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OCR for page 46
Contaminants in Groundwater
Chemical Processes
2
INTROD UCTION
JOHN A. CHERRY, ROBERT W. GILEHAM, an] JAMES F. BARKER
University of Waterloo, Canada
AB STRACT
The movement of most toxic contaminants in groundwater is affected by chemical reactions that cause transfer
of contaminant mass between the liquid and solid phases or conversion of dissolved species from one form
to another. The chemical attenuation of inorganic contaminants occurs mainly by adsorption, precipitation,
oxidation, or reduction. Organic contaminants can be adsorbed or degraded by microbiological processes,
but at present little is known about their behavior, particularly under the anaerobic conditions that are
common in contaminated groundwater.
Field and laboratory studies have established that various toxic heavy metals, transition metals, metalloids,
radionuclides, and other inorganic species can be mobile or immobile in the groundwater zone, depending
on the hydrogeochemical conditions represented by the pH, the redox condition, the ionic strength, the
mineralogy, the solid-phase surface area, and the complexing capacity. Although the importance of chemical
reactions in the attenuation of contaminants is widely recognized, the capabilities for attenuation predictions
are not well developed. This is the case because the chemical processes within dynamic groundwater systems
are complex; consequently, many of the geochemical parameters in predictive models are problematic. The
prediction problem is complicated by the fact that the chemical processes are continually influenced by the
redistribution of dissolved species caused by molecular diffusion and mechanical dispersion. The complexities
of these mixing processes contribute to the difficulties in developing reliable methods for predicting the
chemical behavior of contaminants in the groundwater zone.
Contaminants can enter the groundwater zone from regional
sources such as agricultural fields on which fertilizer or pes-
ticides have been applied or from local sources such as landfills
or waste spills. Within the groundwater zone, migration of the
contaminants is influenced by advection, mechanical disper-
sion, molecular diffusion, and chemical mass transfer. To pre-
dict the behavior of contaminants in groundwater the effects
of each of these influences must be adequately represented in
a model or group of models. This chapter focuses on the pro-
cesses that cause chemical mass transfer and on the manner in
which these processes are represented in models. Emphasis is
46
placed on inorganic contaminants and, in particular, on those
elements or compounds that have concentration limits specified
in drinking-water standards. These include various transition
metals, heavy metals, metalloids, and nonmetallic constituents
such as nitrate. Nearly all of the literature describing the chem-
ical behavior of contaminants in groundwater pertains to these
inorganic constituents.
The number and quantity of organic chemicals that are pro-
duced have increased continuously since World War II. More
than 3,000,000 organic compounds are known to exist and more
than 40,000 are currently manufactured. Many of these are
hazardous or potentially hazardous. In recent years many com-
mon organic chemicals have been recognized to be hazardous
OCR for page 47
Contaminants in Groundwater: Chemical Processes
and relatively mobile in permeable groundwater systems. In
some regions widespread contamination by organic compounds
exists in shallow groundwater. Organic compounds in ground-
water can also cause major changes in the chemical behavior
of inorganic constituents because of inorganic-organic reac-
tions. Because there is a paucity of information on the behavior
of dissolved organic compounds in groundwater, only a brief
review of this subject is presented here. The emphasis in this
chapter is on the chemical behavior of dissolved contaminants
in porous geologic deposits. Only minor consideration is given
to contaminant behavior in fractured rock or in fractured non-
indurated materials. The movement of immiscible organic liq-
uids in subsurface systems is not considered in this chapter.
TRANSPORT WITH CHEMICAL MASS
TRANSFER
A common problem in the evaluation of the existing and future
quality of groundwater is the determination or prediction of
the effect of local sources of contamination. Examples of local
sources are municipal and industrial landfills, industrial la-
goons, roadsalt storage piles, septic sewage systems, mine and
mill wastes, livestock feedlots, and local spills of industrial or
agricultural chemicals. When local degradation of groundwater
quality occurs, it is usually the result of downward movement
into the groundwater zone of leachate from wastes or of spilled
chemicals or waste liquids. As the contaminant solution mi-
grates through the groundwater system, it displaces the original
groundwater. The chemical composition of the zone of contam-
ination continually changes because of dispersion, diffusion,
and chemical reactions. The chemical reactions are driven by
the changes in the chemical conditions caused by dispersion
and diffusion and by the contact of the contaminated water with
the surfaces of minerals and amorphous solids in the medium.
Chemical changes in the zone of contamination may also be
fostered by bacteria in the groundwater system.
There are two main approaches that are commonly used for
the prediction of the behavior of most chemically reactive in-
organic contaminants in groundwater. The first approach in-
volves the incorporation of a simple chemical mass-transfer
term representing adsorption in the advection-dispersion equa-
tion. This approach has the objective of predicting the advance
rate and the shape of the front of a contaminant zone emanating
from a continuous or a temporary source. The second approach
has the objective of predicting the contaminant concentrations
that will occur in the zone of contamination after chemical mass
transfer has caused equilibrium concentrations to be achieved
by precipitation, dissolution, oxidation, or reduction. The first
approach has its origins in applied chemical chromatography,
and the second was adapted from the chemistry of electrolyte
solutions by geochemists interested primarily in seawater and
surface continental waters. The processes of adsorption, dis-
solution, precipitation, oxidation, and reduction usually occur
simultaneously in zones of contaminated groundwater; there-
fore, the use of predictive methods that represent adsorption
with exclusion of the effects of the other processes or the other
processes with the exclusion of adsorption is usually based on
47
convenience of conceptualization and computation rather than
a quest for realism.
Adsorption in Advective-Dispersive Systems
Most currently used models for the prediction of transport of
nonreactive dissolved contaminants in the groundwater zone
are based on the advection-dispersion equation. The devel-
opment of this equation is described by Anderson (Chapter 2
of this volume). Adsorption is usually incorporated into this
equation in a manner based on the assumption that the con-
centration of the contaminant in the solution phase (C) is a
function of the concentration in solid phase (C), or
C = fC.
(3.1)
With this assumption, the advection-dispersion equation for
homogeneous saturated porous media is
dCl~t= Dd2C/6x2— V&Cl~x— f~p/n)Ksj3ClOt, (3.2a)
where x is the flow direction, D is the dispersion coefficient in
the direction of flow, V is the average linear groundwater ve-
locity, p is the dry bulk density of the porous medium, n is
the porosity, t is time, and Ks is the slope of the functional
relationship expressed in Eq. (3.1~. The last term on the right-
hand side of Eq. (3.2) represents the contaminant mass that is
lost from solution as a result of adsorption. When the Ks func-
tion is known, mathematical solutions for Eq. (3.2a) or for its
two- or three-dimensional forms (including representation for
heterogeneous geologic systems) can be obtained by means of
numerical methods. It is important to recognize that Eq. (3.1)
is based on the assumption that equilibrium conditions exist
between the solution-phase and soIid-phase concentrations.
Theoretically, Eq. (3.2a) is only applicable if the reactions are
instantaneous or, in practice, only if the reactions are fast rel-
ative to the groundwater velocity.
In some situations, the relation between C and C is linear,
and thus the slope of the partitioning function (Ks) becomes a
constant and is generally referred to as the distribution coef-
f~cient IKE. In this situation Eq. (3.2a) can be expressed as
dC/6t = D'd2C/6x2— V'~C/6x, (3.2b)
where D' is the effective dispersion coefficient (D' = DIR) and
V' is the rate of advance of the front of the contaminant zone
(V' = VIR) in the absence of dispersion. This front is retarded
relative to the rate of advection of the front of nonreactive
contaminants Ad. R is the retardation factor, defined as V/V'
and is represented by the retardation equation,
V/V' = 1 + (pln)K~.
(3.3)
In the absence of dispersion the front is conceptualized as a
plug-displacement front. In the presence of Gaussian disper-
sion, it represents the 50th percentile concentration level of
an advancing slug of contamination (i.e., the middle of the
dispersed front) emanating from a continuous source. The re-
tardation concept is illustrated schematically in Figure 3.1 for
laboratory conditions and for a hypothetical field situation. For
the field example, the groundwater velocity is assumed to have
little spatial or temporal variability.
OCR for page 48
48
FIGURE 3.1 Schematic illustration of the
retardation concept: the ideal laboratory case
and a hypothetical field case.
When K`' represents the partitioning relation in Eq. (3.3),
the many available analytical and numerical mathematical so-
lutions (e.g., Bear, 1972; Fried, 1975; Finder and Gray, 1977)
for the advection-dispersion equation for nonreactive contam-
inants in saturated homogeneous porous media can be used for
predictive purposes. In these models the velocity of the ad-
vancing front becomes the effective contaminant velocity,
VIR, and the dispersion coefficient becomes the effective dis-
persion coefficient, DIR.
The retardation relation and the use of the distribution coef-
ficient in the advection-dispersion equation were introduced
by Higgins (1959) into the literature on contaminant migration
in groundwater. The use of the distribution coefficient in stud-
ies pertaining to the disposal of radioactive waste became com-
mon in the 1960s and 1970s. More recently this approach has
been used in studies of the behavior in groundwater of various
adsorbed nonradioactive elements, such as transition metals,
JOHN A. CHERRY, ROBERT W. GILEHAM, and JAMES F. BARKER
LABORATORY REPRESENTATION
TRA`:£R INPUT: _
NON REACTIVE CO Into _
r ADSORBED C ~
O _
TRACER ARRIVAL c/c
tC ~ tCH I
FIELD REPRESENTATION
_
IN P UT
to T I ME ~
RETARDATION FACTOR = tC /1C
OUTPUT
PLUG FLOW
,~ ARRIVAL
,, , _
'/NON REACTIVE '/BREAKTHROUGH
1 TRACER ~ ADSORBED TRACER
I I /, /,
to tc* tC
TIME ~
RETARDATION FACTOR = DISTANCE AC / DISTANCE AB
1 r—~ ~ , PLUG F LOW
C `~ I ~ ' PC)s1TIoN
/CO ~ Y~
O . ,
A B
heavy metals, metalloids, and trace organic compounds. A1-
though the convenience of the approach is beyond dispute, its
validity as a means of developing reliable predictions of the
behavior of inorganic contaminants in actual groundwater sys-
tems is questionable in many situations.
The C = fC relation is normally determined in the laboratory
by means of batch tests in which a known mass of the geologic
medium (in particulate form) is immersed in a solution rep-
resenting the leachate or groundwater. The solution contains
a specified concentration of the contaminant of interest. After
agitation of the liquid-solid mixture for a period of hours or
days, the contaminant concentration in solution is determined
and, by difference, the concentration adsorbed on the solids is
known. When this test is repeated using different concentra-
tions of the contaminant in solution, the C = fC relation, which
is known as the adsorption isotherm, is obtained.
There are many possible functional forms of adsorption iso-
OCR for page 49
Contaminants in Groundwater: Chemical Processes
therms, a large number of which are described by Smith (1970~.
However, in studies of trace-level contaminants in geologic
media, isotherms from batch tests usually fit closely to a func-
tional relation known as the Freundlich isotherm,
C = kCa, (3 4)
where k and a are empirical coefficients. If a = 1, the isotherm
is linear, then k = K``, and Eq. (3.3) is applicable. If a ~ 1,
the concentration versus distance profile in the flow direction
is narrow and the contaminant mass in solution advances less
rapidly than would be the case for linear adsorption. If a < 1,
the concentration profile is broad and the contaminant mass in
solution advances more rapidly than in the linear case.
Much of the earlier literature on distribution coefficients and
the retardation equation pertains to cationic radionuclides, such
as radioisotopes of Sr and Cs, present in very low concentra-
tions. The isotherms for these isotopes at low concentrations
are nearly always linear. In recent years batch isotherms have
been determined for various nonradioactive elements such as
Ag, As, Cd, Cr. Pb, and Se. When present at trace concen-
trations under conditions where adsorption rather than pre-
cipitation is the controlling mass-transfer process, these ele-
ments commonly yield isotherms that in some cases are nonlinear
but that can be described by the Freundlich relation. Davidson
et al. (1976) presented the results of numerical advection-dis-
persion-based simulations of the movement of slugs of contam-
inants with different Freundlich isotherms having various val-
ues of a. They concluded that serious errors in predictions of
migration may occur when a linear isotherm is assumed for
contaminants that exist at high concentration.
Limitations of the Adsorption-Isotherm Approach
Limitations and uncertainties are inherent in the use of the
isotherm approach within the framework of the advection-dis-
persion equation. The uncertainties exist because of the het-
erogeneous nature of geologic deposits, because of geochemical
effects, and because of effects caused by the dynamics of the
transport process.
The primary mechanism of adsorption is commonly ion ex-
change, which, for univalent species, can be represented as
X+ + AS = A+ + XS
(3.5)
where X+ is the cationic contaminant in solution, XS is the
contaminant in the adsorbed states on the exchange medium
designated as S. and A+ is the resident cation initially on the
exchange medium. Application of the law of mass action pro-
vides
KeX — tA+ 1~XS]I~X+~tAS],
(3.6)
where KeX is the equilibrium coefficient for the exchange re-
action and the bracketed terms are thermodynamic concentra-
tions or activities. Conversions between activities and concen-
trations for dissolved species can be made using activity
coefficients as described below. For situations when the con-
taminant of interest is present in very small concentrations
relative to the concentration of the exchangeable cation in the
medium and when the exchange between the contaminant (X+ ~
49
and the exchangeable cation (A+) does not cause a significant
change in the activity ratio tA+~/tAS] (designated as r), it is
apparent that Kit is a constant, and
Kd = Kex/r= (XS)I(X+),
(3. 7)
where the quantities in parentheses are concentrations. The
activity coefficients are neglected.
This development of the distribution coefficient provides a
convenient basis for identification of one of the major difficulties
in the distribution-coefficient approach. Consider a situation
where a leachate or spilled liquid enters a groundwater-flow
system. The leachate contains contaminant X+, which exists
in a zone that is retarded relative to the movement of the rest
of the leachate zone. In addition, this zone contains a variety
of dissolved constituents including the competitor cation A+,
which exists in the leachate at concentrations different than the
ambient groundwater. Therefore, each segment of the porous
medium contacted by X+ has been contacted previously by the
major cations (including A+) in the leachate that travels in
advance of the retarded zone that contains X+. Because of this
contact, the occupancy of the exchange sites evolves toward a
new steady-state condition that is not represented in the normal
batch tests used to determine the distribution coefficient. It is
common practice in batch tests to use water that has major-
ion concentrations made up to represent the ambient ground-
water from the field site, or in some cases field samples of
ambient groundwater are used. Even if the aqueous solution
used in the determination of the distribution coefficient by the
batch method has the same composition as the leachate, the
condition in the field is different because the advancing zone
of contamination continuously supplies cations to the exchange
sites. In the normal type of batch test the solids are immersed
in the test solution only once. Using exchange theory and a
mixing-cell model, Reardon (1981) demonstrated that exchange
involving major cations, such as Car+ and Na+, can cause a
gradual change in the ratio r as the zone of contamination
continuously passes through the porous medium. As can be
deduced from Eq. (3.7>, this causes a progressive change in
the Kit for X+ in an advancing zone of X+ contamination. In
recognition of this difficulty, however, batch tests can be con-
ducted in a manner that more closely represents field condi-
tions. In some cases additional laboratory tests combined with
the use of models such as the one described by Reardon can
be used to show that the magnitude of change in r is insignif-
icant relative to the accuracy required of the K,, values for the
particular predictive task.
In some situations, the contact between the porous medium
and the less retarded contaminants can cause alteration of the
exchange properties of the porous medium as a result of pre-
cipitation, dissolution, oxidation, or reduction. For the move-
ment of a species controlled by adsorption to be simulated in
a realistic manner, it would be necessary to simulate the changes
in exchange properties that occur because of all geochemical
influences. This is currently beyond the scope of existing models,
except for the simplest cases involving exchange of major cat-
ions or of trace constituents with exchange properties known
for specified or measurable conditions of pH and major dis-
solved constituents.
OCR for page 50
50
Transport models that represent the exchange of major cat-
ions and that are based directly on the law of mass action were
described by Valocchi et al. (1981), who incorporated exchange
theory into a numerical advection-dispersion model, and by
Dance and Reardon (1982), who used mixing-cell models, and
Schultz and Reardon (1983~. These models have been used
successfully in the simulation of the transport of major cations
in aquifers into which wastewater or tracer solutions were in-
jected.
The use of the isotherm approach rests on the premise that
isotherms can be determined in the laboratory on samples that
are representative of the geologic materials as they exist in the
field. It is generally not feasible to collect samples of geologic
materials and transfer them to the laboratory without, in some
manner, altering their geochemical characteristics. Some of the
more obvious processes that are difficult to control or exclude
during field sampling and during batch testing include the
invasion of oxygen and the degassing of CO2. These processes
can alter the redox condition and pH of the samples, with the
potential to cause significant changes in the concentrations of
the competing cations and to cause alterations in the exchange
characteristics of the solid phase. Oxygen invasion and sample
drying can cause precipitation of iron and manganese hydrox-
ides, which have a strong affinity for adsorption of cationic and
anionic trace contaminants. Studies of batch tests (Hajek and
Ames, 1968; Routson and Serne, 1972) have shown that the
measured Kit can also be influenced by experimental factors
such as the ratio of solution to solids used in the tests. Ralyea
et al. (1980) recommended procedures for determining K,~ val-
ues of radionuclides by batch tests. Although these procedures
would tend to minimize the influence of the foregoing effects,
the effects cannot be entirely eliminated and will generally
contribute to the uncertainty in the predictions generated by
models in which K`` values are used.
The second inherent type of uncertainty in the isotherm
approach to advection-dispersion modeling pertains to the dy-
namic conditions that occur in the porous medium during trans-
JOHN A. CHERRY, ROBERT W. GILEHAM, and JAMES F. BARKER
port. Breakthrough curves that are simulated using the advec-
tion-dispersion model with linear isotherms are symmetrical
or nearly symmetrical, whereas curves obtained from labora-
tory-column experiments are asymmetrical with extended tails.
Reynolds (1978) compiled data from several published studies
in which tracers with linear batch isotherms were used to obtain
breakthrough curves from column experiments. The results
were scaled and plotted on the same graph of dimensionless
concentration versus dimensionless time (Figure 3.2~. The data
for Figure 3.2 represent the work of several investigators and
a variety of sorbates and sorbents. All the breakthrough curves
are asymmetrical and fit within a single narrow band on the
dimensionless graph. This suggests the presence of a common
factor that is currently not accounted for in the advection-
dispersion formulation.
One proposed explanation for the asymmetry of the break-
through curves invokes the presence of physical or chemical
kinetic effects. For example, Cameron and Klute (1977) and
others simulated curves of the type shown in Figure 3.2 by
the addition of an empirically derived first-order kinetic term
to Eq. (3. 2~. Van Genuchten et al. (1974) developed an equation
having a similar form but attributed the kinetic parameters to
the rate-dependent migration of the solute into zones of im-
mobile water. Reynolds et al. (1982) obtained breakthrough
curves for Sr2+ that exhibited the typical asymmetrical form
with extended tails even though the velocities used in the tests
ranged over an order of magnitude. This apparent velocity
independence of the shape of the breakthrough curves casts
doubt on the kinetic explanation of the observed tailing.
lames and Rubin (1978) showed that the transport equation
with a linear-adsorption term accurately predicted the migra-
tion of cations when the flow velocity was so slow that molecular
diffusion was the dominant influence in the dispersion process.
At higher velocities when mechanical mixing was dominant in
the dispersion process, the differences between measured and
calculated breakthrough curves were similar to that repre-
sented in Figure 3.2. The results of James and Rubin at very
NORMALIZED BR EAKTHROUGH CURVES
1 00
080
o
' 060
040
0 20
0.00
it..
a NON- REACTIVE SOLUTE DOMAIN
I I I! 11 !I i! I REACTIVE SOLUTE DOMAIN
Q6 0.8 1.0
1.2 1.4 1.6
EFFECTIVE PORE VOLUMES ( EPV)
1.8 2.0 2.2 2.4
FIGURE 3.2 Breakthrough graphs for nonreactive and adsorbed tracers in column experiments with homogeneous porous media. Each graph
represents numerous sets of experimental data from Gillham and Cherry (1982).
OCR for page 51
Contaminants in Groundwater: Chemical Processes
low velocity are consistent with unpublished diffusion data ob-
tained at the University of Waterloo, using the same materials
as those used by Reynolds et al. (1982~. The evidence suggests
that the discrepancy between measured and predicted curves
is, in some way, the result of the advection process; however,
the manner by which advection causes the discrepancy is not
known. The measured curves can be closely matched with
simulated curves when a kinetic term is included in advection-
dispersion models, but at present the kinetic parameters are
determined arbitrarily by curve fitting. An understanding of
the processes may lead to a less arbitrary means of specifying
values for the parameters, thus improving the predictive re-
liability of the models.
The asymmetrical shape and tailing of breakthrough curves
for adsorbed contaminants have considerable practical impor-
tance for predictions of the first arrivals of contaminants and
for predictions of the rates and effectiveness of purging of zones
of contamination from aquifers by means of pumping-well sys-
tems. For contaminants that are hazardous at very low con-
centration levels, prediction of the behavior of the front or tail
of a contaminant zone (where the contaminant occurs only at
extremely low concentration levels) can be more important than
predictions of the arrival of the center of mass or mid-level
concentrations that are represented by the retardation equation
fEq. (3.3~. The fact that laboratory studies provide evidence
of a lack of predictive capability for behavior of fronts and tails
does not bode well for prediction under field conditions.
Some of the uncertainties that are inherent in the use of
laboratory-determined distribution coefficients in advection-
dispersion models can be avoided if the effects of adsorption
are assessed by field tests using tracers that adsorb and others
that do not react with the porous medium. Values for the
retardation factor (R) are computed from the differences in rates
of passage through the test zone in the aquifer. Goodwin and
Gillham (1982) conducted miniature field tracer tests to obtain
radionuclide retardation factors using a device attached to the
head of hollow-stem augers used in boreholes in sand deposits.
A more common type of test involves the injection of hundreds
or thousands of liters of tracer solution into aquifers. An ex-
ample of a field test of this type is described by Ewing (19591.
These tests are relatively expensive, and, in order to obtain
results in a practical length of time, they are generally only
suitable for application in zones that have moderate or high
hydraulic conductivity and low retardation factors. Although
values for the retardation factor can be estimated from field-
tracer tests, the suitability of these values for use in predictions
that pertain to spatial or temporal scales much different than
those represented by the tracer test is problematic. When
advection-dispersion models are used in the analysis of the
tracer-test data, apparent nonlinear behavior may necessitate
the use of a separate kinetic adsorption term as a means of
obtaining a close match between simulations and the test data
or the use of nonequilibrium physical/chemical or adsorption
parameters within the dispersion term. An example of the latter
approach in a field tracer investigation is provided by Pickens
et al. (1981~. Whether these approaches provide a reliable basis
for prediction remains to be established.
Not many detailed comparisons between migrations of ad-
51
sorbed inorganic contaminants observed in the field and pre-
dicted migrations based on laboratory batch tests exist in the
literature. The few that we are aware of involve cationic radio-
nuclides in groundwater. These comparisons show moderately
good agreement between observed retardation and simulated
retardation based on data from laboratory tests (Ewing, 1959;
Jackson and Inch, 1980; Comer, 1981~. In some situations or-
der-of-magnitude estimates of contaminant retardation are all
that is required for solution of a practical problem, in which
case many of the sources of uncertainty in the predictive ap-
proach may be unimportant.
Precipitation and Solubility Controls
The behavior of many deleterious or toxic inorganic contami-
nants in groundwater is influenced by chemical precipitation
that occurs as the contaminant adjusts to solubility constraints.
The prediction of solubility constraints is based on the law of
mass action and the associated principles of equilibrium-chem-
ical thermodynamics. The objective is to predict the equilib-
rium concentration of the contaminant species of interest under
the conditions that exist or will exist in the groundwater zone.
In the conceptualization of the predictive task, the leachate or
waste solution in some initial state enters the groundwater
system and adjusts by means of various reactions to an equi-
librium state within the groundwater system.
Equilibrium is achieved as a result of one or more of the
following processes: precipitation, dissolution, oxidation, re-
duction, hydrolysis, or complexation. This approach is similar
to that described above for contaminants that are adsorbed in
that both approaches are based on the assumption of equilib-
rium; however, the two approaches are very different. The first
describes the advance of the front of the contaminant zone as
chemical mass transfer continuously occurs in response to a
changing solution-phase concentration, whereas the second de-
scribes the equilibrium concentrations that occur after chemical
mass transfer in a reaction zone has caused equilibrium to be
achieved. According to this conceptualization, chemical mass
transfer occurs in the reaction zone, and in front of the reaction
zone the contaminants are transported at the equilibrium con-
centrations established previously in the reaction zone. This is
the case if the chemical nature of the porous medium does not
change along the flow path. Although it may be possible to
specify the rate of entry of the contaminant solution into the
groundwater system, the chemical-equilibrium approach pro-
vides no information on the rate of advance of the reaction zone
because reaction rates and the availability of reactants are not
included in the description of the system. Even if kinetic terms
were included in the computational routines, so little is known
about the reaction-rate controls in contaminated-groundwater
systems that little additional predictive capability would be
gained. The computational routines for representation of com-
plex equilibrium hydrogeochemical systems have not yet been
incorporated in a practical manner in advection-dispersion
models. Excessive cost of computer time is currently one of
the major limiting factors in the development of this approach
as a practical means of prediction.
The equilibrium relation for a contaminant species controlled
OCR for page 52
52
by precipitation or dissolution is specified as
xX + bB = yY + cC + dD, (3.8)
where X is the inorganic contaminant species in the solution
phase; Y is a mineral or solid amorphous compound in which
the contaminant species is incorporated by precipitation or
from which it is released by dissolution; B. C, and D are other
elements or compounds in solution; and x, y, b, c, and d are
the stoichiometric mole numbers. From the law of mass action,
the equilibrium expression is obtained:
X = fC] · fD]/KeqfB],
(3.9)
where Ken is the equilibrium constant and the quantities within
the brackets are chemical activities, which can be converted
to concentration using well-established conversion relations
(Stumm and Morgan, 19803. If X (in the leachate or spilled
liquid) is initially above the equilibrium concentration when it
enters the groundwater system, adjustment toward equilibrium
will occur by precipitation of mineral or amorphous solids. If
X is below the equilibrium concentration, minerals or amor-
phous solids that contain X as part of the chemical structure
will dissolve when such solids are present in the system. If
they are not present, which is usually the case for most con-
taminants of interest, the condition of undersaturation will per-
sist.
For example, if Pb is the contaminant of interest, and if
PbCO3 is the solid phase that may control the solubility of Pb,
the equilibrium expression would be
Kit `, = ~pb2 + ~ ICON - I,
(3. 10)
where ~CO32-] would depend on the pH and the dissolved
inorganic carbon. The total concentration of dissolved Pb would
be represented by the sum of the concentrations of Pb-+ and
the other species of Pb in solution, which would include com-
plexes and hydrolyzed species (see next section). To determine
which solid phase would be expected to exert the major sol-
ubility control, equilibrium relations for many compounds of
Pb would be considered.
For reactions involving nearly all minerals or amorphous
inorganic compounds of interest, values for the equilibrium
constant at 25°C can be obtained from published listings, or
they can be computed from values of the standard free energies
of reaction in published compilations of thermodynamic data.
When a contaminant in a leachate or waste liquid enters the
subsurface domain, the chemical reactions cause adjustment of
the solution to the new conditions. The reactions are influenced
by the minerals that comprise the porous medium, by the
degree of mixing that occurs between the contaminant solution
and the ambient groundwater, by changes in temperature and
pressure, and, in some cases, by microbial action. To predict
the equilibrium concentration of a particular trace contaminant
by means of Eq. (3.9), it is first necessary to predict the gross
chemical composition of the contaminant solution after the
dominant reactions in the subsurface domain have proceeded
to equilibrium or to some other expected status. The gross
chemical composition is represented primarily by the major
ions such as Na+, Ca2+, Mg2+, C1-, SO42-, and dissolved
inorganic carbon. In many cases K+, Fez+, H2S, NH4+, or
JOHN A. CHERRY, ROBERT W. GILLHAM, and JAMES F. BARKER
NO3- also occur as important components of the gross chemical
composition of the water. The reactions that involve the above
constituents normally determine the ionic strength, pH, and
redox status of the groundwater. These parameters have an
important influence on the solubility of most toxic inorganic
contaminants, which usually constitute only a small percentage
of the total dissolved solids. In practice, the use of Eq. (3.9) is
the last stage in a series of predictions involving numerous
chemical reactions with various assumptions regarding the choice
and status of the reactions. So little is known about the iden-
tities and quantities of reactive minerals in many groundwater
systems that considerable speculation is often used in the rep-
resentation of the suite of reactions that control the gross water
chemistry.
Computer codes, such as the one described by Parkhurst et
al. (1980), are available for performing the computational tasks
in predictions of the gross equilibrium chemistry of ground-
water that reacts with various minerals and amorphous solids.
The available equilibrium constants and free energy data are
generally for 25°C. Extrapolations to the lower temperatures
common for most contaminated groundwater are generally ac-
complished without large uncertainties. Nearly all of the ther-
modynamic data for solid phases is derived for chemically pure
manifestations of the minerals, whereas in groundwater im-
purities are the norm. When solids precipitate in the ground-
water zone, amorphous or poorly crystalline forms commonly
exist at first and then slowly undergo conversion to more crys-
talline forms. The thermodynamic properties for the less crys-
talline forms are not well known. Some reactions for precipi-
tation, dissolution, oxidation, or reduction do not proceed quickly
to equilibrium. These deficiencies notwithstanding, the chem-
ical equilibrium approach for prediction of the concentrations
of hazardous inorganic contaminants in groundwater can often
provide useful estimates of the maximum concentration levels
that can be expected. Estimates of equilibrium concentrations
within a factor of 10 or even 100 are often relevant.
Hydrolysis and Chemical Speciation
The concentrations of toxic inorganic contaminants that are
reported in chemical analyses of groundwater normally rep-
resent the total concentrations of each element in solution. The
concentration limits specified in drinking-water standards also
are expressed in terms of total concentrations of each element
or ion of interest. In aqueous systems, however, most inorganic
contaminants exist in more than one molecular or ionic form.
These forms, or species, can have different valences and, there-
fore, different mobilities in groundwater owing to different
affinities for adsorption and different solubility controls. Knowl-
edge of the distribution of species in solution is therefore nec-
essary for consideration of the behavior of most inorganic con-
taminants in groundwater. The metals and metalloids are
particularly prone to formation of a variety of aqueous species.
These species form as a result of hydrolysis and complexation.
The simple ionic species combine with ligands to form ionic or
neutral-charge aqueous complexes. The major inorganic ligands
in contaminated groundwater are generally C1-, HCO32-, CO2,
and SO42- and in some cases NH3, NO3-, and F-. Even in
OCR for page 53
Contaminants in Groundwater: Chemical Processes
highly contaminated groundwater the metallic elements and
metalloids are rarely present at concentrations that exceed sev-
eral milligrams per liter, whereas the major ligands are com-
monly present at levels of hundreds or thousands of milligrams
per liter. The complexation of a small percentage of a major
ligand with a contaminant can result in the formation of a
complex that represents a large percentage of the total con-
taminant concentration in solution.
Hydrolysis also influences the species of occurrence of a
contaminant in groundwater. For example, dissolved Pb in
water is represented by Pb2+ and various complexes and hy-
drolyzed species,
Pb~total) = Pb2+ + PbCl+ + PbCl + PbSO4 + PbCO3
+ ... + Pb(OH)+ + Pb(OH)2 + Pb(OH>3- + ....
(3.11)
The percentage of the total Pb represented by each ligand
complex depends on the equilibrium constant for each species
and on the concentration of the available ligand. The hydrolysis
species depend on the equilibrium constant and the pH. Com-
puter codes such as that described by Ball et al. (1978) are
available for speciation calculations for many metallic contam-
inants of interest. For some species there is considerable un-
certainty in the values that are currently used to represent the
equilibrium constants. Nevertheless, attempts at evaluating the
species distributions are essential in investigations of the be-
havior of most inorganic contaminants in groundwater. If ad-
sorption or retardation experiments are conducted in the lab-
oratory under conditions where the contaminant of interest
exhibits a much different speciation than would occur under
field conditions, the laboratory result may have limited appli-
cability to field conditions. However, if one of the major species
of an element, such as Pb2+ in the example provided by Eq.
(3.11), is limited to very low concentrations by adsorption or
solubility constraints, the concentrations of complexes will also
generally exist at very low concentrations except in situations
where nearly all of the contaminant mass exists in the com-
plexed form.
Oxidation and Reduction
Redox processes (i.e., oxidation and reduction) are important
because they can cause changes in the mobility of many in-
organic contaminants. Of the 16 inorganic constituents that,
for regulatory purposes, have recommended or mandatory con-
centration limits in drinking-water supplies, 9 of these have
more than one possible oxidation state in groundwater. These
are As, Cr. Fe, Hg, Mn, Se, U. N. and S and are referred to
here as the redox elements. The latter 2 elements occur in
various ionic or molecular forms, such as NO3- and SO42-,
which are included in the drinking-water regulations. Of the
remaining elements listed in drinking-water standards, 4 can
be strongly influenced by redox processes even though they
have only one valence state in aqueous systems; these are Ag,
Cu. Cd, and Zn. The only elements in the drinking-water
regulations that are relatively insensitive to the redox condi-
tions are Cl, F. Ba, and Ra, although in some systems even
Ba and Ra are influenced indirectly by the redox conditions
53
because of reactions with SO~2- and Fe, which are redox de-
pendent.
The ionic or molecular forms of the redox elements in aqueous
systems are commonly deduced from simplified geochemical
models of equilibrium. For illustrative purposes, the results of
such computations are commonly expressed as pe-pH or Eh-
pH diagrams. These diagrams are ubiquitous in the geochem-
ical literature. They are used as guides to the redox status of
the aqueous system. Eh or pe can be used interchangeably
through a direct numerical conversion. Each has a scale that
ranges from negative to positive. Positive values indicate con-
ditions that are more oxidizing, and negative values indicate
conditions that are more reducing.
Figure 3.3 illustrates the potential for Eh and pH to control
the redox state of metals and metalloids in groundwater. Dia-
grams such as these can be used as a conceptual guide to some
of the possibilities of element behavior in groundwater. The
CFAn ( b) CADMIUM
1 ~
-0.5
1.0
0.5
-
O c
-
I,, 0 _
- 0 5 _
-1.0 l l
, , , . . .
1 3 5 7 9
pH
~ ~ , ~ :)
-1.0 , 1 1, 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1
(C ) CHROMIUM
c, (OH) :
~ C'2O'
(d ) ARSENIC
1
_ H2AS O4
l I I I 1 1 1 1 1 1 1 1
11 1 3 5 7 9 11
_.
pH
_ 2O
10
—1O
~20
10
pe
-10
FIGURE 3.3 Eh-pH diagrams for Pb, Cd, Cr, and As for 25°C and
1 atm [from Longmire et al. (1982) except for Cd]. a, Pb-C-S-O-H
system. Activity of dissolved Pb and S at 10-6 m and 1O-3 m, re-
spe.ctively. b, Cd-C-S-Si system. Activity of dissolved Cd at 1O-7 m,
C = 10-2~ m, S = Io-3~ m, Si = 1O-3~ m. c, Cr-O-H system.
Activity of dissolved Cr = 10-6 m. d, As-S-O-H system. Activity of
dissolved As and S at 10-6 m and iO-3 m, respectively.
pe
OCR for page 54
54
boundary lines between stability fields on the Eh-pH diagrams
represent transition domains between the fields rather than
abrupt discontinuities. In fields where solid phases are des-
ignated, the element would be expected to be immobile be-
cause of solubility constraints. In stability fields where an aqueous
species is designated, this species would not be constrained by
solubility if the conditions assumed for the preparation of the
diagram are applicable.
Each of the four elements considered in Figure 3.3 has Eh-
pH domains in which they will be immobile owing to insolu-
bility, if equilibrium conditions are achieved. Under extremely
low redox conditions, Pb and Cd, which have only one oxidation
state in groundwater, form insoluble sulfide minerals. At pH
levels above about 7 or 8 and at redox conditions that are not
so low, these elements form insoluble carbonate minerals. Ar-
senic, which has two possible oxidation states in groundwater,
has sulfide-mineral insolubility under very low redox conditions
and does not form a carbonate mineral under any Eh or pH
conditions. Cr. which also has two possible oxidation states in
groundwater, forms a relatively insoluble oxide under all con-
ditions except at very low pH or high Eh. In the absence of
complexing and where solubility constraints are absent, Pb and
Cd normally occur as divalent cationic species. Pb and Cd are
normally relatively immobile in permeable unfractured
groundwater zones. Nearly all common geologic materials have
a significant capability for cation adsorption.
Of the metals and metalloids that have mandated limits spec-
ified in drinking-water standards, the four that may have the
greatest potential for being relatively mobile in permeable geo-
logic materials under Eh-pH conditions that are common in
shallow groundwater are As, Cr. Se, and U. Under oxidizing
conditions and normal pH conditions (see Figure 3.3) As and
Cr exist as monovalent or divalent anionic species. This is also
the case for Se, which has sulfide-mineral insolubility under
low Eh conditions and has no severe solubility constraints under
oxidizing conditions. In groundwater that has an oxidizing re-
dox condition and has appreciable carbonate alkalinity, U in
the + 6 oxidation state occurs predominantly as anionic com-
plexes such as UO2(CO3~32- and (C03~34-. The mobility of
soluble anionic species in groundwater can be limited to some
degree because of adsorption. Anion adsorption is most likely
to be significant when the geologic materials contain oxides of
iron and aluminum. In general, there have been few field
studies of the mobility or geochemical behavior of these ele-
ments in groundwater.
Although Eh-pH relations provide a framework for consid-
eration of the behavior of redox-sensitive Elements in ground-
water, they do not lead directly to predictions of contaminant
mobility in groundwater. To identify the condition of redox
stability applicable to a particular contaminant that enters a
groundwater system, one must predict the pH, Eh, and major-
ion chemistry that will exist in the contaminated zone. Com-
puter models for equilibrium water chemistry such as the one
described by Parkhurst et al. (1980) can provide predictions of
the pH and Eh, if the initial conditions and the mineral and
amorphous solids that comprise the porous medium are spec-
ified. At present, however, there is little experience with which
to judge the predictive capability of such models for real
JOHN A. CHERRY, ROBERT W. GILEHAM, and JAMES F. BARKER
groundwater systems. The usefulness of these models in many
situations is also limited because of the lack of quantitative
information on the influence of bacteria on redox processes in
groundwater systems. It is known, however, that bacteria have
an influence on the rates at which many important redox re-
actions proceed. A further complication is that reliable natural
or contaminated-groundwater measurements of redox condi-
tions are difficult or impossible to make because of disequilib-
rium or other factors.
In many contaminated-groundwater systems, dispersion can
be a major influence on the redox state of the groundwater.
Contaminated groundwater at waste-disposal sites commonly
has a much lower initial redox state than the ambient ground-
water. Dispersion commonly causes a continual mixing of waters
that are different in chemical composition and in redox status.
As dispersion occurs, the redox and pH conditions may change,
and with these and other effects of dispersion various chemical
reactions involving mineral and amorphous solids take place.
These reactions can cause further changes in pH, the redox
condition, the gross water chemistry, and other factors. The
problem of determining the influence of dispersion on the
chemical mass transfer of contaminants in groundwater is a
particularly difficult one at present because the manner in which
dispersion influences concentration distributions at the field
scale is poorly understood, as indicated by Anderson (Chapter
2 of this volume).
Mineral Dissolution and Acid Consumption
In some situations the dissolution of minerals has a major in-
fluence on contaminant mobility in groundwater. Mineral dis-
solution may cause contaminant concentrations to increase if
the contaminants of interest are released from the minerals as
dissolution occurs, or it may cause the contaminants to be
removed from solution by adsorption or precipitation if the
dissolution of minerals results in changes in the water chem-
istry, which, in turn, cause adsorption or precipitation of other
solid phases. Examples of this latter condition have been ob-
served in aquifers contaminated by acidic leachate (pH 1.5 to
4) from mill wastes in the uranium mining districts of the west-
ern United States and northern Ontario (Taylor, 1980; Morin
et al., 1982~. At sites in the western United States where U-
mill impoundments contain ponded water or pore water at low
pH, the water has exceptionally high concentrations of tran-
sition metals, heavy metals, metalloids, and radionuclides, such
as 226Ra, 2~0Pb, Huh, and 238U. Except for the high radionuclide
concentration, similar conditions exist at many base-metal mines,
where acid-leach milling occurs or where tailings become acidic
because of the oxidation of pyrite.
Monitoring of plumes of contaminated groundwater in shal-
low unconfined sandy aquifers at leaky acidic U-tailings im-
poundments has established that considerable neutralization
commonly occurs as the acidic water moves through the aqui-
fers. This causes the front of the low-pH zone to advance at a
rate that is retarded relative to the advance of mobile constit-
uents such as C1- or SO42-. Neutralization of the acid is at-
tributed primarily to the dissolution of carbonate minerals, such
as calcite or dolomite, in the aquifers. Using calcite as the
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Representative terms from entire chapter:
groundwater zone
Contaminants in Groundwater: Chemical Processes
reactive solid phase, the neutralization process can be repre-
sented as
CaCO3 + 3H+Ca2+ + HCO3- + H2CO3. (3.12)
The main species representing dissolved inorganic carbon have
the equilibrium relation
Keg = tHCO3- ASH+ ]/tH2Co34' (3.13)
where Keq is the equilibrium constant for the dissociation of
H2CO3. Field and laboratory studies have established that on
reaction with calcite, the pH generally rises above 6 and in
some cases above 7. If neutralization causes the pH to rise
above 7 nearly all of the dissolved inorganic carbon occurs as
HCO3-, and if the pH is below 6 the dominant species is
H2CO3. Acid neutralization also occurs as a result of the dis-
solution of aluminosilicate minerals, such as feldspars. It can
also occur by ion exchange, when H+ competes successfully
with other cations for occupancy on exchange sites.
In cases where the acid neutralization occurs primarily by
the dissolution of calcite and where the pH rises above 7, the
rate of advance of the acid front can be represented in a manner
similar to that depicted by Eq. (3.3) for equilibrium adsorption
of trace contaminants:
Va/V = 1/~1 + (IMcaco3IZMH+) = 1/Ra, (3.14)
where Va is the rate of advance of the acid front, MCaco3 is the
number of moles of calcite and dolomite per unit volume of
porous medium, MH+ is the number of moles of H+ in the
water per unit volume of porous medium, and Ra is the acid-
front retardation factor.
Equation (3.14) has the same form as Eq. (3.3), and because
of this similarity the acid-front retardation factor has been in-
corporated directly into the advection-dispersion equation in
the same manner as the retardation factor for adsorption of
trace contaminants. The acid-front retardation factor has been
used in this manner by Haji-Djafari et al. (1979) and Highland
et al. (1981) in numerical advection-dispersion simulations of
the movement of acid fronts at two uranium tailings impound-
ments in Wyoming. Inherent in this approach is the assumption
that the dissolution of calcite or dolomite occurs rapidly so that
local equilibrium exists within the porous medium. If acid neu-
tralization occurs primarily by dissolution of aluminosilicate
minerals, it is much less likely that local equilibrium will occur.
Whether this use of the acid-front retardation factor in advec-
tion-dispersion models provides realistic shapes for acid-front
breakthrough curves cannot be determined from existing field
data and has not yet been evaluated by laboratory experiments.
Because the concentration of acid in solution is not dependent
on the solid-phase acid neutralization capacity, it is unlikely
that the shape of breakthrough curves will be closely repre-
sented by the model. The arrival of the midpoint of the break-
through curve, however, may be adequately represented by
the model.
BEHAVIOR OF ORGANIC CONTAMINANTS
Spurred by the awareness of potential environmental hazards
and the development of sophisticated analytical equipment,
studies of the occurrence and behavior of organic compounds
in contaminated groundwater have been initiated recently. Many
organic compounds are of environmental concern in part per
billion (ppb) or part per trillion (ppt) quantities. Faced with
these problems, research has concentrated on the 120 or so
organic compounds designated as priority pollutants by the
U. S. Environmental Protection Agency.
The major chemical or biochemical processes currently rec-
ognized as having a potential to be significant with respect to
the occurrence and migration of these compounds in hydro-
geologic regimes include sorption, chemical reaction, and bi-
ological reaction.
Sorption
Trace organic solutes, especially those that are nonpolar and
relatively insoluble, tend to be sorbed by sediments and soils.
For some polar organic solutes, these sorption processes are
essentially electrostatic. For less polar organics the process has
not been established but appears to involve weak hydrogen
bonding and, more importantly, solvation or chemical parti-
tioning similar to the distribution of solutes between two im-
miscible solvents. This is similar to the partitioning of solutes
between oil and aqueous phases and is controlled essentially
by solubility.
Sorption of organics on mineral surfaces has been docu-
mented. Field and laboratory research indicates that the sorp-
tion of organic compounds is dominantly onto particulate or-
ganic matter in the sediments. This partitioning of the solute
between the aqueous phase and solid organic matter generally
appears to reach equilibrium rapidly and is reversible; thus it
can be introduced into the chemical mass-transfer term of the
transport equation as a K`~.
The sorption of various nonionic organic solutes at trace con-
centrations onto sediments and soils has been shown to follow
an essentially lirtear isotherm and to be readily reversible for
a broad range of organic solute concentrations (Chiou et al.,
1979; Karickhoff et al., 1979; Means et al., 1980; Schwarzen-
bach and Westall, 1981~. In these studies, the slopes of the
linear isotherms (K,~s), which are referred to by some authors
as partition coefficients (
56
which involves the use of partition coefficients for mixtures of
water and octanol, has been established for soils in which the
organic carbon content exceeds 0.1 percent.
The development of this approach is reviewed by Hansch
and Leo (1979) and others. A linear relationship exists for par-
titioning of organic solutes between sedimentary organic matter
and groundwater and for partitioning between octanol and water.
The octanollwater system can be used as surrogate for the real
groundwater system in describing relative partition coeffi-
cients. The following relationships have been observed for var-
ious organic solute/sediment systems when OC > 0.1 percent:
log Koc = 1.00 log Kow—0.21 (Karickhoff et al., 1979~;
log K`'c = 0.72 log Kow + 0.49 (Schwarzenback and
log K =
oc
Westall, 1981~;
2.00 log K —0.317 (Means et al., 1980~.
OW
Kow is the octanol/water partition coefficient, values of which
can be obtained from Leo et al. (1971) or Hansch and Leo
(1979~. The variation in the slope parameter may be due to
differences in the sediment/soil organic matter or to differences
in the organic solutes investigated.
At least three areas of uncertainty must be investigated be-
fore these partition coefficients can be used with confidence in
predictive models for groundwater systems. One area is the
effect of solute competition for sorption sites. Although Kar-
ickhoff et al. (1979) found that nonpolar organic solutes sorbed
independently at low total solute concentration, it is likely that
at some higher organic solute/solid organic matter ratio sorption
will not be independent. Another problem may occur in or-
ganic-rich groundwaters such as landfill leachates. Trace or-
ganic compounds can become associated with high-molecular-
weight, dissolved organic matter, often humic or fulvic acids
(Schnitzer and Khan, 1972~. These associated organics may not
be available for sorption by matrix organic matter. Perhaps
organic partition coefficients will have to be defined for a three-
phase system: matrix organic matter, aqueous unassociated sol-
utes, and aqueous humic associated solutes. As with inorganic
sorption, there is a need for further theoretical description of
the process and additional field data and field-scale testing for
assessment of the applicability of laboratory results to field
conditions.
The low KoW~KoC relationships described above have been
determined for porous geologic materials that have appreciable
contents of particulate organic carbon, such as 0.1 wt. % or
more. Unfortunately, many very permeable sand or gravel
aquifers may have less than this amount of organic matter. It The major mechanisms of chemical transformation of organic
is these aquifers in which groundwater velocities are often compounds in aqueous systems are photolysis, oxidation, hy-
highest and that have the greatest potential for widespread drolysis, and reduction. In groundwater, photolysis is not sig-
contamination by halogenated hydrocarbons. Whether useful nificant. Callahan (1979) compiled and assessed transformations
predictive relationships can be developed for sand and gravel affecting priority pollutants in aqueous systems. Only a brief,
aquifers with very low contents of organic carbon remains to generalized discussion of the chemical reactions and the state
be determined.
Although numerous investigations have shown that dissolved
organic compounds at trace levels commonly exhibit linear
isotherms, some organic compounds that have a potential to
cause severe contamination of groundwater characteristically
JOHN A. CHERRY, ROBERT W. GILEHAM, and JAMES F. BARKER
have nonlinear isotherms. Notable in this regard are polychlo-
rinated biphenyls (PCBs), which are considered to have a po-
tential to cause adverse health effects if they are present in
drinking-water supplies at concentrations as low as a fraction
of a part per billion (ppb) level. Relative to the levels at which
PCBs are considered to be undesirable in drinking water, the
solubility of PCBs is very large and ranges from approximately
50 ppb (Haque et al., 1974) to about 250 ppb, depending on
the isomer that is being considered. Haque et al., who con-
ducted batch studies of the adsorption of PCBs (Arochlor 1254)
on several soils, obtained nonlinear Freundlich isotherms with
the value of K,~ ranging from 1.1 to 0.81. Griffen et al. (1978),
who conducted batch adsorption studies (Arochlor 1242 and
1254) of three natural soil materials, also obtained nonlinear
Freundlich isotherms with K`~ values of between 1.5 and 0.19.
These investigators noted that adsorption was favored in ma-
terials with higher organic matter and larger surface area. Be-
cause of the nonlinearity, it would be inappropriate to use Eqs.
(3.2b) or (3.3) in the development of predictions of the mobility
of PCBs in groundwater.
The batch adsorption studies by Haque et al. and by Griffin
et al. were done with solution-phase concentrations in the range
of about 5 ppb to several tens or a few hundreds of ppb. These
concentration levels may be viewed as trace levels; however,
they are close to the solubility limits for PCBs in water. Non-
linear isotherms are most likely to occur when the solution-
phase concentration approaches the solubility limit.
Batch adsorption isotherms for many organic pesticides in a
variety of soils are reported in the literature. At concentration
levels much below the solubility limits, the isotherms are typ-
ically linear, and the distribution coefficients are generally large.
Organic pesticides are generally considered to be relatively
immobile in soil, and their use on agricultural land is rarely
regarded as a significant hazard to groundwater resources.
Davidson et al. (1976, 1980), however, have shown that at high
concentrations several organic pesticides have very nonlinear
Freundlich isotherms and, in some cases, low Ks values. Pes-
ticides at high concentrations may occur in groundwater be-
cause of leakage of residual pesticide solutions from used con-
tainers that are deposited in road ditches, sanitary landf~lls, or
dumps. Because millions of these containers are discarded each
year, high pesticide concentrations are cause for concern.
Chemical Reactions
of predictive capability is presented here.
Oxidation often requires the presence of 02, but the reaction
usually involves free radicals, especially OH., peroxy RO2,
alkoxy RO., and singlet oxygen iO2 as the oxidant. The rate of
oxidation of an organic compound (OC) can be expressed in
Contaminants in Groundwater: Chemical Processes
terms of the concentration of oxidants (e.g., tRO2O],(~O2~) and
the individual rate constants (e.g., k, RON:
- dfOCJ/dt = fOCI(kRO. PRO.] + kRO2 fRO2]
+ kHO tHO] + . . . ).
Prediction of the rate constants is possible for most organic
compounds through the use of compiled, empirical structure-
reactivity relationships (Mill, 1980~. Estimates of reaction rates
within a factor of 3-5 are possible. However, the free-radical
content of groundwaters is essentially unknown. Prediction is
then limited to approximations of relative rates of chemical
oxidation.
Chemical structures most susceptible to oxidation include
phenols, aromatic amines, and dienes. Saturated alkyl com-
pounds such as alkenes, halogenated alkenes, alcohols, esters,
and ketones may not be significantly oxidizable in the ground-
water environment.
Hydrolysis usually involves the introduction of a hydroxyl
(OH) group into an organic compound, usually at a point of
unbalanced charge distribution. Hydrolysis often displaces hal-
ogens (X):
R'COOR + H2O ~ R'COOH + ROH,
RX + H2O > ROH + X.
Hydrolysis may be catalyzed by acid (H+), base (OH-), or
metal ions (My thus, the rate of hydrolysis is pH and metal-
ion-concentration dependent. Surface effects may also influ-
ence the rate of hydrolysis. Hydrolysis of atrazine and other
pesticide derivatives is faster when humic material is present.
Wolfe (1980) suggested that such catalytic processes are slow
and reasonable predictions of hydrolysis rates at fixed pH could
be obtained from structure-reactivity relationships that define
kH for the hydrolysis rate relationship:
-d[OC]/dt= [OC]kH.
Mill (1980) felt that such a prediction could be made within a
factor of 2 or 3 for specific chemicals within a family of closely
related molecular structures for which rate data was available.
It is difficult to generalize relative hydrolysis rates, but the
data of Mabey and Mill (1978) indicate that for halogenated
hydrocarbons (RX) the hydrolysis rate increases where X is F-
Cl-Br, where R changes from primary to secondary to tertiary
type, and when allyl or benzyl groups were added.
Reductive dehalogenation involves the removal of a halogen
atom via an oxidation-reduction reaction. Although this mech-
anism is not usually considered in the chemical degradation of
organics it may be operative in low-redox state groundwaters.
The abiological reaction requires mediators, such as Fe+3 or
biological products to accept electrons generated by oxidation
of reduced organics and to transfer these electrons to the hal-
ogenated organic to bring about dehalogenation. Esaac and
Matsumuna (1980) suggested that Eh c 0.35 V is required so
that electrons can be made available for dehalogenation. No
studies have been made to evaluate this mechanism for organic
transformation in groundwaters, but because it is operative
57
under low-Eta conditions, it is a potentially significant process
that warrants evaluation.
Biological Reactions
Enzymatic reactions brought about by the microbes inhabiting
the groundwater environment may be the most important
mechanism for transformation of organic contaminants. This
activity is dominantly bacterial, although yeasts, fungi, and
viruses may also be present. We have measured, by direct
microscopic counting, 106 bacteria per gram of dry, sandy aqui-
fer material in a number of landfill-leachate-contaminated
groundwater systems. Although this is less than the bacteria
content of fertile agricultural soils (108 - 109 bacteria per gram),
it indicates that contaminated-groundwater systems can be mi-
crobially active.
Biodegradation of a broad range of organic compounds has
been demonstrated in laboratory studies of soils, sediments,
and waters. Compounds include pesticides, halogenated hy-
drocarbons, aromatic hydrocarbons, amines, and alcohols. The
broad range of enzymatic activities in actual mixed populations
of microbes permits a broad range of enzyme-catalyzed reac-
tions (oxidation, reduction, hydrolysis, dehydration). Because
the proportion of each species present at any point in space
and time is environmentally dependent, predictions of actual
organic transformation pathways and rates are all but impos-
sible.
The difficulty of prediction is illustrated by the biological
transformation of many chlorinated hydrocarbons, which were
originally thought to be essentially nonbiodegradable (recal-
citrant). It was believed that observed biogradation was limited
to aerobic bacteria because anaerobic bacteria did not have
sufficient energy available for biogradation because of the types
of energy-yielding reactions that they could utilize. Studies of
chlorinated pesticides, trihalomethanes, and other halogenated
hydrocarbons revealed biodegradation in anaerobic environ-
ments. In fact, degradation of trihalomethanes was observed
only in anaerobic systems (Bouwer et al., 19811. In their review
of microbial degradation of organics, Kobayashi and Rittmann
(1982) indicated that anaerobic reductive dehalogenation may
be important in transforming certain classes of organic com-
pounds. Although there is research activity in this area, few
data pertinent to groundwater systems are available at present.
The mechanisms of biodegradation of synthetic, often hal-
ogenated, organic contaminants are not well understood. Mi-
crobes that use organic compounds convert these substrates
into inorganic products (e.g., CO2, H2S) and into cell constit-
uents and often obtain energy for biosynthesis from these re-
actions. The populations responsible for such transformations
increase in numbers of biomass as a result of the introduction
of the organic chemical into the system. This is direct metab-
olism. Degradation of many synthetic organic compounds is
unlikely by this direct utilization mechanism because the re-
quired enzymatic pathways have not been developed by mi-
crobes that have not been previously exposed to such organic
structures. Microorganisms capable of directly utilizing DDT,
2,4,5-T, and many halogenated hydrocarbons, for example,
58
have yet to be isolated. However, these compounds have been
observed to be biodegradable. The mechanism has been termed
cometabolism. Microbes apparently utilize other organic sub-
strate while performing the transformation of the organic con-
taminant.
Alexander (1981) points out two environmental consequences
of cometabolism: (1) the responsible populations do not increase
in number or biomass as a result of the introduction of the
organic compound because it is not utilized for biosynthesis;
and (2) a compound subject to cometabolism is modified slowly,
and products structurally similar to the contaminant accumu-
late because the organism does not possess a sufficient array
of enzymes to bring about its extensive transformation.
Considerable additional research is required before biode-
gradation can be adequately predicted. Correlations of organic
structure and reactivity need to be improved. Models such as
the biofilm-based theory of Rittmann et al. (1980) need to be
improved by incorporation of in situ parameters and by eval-
uation at a number of field sites.
It must be pointed out that microbial processes can also
increase groundwater toxicity as well as reduce it. Biodegra-
dation of toxic organics such as DDT can produce more toxic
intermediate products such as DDE. Also, microbial methy-
lation of metals such as mercury and lead has been shown to
increase the metals' toxicity to other organisms.
Organic Compounds as Complexing Agents
Complexation of some inorganic contaminants, such as trace
metals, by inorganic ligands (e.g., CO32-, OH-) has been dis-
cussed with respect to chemical speciation. Cationic inorganic
contaminants may also be complexed by organic ligands. Many
such complexes are stable in groundwater. Although most pre-
dictive models include inorganic complexes in consideration of
aqueous speciation and transport, the effects of organic ligands
are usually not included. Models considering the possible in-
fluence of simple organic ligands such as NTA, citric acid, and
EDTA on the chemical speciation in seawater or freshwater
have produced conflicting indications of the importance of these
organic ligands. Organic complexation should increase as the
concentration of organic ligands increases and is expected to
be significant in groundwaters with high dissolved organic car-
bon (DOC).
The transport of inorganic contaminants as organic complexes
has been documented. Means et al. (1978) reported unexpected
mobility of 60Co and U from liquid disposal sites at the Oak
Ridge National Laboratory. Apparently, 60Co and some U were
complexed by EDTA that was a constituent in wastes from
decontamination facilities and by dissolved humic substances.
The neutral or anionic species so formed were not subject to
the expected retardation by adsorption onto the soil. Iron, toxic
trace metals, and other radionuclides have been reported to
be transported as organic complexes in groundwaters and sur-
face waters.
Prediction of the extent of organic complexation in contam-
inant transport is limited by the lack of knowledge concerning
the nature and content of organic ligands in groundwaters and
JOHN A. CHERRY, ROBERT W. GILEHAM, and JAMES F. BARKER
by the lack of a comprehensive thermochemical data base for
complex ligands. A majority of organic ligands are natural, high-
molecular-weight, structurally complex fulvic and humic acids.
Assuming that these ligands dominate the groundwater DOC
and that they have properties similar to analyzed material
(Schnitzer and Khan, 1912), the complexing capacity of ground-
waters can be calculated as 1 x 10-2 to 5 x 10-2 meq/L for
each 1 mg/L of DOC. It is expected that most ofthis complexing
capacity is taken by major cations and that only a part is avail-
able to trace metals or radionuclides. The potential for signif-
icant complexation of trace metals, for example, has been dem-
onstrated in high-DOC groundwaters contaminated by landfill
leachate (Knox and Jones, 1979~. Prediction of inorganic con-
taminant complexation by complex organic ligands has been
attempted through models that simplify the organic-metal in-
teraction (Sposito, 1981), but the applicability of these models
for application to contaminated-groundwater systems has yet
to be assessed.
FIELD STUDIES OF CONTAMINATED
GROUNDWATER
Although most of the literature on the chemical behavior of
contaminants in groundwater is based on batch and column
experiments and on thermodynamic models, insight pertaining
to contaminant behavior is also obtained from field investiga-
tions of zones of contamination. In favorable circumstances data
from existing zones of contamination at waste-disposal or chem-
ical spill sites can be used to assess the applicability of predic-
tive models.
There are 10 inorganic constituents for which maximum con-
taminant levels are specified in the National Interim Primary
Drinking Water Regulations of the U. S. Environmental Pro-
tection Agency (1975~. These are As, Ba, Cd, Cr, F, Pb, Hg,
NO3-, Se, and Ag. Except for NO3- and Cr, these constituents
are rarely reported as a cause of significant deterioration of
groundwater quality.
Nitrate is a particularly serious cause of groundwater con-
tamination. NO3-, which has a maximum permissible concen-
tration of 10 mg/L as N2 in drinking water, is a common cause
of groundwater-quality deterioration because nitrogen is a ma-
jor component of agricultural fertilizer and of animal and human
wastes and because nitrogen commonly occurs as NO3-, which
is very soluble and generally unretarded by adsorption. The
widespread nature of NO3- sources and the mobility of NO3-
cause unconfined aquifers to be particularly susceptible to grad-
ual long-term increases in NO3- concentrations. Increases will
probably continue because the quantity of NO3- available for
entry into the groundwater zone will probably not diminish
appreciably in the next few decades. To cause a decrease would
require widespread changes in agricultural practice and in the
management of animal and human sewage.
It is fortunate that in at least some groundwater systems
there exists a biogeochemical process, namely denitrification,
that tends to ameliorate the influence of NO3- inputs to the
groundwater zone. The denitrification process can be expressed
Contaminants in Groundwater: Chemical Processes
schematically as
NO3- ~ organic carbon ~N2 + CO2 + H2O.
According to thermodynamic data, denitrif~cation is expected
to occur only in zones that are essentially devoid of dissolved
oxygen (Lindsay 19791. However, field studies have established
that it will occur in groundwater zones where labile organic
matter and denitrifying bacteria exist even if the groundwater
contains low but measurable concentrations of dissolved oxygen
(Gillham and Cherry, 1978~. Although groundwater from wells
in zones where denitrification apparently occurs contains de-
tectable quantities of dissolved oxygen, it is likely that the
reduction of NO3- takes place in microenvironments in small
pores or on grain surfaces where the redox conditions are most
suitable. Organic matter is a source of energy and of cell carbon
for the bacteria that mediate the reduction process.
The loss of NO3 in the groundwater zone due to denitrifi-
cation cannot be predicted simply from measurements of the
redox condition of groundwater samples. Based on laboratory-
column experiments, Doner and McLaren (1976) developed a
mathematical expression to describe steady-state NO3 loss in
sandy soils due to denitrification. The expression contains sev-
eral parameters, including the mass of N utilized per unit bi-
omass per unit time for maintenance of the bacterial population,
the mass of N utilized per unit biomass in wasted bacterial
metabolism, and the mass of organic matter available to the
bacteria that mediate the reaction. The work by Doner and
McLaren demonstrated the importance of bacteria and organic
matter in the denitrification process. It is unlikely, however,
that mathematical expressions derived from laboratory exper-
iments will be of much use in the development of predictions
of denitrification in the groundwater zone. Whether the mi-
crobiological conditions that are critical to the denitrification
process in the field can be adequately represented in the lab-
oratory remains to be determined. The development of field
techniques for the in situ measurement of the main rate-de-
termining factors is desirable.
Of the other nine inorganic contaminants, Cr has probably
caused the most degradation of groundwater. Cr is used in
many manufacturing processes and is a constituent in sewage
sludge. Cr that leaks from waste lagoons or that leaches from
industrial landfills and sludge-disposal areas or from soil con-
taminated by spills of Cr-rich liquids can pose a hazard to
groundwater quality.
Shallow aerobic groundwater zones are particularly sus-
ceptable to contamination by Cr; + VI) because under oxidizing
conditions the stable Cr species (HCrO3 or CrO2-) are rela-
tively soluble (Figure 3.3) and undergo little retardation by
adsorption in many types of permeable geologic deposits. Perl-
mutter and Lieber (1970) and Pinder (1973) have described a
major occurrence of Cr; +VI) contamination in a sand aquifer
on Long Island, New York. The zone of contamination is ap-
proximately 1400 m long, 350 m wide, and 25 m thick. Shallow
zones of Cr; + VI) contamination in sand aquifers have recently
been one of the main subjects of legal proceedings related to
groundwater protection legislation in the state of Michigan.
Cr(+ III) is not a cause of significant groundwater contamina-
59
tion because the Cr(+ III) is insoluble in water except at low
pH and because it occurs as cationic species that are absorbed.
Of the other eight inorganic contaminants, a few have such
severe solubility limitations that they are generally immobile
in groundwater under normal pH conditions. For example, the
concentration levels at which Ba can occur are limited to very
low levels by the solubility of BaSO4. Two of the contaminants,
Hg and Ag, rarely cause groundwater contamination because
they are uncommon constituents in waste materials deposited
on land and because their mobility in normal groundwater
conditions is probably limited by solubility constraints and ad-
sorption. The remaining contaminants As, Cd, F. Pb, and
Se—have been reported in only a few areas as causes of severe
local groundwater degradation. Some of these contaminants
are particularly prone to causing groundwater contamination
in situations where the pH is much lower or much higher than
the normal range for groundwater. Conditions of extreme pH
in groundwater are common in waste materials at metal or
uranium mines or in ash disposal areas associated with coal-
fired power plants. Se and As, for example, have been reported
at exceptionally high concentrations in groundwater at fly-ash
disposal sites in North Dakota in zones where pH levels are
above 10 (Groenewold et al., 1981~. In contrast, similar inves-
tigations at fly-ash disposal sites in southern Ontario, where
the pH of the groundwater is between 7 and 9.5, have estab-
lished that the concentrations of these elements and the other
elements with maximum permissible limits specified in drink-
ing-water regulations are low (Dodd et al., 19811.
Field investigations of contaminated aquifers at uranium tail-
ings impoundments in Wyoming and in northcentral Ontario
(Morin et al., 1982) have established that the contaminated
zones at low pH (i.e., generally less than about 4.5) invariably
have high concentrations of many transition metals, heavy met-
als, metalloids, and radionuclides, whereas the neutral pH zones
of contamination that exist in advance of the retarded low-pH
fronts rarely have any of these constituents at levels above the
drinking-water limits. This is the case because pH exerts a
dominant influence on the solubility controls and the adsorp-
tion of these contaminants. The hydrogeochemical nature of
sandy aquifers that receive acidic water from tailings impound-
ments can be represented as three main zones: the acidic zone,
the neutralization zone, and the neutral-pH zone. In the neu-
tralization zone, the hazardous contaminants are transferred
from the water phase to the solid phase by precipitation and
adsorption. In the prediction of the movement of metals and
radionuclides in groundwater systems that are receiving acidic,
metal-rich, or radionuclide-rich water, the critical task is the
prediction of the advance of the front of the acidic zone. Field
investigations, e.g., Morin et al. (1982), have shown that low
pH fronts can be greatly retarded because of reactions with
porous media, but the development of a methodology to predict
neutralization-zone behavior at new sites is in the early stages.
A potential cause of groundwater contamination that is an
issue of concern in many communities is municipal landfills.
Detailed monitoring of zones of contaminated groundwater has
taken place at landfills on permeable deposits of sand and gravel
on Long Island (Kimmel and Braids, 1980) in Ontario (Cherry,
60
1983), in Delaware (see Chapter 10 of this volume), and at
many other locations in North America. The investigations in-
dicate that, although major ions such as C1-, HCO3-, Na+,
Ca2+, and Mg2+ and minor constituents such as NH4-, Fe,
and Mn are mobile, the toxic inorganic constituents generally
do not occur at concentration levels above the mandatory drink-
ing-water limits. At the Ontario sites, no values above the limits
were reported for toxic inorganic constituents. At the Long
Island sites only Se occurred in some samples at levels slightly
above the limit. At all of these sites, the pH values of the zone
of contaminated groundwater were near neutral.
The leachate from municipal landfills has high concentrations
of dissolved organic compounds. However, despite the poten-
tial for mobilization of toxic inorganic compounds by complex-
ing with organic compounds, immobility of the heavy metals
and metalloids is the rule rather than the exception.
In some contaminated groundwater, mineral dissolution oc-
curs because the contaminant solution that invades the porous
medium causes reduction of Fe or Mn. This is particularly the
case when the invasion takes place at shallow depth where Fe
and Mn occur in the porous medium as oxides of Fe; + III) and
Mn(+IV). When the water in the reduced state encounters
the oxides, they become unstable and the Fe and Mn go into
solution as Fe(+II) and Mn(+II). Their solubility in the re-
duced oxidation states can be high. Oxides of Fe(+III) and
Mn(+IV) are natural scavengers of metallic elements in the
geochemical environment, and, therefore, their dissolution can
cause the release of elements that accumulated in the oxides
under natural conditions before the invasion of the contaminant
solution. An example of the influence of this source of metals
on the chemistry of contaminated water beneath a landfill is
described by Suarez and Langmuir (19767. The importance of
these oxides as metal scavengers is described by [enne (1968,
1977~.
It has only been in recent years that increased concern for
groundwater quality and the availability of greatly improved
analytical methods for the identification of organic compounds
has resulted in appreciable monitoring of toxic or potentially
toxic organic contaminants in groundwater. This work is re-
vealing widespread contamination of groundwater by organic
chemicals, which indicates significant mobility of many of these
substances through soil and in the groundwater zone. Some
specific examples (Wilson et al., 1981) are occurrences of tri-
chloroethane in many groundwaters in the United States and
Europe as a result of spills, leaks, and disposal of wastes in
soil; occurrence of 1,2-dibromo-3-chloropropane in ground-
water in areas in California, where it was applied to soils as a
nematocide; contamination of a large body of groundwater by
phenol in the vicinity of an accidental spill of this compound
in Wisconsin; and the movement into groundwater of 4,4'-
methylene bis(2-chloroaniline) from a wastewater lagoon in
Michigan.
The processes that control the behavior of specific organic
contaminants in groundwater have been evaluated at only a
few field sites. One such study was conducted during injection
of tertiary-treated sewage effluent in a sand aquifer near Palo
Alto, California (Roberts et al., 1980~. It was observed that
organic trace contaminants under anaerobic conditions were
JOHN A. CHERRY, ROBERT W. GILLHAM, and JAMES F. BARKER
attenuated to varying degrees during the passage of the treated
effluent through the groundwater zone. Of the various low-
molecular-weight halogenated organics studied, chlorobenzene
was most mobile but traveled at a rate of approximately 1/36th
of the rate of nonadsorbed inorganic constituents. Dichloro-
benzene isomers and 1,2,4-trichlorobenzene isomers were ap-
parently more strongly adsorbed than chlorobenzene. Naph-
thalene showed evidence of biodegradation. Another injection
study is under way in southcentral Ontario in an aerobic zone
in a sand aquifer containing a very low solid-phase organic
carbon content (Mackay et al., in press). At this site, 12,000
L of water containing chloride and bromide salt as nonreactive
tracers and containing five toxic halogenated hydrocarbons (o-
dichlorobenzene, bromoform, carbon tetrachloride, hexa-
chloroethane, and tetrachloroethylene) were injected during a
14-h period as a slug into the aquifer. The behavior of the salt
tracers and the organic compounds under natural flow condi-
tions was monitored for more than a year following the injec-
tion. Two of the compounds (carbon tetrachloride and bro-
moform) were found to be quite mobile; they traveled at a rate
of about two thirds of the groundwater velocity. Tetrachloro-
ethylene traveled at a rate of about one third of the groundwater
velocity and o-dichlorobenzene and hexachloroethane at a rate
of less than a quarter of the groundwater velocity. The first
three compounds above were not noticeably biodegraded. There
is a possibility that the latter two compounds were biodegraded
to some degree, however, at most very slowly.
At the above-cited landfills on Long Island and Ontario,
landfill-derived, dissolved organic carbon exists throughout the
zones of contaminated groundwater. Investigations of the iden-
tifiable compounds in this dissolved organic fraction indicated
that many toxic or potentially toxic compounds are mobile in
these groundwater systems (Reinhard et al., in press). Even
with modern analytical techniques, only organic compounds
that comprise as much as 5 to 10 percent of the DOC in zones
of contaminated groundwater at landfill sites are identifiable.
It is expected that toxic organic contaminants are mobile in
groundwater at many municipal landfills situated on permeable
deposits. The dissolved organic fraction in contaminated
groundwater at landfills has a much greater potential to cause
severe contamination of groundwater resources than dissolved
inorganic contaminants.
CHEMICAL REACTIONS AND PERMEABLE
MEDIA
The degree to which chemical reactions can cause attenuation
of contaminants in groundwater can be dependent on the type
of permeable media in which the contaminants occur. In this
chapter the discussions pertain implicitly to nonindurated po-
rous media such as gravel, sand, silt, or clay in which contam-
inants are transported by groundwater flow through the pore
spaces between the grains or particles that comprise the media.
The advection-dispersion theory and the isotherm approach to
predictive modeling of the behavior of adsorbed contaminants
were developed specifically for this type of medium. In many
regions of North America, the occurrence and movement of
Contaminants in Groundwater: Chemical Processes
contaminants in other types of permeable media such as frac-
tured rock or fractured fine-grained, nonindurated deposits are
also important. The effects of chemical processes in these de-
posits can be very different from those that occur in porous
media.
In groundwater systems in fractured crystalline rock in which
the rock matrix is relatively nonporous (such as granite, marble,
and many other types of igneous and metamorphic rocks), the
migrating contaminants contact only the mineral surfaces ex-
posed on the fracture walls and the amorphous geochemical
weathering or alteration products that exist on these surfaces.
Measurement of adsorption parameters for predictive modeling
is particularly difficult for this type of medium because the
measurements must adequately represent the conditions that
occur on the surface of the fractures as they exist in the ground-
water zone. The distribution coefficient expressed in the nor-
mal manner relative to the mass of solids is inappropriate be-
cause rock mass and fracture surface area have no known general
relation. For this reason the distribution coefficient for frac-
tured crystalline rocks has been defined in terms ofthe effective
surface area of reaction in the fractures. There have been few
attempts to determine the adsorptive properties of relatively
undisturbed fracture surface, and there have been no assess-
ments under field conditions of the predictive capabilities of
the models using parameters determined by the existing mea-
surement techniques.
The flow of groundwater in fractured porous media such as
weathered deposits of silt or clay or fractured rocks that have
considerable intergranular porosity such as shale and porous
sandstone commonly occurs through the fractures. Little or no
flow takes place in very porous but relatively impermeable
matrix. As contaminants are transported through the fracture
network, transient chemical-concentration gradients exist be-
tween the water in the fractures and the pore water in the
matrix. These gradients cause the contaminants to diffuse into
the matrix in the frontal part of the contaminant zone and
diffuse out of the matrix in the trailing part. Models of the
advection, dispersion, and diffusion of contaminants in frac-
tured porous media have established that diffusion into the
porous matrix can have a strong influence on contaminant be-
havior (Tang et al., 1981; Grisak and Pickens, 1981~. The surface
area that controls adsorption is generally the surface area con-
tacted by the contaminants in the porous matrix, which is many
orders of magnitude larger than the surface area of the fracture
surfaces. The chemistry of the groundwater in the porous ma-
trix can have a dominant effect on chemical mass transfer by
precipitation or dissolution. The bacterial population that exists
in the porous matrix may be more important than exists on the
fracture surfaces. Research pertaining to the chemical and bio-
chemical behavior of contaminants in fractured porous media
Is In its Infancy.
SUMMARY AND CONCLUSIONS
Most of what is known about the chemical behavior of contam-
61
behavior of inorganic contaminants in groundwater involve the
use of the distribution coefficient, which is incorporated in a
simple retardation term into the advection-dispersion equation,
and the use of thermodynamics-based chemical equilibrium
models. When the contaminant of interest exhibits a linear
equilibrium adsorption isotherm, the retardation relation can
be used to estimate the relative rate of advance of the mid-
concentration position of the front of the contaminant zone.
This approach is applicable in situations where contamination
emanates from a continuous source. It can also be used to
estimate the center of the mass of the contaminant zone in
situations where the contamination originated from a distinct
temporary source. Field experiments have shown that for trace-
level inorganic contaminants controlled by adsorption, such as
some cationic radionuclides, the retardation relation, when ap-
plied in favorable circumstances, provided estimates of relative
velocity within a factor of about 5 or better. The usefulness of
this approach for prediction of the relative advance rates of
toxic nonradioactive inorganic or toxic organic contaminants in
groundwater has not yet been subjected to a comprehensive
evaluation.
When the retardation relation is incorporated into one-di-
mensional advection-dispersion models, simulated break-
through curves for column experiments using adsorbed tracers
that have linear isotherms do not agree closely with experi-
mental data. The experimental breakthrough curves are dis-
tinctly asymmetrical, whereas the simulated curves are sym-
metrical or nearly symmetrical. Hypotheses to account for this
discrepancy have not yet been subjected to comprehensive
evaluation. This area of uncertainty and the uncertainties as-
sociated with sample disturbance and related geochemical ef-
fects do not bode well for predictions of the first arrival of fronts
or of the tails of zones of adsorbed toxic contaminants.
The maximum concentration levels at which many inorganic
contaminants occur in groundwater is controlled by the solu-
bility of minerals or other solids. Thermodynamic-based equi-
librium models are available for prediction of these maximum
concentrations in groundwater in which dissolved organic mat-
ter is not a complicating factor. Although the thermodynamic
data for many of the solid phases and complexes of relevance
are questionable, and although equilibrium is probably not
achieved in many situations, these models can provide useful
order-of-magnitude estimates of maximum possible concentra-
tion levels. The models can be used to assess the possible
influences of various geochemical scenarios on the occurrence
of inorganic contaminants in groundwater. Although the most
advanced equilibrium geochemical models that are currently
available have well-developed computational capabilities for
complex inorganic aqueous systems, they are static models in
that they do not include formal representations of the effects
of advection or dispersion. Advanced models of this type have
not yet been incorporated into advection-dispersion models,
although the effects of advection and dispersion have been
approximated by the use of equilibrium models in combination
with cell models.
The chemical behavior of most toxic inorganic contaminants
inants in groundwater pertains to inorganic contaminants. The depends strongly on the redox and pH conditions of the con-
two main approaches used for the prediction of the chemical tarninated groundwater. When the pH is very low or very high,
62
some heavy metals or metalloids are commonly mobile. A very
low redox condition generally promotes immobility. Field in-
vestigations have established that, at neutral pH, metals and
metalloids are not commonly mobile, except for Cr; + VI) and
Se, which occur in anionic forms and tend to be mobile when
oxidizing conditions prevail. The most critical task in the pre-
diction of the chemical behavior of most toxic inorganic con-
taminants is the prediction of the pH and redox conditions.
This necessitates prediction of the gross water chemistry of the
migrating zone of contamination. In addition to the influence
of geochemical factors, the gross water chemistry is affected
by dispersion, which often causes mixing of waters of different
pH and redox status as the zone of contamination moves through
the groundwater system. The task of predicting the chemical
behavior of reactive inorganic contaminants therefore cannot
be isolated from the problems inherent in predictions of dis-
persion in the groundwater zone.
Dissolved organic contaminants in groundwater can be in-
fluenced by adsorption, oxidation, hydrolysis, or microbial deg-
radation. Laboratory experiments indicate that many trace or-
ganic compounds exhibit linear adsorption isotherms and,
therefore, may be favorable for transport simulation using ad-
vection-dispersion-retardation models. Considerable success
has been achieved in estimating, from solubility data, the dis-
tribution coefficients for adsorption of halogenated hydrocar-
bons by solid organic matter in porous geologic materials that
have appreciable organic matter. The other processes that can
cause attenuation of dissolved organic compounds are much
less amenable to quantification, particularly for anaerobic
groundwater where biological transformations are poorly
understood. The current paucity of information on the chemical
and biochemical behavior of organic compounds in ground-
water is a particularly serious liability in the assessment of
future changes in groundwater quality because entry of organic
compounds to the groundwater zone is now common and be-
cause many toxic organic compounds are relatively mobile.
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