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 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 Łl

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 Ł•%  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 Ł2 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
 Ł2  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
Ł1
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

-------
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.
                                    6-39

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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.
                                     6-41

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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).
                                    7-1

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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).
                                     7-3

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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
                                   7-4

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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.
                                     7-9

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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

                                    7-15

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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.
                                   7-20

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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
                                   7-21

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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.
                                   7-22

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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.
                                   7-28

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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).

                                   7-33

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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
                              7-34

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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
                                   7-40

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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). -
                                    7-41

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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
                                   7-42

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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).
                                  7-46

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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.
                                   7-53

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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
                                   7-54

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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.
                                   7-55

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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.
                                   7-59

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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
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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
                              7-67

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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
                                   7-70

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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
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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.
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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"
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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

-------
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.
                                   7-78

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                                   7-81

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                                   7-82

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                                   7-83

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 ICRP75
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Kato, H., Late Effects  in  Children  Exposed  to the Atomic  Bomb
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                                   7-84

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                                   7-85

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 Miller, R.W.,  Delayed  Effects Occurring  Within the First
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 Morgan, K.Z. and J.E.  Turner, Principles of Radiation
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 Mole, R.H., Carcinogenesis by Thorotrast and Other Sources  of
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 Murphy, D.P.,  The Outcome of 625 Pregnancies  in Women Subject
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National Academy of Sciences - National  Research  Council, The
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of Ionizing Radiations  (BEIR Report), Washington,  D.C.,  1972.
                                   7-86

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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
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            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,
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 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.
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 Straume,  T. and  R. L. Dobson, Implications of New  Hiroshima
 and  Nagasaki Dose  Estimates:   Cancer Risks and Neutron RBE
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 Stein, Z.A. and  M.W. Susser,  The Epidemiology of Mental
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 Schoenberg, editor,  Marcel Dekker,  Inc.,  New York, [in
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 Tabuchi, A., Hirai,  T.,  Nakagawa, S.,  Shimada,  K. and J.
 Fugito, Clinical Findings  on  In Utero  Exposed Microcephalic
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 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
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 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
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UNSCEAR62   United Nations, Report of the United Nations Scientific
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UNSCEAR66   United Nations, Report of the United Nations Scientific
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            United Nations, New York, 1966.

UNSCEAR69   United Nations, Report of the United Nations Scientific
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UNSCEAR72   United Nations Scientific Committee on  the  Effects of.Atomic
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UNSCEAR77   United  Nations Scientific Committee on  the  Effects of  Atomic
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UNSCEAR82   United  Nations Scientific Committee on  the  Effects of  Atomic
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            IX.8, United Nations, New York, 1982.

Up75.       Upton,  A.C.,  Physical Carcinogenesis:   Radiation—History and
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            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
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            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
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 Rontgenstrahlen entwickelt hatte.  Fortschr. Geb.
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 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
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 Whittemore,  A.S.  and A. McMillan, A Lung Cancer Mortality
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 1983.

 World  Health Organization,  International Statistical
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 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,
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 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.
                                  7-92

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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
-------
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
-------
     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
-------
                •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
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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
-------
  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
 Łhrough 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
-------
 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
-------
     *  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
                                      8-16
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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
                                      8-17
-------
 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
                                      8-18
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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
                                       8-19
-------
 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,
                                      8-20
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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.
                                       8-21
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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).
                                       8-25
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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.
                                  8-26
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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
-------
     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
                                   8-29
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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.
                                       8-31
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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 ^«^tjŁ
 with  the  other events in the first year  of post-closure simulation
                                        8-37
-------
            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
 ŁnPrp    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
-------
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
-------
  (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
-------
     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
-------
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
-------
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.
                                        8-43
-------
  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  oŁ  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
0)
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 ŁS«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 nŁn  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

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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
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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
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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
-------
  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
-------
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
-------
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.
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   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.
                                     11-9
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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.
                                    11-10
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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.
                                    11-14
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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.
                                     11-15
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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.
                                    11-16
-------
(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.
                                     11-17
-------
      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.
                                    11-18
-------
(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.
                                    11-19
-------
 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
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 (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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•• DISPOSAL
.TECHNOLOGY ,

1000 •
.-_;,%,- •• •,-,-. t ••-", i.
                               /•^ -in,' ,ri", ' ,  • _

Figure  11-3.  CPG Dose Versus Time for  the Humid Permoable
              Hydrogeologlc  Site
                          ''"        '"''     '   '       !	
-------

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                                                          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-
-------
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              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
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     UJ
     cc
0.5-
         0.0-
         1.0-
        0.5-
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                  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
-------
                                                  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
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c-14" :%;;"J
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C-14 |5
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C-14
C-14
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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
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          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
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      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 Łist>.)
                                          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
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     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
(HEFŁ> 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
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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
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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
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     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
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                                 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
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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
-------
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 LLĄ.  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
-------
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
-------
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
-------
     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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                                         12-14
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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 oŁ 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 (Łor 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
oŁ 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 Ł2 (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
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       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
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       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
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      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.
                                   12-38
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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.
                                   12-39
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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
-------
      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 oŁ 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 oŁ 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.
                                   13-4
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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).
                                   13-5
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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.
                                   13-6
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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
                                   13-7
-------
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.
                                   13-8
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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.
                                   13-9
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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
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                      APPENDIX A

ACRONYMS, ABBREVIATIONS, CONVERSION FACTORS, NOTATION,
                     AND GLOSSARY
                          A-l
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                                   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
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           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
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 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
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  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
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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
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 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
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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
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 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
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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
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 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
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                                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
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-------
                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
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         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
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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
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               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
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               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
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                   APPENDIX D

HYDROGEOLOGIC/CLIMATIC DESCRIPTIONS FOR SPECIFIC
         COMMERCIAL DISPOSAL FACILITIES.
                       D-l
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                            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
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            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
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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)
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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
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 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
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           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
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 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

Ł| 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

                                                                                                       *
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 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
-------
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
-------
                                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
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                                   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* ŁA 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 . oŁ— 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.
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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.
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                           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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