Ally Micuy

DAVIDB. VANCE ARCADIS G&M, Inc. Midland, Texas

JAMESA. JACOBS

Environmental Bio-Systems, Inc.

Mill Valley, California

Based on the origin and susceptibility to microbial interaction, hydrocarbons may be divided into two broad

PROCESS LIMITATIONS OFIN SITUBIOREMEDIATION OF GROUNDWATER 43 classes: petroleum hydrocarbons and complex chlorinated

hydrocarbons.

Petroleum hydrocarbons are largely associated with the production, storage, or use of fuels, lubricants, and chemical feedstocks. Petroleum hydrocarbons have been demonstrated to be biodegradable by numerous species of bacteria. Over a period of 3.5 billion years, bacteria have been able to evolve genetic resources that allow some of them to potentially use petroleum hydrocarbons as a source of food.

Complex industrial hydrocarbons include chlorinated aliphatic and aromatic hydrocarbons, MTBE, pesticides and herbicides, and polymers. These new synthetic compounds have been manufactured for about 100 years. Bacteria have not had time to evolve the genetic information required to utilize them as a source of food. Due to recalcitrance to microbial attack, these complex industrial chemicals are termed xenobiotic.

Successful in situ bioremediation of groundwater has been demonstrated at sites impacted with petroleum hydrocarbons. The most successful method of application has been through stimulation of indigenous microbial species. Microbial stimulation is the process of ensuring that environmental conditions, nutrient availability, and requirements for an electron acceptor are adequate in the contaminated portions of the aquifer.

The most common cause for failure of saturated zone in situ bioremediation is the lack of adequate mass transport of the electron acceptor (usually oxygen). In this regard, the physical setting of the site is critical. Overall permeability and the scale and degree of heterogeneity are the factors governing the advective and diffusional transport rates of contaminants and remediation reagents in the subsurface. If mass transport rates are too low, saturated zone in situ bioremediation is not a viable option.

Given adequate mass transport properties, site-specific microbiological conditions can also impact the process. Unfortunately, the presence of indigenous microbes and efficient mass transport may still prove insufficient for effective bioremediation. Specific reasons for the poor performance of in situ bioremediation systems relate to unoptimized subsurface conditions.

There is uncertainty with regard to the effect of hydro- carbon availability on the effectiveness of biodegradation. Can bacteria degrade hydrocarbons adsorbed to surfaces or degrade hydrocarbons with low levels of solubility? Or must the hydrocarbon be solubilized before it can be biodegraded? Contradictory laboratory evidence and field evidence have been published for both scenarios. With the predominance of evidence indicating that solubilization must take place, degradation reactions with extracellular bacterial exudates are much less likely.

The answer is likely consortia specific and dependent on the ability of the bacteria to synthesize appropriate bio- surfactants. This ability may be absent in some instances. Although petroleum hydrocarbons are amenable to primarily aerobic biodegradation, for it to occur the indigenous bacteria must have the appropriate genetic information. This genetic information is specific and precise. The presence of a specific hydrocarbon will

stimulate the synthesis of an oxygenase enzyme that is expressly configured to react with that stimulating hydrocarbon.

For remediation, indigenous microbes generally possess the genetic information required for appropriate enzyme production and the contaminant will stimulate the pro- duction of those enzymes. General microbial stimulation has the potential to produce a large amount of biomass that may not take part in the biodegradation process and actually cause harm through biofouling and plugging of injection wells, galleries, or surrounding formations. There is potential to lose critical subsurface mass transport capa- bilities.

BACTERIAL TRANSPORT

There are practical limits to the degree of cleanup obtainable using bioremediation. Hydrocarbons at the low parts per million (ppm) level may not be capable of supporting significant levels of microbial activity even under stimulation. Sites with relatively high levels of hydrocarbon impact may actually be better candidates for bioremediation than those lightly impacted at levels slightly above regulatory action levels. Toxicity related to the presence of heavy metals, such as chromium, arsenic, or lead, or low temperature of the groundwater have been observed to inhibit bacterial growth in a variety of settings. Stimulating electron acceptors must also be available at sufficient concentrations; native sulfate concentrations less than about 20 mg/L do not stimulate sulfate reducing bacteria even in the presence of usable carbon substrates. Xenobiotic industrial compounds are often recalci- trant to direct aerobic microbial attack. However, over the last 20 years a biodegradation process termed co- oxidation (or cometabolism) has been successfully demon- strated by researchers. For example, the aerobic degra- dation of trichloroethylene (TCE) has been accomplished using monooxygenase and dioxygenase enzymes produced through the use of petroleum hydrocarbons as a metab- olizable substrate (food source) and stimulus for enzyme production. This general process is termed co-oxidation and the hydrocarbon substrate used as a food source is the cometabolite. Many different hydrocarbon substrates have been observed to stimulate the generation of cooxidation enzymes. The currently known cometabolic substrates fall into two broad classes:

1. Analog substrates, which are hydrocarbons that have a geometry similar to the targeted xenobi- otic compound.

2. Methanotrophic (which is different than methano- genic) microbial systems have proved particularly effective at generating xenobiotic active enzymes. Enzymes with co-oxidizing potential have a strong natural affinity for the hydrocarbon that originally stimulated its generation. The enzyme is genetically tailored to the compound used as a food source. Over 300 mol of methane are required to biodegrade 1 mol of TCE via co-oxidation. The efficiency of the co-oxidation process is extremely poor. Under field conditions where

44 PROCESS LIMITATIONS OFIN SITUBIOREMEDIATION OF GROUNDWATER mass transport is a critical success factor, a 300-fold

decrease in the effectiveness of the reactants in the contaminated zone often can be impractical.

Accurate assessment of potential limiting factors such aerobic terminal electron acceptor (dissolved oxygen), geo- chemical conditions (pH, temperature, conductivity), and macronutrients (orthophosphate and ammonia as nitro- gen) should be documented as part of the bioremediation evaluation process. Bioremediation is a dynamic process requiring monitoring of the hydrological, geochemical, and biological conditions over the life of a project.

NATURAL ATTENUATION: GROUNDWATER REMEDIATION BY NO ACTION

In a time of reappraisal for the allocation of financial resources to environmental action, a question of ever increasing importance is the consequence of no action concerning the release of petroleum or chlorinated hydrocarbons into groundwater. An important portion of that answer comes from the application of site-specific health based risk assessments. However, in instances where human consumption or exposure is not an issue, no action may be a reasonable alternative, even at elevated dissolved contaminant concentrations. The issue then becomes the determination of the consequence of no action under conditions where the sole process for remediation is natural attenuation.

The physical, chemical, geological, and biological processes that take place in a contaminated aquifer are complex. In most instances, a ‘‘native’’ aquifer is in a long-standing state of chemical equilibrium between the groundwater and the geologic matrix through which it flows. The release of anthropogenic hydrocarbons into an aquifer upsets that equilibrium. The dissolved concentration of the contaminant as it migrates through the aquifer is controlled by adsorption, dispersion, volatilization, and degradation. Adsorption affects the overall residence time of the release and dispersion affects the downgradient shape and dissolved concentration in the plume. Only volatilization and degradation contribute to the removal of contaminant from the aquifer, and at low concentrations degradation is the dominant mechanism for attenuation.

The mechanisms for attenuation through degradation can be broadly divided into two categories, biological and abiotic chemical action. This discussion is predicated on relatively ‘‘normal’’ groundwater conditions under which biological action proceeds at a rate orders of magnitude greater than abiotic processes. Extremes of pH, redox conditions, ionic strength, or temperature may make an exception to that generalization. Transformation can be chemically complex, dependent on the environmental conditions described above and affected by aquifer heterogeneity related to granular or fracture variability.

The factor controlling the rate of aerobic degradation is the availability of oxygen and the rate at which it can be introduced into the groundwater (through the groundwater–table interface) or the rate at which oxygen-rich groundwater can pass through the zones of adsorbed contamination. Each pound of petroleum

hydrocarbons requires about 3.08 pounds of oxygen for complete degradation (1).

Typical in situ aerobic decay rates for groundwater are in the range of 35µg/L·d (equivalent to about 0.5 oz/d per cubic yard of aquifer matrix).

Natural attenuation occurs both in the source zone and in the dissolved phase plume. In the source zone, oxygen will be rapidly consumed and portions of the aquifer will then host anaerobic degradation. Anaerobic degradation is limited by the availability of appropriate anaerobic electron acceptors such as nitrate, sulfate, or iron. When their availability is limited, degradation will stop after the production of aliphatic and aromatic organic acids; similarly, at low levels of dissolved oxygen (DO), aerobic degradation may also stop with the production of organic acids. The intrinsic biodegradation process and the alternative terminal electron acceptors are shown in Fig. 1 (2).

Optimum aerobic biodegradation occurs with the dis- solved oxygen above 2 mg/L. Below that, the aerobic degradation rate of aromatic hydrocarbons will decrease dramatically. Conversely, under complete anaerobic condi- tions, nitrate reducing, iron reducing, and suflate reducing bacteria can effectively degrade hydrocarbons. However, at DO concentrations as low as 0.1–0.4 mg/L, anaerobic degradation rates will be reduced to just a few percent of optimum.

Because of all the mechanisms described above, hydrocarbon plumes tend to achieve a stable shape and size even when there is a continuous source of free phase hydrocarbon release. Steady state is achieved when the area of the plume edge is great enough to provide for a natural degradation rate equivalent to the rate of hydrocarbon infiltration. The edges of the dissolved plume do not have enough DO to support optimum rates of aerobic degradation but have too much DO to allow for optimum anaerobic degradation. The interior of a plume will support anaerobic natural attenuation, which is typically limited by the availability of iron in the mineral matrix and sulfate in the native groundwater and to a lesser extent the mineral matrix as well. However, once the source of hydrocarbon has been removed, a dissolved plume will narrow and dissipate from the edges inward, due to the availability of DO from groundwater along those edges.

The selection of a no action natural attenuation option should be based on an appropriate analysis of data gathered during the assessment of the site. First- order decay rates are appropriate for the evaluation of degradation kinetics at low concentrations, less than 1 ppm (an appropriate level to assume at the periphery of a plume). Given first-order decay rates, the analysis has a focus that is twofold—the effect of attenuation over time and the effect over distance.

Attenuation over time is measured at the edges of a plume using concentration measurements gathered repeatedly from specific monitor wells. The minimum recommended time is one year, with quarterly sampling from the selected monitor wells. The data for each well is then semilog plotted as log concentration against time. The slope of the line is the first-order decay constant in percent per day.

PROCESS LIMITATIONS OFIN SITUBIOREMEDIATION OF GROUNDWATER 45

Concentration

Metabolic products

Dominant electron acceptors N₂ Fe(II) CO2 CH4 Organics H2S Aerobic respiration Denitrification Iron (III) reduction Sulfate reduction Methanogenesis Fe(III) O2 NO3− SO42− CO2 pE (mV) +100 −100 0

Intrinsic biodegradation processes

Plume migration/ ground water flow

Source

Figure 1. Intrinsic biodegration process (2). Petroleum hydrocarbons are referred to as organics in the diagram.

Attenuation with distance more accurately incorporates the effects of aquifer heterogeneity. Data for this analysis is obtained from a minimum of three monitoring wells, preferably along the long axis of the plume. This data is semilog plotted as log concentration versus distance. The slope of this line is equal to the decay constant divided by the groundwater velocity.

With this data, decisions can be made based on the site- specific contaminant dynamics under no action natural attenuation. This, in conjunction with a health based risk assessment, can allow for sound decision-making by the business and regulatory community.

In summary, the adoption of a no action alternative is most applicable to the dissolved phase plume only. Except for volumetrically small releases, it will still be necessary to remove or remediate the source zone of an impacted aquifer, after which natural attenuation may be a reasonable approach to the residual dissolved phases. Also implicit in this approach is that no action does not preclude the performance of requisite assessment activity, which can represent a significant long-term liability in some cases. Nonetheless, after proper source abatement, assessment, and analysis, the reliance on natural attenuation mechanisms for the final stages of cleanup is a cost effective and, if properly managed, environmentally sound resolution to aerially extensive dissolved phase hydrocarbon contamination.

NATURAL ATTENUATION: THE EFFECT OF PUMP AND TREAT REMEDIATION

In situ groundwater remediation has matured over the past 25 years, particularly with regard to understanding the dynamics of the interactions between contaminants, the impacted saturated soil matrix, and microbiological activity. Recent interest in the phenomenon of natural

attenuation has served to illustrate the variety of microbial ecosystems that are present in a contaminant plume, each system determined by redox conditions and availability of electron acceptors. While natural attenuation is an attractive alternative to those responsible for groundwater contamination, the regulating communities are more skeptical. The need for proactive groundwater pumping remediation and the efficacy of natural attenuation pose a potentially complex balance that is governed by the subsurface conditions of each individual site. There is no universal applicable rule for the resolution of that balance. It is the responsibility of remediation designers to make those site-specific determinations and to provide the regulating community with information sufficient to support the proactive and the natural attenuation portions of each individual cleanup. Our purpose here is to point out some of the most significant factors impacting that balance.

The first and most dominant control is the nature of the saturated soil matrix. Several factors must be evaluated for remedial design:

1. The degree and scale of sediment heterogeneity, which determines how much and what portion of a contaminated aquifer can be affected with advective groundwater flow. Low permeability regions must rely on diffusional transport, which will dominate the overall remediation rate in the treatment zone. 2. The time of exposure to the contaminant is a

direct function of the impact of heterogeneity described above. The contaminant will diffuse into the nonadvective portions of the aquifer soil matrix. At a minimum, remediation will take as long as the initial exposure. Due to the adsorptive retardation reactions, remediation is likely to take longer than the exposure time.

46 PROCESS LIMITATIONS OFIN SITUBIOREMEDIATION OF GROUNDWATER 3. The geochemical composition of the soil matrix is also

a factor in remedial design. Carbonaceous material and clays have a much higher propensity for the adsorption of organic contaminants. Iron oxides, in turn, have high adsorptive capacity for metal contaminants. Iron and sulfur minerals may be sources of electron acceptors as redox conditions are modified through the interaction of indigenous microbial populations and the contaminant.

4. The background geochemical makeup of the ground- water as well as that in the contaminant plume is an important remedial design factor. Dissolved oxygen, sulfate, nitrate, and iron can all potentially serve as alternate terminal electron acceptors to aid in the degradation of organic contaminants. Ferrous iron, hydrogen sulfide, and carbon dioxide are indicative end products of those reactions.

5. The distribution of the organic contaminant is another factor affecting remedial design. Free phase hydrocarbons should be recovered proactively with an extractive technology (pump and treat) or in the case of CVOCs possibly an in situ chemical oxidation injection technology. In cases where impact is shallow, excavation and disposal is still an extremely viable option.

The treatment of dissolved and adsorbed hydrocarbons is the point at which the balance between proactive remediation and natural attenuation must be determined. One of the most important contributions that a pump and treat system makes to the in situ remediation of contaminated groundwater is plume capture and hydraulic control in the source zone and the core of the dissolved and adsorbed plumes. Background groundwater that is drawn through the plume perpendicular to the natural groundwater flow direction must also be evaluated.

In the past, the focus of pump and treat remediation has been on how it acts to flush and remove the contaminant. The contributions made by recent developments on the mechanisms of natural attenuation reside in the role of electron acceptors present at background concentrations within the aquifer. From the exterior to the interior of a plume, the specific electron acceptor zones are aerobic, denitrification, sulfate/iron reducing, and methanogenic.

The boundary between each specific redox zone and the active electron acceptor is controlled by the kinetics of the degradation process in each zone and the advective transport rates of groundwater through that zone. In most instances, the dominant effect is the groundwater transport rate.

The natural concentrations of these electron acceptors cover a wide range. Natural oxygen levels commonly range from 2 to 8 mg/L. Groundwater sulfate concentrations in soils derived from sedimentary rocks are typically in the 25-mg/L range, with higher values of several hundred mg/L not uncommon. Ferric oxides are commonly present in soils in the range of 0.5–5%; the ability of indigenous iron reducing bacteria to access that iron will vary from location to location. Given adequate permeability and the presence of appropriate electron acceptors, natural

enhancement of pump and treat systems is possible and worth the relatively inexpensive analyses (some of which can be done with field kits) required to evaluate.

NATURAL ATTENUATION: TRANSVERSE DISPERSION AS THE NATURAL DRIVING FORCE

Dispersion is the process by which the interface of contaminated groundwater with native groundwater does not remain abrupt. The leading edge of a contaminant plume will arrive at a given point more rapidly than it would if advection alone were the acting driving force. The mean transport velocity of the contaminant mass remains the same, but concentration gradients are setup. This occurs simultaneously with the phenomenon of the contaminant occupying, with time, an increasing volume of groundwater. There are two fundamental types of dispersion—longitudinal and transverse. Both are accentuated by the inhomogeneous and anisotropic physical configuration of the permeable matrix within a groundwater system. There is also a contribution to both from diffusional transport as well.

Longtitudinal dispersion is caused by differences in groundwater velocity through pore spaces that vary in width or tortuosity. The result is dispersion that occurs along the direction of groundwater flow. Transverse