<&ERfli Ground Water Issue
United States
Environmental Protection
Agency
Metal Attenuation Processes at Mining Sites
Richard T.Wilkin
Background
The EPA Regional Ground Water Forum is a group of
EPA scientists representing Regional Superfund and
Resource Conservation and Recovery Act Offices
(RCRA). The Forum is focused on exchanging informa-
tion related to ground-water characterization, monitor-
ing, and remediation. The application of monitored
natural attenuation (MNA) for inorganic contaminants
in ground water is a topic of concern to the Forum.
The purpose of this Issue Paper is to provide scientists
and engineers responsible for assessing remediation
technologies with background information on MNA
processes at mining-impacted sites. Some of the key
issues concerning the application of natural attenua-
tion for inorganic contaminants are discussed, such
as the geochemical mechanisms responsible for at-
tenuation, attenuation capacity, monitoring parameters,
and evaluating whether attenuated metal and metalloid
contaminants will remain immobile.
Introduction
Acid mine drainage (AMD) is a major source of water con-
tamination in metal-mining and coal-mining districts world-
wide. The causes of AMD are well known. They relate to
the natural weathering of mine wastes and rocks enriched
in metal sulfide minerals. Environmental impacts includethe
destruction of aquatic life and habitats and contamination
of drinking water resources. The most common reactions
that lead to the production of AMD involve the chemical
and biological oxidation of metal sulfides contained in mine
waste heaps, active or abandoned mine workings, or in tail-
ings piles left over from the processing of sulfide ores. The
iron sulfides: pyrite, marcasite and pyrrhotite, are perhaps
the most common sources of AMD production, because
they are ubiquitous in metal sulfide ores and because they
generally are not the target of ore beneficiation processes.
Numerous variables factor into the assessment of potential
For further information contact Richard T, Wilkin (580) 436-8874
[wilkin.rick@epa.gov] at the Ground Water and Ecosystems
Restoration Division of the National Risk Management Research
Laboratory, Office of Research and Development, U.S.
Environmental Protection Agency, 919 Kerr Research Drive, Ada,
Oklahoma 74820.
AMD releases, including the quantity of reactive sulfides,
grain size distribution and grain morphology, bacterial
activity, moisture content, and the availability of dissolved
oxygen or other oxidants (e.g., Jamboretal., 2000; Lowson,
1982; Nordstrom and Southam, 1997; Rigby et al., 2006;
Williamson and Rimstidt, 1994).
Acid mine drainage may form via the interaction of surface
water or ground water with materials enriched in metal
sulfides, such as tailings piles or the underground workings
of deep mines (Figure 1). Production of AMD may occur
during mine operations and may continue for many years
after minesare closed and tailings dams aredecommissioned
from operation; consequently, evaluation of AMD is often a
long-term proposition which usually adds up to high costs
forsite characterization, monitoring and cleanup. Estimates
of the number of sites in the United States affected by AMD
vary widelyfrom200,000toover 550,000 (U.S. Environmental
Protection Agency, 2004). Estimated costs to cleanup
contamination at AMD sites are equally difficult to assess.
One hundred and fifty-six hardrock mining sites were on
or had the potential to be on the National Priorities List
(NPL) for cleanup underthe Comprehensive Environmental
Response, Compensation and Liability Act (CERCLA), with
potential cleanup costs of up to $24 billion dollars (U.S.
Environmental Protection Agency, 2004). Mine sites are
frequently remotely located, which further adds to the costs
of site characterization, remediation, and monitoring.
In some cases, especially where ore host rock is capable
of reacting with acidic drainage, metal concentrations may
attenuate over time and space. A primary control on the
process of metal attenuation at mining-impacted sites is acid
neutralization (Al et al., 2000; Berger et al., 2000). Neutral-
izing capacity in sulfide ore tailings is predominantly from
carbonate minerals (calciteand dolomite) because most non-
carbonate minerals associated with metalliferous deposits
are extremely slow to react and affect pH (e.g., Jambor et
al., 2000). As pH increases, aqueous metal species tend
to precipitate as hydroxide, oxyhydroxide, or hydroxysulfate
minerals (Nordstrom, 1982). In addition, as pH increases
dissolved metals may adsorb onto surfaces of these newly
formed minerals and/orothersurfaces present in the environ-
ment, such as organic matter due to decreasing competition
with protons, decreased surface potential, and increased
hydrolysis of metal ions at circum-neutral pH.
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Figure 1. Impacts of mine workings and mine wastes on ground water and surface water. As oxygenated waters interact
with sulfide minerals, pH decreases and metals are mobilized in the absence of acid-neutralizing rocks and
overburden.
In the context of mining sites, natural attenuation refers
to the observed reduction of contaminant concentrations
and/or contaminant mass flow rates as contaminants are
transported downgradient from their source. At many
mining sites that require cleanup of contaminated ground
water, monitored natural attenuation is not expected to be
relied upon as a sole remedy. The global magnitude of the
acid drainage problem is clear evidence that in most cases
natural processes are incapable of ameliorating the acid-
ity and metal contamination produced by oxidizing sulfide
minerals. However, monitored natural attenuation may be
an effective strategy to augment more active approaches
of remediation (summarized below). In addition, natural at-
tenuation processes often tend to spread contaminants out
in space away from source zones via various mineral-water
reactions. Therefore, it is important to recognize natural
attenuation processes from the perspective of tracking
contaminant transport and fate in the environment. This
Issue Paper provides remedial project managers and other
state or private remediation managers and their technical
support personnel with background information on the vari-
ous physical, chemical, and biological processes of natural
attenuation that may occur at mining sites. This background
information is necessary for preparing sampling plans to
support site characterization, remedy selection, and post-
remedial monitoring efforts.
Treatment Strategies for AMD
Strategies for dealing with AMD include source control
and chemical or biological treatment of contaminated
ground water. Preventing the formation or the migration
of AMD from source zones is generally a favorable option,
but is often difficult to accomplish effectively due to the
large aerial extent of tailings areas and the large volumes
of materials involved. Consequently, cleanup efforts are
usually focused on directly treating impacted ground water
and surface water (U.S. Environmental Protection Agency,
2006). Remediation methods at mining sites can be gener-
ally divided into "active" and "passive" approaches (Johnson
and Hallberg, 2005). Perhaps the most straightforward and
common active approach to treat acidic effluents is through
addition of alkaline materials to raise pH, increase the rate
of ferrous iron oxidation, and cause the removal of metals
and metalloids present in solution via mineral precipitation
or surface adsorption processes. A number of alkaline ma-
terials have been used for active treatment, including lime,
calcium carbonate, sodium carbonate, sodium hydroxide,
and magnesium oxide. Use of these materials can lead
to effective controls on the release of acidic drainage and
dissolved metals. The cost of maintaining direct treatment
facilities is often high. Large volumes of a low-density
sludge result from the reaction between alkaline compounds
and acidic effluents. Moreover, the sludge itself becomes
an environmental concern, both in terms of disposal and
the potential release of contaminants through subsequent
leaching (Jambor et al., 2000).
Figure 2 presents a general classification of passive AMD
treatment systems, which can be broadly grouped into
chemical and biological systems (Neculita et al., 2007).
Treatment systems that rely largely on abiotic chemical pro-
cesses include open limestone channels, anoxic limestone
drains, and successive alkalinity-producing systems. In
open limestone channels, acidic water flows over crushed
limestone or some other alkaline agent. The goals of such
applications are to generate alkalinity, neutralize pH, and
remove soluble aluminum, iron, and manganese via mineral
precipitation (Mukhopadhyay et al., 2007; Ziemkiewicz et
al., 1997). The treatment strategy with anoxic limestone
drains is to neutralize acid-mine drainage using limestone
while maintaining iron in a reduced state (as ferrous iron)
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AMD Passive Treatment
Systems
Chemical
Biological
Successive
Alkalinity
Producing
Systems
Figure 2. Treatment systems for acid mine drainage.
to avoid iron oxidation and the consequent precipitation of
hydrous ferric oxides on the limestone surfaces. Such surface
precipitation or armoring by iron and aluminum precipitates
greatly reduces the effectiveness of the neutralizing material.
In its simplest form, an anoxic limestone drain (ALD) is a
buried, limestone-filled trench that intercepts AMD before it
is exposed to atmospheric O2 (Cravotta and Trahan, 1999;
Hedinetal., 1994). ALD designs are enclosed to minimize
gas flux in contrast to systems such as limestone channels
that are open to the atmosphere. In this way CO2 gas is
retained in the subsurface channel which leads to enhanced
calcium carbonate dissolution and alkalinity production. After
acidic waters pass through the ALD, effluents are exposed
to the atmosphere and hydrous ferric oxides are produced
by the oxidation of ferrous iron to ferric iron. In successive
alkalinity-producing systems, both limestone and organic
matter are used in vertical flow systems to provide alkalinity
production, sulfate reduction, and metal removal (Keplar
and McCleary, 1994).
Biological passive treatment systems for AMD include
bioreactors and constructed wetlands (Figure 2). Sulfate-
reducing passive bioreactors have received recent attention
as promising technologies for AMD treatment (e.g., Alvarez
et al., 2007; Annachatre and Suktrakoolvait, 2001; Costa
and Duarte, 2005; Drury, 1999; Dvorak et al., 1992; John-
son and Hallberg, 2005; Neculita et al., 2007; Steed et al.,
2000). The advantages of this technology are high metal
removal capacity, stable sludge, and low operation costs.
The chemical basis for treatment of AMD by sulfate reduc-
ing bacteria involves microbially-mediated sulfate reduction
coupled to organic matter oxidation. Sulfide precipitation
is the desired mechanism of contaminant removal, but
other mechanisms including adsorption and precipitation of
metal hydroxides occur in passive bioreactors (Neculita et
al., 2007). The efficiency of sulfate-reducing bioreactors is
primarily controlled by the organic carbon source (Coester
et al., 2006; Gibert et al., 2004; Prasad et al., 1999; Zamzow
et al., 2006). Solid-phase testing suggests that organic
substrates with high protein contents or low lignin contents
(e.g., manure) are better capable of supporting bacterial
activity and sustaining contaminant removal (Coester et
al.,2006; Gibert etal., 2004).
Artificial wetlands and biological treatment systems have
been used since the mid-1980s for the treatment of AMD.
The effectiveness of these applications has been variable
and generally difficult to predict (Barton and Karathanasis,
1999; Wieder, 1989; Wildeman and Updegraff, 1997), al-
though recent process-based modeling efforts are beginning
to provide insight into the functioning of these systems that
feedback into system design refinements and operational
improvements (Whitehead et al., 2005). Both aerobic and
anaerobic processes contribute to contaminant removal in
constructed wetland systems. The oxidation of ferrous iron
to ferric iron and the subsequent precipitation of hydrous
ferric oxides is a dominant process that is effective in re-
moving iron and other metals from AMD (Brenner, 2001).
Metal accumulation in these systems also occurs through
precipitation of metal sulfides via the activity of sulfate
reducing bacteria that consume natural organic matter
and sulfate and produce reactive sulfide for metal removal
(Webb etal., 1998).
Permeable reactive barriers (PRBs) and monitored natural
attenuation (MNA) fortreatment of ground water impacted by
AMD can be classified either as chemical or biological pas-
sive treatment systems based on processfunction (Figure 2).
The application of PRBs involves the excavation of a trench
or pit in the flow path of contaminated ground water. The
excavated volume is then filled with reactive materials that
are permeable to allow flow of contaminated ground water
and reaction to remove dissolved contaminants via chemi-
cal or biological processes. Reactive materials that have
been shown to be effective in increasing pH and removing
metals include mixtures of organic carbon, limestone, and
zero-valent iron (Benner et al., 1999; Gibert et al., 2003;
Ludwig et al., 2002; Shokes and Moller, 1999; Waybrant et
al., 1998;Wilkinand McNeil, 2003). Organic carbon-based
PRB systems take advantage of anaerobic microbiological
processes within the PRB to generate alkalinity and remove
metals as sulfides. In zero-valent iron PRB systems, a
variety of abiotic and biotic metal uptake processes are
important in neutralizing acidity and removing metals from
solution. Similarly, MNA involves both biologically medi-
ated processes and abiotic geochemical processes that
are presented in the following sections. Note that many of
the documented natural attenuation processes are actually
strategically enhanced in designed remediation systems,
such as bioreactors.
Background on MNA for Inorganic Contaminants
The term "monitored natural attenuation" refers to the long-
term examination of natural processes with the objective
that such processes will reach site-specific remedial objec-
tives. MNA can be applied in conjunction with other cleanup
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approaches, such as source removal, source control, or
plume control (Rugner et al., 2006). To be considered as
an acceptable option, MNA would be expected to achieve
site remedial objectives within a time frame that is reason-
able compared to that possible by using other more active
remediation technologies as described above. Natural at-
tenuation processes include a variety of physical, chemical,
and biological processes that can act to reduce the mass,
mobility, volume, or concentration of contaminants in ground
water. Attenuation processes important at mining sites
include pH buffering and acid neutralization, adsorption at
the mineral-water interface, mineral precipitation, and dilu-
tion/dispersion (Figure 3).
pH buffering/
acid neutralization
Adsorption
at the mineral-water
interface
Dilution/dispersion
Mineral Precipitation
Figure 3. Processes of natural attenuation at mine-impacted
sites.
EPA's Office of Research and Development is preparing
a technical resource document for the application of MNA
to inorganic contaminants in ground water (see Reisinger
et al., 2005; U.S. Environmental Protection Agency, 2007).
The technical resource document presents a four-tiered
assessment of MNA as a viable remediation option for
selected metal, metalloid, and radionuclide contaminants
encountered in ground water. Components of the tiered
approach include 1) demonstrating contaminant sequestra-
tion mechanisms, 2) estimating attenuation rates and the 3)
attenuation capacity of aquifer solids, and 4) evaluating po-
tential reversibility issues. The technical resource document
is intended to provide a tiered decision-making approach
for determining whether MNA is likely to be an effective
remedial approach for inorganic contaminants in ground
water at a particular location. EPA expects that users of this
document will include EPA and State cleanup programs and
theircontractors, especially those individuals responsible for
evaluating alternative cleanup methods for a given site or
facility. A decision-making approach is provided for evaluat-
ing MNA as a possible cleanup method for contaminated
ground water. Emphasis is placed on developing a more
complete understanding of the site through development of
a conceptual site model that includes an understanding of
the attenuation mechanisms, the geochemical conditions
governing these mechanisms, and indicators that can be
used to monitor attenuation progress (U.S. Environmental
Protection Agency, 2007).
A tiered decision-making approach is an appropriate and
cost-effective way to screen out sites unsuitable for MNA
while collecting the most relevant data at sites that might be
amenable to this technology. Conceptually a tiered assess-
ment of MNA seeks to progressively reduce site uncertainty
as MNA-specific data is collected. MNA for inorganics and
radionuclides is most effectively implemented through four
tiers that require progressively greater information on which
to assess the reasonableness of MNA:
Tier I. Theplumeisnotthreateningpublichealth.isstable,
and some direct evidence of contaminant attenua-
tion exists.
Tier II. The attenuation capacity of the site exceeds the
estimated mass of contaminant at the site.
Tier III. There is strong evidence that attenuation
mechanism(s) will prevail over long periods of
time.
Tier IV. A record of decision including a long-term monitor-
ing plan and other site closure considerations is
developed.
MNA Processes at Mining Sites
Based upon the tiered approach presented above, assess-
ments of MNA at mining sites must demonstrate that chemi-
cal, physical, orbiological processes areoccurringto mitigate
migration of contaminants, that the capacity of the MNA
process exceeds the mass of contaminants in the source,
and that the attenuation processes are sustainable over long
periods of time. As noted previously, these conditions are
frequently not met at mining sites. Although research findings
clearly show that attenuation of contaminants does indeed
occurat mining sites (Table 1), the documented mechanisms
of attenuation are either rate- or capacity-limited so that
contaminants are only partially attenuated or attenuation
occurs over longer flow paths than are acceptable from a
site cleanup perspective. It is equally clear, however, that
many of the attenuation processes important at mining
sites are long-lived, so that a sound understanding of the
factors that control transport and fate of metals in ground
water and across the ground water/surface water interface
can benefit site cleanup efforts.
Acid Neutralization
A primary control on the process of metal attenuation at
mining sites is acid neutralization. Many factors affect
the acid neutralization capacity of a system, including the
type, abundance and reactivity of metal-bearing sulfides in
the ore and waste rock, permeability of the mine workings
or mine tailings, and the ability of the host or surrounding
rocks to consume acidity. Methods are available to predict
whether or not materials will be acid-generating (e.g., U.S.
Environmental Protection Agency, 1994). These methods
provide a numerical accounting with respect to prediction
of acid production and neutralization potential. In general,
materials containing elevated concentrations of carbonate
minerals orthat have elevated inorganic carbon to total sulfur
ratios are the most effective in neutralizing acidity.
Mixing of mine effluents with ambient ground water and
surface water dilutes the dissolved contaminants and can
result in pH increases. In surface waters, dilution and neu-
tralization can occur over spans of meters to many kilome-
ters. Dilution and neutralization are often tied to seasonal
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variations in flow patterns and volumes. Field studies have
documented more effective attenuation of mine effluents
in dry seasons compared to wet seasons (Webster et al.,
1994). Inaddition,field studies and laboratory columntesting
resu Its indicate that mineral assemblages present in tailings
piles, underlying aquifers, and receiving surface waters play
a pivotal role in controlling pH (Blowes and Ptacek, 1994;
Jurjovec et al., 2002; Morin et al., 1988; Sanchez-Espana
et al., 2005). Mineral phases important in buffering pH are
calcite/siderite, aluminum hydroxides, iron hydroxides, and
aluminosilicates.
Role of
The oxidation of iron sulfides in mine wastes results in the
release of iron, sulfate, acidity, and metals to solution. High
aluminum and silica concentrations are also commonly en-
countered in mine effluents and are the result of weathering
of aluminosilicate minerals at low pH. Oxidation and hydro-
lysis reactions can subsequently lead to the precipitation
of a wide array of hydroxide, sulfate, and/or hydroxysulfate
minerals depending on geochemical and biogeochemical
conditions (Nordstrom and Alpers, 1999). These secondary
minerals play important roles in attenuating contaminants
from mine effluents (e.g., Accornero et al., 2005; Casiot et
al., 2005; Fukushi et al., 2003; Gault et al., 2005; Jamieson
et al., 1999; Lee et al., 2002; Lee et al., 2005; Levy et al.,
1997;McCartyetal., 1998; McGregor et al., 1998; Moncur
et al., 2005; Munketal., 2002; Sanchez-Espana etal., 2005;
Sidenko and Sherriff, 2005; Webster et al., 1998; Zanker et
al., 2002). Some of the common secondary minerals found
in association with the weathering of mine wastes, their typi-
cal pH range of formation, and documented contaminant
associations are listed in Table 1.
Table 1. Secondary Minerals Formed from Acid Mine Waters and Contaminant Associations
Mineral
Hydroxides
Goethite
HFO, hydrous ferric
oxide
Gibbsite
Formula
FeOOH
~Fe5HO8-4H2O
AI(OH)3
Typical pH
of
formation
2-4
>5
>5-6
Examples of contaminant
at mine-impacted
settings
Sorption/coprecipitation of Pb (up
to 21 wt%), As (up to 7.7 wt%), Zn
(up to 4.6 wt%), and Cu (up to 2.5
wt%)
Sorption/coprecipitation of As-rich
ferrihydrite; As/Fe=0.02-0.1 , with
10-30% As(lll)
Sorption in the general order of
Pb>Cu>Zn>Ni with increasing pH
Reference
Lee et al. (2005)
Casiot et al.
(2005)
Munk et al.
(2002)
Hydroxysulfates
Alunite
Jarosite
Schwertmannite
KAI3(OH)6(S04)2
KFe3(OH)6(S04)2
Fe808(OH)6(S04)
4-6
2-5
2-4
Precipitation of Al
Coprecipitation with As(V) replac-
ing sulfate in the Jarosite structure
Coprecipitation of Cu, Ni and Zn
Accornero et al.
(2005)
Gault et al.
(2005)
Sidenko and
Sherriff (2005)
Sulfates
Gypsum
Anglesite
Melanterite
CaSO4«2H2O
PbS04
FeSO4«7H2O
>3
>3-4
<2
"Hardpan" precipitate
Precipitation at pH ~3; nanopar-
ticles
Coprecipitation with Zn and Cu;
temporary metal removal in a
highly soluble phase
Moncur et al.
(2005)
Zanker et al.
(2002)
Jamieson et al.
(1999)
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Secondary precipitates can remove contaminants from
impacted waters through adsorption and/or coprecipitation
reactions. Adsorption processes are typically categorized
by the relative "strength" of interaction between the adsor-
bate (species in solution) and the surface or adsorbent.
If solvating water molecules are positioned between the
cation or anion and the surface, the adsorption complex is
referred to as outer sphere and is considered to be weak.
Conversely, if upon adsorption the adsorbate loses waters of
hydration such that there are no water molecules positioned
between the cation or anion and the surface, the adsorption
complex is referred to as inner sphere and is considered
to be strong. The extent to which dissolved contaminants
will sorb to secondary precipitates as outer sphere or inner
sphere complexes will vary as a function of the contaminant
species, the secondary precipitate, pH, particle size and
surface area, and presence of other sorbing species that
may compete for adsorption sites.
Inorganic contaminants may be removed from solution
due to precipitation of an insoluble phase in which the
contaminant represents a major or minor component within
the solid. Examples of secondary precipitates that form in
mine-impacted sites include oxyhydroxides [e.g. ,FeOOH(s)],
hydroxysulfates [e.g., FeeO8(OH)6(SO4)(s)], sulfates [e.g.,
PbSO4(s)], and sulfides [e.g., ZnS(s)]. For each of these
minerals there will be a limited compositional range of
ground-waterchemistry over which precipitation could occur
and formation of these precipitates may compete with other
removal processes such as adsorption (Table 1).
Characterization of secondary precipitates is carried out
by using a variety of tools. Mineralogical identification is
typically made by using powder x-ray diffraction techniques
(XRD). The characterization of particle morphology and
semi-quantitative composition are accomplished using scan-
ning electron microscopy (SEM) and x-ray energy-dispersive
spectrometry (EDX). Analysis of element partitioning to well-
crystalline and poorly-crystalline components of the solid
phase is typically accomplished using selective chemical
extraction procedures. Used in combination, these meth-
ods allow for the identification of attenuation mechanisms
involving secondary minerals (see U.S. Environmental
Protection Agency, 2007). Knowledge about the types of
mineral phases helps to understand the long-term stability
of attenuated metals.
One can also evaluate whether there is a potential for con-
taminant precipitation by evaluating the saturation state
of the ground water with respect to possible precipitate
phases using a saturation state modeling approach. In
order to evaluate whether a ground water is oversaturated,
undersaturated, or at equilibrium with a particular phase,
computer geochemical speciation models are of practical
use. As an example, consider the solubility expression for
lead sulfate (anglesite):
PbSO4(s) = Pb2+ + SO42-
The mass-action expression that applies to the equilibrium
is:
A natural water may or may not be at saturation with respect
to anglesite, depending on whether the phase is indeed
present, available surface area, residence time of water, and
kinetic factors that may impede dissolution and/or precipita-
tion. If we assume equilibrium between water and anglesite,
then the ion activity product, Q, should be the same as the
equilibrium constant, Kf, i.e.,
Q = a
soj-
= K = 10-
where the activity (a) of PbSO4(s) is taken to be 1. Because
ion activity products may vary by orders of magnitude, it is
often more convenient to take the logarithm of the ratio, that
is, to compute the saturation index, SI:
S/= log_ri = 0 at equilibrium
If a water is oversaturated in a particular phase, then the SI
is positive, and there is a thermodynamic driving force for
precipitation to occur. If the water is undersaturated, then
the SI is negative, and the mineral, if present, will tend to
dissolve:
and
S/>0 if oversaturated
SI< 0 if undersaturated.
K=
As previously indicated the stability of a precipitate will
be dictated by the ground-water chemistry. Contaminant
remobilization will occur as a result of dissolution of the pre-
cipitate phase, for example, when log Q/Kt< 0. Precipitate
dissolution may occur due to ground water acidification,
oxidation/reduction of precipitate components, dilution,
or complexation of the precipitate component(s) with dis-
solved species that form more stable compounds. Thus,
it must be recognized that attenuation processes involving
inorganic contaminants are reversible (e.g., Casiot et al.,
2005; Gault et al., 2005; Moncur et al., 2005). Metals taken
up at the mineral-water interface can be released back into
solution. Geochemical modeling of mineral stability and
contaminant adsorption/desorption behavior can provide
insight into contaminant remobilization potential due tofuture
changes in geochemical conditions. However, it must be
noted that thermodynamic databases are often incomplete
and thermodynamic constants for specific compounds
may vary from database to database. Thus, results from
geochemical models must be carefully reviewed. In addi-
tion, the method outlined above ignores rates of mineral
dissolution and precipitation. Again data are often lacking
on the kinetics of biogeochemical processes responsible
for contaminant uptake and remobilization, especially data
that can be applied in field systems to predict the long-term
behavior of contaminants.
of
Microbial processes can play a role in both mobilizing and
attenuating inorganic contaminants at mining sites. For
example, Macur et al. (2001) showed that microbial reduc-
tion of arsenate [As(V)] to arsenite [As(lll)] occurred over
relativelyshorttime scales and result edinenhanced arsenic
mobilization in mine tailings pore water. In addition, iron-
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reducing bacteria may cause contaminant dissociation from
aquifer solids as a consequence of iron oxide dissolution.
Metals and metalloid species associated with secondary
iron-bearing precipitates may be released via the activity
of bacteria under certain conditions (Herbel and Fendorf,
2006; Langer and Inskeep, 2000).
Sulfate-reducing bacteria (SRB), however, have the ability
to reverse the reactions causing acid mine drainage, by at-
tenuating the movement of metals through the precipitation
of sulfide minerals (e.g., Gammons et al., 2005), and by
raising the pH of the water (Tuttleetal., 1969). This process
is recognized in ex situ treatment of acid mine drainage
as previously noted and also in the natural environment
(Church et al., 2007; Kimura et al., 2006; Koschorreck et
al., 2003; Labrenz et al., 2000; Paktunc and Dave, 2002).
The overall sulfate-reduction process can be described by
the reaction:
2CH2O
SO2- + 2H+
H2S
CO
H2O
where CH2O represents organic matter, either in the solid or
aqueous phase. The resulting dissolved hydrogen sulfide
can precipitate with divalent metals in AMD, for example
(M = Cd, Cu, Fe, Ni, Pb, orZn):
H2S + M2+(aq) = MS(s) + 2.W
The mass concentration of reactants involved in sulfate
reduction is usually much larger than the mass concentra-
tion of metals involved in secondary precipitation reactions,
hence these combined reactions can lead to an increase
in alkalinity and the pH of the water, while simultaneously
attenuating divalent metals. Alkalinity produced during the
sulfate reduction process can also drive the precipitation of
carbonate minerals, such as calcite and siderite (Paktunc
and Dave, 2002), and can help neutralize acidity in the
receiving water body.
Most sulfate-reducing bacteria have been considered to
be inactive at low pH (Johnson, 2003). More recent stud-
ies of acid mine drainage systems (both engineered and
natural) have noted that there is some potential for low-pH
sulfidogenesis. For example, in laboratory studies, sulfate
reduction has been shown to occur in solutions as low as
pH 3 in bioreactors using ethanol, methanol, or glycerol
(alone or in various combinations) as an organic substrate
(Kolmert and Johnson, 2001). In addition, in-situ remedia-
tion by sulfate reduction has been shown to occur in acidic
pit lakes and sediments after the pH was raised to 5-6 by
amendment with organic carbon plus lime (Wendt-Potthoff et
al., 2002). In natural AMD systems the reduction of sulfate
to sulfide has been reported at pH values as low as 2-3
(Koschorreck et al., 2003), but there are few reports of the
isolation and/or characterization of acidophillic SRB from
these environments. In a recent study, sediments recovered
from the flooded mine workings of the Penn Mine, a Cu-Zn
mine abandoned since the early 1960s, were cultured for
anaerobic bacteria overarangeofpHfrom4.0to 7.5 (Church
et al., 2007). Phospholipid fatty acid (PLFA) analyses of
Penn Mine sediment showed a high biomass level with a
moderately diverse microbial community structure composed
primarily of iron- and sulfate-reducing bacteria. Cultures of
sediment from the mine produced dissolved sulfide at pH
values near 7 and near 4, forming precipitates of either
iron sulfide or elemental sulfur. Phylogenetic sequences of
Penn Mine sediment and laboratory cultures were closely
aligned to the sulfate-reducing organisms Desu/fosporos/m/s
and Desulfitobacterium (Church et al., 2007). At this site,
sulfate-reducing bacteria play a role in attenuating metals at
moderately low pH. Precipitates of zinc sulfide were identi-
fied in the reducing mine sediments. In the absence of the
bacterial activity, zinc and other metals could be transported
into nearby surface waters.
Characterization of microbiological impacts on natural at-
tenuation processes involves additional tools that can be
used during site characterization efforts. Largely within the
last decade, genetic analyses have been used to identify
microbial communities in environmental samples. Many
of these molecular biological methods rely on 16S rDNA
sequences, such as denaturing gradient gel electrophoresis
(DGGE). DGGE can be used for simultaneous analysis
of multiple samples obtained at various time intervals to
detect microbial community changes, which is an advanta-
geous feature in studying microbial ecology and MNA (U.S.
Environmental Protection Agency, 2007). Examples of the
use of molecular techniques in relation to examinations of
microbiological influences of contaminant behavioral mine
sites are presented in Macur et al. (2001), Druschel et al.
(2004), and Church et al. (2007).
Monitoring
In order to evaluate whether or not natural attenuation pro-
cesses can play a role in achieving site remediation goals,
detailed site investigations are required. Generally, the
necessary investment in site characterization for evaluat-
ing the applicability of natural attenuation is at least or is
more expensive and time consuming than for other site
remediation technologies. On the other hand, where MNA
is applicable, long-term monitoring costs may be less than
for other more active remedial technologies.
The evaluation of natural attenuation in a ground watersystem
involves studies todetermine the location, concentration, and
movement of contaminants in the subsurface. Thus natural
attenuation assessments typically focus on developing site
hydrologic and conceptual models that can be simulated
with a computer geochemical model (see U.S. Environmental
Protection Agency, 2007). Evaluation of natural attenuation
usually involves not only the determination of processes of
attenuation that are currently occurring, but also projects
the sustainability of these processes into the future. Table 2
lists attenuation reactions of selected contaminants and
appropriate parameters that could be examined during
site investigations. Table 3 lists examples of solid-phase
analyses that would likely support MNA assessments. The
use of MNA as part of a site remedial plan will necessarily
require that a long-term monitoring plan be established.
Such plans will be developed to enable decisions regarding
whether or not site remedial objectives are being met, and
to verify that site conditions are not changing in such a way
as to impact the major natural attenuation processes for
contaminants of concern. Long-term monitoring plans should
be developed with well defined triggers that would initiate
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2, Attenuation Reactions and Capacity Parameters for Selected Contaminants
Contaminant
Possible Attenuation Reactions
Relevant parameters
As
Sorption in aerobic environments
Sorption/Precipitation in anaerobic
environments
« Abundance/stability of hosts; typically Fe and Al
(hydr)oxides
• Solid-phase sulfide accumulation; redox buffer capacity,
sulfate reducing activity
Cd
Sorption in aerobic environments
Sorption/Coprecipitation carbonates
Sorption/Precipitation in anaerobic
environments
Abundance/stability of hosts; typically Fe and Al
(hydr)oxides
« Abundance/stability of hosts; may require consideration
of pH buffer capacity
« Solid-phase sulfide accumulation; redox buffer capacity,
sulfate reducing activity
Cu
Sorption in aerobic environments
Sorption/Coprecipitation carbonates
Sorption/Precipitation in anaerobic
environments
• Abundance/stability of hosts; typically Fe and Al
(hydr)oxides
• Abundance/stability of hosts; may require consideration
of pH buffer capacity
« Solid-phase sulfide accumulation; redox buffer capacity,
sulfate reducing activity
Pb
Sorption/Coprecipitation in aerobic
environments
Precipitation as hydroxycarbonate or
sulfate
Sorption/Precipitation in anaerobic
environments
« Abundance/stability of hosts; typically Fe and Al
(hydr)oxides
• Aquifer pH buffer capacity
• Solid-phase sulfide accumulation; redox buffer capacity,
sulfate reducing activity
U
Reductive Precipitation
Sorption
Abundance/reactivity of electron donors
Abundance of hosts; typically metal (hydr)oxides
HI
Sorption in aerobic environments
Sorption
Sorption/Precipitation in anaerobic
environments
• Abundance/stability of hosts; typically Fe and Al
(hydr)oxides
• Abundance of hosts (clays)
« Solid-phase sulfide accumulation; redox buffer capacity,
sulfate reducing activity
Note: Measurement objectives and methodologies to support MNA investigations are documented in U.S. Environmental Protection Agency
(2007). MNA assessments focus on both the aqueous phase and the solid phase in order to identify attenuation pathways.
the implementation of contingency remedial technologies if
natural attenuation processes fail to fulfill expectations.
Conclusions
At most mining sites that require cleanup of contaminated
ground water, MNA is not expected to be a sole remedy.
The magnitude of the acid drainage problem is clear evi-
dence that in most cases natural processes are incapable of
ameliorating the acidity and metal contamination produced
by oxidizing sulfide minerals. Nevertheless, at nearly all
mining sites, natural processes are contributing to varying
degrees and in some cases may contribute significantly
to site remedial goals. Biogeochemical processes can be
particularly important for natural attenuation of some metal
and metalloid contaminants, under specific environmental
conditions.
Cleanup of mining sites, in particular Megasites, is currently
being viewed as a long-term process. This is partly due
to the enormous size, the complexity of contaminants and
sources, and the large volumes of materials encountered at
many mining sites. It is also recognized that the long-term
outcomes of site cleanup programs are extremely difficult
to predict (Gustavson et al., 2007). Effective management
of these sites over long periods of time requires complex
organization of site characterization, technology selec-
tion and utilization, and long-term monitoring. Given that
cleanup expectations at many mining sites are long-term,
it may be appropriate to include an examination of natural
attenuation processes and the role that such processes
play in removing, repartitioning, or otherwise affecting the
fate of contaminants in the environment.
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3. Examples of Solid-phase Analyses to Support MNA Assessments
Method
Powder x-ray diffraction
Microbeam analysis
Wet chemical extractions
Bulk elemental analysis
Batch sorption/column testing
Oxidation/reduction capacity
Biological assays of 16S rDNA
sequences
Most Probable Number (MPN) counts
Objectives
Identification of mineral forms
Analysis of micro-scale distribution and association of contaminants
Evaluation of contaminant associations in the solid-phase
Evaluation of total concentrations of contaminants and other
minor elements
major and
Evaluation of contaminant uptake capacity
Evaluation of redox conditions/buffering capacity
Molecular characterization of microbial populations
Bacterial enumeration
Notice
The U.S. Environmental Protection Agency through its
Office of Research and Development funded and managed
the research described here. It has been subjected to the
Agency's peer and administrative review and has been
approved for publication as an EPA document.
Quality
All research projects making conclusions or recommenda-
tions based on environmentally-related measurements
and funded by the Environmental Protection Agency are
required to participate in the Agency Quality Assurance
(QA) program. This project did not involve the collection or
use of environmental data and, as such, did not require a
Quality Assurance Project Plan.
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