Sandia National Laboratories is developing guidelines that outline
the technical basis for relying on natural attenuation for the
remediation of metals and radionuclide-contaminated soils and
groundwaters at DOE sites for those specific cases where natural
processes are effective at ameliorating soil and groundwater toxicity.
Remediation by natural attenuation (RNA) requires a clear identification
of the specific reaction(s) by which contaminant levels are made
less available as well as considerable long-term monitoring (LTM).
Central to RNA is the development of a conceptual model describing
the biogeochemical behavior of contaminant(s) in the subsurface.
The conceptual model will be used to make testable predictions
of contaminant availability over time. In many cases, comparison
between this prediction and field measurements will provide the
ëbright lineí test of whether RNA is to be implemented.
As a result, development of the conceptual model should guide
site characterization activities as well as long-term monitoring.
Sandia is in the process of developing a protocol and guidance
document to help DOE site managers and their EPA regulators determine
if remediation by natural attenuation is appropriate for their
specific ER projects and, if so, the additional characterization
and analysis required to ensure that the approach is successful.
Natural attenuation is defined as "naturally occurring processes
in the environment that act without human intervention to reduce
the mass, toxicity, mobility, volume, or concentration of contaminants
in those mediaî. These in situ processes include ìbiodegradation,
dispersion, dilution, sorption, volatilization, and/or chemical
and biochemical stabilization of contaminantsî. Natural
attenuation has been extensively documented and is increasingly
relied on for the cleanup of soils and groundwaters contaminated
with fuel hydrocarbons, PAHs, and even chlorinated solvents [1,
2]. Often natural attenuation leads to a net decrease in remediation
costs while providing substantial reductions in contaminant levels,
risks to human health and the environment.
Sites contaminated with metals and radionuclides pose special
problems for application of the natural attenuation approach.
Whereas natural attenuation of organic contaminants means breakdown
by microorganisms, natural attenuation of metals often means sequestering
or transformation by the soil matrix or dilution. Radionuclides,
in turn, might be considered naturally attenuated if their interactions
with soils result in transport times to possible receptors much
greater than their radioactive half life. Although laboratory
and field evidence for the transformation and sequestering of
inorganics is abundant, natural attenuation of metals and radionuclides
has received less emphasis in the regulatory universe relative
to the natural attenuation of organic contaminants. This is changing
with the recent issuance of EPA guidance on monitored natural
attenuation [3].
The increasing recognition of natural attenuation by EPA does
not constitute a change in cleanup goals; nor is it a walk-a-way,
or default solution. The burden of proof remains on the proponent,
not the regulator. EPA expects implementation to require extensive
site characterization, long-term monitoring, risk assessment,
and contingency measures. The net result is that implementation
is site-specific. Typically, natural attenuation has been applied
in combination with more active remedial approaches, at sites
where cleanup levels were not extensively exceeded in the first
place, or after more proactive remediation efforts had been halted.
Generally, contingency plans requiring active remediation were
in place.
Site characterization is typically demonstrated, in decreasing order of importance, through:
If the first criterion is satisfied, further effort is made to
examine the other criteria. On the other hand, in the absence
of historical evidence for reductions in contaminant levels, the
argument for natural attenuation probably cannot be made solely
on the latter two. In the end, the regulators make the decision
whether natural attenuation is applicable.
The future use of the site must be taken into account if remediation
by natural attenuation is considered. RNA often takes longer
to achieve cleanup goals than more active remediation measures.
Land use concerns may consequently bias cleanup towards proactive,
as opposed to passive remediation. Isolated sites with great
distances from potential receptors are, therefore, more likely
to be candidates for RNA compared to sites connected by short
travel times to potential receptors. Some advantages and disadvantages
of RNA are outlined in Table I.
Insert Table I here.
In summary, a number of milestones must be achieved to build support for RNA at a particular site. The source term must be controlled - either treated or removed, to limit subsequent contaminant fluxing. The plume and down-gradient areas must be monitored to establish plume dynamics. If contaminant levels are seen to decrease over time, a conceptual model to account for the decrease should be established, and if possible, refined to provide a basis for making defensible predictions of the long-term evolution of contaminant levels. These milestones are not necessarily easy or cheap to achieve, and in all cases the appropriate regulatory agency should be involved at the earliest stages.
Metals and radionuclides can be removed from soil solutions and
groundwaters by (1) sorption to mineral surfaces and/or soil organic
matter (SOM); (2) formation of insoluble solids; (3) uptake by
plants and organisms; and occasionally (4) through volatilization
(e.g. methylation of mercury). Focusing on the formation of adsorbed
species ('surface complexes'), insoluble solids and uptake by
plants, we note that metal/radionuclide speciation depends primarily
on the ambient biogeochemical conditions of the soil or groundwater:
pH, redox state (electron availability), alkalinity, and the
presence of chelating (e.g. EDTA, natural organic acids) or solid-forming
(e.g. phosphate) ligands are critically important.
At the same time, the sequestering of metals/radionuclides out
of the aqueous phase often makes their engineered extraction problematic.
Corrosive soil leaches, vitrification techniques, and grout curtains
are examples of the extremes which must be gone to in order to
liberate or isolate metals in soils. In many cases the technical
impracticability of metal/radionuclide extraction is a direct
result of the natural attenuation processes. Nevertheless, contaminant
immobilization cannot be assumed - some metals/radionuclides (e.g.,
chromate and pertechnetate) have very little interaction with
the matrix, and can move rapidly through soils and groundwaters.
It is, therefore, necessary to explore sorption, plant uptake,
and solubility in substantial detail. Ideally, this will provide
some basis for identifying the conditions where RNA might be plausible
and where it clearly won't be.
Sorption is particularly effective at limiting the concentrations
of metals/radionuclides that are present in trace quantities.
At high contaminant levels the actual amount of contaminant in
solution is typically determined by the presence of contaminant-containing
insoluble minerals. There are obvious exceptions. For example,
Cs+ and TcO4- form no insoluble
solids.
Sorption can be characterized as being 'reversible' or 'irreversible'.
Contaminants sorbed reversibly to a surface can be desorbed in
response to a decrease in contaminant level in solution. In other
words, the sorbed species remains in contact with the solution
and responds to changes in solution composition. Irreversibly
sorbed species typically do not reequilibrate rapidly with solutions
once sorbed. Irreversible sorption may occur through a combination
of occlusion (overcoating), diffusion into dead-end pores, or
structural collapse of the mineral around the sorbed species.
Because equilibrium desorption cannot always be assumed it is
important to split sorption into forward and backward reactions
(respectively, adsorption and desorption) and treat them separately.
Adsorption is very rapid and typically occurs over time spans
less than a second, but sometimes longer. Adsorption from solution
varies with pH, the type of mineral surface, the amount of surface
coverage, the concentration of the trace element, and the composition
of the soil solution. Ligands which form strong complexes with
the contaminant may either decrease the total amount of sorption,
or form ternary complexes with the surface. At high pH, negative
surface charge is maximal; at low pH positive surface charge is
greatest. As a result cation sorption increases with pH; anion
sorption with decreasing pH.
While adsorption has received the most attention, in many cases
desorption may be the more important control over metal/radionuclide
release at contaminated sites. Routinely the most contaminated
sections of a site are removed and/or stabilized leaving a plume
of contamination behind wherein the contaminants are primarily
sorbed to mineral surfaces. Almost all performance assessment
calculations assume that desorption is reversible. Hence, when
fresh recharge comes into contact with sorbed contaminants the
latter are predicted to instantaneously equilibrate, in effect
setting contaminant levels in solution. In reality desorption
rates are often relatively slow, sometimes vanishingly so. The
actual desorption rate will in many cases determine the net export
of metal/radionuclide toxicity.
Synthetic organic contaminants co-mingled with metals/radionuclides
often give rise to the same observation. NTA, EDTA, and DTPA
are all synthetic organics, which show up at DOE sites. Citrate,
and oxalate are two natural chelating agents of concern as well.
Degradation rates of these chelates typically follows the trend:
citrate ~ oxalate >> NTA > EDTA > DTPA. Estimating
transport of chelated metals and radionuclides requires that the
coupled processes of metal chelation, sorption, and chelate breakdown
be understood.
Table II outlines likely natural attenuation pathways for most
of the radionuclides and metals of concern. Also shown are the
potential caveats which must be kept in mind for each contaminant.
Specifically, we have sought to identify what soil chemical parameters
control the natural attenuation pathway, and, by extension, what
changes in soil chemistry would work against the given natural
attenuation pathways.
Insert Table II here.
Table III outlines the minimal geochemical data needed to determine
if the particular natural attenuation pathways are operative.
Data needs depend primarily on whether the likely fate of the
compound is as a component of an insoluble solid, a sorbed contaminant,
or, possibly, a species occluded on an iron hydroxide or carbonate
mineral surface, or irreversibly sorbed to an interlayer clay
site.
Insert Table III here.
There is currently no protocol for implementing natural attenuation of metals or radionuclides. Typically technical protocols for implementing natural attenuation of organic contaminants follow a format along the following lines.
Natural attenuation of organic contaminants is generally demonstrated
using a wealth of evidence argument pointing to reductions in
contaminant mass. The four most effective components used to
convince a regulatory agency are: evidence of contaminant loss
in the field, variations in electron donor/acceptor levels, appearance
of degradation byproducts, and soil microcosm studies done in
the lab. However, the same approach probably cannot be used for
inorganics. The appearance of byproducts, or variation in acceptor/donor
levels, probably cannot be used to monitor irreversible sorption
or the growth of contaminant-bearing insoluble minerals. When
a contaminant, such as lead, sorbs it will displace some other
cation such as Ca2+, which is likely to be far more
abundant in solution. When Cs+ sorbs to a clay, chances
are that it will be present in only trace amounts, and far less
abundant in solution than the Na+ or K+
it displaces. As a result, while irreversible sorption of trace
contaminants will dramatically affect solution levels of the latter,
changes in other background metal concentrations will more than
likely be minimal. The growth of contaminant-containing hydroxides,
carbonates, and sulfides may also cause undetectable variations
in hydroxide, carbonate, and sulfide levels in solution because
the latter are typically present in initially greater concentrations
than the with which metals they combine.
Standard geochemical codes [e.g., 5, 6] can be used to calculate whether contaminant levels are limited by the formation of an insoluble phase (e.g., Ba2+ by BaSO4 growth). Geochemical modeling to support uptake by sorption is not far enough along to be a stand-alone demonstration of metal sorption. Instead, uptake by sorption can be demonstrated by: 1. Demonstrating that the sorbing phase is present in soils through a solubility calculation or direct observation; and 2. Showing that an appreciable fraction of the compound is associated with that phase. The latter is most directly done through sequential soil leaching procedures which dissolve specific minerals, along with any sorbed material. For example, citrate-Dithionate solutions remove iron hydroxides from soils. Hydroflouric acid removes silicates. H2O2 removes organic matter. Acid acetate buffer solutions remove calcium carbonate.
Natural attenuation, or any other remediation strategy, can only
be assessed with regard to clearly-defined standards. It is important
to consider what objectives can and cannot be attained by natural
attenuation as well as the time scale over which various objectives
may be attained. Environmental quality standards for the subsurface
are defined for both the immobile phase (e.g., soil) and for groundwater.
Sorption processes, although they retard the migration of the
contaminant toward potential receptors, necessarily involve association
of the contaminant with the immobile phase. Soil quality criteria
are commonly defined in terms of the total metal concentration
in the soil. Since metals are naturally-occurring substances,
contamination can only be defined relative to some background
level such as average crustal abundance. If total metal concentration
in the soil is taken as the operative standard, then natural attenuation
can only be applied if some zone of contamination is excluded
from this standard for an extended period or even in perpetuity.
Over the very long term, flushing of contaminated subsurface
material with uncontaminated groundwater may decrease the total
metal concentration in the soil to background levels. It may,
however, be reasonable to define alternative standards for soil
quality that correspond to the bioavailability of soil
metals. Although the determination of the bioavailable fraction
is a complicated problem, it is appropriate to address this question
in the context of the applicability of natural attenuation. Note
that different standards may need to be applied if surficially
contaminated soils are subject to erosion or scouring by wind.
There are a number of technical obstacles, which might potentially
limit the effectiveness of natural processes in controlling contaminant
movement and availability in the subsurface, and consequently,
regulatory acceptance of its implementation. To begin with, unlike
the biodegradation of some organic contaminants (e.g., fuel hydrocarbons),
which results in the contaminant of concern ëgoing awayí,
typically metals and long-lived radionuclides will remain in the
subsurface. (If radionuclides have sufficiently short half-lives
they may ëgo awayí as well). In other words, many
metals and radionuclides may still be present, though unavailable
for biologic uptake. At the same time, dilution may lower contaminant
levels to the point where they are acceptable in a regulatory
sense, though there has been no net reduction in contaminant mass.
The transport of contaminants that exist as components of insoluble
solids or sorbed (reversibly or irreversibly) to mineral surfaces
may, because of the ambient geochemistry, be severely limited.
Contaminants which are strongly sorbed or in solid form in soils
are likely to see much larger volumes of fresh recharge. Consequently,
the potential for dilution is heightened. The immobility of sorbed
and/or solid phase contaminants makes them plausible candidates
for RNA. Critical to such an assessment is a clear understanding
of the sequestering mechanism. Specifically, the speciation of
the contaminant needs to be known for three reasons: 1. to be
able to predict the long-term stability of the sequestering in
the face of possible changes in the ambient geochemistry; 2. to
provide some clues as to how much time must elapse before the
acceptable contaminant availability is achieved, and; 3. to allow
an estimate to be made of the total attenuation capacity of a
given soil/groundwater for the specific contaminant.
The potential for remobilization is a critical obstacle for acceptance
of the remediation of metals and radionuclides. Obviously, time-spans
are important. If remobilization of 90Sr or 137Cs
(half-lives ~ 30 years) occurs over time spans much greater than
a hundred years, a very significant fraction of the radiotoxicity
will have decayed away. For long-lived radionuclides and metals,
dilution may be the only process decreasing potential releases
which might occur with remobilization. It is not hard to imagine
scenarios leading to the remobilization of most, if not all of
the contaminants of concern. Drastic changes in hydrologic conditions
and/or subsurface water chemistry may adversely affect natural
attenuation processes. For example, a natural attenuation remedy
that relies on limited infiltration may be invalidated by irrigation
for agricultural development. Cesium ëirreversiblyí
bound to interlayer clay sites in a soil could be very rapidly
released if ammonium-rich fertilizer were subsequently applied
for agricultural purposes. By the same token, lead sorbed to
iron hydroxides in an initially aerated soil might be released
if the soil became flooded, then anoxic, followed by destabilization
and dissolution of the iron hydroxide host. On the other hand,
the composition ranges of soil and groundwater is typically very
limited, primarily because there are a host of biogeochemical
processes which tend to control the pH, redox state, alkalinity,
etc. of natural waters. Although drastic changes in the compositions
of natural waters are more the exception than the rule, it will
probably be impossible for site-owners to demonstrate that remobilization
will never occur. This is a critical obstacle to the implementation
of natural attenuation for metals and radionuclides.
The respective roles of site characterization and monitoring are
two important considerations for use of RNA. The argument can
be made that site characterization should specifically provide
the means to develop a conceptual model of natural attenuation
and, to the extent possible, calibrate that model so that contaminant
availability can confidently be predicted in the future. Unless
RNA is exceedingly fast (which is often not the case) it will
be difficult to calibrate a kinetic model for RNA given the time
allowed for a site characterization. Ultimately, long-term measurement
might be required. Nevertheless, this should not be confused
with long-term monitoring. Long-term monitoring should, quite
simply, provide the means for assessing whether or not RNA is
working. If the conceptual model for RNA is sufficiently effective
at reproducing measured trends in contaminant levels, it should
allow the frequency of monitoring to be significantly reduced.
An important consideration in evaluating the applicability of
natural attenuation for a given site is its intended land use.
Natural attenuation may be considered as part of the remediation
strategy for a contaminated site or as a component of the permitting
of an existing facility. The latter case necessarily involves
some on-going release of contaminants into the environment and
the relevant question is whether natural attenuation would afford
sufficient protection to human health and the environment. In
the former case, source control is probably (but not necessarily)
a prerequisite to application of natural attenuation.
The efficacy of natural attenuation will depend on numerous factors including the type and extent of primary and secondary contamination (where primary contamination is associated with the original source and secondary contamination with dispersal of contaminants from the source), the hydrologic regime and hydrogeology, subsurface geology, and potential receptors. For a given site, these factors must be evaluated with regard to their likely effects on the sorption and dilution processes by which natural attenuation of metals and radionuclides may be accomplished.
Although natural attenuation encompasses several natural processes,
it is important to recognize that only a few of these processes
are operative for metals and radionuclides. For both metals
and radionuclides, the operative processes are dilution and sorption.
Dilution may occur by dispersion of dissolved contaminants in
groundwater and/or by dilution of dissolved contaminants into
surface water (e.g., upon interception of surface water by contaminated
groundwater). Sorption may be defined generally to include the
processes of adsorption, coprecipitation, precipitation, and diffusion
into the matrix, processes by which solutes become associated
with the immobile, solid phase. Sorption may either be reversible
or slowly reversible. Slowly reversible sorption processes may
be considered as effectively "irreversible" if the time
scale for re-release of the contaminant from the solid phase is
long relative to some time scale of interest or observation.
Slowly-reversible sorption of contaminants from solid phases exposed
to uncontaminated groundwater may also contribute to dilution
of the contaminant. The extent of dilution will be determined
by the rate of contaminant release into solution relative to the
velocity of groundwater flow.
For radionuclides only, radioactive decay is an additional
process contributing to natural attenuation. In some cases, however,
the ingrowth of daughter nuclides may result in an increasing
hazard over time that counterbalances or even outweighs the benefit
due to loss of the parent nuclide.
We see the building of conceptual models for RNA as one of the primary challenges to its successful implementation. The most important sinks for metals and radionuclides in soil and groundwater are fairly well understood (microbiological effects less so). Nevertheless, field-based techniques for demonstrating that contaminants are being taken up into otherwise inaccessible and/or non-bioavailable fractions of the soil matrix are few and far between, and therefore a critical need. SEM, isotope exchange techniques, and soil digestions may provide a means for addressing this need.
1. Wiedemeier T. H., Wilson J. T., Kampbell D. H., Miller R. N., and Hansen J. E. (1995) Technical protocol for implementing intrinsic remediation with long-term monitoring for natural attenuation of fuel contaminant dissolved in groundwater. Air Force Center for Technical Excellence, Technology Transfer Division 1 & 2.
2. Wiedemeier T. H., Swanson M. A., Moutoux D. E., Wilson J. T., Kampbell D. H., Hansen J. E., and Haas P. (1996) Overview of the technical protocol for natural attenuation of chlorinated aliphatic hydrocarbons in ground water under development for the U.S. Air Force Center for Environmental Excellence. EPA Symposium on Natural Attenuation of Chlorinated Solvents.
3. US Environmental Protection Agency (EPA) (1997) Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action, and Underground Storage Tank Sites. USEPA Office of Solid Waste and Emergency Response Directive 9200.4-17.
4. Brady, P. V, Brady, M. V. And Borns, D. J. (1997) Natural Attenuation: CERCLA, RBCAs, and the Future of Environmental Remediation. Lewis Publishers, Boca Raton, Florida.
5. Bethke C. M. (1994) The Geochemist's Workbench™. A Users Guide to Rxn, Act2, Tact, React, and Gtplot. University of Illinois.
6. Wolery T. J. (1983) A computer program for geochemical aqueous
speciation-solubility calculations: user's guide and documentation.
Lawrence Livermore National Laboratory Report UCRL-53414.
Table I. Advantages and disadvantages of RNA
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Table II. Natural attenuation pathways for metals (and other inorganics)[4]
| Chemical | ||
| Pb2+ | Sorption to iron hydroxides, organic matter, carbonate minerals, formation of insoluble sulfides. | Low pH destabilizes carbonates, iron hydroxides. Comingled organic acids and chelates (e.g. EDTA) may decrease sorption. Low EH dissolves iron hydroxides, but favors sulfide formation. |
| CrO42- | Reduction by organic matter, sorption to iron hydroxides, formation of BaCrO4 | Low pH destabilizes carbonates, iron hydroxides. Low EH dissolves iron hydroxides. Are reductants available? |
| As(III or V) | sorption to iron hydroxides, formation of sulfides | Low pH destabilizes carbonates, iron hydroxides. Low EH dissolves iron hydroxides. |
| Zn2+ | sorption to iron hydroxides, carbonate minerals, formation of sulfides | Low pH destabilizes carbonates, iron hydroxides. Comingled organic acids and chelates may decrease sorption. Low EH dissolves iron hydroxides. |
| Cd2+ | sorption to iron hydroxides, carbonate minerals, formation of insoluble sulfides. | Low pH destabilizes carbonates, iron hydroxides. Comingled organic acids and chelates may decrease sorption. Low EH dissolves iron hydroxides, but favors formation of sulfides. |
| Ba2+ | sorption to iron hydroxides, formation of insoluble sulfate minerals | Low pH destabilizes carbonates, iron hydroxides. Low EH dissolves iron hydroxides. What are sulfate levels? |
| Ni2+ | sorption to iron hydroxides, carbonate minerals | Low pH destabilizes carbonates, iron hydroxides. Comingled organic acids and chelates may decrease sorption. Low EH dissolves iron hydroxides, but favors sulfide formation. |
| Hg2+ | formation of insoluble sulfides | Is methylated by organisms |
| NO3- | reduction by biologic processes | |
| Radionuclides | ||
| UO2+2 | sorption to iron hydroxides, precipitation of insoluble minerals, reduction to insoluble valence states | Low pH destabilizes carbonates, iron hydroxides. Comingled organic acids and chelates may decrease sorption. High pH and/or carbonate levels decrease sorption. Low EH dissolves iron hydroxides. |
| Pu(V and VI) | sorption to iron hydroxides, formation of insoluble hydroxides | May move as a colloid. Low EH dissolves iron hydroxides. |
| Sr2+ | sorption to carbonate minerals, formation of insoluble sulfates | Low pH destabilizes carbonates. |
| Am3+ | sorption to carbonate minerals | Low pH destabilizes carbonates. High pH increases solubility of Am-carbonate minerals. |
| Cs+ | sorption to clay interlayers | High NH4+ levels may lessen sorption. How abundant are clays? |
| I- | sorption to sulfides, organic matter | Sorbs to very little else. |
| TcO4- | possible reductive sorption to reduced minerals (e.g. magnetite), forms insoluble reduced oxides and sulfides. | Sorbs to very little else. |
| Th4+ | sorption to most minerals, formation of insoluble hydroxide | may move as a colloid |
| Co2+ | sorption to iron hydroxides, carbonate minerals | low pH destabilizes carbonates. Low EH dissolves iron hydroxides |
Table III. Data Needs for Natural Attenuation of Metals [4]
| Chemical | |
| Pb2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. EH, and if EH is low, sulfide levels. Organic carbon content. |
| CrO42- | EH, electron donor levels, pH (reduction rates are faster at low pH). |
| As(III or V) | EH and, if EH is low, sulfide levels. |
| Zn2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. EH, and if EH is low, sulfide levels. |
| Cd2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. EH, and if EH is low, sulfide levels. |
| Ba2+ | Sulfate levels. |
| Ni2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. EH, and if EH is low, sulfide levels. |
| Hg2+ | EH, and if EH is low, sulfide levels. |
| UO2+2 | Iron hydroxide availability, pH, availability of reducing compound |
| Pu(V and VI) | Iron hydroxide availability, pH, availability of reducing compound |
| Sr2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. |
| Am3+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. |
| Cs+ | Clay content, cation exchange capacity. |
| I- | Metal sulfide mineral content |
| TcO4- | EH, and if EH is low, sulfide levels. |
| Co2+ | Iron hydroxide availability; pH, alkalinity, and Ca2+ levels to answer if calcium carbonate is stable. |