7_biodegradation

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Last update 18.12.2009. Original material by Tuomo Karvonen, 2002.

Department of Civil and Environmental Engineering, Helsinki University of Technologyhttp://civil.tkk.fi

7 BIODEGRADATION/BIOTRANSFORMATION

Biodegradation/biotransformation is biologically mediated reactions - aerobic or anaerobic - that

convert harmful organic compounds to carbon dioxide or to less harmful substance. Two main type

of reactions are discussed here: 1) organic compound is an electron donator (e.g. BTEX compoundsand other aromatic hydrocarbons) and 2) organic compound is electron acceptor (e.g. many

chlorinated solvents - PCE, TCE,..).

1) BTEX-compounds (Benzene, Toluene, Ethylbenzene and Xylene) are aromatic

hydrocarbons which are present in petroleum fuels and are often indicated as the major

contaminants from underground storage tank releases. These compounds are also widely

used as solvents, as raw materials for pharmaceutical, agrochemical, polymer, and explosiveproduction. They are also often identified as ground-water contaminants at manufacturing

facilities and in landfill leachate. Both aerobic and anaerobic degradation are discussed.

2) For chlorinated ethenes and ethanes (perchloroethene PCE, trichloroethene TCE,dichloroethene DCE, vinyl chloride, TCA trichloroethane, dichloroethane DCA and

chloroethane CA) biotransformation via reductive dechlorination is the dominant

biotransformation process at most chlorinated solvent sites. Reductive dechlorination is

assumed to occur under anaerobic conditions and dissolved solvent degradation is assumed

to follow a sequential first-order decay process.

Biodegradation of BTEX-compounds:

Naturally occurring biological processes can significantly enhance the rate of organic mass removal

from contaminated ground water aquifers. Biodegradation of BTEX-compounds is a reaction where

the contaminant (e.g. benzene C6H6) is transformed to carbon dioxide CO2 and water. Biologicallymediated degradation reactions are oxidation/reduction (redox) reactions, involving the transfer of

electrons from the organic contaminant compound to an electron acceptor. Oxygen is the electron

acceptor for aerobic metabolism whereas nitrate, ferric iron, sulphate and carbon dioxide serve aselectron acceptors for alternative anaerobic pathways. Early conceptual models of natural

attenuation were based on the assumption that the anaerobic degradation pathways were too slow tohave any meaningful effect on the overall natural attenuation rate at most sites. Accordingly, most

field programs focused only on the distribution of oxygen and contaminants, and did not measure

the indicators of anaerobic activity such as depletion of anaerobic electron acceptors or

accumulation of anaerobic metabolic by-products. Recent research suggests that hydrocarbons are

degraded both aerobically and anaerobically in subsurface environments (Rafai et al., 1998).

Figs below list the redox reactions for benzene, toluene, ethyl benzene, and xylene (BTEX). In the

presence of organic substrate and dissolved oxygen, microorganisms capable of aerobic metabolism

will predominate over anaerobic forms. However, dissolved oxygen is rapidly consumed in the

interior of contaminant plumes, converting these areas into anoxic (low oxygen) zones. Under these

conditions, anaerobic bacteria begin to utilize other electron acceptors to metabolize dissolved

hydrocarbons. The principle factors influencing the utilization of the various electron acceptors

include: 1) the relative biochemical energy provided by the reaction; 2) the availability of individual

or specific electron acceptors at a particular site; and 3) the kinetics (rate) of the microbial reactionassociated with the different electron acceptors.
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7.1 Benzene

The oxidation/reduction reactions for benzene are shown below (Rafai et al. 1998):

Oxidation reaction:

Reduction reaction:

Overall reaction:

7.2 Toluene

The oxidation/reduction reactions for toluene are shown below (Rafai et al. 1998):

Oxidation reaction:

Reduction reaction:


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Overall reaction:

7.3 Ethylbenzene and xylene

The oxidation/reduction reactions for ethylbenzene and xylene are shown below (Rafai et al. 1998):

Oxidation reaction:

Reduction reaction:


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Overall reaction:

The transfer of electrons during the redox reaction releases energy which is utilized for cell

maintenance and growth. The biochemical energy associated with alternative degradation pathways

can be represented by the redox potential of the alternative electron acceptors: the more positive theredox potential, the more energetically favourable is the reaction utilizing that electron acceptor.

With everything else being equal, organisms with more efficient modes of metabolism grow fasterand therefore dominate over less efficient forms.

________________________________________________________________________________Table 7-1.

Based solely on thermodynamic considerations, the most energetically preferred reaction should

proceed in the plume until all of the required electron acceptor is depleted. At that point, the nextmost-preferred reaction should begin and continue until that electron acceptor is gone, leading to a


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pattern where preferred electron acceptors are consumed one at a time, in sequence. Based on these

principles, one would expect to observe monitoring well data with "no-detect" results for the moreenergetic electron acceptors, such as oxygen and nitrate, in locations where evidence of less

energetic reactions is observed (e.g. monitoring well data indicating the presence of ferrous iron). Inpractice, however, it is unusual to collect samples from natural attenuation monitoring wells that are

completely depleted in one or more electron acceptors. Two processes are probably responsible forthis observation (Rafai et al. 1998):

1. Alternative biochemical mechanisms having very similar energy potentials (such as aerobic

oxidation and nitrate reduction) may occur concurrently when the preferred electron

acceptor is reduced in concentration, rather than fully depleted. Similarly noting the nearly

equivalent redox potentials for sulphate and carbon dioxide (-220 volts and -240 volts,

respectively) one might expect that sulphate reduction and methanogenic reactions may also

occur together.

2. Standard monitoring wells, having 1.5 to 3 m screened intervals, will mix waters from

different vertical zones. If different biodegradation reactions are occurring at different

depths, then one would expect to find geochemical evidence of alternative degradationmechanisms occurring in the same well. If the dissolved hydrocarbon plume is thinner thanthe screened interval of a monitoring well, then the geochemical evidence of electron

acceptor depletion or metabolite accumulation will be diluted by mixing with clean waterfrom zones where no degradation is occurring.

7.4 Kinetics of the Biodegradation Reactions

Aerobic biodegradation can be simulated as an "instantaneous" reaction that is limited by the

amount of electron acceptor (oxygen) that is available. The microbial reaction is assumed to occur

at a much faster rate than the time required for the aquifer to replenish the amount of oxygen in theplume. Although the time required for the biomass to aerobically degrade the dissolved

hydrocarbons is on the order of days, the overall rate that groundwater is replenished in most

plumes is on the order of years or tens of years. For example, microcosm data presented by Davis etal. (1994) show that microbes that have an excess of electron acceptors can degrade concentrations

of dissolved benzene (~1 mg/L) very rapidly. In the presence of a surplus of oxygen, aerobicbacteria can degrade dissolved benzene in about 8 days, which can be considered "instantaneous"

compared to years required for flowing ground water to replenish the plume area with oxygen.Recent results from the Air Force Natural Attenuation Initiative indicate that the anaerobic

reactions, which were originally thought to be too slow to be of significance in ground water, can

also be simulated as instantaneous reactions (Newell et al., 1995). For example, Davis et al. (1994)

also ran microcosms with sulphate reducers and methanogens that indicated that benzene could bedegraded within a couple of weeks time frame (after acclimation). When compared to the time

required to replenish electron acceptors, the anaerobic reactions can also be considered to be

instantaneous at many sites. This conclusion is supported by observing the pattern of anaerobic

electron acceptors and by-products along the plume at natural attenuation research sites.


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7.5 Biodegradation of chlorinated solvents (chlorinated ethenes and ethanes)

"Natural Attenuation" refers to naturally-occurring processes in soil and ground-water environments

that act without human intervention to reduce the mass, toxicity, mobility, volume, or concentration

of contaminants in those media. Biotransformation can often be a dominant process in the naturalattenuation of chlorinated solvents. At chlorinated solvent contaminated sites, most of the solvent

degradation occurs by reductive dechlorination (U.S.EPA, 1998). Reductive dechlorination is a

microbially-mediated reaction whereby a chlorine atom on the chlorinated solvent is replaced by a

hydrogen atom (Vogel and McCarty, 1987). In many bioremediation processes, an organic

contaminant (such as benzene) acts as an electron donor and another substance (such as oxygen,

nitrate, etc.) acts as the electron acceptor. However, during reductive dechlorination, hydrogen acts

as the electron donor and halogenated compounds, such as chlorinated solvents, act as electron

acceptors and thus become reduced, as shown in the following half reaction (Aziz & al. 2000):

R-Cl + H+ + 2e- R-H+ Cl-

Figure 7-1 shows the reductive transformation pathways for the common chlorinated aliphatics.

Fig. 7-1. Reductive dechlorination pathways for common chlorinated aliphatic hydrocarbons (after

Vogel and McCarty, 1985; Vogel and McCarty, 1987).

Fig. 7-2. Reductive dechlorination of perchloroethene PCE to trichloroethene TCE.


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Fig. 7-3. Reductive dechlorination of chlorinated ethanes.

For biological reductive dechlorination to occur, the following conditions must exist:1. The subsurface environment must be anaerobic and have a low oxidation-reduction potential

(ORP).

2. Chlorinated solvents that are amenable to reductive dechlorination must be present.

3. A population of dechlorinating bacteria must be present.

4. An adequate supply of fermentation substrates to produce dissolved hydrogen must be

present.

The environmental chemistry and the ORP of a site play an important role in determining whether

reductive dechlorination will occur. Based on thermodynamic considerations, reductivedechlorination will occur only after both oxygen and nitrate have been depleted from the aquifer,

because oxygen and nitrate are more energetically favourable electron acceptors than chlorinatedsolvents when hydrogen is the electron donor (U.S. EPA, 1998). According to Rafai et al. (1998),

the role of hydrogen as an electron donor during reductive dechlorination is now widely recognized

as a key factor governing the dechlorination of chlorinated compounds. The hydrogen is produced

in the terrestrial subsurface by the fermentation of a wide variety of organic compounds including

anthropogenic compounds such as petroleum hydrocarbons and natural organic matter. Hydrogen is

then used by the dechlorinating bacteria as an electron donor. Although the chlorinated solvents

degrade primarily via reductive dechlorination, which occurs under highly reduced anaerobic

conditions, some of the chlorinated solvents may degrade under aerobic conditions. TCE, c-DCE

and VC degrade cometabolically (McCarty and Semprini, 1994) and VC (Hartmans et al., 1985;

Hartmans and de Bont, 1992) and possibly c-DCE (Bradley and Chapelle, 1998) can be directly


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oxidized to carbon dioxide under aerobic conditions. PCE has not been found to degrade aerobically

(McCarty and Semprini, 1994).

7.6 Modelling of biodegradation of BTEX-compounds

Biodegradation Models for BTEX-compounds

To apply an electron-acceptor limited kinetic model, such as the instantaneous reaction model, the

amount of biodegradation that the groundwater that moves through the source zone can supportmust be calculated. The conceptual model is that:

1. Ground water upgradient of the source contains electron acceptors;2. As the upgradient ground water moves through the source zone, hydrocarbons in NAPLs

(Non-Aqueous Phase Liquids) and contaminated soil release dissolved hydrocarbons (in the

case of petroleum sites, BTEX);3. The biological reactions continue until the available electron acceptors are consumed

4. The total amount of available electron acceptors available for biological reactions can be

estimated by: a) calculating the difference between the upgradient wells and source zone

wells for oxygen, nitrate, and sulphate; and b) measuring the production of by-products

ferrous iron and methane in the source zone;

5. Using stoichiometry, a utilization factor can be developed to convert the mass of oxygen,

nitrate, and sulphate consumed to the mass of dissolved hydrocarbon that is used in the

biodegradation reactions. Similarly, utilization factors can be developed to convert the mass

of metabolic by-products that are consumed to the mass of dissolved hydrocarbon that are

used in the biodegradation reactions.

6. For a given background concentration of an individual electron acceptor, the potentialcontaminant mass removal or "biodegradation capacity" depends on the "utilization factor"

for that electron acceptor. Dividing the background concentration of an electron acceptor byits utilization factor provides an estimate (in concentration units) of the assimilative capacity

of the aquifer by that mode of biodegradation. When the available electron acceptor/by-product concentrations (Step 4) are divided by the appropriate utilization factor (Step 5), an

estimate of the "biodegradation capacity" of the groundwater flowing through the source

zone and plume can be developed.

Example: Utilization factor for benzene:

Oxygen: 1 mole benzene reacts with 7.5 moles oxygen or (6x12 + 6) grams of benzene react with

(7.5x32) grams of oxygen; 78 gms benzene react with 240 gms of oxygenUtilization Factor = 240/78 = 3.08

Nitrate 1 mole benzene reacts with 6 moles nitrate: Utilization Factor = 372.06/78 = 4.77


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7.7 Conceptual Model for Biodegradation (BIOPLUME III as an example; Rafai et al. 1999)

The main electron acceptors include oxygen for aerobic biodegradation and nitrate, iron (III),sulphate, and carbon dioxide for anaerobic biodegradation. Manganese has also been identified as

an anaerobic electron acceptor; however, manganese has not been incorporated into the current

version of BIOPLUME III. The conceptual model used in BIOPLUME III to simulate thesebiodegradation processes tracks six plumes simultaneously: hydrocarbon, oxygen, nitrate, iron (II),sulphate, and carbon dioxide. Iron (III) is input as a concentration matrix of ferric iron in the

formation. Once ferric iron is used for biodegradation, BIOPLUME III simulates the production and

transport of ferrous iron. Biodegradation occurs sequentially in the following order:

Oxygen Nitrate Iron (III) Sulphate Carbon Dioxide

The biodegradation of hydrocarbon in a given location using nitrate, for example, can only occur if

oxygen has been depleted to its threshold concentration at that location. Three different kinetic

expressions can be utilized for the biodegradation reaction for each of the electron acceptors:

1. First-order decay

2. Instantaneous reaction

3. Monod kinetics

The first-order decay model implemented e.g. in BIOPLUME III (Rafai et al.) for any of the

electron acceptors is limited by the availability of the electron acceptor in question. In other words,the model allows the first-order reaction to take place up to the point that the electron acceptor

concentration available in the aquifer has been depleted. The Monod kinetic model assumes aconstant microbial population for each of the aerobic and anaerobic reactions and does not simulate

the growth, transport and decay of the microbial population in the subsurface.

7.7.1 Monod kinetics (Michaelis-Menten kinetics)

A popular expression for simulating biological processes (e.g. biodegradation) is the Monod-

function referred also to as Michaelis-Menten kinetics:

=MAXC/(Kc+C) (7-1)

where is the growth rate of biomass [1/d], MAX is the maximum growth rate [1/d], C isconcentration of growth-limiting substrate [mg/l] and Kc is half-saturation coefficient [mg/l] which

is the concentration which allows the microorganism to grow at half the maximum growth rate.

In groundwater, the Monod growth function is related to the rate of decrease of an organic

compound. The reduction of contaminant concentrations using Monod kinetics can be expressed as

(Wiedemeier et al. 1999):

dC/dt = MtMAXC/(Kc+C) (7-2)

where C is the contaminant concentration, Mt is the total microbial concentration, MAX is themaximum contaminant utilization per unit mass microorganisms and Kc is the contaminant half-

saturation constant.


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7.7.2 Limitations

The limitations/difficulties involved in the modelling of biodegradation of BTEX-compounds

1. It is difficult to take into account the selective or competitive biodegradation of thehydrocarbons. This means that hydrocarbons are generally simulated as a lumped organic

which represents the sum of benzene, toluene, ethyl benzene or xylene. If a singlecomponent is to be simulated, the user would have to determine how much electron acceptor

would be available for the component in question.

2. The conceptual models for biodegradation are simplifications of the complex biologically

mediated redox reactions that occur in the subsurface.

Exercise 7-1

Consider biodegradation of benzene in saturated groundwater. Biodegradation follows first-order

reaction:

dC/dt = -kBC and initial concentration C(t=0) = 2000 g/l

Biodegradation coefficient is assumed to be constant; kB = 0.005 d-1

. Other processes are neglected

here. Exact solution is

C(t) = exp(-kBt)*C(0)

a) How long time does it take that concentration falls below the drinking water limit 5 g/l?b) Degradation equation can be solved numerically by approximating the differential equation

as follows:

[Ct+1

- Ct

]/ t = - kB[ Ct+1

+ (1- )Ct

] (7-3)

where Ct+1 is the unknown value at time t+t, Ct is the known value at time t is the coefficient of

implicity; if =0, the solution is explicit, if =0.5, the approximation is called Crank-Nicolson and

if =1, the solution is fully implicit.

What is the error (%) of the numerical solution after 3600 days as compared to the analytical

solution when a) explicit, b) Crank-Nicolson and c) fully implicit solution is used. Time step t=30d.

c) Optimal value for the weighting coefficient OPT can be calculated using the method givenin section 5.2.6 Computation of first-order reaction (biodegradation).

1) What is the optimal value for if input data from b) are used?

2) What is the optimal value for if time step is 300 d? Show that the numerical solution

method gives accurate solution if optimal value for is used for both t=30 d, and

t=300 d.

Exercise 7-2

Calculate utilization factors for toluene both in aerobic degradation and in anaerobic degradation

reactions (nitrate, manganese, iron(III), sulphate and methanogenic reaction).


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7.8 Sequential model for chlorinated solvents

Reductive dechlorination can be modelled as a sequential first-order decay process. This means that

a parent compound undergoes first-order decay to produce a daughter product and that product

undergoes first-order decay and so on. Generally, the more highly chlorinated the compound, themore rapidly it is reduced by reductive dechlorination (Vogel and McCarty, 1985; Vogel and

McCarty, 1987).

Kinetics of Sequential First Order Decay

Although the chlorinated ethenes primarily degrade biologically, chlorinated ethanes can degrade

both biologically and abiotically. For chloroethane (CA), abiotic decay to ethanol occurs much

more rapidly than biotransformation to ethane. The abiotic decay of 1,1-DCA is slow relative tobiotransformation. 1,1,1-TCA can degrade abiotically to both acetic acid (by hydrolysis) and to 1,1-

DCE (by elimination) (Vogel and McCarty, 1987).

7.9 Chlorinated Ethenes

The reaction rate equations describing the sequential first order decay of the chlorinated ethenes are

shown below:

(7-4)

where 1, 2, 3, 4 and E, are the first order biotransformation rate coefficients, y1, y2, y3, and y4are the daughter: parent compound molecular weight ratios and CPCE, CTCE, CDCE, CVC and CETH are

the aqueous concentration of PCE, TCE, DCE, vinyl chloride VC, and ethene, respectively. From

these expressions, it is clear that TCE, DCE, and VC are simultaneously being produced and

degraded, which often results in net accumulation before observed degradation. Furthermore, these

reaction expressions cause the reactive transport equations to be coupled to each other as discussed.

General transport equation for chlorinated solvents:

(7-5)

so that r1=rPCE, r2=rTCE, r3=rDCE, r4=rVC, r5=rETH.

Transport of chlorinated ethenes can be calculated by solving a set of coupled partial differential

equations to describe the reactive transport of chlorinated solvent species, such as PCE, TCE, DCE,

VC and ETH, in saturated ground-water systems. The equations describe one-dimensional

advection, three-dimensional dispersion, linear sorption, and sequential, first-order


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biotransformation. All equations, except the first, are coupled to a parent species equation through

the reaction term as shown below:

(7-6)

where C1, C2, C3, C4 and C5 concentrations of PCE, TCE, DCE, VC, and ETH, respectively [mg/L],

Dx , Dy and Dz are the hydrodynamic dispersion coefficients [m2/d], vs is the seepage velocity [m/d],

i, i=1,..,5 are the first-order degradation coefficients [1/d], yi, i=1,,5 are the yield coefficients(dimensionless values, for example, y1 would represent the mg of TCE produced per unit mg of

PCE destroyed); and Ri, i=1,,5 are respective retardation factors.

7.10 Chlorinated Ethanes

The following are the rate expressions for the degradation of the chlorinated ethanes.

(7-7)

where 5, 6, and 7 are the first order biotransformation rate coefficients, A is the abiotic ratecoefficient for chloroethane, y5 and y6 are the daughter: parent compound molecular weight ratios

and CTCA, CDCA and CCA and the concentration of 1,1,1-trichloroethane, 1,1-dichloroethane and

chloroethane, respectively.

Yield constants (i.e., y1, y2,y6) are incorporated to account for molecular weight differences

between parent and daughter compounds. The constants are necessary because kinetic expressionsare valid on a molar basis only.

Exercise 7-3Compute the yield coefficients of chlorinated ethenes and ethanes y1,..y6.


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Exercise 7-4Use EXCEL to create a model for calculating the transport and sequential biodegradation of

chlorinated ethenes. Compute advection, adsorption (linear isotherm) and biodegradation. Assume1-D case, neglect dispersion in all directions (Dx=Dy=Dz=0) and use only one storage to describe

the aquifer (i.e. an ordinary differential equation is used for each compound). Input data: Pore water velocity v = 0.5 m/d

Aquifer length dx=15 m

Retardation factor R same for all compounds; R=3.5

Biodegradation coefficients: PCE = 0.0055 1/d, TCE = 0.0041 1/d, DCE =0.0022 1/d, VC =

0.0018 1/d and ETH=0.0.

Concentration is zero for all compounds at time t=0

Source area concentrations (constant) are: CPCE = 10 mg/l, CTCE = 40 mg/l, CDCE = 100 mg/l,CVC = 5 mg/l, CETH = 0.1 mg/l

Yield coefficients y1,..y4 as computed in Ex. 7-3.

Calculate a period of five years using explicit solution method and select a time step length that

ensures numerical stability of the solution.

Note! Input concentration of the aquifer is the source area concentration.

7.11 Remediation by Natural Attenuation (RNA)

7.11.1 Fundamentals of Natural Attenuation (Intrinsic Remediation)

Naturally occurring biological processes can significantly enhance the rate of organic mass removal

from contaminated aquifers. Biodegradation research performed in USA by universities,

government agencies, and other research groups has identified several main themes that are crucial

for future studies of natural attenuation (Rafai et al. 1998):

1. The relative importance of groundwater transport vs. microbial kinetics is a key

consideration for developing workable biodegradation expressions in models. Results from

the United Creosote site (Texas) and the Traverse City Fuel Spill site (Michigan) indicatethat biodegradation is better represented as a macro-scale wastewater treatment-type process

than as a micro-scale study of microbial reactions.

2. The distribution and availability of electron acceptors control the rate of in-situbiodegradation for most petroleum release site plumes. Other factors (e.g., population ofmicrobes, pH, temperature, etc.) rarely limit the amount of biodegradation occurring at these

sites.

Naturally-occurring processes in soil and ground-water environments reduce the mass, toxicity,

mobility, volume, or concentration of contaminants in those media. These in-situ processes include

biotransformation, dispersion, dilution, adsorption, volatilization, and chemical or biological

stabilization or destruction of contaminants (U.S. EPA, 1998).


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7.11.2 How to use models in Remediation by Natural Attenuation?

Models are generally used to answer a number of questions regarding natural attenuation:

1. How long will the plume extend if no engineered/source controls are implemented?

2. How long will the plume persist until natural attenuation processes completely dissipate thecontaminants?

3. How long will the plume extend or persist if some engineered controls or source reduction

measures are undertaken (for example, free phase removal or residual soil contamination

removal)?

The model can also be used to simulate bioremediation of BTEX-hydrocarbons in ground water by

injecting electron acceptors (except for iron(III)) and can also be used to simulate air sparging for

low injection air flow rates. Finally, the model can be used to simulate advection, dispersion, and

sorption without including biodegradation.

The models will predict the maximum extent of dissolved-phase plume migration, which may then

be compared to the distance to potential points of exposure (e.g. drinking water wells, ground-waterdischarge areas, or property boundaries).

Models can be used in RNA studies in two ways:

1. As a screening-level model to determine if RNA is feasible at sites contaminated withBTEX-compounds or with chlorinated solvents. Model is intended to be used as a

screening-level model to determine if natural attenuation is occurring at sufficient rates at asite to warrant a full natural attenuation study. Ideally, site-specific biotransformation rate

constants should be employed, but literature values can be used if measured rate constants

are unavailable.

2. As an RNA ground-water model to address selected chlorinated solvent problems

A mathematical model is an appropriate tool at sites where simplifying assumptions (e.g. uniform

ground-water flow, a vertical plane source, first-order decay) can be made so that the resulting

simulations provide useful information for the problem being addressed.

In any modelling study, it is recommended that proper care should be used to select the model that

is best suited to 1) the source, hydrogeology, and biotransformation processes present at the site

and, 2) the type of problem being addressed (e.g. screening of alternatives, providing supporting

evidence of natural attenuation, developing detailed design information).

A screening model has the following limitations:1. It assumes simple ground-water flow conditions. The model should not be applied where

pumping systems create a complicated flow field. In addition, the model should not be

applied where vertical flow gradients affect contaminant transport.

2. A screening tool assumes uniform hydrogeologic and environmental conditions over the

entire model area. Moreover, it assumes constant source, hydrogeological, and biological

property values for the entire model area and, therefore, simplifies actual site conditions. For

this reason, the screening model should not be applied where extremely detailed, accurate

results that closely match site conditions are required. More comprehensive numerical

models should be applied in such cases.


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7.11.3 Natural Attenuation Lines of Evidence

To support remediation by natural attenuation, it must be scientifically demonstrated that

attenuation of the site contaminants is occurring at rates sufficient to be protective of human healthand the environment. According to the "Technical Protocol for Evaluating Natural Attenuation of

Chlorinated Solvents in Ground Water" (U.S. EPA, 1998), three lines of evidence can be used tosupport natural attenuation of chlorinated solvents including:

1. Observed reductions in contaminant concentrations along the flow path downgradient from

the source of contamination

2. Documented loss of contaminant mass at the field scale using:

a) Chemical and geochemical analytical data including decreasing parent compound

concentration, increasing daughter compound concentrations, depletion of electron

acceptors and donors, and increasing metabolic by-product concentrations; and/or

b) A rigorous estimate of residence time along the flow path to document contaminant

mass reduction and to calculate biological decay rates at the field scale.

3. Laboratory microcosm or field data that support the occurrence of biotransformation and

give rates of biotransformation.

At a minimum, the investigator must obtain the first two lines of evidence or the first and third linesof evidence. The second or third line of evidence is crucial because it provides biotransformation

rate constants. These rate constants can be used in conjunction with other fate and transportparameters to predict contaminant concentration and to assess risk at a downgradient point of

exposure (U.S. EPA, 1998). Compared to fuel hydrocarbon plumes, use of natural attenuation as astand-alone remedy for chlorinated solvent plumes is appropriate for a much lower percentage of

plumes, because of their longer plume lengths.

7.12 Ground Water Characterization measurements (According to Rafai et al., 1998)

Dissolved Oxygen

Dissolved oxygen is the most thermodynamically favoured electron acceptor used in the

biodegradation of fuel hydrocarbons. Dissolved oxygen concentrations are used to estimate the

mass of contaminant that can be biodegraded by aerobic processes. As a rule, the stoichiometric

ratio of dissolved oxygen consumed by microbes to destroyed BTEX compound is 1.0 mg/L of

dissolved oxygen consumed to approximately 0.32 mg/L of BTEX compounds destroyed. During

aerobic biodegradation, dissolved oxygen levels are reduced as aerobic respiration occurs. Also,anaerobic bacteria (obligate anaerobes) generally cannot function at dissolved oxygen levels greater

than about 0.5 mg/L. Therefore, higher values of dissolved oxygen indicate that aerobic

biodegradation is likely at work.

Redox potential

The oxidation/reduction (redox) potential of ground water (EH) is a measure of electron activity and

is an indicator of the relative tendency of a solution to accept or transfer electrons. Redox reactions

in ground water are usually biologically mediated and therefore, the redox potential of a ground

water system depends upon and influences rates of biodegradation. Knowledge of the redox

potential of ground water is also important because some biological processes only operate within a

prescribed range of redox conditions. Knowledge of the redox potential of ground water can be used

as an indicator of certain geochemical activities such as sulphate reduction. The redox potential of

ground water generally ranges from -400 millivolts [mV] to 800 mV. Redox potential can be used

to provide real time data on the location of the contaminant plume, especially in areas undergoinganaerobic biodegradation. Mapping the redox potential of the ground water while in the field allows


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the field scientist to determine the approximate location of the contaminant plume. To map the

redox potential of the ground water while in the field it is important to have at least one redoxmeasurement (preferably more) from a well located upgradient of the plume. The redox potential of

a ground water sample taken inside the contaminant plume should be somewhat lower than thattaken in the upgradient location. Redox potential measurements should be taken during well

purging and immediately before and after sample acquisition using a direct-reading meter. Becausemost well purging techniques can allow aeration of collected ground water samples (which can

affect redox potential measurements), it is important to minimize potential aeration.

pH, Temperature, and Conductivity

Because the pH, temperature, and electric conductivity of a ground water sample can change

significantly within a short time following sample acquisition, these parameters must be measured

in the field in unfiltered, unpreserved, "fresh" water collected by the same technique as the samples

taken for laboratory analyses. The measurements should be made in a clean glass container separate

from those intended for laboratory analysis and the measured values should be recorded in the

ground water sampling record. The pH of ground water has an effect on the presence and activity of

microbial populations in ground water. This is especially true for methanogens which may be activeafter all aerobic, sulphate reduction, and nitrate reduction degradation has taken place. Microbescapable of degrading petroleum hydrocarbon compounds generally prefer pH values varying from 6

to 8 standard units. Electric conductivity is a measure of the ability of a solution to conductelectricity. For ground water, conductivity is directly related to the concentration of ions in solution,

increasing as ion concentration increases. Like chloride, conductivity is used to ensure that ground

water samples collected at a site are representative of the water comprising the saturated zone in

which the dissolved-phase contamination is present. If the conductivities of samples taken from

different sampling points are radically different, then the waters may be from different

hydrogeologic zones. Ground water temperature directly affects the solubility of oxygen and other

geochemical species. The solubility of dissolved oxygen is temperature dependent, being more

soluble in cold water than in warm water. Ground water temperature also affects the metabolic

activity of bacteria. Rates of hydrocarbon biodegradation roughly double for every 10C increase in

temperature ("Q"10 rule) over the temperature range between 5 and 25C. Ground water

temperatures less than about 5C tend to inhibit biodegradation, and slow rates of biodegradation

are generally observed in such waters.

Alkalinity

The total alkalinity of a ground water system is indicative of waters capacity to neutralize acid.

Alkalinity is defined as the net concentration of strong base in excess of strong acid with a pure CO 2-water system as the point of reference (Domenico and Schwartz, 1990). Alkalinity results from the

presence of hydroxides, carbonates, and bicarbonates of elements such as calcium, magnesium,sodium, potassium, or ammonia. These species result from the dissolution of rock (especially

carbonate rocks), the transfer of CO2 from the atmosphere, and respiration of microorganisms.

Alkalinity is important in the maintenance of ground water pH because it buffers the ground water

system against acids generated through both aerobic and anaerobic biodegradation processes.

Nitrate

In the hierarchical order of processes occurring in the microbiological treatment zone, after

dissolved oxygen has been depleted, nitrate may be used as an electron acceptor for anaerobic

biodegradation. Nitrate concentrations are used to estimate the mass of contaminant that can be

biodegraded by denitrification processes. By knowing the volume of contaminated ground water,

the background nitrate concentration, and the concentration of nitrate measured in the contaminatedarea, it is possible to estimate the mass of BTEX lost to biodegradation. Stoichiometrically, each 1.0


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mg/L of ionic nitrate consumed by microbes results in the destruction of approximately 0.21 mg/L

of BTEX compounds. Nitrate concentrations are a direct input parameter to the Platform.

Sulphate and Sulphide Sulphur

After dissolved oxygen and nitrate have been depleted in the microbiological treatment zone,

sulphate may be used as an electron acceptor for anaerobic biodegradation. This process is termedsulphanogenesis and results in the production of sulphide. Sulphate concentrations are used as an

indicator of anaerobic degradation of fuel compounds. By knowing the volume of contaminated

ground water, the background sulphate concentration, and the concentration of sulphate measured in

the contaminated area, it is possible to estimate the mass of BTEX lost to biodegradation through

sulphate reduction. Stoichiometrically, each 1.0 mg/L of sulphate consumed by microbes results in

the destruction of approximately 0.21 mg/L of BTEX. Sulphate concentrations are a direct input

parameter for the Platform.

Ferrous Iron

Ferric iron is also used as an electron acceptor during anaerobic biodegradation of petroleum

hydrocarbons after nitrate or sulphate depletion, or some times in conjunction with them. Duringthis process, ferric iron is reduced to the ferrous form which may be soluble in water. Ferrous ironconcentrations are used as an indicator of anaerobic degradation of fuel compounds. By knowing

the volume of contaminated ground water, the background ferrous iron concentration, and the

concentration of ferrous iron measured in the contaminated area, it is possible to estimate the mass

of BTEX lost to biodegradation through ferric iron reduction. Stoichiometrically, the degradation of

1 mg/L of BTEX results in the production of approximately 21.8 mg/L of ferrous iron during ferric

iron reduction. Iron concentrations are used as a direct input parameter to the Platform. The

equivalent amount of Ferric Iron is estimated from the measured Ferrous Iron, which is used as

model input. BIOPLUME III simulates the hydrocarbon reduction and the corresponding Ferric

Iron depletion.

Carbon Dioxide

Metabolic processes operating during biodegradation of fuel hydrocarbons result in the productionof carbon dioxide (CO2). Accurate measurement of CO2 produced during biodegradation is difficult

because carbonate in ground water (measured as alkalinity) serves as both a source and sink for freeCO2. If the CO2 produced during metabolism is not removed by the natural carbonate buffering

system of the aquifer, CO2 levels higher than background may be observed. Comparison ofempirical data to stoichiometric calculations can provide estimates of the degree of microbiological

activity and the occurrence of in situ mineralization of contaminants.

MethaneDuring methanogenesis (an anaerobic biodegradation process), carbon dioxide (or acetate) is used

as an electron acceptor, and methane is produced. Methanogenesis generally occurs after oxygen,

nitrate, and sulphate have been depleted in the treatment zone. The presence of methane in ground

water is indicative of strongly reducing conditions. Because methane is not present in fuel, the

presence of methane in ground water above background concentrations in contact with fuels is

indicative of microbial degradation of fuel hydrocarbons. Methane concentrations can be used to

estimate the amount of BTEX destroyed in an aquifer. By knowing the volume of contaminated

ground water, the background methane concentration, and the concentration of methane measured

in the contaminated area, it is possible to estimate the mass of BTEX lost to biodegradation through

methanogenesis reduction. The degradation of 1 mg/L of BTEX results in the production of

approximately 0.78 mg/L of methane during methanogenesis. Methane concentrations are used asan indirect input parameter to the Platform. The equivalent amount of CO 2 is estimated from the


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measured methane, which is used as model input. BIOPLUME III simulates the hydrocarbon

reduction and the corresponding CO2 depletion.

Chloride

Chloride is used to ensure that ground water samples collected at a site are representative of the

water comprising the saturated zone in which the dissolved-phase contamination is present (i.e. toensure that all samples are from the same ground water flow system). If the chloride concentrations

of samples taken from different sampling points are radically different, then the waters may be from

different hydrogeologic zones.

Total Petroleum Hydrocarbons and Aromatic Hydrocarbons

These analytes are used to determine the type, concentration, and distribution of fuel hydrocarbon in

the aquifer. Of the compounds present in most gasolines and jet fuels, the BTEX compounds

generally represent the regulatory contaminants of concern. For this reason, these compounds aregenerally of significant interest in the fate and transport analysis. At a minimum, the aromatic

hydrocarbon analysis (Method SW8020) must include the BTEX compounds and the

trimethylbenzene and tetramethylbenzene isomers. The combined dissolved-phase concentrations ofBTEX, trimethylbenzene, and tetra-methylbenzene should not be greater than about 30 mg/L for aJP4 spill. If these compounds are found in concentrations greater than 30 mg/L then sampling errors

such as emulsification of NAPL in the ground water sample have likely occurred and should be

investigated.

Exercise 7-5

Landfill bottom liner is composed of a clay soil: the thickness of the layer is 1 m, hydraulic

conductivity is K=1.0*10-9

m/s, porosity n=0.5, dry bulk density = 1.5 kg/dm3 and combineddiffusion/dispersion coefficient for benzene is D=2.5*10

-10m

2/s. The hydraulic gradient is assumed

to be unity, i.e. dH/dz=1.0. Adsorption isotherm for benzene is linear. See also section 4.5 Case

Study: landfill bottom liner.

Transport of benzene through the clay liner into the underlying groundwater zone is computed by

solving numerically the advection -diffusion/dispersion - adsorption - biodegradation equation:

(7-8)

Initial and boundary conditions:

C(x,0) = 0.0

C(0,t) = 100 g/l (the limiting value of benzene in drinking water is 5 g/l)dC(L,t)/dz = (v/R)C (L = 100 cm) (advection outflow)

The adsorption is taken into account using the linear isotherm, i.e.

S = KdC

which leads to the final partial differential equation to be solved:


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Department of Civil and Environmental Engineering, Helsinki University of Technology

(7-9)

where

In this case the distribution coefficient Kd is 0.67 l/kg. The equation was solved numerically and

after 15 years a steady-state situation was attained. In steady-state the penetration of benzene was

stopped and the influence of advection, diffusion and dispersion equals the biodegradation rate. In

the numerical solution the nodal distance is dz=2.0 cm.

The table shown below lists the steady-state concentration distribution as a function of depth z

between 0..49 cm.

Problem: calculate an estimate of the biodegradation coefficient kB based on the available data (unit

of kB is d-1

).

7_biodegradation

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