend to decrease the subsequent
leaching rate. This secondary effect will also tend to slow down the
increase of the leaching rate.
12-21
-------
{/>
§
Q
I
11.0
10.0
0.0
8.0
7.0
6.0
5.0
4.0
3.0
2.0
1.0
0.0
vjr&zr*,
I
p
f^^r9t
_ Calculated maximum 1
4(^ body dose for C-1 4 |
Dl
" 1.
r
V
•trlbutlon coeffic
trench matorlal
host soil
aquif*r
ients for:
•
1 . Degree of trench cap failure
2. Distance from trench bottom to aquifer
3. Length of disposal site
4. Ground-water velocity in host soil
5. Ground-water velocity in aquifer
r—~Mr~t
0.0
0.5
1.0
1.5
2.0
2.5
3.0
3.5
NORMALIZED DOSE, x/u.
Figure 12-8. Results of Uncertainty Analysis for C-148
Standard Deviation = 8%
12-22
-------
4.0
3.0
-------
For the mobile radionuclides, the secondary effect is less sensitive
than the primary effects. The sensitivity of the secondary effect
relative to primary effects decreases with the increase in the
radionuclide distribution coefficients. This phenomena can be seen from
the results of the uncertainty analysis. The uncertainty of the maximum
concentration analysis for C-14 (having the most probable distribution
coefficient of 0.01 ral/g) is less than that for 1-129 (having the most
probable distribution coefficient of 3.0 ml/g). To demonstrate this
tendency, an uncertainty analysis for Ra-226 was also conducted, and the
results are presented in Figure 12-10. The results indicate that the
standard deviation for the analysis is 65 percent, which is larger than
the uncertainty for 1-129 and far greater than the uncertainty for C-14.
This is because Ra-226 is a relatively long half-life radionuclide
(similar to C-14 and 1-129) and has a medium capacity of desorption
(distribution coefficient = 220 ml/g), which is much greater than the
distribution coefficients for C-14 and 1-129.
The half-life of the radionuclide will also significantly influence
the uncertainty of the maximum concentration analysis because of the
secondary effect from the retardation of the radionuclide in the
aquifer. To demonstrate this tendency, an uncertainty analysis was
conducted for H-3 to represent radionuclides with high mobility and a
relatively short half-life. The results are presented in Figure 12-11.
The results show that despite the mobile nature of H-3, an uncertainty of
47 percent could be expected, which is much greater than the uncertainty
for C-14 having similar mobility. The difference is primarily due to the
difference in half-lives.
12.4.4 Summary
As indicated in Section 12.4.3,, the predominant radionuclides that
contribute the major portion of the maximum CPG dose are C-14 and 1-129;
the uncertainties of the results of the maximum CPG dose analyses for
C-14 and 1-129 are 8 percent and 26 percent, respectively. By
considering the additional uncertainty in the estimation of the
probability density distribution of each input.parameter, one may claim
that the upper bound of the uncertainties of the results of maximum CPG
dose analyses due to the geosphere transport is on the order of
10 percent of its mean value'for C-14 and 40 percent of its mean value
for 1-129.
12,5 Uncertainty Due to Transport in the Food Chain Pathway
Given a model for predicting the concentration of radionuclides in
foods, the effect of uncertainties in the parameters can be determined.
There are some important limitations to this approach, however. First,
any empirical model is, at best, descriptive of observations, but is
neither exact nor complete. To the extent that significant processes are
not included in the model, their contributions to the uncertainty cannot
be determined. Second, the appropriate distributions, and inter-
dependence of the model parameters, are seldom well known. While one can
12-24
-------
55
in
Q
m
<
00
o
QC
Q.
4.0
3.0
2.0
1.0
Distribution coefficients for:
1. trench materials
2. host soil
3. aquifer
Degree "of trench cap failure
Distance from trench bottom to aquifer
Length of disposal site
Ground-water velocity In host soil
5. Ground-water velocity In aquifer
Calculated maximum
body dose for RA-226
0.5
1.0
1.5
2.0
2.5
3.0
NORMALIZED DOSE, X/n
Figure 12-10. Results of Uncertainty Analysis for Ra-226,
Standard Deviation * 65%
12-25
-------
4.0
55
z
til
Q
CO
<
CO
o
cc
Q.
3.0
2.0
1.0
Distribution coefficients for:
1. trench materials
2. host soil
3. aquifer
1. Degree of trench cap failure
2. Distance from trench bottom to aquifer
3. Length of disposal site
4. Ground-water velocity In host solG
5. Ground-water velocity In aquifer
Calculated maximum
body dose for H-3
NORMALIZED DOSE,
Figure 12-11. Results of Uncertainty Analysis for H-3,
Standard Deviation = 47%
12-26
-------
obtain insight into the uncertainties, a quantitative estimate of overall
uncertainty may be more representative of opinion than of objective
fact. The following discussion will consider uncertainties associated
with two models, the deposited activity model for radionuclides and the
specific activity model for C-14. The discussion in this section is
based primarily on material from a National Council on Radiation
Protection and Measurements (NCRP) publication (NCRP84).
12.5.1 Interception
The deposited activity model presumes that a fraction, fr, of the
depositing flux is intercepted by vegetation and incorporated into the
associated crop. The remainder is considered to deposit on the soil,
where it may subsequently be available for uptake by the root system.
For forage or leafy vegetable crops, the interception fraction is
strongly dependent on the areal density, Yv, of the crop. In this
case, the quantity fr/Yv can reasonably be considered as lognormally
distributed with a median (geometric mean or gm) of 1.8 m2/kg (dry
weight) and a geometric standard deviation (gsd) of 1.6. For other
crops, the relationship between fr and Yv is more case specific.
While a default value of fr, such as 0.2 or 0.25, is frequently used
for irrigation spray or particulate deposition on these crops, there is
no consensus as to what the distribution of fr should be; however, a
gsd of 2 might be considered reasonable.
12.5.2 Crop Yield
Strictly speaking, Yv is the total areal density of the
above-ground portion of a crop. For many crops, the edible portion may
be only about one-third of this quantity. A typical value for Yv is
about 2 kg/m2. Again, a gsd of 2 might be considered reasonable for
estimating the uncertainty in this parameter.
12.5.3 Weathering Half-Life
The weathering half-life, tw, varies with the chemical form of the
depositing radionuclide, crop type, stage of development, and the
processes affecting removal. As customarily measured, it also includes
the effect of dilution caused by plant growth. A nominal value of
14 days with a gsd of 1.6 is representative. For radionuclides with
half-lives substantially longer than 14 days, weathering is the principal
loss mechanism.
12.5.4 Other Parameters
The time of exposure during the growing season generally is
significantly longer than the weathering half-life and therefore its
uncertainty is not an important factor in the model. Similarly, the
uncertainties in the bulk and surface density of soil do not make
substantial contributions to the overall uncertainty.
12-27
-------
12.5.5 Uptake from Soil
Generally, direct deposition onto plant surfaces is a much more
important contaminating mechanism than uptake from soil, if trench
rather than spray irrigation is used, however, uptake from soil would be
important because there would be no direct deposition on vegetation.
Soil-to-plant transfer factors show a wide range—a gsd of 4 is typical.
Since soil-to-plant transfer is really soil-to-soil water-to-plant
transfer and since the soil-to-soil water distribution factor, Kd, can
vary widely, the large uncertainty in uptake from soil is not
surprising. Another significant contributor to the uncertainty in uptake
from soil is the environmental removal rate of radionuclides from soil.
Since leaching, a principal consideration, is highly dependent on Kd,
there is a strong correlation between the transfer factor and the
leaching rate. Overall, a gsd of 5 might be considered reasonable for
deposition to soil-to-plant transfers.
12.5.6 Transfers to Milk and Meat
Intakes of feed and water by animals can best be determined on a
site-specific basis. While typical values can be assigned, the
uncertainties in these values may represent differences in specific
management practices rather than random variation. The transfer factor
for iodine from feed to milk, fm, has a gsd of about 1.7. other
radionuclides would have comparable uncertainties. Taking into account
other uncertainties such as that in the milk production rate, the overall
transfer from feed to milk could reasonably be assigned a geometric
standard deviation (gsd) of about 2. Similarly, on the basis of data for
cesium, the uncertainty in the meat transfer coefficient, ff, can be
considered to be represented by a gsd of about 2.3. Considering other
associated uncertainties, an overall gsd of about 3 would be reasonable
for feed-to-meat transfers.
12.5.7 Carbon-14
Because atmospheric carbon dioxide is the primary source of carbon
in plants, there is little uncertainty in the transfer of carbon-14 to
plants and animal products. At equilibrium, plants and animals will have
the same specific activity as the atmosphere to which they are exposed.
What uncertainty there is has to do with considerations affecting the
atmospheric concentration where the different crops providing food and
feed are grown. Such considerations have much more to do with scenario
considerations than with model uncertainty.
12.5.8 Summary
The overall uncertainties in food chain models can be considerable.
The uncertainty for leafy vegetables and pasture feed can be represented
by a geometric standard deviation (gsd) of about 2.3. For other produce
where direct deposition is the source of contamination, a gsd of 3.2
12-28
-------
would be appropriate. For food products where soil-to-crop transfer is
the dominant contamination mechanism, a gsd of about 5 would be
reasonable unless site-specific parameters can be used. The gsd values
for transfers of activity directly deposited onto forage to milk and meat
would be about 3 and 3.8. respectively. For transfers from other feeds,
these values would increase to about 3.8 and 4.4, respectively. With
specific activity models such as that used for C-14, the principal
uncertainties are usually associated with the postulated scenario rather
than model assumptions. In any case, the uncertainty estimates in this
section should not be considered as authoritative, verified values but as
aids to evaluating the potential significance of the food pathways.
12.6 Uncertainty Due to Estimation of Organ Doses
As mentioned in Chapter 6, the primary sources of uncertainty in
estimating doses to organs of individuals exposed through ingesting or
inhaling radionuclides are associated with: (1) ICRP model formulation
and (2) parameter variability caused by measurement and sampling errors
or natural variations. It was also mentioned that the Agency's ability
to quantify these uncertainties is extremely limited because of the lack
of experimental data.
This difficulty can be attributed to several factors. First, most
of the ICRP models for estimating doses to organs of individuals in the
general population were developed from animal experiments; the metabolic
behavior of radionuclides in animals and man often differs
significantly. Differences are also observed in the anatomical structure
o£ organs and tissues in animals and man. To quantitatively determine
the uncertainties associated with using animal-based models requires
extensive animal and human data and a means for properly extrapolating
animal results to humans. Data and methods are both lacking in this
area. Second, most of the assumptions used in the ICRP modeling approach
(e.g., for handling ingrowth of radioactive daughters, for relating
similar nuclides with different metabolic patterns, or for estimating
doses to organs consisting of heterogeneous cell populations) have not
been properly tested and verified. Generally, the experimental data
supporting these assumptions are very sparse as well.
In addition, for those models that are assumed from the outset to be
correct, considerable uncertainties are expected in estimating organ
doses because of the variability in anatomical and physiological
parameters. Parameter variability primarily relates to age differences
in the general population. The parameters employed for EPA modeling
purposes were obtained from persons with anatomical or metabolic
characteristics similar to "Reference Man" and represent "best estimates"
or "average" values from parameter distributions. The parameter values
are normally scaled for other age groups in the general population. This
method ignores the recognized variability among individuals, and it
automatically introduces bias when extending these models to other
members of the population. Many of the parameters used for estimating
12-29
-------
doses bo organs, such as radionuclide intake rates (I), organ masses (m),
blood transfer factors (f^), organ uptake rates (f2), and
biological half-lives of ingested radionuclides, vary with age. In
addition, considerable variability in these parameters can exist among
individuals of the same age group. To properly ascertain the magnitude
of this uncertainty requires knowing how these parameters vary with age
and obtaining a parameter distribution for each age group. Again, there
are limited data upon which to base such an analysis.
If we restrict our attention to an "average" individual, some
parameter uncertainties will be greatly reduced, and the overall
uncertainty may be better obtained. In particular, for radioiodine, the
variability of the target organ (thyroid) mass is quite large, especially
when all age groups are considered; nevertheless, the average thyroid
mass is known to be within perhaps _+ 20 percent. The major sources of
error with respect to 1-129 dosimetry appear to be related to assumptions
regarding intake volume and f^. Actual average daily water intake
is probably between 1 and 1.5 L, rather than 2 L, as assumed here. Based
on more recent studies, the assumed value for £2 (0.3) is probably
high, perhaps by a factor of 2. Both of these errors would tend to bias
the dose estimates high. Thus, the dose conversion factor for 1-129
should be regarded as an upper bound "conservative" estimate.
The major source of dosimetric uncertainty with respect to C-14 is
uncertainty over retention time in the body. The models used here assume
that the C-14 released from the waste site into ground water and
subsequently ingested in drinking water is handled by the body like
carbon ingested in food. This assumption is highly conservative.
Carbon-14 ingested in an inorganic (carbonate/bicarbonate) form will be
rapidly eliminated from the body through exhalation. Furthermore,
organic C-14 compounds originating from the waste site may not, for the
most part, be in a form which the body can utilize as a carbon source;
hence/ the average retention time may also be low for ingested organic
compounds. In conclusion, the dose conversion factors for C-14 should be
regarded as upper bounds and may have overestimated the actual average
dose.
12.7 Uncertainty Due to Health Effects Conversion Factors
The uncertainties in the risk estimates for radiogenic cancer are
discussed in Section 7.5. The chief sources of uncertainty associated
with a uniform whole-body dose of low-LET radiation to the general
population, and their estimated magnitudes, are summarized in
Table 7-10. The estimated combined uncertainty, due to all sources,
encompasses the range from 23 to 160 percent of the central estimate.
Based on the central estimate of 395 fatal cancers per million person-rad
(see Table 7-3), the overall uncertainty range is 91 to 630 fatal cancers
per million person-rad. For risks to individual organs, the percent
uncertainties may be much larger.
12-30
-------
The uncertainties in the risk estimates for radiation-induced
genetic effects are discussed in Sections 7.6.4 through 1.6.1. A list of
the sources of uncertainty and their magnitudes is given in Table 7-17.
As noted in Section 7.6'.7, the EPA genetic risk estimate is believed to
be uncertain by about a factor of 4 either way. Based on limited human
data, however, it is more likely to be on the conservative (high) side.
12.8 Uncertainty of the Overall Health Effects and Maximum CPG Dose
Analyses
Based on the uncertainty analyses discussed for each of: the
components in Sections 12.2 through 12.7, one may calculate the overall
uncertainties using the method to be presented in this section. Since
uncertainties for all of the components are not quantifiable, the
analyses of the overall uncertainty of health effects and maximum CPG
dose cannot be performed accordingly. Nevertheless, the analysis
calculates an example of the overall uncertainty by using the quantified
uncertainties and the assigned uncertainties made for those components
that are not quantifiable. The selected example represents the disposal
technology specified by the 10 CFR 61 regulations (NRC82a) applied to a
site in a humid permeable hydrogeological setting.
12.8.1 Method of Analysis
For the purpose of the uncertainty analysis, the cumulative health
effects and the maximum CPG dose analyses may be expressed, respectively,
as:
and
where:
HE =
IR =
AT =
MC =
FCC =
DCF =
HE = [IR] [AT] [FCC] [DCF] [HCF] (12-3)
BD = [IR] [MC] [FCC] [DCF] [BDC] (12-4)
the cumulative health effects;
the radionuclide inventory in the waste;
the cumulative radioactivities being discharged into the regional
river basin based on the unit curie of disposal;
the maximum concentration of radionuclides due to the unit curie
disposal;
the food chain factor; _j
the organ dose conversion factors; >
12-31
-------
HCF * the health risk conversion factor;
BD s the maximum CPG dose;
BDC s the CPG dose conversion factor; and
[] designates a random variable.
Equations 12-3 and 12-4 are a linear multiplicative chain of
independent parameters. Therefore, the overall uncertainty (geometric
standard deviation) of the assessment can be calculated from the standard
deviation for each individual parameter without undergoing a time
consuming calculation of joint probability functions. However, in order
to apply this method, one has to assume that the probability density
distribution for each component as discussed in Section 12.1.3 is in the
form of a log-normal distribution. For the purpose of this analysis,
this assumption is thought to be acceptable, judging from the results of
observations on the distribution of general environmental parameters.
The mean value and its standard deviation for the overall assessment may
be calculated by employing the Theorem of Variance for the joint
distribution of random variables as:
DCF
rHCF
and
(12-5)
MC
FCC
DCF
HCF
for the cumulative health effect analysis; and
and
BD
BD~ 4*
DCF
(12-6)
FCC
BDC
for the maximum CPG dose assessment. In the above equations:
V = the log transformed mean value; and
cr » the log transformed standard deviation; and
the subscripts HE, IR, FCC, DCF, HCF, BD, MC, and BDC are the same as
those defined for Equations 12-3 and 12-4.
12-32
-------
Therefore, the uncertainty of the assessment can be calculated from
Equations 12-5 or 12-6, if the mean values and the standard deviations
for all the components of the various calculations are defined.
12.8.2 Uncertainties of assessment Components
As was stated in Sections 12-2 through 12-7, the quantification of
the uncertainty for each component is extremely difficult to obtain and
the uncertainty was not quantified except for the components of the
transport through the geosphere and health effects conversion factors.
This impeded the quantification of the overall uncertainties for the
cumulative health effects analysis and the maximum CPG dose analysis.
Nevertheless, the best estimate on the uncertainties for each component,
through expert judgment, was made to obtain an order of magnitude
quantification for the overall uncertainties of the assessments. The
analyzed and estimated standard deviations for C-14 and 1-129 for each
component, together with the mean values extracted from the results of
the assessment using PRESTO-EPA models, are listed in Tables 12-1 and
12-2 for the cumulative health effects analysis and the maximum CPG dose
analysis, using the example case of the disposal technology specified by
the 10 CFR 61 regulations at a humid permeable site.
12.8.3 Results of the Overall Uncertainty Analysis
As indicated in Section 12.8.2, the quantification of the
uncertainties for all components of the uncertainty analysis cannot be
conducted at this time. Quantification of the uncertainties for the
radionuclide transport through the geosphere and for the heialth effects
conversion factors has been obtained from detailed analyses!, while best
expert judgment is used to estimate the uncertainties for the remaining
components. These results were presented in Tables 12-1 arid 12-2 and
served as input to the overall assessment of uncertainties carried out
with Equations 12-5 and 12-6.
Parallel analyses are also conducted for the minor contributors,
1-129 for the cumulative health effects analysis and C-14 for the maximum
CPG dose analysis. The results of the analyses indicated that the upper
bound, or the uncertainties for the cumulative health effects analysis
for both C-14 and 1-129, is identical and equal to 161 percent of the
central value, while the uncertainties for the maximum CPG dose analysis
are 147 percent of the central value for 1-129 and 132 percent of the
central value for C-14.
r/
The combined results of the assessment for the 10 CFR 61 disposal
technology applied to a site in a humid permeable region are presented in
Table 12-3. Note that these results reflect the combined effects from
both C-14 and 1-129. It is also interesting to note that the major share
of fatal health effects (about 98 percent) is estimated to originate from
C-14, while 1-129 is the major contributor (92 percent) for maximum CPG
dose.
12-33
-------
Table 12-1. Summary of estimated mean and standard deviation
for cumulative health effects analysis for C-14
Items
Estimated Estimated
value Unit standard deviation
Radionuclide
Inventory
Cumulative Activities
Discharged
Food Chain
Factor
Dose Equivalent
Conversion Factor
Health Risk
Conversion Factor
6.681F.+02 Ci
6.681E+02 (100%)
(Ci/yr)
5.984E-01 xlO.OOO yr/Ci 0.00 (0%)
man-pCi/vr
1.604E+04 Ci/yr
0.00
(0%)
1.540E-06 mrem/yr/pCi/yr 9.240E-07 (60%)
death/vr
2.806E-01 man-mrem/yr 2.370E-01 (60%)
12-34
-------
Table 12-2. Summary of estimated mean and standard deviation
for maximum CPG dose analysis for 1-129
Items
Estimated
value
Unit
list i mated
standard deviation
Radionuclide
Inventory
Maximum Concentration
Factor
Food Chain
Conversion Factor
Dose Equivalent
Conversion Factor
7.899E+00 Ci
2.609E-09 Ci/rtrVci
7.899E-I-00 (100%)
1.044E-09 (40%)
4.819E+11 pCi/yr/Ci/m3 0.00
(0%)
8.568E-04 mrem/yr/pCi/yr 5.141E-04 (60%)
Table 12-3. Results of uncertainty analyses: 10 CFR 61
technology at a humid permeable site
Analyses
Calculated value
Standard deviation
Fatal Health Effects
Maximum CPG Dose
3.9 deaths
9.2 mrem/yr
6.3 deaths (161%)
13.4 mrem/yr (146%)
12-35
-------
Based on the above results (Table 12-3), the upper bounds of the
results of analysis for the analyzed scenario are estimated to be 10.2
cumulative fatal health.effects and 22.6 mrem/yr for the maximum CPo'dose.
The cumulative health effects presented in Chapter 9 for the
analyzed scenario are the combined result of fatal cancers and serious
genetic effects. Serious genetic effects comprise, in general, only a
small fraction of total health effects (fatal plus serious genetic).
Therefore, for the purpose of the uncertainty analysis, one may
reasonably assume that the uncertainty for the serious genetic effects is
the same as that for the fatal cancers. Thus, the upper bound of the
combined total population health effects (fatal plus serious genetic) is
calculated to be 11.5 health effects, with a central value of 4.4 health
effects.
12.9 Conclusion
Despite the difficulty of quantifying the uncertainty of the
cumulative health effects and the maximum CPG dose, an effort was made to
quantify the uncertainties. Since raultidisciplinary processes were
involved, the analysis was divided into five components to permit the
uncertainty analysis for each component to be conducted by experts in the
field. The results of the analysis for each component are summarized as
follows:
12,9.1 Source Term Concentration ,
The analysis centered on C-14 and 1-129 because these predominant
radionuclides are projected to be the major contributors to the
cumulative health effects and the maximum CPG dose, since the
concentrations of these radionuclides, particularly 1-129, are in a trace
amount, the accuracy of their measurement was found to be poor.
It is believed that the uncertainty of the C-14 and 1-129
concentrations in daily samples collected from each'waste stream may vary
considerably from one to another. However, the uncertainty of the
concentrations of all radionuclides from all waste streams for 20 years
is expected to be much smaller than that for any daily sample of any
waste stream. Based upon a detailed evaluation of the EPA LLW source
terra (Gr86) and an-extensive study of the occurrence of C-14 and 1-129 in
waste samples (C185), a qualitative uncertainty of a factor of 2 was
assigned to C-14 and 1-129.
12.9.2 Radionuclide Transport in Geosphere
It was found that the major uncertainty for the geosphere transport
will be dominated by the selection of site scenarios. This finding
occurs because the amount of available dilution water in the underlying
aquifer will greatly affect the results of the assessment. Any
uncertainties resulting from these site scenarios were not considered
12-36
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because the risk assessments were concentrated on typical sites for three
distinct hydrogeological settings. When this simplification is imposed,
the major uncertainties will be governed by the rate of infiltration, the
distribution coefficients, and the distance of transport.
Fortunately, the model output parameters are controlled by
integrated effects either over an extremely long period of time (for the
cumulative health effects assessment) or over the entire disposal area
(for the maximum CPG dose assessment), which have greatly converged the
uncertainties of the output parameters, at least for mobile
radionuclides. Our results concluded that the uncertainty of the
cumulative C-14 and 1-129 radionuclide releases for the health effect
assessment is near zero, and the uncertainty of the maximum CPG dose
assessment for 1-129 and C-14 is approximately 40 percent and 10 percent
of the mean values, respectively.
12.9.3 Radionuclide Transport through the Food Chain
Radionuclide transport through the food chain pathway is divided
into two categories, the drinking water pathway and the nondrinking water
pathway. The radionuclide transport in the food chain for nondrinking
water includes the deposition from the atmosphere or from irrigation
water, plant uptake from soil, transfer to milk and meat, and finally,
human consumption.
When the overall uncertainties of the transport through the food
chain are considered, the uncertainties resulting from the nondrinking
water pathway are negligible. Whereas the drinking water pathway
accounts for the major portion of exposures, C-14 and 1-129 account for
approximately 99 percent of the total exposure from the analyzed
scenario, which uses the 10 CFR 61 technology in a humid permeable site.
Therefore, the uncertainties of radionuclide transport through the food
chain pathway are dominated by the uncertainty of the daily consumption
of drinking water.
Furthermore, since the PRESTO-EPA model calculates the radiation
exposure to average persons in the United States, the variation in per
capita consumption of drinking water resulting from individual
differences should not be considered part of the uncertainty. Therefore,
the overall uncertainty of the radionuclide transport in food chain
pathways for the analyzed scenario is small and can be neglected.
12.9.4 Organ Dose Conversion Factor
The primary sources of uncertainty in estimating doses to organs of
individuals exposed to radionuclides are associated with: (1) ICRP model
formulation and (2) parameter variability caused by measurement and
sampling errors or natural variations. The quantification of these
uncertainties is extremely difficult because of the lack of experimental
data. Therefore, for many radionuclides of interest no quantitative
12-37
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statement can be made about the uncertainties associated with the use of
EPA dose conversion factors. Data are lacking and more research is
needed to test models and determine the variability in model parameters.
12.9.5 Health Effects Conversion Factor
The estimated overall uncertainty resulting from all sources
encompasses the range from 23 to 160 percent of the central value. Based
on the central estimate of 395 fatal cancers per million person-rad, the
overall uncertainty range is from 91 to 630 fatal cancers per million
person-rad. For risk to individual organs, the percent uncertainty may
be much larger.
12.9.6 Results of Overall Uncertainty Analysis
Realizing that the uncertainties for all of the components are not
quantifiable, the analysis calculated an overall upper bound of the
results for a selected scenario by using the best estimated uncertainties
for each component. The analyzed scenario, which applied the disposal
technology specified in 10 CFR 61 for the humid permeable site, was
selected. The results of the analyses for combined effects from C-14 and
1-129 are 161 percent of the central value of 3.9 deaths for the
cumulative fatal health effects over 10,000 years, and 146 percent of the
central value of 9.2 mrem/yr for the maximum CPG dose.
The upper bounds of the combined results for the analyzed scenario
are estimated to be 11.5 total health effects (10.2 for fatal health
effects and 1.3 for the genetic effects) for the cumulative health
effects analysis and 22.6 mrem/yr (21.0 rarem/yr from 1-129 and
1.6 mrem/yr from C-14) for the maximum CPG analysis.
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REFERENCES
CL85 Cline, J. E., Noyce, J.R., Coe, L.J. and K.W. Wright, Assay of
Long-Lived Radionuclides in Low-Level Wastes from Power Reactors,
U.S. Nuclear Regulatory Commission, NUREG/CR-4101, April 1985.
DOE86 U.S. Department of Energy, Integrated Data Base for 1986:
Fuel and Radioactive Waste Inventories, Projections, and
Characteristics, DOE/RW-0006, Rev. 2, September 1986,,
Spent
Gr86 Gruhlke, J. M., Neiheisel, J. and L. Battist, Estimates of the
Quantities, Form and Transport of Carbon-14 in Low-Level
Radioactive Wastes, EPA 520/1-86-019, U.S. Environmental
Protection Agency, Office of Radiation Programs, Washington, D.C.
20460, September 1986.
Hu86 Hung, C.Y., An Optimum Model for Application to the Assessment of
Health Effects Due to Land Disposal of Radioactive Wastes,
Proceedings of Nuclear and Chemical Waste Management, Vol. 6, 1986.
Hu87 Hung, C. Y., A Critical Evaluation of the Uncertainty in the
Population Health Effects Analysis from the Release and
Groundwater Transport Model used in PRESTO-EPA for LLW Standard,
Technical Report, in press, Environmental Protection Agency,
Washington D. C., 1987.
Ke85 ICempf, C.R., Alternatives for Packaging C-14 Waste: C-14 Generator
Survey Summary, Brookhaven National Laboratory, Report A-3172,
1985.
NCRP84 National Council on Radiation Protection and Measurements,
Radiological Assessment: Predicting the Transport,
Bioaccumulation, and Update by Man of Radionuclides Released to
the Environment. NCR? Report No. 76, Be the sd a-, Md., March 1984.
NRC81 U.S. Nuclear Regulatory Commission, Draft Environmental Impact
Statement on 10 CFR Part 61, Licensing Requirements for Land
Disposal of Radioactive Waste, Volumes 1-4, NUREG/CR-0782,
September 1981.
NRC82a U.S. Nuclear Regulatory Commission, Final Environmental Impact
Statement on 10 CFR Part 61 Licensing Requirements for Land
Disposal of Radioactive Wastes, Volumes 1-3, NUREG-0945, November
1982.
NRC82b U.S. Nuclear Regulatory Commission, Licensing Requirements for
Land Disposal of Radioactive Waste, 10 CFR 61, Federal Register,
47 (248):57446-574788, December 27, 1982.
NRC86 U.S. Nuclear Regulatory Commission, Update of Part 61 Impacts
Analysis Methodology, 3 volumes, NUREG/CR-4370, January 1986.
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Chapter 13: PREDISPOSAL WASTE MANAGEMENT OPERATIONS
13.1 introduction
EPA is proposing an annual dose limit for the CPG for LLW management
operations prior to disposal. Predisposal management operations include
preparation of the waste for disposal, i.e., compaction, incineration,
solidification, packaging, handling, storage, and placement. These
activities could be carried out, for example, by LLW generators (power
plants, industries, hospitals, medical centers, or DOE sites), at or
adjacent to operating LLW disposal facilities, or at regional LLW
processing facilities designed to serve a State or an entire Compact.
The following analysis assesses the potential exposures to the public
from radiological releases during predisposal management operations.
Waste generators increasingly are opting for volume reduction and
waste processing to meet NEC's waste stabilization requirements (NRC82),
to reduce disposal costs, and to stay within volume limits imposed by
host States for existing disposal facilities (LLR86). This trend in the
processing of LLW is being met in a number of ways. Large LLW generators
are building their own processing facilities and small generators are
being serviced by mobile processing units (i.e., compactors,
solidifiers). States and commercial companies are planning to establish
regional facilities solely for processing LLW, with the processed LLW
then shipped to a disposal site.
These trends toward (1) widespread processing of wastes at a large
number of diverse facilities and generators, and (2) possible long-term,
aboveground storage present an area in which environmental protection
standards for certain releases from these activities are lacking. In
some cases, the processing and storage of LLW done at various uranium
fuel cycle facilities, such as power reactors and fuel fabrication
plants, will be covered under the 40 CFR 190 standards (EPA77a). These
standards encompass all activities carried out at these facilities, and
many LLW processing and storage operations at these facilities use the
same techniques as are found at any LLW processing facility, such as
evaporation, incineration, compaction, handling, and storage. Several of
these operations were evaluated for the UFC standards (EPA73a,b). In
addition, atmospheric releases of radionuclides from LLW processing and
storage operations at those DOE- and NRC-regulated facilities are covered
by the Clean Air Act standards pursuant to 40 CFR 61 (EPA85). These
standards also encompass all airborne activities at these facilities,
which include many of the same LLW processing and storage operations as
are carried out at specific LLW processing facilities (EPA84). However,
releases through such other pathways as water and gamma exposure from
processing operations, as well as long-term storage at NRC-regulated
large, away-from-generator or central processing facilities and at DOE
facilities, would not be covered under 40 CFR 61.
13-1
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EPA has not performed the comprehensive quantitative risk assessment
necessary to determine the health impacts of the various predisposal
management operations. However, some limited analysis has been done on
operational spillage (see Section 13.4.1). Such comprehensive
assessments would first require identification of potential exposure
pathways that are not already limited by existing regulations or
standards. For example, since the proposed standard represents a limit
on the cumulative dose through all pathways, the contribution of the air
pathway, even though limited by the Clean Air Act emissions standards,
would need to be further quantified, including exposure resulting from
surface spillage, followed by resuspension and offsite transport.
It is theoretically possible that offsite contamination could occur
as a result of spillage and surface runoff during a rainstorm (or
flood). Finally, if waste treatment or storage vessels are located near
the boundary of the site, external direct gamma radiation could also
cause exposure to individuals at the site boundary.
Therefore, the proposed predisposal management standard would
probably result in actions to control spillage (e.g., by the use of good
housekeeping practices and proper design of handling equipment) or to
limit direct gamma radiation (e.g., by placing storage facilities away
from the facility boundary).
13.2 Basic Assumptions
In this analysis, we have examined the most likely major steps in
the management of LLW. They include: evaporation, incineration, liquid
storage, packaging, solid waste storage, compaction, and solidification
processes.
As indicated earlier, many of the operations already take place at
various generators' facilities, and in some cases these operations were
analyzed in connection with the UFC 40 CFR 190 and CAA 40 CFR 61
standards (EPA73a,b, EPA84). The data presented here come from reports
of the DOE, NRG, and EPA. Some of the data deals with hypothetical
generic facilities; other data are concerned with actual operations at
DOE or commercial facilities.
The basic assumption underlying this analysis is that the major
radiation dose to the critical population group from the facilities is
through airborne discharges to the atmosphere based on present
practices. Some gamma exposure could be present and some minimal liquid
releases could occur. In almost all cases, however, operations today
recycle many liquids for use, and waste liquids are usually solidified
and disposed of as solids.
The exposure pathways, demography, and other parameters, as well as
the mathematical models relating dose to man for the estimated
radionuclide releases from the generic facilities, are described in
13-2
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DOE79. (These documents were reviewed by EPA in 1979 and were found to
be adequate.) During many waste management operations, some o£ the
radionuclides in the wastes are released as volatile gases and
particulates. Before these gases and particulates are releaised to the
atmosphere, they are routed to treatment systems designed to remove the
majority of the radionuclides. Those releases cited from specific
facilities are discussed in EPA84 and include LLW operations, such as
evaporation and incineration and LLW disposal sites. The maximum annual
CPG doses are based on hypothetical area residents whose habits would
tend to maximize the dose.
Several major factors that can affect the potential radiation dose
to the CPG and populations as a result of release of radionuclides to the
atmosphere are as follows: proximity to the plant, the pathways by which
the radionuclides can reach people, the length of time during which the
radionuclides continue to pose a health hazard, decay time,
meteorological factors, facility capacity, and off-gas treatment.
13.3 General Air Emissions Pathway
A review was made of the previous evaluations by EPA, in connection
with its regulations for radionuclide air emissions (40 CFR 61). The
results of this review are given in this section (EPA84, EP?i85).
13.3.1 Department of Energy Facilities
The DOE administers many government-owned, contractor-operated
facilities that emit radionuclides to the air. Operations at these
facilities include research and development; production of nuclear
weapons; enrichment o£ uranium and production of plutonium for nuclear
weapons and reactors; and processing, storing, and disposing of
radioactive wastes. Not all of these operations take place at all sites,
of course. Certain of these facilities are on large sites, some of which
cover hundreds of square miles in remote areas; several States are host
to such sites. Some smaller facilities resemble typical industrial sites
and are located in suburban areas. As indicated earlier, many of these
facilities use, or are expected to use, the same processing, management,
and storage techniques as one would expect to find at a large commercial
centralized LLW processing center.
Each facility differs in emission rates, site size, nearby
population densities, and other parameters that directly affect the
offsite dose from radionuclide emissions. Many different radionuclides
are emitted to the atmosphere. Six sites have multipurpose operations
spread over very large areas. Another 12 or so sites are primarily
research and development facilities located in more populated areas.
Reactor and accelerator operations at these sites may release radioactive
noble gases and tritium; other operations may release small amounts of
other radionuclides. several facilities are primarily engaged in weapons
development and production, and may release small amounts of tritium and
13-3
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certain long-lived radionuclides. Finally, two sites are dedicated
entirely to gaseous diffusion plants that enrich uranium for use in
commercial electric power reactors and for defense purposes. They
primarily emit uranium.
At 15 of the smaller DOE facilities, which are considered as a group
in the Radionuclides Emissions BID (EPA84) because of their relatively
small health impact, the doses to the nearby individuals are estimated to
be considerably less than 1 mrem/yr. These small doses were also
reported at less than 1 mrem/yr in a previous report (EPA77b).
A second group, which consists of the 13 facilities having the
largest emissions of radionuclides, was studied in more detail. The
collective dose to the populations living around these sites is also low,
no higher, than about 10 person-rem after 1 year of site operation.
The doses from these facilities to the CPG are generally estimated
to range from 2 to 10 mrem/yr, although two facilities indicated doses of
greater than 25 rarem/yr. These exposure results reflect all operations
(resulting in airborne releases) carried on at these sites, and the major
releases are those from the principal activities carried on at these
facilities, e.g., reactor operation, fuel reprocessing, enrichment, etc.
Therefore, the various LLW processing, management, and storage operations
carried on at the sites contribute only a small percentage of the
radioactivity to the offsite population. A rough estimate would be 10 to
15 percent. Therefore, it is expected that the doses to the CPG from LLW
operations at these DOE facilities will also be a small percentage
(probably several orders of magnitude less) of that reported for the
overall health impact from air emissions from all operations on that
site. Chapter 3 lists the various DOE facilities throughout the U.S.,
and also presents the volume and radionuclide characteristics of the
various LLW generated and disposed of at these facilities.
13.3.2 Nuclear Regulatory Commission-Licensed
and Non-DOE Federal Facilities
NRC-licensed and non-DOE Federal facilities include research and
test reactors, shipyards, the radiopharmaceutical industry, and other
research and industrial facilities. This category includes both
facilities licensed by NRC and those licensed by a State under an
agreement with NRC. These facilities number in the thousands and are
located in all 50 States. Uranium fuel-cycle facilities are not included
because radionuclide emissions from these facilities are limited by EPA
standards (40 CFR 190). See the discussion in Sections 13.1 and 13.2.
The principal differences among these various types of activities are
their emission characteristics and rates, their sizes, and the population
densities of the surrounding areas.
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The vast majority of NRC-licensed and non-DOE Federal facilities
emit relatively small quantities of radionuclides, which cause
correspondingly low doses to people living nearby. From EPA studies and
contractor-supported analysis, the maximum radiation doses from these
facilities were less than 1 mrem/yr, with the total dose to the
population living around a site rarely exceeding 1 or 2 person-rem/yr of
operation (EPA77b, EPA84). Various LLW processing, management, and
storage operations are also carried out at these facilities. In many
cases, the quantities of waste and hence the waste management operations
are small. Chapter 3 presents a description of a number of the waste
streams generated by these different operations.
Waste management and storage operations take place at all facilities
where radionuclides are used, and the size of the waste operations can be
either small or large depending on the annual throughput of materials and
waste generated. We want to emphasize that these doses are calculated
from all operations taking place at a specific facility. It is expected
that the doses from the various LLW management operations will be only a
small percentage of the overall amount, in many cases probably less than
10 to 20 percent.
13.3.3 Air Emissions from Compaction
DOE' estimates of doses to the CPG from gaseous effluents released
from a generic model fuel bundle residue compaction facility are in the
range of 1E-11 to 1E-09 mrem/yr (DOE79).
13.3.4 Air Emissions from Incineration
A DOE estimate of doses to the CPG from gaseous effluents released
from a generic model solvent incineration facility was in the range of
1E-09 to IE-OS mrem/yr, whereas for a generic model LLW incineration
facility the range was 1E-18 to 1E-07 mrem/yr (DOE79). Further generic
analysis for intermediate level waste (or what might be considered
greater-than-Class C waste) found the dose to the CPG ranged between
IE-OS and 20 mrera/yr.
Another environmental impact analysis (Ph84) of incineration of
institutional LLW indicated that the radionuclide air emissions due to
incineration are very small and that either the CPG organ doses or
whole-body doses will also be small (less than 0.001 mrem/yr).
in a generic licensing report to the NRC, the Newport News
Industrial Corporation and Energy Incorporated presented an environmental
impact analysis of their incinerator system for LLW volume reduction at
nuclear power plants. The environmental analysis is based on a system
capable of processing up to 1,200 m3/yr and incinerating up to 91 kg/h
of LLW. Their analysis indicated that the maximum dose to the CPG would
be less than 0.1 mrem/yr (En77).
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13.3.5 Packaging
DOE estimates of doses to the CPG from a generic model packaging
facility are in the range of 1E-12 to 1E-06 mrem/yr (DOE79).
DOE also analyzed a generic package facility for intermediate-level
waste (or what might be considered greater-than-Class C waste) and found
the doses to the CPG to be in the range of 1E-04 to 4 mrem/yr (DOE79).
13.3.6 Solidification
DOE estimates for the immobilization of LLW by bitumen and cement
solidification systems indicate doses to the CPG should be in the range
of IE-OS to 0.1 mrem/yr (DOE79).
13.3.7 Storage
DOE also considered storage as an option for its many generic
processing operations, such as those discussed in Sections 13.3.3 through
13.3.6. The analysis of doses to the CPG was in the same range as that
indicated for the previous packaging, solidification, and incineration
operations, 1E-18 to 0.1 mrem/yr. For intermediate-level radioactive
wastes, the dose analysis resulted in a range of IE-OS "to 1 mrem/yr
(DOE79).
13.4 LLW Disposal Facilities Operations
Normal operational releases from an LLW disposal facility can
potentially occur through small spills and releases resulting from normal
waste handling and disposal operations. Releases have also occurred at
some existing sites as a result of water management programs involving
evaporation and treatment of trench leachate. Since the need for active
maintenance programs is expected to be eliminated in the future, releases
resulting from such programs were not analyzed. Also considered was the
processing of waste at a regional processing center, which for purposes
of analysis is assumed to be located at the disposal facility.
13.4.1 Operational Spillage
Small leaks and spills from waste containers during normal
operations can potentially be released to the air or contaminate the
ground surface, which can then be carried from the site by the actions of
wind or precipitation runoff. It is believed that the contamination of
the ground surfaces at the Maxey Flats facility was caused by earlier
cases of inadequate waste handling and site maintenance procedures. It
is known that waste packages delivered to the facility frequently failed
to properly contain the waste within the packages and/or ruptured during
emplacement operations.
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At currently operating facilities, however, considerably more
attention is being paid to minimizing potential surface contamination.
For example, disposal facilities in operation have procedures to survey
facility areas on a routine basis, as well as when possible contamination
is suspected. Allowable contamination limits have been established at
operating facilities, in addition, monitoring programs at all operating
facilities have been improved and routinely sampled for onsite surface
contamination.
Of interest are environmental monitoring results for the Barnwell,
South Carolina, disposal facility. This facility accepts approximately
50 to 70 percent of the LLW in the country. Given the large volume of
waste received at the facility, most of the operational impacts
associated with LLW disposal would be expected to be associated with this
facility. For example, the concentrations of Co-60 and Cs-137 measured
onsite are within the range of measurements of samples collected offsite
(NRC81).
Thus, there appear to be no significant releases of radionuclides
from the operating sites from surface contamination. This is principally
due to the increased attention paid by facility operators,to minimizing
facility contamination. The practice of delivering bulk liquids to
disposal facilities for solidification has been discontinued. All
disposal facilities have license conditions that restrict wastes
delivered to the disposal facilities to dry solids, and include
restrictions on the amount of free-standing liquids allowed in the
waste. Compliance with DOT regulations is also required. Improvements
in waste form and packaging required by 10 CFR 61 will also reduce the
potential for surface contamination and subsequent release to offsite
areas.
Since releases during normal operations caused by spills have not
been significant and are not expected to be significant in the future,
NEC conducted no detailed analysis of these potential pathways of release
and potential public impacts in its EIS for 10 CFR 61 (NRC81). However,
in EPA's risk assessment methodology, spillage was taken into account
(see Chapters 8 and 11). EPA's results indicated that any operational
spillage would account for less than 0.1 mrem/yr to any offsite
individual.
13.4.2 Operational Airborne Emissions
Analysis done at four LLW disposal sites for exposures to offsite
individuals from airborne migration indicated less than 0.2 mrem/yr
(Ad78, FBD78a, FBD78b, FBD78c).
13.4.3 Evaporator Operation
A site investigation and analysis by EPA of the LLW evaporation
system operation in 1974-1975 at the Maxey Flats burial site indicated
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that the maximum individual dose rate received by the nearest resident
was less than 3 rarem/yr (Mo77). 2\n NRC computer model analysis of the
Maxey Flats evaporator for both airborne and aquatic pathways found the
individual dose rate tp be less than 0.01 mrem/yr (Ad78). The difference
in these two analyses is probably based on parameter data used,
especially annual throughput of materials evaporated.
13.4.4 Offsite Gamma Radiation
Offsite external radiation monitoring is done by South Carolina for
the Barnwell waste disposal site. During the period 1983-1985, the
annual background gamma exposure rate at stations not influenced by the
Barnwell LLW burial site was 6E+01 mR/yr. The annual gamma exposure
(background included) rates at the Barnwell exclusion fence and at
stations less than 8 kilometers were 6.6E+01 mR/yr and 5.9E+01 mR/yr.
Thus, the exposure rates at the exclusion fence and the other stations
appear to be within background levels. The operations of the Barnwell
LLW burial site do not appear to influence the annual gamma exposure
levels in the immediate vicinity (SC86).
13.5 Regional Processing Facility
One of the viable options addressed was that of processing waste on
a regional basis at a central processing facility. Such a facility could
be located at or be separate from the disposal facility.
Such waste processing activities can lead to potential airborne
releases of radionuclides and subsequent exposures to the public in the
neighborhood of the regional processing facilities. NRC analyzed the
potential population exposures due to the assumed operation of a central
waste processing facility (an incinerator) that was co-located with the
disposal facility. These exposures were estimated to be approximately
2 person-rera/yr, arising from the assumed incineration of 100,000 nr* of
combustible trash per year. The total population assumed to be exposed
was 480,000 within an 80-kilometer radius of the processing facility.
This would be in the neighborhood of less than 0.01 mrem/yr to the
nearest individual (NRC81).
13.6 summary
Analyses for the various pre-disposal LLW management and storage
operations have not been extensive and, as shown by our review, have been
somewhat fragmented. In some cases though, the analyses has been very
detailed and based on actual operating data (e.g., the data collected by
EPA), while the generic data presented by DOE are based on existing or
available technology applied to proposed facilities. It must be kept in
mind, however, that the doses presented here are related to specific
sizes of facilities, although most studies have shown that these sizes
are probably the optimum. It is not likely that the waste volume
processed in super-sized facilities would be two to three times what is
being planned or handled today.
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In dealing with waste management processing and storage operations,
only a few major alternatives generally will be used. Storage is a
common operation and shielding to prevent gamma exposure is an acceptable
and widely used technique. Evaporation is a well known process and has
been used in the nuclear industry since its inception. Liquids are
usually evaporated and the residues, along with other semiliquids, are
solidified. The solidification process is also well known and has been
in use at nuclear installations since the 1960's (HO76).
General trash is commonly packaged, either with or without
compaction depending on the materials. Compaction is being used more
frequently for radioactive waste generation (DOE79, Jo86). In essence,
packaging consists of containing general trash in steel drums or boxes
for interim storage or disposal.
Incineration is another important technique for treating LLW.
Incineration consists of burning the waste and treating the off-gas for
removal of radionuclides and other noxious materials, thereby reducing
the waste volume and rendering it noncombustible (DOE79, Jo86). In
determining the health impacts from incineration versus direct disposal,
the use of incineration considerably reduces the health effects and CPG
doses over direct disposal (in our analyses, by a factor of 2 - see
Chapter 10, Section 10.7.1(8)).
The underlying reason for this reduction in health impacts is that
incineration transforms those radionuclides that are volatile (in many
cases the majority of the radionuclides present in the waste) from being
present in the water ingestion pathway, as the result of direct disposal
of waste without incineration, to being present in the air inhalation
pathway. In the airborne pathway, the radionuclides are diluted
considerably and the body response health impact from inhalation is
always much less than when the material is ingested, as through the water
pathways.
In summary, potential releases from the airborne pathway and the
waterborne carry-off pathway from.contaminated surfaces are expected to
be on the order of a few mrem/yr. Even when combining these various
operations, the overall CPG should be less than 10 mrem/yr for processing
Class A, B, and C wastes. Where greater-than-Class C waste is processed,
improved technology and techniques may be required to keep the CPG doses
to less than 20 to 25 mrera/yr.
Overall, it is felt that these types of CPG exposures can be further
reduced by:
• the continued practice of strict housekeeping procedures to
maintain potential contamination of equipment and surfaces to
ALARA levels; and
• improvements in waste form and packaging.
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REFERENCES
Ad78 Adam, J.A. (USNRC) and V.L. Rogers (FBDU), A Classification
System for Radioactive Waste Disposal - What Waste Goes Where?,
NUREG-0456 (FBDU-224-10), U.S. Nuclear Regulatory Commission,
Washington, D.C., June 1978.
DOE79 U.S. Department of Energy, Environmental Aspects of Commercial
Radioactive Waste Management, 3 Volumes, DOE/ET-0029, Washington,
D.C., May 1979.
En77 Energy Incorporated and Newport News Industrial Corporation,
Topical Report RWR-1™ Radwaste Volume Reduction System,
EI/NNI-77-7-NP, June 24, 1977.
EPA73a U.S. Environmental Protection Agency, Environmental Analysis of
the Uranium Fuel Cycle, Part I - Fuel Supply.
EPA-520/9-73-003-B, Washington, D.C., October 1973.
EPA73b U.S. Environmental Protection Agency, Environmental Analysis of
the Uranium Fuel Cycle, Part II - Nuclear Power Reactors,
EPA-520/9-73-003-C, Washington, D.C., November 1973.
EPA77a U.S. Environmental Protection Agency, Environmental Radiation
Protection Standards for Nuclear Power Operations, 40 CFR 190,
Final Rule, Federal Register, 42_(9):2858-2861, January 13, 1977.
EPA77b U.S. Environmental Protection Agency, Radiological Quality of the
Environment in the United States, 1977, EPA-520-1-77-009,
Washington, D.C., September 1977.
EPA84 U.S. Environmental Protection Agency, Background Information
Document for Final Rules, 2 Volumes, EPA-520/1-84-022-1,2,
Washington, D.C., October 1984.
EPA85 U.S. Environmental Protection Agency, National Emission Standards
for Hazardous Air Pollutants, Standards for Radionuclides,
40 CFR 61, Federal Register, 50(25):5190-5200, February 6, 1985.
FBD78a Ford, Bacon and Davis Utah, Inc., Compilation of the Radioactive
Waste Disposal Classification System Data Base, Analysis of the
West Valley Site, Prepared for the U.S. Nuclear Regulatory
Commission, FBDU-247-01, Salt Lake City, Utah, September 1978.
FBD78b Ford, Bacon and Davis Utah, Inc., Compilation of the Radioactive
Waste Disposal Classification System Data Base, Analysis of the
Hanford Commercial Site, Prepared for the U.S. Nuclear Regulatory
Commission, FBDU-247-002, Salt Lake City, Utah, October 1978.
13-10
-------
FBD78c Ford, Bacon and Davis Utah, Inc., Compilation of the Radioactive
Waste Disposal Classification System Data Base, Analysis of the
Barnwell Site, Prepared for the U.S. Nuclear Regulatory
Commission, FBDU.-247-004, salt Lake City, Utah, November 1978.
Ho76 Holcomb, W.F. and S.M. Goldberg, Available Methods of
Solidification for Low-Level Radioactive Wastes in the United
States, Technical Note ORP/TAD-76-4, U.S. Environmental
Protection Agency, Washington, D.C., December 1976.
Jo86 Jolley, R.L., et al., Low-Level Radioactive Waste from Commercial
Nuclear Reactors, Volume 2, Treatment, Storage, and
Transportation Technologies and Constraints, DOE Report
ORNL/TM-9846/V2, Oak Ridge National Laboratory, Tennessee, May
1986.
LLR86 Low-Level Radioactive Waste Policy Amendments Act of 1985, Public
Law 99-240, January 15, 1986.
Mo77 Montgomery, D.M., H.E. Kolde, and R.L. Blanchard, Radiological
Measurements at the Maxey Flats Radioactive Waste Burial Site -
1974 to 1975, U.S. Environmental Protection Agency,
EPA-520/5-76-020, January 1977.
NRC81 U.S. Nuclear Regulatory Commission, Draft Environmental Impact
Statement on 10 CFR 61 Licensing Requirements for Land Disposal
of Radioactive Waste, NUREG-0782, Volume 2, Washington, D.C.,
September 1981.
NRC82 U.S. Nuclear Regulatory Commission, Licensing Requirements for
Land Disposal of Radioactive Waste, 10 CFR 61, Federal Register,
47(248)-.57446-57482, December 27, 1982.
Ph84 Philip, P.C., S. Jayaraman and J. Pfister, Environmental Impact
of Incineration of Low-Level Radioactive Wastes Generated by a
Large Teaching Medical Institution, Health Physics,
46(5):1123-1126, May 1984.
SC86 South Carolina Department of Health and Environmental Control -
Bureau of Radiological Health, 1983-1985 Summary Report on
Radiological Environmental Monitoring Around Chem-Nuclear
Systems, Inc., Columbia, S.C., October 1986.
13-11
-------
-------
APPENDIX A
ACRONYMS, ABBREVIATIONS, CONVERSION FACTORS, NOTATION,
AND GLOSSARY
A-l
-------
APPENDIX A
Page
A.I Acronyms A_3
A. 2 Metric-to-English Conversion Factors .... A-7
A.3 Scientific Notation A_9
A.4 Glossary o A-ll
A-2
-------
APPENDIX A: ACRONYMS, ABBREVIATIONS, CONVERSION FACTORS,
NOTATION, AND GLOSSARY
A.I Acronyms
AEA ' Atomic Energy Act of 1954, as amended
AEC U.S. Atomic Energy Commission
AECB Atomic Energy Control Board of Canada
AIF Atomic industrial Forum
ALAP As low as practicable
ALARA As low as reasonably achievable
ANL Argonne National Laboratory
BEAR Biological Effects of Atomic Radiation
BEIR Biological Effects of Ionizing Radiation
BID Background Information Document
BRC Below Regulatory Concern
BWR Boiling water reactor
CC Concrete canister
CFR Code of Federal Regulations
CPG critical population group
CRCPD Conference of Radiation Control Program Directors
CW Consumer waste
DGD Deep geologic disposal
DOD U.S. Department of Defense
DOE U.S. Department of Energy
DOT U.S. Department of Transportation
DREF Dose rate effectiveness factor
DWI Deep-well injection
EIS Environmental Impact statement
EMCB Earth-mounded concrete bunker
EPA U.S. Environmental Protection Agency
ERDA U.S. Energy Research and Development Administration
FRC Federal Radiation Council
GI Gastrointestinal
GW(e) Gigawatts of electric power
HANF Hanford, Washington
HECF .Health effects conversion factors
.HEW U.S. Department of Health, Education, and Welfare
HF Hydrofracture
HIC High-integrity container
HLW High-level radioactive waste
IAEA international Atomic Energy Agency
ICRP International Commission on Radiological Protection
IDD intermediate-depth disposal
INEL Idaho National Engineering Laboratory
ISD Improved shallow-land disposal
L Pulmonary lymph
LANL Los Alamos National Laboratory
LET Linear energy transfer
LLI Lower large intestine
A-3
-------
LLRWPA Low-Level Radioactive Waste Policy Act
LLW Low-level radioactive waste
LQ Linear quadratic
LWR Light-water reactor
MD Municipal dump
MIRD Medical Internal Radiation Dose
MTHM Metric tons of heavy metal
NARM Naturally occurring and accelerator-produced radioactive materials
NAS National Academy of Sciences
NCHS National Center for Health Statistics
NCRP National Council on Radiation Protection and Measurements
N-P Naso-pharyngeal
NRG U.S. Nuclear Regulatory Commission
NRPB National Radiological Protection Board
NTS Nevada Test Site
OMB Office of Management and Budget
ORNL Oak Ridge National Laboratory
ORP EPA's Office of Radiation Programs
P Pulmonary
PWR Pressurized water reactor
RBE Relative biological effectiveness
RFP Rocky Flats Plant
S Stomach
SAB Science Advisory Board
SF Suburban sanitary landfill
SI Suburban sanitary landfill with incineration
SI Small intestine (Chapter 6 only)
SLD Shallow-land disposal
SLF Regulated sanitary landfill
SRP Savannah River Plant
SS Source and special nuclear material
T-B Tracheo-bronchial
TRU Transuranic
TSCA Toxic Substances Control Act
UF Urban sanitary landfill
UI Urban sanitary landfill with incineration
UIC Underground injection control
ULI Upper large intestine
UNSCEAR United Nations Scientific Committee on the Effects of Atomic
Radiation
USGS U.S. Geological Survey
WL Working level
WLM Working level month
A-4
-------
A.2 Metric-to-English Conversion Factors
Efforts have been made to use metric units of measure throughout this
volume of the EIS. Meteorological data and calculations are examples of
subject areas commonly reported in the metric system. To assist the reader in
converting from metric values to the more familiar English values, the
., following conversion table is provided.
To Convert from
Centimeters (cm)
Centimeters (cm)
Cubic centimeters (cm3)
Cubic meters (m3)
Degrees Centigrade (°C)
Grams (g)
Grams (g)
Hectare (ha)
Kilograms (kg)
Kilometers (km)
Liter (L)
Liter (L)
Meter (m)
Meters per second (m/s)
Milligrams (mg)
Milliliters (mL)
Millimeter (mm)
Square meter
Tonne (t)
To
inches (in)
Feet (ft)
Cubic feet (ft3)
Cubic feet (ft3)
Degrees Fahrenheit (°F)
ounces (oz)
Pounds (Ib)
Acres
Pounds (Ib)
Miles (mi)
Cubic feet (ft3)
Gallons (gal)
Feet (ft)
Miles per hour (mi/h)
Ounces (oz)
Ounces (oz)
Inches (in)
Square feet (ft2)
Kilograms (kg)
Multiply by
0.394
0.0328
0.0000353
35.314
*
0.0353
0.00220
2.471
2.204
0.621
0.0353
0.264
3.281
2.237
0.000035
0.0338
0.0394
10.764
1,000
*°F = (°C x 9/5) + 32
A-5
-------
A.3 Scientific Notation
The conventional notation, when dealing with very large or very small
numbers, is awkward and cumbersome. Writing 0.000000000000001, for example
is undesirable, as is calling this number "a millionth of a billionth "
Another system would indicate the above number as 1 x 10~15. This notation
then can be converted back to the original number by moving the decimal point
according to the power of ten. If the power of ten is positive, for example,
the decimal is moved right the number of places indicated by the power if
the power of ten is negative, the decimal is moved left the number of places
indicated by the power. An example of a positive and negative power of ten
follows:
1.25 x 105 = 125000
1.25 x 10~4 = 0.000125
The notation system used in this volume of the EIS utilizes a value
followed by the letter E. After the E is another number, which represents a
power of ten. The number 1.055E+03, for example, is 1.055 x 103. The
number 1.08E-08 is identical to 1.08 x 10~8.
Prefixes are often added to units (such as curies or grams) to indicate
the magnitude of the value. Prefixes used in this statement, their values,
and their abbreviations are as follows:
Prefix
giga
mega
kilo
centi
railli
micro
nano
pico
femto
Value
1,000,000,000
1,000,000
1,000
0.01
0.001
0.000001
0.000000001
0.000000000001
0.000000000000001
Symbol
G
M
k
c
m
v
n
P
f
Thus,
and
1 kilogram (kg) = 10 grams = 1,000 grams
--6
1 microcurie (pci) = 10 curie = 0.000001 curie.
A-6
-------
A.4 Glossary
activation product: An element made radioactive by bombardment with neutrons,.
pro'tons, or other nuclear particles.
alpha particle:
aquifer:
arid site:
barrier: (natural
or engineered)
beta particle:
biointrusion
barriers:
biosphere:
buffer zone:
compaction:
Positively charged particle emitted by certain radioactive
materials. It is made up of two neutrons and two protons,
identical to the nucleus of a helium atom. It is the
least penetrating type of ionizing radiation.
A water-bearing formation below the surface of the earth
that can furnish an appreciable supply of water for a well
or spring.
A term often applied to a shallow-land waste disposal site
located in an area that receives very little annual
precipitation, typically less than 25 cm/yr. In these
sites there is little potential for radionuclide transport
by rainwater moving downward through the soil.
A material object or substance that delays or prevents
migration of water and/or radionuclides into the general
environment.
A subatomic particle emitted from a nucleus during
radioactive decay with a single electrical charge. A
negatively charged beta particle is identical to an
electron. A positively charged beta particle is called a
positron.
An engineered barrier designed to prevent plant roots or
burrowing animals from coming into contact with buried
waste, and thereby prevent transport of radionuclides by
these vectors.
That portion of the Earth's environment inhabited by any
living organisms. It comprises parts of the atmosphere,
the hydrosphere (ocean, seas, inland waters, and
subterranean waters), and the lithosphere.
An area surrounding a nuclear facility (e.g., a waste
disposal site) established to provide an isolation area
between the facility and places used by or accessible to
the public.
The reduction in bulk volume of a material; hence, an
increase in its density (weight per unit volume), by
application of external pressure. Often it is an
economical way to aid in the safe handling of low-level
solid wastes.
A-7
-------
conditioning of
waste:
containment:
contamination,
radioactive:
controlled area:
criteria:
critical organ:
critical pathway:
Those operations that transform waste into a form suitable
for transport and/or storage and/or disposal. The
operations may include converting the waste to another
form, enclosing the waste in containers, and providing
additional packaging.
The confinement of radioactive material in such a way that
it is prevented from being dispersed into the environment
or is released only at a specified rate.
The presence of a radioactive substance or substances in or
on a material or in a place where they are undesirable or
could be harmful.
an area- into which access is limited and personnel are
subject to appropriate controls (such as individual
assessment of dose and special health supervision).
Principles or standards on which a decision or judgment
can be based. They may be qualitative or quantitative.
The most exposed human organ or tissue or the organ of
interest in an analysis, whichever is appropriate.
The dominant environmental pathway through which a given
radionuclide reaches humans.
critical population For a given radiation source, the members of the public
group (CPG): whose exposure is reasonably homogeneous and is typical of
individuals receiving the highest effective dose
equivalent or organ dose equivalent (whichever is
relevant) from the source.
cumulative
population
health effects:
curie (ci):
daughter:
decay product:
Fatal cancers or serious genetic effects (i.e.,
disorders and traits that cause serious handicap at
some time during lifetime).
A unit rate of radioactive decay; the quantity of any
radionuclide that undergoes 3.7 x 1010 (3.7E+10)
disintegrations/second. Several fractions of the curie
are in common usage, i.e., millicurie, picocurie, etc.
Synonym for radioactive decay product.
A nuclide (daughter) resulting from the radioactive
disintegration of a radionuclide (parent), being formed
either directly or as the result of successive
transformations in a radioactive series. Also called a
daughter. Decay products may be stable or radioactive.
A-8
-------
deep-well
injection:
disposal:
distribution
coefficient:
documentation:
dose, radiat ion:
dose assessment:
dose equivalent:
dosimetry:
effective dose
equivalent:
effective half-
life (t1/2):
electron volt (eV)
engineered
storage:
The discharge of liquid wastes via deep wells into ,
permeable but confined geological formations deep
underground as a means of isolating the wastes from the
human environment.
The permanent isolation of radioactive waste from the
accessible environment whether or not recovery is possible.
The ratio of the concentration of a radionuclide (Ci/g)
adsorbed by a solid to the concentration of the same
radionuclide (Ci/mL) in solution (water) when the liquid
and solid are in contact and the respective radionuclide
concentrations have reached equilibrium.
Written, recorded, or pictorial information describing,
defining, specifying, reporting, or certifying activities,
requirements, procedures, or results.
The amount of energy imparted to matter by ionizing
radiation per unit mass of the matter, usually expressed
as the unit rad, or in SI units, 100 rad = 1 gray (Gy).
An estimate of the radiation dose to an individual or a
population group usually by means of predictive modeling
techniques, sometimes supplemented by the results of
measurements.
A term used to express the effective radiation dose when
modifying factors have been considered; the product of
absorbed dose multiplied by a quality factor multiplied by
a distribution factor. It is expressed numerically in
rems, or in SI units, 100 rems = 1 sievert (Sv).
Quantification of radiation doses to individuals or
populations resulting from specified exposures.
The sum of risk-weighted dose equivalents to a specified
set of organs, jiormalized to the risk to the whole body. ,
The time required for one-half of a radioactive material
originally present in the body to be removed by biological
clearance and radioactive decay.
A unit of energy equivalent to the energy gained by an
electron in passing through a potential difference of one
volt.
A method of radioactive waste storage utilizing sealed
containers placed in any of a variety of structures
especially designed to protect the integrity of containers
from accidents and environmental processes.
A-9
-------
environmental Mathematical descriptions of the movement of radionuclides
transfer models: through the environment to an end point (usually to man).
evapotranspiration: The sum total of water lost from the land by evaporation
and plant transpiration.
fissile:
fission:
fission products:
fuel cycle:
gamma ray:
general
environment:
geometric mean:
Any nucleus capable of undergoing fission due to
interaction with neutrons.
The splitting of a heavy nucleus into approximately equal
parts (which are nuclei of lighter elements), accompanied
by the release of a relatively large amount of energy.
Fission can occur spontaneously, but usually is caused by
nuclear absorption of gamma rays, neutrons, or other
particles.
The nuclides resulting from the fission of heavy elements.
The series of steps involved in supplying fuel for nuclear
power reactors. It includes mining, refining, the
original fabrication of fuel elements, their use in a
reactor, chemical processing to recover the fissionable
material remaining in the spent fuel, re-enrichment of the
fuel material, and refabrication into new fuel elements.
High-energy, short-wavelength electromagnetic radiation
emitted from the nucleus of a decaying radionuclide.
Gamma radiation frequently accompanies alpha and beta
emissions and always accompanies fission. Gamma rays are
very penetrating and are most effectively stopped by dense
materials.
The total terrestrial, atmospheric, and aquatic
environments outside sites within which any activity,
operation, or process under the authority of the Atomic
Energy Act of 1954, as amended, is conducted.
The Nth root of the product of a set of N positive
numbers; equivalently, the exponential of the arithmetic
mean of their logarithms.
geometric standard The exponential of the standard deviation of the
deviation: logarithms of a set of positive numbers.
geosphere:
ground water:
The solid portion of the earth, synonymous with the
lithosphere.
Subsurface water within a zone of saturation.
A-10
-------
ground-water
transport:
health impacts:
heavy metal:
high-level radio-
active waste:
humid site:
hydraulic
conductivity:
hydrofracture
process:
hydrogeology:
hydrology:
immobilization
of waste:
The principal means by which radionuclides can be
mobilized from an underground repository and moved into
the biosphere. Avoiding such transport is the basis for
selecting and designing disposal systems.
For the purpose of this analysis, health impacts consist
of cumulative population health effects and maximum CPG
risk.
All uranium, plutonium, or thorium placed into a nuclear
reactor.
Waste whose radioactivity is predominantly characterized
by high-energy radiation; consists of the by-products of
nuclear reactors and wastes generated by spent fuel
processing operations of the nuclear fuel cycle. These
are highly radioactive materials resulting from the
reprocessing of spent nuclear fuel, including liquid waste
produced directly in reprocessing and any solid material
derived from such liquid waste.
An area from which annual precipitation exceeds water loss
by evaporation; hence, there is a significant downward
flux of moisture through the soil which could transport
radionuc1ides.
Ratio of flow velocity to the gradient of driving force
for viscous flow under saturated conditions of a specified
liquid in a porous medium.
A process for permanent disposal of radioactive liquid
waste in which wastes in the form of a slurry containing
hydraulic binders (grouts) are injected by means of
fracturing into a deep underground formation (such as a
nearly impermeable shale formation) considered to be
isolated from the surface. The slurry solidifies in situ,
ensuring fixation of the waste.
The study of the geological factors relating to the
Earth's water.
The study of all waters in and upon the Earth. It
includes underground water, surface water, and rainfall,
and embraces the concept of the hydrological cycle.
Conversion of a waste to a solid form that reduces the
potential for migration or dispersion of radionuclides by
natural processes during storage, transport, and disposal.
A-ll
-------
incineration:
incinerator ash:
ingest:
The process of burning a combustible material to reduce
its volume and yield an ash residue.
The' residue remaining after burning waste in a specially
designed unit. The volume of radioactive ash will be much
less than that of the original waste, and the ash will
usually be incorporated into a solid matrix for disposal.
Take into the body by way of the digestive tract.
ionizing radiation: Any electromagnetic or particulate radiation capable of
producing ions, directly or indirectly, in its passage
through matter.
irregularly
inherited
disorders:
isotope:
light-water
reactor (LWR):
linear energy
transfer (LET)
lognormal
distribution:
management and
storage:
maximum CPG risk:
member of the
public:
Genetic conditions with complex causes, constitutional and
degenerative diseases, etc.
One of two or more atoms with the same atomic number (the
same chemical element) but with different atomic weighrs.
Isotopes usually have very nearly the same chemical
properties, but some have somewhat different physical
properties.
A nuclear reactor whose heat removal system is based on
the use of ordinary water as the moderator and reactor
coolant.
The rate at which charged particles transfer their energy
to the atoms in a medium; expressed as energy lost per
distance traveled in the medium.
A normal distribution (i.e., bell-shaped, symmetrical,
and of infinite extent) of the logarithms of a set of
numbers.
All activities, operations, or processes, administrative
and operational, except for transportation, conducted to
prepare radioactive wastes for storage or disposal, the
storage of any of these materials, or activities
associated with disposal of these wastes.
The probability of contracting a fatal cancer or serious
genetic effect, and which is based on the maximum annual
effective whole-body dose equivalents to the CPG.
Any individual who is not engaged in operations involving
the management, storage, and disposal of materials
regulated by these standards. A worker so engaged is a
member of the public except when on duty at a site
regulated by these standards.
A-12
-------
monitoring:.
neutron:
nonstochastic
effect:
The methodology and practice of measuring levels of
radioactivity either in environmental samples or en route
to the environment. Examples include ground-water
monitoring, gaseous effluent (stack) monitoring, and
personnel monitoring.
An uncharged elementary particle with a mass slightly
greater than that of a proton, and found in the nucleus of
every atom heavier than hydrogen. Neutrons sustain the
fission chain reaction in a nuclear reactor. Bombardment
of materials by neutrons can cause them to become
radioactive.
Those health effects that increase in severity with
increasing dose and usually have a threshold.
operational period: The period during which a nuclear facility is being used
for its intended purpose until it is shut down and
decommissioned.
pathways model:
rad (radiation
absorbed dose)
radioactive decay:
radioactive waste;
radioactivity:
radionuclide:
relative
biological
effectiveness
(RBE):
A mathematical description, usually in the form of a
computer algorithm, that allows estimation of the
magnitude and direction of possible radionuclide transport
vectors.
A measure of the energy imparted to matter by ionizing
radiation; defined as 100 ergs/g. A millirad (mrad) is
1E-03 of a rad. In SI units, 100 rad = 1 gray (Gy).
A process whereby an atom emits particles or excess
energy. This emission is referred to as radioactivity.
The energy is usually in the form of alpha or beta
particles, gamma or X rays, or neutrons.
Any material that contains or is contaminated with
radionuclides at concentrations or radioactivity levels
requiring regulation by the competent authorities and for
which no use is foreseen.
The property of certain nuclides of spontaneously emitting
alpha or beta particles or gamma or X-radiation, or of
undergoing spontaneous fission.
A radioactive nuclide.
The ratio of the dose (rad) of high-LET radiation to the
dose of low-LET radiation, which expresses the
effectiveness of high-LET compared to low-LET radiation in
causing the same biological endpoint.
A-13
-------
rera (roentgen
equivalent man):
retrievability:
risk:
risk analysis:
risk projection:
roentgen (R):
saturated zone:
sensitivity
analysis:
sha1low-ground
disposal:
A measure of dose equivalence for the biological effect
of radiations of different types and energies on man
compared to the effect of X rays, in Si units,
100 rem = 1 sievert (Sv).
The capability to remove waste from where it has been
stored.
For the purposes of radiation protection, the probability
that a given individual will incur any given deleterious
effect as a result of radiation exposure.
An analysis of the risks associated with a technology
wherein the possible events and their probabilities of
occurrence are considered, together with their potential
consequences, the distribution of these consequences
within the affected population(s), the time factor, and
the uncertainties of these estimates.
Absolute - risk projection based on the assumption that
the excess risk from radiation exposure adds to the
underlying (base-line) risk by a constant increment
dependent on dose but independent of the base-line risk.
Relative - risk projection based on the assumption that
the excess risk from radiation exposure is proportional to
the base-line risk.
A unit of measurement of exposure to gamma or x rays in
air, equivalent to an absorbed dose in tissue of
approximately 0.9 rad. The milliroentgen (mR) is 1E-03 of
a roentgen.
A subsurface zone in which all the interstices are filled
with water under pressure greater than that of the
atmosphere. , This zone is separated from the unsaturated
zone, i.e., zone of aeration, by the water table.
An analysis of the variation of the solution of a problem
with changes in the values of the variables involved. For
example, in simple parameter variation, the sensitivity of
the solution is investigated for changes in one or more
input parameters within a reasonable range about selected
reference or mean values.
Disposal of radioactive waste, with or without engineered
barriers, above or below the ground surface, where the
final protective covering is of the order of a few meters
thick.
A-14
-------
solidified waste,
radioactive:
Liquid waste or otherwise mobile waste materials (ion
exchange resins, etc.) that have been immobilized by
incorporation (either physical or chemical) into a solid
matrix by some specific treatment.
spent nuclear fuel: Any nuclear fuel removed from a nuclear reactor after it
has been irradiated and whose constituent elements have
not been separated by reprocessing.
stochastic effect:
storage:
surface water:
target:
teratogenesis:
transmissivity,
hydraulic:
transuranic waste:
unsaturated flow:
unsaturated zone:
volume reduction:
A health effect for which the probability of occurrence is
a function of the dose received, but for which the
severity of the effect is independent of the dose received.
Placement of radioactive wastes with planned capability to
readily retrieve such materials.
Water that fails to penetrate into the sub-soil and flows
or gathers on the surface of the ground.
Material subjected to particle bombardment or irradiation
in order to induce a nuclear reaction.
Production of congenital abnormalities or defects by
irradiation of the fetus.
Rate at which water is transmitted through a unit width
of aquifer under a unit hydraulic gradient. It is
expressed as the product of the hydraulic conductivity and
the thickness of the saturated portion of the aquifer.
Waste containing more than 100 nanocuries of
alpha-emitting transuranic isotopes, with half-lives
greater than 20 years, per gram of waste.
The flow of water in undersaturated soil by capillary
action and gravity.
A subsurface zone in which at least some interstices
contain air or water vapor, rather than liquid water.
Also referred to as "zone of aeration." (See: saturated
zone.)
A treatment that decreases the physical volume of a
waste. Volume reduction is used to facilitate subsequent
handling, storage, transportation, or disposal of the
waste. Typical treatments are mechanical compaction,
incineration, or evaporation. Volume reduction results in
a corresponding increase in radionuclide concentration.
A-15
-------
working level (WL): Any combination of short-lived radon daughters (through
Po-214) per liter of air that will result in the emission
of 1.3E+05 MeV of alpha energy. An activity concentration
of 100 picocuries per liter of Rn-222 in equilibrium with
its daughters, corresponds approximately to one WL. A
working level month (WLM) is an exposure to a
concentration of one WL for 170 hours (about 21 work days).
X ray:
Penetrating electromagnetic radiation whose wavelengths
are shorter than those of visible light, in nuclear
reactions, it is customary to refer to photons originating
in the nucleus as gamma rays, and those originating in the
extranuclear part of the atom as X rays.
A-16
-------
APPENDIX B
NRC LOW-LEVEL RADIOACTIVE WASTE CLASSIFICATION SYSTEM
B-l
-------
APPENDIX B: NEC LOW-LEVEL RADIOACTIVE WASTE CLASSIFICATION SYSTEM
The following is the U.S. Nuclear Regulatory Commission's waste
classification as set forth in Title 10, Code of Federal Regulations,
Part 61.55.
S61.55 Waste Classification
(a) Classification of waste for near surface disposal.
(1) Considerations. Determination of the classification of
radioactive waste involves two considerations. First, consideration must
be given to the concentration of long-lived radionuclides (and their
short-lived precursors) whose potential hazard will persist long after
such precautions as institutional controls, improved waste form, and
deeper disposal have ceased to be effective. These precautions delay the
time when long-lived radionuclides could cause exposures. In addition,
the magnitude of the potential dose is limited by the concentration and
availability of the radionuclide at the time of exposure. Second,
consideration must be given to the concentration of shorter-lived
radionuclides for which requirements on institutional controls, waste
form, and disposal methods are effective.
(2) Classes of waste.
(i) Class A waste is waste that is usually segregated from other
waste classes at the disposal site. The physical form and
characteristics of Class A waste must meet the minimum requirements set
forth in §61.56(a). if class A waste also meets the stability
requirements set forth in §61.56(b), it is not necessary to segregate the
waste for disposal.
(ii) Class B waste is waste that must meet more rigorous
requirements on waste form to ensure stability after disposal. The
physical form and characteristics of Class B waste must meet both the
minimum and stability requirements set forth in §61.56.
(iii) Class C waste is waste that not only must meet more rigorous
requirements on waste form to ensure stability but also requires
additional measures at the disposal facility to protect against
inadvertent intrusion. The physical form and characteristics of class C
waste must meet both the minimum and stability requirements set forth in
S61.56.
(iv) Waste that is not generally acceptable for near-surface
disposal is waste for which waste form and disposal methods must be
different, and in general more stringent, than those specified for Class
C waste. In the absence of specific requirements in this part, proposals
for disposal of this waste may be' submitted to the Commission for
approval, pursuant to §61.58 of this part.
B-2
-------
(3) Classification determined by long-lived radionuclides.
radioactive waste contains only radionuclides listed in Table 1,
classification shall be determined as follows:
If
(i) If the concentration does not exceed 0.1 times the value in
Table 1, the waste is Class A.
(ii) If the concentration exceeds 0.1 times the value in Table 1,
but does not exceed the value in Table 1, the waste is Class C.
(iii) If the concentration exceeds the value in Table 1, the waste
is not generally acceptable for near-surface disposal.
(iv) For wastes containing mixtures of radionuclides listed in
Table 1, the total concentration shall be determined by the sum of
fractions rule described in paragraph (a)(7) of this section.
(4) Classification determined by short-lived radionuclides. If
radioactive waste does not contain any of the radionuclides listed in
Table 1, classification shall be determined based on the concentrations
shown in Table 2. However, as specified in paragraph (a)(6) of this
section, if radioactive waste does not contain any nuclides listed in
either Table 1 or Table 2, it is Class A.
(i) If the concentration does not exceed the value in Column 1,
the waste is Class A.
(ii) If the concentration exceeds the value in Column 1, but does
not exceed the value in column 2, the waste is Class B.
(iii) If the concentration exceeds the value in Column 2, but does
not exceed the value in Column 3, the waste is Class C.
(iv) If the concentration exceeds the value in Column 3, the waste
is not generally acceptable for near-surface disposal.
(v) For wastes containing mixtures of the nuclides listed in Table
2, the total concentration shall be determined by the sum of fractions
rule described in paragraph (a)(7) of this section.
(5) Classification determined by both long- and short-lived
radionuclides. If radioactive waste contains a mixture of radionuclides,
some of which are listed in Table 1 and some of which are listed in
Table 2, classification shall be determined as follows:
(i) If the concentration of a nuclide listed in Table 1 does not
exceed 0.1 times the value listed in Table 1, the class shall be that
determined by the concentration of nuclides listed in Table 2.
B-3
-------
Table 1.
Radionuclide
Concentration
curies per
cubic meter
C-14
C-14 in activated metal
Ni-59 in activated metal
Nb-94 in activated metal
Tc-99
1-129
Alpha-emitting transuranic nuclides with
half-life greater than five years
Pu-241
Cm-242
8
80
220
0.2
3
0.08
^100
are nanocunes per gram.
B-4
-------
Table 2.
Radionuclide
Concentration, curies
per cubic meter
Col. 1 Col. 2 Oil. 3
ABC
Total of all nuclides with less than 5-year
half-life
H-3
Co-60
Ni-63
Ni-63 in activated metal
Sr-90
Cs-137
700
40
700
3.5
35
0.04
1
(')
(1)
(1)
70
700
150
44
(')
(1)
(1)
700
7000
7000
4600
hhere are no limits established for these radionuclides in Class B or
C wastes. Practical considerations such as the effects of external
radiation and internal heat generation on transportation, handling, and
disposal will limit the concentrations for these wastes. These wastes
shall be Class B unless the concentrations of other miclides in Table 2
determine the waste to be the Class C independent of these nuclides.
B-5
-------
(ii) If the concentration of a nuclide listed in Table 1 exceeds
0.1 times the value listed in Table 1 but does not exceed the value in
Table 1, the waste shall be Class C, provided the concentration of
nuclldes listed in Table 2 does not exceed the value shown in Column 3
of Table 2.
(6) Classification of wastes with radionuclides other than those
listed in Tables 1 and 2. If radioactive waste does not contain any
nuclides listed in either Table 1 or 2, it is Class A.
(7) The sum of the fractions rule fbr mixtures of radionuclides.
For determining classification for waste that contains a mixture of
radionuclides, it is necessary to determine the sum of fractions by
dividing each nuclide's concentration by the appropriate limit and
adding the resulting values. The appropriate limits must all be taken
from the same column of the same table. The sum of the fractions for
the column must be less than 1.0 if the waste class is to be determined
by that column. Example: A waste contains Sr-90 in a concentration of
50 Ci/ra3 and Cs-137 in a concentration of 22 Ci/m3. Since the
concentrations both exceed the values in column 1, Table 2, they must
be compared to Column 2 values. For Sr-90 fraction, 50/150 = 0.33; for
Cs-137 fraction, 22/44 = 0.5; the sum of the fractions = 0.83. Since
the sum is less than 1.0, the waste is Class B.
(8) Determination of concentrations in wastes. The concentration
of a radionuclide may be determined by indirect methods such as use of
scaling factors which relate the inferred concentration of one
radionuclide to another that is measured, or radionuclide material
accountability, if there is reasonable assurance that the indirect
methods can be correlated with actual measurements. The concentration
of a radionuclide may be averaged over the volume of the waste, or
weight of the waste if the units are expressed as nanocuries per gram.
B-6
-------
REFERENCE
NRC82 U.S. Nuclear Regulatory Commission, Licensing Requirements
for Land Disposal of Radioactive Waste, 10 CFR 61, Federal
Register, 47(248):57446-57482, December 27, 1982.
B-7
-------
-------
APPENDIX C
INPUT PARAMETERS AND PARAMETER VALUES USED
IN THE PRESTO-EPA ANALYSES
C-l
-------
TABLE OF CONTENTS
C.I Input Parameters and Parameter Values Used ia the
PRESTO-EPA Analyses
C-3
C.2 Additional Information for PRESTO-EPA-BRC Parameters . . . C-23
REFERENCES C-25
C-2
-------
APPENDIX C:
INPUT PARAMETERS AND PARAMETER VALUES! USED
IN THE PRESTO-EPA ANALYSES
C.I Input Parameters and Parameter Values
Used in the PRESTO-EPA Analyses
A listing of the input parameters in the PRESTO-EPA codes and the
values used in the LLW analyses are summarized in this appendix. A
complete listing and short description of each of the input: parameters is
contained in Table C-l. This table also contains parameter values for
those parameters whose values remain constant over the various analyses.
These are followed by listings of parameter values that vary by setting,
waste form, disposal method, and radionuclide (Tables C-2 through C-7).
Input parameters for the PATHRAE-EPA code are not included in this
Appendix. For information pertaining to the PATHRAE-EPA input
parameters, see the User's Manual (EPA87f).
Some of the input parameters are complicated factors whose
description is beyond the scope of this summary. Examples are the
resuspension factors (REl, RE2, and RE3) in Table C-2 and the cap
performance factors (NYRl, NYR2, PCT1, and PCT2) in Table C-4. The
various PRESTO methodology and user's manuals (EPA87a through EPA87e)
should be consulted for a complete description. Values for some
parameters are based on various equations. These equations are listed in
Table C-8. Additional information for some important PRESTO-EPA-BRC
input parameters is contained in Section C.2.
Abbreviations are used throughout the data tables.
key identifies these abbreviations:
The following
Key
Setting
Disposal Method
Shallow Options:
Deep Options:
HP - Humid Permeable
AP - Arid Permeable
HI - Humid Impermeable
CS - Conventional Shallow
IS - Improved Shallow (10 CFR 61)
ID - Intermediate Depth
SL - Sanitary Landfill
EM - Earth-Mounded Tumulus
CB - Concrete Bunker
CC - Concrete Canister
HF - Hydrofracture (Solidified Waste Form)
DG - Deep Geologic (Solidified Waste Form)
DI - Deep-Well Injection (Absorbing Waste Form)
C-3
-------
BRC Options:
Waste Form
- Municipal Dump
- Suburban Sanitary Landfill
- Suburban Sanitary Landfill w/lncineration
- Urban Sanitary Landfill
- Urban Sanitary Landfill w/lncineration
- Absorbing Waste
- Activated Metal
- Solidified Waste
- Trash
MD
SF
si
UF
UI
AW
AM
SW
TR
I/S - incinerated/Solidified
ASH - Incinerated
HIC - Emplaced in High Integrity Containers
C-4
-------
Table C-l. -Listing and description of all input parameters and the values for
parameters which remain constant over the analyses
Parameter
Hydrogeologic
SINFL
DTRAQ
SSAT
RESAT
DWELL
GWV
AQTHK
AQDISP
PORA
PORV
PERHV
DWS
VWV
HGRAD
ALV
ALH
BDENV
RAINF
ERODF
SEDELR
PORS
BDENS
STFLOW
ADEPTH
PD
PPN
RUNOFF
SEEP
STPLNG
COVER
CONTROL
EXTENT
Explanation
Parameters
Infiltration rate for non-cap areas (m/yr)
Distance from trench bottom to aquifer (m)
Fraction of saturation (default equal to 0)
Fraction of residual saturation
Distance from trench to well (m)
Ground-water velocity (m/yr)
Aquifer thickness at well (m)
Dispersion angle of pollutant plume (rad)
Aquifer porosity
Sub-trench porosity
Sub-trench permeability (m/yr)
Distance from well to basin stream (m)
Vertical water velocity (m/yr)
Hydraulic gradient (dimensionless)
Dispersivity in confining stratum (m)
Dispersivity in the aquifer (m)
Density of confining stratum (g/cm3)
Rainfall factor (R)
Soil-erodibility factor (tons/acre-R)
Sediment delivery ratio
Porosity of surface soil
Bulk density of soil (g/cm3)
Stream flow rate (nrVyr)
Active depth of soil (m)
Distance from trench to local stream (m)
Total annual precipitation (deep option)
Fraction of precipitation that runs off
Fraction of precip. that becomes deep
infiltration (deep option)
Slope steepness-length factor
Crop management factor
Erosion control practices factor
Cross-slope extent of surface spillage (m)
2
Values used in code
POP
A
C
0
A
A
A
A
A
A
A
A
A
N/A
N/A
N/A
N/A
N/A
A
A
1.0
A
A
. A
0.1
A
N/A
A
N/A
' A
A
A
C
CPG
A
C
0
A
C
A
A
A
A
A
A
N/A
0
0
0
0
0
A
A
1.0
A
A
A
0.1
C
N/A
A
N/A
A
A
A
C
DEEP
A
C
0
C
A
A
A
A
A
C
A
A
C
1.0
C
0.3
C
A
A
1.0
A
A
A
0.1
A
0
A
0
A
A
A
C
BRC
A
C
0
A
1.61E+3
A
A
A
A
A
A
3220
N/A
N/A
N/A
N/A
N/A
A
A
1.0
A
A
A
0.1
50
N/A
A
N/A
A
A
A
0.45
Engineered Parameters
NYR1, NYR2
PCT1
PCT2
Beginning and ending year of cap failure, respectively
Beginning percentage of cap failure at NYR1
Ending percentage of cap failure at NYR2
C
C
C
C
C
C
N/A
N/A
N/A
1,40
0
0.3
C-5
-------
Table C-K Listing and description of all input parameters and the values for
parameters which remain constant over the analyses (continued)
Parameter
Explanation
2
Values used in code
POP
CPG
DEEP
BRC
Engineered Parameters (continued)
TAREA
TDEPTH
OVER
CFT1
DCFT
Waste-Related
PORT
DENCON
RELFAC
IOPVWV
TRAM
SOAH
STAH (I)
ATAH (I)
DECAY (I)
SOL (I)
CON (I)
XK01 (I)
XKD2 (I)
XK03 (I)
XKD4 (I)
Waste facility surface area (m^)
Depth of operating trench (m)
Thickness of trench overburden (m)
Number of years before waste containers fail
Number of years after CFT1 that containers fail fully
Parameters
Porosity of material in trench
Density (mean) of waste materials (g/cm^)
Annual fraction of trench inventory released
Option for forming into waste
(-1 = no farming; 0 = farming)
Amount of each nuclide in trench at t = 0 (Ci)
Amount of surface spillage of each nuclide (Ci)
Amount of each nuclide placed in stream (Ci)
Amount of each nuclide placed in air (Ci)
Radiological decay constant Cyr~b
Radionuclide solubility (g/ml) (if LEOPT = 5)
Global health effects conversion factor (HE/Ci)
Surface Kd for nuclide I (ml/g)
Waste Kd for nuclide I (ml/g)
Vertical zone Kd for nuclide I (ml/g)
Aquifer Kd for nuclide I (ml/g)
C
C
C
B
B
B
B
B
-1
1.0
E
0
0
E
N/A
E
E
E
E
E
C
C
C
C
C
' B
B •
N/A
-1
N/A
N/A
N/A
N/A
E
N/A
N/A
E
E
E
E
C
C
C
B
B
B
B
B
-1
••
1.0
E
0
0
E
N/A
E
E
E
E
E
0.2
6.6
0.6
0
0
B
B
1.0*
-1
1.0
E
0
0
E
N/A
E
E
E
E
E
Atmospheric Pathway Parameters
H
VG
U
VO
XG
HLID
ROUGH
FTWIND
CHIQ
RE1, RE2, RE3
RR
FTHECH
Atmospheric source height (m)
Contaminated soil particles' fall velocity (m/s)
Annual average wind speed (m/s)
Deposition velocity (m/s)
Distance from trench to local population (m)
Height of inversion layer (m)
Hosker's roughness parameter (m)
Fraction of time wind blows toward population
User-specified dispersion coefficient
Resuspension rate equation factors
Resuspension rate from fanning (sec~b
Resuspension rate modifier
1.0
A
A
A
A
300
0.01
A
A
A
0
0
1.0
A
A
A
C
300
0.01
A
0
A
0
0
1.0
A
A
A
A
300
0.01
A
A
A
0
0
1.0
A
A
A
C
300
0.01
C
C
A
C
0.24
*0.25 in AP setting
C-6
-------
Table C-K Listing and description of all input parameters and the values for
parameters which remain constant over the analyses (continued)
Parameter
Explanation
1
Values used in code
POP
CPG
DEEP
BRC
Atmospheric Pathway Parameters (Continued)
IT Stability class formulation (1 = Pasquill-Gifford)
IS Stability category indicator (4 = neutral stability)
FTWND2 Fraction of time wind blows toward population of
interest, during incineration
CHIQ2 Dispersion coefficient during incineration
RHECH Dust resuspension for onsite operations (kg/sec)
RING Waste incineration rate (kg/sec)
POPG Number of equivalent full-time, full-exposure onsite
visitors exposed to gamma
POPDST Number of equivalent full-time, full-exposure
workers and visitors exposed to dust
Source Term Parameters
RELFRAC (1) Release fraction for absorbing waste (CPG)
RELFRAC (2) Release fraction for activated metals (CPG)
RELFRAC (3) Release fraction for the trash (CPG)
RELFRAC (4) Release fraction for solidified waste (CPG)
RELFRAC (5) Release fraction for incin./solid. waste (CPG)
FRTRSH Fraction of waste that is not in water-tight
containers (CPG)
FTRAB Fraction of trash that is absorbing waste (CPG)
SPLAW Spillage fraction for absorbing waste (CPG)
SPLAH Spillage fraction for activated metal (CPG)
SPLTR Spillage fraction for trash (CPG)
SPLSW Spillage fraction for solidified waste (CPG)
SPLIS Spillage fraction for incin./solid. waste (CPG)
CIAW (I) Absorbing waste activity (Ci) (CPG)
CIAM (I) Activated metal activity (Ci) (CPG)
CITR (I) Trash activity (Ci) (CPG)
CISW (I) Solidified waste activity (Ci) (CPG)
CIIS (I) Incinerated/solidified activity (Ci) (CPG)
FVOLAT (I) Fraction of each radionuclide released to the
atmosphere through incineration (BRC)
Biological Pathway Parameters
WWATL Fraction of irrigation supplied by contaminated well
(1 = 1001)
WWATA Fraction of water for livestock supplied by contaminated
well
1
4
N/A
N/A
N/A
N/A
N/A
1
4
N/A
N/A
N/A
N/A
N/A
1
4
N/A
N/A
N/A
N/A
N/A
1
4
C
C
C
C
C
N/A
A
A
N/A
A
A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
0
E
E
E
E
0.555
1.0
C
0
C
0
0
E
E
E
E
E
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
N/A
F
A
A
A
A
C-7
-------
Table CM. Listing and description of all input parameters and the values for
parameters which remain constant over the analyses (continued)
Parameter
Bioloqical
WWATH
SWATL
SWATA
SWATH
Y1, Y2
PP
XAHBWE
TE1, TE2
TH1
TH2
TH3
TH4
TH5
TH6
FP
FS
QFC
QFG
TF1, TF2
TS
ABSH
P14
XRTH
RTGR
FI
WIRATE
QCW. QGW,
QBW
ULEAFY
UPROO
Explanation
Pathway Parameters (Continued)
Fraction of water for humans supplied by contaminated
well
Fraction of irrigation water supplied by contaminated
stream
Fraction of water for livestock supplied by contaminated
stream
Fraction of water for humans supplied by contaminated
stream
Agricultural productivity for pasture grass and
other consumed vegetation, respectively (kg/rn^)
Surface density of farmed soil (kg/m3)
Weathering removal decay constant (h"1)
Period of time pasture grass or vegetables are
exposed to contaminated air while in fields (h)
Delay time between harvest and consumption
of: pasture grass
Stored feed
Leafy vegetables (maximum individual)
Produce (maximum individual)
Leafy vegetables (population)
Produce (population)
Fraction of year animals graze on pasture grass
Fraction of daily feed which is fresh grass
Amount of feed consumed daily by cattle (kg)
Amount of feed consumed daily by goats (kg)
Transport time from animal feed via milk to maximum
individual and general population (h)
Time between slaughter and consumption (h)
Absolute humidity of atmosphere (g/m3)
Fractional equilibrium ratio for C-14
Maximum root depth for onsite farming (m)
Root growth rate constant (yr-1)
Fraction of year that crops are irrigated
Irrigation rate (L/rn^-h)
Amount of water consumed by milk cows, milk goats,
and beef cattle, respectively (L/d)
Human uptake of leafy vegetables (kg/yr)
Human uptake of produce (kg/yr)
2
Values used in code
POP
A
A
A
A
A
240
0.0021
720
1440
A
2160
24
1440
336
336
A
A
50
6
48
96
A
A
1.0
0
0
A
A
60, 8
50
A
A
CPG
A
A
A
A
A
240
0.0021
720
1440
A
2160
24
1440
336
336
A
A
50
6
48
96
A
A
1.0
0
0
A
A
60, 8
50
A
A
DEEP
A
A
A
A
A
240
0.0021
720
1440
A
2160
24
1440
336
336
A
A
50
6
48
96
A
A
1.0
0
0
A
A
60, 8
50
A
A
BRC
A
A
A
A
A
240
0.0021
720
1440
A
2160
24
1440
336
336
A
A
50
6
48
96
A
A
1.0
0
0
A
A
60, 8
50
A
A
C-8
-------
Table C-l. Listing and description of all input parameters and the values for
parameters which remain constant over the analyses (continued)
Parameter
Bioloqical
UCMILK
UGMILK
UMEAT
UWAT
UAIR
POP
RA (I)
RW (I)
BV (I)
BR (I)
FHC (I)
FH6 (I)
FF (I)
Explanation
Pathway Parameters (Continued)
Human uptake of cow milk (L/yr)
Human uptake of goat milk (L/yr)
Human uptake of meat (kg/yr)
Human uptake of drinking water (L/yr)
Inhalation rate (m3/yr)
Local population size
Radionuclide retention fraction for air
Radionuclide retention fraction for irrigation
Radionuclide soil-to-plant uptake factor for
vegetative parts
Radionuclide soil-to-plant uptake factor to grain
Nuclide forage-to-milk transfer factor for cows
Nuclide forage-to-milk transfer factor for goats
Nuclide forage-to-beef transfer factor
Values used in code
POP
A
0
A
A
8035
A
2.0E-1
2.5E-1
E
E
E
E
E
CPG
A
10
A
A
8035
1.0
2.0E-1
2.5E-1
E
E
E
E
E
DEEP
A
0
A
A
8035
A
2.0E-1
2.5E-1
E
E
E
E
E
2
BRC
A
0
A
A
8035
C
2.0E-1
2.5E-1
E
E
E
E
E
Administrative Parameters
MAXYR
NONCLD
IDISP
IPRT1,
IPRT2
IDELT
LINO
IAVG1 ,
IAVG2
IAQSTF
ITWO
CPRO
INTYR
The number of years simulation will run
The number of nuclides (I) used in run
Indication variable for mode of disposal (1 = CPG)
Beginning and ending year of printed summaries
Time step between printed summaries
Option parameter to specify max. indv. or pop. H.E.
for DARTAB (1 = POP, 0 = max. indv.)
Beginning and ending years for averaging nuclide
concentration values
Control parameter for flow from aquifer to basin
(0 = yes, - = no)
Secondary year for which organ dose summary table
will be printed
Fraction of unused local water flowing to basin
(0 = 100%)
Maximum number of years for which detailed output summaries
will be printed, in addition to 1,000-year summary
10,000
40
N/A
0
1000
100
1
0
1000
0
N/A
0
100,
500
1000
40
1
0
1000
100
0
1
1000
'-
-
0
N/A
10,000
40
C
0
10,000
1000
1
1
10,000
—
—
0
1000,
5000
10,000
40
N/A
0
1000
100
1
1
1000
0
N/A
0
100,
500
C-9
-------
Table
-------
Table C-l. Listing and description of all input parameters and the values for
parameters which remain constant over the analyses (continued)
Key for Table C-l
A - Input parameter which varies by setting (see Table C-2).
B - Input parameter which varies by waste form (see Table C-3).
C - Input parameter which varies by disposal method (see Table C-4).
E - Input parameter which varies by radionuclide (see Tables C-5 through C-7).
F - See additional information relating to PRESTO-EPA-BRC, in Section C.2 of this Appendix.
1 - Explanations are in summary form; for a more detailed description of each input parameter, see the
various PRESTO-EPA Methodology and User's Manuals (EPA87a through EPA87e).
2 _ POP - PRESTO-EPA-POP (Population Health Effects Code)
CPG - PRESTO-EPA-CPG (Critical Population Group Dose Analysis Code)
DEEP - PRESTO-EPA-DEEP (Health Effects from Deep Disposal Code)
BRC - PRESTO-EPA-BRC (Health Effects from Unregulated Disposal Code)
3 - Parameters from the INFIL data set (sub-program in PRESTO-EPA). See PRESTO-EPA Methodology Manual
for details (EPA82a,b).
N/A - Not Applicable.
Oil
-------
Table C-2. Input parameters and parameter values which vary by setting
{parameters listed with an "A" in Table C-l)
H
to
Parameter
WWATL
WWATA
WWATH
SWATL
SWATA
SWATH
SINFL
RSAT
DWS^
DWELL W
GWV
AQTHK
AQDISP
PORA
PORVa
PERHV
VG
U
VD
XG^
FTWIND(a)
CHIQ^
RE1
RE2
RE3
RAINF
ERODF
STPLNG
COVER
CONTRL
PORS
BDENS
Explanation
Fraction of irrigation water supplied by contaminated well (1 = 100%)
Fraction of water for livestock supplied by contaminated well
Fraction of water for humans supplied by contaminated well
Fraction of irrigation water supplied by contaminated stream
Fraction of water for livestock supplied by contaminated stream
Fraction of water for humans supplied by contaminated stream
Annual infiltration rate for non-cap portions of site (m/yr)
Fraction of residual saturation
Distance from well to basin stream (m)
Distance from trench to nearest well (m)
Velocity of ground water in aquifer (m/yr)
Thickness of aquifer at well location (m)
Dispersion angle of pollutant plume in aquifer (radians)
. Aquifer porosity
Sub-trench porosity
Sub-trench permeability (m/yr)
Fall velocity of contaminated soil particles due to gravity (m/sec)
Annual average wind speed toward critical population (m/sec)
Deposition velocity (m/sec)
Distance from trench to population of interest (m)
Fraction of time wind blows toward population of interest
(values from.RADE program which reflect weighted population)
User-specified dispersion coefficient
Factors in resuspension rate equation
Factors in resuspension rate equation
Factors in resuspension rate equation
Rainfall factor
Soil-erodibility factor (tons/acre x R) R = RAINF
Slope steepness - length factor
Crop management factor
Erosion control practices factor
Porosity of the surface soil
Bulk density of the soil (g/cm3)
Value
HP
0
0.5
1.0
0
0
0
0.43
0.17
457
457
27.8
30.5
0.3
0.39
0.35
2.2
0.01
2.01
0.01
480 .
0.4458
3.856E-5
1E-6
-0.15
1E-11
250
0.23
0.27
0.30
0.30
0.39
1.6
by setting
AP
1
1
1
0
0
0
0
0.03
30,000
29,000
90
37
0.3
0.40
0.40
63.4
0.027
4.8
0.027
29,000
19.477
5.186E-8
1E-4
-0.15
1E-9
20
0.5
0.26
0.30
0.40
0.30
1.55
HI
0
0
0
0.1
0.1
1.0
3.0E-3
0.31
250
250
0.03
11
0.1
0.25
0.32
0.019
0.01
5.0
0.01
7240
0.288
3.25E-7
1E-6
-0.15
l.OE-10
100
0.19
0.54
0.10
1.0
0.30
1.49
-------
Table C-2. Input parameters and parameter values which vary by setting
(parameters listed with an "A" in Table C-l) (continued)
o
H
w
Parameter
STFLOW
PD(a)
RUNOFF
TH1
Yl
Y2
FP
FS
TS
ABSH
FI
WIRATE
ULEAFY
UPROD
UCHILK
UHEAT
UWAT
TWT
SLOP
XKI
EGSG
EPSP
YGHAX
XDE
XKE
YPI
YGI
Explanation
Annual flow rate of nearest stream (m3/yr)
Distance from trench to nearest stream (m)
Fraction of annual precipitation that runs off annually
Delay time between harvest and consumption of pasture grass (h)
Agricultural productivity for pasture grass (kg/m2)
Agricultural productivity for consumed vegetation (kg/m2)
Fraction of year animals graze on pasture grass
Fraction of animal's feed that is fresh grass during period
animals are in pasture
Length of time between slaughter of animal and consumption
of meat (h)
Absolute humidity of atmosphere (g/m3)
Fraction of year that crops are irrigated
Irrigation rate (L/m2 - h)
Human uptake of leafy vegetables (kg/yr)
Human uptake of produce (kg/yr)
Human uptake of cow milk (L/yr)
Human uptake of meat (kg/yr)
Human uptake of drinking water (L/yr)
Population in local area for first 1,000 years
Trench width (m)
Trench cap slope
Permeability of trench cover (m/h)
Porosity in gravity zone
Porosity in pellicular zone
Trench cap thickness (m)
Equivalent upward diffusivity (m/h)
Equivalent upward hydraulic conductivity (m/h)
Pellicular infiltration capacity (m/h)
Gravity infiltration capacity (m/h)
Value
HP
3.57E-f5
460
0.29
0
0.67
0.65
1.0
0.83
480
9.9
0.40
0.015
14.0
88.5
89.4
62.8
481.9
25
30.5
0.01
0.02
0.25
0.24
1.2
3.5E-4
1.4E-6
0.01
1.2
bv setting
AP
1000
4000
0.005
0
0.04
0.76
0.47
1.0
480
4.4
0.65
0.114
1C C
16.5
94.2
122.7
61.6
467.9
15
12.2
0.0
4.0
0.35
0.03
1.5
2.0E-3
1.0E-4
0.1
1.5
HI
3.65E+8
50
0.56
330
0.336
0.56
OOT
.27
0.10
336
6.4
0.08
0.042
1O Q
la. 9
84.9
112.3
62.1
391.6
4,285
OT O
87.8
0.25
3.6E-5
\J * V 1
0>n
.47
31
.1
8.0E-5
9.0E-5
OAl
.01
3.1
aValues only for POP - see Table C-1.
^ee key on page C-3.
-------
Table C-3. Input parameters and parameter values which vary by waste form
(parameters listed with a "B" in Table C-l)
Parameter
LEOPT3
PORT
DENCON
CFT1b
DCFTb
RELFAC
Explanation Value bv waste fom
AW TR AH SW I/S ASH HIC
Leaching option (1 = total contact, 225552 N/A
2 = immersed fraction, 5 = release
fraction). See PRESTO methodology
manual for more information (EPA85a).
Porosity of material within trench 0.4 0.6 0.5 0.2 0.2 0.35 N/A
Mean density of waste in trench 0.8 0.8 3.5 1.8 2.0 0 89 N/A
(g/cm3)
Number of years before waste container 20 0 0 20 20 0 300
starts to fail
Number of years after CFT1 that all 50 0 0 50 50 0 0
containers have failed
Annual fraction of waste inventory N/A N/A * * * N/A N/A
released
HF DI DG
5 1 5
0.25 0.13 0.295
1.645 2.102 1.70
0 0 20
0 0 50
l.OE-4 0.005 l.OE-6
a Values for POP and BRC only - see Table C-l.
b Values for POP and DEEP only - see Table C-l.
^ee key on page C-3.
N/A - Not Applicable.
*See Table C-8.
-------
Table C-4. Input parameters and parameter values which vary by disposal method
(parameters listed with a "C" in Table C-l)
Parameter
NYRl
NYR2
PCT1
PCT2
Explanation
Beginning year of cap failure
Ending year of cap failure
Fraction of cap failed in year NYRl
Fraction of cap failed in year NYR2
CS
100
300
0
0.2
IS
100
300
0
0.1
ID
100
300
0
0.1
SL
1
40
0
0.3
EH
100
300
0
0.15
Input
CB
100
300
0
0.075
values bv disposal method
CC
100
300
0
0.075
HF
0
0
0
0
DG
0
0
0
0
DI
0
0
0
0
HD
1
40
0
0.3
SF
1
40
0
0.3
SI
1
40
0
0.3
Ul-
1
40
0
0.3
Ul
1
40
0
0.3
- between NYRl and NYR2 a linear
interpolation between PCT1 and PCT2
determines fraction of cap that
has failed
i
{TAREA Total combined facility surface area
TDEPTH Nominal depth of trench in shallow
options, and waste thickness in deep
options (m)
OVER Thickness of -trench overburden (m)
EXTENT Surface length of trench as disposal area
parallel to stream (on a unit volume basis) (m)
DTRAQ Distance from trench bottom to nominal
aquifer depth (m)
PORV Sub-trench porosity
DWELL Distance from trench to well (m)
7.0 12 16 2.6 9.5
2.0 5 10 0.6 2.0
0.45 0.38 0.41 0.71 0.36 0.41
N/A N/A N/A N/A N/A
* * . * * *
0.167
8.0
2.0
0.41
*
N/A
*
0.133
11.5
4.0
0.36
*
N/A
*
2
0.5
300
1.414
261
0.32
N/A
0.33
3.0
300
0.574
176
0.25
N/A
0.0167
60
900
0.129
848
0.20
N/A
0.
6-
0.
0.
*
2
6
6
45
N/A
N/A
0.2
6.6
0.6
0.45
*
N/A
N/A
0.
6.
0.
0.
*
2
6
6
45
N/A
N/A
0.2
6.6
0.6
0.45
*
N/A
N/A
0.2
6.6
0.6
0.45
*
N/A
N/A
*See Table C-8.
-------
Table C-4. Input parameters and parameter values which vary by disposal method
(parameters listed with a "C" in Table C-l) (continued)
Parameter Explanation
PD
CFT1
DCFT
Distance from trench to local stream (m)
Number of years before waste containers fail
Number of years after CFT1 that containers
fail fullM
Input values bv disposal method
CS
*
20
50
IS
*
20
50
ID
*
20
50
SL
*
20
50
EH
*
20
50
CB
*
100
200
CC
*
100
200
HF
N/A
0
0
DG
N/A
20
50
DI
N/A
0
0
HD
N/A
N/A
N/A
SF
N/A
N/A
N/A
SI
N/A
N/A
N/A
UF
N/A
N/A
N/A
UI
N/A
N/A
N/A
XG Distance from trench to local population
- atmospheric (m)
* SPLAW Spillage fraction for absorbing waste
(CPG only)
********** 1.6E+4- 1.8E+4 1.8E+4 2.2E+4 2.2E+4
1E-7 1E-7 1E-7 1E-3 1E-7 1E-7 1E-7 1E-7 N/A N/A N/A N/A N/A N/A N/A
SPUR
POPG
Spillage fraction for trash waste (CPG only)
Number of equivalent full-time, full-exposure
onsite visitors exposed to gamma (BRC only)
1E-7
N/A
1E-7 1E-7
N/A N/A
1E-3 1E-7 1E-7 1E-7 1E-7 N/A N/A N/A N/A N/A N/A N/A
N/A N/A N/A N/A N/A N/A N/A 5E-7 4E-7 2E-6 3E-7 1E-5
POPDST Number of equivalent full-time, full-exposure N/A
workers and visitors exposed to dust (BRC only)
RINC Waste incineration rate (kg/h)(BRC only) N/A
IDISP Indicator variable for the mode of disposal N/A
VWV Vertical water velocity (m/yr) N/A
ALV Dispersivity in confining stratum (m) N/A
N/A N/A N/A N/A N/A N/A N/A N/A N/A 1E-6 1E-6 3E-5 1E-6 1E-5
N/A N/A N/A N/A N/A N/A N/A N/A N/A 0
N/A N/A N/A N/A N/A N/A 3 4 2 N/A
N/A N/A N/A N/A N/A N/A 1.0 5 0.5 N/A
N/A N/A N/A N/A N/A N/A 5 25 40 N/A
0
N/A
N/A
N/A
1E-6
N/A
N/A
N/A
0
N/A
N/A
N/A
1E-6
N/A
N/A
N/A
-------
Table C-4. Input parameters and parameter values which vary by disposal method
(parameters listed with a "C" in Table C-l) (continued)
Input values by disposal method
Parameter
Explanation
CS IS ID SL EH CB CC HF DG DI HD
SF
SI
UF
UI
BDENV Density of confining stratum (g/cm3)
RSAT Fraction of residual saturation
POP Local population size
CHIQ2 Dispersion coefficient during incineration
(BRC only)
FTWIND2 Fraction of time wind blows toward
population during incineration (BRC only)
RMECH Dust resuspension for onsite operations
(BRC only)
RR
Resuspension rate for farming (BRC only)
N/A N/A N/A
N/A N/A N/A
1.75E5 1.75E5 1E6
F F F
N/A
N/A
1E6
F
N/A N/A N/A N/A N/A N/A N/A 2.3 1.85 1.5 N/A
N/A N/A N/A N/A N/A N/A N/A 0.31 0.02 0.1 N/A
N/A N/A N/A N/A N/A N/A N/A N/A N/A N/A 6E4
N/A N/A N/A N/A N/A N/A N/A N/A N/A N/A F
N/A N/A N/A N/A N/A N/A N/A N/A N/A N/A F F F F F
N/A N/A N/A N/A N/A N/A N/A N/A N/A N/A F F F F F
N/A N/A N/A N/A N/A N/A N/A N/A N/A N/A 5E-4 4E-4 4E-4 4E-4 4E-4
*See Table C-8.
ISee keu on pag@ G-3,
N/A - Not Applicable.
F - See additional information relating to PRESTO-EPA-BRC, in Section C.2 of this Appendix.
-------
Table C-=5. Input parameters and parameter values which vary by radionuclide
(parameters listed with an "E" in Table C-l)
Parameter
TRAM(I)
SOAH(I)
STAH(I)
ATAH(I)
DECAY (I)
CON (I)
XKDHI)
XKD2(I)
XKD3U)
XKD4U)
RA(I)
RW(I)
BV(I)
BR(I)
FHC(I)
FHG(I)
FF(I)
Explanation
Amount (Ci) of each radionuclide found in the trench at the
beginning of the simulation. One curie is assumed, with
constant rate of deposit and decay over 20-year operating
life of site
Amount of spillage onto the surface that exists at the
beginning of simulation, as a fraction of TRAM
Amount (Ci) of radioactivity placed into stream at beginning
of simulation
Amount (Ci) of radioactivity placed in air above trench at
beginning of simulation
Radiological decay constant (yr~b
Conversion factor for basin health effects (health effects/
Ci released)
Surface Kd of radionuclide I (ml/g)
Waste Kd of radionuclide I (ml/g)
Vertical zone Kjj of radionuclide I (ml/g)
Aquifer Kd of radionuclide I (ml/g)
Radionuclide retention fraction for air
Radionuclide retention fraction for irrigation
Radionuclide soil-to-plant uptake factors for vegetative
parts (d/kg)
Radionuclide soil-to-plant uptake factors for grain (d/kg)
Radionuclide forage-to-milk transfer factors for cows (d/L)
Radionuclide forage-to-milk transfer factors for goats (d/L)
Radionuclide forage-to-beef transfer factors (d/kg)
Input
value
Table C-6
l.OE-7*
0
0
Table C-6
Table C-7
Table C-7
Table C-7
Table C-7
Table C-7
2.0E-1
2.5E-1
Table C-6
Table C-6
Table C-6
Table C-6
Table C-6
*1.0E-3 for sanitary landfill scenarios.
C-18
-------
Table C~6. Input parameters and parameter values which vary by radionuclide
(parameters listed with an "E" in Table C-l)
Parameter values
Radionuclide
Hydrogen-3
Carbon-14
Manganese-54
Iron-55
Nickel -59
Cobalt-60
Nickel -63
Strontium-90
Niobium-94
Technetium-99
Ruthenium- 106
Antimony-125
Iodine-129
Cesium-134
Cesium-135
Cesium-137
Cerium-144
Europium-154
Radium-226
Uranium-234
Uranium-235
Neptunium- 23 7
Uranium-238
Plutonium-238
Plutonium-239
Plutonium-241
Americium-241
Plutonium-242
Americium~243
Curium-243
Curium-244
TRAH (Ci)1
6.17E-1
1.00
9.52E-2
2.21E-1
1.00
3.76E-1
1.00
8.02E-1
1.00
1.00
1.04E-1
2.26E-1
1.00
1.77E-1
1.00
8.10E-1
9.06E-2
6.84E-1
1.00
1.00
1.00
1.00
1.00
1.00
1.00
6.36E-1
1.00
1.00
1.00
8.20E-1
7.06E-1
SOAH (Ci)2
6.17E-8
l.OOE-7
9.52E-9
2.21E-8
l.OOE-7
3.76E-8
l.OOE-7
8.02E-8
l.OOE-7
l.OOE-7
1.04E-8
2.26E-8
l.OOE-7
1.76E-8
l.OOE-7
8.10E-8
9.06E-9
6.84E-8
l.OOE-7
1 .OOE-7
l.OOE-7
l.OOE-7
l.OOE-7
l.OOE-7
l.OOE-7
6.36E-8
l.OOE-7
l.OOE-7
l.OOE-7
8.20E-8
7.06E-8
DECAY (yr-1)
5.64E-02
1.21E-04
8.09E-01
2.57E-01
8.66E-06
1.32E-01
7.53E-03
2.42E-02
3.47E-05
3.25E-06
6.89E-01
2.50E-01
4.08E-08
3.36E-01
2.30E-07
2.31E-02
8.90E-01
4.33E-02
4.34E-04
2.83E-06
9.85E-10
3.30E-07
1.55E-10
7.90E-03
2.87E-05
5.25E-02
1.51E-03
1 .83E-06
9.40E-05
2. 17E-02
3.94E-02
BV
4.80
5.50
2.50E-1
4. OOE-3
6. OOE-2
2. OOE-2
6. OOE-2
2.50
2. OOE-2
9.50
7.50E-2
2.00E-1
1.00
8. OOE-2
8.00E-2
8. OOE-2
l.OOE-2
2.50E-3
1.50E-2
8.50E-3
8.50E-3
4.30E-3
8.50E-3
4.50E-4
4.50E-4
4.50E-4
5.50E-3
4.50E-4
5.50E-3
8.50E-4
8.50E-4
BR
4.80
5.50
5. OOE-2
1 .OOE-3
6. OOE-2
7. OOE-3
6. OOE-2
2.50E-1
5. OOE-3
1.50
2. OOE-2
3. OOE-2
1.00
3. OOE-2
3. OOE-2
3. OOE-2
4. OOE-3
2.50E-3
1.50E-3
4. OOE-3
4. OOE-3
4.30E-3
4.00E-3
4.50E-5
4.50E-5
4.50E-5
2.50E-4
4.50E-5
2.50E-4
1.50E-5
1.50E-5
FHC
l.OOE-2
1.20E-2
3.50E-4
2.50E-4
1. OOE-3
2.00E-3
1. OOE-3
1.50E-3
2. OOE-2
1 .OOE-2
6. OOE-7
l.OOE-4
l.OOE-2
7. OOE-3
7. OOE-3
7. OOE-3
2.00E-5
4 2.00E-5
4.50E-4
6.00E-4
6.00E-4
5.00E-6
6.00E-4
l.OOE-7
l.OOE-7
l.OOE-7
4. OOE-7
l.OOE-7
4. OOE-7
2.00E-5
2.00E-5
FMG
1.70E-1
l.OOE-1
2.50E-4
1.30E-4
6.70E-3
1. OOE-3
6.70E-3
1.40E-2
2.50E-3
2.50E-2
1.30E-4
1.50E-3
3.00E-1
3.00E-1
3.00E-1
3.00E-1
5.00E-6
2.00E-5
5.00E-6
5.00E-4
5.00E-4
5.00E-6
5.00E-4
1.50E-6
1.50E-6
2.50E-6
0.0
1.50E-6
0.0
0.0
0.0
FF
1.20E-2
3. 10E-2
4.00E-4
2.00E-4
6. OOE-3
2. OOE-2
6. OOE-3
3.00E-4
2.50E-1
8.50E-3
2. OOE-3
1. OOE-3
7. OOE-3
2. OOE-2
2. OOE-2
2. OOE-2
7.50E-4
4.80E-3
2.50E-4
2.00E-4
2.00E-4
5.50E-5
2.00E-4
5. OOE-7
5. OOE-7
5. OOE-7
3.50E-6
5. OOE-7
3.50E-6
3.50E-6
3.50E-6
bne curie of each nuclide is disposed of. These values assume the disposal rate is constant over
the 20-year life of the site, with decay during that period.
2Spillage fraction is l.OE-7 of the value of TRAM (l.OE-3 for sanitary landfill scenarios, 0.2 for
BRC municipal dump scenarios, and 0.1 for BRC sanitary landfill scenarios).
1 C-19
-------
Table C-7.
Input parameters and parameter values which vary by radionuclide and setting
(parameters listed with an "E" in Table C-1)
Radionuclide
At\U I
H-3
C-14
Fe-55
Ni-59
Co-60
Ni-63
Sr-90
Nb-94
Tc-99
Ru-106
Sb-125
-129
^ Cs-134
0 Cs-135
Cs-137
Ba-137m
Eu-154
Ti-208
Po-210
Pb-210
Pb-212
Bi-214
Pb-214
Ra-226
Th-228
Ac-228
Ra-228
Th-232
U-234
U-235
Np-237
U-238
0.01
0.01
6000
150
55
150
150
350
0.5
220
45
3
100
• 100
100
100
4000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
750
750
5
750
Humid Permeable Setting (HP)
XKD2
0.01
0.01
"50
50
50
50
30
70
0.5
70
45
3
100
100
100
100
4000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
750
750
5
750
XK03
0.01
0.01
6000
150
55
150
150
350
0.5
220
45
3
1000
1000
1000
1000
2000
60,000
220
220
60,000
220
220
220
60,000
220
220
50,000
750
750
5
750
XKD4
0.01
0.01
.6000
150
55
150
20
350
0.5
220
45
3
500
500
500
500
4000
60,000
220
220
60,000
220
220
220
60,000
220
220
50,000
750
750
5
750
CON (I) XK01
5.36E-6 0.01
5.39E-3 0.01
3.98E-4 2000
1.95E-5 3000
4.85E-4 5000
4.81E-5 3000
2.70E-4 150
8.41E-2 350
1.44E-4 0.1
6.81E-4 220
4.10E-7 45
5.64E-3 0.1
7.81E-2 5000
8.01E-3 5000
5.35E-2 5000
5.35E-2 5000
.34E-4 4000
.09E-3 60,000
.38E-1 220
.44E-1 220
.55E-3 60,000
.16E-5 220
2.94E-5 220
3.03E-2 220
2.59E-3 60,000
3.27E-4 220
2.40E-2 220
4.56E-3 60,000
1.78E-4 6
2.12E-4 6
2.80E-1 5
2.22E-5 6
Arid Permeable Setting (API i
XKD2
0.01
0.01
50
50
50
50
30
70
0.1
70
45
0.1
2000
2000
2000
2000
2000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
6
6
5
6
XKD3
0.01
0.01
2000
3000
5000
3000
150
350
0.1
220
45
0.1
10,000
10,000
10,000
10,000
4000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
6
6
5
6
XKD4
0.01
0.01
2000
3000
5000
3000
150
350
0.1
220
45
0.1
5000
5000
5000
5000
4000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
6
6
5
6
CON (I) XK01
6.43E-6 0.01
5.39E-3 0.01
3.72E-4 1500
4.49E-4 150
7.07E-2 40
3.07E-4 150
5.97E-3 30
1.67E+0 350
1.46E-1 0.033
8.54E-5 220
4. 10E-7 45
1.17E+0 0.01
7.41E-2 200
2.79E-2 200
1.46E-1 200
1.46E-1 200
1 .04E-1 4300
2.94E+0 60,000
2.19E-1 220
4.33E-1 220
1.56E-1 60,000
1.29E+0 220
2.39E-1 220
7.38E-2 220
7.08E-3 60,000
8.79E-1 220
5.34E-2 220
8.82E-3 60,000
1.19E-3 50
1.57E-1 50
3.63E-1 5
1.07E-3 50
tumid Impermeable Setting (H"
XKD2 XKD3 XK04
0.01
0.01
50
50
40
50
30
70
0.033
70
45
0.01
200
200
200
200
2000
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
50
50
5
50
0.01
0.01
1500
150
40
150
30
350
0.033
220
45
0.01
250
250
250
250
4300
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
50
50
5
50
0.01
0.01
1500
150
40
150
30
350
0.033
220
45
0.01
250
250
250
250
4300
60,000
220
220
60,000
220
220
220
60,000
220
220
60,000
50
50
5
50
CON (I)
4.49E-6
5.38E-3
3.94E-4
1.87E-5
6.80E-4
4.55E-5
2.48E-4
8 55E-2
2.52E-4
5.94E-4
2.49E-5
6.05E-3
7.77E-2
7.97E-3
5.32E-2
5.32E-2
4.55E-4
3.70E-3
1.35E-1
1.36E-1
1.60E-3
1.20E-3
2 47E-4
2 77E-2
2.33E-3
1.11E-3
2 -20E-2
4. 10E-3
1 56E-4
3.25E-4
2 76E-1
1.55E-4
-------
Table C-7. Input parameters and parameter values which vary by radionuclide and setting (parameters listed with an
"E" in Table C-l) (continued)
Himirl Permeable Sett i no (HP)
Radionuclide
Pu-238
Pu-239
Pu-241
Am-241
Pu-242
Am-243
Qn-243
Om-244
XKD1
3500
3500
3500
80,000
3500
80,000
3300
3300
XKD2
700
700
700
80
700
80
700
700
XKD3
3500
3500
3500
80,000
3500
80,000
3300
3300
XKD4
3500
3500
3500
80,000
3500
80,000
3300
3300
CON (I)
3.96E-3
3.73E-2
6.65E-4
4.45E-2
3.54E-2
8.21E-2
3.95E-2
8.81E-3
Arid Permeable Setting (AP)
XKD1
2000
2000
2000
2E+5
2000
2E+5
3300
3300
XKD2
700
700
700
80
700
80
700
700
XKD3
2000
2000
2000
2E+5
2000
2E+5
3300
3300
XKD4
2000
2000
2000
2E+5
2000
2E+5
3300
3300
CON (I)
6.97E-2
7.86E-3
1 . 19E-3
1.42E-1
7.67E-3
1.77E-1
9.03E-2
5.62E-2
Humid Impermeable Setting (HI)
XKD1
1800
1800
1800
4700
1800
4700
3300
3300
XKD2
700
700
700
80
700
80
700
700
XKD3
1800
1800
1800
4700
1800
4700
3300
3300
XKD4
"1800
1800
1800
4700
1800
4700
3300
3300
CON(I)
3.35E-2
3.65E-3
5.79E-4
5.72E-2
3.46E-3
7.66E-2
3.54E-2
2.98E-2
o
to
H
-------
'Table C-8. Equations relating to various input parameters
• RELFAC - Annual fraction of waste inventory released (POP only)
RELFAC s (site factor) x (waste factor) x (disposal factor)
Site factors Waste factor Disposal factor
1.0
1.0
0.25
HI
HP
AP
0.0
0.0
1.0
1.0
1..0
TR
AW
AH
SW
I/S
l.OE-03
4.0E-04
2.0E-04
2.0E-04
4.0E-04
5.0E-05
l.OE-04
SL
CS
IS
ID
EH
CB
CC
• DTRAQ - Distance from trench bottom to nominal aquifer depth (m)
DTRAQ = (site factor) - TDEPTH
Site factors
21.6 HP
87 AP
28 HI
• DWELL - distance from trench to well (m) (CPG only)
A
DWELL =
100
where, A
waste volume (m )
(TDEPTH - OVER)* P.E.
where, P.E. = placement efficiency = 0.5
0.68
0.27
0.80
SL, CS, HF, DG, DI
EH, CC
CB
IS, ID
• PD - distance from trench to local stream (m) (CPG only)
A
PD = PD*
where PD* =
460
4000
100
HP
AP
HI
• XG - distance from trench to local population - atmospheric (m) (CPG
only)
XG = PD + 1
C-22
-------
C.2 Additional Information for PRESTO-EPA-BRC Parameters
The following parameters are important to the BRC risk analysis.
in the*PRESTO-EPA-BRC analysis of collective population exposures,
the effects of waste form (En84) on radionuclide transport are reflected
primarily in the variables' waste porosity (PORT) and waste density
(DENCON).
PORT (unitless)
DENCON (gm/cc)
Trash
0.60
0.80
Ash
0.35
0.89
The release rate of the radionuclides out of the trench, in
individual BRC waste streams, has been simulated in a dynamic release
submodel through the use of distribution coefficients (XKD2). The
submodel assumes that the total mass of radionuclides contaminating the
waste will be in two forms, solids and dissolved solids, after water
infiltrates through the trench. The radionuclides in the solid phase will
remain stationary and those in the dissolved phase will become mobile.
Other radionuclide-specific input data requirements related to waste
form and processing include DECAY (I) and FVOLAT (I). DECAY (I) is the
radiological decay constant (year"1) for each radionuclide I.
For scenarios involving incineration of wastes at the sanitary
landfill or pathological incinerator prior to disposal, the variable
FVOLAT (I) is employed. This variable is defined as the fraction of each
radionuclide released to the atmosphere through the incineration process.
Values for FVOLAT (Oz84) are listed below:
Radionuclide FVOLAT
Hydrogen-3 , 0.90
Carbon-14 0.75
Technetium-99 0.01
Ruthenium-106 0.01
Iodine-129 0.01
For other isotopes, the value of FVOLAT is 0.005 (sanitary landfill
incinerator) or 0.0025 (pathological incinerator)(Oz84).
Incinerator control efficiencies for radionuclides present in
surrogate BRC waste streams are defined by the expression::
Control Efficiency (I) = 1 - FVOLAT (I)
C-23
-------
Those radionuclides that escape through the stack are subjected to
atmospheric dispersion and may reach the local population. The rate of
incineration used in the modeling is continuous over the 20 years of
operation of the SF.
After incineration, the ash and rubble are landfilled, using methods
similar to those described in Section 4.3.2. The major differences from a
computer simulation viewpoint are the porosity (0.35) and density
(0.89 g/cra3) for ash after being placed in the trench.
C-24
-------
REFERENCES
En84 Envirodyne Engineers, Inc., Radiation Exposures and Health Risks
Resulting from Less Restrictive Disposal Alternatives for Very
Low-Level Radioactive Wastes, prepared for U.S. Environmental
Protection Agency, Contract No. 68-02-3178, Work Assignment 20,
1984.
EPA87a U.S. Environmental Protection Agency, in press, PRESTO-EPA-POP: A
Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Code, Volume I, Methodology Manual, RAE-8706-1, Rogers
and Associates Engineering Corporation, Salt Lake City, Utah, 1987.
EPA87b U.S. Environmental Protection Agency, in press, PRESTO-EPA-POP: A
Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Code, Volume II, User's Manual, RAE-8706-2, Rogers and
Associates Engineering Corporation, Salt Lake City, Utah, 1987.
EPA87c U.S. Environmental Protection Agency, in press, PRESTO-EPA-
DEEP: A Low-Level Radioactive Waste Environmental Transport and
Risk Assessment Code, Documentation and User's Manual, RAE-8706-3,
Rogers and Associates Engineering Corporation, Salt Lake City,
Utah, 1987.
EPA87d U.S. Environmental Protection Agency, in press, PRESTO-EPA-CPG: A
Low-Level Radioactive waste Environmental Transport and Risk
Assessment Code, Documentation and User's Manual, RAE-8706-4,
Rogers and Associates Engineering Corporation, Salt Lake City,
Utah, 1987.
EPA87e U.S. Environmental Protection Agency, in press, PRESTO-EPA-BRC: A
Low-Level Radioactive Waste Environmental Transport and Risk
Assessment Code, Documentation and User's Manual, RAE-8706-5,
Rogers and Associates Engineering Corporation, Salt Lake City,
Utah, 1987.
EPA87f U.S. Environmental Protection Agency, in press, PATHRAE-EPA: A
Performance Assessment Code for the Land Disposal of Radioactive
Wastes, Documentation and User's Manual, RAE-8706-6, Rogers and
Associates Engineering Corporation, Salt Lake City, Utah, 1987.
Oz84 Oztunali, 0. I. and G. Roles, De Minimus Waste Impacts Analysis
Methodology, for USNRC, NUREG/CR-3585, prepared by Dames S, Moore
for the U.S. Nuclear Regulatory Commission, February 1984.
C-25
-------
-------
APPENDIX D
HYDROGEOLOGIC/CLIMATIC DESCRIPTIONS FOR SPECIFIC
COMMERCIAL DISPOSAL FACILITIES.
D-l
-------
-------
TABLE OF CONTENTS
Page
D.I Barnwell D~3
D.I.I General Geology D-3
D.I.2 Hydrogeology D-4
D.I.3 Surface Water Hydrology D-7
D.I.4 Climatic Setting D-10
D.I.5 Hydrogeologic Pathways D-ll
D.2 Beatty D~13
D.2.1 General Geology D-13
D.2.2 Hydrogeology D-15
D.2.3 Surface Water Hydrology D-18
D.2.4 Climatic Setting D-18
D.2.5 Hydrogeologic Pathways D-21
D.3 West Valley D-22
D.3.1 General Geology D-22
D.3.2 Hydrogeology. D-26
D.3.3 Surface Water Hydrology D-29
D.3.4 Climatic Setting D-31
D.3.5 Hydrogeologic Pathways D-32
REFERENCES. . , D~33
D-2
-------
-------
APPENDIX D: HYDROGEOLOGIC/CLIMATIC DESCRIPTION FOR
SPECIFIC COMMERCIAL DISPOSAL FACILITIES
The descriptions of general geology, hydrogeology, surface water
hydrology, climatic settings, and potential hydrogeologic pathways for
commercial disposal facilities located at Barnwell, Beatty, and West
Valley are presented in a general, qualitative manner. These sites are
generally representative of conditions in those regions, and much is
already known concerning site-specific conditions.
D.I. Barnwell
The Barnwell Low-Level Radioactive Waste Disposal facility is
located in Barnwell County, South Carolina, approximately 60 km southeast
of Augusta, Georgia, and along the eastern boundary of the Biarnwell
Nuclear Fuel Plant (FB78).
D.I.I General Geology
The Barnwell facility is located in the Southern Atlantic Coastal
Plain Province, approximately 65 km southeast of the Fall Line that
separates the Piedmont Plateau of the southern Appalachians from the
coastal plain sediments. The Barnwell site lies on the Brandywine
Pleistocene Coastal Terrace which has gently rolling topography cut in
Tertiary sedimentary rocks. Figure D-l is a generalized
northwest-southeast cross-section that shows how the Coastal. Plain
deposits lap onto the crystalline rocks of the Appalachian Piedmont.
Cenozoic and Mesozoic sedimentary rocks thin to the northwest and thicken
to the east and southeast toward the Atlantic Ocean (Fell).
Of the geologic formations that occur beneath the Barnwell plant
area, the Triassic red bed sequence in the area was deposited in a fault
basin like those created in the Northern Coastal Plain in the
mid-Atlantic New England area (Well, FB78).
The rocks younger than the Precambrian-Paleozoic sequence are a
varied sequence of clastic sedimentary rocks displaying a vairiety of size
and sorting ranges. Most units contain some amount of clay and silt and
the Eocene rocks have some thin limestone beds present. The "cleanest"
unit is the Tuscaloosa Formation, a quartzose, arkosic sand unit with
intervening beds of kaolinitic clay. Recent to Pliocene deposits consist
of alluvium and gravelly terrace deposits in stream valleys.
D.I.2 Hvdroqeology
Ground water is found in varying amounts and qualities in almost all
sedimentary formations in the area. The water table in the Barnwell area
occurs in the Hawthorne Formation, although this unit is a poor aquifer
and is not generally suitable for even domestic use except where sand or
D-3
-------
X
Q.
UJ
Q
244
122
122
244
366
488
NW
PLANT
BOUNDARY'
- AIKEN
COUNTY
MARINE SEDIMENTS
OF EOCENE AGE
ESTUARINE OR ALLUVIAL OC
SEDIMENTS OF v*1"
MIOCENE AGE
V
PLANT-
^BOUNDARY
AIKEN BARNWELL
COUNTY'COUNTY
SEDIMENTS
LATE CRETACEOUS
244
122
122
244
366
488
DISTANCE, km
Figure D-1. Profile of Geologic Formations Beneath the
Savannah River Plant (Fe77)
-------
gravel channels occur. The water table fluctuates seasonally, but is
approximately 23 m below the surface at Barnwell. The ground-water flow
pattern is generally north to south across the facility, but rises closer
to the surface in the central area, where a large number of burial
trenches have been constructed and filled. It is common to have enhanced
infiltration in disturbed areas like landfills, and this pattern appears
to be true at Barnwell because this same area does not have a
corresponding topographic high. The other sedimentary aquifers beneath
the facility are the Barnwell Formation, the McBean-Congaree aquifer, and
the Tuscaloosa Formation. The Tuscaloosa Formation is the principal
regional aquifer in the area and lies approximately 100 m below the
Barnwell site. The artesian system may yield as much as 7,570 L/min to
municipal and industrial wells (FB78). The McBean and Congaree Formation
consists of an alternating sequence of sands, marls, clays, and
limestone. The sand and limestone beds are water bearing and supply water
for industrial and municipal supplies in the area around Barnwell. These
two formations discharge via springs and seep directly to surface water
drainages such as Lower Three Runs Creek and the Savannah River to the
south. The Eocene Barnwell Formation is a clayey sand to sandy clay and
is not a significant source of water supply but is used for limited,
rural, domestic supplies.
The character and thickness of the unsaturated zone is of particular
interest for a LLW facility because this is the medium through which
leachate from the trenches will flow. At the Barnwell site, the
unsaturated zone has a thickness of approximately 9 to 15 m. Given the
6- to 7-m depth of the trenches, this thickness is great enough to prevent
trench flooding except in an extremely wet year. The material in the
unsaturated zone is composed of sand and clay from the Hawthorn
Formation. The average permeability of these sediments is 8E-05 m/min
(NRC82), which corresponds to a silty sand. The grain size distribution
supports the permeability figures because the sediments are composed of
75 percent sand and 25 percent silt and clay (principally kaolinite). The
distribution coefficients (Kd) for most of the radionuclides at the
Barnwell site are generally lower than those of the montmorillonite and
zeolite-rich western soil (Ne83). An exception, however, is uranium,
which may have greater retention because of less bicarbonate concentration
in the ground water (Wo83).
D.I.3 Surface Water Hydrology
Two major river systems are in the Barnwell area - the Savannah River
to the south and the Salkehatchie River to the northeast. The Barnwell
facility is located on the edge of the Lower Three Runs Creek watershed, a
southerly flowing tributary of the Savannah River, and only 1 to 2 percent
of the area is within the Salkehatchie watershed. There are numerous
swamps on the Brandywine Terrace, many of them in the Carolina bays
geomorphic features. The bays are local, circular depressions with closed
drainage systems that often support swamps or ponds (NRC82). The closest
perennial stream is Mary's Creek, a tributary to Lower Three Runs Creek,
D-5
-------
about 1 km south of the facility (McD84). West of the Barnwell facility
on the Savannah River Plant site, Lower Three Runs Creek is dammed to form
a large lake, Par Pond, the largest impoundment in the area, covering over
110 km2. Flow into Lower Three Runs Creek is controlled by the
discharge system at Par Pond, surface drainage from the Barnwell site
would not impact Par Pond, but would most likely flow south to Mary's
Creek, and from there to Lower Three Runs Creek and the Savannah River. A
summary of USGS data that characterize flow rates and drainage areas for
Lower Three Runs Creek and the Savannah River is presented in Table D-l.
Generally, the sandy soil and surficial material in the Barnwell
Formation aids infiltration and controls surface runoff except during
extreme precipitation events such as hurricanes and thunderstorms. Most
precipitation infiltrates through the unsaturated zone to the water table
and then moves laterally to the surface discharge system. Some
water-holding soils occur in the Carolina bays area, but these soils
account for less than 10 percent of the area at Barnwell.
D.I.4 Climatic Setting
The climate at Barnwell can be characterized as a warm, humid type
with all seasons represented. The main chain of the Appalachians protects
the area from the more severe winters of the Tennessee Valley, but the
humid, semitropical summers of the southeast are not moderated by any
local topography or geographic feature. The most extensive data on
climate of the area have been collected at the nearby SRP and at a Class A
weather station at Augusta, Georgia. The following climatic summary is a
synthesis of information derived from NRC82, Fe77, FB78, CN80, LE71, and
NOAA80.
Climatic features that directly affect the transport of radionuclides
through air, ground water, and surface water pathways include wind speed
and direction, atmospheric stability, mixing depth, temperature, humidity,
rainfall, and solar radiation (sunshine). The general climate at Barnwell
is, except for the hot and humid summers, moderate, winters are mild with
little snow and spring and fall have temperate weather. Severe weather is
not unknown, with tornados, hurricanes, and hailstorms occurring with
regularity. Precipitation is distributed fairly evenly throughout the
year and the average annual humidity is 66 percent.
The average daily temperature ranges from 1 to 33°C, with extremes of
-15°C to +41°C. The average relative humidity ranges from 45 to
92 percent. The average annual rainfall at Barnwell is about 1.2 m/yr.
The propensity for flooding is low at Barnwell, but surface erosion could
result from extremely heavy precipitation events.
The average annual atmospheric mixing depth is 938 m (NRC82). The
prevailing wind at the SRP is from the southwest with a secondary
direction from the northeast (McD84). Data at SRP indicate that wind
speeds less than 2 m/sec occur 15 percent of the time.
D-6
-------
Table D-1. Sunmary of discharge data for Lower Three Runs Creek
and the Savannah River (USGS81)
Lower Three
Runs Creek
below Par
Pond at
Savannah
River
Lower Three Savannah Savannah
Runs Creek River at River at
near Snelling, Augusta, Clyo,
South Carolina Georgia Georgia
Location
Lati tude
Longi tude
Drainage Basin
Area, krn^
Average Stream
Flow, nrVsec
Maximum Stream
Flow, nrVsec
Minimum Stream
Flow, rn^/sec
33°,14',07"
81°,31',60"
90.4
0.93
4.31
0.05
33°, 0',35"
81°,28',50"
153.6
2.71
21
0.45
19,446
292
9,930
18
32°,31',33"
81°,15',45"
25,511
346
7,660
55
D-7
-------
D.I.5 Hydroqeologic Pathways
The host soils at the Barnwell LLW disposal site are moderately
permeable and well drained, with a low natural attenuation as described
earlier. Although measurable water is found in the trench only after
prolonged storm events, samples of this trench water indicate that
oxidizing conditions are present and that microbiological action is taking
place to reduce levels of organic components in the waste. The primary
potential pathway from the waste at Barnwell is leaching and drainage from
the trenches, migration to the water table in the Hawthorn Formation, and
flow down-gradient to Mary's Spring and Mary's Creek. Leached radioactive
material would then be available for human or animal ingestion or plant
uptake through contaminated surface or ground water.
The ground-water velocity (approximately 27.8 m/yr) would give a
travel time of approximately 36 yr for those highly mobile radionuclides
to migrate from the trenches to Mary's Creek after the material had
infiltrated to the aquifer. Assuming placement of a well midway between
the trenches and Mary's Creek, contaminants could reach the well within
approximately 18 yr. Vertical flow downward to the Barnwell Formation and
the McBean-Congaree aquifer would also be possible because the hydrologic
head in this area drives water from the Hawthorn to the McBean-Congaree.
The pathways to humans through this system would be longer in space and in
time. The most likely avenue would be through surface water discharge,
although the possibility exists of flow to an irrigation well drilled to
the McBean-Congaree.
The possibility that either of these scenarios will occur under
current demographic distribution is remote because Mary's creek and Lower
Three Runs Creek are, in this area, on SRP land. As long as this land
remains part of this facility, dilutions and decay will ensure that LLW
leachate will not significantly impact the local population. However,
health effects to the population of communities that use the Savannah
River downstream of the Barnwell disposal facility for human and animal
consumption and/or agricultural irrigation are possible if contaminants
escape the disposal facility and migrate into the Savannah River.
D.2 Beatty
The Beatty Low-Level Radioactive Waste Disposal facility is a 32.4-ha
tract located about 18.4 km southeast of Beatty, Nevada, and midway
between Beatty and Lathrop Wells, Nevada.
D.2.1 General Geology
The Beatty site is on the upper northeast border of the northwest
part of the Amargosa Desert. The Amargosa Desert is a large, northwest
trending valley that is both a topographic and hydrographic province. The
LLW disposal area of the valley is bounded on the northeast by Bare
Mountain and on the southwest by the Grapevine and Funeral Mountains. At
D-8
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the head of the valley to the northwest, the Bullfrog Hills separate the
Amargosa Desert from Sarcobatus Flats.
The Amargosa River is the principal drainage and enters the valley
from Oasis Valley and Amargosa Narrows. The town of Beatty is in Oasis
Valley. The Amargosa River is an intermittent stream that flows southeast
along the desert valley floor immediately west of the site. Water rarely
flows past the site except in floods, and surface-water flaw is seldom
seen less than 16 km from the facility (Cle62). The main tributaries of
the Amargosa River are Beatty Wash, Forty Mile Canyon Wash, and Carson
Slough.
The topography of the disposal site is nearly flat or gently sloping
(10 m vertical/ 1.6 km horizontal) toward the Amargosa River. The
southwest side of Bare Mountain, northeast of the site, is a pediplain
with armored gravels protecting the erosional surface. Caliche zones are
not present in the soil horizon near Beatty, thus allowing direct
infiltration to take place unimpeded by a dense soil horizon. Numerous
intermittent valleys are present along the slope, but only receive
moisture during snowmelt and convective storms. The Beatty site lies
between the Amargosa River channel and a secondary drainage channel, along
the north side of U.S. Highway 95, which carries the drainage from these
intermittent streams to the Amargosa River.
The Amargosa Desert, like most bolson valleys in the Great Basin,
consists of a fault-controlled Pleistocene and Tertiary valley fill
overlying volcanic and sedimentary basement rocks. Bare Mountain, north
and east of the disposal area, is composed mainly of Paleozoic carbonates
and metasediments and various Tertiary volcanics, and rises over 915 m up
from the valley floor to elevations greater than 1,800 m.
The valley floor contains a variety of alluvial materials ranging
from silt through gravels. The soil zone is usually capped by a thin,
low-density soil zone comprised of a large number of air vesicles between
the soil particles, with lag sands and gravels armoring the surface. Soil
moisture is estimated at 6 to 10 percent. Few deep wells are located near
the facility, but analysis of the one deep well at the site, coupled with
regional and geophysical analyses, indicates that between 150 and 180 m of
valley fill may be present on a largely irregular bedrock surface. Many
Great Basin valleys of this type are fault controlled, and a large fault
probably lies along the northeast side of the valley along the Bare
Mountain front.
The valley fill material consists of pebbles, cobbles, and boulders
representing the full range of bedrock units, namely, sandstone,
siltstone, dolomite, limestone, shale, phyllite, schist, quartzite, and
marble. The bedrock beneath the valley appears to be a quartzite similar
to that found on Bare Mountain, as evidenced by samples collected near the
base of the one deep well drilled on the property (Cle62).
D-9
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D.2.2 Hydroqeology
The hydrogeologic area of interest with respect to the low-level
disposal site at Beatty is the 500 feet of valley fill material on which
the site is located. Material filling Great Basin bolson valleys is
largely heterogeneous, having been derived from mudflows, floods, and
other desert sediment transport mechanisms. Permeabilities in such
material are usually estimated by statistical analyses which predict zones
of percent permeable material. Based on such work to the east in the Ash
Meadows flow system, this portion of the Amargosa Desert valley fill can
be expected to have an estimated average transmissivity of approximately
1.9E+05 L/da/ra. Typically, the water table is 91 to 98 m below land
surface, with unsaturated zone moisture contents of 15 to 20 percent of
saturation. A depth to water of 87 m was reported in a well at the Beatty
facility (Wa63). The casing in the well was perforated from 138 to 150 m
and 156 to 175 ra below the surface (Cle62). If the casing was properly
installed and if the water levels are correct, then an artesian head
exists in the deep sediments of the valley fill. The major clay section
of the well, from 111 to 99 m, could serve in part as a confining bed for
this artesian water. Clebsch refers to an upper and lower aquifer and
notes that the water levels in these zones are 6 m apart, with the higher
potential measured in the upper section (Cle62). This would indicate a
potential for downward flow in this portion of the alluvial fill. Both
"aquifers" are beneath the clay layer noted in the well, which would
insulate the water-bearing sections from direct access from recharge.
However, because no other detailed well logs are available nearby, the
extent of the clay layer is unknown. Thus, the aquifer zones may only be
semiconfined locally, a condition commonly found in Great Basin valley
fill material.
The general direction of ground-water flow in the Amargosa Desert
valley fill is from northwest to southeast. The Oasis Valley area and the
Spring Mountains are the major recharge sources to both the valley fill
and bedrock systems in the area. A large recharge source exists in the
Araargosa River north of Beatty, and the flow is channeled through Amargosa
narrows into the valley fill of the Amargosa Desert (Wh79).
Some desert valleys in the Great Basin contain permeable sedimentary
rocks beneath the valleys. The Amargosa Desert has been shown to be the
regional discharge area for a large regional flow system. At present, no
deep bedrock wells are available to assess the bedrock beneath the Beatty
site to determine whether formations there are part of a regional
hydrologic system (Cle62, Ni82). However, a bedrock flow system appears
to be beneath the Araargosa Desert and it is part of the oasis Valley-Forty
Mile Canyon ground-water basin that receives water from the Pahute
Mesa-Timber Mountain area (Ba72). Based on deep drilling data at the NTS,
bedrock units carrying the water could be either Paleozoic carbonates,
such as are found in other regional aquifers in Nevada, or fractured
Tertiary volcanics. The Amargosa Desert is a regional discharge area much
the same as the Ash Meadows area to the southeast (Ba72). However, no
D-10
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major springs in the area have been specifically tied to the Pahute Mesa
flow system. The total discharge from the Amargosa Desert region is
estimated at 3E+07 m3 per year, the majority of which is discharge
through evapotranspiration (Va63).
Based on well tests (Cle62) and a regional gradient of 5.7E-03, it is
estimated that a ground-water velocity in the valley fill is 1.22 m/da, a
rather high value. The tests also indicate transmissivity ranges from
1,500 to 19,000 L/da/m. The closest producing wells are aibout 25 km
east-southeast of the disposal facility, and they produce 1,100 to 3,800
L/min from the valley fill.
D.2.3 Surface Water Hydrology
The Amargosa River does not flow perennially in the airea of the
disposal site, with the closest gauging station at Beatty, 0.16 km below
Amargosa Narrows. Table D-2 shows the USGS statistics for the Amargosa
River near Beatty, Nevada. Stream flow, when recorded, was usually the
result of local, high-intensity storms. These storms can provide a large
discharge as indicated by the maximum recorded discharge of 120 m3/sec.
Other surface streams in the area only flow during storms and
snowmelt. Most surface flow readily infiltrates the ground, and
continuous flow is not usually seen on the Amargosa River except
immediately adjacent to some perennial springs that are a long distance
downstream of Beatty.
D.2.4 Climatic Setting
The climate at Beatty is a warm-to-hot, arid desert climate'
characterized by low humidity and large seasonal temperature
fluctuations. The area generally has low precipitation and high
evaporation. There is a seasonal precipitation variation,, with the winter
months being the wettest. The higher elevations near the Amargosa Desert,
such as Spring Mountains, have much more precipitation than the lowlands.
The average annual precipitation is estimated at 5 to 13 cm/yr (Cle62).
During the year, the winter precipitation brings the most moisture to the
area, with low pressure storm impulses from the Pacific Ocean the most
common event. Every few years, major storms from the Guli: of Mexico
account for especially wet weather in winter and the early spring months.
These storms serve to raise the total annual precipitation significantly
in those years. Summer precipitation is not uniformly disstributed with
respect to area. The localized convective storms that develop during
these months can cause high amounts of precipitation in isolated areas and
leave other locations dry and unaffected. Most of this moisture
originates from the south and southeast (Wa63).
The average monthly temperatures in the Beatty/Lathrop Wells area
(Amargosa Desert) range between 3°C and 29°C. The recorded extremes of
temperatures are -17°C to 46°C in the Beatty/Lathrop Wells area.
D-ll
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Table D-2. Summary of discharge data for Amargosa River near
Beatty, Nevada (USGS68)
Location:' Latitude, 36° 52' 55"
Longitude, 116° 45' 05"
Drainage Base Area: 1217 krn^
Average Stream Flow: None most of the year
Maximum Stream Flow: 120 nrVsec
Minimum Stream Flow: 0
D-12
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According to USGS records, rainfall at Beatty averages 0.117 m/yr.
The combination of high temperatures and low humidity in the Amargosa
Desert means that the Beatty area has a high evaporation rate. Clebsch
reports a conservative estimate of 25 m/yr of evaporation. The highest
evaporation is in the summer months and the lowest in the winter months
(Cle62).
The high evaporation and low rainfall, coupled with the moderate
permeability of the valley fill, indicate that flooding of burial
trenches is not a problem at the Beatty facility. The main
climate-related problem would be erosion from high intensity storms.
D.2.5 Hvdroqeologic Pathways
The remote location suggests very few pathways to humans exist at
the Beatty facility in the short-term period. If water were to leak from
the trenches, the high evaporation rate could retard the water and
radionuclides from migrating downward. During the course of migration,
the radionuclides would also be adsorbed in the clay in the valley fill
sediments. Once in the flow system, the velocity is fairly rapid in the
sediments at Beatty, although the closest water use is miles away and
local populations are small. However, because of the scarcity of water
available for human and animal consumption and agricultural irrigation in
the southwest region, a large percentage of the potentially contaminated
aquifer is thought to be utilized by downstream populations.
The other possibilities of release from Beatty would be through wind
erosion on disturbed trenchland and ground disturbance due to earthquake
activity.
D.3 West Valley
The West Valley Low-Level Radioactive Waste Disposal facility is
located at the West Valley Nuclear Service Center in western New York,
48 km south of Buffalo in Cattaraugus County.
D.3.1 General Geology
The West Valley site location is on the gently sloping flank of a
bedrock ridge. Local elevation is about 419 m above sea lesvel, and local
drainage is north toward Buttermilk Creek, a tributary of Cattaraugus
Creek, which flows into Lake Erie.
The West Valley LLW waste disposal site is located in the Allegheny
Plateau physiographic province. A pre-existing erosional surface was
moderately to deeply dissected during- the Pleistocene era, and a highly
variable thickness of till, outwash, and glacial lake deposits up to
180 m thick was deposited on the area. The waste disposal site is
located in a thick sequence of till gravel and glacial lake deposits
estimated at from 90 to 150 m thick (Pr77, FB78). The area is underlain
D-13
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with a thick, flat-lying sequence of shales, siltstones, limestones, and
sandstones. Except for the glacial deposits, rocks younger than
Pennsylvanian are usually not reported in the area. The Paleozoic
sedimentary sequence may be as much as 2,745 ra thick and rests on
crystalline Precambrian sequence at depths estimated in the 2,450-m
range. The only bedrocks exposed in the West Valley area are the shales
and siltstones exposed on Buttermilk creek Valley. These rocks are in
the Machias Formation of the Upper Devonian Canadaway Group (EPA77).
Little is known of the potential bedrock aquifers in the disposal
site area. Associated bedrock units are poorly productive, and the low
quantities of water observed are from brackish to brine in quality.
Carbonates in the lower Devonian have produced large amounts of water in
wells several hundred meters deep, but the quality of this water is
variable, being fresher near recharge areas well to the west of the West
Valley area (FB78).
The major stratigraphic materials of interest at West Valley are the
complex series of Pleistocene glacial deposits that overlay the bedrock
in the area. The dominant glacial topographic feature is Buttermilk
Creek Valley, which contains from 2 to 170 m of glacial deposits.
Originally, Buttermilk Creek was a deep bedrock valley, but with
successive Pleistocene glaciations, the valley has been filled with a
heterogeneous series of glaciofluvieil material. The principal deposits
found in the area are (EPA77):
• Till—a very fine-grained, compact, dark blue-gray,
heterogeneous mixture of clay and silt, containing minor
amounts of sand and stones;
• Coarse granular deposits—a mixture of sand and pebbles up to
several inches in diameter, which also contains minor amounts
of silt and clay;
• Outwash—coarse, granular deposits of stratified, well-sorted
sands and gravels. Some deposits appear to be thin-bedded
units of both sand and gravel; and
• Lake deposits—fine-grained, thin-bedded sands, silts, and
clays with minor amounts of fine pebble.
The West Valley LLW disposal facility was constructed in the
glacial till (Figure D-2). The till has been leached of calcium
carbonate and is oxidized and weathered up to 5 m below the surface,
which is littered with pebbles, cobbles, and boulders. Unweathered
till extends down to about 27 m below the surface and lies on an
unknown thickness of glacial lake deposits. On average, the till
contains 50 percent clay, 27 percent silt, 10 percent sand, and
13 percent gravel, coarse, granular surficial material found above the
first "tight" till is thin (0-7.6 m thick) and discontinuous, often
isolated and perched by local stream valleys.
D-14
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- 480
o
H
(Jl
LEGEND
DC
$ HOLOCENE GRAVELS
£| BURIAL TILL
GRAVEL
SAND
VARVED CLAYS (RECESSIONAL LACUSTRINE)
OLDER TILL
DEVONIAN BEDROCK
DC
120 240 360
i i i
METERS
Figure D-2. Cross-Section of Glacial Deposits at the
West Valley Disposal Site (FB78)
- 450
- 420
- 390
- 360
- 330
- 300
- 270
- 240
m
m
0
m
m
*
-------
D.3.2 Hydroqeology
The principal ground-water occurrences and movement in the West
Valley area are in the glacial deposits. While some bedrock units can
and are to be utilized as aquifers, use of these units as water sources
depends on their proximity to the surface and the lack of sufficient
surface and shallow ground-water supplies. The presence of relatively
impermeable and low yielding Upper Devonian siltstones and shales
immediately below the LLW disposal facility makes bedrock interaction
with leachate from the facility unlikely. There is presently no
evidence of a hydrologic connection between the till and bedrock,
except possibly in the weathered zone at the top of the bedrock as
described below.
Several water-bearing zones have been noted in the glacial deposits
at West Valley: principally an artesian unit 12 to 15 m below land
surface; an outwash zone from 30 to 38 ra; and a poorly defined,
apparently permeable zone at depths greater than 60 m, which may
possibly be used for individual supply (FB78).
In addition to the confined water-bearing zones, an unconfined
water table unit occurs in the coarse, granular, near-surface
sediments. As with most unsaturated units, the water-table surface
mimics the topography. This unit is the one most directly connected to
evapotranspiration and direct discharge to streams. The
transraisslbility of this material is very low and is estimated to be
approximately l,300L/da/m with a 25 percent porosity (EPA77).
The shallow artesian unit is found between Buttermilk Creek and
Frank's Creek. This condition, however, does not occur in the disposal
site area. The unit is confined by the upper till, and the depth to
water varies between 2 and 6 m below land surface (EPA77). This unit
is of minor importance and is not found directly beneath the LLW
disposal area.
The third water-bearing zone is a confined unit that is present at
the base of the till and comprises the base of the gravelly till and
the top of the fractured and weathered bedrock. Not much is known
about this unit, but it is believed to be sufficiently permeable to
support low-yield (4 to 40 L/min) wells in the area (FB78).
The silty, clayey till in which the LLW trenches are emplaced is
not a water-yielding zone. It is, however, saturated and allows only
extremely slow water movement. Slug tests conducted by the USGS in
observation wells were analyzed by a variety of methods to arrive at
horizontal till permeabilities of 2E-08 to 6E-08 cm/sec. Distortion
and disturbance in the stratification due to loading and clay expansion
do not appear to seriously affect the permeability of the till.
D-16
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The oxidized till has a higher overall permeability (up to
10 times greater) than the unweathered till. The greater permeability
may be in part due to the presence of numerous fractures in the
oxidized till to a depth of about 5 m. All the till material is
anisotropic on a gross scale, with the unweathered till having a
horizontal permeability on the order of 100 times greater than the
vertical permeability. Silty lenses may occur in the till, which, if
they intersect trenches, may conduct water away from the trenches
toward surface drainages. Coarse sand lenses are reported to exist on
some trench walls but not in others (EPA77). Even so, the permeability
is very low, with an estimated 160 to 200 yr being the ground-water
travel time to the nearest surface drainage, Buttermilk Creek, where
springs do exist along with the valley walls. Unweathered till without
sand or gravel lenses is less strongly anisotropic than the weathered
till, with very low vertical and horizontal permeabilities. All
hydraulic gradients observed at the burial site indicate thait the
potential for migration from the trenches under undisturbed conditions
is down and away from the trenches toward surface water drainages or
the deeper glacial deposits (Pr77).
The main source of past leakage and contamination at West Valley
was cap failure and the consequent filling and overflowing of some of
the trenches. Reworking of the covers and pumping operations resulted
in spillage and radionuclide redistribution, causing contamination of
the soil and weathered till zone over most of the site (ClaSl).
D.3.3 Surface Water Hydrology
The West Valley area falls within the Cattaraugus Creek drainage
basin, and the closest tributary to the site is Frank's creek, which is
a tributary to Buttermilk Creek. Several small, marshy areeis and minor
drainages can be found on the northwest side of the facility, but these
will probably not be a factor in any future excursions. Buttermilk
Creek is a tributary to Cattaraugus Creek, which flows westerly about
65 km to Lake Erie. The main drainages flow over glacial deposits that
fill deeply incised bedrock valleys. The data are summarized in
Table D-3.
Seasonal variations of flow at the Gowanda and Buttermilk creek
gauging stations are quite similar, with high flows occurring in fall
and spring and low flows in summer. Flow is somewhat controlled by
surface impoundments, and direct ground-water discharge from deep
water-bearing zones forms a small percentage of surface-water flow.
Precipitation, snowmelt, and soil saturation along the stream beds are
the main sources of surface water in the area.
Streams in the northeast are subjected to periodic flooding, an
important consideration at West Valley because of its past history of
contamination leakage. A review of U.S. Army Corps of Engineers
records indicates that even a 100-yr flood on Buttermilk Creek would
D-17
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Table D-3. Sunmary of USGS discharge data for Buttermilk Creek
and Cattaraugus Creek (USGS81)
Cattaraugus
Creek at
Gowanda,
New York
Buttermilk
Creek near
Springville,
New York
Location: Latitude
Longi tude
Drainage Basin
Area, krn^
Average Stream Flow
rn^/sec
42° 27' SO-
TS0 56' 10"
1118
20.9
42° 28' 21"
78° 39' 54"
76
1.32
Maximum Stream Flow
m?/sec
Minimum Stream Flow
rn^/sec
98
0.17
111
0.06
D-18
-------
not affect the West Valley disposal site (FB78). The disposal site
area is well above local flood plains. Flash flood surface runoff is
not as common in the east as in the west where different soil types and
infiltration rates prevail. Nonetheless, because high intensity
thunderstorms can occur in the Northeast, care should be taken in
grading the disposal facility to prevent localization of surface
drainage in the area around the trenches. Dikes around the trenches
should also be used to control trench overflow in the event of future
cap failure due to a combination of seal cracking and heavy rain. Two
water supply earthen dams are located at the southeastern part of the
Nuclear Service Center. Their overflow is below the highest elevation
at the facility. Therefore, in the event of dam failure during an
extreme precipitation event, flooding would be contained in the
Buttermilk Creek Valley and pose no danger to the disposal facility.
D.3.4 Climatic Setting
The prevailing climate at West Valley is a cool humid type with
approximately 1.-2 ra of annual precipitation, much of which occurs as
snow. The nearby Great Lakes Region and local topography have an
influence on weather .in the area, which is subject to the "la,ke
effect," which can produce up to 380 cm of snowfall per year in western
New York. The average annual temperature ranges from -18°C to 32°C,
with a mean of 7.2°C. Winds occur with an average speed of a;bout
19 km/h from the southwest (NOAA79, Ri78).
The area has high rainfall and a high percentage of overcast and
partly cloudy weather which minimizes direct evaporation (FB78). Pan
evaporation in this area of the United States is 89 to 102 cm/yr.
Vegetation flourishes in this climate, and plant transpiration is
expected to be high most of the year, but may exceed soil moisture
availability during hot, dry periods. The high annual rainfall usually
ensures a low incidence of soil moisture deficiency.
D.3.5 Hvdroqeologic Pathways
The potential major hydrogeological pathway that could occur at
the West Valley site is the overflow of trench water, with subsequent
contamination of the ground surface. The radionuclides that
contaminate the ground surface could potentially be transported into
the local and/or downstream water supplies if communities or
individuals draw their water from surface streams. The abovementioned
trench-water overflow is potentially caused by the extremely low
hydraulic conductivity of the host formation and rainwater infiltration
allowed by excessive trench cap failure.
Sand lenses have been observed underneath some disposal trenches;
however, there is no evidence that these sand lenses are connected to
any ground water. Therefore, the potential radionuclide transport
pathway from the trench to the biosphere through this pathway was not
considered for this analysis.
D-19
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REFERENCES
Ba72 Bateraan, R.L., et al., Development and Management of Ground
Water and Related Environmental Factors in Arid Alluvial and
Carbonate Basins in Southern Nevada, Desert Research Institute,
Center for Water Resources Research, Project Report No. 18,
1972.
CN80 Chera-Nuclear Systems, Inc., Environmental Assessment for
Barnwell Low-Level Radioactive Waste Disposal Facility,
Columbia, S.C., 1980.
ClaSl Clancy, J.J., D.F. Gray, and O.I. oztunali, Data Base for
Radioactive Waste Management; Volume 1, Review of Low-Level
Radioactive Disposal History, U.S. Nuclear Regulatory
Commission, NUREG/CR-1759, November 1981.
Cle62 Clebsch, A. Jr., Geology and Hydrology of a Proposed Site for
Burial of Solid Radioactive Waste Southeast of Beatty, Nye
County, Nevada, in: U.S. Atomic Energy Commission, 1968,
WASH-1143, 1962.
EPA77 U.S. Environmental Protection Agency, Summary Report on the
Low-Level Radioactive Waste Burial Site, West Valley, New York
(1963-1975), EPA-902/4-77-010, 1977.
FB78 Ford, Bacon and Davis Utah, Inc., Compilation of the
Radioactive Waste Disposal Classification System Data Base;
Analysis of the West Valley Site: Task Report for U.S. NRG,
FDU-247-01, Salt Lake City, Utah, 102 pp., and Compilation of
the Radioactive Waste Disposal Classification System Data Base;
Analysis of the Barnwell Site: Task Report for U.S. NRC,
FDU-247-04, Salt Lake City, Utah, 1978.
Fe77 Fenimore, J.W. and R.L. Hooker, The Assessment of Solid
Low-Level Waste Management at the Savannah River Plant:
Savannah River Laboratory Report, DPST-77-300, 1977.
LE71 Law Engineering and Testing Co., Atlanta, Ga. Report on
Geologic and Hydrologic Studies near Snelling, South Carolina,
Study Conducted for Chem-Nuclear Systems, Inc., Letco Job
No. 6605, 1971.
McD84 McDonald, B.B., General Description of the Low-Level
Radioactive Waste Burial Facility near Barnwell, South
Carolina, U.S. Geological Survey, Columbia, S.C., 1984.
Ne83 Neiheisel, J., Prediction Parameters of Radionuclide Retention
at Low-Level Radioactive Waste Sites, U.S. Environmental
Protection Agency, Office of Radiation Programs, EPA
520/1-83-0125, Washington, D.C., 1983.
D-20
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N182
NOAA79
NOAA80
NRC82
Pr77
Nichols, D., U.S. Geological Survey, Personal Communication,
1982.
National Oceanic and Atmospheric Administration, National
Climatological Summary Tables: National Climatic Data Center,
Asheville, N.C., 1979.
National Oceanic and Atmospheric Administration, Local
Climatological Data Annual Services for 1979, Environmental
Data and Information Services, National Climatic Center,
Asheville, N.C., 1980.
U.S. Nuclear Regulatory Commission, Environmental Assessment
for the Barnwell Low-Level Waste Disposal Facility,
NUREG-0879, 1982.
Prudic, D.E. and A.D. Randall, Ground-Water Hydrology and
Subsurface Migration of Radioisotopes at a Low-Level Solid
Radioactive Waste Disposal Site, West Valley, New York, U.S.
Geological Survey, Open File Report 77-566, 1977.
Ri78
USGS68
USGS81
Wa63
Riffner, J.A., ed., Climates of the States,
Detroit, Michigan, 1978.
Gale Research Co.,
Wh79
Wo83
U.S. Geological Survey, Water Resources Data for Nevada,
Part 1, Surface Water Records and Water Resources Data for New
York, Part 1, Surface Water Records, 1968.
U.S. Geological Survey, Water Resources Data for New York and
Water Resources Data for South Carolina, 1981.
Walker, G.E. and T.E. Eakin, Geology and Ground Water of
Amargosa Desert, Nevada-California: Nevada Department of
Conservation and Natural Resources, Ground Water Resources -
Reconnaissance Series Report 14, 1963.
White, A.P., Geochemistry of Ground Water Associated with
Tuffaceous Rocks, Oasis Valley, Nevada, U.S. Geological
Survey, Professional Paper 712-E, 1979.
Wolfsburg, K., et al.. Research and Development Related to
Sorption of Radionuclides on Soils, Los Alamos National
Laboratory, LA-UR-83-800, Los Alamos, New Mexico, 1983.
D-21
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APPENDIX E
A DESCRIPTION OF THE RADRISK AND CAIRO
COMPUTER CODES USED BY EPA TO ASSESS
DOSES AND RISKS FROM RADIATION EXPOSURE
E-l
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APPENDIX E
Page
E.I Introduction E~3
E.2 Life Table Analysis to Estimate the Risk of Excess Cancer. . . . E-3
E.3 Risk Analysis Methodology
References •
E-5
E-8
E-2
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APPENDIX E:
A DESCRIPTION OF THE RADRISK AND CAIRO
COMPUTER CODES USED BY EPA TO ASSESS DOSES
AND RISKS FROM RADIATION EXPOSURE
E.I Introduction
This appendix provides a brief overview of the RADRISK (Du80) and
CAIRO (Co78) computer codes used by the EPA to assess the health risk
from radiation exposures. It describes the mechanics of the life table
implementation of the risk estimates derived in Chapter 7.
E.2 Life Table Analysis to Estimate the Risk of Excess Cancer
Radiation effects can be classified as stochastic or nonstochastic
(NAS80, ICRP77). For stochastic effects, the probability of occurrence
of the effect, as opposed to the severity, is a function of dose;
induction of cancer, for example, is considered a stochastic effect.
Nonstochastic effects are those health effects for which the severity of
the effect is a function of dose; examples of nonstochastic effects
include cell killing, suppression of cell division, cataracts, and
nonmalignant skin damage.
At, the low levels of radiation exposure attributed to radionuclides
in the environment, the principal health detriments are the induction of
cancers (solid tumors and leukemia) and the expression, in later
generations, of genetic effects. In order to estimate these effects,
instantaneous dose rates for each organ at specified times are calculated
for use in a subroutine adaptation of CAIRO contained in the RADRISK
code. This subroutine uses annual doses derived from these dose rates to
estimate the number of incremental fatalities in the cohort due to
radiation-induced cancer in the reference organ. The calculation of
incremental fatalities is based on estimated annual incremental risks,
computed from annual doses to the organ, together with radiation risk
factors such as those given in the 1980 NAS report, BEIR-3 (NAS80).
Derivation of the risk factors in current use is discussed in Chapter 7.
An important feature of this methodology is the use of actuarial
life tables to account for the time dependence of the radiation insult
and to allow for competing risks of death in the estimation of risk due
to radiation exposure. A life table consists of data describing
age-specific mortality rates from all causes of death for a given
population. This information is derived from data obtained on actual
mortality rates in a real population; mortality data for the U.S.
population during the years 1969-1971 are used throughout this study
(HEW75K
The use of life tables in studies of risk due to low-level radiation
exposure is important because of the time delay inherent in radiation
risk. After a radiation dose is received, there is a minimum induction
E-3
-------
period of several years (latency period) before a cancer is clinically
observed. Following the latency period, the probability of occurrence of
a cancer during a given year is assumed to be constant for a specified
period, called a plateau period. The length of both the latency and
plateau periods depends upon the type of cancer.
During or after radiation exposure, a potential cancer victim may
experience years of life in which he/she is continually exposed to risk
of death from causes other than incremental risk from radiation
exposure. Hence, some individuals in the population will die from
competing causes of death, and are not victims of radiation-induced
cancer.
Each member of the hypothetical cohort is assumed to be exposed to a
specified activity of a given radionuclide. In this analysis, each
member of the cohort annually inhales or ingests 1 pCi of the
radionuclide, or is exposed to a constant external concentration of
1 pCi/cra3 in air or 1 pci/cm2 on ground surfaces, since the models
used in RADRISK are linear, these results may be scaled-to evaluate other
exposure conditions. The cohort consists of an initial population of
100,000 persons, all of whom are simultaneously liveborn. In the
scenario employed here, the radiation exposure is assumed to begin at
birth and continue throughout the entire lifetime of each individual.
No member of the cohort lives more than 110 yr. The span from 0 to
110 yr is divided into 9 age intervals, and dose rates to specified
organs at the midpoints of the age intervals are used as estimates of the
annual dose during the age interval. For a given organ, the incremental
probability of death due to radiation-induced cancer is estimated for
each year using radiation risk factors and the calculated doses during
that year and relevant preceding years. The incremental probabilities of
death are used in conjunction with the actuarial life tables to estimate
the incremental number of radiation-induced deaths each year.
The estimation of the number of premature deaths proceeds in the
following manner. At the beginning of each year, m, there is a
probability PN of dying during that year from nonradiological causes,
as calculated from the life table data, and an estimated incremental
probability PR of dying during that year due to radiation-induced
cancer of the given organ. In general, for the m-th year, the
calculations are:
M(ra) - total number of deaths in cohort during year m,
= [PN(m) + PR(m)] x N(m)
Q(m) - incremental number of deaths during year m due to
radiation-induced cancer of a given organ,
= PR(m) x N(m)
E-4
-------
N(m+l) = number of survivors at the beginning of year m + 1
= N(m) - M(m)
PR is assumed to be small relative to PN, an assumption that is
reasonable only for low-level exposures such as those considered here
(Bu81). The total number of incremental deaths for the cohort is then
obtained by summing Q(m) over all organs for 110 yr.
In addition to providing an estimate of the incremental number of
deaths, the life table methodology can be used to estimate the total
number of years of life lost to those dying of radiation-induced cancer,
the average number of years of life lost per incremental mortality, and
the decrease in the population's life expectancy. The total number of
years of life lost to those dying of radiation-induced cancesr is computed
as the difference between the total number of years of life lived by the
cohort assuming no incremental radiation risk, and the total number of
years of life lived by the same cohort assuming the incremental risk from
radiation. The decrease in the population's life expectancy can be
calculated as the total years of life lost divided by the original cohort
size (N(l)=100,000).
Either absolute or relative risk factors can be used. Absolute risk
factors, given in terms of deaths per unit dose, are based on the
assumption that there is some absolute number of deaths in a population
exposed at a given age per unit of dose. Relative risk factors, the
percentage increase in the ambient cancer death rate per unit dose, are
based on the assumption that the annual rate of radiation-induced excess
cancer deaths, due to a specific type of cancer, is proportional to the
ambient rate of occurrence of fatal cancers of that type. Either the
absolute or the relative risk factor is assumed to apply uniformly during
a plateau period, beginning at the end of the latent period,.
The estimates of incremental deaths in the cohort from chronic
exposure are identically those that are obtained if a corresponding
stationary population (i.e., a population in which equal numbers of
persons are born and die in each year) is subjected to an acute radiation
dose of the same magnitude. For example, the total person-years lived by
the 1970 life table cohort is approximately 7.07 million, the estimates
of incremental mortality in the cohort from chronic irradiation also
apply to a 1-year dose of the same magnitude to a population of this
size, age distribution, and age-specific mortality rates. More precise
life table estimates for a specific population can be obtained by
altering the structure of the cohort to reflect the age distribution of a
particular population at risk.
E.3 Risk Analysis Methodology
Risk estimates in current use at EPA are based on the 1980 BEIR-3
report of the National Academy of Sciences Advisory Committee on the
E-5
-------
Biological Effects of Ionizing Radiation (NAS80). The form of these risk
estimates is, to some extent, dictated by practical considerations,
e.g., a desire to limit the number of cases that must be processed for
each environmental analysis and a need to conform to limitations of the
computer codes in use. For example, rather than analyze male and female
populations separately, the risk estimates have been merged for use with
the general population; rather than perform both an absolute and a
relative risk calculation, average values have been used.
The derivation of the risk estimates from the BEIR-3 report is
presented in Chapter 7. A brief outline of the general procedure is
presented below. Tables referenced from Chapter V of NAS80 are
designated by a V prefix.
The total number of premature cancer fatalities from lifetime
exposure to 1 rad per year of low-LET radiation is constrained to be
equal to the relative risk value (403 per million person-rad) given in
Table V-25 of the BEIR-3 report for the L-L and IFE models for leukemia
and solid cancers, respectively (NAS80).
(2) For cancers other than leukemia and bone cancer, the age and
sex-specific incidence estimates given in Table V-14 were multiplied by
the mortality/ incidence ratios of Table V-15 and processed through the
life table code at constant, lifetime dose rates of 1 rad/yr. The
resulting number of deaths are averaged, using the male/female birth
ratio, and proportioned for deaths due to cancer in a specific organ as
described in chapter 7. These proportional risks are then used to
allocate the organ risks among the 358.5 deaths per million person-rad
remaining after the 44.5 leukemia and bone cancer fatalities (Table V-17)
are subtracted from the 403 given in Table V-25.
(3) The RADRISK code calculates dose rates for high- and low-LET
radiations independently. A quality factor of 20 has been applied to all
alpha doses to obtain the organ dose equivalent rates in rem per year
(ICRP77). For high-LET radiation risk estimates, the risk from alpha
particles is considered to be eight times that for low-LET radiation to
the same tissue except for bone cancer, for which the risk coefficient is
20 times the low-LET value. Additional discussion was included in
Chapter 7.
A typical environmental analysis requires that a large number of
radionuclides and multiple exposure models be considered. The RADRISK
code has been used to obtain estimates of cancer risk for unit intakes of
about 200 radionuclides and unit external exposures by approximately 500
radionuclides. The calculated dose rates and mortality coefficients
described in the preceding sections are processed through the life table
subroutine of the RADRISK code to obtain lifetime risk estimates. At the
low levels of contamination normally encountered in the environment, the
life table population is not appreciably perturbed by the excess
radiation deaths calculated and I since both the dose and risk models are
E-6
-------
linear, the unit exposure results may be scaled to reflect excess cancers
due to the radionuclide concentrations predicted in the analysis of a
specific source.
As noted in the discussion of the life table analysis, risk
estimates for chronic irradiation of the cohort may also be applied to a
stationary population having the same age-specific mortality rates as the
1970 U.S. population. That is, since the stationary population is formed
by superposition of all age groups in the cohort, each age group
corresponds to a segment of the stationary population with the total
population equal to the sum of all the age groups. Therefore, the number
of excess fatal cancers calculated for lifetime exposure of the cohort at
a constant dose rate would be numerically equal to that calculated for
the stationary population exposed to an annual dose of the same
magnitude. Thus, the risk estimates may be reported as a lifetime risk
(the cohort interpretation) or as the risk ensuing from an annual
exposure to the stationary population. This equivalence is particularly
useful in analyzing acute population exposures. For example, estimates
for a stationary population exposed to annual doses which vary from year
to year may be obtained by summing the results of a series of cohort
calculations at various annual dose rates.
E-7
-------
REFERENCES
Bu81 Hunger, B.H., Cook, J.R. and M.K. Barrick, Life Table Methodology for
Evaluating Radiation Risk: An Application Based on Occupational
Exposures, Health Physics, 40(4): 439-455.
Co78 Cook, J.R., Bunger, B. and M.K. Barrick, A Computer Code for Cohort
Analysis of Increased Risks of Death (CAIRO), EPA Report 520/4-78-012,
U.S. Environmental Protection Agency, Washington, D.C., 1978.
Du80 Dunning, D.E. Jr., Leggett, R.W. and M.G. Yalcintas, A Combined
Methodology for Estimating Dose Rates and Health Effects from Exposure
to Radioactive Pollutants, Report ORNL/TM-7105, Oak Ridge National
Laboratory, Tennessee, 1980.
HEW75 U.S. Department of Health, Education, and Welfare, 1975, U.S.
Decennial Life Tables for 1969-1971, Vol. 1, No. 1, DHEW Publication
No. (HRA) 75-1150, Public Health Service, Health Resources
Administration, National Center for Health Statistics, Rockville,
Maryland.
ICRP77 International Commission on Radiological Protection, 1977,
Recommendations of the International Commission on Radiological
Protection, Ann. ICRP, Vol. 1, No. 1, Pergamon Press, 1977.
NAS80 National Academy of Sciences - National Research Council, The Effects
on Populations of Exposures to Low Levels of Ionizing Radiation,
Committee on the Biological Effects of Ionizing Radiations (BEIR
Report), Washington, D.C., 1980.
E-8
-------
APPENDIX F
MAXIMUM CPG DOSES FOR BRC WASTE DISPOSAL SCENARIOS
F-l
-------
APPENDIX F: MAXIMUM CPG DOSES FOR BRC WASTE DISPOSAL SCENARIOS
This Appendix details, in the form of tables, the maximum annual
dose to the CPG, the radionuclide providing the major dose, and the year
in which the maximum CPG dose occurs for each of the ten major pathways
(see Table 10-2) at each of the three hydrogeologic/c1imatic settings for
the 15 BRC waste disposal scenarios. Section 4.4 describes the scenarios
in detail.
The CPG doses are calculated over a 10,000-year time span, during
which the maximum individual in any given year may be either an onsite
worker, an onsite resident, or an offsite resident. As explained in
Section 10.7.2, the onsite worker while employed at the BRC waste
disposal facility is also considered a member of the general public. The
onsite resident is a member of the general public living onsite and/or
growing crops for human consumption. The offsite resident is a member of
the general public who lives away from the BRC waste disposal site, but
is subjected to the various pathways capable of transporting
radionuclides to the human population. The pathways can be separated
into onsite and offsite workers and residents as follows (see also
Section 8.5.4):
Onsite Worker Pathways (Pre-closure)
(1) Direct gamma exposure
(2) Dust inhalation
Onsite Resident Pathways (Post-closure)
(1) Food grown onsite
(2) Biointrusion
Offsite Resident Pathways (Post-closure)
(1) Ground water to river
(2) Ground water to well
(3) Surface erosion
(4) Facility overflow or bathtub effect
Offsite Resident Pathways (Pre-closure)
(1) Spillage
(2) Atmospheric inhalation.
Three time periods are involved in the CPG analysis, in all cases
it is assumed that the disposal site has a maximum 20-year inventory of
BRC wastes, with radioactive decay taken into consideration. The 0 year
represents the last year of pre-closure. in the 0 year, the CPG is to
the onsite worker and some offsite residents.
F-2
-------
The year 1 represents the first year of the post-closure phase and
the CPG is to onsite residents. Finally, there are the variable years,
beyond year 1 of the post-closure phase, in which the CPG is to offsite
residents.
The ground-water to river, the bathtub effect, and spillage pathways
are applicable only to the humid impermeable hydrogeologic/climatic
setting. This is because this setting deals solely with surface water
flow, whereas the other two settings use the ground-water migration
pathways.
The erosion pathway for the arid permeable setting does not appear
within the 10,000-year analysis performed.
For the scenarios involving urban demographic disposal settings, it
was assumed that there would be no food grown onsite after site closure.
The four scenarios, Tables F-12 through F-15, were analyzed for
reference purposes only. These four scenarios were not relevant to the
analysis for regulatory considerations and were used for comparison
purposes only.
F-3
-------
*J
Table F-1. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 1. Three-Unit
Pressurized-Water Power Reactor Conplex - Municipal
Dump (PWR-MD)
Site pathways
GW to River
GW to Wei 1
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Humid
Dose
(mrem/yr)
7.5E-07
3.1E-02
2. 1E-03
2.8E-06
2. 1E-04
3.2E-01
l.lE-iOO
1.2E+01
3.2E-02
2. 1E-06
Impermeable
Nuclide
1-129
1-129
Co-60
Pu-239
1-129
Cs-137
Cs-137
Co-60
Am-241
Am-241
Site
Year
3060
3060
0
3900
100
1
1
0
0
0
Humid
Dose
(mrem/yr)
*
1.5E-01
*
3.3E-03
*
1.9E-01
6.4E-01
1.2E+01
3.2E-02
4.4E-06
Permeable Site
Nuclide Year
1-129 184
Pu-239 3900
Cs-137 1
Cs-137 1
Co-60 0
Am-241 0
Am-241 0
Arid
Dose
(mrem/yr)
*
7.3E-04
*
*
*
2.7E-01
9.0E-01
1.2E+01
3.2E-02
5.0E-06
Permeable Site
Nuclide Year
1-129 556
>13000
Cs-137 1
Cs-137 1
Co-60 0
Am-241 0
Am-241 0
*Pathway not applicable.
Note: In all tables, GW means ground water.
-------
I
en
Table F-2. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 2. Two-Unit
Boiling-Water Power Reactor Complex - Municipal
Dump (BWR-HD)
Site .pathways
GW to River
GW to Wei 1
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Humid
Dose
(mrem/yr)
1.7E-06
7.1E-02
3.5E-03
1.8E-06
4.7E-04
8.1E-01
2.7E+00
1.1E+01
1 . 1E-02
1 OC AC
Impermeable
Nuclide
1-129
1-129
Cs-137
1-129
1-129
Cs-137
Cs-137
Co-60
Co-60
\AJ~WW
Site
Year
3060
3060
0
3900
100
1
1
0
0
f\
\s
Humid Permeable Site
Dose Nuclide Year
(mrem/yr)
*
3.3E-01 1-129 184
*
1.9E-03 1-129 3900
*
4.8E-01 Cs-137 1
1.6E+00 Cs-137 1
1.1E+01 Co-60 0
1 . 1E-02 Co-60 0
O QC AC fV* £A A
Arid Permeable
Dose Nuclide
(mrem/yr)
*
1.7E-03 1-129
*
*
*
6.7E-01 Cs-137
2.2E+00 Cs-137
1.1E+01 Co-60
1.1E-02 Co-60
Site
Year
556
>13000
1
1
0
0
n
*Pathway not applicable.
-------
Table F-3. Haxiroura annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 3. University
and Medical Center Complex - Urban Sanitary Landfill
(LUHC-UF)
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Humid
Dose
(mrem/yr)
7.3E-07
1.5E-03
5.4E-04
1.1E-06
6.3E-04
*
*
1.8E-01
1.6E-04
1.1E-09
Impermeable Site
Nuclide Year
C-14 2360
C-14 2360
Cs-137 0
C-14 3900
C-14 100
Co-60 0
Am-241 0
H-3 0
Humid Permeable Site
Dose Nuclide Year
(mrem/yr)
*
1.2E-01 C-14 100
*
1 . 7E-03 C-14 3900
*
*
*
1.8E-01 Co-60 0
1.6E-04 Am-241 0
2.3E-09 H-3 0
Arid
Dose
(mrem/yr)
*
6.4E-05
*
*
*
*
*
1.8E-01
1.6E-04
2.6E-09
Permeable Site
Nuclide Year
C-14 247
>13000
COr-60 0
Am-241 0
H-3 0
*Pathway not applicable.
-------
Table F-4. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 4. Metropolitan
Area and Fuel-Cycle Facility - Suburban Sanitary
Landfill (MAFC-SF)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
1.4E-06
1.8E-02
5.2E-04
4.1E-06
5.7E-04
2.3E-02
7.6E-02
8.9E-01
5.3E-02
8.QE-05
Nuclide
C-14
C-14
Cs-137
U-234
C-14
Cs-137
Cs-137
Co-60
U-234
y_234
Year
2360
2360
0
3900
100
1
1
0
0
Q
Dose
(mrem/yr)
*
1.1E+00
*
5.6E-03
*
1.4E-02
4.6E-02
8.9E-01
5.3E-02
1.7E-05
Nuclide Year Dose
(mrem/yr)
*
C-14 47 1.4E-04
*
U-234 3900 *
*
\
Cs-137 1 1.9E-02
Cs-137 1 6.4E-02
Co-60 0 8.9E-01
U-234 0 5.3E-02
\\J31A n o nc AC
w^fc-w-r v t. • Wl_ — Vtf
Nuclide Year
C-14 231
>13000
Cs-137 1
Cs-137 1
Co-60 0
U-234 0
11 OO^ A
4J— t*JT . V
*Pathway not applicable.
-------
00
Table F-5. Haximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 5. Metropolitan
Area and Fuel-Cycle Facility - Suburban Sanitary
Landfill with Incineration (HAFC-SI)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
8.0E-07
6.6E-02
2.8E-04
3.4E-06
1.5E-04
9.0E-02
3.0E-01
5.4E+00
2.1E-01
3.6E-02
Nuclide
C-14
C-14
Cs-137
U-234
C-14
Cs-137
Cs-137
Co-60
U-234
U-234
Year
2320
2320
0
3900
100
1
1
0
0
0
Dose
(mrem/yr)
*
l.BE-tfO
*
4.5E-03
*
5.5E-02
1.8E-01
5.4E+00
2.1E-01
6.0E-02
Nuclide Year Dose
(mrem/yr)
*
C-14 25 9.4E-05
*
U-234 3900 *
*
Cs-137 1 7.7E-02
Cs-137 1 2.6E-01
Co-60 0 5.4E+00
U-234 0 2.1E-01
U-234 0 3.7E-02
Nuclide Year
C-14 224
>13000
Cs-137 1
Cs-137 1
Co-60 0
U-234 0
U-234 0
"Pathway not applicable.
-------
Table F-6. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 6. Two-Unit
Power Reactor, Institutional, and Industrial
Facilities - Municipal Dump (PWRHU-HD)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
5.0E-07
2.0E-02
1.6E-03
2.0E-06
2.5E-04
2.4E-01
8.1E-01
8.8E+00
2.1E-02
1.2E-06
Nuclide
1-129
1-129
Co-60
Pu-239
1-129
Cs-137
Cs-137
Co-60
Am-241
Am-241
Year
3060
3060
0
3900
100
1
1
0
0
0
Dose
(mrem/yr)
*
6.7E-01
*
2.5E-03
*
1.5E-01
4.8E-01
8.8E+00
2. 1E-02
2.6E-06
Nuclide Year Dose
(mrem/yr)
*
C-14 32 4.9E-04
*
Pu-239 3900 *
*
Cs-137 1 2.0E-01
Cs-137 1 6.8E-01
Co-60 0 8.8E+00
Am-241 0 2.1E-02
Am-241 0 2.9E-06
Nuclide Year
1-129 556
>13000
Cs-137 1
Cs-137 1
Co-60 0
Am-241 0
Am-241 0
*Pathway not applicable.
-------
Table F-7. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 7. Uranium
Hexafluoride Facility - Municipal Dump (UHX-HD)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
H
O
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
9.0E-09
3.7E-04
5.9E-05
2.7E-06
1.5E-05
1.2E-04
4.2E-04
2.4E-02
1.3E-01
6.7E-06
Nuclide
U-234
U-234
U-234
U-234
U-234
U-234
U-234
U-235
U-234
U-234
Year
22200
22200
0
3900
100
1
1
3900
0
0
Dose Nuclide Year
(mrem/yr)
*
1.2E-03 U-234 22200
*
3.6E-03 U-234 3900
*
1.1E-04 U-234 . 1
3.8E-04 U-234 1
2.4E-02 U-235 3900
1.3E-01 U-234 0
1.5E-05 U-234 0
Dose Nuclide Year
(mrem/yr)
*
4.7E-05 U-234 22200
*
* >13000
*
1.3E-04 U-234 1
4.4E-04 U-234 1
2.6E-03 U-235 >10000
1.3E-01 U-234 0
1.6E-05 U-234 0
*Pathway not applicable.
-------
Table F-8. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 8. Uranium
Foundry - Municipal Dump (UF-MD)
Site pathways
GW to River
GW to Wei 1
Spillage
Erosion
Bathtub
Food Ons ite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Humid
Dose
(mrem/yr)
4.2E-09
1.7E-04
1.8E-05
.8.5E-07
4.7E-06
3.8E-05
1.3E-04
4.9E-03
3.9E-02
1.1E-05
Impermeable
Nuclide
U-238
U-238
U-238
U-238
U-238
U-238
U-238
U-235
U-238
U-238
Site
Year
>7E+06
>7E+06
0
3900
100
1
1
3900
0
0
Humid Permeable Site
Dose Nuclide Year
(mrem/yr)
*.
5.7E-04 U-238 >7E+06
*
1.1E-03 U-238 3900
*
3..5E-05 U-238 1
1.2E-04 U-238 1
4.9E-03 U-235 3062
3.9E-02 U-238 0
2.4E-05 U-238 0
Arid Permeable
Dose Nuclide
(mrem/yr)
*
2.2E-05 U-238
*
*
*
4. 1E-05 U-238
1.4E-04 U-238
4.7E-04 U-235
3.9E-02 U-238
2.7E-05 U-238
Site
Year
>7E+06
>13000
1
1
>13000
0
0
*Pathway not applicable.
-------
Biointrusion
Table F-9. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 9. Large
University and Medical Center with Onsite
Incineration and Disposal (LURO-3)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
tjj Erosion
i
w Bathtub
Food Onsite
Dose
(mrem/yr)
1.1E-06
5.4E-01
1.4E-05
6.4E-08
8.9E-05
*
Nuclide
C-14
C-14
Cs-137
C-14
C-14
Year
2320
2320
0
>8400
100
Dose
(mrem/yr)
*
2.4E+00
*
l.OE-04
*
*
Nuclide Year Dose
(mrem/yr)
*
C-14 16 1.3E-04
*
C-14 >6600 *
*
*
Nuclide Year
C-14 221
>28000
Direct Gamma
Dust Inhalation
Atmosphere
1.6E-01
7.6E-03
1.3E-03
Co-60
flm-241
H-3
0
0
0
1.6E-01
7.6E-03
3.2E-03
Co-60
flm-241
H-3
0
0
0
1.6E-01
7.6E-03
1.3E-03
Co-60
Am-241
H-3
0
0
o •
*Pathway not applicable.
-------
Table F-10. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 10. Large
Metropolitan Area with Consumer Wastes - Suburban
Sanitary Landfill with Incineration (LMACW-SI)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
7
H
W
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Ons ite
Biointrusion
Direct Gamma
Dust Inhalation
Dose
(mrem/yr)
7.1E-07
6.1E-02
6.9E-04
2.1E-06
2.4E-04
2.1E-01
6.8E-01
2.1E+01
3.9E-02
i . o£— v j
Nuclide
1-129
1-129
Co-60
Pu-239
1-129
Cs-137
Cs-137
Co-60
Am-241
mil— e.n i
Year Dose Nuclide Year
(mrem/yr)
2940 *
2940 1.1E+00 C-14 25
0 *
3900 2.5E-03 Pu-239 3900
100 *
1 1.2E-01 Cs-137 1
1 4.1E-01 Cs-137 1
0 2.1E+01 Co-60 0
0 3.9E-02 Am-241 0
Dose Nuclide Year
(mrem/yr)
*
7.3E-04 1-129 553
*
* >13300
*
1.7E-01 Cs-137 1
5.7E-01 Cs-137 1
2. 1E+01 Co-60 0
3.9E-02 Am-241 0
^Pathway not applicable.
\
-------
Table F-11. Haximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 11. Large
Metropolitan Area with Consumer Wastes - Urban
Sanitary Landfill with Incineration (LHACW-UI)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
-• =
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
5.5E-07
7.5E-03
9.0E-04
2.5E-06
3.5E-04
*
*
4.4E+00
1.3E-02
5.8E-04
Nuclide
C-14
C-14
Cs-137
Pu-239
C-14
Co-60
Am-241
Am-241
Year
2320
2320
0
3900
100
0
0
0
Dose
(mrem/yr)
*
4.9E-01
*
3.1E-03
*
*
*
4.4E+00
1.3E-02
9.4E-04
Nuclide Year Dose
(mrem/yr)
*
C-14 46 3.0E-04
*
Pu-239 3062 *
*
*
*
Co-60 0 4.4E-HOO
Am-241 0 1.3E-02
Am-241 0 6.0E-04
Nuclide Year
1-129 562
>13300
Co-60 0
Am-241 0
Am-241 0
*Pathway not applicable.
-------
Table F-12. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 12. Consumer
Product Wastes - Suburban Sanitary Landfill (CW-SF)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
^ Bathtub
in
Food Ons ite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
6.5E-11
9.3E-07
2.6E-05
5.1E-08
1.5E-06
1.3E-05
4.5E-05
3.9E-06
1.8E-03
2.7E-10
Nuclide
Np-237
Np-237
H-3
Am-241
Am-241
Am-241
Am-241
Am-241
Am-241
Am-241
Year
>7E+06
>7E-f-06
0
3900
100
1
1
3900
0
0
Dose
(mrem/yr)
*
4.3E-02
*
6.2E-05
*
1.4E-05
4.6E-05
3.9E-06
1 .8E-03
6.0E-10
Nuclide Year Dose
(mrem/yr)
*
H-3 23 2.6E-08
*
Am-241 3900 *
*
Am-241 1 1.5E-05
Am-241 1 5.1E-05
Am-241 3062 1.4E-09
Am-241 0 1.8E-03
Am-241 0 6.7E-10
Nuclide Year
Np-237 >7E+06
>13000
Am-241 1
Am-241 1
Am-241 >13000
Am-241 0
Am-241 0
*Pathway not applicable.
Note: Scenarios 12, 13, 14, and 15 are reference scenarios where the waste streams are already deregulated.
-------
H
Table F-13. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 13. Consumer
Product Wastes - Urban Sanitary Landfill (CW-UI)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Ons ite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
1.5E-10
3.8E-07
1.4E-04
2.9E-07
8.5E-06
*
*
3.7E-06
1.7E-03
3.6E-10
Nuclide
Np-237
Np-237
H-3
Am-241
Am-241
Am-241
Am-241
Am-241
Year
>7E+06
>7E+06
0
3900
100
3900
0
0
Dose
(mrem/yr)
*
1.7E-02
*
3.5E-04
*
*
*
3.7E-06
1.7E-03
7.8E-10
Nuclide Year Dose
(mrem/yr)
*
H-3 23 6.1E-08
*
Am-241 3900 *
*
*
*
Am-241 3062 1.3E-09
Am-241 0 1.7E-03
Am-241 0 8.7E-10
Nuclide Year
Np-237 >7E+06
Am-241 >13000
Am-241 0
Am-241 0
*Pathway not applicable.
-------
H
Table F-14. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 14. Large
University and Hedical Center with Onsite
Incineration and Disposal (LURO-1)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways
GW to River
GW to Well
Spillage
Erosion
Bathtub
Food Onsite
Biointrusion
Direct Gamma
Dust Inhalation
Atmosphere
Dose
(mrem/yr)
*
*
*
*
*
*
*
*
*
3. 1E-04
Nuclide Year Dose
(mrem/yr)
*
*
*
*
*
*
*
*
*
H-3 0 7.6E-04
Nuclide Year Dose Nuclide Year
(mrem/yr)
*
*
*
*
*
*
*
*
*
H-3 0 3.2E-04 H-3 0
*Pathway not applicable.
-------
Table F-15. Maximum annual CPG dose, dominant radionuclide,
and year of occurrence for Scenario 15. Large
University and Medical Center with Onsite
Incineration and Disposal (LURO-2)
Humid Impermeable Site
Humid Permeable Site
Arid Permeable Site
Site pathways Dose Nuclide Year Dose Nuclide Year
(mrem/yr) (mrem/yr)
GW to River 1.8E-05 C-14 2320 *
GW to Well 9.1E+00 C-14 2320 4.0E-rtl C-14 16
Spillage 2.3E-05 C-14 0 *
Erosion 1.1E-06 C-14 >8400 1.7E-03 C-14 >6600
i
S Bathtub 1.4E-03 C-14 100 *
Food Onsite * *
Biointrusion * *
Direct Gamma * *
Dust Inhalation 3.8E-06 H-3 0 3.8E-06 H-3 0
Atmosphere 1.5E-04 H-3 0 3.8E-04 H-3 0
Dose Nuclide Year
(mrem/yr)
*
2.2E-03 C-14 221
*
* >28000
*
*
*
*
3.8E-06 H-3 0
1.6E-04 H-3 0
*Pathway not applicable.
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\,
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