United States
Environmental Protection
Agency
Office of
Research and
Development
Office of Solid Waste
and Emergency
Response
EPA/540/S-92/OOI
January 1992
&EPA Ground Water Issue
Chemical Enhancements to Pump-and-Treat Remediation
Carl D. Palmer* and William Fish*
Background
Conventional pump-and-treat technologies are among the
most widely used systems for the remediation of
contaminated ground water. Within recent years it has
become recognized that these systems can require
protracted periods of time to make significant reductions in
the quantity of contaminants associated with both the liquid
and solid phases which constitute the subsurface matrix.
Evaluating the effectiveness of pump-and-treat remediations
at Superfund sites, as well as attempting to improve this
effectiveness, are issues identified by the Regional
Superfund Ground Water Forum as a concern of Superfund
decision makers. The Forum is a group of ground-water
scientists and engineers, representing EPA's Regional
Superfund Offices, organized to exchange the most recent
information related to ground-water remediation at Superfund
sites.
Recent research has led to a better understanding of the
processes involved in the transport and transformation of
contaminants in the subsurface. While some of these
processes are not readily amenable to enhanced removal by
ground-water extraction, others suggest that there are
available techniques to increase the efficiency of these types
of remediation systems. The intent of this document is to
explore the use of chemical enhancement to improve ground-
water remediation efficiencies using pump-and-treat
technologies, and point out arenas of contamination where
such techniques are not practical.
For further information contact John Matthews, Chief,
Applications and Assistance Branch, RSKERL, FTS 743-
2408, or Bert Bledsoe, Project Officer, RSKERL, FTS 743-
2324. Both may be reached at 405-332-8800.
Summary and Conclusions
Recognition that conventional pump-and-treat remediation
often requires lengthy periods of time to achieve clean-up
objectives will encourage professionals involved in site
remediation to contemplate alternative methods of aquifer
restoration. Some form of chemical enhancement for pump-
and-treat will likely be an alternative considered for many
waste-site cleanups. Although chemical enhancement of
pump-and-treat may be a means of accelerating aquifer
remediation, there are many aspects of chemical
enhancement that need to be known before these techniques
can be successfully implemented.
Not all waste sites are amenable to chemical enhancement
methods. In particular, if tailing in the concentration-versus-
time curves for the extraction wells is dominated by physical
processes, then chemical enhancement methods will have no
advantage over conventional pump-and-treat. Knowledge of
the relative contributions of chemical and physical processes
limiting pump-and-treat are needed during the early stages of
site Remedial Investigations to ascertain the general
usefulness of chemical enhancement.
Even when it is known that physical processes contribute little
to the tailing, specific knowledge is needed about the
chemical processes that contribute to tailing at a particular
site. Only then can potential chemical agents that are likely
to influence these processes be identified. The reactive
agents may be chosen to compete with the contaminants for
adsorption sites, complex the contaminant, change the redox
state of the contaminant, change the solvation properties of
the ground water, act as a surfactant, ionize the contaminant,
or substitute for the contaminant in a precipitate. If the
reactive agents are chosen on the basis of incorrectly-
identified limiting processes, there is a risk that the reactive
agents will provide no net benefit and may even prolong
remediation.
Even when a reactive agent is found that specifically
addresses the limiting chemical process, other considerations
must be investigated to assure successful implementation.
The key areas of concern in any chemical enhancement
method are 1 ) the delivery of the reactive agent to those
'Environmental Science and Engineering, Oregon Graduate
Institute of Science and Technology, Beaverton, OR.
Superfund Technology Support Center for
Ground Water
Robert S. Kerr Environmental
Research Laboratory
Ada, Oklahoma
Technology innovation Office
Office of $o8<| Waste and Emergency
Reeporwev us EPA, Washington, D,C.
Walter W. KovaUck, Jr., Ph.D.
Director
) Printed on Recycled Paper
-------
areas of the aquifer where it is needed, 2) the enhanced
removal of the target contaminants by the reactive agent, 3)
the removal of the reactive agent from subsurface, 4) the
impact of the reactive agent on the treatment of the target
contaminant and the volume of sludges to be disposed.
Additional site characterization, treatability tests, and design
studies must be conducted to address each of these
important aspects of a chemical-enhancement program. At
some sites, implementation of chemical enhancement will
require additional capital expenditures for wells and treatment
facilities. The advantages and disadvantages, including the
additional costs, of chemical enhancements need to be
compared with other methods of remediation, such as
conventional pump-and-treat.
While many individual components are associated with the
implementation of chemical enhancement to pump-and-treat,
they all should be investigated. If one aspect of this process
fails, the entire system can fail. While such failure is not
necessarily a disaster (conventional pump-and-treat can
continue), it is a waste of resources that could be utilized for
more beneficial uses. It is believed that these issues must be
addressed and a reasonable probability of success
demonstrated in all aspects of a chemical enhancement
system before it is implemented.
Introduction
The recognition that ground water in many areas of the U.S.
is contaminated has brought about demands that the quality
of these aquifers be restored. At Superfund sites, the initial
cleanup is accomplished in a relatively short time by
removing sources of contamination from the surface,
removing highly contaminated shallow soil, and in some
cases installing a low-permeability cap. In contrast,
remediation of the ground water beneath a site is often an
inexact process requiring years to complete.
A common method for aquifer remediation is to withdraw the
contaminated water from the aquifer and treat it on-site. The
treated water may then be returned to the aquifer, discharged
to surface water, or transferred to a public water treatment
plant. Such "pump-and-treat" technology is widely used for
remediating aquifers (Palmer et al., 1988) with about 68% of
the Records of Decision identifying it as the method of
remediation (Travis and Doty, 1989). However, at many sites
pump-and-treat technology will require decades of costly
operation to achieve the desired levels of cleanup. Extended
periods for remediation are highly undesirable because the
operation and maintenance costs associated with the
remediation can be large, and, in many cases, otherwise
valuable land cannot be used for any economic purpose.
The great costs of cleanup make it essential to investigate
technologies that may speed up remediation. One such
technology is the injection of chemical constituents, "reactive
agents", that improve the rate of removal of contaminants
from the subsurface. The applicability of such "chemical
enhancement" technology and the specific chemicals that can
be used depend on the processes that control the slow
'tailing" of contaminant concentrations in the extraction wells.
Not all processes leading to lengthy remediations can be
corrected by chemical enhancement. However, certain
problematic types of contamination maybe amenable to well
conceived applications of reactive agents.
The limitations of aquifer remediation by conventional pump-
and-treat will encourage engineers, scientists, and regulators
to propose various chemical enhancement methods for the
remediation of particular sites. While these proposed
methods must be evaluated with regard to specific site
conditions, there are general concepts applicable to all
chemical enhancement methods. This document is intended
to 1 ) outline these general concepts, 2) pose key questions
that should be answered before any chemical-enhancement
scheme is initiated, 3) stimulate discussion on the merits and
limitations of chemical enhancement methods, and 4) focus
research on particularly problematic areas of chemical
enhancement.
Processes Affecting Pump-and-Treat Remediation
A major concern in pump-and-treat operations is that
contaminant concentrations within the extraction wells will
decline at a progressively slower rate as pumping continues.
When the rate of decline becomes small and the contaminant
concentrations are still above the target cleanup levels, an
extraction well is said to exhibit 'tailing" (Fig. 1). Contaminant
concentrations may have dropped several orders of
magnitude, but they remain above the target clean-up level
despite a considerable period of pumping. A great
uncertainty in pump-and-treat operations is the time required
for these tailing concentrations to decrease below the target
clean-up levels. Reasonable estimates of clean-up times
under these conditions require an understanding of the
physical and chemical processes that can cause tailing: 1)
the differing amounts of time required by contaminated
waters to flow along different streamlines from the irregular
boundary of the plume to the extraction wells, 2) multiple
rates of mass transport within spatially variable sediments, 3)
limited mass transfer from reserves of nonaqueous phase
liquids and solid phase mineral precipitates, and 4) slow
resorption reactions (Keely et al., 1987; Keely, 1989).
MAX
HI
O
O
O
WITHOUT
"TAILING"
"TAILING"
PHENOMENON
"RESIDUAL"
:ONCENTRATION
TIME FROM INITIATION OF EXTRACTION
Figure 1. Concentration versus lime curve for an extraction well with
continuous pumping (after Keely et al., 1987).
-------
Physical Causes of Tailing
Ground water entering an extraction well is a mixture of
waters that have traveled along multiple subsurface pathways
between the edge of the contaminant plume and the well.
The time required for contaminated ground water to travel
along these different flow paths is controlled by 1) placement
of the extraction wells relative to the contaminant boundaries,
2) the extraction rate, 3) the aquifer porosity, 4) the
magnitude and direction of the natural hydraulic gradient, and
5) the location and types of hydraulic boundaries. As an
example, consider a single extraction well in an aquifer with a
natural hydraulic gradient of 0.007 towards the east (Fig. 2).
If the edge of the contaminant plume is to the west, then only
a portion of the plume's edge is captured by the extraction
well. The flow paths along the outside of the capture zone for
the well have a greater distance to travel and are influenced
by lower hydraulic gradients than the flow paths near the
center of the capture zone. As a consequence of these
variable residence times within each stream tube, the
concentration-versus-time curve for the extraction well
exhibits substantial tailing even in the absence of chemical
reaction (Fig. 3).
In heterogeneous porous media, ground water in higher
permeability layers has greater velocities than water within
the lower permeability zones. The higher permeability
pathways are not necessarily sand or gravel nor are the
lower permeability zones necessarily silts or clays; it is
sufficient if one region possesses greater hydraulic
conductivity relative to the adjacent materials. When the
contrast in hydraulic conductivity between these zones is
large, the advective component of transport through the
lower permeability lenses becomes small. As contaminants
are transported through such a heterogeneous aquifer, they
are advected along the high permeability layers and diffuse
into the lower permeability layers. Such an advection-
diffusion process can affect the concentration of
contaminants within higher permeability layers (Gillham et al.,
1984; Sudicky et al., 1985). If aqueous contaminants have
been present over many years, their concentration in the
lower permeability layers can equal the concentrations in the
higher permeability zones. During pump-and-treat
remediation, contaminants in the high permeability layers are
removed more quickly than from the lower permeability
layers. These variable rates of advective transport create
concentration gradients between zones of contrasting
permeability and cause the slow diffusion of contaminants
from the low permeability zones to the high permeability
zones where they can be pumped to the surface (Fig. 4).
Thus, the contaminant concentrations in the extracted water
are initially high as the more permeable layers are flushed.
At later times, the concentration in the extracted water is
limited by the rate of diffusion of the contaminants into the
high permeability zones (Fig. 5). If pumping is discontinued,
the velocities within the high permeability layers decrease
and the concentration of contaminants within these zones
increase (Fig. 6) because of the greater residence time of a
parcel of water within the contaminated portion of the aquifer.
The main point is that, in most cases, lengthy tailing-off of
contaminant concentrations in extraction wells is at least
partly due to physical attributes of the system that cannot be
ameliorated by injections of chemical agents. Thus, chemical
enhancement cannot be expected to eliminate all unexpected
delays in pump-and-treat removal. However, when the rates
of chemical mass transfer from contaminant reserves in an
aquifer are the primary limitations on removal, then the use of
reactive agents may substantially enhance remediation.
CAPTURE ZONE
ğ. DIRECTION OF NATURAL
GROUNDWATER FLOW
STAGNATION
POINT
EXTRACTION
WELL
Figure 2. Flow lines from the edge of a contaminant plume toward an
extraction well.
0.8
in
o
o
o
UJ
LU
-------
ADVECTION
' ff~ff f * TTTTTTT" * "<"TT\
MOtECULAR DIFFUSION
LOW PERMEABILITY LENSES '
FTT\
HIGH PERMEABILITY STRATA
^ nmn rtHt/itABii
MOLECULAR DIFFUSION
ADVECTION
.ğ>*./
Figure 4. Heterogeneous porous medium with advection through the
low permeability zones and mass transfer by molecular diffusion from
the lower permeability lenses.
Z 10"
<
DC
Z
LU
10
O
u
> 10"
3
LU
10
-4
1234
PORE VOLUMES
o.
5
CL
ON
g
<
DC
UJ
o
o
o
OFF
MAX
o -
"RESIDUAL"
CONCENTRATION
CESSATION
OF PUMPING
(CLOSURE?)
*1 '2
TIME FROM INITIATION OF EXTRACTION
Figure 6. Concentration versus time for an extraction well that is turned
oH at time I,.
Chemical Causes of Tailing
At many sites, some or even most of the contaminant mass
will not be dissolved in the ground water but will be present
as 1) adsorbed species, 2) precipitates, or 3) nonaqueous
phase liquids (NAPLs). " These reserves of matrix-associated
contaminants and contaminants in the immobile fraction of
the NAPLs cannot be directly extracted by pump-and-treat:
they must transfer from the solid or NAPL to the ground
water before they can be removed. If the equilibrium
concentration in the ground water is small relative to the total
mass of contaminant in the soil or if the rate of mass transfer
is small relative to the ground-water velocity, then large
quantities of water must pass through contaminated sections
of aquifer before it is remediated.
If reactions between solutes and stationary phases are rapid
relative to the flow rate, equilibrium partitioning can be
assumed. However, rapid equilibration does not translate to
rapid removal rates if the equilibrium concentration in solution
is very low. The retention of contaminants by mineral
surfaces and microbial cell walls, ion exchange reactions with
clays, and the partitioning of organic contaminants between
soil organic matter and the ground water can significantly
increase the time required for remediation of contaminated
aquifers. For hydrophobic, nonpolar organic compounds,
resorption can often be represented by linear isotherms
(e.g., Chiou et al., 1979). In the absence of free product
(NAPL), the number of pore volumes required to remove the
organic contaminant from a homogeneous aquifer is
approximately equal to the retardation factor, R,
Figure 5. Concentration versus time for removal of contaminants from a
ideally layered aquifer. The layers are assumed to be 10 cm thick, the
length is 10 m, the retardation factor is unity, and the diffusion
coefficient is 10 Bcrrr7s. Calculated using the equations given by
Sudickyand Frind (1962).
(1)
-------
where pb is the dry bulk density of the soil, n is porosity, foe is
the fraction of organic carbon in the soil (mass of carbon/
mass of soil), and Koc is the partition coefficient for the
contaminant into soil organic carbon (mass per unit mass of
carbon/equilibrium concentration in water). A compound with
a large Koc value can have a large retardation factor even in a
soil with a small to a moderate amount of organic carbon.
Thus, many pore volumes of water must be flushed through
the soil to remove such hydrophobic organic contaminants.
In some cases equilibrium partitioning may not be applicable.
Laboratory tests had shown that weeks are required to
achieve equilibrium concentrations in laboratory experiments
with sediments (Hamaker and Thompson, 1972; Coates and
Elzerman, 1986; Karickhoff, 1980). Resorption of pyrene,
hexachlorobenzene, and pentachlorobenzene from river
sediments requires days to weeks (Karickhoff and Morris,
1985). If such rates of resorption are slow relative to the
rate of ground water flow, then equilibrium concentrations
may not be attained during pump-and-treat and tailing in the
concentration-versus-time curve can result.
Although the linear adsorption model is adequate for
describing the adsorption equilibria of many nonpolar,
hydrophobic organic contaminants (Chiou et al., 1979), it
does not adequately describe the behavior of organic or
inorganic ions over a wide range of pH and adsorbate
concentrations. The adsorption of ionic solutes is often
represented by an adsorption isotherm. An adsorption
isotherm is a plot of the contaminant concentration on the soil
versus the equilibrium solution concentration of the
contaminant. Adsorption isotherms are defined according to
their general shape and mathematical representation. For a
Langmuir isotherm, the concentration on the soil increases
with increasing concentration in the ground water until a
maximum concentration on the soil is reached (Fig. 7). The
isotherm can be represented by the equation:
S = S
KG
max
,1 + KC
(2)
where S (mass/mass) is the concentration on the soil, S
(mass/mass) is the maximum concentration on the soil, Kmax
((length) Ymass) is the Langmuir adsorption constant, and C
(mass/( length)3) is the concentration in the ground water. A
Freundlich (or Kuster) isotherm is given by the equation:
LU M
oĞ
o S
o "5T
uj ra
8
i
n
1 + KC
AQUEOUS CONCENTRATION
Figure 7. A Langmuir isotherm.
O
p
I
o .2
UJ %
h- tr
m I
DC Ğ
g 8
< 1:
b < 1
S = KC
AQUEOUS CONCENTRATION
S = KC
(3)
Figure 8. Freundlich isotherms, A linear isotherm is the special case
for which the exponent, b, is equal to unity,
where K is the Freundlich adsorption constant and b is a
positive parameter. The shape of a Freundlich isotherm
depends on the value of b. If b is greater than 1.0, the
isotherm becomes steeper with increasing concentrations in
the ground water. If b is less than 1.0, the isotherm becomes
steeper at lower concentrations (Fig. 8). A linear isotherm is
a special case of the Freundlich isotherm where the
parameter b is equal to unity. At constant pH, cations tend to
follow Freundlich isotherms while anions tend to follow
Langmuir isotherms (Dzombak, 1986, Dzombak and
Morel, 1990).
Adsorption isotherms are useful for illustrating the
dependence of the solid phase concentration of the
contaminant on the aqueous phase concentration of the
contaminant at a given pH. However, adsorption of inorganic
ions is pH-dependent and the form of the isotherm should be
known over the entire pH range likely to be found at a site.
Sometimes this pH-dependence is presented as the fraction
of the contaminant adsorbed versus pH or a "pH-edge" (Fig,
9). For cations, the pH-edge for most minerals show little or
no adsorption at low pH. As pH increases, the portion of the
contaminant that is adsorbed increases until the fraction is
-------
unity (provided that the mass of contaminant does not
exceed the available adsorption sites). The pH-edges for
anions are the opposite to those for cations. There is little or
no adsorption at higher pH but as pH decreases the fraction
of the anion that is adsorbed increases to unity or to the ratio
of the mass of sites to the mass of contaminant if the amount
of contaminant exceeds the number of available sites. For
either cations or anions, the shape and position of the pH-
edge depends on the specific mineral surface and ions under
consideration.
100
HQ 80
LLJJQ 60
OC f\ 40
s§20
^ 0
100
h-3 8°
LUflC 60
DC (/) 40
Q-5 20
0
CATIONS
INCREASING ADSORBENT
ANIONS
INCREASING ADSORBENT
Figure 9. pH adsorption edges for cations and anions. The fraction of
contaminant adsorbed will not reach unity if the sites are saturated.
Adsorption processes can also be modeled using surface
complexation models (e.g., Stumm et al., 1976; Schindler,
1981; Schindler and Stumm, 1987; Dzombak and Morel,
1990). The key advantage of this type of approach is that it
has a foundation in chemical theory allowing the results to be
extended beyond the exact test conditions. The dependence
of the amount of adsorption on the pH of the solution and the
competition between several adsorbates for the adsorption
sites are, in principle, accounted for in such a model. The
disadvantages of the surface complexation model are the
lack of a consistent set of equilibrium constants and the
potential lack of linear additivity when multiple adsorbents are
present. The first limitation is being overcome through
compilations of consistent data sets for oxide surfaces. At
this time, there does exist a consistent set of adsorption
constants for adsorption onto hydrous ferric oxide (Dzombak,
1986; Dzombak and Morel, 1990) based on a two-layer
surface complexation model. Such data sets need to be
derived for other oxide surfaces as well.
The general concepts of ion adsorption can be applied to
anticipate some of the behavior of contaminants during
pump-and-treat remediation. The rate of removal of ionic
contaminants under acidic conditions can be substantially
different than under neutral or alkaline conditions. Adsorption
of anions, for example, is more likely to be a problem at lower
pH than at more neutral or alkaline conditions. Furthermore,
there may be changes in the amount of adsorption during
remediation as acid or alkaline waters are returned to more
neutral pH conditions. In all of these adsorption models, ionic
contaminants follow nonlinear isotherms. The partition
coefficient equals the slope of the adsorption isotherm. As
aqueous contaminant concentrations decrease during
remediation, the slope of the isotherm, hence the retardation,
changes. In most cases, the retardation will increase with
decreasing concentrations making it more difficult to
decrease intermediate concentrations below the maximum
contaminant level (MCL) than to decrease the initially high
concentrations to intermediate levels.
Large reserves of inorganic contaminants maybe formed as
the result of the precipitation of crystalline and amorphous
materials within the soils. For example, one concern is the
potential effect of a reserve of solid BaCr04within aquifer
systems contaminated with hexavalent chromium (Cr(VI))
(Palmer and Wittbrodt, 1990). As Cr(VI)-laden waters enter
the subsurface, natural Ba may react with aqueous
chromate (CrO) to precipitate a reserve of BaCr04. In many
cases, the size of the reserve will be limited by the availability
of Ba'in the soil rather than the mass of chromate spilled.
During the inital phases of a pump-and-treat remediation,
ground water containing high concentrations of Cr(VI) in
excess of available Ba2*are removed and the concentrations
in the extraction wells quickly decrease (Fig. 10). At some
point, BaCrO becomes the principal source of Cr(VI) in the
pore water. The Cr(VI) concentration will remain relatively
constant for as long as there is BaCrOjemaining in the soil.
Using equilibrium concepts, Palmer and Wittbrodt (1990)
estimated that 25 to 50 pore volumes are required to remove
Cr(VI) from soils at a hard chrome plating facility. If
equilibrium is not obtained and kinetic processes control the
amount of BaCrO4dissolved as the ground water passes
through the soil, then more pore volumes would be required.
Similar volubility limitations may occur for other inorganic
contaminants. The effect of precipitates on efficacy of pump-
and-treat remediation depends on the volubility of the mineral
phase. The most troublesome mineral precipitates are those
with solubilities low enough to create a relatively large
contaminant reserve, yet with solubilities large enough to
exceed the target clean-up levels. A complicating factor is
substitution of contaminants in the crystalline structure of
other minerals. The degree of substitution affects the
equilibrium concentration of the contaminant. Regardless of
whether the contaminant has been precipitated in pure or
substituted mineral phases, if the rate of dissolution is slow
relative to the velocity of the ground water, then the time
required for the removal of the contaminant from the
subsurface will be greater than when equilibrium conditions
have been achieved.
-------
When pump-and-treat remediation is predominantly limited by
chemical processes that restrict the transfer of mass from
these contaminant reserves to the ground water, chemical
enhancement to pump-and-treat should be considered.
Although the choice of a reactive agent that will greatly
enhance contaminant removal is a primary concern, there are
several other factors that must be considered before
implementation of a chemical-enhancement program.
MAX
<
QC
UJ
O
o
O
CONTAMINANT
CONCENTRATIONS
CONTROLLED
BY SOLUBILITY
SOLID PHASE
RESERVE
DEPLETED
tl t2
TIME FROM INITIATION OF INJECTION
Figure 10. Concentration versus time for an extraction well in a
formation that contains a solid phase precipitate.
Nonaqueous-phase liquids can also provide large reserves of
contaminants in the subsurface. For example, if a cubic
meter of soil with a 35% porosity contains trichloroethylene
(TCE) at 20% residual saturation, then approximately 270
pore volumes must pass through the soil and reach
equilibrium with the TCE (1100 mg/L) before the solvent is
removed from the soil by dissolution. Sandbox experiments
with perchloroethylene (Anderson, 1988; Anderson et al.,
1992) suggest that this equilibrium is achieved very quickly
as the water passes through fingers of residual solvent.
Longer periods of time are required to remove solvents when
they are present in pools (e.g., Johnson and Pankow, 1992;
Anderson et al., 1992). Using the equations given by Hunt et
al. (1988) and reasonable choices of transport parameters, if
can be shown that only the water that passes very close to
the edge of the solvent pool is likely to reach equilibrium
concentrations with the solvent while the concentrations
further above the pool are limited by the rate of mass transfer
from the pool to the bulk aquifer (Fig. 11). The average
concentration of the solvent measured in a monitoring well
with a two-meter length of screen placed just above the pool
will increase across the length of the pool over which the
ground water has flowed (Fig. 12). If the 103 kg of TCE used
in the previous example is distributed in a 20 cm thick pool
below the cubic meter of soil, the average concentration of
TCE in the groundwater exiting from the block of soil is 28.6
mg/L and 10,200 pore volumes must pass through the
aquifer before the TCE is removed. Thus, it takes
approximately 39 times longer to remove solvent from a pool
than to remove the same mass of solvent from residual
saturation.
0.30
^0.25
O
O
a
u
o
CD
0.20
0.15
0.10
. 0.05
0.00
GROUNDWATER VELOCITY = 0.1 m/d
TRANSVERSE DISPERSIVITY = 0.0003 m .
EFFECT. DIFFUSION COEFF. = 2.3E-5 m!/d
0.0 0.2 0.4 0.6 0.8
RELATIVE CONCENTRATION
Figure 11. Concentration of a contaminant at different elevations above
a DNAPL pool for different pool lengths (after Johnson and Pankow,
1992).
z 0.05
O
< 0.04
UJ
o
o
UJ
CC
0.03
0.02
0.01
0.00
Average Concentration Over
2.0 m Interval Above Pool
MONITORING
WELL
n
NAPL ğ
POOL ĞĞ,,,
^ -
-p-
2.0m
|
ALONG POOL ~~"1
0 2 4 6 8 10
DISTANCE ALONG POOL (m)
Figure 12. Average concentration of a contaminant over a 2-m interval
above a DNAPL pool (after Johnson and Pankow, 1992).
-------
Chemical Enhancements for Pump-and-Treat
Remediations
If chemical enhancement of pump-and-treat is to be
successful, four key areas must be satisfactorily addressed in
the design: 1 ) delivery of the reactive agent to where it is
needed within the aquifer, 2) the interaction between the
reactive agent and the contaminant, 3) the removal of the
contaminant and the reactive agent from the subsurface, 4)
the treatment of the extracted water and disposal of the
resulting sludges (Fig. 13).
INJECTION OF
REACTIVE
AGENTS
DISPOSAL
RECOVERY OF
REACTIVE AGENT
TREATMENT
EXTRACTION OF
REACTIVE AGENTS
& CONTAMINANTS
I DELIVERY | ğ I REACTION I - I REMOVAL I
Figure 13. Schematic representation of chemical enhancement ofa
pump-and-treat operation. Key areas of concern are shown in boxes.
In some cases, the reactive agent will be recovered and re-used.
Delivery
Delivery of the reactive agents to the areas within the aquifer
where they are needed to enhance the removal of
contaminants can be a complex process. Reactive agent
solutions must be injected without clogging the aquifer near
the injection well with particles and chemical precipitates.
The ground water containing the reactive agents must then
move in some reasonable period of time to the contaminated
portion of the aquifer. The rate, mode, and scheduling of the
injection and pumping must be designed such that the
reactive agent reaches those areas in a relatively short
period of time. Many of these processes are influenced by
the heterogeneities within the aquifer.
Clogging of Injection Wells
The clogging of injection wells is a common problem.
Clogging can be the result of the physical filtration of
suspended particles at the well interface, the formation of
inorganic precipitates, or the growth of microorganisms. As
more materials are entrapped, precipitated, or grown in the
pore space adjacent to the well face, they occupy increasing
amounts of the pore space and severely reduce the hydraulic
conductivity (Palmer and Cherry, 1984). A plot of the ratio of
the new permeability, k, to the original permeability, k0,
versus the new porosity,6, to original porosity,90, using the
Carmen-Kozeny model (Carmen, 1937)'and'the Rumpf-
Gupte (1971) model (Fig. 14) illustrates that small changes in
porosity can result in order of magnitude reductions in
permeability. Because this reduction in permeability is
immediately adjacent to the well screen, it is generally
manifested as a reduction in well efficiency.
The mechanisms responsible for well clogging dictate the
actions to avoid the problem. Additives may be required to
prevent precipitation. Problems with particles in the injection
water can be overcome by flocculation and removal with filter
presses. Microbiological activity can be inhibited by the
removal of nutrients or dissolved oxygen. Otherwise,
periodic treatment such as surging, jetting, development, and
acid treatment of the injection well may be required. These
options need to factored into estimates of capital costs of the
remedial design as well as the operation and maintenance
costs.
0.1
0.01
0.001
Carmen (1937), n0 =0.47
I Carmen (1937), nQ =0.20
Rumpf and Gupte (1971)
0.9 0.8
0.7
0.6 0.5 0.4 0.3
n/n,
Figure 14. Ratio of permeability to initial permeability versus porosity to
initial porosity using several empirical models (after Palmer and Cherry,
1964).
Transport of the Reactive Aaent to Contaminated Areas
A fundamental problem with chemical enhancements to
pump-and-treat remediation is getting the reactive agent to
-------
the contaminated portions of the aquifer so that it can interact
with the contaminants to facilitate their removal. Two major
problems should be considered: 1) differential flow paths
governed by well hydraulics, and 2) mass transfer between
heterogeneities that is governed by molecular diffusion. The
paths the injected water follows depend on the hydraulic
properties of the medium, the aquifer thickness, the natural
hydraulic gradient, the rate of injection, the placement and
pumping (injection) rate of other nearby wells, and the
location and type of hydraulic boundaries in the vicinity of the
site. Some of these effects are illustrated for homogeneous,
isotropic aquifers with stream function calculations (Fig. 15).
The two wells in Fig. 15A are 50 feet apart. The up-gradient
well injects fluid at a rate of 0.5 gpm while the down-gradient
extraction well removes water at the same rate. Not all of the
stream lines from the injection well are captured by the
extraction well. In this particular case, 20% of the injection
fluid continues to be transported through the aquifer. If the
wells are placed 25 feet from one another (Fig. 15B), then
less of the injected fluid is lost (159'.) but a much smaller
volume of the aquifer is remediated with the chemical
extractant. If the rate of injection is reduced to 0.25 gpm, all
of the injected fluid is captured by the extraction well (Fig.
15 C); however, the volume of affected aquifer is diminished
even further.
150
100
50
LJJ
O
52 -so
Q
-100
-150
NATURAL -
HYDRAULIC
GRADIENT
EXTRACTION
WELL
INJECTION
WELL
-150 -100
-50 0 50
DISTANCE (ft)
100 150
Figure 15a. Flow lines between an injection well and an extraction well
located 50 feet apart and with a natural gradient of 0.007 to the north.
Extraction rate and injection rate are both 0.5 gpm.
While, in principle, the distribution of wells and the rate of
injection and extraction could be optimized, in practice, it may
prove to be difficult. Firstly, the factors to be optimized are
not universally agreed upon. Possibilities may include
minimizing the number of wells, the total cost, or the time for
cleanup, or maximizing the mass of contaminant removed
per unit time. Secondly, none of the necessary parameters
are known with certainty. For example, concentration
isopleths are known only within some range of distance, the
variation in the hydraulic parameters have not been
measured, and economic factors such as interest rates and
costs over the next few years are based on extrapolation of
current conditions.
150
100
^ 50
UJ
O 0
-50
O
-1oo
-150
NATURAL
HYDRAULIC
GRADIENT
EXTRACTION
WELL
INJECTION
WELL
-150 -100
-50 0 50
DISTANCE (ft)
100 150
Figure 15b, Flow lines between an injection well and an extraction well
located 25 feet apart and with a natural gradient of 0.007 to the north.
Extraction rate and injection rate are both 0.5 gpm.
150
100
^ 50
£
UJ
O 0
-50
Q
-100
-150
NATURAL -
HYDRAULIC
GRADIENT
EXTRACTION
WELL
INJECTION
WELL
I | | 1
-150 -100
-50 0 50
DISTANCE (ft)
100 150
Figure 15c. Row lines between an injection well and an extraction well
located 25 feet apart and with a natural gradient of 0.007 to the north
Extraction rate is 0.5 gpm while the injection rate is 0.25 gpm.
-------
Aquifer heterogeneity is an important factor in the rate of
transport of dissolved constituents. As a reactive agent is
injected, it is advected along the higher permeability
pathways through the aquifer. When the hydraulic
conductivity of the low permeability lenses is two or more
orders of magnitude less than that of the high permeability
zones, transport of the reactive agent into the lower
permeability zones is controlled by molecular diffusion. The
concentration of the reactive agent in the higher permeability
zones should be maintained long enough to permit the
diffusion of the reactive agent into the lower permeability
lenses. The time it takes the reactive agent to diffuse into the
lower permeability lenses depends on the diffusion
coefficient, the concentration, and the soil-water interactions
of the reactive agent.
The process of diffusion can be complex in a multicomponent
system. If a completing agent is used to enhance the pump-
and-treat remediation, the complex initially forms in the
higher permeability zone, creating a concentration gradient
between the higher permeability layers and the lower
permeability layers. Thus, not only does the reactive agent
diffuse into the lower permeability zones, but so may the
complex. The extent to which this affects the time for
remediation depends upon the relative rates of diffusion of
the solutes into and out of the lower permeability lenses and
the relative rates of reaction for the important chemical
processes.
Modes of Injection
The above discussion implicitly assumes that the reactive
agent is continuously injected into the aquifer, but there are
several options for injecting the fluid as well as for adding the
reactive agent. The water may be injected continuously, as a
series of pulses, or as a slug. With continuous fluid injection
the reactive agent may be added to the injection fluid
continuously or it may be added in a pulsed or slug mode.
For pulsed injection of the fluid, the reactive agent maybe
injected with each pulse or as a slug into a single pulse of
fluid. For a slug injection of fluid, the reactive agent can be
added as a single slug. Thus, there are six general
combinations of fluid and reactive agent injection modes (Fig.
16). The advantages and disadvantages of these modes of
injection (Table 1 ) were used to determine the suitability of
each method for chemical enhancements to pump-and-treat
(Fig. 16).
The modes most likely to be useful are continuous fluid
injection with either continuous or pulsed addtition of the
reactive agent. The use of the slug mode for the injection of
water is not practical; a relatively small slug of water injected
into the aquifer does not affect a large volume of the aquifer.
The major concern is the volume of aquifer over which the
injected water is transported and how long it takes to
accomplish this distribution. While continuous injection may
provide the shortest time period over which it takes to
distribute the fluid over a given volume of aquifer, there may
be inherent economic benefits of using a pulsed mode. If
staff must be present during injection, then coordinating the
injection pulses with personnel shifts may be advantageous.
A key advantage of continuous addition of the reactive agent
during continuous fluid injection is that relatively high
concentrations of the agent can be maintained within the
higher permeability zones of the aquifer. This creates a large
concentration gradient between the higher permeability
lenses and the lower permeability lenses, potentially reducing
the time to remove the contaminants from those zones.
However, this means that a larger mass of the reactive agent
is required. A slug addition of the reactive agent would
require less mass; however, the concentrations may
decrease with time because of the nonlinear relationship
between the adsorbed and aqueous concentrations. Pulsed
addition would repeatedly increase the concentration
gradients between the higher and lower permeability zones,
however, the gradients may be locally reversed between the
passing pulses. The system would have to be more carefully
designed to ensure that the net direction of diffusion of the
reactive agent is into the low permeability zones.
REACTIVE
AGENT
ADDITION
MODE
CONTINUOUS
PULSED
SLUG
FLUID INJECTION MODE
CONTINUOUS
good
good
fair
PULSED
N.A.
good
fair
SLUG
N.A.
N.A.
poor
N.A. = not applicable
Figure 16. Relative rating for different combinations of the modes of
injection of the fluid and reactive agent.
Timing of Injectibn 'of the Reactive Agent
Chemical enhancement can be initiated at any time during
the pump-and-treat remediation. Injection of the reactive
agent may begin 1 ) as the extraction program begins, 2) after
the concentration-versus-time curve significantly flattens, or
3) at some time intermediate to the first two.
An advantage of initiating injection of the reactive agent at
the same time that extraction of the contaminated water
begins is that it provides the earliest start to the chemical
enhancements. The flow paths over which the reactive agent
must travel remain approximately the same and therefore
any delay in initiating the injection also delays aquifer
cleanup. However, several problems may arise suggesting
that this method may not always be the most advantageous.
If high concentrations of contaminant are still within the
aquifer, then a greater concentration of the reactive agent
may be required to remove this material. The cost of the
additional reactive chemicals should be compared with the
cost of a longer extraction time. Geochemical interactions
within the contaminant plume can result in elevated
concentrations of solutes other than the contaminant (e.g.,
10
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Table 1. Modes of injection for chemical enhancement to pump-and-treat
FLUID
INJECTION
MODE
ADVANTAGES
DISADVANTAGES
CONTINUOUS
SLUG
PULSED
ACTIVE AGENT
INJECTION
MODE
CONTINUOUS
SLUG
PULSED
Fluid distributed over wide area.
Less maintenance of pumping schedules.
Less volume of water.
Less potential dogging of wells.
Can be developed around working
schedules.
Maintain concentration in high
permeability zones allowing for
diffusion into low permeability zones.
Requires less mass of active agent.
Less total mass of active agent.
Can be planned around work schedules.
Allows for sufficient time for
diffusion.
Greater potential for clogging of
screens.
Greater pumping costs.
Fluid distributed over very small volume
of aquifer.
Requires more design to insure injection
and off periods are balanced relative to
natural groundwater flow.
Requires more mass of active agent,
May not allow sufficient time for
diffusion into low permeability lenses.
Concentration decreases with
time/distance which can reduce
effectiveness of the active agent.
Requires greater maintenance/control.
Requires more analysis to insure that
injection and off periods are of
sufficient length.
Fe and Mn in anoxic plumes). High concentrations of these
solutes may severely impede the effectiveness of the reactive
agent, thereby requiring a greater mass of reactive agent to
be injected into the aquifer. In some cases, these elevated
levels of solutes may result in precipitation of the reactive
agent and clogging of the aquifer.
Some of these problems can be avoided by using
conventional pump-and-treat until the rate of decline in the
concentration of the contaminant is low. As the
concentrations of the contaminant decline, the high levels of
other solutes present within the contaminated area are also
likely to decline toward background levels. At lower
concentrations of interfering solutes, the reactive agent
interacts more efficiently with that portion of the contaminant
that is most difficult to remove by conventional pump-and-
treat. However, it may require the removal of several pore
volumes over several years to reach these lower
concentration levels so added costs for operations and
maintenance may be incurred.
The third possibility is to initiate chemical enhancement at
some time between the initial start-up of the pump-and-treat
system and the time it takes for the concentrations to level
off. Deciding the optimal time requires a more sophisticated
analysis then either of the previous two choices, however,
some simple criteria may serve as guides. For example, if
there is concern over the precipitation of a solid phase, the
criterion may be the time at which the concentration of the
interfering ion decreases below the critical concentration
computed from the volubility limit and the concentration of the
reactive agent. While this is not necessarily the "optimal"
time, it may serve as a practical estimate.
Rptarrlatinn anrl thp Rate of Transport nf thp
Agent
The reactive agent must travel through the aquifer almost as
quickly as the water. If the reactive agent is significantly
retarded, then it may take longer for the reactive agent to be
transported to the target areas of the aquifer than it takes to
11
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remove the contaminant from the subsurface without
chemical enhancement. Thus, there is a possible paradox
here: the reactive agent must react within the subsurface to
enhance the removal of the contaminant, yet it must not be
retarded. However, careful consideration of the chemistry of
the system may allow this paradox to be resolved so that
both objectives are achieved.
If the reactive agent is chosen to compete with the
contaminant for adsorption sites, both of these objectives can
be realized by controlling the concentration of the reactive
agent. For example, if the adsorption of the reactive agent
follows a Langmuir-type adsorption isotherm, the amount of
retardation is insignificant if the concentrations are high
enough to saturate all of the available adsorption sites in the
soil. If a completing agent is utilized to enhance the removal
of the contaminant, then it should be chosen so that neither
the agent nor the complex are significantly retarded.
Reactive agents that are reducing or oxidizing agents maybe
retarded as they react with materials other than the
contaminant. Injection may have to simply continue until all
of the material between the injection point and the extraction
well that reacts faster than the contaminant is titrated from
the aquifer. The amount of oxidant or reductant required to
titrate the soil can be estimated from the oxidation and
reduction capacities of the soil (Barcelona and Helm, 1991).
In some cases, if something is known about the reaction
rates for these redox reactions, some control can be obtained
through control of the ground-water velocities (i.e., by
controlling the rates of injection and extraction).
Reactive Agent - Contaminant Interactions
Different reactive agents can be chosen (Table 2) depending
upon the processes that control the tailing in the
concentration-versus-time curve for the extraction wells. The
reactive agent may compete with the contaminant for
adsorption sites, complex the contaminant, change the redox
state of the contaminant, change the solvation properties of
the groundwater, act as a surfactant, ionize the contaminant,
or substitute for the contaminant in a precipitate. These
possibilities are not necessarily exclusive of one another. For
example, a reactive agent may change the redox state of the
contaminant and then form a complex with the altered form.
Completion for Adsorption Sites
If the tailing of the concentration-versus-time curves for the
extraction wells are controlled by adsorption processes, the
reactive agent can be chosen to compete for the adsorption
sites. Such competition is most likely to be effective for ionic
solutes and least effective in displacing neutral organic
molecules partitioned into soil organic matter.
The general concepts of ion adsorption can be applied to
anticipate some of the constraints on the use of reactive
agents to enhance pump-and-treat remediation. The
adsorption of ions onto hydrous ferric oxide may be used as
a model for appreciating the qualitative effects. Ultimately,
laboratory tests utilizing the site contaminants and geologic
materials should be performed. In general, we expect
competition to be significant only when the adsorption sites
are near saturation. An ionic contaminant can be easily
displaced by a reactive agent with similar adsorption
properties if the concentration of the reactive agent is
sufficient to saturate the adsorption sites. Ionic contaminants
can also be displaced by a reactive agent with a lower
adsorption affinity but only if the agent is present in great
excess of the total number of adsorption sites in the soil.
Table 2. Reactive agents - contaminant interactions
Competition for adsorption sites
Complexation of the contaminant
Cosolvent effects
Enhanced mobilization and solubilization by surfactants
Oxidation
Reduction
Precipitation/Dissolution
lonization
Complexation
The reactive agent may be effective in forming aqueous
complexes with an ionic contaminant. The aqueous
complexes are not expected to be adsorbed as readily as the
noncomplexed contaminant, therefore they are more mobile
and relatively easy to remove by pump-and-treat technology.
For example, James and Bartlett (1 983) found that citric and
diethylenetriaminepentaacetic (DTPA) acids complexed
Cr(lll) sufficiently to maintain elevated solution concentrations
at pH 7.5 and 6.5, respectively. These organic acid anions
can also contribute to the removal of chromium by competing
with chromate for adsorption sites at oxide surfaces. Citrate
may also contribute to the reduction of the Cr(VI) to Cr(lll)
which can then be complexed by another citrate molecule.
While there are many such potential chelators, those that are
environmentally safe enough to be used in aquifers are
relatively weak and non-specific in their binding action. Two
key consequences of these properties are that 1) the chelator
must be present in great excess of the contaminant
concentrations and 2) high concentrations of common,
nonhazardous soil constituents such as Fe, Mn,and Al may
also be removed (e.g., Grove and Ellis, 1980; Norvell, 1984) .
The presence of these constituents in the waste stream may
substantially increase the costs of treatment and disposal
over conventional pump-and-treat. Some chelating agents,
such as citrate, can be utilized as substrate by bacteria in the
subsurface. As the concentrations of contaminants decrease
below toxic levels around injection wells, bacterial growth
may increase to the level where the increased biomass clogs
the aquifer and wells.
In some cases, the adsorption properties of the soil matrix
may be altered by the use of chemical extractants. The
removal of iron and aluminum oxide surfaces should
decrease the adsorption density of the geologic materials.
Zachara et al, (1988b); however, found the adsorption of
12
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chromate onto kaolinite increased with treatment with
dithionate-citrate-bicarbonate (DCS) or hydroxylamine-
hydrochloride (NHZOHHCI) solutions. The reasons for this
increase in adsorption are not clear.
Treatability studies should be conducted to determine not
only the efficacy with which contaminants are removed by
such chelators but also to estimate the total load of metals
that must be treated and disposed, and the potential
increases in biomass. The cost of these additional loads
must then be compared with the costs of conventional pump-
and-treat remediation.
Cosolvents
The rate of removal of hydrophobic organic contaminants is
often limited by their relatively low volubility in water.
However, the solubilities of many of these contaminants are
much greater in other solvents. Theoretical models
suggesting an exponential decrease in the amount of
adsorbed organic contaminant with increasing fractions of
water miscible solvents (Rao et al., 1985; Woodburn et al.,
1986) have been substantiated in laboratory experiments for
several organic compounds (Rao et al., 1985; Nkedi-Kizza et
al., 1985, 1987; Mahmood and Sims, 1985; Woodburn et al.,
1986; Fu and Luthy, 1986a, 1986b; Zachara et al., 1988a).
For example, the adsorption coefficient for anthracene in
methanol-water mixtures decreased by four orders of
magnitude as the fraction of methanol was increased from O
to 1 (Nkedi-Kizza et al., 1985). The injection of cosolvents
may therefore be expected to drastically increase the
volubility and decrease the retardation factors for these
organic compounds thereby facilitating their removal from the
subsurface. Cosolvents that are used as substrate by
microbes may have the added advantage of promoting co-
metabolism of primary contaminants. Small amounts of
biodegradable cosolvent that are difficult to remove from the
subsurface will be of less concern because of their eventual
transformation. Thus, cosolvents, such as alcohols, are
potentially effective reactive agents for chemical
enhancement to pump-and-treat of hydrophobic organics.
However, some consequences of cosolvent injection maybe
less desirable.
The order-of-magnitude decreases in adsorbed contaminants
are generally achieved with cosolvent concentrations greater
than 20%. Fluids containing this amount of cosolvent will
have densities and viscosities that differ substantially from
the ground water. Thus, the transport behavior of these
fluids is more complex and more difficult to predict than for
fluids with homogeneous properties. Cosolvent interaction
with clays in the aquifer matrix may either increase or
decrease the permeability of the soil. Cracks have appeared
in soils treated with methanol (Brown and Anderson, 1982).
The formation of such high permeability pathways maybe
particularly troublesome at sites where dense nonaqueous
phase liquids (DNAPLs) are present. Cosolvents such as
methanol can serve as substrate for subsurface microbes
resulting in biofouling of the aquifer. Biotransformation may
substantially alter the geochemistry of the aquifer and
promote the reductive dissolution of Fe and Mn oxides.
These metals can create problems with well clogging and
interfere with surface treatment. Also, additional treatment
facilities must be constructed for the separation of the
cosolvent from the water. These facilities incur capital
expenditures as well as operation and maintenance costs.
Surfactants
A surfactant adsorbs to interfaces and significantly decreases
the interracial tension (Rosen, 1978). This property of
surfactants has made these chemicals useful in enhanced oil
recovery and several researchers have proposed their use in
the remediation of NAPL-contaminated sites (e.g., Ellis et al,
1985). In general, surfactants are composed of a
hydrophobic moiety, often a long chain aliphatic (Cloto C20)
group, and a hydrophilic moiety (Fig. 17) that can be anionic,
cationic, nonionic, or zwitterionic (i.e., possess both positive
and negative charges). The orientation of the surfactant
molecules at an interface can reduce the interracial tension
and alter the wetting properties of the soil matrix. When the
interface is a nonaqueous phase liquid, the lowering of the
interracial tensions decreases the capillary forces keeping the
NAPL in place and results in greater mobility of the NAPL.
HYDROPHILIC
MOIETY
ALKYLBENZENE SULFONATE
CH, CH, CH, CH
CH3CHCH2CHCH2CHCH2CH
HYDROPHOBIC MOIETY
Figure 17. The surfactant alkylbenzene sulfonate.
For enhanced oil recovery, increased mobility of the NAPL
allows a greater fraction of the available oil to be pumped to
the surface. In the case of NAPLs that have a greater
density than water (DNAPLs), increased mobility is not
necessarily desirable. Once mobilized, the DNAPL may
migrate deeper into the aquifer. If the DNAPL migrates into
areas that were previously uncontaminated, additional wells
and pumps will be required and the costs of remediation will
increase accordingly.
Surfactants can also promote the solubilization of
hydrophobic organic contaminants. Above a critical
concentration known as the "critical micelle concentration,"
colloidal-size micelles can form by the aggregation of the
monomeric surfactant molecules. In water, the micelles
form by the hydrophobic moieties grouping together in the
core of the micelle, and the hydrophilic moieties orienting
toward the surface of the micelle (Fig. 18). Hydrophobic
organic contaminants partition into the hydrophobic core
of the micelle thereby increasing the volubility of the
organic contaminant.
13
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MICELLE FORMATION
HYDROPHOBIC
CORE
HYDROPHILIC
SURFACE
Figure 18. Aggregation of surfactant molecules into a micelle.
Although the successful application of surfactants to
enhanced oil recovery has been demonstrated, transfer of
this knowledge to aquifer remediation is not direct.
Surfactants used for enhanced oil recovery are chosen on
the basis of temperatures and salinities that are much higher
than those at most hazardous waste sites. To achieve the
desired behavior, the surfactant must be chosen for the
solvent under the conditions of use (Rosen, 1978). Incorrect
surfactant formulations may result in high-viscosity
macroemulsions that are difficult to remove. The surfactant
can alter the wetting properties of the soil matrix and cause
the NAPL to become the wetting phase. The NAPL would
then occupy the smaller pores of the soil matrix, thereby
exacerbating clean-up efforts. The toxicity and potential
biodegradation of surfactants that will remain in the aquifer
following NAPL removal is of great concern in shallow
aquifers.
The use of surfactants for aquifer remediation looks
promising; however, there is little experience in their
application. Laboratory experiments have demonstrated
enhanced removal of anthracene and biphenyls (Vignon and
Rubin, 1989), petroleum hydrocarbons (Ellis et al., 1985),
DDT and trichlorobenzene (Kile and Chiou, 1989), automatic
transmission fluid (Abdul et al., 1990), and PCBS (Ellis et al.,
1985; Abdul and Gibson, 1991). Surfactant mixtures that
specifically address the needs for aquifer remediation need to
be developed and tested in the field as well as in the
laboratory. When DNAPLs are present, mixtures that
increase solubilization more than mobilization may be
desired.
Qxidants-Reductants
The addition of a reactive agent that changes the oxidation
state of a contaminant is potentially useful for 1 ) decreasing
the toxicity of the contaminant, 2) increasing its mobility, or 3)
increasing its susceptibility to completing agents. For
example, chromium can be reduced from the more toxic
Cr(VI) to the less toxic Cr(lll). The oxidation of selenite
(Se(IV)) to selenate (Se(VI)) results in a solute that is less
toxic and more mobile. However, oxidants and reductants
are not specific and must .therefore, be in excess of the
amount of contaminant. This will locally alter redox
conditions within the aquifer and may result in the
precipitation of solid phases that may clog the aquifer and
injection/extraction wells or mobilization of other metals that
must be handled in the treatment train.
The rate of reaction is an important factor in considering an
oxidation or reduction reaction to facilitate the removal of a
contaminant from the subsurface. Often the rates are
strongly dependent on pH. For example, rate of reduction of
Cr(VI) by ferrous iron varies with {H*}3(Wiberg, 1965).
Precipitation-Dissolirtion-lonfzatiorl
At metal-contaminated sites, it maybe possible to add a
chemical constituent that will cause the precipitation of the
contaminant in a solid phase with very low volubility. For
example, the neutralization of acid mine waters by
carbonate-buffered solutions will cause the precipitation of
metal-oxides, hydroxides, and carbonates. Pb2*can
precipitate as a relatively insoluble PbC03phase. While this
may reduce the risk of contaminant concentrations of
exceeding the MCL, it does not remove the metals from the
site. The precipitates can continue to act as long-term, low-
level sources and the contaminants may still enter the
biosphere through plant root systems or erosion. In addition,
the precipitation of metal oxyhydroxides and carbonates can
cause clogging of the aquifer and severe reduction in well
efficiency.
Remediation of contaminated sites by conventional pump-
and-treat may often be limited by dissolution of sparingly
soluble mineral phases and nonaqueous phase liquids.
Reactive agents that increase the volubility of these phases
will release the contaminant into solution where it can be
removed via an extraction well. For example, many phenolic
compounds can be ionized at higher pH (e.g., Palmer and
Johnson, 1992). The use of a base as a reactive agent will
enhance the volubility of the phenolic phase and decrease
the retardation factor of the dissolved compounds. If
cadmium is being released into solution from CdCO3, the
addition of acid can dissolve the carbonate mineral phase
and bring the Cd2*into solution. However, such treatments
are not selective and other ions including Fe, Al, and SiO2
will be added to solution. These ions may interfere with
treatment processes and increase the volume of sludge to
be disposed. The natural buffering capacity of the aquifer
will require that the concentration of injected acid or base be
in excess of the amount of contaminant in the subsurface.
At metals-contaminated sites where remediation is limited by
the presence of a sparingly soluble mineral phase, it may be
possible to release the contaminant more rapidly by the
addition of ion that will substitute for the contaminant within
the mineral phase. This is most likely to be applicable where
the availability of one of the counter ions in the solid phase is
limited. By scavenging the counter ion into another solid
phase, the contaminant will be released into solution where it
can be easily removed. For example, if BaCrOJimits the
remediation of chromium-contaminated sites, the injection of
high levels of su If ate would precipitate BaS04and increase
14
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the volubility of the BaCr04component, thereby allowing the
removal of the Cr(VI) in fewer pore volumes than by
conventional pump-and-treat.
Such chemical enhancement methodologies require specific
knowledge about the mineral phases limiting aquifer
remediation and the geochemistry of the ground waters on
site. Detailed geochemical studies will increase the cost of
site investigation and feasibility studies but may more than
pay for themselves if efficient removal methods can be
identified and problems associated with the implementation
and operation of the clean-up effort are avoided.
Removal of the Contaminants and Reactive Agents
from the Subsurface
The basic concept of chemical enhancement to pump-and-
treat is to increase the mobility of the contaminants in the
subsurface so they may be more easily removed via
extraction wells. Removal of contaminants from the
subsurface in a chemical enhancement scheme, therefore,
requires decisions about the density, placement, and
pumping rates of these extraction wells. Several aspects of
such an enhanced extraction system design will be similar to
those utilized in conventional pump-and-treat remediations.
For example, the contaminants must be contained within the
capture zones of the extraction wells. However, key
differences between conventional pump-and-treat and
chemical-enhancement extraction system designs need to be
explored. For example, in conventional pump-and-treat,
contaminants are expected to be retarded. Therefore, the
density of the well system is typically chosen to shorten travel
times and thereby decrease the time for remediation to the
extent possible. In chemical-enhancement methods, the
reactive agent is used to make the contaminants behave
more like a nonreactive tracer. Therefore, it may be possible
to utilize a lower density of extraction wells to achieve an
effective removal of the mobilized contaminant. Another
important consideration in the design of an extraction system
for chemical enhancement is the coordination with the
injection of the reactive agent. As described above, there
can be advantages to initially removing high levels of
contaminants by conventional pump-and-treat before
initiating the injection of the reactive agent. Also, pumping
rates for the extraction wells may be adjusted during the
injection to aid in the distribution of the reactive agent within
the aquifer.
Use of a reactive agent in a pump-and-treat scheme
introduces one or more new chemical constituents into the
subsurface. To be effective, reactive agents generally must
be added to an aquifer at non-trace concentrations. Even if
the reactive agent is harmless to human health, state and
federal regulations will often require that concentrations of
the reactive agent be lowered to some permissible level.
Removal of the agent then involves all of the problems
encountered in the removal of the original contaminant and in
some cases the agent may even be more difficult to remove.
For example, if the reactive agent is a solute that is used to
compete with the contaminant for the adsorption sites, then
the reactive agent must have a greater affinity for the
adsorbent; but this greater affinity also makes it more difficult
to remove from the subsurface. There still maybe a net
benefit if the target clean-up level for the reactive agent is
greater than for the contaminant. It can also be argued that
the net risk is reduced because the reactive agent must, by
any reasonable choice, be less toxic than the original
contaminant.
One complication that may arise during the removal of the
reactive agent from the subsurface is clogging of the screen
and filterpack as waters are mixed at the extraction wells.
This problem is likely to be most acute when the reactive
agent changes the redox conditions in the subsurface. As
oxidized waters mix with reduced waters that contain iron,
precipitates may clog the screen, pipes, and treatment tanks.
Treatment and Disposal
The previous sections have outlined many of the technical
considerations that must be addressed to implement an
effective chemical enhancement strategy. However, even a
chemical enhancement plan that is completely satisfactory in
terms of subsurface deployment and removal of solutions
may still present technical difficulties in the treatment and
handling of the extracted wastes. Three broad categories of
post-extraction problems are discussed in this section: the
effects of the reactive agent on the treatment of the target
contaminants, the removal of the reactive agent from the
waste stream before discharge, and the recovery and reuse
of the reactive agent.
Removal of the Reactive Agent before Discharge
As described above, the use of a reactive agent in a pump-
and-treat scheme introduces one or more new chemical
constituents in non-trace levels into the water brought to the
surface. Extracted water will, therefore, contain substantial
quantities of the reactive agent, probably in excess of the
target contaminants. State and federal regulations will often
require that concentrations of the reactive agent be lowered
to some permitted discharge level. For example, if
phosphate is used as an extractant, standards may restrict
the permissible concentration in discharges from the
treatment facility, even if the treated wastes are routed into a
municipal sewage treatment system. Limits on phosphate
discharges can be anticipated in localities in which
eutrophication is a problem in waters receiving regional waste
waters.
If the levels of reactive agent in discharges from the site are
regulated, then the treatment plan must explicitly include a
means of removing the agent. In many cases, the most
efficient system will effect the removal of the reactive agent
simultaneously with the treatment of the targeted
contaminants. For example, if phosphate were used to
enhance chromate removal, then the neutralization step in a
treatment process could be modified to induce the
precipitation of much of the phosphate. Phosphate and
reduced Cr would precipitate in the same step and could be
removed collectively in the sludge.
For the specific system mentioned, note that the treatment
procedure would need to be modified. Removal of Cr3*alone
can be effected by addition of NaOH to achieve a basic pH.
The resulting sludge then contains a mixture of Cr
hydroxides, probably coprecipitated with by-products of the
reductant step. For instance, if bisulfife is the reducing
agent, sulfate and bisulfite will constitute part of the sludge.
15
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If phosphate is present in the extracted water, some
phosphate is likely to coprecipitate with Cr3*. However, the
concentration of phosphate remaining in solution in such a
complex system would be difficult to accurately predict and
tests would be necessary to find the optimal pH for
phosphate precipitation.
It is possible that pH adjustment alone would not precipitate
sufficient phosphate. In that case, an additional treatment
reagent would be needed; for example, substituting Ca(OH)2
for some or all of the NaOH in the neutralization step. Ca-
phosphates are relatively insoluble and would strip out much
of the phosphate. However, pilot studies would be needed to
ensure that the sludges produced behaved in the desired
fashion. The presence of phosphate might decrease the
density of the sludge so that longer settling times are
required. Furthermore, the presence of Ca2*could lead to
the buildup of scale in unexpected parts of the system. In
general, it is wise to bear in mind that the treatment process
will be efficient only if it is regarded as a coordinated
chemical system in which the alteration of one part can cause
dramatic changes in the behavior of another part.
If the reactive agent is successfully coprecipitated with the
target contaminants, the total volume of sludge sent to
disposal will increase correspondingly. Although the reactive
agent may be harmless, once it is commingled with a toxic
waste, the entire volume could be classified as hazardous
and the cost of disposal assessed accordingly. Thus, while
the removal of contaminant and reactive agent in a single
step may save operation or capital costs, the increased
volume of sludge to be landfilled over the case where no
reactive agent is used, will generate costs that will offset
some of the savings. The cost of testing and design of the
removal system for the reactive agent must also be factored
into the economic analysis.
In some cases, it will be desirable or essential to remove the
reactive agent in a separate stage of the treatment system.
One reason for a separate treatment step is to achieve a
better removal of the reactive agent than could be
conveniently achieved in a single step. In the example
above, it might be disirable to optimize Cr3*precipitation
without regard for phosphate and then strip phosphate out of
the supernatant with Ca(OH)2or alum treatment in a
subsequent step. If the Cr3*precipitation step could be
designed to minimize phosphate coprecipitation, and if the
phosphate sludge were sufficiently free of Cr to be classified
as non-hazardous, the two-stage removal would have the
added advantage of minimizing the volume of hazardous
solids for disposal.
Interference of Reactive Agent with Treatment
Processes
Even if regulations do not require the removal of reactive
agents, it may be necessary to remove them from the
process stream. Some reactive agents may be harmless to
humans or to the environment, but they nonetheless may
have chemical properties that alter the behavior of the
contaminants in the waste stream. For example, many
conceivable reactive agents would function by completing the
target contaminants and enhancing their volubility.
Specifically, citrate or oxalate salts might be used to bind up
and mobilize metal ions. The same solubilization of metal
ions that is desirable in the extraction step may become a
major headache in the treatment step.
As discussed above, most chelators that are environmentally
safe enough to be used in aquifers, such as malonate,
succinate, and citrate, will be relatively weak and non-specific
in their binding action. A weak affinity for the target metal
means that the chelator must be present in great excess and
will be found in corresponding excess in the extracted water.
The excess chelator may interfere with one or more
segments of the treatment process by binding to the target
contaminant.
The most obvious interference would be the inhibition of
precipitation. The soluble metal-chelator complex will not
readily precipitate out of solution. Lowering the pH will
dissociate most metal complexes (such as citrate or oxalate),
but metal ions will be soluble at the lower pH. Raising the pH
will favor precipitation, but the higher pH also favors the
binding of most chelators. Very caustic pH levels maybe
required to induce precipitation of metals in the presence of
excess chelator. An alternative scheme is to remove the
chelator from solution before the metal precipitation step. An
organic chelator such as citrate could be degraded by
biological treatment. An inorganic chelator such as
polyphosphate would not biodegrade and would be difficult to
remove economically by chemical means. Any process that
requires a separate treatment step for the chelator will have
greater operation and capital costs than the corresponding
process in the absence of that step.
The nonspecificity of chelators used as reactive agents
creates another substantial problem for treatment processes,
Aquifer materials may contain large amounts of naturally-
occurring metals that may be solubilized by the chelator. Iron
and manganese will be the most important in many aquifers,
but copper, zinc, aluminum, and other metals may be
extracted to varying degrees. The solution brought to the
surface will contain not only free chelator, but also substantial
quantities of chelator bound to non-target metals. In some
situations, the amount of chelator actually bound to the target
metal(s) may be only a small fraction of the total chelator in
the extracted water. The additional metals would not be
solubilized in the absence of chelator, so these metals are a
specific feature of chemical enhancement.
Large quantities of iron and manganese in the extracted
water will require special attention in the treatment process.
Precipitation of the target metals will be accompanied by the
precipitation of substantial amounts of iron and manganese
hydroxides. The sludge volume will be correspondingly
increased, with the concomitant elevation of disposal costs.
Furthermore, the behavior of iron in particular is apt to be
different from that of many target metals. Iron is an
especially insoluble metal (under aerobic conditions) and may
therefore precipitate out of the waste stream before the
target metal precipitates. If this precipitation can be
anticipated and controlled, it may simplify the treatment
process. However a more likely scenario is that iron
precipitation will be somewhat unpredictable and will
occur in inopportune sections of the treatment facility,
causing plugging or fouling of the equipment and interfere
with the treatment process by coprecipitating the
contaminants.
16
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Pornuorw anrl
nf tho Poartiuo
Agent
Site Characterization for Chemical Enhancements
In some cases, it maybe beneficial to remove the reactive
agent from the waste stream so that if can be re-injected into
the aquifer and re-used for additional extraction of the
subsurface contaminants. Such re-use may be particularly
advantageous when the reactive agent is expensive or when
it must be removed because of concerns about interference
with treatment or because of regulatry requirements on
discharges of the treated water. However, the methods that
may be used to extract the reactive agent from the waste
stream for re-use may not be the optimal methods for
removal for other purposes.
Effects of Rapid and Concentrated Extraction
One of the assumptions underlying the use of chemical
enhancement in a pump-and-treat operation is that the rapid
extraction of concentrated waste solutions is beneficial to the
clean-up operation. Of course, this will be true at many
sites because annual operation and maintenance costs are
directly reduced by a more rapid removal of contaminants.
Furthermore, more concentrated wastes may be easier to
treat than dilute wastes. However, an accurate economic
analysis of the various options available in restoring an
aquifer should consider all costs associated with chemical
enhancement. Besides the above-mentioned increases in
research and development and operation and maintenance
associated with chemical enhancement, there may be
added costs due to the rapid extraction and concentration
of wastes.
Careful consideration should be given to the capital costs of
designing facilities that can efficiently handle and treat large
volumes of a target contaminant in a short time. Injection/
extraction wells, settling basins, sludge pumps, metering
pumps, and other facilities may need to be greatly expanded
to handle the concentrated waste load. Although the system
will be operated for a shorter period of time there will be a
tradeoff between increased capital costs and lower operation
and maintence costs. It may be cheaper to run a pump-and-
treat operation for ten years with a small facility, rather than
build a much larger facility that will only need to operate for
one or two years.
Furthermore, planning in the pilot stages should give careful
attention to the performance of the treatment process at
different contaminant concentrations. If the process is
especially efficient and reliable at the high concentrations
attainable only with chemical enhancement, then the
additional costs will be mitigated. If, however, the
treatment process becomes more difficult or unreliable when
contaminants are very concentrated, then the use of
chemical enhancement may be contraindicated. Special
care is needed if a biological treatment step is anticipated.
Microbes generally thrive at higher substrate concentrations,
but if higher contaminant levels lead to toxic levels of the
target contaminant or some secondary constituent, then
the biological treatment may fail altogether. Many of the
problems of high concentration can be circumvented by
appropriate dilution, but this is a feature that should be
anticipated and incorporated into the design.
Rational implementation of a chemically enhanced pump-
and-treat remediation will require many of the same
characterization and testing methods required for
conventional pump-and-treat operations. Physical
hydrogeological parameters such as hydraulic conductivity,
the potentiometric surface, and porosity (Table 3) can be
obtained using the methods outlined by Mercer and Spalding
(1992a, b), Palmer and Johnson (1989), Rehm et al. (1985),
and Ford et al. (1984). These physical parameters can be
used in modelling studies to ascertain the feasibility of getting
the reactive agent to the contaminated areas within a
reasonable period of time while maintaining a capture zone
for the contaminant and the reactive agent. The results of
such studies should help identify the optimum injection
concentrations, the number of wells, and their location.
If chemical enhancements are to be considered, greater
effort must be placed on the chemical characterization of the
site. In particular, the key chemical processes that limit
pump-and-treat remediation must be identified if the proper
type of reactive agent is to be chosen, Important chemical
processes and their characterization have been recently
addressed by Boulding and Barcelona (1992a,b,c), Palmer
and Johnson (1992), and Palmer and Fish (1992).
Table 3. Physical-hydrogeological and chemical parameters that
should be identified during site characterization for chemical
enhancements to pump-and-treat remediation
Physical-Hydrogeologlcal Parameters
bulk density
porosity
hydraulic conductivity
storativity
potentiometric surface
site boundary conditions
ground water-surface water interactions
infiltration rates
leakage from adjacent aquitards
Chemical Parameters
pH
redox conditions
contaminant concentrations and spatial distribution
non-contaminant concentrations
oxidation capacity of the aquifer
reduction capacity of the aquifer
organic contaminant partition coefficients
ionic adsorption parameters
Several approaches must be used to determine the chemical
processes limiting pump-and-treat remediation. For ionic
solutes, adsorption tests are important for quantifying the
fraction of the solute adsorbed onto the surfaces of the soil
as a function of pH and the aqueous concentration of the
contaminant. The potential for mineral controls can be
identified by calculating mineral saturation indices using
geochemical models such as MINTEQ (Felmy et al., 1984)
and may be verified through x-ray diffraction or electron
17
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microscopy. Oxidation and reduction tests (e.g., Barcelona
and Helm, 1991 ) are useful for determining the amounts of
oxidant and reductant necessary to alter the redox state of a
contaminant in the subsurface. Bench-scale tests to
measure the increase in solute concentrations following the
addition of proposed reactive agents can provide information
about potential compositions of interfering solutes entering
the treatment train. Treatment studies using water
compositions based on these tests can be used to determine
potential problems and test proposed solutions to the
treatment process.
For neutral organic contaminants, batch-sorption tests can be
conducted to determine the fraction of the contaminant
partitioned into the soil organic matter. However, at most
sites, the partitioning can be determined from published
values of the Kocof the contaminant (e.g. Maybey et al.,
1982; Montgomery and Welkom, 1989) and the fraction of
organic carbon in the soil. This approach is simpler than
batch experiments; however, it does require that the f of
the soil be measured. When nonaqueous phase liquid are
present, they are the limiting factor in site remediation; pools
of NAPLs are more problematic than NAPLs retained at
residual saturation in the soil. Again, proposed reactive
agents should be tested at the bench scale and the
treatability of the extracted water tested before pilot testing
and implementation of the chemical enhancement operation.
The tests discussed above are generalizations of a few that
can be conducted. The specific tests required at a site
depend on the target contaminants and the nature of the soil
materials from which they must be extracted. Utilizing the
knowledge from laboratory studies and the experience from
other hazardous waste sites will be important in directing the
type of tests that need to be conducted. Unfortunately, at
this time, there have been few field demonstrations of
chemical enhancement methods from which to obtain such
experience.
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