-------
It has been found that the flux during the active life of the facility is
greater than that during postclosure, and will, therefore, be the limiting
factor. From Equation 5-6:
0.17 v (0.994)
T-20
(5-16)
where:
qai - allowable flux during the active period for contaminant i
(g/ma-sec)
v = windspeed (m/sec)
T = temperature (°C)
H^ = Henry's Law Constant (dimensionless)
Cli = concentration of contaminant in the sludge liquid (mg/Sl)
MWi = molecular weight of contaminant
Combining Equations 5-6 and 5-15 for the criteria case yields:
0.17 v (0.994)1"-20 H^Cli
CET/SRR =
(5-17)
which can be solved for Cl.. to give Cli£T, the limiting sludge liquid
concentration:
_
C1iET ~ SRR (0.17 v)(0.994)T-20 H-J
(5-18)
5.4. INPUT PARAMETER REQUIREMENTS
5.4.1. Fate and Transport: Pathway Data.
1. Vertical Term for Transport (V) — It is conservatively assumed
that atmospheric conditions are stable. Therefore, V will
always be equal to 1.
2. Lateral Virtual Distance (Xy) — Equal to X0 [Cot
(A0'/2)]//iT, where A01 is the sector width in
radians. It is assumed that the sector width is 0.393 (22.5°);
therefore, Xy = 2.84X0.
3. Average Wind Speed (v) — Obtained from local weather station.
4. Average Air Temperature (T) — Obtained from local weather
station.
5-15
-------
5. Air-Filled Porosity of Cover Soil (na) — It is assumed that
cover soils will be drained to field capacity. Therefore, the
air-filled porosity is assumed to be equal to the effective
porosity (ne). Values for effective porosity can be obtained
from Table 4-5.
6. Porosity of Cover Soil (n) — Can be measured in the laboratory
or obtained from Table 4-5.
7. Cover Thickness (tc) — Obtained from site design or operating
procedures.
8. Length or Width of Source (X0) — Obtained from site map or
plans. It is assumed that source areas are square. For active,
uncovered case, Xo is equal to the square root of the area of
an individual landfill cell. For postclosure, covered case,
X0 is equal to the square root of the area of the overall
landfill.
9. Distance from Center of Source to Receptor (r1) — Obtained
from site plans, r1 is taken as the sum of one-half the width
of the total landfill area (X0/2) plus the width of the
buffer area from the landfill area to the property boundary.
10. Standard Deviation of the Vertical Concentration Distance
(oz) — Atmospheric conditions are assumed to be stable.
5.4.2. Fate and Transport: Chemical-Specific Data.
1. Contaminant Concentration in Sludge Liquid (C-j) — Derived
directly for a contaminant by applying the TCLP. Alternately,
C-j can be calculated from the dry weight contaminant
concentration, Cs, the organic carbon distribution
coefficient for the contaminant, K0ct the fraction of organic
carbon in the sludge solids, foc, and the solids content of
the sludge, S.
2. Henry's Law Constant (H'j -- Obtained from Appendix B or
derived directly.
3. Molecular Weight of Contaminant (MW) — Obtained from the
literature.
5.4.3. Health Effects Data. A reference air concentration (RAC, in
3
mg/m ), will be defined as an ambient air concentration used to evaluate
the potential for adverse effects on human health as a result of sludge
landfilling. That is, for a given landfill site, and given the practice
5-16
-------
definitions and assumptions stated previously in this methodology, the cri-
terion for a given sludge contaminant is that concentration in the sludge
that cannot be exceeded, and is calculated to result in air concentrations
below the RAC at a compliance point downwind from the site. Exceeding the
RAC would be a basis for concern that adverse health effects may occur in a
human population in the site vicinity.
RAC is determined, based upon contaminant toxicity and air inhalation
rate, from the following general equation:
Reference Air Concentration: RAC = I /I (5-19)
pa
where I is the acceptable chronic pollutant intake rate (in nig/day) based
on the potential for health effects and I is the air inhalation rate (in
a
mVday). This simplified equation assumes that the inhaled contaminant
is absorbed into the body via the lungs at the same rate in humans as in the
experimental species tested, or between routes of exposure (e.g., oral and
inhalation). Also, this equation assumes that there are no other exposures
of the contaminant from other sources, natural or manmade. I varies
according to the pollutant evaluated and according to whether the pollutant
acts according to a threshold or nonthreshold mechanism of toxicity.
5.4.3.1. THRESHOLD-ACTING TOXICANTS — Threshold effects are those
for which a safe (i.e., subthreshold) level of toxicant exposure can be
estimated. For these toxicants, RAC is derived as follows:
/RfD x bw
Reference Air concentration: KAU =
where:
RfD = reference dose (mg/kg/day)
bw = human body weight (kg)
/RfD x bw\
\ RE /
TBI
(5-20)
5-17
-------
TBI = total background Intake rate of pollutant from all other sources
of exposure (mg/day)
Ia = air inhalation rate (ma/day)
RE « Relative effectiveness of inhalation exposure (unitless)
The definition and derivation of each of the parameters used to estimate RAC
for threshold-acting toxicants are further discussed below.
5.4.3.1.1. Reference Dose (RfD) —When toxicant exposure is by
ingestion, the threshold assumption has traditionally been used.to establish
an acceptable daily intake, or ADI. The Food and Agricultural Organization
and the World Health Organization have defined ADI as "the daily intake of a
chemical which, during an entire lifetime, appears to be without appreciable
risk on the basis of all the known facts at the time. It is expressed in
milligrams of the chemical per kilogram of body weight (mg/k;g)" (Lu, 1983).
Procedures for estimating the ADI from various types of toxicological data
were outlined by the U.S. EPA in 1980 (U.S. EPA, 1980c). More recently the
Agency has preferred the use of a new term, the "reference dose," or RfD, to
avoid the connotation of acceptability, which is often controversial.
The RfD is an estimate (with uncertainty spanning perhaps an order of
magnitude) of the daily exposure to the human population (including
sensitive subgroups) that is likely to be without appreciable risk of
deleterious effects during a lifetime. The RfD is expressed in units of
mg/kg bw/day. The RfD is estimated from observations in humans whenever
possible. When human data are lacking, observations in animals are used,
employing uncertainty factors as specified by existing Agency methodology.
RfD values for noncarcinogenic (or systemic) toxicity have been derived
by several groups within the Agency. An Intra-Agency Work Group verifies
each RfD, which is then loaded onto the Agency's publically available
5-18
-------
Integrated Risk Information System (IRIS) database. Most of the
noncarcinogenic chemicals that are presently candidates for sludge criteria
for the landfill pathway are included on the Agency's RfD list, and thus no
new effort will be required to establish RfDs for deriving sludge criteria.
For any chemicals not so listed, RfO values should be derived according to
established Agency procedures (U.S. EPA, 1988),
5.4.3.1.2. Human Body Weight (bw) and Air Inhalation Rate (I ) —
a
An assumption of 20 m3 inhalation/day by a 70-kg adult has been widely
used in Agency risk assessments and will be used in this methodology for
adults. Table 5-3 shows values of I for a typical man, woman, child and
a
infant with a typical activity schedule, as defined by the International
Commission on Radiological Protection (ICRP, 1975). Additional values have
been derived for an adult with the same activity schedule, but using upper
limit rather than average assumptions about respiration rates for each
activity, and for an adult with normal respiration rates, but whose work is
moderately active and who practices 1 hour of heavy activity (i.e.,
strenuous exercise) per day (Fruhman, 1964, as cited Diem and Lentner, 1970;
Astrand and Rodahl, 1977,, as cited in Fiserova-Bergerova, 1983).
Representative body weights have been assigned to each of these individuals
to calculate a respiratory volume-to-body weight ratio. (These ratios have
been derived for illustrative purposes only.) The resulting ratio values
range from 0.33 to 0.47 m3/kg/day, all of which exceed the ratio value
of 0.29 m3/kg/day estimated from the 70-kg adult inhaling 20 ma/day,
as used currently by the Agency. Therefore, the typically assumed values
for adults may underestimate actual exposure. In cases where children or
infants are known to be at specific risk, it may be more appropriate to use
values of bw and I for children or infants.
a
5-19
-------
in
O.O
- —
o 8
o\
•8
I rtn,
o +-
— CL-
8.
5-20
-------
5.4.3.1.3. Total Background Intake Rate of Pollutant (TBI) — It is
important to recognize that sources of exposure other than sludge disposal
practices may exist, and that the total exposure should be maintained below
the RfD. Other sources of exposure include background levels (whether
natural or anthropogenic) in drinking water, food or air. Other types of
exposure, due to occupation or habits such as smoking, might also be
included depending on data availability and regulatory policy. These expo-
sures are summed to estimate TBI.
Data for estimating background exposure usually are derived from
analytical surveys of surface, ground or tap water, from FDA market-basket
surveys and from air-monitoring surveys. These surveys may report means,
medians, percentiles or ranges, as well as detection limits. Estimates of
TBI may be based on values representing central tendency or on upper-bound
exposure situations, depending on regulatory policy. Data chosen to esti-
mate TBI should be consistent with the value of bw. Where background data
are reported in terms of a concentration in air or water, ingestion or
inhalation rates applicable to adults or children can be used to estimate
the proper daily background intake value. Where data are reported as total
daily dietary intake for adults and similar values for children are unavail-
able, conversion to an intake for children may be required. Such a conver-
sion could be estimated on the basis of relative total food intake or rela-
tive total caloric intake between adults and children.
For example, in deriving the National Emission Standard for mercury, the
average dietary contribution of 10 pg Hg/70 kg/day was subtracted from the
assumed threshold of 30 jjg/70 kg/day to give an allowable increment from
inhalation exposure of 20 yg/70 kg/day. An assumed inhalation volume of
5-21
-------
20 m /day for a 70-kg man was then applied to derive an allowable
ambient air concentration of 1 yg Hg/m3 (U.S. EPA, 1984a). For the
purposes of this methodology, however, TBI should be an estimate of
background exposure from all sources, including inhalation.
As stated in the beginning of this subsection, the TBI is the summed
estimate of all possible background exposures, except exposures resulting
from a sludge disposal practice. To be more exact, the TBI should be a sum-
med total of all toxicologically effective intakes from all nonsludge
exposures. To determine the effective TBI, background intake values (131)
for each exposure route must be divided by that route's particular relative
effectiveness (RE) factor. Thus, the TBI can be mathematically derived,
after all the background exposures have been determined, using the following
equation:
BI (food) BI (water) BI (air) BI (nV
TBI - RE (food) + RE (water) + RE (air) + •" + RE (n)
(5-21)
where:
TBI = total background intake rate of pollutant from all other
sources of exposure (mg/day)
BI = background intake of pollutant from a given exposure route,
indicated by subscript (mg/day)
RE = relative effectiveness, with respect to inhalation exposure,
of the exposure route indicated by subscript (unitless)
5.4.3.1.4. Fraction of Inhaled Air from Contaminated Area — It is
recognized that an individual exposed to air emissions from a landfill may
not necessarily remain in the landfill proximity for 24 hours/day. However,
if it is assumed that residential areas may be contaminated, it is likely
that less mobile individuals will include those at greatest risk. Therefore,
it is reasonable to assume that 100% of the air breathed by the ME Is will be
from the area of the landfill.
5-22
-------
5.4.3.1.5. Relative Effectiveness of Exposure (RE) — RE is a
unitless factor that shows the relative toxicological effectiveness of an
exposure by a given route when compared to another route. The value of RE
may reflect observed or estimated differences in absorption between the
inhalation and ingestion routes, which can then significantly influence the
quantity of a chemical that reaches a particular target tissue, the length
of time it takes to get there, and the degree and duration of the effect*
The RE factor may also reflect differences in the occurrence of the critical
toxicological effects at the portal of entry. For example, carbon tetra-
chloride and chloroform were estimated to be 40% and 65% as effective,
respectively, by inhalation as by ingestion based on high-dose absorption
differences (U.S. EPA, 1984b,c). In addition to route differences, RE can
also reflect differences in the exposure matrix. For example, absorption of
nickel ingested in water has been estimated to be 5 times that of nickel
ingested in the diet (U.S. EPA, 1985d). The presence of food in the
gastrointestinal tract may delay absorption and reduce the availability of
orally administered compounds, as demonstrated for halocarbons (NRC, 1986).
Physiologically based pharmacokinetic (PB-PK) models have evolved into
particularly useful tools for predicting disposition differences due to
exposure route differences. Their use is predicated on the premise that an
effective (target-tissue) dose achieved by one route in a particular species
is expected to be equally effective when achieved by another exposure route
or in some other species. For example, the proper measure of target-tissue
dose for a chemical with pharmacologic activity would be the tissue concen-
tration divided by some measure of the receptor binding constant for that
chemical. Such models account for fundamental physiologic and biochemical
5-23
-------
parameters such as blood flows, ventilatory parameters, metabolic capacities
and renal clearance, tailored by the physicochemical and biochemical prop-
erties of the agent in question. The behavior of a substance administered
by a different exposure route can be determined by adding equations that
describe the nature of the new input function. Similarly, since known
physiologic parameters are used, different species (e.g., humans vs. test
species) can be modeled by replacing the appropriate constants. It should
be emphasized that PB-PK models must be used in conjunction with toxicity
and mechanistic studies in order to relate the effective dose associated
with a certain level of risk for the test species and conditions to other
scenarios. A detailed approach for the application of PB-PK models for
derivation of the RE factor is beyond the scope of this document, but the
reader is referred to the comprehensive discussion in NRC (1986). Other
useful discussions on considerations necessary when extrapolating route to
route are found in Pepelko and Withey (1985) and Clewell and Andersen (1985).
Since exposure for the vapor pathway is by inhalation, the RE factors
applied are all with respect to the inhalation route. Therefore, the value
of RE in Equation 5-20 gives the relative effectiveness of the exposure
route and matrix on which the RfD was based when compared to. inhalation of
contaminated air. Similarly, the RE factors in Equation 5-21 show the
relative effectiveness, with respect to the inhalation route, of each back-
ground exposure route and matrix.
An RE factor should only be applied where well-documented, referenced
information is available on the contaminant's observed relative effective-
ness or its pharmacokinetics. When such information is not available, RE is
equal to 1.
5-24
-------
5.4.3.2. CARCINOGENS — For carcinogenic chemicals, the Agency con-
siders the excess risk of cancer to be linearly related to dose, except at
high-dose levels (U.S. EPA, 1986a). The threshold assumption, therefore,
does not hold, as risk diminishes with dose but does not become zero or
background until dose becomes zero.
The decision whether to treat a chemical as a threshold- or nonthreshold-
acting (i.e., carcinogenic) agent depends on the weight of the evidence that
it may be carcinogenic to humans. Methods for classifying chemicals as to
their weight of evidence have been described by U.S. EPA (U.S. EPA, 1986a),
and most of the chemicals that presently are candidates for sludge criteria
have recently been classified in Health Assessment Documents or other
reports prepared by the U.S. EPA's Office of Health and Environmental
Assessment (OHEA), or in connection with the development of recommended
maximum contaminant levels (RMCLs) for drinking-water contaminants (U.S.
EPA, 1985e). To derive values of the reference air concentration (RAC), a
decision must be made as to which classifications constitute sufficient
evidence for basing a quantitative risk assessment on a presumption of
carcinogenicity. Chemicals in classifications A and B, "human carcinogen"
and "probable human carcinogen," respectively, have usually been assessed as
carcinogens, whereas those in classifications D and E, "not classifiable as
to human carcinogenicity because of inadequate human and animal data" and
"evidence of noncarcinogenicity for humans," respectively, have usually been
assessed according to threshold effects. Chemicals classified as C,
"possible human carcinogen," have received varying treatment. For example,
lindane, classified by the Human Health Assessment Group (HHAG) of the U.S.
EPA as B2~C, or between the lower range of the B category and category C,
5-25
-------
has been assessed using both the linear model for tumorigenic effects (U.S.
EPA, 1980b) and based on threshold effects (U.S. EPA, 1985e). Table 5-4
gives an illustration of these U.S. EPA classifications based on the
available weight of evidence.
Using the weight-of-evidence classification without noting the
explanatory material for a specific chemical may lead to a flawed conclu-
sion, since some of the classifications are exposure-route dependent.
Certain compounds (e.g., nickel) have been shown to be carcinogenic by the
inhalation route, but not by ingestion. The issue of whether or not to
treat an agent as carcinogenic by ingestion remains controversial for
several chemicals.
If a pollutant is to be assessed according to nonthreshold, carcinogenic
effects, the RAC is derived as follows:
Reference Air Concentration: RAC =
(5-22)
where:
q-j* = human cancer potency [(mg/kg/day) 1]
RL = risk level (unitless) (e.g., l(r5, 1CT6, etc.)
bw = human body weight (kg)
RE - relative effectiveness of inhalation exposure (unitless)
Ia - air inhalation rate (m3/day)
TBI = total background intake rate of pollutant (mg/day), from
all other sources of exposure
The RAC, in the case of carcinogens, is thought to be protective since the
q * is typically an upper-limit value (i.e., the true potency is consid-
ered unlikely to be greater and may be less). The definition and derivation
5-26
-------
TABLE 5-4
Illustrative Categorization of Evidence Based on Animal and Human Data*
Animal Evidence
Human
Evidence
Sufficient
Limited
Inadequate
No data
No evidence
Sufficient
A
Bl
82
B2
B2
Limited
A
Bl
C
C
C
Inadequate
A
Bl
D
0
D
No Data
A
Bl
D
D
D
No
Evidence
A
Bl
D
E
E
*The above assignments are presented for illustrative purposes. There may
be nuances in the classification of both animal and human data indicating
that different categorizations than those given in the table should be
assigned. Furthermore, these assignments are tentative and may be modified
by ancillary evidence. In this regard, all relevant information should be
evaluated to determine if the designation of the overall weight of evidence
needs to be modified. Relevant factors to be included along with the tumor
data from human and animal studies include structure-activity relationships;
short-term test findings; results of appropriate physiological, biochemical
and toxicological observations; and comparative metabolism and pharmaco-
kinetic studies. The nature of these findings may cause an adjustment of
the overall categorization of the weight of evidence.
5-27
-------
Cancer Potency (q*) — For most carcinogenic
of each of the parameters used to estimate RAC for carcinogens are further
discussed in the following subsections.
5.4.3.2.1. Human
chemicals, the linearized multistage model is recommended for estimating
human cancer potency from animal data (U.S. EPA, 1986a). When
epidemiological data are available, potency is estimated based on the
observed relative risk in exposed vs. nonexposed individuals, and on the
magnitude of exposure. Guidelines for use of these procedures have been
presented in the U.S. EPA (1980c, 1985e) and in each of a series of Health
Assessment Documents prepared by OHEA (e.g., U.S. EPA, 1985d). The true
potency value is considered unlikely to be above the upper-bound estimate of
the slope of the dose-response curve in the low-dose range, and it is
expressed in terms of risk/dose, where dose is in units of mg/kg/day. Thus,
q,* has units of (mg/kg/day) 1. OHEA has derived potency estimates
for each of the potentially carcinogenic chemicals that are presently
candidates for sludge criteria. Therefore, no new effort will be required
to develop potency estimates to derive sludge criteria.
5.4.3.2.2. Risk Level (RL) — Since by definition no "safe" level
exists for exposure to nonthreshold agents, values of RAC are calculated to
reflect various levels of cancer risk. If RL is set at zero, then RAC will
be zero. If RL is set at 10 6, RAC will be the concentration which, for
lifetime exposure, is calculated to have an upper-bound cancer risk of one
case in one million individuals exposed. This risk level refers to excess
cancer risk, i.e., over and above the background cancer risk in unexposed
individuals. By varying RL, RAC may be calculated for any level of risk in
the low-dose region, i.e., RL <10~2. Specification of a given risk
5-28
-------
level on which to base regulations is a matter of policy. Therefore, it is
common practice to derive criteria representing several levels of risk
without specifying any level as "acceptable."
5.4.3.2.3. Human Body Weight (bw) and Air Inhalation Rate (I ) —
9
Considerations for defining bw and I are similar to those stated in Sec-
3
tion 5.4.3.1.2. The HHAG assumes respective values of 70 kg and 20
mVday to derive unit risk estimates for air, which are potency
estimates transformed to units of (yg/m )
3.-1
As illustrated in
Table 5-3, exposures may be higher in children than in adults when the
ratios of inhalation volumes to body weights are compared. However, because
exposure is lifelong, values of bw and I are usually chosen to be
a
representative of adults.
5.4.3.2.4. Relative Effectiveness of Exposure (RE) — RE is a unit-
less factor that shows the relative toxicological effectiveness of an expo-
sure by a given route when compared to another route. The value of RE may
reflect observed or estimated differences in absorption between the inhala-
tion and ingestion routes, which can significantly influence the quantity of
a chemical that reaches a particular target tissue, the length of time it
takes to get there, and the degree and duration of the effect. The RE
factor may also reflect differences in the occurrence of critical toxico-
:logical effects at the portal of entry. For example, carbon tetrachloride
and chloroform were estimated to be 40% and 65% as effective, respectively,
by inhalation as by ingestion based on high-dose absorption differences
(U.S. EPA, 1984b,c). In addition to route differences, RE can also reflect
differences in the exposure matrix. For example, absorption of nickel
ingested in water has been estimated to be 5 times that of nickel ingested
5-29
-------
in food (U.S. EPA, 1985d). The presence of food in the gastrointestinal
tract may delay absorption and reduce the availability of orally
administered compounds, as demonstrated for halocarbons (NRC, 1986).
PB-PK models have evolved into particularly useful tools for predicting
disposition differences due to exposure route differences. Their use is
predicated on the premise that an effective (target-tissue) dose achieved by
one route in a particular species is expected to be equally effective when
achieved by another exposure route or in some other species. For example,
the proper measure of target-tissue dose for a chemical with pharmacologic
activity would be the tissue concentration divided by some measure of the
receptor binding constant for that chemical. Such models account for
fundamental physiologic and biochemical parameters such as blood flows,
ventilatory parameters, metabolic capacities and renal clearance, tailored
by the physicochemical and biochemical properties of the agent in question.
The behavior of a substance administered by a different exposure route can
be determined by adding equations that describe the nature of the new input
function. Similarly, since known physiologic parameters are used, different
species (e.g., humans vs. test species) can be modeled by replacing the
appropriate constants. It should be emphasized that PB-PK models must be
used in conjunction with toxicity and mechanistic studies in order to relate
the effective dose associated with a certain level of risk for the test
species and conditions to other scenarios. A detailed approach for the
application of PB-PK models for derivation of the RE factor is beyond the
scope of this document, but the reader is referred to the comprehensive
discussion in NRC (1986). Other useful discussions on considerations
necessary when extrapolating route to route are found in Pepelko and Withey
(1985) and Clewell and Andersen (1985).
5-30
-------
Since exposure for the vapor pathway is by inhalation, the RE factors
applied are all with respect to the inhalation route. Therefore, the value
of RE in Equation 5-22 gives the relative effectiveness of the exposure
route and matrix on which the q * was based when compared to inhalation of
contaminated air. Similarly, the RE factors in Equation 5-21 show the
relative effectiveness, with respect to the inhalation route, of each back-
ground exposure route and matrix.
An RE factor should only be applied where well-documented, referenced
information is available on the contaminant's observed relative effective-
ness or its pharmacokinetics. When such information is not available, RE is
equal to 1.
5.4.3.2.5. Total Background Intake Rate of Pollutant (TBI) — It is
important to recognize that sources of exposure other than sludge disposal
practices may exist, and that the total exposure should be maintained below
the determined cancer risk-specific exposure level. Other sources of
exposure include background levels (whether natural or anthropogenic) in
drinking water, food or air. Other types of exposure, due to occupation or
habits such as smoking, might also be included depending on data availabil-
ity and regulatory policy. These exposures are summed to estimate TBI.
Data for estimating background exposure usually are derived from analyt-
ical surveys of surface, ground or tap water, from FDA market-basket sur-
veys, and from air-monitoring surveys. These surveys may report means,
medians, percentiles or ranges, as well as detection limits. Estimates of
TBI may be based on values representing central tendency or on upper-bound
exposure situations, depending on regulatory policy. Data chosen to esti-
mate TBI should be consistent with the value of bw. Where background data
5-31
-------
are reported in terms of a concentration in air or water, ingestion or
inhalation rates applicable to adults can be used to estimate the proper
daily background intake, value. For certain compounds (e.g., nickel) that
have been shown to be carcinogenic by the inhalation route, but not by the
ingestion route, the TBI should not include background exposure from the
ingestion route. Thus, in such a case only background exposures from other
air emission sources should be included in the TBI.
As stated in the beginning of this subsection, the TBI is the summed
estimate of all possible background exposures, except exposures resulting
from a sludge disposal practice. To be more exact, the TBI should be a sum-
med total of all toxicologically effective intakes from all nonsludge expo-
sures. To determine the effective TBI, background intake values (BI) for
each exposure route, must be divided by that route's particular relative
effectiveness (RE) factor. Thus, the, TBI can be mathematically derived,
after all the background exposures have been determined, using the following
equation:
_ BI (food) BI (water) BI (air) BI (n)
TBI ~ RE (food) + RE (water) + RE (air) + •" "''RE (n)
(5-23)
where:
TBI = total background intake rate of pollutant from all other
sources of exposure (mg/day)
£
BI = background intake of pollutant from a given exposure route,
indicated by subscript (mg/day)
RE ='relative effectiveness, with respect to inhalation exposure,
of the exposure route indicated by subscript (unitless)
5.4.3.2.6. Fraction of Inhaled Air From Contaminated Area — It is
recognized that an individual exposed to air emissions from a landfill may
not necessarily remain in the landfill proximity for 24 hours/day. However,
5-32
-------
if it is assumed that residential areas may be contaminated, it is likely
that less mobile individuals will include those at greatest risk. There-
fore, it is reasonable to assume that 100% of the air breathed by the MEIs
will be from the area of landfill.
5.5. SITE-SPECIFIC APPLICATION
This section presents sample calculations for determining the vapor
exposure resulting from landfilling of sludge. In the following, calcula-
tions are first made for a particular landfill on a site-specific applica-
tion and then an example is given for calculating maximum allowable contami-
nant levels in sludge. Benzene, because it is a volatile contaminant of
concern, is used for the example calculations. For the examples, data de-
scribing the occurrence and concentration of benzene in sludge are taken
from U.S. EPA (1985a). The pathway and chemical parameters used in the
calculations are summarized in Table 5-5. , Data describing waste sites are
values assumed to represent reasonable cases. In actual practice, the data
used in the calculations would be those measured or collected by the
applicant.
Assume operating procedures include excavation of a 4- by 16-m trench,
disposal of three daily 0.5-m lifts in each trench, application of a daily
cover of 0.3 m soil and application of a final cover of 1.0 m soil. Assume
that 67% of the total disposal site area is available for trenches (Table
5-6).
5.5.1. Tier 1 Calculation. The Tier 1 calculation involves comparing the
equilibrium vapor concentration of the constituent with the reference air
concentration (RAC). This approach represents the worst possible case with
no allowance made for atmospheric dilution, dispersion or degradation. The
5-33
-------
TABLE 5-5
Input Parameters for Example Calculations: Vapor Loss
Fate and Transport: Pathway Data
1. Vertical term for transport
2. Lateral virtual distance
3. Average windspeed
4. Average air temperature
V = 1
Xy = 2.84xo
xo = 22.64 m active
1361.46 m postclosure
v = 2 m/sec
T = 15°C
5. Air-filled porosity of cover soil na = 0.1
6. Porosity of cover soil
7. Cover thickness
8. Length of source
9. Distance from center of source
to receptor
10. Standard deviation of the
vertical concentration distance
n = 0.4
Tc
0.3 m active,; 1.0 m post-
closure
X0 = 8 m active, 480 m post-
closure
r1 = 340 m
az =6.2 m
Fate and Transport: Chemical-Specific Data (Benzene)
11. Contaminant concentration in
sludge liquid
12. Henry's Law Constant
13. Molecular weight of contaminant
Health Data (Benzene)
14. Reference concentration in air
X = 0.0083 mg/9.
H1 = 0.24
MW = 78
RAC = 6.73xlO~a vg/m3
5-34
-------
TABLE 5-6
Supporting Sludge landfill Characteristics
Daily disposal rate
Trench dimensions
Depth of daily fill
Life of facility
Total trench area
Total disposal site area
10 dry metric tons/day
4 m by 16 m
0.5 m
20 years
156,000 m2
234,000 m2
5-35
-------
equilibrium vapor concentration is taken as the product of the Henry's Law
Constant of the constituent and the liquid phase concentration of the
constituent.
The liquid phase constituent concentration can be obtained in several
ways. If the leachate extraction procedure is used, the liquid concentra-
tion will be determined directly from the procedure. If the analytical
results are expressed in terms of dry weight, it will be necessary to con-
vert the dry weight results to an equivalent liquid phase concentration
accounting for partitioning between the liquid and solid phases. This is
accomplished with Equation 5-24:
cdrv s
Ci =
(5-24)
where:
Koc foe
C-|
~ concentration of contaminant in sludge liquid (mg/9.)
= dry weight concentration of contaminant in sludge (mg/kg)
S = solids content of sludge (kg dry solids/kg total wet
sludge)
Koc = organic carbon distribution coefficient
mg contaminant/kg organic carbon
mg contaminant/5, sludge liquid
foc = organic carbon content of sludge (kg organic carbon/
kg sludge solids)
Y8, = density of sludge liquid (kg sludge liquid/8. sludge liquid)
For the example calculation, the mean dry weight concentration of benzene
in sludge, 0.326 mg/kg, reported in U.S. EPA (1985a) is used. The organic
carbon distribution coefficient for benzene is 74.2 a./kg (U.S. EPA, 1985a).
Assuming a solids content of 30% for dewatered sludge, an organic carbon
content of 50% for the sludge solids and a density of 1 kg/a, for the sludge
liquid, the equivalent liquid concentration is the following:
5-36
-------
Cl =
(0.326 mg/kgH0.30 kg/kg)
(74.2 !l/kg)(0.5 kg/kg)(0.30 kg/kg) + (1-0-30) kq/kq
1.0 kg/I,
= 0.0083 mg/8.
The Henry's Law Constant for benzene is then used to calculate the vapor
concentration in equilibrium with the liquid concentration:
Cv = H C
1
(5-25)
From Appendix B, the dimensionless Henry's Law Constant for benzene is 0.24.
The equilibrium vapor pressure is:
Cv = (0.24) (0.0083 mg/fc) = 0.0020 mg/!l = 2.0 mg/m3
The RAC for the carcinogen benzene is derived using Equation 5-22:
RAC =
/RL x bw \
\q!* x REJ
- TBI
*' la
The risk level (RL), the body weight (bw), and the daily inhalation volume
(Ia) are set for this example at 10~6, 70 kg and 20 m3, respec-
tively. The relative effectiveness factor (RE) is set at 1. The human
cancer potency for benzene has been determined by the U.S. EPA to be
5.2xlO~2 (mg/kg/day)"1. Current total background intake (TBI) of
benzene from all other sources (i.e., except from landfilling of sludges)
Ms
has not been determined for 1986, but for illustrative purposes a TBI of
zero is used here to derive an example RAC. Determination of an RAC for a
specific landfill site should be based on a current local assessment of TBI.
RAC
fc
10~6 x 70 kg
.2x10-2 (mg/kg/day)
= 6.73 x 10~s mg/m3
= 6.73 x 10~2 yg/m3
= 0.0673 yg/m3
) -o
-1 X I/
* 20 ma
5-37
-------
The above vapor concentration is then compared to the reference value for
benzene, RAC = 6.73xlO~2 yg/m3. Since 2 mg/ma » 6.73xlO~2 vg/m3 ,
it is necessary to proceed to Tier 2.
5.5.2. Tier 2 Calculation. The Tier 2 methodology involves estimating
the flux of contaminant out of the landfill and using an atmospheric
dispersion model to predict the atmospheric concentration of contaminant
downwind of the site. The long-term average downwind concentration is then
compared with the RAC. .
The first step of the Tier 2 methodology is to calculate the flux rate
of contaminants during the active life of the trench (i.e., before emplace-
ment of final cover). The flux rate for the active period of the cell is
taken as the time-weighted average of the flux rate with no cover and the
flux rate with temporary daily cover. Under the assumed operating condi-
tions, each trench will be active for a 3-day period. On each of the 3
days, a lift of sludge will be applied followed by a temporary soil cover.
It is assumed that the sludge will be uncovered for 4 hours each day.
Therefore, the fraction of time that the sludge is uncovered is
(3x4)/(3x24) = 0.17. The fraction of time that the sludge is covered by
temporary cover is .( 3x20) /( 3x24) = 0.83.
The flux rate during the portion of time that wastes are uncovered is
calculated using Equation 5-6:
T-20 /"""""
q . = [0.17 v (0.994) H.C,.]/YMWI
CM III
Based on the parameter values provided in Tables 5-4 and 5-5, the resulting
contaminant flux rate is:
q • =
31
m/sec)(0.994)15~2°(0.24)(0.0083
= 7.9 x 10~s g/m2-sec
5-38
-------
The flux rate during the portion of time that wastes are covered by
temporary cover is calculated using Equation 5-8:
2
q = [9.2 x 10~5 n!0/3 (1 .006)T~20
pi
a
TC n
As shown in Table 5-5, the thickness of the temporary soil cover, T , is
\f
0.3m. The air-filled porosity of the soil cover, n , and the total
9
porosity of the soil cover, n, are assumed to be 0.1 and 0.4, respectively.
The resulting contaminant flux rate is:
qp. = [(9.2 x 10~5)(0.1)10/3 (1.006)15"20 (0.24) (0.0083) ]/|VMWI (0.3)(0.4):
qp1 = 2.0 x 10~10 g/m2-sec
The time-weighted flux for the active period is then:
(7.9 x 10~5 g/m2-sec)(0.17) + (2.0 x 10"10 g/m2-sec)(0.83)
= 1.3 x 10~5 g/m2-sec
The second step of the Tier 2 methodology is to calculate the flux rate
of contaminants during the postclosure period. The flux during the post-
closure period is calculated using Equation 5-8:
'Pi
[9.2xlO-5na10/3(1.006)T-2°
c n2
The values of variables used will be the same as those used to calculate flux
through the temporary cover, except for the cover thickness, T . For post-
closure, the cover thickness is 1.0 m. The resulting postclosure flux is:
qp. = [9.2 x 10
5
(1.006)15~20 (0.24)(0.0083) ]/|
|V78~(1)(0.4)2|
= 5.8 x 10 g/m -sec
The long-term average exposure level is based on exposure to contami-
nants over a 70-year period. The fluxes for the active and postclosure
cases, therefore, must be adjusted to reflect the period of exposure. For
the active case, one landfill cell will be open at any time during the
entire 20-year active period. The long-term average flux associated with
the active portion of the fill is then 20/70 times the active flux or:
5-39
-------
(20/70)(1.3 x 10~s g/m2-sec) = 3.71 x 10~6 g/m2-sec
The long-term average flux associated with postclosure will be based on the
average postclosure life of the landfill cells. That is, all cells will be
closed at least 50 years and the maximum postclosure period wiII be 70 years.
Because the postclosure period varies linearly from 50 to 70 years, the
average period of 60 years was used. The long-term average Flux associated
With postclosure will then be 60/70 of the postclosure flux or:
(60/70)(5.8 x 10"11 g/m2-sec) = 5.0 x 10"11 g/m2-sec '
Because the active and postclosure fluxes involve different source areas,
the fluxes could not be summed to obtain a single long-term average flux to
calculate exposure. That is, the active flux is associated with the area of
one landfill cell, while the postclosure flux is associated with the area of
the total disposal site. Therefore, the downwind concentrations associated
with each flux were calculated.and summed to obtain the average exposure.
Downwind exposure concentrations were calculated using Equation 5-9:
X02Q V
C(r,0) - (2.032 x 10*) [—
Xy) v oz
The above equation is for the concentration along the center line of the
plume, which represents a worst case.
For the active-period exposure, the side of the source area, X , is
108 m. In determining the distance from the source center to the receptor,
r1, it was assumed that the active cell was always located in the center of
the disposal area. The distance was then taken as one-half the square root
of the area of the disposal site plus a buffer zone distance, assumed to be
100 m.
r1 = (0.5) (234,000 m2) + 100 m = 340 m
5-40
-------
The lateral virtual distance, X. was calculated using Equation 5-13:
= 8(2.84) = 22.69 m
a was calculated for a worst-case condition corresponding to stable
atmospheric conditions. For a downwind distance corresponding to the dis-
tance r1 (0.34 km), a was calculated using the data presented in
Table 5-1.
oz = 14.456 (0.34)0'78407 = 6.2 m
For stable atmospheric conditions and a contaminant release height of 0 m, L
is infinite and therefore the vertical term, V, is equal to 1 (Equation
5-11). For the assumed wind speed of 2 m/sec, the resulting downwind con-
centration is:
C (r,0) = (2.032 x
= (2.032 X 10s)(8)2 (3.71
(340 + 22.69)(2)(6.2)
,' 3
0.11 vg/m
For the postclosure period, the side of the source area, X , is 480
o
m. The distance from the source center to the receptor, r1, is the same as
for the active case. As for the active case, the lateral virtual distance,
X , was calculated using Equation 5-13:
Of
= 1361.46 m
5-41
-------
As with the active case, <*z was selected for a worst-case condition cor-
responding to stable atmospheric conditions. Because the downwind distance
is the same as for the active case, o will be 6.2 m. The vertical
term, V, will also be equal to 1 , as in the active case. Using the above
data and the assumed wind velocity of 2 m/sec, the downwind contaminant
concentration is:
C(rt0) • 2.032 X10*
2.032 x IP6 (480)2 (5.0 x 10~11)(1)
(340 + 1361.46)(2)(6.2)
= 1 .1 x 10~3 yg/m3
The above results show that the exposure due to the postclosure release
is negligible compared to the exposure due to active release. The total
exposure concentration for comparison to the reference level will then be
the active exposure, or 0.11 ng/m3. This is compared to the reference
air concentration, RAC = 6.73xlO~2 yg/m3, for benzene.
5.6. NATIONAL CRITERIA SITE-SPECIFIC APPLICATION
To establish sludge concentration criteria for volatile contaminants, it
is necessary to operate the methodology provided here in a reverse mode.
That is, the RAC must be taken as input to determine the maximum allowable
concentration in the source sludge. From Equation 5-14, the compliance
point concentration is defined as:
C = Q x SRR
Since SRR is characteristic of a site and not concentration dependent, a
single value can be calculated for a representative site. When C is set at
C , the long-term effects threshold concentration, the allowable
long-term average flux Q is defined as:
Q = CET/SRR
5-42
(5-26)
-------
The flux during the active life of the facility was shown in Section 5.5.2.
to be far greater than that during postclosure, and therefore the latter may
be set equal to zero in the calculation of criteria. From Equation 5-6:
0.17 v (0.994)T-20 HjClj
Qai = /-——
y MWi
where:
qai = flux during the uncovered period (g/m2-sec)
v = windspeed (m/sec)
T = temperature (°C)
Hi = Henry's Law Constant (dimensionless)
Cli = concentration of contaminant in the sludge liquid (mg/s.)
MWi = molecular weight of contaminant
The average flux during the human lifetime is determined by adjusting the
uncovered period flux, q ., for the fraction of time that the sludge is
uncovered (0.17) during the facility active life, and for the assumed total
active life of the facility (20 years) during the human lifetime (70 years),
as described in Section 5.5.2. The resulting relationship is as follows:
(5-27)
Combining Equations 5-6, 5-26 and 5-27 for the criteria case yields:
0.17 v (0.994)1"-20
Q = q . x 0.17 x (20/70)
31
CET/SRR
x 0.17 x
(5-28)
The latter can be solved for Cl to give Cl , the limiting sludge
liquid concentration:
CET
CliET =
70
SRR (0.17 v)(0.994)T-20Hi 0-17 20
(5-29)
For benzene, the MW
6.73xlO~2 pg/m3. Therefore:
6.73 x IP"2 C/78)
is 78 g/mol, the H. is 0.24 and the RAC is
CliET =
(3xlO*)(0.17)(2)(0.994)
= 4.9 x 10~3 mg/8,
ic_
ID
1 70
x - x— (5-30)
(0.24) 0.17 20
5-43
-------
-3
The national criteria would, therefore, be set at 4.9x10 " mg/9. in
leachate. Based on Equation 5-24 and assuming a solids content of 30% for
dewatered sludge, an organic content of 50% for the sludge solids and a
density of 1.0 kg/8, for the sludge liquid, the corresponding dry weight
concentration of benzene in the landfilled sludge would be 0.19 mg/kg.
5-44
-------
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U.S. EPA. 1984c. Health Assessment Document for Chloroform. Office of
Health and Environmental Assessment, Environmental Criteria and Assessment
Office, Research Triangle Park, NC. EPA 600/8-84-004A. NTIS PB 84-195163.
U.S. EPA. 1984d. Guidelines for Deriving Numerical Aquatic Site-Specific
Water Quality Criteria by Modifying National Criteria. Environmental
Research Lab., Duluth, MN. EPA/600/3-84/099. NTIS PB 85-121101/REB.
6-9
-------
U.S. EPA. 1985a. Environmental Profiles and Hazard Indices for Constitu-
ents of Municipal Sludges: Office of Water Regulations and Standards,
Washington, DC.
U.S. EPA. 1985b. Technical Support for Development of Guidance on Hydro-
geologic Criterion for Hazardous Waste Management Facility Location. Office
of Solid Waste, Washington, DC.
U.S. EPA. 1985c. Health Assessment Document for Polychlorlnated D1benzo-j>-
dloxln. Office of Health and Environmental Assessment, Environmental
Criteria and Assessment Office, Cincinnati, OH. EPA 600/8-84/014F. NTIS PB
86-122546.
U.S. EPA. 1985d. Drinking Water Criteria Document for Nickel,, Prepared by
the Office of Health and Environmental Assessment, Environmental Criteria
and Assessment Office, Cincinnati, OH, for the Office of Drinking Water,
Washington, DC. EPA/600/X-84/193. NTIS PB86-117801.
U.S. EPA. 1985e. National Primary Drinking Water Regulations; Synthetic
Organic Chemicals, Inorganic Chemicals and Microorganisms; Pro- posed Rule.
(40 CFR Part 141) Federal Register 50(219): 46936-47022.
U.S. EPA. 1986a. Guidelines for Carcinogen Risk Assessment. Federal
Register 51(185): 33992-34003.
U.S. EPA. 1986b. Guidelines for Health Risk Assessment of Chemical
Mixtures. Federal Register 51(185): 34014-34025.
6-10
-------
U.S. EPA. 1988. Reference Dose (RfD): Description and Use 1n Health Risk
Assessments. Integrated Risk Information System (IRIS). Online.
Intra-Agency Reference Dose (RfD) Work Group, Office of Health and
Environmental Assessment, Environmental Criteria and Assessment Office,
Cincinnati, OH. February.
U.S. EPA. 1989. Development of Risk Assessment Methodology for Land
Application and Distribution and Marketing of Municipal Sludge. Prepared by
the Office of Health and Environmental Assessment, Environmental Criteria
and Assessment Office, Cincinnati, OH, for the Office of Water Regulations
and Standards, Washington, DC. EPA/600/6-89/001. NTIS PB90-135740/AS.
Van Genuchten, M.T. 1985. Convective-dlspersive transport of solutes
Involved 1n sequential first-order decay reactions. Comput. Geosd.
11: 129-147.
Van Genuchten, M.T. and W.J. Alves. 1982. Analytical Solutions-of the One-
Dimensional Convective-D1spers1ve Solute Transport Equation. U.S. Dept.
Agrlc. Tech. Bull. No. 1661. 151 p. As referenced in Van Genuchten, M.T.
1985. Convective-d1spers1ve transport of solutes involved 1n sequential
first-order decay reactions. Comput. Geosci. 11:129-147.
Van Genuchten, M.T. and P.J. Wierenga. 1976. Mass transfer studies In
sorb- 1ng porous media. I. Analytical solutions. Soil Sci. Am. 0. 40:
473-480. AT123D: Analytical Transient One-, Two-, and Three-D1mens1onal
Simulation of Waste Transport in the Aquifer Systems. ORNL-5602,
Environmental Sciences Division, Pub. No. 1439.
6-11
-------
Wang, S.T., A.F. McMillan and B.H. Chen. 1977. Analytical model of disper-
sion in tidal fjords. J. Hydraulic Div., ASCE. 103: 737-751. As
referenced in Yen, G.T. 1981. AT123D: Analytical Transient One-, Two-,
and Three-Dimensional Simulation of Waste Transport in the Aquifer System.
ORNL-5602, Environmental Sciences Division, Pub. No. 1439. Supplied at
earlier date.
Warrick, S.T., J.W. Biggar and D.R. Nielsen. 1971. Simultaneous solute and
water transfer for an unsaturated soil. Water Resour. Res. 7: 1216-1225.
As referenced in Yeh, G.T. 1981. AT123D: Analytical Transient One-, Two-,
and Three-Dimensional Simulation of Waste Transport in the Aquifer System.
ORNL-5602, Environmental Sciences Division, Pub. No. 1439. Supplied at
earlier date.
Yeh, G.T. 1981. AT123D: Analytical Transient One-, Two-, and Three-Dimen-
sional Simulation of Waste Transport in the Aquifer System. ORNL-5602.
Environmental Sciences Div., Pub. No. 1439. Oak Ridge National Laboratory,
Oak Ridge, TN.
Yeh, G.T. and Y.J. Tsai. 1976. Analytical transient three-dimensional
modeling of effluent discharges. Water Resour. Res. 12: 533-540. As
referenced in Yeh, G.T. 1981. AT123D: Analytical Transient One-, Two-,
and Three-Dimensional Simulation of Waste Transport in the Aquifer System.
ORNL-5602, Environmental Sciences Division, Pub. No. 1439.
6-12
-------
APPENDIX A
COLUMN METHOD FOR DETERMINING RETARDATION FACTOR (RF)
AND DISTRIBUTION COEFFICIENT (Kd)
A-l
-------
A.I. SCOPE AND APPLICATION
The column methods described herein can be used to experimentally
determine the velocity of a contaminant through a column of porous
soil/rock. The method is directed to measurement of a retardation factor
(the ratio of water velocity to contaminant velocity). A distribution
coefficient can subsequently be derived based on the porosity and density of
the soil/rock matrix. The method is applicable to any porous; media through
which water-borne contaminants may flow. Water is passed through a column
of the porous media on a once-through or recirculating basis. Contaminant
is introduced continuously or as a spike. The time of travel for the
contaminant is determined by measuring contaminant in effluent volumes. The
result is compared to the velocity of water through the column. The ratio
of the two values is defined as the retardation factor.
A. 2. THEORY
RF = V /V
gw c
The column method, which measures the migration velocity of a
contaminant (V ) relative to the groundwater velocity (V ), provides a
c gw
retardation factor (RF) according to the following equation:
(A-l)
However, when a measurement is made to determine the value of a particular
contaminant retardation factor in a rock/groundwater system, the solution's
chemical composition (including pH, Eh, cations and anions), the rock's
characteristics (chemical composition, mineralogy, surface area, cation and
anion exchange capacities) and the equilibrium between rock and groundwater
should also be considered. These parameters are important because they can
greatly affect the measured value of RF.
A-2
-------
Two common expressions used to describe equilibrium adsorption reactions
are:
S =
abC
and
1 -i- aC
J/n
(A-2)
(A-3)
S = KC
where
S = contaminant concentration sorbed on the rock (yg/g)
C = contaminant concentration in solution (yg/ma.)
and a, b, K and n are constants.
These equations (Equation A-2 after Langmuir; Equation A-3 after
Freundlich) may describe the relationship between S and C for a given solid
and solution composition at a constant temperature (often called adsorption
isotherms). Both equations are commonly used for an empirical description
of experimental adsorption data.
When contaminant concentrations are small, such that aC is <1 in
Equation A-2, the isotherm equation reduces to:
S = abC (A-4)
The exponential constant 1/n in Equation A-3 is usually close to unity,
and that equation, too, can be approximated by:
S = KC (A-5)
Both Equations, A-2 and A-3, can then be approximated by:
S = KdC (A-6)
where the constants ab and K can be taken as the Kd. It is important to
remember that Equation A-6 is usually an approximation and that it holds
only under the conditions mentioned above.
A-3
-------
When a linear isotherm, such as that given by S = KdC, can be used to
describe the adsorption reaction, the transport equation for a contaminant
in equilibrium with both rock and water in a one-dimensional porous medium
flow path is:
where
D
x
e
!£
at
= D
3x2
(A-7)
ax
gw
dispersion coefficient (cm /sec)
the distance along the flow path
groundwater velocity (average pore-water velocity, cm/sec)
JW
3
b = bulk density (g/cm )
e = porosity of the porous medium
The expression that relates the Kd to the retardation factor,
RF = [1 + (b/6) Kd] (A-8)
can be substituted into Equation A-7 to obtain a simpler form for the
transport equation:
!£ = D i!£
at 3x2
+ V,
gw
ac
ax
(A-9)
If a spike of contaminant is added to the groundwater as it enters the
column, adsorption delays the elution of the peak until RF pore volumes have
been eluted. The pore volume or void volume of a column is given by the
porosity of the porous media (e) times the total column volume (CV). The
retardation factor can then be calculated from the ratio of the volume
required to elute the contaminant's peak (or maximum concentration) to the
pore volume of that column.
If the porosity is unknown, it can be calculated from:
6 = 1 - b/p (A-10)
A-4
-------
where p is the average density of the individual particles used to pack
the column.. An experimental check on the calculated pore volume can be
obtained by the elution of a nonsorbing element, which will have a maximum
concentration at exactly one pore volume.
In summary, the transport equation for contaminant migration used in
most safety assessment models utilizes a linear adsorption isotherm
(Equation A-6). Adsorption of the contaminant results in a lower migration
velocity for the contaminant than that of the groundwater: Vr = V /RF.
gw
Generally, this is true only when the groundwater composition, rock chemical
composition and temperature do not vary (i.e., they are at equilibrium).
A.3. INTERFERENCES
Interferences of two types may occur in the column method:
(1) interference in the analysis of eluent for the contaminant of interest,
and (2) interaction of the contaminant with the apparatus or column
material. In the former case, interferences are identified in the methods
prescribed for conducting the analysis required to monitor water for the
contaminant. In the latter case, tubing, pumps and column materials must be
selected that are compatible with the contaminant of interest. If
compatibility cannot be determined from analytical laboratory or materials
handling handbooks, a simple laboratory test should be conducted as a blank
run. Results of the blank run will indicate if the apparatus itself is
retarding or removing contaminant.
A-5
-------
A.4. APPARATUS AND MATERIALS
Equipment requirements vary with the selection of high- or low-pressure
systems in a single-pass or recirculating mode.
A.4.1. CONTACT COLUMN
A contact column is required to hold the soil/rock matrix during the
contact period. A typical low-pressure configuration is depicted in Figure
A-l. The column must be constructed of material that will withstand the
intended operating pressures and not interact with the groundwater, the
contaminant or the soil/rock matrix. For low-pressure experiments, a clear,
inert plastic is desirable because it permits direct observation of the
column, which will help identify problems with changes in the porous media
or bubble entrapment. The upflow configuration is preferred to facilitate
bubble migration out of the column. A double layer of screen (inert
material such as plastic) should be placed at the ends of the column to
disperse flow and reduce the end-cap volume while holding the matrix in
place.
The column diameter should be at least 30 times the average particle
size of the porous media. The column length should be at least 4 times the
column diameter. The column volume should also be selected such that
uncertainty about the volume of end-caps and tubing does not greatly affect
the estimate of pore volume.
A.4.2. SYSTEM LAYOUT -- LOW-PRESSURE METHODS
Low-pressure column studies require the use of a fluid reservoir, a
fluid delivery system, a column and an effluent collection system. Contact
with groundwater may be accomplished in a single pass or through use of a
recirculating system.
A-6
-------
Effluent Solution
Tubing
Tubing Connector Nipple
O Ring
End Cap
Screen
Column Body
Influent
FIGURE A-l
A Detailed View of the Column Used for Low-Pressure
Column Retardation Studies (Single-Pass or Recirculating)
A-7
-------
A.4.2.1. Single-Pass Column Method. A schematic of the apparatus needed
for a single-pass, low-pressure column method is illustrated in Figure A-2.
The reservoir can be constructed of any suitable, nonreacting material for
maintaining influent solution. If volatile contaminants are to be studied,
an open reservoir will not be suitable unless contaminants are injected in
line as a spike. If steady feed methods are employed, a diaphragm system
may be required to prevent volatile losses to the atmosphere.
The groundwater velocity through the column is controlled by the
hydraulic head gradient [pressure difference between the column's inlet and
outlet (AH) divided by the column length (L)] and the hydraulic
conductivity of the porous media (K) according tot'
AH
L
Vgw - K
(A-ll)
A pump is not required for/nonvolatil.e. systems if the reservoir is elevated
above the column outlet: Such a gravity feed system is practical for heads
of up to 50 cm of water. At greater heads, the physical dimensions of the
apparatus become limiting, and a pump is more desirable.
The hydraulic conductivity of the soil/rock matrix may also constrain
the size/configuration of the apparatus. If small columns (~5 cm) are
employed at a head of H = 50 cm water, the practical upper limit to the
hydraulic head gradient for a gravity feed system is +H/L = 10 cm water/cm
of column. The minimum velocity (Equation 4-10) should be 3x10
cm/sec, which limits the system to samples having values of K>3xlO
cm/sec. Less permeable media (K<10~5 cm/sec) will require a pump.
Low-pressure syringe and peristaltic pumps are available that will maintain
flow rates over a range suitable, for :controlling velocities in experiments
on relatively permeable'columns.' / ..,•..• •> , :
A-8
-------
Groundwater
Reservoir
Column
Apparatus
Spike
Injection
Valve
nnnnn
FIGURE A-2
Apparatus Needed for a Low-Pressure, Single-Pass
Column Retardation Study
A-9
-------
The effluent fraction collector can be obtained commercially or may
consist of a test-tube rack with tubes that are changed manually. If an
automated collector is employed, it should be adjusted to receive
small-volume increments (i.e., v
-------
Groundwater
Reservoir
Pump
88
I
Sampling
Port
Column
FIGURE A-3
Apparatus Needed for a Low-Pressure Recirculating
Column Retardation Study
©Registered Trademark of E.I. duPont deNemours and Co., Wilmington, DE.
A-l 1
-------
Groundwater
Reservoir
High-
Pressure
Pump
Spike
Injection
Valve
Confining
Pressure
Pump
"Pressure*
Transducer
Compressed
Air
Pressure
Transducer
Throttle
Valve
FIGURE A-4
Apparatus Needed for a High-Pressure
Column Retardation Study
®Registered Trademark of E.I. duPont deNemours and Co., Wilmington, DE.
A-12
-------
The influent pump must maintain high pressures to force liquid through
low-permeability samples at a relatively constant velocity. The maintenance
of a constant velocity is complicated by fluctuations in permeability over
time. Constant-flow-rate pumps can accommodate decreases in permeability by
increasing the pressure gradient along the column. However, a maximum
pressure-setting control is necessary for safety considerations.- When that
pressure is reached, further declines in permeability will result in
decreased groundwater velocity.
High-pressure systems are often applied for rock systems of low
permeability. When rock cores are sufficiently impermeable, the groundwater
may flow around the core down the edges of the column rather than through
the sample. To prevent such short ciWcuiting, the core can be cast in an
epoxy jacket that bonds to the rock surface(and forms a column wall. Spike
injections of contaminant are most commonly employed in high-pressure
systems.
A.5. REAGENTS
A.5.1. GROUNDWATER
To the extent possible, groundwater representative of the site of
interest should be utilized. If natural groundwaters are not available,
they can be synthesized based on key parameters such as total dissolved
solids, conductivity, ionic strength, pH, Eh and total organic carbon.
Barring the availability of good data, distilled water can be employed to
represent meteoric water. Regardless of the source, the water should be
analyzed to determine the presence or absence of the contaminant of interest.
A-13
-------
Special attention must be directed to maintaining the redox or Eh status
of the leaching solution. The solubility of metals is greatly affected by
changes in redox potential because of the presence of species couples, such
as S /S04 2, which can produce low-solubility metal salts (i.e., the
sulfides). The dissolution and/or evolution of gasses, especially
atmospheric oxygen, can greatly .affect redox potential. As a consequence,
measures should be taken to maintain leaching solutions at the desired redox
potential values. Common measures include:
o Purge oxygen from the air space above leaching solutions by
maintaining a nitrogen blanket.
o Employ a redox buffer in the leaching solution. One such buffer
is the pyrogallol-fe+2 complex. The concentration of the
two species is. selected on the basis of the desired Eh level.
o Prepare the leaching solution fresh daily and monitor Eh before
and after use of each batch.
A.5.2. CONTAMINANT
A clean source of the contaminant of interest is required to prepare
spikes or continuous-feed solutions. Certified materials should be
utilized. Spikes should be prepared as aqueous solutions prior to injection
to eliminate problems with solution kinetics. A purity check is advised
here. For organics, shelf-life is limited and, therefore, purity checks
should be conducted periodically.
A.6. SAMPLE COLLECTION, PRESERVATION AND HOLDING
Samples should be collected serially with a fraction collector or by
manual replacement of sample vials at the effluent port. Change-out time
should be selected to accumulate a sample volume <1/20 of a pore volume.
A-14
-------
Sample preservation should be done as normally prescribed for the
contaminant of interest. If preservatives are indicated, the proper amount
should be added to the sample vial and, where necessary, calculations made
to account for the added volume of fluid.
Special precautions are required for collection/preservation of volatile
contaminants. In the case of cyanides, an alkaline receiving solution in
the sample vial can be used to prevent vapor loss. For organic volatiles,
direct feed to the analytical instrument or provisions for collection in a
closed container are necessary. Holding times should be minimized.
A.7. PROCEDURE
Select the system configuration on the basis of the materials of
interest and the availability of apparatus. High-pressure systems are
required if low-permeability matrices such as rock cores are to be
evaluated.
Assemble the system sizing the column so that diameter is >30 times the
maximum particle diameter and column length is >4 times column diameter. In
all cases, the column volume should be greater than the dead volume (sum of
tubing, end-caps, sample-holding screens, etc.).
If an intact core is to be evaluated, the column must be fitted to the
core in such a manner that side flow is minimized. For low-permeability
cores, an epoxy jacket may be cast around the core. For loose aggregates,
the material must be added to the column and packed to a density
representative of natural conditions. This can be accomplished mechanically
or by repeated pulses with uncontaminated groundwater.
A-15
-------
The height of the groundwater reservoir or the pump size/speed should be
selected to accomplish the desired groundwater velocity. To reduce the
effects of diffusion, select conditions such that:
Vgw > 1.6 x 10 /L cm/sec
where
V
gw
groundwater velocity in cm/sec
L = length of the column in cm
Calculate the number of mass transfer units (n) according to:
(b) (Kd) (L) (Sk)
n =
(e) (Vgw)
(A-12)
where
b
Kd
bulk density of the soil/rock matrix (g/cm3)
distribution coefficient in
contaminant/ms. groundwater)
contaminant/g soil)/(jjg
gW
length of the column in cm
porosity of the soil matrix (dimensionless)
groundwater velocity in cm/sec
Sfc = sorption rate constant (sec"1)
to determine if equilibrium is to be expected. In general, 90% of
equilibrium is attained when n = 20, while only 50% is reached when n = 3.
If a single-pass system is employed, make up a spike solution such that
the concentration approximates contaminant levels of interest and has a
total spike volume <10% of the total pore volume.
Activate the flow system and observe until flow conditions are steady.
Activate the sample collection system. Inject the spike and note the time
of injection. Analyze effluent samples and determine the time of passage
for the centroid of the peak. Calculate the retardation factor (RF) as:
= vn .-/effective pore volume
(J. b
(A-13)
A-16
-------
where VQ 5 is the volume eluted when 50% of the total spike has passed
(the centroid of the spike).
If a constant-feed system is employed, the feed water should be brought
to the desired contaminant concentration and allowed to equilibrate. The
system is then activated with the contaminated groundwater feed and the
effluent analyzed until the effluent concentration is one-half the influent
concentration (C = 1/2 C ). The volume of eluent at the time C = 1/2 C
o o
is defined as V_ _ and can be used to calculate RF according to:
RF = VQ 5/effective pore volume (A-14)
If a recirculating system is employed, the feed groundwater is brought
to the desired contaminant concentration and flow initiated. The effluent
is monitored until effluent concentrations are equivalent to the influent.
At that time, the volume and concentration of eluent are measured to
determine the total mass of contaminant adsorbed on the column. This is
used to calculate the distribution coefficient (Kd) according to
Kd = S/Cf (A-15)
where
S = concentration of contaminant on soil/rock (ng/g) or
determined by mass of contaminant removed over mass
of the soil/rock core
C = concentration of feed water
RF is then calculated from Equation A-8.
A.8 CALCULATIONS
Methods for determining the distribution coefficient from column
adsorption studies depend on the contact system employed. If a spike feed is
utilized, it is necessary to determine when half of the mass of contaminant
A-17
-------
has passed through the column. This is accomplished through analysis of
effluent concentration data. Each sample of effluent is analyzed for the
contaminant and results plotted in terms of concentration (vertical axis)
and column or pore volumes of effluent (horizontal axis). The spike will
appear as a peak in the effluent with width and height determined by the
column dimensions, water velocity and attenuation. The area under the peak
represents the total mass of contaminant in the effluent. If the peak is
symmetrical, the centroid lies at a point directly below the maximum
concentration. The cumulative pore volume at that point is defined as
V , or the volume required for half of the spike to pass from the
0.5
column. If the peak is not symmetrical, the centroid must be located. The
centroid is defined as the vertical line dividing the area under the curve
into two equal portions. Once again, the intersection of the vertical line
with the horizontal axis defines V . The two cases are illustrated in
\j • 3
Figure A-5. The retardation factor (RF) is calculated from Equation A-8.
When a constant-feed system is utilized, the plot of concentration and
effluent volume represents a breakthrough curve, as illustrated in
Figure A-6, rather than a peak. For this system, VQ 5 is selected as the
volume at which effluent concentrations are half of the feed concentrations,
or C/C = 0.5. Once again, Equation A-8 is applied to determine the
o
retardation factor.
Once the value for RF has been determined, the distribution coefficient
(Kd) can be calculated from the conversion of Equation A-8:
Kd = (RF - l)/e/b (A-16)
where
e = porosity of the soil column (dimensionless)
b = bulk density of the soil in the column (g/cm3)
A-18
-------
O
.80-
.70-
.60-
.50-
.40-
.30-
.20-
.10-
B. Symmetrical Spike
Effluent Volume
FIGURE A-5
Selection of VQ from Spike Elution Data
A-19
-------
o
§
10
Effluent Volume
FIGURE A-6
Selection of V. _ from Continuous-Feed Data
U. D
A-20
-------
If a recirculating system is employed, Kd can be determined directly.
The influent and effluent lines are analyzed continuously for the
contaminant of interest, and the eluent volume is monitored. The
concentration and volume are recorded at the time when influent and effluent
concentrations are equivalent. The total mass of contaminant (M ) in the
system is defined as:
M = V C
T wo
(A-16)
where
VM = total volume of ?solution in the apparatus (9.)
CQ = initial concentration of contaminant in the solution (mg/8,)
The mass of dissolved contaminant (M_) at the end of the procedure is
defined as:
Mc = V C.
S w f
(A-18)
where Cf = final concentration of contaminant in solution (mg/a,).
Therefore, the absorbed mass of contaminant (Ma) is defined as:
Ma = M - M
I O
- Vw
-------
Combining Equations A-18, A-19b, A-20 and A-21, the distribution coefficient
(Kd) is calculated as:
Kd = (V /V )[C - CJ/C_]/(b/e)
W S 0 ft
(A-22)
A.9. REFERENCES
Material in this appendix is derived from the following references:
Relyea, J.F. 1981. Status report: Column method for determining
retardation factors. U.S. Dept. of Energy, Richland, WA. PNL--4031, UC-70.
Relyea, J.F. 1982. Theoretical and experimental considerations for the use
of the column method for determining retardation factors. Radioactive Waste
Management and the Nuclear Fuel Cycle. 3: 151-156. (Modified)
A-22
-------
APPENDIX B
INPUT PARAMETERS FOR CONTAMINANTS OF INTEREST
B-l
-------
B.I. DISTRIBUTION COEFFICIENTS
Distribution coefficients are required to determine how a contaminant
will partition itself between the soil particles and the soil water. The
distribution coefficient (Kd) is defined as:
Kd = S/C (B-l)
where
S = concentration of contaminant on soil (mg/kg)
C = concentration of contaminant in water (mg/S,)
The concept of Kd is a gross simplification of attenuation of inorganic
contaminants in soil. Precipitation chemistry is an important factor in
attenuation over and above adsorption and exchange. Precipitation does not
yield a solution concentration in proportion to the mass of contaminant in
the system. As a consequence, the use of a Kd is most valid at low
contaminant concentration levels where contaminants do not exceed solubility
thresholds.
For organics, the Kd concept is more broadly useful because adsorption
accounts for most soil attenuation. In the case of organics, Kd is
calculated from the distribution as a function of organic carbon content of
the soil (K ) and the fraction of soil (f ) consisting of organic
oc oc
matter as follows:
Kd = (K )(f ) (B-2)
oc oc
If values for K have not been determined experimentally, equations are
available that relate K to octanol/water partition coefficient data —
oc
(K ) or solubility.
ow
Table B-l is provided to assist the analyst in selecting Kd values for
contaminants of interest. Values for, inorganic contaminants were derived
B-2
-------
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B-6
-------
from the literature for sandy and sandy-Toam .soils. No difference is
anticipated between unsaturated and saturated soils. The analyst should
select the soil condition most closely matched to soils found on the site
for selection of the Kd. .,...'
For organic contaminants, the Kd is a function of organic content in
soil. As a consequence, the analyst has two options:
1. If the organic content of the soil on the site is known, the
Koc value should be selected from Table B-l and the Kd
calculated from Equation B-2.
2. If the organic content of the soil on the site is not known,
the soil classification should be matched with those soil types
provided in Table 8-1 and the associated Kd value selected.
It is assumed that subsoils in the aquifer will not have organic matter and,
therefore, the Kd for organics in the saturated zone is equated to zero.
This is conservative in that research suggests that at low organic levels
(i.e., <0.1%), organics interact with clay minerals. However, these
interactions are not well understood and no means of prediction is currently
available. Therefore, retention in the saturated zone is not considered at
this time.
Whenever specific Kd values are available for the on-site soil, they
should be employed in place of the values provided in Table B-l. Use of
such data should be accompanied by detailed documentation on how they were
derived. ,; :
B.2. HENRY'S LAW CONSTANTS
The Henry's Law Constant allows one to calculate vapor concentrations
over a solution as a function of the contaminant's concentration in the
solution. If Henry's Law Constants (H) have not been derived
experimentally, they are estimated according to:
B-7
-------
H + Pvp/S (B-3)
where
P » vapor pressure of contaminant (atm)
S = solubility of contaminant in water (mol/m3)
Both S and P need to be measured at the same temperature. Hence, if
vapor pressures are given at a different temperature, they must be
adjusted. The Henry's Law Constant can also be determined from
thermodynamic data describing the free energy of solution if such data are
available. This approach considers the reaction for dissolution of a gas:
A(l) = A(g) (B-4)
The equilibrium constant for the above reaction is:
Ks ,=
where (A(g)} is the activity of constituent A in the atmosphere and
(A(l)} is the activity in solution. By definition, the activity in the
atmosphere is equal to the partial pressure and the activity in solution is
equal to the concentration in solution, multiplied times an activity
coefficient. Equation B-5 can then be rewritten as:
PA
Ks = _ ft
where
P.
(B-6)
, [A(1)JYA
= partial pressure of constituent A (atm)
[A(l)] = concentration of A in solution (M/S.)
Y. = activity coefficient of A (dimensionless)
Because PA/[A(1)] is the Henry's Law Constant, Equation B-6 can be
rewritten as:
H = Ks
(B-7)
B-8
-------
TABLE B-2
Activity Coefficients for Species of Various Charge
for Various Ionic Strengths
I (M/fc)
0.01
0.02
0.03
0.04
0.05
0.06
0.07
0,08
0.09
0.10
0.12
0.14
0.16
0.18
0.20
0.22
0.24
0.26
0.28
0.30
0.32
Y0a
1.00
1.00
1.01
1.01
. 1.01
1.01
1 .02
1.02
1 .02
1 .02
1.03
1.03
1.04
1.04
1.05
1.05
1.06
1.06
1.07
1.07
1.08
Y±lb
0.901
0.867
0.844
0.825
0.810
0.797
0.786
0.776
0,767
0.776
0.764
0.755
0.747
0.740
0.734
0.728
0.724
0.720
0.716
0.713
0.710
Y±2C
0.658
0.565
0.507
0.464
0.431
0.404
0.382
0.362
0.346
0.363
0.325
0.324
0.311
0.299
0.290
0.281
0.274
0.268
0.263
0.258
0.254
B-9
-------
TABLE B-2 (cont.)
I (M/!l)
0.34
0.36
0.38
0.40
0.42
0.44
0.46
0.48
0.50
Yoa
1.08
1.09
1.09
1.10
1.10
1.11
1.11
1.12
1.12
Y±l
0.708
0.706
0.704
0.702
0.700
0.699
0.698
0.697
0.696
Y±2C
0.251
0.248
0.245
0.243
0.241
0.239
9.238
0.236
0.235
- Activity coefficient of uncharged species
« Activity coefficient of singly charged species
Cyj-2 - Activity coefficient of doubly charged species
B-10
-------
The activity coefficient depends on the ionic strength of the solution.
Representative values are given in Table B-2.
The equilibrium constant, Ks, can be calculated from the free energy of
reaction B-4:
-------
For charged species and 0.1 < I < 0.5:
-logy = 0.5 Z= / VI
I
- 0.21)
(Butler, 1964)
For unchanged species
log YQ = KI
where K is a constant. Unless otherwise given, K = 0.10 as suggested by
Butler (1964).
Values for H and H' were found in the literature or derived for the
contaminants of interest and are listed in Table B-3. The references
indicate where the values or the inputs for derivation of values were
obtained. The notes specify the method of derivation when published values
were not found.
B.3. POROUS MEDIA HYDROL06IC PROPERTIES
The methodology for evaluating disposal of municipal sewage sludges
requires the input of various site-specific values related to hydrologic
flow in soils and other geologic media. Some of these values must be
determined by direct measurement, while others can be selected from reported
values for given soil types or aquifer media. The following tables and
figures present typical values to assist the applicant and/or reviewer in
determining the reasonableness of values derived for specific applications:
o Table B-4 provides typical values for the slope df the moisture
retention curve for soils that may be found in the unsaturated
zone.
o Figure B-l provides ranges of values for saturated hydraulic
conductivity of different aquifer media.
o Table B-5 provides ranges of values for porosity of unsaturated
and saturated zone media.
B-l 2
-------
TABLE B-3
r™c+an
Constants
Con^tants <"> an<* Dimensionless Henry's Law
1) (Assumed Temperature: 20°C)a for Selected Contaminants
Contaminant
Aldrin
Arsenic
Benzene
Benzo(a)anthracene
Benzo(a)pyrene
Bis(2-ethylhexyl)
phthalate
Carbon tetrachloride
Cadmium
Chlordane
Chloroform
Chromium
Cobalt
Copper
Cyanide
DDT/ODE/ODD
2,4-Dichlorophenoxy-
acetic acid
Dieldrin
Dimethylnitrosamine
Fluoride
H
(atm-ma/mol)
1.4xlO~5
NV
5.5xlO-3
NV
NV
1.0
2.3xlO-2
NV
0.59
4.8X10-3
NV
NV
NV
1.9xlO~3
3.8x10-3
9xlO~s
2x1 0~7
4.9X10"4
(PH = 6):
1x10-'
(PH = 7):
IxlO-e
H1
(Dimensionless)
6.1x10-*
NV
2.4X1Q-1
NV
NV
40
9.7xlO-i
NV
24
2.0X10-1
NV
NV
NV
0.082
1.7xlO-3
3.7X10-3
8.9xlO-6
2.0xlO-2
(PH =6):
4.4x10-6
(PH = 7):
4.4x10-7
Reference
Lyman et al., 1982
Lyman et al., 1982
U.S. EPA, 1985afa
U.S. EPA, 1985ab
Lyman et al., 1982
U.S. EPA, 1985ab
Lyman et al., 1982
c/ for 10°C
Lyman et al., 1980
Oawson et at. , 1980b
Lyman et al., 1982
Dawson et al., 1980b
c/ for 10°C
B-13
-------
TABLE B-3 (cont.)
Contaminant
Heptachlor
Hexachlorobenzene
Hexachl orobutadi ene
Iron
Lead
Lindane
Malathion
Mercury
Methyl ene bis
(2-chloroaniline)
Methyl ene chloride
Methyl ethyl ketone
Molybdenum
Nickel
Nitrate
Pentachlorophenol
Phenanthrene
Phenol
Polychlorinated
biphenyls:
Aroclor 1242
Aroclor 1254
Aroclor 1248
Aroclor 1260
H
(atm-ma/mol)
7x10-=
3.7xlO-s
3.73
NV
NV
4.8x10-'
1.2X10-7
l.lxlO-2
5.1x10-7
3xlO-3
20.8
NV
NV
NV
3.4x10-*
3.9xlO-s
3xlO-7
5.6x10-*
2.7xlO-3
3.5xlO-3
7.1xlO-3
H1
(Dimensionless)
2.9X10-3
1.5X10-3
122
NV
NV
2.2xlO-5
5xlO-6
4.8X10"1
2.1X10-6
K3X10-1
900
NV
NV
NV
1.5x10-*
1.7xlO-3
1.2xlQ-s
2.4xlO-2
1.2X10-1
1.6X10-1
S.OxlQ-1
Reference
Dawson et al., 1980b
U.S. EPA, 1985ab
Verschueren, 1983b
Lyman et al., 1982
Dawson et al., 1980b
Lymam et al., 1982
U.S. EPA, 1985ab;
SRI, 1984&
Lymari et al . , 1982
U.S. EPA, 1985ab
Lyman et al., 1982
Lyman et al., 1982
U.S. EPA, 1985ab
Lyman et al., 1982
Lyman et al . , 1982
Lyman et al., 1982
Lyman et al . , 1982
Selenium
NV
NV
B-14
-------
TABLE B-3 (cont.)
Contaminant
Tetrachloroethylene
Toxaphene
Trichloroethylene
Tricresyl phosphate
Vinyl chloride
Zinc
H
(atm-m3/mol)
8.3xlO"3
5.4x10-2
1x10-2
1.5x10-2
2.4
NV
H1
(Dimensionless)
3.4xlO-i
2.2
4.2X10-1
0.61
99
NV
Reference
Lyman et al.,
Dawson et al . ,
Lyman et al.,
MSOSb
Lyman et al.,
1982
198Qb
1982
1982
Henry's Law Distant can be estimated by (Thibodeaux,
H, = 16 Pv (MW)
where
H'
PV
(MW)
SOL
T
(SOL) T
Henry's Law Constant (cm3/cm3)
saturation vapor pressure of the contaminant (mm Hg)
molecular weight of the compound (g/g mol)
contaminant's solubility in water (ppm)
ambient temperature (°K)
NV = A calculation of Henry's Law Constants for these materials is
not meaningful. No measurable vapor levels are anticipated.
B-15
-------
TABLE B-4
Typical Values for Slope of Soil Moisture Retention Curve*
Soil Texture
Clay
Silty clay
Silty clay loam
Clay loam
Sandy clay loam
Sandy silt loam
Silty loam
Sandy loam
Loamy sand
Sand
Value for Curve (b)
11.7
9.9
7.5
8.5
7.5
5.4
4.8
6.3
5.6
4.0
*Source: Hall et al., 1977
B-16
-------
cb
11
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Unconsolidated /f /f K K K
Deposits ^ (darcy) (cm2) (cm/s) (m/s) (qai/dav/
ECQ 1
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rJQ "D
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Conversion
1
Factors
c in
^2-» -!
2 o = J
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Permeability, k*
ft2
darcy
m/s
ft/s
gal/day/ft2
cm
r
9.29 X
1.02 X
3.11 X
5.42 X
in ft^ miiltin
110°-9 !'
10~3 1
08
06
m
ft
z
|-10S
-104
-103
-102
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•1
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io-3
io-4
10~5
io-6
10~7
r-10-3
-io-4
-io-5
-io-6
-io-7
-io-8
-io-9
io-10
io-11
io-12
io-13
io-14
io-15
r-102
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-1
-io-1
•IO-2
-io-3
-io-4
-io-5
io-6
10~7
io-8
10~9
io-10
-1
r-106
-io-2
-io-3
-io-4
"10~5
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-io-7
io-8
io-9
io-10
io-11
10~12
0-8 g-16 1Q-11 Q-13
-10s
•104
-103
-102
10
I
_1
0
0"z
0~3
.
0
j-5
0~6
O"7
Hydraulic conductivity, K
darcy m/s ft/s gal/day/ft2
X 10~3 1.01 >
X 10~11 1
V
10-4 3.35 X
10~10 5.83 X
10"
6
io-7
io-13
^ .
A \.
C108 9.80 X102 3.22 X103 1.85 X 109
C 1010 9.11 X 105 2.99 X 10s 1.71 X 1012
5 y.oo X 10 3.17 X 10 1.82 X 10
3.15 X104 3.05 X10~1 '1 574 X105
5.49 X10~2 4.72 X 10~7 1.74 X10~6 1
FIGURE B-l
Representative Values for Saturated Hydraulic Conductivity
Source: Freeze and Cherry, 1979
B-l 7
-------
.TABLE, B-5
Porosity Values for Porous Media
A. Representati ve Values for Porosity
Material
Coarse gravel
Medium gravel
Fine gravel
Coarse sand
Medium sand
Fine sand
Silt
Clay
Porosity
28%
32%
34%
39%
39%
43%
46%
42%
B. Effective Porosities for General Hydrogeologic Classifications*
Generic Classification
Effective Porosity
(Dimensionless)
Fractured Crystalline Silicates
Fractured and Solutioned Carbonates
Porous Carbonates
Porous Silicates
Porous Unconsolidated Silicates
Fractured Shale
0.01
0.10
0.10
0.01
Average Value
0.16
0.01
*Source: Shafer et a!., 1984
B-18
-------
8.4. GEOCHEMICAL CONSIDERATIONS
The following series of figures are provided to convert unsaturated zone
contaminant concentratipns to resulting saturated zone concentrations based
on geochemical interactions. Each figure addresses a specific inorganic
contaminant (arsenic, 3.1; mercury, 3.2; lead, 3.3; copper, 3.4; and nickel,
3.5). Six curves are provided for each contaminant (a-f) depicting
relations for a different set of pH and Eh conditions. The pH values
included are 6.0 and 7.0. Eh values are -200 mv, +150 mv and +500 mv.
Directions for use of the curves can be found in Section 4.3.3.1.
B-19
-------
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B.4. .REFERENCES
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Chemical Rubber Co. Yearly. Handbook of Chem1s,try and Physics. Cleveland,
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Freeze, R.A. and 3.A. Cherry. 1979. Groundwater. Prentice-Hall, Englewood
CUffs, NJ. .. •._••-.• ,.- -;•.' •• •-,. ••• <•• : • • •
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O'Melia, C.R. and W. Stumm. 1967. Aggregation of silica dispersions by
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*U.S. GOVERNMENT PRINTING OFFICE: 19 3 0 . 7 * 8 . 1 5 9/2 0 17 9
B-51
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