Loannis Paspaliaris Nymphodora Papassiopi Anthimos Xenidis and Yung Tse Hung

CONTENTS

14.1 Introduction 519

14.1.1 Contamination of Soils 519

14.1.2 Remediation Technologies 521

14.2 Soil Vapor Extraction 523

14.2.1 General Description 523

14.2.2 Design Considerations 524

14.3 Bioremediation 534

14.3.1 Introduction 534

14.3.2 Principles of Bioremediation 535

14.3.3 Engineering Factors 537

14.3.4 In Situ Methods for the Biological Treatment of Organic Contaminants 539

14.3.5 Ex Situ Biological Treatment 545

14.4 Phytoremediation 546

14.4.1 General Description 546

14.4.2 Phytoremediation Mechanisms 547

14.4.3 Design Considerations 553

14.5 Soil Washing 560

14.5.1 General Description 560

14.5.2 Design Considerations 562

14.6 In Situ Soil Flushing 563

14.6.1 General Description 563

14.6.2 Design Considerations 564

Nomenclature 565

References 567

14.1 INTRODUCTION 14.1.1 Contamination of Soils

Soil can be defined as the top layer of the Earth's crust, consisting of mineral particles, organic matter, water, air, and living organisms. As the interface between the Earth's atmosphere and the lithosphere, soil performs a number of diverse functions essential for life preservation and human activities; it is the substrate necessary for the growth of plants and animals and the basis for all agricultural production, and it serves as a protection and filtering layer necessary for clean ground-water supplies. The rate of soil formation and regeneration is very slow, so soil is practically a nonrenewable resource. In view of the high rates of soil degradation, it has become essential that soil resources be protected against the factors that degrade its quality and limit its availability. Human activities can greatly affect the geochemical cycles of soil constituents, resulting in the contamination

of soil with heavy metals and other toxic compounds. Soil contamination is mainly the result of improper environmental management in chemical industries, mining and mineral processing operations, industrial waste disposal sites, municipal landfills, and other facilities, both during operation and after closure. Additionally, widespread soil contamination may occur as a result of emissions from transport and industry, which re-deposit onto the soil surface, as well as from overuse of agricultural chemicals. The result of this diffuse soil contamination is the accumulation of the various contaminants in the soil surface layer and their dissolution and transportation into deeper soil layers and groundwater under the effect of the infiltrating water. In some cases, uncontrolled urban expansion has led to changes in land use, and former mining or industrial sites have been gradually transformed into residential, recreational, or even agricultural areas; in these cases, contaminated land may pose a high risk to human health and agricultural production.

Soil contamination was not perceived as a problem until the 1970s, when incidents in the U.S. and Europe (Love Canal, NY; Times Beach, MO; Lekkerkerk, the Netherlands) awakened public awareness about the serious threats posed to human health and the environment by abandoned or improperly managed hazardous wastes. In response to the growing public concern, the U.S., the Netherlands, and a number of other European countries started a systematic effort beginning in 1980 to identify potentially contaminated sites, assess the level of contamination, establish priorities for remediation based on risk assessment studies and gradually implement the required remedial actions.

In the U.S., three federal programs are currently in progress for identifying and cleaning up contaminated sites1:

1. In 1980, Congress passed the Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA). Commonly known as Superfund, the program under this law is focused on the remediation of abandoned or uncontrolled hazardous waste sites. Since 1980, Superfund has assessed nearly 44,400 sites. To date, 33,100 sites have been removed from the Superfund inventory to aid their economic redevelopment, and 11,300 sites remain active with the site assessment program or are included in the National Priorities List (NPL) for the implementation of remedial actions. By September 2000, 1509 sites were included in the NPL with ongoing or completed cleanup activities.

2. The second program is directed at corrective actions at currently operating industrial facilities. This program is authorized by the Resource Conservation and Recovery Act (RCRA) of 1980 and its subsequent amendments. At the time of writing, there are no statistical data about the progress of this program. Approximately 2000 sites were included in the RCRA Corrective Action Baseline by the end of September 2007. Amongst these sites, remedy constructions were completed for 560 sites and remedy decisions were made for 726 sites.

3. The third cleanup program, also authorized by the RCRA, addresses contamination resulting from leaks and spills (mainly petroleum products) from underground storage tanks (USTs). This law has compelled cleanup activities at many UST sites. By February 1999, over 385,000 releases had been reported, 327,000 cleanup projects initiated, and 211,000 projects completed.

Many policies and practices have been adopted by European countries for the management of contaminated sites. Information about the various national polices, the technical approaches for risk assessment, and the progress of rehabilitation activities in Europe has been compiled in the framework of two European networks—CARACS (Concerted Action for Risk Assessment for Contaminated Sites) and CLARINET (Contaminated Land Rehabilitation Network for Environmental Technologies)—which were funded by the European Commission. A detailed description of European national policies can be found in relevant publications2,3 and in the CLARINET website (http://www.clarinet.at).

Table 14.1 summarizes the available data related to the registration, assessment, and remediation of contaminated sites in the U.S. and several European countries. The number of sites presented

TABLE 14.1

Available Data for the Registration, Assessment, and Remediation of Contaminated Sites in the U.S. and Europe

TABLE 14.1

Available Data for the Registration, Assessment, and Remediation of Contaminated Sites in the U.S. and Europe

Number of Sites

Suspected

Confirmed

Cleanup Initiated

Country

Contamination

Contamination

or Completed

Data Till

U.S., Superfund

44,400

11,300

1,509

2000

U.S., UST

385,000

211,000

1999

U.K.

100,000

1995

The Netherlands

110,000

1998

Belgium

10,500

86

1998

France

250,000

896

125

1997

Spain

18,000

4,900

77

1995

Italy

9,000

1,570

1997

Germany

300,000

1997

Austria

2,476

145

97

1999

Switzerland

50,000

3,000

200

1998

Denmark

14,500

4,048

800

1997

Norway

3,350

2,100

99

1999

Finland

25,000

1,200

1995

Sweden

20,000

12,000

1999

Hungary

10,000

200

1998

Czech Republic

12,000

1,000

210

1998

Source: From NATO/CCMS, Evaluation of Demonstrated and Emerging Technologies for the Treatment and Clean Up of Contaminated Land and Groundwater, NATO CCMS Pilot Study, Phase III, 1999 Annual Report, EPA 542/R-99/007, no. 235, 1999; Ferguson, C. and Kasamas, H., Eds., Risk Assessment for Contaminated Sites in Europe, Vol. 2. Policy Frameworks, LGM Press, Nottingham, UK, 1999. With permission.

Source: From NATO/CCMS, Evaluation of Demonstrated and Emerging Technologies for the Treatment and Clean Up of Contaminated Land and Groundwater, NATO CCMS Pilot Study, Phase III, 1999 Annual Report, EPA 542/R-99/007, no. 235, 1999; Ferguson, C. and Kasamas, H., Eds., Risk Assessment for Contaminated Sites in Europe, Vol. 2. Policy Frameworks, LGM Press, Nottingham, UK, 1999. With permission.

in the table changes yearly, because the entire process is in a state of continuous progression. It has been suggested that the real extent of the problem has become clear only recently. For example, in 1980 about 350 sites in the Netherlands were thought to be contaminated. This number increased to 1600 in 1986 and 110,000 in 1999. The estimated costs1 for rehabilitation of these sites was 0.5 billion Euros in 1980, 3 billion Euros in 1986, and between 15 and 25 billion Euros in 1999.

14.1.2 Remediation Technologies

Until recently, a common practice for the remediation of contaminated sites was to excavate the contaminated soil, replace it with clean soil, and then dispose of the contaminated material at municipal waste landfills. This practice, however, was gradually discouraged by the environmental authorities, which issued very strict regulations for landfilling and increased the corresponding disposal costs. In many industrial countries, the cost for disposal in a municipal waste landfill ranges from 80 to 150 USD/ton. If contaminated soil is characterized as hazardous waste, landfilling in state-of-the-art hazardous waste landfills may cost4 between 500 and 800 USD/ton. The high disposal costs and the limited availability of clean soil has led to the development of alternative remediation methods, which permit the reuse of treated soil following the removal or immobilization of contaminants.

Soil remediation technologies can be classified according to the type of treatment processes taking place5-7:

1. Biological processes. These are based on the use of living organisms (e.g., microorganisms or plants).

2. Chemical processes. These destroy, fix, or remove toxic compounds by using one or more types of chemical reactions.

3. Physical processes. These separate contaminants from the soil matrix by exploiting physical differences between the soil and the contaminants (e.g., volatility) or between contaminated and uncontaminated soil particles.

4. Solidification and stabilization processes. These immobilize the contaminants through physical or chemical processes. Solidification involves the entrapment of contaminants into a consolidated mass and stabilization is the conversion of contaminants to a chemical form that is less available.

5. Thermal processes. These exploit physical and chemical processes at elevated temperatures.

Another classification of remediation technologies describes where the action is taking place. Ex situ methods are those applied to excavated soil and in situ processes are those applied to the soil in its original location. On-site techniques are those that take place on the contaminated site; they can be either ex situ or in situ. Off-site processes treat the excavated soil in fixed industrial facilities, away from the contaminated site.

The following categories of technologies are predominately ex situ:

1. Soil washing and related chemical treatment techniques

2. Solidification-stabilization

3. Thermal processes

4. Vitrification

5. Bioremediation using landfarming or biopile techniques

The most common in situ technologies are as follows:

1. Soil vapor extraction (SVE)

2. Air sparging

3. In situ bioremediation techniques combined with SVE and air sparging

4. Soil flushing

5. Electroremediation

6. Phytoremediation

Currently, most remediation projects are carried out using ex situ technologies, both in the U.S. and in Europe. However, there is an increasing trend toward the application of in situ technologies because of their considerable advantages over ex situ techniques, such as less disturbance of the site, lower treatment costs, and so on.

Published data for the cost of remediation technologies are highly variable. One reason for this variability is that remediation costs depend on several case-specific parameters, such as type of contaminants, geotechnical and geochemical characteristics of the soil matrix, and the hydrogeol-ogy of the site for in situ techniques. Differences in the reported cost data for the same technology between two countries may also reflect a different degree of commercialization for the specific technology. Indicative cost ranges for characteristic remediation technologies are presented in Table 14.2, based on the U.S. and European Union (EU) experiences.

This chapter presents a detailed description of five technologies: soil vapor extraction, bioreme-diation, phytoremediation, soil washing, and soil flushing. Information about other categories of proven or emerging technologies is available on several websites. An overview of the technologies currently applied in the U.S., with detailed cost and performance data from characteristic case studies, can be found at the FRTR (Federal Remediation Technologies Roundtable) website (http:// www.frtr.gov). Detailed information on several soil remediation technologies can also be found on the United States Environmental Protection Agency's (U.S. EPA) Cleanup Information site (http:// www.clu-in.org).

TABLE 14.2

Indicative Costs of Remediation Technologies

Remediation Technology

Range of Costs in the U.S.a (USD/t) Range of Costs in the EUb (Euro/t)

Bioremediation Soil washing

Stabilization-solidification

Thermal treatment

Incineration

Soil vapor extraction

Phytoremediation

50-150 80-120 240-340 120-300 200-1500 20-220 10-35

20-40 20-200 80-150 30-100 170-350 20-60

aSource: Schnoor, J.L., Phytoremediation. Technology Evaluation Report TE-98-01, Ground-Water Remediation Technologies Analysis Center, Pittsburgh, PA, 1997. With permission. bSource: Vic, E.A. and Bardos, P., Remediation of Contaminated Land. Technology Implementation in Europe, Federal Environmental Agency, Austria. CLARINET Report, available at www.clarinet.at, 2002. With permission.

14.2 SOIL VAPOR EXTRACTION 14.2.1 General Description

Soil vapor extraction (SVE) is a relatively new yet widely applied technology for the remediation of soils contaminated with volatile organic compounds (VOC) in the unsaturated zone above the water table (vadose zone). The process consists of generating an airstream through the contaminated soil subsurface in order to enhance the volatilization of organic contaminants and thus remove them from the soil matrix.9-13

Figure 14.1 presents the main components of a typical in situ SVE system.9,10 Vertical extraction wells are installed inside the contaminated zone at appropriate distances from one another. The SVE wells are typically constructed of PVC pipe, with a screened interval, which is placed within the contaminated zone. The wells are connected to blowers or vacuum pumps, which induce a continuous airflow through the pores of the unsaturated zone. The soil surface is sometimes covered with an impermeable seal, made from high density polyethylene (HDPE) or bentonite clay for example, to prevent the vertical influx of air from the surface, which might cause short-circuiting problems, and promote horizontal gas flow through the contaminated area. The airstream, which contains the contaminant vapors, passes initially through an air-water separation unit to remove the entrained moisture and is then directed to the gas treatment unit, where the contaminants are thermally destroyed or removed by adsorption.

There are three main prerequisites for the successful application of SVE technology:

1. The contamination should be trapped in the vadose zone.

2. The contaminants should have high volatility.

3. The contaminated zone should have high permeability.

A general simple rule is that SVE can be applied successfuly for contaminants with vapor pressure greater than 0.5 mmHg and for soils with air permeability coefficients ranging between 1 x 10 2 and 1 x 10-5 cm/s.11

Many modifications and additional treatment options have been proposed to enhance the performance and extend the applicability of SVE systems, examples of which include the following:

1. Pumping of the groundwater to lower the water table and enlarge the vadose zone, with simultaneous treatment of contaminated groundwater.10

2. The combination of SVE with air sparging technology. Air sparging involves the injection of air into the saturated zone of contaminated groundwater. The air bubbles enhance the

Water table

FIGURE 14.1 Schematic representation of an SVE system.

Water table

FIGURE 14.1 Schematic representation of an SVE system.

volatilization of dissolved contaminants, especially those with low solubility in water, and then migrate upward to the vadose zone to be captured by the SVE system.14-17

3. Combination of SVE with the bioventing technology.17-19 Bioventing uses a system configuration similar to SVE but with a different objective. In bioventing, the induced airflow aims to provide sufficient oxygen for the aerobic biodegradation of contaminants. It is thus possible to remove contaminants with relatively low volatility and high biodegradability.

4. Thermal enhancement of volatilization.19-21 Volatility of contaminants increases greatly with temperature, so several techniques have been developed to raise soil temperature, including the injection of hot air or steam, electrical resistance heating, and radio frequency heating.

Soil vapor extraction has become a very popular technology since the mid-1990s, because it has several important advantages:

1. It is an in situ technology and can even be applied below existing buildings, roads, and so on, thus causing minor disturbance to ongoing site operations.

2. The whole installation may be achieved using low-cost and easily available equipment, and the operation of the system is quite simple.

3. Although it is focused on the treatment of volatile contaminants trapped in the vadose zone, SVE can be integrated easily with other technologies targeting the remediation of groundwater or less volatile compounds, and this flexibility enables the application of the technology to a broad range of sites.

14.2.2 Design Considerations

The most important parameters for the preliminary design of an SVE system are the VOC concentration in the extracted air, the air flow rate, and the radius of influence of each extraction well. These parameters determine the number of wells that must be installed to remediate the whole contaminated area, the time required to obtain the cleanup goals, the size and characteristics of the gas treatment facility and auxiliary equipment, and finally the cost of the whole remediation project.

The design of SVE systems can be based on relatively simple mathematical models that describe the two basic phenomena governing the performance of SVE technology: the phase distribution of the organic contaminants and the characteristics of the airflow in the vadose zone.11-13,22,23 A simplified modeling approach, providing valuable tools for preliminary design calculations, will be presented in the following sections.

14.2.2.1 Phase Distribution of Organic Contaminants in the Vadose Zone

Organic contaminants can be present in the vadose zone in four distinct phases (Figure 14.2):

1. As an immiscible organic liquid retained by capillary forces in the pore space between the soil particles. This free organic phase is often referred to with the abbreviation NAPL (nonaqueous phase liquid).

2. As dissolved compounds in soil pore water.

3. As an adsorbed film on the surface of soil particles.

4. As vapor in soil air present in the pore space.

The distribution of a contaminant among the four phases depends on (1) the physical and chemical properties of the compound and (2) the characteristics of the soil, and can be described by relatively simple equations (see Table 14.3).

Absorbed on soil particles air in soil moisture

Dissolved

Free NAPL

Absorbed on soil particles

SOLID

AQUEOUS

Sorption Four-phase model

SOLID

AQUEOUS

Sorption Four-phase model

Phases of contaminants present in soil matrix in soil moisture

Dissolved

Sorption Three-phase model

FIGURE 14.2 Phase distribution of organic contaminants in the vadose zone. The solid arrows in the three-and four-phase models represent the equilibria taken into consideration in the equations of Table 14.3.

TABLE 14.3

Basic Equations and Required Data for Calculating the Phase Distribution of Contaminants under Equilibrium Conditions

In the presence of NAPL in the soil matrix

Without NAPL in the soil matrix

Ct = pbCs + ®wCw + + ®orCor

(1)

Ct = pbCs + 9wCw + 9aCa

(7)

Ca = P° x X x 7 x MWAR7)

(2)

Ca = KH X Cw

(8)

Co = m x por x 106

(3)

Cs = Koc x foc x Cw

(5)

Cw = C x X x 7

(4)

0t = 9w + 9a

(9)

Cs = Koc x foc x Cw

(5)

et = ew + ea + e„

(6)

Phase distribution values for calculation

Cs = adsorbed concentration of contaminant in the soil particle (mg/kg) Cw = dissolved concentration in pore water (mg/L) Ca = vapor concentration in pore air (mg/L) Cor = concentration of contaminant in NAPL (mg/L) 8a = pore volume occupied by the gas phase (L/L) 8or = pore volume occupied by NAPL (L/L) Required data

Contaminant propertiesa

MW = molecular weight (g/mol)

P° = vapor pressure of the compound (mmHg)

CW = water solubility (mg/L)

KH = Henry's constant (dimensionless)

Koc = organic carbon partitioning coefficient (L/kg)

Soil characteristicsb pb = soil bulk density (kg/L)

0w = pore volume occupied by water (L/L)

foc = fraction of organic carbon in soil

Contamination data0

Ct = total quantity of contaminant per unit soil volume (mg/L) por = specific density of the NAPL mixture (kg/L) m = mass fraction of contaminant in the NAPL mixture X = moles of contaminant in the NAPL mixture = activity coefficient of contaminant in NAPL

aData for these properties and for a long list of organic compounds can be found in several environmental engineering textbooks and handbooks.910,24-27

bSoil characteristics and data related to the concentration levels and the composition of organic contaminants should be collected during the investigation of the specific contaminated site.

When a single organic compound is present in the soil matrix as NAPL, its concentration in soil air (Ca) can be directly calculated from the vapor pressure of this compound (P° ) and the Ideal Gas Law:

where MW is the molecular weight of the compound, R is the ideal gas constant, and T is the absolute temperature. For a mixture of compounds, such as gasoline, the partial pressure (P) of each constituent i in the soil air depends on the composition of the mixture according to Raoult's Law:

where P° is the vapor pressure of the pure constituent, % is the mole fraction of the constituent, and Yi is the activity coefficient, representing the deviation from the properties of an ideal mixture.

Temperature has a strong influence on the vapor pressure of the contaminants. This effect can be described by the Clausius-Clapeyron equation:

where P is the vapor pressure at T, P ° the vapor pressure at To, and X is the molar heat of vaporization.

In the presence of NAPL, the concentration of contaminants in the soil moisture (Cw) can be calculated simply from the solubility of the compounds (equation 3 in Table 14.3). Adsorption of contaminants to the soil particles is a much more complex phenomenon, which depends both on contaminant properties and on soil characteristics. The simplest model for describing adsorption is based on the observation that organic compounds are preferentially bound to the organic matter of soil, and the following linear equation is proposed for calculating the adsorbed concentration (Cs):

where Koc is the organic carbon partitioning coefficient of the contaminant and foc is the fraction of organic carbon in the soil.

When the SVE technology is applied in a contaminated site, the NAPL is gradually removed. Towards the end of the remediation and when NAPL is no longer present, a three-phase model should be considered to calculate the phase distribution of contaminants (see Table 14.3). In this case, the vapor concentration in pore air (Ca) is calculating using the Henry's Law equation (Equation 14.5), which describes the equilibrium established between gas and aqueous phases:

where KH is the Henry's Law constant of the contaminant. Note, however, that during this phase the process is often governed by nonequilibrium rate-limiting conditions.

14.2.2.2 Basic Airflow Equations

The movement of air in the subsurface during the application of SVE is caused by the pressure gradient that is applied in the extraction wells. The lower pressure inside the well, generated by a vacuum blower or pump, causes the soil air to move toward the well. Three basic equations are required to describe this airflow: the mass balance of soil air, the flow equation due to the pressure gradient, and the Ideal Gas Law.

The mass balance of soil air may be described by the classic continuity equation for compressible fluids:

3Rl = p(pA) + d(PaUy) + d(PaU) dt 1 dx dy dz where 0a is the pore volume occupied by the gas phase, pa is the density of air, which is not constant due to air compressibility, and ux is the air velocity in the x-direction.

For a radial flow from a circumference of radius r toward the well, Equation 14.6 may be simplified as follows:

a dt dr

The air velocity due to the pressure gradient can be described by Darcy's Law:

m dr where K is the intrinsic permeability of soil, which is independent of the fluid properties, |J.a is the viscosity of air, and dP/dr is the pressure gradient in the radial r direction.

Finally, the Ideal Gas Law can be used to describe the relationship between air density and pressure:

Ka RT

where MW is the molecular weight of air, R is the Ideal Gas Law constant, and T is the absolute temperature.

Combining Equations 14.7-14.9, a differential equation, with pressure as the single variable, can be derived:

a dt ma dr

Under steady-state conditions, equation 14.10 has a simple analytical solution, which allows the calculation of the pressure Pr at several radial distances from the well:

where Pw is the pressure at the extraction well, Rw is the radius of the well, R1 is the radius of influence of the well, and P1 is the pressure at distance R1.

14.2.2.3 Radius of Influence and Number of Wells

Equation 14.11 introduces the notion of radius of influence, which is one of the important design parameters of SVE systems. Theoretically, the maximum radius of influence of a well is the distance at which the pressure becomes equal to the ambient atmospheric pressure, i.e., P: = Patm. In practice, RI is determined as the distance at which a sufficient level of vacuum still exists to induce airflow, e.g., 1% of the vacuum in the extraction well.912 The extraction wells are usually constructed using pipes with a standard radius, e.g., Rw = 5.1 cm (2 in.) or 10.2 cm (4 in.), and the vacuum applied in the wells typically ranges from 0.05 to 0.15 atm, i.e., Pw = 0.95-0.85 atm.912 If the vacuum required in the radius of influence is 1% of the vacuum in the extraction well, the

FIGURE 14.3 Determination of the required number of wells from the radius of influence.

corresponding PI values will range from 0.9985 to 0.9995 atm. The radius of influence RI is usually determined with preliminary field tests. Vacuum is applied in a test extraction well and the pressure Pr is measured in a monitoring well, installed at a distance r from the well. In practice, pressure drawdown is monitored at two or three points at varying radial distances from the well. Using the field-test data and Equation 14.11, it is possible to determine the radius of influence of the well RI at various operating vacuum values Pw.

Once the radius of influence has been determined, the number of wells Nwells required to remediate the entire contaminated area can be calculated from Equation 14.12:

where Acontam is the surface area corresponding to the contaminated zone. The factor 1.2 is arbitrarily chosen to account for the overlapping of the areas of influence between the wells and the fact that peripheral wells may reach outside the contaminated zone (Figure 14.3).12

14.2.2.4 Air Flow Rates

The flow rate of extracted air can be determined by considering the air velocity, as determined by Darcy's Law (Equation 14.8), and the radial distribution of pressure (Equation 14.11). The solution for air velocity as a function of the radial distance is given in Equation 14.13:

Pw

K

[r x ln(RwlRi)]

1 -

r p ^

2

x

p

w

Using Equation 14.13, one can easily calculate the volumetric flow rate Qw of the air extracted from the well:

ma ln(RjRi)

where uw is the velocity at the wellbore and H is the thickness of the vadose zone though which air is removed. The volumetric flow rate Qw corresponds to the pressure Pw of the well. To convert this flow rate to equivalent standard conditions, the following relationship can be applied:

Patm

It is obvious from Equation 14.14 that the most important parameter determining the volumetric air flow rate Qw is the intrinsic permeability K of soil. At this point it is important to stress the difference between water permeability (or hydraulic conductivity) kw, air permeability ka, and intrinsic permeability K. In most cases, when permeability data are provided for a type of soil or geological formation, these data are based on hydraulic conductivity measurements and describe how easily the water can flow through this formation. However, the flow characteristic of a fluid depends greatly on its properties, e.g., density p and viscosity p,. Equation 14.16 describes the relationship between permeability coefficient k and fluid properties p and k = K ¥ ^^, (14.16)

m where K is the geometric or intrinsic permeability of the soil, which depends only on the geometric characteristics of the soil (e.g., particle size distribution), and g is the gravity acceleration constant (g = 9.81 m/s2). Note that water and air permeability coefficients have units of velocity (cm/s), but K has units of surface (cm2).

When the hydraulic conductivity kw of a soil is known, one can easily estimate the corresponding values of intrinsic and air permeabilities, taking into consideration the properties of water and air under usual environmental conditions: e.g., pw = 1.0 g/cm3, = 1 x 10~2 g/(cms), pa = 1.2 x 10~3 g/cm3, and 14 = 1.83 x 10 4 g/(cms) (T = 20°C, P = 1 atm). For instance, a soil with hydraulic conductivity kw = 1 x 10 3 cm/s has an approximate intrinsic permeability of K = 1 x 10"8 cm2, and its permeability to airflow under normal conditions will be ka = 6.6 x 10 5 cm/s.

The airflow equations presented above are based on the assumption that the soil is a spatially homogeneous porous medium with constant intrinsic permeability. However, in most sites, the vadose zone is heterogeneous. For this reason, design calculations are rarely based on previous hydraulic conductivity measurements. One of the objectives of preliminary field testing is to collect data for the reliable estimation of permeability in the contaminated zone. The field tests include measurements of air flow rates at the extraction well, which are combined with the vacuum monitoring data at several distances to obtain a more accurate estimation of air permeability at the particular site.

14.2.2.5 Removal Rate of Contaminants and Required Cleanup Time

The contaminants removal rate Rrem can be calculated by multiplying the flow rate of air extracted from all the wells by the concentration of contaminant in the soil air Ca:

The required cleanup time Tclean is directly related to the removal rate:

where Mspill is the estimated total amount of spill.

Equations 14.17 and 14.18 are very simple, but the accuracy of the predictions depends greatly on the realistic estimation of Ca, which varies with time during the operation of the SVE system. For the start of the SVE project and considering that the free organic phase, NAPL, is present in the subsurface, a first approximation is to calculate Ca from the vapor pressure data of the contaminants (equation 2 in Table 14.3 or Equation 14.1). The actual concentration, however, will be lower than this value for two main reasons: (1) the extracted airstream does not pass only through the contaminated zone and (2) limitations on mass transfer exist. An effectiveness factor n should be considered to take into account the effect of these phenomena on removal rates. The value of this factor can be determined by comparing the calculated concentration with data obtained from the preliminary pilot tests at the site:

Ca,equil where Ca,field is the concentration in extracted air measured during the field tests and Ca,equil is the value calculated from the vapor pressure data.

Practical experience from the application of SVE at sites contaminated with a single type of contaminant (e.g., trichloroethylene, TCE) indicates that the removal of contaminants follows a trend in two distinct phases. During the initial phase, which covers the period from the project startup to the exhaustion of NAPL in the subsurface, the removal rate is almost linear. The second phase is characterized by a constant decrease in removal rates.

This trend can be explained with the following mechanism. In the presence of NAPL, the extracted vapor concentration depends mainly on the vapor pressure of the contaminant. After the disappearance of free NAPL, the extracted vapor concentration becomes dependent on the partitioning of contaminants among the three other phases (see Table 14.3). As the air passes through the pores, the dissolved contaminants volatilize from the soil moisture to the gas phase, causing the desorption of contaminants from the surface of soil particles into the aqueous phase. As a result, the concentration in all three phases decreases, with a consequent decrease in removal rates.

For the initial linear phase of remediation, the pilot test data and Equations 14.17 to 14.19 can provide relatively good predictions for the required cleanup time. For the second phase, it is necessary to use more sophisticated models combining airflow, equilibrium, and mass transfer Equations 14.13 to 14.16, in order to obtain sufficiently accurate predictions. To obtain a first rough estimation, the methodology proposed by Kuo12 can be applied. Kuo's approach is based on the observation that the VOC concentrations of extracted air decrease exponentially with time during the second stage of remediation. To simulate the exponential decrease in removal rates, the following procedure is suggested:

1. The mass of contaminant that must be removed during the second stage is divided into two or three equal parts, corresponding to successive cleanup time intervals.

2. Initial Cai and final Caf vapor concentrations are calculated for each interval using the phase distribution equations in the absence of NAPL (see Table 14.3).

3. A mean vapor concentration Cam representing each time interval is determined from the geometric average of the two concentrations:

4. The successive cleanup time intervals are calculated using the mean concentration values, and they are summed to determine the total required time.

This procedure is illustrated in the practical example presented in Section 14.2.2.6.

Note that the initial linear phase is observed only in sites containing a single contaminant. For sites contaminated with mixtures of contaminants there is a decreasing rate of removal from the beginning of the project due to the different volatility of the components. The more volatile constituents are extracted with a higher rate from all the phases, and as a consequence the total VOC content of extracted air decreases constantly with time. This effect should be considered during the design phase.

14.2.2.6 A Practical Example

A tank containing 20 m3 toluene ruptures, contaminating an area of 1250 m2 in the vadose zone with an average depth of 4 m. The soil in the subsurface has the following characteristics: bulk density pb = 1.7 g/cm3, total porosity 0t = 0.4, moisture corresponding to porosity 0w = 0.2, and organic carbon content foc = 0.01.

Owing to the high volatility of toluene (vapor pressure P o = 22 mmHg), the decision was to use SVE technology. Preliminary field tests were conducted in the area using an extraction well with Rw = 5.1 cm (2 in.) and total perforated length inside the contaminated zone H = 4 m. The tests were carried out applying a vacuum of 0.1 atm (i.e., Pw = 0.9 atm) in the well, and the pressure was measured at a distance of 6 m and found to be 0.99 atm after reaching steady-state conditions. The flow rate of extracted air, as measured in the exhaust of the vacuum pump, was Qw = 0.2 m3/min, and the air contained 78 mg/L toluene. The temperature of the subsurface was 25°C.

To determine some important design parameters for this SVE project, the following procedure could be used:

Step 1: Obtain the physicochemical data of the compound of concern. Important sources for this type of data are references 9, 10, and 24-27. From tables included in Reference 9, the following properties of toluene were obtained: MW = 92.14 g/mol, P o = 22 mmHg = 0.0289 atm, KH = 0.276 (dimensionless), C = 490 mg/L, log Koc = 2.06, and por = 0.866 g/cm3.

Step 2: Calculate the initial distribution of toluene in the subsurface. The initial distribution of toluene can be calculated using equations 1 to 5 from Table 14.3 and taking into consideration that the organic phase is a pure compound, i.e., X = 1, m = 1, and y = 1. The total quantity of contaminant per unit soil Ct can be estimated from the known amount of spill Mspill and the volume of the contaminated zone:

Ct = Mspill = 2°,°00L x 0.8663kg/L = 17,320kg = 3^kg/m3 = 3464mg/L.

The concentrations and the mass distribution of toluene in the four phases, as calculated from this set of equations, are presented in Table 14.4. As seen in the table, the major part of the toluene, i.e., 68.9%, remains in the vadose zone as free NAPL, 27.6% is adsorbed on the surfaces of solid particles, and only 3.5% is distributed between the aqueous and gas phases. Free NAPL occupies only a small part of the available pore volume, and it is not expected to disturb the movement of air through the contaminated zone.

Step 3: Calculate from the field test data the radius of influence, the required number of wells, and the required capacity of the gas treatment facility. The radius of influence R: can be calculated from equation 14.11 using the pressure monitoring data at r = 6 m while considering that the minimum required vacuum at R: should be 0.001 atm, i.e., P: = 0.999 atm. With these values R: is found to be 9.91 m. The number of wells is calculated from equation 14.12, Nwells = 4.86. This means that five wells must be installed to remediate the entire contaminated area. Once the number of wells has been determined, the required capacity of the gas treatment facility can be defined from the flow rate data obtained during the field tests. In this case, the gas treatment unit should be able to treat Nwells x Qw = 5 x 0.2 = 1.0 m3/min of toluene-laden air.

TABLE 14.4

Concentrations, Mass Distribution of Toluene, and Volume Occupied by the Four Phases in the Vadose Zone

Toluene concentrations Toluene mass distribution Volume of four phases

Total

3464 mg/L 17.32 t 5000 m3

Free NAPL

Aqueous

986 m3

Solids

563 mg/kg 4.78 t 3000 m3

The flow rate data can also be used to estimate the permeability of the subsurface. The required additional parameter is the value of air viscosity, i.e., |j.a = 1.83 x 10 4 g/(cm s). The intrinsic permeability of soil is calculated from equations 14.16 and 14.15 and is found to be K = 1.34 x 10 8 cm2. Care should be taken to perform the appropriate unit conversions when using Equation 14.15.

Step 4: Estimate the effectiveness factor ^ for the removal and the cleanup time required to obtain a residual toluene concentration of 150 mg/L. The phase distribution calculations carried out in Step 2 indicate that the equilibrium concentration of toluene in the gas phase is Ca,equil = 109 mg/L (see Table 14.4). The concentration measured in the extracted air during the field tests is lower, at Ca,field = 78 mg/L, indicating that the removal effectiveness is limited either as a result of mass transfer phenomena or the existence of uncontaminated zones in the airflow pattern. The corresponding effectiveness factor is n = 78/109 = 0.716.

The amount of toluene that must be removed from the soil Mrem can be calculated by considering the initial total amount of spill Mspill and the residual acceptable quantity corresponding to the cleanup objectives Mfinal:

Mrem = Mspill - Mfinal = 17.32 t - (150 g/m3) x 5000 m3 x (10-6 t/g) = (17.32 - 0.75) t = 16.57 t.

The removal of toluene is assumed to take place in two stages. The first stage corresponds to the removal of free NAPL, which, according to the phase distribution calculations (Step 2; Table 14.4) represents a mass of Mrem1 = 11.94 t. The second stage corresponds to the removal of toluene, which is distributed among the other three phases, and represents a mass of Mrem2 = 16.57 - 11.94 = 4.63 t.

As this site is contaminated with a single compound, the removal of free NAPL is expected to follow a linear trend with constant removal rate. The required time can be calculated from Equations 14.17 and 14.18:

1 h¥ NWen ¥ qw ¥ CWil 0.716 ¥ 5 ¥ (0.2m3/min) ¥(109g/m3) ¥ (1440d/min) .

The second stage of treatment is assumed to follow an exponential decrease in removal rates. Applying the approach of Kuo, this stage is divided into two time intervals, T2-1 and T2-2, representing the successive removal of equivalent amounts of toluene, Mrem2-1 = Mrem2-2 = 2.315 t. The initial theoretical concentration in the gas phase for the time interval T2-1 is equal to the vapor pressure of toluene, Cai = 109 mg/L. The final vapor concentration for this interval Ca,f can be calculated from the total residual concentration Ct,f and the phase distribution equations 5 and 7-9 in Table 14.3:

,, (Mspill - Mem1 - Mrem2-1) (17.32 -11.94 - 2.315)t n.„¥ln-3t/ 3 „

tf 5000m3 5000m3

The mean vapor concentration Ca,m for the time interval T2-1 is calculated from the geometric average of Cai and Ca,f (Equation 14.20), i.e., Ca,m = 91.4 mg/L, and the required treatment time from Equations 14.17 and 14.18:

2-1 Nwell x Qw x Caequil 0.714 x 5 x (0.2 m3/min) x (91.4 g/m3) x (1440 d/min)

The same procedure is applied for the last time interval T2-2, and the following values are calculated:

Cai = 76.7 mg/L, Caf = 18.8 mg/L, Cam = 37.9 mg/L, T2-2 = 59.4 d The total cleanup time, as estimated with this approach, will be

Tclean = T + T%x + TM = 106.5 + 24.6 + 59.4 = 190.5 d

As seen from these calculations, the removal of free NAPL, representing almost 70% of the total toluene spill, takes approximately 106 days. The operation of the SVE system should continue for an additional 84 days in order to achieve the cleanup objectives and remove the final 30% of the toluene spill.

14.3 BIOREMEDIATION 14.3.1 Introduction

The bioremediation techniques exploit the biological activity of microorganisms to degrade or detoxify environmentally hazardous compounds. Traditionally, biological treatment has been applied for the remediation of sites contaminated with organic contaminants. Most organic compounds can be degraded through the action of appropriate microbial communities towards more simple and less harmful inorganic or organic molecules. The degree of degradation determines whether mineralization or biotransformation has occurred. Mineralization is the complete degradation of organic compounds into inorganic final products, such as carbon dioxide and water, whereas biotransformation is the partial degradation of the compound to more simple organic molecules.

Bioremediation is not restricted only to biodegradable organic contaminants. New techniques are currently under development for the bioremediation of metal-contaminated sites. Microbial activity can alter the oxidation state of some elements, reducing or increasing their mobility, and this transformation can be used for remediation purposes.

Bioremediation systems in operation today rely on microorganisms indigenous to contaminated sites. The two main approaches, based on the actions of native microbial communities, are biostimulation and intrinsic bioremediation. In biostimulation, the activity of native microbes is encouraged, creating (in situ or ex situ) the optimum environmental conditions and supplying nutrients and other chemicals essential for their metabolism. The vast majority of bioremediation projects are based on this biostimulation approach. Intrinsic bioremediation is a remedial option that can be applied when there is strong evidence that biodegradation will occur naturally over time without any external stimulation; i.e., a capable microbial community exists at the site, the required nutrients are available, and the environmental conditions are favorable. An additional prerequisite is that the naturally occurring rate of biodégradation is faster than the rate of contaminant migration towards sensitive environmental receivers, e.g., a well used for abstraction of drinking water. In that case, and if sufficient supportive data are provided, the regulatory authorities may issue a permit to pursue the intrinsic bioremediation option for a particular site. This remediation strategy is not a "no action" alternative. It requires the design and implementation of a systematic monitoring procedure to follow closely the progress of this natural process and prevent any undesirable side effects, such as the generation of toxic bioproducts due to unexpected changes in redox conditions.

In some cases the indigenous microorganisms are not able to degrade or detoxify the specific contaminants to acceptable levels. The use of nonnative microbes or even genetically engineered microorganisms especially suited to degrading the contaminants of concern is another bioremediation option known as bioaugmentation, that is currently under development. An important research effort has been devoted since the mid-1990s to discover microbial species capable of destroying or detoxifying specific hazardous pollutants, and to isolate them in pure cultures in order to exploit their efficiency in bioremediation projects. Such pure specific degrading microbial populations have been successfully used for the treatment of contaminated soils under laboratory conditions, but to date there are no known cases of full-scale projects applying the bioaugmentation principle.

Regardless of whether the microbes are native or artificially introduced into the soil, it is important to understand the mechanisms by which they degrade or detoxify hazardous pollutants through their metabolic activity. Understanding these mechanisms is essential for the proper design of bioremediation systems that provide the optimum conditions and the required nutritional supplements for the specific microbial process.

14.3.2 Principles of Bioremediation 14.3.2.1 Basic Microbial Metabolism

The microbial degradation of organic contaminants occurs because the organisms can use the pollutants for their own growth and maintenance. A microbial cell operates two critical types of metabolic processes, referred to as anabolic (cell-building) and catabolic (energy-releasing) processes. Anabolic processes involve the production of new cells and require a source of carbon, which is the most important constituent of cellular mass. Catabolic processes are energy-producing chemical reactions and require a source of energy.

Organic contaminants are used by microorganisms both as a source of carbon and as a source of energy. The microbes gain energy from the contaminants through their oxidation, which involves the breaking of chemical bonds and transfer of electrons away from the contaminant. To complete the chemical reaction, another compound is needed to receive the electrons. The contaminant, which is oxidized, is called the electron donor and the chemical, which is reduced, is called the electron acceptor. The microorganisms use the energy produced from these electron transfers to build new cells or simply to maintain the existing cells. The electron donor and the electron acceptor are essential for cell growth and maintenance and are commonly called the primary substrates.

Depending on the type of electron acceptor, the metabolic modes are broadly classified into three main categories: aerobic respiration, anaerobic respiration, and fermentation. Aerobic respiration is the term used to describe the metabolism in which molecular oxygen (O2) serves as the electron acceptor. Many microorganisms follow the mode of aerobic respiration, and most bioremediation projects exploit this particular type of metabolism. There is, however, a wide variety of microorganisms that are able to survive and grow under anaerobic conditions using several inorganic or organic compounds other than oxygen as electron acceptors. This form of metabolism is called anaerobic respiration. The most commonly used electron acceptors under anaerobic conditions are nitrates (NO3) and sulfates (SO4), which are soluble constituents in the aqueous phase, and the oxidized forms of iron (Fe[III]) and manganese (Mn[IV]), which are common constituents of soil particles, mainly in the form of oxides. A type of metabolism that can play an important role under strictly

TABLE 14.5

Typical Benzene Biodegradation Reactions under Various Electron Acceptor and Redox Conditions

Indicative Redox Electron

TABLE 14.5

Typical Benzene Biodegradation Reactions under Various Electron Acceptor and Redox Conditions

Indicative Redox Electron

Conditions, Eh

Acceptors

Biodegradation Reactions

Refs.

> +200 mV

O2

CA, H

h 7.5 O2 ^ 6 CO2 + 3 H2O

[28]

< +200 mV

NO3-

QH, H

h 6 NO3T + 6 H+ ^ 6 CO2 + 3 N2 + 6 H2O

[29]

< 0 mV

Fe(III)

CA H

h 30 Fe3+ + 12 H2O ^ 6 CO2 + 30 Fe2+ + 30 H+

[30]

< -100 mV

SO42-

CA H

h 3.75 SO4- + 7.5 H+ ^6 CO2 + 3.75 H2S + 3 H2O

[31]

< -200 mV

QHs

CA h

h 12 H2O ^ 2.25 CO2 + 3.75 CH4

[32]

anaerobic conditions is fermentation. During fermentation there is no need for an external electron acceptor, because the organic contaminant serves as both electron donor and electron acceptor.

The typical biodegradation reactions under various electron acceptor conditions are presented in Table 14.5 for the simple case of benzene. Which type of electron acceptor will be used is closely related to the prevailing redox conditions. Under aerobic conditions, with redox potential greater than 200 to 220 mV, biodegradation is mainly performed by aerobic microorganisms. When oxygen is depleted but the redox potential remains relatively high, biodegradation can proceed through the metabolic activity of nitrate-reducing bacteria. The Fe(III) oxides of soil can be used as electron acceptors over a wide range of redox values, depending upon their crystallinity. Finally, sulfate-reducing and methanogenic bacteria are active only under strongly reducing conditions.

14.3.2.2 Co-Metabolism

In some cases, microorganisms can transform a contaminant, but they are not able to use this compound as a source of energy or carbon. This biotransformation is often called co-metabolism. In co-metabolism, the transformation of the compound is an incidental reaction catalyzed by enzymes, which are involved in the normal microbial metabolism.33 A well-known example of co-metabolism is the degradation of (TCE) by methanotrophic bacteria, a group of bacteria that use methane as their source of carbon and energy. When metabolizing methane, methanotrophs produce the enzyme methane monooxygenase, which catalyzes the oxidation of TCE and other chlorinated aliphatics

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