Natural Attenuation of Contaminants in Soils

Natural Attenuation of Contaminants in Soil Natural Attenuation of Contaminants in Soil Raymond N. Yong Catherine N. ...

Author: Raymond N. Yong | Catherine N. Mulligan

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Natural Attenuation of Contaminants in Soil

Natural Attenuation of Contaminants in Soil Raymond N. Yong Catherine N. Mulligan

LEWIS PUBLISHERS A CRC Press Company Boca Raton London New York Washington, D.C.

This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.”

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CHAPTER 1 Natural Attenuation and Ground Contamination 1.1 INTRODUCTION Ground contamination by various forms of pollutants is a fact of life in this highly industrialized age. Some of the major sources of pollutants are indiscriminate, inadvertent and uncontrolled discharge of liquid wastes, poor management and control of chemical and other processes, inadequate and improper liquid waste and runoff collection systems, disposal and improper storage or containment of hazardous wastes, runoff from nonpoint source pollution surfaces, and spills and illicit dumping. It has become clear that costs for remediation of contaminated sites can reach such proportions that remediation with available conventional or even innovative intervention techniques cannot be afforded. Russell et al. (1991) suggest that over the next three decades, the anticipated expenditure required to clean up the sites containing hazardous wastes may be in the order of between $370 billion and $1.7 trillion. The 1993 U.S. Environmental Protection Agency (U.S. EPA, 1999) report estimated that the number of contaminated sites, including those on the national priority list, totaled about 330,000. One can only speculate on the hundreds of billions of dollars required to remediate all of these sites. It is small wonder that one would want to ﬁnd methods for cleanup of contaminated sites that would not require such huge expenditures. Accordingly, attention is being directed to the use of the natural attenuation capability of soils as a possible remediation technique. The threats posed by pollutants in contaminated ground are normally considered in relation to their impact on the health of biotic receptors and the environment. To minimize or eliminate the threats, removal of the pollutants is an obvious course of action. The conventional thinking is that removing the sources of the threats would eliminate the threats. Several factors are to be considered. These can be grouped into two broad categories: • Land use requirements, intended use of remediated land (decontaminated site) and permissible use • Cost effectiveness of physical or chemical removal of pollutants from the contaminated site to meet land use requirements, intended use or permissible use

1

2

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Experience has shown that complete removal of pollutants from a contaminated site cannot always be accomplished without considerable expense. In many cases, complete removal of pollutants cannot be achieved at all. Several options present themselves if one recognizes that total removal of pollutants is not always required — so long as one can eliminate the health and environmental threats posed by these pollutants. These include: • Immobilizing the pollutants so they stay ﬁxed to the soil particles (soil’s solids) and cannot migrate or diffuse into the surroundings • Ensuring that the immobilized pollutants do not produce leachates harmful to the environment and biotic receptors • Reducing the concentration of volume and mass of the pollutants • Reducing the toxicity level of the pollutants • Eliminating the pollutant path in the source-receptor-path (SRP) chain of communication

The net result of these actions will establish (1) avoidance of groundwater contamination, (2) minimization of pollutant migration, (3) elimination of interaction of pollutants (source) and biotic receptors and (4) rehabilitation of the contaminated land to a level consistent with land use requirements. 1.1.1

Natural Attenuation

In the latter part of this chapter and in the next few chapters, we will be developing the various aspects of the soil–pollutant interaction mechanisms that contribute to the attenuation of contaminants and pollutants. We ﬁrst consider the concept or property of natural attenuation in the context of its role in the management of ground pollution — i.e., pollutants in a contaminated site. Reduction of toxicity and concentration of pollutants to minimal and acceptable levels are goals that must be achieved in remediation of contaminated land. The assimilative capability of soils and the various reactive processes in soils have the potential to achieve such goals. The natural attenuation of contaminants, as a process in soils, can be a useful tool in the management of ground pollution. However, it is recognized that the processes and reactions in soils that produce or result in attenuation of contaminants and pollutants differ among soils. These processes and reactions may even differ among different locations of the same soil. The capability for natural attenuation of contaminants and pollutants (i.e., natural attenuation capacity) of soils varies from one soil to another. As we will see in the discussions to follow, a signiﬁcant factor in the development of the natural attenuation capacity of soils comes from the composition of the soil and the immediate environment, i.e., the soil-water system as a geoenvironmental feature. The knowledge that many soils have the ability to naturally attenuate contaminants has led to at least four structured approaches to the utilization of this natural attenuation capacity (see Chapter 8 for detailed discussion of these approaches), including:

NATURAL ATTENUATION AND GROUND CONTAMINATION

3

•

Development and use of soil-engineered barrier systems that exploit and enhance attenuation capacity. These can take the form of clay-liner and clay-barrier systems and reactive treatment walls placed in the ground to intercept pollution plumes. One could call this approach engineered natural attenuation (EngNA). • Utilizing the natural attenuation capacity of soils as a remedial tool to manage the transport of pollution plumes. This approach is quite often perceived as a passive remediation technique and is generally identiﬁed as the monitored natural attenuation (MNA) procedure for remediation of soils. • Enhancing the natural attenuation capacity of the soils by biostimulation or bioaugmentation (see Chapter 8). This is sometimes called enhanced natural attenuation (ENA). • Using the natural attenuation capacity of soils as part of a treatment procedure. A good example of this use is in the ﬁnal cleanup treatment in a treatment train process where other remediation techniques are initially used to remove the major portions of the pollutant load. The overall remediation process is sometimes identiﬁed as treatment trains, or layered treatments.

The increased awareness and acceptance of the natural attenuation capacity of soils has reached the point where many jurisdictions and regulatory agencies are adopting (or have adopted) this concept of passive remediation for remediation of contaminated land. It follows that a necessary requirement for adoption of this passive remediation technique is that the soil possesses the qualities that will interact with the pollutants to allow this type of technique to be applied. To ensure that the process is effective, monitoring guidelines and criteria have been (or are being) established. The complete process of MNA is now a standard application procedure. To properly consider the MNA approach to remediation of contaminated land, we need to examine the pros and cons. The pros consist of two principal items — decreased cost of remediation and minimal physical intrusion on the contaminated site. The cons also consist of two major items — time requirements for the attenuation treatment to be effective and the long-term monitoring program with all its associated tests on recovered samples. For management purposes, it is useful to construct an analytical computer prediction model that would permit one to determine whether the MNA process is meeting set objectives and targets. Chapter 8 discusses the detailed protocols associated with implementation of MNA and the other adaptations of the natural attenuation process. Modeling of remediation by the MNA process requires identiﬁcation and understanding of the processes that participate in and control contaminant attenuation in the MNA remediation scheme. To better understand how the process works, and to better appreciate what conditions render the process effective, it is necessary to understand how the interactions between pollutants and soil solids, and their reactions, lead to attenuation. To do so, we need to understand the mechanisms and processes associated with chemical mass transfer (of pollutants), and also the reactions associated with the interactions of microorganisms and organic chemicals. These mechanisms and processes are examined in some detail in other chapters. We start, however, by looking at the land environment and land use, since these are the bounding parameters.

4

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

1.2 LAND USE Concerns regarding the preservation of the quality and integrity of the environment are generally directed toward the forces and activities that can degrade the environment. This book focuses on the land environment, and in particular on the sustainability of land use. We consider land as a resource base. Maintenance and enhancement of effective use of the land and its ability to provide the resources and environment for habitation are essential if the resource base is to be preserved. This chapter provides a general introduction to the many problems and issues arising as a result of ground contamination. The details of the interaction processes between soil particles and contaminants (pollutants), and other aspects of concern are dealt with in chapters 3–7.

1.2.1

Land Suitability and Use

When we begin to examine the issues of land environmental impact, it is useful to consider land as the basic component of a land environment. We deﬁne land to include the physical natural landforms and receiving waters contained therein, together with the underlying geology. The major concerns in respect to a land environment relate directly to “what deﬁnes a sustainable environmental resource base,” as articulated in the 1992 Rio Declaration on Environment and Development (U.N. Department of Public Information, 1993). In essence, the land environment is a major resource base upon which activities instituted by humans will provide the means to meet the needs of the global population. The major elements that deﬁne the land in the context given above are contained within the biosphere. Two major groups are identiﬁed as follows: • Terrain features and factors: linear features, topography, physical attributes, watershed, local hydrology, receiving waters, water, surface layer and vegetative cover • Interface and subsurface: subsurface soil and soil-water system, underlying geology, groundwater and aquifer regime

Types of land use range from natural virgin forests to grazing and cultivated lands to highly developed lands designed to meet urbanization and infrastructure requirements. The extent of human intervention in changing original land suitability for purposeful land use depends on: • Original land conditions and suitability, i.e., site-speciﬁc issues regarding land capabilities • Demand, planning and socioeconomic factors • Requirements established to satisfy environment and land resource protection • Development practices and technological capabilities

In spite of the extent and intent of human intervention, purposeful land use will be conditioned by initial land suitability.

NATURAL ATTENUATION AND GROUND CONTAMINATION

1.2.2

5

Land Use, Ground Contamination and Sustainable Development

A prominent factor in the degradation of land use is ground contamination by pollutants. The presence of pollutants in contaminated sites poses serious threats to human health, the immediate ecosystem and the environment. There is a great need to avoid ground contamination and to remediate and rehabilitate contaminated ground. To that extent, development of any piece of land must conform to practices that would ensure protection and preservation of the land environment and its natural resources. This is the essence of the geoenvironmental compartment of sustainable development. We recognize that in order to meet the needs of an ever increasing world population there will be increased needs for an adequate supply of goods and services to feed, shelter and clothe the population. The 27 principles articulated in the 1992 Rio Declaration show the need for protection and maintenance of environmental quality while meeting the needs of the global population. These principles were reinforced in the 2002 World Summit on Sustainable Development. Principles 1, 3 and 4 of the Rio Declaration state that • “Human beings are at the centre of concerns for sustainable development. They are entitled to a healthy and productive life in harmony with nature.” (Principle 1) • “The right to development must be fulﬁlled so as to equitably meet developmental and environmental needs of present and future generations.” (Principle 3) • “In order to achieve sustainable development, environmental protection shall constitute an integral part of the development process and cannot be considered in isolation from it.” (Principle 4)

At the heart of sustainable development is a healthy and sustainable environment. To properly address the problems and issues of ground contamination and progressive advance of pollutant plumes, the geoenvironmental compartment of sustainable development requires a knowledge of the linkages between humans and a healthy, robust and sustainable land environment. Figure 1.1 shows these linkages and identiﬁes some of the major issues and land environment impacts. The following observations regarding the major land environment issues shown in the ﬁgure are pertinent: • Waste generation and pollution: Wastes generated from the production of goods and services will ultimately ﬁnd their way into one of three disposal media: (1) receiving waters, (2) atmosphere or (3) land. Land disposal appears to be the most popular method for waste containment and management. Its impacts include ground contamination by pollutants. • Depletion of agricultural lands: This will arise because of increased urbanization and industrialization pressures, infrastructure development, exploitation of natural resources and use of intensive agricultural practices. The loss of agricultural lands places greater emphasis and requirement on higher productivity per unit of agricultural land. The end result of this is the development of high-yield agricultural practices. One of the notable effects is soil quality loss. To combat this, there is an inclination to use more soil amendments and other means to enhance yield. A

6

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Geoenvironmental Compartment of SUSTAINABLE DEVELOPMENT Anthropogenic activities associated with Urbanization, Industrialization, Production and Exploitation for Production of Goods and Services (Food, Shelter, Clothing)

Some Major Land Environment Issues Waste generation and pollution Depletion of agricultural lands Increased demand on natural resources – water, ocean, forestry and mining assets Greenhouse gases and climate change Desertification and watershed depletion Photosynthesis, biological magnification

Figure 1.1

Some Major Impacts from Urbanization, Industrialization, Exploitation and HighYield AgriculturalPractices Wastes, pollution, discharges, runoffs, soil -quality loss, groundwater and aquifer deterioration and surface water quality emissions, non-point source pollution, surface hydrology changes, erosion, watershed loss, wastepiles, tailings ponds, acid mine drainage, land use deterioration

Some major land environment issues and impacts resulting from activities associated with urbanization, industrialization, production and exploitation for production of goods and services.

resultant land environment impact from such practices is run-offs and nonpoint source pollution. • Increased demand on natural resources: The issues related to exploitation activities fall into the categories of (1) land and surface degradation associated with mining and forestry activities and (2) water supply and utilization. Surface hydrology changes, erosion, watershed loss, tailings and sludge ponds, acid mine drainage, etc. are some of the many land environment impacts. • Greenhouse gases, climate change desertiﬁcation: To a very large extent, these are consequences of industrialization, urbanization and production. Their impact on the land environment can be felt for example in acid rain (and snow) interaction with soil and undesirable changes in photosynthesis processes and erosion of coastal areas owing to increasing water levels. • Photosynthesis and biological magniﬁcation: The processes associated with photosynthesis are important since they produce about 20% of the available oxygen in the atmosphere. Desertiﬁcation, deforestation and many of the activities associated with mineral and other natural resource exploitation degrade the capability of land and aquatic plants to engage in these processes. Biological magniﬁcation, which concerns the concentration of toxic elements or pollutants by plants and such biotic receptors as aquatic organisms and animals, is a problem that needs to be addressed when considering the containment and management of pollutants.

Figure 1.1 shows that if a healthy and robust land environment is considered to be at the heart of the geoenvironmental compartment of sustainable development, land environment impacts need to be mitigated. Ground and groundwater contami-

NATURAL ATTENUATION AND GROUND CONTAMINATION

7

nation from, for example, wastes, pollutant discharges, runoffs, acid mine drainage, emissions and landﬁlls are signiﬁcant problems. The goals and objectives of sustainable development insofar as the geoenvironmental compartment is concerned are designed to ensure that these issues are properly addressed and that the impacts are mitigated. At the center of many of the issues and impacts is ground contamination by pollutants.

1.3 CONTAMINATED GROUND A contaminated piece of ground is generally identiﬁed as a contaminated site. This designation normally indicates a piece of ground that has been found to contain toxic substances that can cause serious injury to the immediate environment and its biotic receptors. These toxic substances are most often referred to as pollutants. Ground contamination by pollutants commonly occurs in connection with anthropogenic activities such as those shown in Figure 1.1, i.e., activities associated with the range of industries involved with the production of goods and services — including municipal and domestic services. Inadvertent spills, liquid and solid waste discharges and waste containment and management strategies are some sources of pollutants. While one might sometimes consider industrial and municipal solid waste substances to be nonhazardous or toxic in their solid form, the same cannot be conﬁdently said for the leachates derived from these solids. It is these leachates that have the ability to move in the ground (soil substrata) and it is these leachates that contain both pollutants and nonpollutants (contaminants) that pose threats to the environment and human health. The term contaminant, which is used in general discussions of the transport of dissolved solutes contained in the leachate, is explained in greater detail in the next section. The need to avoid ground contamination is obvious and is a signiﬁcant consideration in the management of municipal and industrial waste facilities. We must nevertheless accept that: • There are countless contaminated sites in existence. • Waste landﬁlls will continue to constitute part of the waste disposal equation. • Nonpoint source pollution and inadvertent and ill-advised operations will continue to exist.

Several techniques are available to manage and control contaminant and pollution plumes to minimize environmental and health impacts. These include construction of impermeable barriers and liner systems for containment facilities, remediation techniques designed to remove or reduce (attenuate) the pollutants in the ground, and passive procedures relying on the properties of the ground to reduce contaminant concentrations in leachate streams and pollution plumes. Since greater attention is being paid to natural attenuation, it is useful to examine its basic essentials and the fundamental processes involved. With a better understanding and appreciation of these processes, we can, if needed, increase the effectiveness of this passive technique with some very simple procedures (as developed in Chapter 8).

8

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

1.3.1

Control of Ground Contamination through Attenuation

The reduction of toxicity and concentration during contaminant plume and contaminated groundwater transport in soils is generally identiﬁed as contaminant attenuation or, when used in the context of contaminants and pollutants, simply attenuation. In essence, with the exception of toxicity reduction, contaminant attenuation is the result of the assimilative action of soils. That is, the assimilative action of soils includes physical and physico-chemical partitioning mechanisms and chemical and biologically mediated reactions. The actions of microorganisms leading to toxicity reduction, for example biotransformation, do not ﬁt in the category of assimilative capacity of soil. Strictly speaking, toxicity reduction should be considered a separate category. The amount of contaminant assimilation is often referred to as the assimilative capacity of a soil. This capacity is a direct function of the contaminant-soil interactions, which include a combination of physical, chemical and biologically mediated reactions in the soil. The nature and extent of the reactions are dependent on the physical and chemical properties of both contaminants and soil-water systems. Traditionally, attenuation has been considered almost exclusively in connection with remediation of contaminated sites. More recently, however, attenuation has been considered as part of a design tool to be used in the management of leachate plumes and controlled disposal facilities. In addition to the processes occurring within the soil-water system, dilution of contaminant concentrations by groundwater or external water sources also contributes to the attenuation of contaminants. We can therefore consider the many interactions and processes between contaminants, soil solids and soil water that contribute to attenuation to include not only dilution but also the combination of physical, chemical and biologically mediated reactions and combinations thereof. These interactions and processes that provide the soil with its attenuation capability or potential for contaminant attenuation are some of the principal considerations in: 1. The design of engineered barrier systems 2. Management and control of contaminating leachate plumes 3. Remediation and management of contaminated sites

We can view contaminant attenuation as the net result of the sum of processes occurring in the soil-water system that reduce the toxicity and concentration of contaminants in transport through or interaction with these soils. These processes, known as attenuating processes, include: 1. Abiotic and biotic reactions, transformations and degradations 2. Partitioning via sorption, volatilization and accumulation 3. Dispersion and dilution

The ﬁrst two are processes that occur as a direct result of interactions between soil solids, pore water, pore space and contaminants and interactions within the pore water itself. Strictly speaking, the third process is not an interaction process between soil solids and contaminants, since this is the result of mixing of the contaminants

NATURAL ATTENUATION AND GROUND CONTAMINATION

9

in the leachate plume with groundwater. A few key points are worthy of note in consideration of the overall process of attenuation of contaminants: • The attenuation processes that result in the decrease of contaminant concentrations are processes that should be applicable to the gamut of contaminants and soils. • The quantiﬁcation of attenuation, often expressed as the degree of attenuation, is contaminant speciﬁc, i.e., speciﬁc to a particular contaminant or a suite of contaminants. • The degree of attenuation of a speciﬁc contaminant or suite of contaminants is in turn speciﬁc to a particular soil, i.e., speciﬁc to the subsurface soil within which transport of the leachate plume occurs. • It is more than likely that the same degree of attenuation of a particular contaminant or suite of contaminants in a speciﬁc soil type will not be obtained if the soil type is changed. The degree of attenuation may be lesser or greater. • For proper matching between contaminants and soil composition to ensure optimum attenuation, we need to develop a better understanding of the many attenuation processes that are the outcome of interactions between contaminants, soil solids and the host environment.

1.3.2

Pollutants, Contaminants, Groundwater and Pore Water

Since the terms pollutants, contaminants, groundwater and pore water will be used throughout this book, it would be useful to deﬁne them and to indicate how these are used in this book. We deﬁne contaminants to mean those substances (solutes, chemicals, etc.) that are not part of the initial composition of a natural soil material. These are generally introduced into the soil as a result of regional and environmental factors and anthropogenic activities. A simple example of regional and environmental processes contributing to contamination of a soil is the leaching of salts from the surface layers into the lower soil layers. By and large, we will use the term contaminants when the generic sense is deemed to be more appropriate. The term pollutant is used to mean a contaminant that has been identiﬁed as a threat to human health and the environment because of its nature — as opposed to its concentration. The toxic chemicals or compounds listed in the U.S. EPA Priority Pollutants List are good examples of pollutants associated with industrial activities that have been found in contaminated sites. These include chlorinated benzenes and ethanes, chlorinated phenols and naphthalenes, pesticides, polychlorinated biphenyls (PCBs) etc. Materialization of threat to human health and the environment can occur through direct and indirect pathways as shown in Figure 1.2. We need to consider the SRP concept in evaluating the level of threat to biotic receptors. Physical contact, exposure and inhalation are good examples of direct threats to humans. An example of an indirect threat is the contaminated food chain. The term pollutants will more often be used in our discussions for the speciﬁc cases under discussion and to emphasize the issues being considered. It is not uncommon to ﬁnd the terms groundwater and pore water used interchangeably in the technical literature. However, because we will be concerned with interactions between contaminants and soil solids occurring in the water within the pores of the soil in our discussion of the attenuation processes, it will be useful to

10

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Inh

ala

Atmosphere

Evap

De lan pos it d an ion d w on at er

tion

and

dire

ct c ont a

ct

orati on

Up

ta

ke

Aquatic and terrestrial biota

t Indirecct conta stion … … inge ct ta on c on ct sti re e i D ing d an

Contaminated ground with pollutant plume

AQUIFER Figure 1.2

Schematic showing contaminated ground with a pollutant plume as the potential source of health and environmental threats. Pollutants carried into the atmosphere by evaporation and deposited on land and on receiving waters are also potential threats via air, land and water pathways.

differentiate between these two terms. Strictly speaking, while all the water contained in the soil pore spaces is groundwater, it is convenient to consider that from a mechanistic point of view this water (i.e., groundwater) consists of both bulk water and pore water. In a simplistic sense, this mechanistic view separates the water in the pores into two parts (see Chapter 3): (1) pore water (Gouy region deﬁning the diffuse ion layers), in which the interaction processes between dissolved solutes and the active surfaces of the soil solids occur — contributing to contaminant attenuation and (2) bulk water, which consists of water in the regions close to and beyond the boundaries of the diffuse double layers, i.e., in the regions where Gouy forces are minor or nonexistent. We should emphasize that there is no clear demarcation between bulk water and pore water and that the boundary separating the two is a diffuse region.

1.4 POLLUTANTS AND GROUND CONTAMINATION Ground contamination occurs as a result of (1) natural processes such as runoff and leaching of surface layer contaminants into the subsoil layers and (2) anthropogenic activities associated with industrialization, urbanization and exploitation resulting for example in waste generation and discharge of waste streams. The latter is normally assumed to be the greater source of pollutants in contaminated ground.

NATURAL ATTENUATION AND GROUND CONTAMINATION

11

The activities associated with manufacturing, processing and storage, in relation to the industries that provide goods and services, are largely responsible for the major sources of pollutants found in contaminated sites. Other sources of pollutants in contaminated ground come from leachates emanating from landﬁlls (Figure 1.3), unregulated dumping and discharges, leaking storage tanks, application of biocides and fertilizers, to name a few. For example, metal extraction, reﬁning and production generate signiﬁcant amounts of waste. Processes such as steel pickling, metal degreasing and ﬁnishing, anodizing, and galvanization will discharge, respectively, spent ferrous pickle liquor, alkaline cleaning agents, acid strip solutions, spent caustic baths and spent chromic and sulfuric acids. These waste streams are sources of toxic chemicals and substances that will be regarded as pollutants when they ﬁnd their way into the soil as improper or inadvertent discharges. Table 1.1 shows various kinds of pollutants obtained in the industrial waste streams such as the example of metal manufacture and production discussed above. The table illustrate the typical types of pollutants found in respect to manufacture, processing and production of various goods. Instead of seeking a more detailed or comprehensive listing of the various inorganic and organic pollutants generated, it is more instructive to see what kinds of pollutants have been found in contaminated sites. Some of the more common pollutants found in many contaminated sites reported in the literature and in ﬁeld reports include heavy metals (compounds of lead, mercury, copper, cadmium and chromium), petroleum hydrocarbons (PHCs), halogenated and nonhalogenated organic chemicals and solvents and PCBs. An indication of some of the sources of contaminants found in U.S. Superfund sites is shown in Figure 1.3. The common types of pollutants found in contaminated ground can be divided into two groups: 1. Heavy metals (HMs): These are elements that easily lose electrons to form positive ions, and although there are 39 elements classiﬁed as HMs, the majority of these are not found in signiﬁcant quantities in contaminated ground. As noted in Table 1.1, the HMs are generally obtained as waste streams from mining, ore reﬁning, metal producing and electroplating industries. The common ones that have been reported in contaminated ground include cadmium, chromium, copper, iron, lead, mercury, nickel, silver, tin and zinc. They can exist in their elemental forms or as compounds in the pore water or attached (partitioned) to the soil solids (see Chapter 4 for a detailed discussion). The interaction mechanisms leading to partitioning of HMs and their subsequent desorption need to be carefully identiﬁed when one seeks to determine the attenuation of these pollutants. 2. Organic chemicals and compounds: Activities associated with industrial production and use of petroleum products, solvents, pesticides and explosives (Table 1.1) are some of the major sources of organic chemicals and compounds found in contaminated sites. Production of industrial intermediates using aliphatic and aromatic compounds can also be included in the above. The chemicals and compounds listed in the Toxicity Characteristics Leaching Procedure (TCLP), a test based on the EPA method 1311 with regulatory levels limit for characterization of a chemical as toxic are good examples of the chemicals of concern. These chemicals include benzene, carbon tetrachloride, chlordane, chlorobenzene, ocresol, m-cresol, r-cresol, 1,4-dichlorobezene, 1,2-dichloroethane, 1,1 dichloro-

12

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Waste Landfill

Pollutant source (spill)

Pollutant plume

Leachate (pollutant) plume

Underground storage tank

Flow direction

BEDROCK

Figure 1.3

Schematic showing (1) leachate plume with pollutants, (2) pollutant plume from a leaking underground storage tank and (3) a surface pollutant source (spill).

ethylene, 2,4 dinitrotoluene, endrin, heptachlor, hexachlorobenzene, hexachloro1,3-butadiene, hexachloroethane, lindane, methoxychlor, methyl ethyl ketone, nitrobenzene, pyridine, tetrathcloroethylene, toxaphene, trichloroethylene, 2,4,5trichlorophenol, 2,4,6-trichlorophenol, silvex and vinyl chloride. Chapter 5 addresses the transport and fate of many of these chemicals.

1.5 CONTROL AND MANAGEMENT OF POLLUTION PLUMES The transport of pollutants in the soil substrate, whether by movement of pollutant-containing leachates or as pollution plumes emanating from a pollution source (Figure 1.3), must be controlled and managed to ensure that the pollutants do not become serious threats to the environment and biotic receptors. Two particular points need to be stressed: • SRP: Identiﬁcation of the source and receptor are obvious requirements. However, if there is no pathway from source to receptor, one can argue that no threat exists. • Mobility of pollutants: SRP notwithstanding, if the pollutants are not mobile their threat will be minimal.

In the case of the waste landﬁll shown in the top left part of Figure 1.3, the pollution plume emanating from the bottom of the landﬁll is assumed to be due to leakage because of an inadequate or damaged containment liner system, i.e., whatever system has been used to contain the leachate within the landﬁll has not proven

NATURAL ATTENUATION AND GROUND CONTAMINATION

13

Table 1.1 Typical Pollutants Obtained from Various Industrial Waste Streams Sources for Major Waste Streams

Typical Industries

Typical Pollutants

Metal Production: Industries and Users Electronic industry; metal manufacture and goods; metallurgical activities; machinery production and use

Steel pickling; metal ﬁnishing; metal cleaning; metal surface treatment; galvanizing; plating; anodizing; metal working and production; smelting

Spent acids (sulfuric, chromic, pickle liquor, etc.); spent alkalis; spent reagents; metal sludges; metal residues; solvents

Chemical Production: Industries and Users Pharmaceuticals Production of chemicals and feedstock, e.g., inks, plastics, petrochemicals, explosives, pesticides, insecticides, fungicides, soaps, detergents, petrochemicals, paints and varnishes, glues, etc.

Chemical manufacture and synthesis; pigments; ﬁne chemicals manufacture; pharmaceuticals manufacture; petroleum reﬁning; production of chemical intermediates, phenols, plastics, rubbers; etc.

Spent acids and alkalis; metal sludges; chlorinated solvents; aromatic amines; nitrated phenylamines; chlorinated hydrocarbons; halogenated and nonhalogenated solvents; petroleum hydrocarbons; etc.

Agriculture, Pulp and Paper Products, Printing Agriculture and animal husbandry; horticulture; paper and cardboard production; printing and photography

Application of fertilizers, fungicides, herbicides; food processing; livestock wastes; photographic and printing processes

Phenolics, carbamates, benzoic acids, organochloride compounds, phenoxyacids, triazines, substituted ureas, etc.; nonhalogenated solvents, methanol, acetone, various glycols, ketones and acetates, cyclohexane, xylene, etc.

Petroleum, Gas, Consumer Goods: Production Petroleum, oil and gas extraction and production; textile and clothing industry; leather and woodworking industries

Gas works; oil and petroleum reﬁning; wood preservatives; textiles and leather production

Spent caustic; acid tars; heavy metal sludges; oil sludges; distillation residues; halogenated and nonhalogenated solvents; etc.

to be effective or secure. The transport of the pollutants in the soil substratum for any of the three examples shown in the ﬁgure occurs as an expansion or advance of the pollution plumes. The net result is a spreading of the threat to the environment and human health. 1.5.1

Engineered Barrier-Liner Systems

For effective prevention of pollutant transport through the soil, engineered containment barriers that line waste impoundment facilities are popular options. These

14

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

engineered containment barriers serve to (1) offer a competent impermeable barrier between the wastes contained within the impoundment facility and the surrounding ground and (2) provide a means for attenuating the contaminants passing through the barrier as a designed leachate control and management feature. The design, construction and management of such types of waste disposal landﬁll systems can be complex, depending on the nature of the waste being contained. An effective leachate collection system is vital to the successful management of waste disposal landﬁll facilities. Figure 1.4 shows a waste landﬁll containment system with a typical cover system and two kinds of bottom engineered barrier-liner systems. The dimensions of the layers of the barrier-liner system will vary according to the type of waste being contained, underlying geology and hydrogeology, local conditions and regulations, and requirements mandated by the regulatory authorities. Engineered barrier-liner systems are useful for containment of leachates — so long as the barriers can function according to design requirements. For waste disposal in landﬁlls, engineered barrier-liner systems are the basic building blocks of a successful safe and secure landﬁll. This method of control and management of pollutant plumes does not allow the formation of the kind of pollutant plume shown in the top left portion of Figure 1.3 since the leachates are contained within the landﬁll facility. There are situations, however, where the barrier-liner system has been designed to act as an attenuating barrier, as shown for example in the right-hand bottom liner system in Figure 1.4. When such is the case, the barrier is generally an engineered clay barrier designed to attenuate the contaminants in leachate transport through the clay barrier (see Chapter 8). The requirements in regard to the properties and characteristics of the clay material used in the barrier are no different in principle from those generally used to evaluate the natural attenuation capacity of soils. The signiﬁcant difference between engineered attenuating systems and systems that rely on natural attenuation is the capability of adding or subtracting the various elements considered relevant to the attenuation process. The subject of landﬁlls and containment systems, which has been well treated in dedicated textbooks, is outside the purview of this book. 1.5.2

Constructed and Emplaced Barriers

In situations where leachates, pollutant plumes and pollution sources already exist, as for example in the cases shown in Figure 1.3, construction of engineered barrier-liner systems may not be the most expedient or economic remedy against the spread of pollution plumes. Other than the engineered barrier-liner systems that are design requirements for waste landﬁlls, techniques using isolation barriers, for example, separation walls constructed with sheet piles or similar techniques, have been used (Figure 1.5). By and large, isolation-separation barriers that isolate the pollutants from further transport into the surrounding ground are used (1) as a secure but temporary containment to await future remediation treatment of the contained pollutants or (2) to provide a secure remediation treatment facility such as a pumpand-treat procedure or a contained bioreactor system (see Mulligan, 2002). Attenuation barrier systems inserted into the ground to intercept further progression of pollution plumes (right-hand portion of Figure 1.5) rely on the properties of

NATURAL ATTENUATION AND GROUND CONTAMINATION

Figure 1.4

15

Waste landﬁll system showing typical top cover and bottom barrier-liner systems.

the barrier material to attenuate the contaminants through partitioning processes and transformations (Yong, 2000). The underlying principles of such barrier systems, called permeable reactive walls or barriers, are similar to those demonstrated in the natural attenuation capacity of soils (see Chapter 8). 1.5.3

Dilution, Retardation, Retention and Attenuation

We mentioned in Section 1.3.1 that the attenuation of contaminants and pollutants, owing to the assimilative processes of soils, refers to their reduction during transport in soils. We further indicated that reduction of concentrations can be accomplished by (1) dilution from mixing with uncontaminated groundwater, (2) interactions and reactions between contaminants and soil solids that can lead to partitioning of the contaminants between the soil solids and pore water and (3) transformations that reduce the toxicity threat posed by the original pollutants. Short of overwhelming dilution with groundwater, it is generally acknowledged that partitioning is by far the more significant factor in attenuation of contaminants or pollutants. While the details of the processes leading to partitioning will be discussed in Chapters 4 and 5, we ﬁrst need to focus on the broader issues of temporary and permanent partitioning of contaminants. Figure 1.6 shows a schematic of various pollutant pulses emanating from the source shown on the left-hand side of the drawing. The three mechanisms (dilution, retardation and retention) are all assumed to be solely responsible for pollutant attenuation. Beginning with the dilution pulse, by assuming a homogeneous and uniform dilution of the pollutants in the original

16

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Contaminated Site

Permeable reactive barrier to attenuate contaminants in plume flowing through the barrier

Impermeable isolation containment barrier system created to contain pollutants within the contaminated site (a)

Figure 1.5

Impermeable Clay Stratum

(b)

Schematic diagram showing two examples of control of pollutants in soil. (a) Containment of the pollutants in the contaminated site using an impermeable wall surrounding the pollution plume. (b) An intercepting permeable reactive wall placed so that the advance of the pollution plume into and through the wall will result in attenuation of the pollutants.

pollutant pulse load, a rectangular pulse shape is obtained. Arguably, this is a simplistic assumption. To illustrate the signiﬁcance of distinguishing between the three mechanisms of pollutant attenuation, we consider the target distance x from the pollutant source (Figure 1.6) to represent a hypothetical regulatory decision point. A not uncommon regulatory control standard or criterion would require that the concentration of pollutants passing or reaching this control target distance x must not exceed a speciﬁed concentration — generally identiﬁed as a critical concentration or target concentration. This criterion says nothing about the total load delivered to this point over some time interval. Another kind of control requirement could specify that the total pollutant load delivered to point x or reaching a distance x away from the source over a speciﬁed time interval should not exceed some critical limit. It is this control requirement that will place a demand on attenuation mechanisms that require permanent partitioning of the pollutants during transport through the soil substratum. This limiting pollutant load is important when the point of delivery (target or receptor) of pollutants is a water resource or an environmentally sensitive piece of land. The simplistic approach adopted for pollutant attenuation through dilution assumes that dilution of the original pollutant load will ultimately deliver the total pollutant load over some undeﬁned time period. Similarly, since retardation mechanisms are thought to rely on temporary reversible sorption processes and on physical

Pollutant pulse load (area defines total pollutant load)

Retardation Pulse

17

Target distance x

NATURAL ATTENUATION AND GROUND CONTAMINATION

Pollutant concentration

Retention pulse

Dilution pulse Source Distance from source

Figure 1.6

Schematic showing the progress of the various pollutant pulses. Note that the areas of the dilution and retardation pulses are constant and that they are equal to the original pollutant pulse load. Because of retention mechanisms, the retention pulses will decrease in size as one progresses farther from the pollutant source.

constraints to retard movement of the pollutants, the areas in the retardation pulse shapes shown in Figure 1.6 are similar to the original rectangular shape shown on the left. If the sorption mechanisms do not permit easy desorption or removal of the sorbed pollutants, retention of the pollutants occurs. Under such circumstances, the mechanism of retention as an attenuating tool would provide for decreasing concentrations of pollutants as transport of the pulse continues away from the source, as shown in Figure 1.6. The likelihood of only one mechanism’s being solely responsible for attenuation of pollutants in transport in the soil is very remote. Probably all mechanisms will participate — to varying degrees — in the attenuation of pollutants. It is obvious, however, that the greatest beneﬁt obtained from the attenuating capacity of soils is when irreversible sorption, e.g., retention, is the primary mechanism of attenuation. For such retention to be obtained, it is necessary to understand the various partitioning processes and the role of soil composition and structure in interactions with the pollutants.

1.6 NATURAL ATTENUATION AND REGULATORY ATTITUDES Natural attenuation refers to the situation when attenuation of contaminants results because of the processes that contribute to the natural assimilative capacity

18

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

of soil. This means that contaminant attenuation occurs as a result of the natural processes occurring in the soil during the contaminant-soil interaction. Broadly speaking, therefore, natural attenuation refers to natural processes occurring in the soil that reduce the toxicity or concentration of the contaminants. Therefore, natural attenuation will always be italicized when used in the context of processes. NA will be used for the remediation treatment option. The term contaminants is used instead of pollutants because the attenuation processes apply to a broad range of contaminants and not solely to pollutants. These natural processes of contaminant attenuation include dilution, partitioning of contaminants and transformations. They involve a range of physical actions, chemical and biologically mediated reactions and combinations of all of these. Monitored natural attenuation (MNA) has been deﬁned in U.S. EPA-SAB (2001) as “a remediation approach based on understanding and quantitatively documenting naturally occurring processes at a contaminated site that protect humans and ecological receptors from unacceptable risks of exposure to hazardous contaminants.” The report further indicates that this is “a knowledge-based remedy because instead of imposing active controls, as in engineered remedies, scientiﬁc and engineering knowledge is used to understand and document naturally occurring processes to clearly establish a causal link.” MNA, because of its adherence to “remedy by natural processes,” necessitates an understanding of the principles involved in the natural processes that contribute to the end result. Monitoring of the pollutant plume at various positions away from the source is a key element of the use of MNA. Remembering that MNA is a pollutant- and soil-speciﬁc phenomenon, one generally tracks a very limited number of pollutants, speciﬁcally the ones considered to be the most noxious. Historically, more attention has been paid to documenting the properties and characteristics of the pollutants. The pollutants tracked have primarily been the organic chemicals including chlorinated solvents (perchlorethylene (PCE), dichloroethylene (DCE) and trichloroethylene (TCE)) and hydrocarbons such as benzene, toluene, ethyl benzene and xylene (BTEX). However, more attention is now being paid to the nature of the soils involved — not only from the viewpoint of initial soil composition and potential interactions, but also from the perspective of the role of soil fractions. We need to learn about the role of the various soil fractions in relation to processes involving electron transfer and partitioning of contaminants. These aspects are discussed in Chapters 4 and 5. When active controls or agents are introduced into the soil to render attenuation more effective, this is called enhanced natural attenuation. This is to be distinguished from engineered natural attenuation (EngNA), which is probably best illustrated by the permeable reactive barrier shown in Figure 1.5 and the barrier-liner system shown as the bottom right of Figure 1.4. Enhanced natural attenuation (ENA) refers to the situation where, for example, nutrient packages are added to the soil system to permit enhanced biodegradation to occur or where catalysts are added to the soil to permit chemical reactions to occur more effectively. ENA could include biostimulation and bioaugmentation. These aspects will be discussed in greater detail in Chapter 8. Apart from the considerable interest by regulatory agencies in application of MNA to remediation of contaminated sites (U.S. EPA-SAB, 2001; NRC, 2000; U.S.

NATURAL ATTENUATION AND GROUND CONTAMINATION

19

EPA, 1999), the use of MNA, ENA and EngNA for management and control of leachate plumes is now being seriously considered for adoption by many regulatory agencies. In particular, with the use of ENA and EngNA, regulatory agencies can adopt performance criteria and control over construction of landﬁlls, remediation efﬁciencies of contaminated sites, discharge scenarios, etc. Speciﬁcation of limiting pollutant concentrations at various distances from pollution sources, as discussed previously, together with compulsory monitoring at those positions provide regulatory agencies with the tools to track performance and adherence to mandated guidelines. 1.6.1

Monitored Natural Attenuation, Protocols and Management Requirements

Since MNA may not yet have been accepted by many regulatory agencies and jurisdictions, implementation of the concept as a remedial tool requires one to adhere to a strict set of protocols. These protocols, which are generally issued by an oversight body (e.g., regulatory commission), are designed to ensure de minimus practice requirements for protection of, as Yong says, “the public good and the environment” at all times. They are especially essential if one chooses to implement remediation of contaminated sites using the MNA procedures. The protocols are designed to require detailed information on the nature and properties of the pollutants — in addition to the development of a good understanding of the mechanisms contributing to the natural attenuation capacity of the soil. Chapter 8 provides the details of the kinds of protocols for MNA and the various elements needed to satisfy the lines of evidence required in these protocols. An integral part of the protocols is the implementation of long-term monitoring of the subsurface environment in the contaminated site and in the region surrounding the site. Placement of monitoring wells in the contaminated zone and downstream of the pollution plume will ensure capture of information regarding the nature of the pollutants in the pollution plume. When one combines the monitoring information with (1) previously obtained detailed site characterization, (2) a good understanding of the processes that contribute to the attenuation properties of the soil and (3) predictions of transport and fate of the pollutants from properly developed models, one will obtain a clear picture of the effectiveness of the attenuation process. The protocols for this process are described in Chapter 8. Modeling of the performance of the MNA process at the speciﬁc site of application is needed — not only to provide some indication of the effectiveness of the process, but also to inform one of the likely changes in the performance characteristics of the process. The models developed are essentially transport and fate models. They are intended to provide support for the lines of evidence. The models provide the evidence of performance needed to evaluate the effectiveness of the MNA process. Lines of evidence consist of several parts. One part could be in the form of corroborating laboratory and bench-type treatability tests. Another part could include information on the hydrogeology of the setting. Other parts could deal with the performance of pollutants and the composition and properties of the pollutants at various locations and times.

20

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Detailed protocols and guidelines are generally issued by regulatory or oversight agencies to cover general and speciﬁc pollutants. These detailed protocols, which can be site speciﬁc or general, are designed to cover the essential requirements, procedures, standards, criteria, etc. that need to be used in determining the lines of evidence and the effectiveness of the MNA process. Since insufﬁcient attention is paid to the role of the soil-water system — particularly in regard to the various soil compositional factors and the role of soil in attenuating the pollutants — it is useful to observe that the attenuation process requires two partners: (1) pollutants and (2) the soil-water system. How they interact to successfully promote attenuation depends on the actions of both partners. Paying attention only to one partner (pollutants) denies one the capability to fully exploit the process. The keys to successful use of the natural attenuation potential of soils are: • Compatibility between contaminants and soil composition: Proper matching of the contaminants and soils is required to obtain the necessary reactions to obtain (1) effective and optimum partitioning of the contaminants with the soil solids and (2) transformations and alterations that will reduce toxicity of contaminants. • Speciﬁcation of target concentrations or critical concentrations: Without knowledge of target or critical concentrations, i.e., limits beyond which serious threats to the environment and biotic receptors occur, it would not be possible to determine whether attenuation would be a viable tool for pollution management. These limits are obviously pollutant or contaminant speciﬁc and location speciﬁc. • Proper monitoring system: The need for judicious placement of monitoring wells to monitor leachate quality and capture the necessary samples for testing is obvious. Wells upstream, within the contaminated zone and downstream of the contaminant plume will be necessary. • Continuous analysis of samples retrieved from monitoring stations in a welldeveloped monitoring protocol: This includes quality testing of both pore water and soil solids to determine partitioning and mobile concentration of pollutants and to determine natural attenuation rates and avoid impact on potential receptors.

Since MNA is dependent on information, i.e., it is a knowledge-based method of management of pollutant transport in the soil, the requirement for a complete understanding of what is happening is obvious. The same requirements are also obvious for ENA and EngNA. For ENA and EngNA, however, the “what is happening” part of the equation will provide one with the opportunity to implement further enhancements or changes for the ENA system, and for the EngNA, assessment of the capability of the engineered system can be easily made.

1.7 CONCLUDING REMARKS To answer the question of why one should want to, or need to, know the basic processes contributing to the attenuation of contaminants in soil, we have laid the basis for such a requirement on the need to maintain the quality of the land envi-

NATURAL ATTENUATION AND GROUND CONTAMINATION

21

ronment. The following points have been made in support of the need to obtain a better understanding of the attenuation processes in soil: Maintenance of land quality and land use: The concept of sustainable land use has been introduced. Sustainable land environment and sustainable development: The impacts resulting from development of our land environment for urbanization, industrialization, production and exploitation to meet the requirement for provision of food, shelter and clothing must be mitigated to ensure that we maintain a sustainable land environment. This recognizes the fact that land is a natural resource base. Assimilative capacity of soil: The assimilative capacity of soil for contaminants is a fundamental element in the natural attenuation of contaminants in soil. The contributing factors and interactions in the natural attenuation process include dilution and reactions in the soil that are physically, chemically and biologically mediated. Use of natural attenuation for management of pollutant transport and transmission in soil: Natural attenuation can be used to manage and control the transport of pollutants in soil in at least three ways. These include MNA, ENA and EngNA, which have various beneﬁts and are used as effective tools in the control and management of contaminant and pollutant leachate plumes, especially in speciﬁc bioremediation schemes.

REFERENCES Mulligan, C.N., Environmental Biotreatment, Government Institutes Inc., Rockville, MD, 2002, 395 pp. NRC (National Research Council), Natural Attenuation for Groundwater Remediation, Committee on Intrinsic Remediation, Water Science and Technology Board and Board on Radioactive Waste Management, Commission on Geosciences, Environment and Resources, National Academy Press, Washington, D.C., 2000, 274 pp. Russell, M., Colglazier, E.W., and English, M.R., Hazardous Waste Remediation: The Task Ahead, Waste Management Research and Education Institute, University of Tennessee, 1991. United Nations Department of Public Information, Agenda 21: Programme of Action for Sustainable Development, United Nations Publication, Sales No. E.93.1.11, 1993. U.S. EPA, Cleaning up the Nation’s Waste Sites: Markets and Technology Trends, U.S. EPA Ofﬁce of Solid Waste and Emergency Response, Washington, DC, 1991. U.S. EPA, Recent Developments for In Situ Treatment of Metal Contaminated Soils, EPA 542-R97-004, Ofﬁce of Solid Waste and Emergency Response, Washington, DC, 1997. U.S. EPA, NATO/CCMS Pilot Study: Evaluation of Demonstrated and Emerging Technologies for the Treatment and Clean Up of Contaminated Land and Groundwater (Phase III) EPA 542-R-99-008, 1999. U.S. EPA-SAB, Monitored Natural Attenuation: USEPA Research Programme: An EPA ofﬁce of the Administrator, Science Advisory Board Review, EPA-SAB-EEC-01-004, 2001. Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate, and Mitigation, CRC Press, Boca Raton, FL, 2000, 307 pp.

CHAPTER 2 Soil Composition and Transmission Properties 2.1 INTRODUCTION A soil-water system is a soil mass comprising soil solids and pore water contained therein. The nature of the soil solids and the chemistry of the pore water are distinct pieces of the overall material that combine to (1) form a soil mass and (2) provide the soil mass with its distinct properties and characteristics. The term soil-water system is used as a general term to denote a soil mass and more often in speciﬁc discussions concerning the interactions that involve the chemical nature of the pore water, the soil solids and inﬂuent contaminants. These will be discussed in greater detail in Chapter 3. For this chapter, we will be concerned with the nature of soil, and in particular, we will want to determine what constitutes a soil-water system and what endows the soil with the properties that are of importance in the transmission of ﬂuids through the soil. It is not uncommon to see the term soil used in the literature to mean a soil mass, and hence, a soil-water system (i.e., soil solids and pore water). In this book, we will also use the term soil to refer to a soil mass, and when it is necessary to emphasize the role of soil particles in interactions in the soil, we will use the term soil-water system. The subjects of interest in this chapter deal directly with the many features of a soil that inﬂuence the transport and partitioning of contaminants in soils. This requires us to develop an understanding of what it is that constitutes a soil. In particular, we need to appreciate how the various physical and physico-chemical properties and characteristics of soils impact the development of the assimilative capacity of soils. In essence, therefore, we should realize that the attenuation processes in soil involve direct interactions between the contaminants and soil solids and interactions with the dissolved solutes in the pore water. The inﬂuence of the nature of the soil-water system in developing contaminant attenuation needs to be determined. In particular, it would be very useful to know how soil composition and soil type inﬂuences contaminant attenuation. In other words, we need to determine those pertinent soil features, properties and characteristics that are important in

23

24

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

controlling of the interactions between the soil solids and contaminants that result in the attenuation of contaminants. 2.1.1

Soil Properties and Natural Attenuation

To illustrate the role of soil composition in the determination of natural attenuation of contaminants, one might conduct a simple laboratory experiment along the lines of an adsorption isotherm test (see Chapter 3) to determine whether the amount of contaminants retained by a soil would be different if the soil composition was changed. The simplest test would involve a soil suspension where the suspended soil particles (i.e., soil solids) were allowed to interact directly with a chosen contaminant. Figure 2.1 shows the adsorption test results for a soil suspension using an illitic clay soil consisting primarily of illite clay minerals with about 5.7% carbonates and 1.2% amorphous oxides. Other soil solids included small detectable proportions of chlorite, quartz, feldspar and calcite minerals. To avoid introducing factors such as competition between other contaminants, a single contaminant species, lead in the form of Pb(NO3)2, was used as the test contaminant. The adsorption test results for the natural illitic clay soil shown in Figure 2.1 are portrayed in terms of the concentration of Pb adsorbed or retained by the soil solids in relation to the concentration of Pb remaining in the aqueous phase of the soil solution (supernatant). The test involved suspending the soil solids in an aqueous phase with varying amounts of Pb. One could use, for example, a 1:20 proportion of soil solids:aqueous phase (by weight) as a soil solution. This allows one to assume that if complete dispersion of all the soil solids is obtained, then all the suspended and dispersed soil solids will have an equal opportunity to interact with the Pb. The bottom curve in the ﬁgure shows Pb adsorption by the natural illitic clay soil. By only removing the carbonate and amorphous oxides, while leaving all the other soil solids in suspension, we see from the ﬁgure that the Pb adsorption curve is signiﬁcantly higher, i.e., more Pb is adsorbed by the carbonate and oxide-free soil. The changes in the surface features of the suspended solids owing to removal of the carbonates and oxides have contributed to signiﬁcant changes in the retention characteristics of the carbonate and oxide-free soil. The later parts of this chapter and Chapter 3 will provide further details on the contributions of soil structure and surface properties on the retention characteristics of the soil solids. Continuing with the same type of experiment, one could retain the same soil suspension and vary the type of contaminant. The results shown in Figure 2.2 are for the natural illitic clay in separate interactions with equal amounts of Pb(NO3)2 and Cu(NO3)2. The results suggest that more Pb is retained than Cu. When mixtures of various contaminants are involved, the picture becomes more complicated. However, the message is not likely to be changed, i.e., the nature of the soil is important in the determination of how well and how much of any contaminant is retained in the soil-water system. These issues will be discussed in greater detail in this chapter and in the next chapter.

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

25

Retained Pb concentration (mmoles/100 g soil)

16 Pb(N03 )2 and illitic soil 14 Soil suspension pH ~ 5.0 12 10 Illite clay soil with carbonates and oxides removed

8 6 4

Illite clay soil

2 0 0

10

20

30

40

50

Equilibrium Pb concentration in supernatant (mmoles/L) Figure 2.1

Lead (Pb) retention by soil solids in a soil suspension test using a lead nitrate solution [Pb(NO3)2] as the source of contaminant Pb. Soil used in the soil suspension is an illitic soil with about 5.7% carbonates and about 1.2% amorphous oxides by weight. (Data from Yong, R.N. and MacDonald, E.M., Inﬂuence of pH, metal concentration, soil component removal on retention of Pb and Cu by an illitic soil, in Adsorption of Metals by Geomedia, Jenne, E.A., Ed., Academic Press, San Diego, CA, 1998, Ch. 10, pp. 230–254.)

2.2 NATURE OF SOIL The major sources of soils are rock and decomposed organic matter. Physical and chemical weathering of rock are two of the natural processes of disintegration of rock that ultimately result in the production of soil. Physical weathering of rocks involves processes that produce cracking of the rocks owing to heating and cooling since rock components have different coefﬁcients of expansion. Other physical weathering processes include expansion of water in rock ﬁssures as a result of freezing and abrasion owing to movement of ice and water. Much smaller particles are obtained as a result of the breakdown of the rocks or minerals. On the other hand, chemical weathering involves loss of some constituents, addition of protons and rearrangement of constituents into new materials. The rates of these processes are determined by temperature, moisture, composition, potential for leaching and composition of the leaching solution. Chemical processes involved include hydrolysis, hydration, carbonation, oxidation, reduction and solution. Protons in the leaching solution can be supplied from CO2 in the air or from decaying vegetation or from other acids added to the system.

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Retained Pb or Cu concentration (mmoles/100 g soil)

26

16 * Pb retention

14

* *

12 10

*

8

*

Cu retention

6 *

4 *

Illite clay soil with carbonates and oxides removed

2

Soil suspension pH ~ 5.0

0 0

10

20

30

40

50

Equilibrium Pb or Cu concentration in supernatant (mmoles/L)

Figure 2.2

Lead (Pb) or copper (Cu) retention by soil solids in a soil suspension test using either a lead nitrate solution [Pb(NO3)2] or a copper nitrate solution [Cu(NO3)2] as the source of contaminant Pb or Cu. Soil used in the soil suspension is an illitic soil with carbonates and amorphous oxides removed. (Adapted from Yong, R.N. and MacDonald, E.M., Inﬂuence of pH, metal concentration, soil component removal on retention of Pb and Cu by an illitic soil, in Adsorption of Metals by Geomedia, Jenne, E.A., Ed., Academic Press, San Diego, CA, 1998, Ch. 10, pp. 230–254.).)

We recognize that the ﬁrst requirement for soil formation is physical and/or chemical breakdown of parent rock. Following initial formation, weathering factors contribute signiﬁcantly to the formation of what we see as soils in place. We should also be aware that transport factors such as wind, water, and glacial activity move soils to different locations. Accordingly, we should take into account that these agents are also responsible for determining what we see as soils in any location. Regardless of whether initial soil development results from in situ processes of physical weathering or from transported means, subsequent weathering processes produce the ultimate product of soils in place. The weathering rate depends on such factors as: 1. Parent rock: texture and composition, particle sizes in the rock, permeability, solubility, intrinsic strength and exposed surface area. The extent of inﬂuence of parent rock on soil type is dependent on the microclimate and microenvironment. While, for example, composition and texture are important at the early stages of rock weathering, their inﬂuences are not long lasting in regions where high temperatures and high humidity prevail because of the high rates of reactions involving rock material and energy. As one enters into regions where the rates of reactions are lesser, the inﬂuences of composition and texture of the parent rock on weathering products are

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

27

longer lasting. It is useful to note that because of different microclimates and microenvironments, the same parent rock can produce different soil types. Alkali and alkaline contents in parent rocks are important determinants in weathering products. Rocks that contain alkalis and alkaline cations can produce a greater variety of weathered soil products than those that do not. In the latter instance, the weathering products include lateritic material and kaolinites. 2. Immediate environment and time: groundwater presence and position, topography, composition of groundwater, organic matter, oxygen and other gases, temperature, pressure, microclimate, and local environmental changes. As one would intuitively anticipate, both the amount of water available and heat are big factors in the chemical weathering of parent rock mass. Warm and humid climates are therefore very instrumental in producing correspondingly high weathering of the minerals in the parent rock mass. The presence and amount of organic matter are also very important factors in the control and rate of weathering. If organic matter is allowed to remain, as for example in cool environments, high humidity would produce conditions that would successfully promote reactions between the organic acids and parent rock. However, if the environment is changed to a hot humid environment, oxidation of the organic matter would preclude the formation of organic acids, resulting in little or no reaction between the organic matter and the parent rock.

2.2.1

Soil Material and Classiﬁcation

Soil material obtained solely through physical breakdown of rock will in essence reﬂect the composition of the parent rock. When chemical weathering occurs, the extent of the weathering process will determine the degree to which parent rock composition remains. In the natural sequence of time, it is inevitable that chemical weathering processes will follow any physical breakdown of parent rock — simply because of the fact that microclimatic and microenvironmental forces will act on the rock fragments. This is generally referred to as forces exercised as a result of regional controls. It should be mentioned that other than forces associated with microclimatic and microenvironmental factors, other signiﬁcant regional controls are those of anthropogenic origin, e.g., excavation and soil manipulation. Taking all of these into account, the nature and character of a soil is therefore best deﬁned by geologic origin and regional controls. The weathering processes responsible for formation of the nature of soils seen “in-place” after deposition are much the same as those described for chemical weathering of rocks. These processes, which include hydrolysis, oxidation, reduction and carbonation, are brought into play when water, organic matter, carbon dioxide and oxygen interact with rock material containing iron in the form of sulphides or oxides in addition to such minerals as silica, quartz, feldspar, hornblende and olivine. For example, one of the more signiﬁcant reactions is carbonation. When this process occurs with silicates, silica is liberated, producing either quartz or colloidal silica. Another reaction involving carbonation relates to the combination of carbonic acid with bases, which will result in the formation of carbonates. Distinguishing between the various kinds of soil materials requires the application of a classiﬁcation scheme that identiﬁes the materials according to certain

28

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

characteristic features. One of the primary aims in classifying a soil is to be able to communicate the proper soil descriptive and characteristic information and have it properly accepted. This is particularly important in soil identiﬁcation because geologists, soil scientists, soil engineers, environmental engineers, pedologists and many others may have different views as to what is necessary in describing a soil. For example, soil engineers require a good understanding of the physical and mechanical properties of the soil materials. However, this type of information is less important to pedologists, who are more interested in the soil material as weathered and unaltered unconsolidated material and particularly in the development of the soil proﬁle. For the purposes of contaminant attenuation evaluation and assessment, the ﬂuid and gaseous transmission properties of soil are of prime importance. These determine the rate at which the aqueous phase will be able to move through the soil, and by extension, the time that the contaminants will interact with the soil solids. The other distinguishing properties and characteristics of soil materials required relate directly to those that are prominently featured in the processes resulting in chemical mass transfer. These will be discussed in detail in the later parts of this chapter and in Chapter 3. Here we will consider the simpler means for classiﬁcation of soils and leave the detailed classiﬁcation schemes associated with attenuation processes to be addressed at the appropriate time. Other than a straightforward visual classiﬁcation that distinguishes between different types of soil materials, a simple and uncomplicated procedure for soil classiﬁcation is by means of the characteristic grain-size or particle-size distribution of the soil material. Methods for determination of particle sizes can be found in the various soil testing manuals for the different technical and scientiﬁc disciplines such as highway engineering practice, soil mechanics, soil physics, agriculture, etc. The size distribution of particles (soil solids) inﬂuences the physical, chemical and biological properties of soils. Whereas the larger granular particles such as sands, coarse silts and gravels can be easily visualized as forming a basic granular skeletal structure for the soil, and contributing directly to the mechanical properties of the soil, it is the material contained in the included voids that is responsible for much of the physico-chemical and chemical properties of the soil. The principal groups of soils discriminated solely on the basis of size distinction, shown for example in Figure 2.3, include (1) gravels for particle sizes or average diameters greater than 2 mm, (2) sands for particle sizes between 0.02 and 2 mm, (3) silts for particle sizes between 0.002 and 0.02 mm, (4) clays for particle sizes less than 0.002 mm and (5) colloids for particles with effective diameters less than 0.001 mm but larger than molecular size (10-6 mm). The effective diameter is determined from the settling velocity of the particles in water. While particle-size classiﬁcation uses the term clay to refer to soils with particle sizes of less than 0.002 mm effective diameter, the term clay is often used to refer to clay minerals obtained as a result of chemical weathering of rocks. These clay minerals are oxides of aluminum and silicon with smaller amounts of metal ions substituted within the crystal. To avoid confusion, the term clay size should be used to refer to the particle-size clay and clay minerals to mean those minerals derived from chemical weathering of rock. The clay-sized fraction of most soils contains

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

Clay

Silt

29

Fine -- Medium -- Coarse Sand

Gravel

100

Percent finer by weight

80 Clay 60 Sand 40 Clayey silt 20

0

0.001

0.01

0.1

1.0

10

Effective or average particle diameter, mm Figure 2.3

Particle-size distribution curve showing general classiﬁcation scheme based on particle-size discrimination.

clay minerals. The term clay soil is used to refer to a soil mass that contains claysized soil particles that include both clay minerals and nonclay minerals. There is some disagreement as to what the limiting size boundary between sands and silts should be. For example, there are some who choose to use 0.006 mm as the distinguishing boundary between ﬁne sands and coarse silts. The particle size clssiﬁcation scheme, however, that distinguishes between the different groups based upon changes in one or two orders of magnitude of particle size (see the upper scale in Figure 2.3), is often considered to be the simplest classiﬁcation scheme. It is not uncommon to ﬁnd subgroup classiﬁcations added to the general grouping, such as coarse, medium and ﬁne sands and silts. The three example particle size distribution curves shown in Figure 2.3 illustrate typical curves for clay soils, clayey silts and sands. The primary aim of a particle-size soil classiﬁcation scheme is to provide a basic understanding of soil type, and classiﬁcation schemes must be considered as application speciﬁc, i.e., speciﬁc to the role of the soil being tested. 2.2.2

Soil Composition

The previous sections have shown that soil materials are derived from the weathering of parent rock and are obtained on site as residual or transported soil. The soil solids that make up a soil material include various types of inorganic and organic solids. These different types, known as soil fractions, combine in various conﬁgu-

30

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

rations (1) according to the manner in which the soil is formed and (2) consistent with the regional controls. Figure 2.4 shows the major inorganic and organic soil fractions in a typical soil material generally known as a soil unit. The three separate phases in a typical soil unit are gaseous, ﬂuid and solid phases. The gaseous phase generally consists of air and gases, and the ﬂuid phase contains water and dissolved solutes. The soil solids consist of soil fractions, as shown in Figure 2.4. These include: • Soil organics: humic substances, polysaccharides • Inorganic noncrystalline material: hydrous oxides of iron, aluminum and silicon; allophanes • Inorganic crystalline material: oxides and hydrous oxides of iron, aluminum and silicon; primary and clay minerals; carbonates, sulfates, phosphates and sulphides

The amounts of each of the various soil fractions and the gaseous and ﬂuid phases present in any soil-water system are dependent on the formational characteristics of the soil. Source rock material, weathering regime and regional controls are the primary factors in determining the composition of the soil. The importance of soil composition in respect to the development of soil properties and inﬂuence on behavior cannot be overstated. The two ends of the particle-size distribution classiﬁcation provide us with the examples of cohesionless granular soils at one end and colloidal and cohesive clay soils at the other. Granular soils that are cohesionless soils are also known as coarse-grained soils. These consist of those soil materials falling into the categories of sands and gravels. They lack cohesion between individual particles — generally known as soil grains because of their shape — and hence are unable to stand freely in a soil mass without the aid of side support in the dry granular state. It is not uncommon to ﬁnd primary minerals, i.e., minerals present in the source rocks, as the principal components constituting these granular soils. The common childhood experience of being unable to form sand castles with dry sand but with some success with wetted sand is the best illustration of the cohesionless nature of the material. The capillary action between the wetted sand grains provides the resultant cohesive action between grains. The granular nature of these soils makes them good candidates for sturdy and stable soil masses when optimum packing of the soil grains is obtained. The granularity of the soils produces packing of soil grains that will result in higher hydraulic conductivities in comparison to those obtained in soils that have much smaller particle sizes and little or no particle granularity features. While pure cohesionless or coarse-grained soils contribute little or nothing directly to the assimilative capacities of soils, coarse-grained soils that contain ﬁne-grained soil particles have some ability to inﬂuence contaminant attenuation through control of the basic soil structure. These ﬁne-grained soils are generally clays and silts. In the wetted state, these ﬁne-grained soils exhibit bonding between particles as a result of the physicochemical properties of the soils. It is these physico-chemical properties that are of signiﬁcance in the development of the assimilative capacity of soils. Soil composition refers to the makeup of a soil. A typical soil contains such soil fractions as clay minerals, the various oxides and hydrous oxides, humic material or soil organic matter, carbonates and primary minerals. The proportions of any of these in a soil unit vary according to geologic origin, weathering history, regional

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

IDEALIZED SOIL UNIT

31

Silt particle

Carbonate

Clay mineral particle

Crystalline and amorphous oxides, hydrous oxides

Figure 2.4

Soil organic matter (SOM)

Ped (fabric unit)

An idealized typical soil unit in a soil mass consisting of various soil fractions. The positions of the various fractions and the conﬁguration deﬁne the structure of the soil.

controls and human intervention. Particle sizes in a clay soil vary from the larger sizes within the silt range to colloidal particles. Generally, primary minerals constitute the coarse-grained particles, with clay minerals, amorphous materials and soil organics occupying the lower end of the particle-size spectrum. The proportions of primary minerals in clay soils, as seen for example in the particle-size distribution curve for clay soils in Figure 2.3, are generally relatively small. Because of the small particle sizes and hence larger speciﬁc surface areas of the ﬁner-grained soil fractions, these are the main participants in the attenuating processes. Since soil is formed from the physical and chemical weathering of rocks, and since the composition of rocks varies considerably, we can anticipate that uniformity in soil composition would be a most uncommon phenomenon. The schematic soil unit in Figure 2.4 depicts a particular combination of the many soil fractions into microstructural units and individual fractions as shown in the diagram. By changing the proportions of any or all of the different soil fractions and the conditions of soil formation, one would have an almost inﬁnite number of different microstructural units and series of conﬁgurations that deﬁne the structure of the soil unit shown in Figure 2.4. In essence, soil composition, which includes various soil fractions, is a fundamental feature of a soil that has considerable impact on the development of the structure (i.e., macrostructure) of the soil and the various physical and physicochemical properties of the soil. These aspects will be discussed in greater detail in the following sections.

2.3 SOIL FRACTIONS, COMPOSITION AND ATTENUATION Before considering the attenuation processes and the interaction mechanisms between soil solids and contaminants in detail, it is necessary to sketch out the main

32

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

soil features that are central to the interactions in contaminant attenuation. It has been well demonstrated in laboratory exercises and ﬁeld experiences that clay soils are the best type of soil material for attenuating contaminants. What we need to determine is what particular sets of characteristics and properties of such soils are involved in the attenuation processes and how these relate to soil composition and structure. The idealized picture in Figure 2.5 shows a contaminant stream passing through the representative soil unit shown in Figure 2.4. In a very simplistic fashion, this picture indicates that in passing through the unit soil mass, the concentration of contaminants in the leachate stream is signiﬁcantly reduced — implying thereby that a large proportion of the contaminants are left in the unit soil mass. The two attenuating mechanisms that would be responsible for the reduction in concentration of contaminants are partitioning and physical hindrance. 1. Partitioning: We can assume initially that partitioning occurs as a result of interactions between contaminants and soil solids resulting in chemical mass transfer. This leads to the conclusion that the surfaces of the soil solids are reactive, i.e., that the surface forces associated with the soil solids’ surfaces can react chemically with the contaminants to achieve transfer of the contaminants. 2. Physical hindrance: Availability of voids and channels for ﬂow of solutes and contaminants in the contaminant stream depends on the structure of the unit soil mass. Arrangement of soil particles and peds or clusters of particles must allow for adequate void structure and spatial continuity of voids.

Contaminant Stream -- In

Figure 2.5

Contaminant Stream -- Out

Pictorial diagram illustrating contaminant sorption onto clay soil solids in transport of contaminant stream through the unit soil mass. Bottom diagram shows transfer of contaminants onto clay particle surface as a result of contaminant interaction with the reactive clay particle surface.

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

33

The nature of the reactive surfaces of the clay soil solids and the amount of reactive surfaces available are essential features in the attenuation process. We have stressed the use of clay soils because of their ﬁne-grained nature. The speciﬁc surface area (SSA) of ﬁne-grained soils, i.e., surface area per unit weight of soil particles, is much higher in comparison to coarse-grained soils, and in the case of clay soils, the soil fractions such as the clay minerals, amorphous materials and soil organics have signiﬁcantly large SSAs. In addition to the large SSAs, these surfaces are highly reactive, i.e., the nature of the surfaces are such that they can provide the proper sorption mechanisms for contaminant attenuation. The other important point to consider is that the proportions of the various soil fractions and their distribution produce soil structures unique to the conditions of that particular unit soil formation, i.e., geologic origin and regional controls. The inﬂuence of soil structure on contaminant attenuation is seen through: • Availability of particle surfaces: The amount of reactive surfaces available for interaction with the contaminants depends on the arrangement of the soil particles. If more peds and other kinds of microstructural units are formed, this will diminish the availability of exposed reactive surfaces of particles, and by extension, this will likely reduce the amount of contaminants that will be sorbed by the particles. • Structure of voids and spatial connection: Continuity in void spaces provides a channeling effect that permits ﬂow of the contaminant stream. Restrictions in void spaces impede ﬂow, and if the restrictions are severe, physical blocking of contaminant movement through the restriction can occur, essentially creating a dam effect.

To better appreciate the role of composition in the development of soil structure, we need to understand the controls on soil structure formation. This requires us to determine how the soil fractions such as clay minerals, soil organics, oxides and hydroxides of iron or manganese interact and the various forces that participate in the production of microstructures and ﬁnal soil macrostructure, as will be seen in Section 2.5. We are not ignoring the contributions of the primary minerals such as quartz and feldspar to the macrostructure of soils. These are important constituents of soils and contribute signiﬁcantly to the overall strength and stability of soils. However, in respect to the processes that contribute to attenuation of contaminants, these primary minerals play a relatively minor role and are therefore not included in the discussions to follow.

2.4 CLAY MINERALS AND SOIL FRACTIONS The clay-sized fractions of most soils contain clay minerals. Clay minerals are very important because the reactivities of the mineral particles’ surfaces are critical to many geochemical processes. The clay minerals are secondary minerals that are obtained as the product of chemical weathering of the primary minerals found in metamorphic and sedimentary rocks. We see, for example, that the alteration of a Ca-feldspar (plagioclase) to kaolinite can be written as a chemical reaction of end products:

34

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

CaAl2Si2O8 + 3H2O + CO2 Æ Al2Si2O5(OH)4 + Ca2+ + 2HCO3– The ﬁrst term on the left-hand side of the preceding equation represents feldspar, and the ﬁrst term after the arrow represents kaolinite. The feldspar is dissolved and the components reform into a new mineral species (kaolinite). The clay minerals, formed from dissolution of primary minerals, depend on the composition of the solution. Much depends upon which ions are lost by leaching in the dissolution process and subsequent to the process. With minimal or restricted leaching, cations remain in solution and 2:1 clay minerals such as montmorillonite or illite are formed. When silica is leached, 1:1 minerals such as kaolinite are formed. However, if silica leaching is excessive, one will obtain iron and aluminum oxides. Thus it is possible for a feldspar to weather ﬁrst to montmorillonite and then to kaolinite, or directly to kaolinite depending upon the conditions during weathering. Clay minerals consist of oxides of aluminum and silicon with small amounts of metal ions substituted within the crystal structure of the minerals. The aluminumoxygen and silicon-oxygen combinations form the basic molecular structural units that combine in various conﬁgurations to make up the different types of clay minerals found in clay soils. As a result, the clay minerals obtained are generally identiﬁed as aluminosilicate minerals or layer silicates (phyllosilicates). Figure 2.6 shows the basic silica tetrahedral and the Al-, Fe- or Mg-octahedral structural unit. The tetrahedral and octahedral molecular structural units are stacked together to form unit cells as shown in Figure 2.5 for the 1:1 and 2:1 unit cells. The 1:1 and the 2:1 refer to one tetrahedral structural unit and one octahedral structural unit for the former, and two tetrahedral structural units and one octahedral structural unit for the latter. These constitute the basic building blocks for the various clay minerals, as will be seen in the next few sections. The basic unit cells are linked laterally to form unit layers as shown in Figure 2.6. These unit layers consist of the basic tetrahedral sheet on top of the octahedral sheet for the 1:1 layer silicate and an octahedral sheet sandwiched in between two tetrahedral sheets for the 2:1 layer silicate. A clay mineral particle is formed from a stack of vertically linked unit layers. The type of linking, together with the nature and distribution of the ions populating the tetrahedral and octahedral units, characterizes the nature and properties of the clay mineral obtained. 2.4.1

Kaolin, Kaolinite and Kandite

Historically, the term kaolin has been used to denote the mineral and the mineral group in the parent rock. To avoid confusion, the term kaolin is now used to refer to the mineral group in the parent rock, and the term kaolinite is used to mean the mineral itself. The name kandite is used to denote the dioctahedral group of 1:1 layer silicates, of which kaolinite is the best known. In this dioctahedral group, only two thirds of the octahedral positions are ﬁlled in the octahedral sheet. The clay minerals found in this group are kaolinite, halloysite, dickite and nacrite. Pure kaolinite minerals, for example, are good source materials for usages such as adsor-

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

Figure 2.6

35

Basic silica tetrahedral and Fe, Al or Mg octahedral structure and sheet.

bents, paint formulations and paper coatings (Grim, 1962). This clay mineral is perhaps the most commonly found mineral in warm and moist climates. Acid leaching of the feldspars and micas in parent rocks constitutes a major source for kaolinites. Other sources of kaolinite could be the resilication of aluminum-rich materials by hydrothermal alteration (Weaver and Pollard, 1973) and precipitation of gels or solutions of silica and alumina (Dixon, 1977). As shown in the top righthand portion of Figure 2.6, the typical structure for 1:1 layer silicates consists of uncharged tetrahedral and octahedral sheets forming the basic unit layer with thickness of about 0.7 nm. The unit cell for the basic unit layer (Figure 2.7) is called a triclinic unit cell. Stacking of these unit layers creates the kaolinite mineral with the thickness of the unit layer as the repeat spacing between unit layers. Generally, the tetrahedral positions in the tetrahedral sheet are occupied by Si ions. As noted previously, the octahedral sheet has two thirds of the octahedral positions occupied by Al ions — typical of a gibbsite structure. Kaolinites are platey in morphology, with varying particle sizes dependent on the amount of disorder in the unit layers. In general, the average equivalent diameter of a typical particle is about 10 times that of the thickness of the particle. It has been argued that kaolinite refers to the situation where fully ordered minerals that show triclinic symmetry exist. Furthermore, the argument states that the term kandite should be used in place of kaolinite when doubt exists. For this book, the term kaolinite will be used to refer to the mineral, and the term kandite to refer to the group of clay minerals that are classiﬁed as dioctahedral

36

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Figure 2.7

1:1 layer silicate structural arrangement. The octahedral sheet with two thirds of the available positions ﬁlled with Al ions is a gibbsite sheet.

1:1 layer silicates. As stated previously, this group includes halloysites, dickites and nacrites in addition to kaolinites. Dickites and nacrites are variants of the basic kaolinite structure, with dickite being more crystalline than kaolinite. Both dickite and nacrite are large crystalline minerals and are relatively rare. Halloysite has the same basic unit layer as the kaolinite mineral except that a monomolecular layer of water separates the contiguous unit layers. Instead of the regular planar particle morphology, halloysites tend to assume tubular shapes. 2.4.2

Illites, Micas and Mixed-Layer Clays

Illites belong to the mica mineral structural group and are platey with variable thicknesses. Micas have a 2:1 layer structure for the basic unit layer, with cations in the interlayer separating the basic unit layers. These cations are called interlayer ions and can consist of potassium, sodium and calcium, with potassium being the more common interlayer ion. The various types of mica are distinguished by the population of the octahedral sheet and the distribution of silicon ions in the tetrahedral sheet. Historically, there has been considerable debate and discussion concerning mica nomenclature. The concern has been in respect to the recognition of micas as sources for vermiculites and mixed-layer minerals that represent various alteration products of the physico-chemical processes in the development of minerals from micas to smectites.

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

Basic unit layer; thickness = 1.0 nm

Basic unit cell

2:1 layer silicates

37

Montmorillonite Exchangeable cations in interlayer spaces

Clay mica

oo oo

oo oo

oo oo

oo oo

oo oo

oo oo

Potassium ions in hexagonal holes of silica sheets

Chlorite Interlayer hydroxide sheet

Figure 2.8

2:1 layer silicate basic unit layer. The different minerals shown in the diagram are distinguished by the nature of the interlayers. The interlayer hydroxide sheet separating the basic unit layers in the chlorite mineral may be a brucite or gibbsite sheet. However, what distinguishes this sheet from the octahedral sheet is that it does not share its atoms with the adjacent tetrahedral sheets.

For the purpose of this book, we will recognize illites as those hydrous clay micas that do not ordinarily expand from a 1.0-nm basal spacing (Grim et al., 1937). They are often referred to as clay micas or hydromicas. They have a 2:1 layer structure as shown in Figure 2.8, with potassium occupying positions in the interlayer. These interlayer ions are not exchangeable. Ordinarily, illites are nonswelling since they do not contain any expanding layer in their structure. However, there are illites that contain swelling layers. These are illites that are interlayered with montmorillonite and are called interlayered illites. Soils that contain these kinds of illites are identiﬁed as mixed-layer clays. Interlayered illites, i.e., illite-montmorillonite mixed-layer minerals, are perhaps the most common type found in mixed-layer clays. Another type of mixed-layer mineral often encountered in mixed-layer clays is the interlayered chlorite. This is a chlorite-montmorillonite mixed-layer mineral and is also sometimes called a swelling chlorite. 2.4.3

Vermiculites, Interlayered Vermiculites and Chlorites

The 2:1 layer structure that deﬁnes the unit layer structures of vermiculites and chlorites is similar to the micas. What distinguishes vermiculites and chlorites from

38

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

micas is primarily the nature of the octahedral sheets in the unit layer and the interlayer between the unit layers. Vermiculites are obtained as alteration products, more often from biotite micas and quite often from such rocks as granites and gneiss. Removal of the potassium in the interlayers of biotite micas is seen as a principal alteration process producing the vermiculite crystal. In the dry state, the repeat spacing between unit layers is 1.0 nm. The repeat spacing expands to a maximum of about 1.47 nm when vermiculite is exposed to water because of the two molecular layers of water in the interlayer associated with the interlayer magnesium ions. Interlayered vermiculites are more likely to be a combination of vermiculite with mica or chlorite. Removal of the potassium from illites, biotites, muscovite and the hydroxyl sheet in chlorite result in the production of vermiculite interstratiﬁed with mica and chlorite layers (Weaver & Pollard, 1973). Because chlorites appear to have considerable isomorphic substitutions in the octahedral sheet of the unit layer, the term chlorite-type mineral is perhaps more appropriate in classifying the type of mineral structure shown in Figure 2.8. The octahedral sheet that forms the interlayer can be a brucite layer or a gibbsite layer. When Al ions ﬁll two thirds of the available positions in the interlayer hydroxide sheet, this interlayer is known as the gibbsite layer, with a chemical formula Al2(OH)6. When Mg is in the octahedral sheet, the interlayer is known as the brucite layer, with a chemical formula Mg3(OH)6. The interlayer hydroxyl sheet does not have a plane of atoms that can be shared with the adjacent tetrahedral sheets. 2.4.4

Smectites

Smectites are dioctahedral and trioctahedral hydrous aluminum silicate clay minerals containing magnesium and calcium. The dioctahedral minerals that are generally obtained as the result of transformation and weathering processes of volcanic material and igneous rocks are montmorillonite, beidellite and nontronite. The trioctahedral minerals that are essentially obtained or inherited from the parent material and that are not commonly found as soil fractions in soils are saponite, sauconite and hectorite. The term bentonite has sometimes been used in the literature dealing with clayengineered barriers to mean a swelling clay composed of a signiﬁcant amount of montmorillonite. Most often, it is assumed that this would be sodium montmorillonite. Unless there is a speciﬁc detailed characterization of the montmorillonite, this would be a serious error. Bentonites fall into the class of dioctahedral smectites (Figure 2.8). These are obtained as the alteration products of volcanic ash and are composed primarily of montmorillonite with some measurable proportion of beidellite. These proportions vary depending on the source of the bentonite. Beidelites and nontronites are distinguished from montmorillonites by octahedral substitutions, with beidelites being the aluminum-rich smectites and nontronites the iron-rich smectites. The interlayer cations shown in the montmorillonite picture in Figure 2.8 determine the hydrated basal spacing. The ability of these cations to take in water is one of the principal reasons for using montmorillonites in the construction of engineered barriers designed to impede transport of leachates.

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

2.4.5

39

Carbonates

Some of the various types of carbonate minerals found in soils and groundwater include calcite (CaCO 3 ), magnesite (MgCO 3 ), siderite (FeCO 3 ), dolomite (CaMg(CO 3 ) 2 ), trona (Na 3 CO 3 HCO 3 ·H 2 O), nahcolite (NaHCO 3 ) and soda (Na2CO3·10H2O). Calcite (CaCO3) is probably the most common form of the carbonate mineral found in soils. Because of its inﬂuence on the pH of the soil, its presence and distribution in soils is very important since many of the chemical processes and reactions occurring in soils are pH sensitive. If CO2 is present in the pore water, dissolution of calcite will be likely, according to the following relationship: CaCO3 + CO2 + H2O ´ Ca2+ + 2HCO3– Carbonate minerals are thought to be good adsorbers of heavy metals and phosphates. The relatively high solubilities of carbonates and sulfates compared with the layer silicates and the aluminum/iron oxides, hydroxides and oxyhydroxides, mean that their presence in large amounts is mainly conﬁned to regions where limited leaching and high evaporation occur — typically arid and semi-arid regions. 2.4.6

Soil Organics

Although soil organic matter composes anywhere from 0.5 to 5% by weight of a typical soil mass, its importance in the development of soil stability and processes associated with contaminant attenuation cannot be overstated — even at such low proportions. Among the beneﬁcial effects are (Hayes & Swift, 1985) formation and maintenance of good soil structure and improved water retention. Although peat material is technically a soil organic matter, it is considered an exception to this proportionate distribution since it can, by itself, constitute 100% of a surface mass. Soil organic matter can originate from vegetation and/or animal sources and is generally classiﬁed on the basis of degradation (Hayes & Swift, 1985) as unaltered organics and transformed organics. Unaltered organic materials are fresh and relatively fresh organic materials that have yet to show the results of transformation processes. Transformed organic materials by deﬁnition refer to organic materials that show no morphological resemblance to the parent material and do not exhibit properties and characteristics of the parent material. Transformation occurs as a result of chemical and biologically mediated reactions. According to Flaig et al. (1975) these include demethylation and oxidative, reductive and other electron transfer reactions catalyzed by enzymes. Amorphous soil organics are highly aromatic polymers that are rich with functional groups. They consist of humic acids, fulvic acids, humins and decayed material such as polysaccharides, lignins and polypeptides. The term amorphous in organic chemistry bears a different meaning from that used in this text. Amorphous materials in soils refer to oxides, hydrous oxides of iron and aluminum, and oxides of silicon that do not exhibit any crystalline structure. In relation to transformed organics, the term amorphous is speciﬁcally used to indicate that the transformed organics no longer exhibit properties and characteristics of the parent material.

40

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Characterization of these amorphous soil organics is generally obtained through alkali treatment procedures on soil organics extracted from the parent soil as shown in Figure 2.9. The standard procedure for determination of humic and fulvic acids, humins, and nonhumic fractions follows the protocols shown in the left-hand branch of the schematic shown in the ﬁgure The protocols shown as the right-hand branch of the schematic have been used by Yong and Mourato (1988) to determine the presence of polysaccharides in the nonhumic fraction. As seen from the ﬁgure, one deﬁnes the constituents of the soil organic matter as follows: • Humic acids: soluble in bases but precipitate in acids • Fulvic acids: soluble in both bases and acids • Humins: insoluble in acids and bases

Classiﬁcation of polysaccharides as a humic substance sometimes poses a dilemma. Polysaccharides are long molecules that are obtained as by-products of microbial metabolism, generally synthesized as a by-product from the breakdown of animal or plant-derived organics. Microbial polysaccharides are synthesized within soil by active microbial populations. Depending on whether polysaccharides are extracted with the fulvic fraction, these carbohydrates are often incorporated as fulvic-related material. Thus, according to the basic deﬁnition of a humic or amorphous substance, polysaccharides do not fall in the class of nonidentiﬁable degraded organics since these polymeric carbohydrates have a deﬁnitive structure. 2.4.7

Oxides, Hydroxides and Oxyhydroxides

Oxides, hydroxides and oxyhydroxides are products of weathering of parent rock material. Good examples are the oxides and hydroxides of iron that are obtained as a result of weathering of iron-bearing silicates. The iron bound in the silicates that is released through a combination of hydrolytic and oxidative reactions precipitates as iron oxide or iron hydroxide because of the low solubility of Fe3+ in the normal pH environment of the soil (Schwertmann and Taylor, 1977). It is not uncommon to ﬁnd the oxides, hydroxides and oxyhydroxides grouped under the common name of oxides in the literature. The main structural conﬁguration of oxides is octahedral. The oxides of aluminum obtained primarily from weathering of alumino-silicate minerals show octahedral sheets containing OH– ions with two thirds of the positions occupied by Al3+ ions. Depending on how the OH– ions in the octahedral sheets are positioned, one can obtain, for example, either gibbsite or bayerite. When the OH– ions in each of the octahedral sheets that are stacked on top of each other are directly opposite to each other (i.e., between the stacked sheets) and bonded by hydrogen bonds, the structure of gibbsite is obtained. However, when the OH– ions in each of the octahedral sheets are positioned in the space formed by the OH– ions in the opposing stacked sheet, one obtains a closely packed OH– ion conﬁguration that is representative of the structure of bayerites. A primary feature of the oxides of iron, aluminum, manganese, titanium and silicon that distinguishes them from layer silicate minerals is their surfaces, which essentially consist of broken bonds. The low solubilities of the oxides of aluminum,

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

41

SOIL EXTRACT organics with 0.5 N HCl for 23 h under nitrogen atmosphere Centrifuge

Precipitate

Buffer to pH 10 with 10 N NaOH

Buffer to pH 2 with 8 N HCl

Add

HUMIN

Supernatant

Centrifuge

Precipitate

Soluble

Centrifuge

ethanol

EXTRACT with 0.5 N NaOH 23 h nitrogen atmosphere Insoluble

Supernatant

Centrifuge

Centrifuge Supernatant

Precipitate

FULVIC ACID

HUMIC ACID

Redissolve in alkali and add electrolyte

Precipitate NON HUMIC fraction

HUMIN Supernatant Discard

Treat with ethanol

Precipitate GRAY HUMIC ACID

Soluble

BROWN HUMIC ACID

Soluble HYMATOMELANIC ACID

Residue

Figure 2.9 Extraction and treatment technique for classiﬁcation of soil organic matter (Adapted from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, 2000, 307p.)

iron and manganese in the general pH range found in most natural soils mean that they are more common than the oxides of titanium and silicon. Iron oxides are the most common form of oxides found in soils. As with most of the other oxides, because of their surface charge characteristics, the amorphous forms of the oxides form coatings around particles with net negatively charged surfaces. The net result is a change in the charge characteristics of the soil particles and a change in the physical, chemical and interaction properties of the soil. Typically, interaction of the oxide surfaces with water occurs between the broken (i.e., unsatisﬁed) bonds on the oxide surfaces and the hydroxyl groups of dissociated water molecules. It is important therefore to distinguish between the mineral and amorphous form of oxides since distribution of the oxides in the soil depends on the form. The surface charges of the amorphous materials coating the other soil solids are pH dependent, and hence the nature of the interactions between the coated soil particles and water are conditioned by the pH of the medium.

2.5 PHYSICAL CHARACTERISTICS AND PROPERTIES The physical characteristics and properties of soils cover the range from the porosity of the soil to its strength and integrity. In this section, we will be concerned with those properties and characteristics that have immediate and direct bearing on

42

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Figure 2.10

Schematic diagram showing the major physical soil factors involved in controlling contaminant transport through the unit soil mass.

the contaminant attenuation capability of soils. In essence, we need to look at those properties and characteristics of soils that are directly involved with, or contribute to, the interactions between soil particles and contaminants. The sketch shown in Figure 2.10 offers a view of the major pertinent physical factors involved in the interactions. These are (1) porosity n or void ratio e, (2) continuity of voids, (3) soil structure, (4) hydraulic conductivity and (5) amount of “wetted” surface area. All of these properties are directly related to the density of the soil mass. In essence, all of these properties are dependent on particle packing and arrangement, i.e., soil structure. This follows from the discussion in Section 2.3. From the point of view of soil structure characterization, Bouma (1992) states that (1) the static parameters describing soil structure include soil bulk density, porosity and equivalent pore size distribution, and (2) the dynamic characteristics of the soil structure can be obtained by determining the permeability (for water and air) of the soil. 2.5.1

Soil Composition and Soil Structure

The role and inﬂuence of soil structure on the properties and performance of clays have been studied from many different perspectives, depending upon soil end use or function. Early studies on soil structure in geotechnical engineering, for example, provided us with descriptions of ﬂocculent, honeycomb and “cardhouse” structures (Terzaghi & Peck, 1948). Studies such as those reported by Lambe (1953, 1960), Pusch, (1966) and Yong and Warkentin (1966) to a large extent paid attention to the contributions made by the different clay minerals in combination with other soil fractions on the engineering properties and performance of these soils. These properties were primarily those that concern the stability and hydraulic conductivity of the clay soils. Soil structure consists of a combination of individual coarse- and ﬁne-grained particles and microstructures in various arrays, as shown schematically in Figure 2.4. In clay soils, microstructures predominate. These microstructures, which are packets of clay particles grouped tightly together through bonding forces such as forces of

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

43

1200 10 meq/L NaHCO3

Rate of Shear, D per sec.

1000

1 meq/L NaCl

o

800 o

#

#

300 meq/L NaCl

600 # o 400 # 3000 meq/L NaCl

200 o # 0 0

Figure 2.11

3 meq/L NaCl

I I tb tot 1 2 Shear Stress t , Pa

3

4

5

Rheograms for kaolnite soil suspension showing development of microstructural units as salt concentration increases — as might be deduced by the development of shear stress at very low rates of shear.

attraction between particles have been referred to as, peds, ﬂoccs and clusters. The types or properties of these microstructures, together with their arrangement in a soil mass, are factors that contribute signiﬁcantly to the physical and interactive (chemically reactive) properties of the clay soil. In particular, these microstructures have the capability to inﬂuence not only hydraulic conductivity but also partitioning of contaminants in the transport through the soil. This is because of the amount of surface areas exposed to the contaminants and because the size and arrangement of the microstructural units or aggregate groups control the nature of the pore channels through which the contaminant in the pore water is transported. The microstructures in clay soils, which are often referred to as peds, clusters, domains, aggregate groups and tactoids, vary in size and particle population. Quite often, composition and formational features such as origin and regional controls are signiﬁcant inﬂuences on the development of the type of microstructures. For convenience, these microstructures will be referred to as microstructural units (i.e., mus for plural and mu for a single unit). The mu’s form the building blocks for the macrostructure of most clay soils. Figure 2.11 shows the results of viscosity experiments on kaolinite soil suspensions conducted to demonstrate the development of mus as one changes the local environment of soil particle deposition. In this case, we allow the kaolinite particles to ﬂocculate as we manipulate the resultant forces of interaction between the particles by introducing electrolytes. The results of these manipulations are the rheograms shown in Figure 2.11 obtained for kaolinite suspension with 9.1% w/w solids concentration at various concentrations of NaCl and a rheogram for a 4.9% w/w kaolinite solids concentration in a 10 meq/L NaHCO3

44

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

solution. The shear stress, to, identiﬁes the critical shear stress needed to initiate ﬂow of the soil suspension, and the Bingham yield stress, tb, for any one rheogram is determined by the intercept on the abscissa of the linear portion of the ﬂow curve. Using h to refer to the slope of the linear portion of the rheogram and D to indicate the shear rate, the shear stress t in the linear range of the rheogram is given by: t = tb + hD . The Bingham yield stress, tb, is a measure of the bond strength between microstructural units, and the rate-dependent term, hD, is the resistance offered by the interparticle forces to shear displacement. The rheogram results of different soil compositions shown in Figures 2.12 and 2.13 demonstrate the inﬂuence of composition on development of MUs. These include the use of kaolinite with polysaccharides representing nonhumic organics (Figure 2.12) and kaolinites and illites with iron hydroxides, Fe2O3 (Figure 2.13). The polysaccharide xanthan has been described by Yong and Mourato (1990) as double stranded, anionic and unbranched, with dimensions equivalent to 4 mm. While the to and tb values for the xanthan-kaolinite mixture shown are lower than the kaolinite suspension in 3000 meq/L NaCl, Yong (2003) has noted that as the proportion of xanthan added was increased progressively to 1.0% w/w, the to and tb values also increased progressively. What is signiﬁcant about the addition of xanthan is that even with very small proportions of the polysaccharide, the anionic character of the polysaccharide provides a bridging bond between the positive edge charges of the kaolinite particles and domains, packets or tactoids. These bridging bonds are capable of contributing greater attachment strength between units. 800 9.1% kaolinite 1 meq/L NaCl

Rate of Shear D per sec.

600

14% kaolinite 1 meq/L NaCl

9.1% kaolinite 3000 meq/L NaCl

400 0.2% Xanthan only 200 0.2% Xanthan plus kaolinite 0

Figure 2.12

I 0 to

I tb

1 2 Shear Stress t , Pa

3

4

Rheograms for kaolinite in 1 meq/L NaCl and 3000 meq/L with addition of Xanthan (polysaccharide).

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

45

2.5 Iron oxide Kaolinite plus iron oxide

Bingham Yield Stress, Pa

2

1.5

Kaolinite

1

Illite plus iron oxide

0.5 Illite 0 3

4

5

6

7

8

9

10

pH

Figure 2.13

Bingham yield stress curves for kaolinite, illite and iron oxide and their mixtures.

Iron hydroxides, Fe2O3, added to the clay minerals kaolinites or illites creates interactions between the soil fractions and results in formation of different kinds of MUs, i.e., different from MUs created solely with single-species minerals or single-type fractions. The addition of a very small proportion of polysaccharide, for example (Figure 2.12), does not essentially change the monomineralic nature of the soil. The primary change as we have seen is in regard to the development of bridging bonds that are the basic building blocks for the microstructures. However, when iron hydroxide as a soil fraction is mixed with a pure clay mineral, for example kaolinite or illite, the kinds of MUs developed are different. This is because the nature of the coatings on the mineral particles depends on the sequence and conditions of mixing of the soil fractions and on the pH of the soil-water system. Figure 2.13 shows the variation in Bingham yield stress tb in relation to pH for kaolinite, illite and iron hydroxide, together with mixtures of kaolinite with iron hydroxide and illite with iron hydroxide, prepared initially at pH 9.5. The reasons for the signiﬁcant differences for some of the mixtures will be discussed in a later section when we deal with the nature of the surface reactive forces of the soil fractions. 2.5.2

Soil Structure and Transmission Properties

The transmission property of a soil refers to those properties that participate in the transport of leachates and other ﬂuids through the soil. These are generally considered to fall in the class of transport processes that control natural attenuation and are essentially described by the permeability of the soil to aqueous and gaseous

46

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

phases. The factors that affect hydraulic conductivity are divided into two groups: (1) those that pertain to the ﬂuid phase and (2) those that concern the soil solids and soil structure. As we pointed out previously, a unit soil mass is made up of an almost inﬁnite number and arrangement of soil particles and peds (fabric units), as shown for example in Figure 2.4. Accordingly, it would not be surprising that soils with similar compositions can have different densities and correspondingly different hydraulic conductivities. For that matter, it should also not be surprising to learn that soils with similar densities and composition could also possess differing hydraulic conductivities — attributable to differing soil macro- and microstructures. Soil permeability to water (aqueous phase) is measured as hydraulic conductivity for saturated soils and diffusivity for unsaturated soils. For saturated soils the permeability is commonly expressed in terms of a Darcy permeability coefﬁcient, k, which is generally obtained via application of the Darcy model as a means of analysis of hydraulic conductivity data. From laboratory permeability measurements, the Darcy coefﬁcient k is obtained from the relationship v = ki = k(DhDL). The hydraulic gradient i is the ratio of the hydraulic head Dh and DL, the spatial distance. This form of expression of k does not consider the impact of the properties of the permeant on hydraulic conductivity, nor does it take into account the structure of the soil (micro- and macrostructure). To accommodate the inﬂuences of the permeant and soil structure properties such as permeant viscosity, soil-voids’ features, tortuosity and shape of pore space cross section, a different type of permeability coefﬁcient is needed. To obtain this new coefﬁcient, the Poiseuille relationship for ﬂow through ﬁne-bore tubes shown in Equation 2.1 is used in an adapted form for determination of the link between soil structure and soil permeability.

v* =

r 2 g Dy 8h Dl

(2.1)

where:

g

v* = the mean effective ﬂow velocity through a narrow tube of radius r, and h = the density and viscosity of ﬂuid or permeant, respectively, and Dy = the potential difference between the ends of a tube of length Dl.

To account for the inﬂuence of the properties of pore channels deﬁned by the structure of a soil, and the fact that the wetted soil particles’ surface area is controlled by the structure of the soil, Yong (2003) has used a modiﬁcation of the combined form of the Poiseuille and Kozeny-Carman (K-C) model:

v = k *i =

Cs n 3 g DY hT 2 Sw2 Dl

(2.2)

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

47

where • k* = the permeability coefﬁcient that considers permeant and soil structure properties,

=

Cs n 3 g hT 2 Sw2

• Cs = the shape factor, and has values ranging from 0.33 for a strip cross-sectional face to 0.56 for a square face. This factor accounts for the fact that the crosssectional face of any of the pore spaces in the soil mass is highly irregular and allows one to choose a typical value for a representative pore cross-section area. Yong and Warkentin (1975) have suggested that a value of 0.4 for Cs can be used as a standard value — with a possible error of less than 25% in the calculations for an applicable value of k*. • i = the hydraulic gradient, which is the ratio of the potential difference Dy between the entry and exit points of the permeant and the direct path length Dl of the soil mass being tested. • T = the tortuosity, which is the ratio of effective ﬂow path Dle to thickness of test sample Dl and which is quite often taken to be ª ÷2. • g and h = the density and viscosity of the permeating ﬂuid, respectively. • n = the porosity of the unit soil mass. • Sw = the wetted surface area per unit volume of soil particles.

Equation 2.2 contrasts with the standard K-C relationship in that the wetted surface area consideration in the K-C model assumes that Sw = S(1-n) and that the n . S(1 - n) This gives us the relationship for k from the K-C model as radius r of the Poiseuille tube is: r =

k=

Cs gn 3 hT 2 S 2 (1 - n) 2

(2.3)

where S = the speciﬁc surface area of soil. With this measure of surface area, we assume that all particle surfaces are in contact with the permeating ﬂuid. Experience with the use of the classical K-C relationship has been good when the pore sizes are relatively uniform, i.e., when the particle sizes are relatively uniform. However, experience has also shown that when pore sizes vary signiﬁcantly, and especially when there is a high proportion of ﬁne fractions in the soil, the relationship does not hold well. This should be intuitively obvious because the speciﬁc surface area of the soil will be proportionately higher because of the presence of the ﬁnes. The adaptation introduced by Yong and Mulligan (2003) considers the surface area of the particles in terms of only the wetted surfaces, i.e., Sw. This allows for compositional differences and soil structure differences that can impact severely on the distribution of pore sizes and the availability of soil particle surfaces for direct interaction with the permeant. We see from Equation 2.2 that Cs, T and S are soil

48

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

10-7 De cre as i po ng c ros alc ity ula ted n

Permeability Coefficient k*, cm/sec

10-8

10-9 n = 0.4

10-11

Inc rea sin po g ca ros lcu ity l n ated

10-12

10-13 1000

Figure 2.14

n = 0.3

10-10

10000 Wetted Surface Area Sw , cm2/cm3

100000

Variation of permeability coefﬁcient k* with wetted surface area Sw and calculated porosity n.

property parameters that are dependent on soil composition and soil structure. These can be expressed as a parameter b = Cs/(TS)2 . Along the same lines, the density and viscosity of the permeating ﬂuid, g and h, respectively, are properties of the permeant and can be described by a parameter m = g/h. Using these parameters, the relationship for k* can be expressed as k* = mbn3. Assuming that the physical properties of a leachate permeant are not too far distant from those of water at about 20ºC, we can compute m directly. Further assuming a tortuosity, T, value of ÷2, and Cs = 0.4, the graphical relationships shown in Figures 2.14 and 2.15 can be obtained. These graphs show the relationship between the soil permeability expressed as a coefﬁcient k* and the amount of surface area wetted in ﬂuid ﬂow through the soil — all of which are determined in relation to the porosity of the soil. The wetted surface area is the surface area of the pore channels in the soil through which ﬂuid ﬂow occurs.Yong and Mulligan (2002) have provided an example of how the chart in Figure 2.15 could be used to provide information on the wetted surface area and the inﬂuence on soil structure in the control of hydraulic conductivity. The properties of the laboratory and ﬁeld soils are shown in Table 2.1, and their locations on the wetted surface chart are shown in Figure 2.15. The wetted surface areas Sw are seen to be small fractions of the speciﬁc surface area of the various soils. This tells us that the microstructure of the soils plays a dominant role in the development of ﬂow channels. Denoting the surface area of the soil solids comprising a unit volume of the soil being permeated as SSAv , we can deﬁne the ratio Sw/SSAv as the wetted surface ratio (WSR). The WSR provides an indication of the microstructure of the soil and the extent of soil particle surfaces available for interaction with the ﬂuid used for the permeability test, i.e., surface area wetted during hydraulic ﬂow (Yong & Mulligan, 2003).

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

49

Effective Permeability Coefficient k*, cm/sec

10-05 3

2 m 00 cm /c S w = 10

10-06 3 m2 /cm 3000 c

Kaolinite 10-07 Mudrock

3 m2 /cm 7000 c

Glacial till Alluvium

10-08

3 2 cm /cm 14000

Lachenaie Illite Micaceous

10-09 0.2

0.3

0.4

0.5

0.6

Calculated Porosity n Figure 2.15

Variation of permeability coefﬁcient k* with calculated porosity in relation to wetted surface area Sw.

Table 2.1 Speciﬁc Surface Area (SSA) and Cation Exchange Capacity (CEC) for Some Soil Samples

Soil Sample

SSA (m2/g)

CEC (meq/100 g)

Void Ratio, e

Permeability (coeff. k, 10-8 cm/sec)

Kaolinite Illite Natural clay (Lachenaie) Natural micaceous soil Mudrock (soil) Glacial till (ﬁne fraction) Estuarine alluvium

12.0 81.0 80.0 200.0 46.4 69.6 83.5

8.0 50.0 60.0 13.0 11.9 23.9 37.7

0.95 0.46 0.81 0.48 0.33 0.43 0.50

2.0 0.9 0.7 0.2 4.0 3.5 2.2

2.5.3 Darcy’s Law, Low Water Contents and Unsaturated Flow There has been some signiﬁcant recent evidence concerning the lack of proportionality between ﬂow velocities and hydraulic gradients in experimental studies of ﬂow in clay soils, particularly for low water contents in saturated soils and in partly saturated soils (Philip, 1957; Lutz & Kemper, 1959; Hansbo, 1960; Yong & Warkentin, 1966). There is experimental evidence that at low water contents linear proportionality between ﬂow velocity and hydraulic gradient is not established until some critical gradient is reached. Figure 2.16 shows the Darcy model with the linear proportional relationship between the ﬂow velocity and the hydraulic gradient, i,

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Velocity v

50

v=

ki

Linearity between v and i Critical gradient ic

Hydraulic gradient, i = g⑂/d⑂ Apparent threshold gradient ia Figure 2.16

Flow velocity v in relation to the hydraulic gradient i. The Darcy model shows the linear proportional relationship between v and i passing through the origin of axes.

passing through the origin. This relationship appears to work satisfactorily for coarsegrained soils and for soils that have somewhat rigid soil structure and geometries. For clay soils, and especially for clays with high speciﬁc surface areas, the linear relationship is not evident at low water contents — even in the fully saturated state. This is shown in Figure 2.16, where linearity is not established until after one reaches a certain hydraulic gradient. This is deﬁned as the critical gradient ic, and the intercept of the extended linear relationship on the abscissa is identiﬁed as the apparent threshold gradient ia. To examine the applicability of the Darcy model for analysis and prediction of unsaturated ﬂow, we begin with the Hagen-Poiseuille relationship, which describes the mean velocity v of an –incompressible ﬂuid through N number of circular capillary tubes of mean diameter d and effective length Dle: _

v d 2 Dp v = max = 2 32m Dle

(2.4)

where m = ﬂuid viscosity and Dp = pressure. Douglas and Yong (1981) have shown that if one considers that portion of the pore spaces ﬁlled with water, using the expression nw to indicate the porosity associated with water-ﬁlled pores, the mean velocity v will now be expressed as the capillaric velocity vc to reﬂect the reality of ﬂow only in the water-ﬁlled pores:

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

51

1015 䡵

D夝

夝夝

o

Friction factor, f

1014

夝䡵

䡵

o Do 䡵 Distilled water

D

D 3.0 x 10-4N NaCl

1013

o䡵 o o

o 1.4 x 10-3N NaCl

夝

1.0 x 10-1N NaCl

1012 10-10

10-9

10-8

Reynolds’ number Re Figure 2.17

Relationship between friction factor f and Reynolds’ number for partially hydrolyzed pure kaolinite at different NaCl concentrations. (Data from Douglas, E. and Yong, R.N., Validity of Darcy’s law for one-dimensional, unsaturated ﬂow in kaolinite-water-NaCl systems, Trans. ASAE, 24, 657–662, 1981.) _

1 d 2 nw Dp vc = 96 mT 2 Dl e

(2.5)

where T = tortuosity =l/le, nw = q/1 - q and q = volumetric water content. They further show from dimensional analysis that the friction force factor f is proportional to the ratio of the frictional forces and the viscous force. This can be obtained in terms of the inverse of the Reynolds’ number as f = C/Re. Expressing the Reynolds’ -

number as Re = r dn/m, we obtain the friction force factor f = cm/r -dn , where c is a constant, and r is the ﬂuid density. The results from a typical set of desorption tests conducted by Douglas and Yong (1981) and from a partially hydrolyzed pure kaolinite soil with varying NaCl concentrations shown in Figure 2.17. Note that the abscissa is expressed as a logarithmic scale. The graph shows the relationship between the friction factor f and Reynolds’ number Re. We can draw several conclusions from this relationship — besides the obvious one, which essentially points out the nonlinearity between ﬂuid ﬂux and ﬂuid pressure gradient. It would be very interesting to speculate on the reasons for the nonlinearity. With the varying salt concentrations, different soil structures will be obtained, and hence we will have differing microstructures. The result of variable microstructures, combined with the reactive surfaces of the particles, will provide for a combination of equivalent drag

52

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

resistance forces that would not support the linear relationship between ﬂuid ﬂux and hydraulic gradients implicit in the Darcy model.

2.6 SURFACE PROPERTIES OF SOILS To develop the full potential of a soil for sorption or assimilation of contaminants, it is evident from Section 2.1.1, and from intuitive reasoning, that the soil fractions constituting a soil mass should have maximum contact with the contaminants carried in the soil water (pore water). Together with maximum contact, the soil should also have maximum uptake capability. We deﬁne uptake in this instance to mean chemical mass transfer resulting from the mechanisms, processes and/or reactions between the contaminants in the pore water and the soil solids’ surfaces. If the mechanisms relevant to contaminant assimilation and attenuation in a soil-water system are to be fully utilized, it follows that all the available soil solids’ surfaces need to be involved in the interactions with the contaminants. This is why soil structure and the presence of microstructural units are important factors in the development of the assimilative capacity of soils. The greater the number of soil particles that participate in forming MUs, the lesser will be the available soil particles’ surfaces for interaction with contaminants in the leachate stream. When we have maximum surface area contact, uptake of contaminants increases because there are more reactive sites for interaction with the contaminants. When we combine these with pore or void space restrictions, hindrance partitioning of the contaminants occurs, as discussed brieﬂy in the ﬁrst part of Section 2.3. By this, we mean that when contaminants are removed from the pore water through interaction with the soil solids, a process of partitioning is said to have occurred. The mechanisms and processes leading to partitioning are discussed in Chapters 4 and 5, respectively, for inorganic and organic contaminants. 2.6.1 Reactive Surfaces of Soil Fractions We deﬁne reactive surfaces to mean those surfaces that by virtue of their properties are capable of reacting physically and chemically with solutes and other dissolved matter in the pore water. Chemically reactive groups (molecular units) associated with the surfaces of the various soil fractions render the surfaces of these fractions reactive. These reactive groups are deﬁned as surface functional groups. The literature often refers to these simply as functional groups. The soil fractions with reactive surfaces include layer silicates (clay minerals), soil organics, hydrous oxides, carbonates and sulfates. The surface hydroxyls (OH group) are the most common surface functional group in inorganic soil fractions (soil solids) such as clay minerals with disrupted layers (e.g., broken crystallites), hydrous oxides and amorphous silicate minerals. These surfaces are often referred to as hydroxylated surfaces. The nature of the reactive surfaces of clay minerals can be traced directly to the structure of the layer lattice structures shown in Figure 2.6. Taking the group of clay minerals with the 1:1 structure such as the kaolinite shown in Figure 2.7, we have siloxane and gibbsite surfaces on opposite basal surfaces of the particles. The siloxane surface is characterized by the basal plane of oxygen atoms that bound the

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

53

tetrahedral silica sheet. These basal planar surfaces are typical of minerals whose structures have bounding tetrahedral sheets. Thus, while kaolinites possess one siloxane surface by virtue of the 1:1 structure, the illites, montmorillonites and vermiculites show siloxane-type surfaces on both bounding surfaces. These siloxanetype surfaces are reactive surfaces because of the structural arrangement of the silica tetrahedra and the nature of the substitutions in the layers. The regular structural arrangement of interlinked SiO4 tetrahedra, with the silicon ions underlying the surface oxygen ions, produces cavities that are bounded by six oxygen ions in a ditrigonal formation. Where no substitution of the silica in the tetrahedral layer and lower valence ions in the octahedral layers occurs, the surface may be considered to be free of any resultant charge. However, if replacement of the ions in the tetrahedral and octahedral layers by lower valence ions occurs through isomorphous substitution, we obtain resultant charges on the siloxane surface and hence a reactive surface. The resultant charges on these surfaces owing to isomorphous substitution are generally positive in nature. The broken edges of the kaolinite particles have Al and Si centers that have hydroxyl terminals, i.e., they are terminated by hydroxyls. The aluminols can accept or donate protons. According to Sposito (1984), adsorption of water onto these aluminol sites will produce Lewis acid sites. More detailed discussions on reactive surfaces can be found in Chapter 3 where we discuss reactions between soil particles and solutes in the pore water. Unlike the clay minerals, a greater variety of surface functional groups exists in soil organic matter. Carbon constitutes the backbone structure of all soil organics. It can be combined in saturated or nonsaturated rings or chains. Carbon and nitrogen can combine with oxygen and/or hydrogen to form surface functional groups. These functional groups essentially control the properties and reactions of the organic molecules. The schematic diagram (Yong, 2000) in Figure 2.18 shows the common functional groups associated with soil organic matter. These include hydroxyls, carboxyls, phenolics and amines. These surface functional groups are organic molecular units and are considered to be part of the organic matter itself. The wide ranges and values of the proportions of each kind of functional group are to a very large extent due to differences in soil organic matter composition, i.e., source material, degradation, extraction and testing procedures. The values shown in Table 2.2 provide a good example of the range of values for some of the typical functional groups. Carboxyls and phenolic groups have generally been considered as the more important functional groups in soil organics, particularly in respect to the development of their CEC and chelating capability. Functional groups have the ability to protonate or deprotonate depending on the pH of the surrounding medium. They develop positive or negative charges depending on the pH of the soil and their respective pKa and pKb (i.e., their respective log acidity and log basicity constants). 2.6.2 Surface Charge Density We have seen from the previous section that silanol surfaces associated with clay particles exhibit resultant positive charges. Additionally, we can see from Table 2.2 and especially Figure 2.16 that the various kinds of functional groups associated

54

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Soil organic matter Hydroxyl

Carboxyl Amine

COOH

-

OH

-

Carbonyl

+

NH x

CO

+

O Methoxyl O

OH

+

CH 3

-

O

Phenolic

+

Quinone

Figure 2.18

Schematic diagram showing the common types of functional groups associated with soil organic matter (from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, 2000, 307p.).

Table 2.2 Typical Composition and Common Functional Groups for Fulvic Acid, Humic Acid and Humin

Percent Percent Percent Percent Percent Percent Percent Percent Percent

carbon content oxygen hydrogen carbonyl carboxyl quinone ketones alcoholic OH phenolic OH

Fulvic Acid

Humic Acid

Humin

40–50 40–50 4–7 Up to 5 1– 6 2" 2" 2.5–4 2–6

50–60 30–40 3–6 Up to ~4 3–10 1–2 1–4 Up to 2 Up to ~4

50–60 30–35 NA NA NA NA NA NA NA

Note: NA = not applicable Source: From Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, 2000, 307p.

with soil organic matter also exhibit negative or positive charges. The carboxyl, phenolic and hydroxyl groups possess negative charges as shown in Figure 2.16. Hydrous oxides, oxides and hydroxide sheets such as brucite and gibbsite possess charges owing to association and dissociation of protons. The charges developed on

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

55

the surfaces of the oxides or hydroxide sheets are due to amphoteric dissociation of the surface hydroxyl groups or from adsorption of H+ and OH– ions (Gast, 1977). The nature of these charges is dependent on both the pH and concentration of electrolytes in the pore water. We will discuss all of these further in Chapter 3 when we address the interactions of soil particles and soil fractions with water. The total number of electrostatic charges on the clay particles’ surfaces divided by the total surface area of the particles involved provides us with the quantitative determination of the surface charge density. The common procedure is to express this surface charge density in terms of its reciprocal, as shown in Table 2.3. The hydrous oxides such as goethite [-FeOOH ] and gibbsite [-Al(OH)3] have been omitted from the table because the range of values for these types of soil fractions are dependent upon (1) their structure, (2) the speciﬁcally adsorbed potential-determining ions and (3) the pH of the pore water. 2.6.3 Speciﬁc Surface Area (SSA) We have stated previously that the amount of surface area presented by the soil particles or soil fractions for interaction with dissolved solutes or contaminants in the pore water is an important consideration — particularly if the surfaces are highly reactive. This indicates that providing more reactive sites that can react with the contaminants would produce greater interactions between the soil particles and the contaminants. The hope is that these reactions are such that they will result in assimilation of contaminants and thus result in contaminant attenuation. While we can theoretically deﬁne or calculate the surface area of individual particles assuming ideal shapes such as planar surfaces, cubes, spheres and cylinders or conical shapes, the measurement of surface areas of soil particles in any representative elementary volume (REV) requires the use of an adsorbate. It is intuitively obvious that soil particles with planar shapes exhibit the greatest surface area per unit volume or per unit mass. This means that clay minerals will have the largest surface areas in comparison to other soil fractions. Using the a and b dimensions of a unit cell for a dioctahedral 2:1 mineral such as a smectite and Avogadro’s number of such unit cells, Greenland and Mott (1985) calculated the speciﬁc surface area of such minerals to be 757 m2/g. This refers to all the surfaces available for interaction with water or other ﬂuids. Figure 2.19 gives a highly simpliﬁed schematic of available surfaces in relation to some clay minerals and the role of microstructure in modiﬁcation of available surfaces. Cleavage of the dioctahedral 2:1 mineral into the component unit layers allows all the surfaces of the unit layers to interact with pore water (Figure 2.19 “A”). This is one of the characteristics of swelling soils that make them attractive as part of a buffer or liner material for containment of highlevel nuclear wastes in repositories. As noted in Figure 2.19 “B,” the intact mineral particle (e.g., a kaolinite mineral particle) has considerably less surface area. When combined with other particles in a microstructural unit, the total available surface area becomes less than the sum of all the particles involved because of inaccessibility to ﬂuid conductivity. Wetting of the pore channels deﬁned between microstructural units is conﬁned to those surfaces that deﬁne the pore channels used for ﬂuid movement. This is particularly signiﬁcant

56

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Table 2.3 Charge Characteristics, SSA and CEC for Some Clay Minerals

Soil Fraction Kaolinite

Range Cation of Exchange Capacity Surface Charge (meq/ (CEC), Area 100 g) meq/100 g (m2/g)

Reciprocal of Charge Density (nm2/ Isomorphous charge) Substitution

5–15

10–15

5–15

0.25

Clay micas and chlorite

10–40

70–90

20–40

0.5

Illite

20–30

80–120

20–40

0.5

Montmorillonite1

80–100

800

80–100

1.0

100–150

700

100–150

1.0

Vermiculite2

Source of Charges

Dioctahedral; Surface silanol, 2/3 of positions edge silanol ﬁlled with Al and aluminol groups (ionization of hydroxyls and broken bonds) Dioctahedral: Silanol groups, Al for Si plus Trioctahedral or isomorphous mixed Al for substitution Mg and some broken bonds at edges Usually Isomorphous octahedral substitution, substitution Al silanol groups for Si and some edge contribution Dioctahedral; Primarily from Mg for Al isomorphous substitution, with very little edge contribution Usually Primarily from trioctahedral isomorphous substitution Al substitution, for Si with very little edge contribution

1

Surface area includes both external and intralayer surfaces. Ratio of external particle surface area to internal (intralayer) surface area is approximately 5:80. 2 Surface area includes both external and internal surfaces. Ratio of external to internal surface area is approximately 1:120. Note: Ratios of external:internal surface areas are highly approximate since surface area measurements are operationally deﬁned, i.e., they depend on the technique used to determine the measurement. Source: Yong, R.N. and MacDonald, E.M., Inﬂuence of pH, metal concentration, soil component removal on retention of Pb and Cu by an illitic soil, in Adsorption of Metals by Geomedia, Jenne, E.A., Ed., Academic Press, San Diego, CA, 1998, Ch. 10, pp. 230–254.

in the control of hydraulic conductivity, as discussed in Section 2.5 with respect to Sw, the wetted surface area. Laboratory measurements are often made to determine the speciﬁc surface area of soils since theoretical calculations are not only tedious but also unrealistic if the soil contains different soil fractions of varying proportions and sizes. The

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

Figure 2.19

57

Surface areas of mineral particles and reduction of available surface areas owing to microstructural units and aggregation into macrostructure.

general procedure is to determine the amount of gas or liquid (adsorbate) that forms a monolayer coat on the surfaces of the particles. The choice of the adsorbate and the availability of soil particles in a totally dispersed state are the two most important factors in any laboratory measurement of the SSA of a soil sample. Accordingly, this makes a laboratory determination of the SSA an operationally deﬁned property; i.e., dependent on adsorbate used and degree to which the soil has been properly dispersed, one could easily obtain different SSA values for the same soil sample. When the adsorbate is a gas, the amount of gas (e.g., number of molecules of nitrogen) sorbed by the soil particles is dependent on the partial pressure of the gas and on the temperature. In the test procedure, more than one layer of gas is sorbed by the soil particles. To determine the amount (volume) of gas equivalent to a sorbed monomolecular layer of gas, the relationship developed by Brunauer et al. (1938) known as the BET equation for multilayer sorption is probably the best known relationship used. For adsorbates that are polar ﬂuids, the techniques described by Mortland and Kemper (1965) for ethylene glycol and by Carter et al. (1986) for ethylene glycol-monoethyl ether are commonly used. Since all the techniques require a uniform monolayer coating by the adsorbate of all the soil particles, it is easy to see that the values of SSA determined are directly dependent on the procedure used and the analytical technique used to reduce the data obtained. These can all lead to differences in results. Failure to provide for a totally dispersed state of soil particles and failure to obtain uniform monolayer distribution or coating of the dispersed particles are two of the more signiﬁcant factors. A further consideration is the presumption that the interactions

58

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

between pore water and the soil particles’ surfaces mimic those exhibited by the adsorbates.

2.7 CONCLUDING REMARKS To recognize whether natural attenuation of contaminants in soils can occur or how effective the process is, it is important to realize that the interactions or processes that occur in the soil depend not only on the properties and characteristics of the contaminants but also on the properties and characteristics of the soils. All too frequently, assessments and evaluations of natural attenuation of contaminants overlook the fact that the abiotic and biotic reactions between contaminants and soils need the same detail in determination of the various attributes of the soil that participate in those reactions. A large factor in determining whether natural attenuation can be successfully applied as a remediation technique is the consistency or uniformity of the soil within which transport of the contaminants will occur. In describing and predicting transport of contaminants in soils, we need to use transport coefﬁcients or values that describe the transmission characteristics or properties of the soil, e.g., hydraulic conductivity or void ratio. Most often, a single value is used to characterize a soil, even though it is known that the characteristic values vary with space and time in the soil. It is not unusual to ﬁnd that holes dug to measure the depth of shallow water tables may show values of 1 and 3 m in two holes only 3 m apart. While gradients of this magnitude should produce a level water table, obviously, instantaneous water table levels are very different. Measurements of saturated hydraulic conductivity at 1 m depth over an area of 1000 m2 typically show a range of values that vary at least one order of magnitude. One needs to question the validity of an average of these numbers, or the validity of a single measurement? It is not unusual to have measurements of penetration of a chemical into the soil show the considerable variability — with highly variable concentrations at adjacent points only a few meters apart. Evidence of this magnitude of variability is one of the reasons why it is difﬁcult to use mechanistic or deterministic models to describe transport of contaminants in soils and why statistical or probabilistic models offer more promise for future developments. Care needs to be exercised in ascribing distribution functions to the various properties and characteristics of soils. Experience has shown that the values of hydraulic conductivity and bulk densities for a typical site occupying for example, 1 hectare, have different distribution functions. Bulk density values are generally symmetrically distributed about the mean in a Gaussian or normal distribution. This permits calculations of the mean and standard deviation. The same cannot be said for the hydraulic conductivity of the soil in the 1-hectare site. Most often, one will obtain values of hydraulic conductivity that are not symmetrically distributed. This indicates that Gaussian statistics are generally not valid. Instead, distributions such as the Poisson may be more useful. It is often argued that the modal value rather than the mean should be used for characteristics not ﬁtted with Gaussian or normal distribution. The concept of modal characteristics of a soil for classiﬁcation takes it cue from experience that suggests

SOIL COMPOSITION AND TRANSMISSION PROPERTIES

59

that this is the appropriate scheme for classiﬁcation of most natural things with large variability. While a prediction of the disappearance of a chemical from a volume of soil may depend upon the modal value of hydraulic conductivity, contamination of a particular site depends upon the larger value. This chapter deals with the nature of soils and pays particular attention to those soil fractions that contribute directly to the assimilative capacity of soils. Only those properties of soils deemed pertinent in the various processes associated with transport of leachates in the soils are considered. While the many other physical properties of soils such as shear strength, compactibility, consolidation, etc. are to some extent important in considerations of buffer-liner and barrier systems for landﬁlls and impoundments, they are not addressed here since they are secondary to the interactions considered in transport processes. The main items and issues considered include • We need to determine pertinent soil features (origin, geologic and regional controls, soil fractions and compositional control on development of soil structure), properties and characteristics that are important in control of the interactions between the soil solids and contaminants that result in the attenuation of contaminants. • Chemically reactive groups, deﬁned as surface functional groups, associated with the surfaces of the various soil fractions make these surfaces reactive. The presence of soil particles with reactive surfaces creates the situation where these particles react chemically with solutes and other dissolved matter in the pore water. The soil fractions with reactive surfaces include layer silicates (clay minerals), soil organics, hydrous oxides, carbonates and sulfates. • Mechanisms associated directly with contaminant assimilation and attenuation need to be optimized in a soil-water system. This means that all the available soil solids’ surfaces should be made available, resulting in maximization of surface area contact with the solutes in pore water and uptake — i.e., partitioning of the contaminants. • To determine the reactive surfaces of soil particles, the techniques used lead to results that are operationally deﬁned. Values of SSA are directly dependent on the procedure used and the analytical technique used to reduce the data obtained.

REFERENCES Bouma, J., Effect of soil structure, tillage and aggregation upon soil hydraulic properties, in Interacting Processes in Soil Science, Wagenet, R.J., Baveye P., and Stewart, B.A., Eds., Lewis Publishers, Boca Raton, FL, 1991, pp. 1–36. Brunauer, S., Emmett, P.H., and Teller, E., Adsorption of gases in multimolecular layers, J. Am. Chem. Soc., 60, 309–319, 1938. Carter, D.L., Mortland, M.M., and Kemper, W.D., Speciﬁc surface, in Methods of Soil Analysis, Part. 1: Physical and Mineralogical Methods, A Klute, Ed., Monograph 9, Am. Soc. Agron., 1986. Dixon, J.B., Kaolinite and serpentine group minerals, in Dixon, J.B. and Weed, S.B., Eds., Minerals in Soil Environments, Soil Science Society of America, 1977, pp. 357–403.

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Douglas, E. and Yong, R.N., Validity of Darcy’s law for one-dimensional, unsaturated ﬂow in kaolinite-water-NaCl systems, Trans. ASAE, 24, 657–662, 1981. Flaig, W., Beutelspacher, H., and Reitz, E., Chemical composition and physical properties of humic substances, in Soil Components, Gieseking, J.E., Ed., Springer-Verlag, Berlin, 1, 1–219, 1975. Follet, E.A.C., The retention of amorphous, colloidal “ferric hydroxide” by kaolinites, J. Soil Sci., 16, 334–341, 1965. Gast, R.G., Surface and colloid chemistry, in Minerals in Soil Environments, Dixon, J.B. and Weed S.B., Eds., Soil Science Society of America, WI, 1977, pp. 27–44. Greenland, D.J. and Mott, C.J.B., Surfaces of soil particles, in The Chemistry of Soil Constituents, Greenland, D.J. and Hayes M.H.B., Eds., John Wiley and Sons, New York, 1985, 409 pp. Grim, R.E., Applied Clay Mineralogy, McGraw-Hill, New York, 1962, 422 pp. Grim, R.E., Bray, R.M., and Bradley, W.F., The mica in argillaceous sediments, Am. Mineral., 22, 813–829, 1937. Hansbo, S., Consolidation of clay, with special reference to inﬂuence of vertical sand drains, Proc. Swed. Geotech. Inst.,18, 41–61, 1960. Hayes, M.H.B. and Swift, R.F., The chemistry of soil organic colloids, in The Chemistry of Soil Constituents, Greenland, D.J. and Hayes M.H.B., Eds., John Wiley & Sons, New York, 1985, 409 pp. Lambe, T.W., The structure of inorganic soil, Proc. ASCE, 315., 1953. Lambe, T.W., The structure of compacted clay, J. Soil Mech. Foundation Div. ASCE, 84, (SM2), 34., 1958. Lambe, T.W., A mechanistic picture of shear strength in clay, Proc. ASCE Research Conference on Shear Strength of Cohesive Soils, 1960, pp. 555–580. Lutz, J.F. and Kemper, W.D., Intrinsic permeability of clay as affected by clay-water interactions, Soil Sci., 88, 83–90, 1959. Mortland, M.M. and Kemper, W.D., Speciﬁc surface, in Methods of Soil Analysis, Monograph 9, Am. Soc. Agron., 1965, pp. 532–544. Ohtsubo, M., Yoshimura, A., Wada, S-I, and Yong, R.N., Particle interaction and rheology of illite-iron oxide complexes, Clays Clay Miner., 39, 347–354, 1991. Philip, J.R., The physical principles of soil water movement during the irrigation cycle, Congr. Inter. Comm., Irrigation and Drainage, 8, 125–154, 1957. Pusch, R., Quick clay microstructure, J. Eng. Geol., 3, 433–443, 1966. Schwertmann, U. and Taylor, R.M., Iron oxides, in Minerals in Soil Environments, Dixon, J.B. and Weed, S.B., Eds., Soil Science Society of America, Madison, WI, 1977, pp. 145–180. Sposito, G., The Surface Chemistry of Soils, Oxford University Press, New York, 1984, 234 pp. Terzaghi, K. and Peck, R.B., Soil Mechanics in Engineering Practice, Wiley & Sons, New York, 1948, 566 pp. Weaver, C.E. and Pollard, L.D., The Chemistry of Clay Minerals, Elsevier, Amsterdam, 1973, 213 pp. Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, 2000, 307p. Yong, R.N., Inﬂuence of microstructural features on water, ion diffusion and transport in clay soils, Appl. Clay Sci., 2003, 23, 3–13. Yong, R.N. and MacDonald, E.M., Inﬂuence of pH, metal concentration, soil component removal on retention of Pb and Cu by an illitic soil, in Adsorption of Metals by Geomedia, Jenne, E.A., Ed., Academic Press, San Diego, CA, 1998, Ch. 10, pp. 230–254.

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Yong, R.N. and Mourato, D., Extraction and characterization of organics from two Champlain Sea subsurface soils, Can. Geotech. J., 25, 599–607, 1988. Yong, R.N. and Mourato, D., Inﬂuence of polysaccharides on kaolinite structure and properties in a kaolinite-water system, Can. Geotech. J., 27, 774–788, 1990. Yong, R.N. and Mulligan, C.N., The impact of clay microstructural features on the natural attenuation of contaminants, Proc. Workshop Clay Microstructure and Its Importance to Soil Behaviour, Lund, Sweden, 2002. Yong, R.N and Mulligan, C.N. (2003) The impact of clay microstructural features on the natural attenuation of contaminants. Appl. Clay Sci., 23, 179–196. Yong, R.N. and Warkentin, B.P., Introduction to Soil Behavior, MacMillan, New York, 1966, 451 pp. Yong, R.N. and Warkentin, B.P., Soil Properties and Behavior, Elsevier, Amsterdam, 1975. 449 pp.

CHAPTER 3 Soil-Water Systems and Interactions 3.1 INTRODUCTION In Chapter 2, we deﬁned a soil-water system to mean a soil mass that includes soil solids and the pore water contained therein. We should add that this also means the inclusion of a gaseous phase — since not all soils are fully saturated with a ﬂuid phase at all times. We have discussed (1) the nature of soil, (2) what constitutes a soil-water system and (3) what physical properties are inﬂuential in the control of the hydraulic conductivity of soils. We will be interested in continuing the discussion of the nature of soil by examining the interactions between soil particles (soil solids) and water. Why? Because an understanding of the various mechanisms of interaction between the soil particles and the electrolytes in water will tell us how the reactive surfaces of the soil particles react with the chemical nature of the water in the soil pores. In particular, we need to develop a wider appreciation of the interactions between contaminants and soil particles so that we can determine the type of surface complexation model that best describes the retention and attenuation of contaminants in the soil-water system.

3.2 FUNCTIONAL GROUPS AND ELECTRIC CHARGES One of the primary attributes of clay soils with respect to contaminant attenuation is the capability of the soil to promote interactions between the contaminants and the soil fractions that make up the clay soils. Interaction between contaminants and soil fractions resulting in physical adsorption of the contaminants (physisorption) by the soil fractions is by and large electrostatic in nature — attraction established between positive and negative charges. Other interactions involving ionic and covalent bonding that result in the chemical mass transfer of contaminants are considered as chemical adsorption (chemisorption). The basic structure of many of the soil fractions and their surface functional groups provides us with the basis for formation

63

64

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

of complexes between the functional groups and the contaminants in the pore water. The nature and extent of surface complexation depend on both the reactive properties of the soil particles and the contaminants themselves. 3.2.1

Clay Mineral Particles

The discussion in Section 2.6.1 on the reactive surfaces of layer lattice clay mineral particles such as kaolinites, illites, chlorites, etc. showed that isomorphous substitution of the ions in the tetrahedral and octahedral layers by lower valence ions results in the development of electric charges on the siloxane surfaces. Electric charges can also be developed on the faces of the edges of clay particles owing to broken bonds at the edges of the particles. This results in the production of hydrous oxide types of edge surfaces. For example, the broken octahedral sheets in a kaolinite mineral particle will provide for Lewis acid sites [Al(III)·H2O], which can bind OH groups in single coordination. Figure 3.1 shows a typical kaolinite particle with: • A siloxane upper bounding surface where both silanol [SiOH] and siloxane [SiOSi ] functional groups can exist together on the surfaces of the silica tetrahedra. At this surface, the OH groups on the silica surface become the centers of adsorption of the water molecules. The surface silanol groups are weak acids. They will bond with internal silanol groups, which have the capability of establishing hydrogen bonding with water. When surface silanol groups dominate, the surface will be hydrophilic. However, when siloxane groups dominate, the surface will be hydrophobic. • A gibbsite lower bounding surface. The aluminol groups in this gibbsite lower bounding surface do not appear to affect the net negative charge distributed on the bounding surface (Greenland and Mott, 1985). • A particle edge containing silanol and aluminol groups. The Al3+ in the exposed edges of the octahedral sheets will complex with both H+ and OH– in the coordinated OH groups. However, the Si4+ will complex only with OH–.

In summary, we note that the various functional groups at the basal and edge surfaces of the inorganic soil fractions, together with isomorphous substitutions in the lattices of the layer-lattice clay minerals, result in the development of negative and positive charges distributed on the surfaces and edges of the soil particles. Table 2.3 in Chapter 2 provides us with the reciprocal of the surface charge density for the more common clay minerals. 3.2.2

Oxides and Soil Organic Matter

We noted in the previous chapter that the nature and magnitude of charges for hydrous oxides and oxides are dependent upon their basic structure and the pH of the immediate environment characterizing the system. For convenience, we will use the name oxides to include the hydrous oxides, oxyhydroxides and oxides. A signiﬁcant distinguishing feature of the various oxides such as iron, aluminum, man-

SOIL-WATER SYSTEMS AND INTERACTIONS

65

Siloxane surface as upper bounding surface, with silanol groups and excess negative charge from isomorphic substituted octahedral sheets distributed over the surface oxygens.

Aluminol groups on gibbsite lower bounding layer

Silanol and Aluminol Groups At Edge Surfaces. .... 4+

–

3+

complexes with both OH

Si complexes only with OH in silanol groups. Al

+

and H in aluminol groups.

Figure 3.1

Repeating

Siloxane bounding surface

–

Gibbsite layer as bounding surface

Silanol and aluminol groups on bounding and edge surfaces of kaolinite mineral particle (from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate, and Mitigation, CRC Press, Boca Raton, FL, 2000.)

ganese, titanium and silicon is the fact that their surfaces essentially consist of broken bonds — in contrast to the clay minerals described in the previous section. The surfaces of the oxides of iron and aluminum, for example, show coordination to hydroxyl groups, which protonate or deprotonate depending upon the pH of the surrounding medium. This means that interaction of the oxide surfaces with water is between the unsatisﬁed bonds on the oxide surfaces and the hydroxyl groups of dissociated water molecules. Both negative and positive charges exist on the oxide surfaces, the predominance of each being dependent on the pH of the system. When the sum of the negative charges equals the sum of the positive charges, the point of zero charge (pzc) is attained. This means one could switch from a net positive charge to a net negative charge when one changes the pH of the system from pH ranges below the pcz to values above the pcz. We have seen from Figure 2.16 in Chapter 2 that carbon and nitrogen combine with oxygen and/or hydrogen to form the various types of surface functional groups associated with soil organic matter. These can protonate or deprotonate depending on the aqueous environment pH. They develop positive or negative charges depending on the pH of the soil and their respective acidity or basicity constants (i.e., pKa or pKb). The acidic properties associated with the soil organics are due to the carboxyl group and the hydrogen in the oxygen-containing functional groups that can be dissociated. The carboxyl and phenolic OH groups are considered to be responsible for a signiﬁcant portion of the source of negative charge. Hayes and Swift (1985) report a range of 2 to 4 meq/g for soil organics —compared with the charge range of 0.01 to 2 meq/g for clay minerals.

66

3.2.3

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Negative and Positive Charges, Charge Reversal and Net Surface Charges

From the previous discussions, we note that there are two kinds of negative charges associated with the surfaces of soil particles: constant or permanent negative charges and pH-dependent negative charges. The constant or permanent negative charges are attributed to isomorphous substitutions and/or site vacancies in the layerlattice structures of the clay minerals and noncrystalline hydrous oxides. In the case of pH-dependent negative charges, it is important to distinguish between inorganic and organic soil fractions as contributors to these charges. The functional groups of soil organic matter such as carboxyls and phenolic groups contribute signiﬁcantly to the pH-dependent negative charges. For those inorganic fractions, there are both nonhydrolyzable and hydrolyzable types. The nonhydrolyzable pH-dependent negative charges are associated with the hydrous-oxide coatings on clay mineral particles — in contrast to the hydrolyzable pH-dependent negative charges obtained from the structural OH2 groups that result from protonation of hydroxyls at defects from removal of structural Al, Fe and Mg (Huang and Jackson, 1966). Before ending this brief discussion on pH-dependent negative charges, we should note that strictly speaking, we should refer to the surfaces associated with these charges as variablecharge surfaces — as opposed to pH-dependent surfaces. The soils that show such characteristics are identiﬁed as variable-charge soils. The positive charges associated with the surfaces of soil fractions are most commonly found in the hydrous oxides and hydroxides of Al, Fe and Mn and in the broken edges of clay minerals. Charge reversal refers to the situation in which the net charges on a particle surface (i.e., charge density) change from a positive to negative or vice versa when one progresses from a system pH below the pcz to above the pcz. This is a characteristic of soil fractions such as hydrous oxides and kaolinites. We should distinguish between the pzc, and the isoelectric point (iep). We previously deﬁned the pcz to mean the instance (pHpzc) when the positive and negative surface charges are equal in number. In essence, the pcz is the pH at which titration curves of the candidate soil solution intersect at the zero point of adsorption densities, i.e., when the sum of the adsorption densities of H+, (GH) and OH– (GOH) is zero. We recall that in portraying the data from titrations, the curves begin on the positive portion of the {GH – GOH} ordinate at low pH values and progressively decrease to minus portion as the pH increases as seen in Figure 3.2 for a kaolinite soil at constant KCl concentration and with Pb2+ added as Pb(NO)3. The iep refers to the pH at which the electric potential developed at the solidliquid interface, as a result of movement of colloidal particles in one direction and counterions in the opposite direction, becomes zero. At this point, a balance between positive and repulsive energies can be obtained. While one can sometimes obtain accord between the pH pzc and pH iep, this is unusual since these values are operationally deﬁned. 3.2.4

Electriﬁed Interface and Interactions

The net charges on the surfaces of the clay particles constitute an electriﬁed interface in reaction with a solution containing anions and cations. Because the net

SOIL-WATER SYSTEMS AND INTERACTIONS

67

2 2+

Pb concentration 0 ppm 100 ppm

1.5

300 ppm 400 ppm

- GOH (cmol/kg)

1

0.5

0

GH

+

-0.5 Kaolinite pH pzc between 4.1 and 4.2

-1

-1.5 2

3

4

5

6

7

8

9

10

11

pH

Figure 3.2

Titration curves for kaolinite at constant KCl concentration and with Pb2+ added in the form of Pb(NO)3. (From Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate, and Mitigation, CRC Press, Boca Raton, FL, 2000).

charge on clay particle surfaces is more likely to be negative, cations in solution are attracted to the reactive surfaces. The nature of the distribution of the various ions in the solution is conditioned by the various reactions between these ions and the charged surfaces. These reactions are both electrostatic and chemical. They apply to contaminants introduced into the solution and are of extreme importance in the determination of the natural attenuation (assimilative) capability of the soil. The interactions generated are portrayed in a general simpliﬁed scheme shown in Figure 3.3. The partly hydrated cations and anions in the inner Helmholtz plane are potential determining ions (pdis) and are bonded to the reactive surfaces by ionic and covalent bonds. pdis contribute directly to the charge and potential on the surfaces of reactive particles. Although there are some minor differences in details concerning the nature and distribution of the various ions adjacent to the reactive surfaces (e.g., Stern, 1924; Grahame, 1947; Kruyt, 1952; Sposito, 1984; Greenland and Mott, 1985; Ritchie and Sposito, 2002), there is general agreement on (1) the altered or structured water layers adjacent to the reactive surfaces and (2) the swarm of counterions forming a diffuse layer of ions. The modiﬁcation of the Stern layer model by Grahame (1947) to include the inner Helmholtz plane (ihp) and outer Helmholtz plane (ohp) as shown in Figure 3.3 is depicted in a more generic form in Figure 3.4 to illustrate the results of various interactions emanating outward from the surface of the reactive particle. The structured water adjacent to the surface of the particle is due to the speciﬁcally adsorbed ions at the interface. Ritchie and Sposito (2002) maintain that the complexes formed

68

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Clay particle with reactive surface

Plane of center of specifically sorbed anions, IHP

Partly hydrated cation

Cation

Water molecule Anion Plane of center of first layer of hydrated cations, OHP

Plane of partly hydrated specifically sorbed cations, IHP OHP = outer Helmholtz plane; IHP = inner Helmholtz plane

Figure 3.3

Speciﬁcally sorbed ions, IHP and OHP. Left-hand drawing shows partly hydrated speciﬁcally sorbed cations and fully hydrated cations. Right-hand drawing shows anions as speciﬁcally sorbed anions and hydrated cations. The speciﬁcally sorbed ions are potential determining ions (pdi’s).

between the surface functional groups and the ions at the interface are inner sphere complexes. This assumes direct contact between them without interruption from any water molecule. When a layer of water molecules interrupts contact between these, the complexes are classiﬁed as outer sphere complexes. The IHP and OHP shown in Figure 3.3 identify the positions of the inner sphere and outer sphere complexes with corresponding distances of c and b, respectively (Figure 3.4). The thickness of the Stern layer is obtained as d = c + b. Electrostatic bonding mechanisms for the counterions beyond the Stern layer required to satisfy the net negative charge of the reactive particles result in the formation of a diffuse ion layer. The surface charge ss associated with the surface of the particle shown in Figure 3.4 is balanced by the sum of the Stern layer charge sd. at the ohp and the diffuselayer charge sddl. The surface potential ys, which is also associated with the surface of the particle, varies with electrolyte concentration and the nature of the charge of the soil particle, i.e., constant charge surface or a pH-dependent charge surface. This surface potential drops from ys at the surface to a potential yd at the ohp, as shown in Figure 3.4. The potential y (electric potential) beyond the ohp can be described by the Gouy Chapman diffuse double-layer model, as will be seen in the next section. From the viewpoint of electrokinetics, yd is considered to be equal (or almost equal) to the zeta potential z.

SOIL-WATER SYSTEMS AND INTERACTIONS

69

The Coulombic interaction established between the ions and forces associated with the negative charged sites on the reactive surfaces of the particles in the ihp can be calculated as Eihp (Yong, 2000): Eihp =

zi e 2 eR

(3.1)

For the interaction energy Eohp in the OHP, the four energy components (Yong, 2000) include (1) Coulombic interaction energy Ec, (2) ion-dipole interaction Eid, (3) dipole-dipole interaction Edd and (4) dipole-site interaction Eds. Eohp =

zi e 2 me + eR er 2

Ê Dn m 2 ˆ me 1 Á ˜+ er 3 ¯ er12 Ë

(3.2)

where zi = valence of the ith species of ion in the bulk solution, e = electronic charge, e = dielectric constant, R = distance between the center of ion i and the corresponding negative charge site on the reactive particle, r = sum of the radii of the ion and water molecule, r1 = distance between center of the dipole and corresponding negative charge site on the particle, m = dipole moment of a water molecule = 10–10 ESU cm, and Dn = geometrical factor (0.334 for three water molecules and 1.188 for six water molecules). 3.2.5

Interactions and Surface Complexation Models

The degree of interaction between contaminants and soil particles and the degree of interaction between proximate particles depend on how one views the interactions immediately adjacent to the reactive surfaces. The energy status of a representative element of a soil-water system is deﬁned by the level of interactions established. The total potential R shown in Figure 3.4 is a characteristic of the interaction energy. Evaluation of the complexes formed between the surface functional groups of the soil particles and the ions in the pore water can be performed using a variety of surface complexation models such as the single-layer, double-layer and triple-layer models. The details of the more popular complexation models can be found in several texts, e.g., Kruyt (1952), Singh and Uehara (1986), Bockris and Reddy (1970) and Yong (2000). The Gouy-Chapman diffuse double layer (GC-DDL) model is perhaps the most familiar model of the interaction and distribution of ions. The model provides the capability of computing the potential y shown in Figure 3.4 in relation to the distance d from the charged particle surface. Because of bonding processes associated with chemical bonding and complexation in the d region shown in Figure 3.4, the simple electrostatic interaction calculations do not apply in the vicinity of the charged particle surfaces, i.e., the GC-DDL model does not provide an accurate description of y immediately adjacent to the charged particle surface. The following assumptions and conditions for determination of y are generally used:

70

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Inner Helmholtz Plane (ihp)

Surface of clay particle

Outer Helmholtz Plane (ohp)

;s Diffuse ion layer

;@

Clay mineral particle with reactive surface

Potential ;

;

Electrified interface

?

>

Distance d from particle surface

@ Figure 3.4

Generalized model of electriﬁed interface with aqueous solution containing dissolved solutes.

• Interaction between the charged ions and solutes (cations and anions) in solution and the charged (soil) particle surfaces are Coulombic in nature. • The ions in solution are considered to be point-like in nature, i.e., zero-volume condition. • The density of the charges r owing to the assumed point-like ions that contribute to the interactions (i.e., space charge density) can be described by the Boltzmann distribution.

The relationship for y can be obtained (e.g., Kruyt, 1952; van Olphen, 1977; Yong, 2000) from

y=-

Êd 2kT ln cothÁ e Ë2

8pe 2 zi2 ni ˆ ˜ ekT ¯

(3.3)

where k = Boltzmann constant, ni = concentration of the ith species of ion in the bulk solution, and T = temperature. The relationship between the surface charge density ss and surface potential ys is

SOIL-WATER SYSTEMS AND INTERACTIONS

71

1

ze Ê 2n ekT ˆ 2 ss = Á i ˜ sinh i y s Ë p ¯ 2kT 3.2.6

(3.4)

Applications and Chemical Speciation

For those whose interests lie primarily in “knowing” whether natural attenuation of contaminants can or will occur in candidate sites and soils, there are controlled laboratory attenuation tests that provide direct experimental evidence on samples obtained from the candidate sites. How representative these tests are and whether the results can be replicated in the ﬁeld are questions that cannot be easily answered without a better understanding of the interactions and processes involved. For those who are interested in determining (1) the degree (scale) of natural attenuation, (2) the timescale required to reach certain levels of attenuation, (3) the effectiveness of natural attenuation and (4) the lateral extent or plume generation required to reach satisfactory attenuation remediation criteria, a knowledge of the basic interactions between contaminants and soil particles will provide the tools to evaluate and predict transport performance, partitioning and chemical mass transfer. For abiotic processes, much of the information required can be obtained via determination of chemical speciation. Prediction of chemical speciation in respect to contaminants in soils, using specially developed computer models, has received considerable attention in recent years. These models, which include PHREEQE (Parkhurst et al., 1980), PHREEQM (Appelo and Postma, 1993), GEOCHEM (Sposito and Mattigod (1979), SOILCHEM (Sposito and Coves, 1988), EQ3 (Wolery, 1983) and MINTEQA2 (Allison et al., 1991), are well described by their respective developers. The key elements of the models include calculation procedures based on assumption of thermodynamic equilibrium for speciation of soluble complexes, dissolution and precipitation of minerals, and sorption of ions onto soil particles’ surfaces using a variety of surface complexation models. Knowledge of the outcome of competition between the functional groups on the reactive surfaces of soil particles and the various species of ions in solution for formation of complexes is fundamental to the successful determination of attenuation via soil assimilation.

3.3 SOIL-WATER ENERGY CHARACTERISTICS Interactions between soil solids, water and the various dissolved solutes in the pore water can be characterized in terms of energy relationships. These soil-water energy relationships are generally known as soil-water characteristics. They essentially inform one about the water-holding capacity of a soil. As one might expect, the chemistry of the pore water and the surface properties of the soil solids are key factors in the demonstration of the water-holding capacity of a soil — with the latter being more signiﬁcant. A useful demonstration of this can be seen by performing a simple experiment with a Buchner-type pressure membrane system

72

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Open to atmosphere

A Directional valve

Air pressure

Cover plate Water (test fluid) Soil sample Open to atmosphere Ceramic porous stone

B

Capillary tube

To vacuum pump system

Figure 3.5

Buchner-type pressure-tension apparatus for determination of soil water–holding capacity. Position of directional valves A and B shown on diagram are for pressure membrane-type test.

(Figure 3.5) containing different kinds of soils permeated with water with and without dissolved solutes. For the situation shown in Figure 3.5, the directional valves are oriented to allow pressure to be applied to the test ﬂuid (generally water). Switching the valves to allow for water extraction via a vacuum pump system attached to the outlet would also permit determination of the water-holding characteristics of the test soil. Figure 3.6 illustrates the differences that might occur in the general water-holding pattern (water retention curves) for three typical kinds of soils: clay, loam soil and sand. To understand the concepts of water-holding capacity, water retention, energy relations between soil solids and water, soil-water potential y and how all of these relate to reactive surfaces, assimilation of contaminants, and partitioning, it is necessary to restate a few fundamental points: • The reactive surfaces of soil fractions or soil solids provide the sites for reactions between the contaminants (ions, dissolved solutes, etc.) in the pore water of the soil. • The nature of the bonding forces between the contaminants and the soil solids and the energy of interaction established depend on both the surface functional groups associated with the soil solids and the contaminants. • The water-holding capacity of soil and the water retention characteristics of soils are direct functions of the energy of interaction between soil solids and water. These tell us about the forces and interactions between water and the dissolved solutes in the water.

SOIL-WATER SYSTEMS AND INTERACTIONS

Figure 3.6

73

Typical desorption curves obtained using pressure-membrane apparatus.

We go back to Figure 3.6 and consider the bottom water retention curve for sand to highlight some of the preceding fundamental points. First of all, we understand that in the absence of surface functional groups, the primary mechanism for water retention in sands is via capillary forces. If we conduct a simple “height of capillary rise” experiment using a vertical column of clean sand, it will become clear that the height of capillary rise h in the column of sand is determined by the density or packing of the soil solids (i.e., by the effective radius r of the pores in the soil). The relationship between h and r is h = 2swcosa/rgw, where sw = surface tension of water, gw = density of water, and a = contact angle between the wetted soil-particle surface and water. This, of course, presumes that the shape of the soil voids can be well approximated by capillary tubes of consistent bore, which is clearly not the case. Nevertheless, the capillary tube model tells us that the smaller the value of r, the greater the height of capillary rise, h. This is important since it gives us a perception of a simple water-holding fact associated with particles that do not have reactive surfaces. Furthermore, this also tells us why the desorption test procedure using a system such as that depicted in Figure 3.5 and the water retention curves shown in Figure 3.6 are important tools in the characterization of soil water-holding relationships. To describe the water-holding capability of the sand, we can deﬁne a capillary potential, yc, as a measure of the energy by which water is held by the soil particles by capillary forces. This recognizes that the nature of the soil voids do not resemble uniform capillary tubes. Buckingham (1907) has deﬁned this as the potential owing to capillary forces at the air-water interfaces in the soil pores holding water in the soil. It is a measure of the work required per unit weight of water to remove the

74

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

water from the mass of soil. The various contact angles established between water with the wetted soil particles will be directly related to soil type and soil density. For the other two water retention curves shown in Figure 3.6, i.e., the curves for the loam soil and clay, it is obvious that the surface forces associated with the soil solids, discussed in Sections 3.2.4 and 3.2.5, contribute signiﬁcantly to the increase in water retention and water holding capacity of the soils. A useful way to view the energy relationships between soil particles and water is to consider the work required to move water into and out of a soil mass — using the pore water in soil as the frame of reference. Accordingly, we deﬁne the total work required to move the water into and out of the soil as the soil-water potential y. The importance of the frame of reference is best illustrated by considering the height of capillary rise in the sand column previously described. Using the soil particles as the frame of reference, the capillary potential in respect to the soil solids — instead of the soil pore water — is best described by using soil suction to indicate the role of the soil solids in moving water into or out of the soil. The forces required to move the water are expressed in positive units — as opposed to those associated with the capillary potential yc, which uses the pore water as the frame of reference. In the case of the capillary potential yc, the magnitude of the units is the same as the soil suction but opposite in sign as shown, for example, in Figure 3.6. The yc is a negative quantity (in terms of the pore water). The curve for the sand sample shows that the yc becomes less negative in value as the water content increases. 3.3.1

Components of Soil-Water Potential

The soil-water potential y, which characterizes the energy state of soil water, varies throughout a soil mass because the forces acting on the soil water at any one point in the soil are never uniformly distributed. Using a pool of free water at the same elevation and temperature of a soil mass under one atmospheric pressure as a reference base, we can deﬁne y as the total work required to move a unit quantity of water from the reference pool to the point under consideration in the soil. It is a negative number. The different components that make up a total y include two major components and three components that are generally deemed to be of lesser importance. These two major components of the y, which have the greatest inﬂuence on the water-holding capacity of soils, are the matric ym and osmotic yP potentials. They are described as follows: • Matric potential ym. This is a property of the soil matrix and pertains to sorption forces between soil fractions and soil-water. For granular materials, the matric potential is the capillary potential. However, for clay soils, this is not the case because of microstructural units, surface functional forces and reactive surfaces (see Figure 3.7). • Osmotic potential yP or solute potential ys. For nonswelling soils, the osmotic potential is the solute potential, yP = ys = nRTc, where n = number of molecules per mole of salt, c = concentration of the salt, R = universal gas constant and T = absolute temperature. In swelling soils, the yP is used in recognition of the singular characteristic of intralamellar or interlayer swelling such as that demonstrated by montmorillonites. For nonswelling clay soils, published literature shows interchangeable usage of the ys and the yP to mean the same thing.

SOIL-WATER SYSTEMS AND INTERACTIONS

75

“Exploded” view of idealized four-particle unit Simplified edge view of idealized four-particle unit

Clay particle Interaction of overlapping diffuse ion-layers Energy characteristic defined by osmotic or solute potential OF Hydration water layer Energy characteristic defined by matric potential

Om

Figure 3.7

Schematic drawing showing exploded view of idealized four-particle unit illustrating hydration water layer and interaction of overlapping diffuse ion layers. The location of the associated components of soil-water potential (matric and osmotic/solute) is shown.

In swelling soils such as those used as buffer material in waste isolation barriers, water uptake by an initially dry soil is into the interlayer spaces. (The diagram for montmorillonite in Figure 2.8 in Chapter 2 shows the interlayer spaces.) Water uptake into the interlayer spaces is initially due to hydration forces. This hydration water resides in the ihp and ohp (Figures 3.3 and 3.4) and is commonly accepted to be a different form of water structure. The volume expansion associated with this hydration water is commonly deﬁned as crystalline swelling. This mechanism of water uptake associated with the matric potential ym does not require the capillary forces or processes related to the presence of air-water interfaces. It should be noted that interlayer or interlamellar swelling beyond crystalline swelling stems from interactions described by the osmotic or solute potential yP or ys.

3.4 CHEMICAL REACTIONS IN PORE WATER The surface functional groups associated with the various soil fractions and the ions and other dissolved solutes such as naturally occurring salts in the pore water

76

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

of a soil react chemically when brought together as a wet soil mass. For convenience, we will refer to the various soil fractions as soil solids. To understand the mechanisms leading to chemical mass transfer of contaminants and pollutants (partitioning and attenuation) it is necessary to bear in mind that the chemistry of the pore water is intimately linked to the chemistry of the pollutants/contaminants and the reactive surfaces of the soil solids. Interactions between contaminants, pollutants and soil solids involve many different sets of chemical reactions — including biologically mediated chemical reactions. The pH of the soil-water system and the other dissolved solutes in the pore water inﬂuence the various interaction mechanisms such as acidbase reactions, speciation, complexation, precipitation and ﬁxation. 3.4.1

Acid-Base Reactions and Hydrolysis

The pore water in a wet soil, without dissolved solutes and other contaminant ions, is by itself a solvent that can be either a protophillic or a protogenic solvent. It can function as an acid or as a base. Through self-ionization, it can produce a conjugate base OH– and a conjugate acid H3O+. The deﬁnition of an acid as a aqueous substance that dissociates to produce H+ ions and a base as an aqueous substance that dissociates to produce OH– ions is better viewed in terms of proton donation or acceptance. This avoids the consideration of acids and bases as aqueous substances and allows the categorization of nonaqueous substances in terms of acids and bases. The Brønsted-Lowry deﬁnition of an acid as a substance that has a tendency to lose a proton (H+) and a base as a substance that has a tendency to accept a proton allows us to do just that. This (Brønsted-Lowry) acid-base concept considers an acid as a proton donor (protogenic substance) and a base as a proton acceptor (protophillic substance). Since water has the capability to both donate and accept protons (i.e., both protogenic and protophillic), it is called an amphiprotic substance. Hydrolysis, which is an acid-base reaction, refers to the reaction of H+ and OH– ions of water with the solutes and ions present in the pore water and is basically a neutralization process. The presence of ionized cations and anions associated with the soil solids in a soil-water system results in pH levels in the soil-water system that vary from below neutral to above neutral pH values dependent on the strength of ionization of the ions. Hydrolysis reactions in a soil-water system continue as long as the reaction products are removed from the system, for example, through processes associated with precipitation, complexation and sorption. 3.4.2

Oxidation-Reduction (Redox) Reactions

Abiotic and biotic oxidation-reduction (redox) reactions occur in the pore water in soils. Oxidation-reduction reactions involve the transfer of electrons between the reactants. The activity of the electron e_ in the chemical system typiﬁed by the reactive soil particles and contaminants in the pore water is of particular importance. A measure of this electron activity is the redox potential Eh and is given as Eh = pE Ê Ë

2.3 RT ˆ F ¯

SOIL-WATER SYSTEMS AND INTERACTIONS

77

E = electrode potential, R = gas constant, T = absolute temperature and F = Faraday constant. The mathematical term pE is the negative logarithm of the electron activity e–. The relationship between Eh and pE at a standard temperature of 25°C is RT ˆ ai,ox Eh = 0.0591 pE = E 0 + Ê ln Ë nF ¯ ai,red

(3.5)

where E0 refers to the standard reference potential, n = number of electrons, and the subscripts for the activity a refer to the activity of the ith species in the oxidized (ox) or reduced (red) states. The redox capacity is a measure of the number of electrons that can be added or removed from the soil-water system without a measurable change in the Eh or pE and is comparable to the buffering capacity, which measures the amount of acid or base that can be added to a soil-water system without any measurable change in the system pH. Because bacteria in the soil utilize oxidation-reduction reactions to extract the energy required for growth, they essentially function as catalysts for reactions involving molecular oxygen and soil organic matter and/or organic chemicals. Electron transfer in a redox reaction is generally accompanied by proton transfer. Redox reactions result in the decrease or increase in the oxidation state of an atom in the case of inorganic solutes. This has signiﬁcance particularly for those pollutants that have multiple oxidation states (e.g., chromium and arsenic), where increases or decreases in the oxidation state alter the toxicity of the inorganic pollutant. In the case of organic chemical pollutants, redox reactions result in the gain or loss of electrons in the chemical.

3.5 INTERACTIONS, EXCHANGES AND SORPTION The interactions between dissolved solutes such as ions and compounds (pollutants and nonpollutants) in the pore water and the soil solids, described in previous sections, can lead to a host of very interesting phenomena such as transformations, precipitations, dissolution, exchanges, complexation and sequestering — to name a few. For convenience, we will refer to all the dissolved solutes, ions and compounds in the pore water as contaminants. Many of these processes result in partitioning of the contaminants, i.e., chemical and physical mass transfer of the contaminants from the pore water to the surfaces of the soil solids. In general, the primary mechanism for partitioning is sorption by the soil solids. The forces of interaction involved in sorption include short-range chemical forces such as covalent bonding and longrange forces such as electrostatic forces. The literature commonly considers sorption as the primary mechanism involved in partitioning of contaminants. This is a convenient means to avoid the tedious task of trying to determine which of the various processes such as physical adsorption, chemical adsorption (chemisorption) and precipitation are involved in the partitioning process.

78

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Specifically adsorbed counterions in this zone ihp

Surface of clay particle

ohp Diffuse ion layer

Os

Non-specifically adsorbed counterions in this zone

Clay mineral particle with reactive surface

O@

O

O

Electrified interface

?

>

Distance d from particle surface

@ Figure 3.8

Generalized model of electriﬁed interface with aqueous solution showing location of speciﬁcally and nonspeciﬁcally adsorbed counterions. IHP, inner Helmholtz plane; OHP, outer Helmholtz plane.

We deﬁne physical adsorption of contaminants in the pore water by soil solids as the attraction of the pollutants to the surfaces of the soil solids in response to the charge deﬁciencies of the soil fractions (i.e., soil solids). In our discussion on the diffuse double-layer (DDL) model (Section 3.2.5), we noted from the reactive surfaces of the soil solids that counterions are attracted to these reactive surfaces because of the requirement for electroneutrality. Cations and anions are speciﬁcally or nonspeciﬁcally adsorbed by the soil solids depending on whether their interactions are in the diffuse ion-layer or the Stern layer (as shown in Figure 3.8). Counterions in the diffuse ion-layer are nonspeciﬁcally adsorbed. They are held primarily by electrostatic forces and reduce the magnitude of the potential y but do not inﬂuence the sign of y. They are sometimes referred to as indifferent ions. Adsorption of alkali and alkaline earth cations by the clay minerals is a good example of nonspeciﬁc adsorption. Adsorption of cations is related to their valence and crystalline unhydrated and hydrated radii. Cations with smaller hydrated size or large crystalline size are preferentially adsorbed, everything else being equal. Speciﬁc adsorption of counterions occurs within the Stern layer (Figure 3.8). The forces involved are those associated with the electric potential within the Stern layer. Speciﬁc adsorption of cations involves bonding by covalent bonds through the O and OH groups. Sposito (1984) refers to speciﬁc adsorption as inner-sphere surface complexation of the ions in solution by the surface functional groups of the soil fractions. Speciﬁcally adsorbed ions are referred to as speciﬁc ions. They have

SOIL-WATER SYSTEMS AND INTERACTIONS

79

the ability to inﬂuence the sign of y. Examples of speciﬁcally adsorbed cations include the adsorption of various heavy metals such as Pb, Cu, Cr and Cd by the oxides and hydrous oxides of Al, Fe and Mn. Physical or nonspeciﬁc adsorption of anions is held by electrostatic or Coulombic forces in the diffuse double layer or in the Stern layer. Generally, this kind of sorption is considerably less than the adsorption capacity for cations since the soil fractions that have positive charge sites are primarily the oxides and edges of some clay minerals. Adsorption of anions is inﬂuenced by the pH of the soil-water system and the electrolyte level. Speciﬁc adsorption of anions is more a ligand exchange reaction than a sorption and involves anion displacement of OH– from the surface and incorporation as a ligand in the coordination of the structural cations (Bolt, 1978). 3.5.1

Bonding and Sorption Mechanisms

Bonds are developed as a result of interactions between charged particles of the various soil fractions and also between these particles and the charged contaminants. This is one of the principal means for sorption of contaminants. The interactions between positively and negatively charged atoms and molecules result in development of interatomic bonds including ionic, covalent and coordinate covalent bonds. In ionic bonding, electron transfer between atoms results in an electrostatic attraction between the resulting oppositely charged ions. Covalent bonding involves more or less equal sharing of electrons between the partners. In coordinate covalent bonding, the shared electrons originate only from one partner. Other interactions such as ion-ion interactions between the soil solids and contaminants are Coulombic in nature. A signiﬁcant portion of the bonding of contaminants or even the bonding of different types of soil solids occurs as a result of interactions between nuclei and electrons. These are basically electrostatic in nature. For example, electrical bonding occurs (1) between the negative charges on clay mineral surfaces and positive charges on organic matter and (2) between negatively charged organic acids and positively charged clay mineral edges. The hydrogen bonds formed between soil-organic matter and clay particles are electrostatic or ionic bonds. Bonding between the oxygen from a water molecule and the oxygen on the clay particle surface is strong in comparison with other bonds between neutral molecules. This type of bonding is not only responsible for holding layers of clay minerals together, but also in bonding organic molecules to clay surfaces and holding water at the clay particles’ surfaces. The presence of polyvalent exchangeable cations permits adsorption of organic anions such as those in organic chemicals to clay mineral surfaces through the formation of polyvalent bridges. Interactions such as those between instantaneous dipoles and dipole-dipole interactions also produce forces of attractions that may be responsible for sorption of contaminants. For nonpolar molecules (e.g., organic chemicals), these types of interaction are the most common type of sorption bonding mechanism established between nonpolar molecules such as organic chemicals and soil solids. The interaction forces developed are categorized as van der Waals forces. These include (1) Keesom, forces developed as a result of dipole orientation; (2) Debye, forces developed owing to

80

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

induction; and (3) London dispersion forces. Another type of bond is the steric bond, which develops as a result of ion hydration at the surface of the soil solids. Chemical adsorption, or chemisorption, refers to high-afﬁnity, speciﬁc adsorption that occurs in the inner Helmholtz layer (see Figure 3.8) through covalent bonding. The ions penetrate the coordination shell of the structural atom and are bonded by covalent bonds via O and OH groups to the structural cations. The valence forces bind atoms to form chemical compounds of deﬁnite shapes and energies. Chemisorbed ions can inﬂuence the sign of y. They are called potential determining ions (Pdis) and are also sometimes referred to as high-afﬁnity speciﬁcally sorbed ions. 3.5.2

Cation Exchange

Ion exchange can occur between the ions in the diffuse ion layers and the surfaces of the reactive soil particles because of the charge imbalance. In a soil water system, cation exchange occurs when positively charged ions in the pore water are attracted to the surfaces of the clay fractions because of the need to satisfy electroneutrality. This process, which is stoichiometric and responds to the need for electroneutrality in the system, requires the replacing cations to satisfy the net negative charge imbalance of the charged reactive surfaces of the soil particles. Exchangeable cations are those cations associated with the charged sites on the surfaces of (primarily) clay particles. By and large, the number of charged sites that are considered as exchange sites are determined by isomorphous substitution in the layer-lattice structure of the clay minerals. The quantity of exchangeable cations held by the soil is called the cation-exchange capacity (CEC) of the soil and is generally equal to the amount of negative charges. It is usually expressed as milliequivalents per 100 g of soil (meq/100 g soil). Although exchangeable cations are usually associated with clay minerals, we must note that other soil fractions also contribute to the exchange capacity of a soil. Amorphous materials and natural soil organics also contain surface functional groups that contribute to cation exchange. Measured values for CEC can range from 15 to 24 meq/100 g soil for Fe-oxides, from 10 to 18 meq/100 g soil for Al-oxides and from 20 to 30 meq/100 g for allophanes. The nature and distribution of the oxygen-containing functional groups of soil organic matter (SOM) are also inﬂuential in cation exchange. In particular, the carboxyl and phenolic functional groups appear to be the major contributors to the CEC of these soils. Appelo and Postma (1993) have linked the CEC of a soil to clay content, clay minerals, soil organic matter and oxides/hydrous oxides. They provide CEC values of up to 100 meq/100 g soil for goethites and hematites and from 150 to 400 meq/100 g soil for organic matter at a pH of 8. The empirical relationship cited by Appelo and Postma (1993), which relates the CEC to the percentage of clay less than 2 mm and the organic carbon is given as follows: CEC (meq/100g soil) = 0.7 Clay% + 3.5 OC% where Clay% refers to the percentage of clay less than 2 mm and OC% refers to the percentage of organic carbon in the soil.

SOIL-WATER SYSTEMS AND INTERACTIONS

81

The predominant exchangeable cations in soils are calcium and magnesium, with potassium and sodium being found in smaller amounts. Exchangeable cations can be replaced by another of equal valence, or by two of half the valence of the original one. When the exchange cations possess the same positive charge and similar geometries as the replacing cations, the following relationship applies: Ms/Ns = Mo/No = 1, where M and N represent the cation species, and the subscripts s and o represent the surface and the bulk solution. This exchange mechanism plays a signiﬁcant role in the partitioning of heavy metals. We will see more of this when we discuss partitioning and attenuation of heavy metal contaminants. For soil fractions that have net surface charges dependent on pH, values of CEC will vary depending on the pH of the system. The soil fractions that are included in this list are kaolinites, natural soil organics and the various oxides or amorphous materials. In kaolinites, for example, the values of CEC can vary by a factor of 3 between the CEC at a pH of 4 (CEC = 2) and a pH of 9 (CEC = 6). Higher variations can be expected for oxides because the proportion of pH-dependent charges is much higher for the oxides than the proportion of pH-dependent edge surface charges to planar surface charges in kaolinites. The technique used for measurement of CEC in soils is such that operational differences will produce differing results, thus rendering the resultant CEC measurement as an operationally deﬁned quantity. The condition that cation sorption should occur on all available sites requires one to determine that all sorption sites are occupied. One should note that incomplete dispersion of soil particles and ﬂocs will produce the situation where sorption sites are rendered inaccessible. Operator technique and test conditions are important factors in the determination of CEC. Saturation and subsequent exchange in the interlayers of layer-lattice clay minerals is an issue that highlights the problem. We also need to be aware that the reactions between the saturating cation solution and soil fractions can produce results that would be erroneous. Using ammonium acetate (NH4OAc) as a saturation ﬂuid, for example, for soils with signiﬁcant amounts of carbonates can cause dissolution of CaCO3 and gypsum. This would result in extraction of excessive amounts of Ca2+ by NH4+ , thereby producing CEC measurements that would be unreasonably high.

3.6 CHEMICAL BUFFERING AND PARTITIONING The chemical buffering capacity of soils is one of the means utilized in natural attenuation of contaminants. The chemical buffer capacity of soil, which is the reciprocal of the slope of the titration curve of the soil, is deﬁned as the number of moles of H+ or OH– that must be added to raise or lower the pH of 1 kg of the soil by 1 pH unit. Figure 3.9 shows the titration curves for two clay soils: a montmorillonite and an illite. The soil buffer capacity shown in the diagram is determined from the negative inverse slope of the respective titration curve and plotted in relation to pH. Expressing the buffering capability b of a soil in terms of changes in the amount of hydroxyl ions (OH–) or hydrogen ions (H+) added to the system and in respect to the resulting pH changes, we obtain b = (dOH-/dpH) = (dH+/dpH).

82

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

100 9

80

Buffer capacity, cmol/kg soil 60 40

Buffer capacity vs. pH for montmorillonite

8

..

7 6

20

pH

.

. .

5 4

0

Buffer capacity vs. pH for illite

.

3 2 Titration curve for montmorillonite

1

Titration curve for illite

0 0

Figure 3.9

20

40

60 80 100 120 140 160 Amount of acid added, cmol/kg soil

180

200

Titration and buffer capacity curves for illite and montmorillonite soils. Note that the horizontal axis for buffer capacity curves begins with zero value on the righthand side, and the relationships are given with respect to the pH ordinate axis.

We can obtain a clearer picture of how well a soil can perform as a chemical buffer by viewing its buffering capacity curve. Figure 3.9 shows that when the pH of an illite soil-water system is greater than 4, its capacity for chemical buffering is higher than montmorillonite. This is very interesting and informative since the illite soil has a smaller CEC than the montmorillonite soil. What this tells us is that the high resistance in pH change in illite is due not only to adsorption of H+ onto the exchange sites on the clay particles, but also to the neutralization of H+ by the carbonates in the illite soil. 3.6.1

Partitioning, Adsorption Isotherms and Distribution Coefﬁcients

One method for determination of partitioning of contaminants in a leachate stream or in the pore water by sorption mechanisms is to conduct batch equilibrium adsorption isotherm tests. Because the procedure is conducted with soil solutions, the results obtained are more than likely indicative of maximum sorption partitioning since the soil particles in the soil solution are considered to be completely dispersed. All the soil particle surfaces in the soil solution are exposed and available for sorption of the contaminants in the aqueous phase. Figure 3.10 shows the typical shapes of more popular adsorption isotherms characterized as high-afﬁnity-type, constant-type (linear adsorption curve), Freundlich-type and Langmuir-type isotherms. The relationships between the concentration of solutes

High affinity type

83

k 3s s* = 1 + k 4s

Langmuir type

=k

2sm

Freundlich type

s*

Concentration of solutes sorbed, s*

SOIL-WATER SYSTEMS AND INTERACTIONS

= s*

s k1

kn and m are constants, n = 1 to 4 Constant adsorption

Equilibrium concentration of solutes in solution, s

Figure 3.10

Different types of adsorption isotherms obtained from batch equilibrium tests.

adsorbed, s*, and the equilibrium concentration of solutes in the aqueous solution, s, are given adjacent to the respective curves. The constants kn (n = 1 … 4) and m are generally obtained from experiments and data ﬁtting procedures. It is important to take note of the fact that both the constant-type and Freundlich-type isotherms predict limitless adsorption and must be used with deﬁned limits based on experimental information. The use of adsorption isotherms will be discussed in greater detail in the next few chapters dealing with attenuation of heavy metals. The distribution coefﬁcient kd is generally obtained as the slope of the adsorption isotherm. For constant-type adsorption isotherms, kd = k1 (Figure 3.10). For Freundlich-type and Langmuir-type adsorption isotherms, kd is generally taken as the initial tangent, i.e., initial slope deﬁned by the curves shown in the ﬁgure However, it is not uncommon to also choose other points on the curve to deﬁne the slope of the curve. Various kd values for different kinds of soils in relation to different pollutants and under various conditions should be obtained when predictions of transport and attenuation of contaminants need to be made. However, it must be noted that the use of batch equilibrium testing where soil solutions are used to determine the adsorption isotherms and kd do not in any way represent compact soils found in the ﬁeld. Soils in the ﬁeld are generally in compact form and do not present individual soil particles for interaction with contaminants in the pore water. Tests are needed that would expose the compact soils to contaminants in the pore water, i.e., leaching tests conducted on soil columns (discussed in detail in the next few chapters).

84

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

3.7 WATER MOVEMENT IN SOILS Up to this point, we have considered the situation where interactions between soil solids and pore water are evaluated under conditions that involve the static state of soil water; i.e., no pore water movement is involved. Pore water in soils is rarely in a static state. Additions and subtractions of pore water (or groundwater) arise owing to rainfall, snow melt, evaporation, transpiration, condensation, irrigation and drainage, to name a few types of events that contribute to or subtract from the water content in soils. Distribution and migration of the water accompanying any of the previously mentioned events occurs in response to the many ﬂuxes arising from the internal energy of the water itself and from internal and external mechanisms and driving forces owing to thermal, ionic, osmotic, gravitational, hydraulic and other thermodynamic gradients. These gradients, forces and ﬂuxes, together with the transmission properties of the soil, determine the rate and amount of water movement in the soil. Except for liquid pollutants, we consider the movement of all nonliquid pollutants in soils to be directly associated with the presence and movement of pore water in soils. We can easily separate the movement of water in three categories that are differentiated by the initial soil condition: (1) uptake of water by dry soil; (2) water movement in partly saturated soil, sometimes referred to as unsaturated moisture movement or unsaturated ﬂow; and (3) water movement in fully saturated soils. We have already considered the movement of water in fully saturated soils in Section 2.5.2 in Chapter 2 in our discussions of the transmission properties of soils. In this section, we will be concerned with water uptake into dry soils and water movement in partly saturated soils. We refer to the wetting of a dry soil as water uptake when it is clear that the water is being drawn into the soil by the internal forces of the soil. In sands, for example, these internal forces would clearly be associated with the capillary forces. In clay soils, however, internal forces are associated with both the matric and solute potentials ym and ys. 3.7.1

Water Uptake by Dry Clay Soil

We started to discuss the phenomenon of water uptake in Section 3.3.1 with respect to the role of the matric potential ym. The characteristics of water uptake (i.e., wetting) of a dry clay soil are different in nonswelling and swelling soils beyond the hydration layer. The primary mechanism involved in initial wetting of dry soil solids’ surfaces is the action arising from the adsorption energy of water. The thickness of the hydration water adsorbed onto the surfaces of the soil particles owing to this mechanism does not exceed 1 nm for nonswelling soils. This water is generally called crystalline water to reﬂect the fact that the structure of this water is unlike that of ordinary bulk water. For swelling soils such as the 2:1 dioctahedral series of alumino-silicate clays (e.g., montmorillonites and nontronites; see Figure 2.8 in Chapter 2), water uptake in response to the ym can seemingly exceed 1 nm. The further uptake of water (beyond the hydration layer) is not in response to the ym but is in fact due to the mechanisms represented in the DDL models. If we refer back to Figure 3.7, we can

SOIL-WATER SYSTEMS AND INTERACTIONS

85

see this illustrated in the form of the respective energy characteristics. The interaction of the overlapping diffuse ion layers and the associated energy characteristic deﬁnes the swelling phenomenon. When this occurs, the soil volume expansion that occurs is called swelling, thus giving the name of swelling soils to such soils. The interlayer or interlamellar expansion owing to crystalline water uptake (hydration) is a function of the layer charge, interlayer cations, properties of adsorbed liquid and particle size. The energy characteristic associated with this is shown in Figure 3.7. Water uptake beyond hydration that is due to double-layer forces results in increases in interlayer separation space in proportion to

1----s

, where s is the electrolyte concentration in the

liquid phase. Because of the popular use of bentonites for barrier and buffer-liner systems, it is important to pay attention to what is happening in the interlayer spaces since these interactions play a dominant role in the assimilative processes. Furthermore, the characteristics of water uptake and uptake of pore water containing contaminants differ somewhat from those of nonswelling soils. In particular, we need to be aware of the energy requirements associated with the movement of water in the interlayers and between particles. Predictions of water and solute movement that rely on speciﬁcation of the solute and matric potentials ys and ym must recognize how these measurements are made and their relevance in relation to their control in water uptake and movement. 3.7.2

Unsaturated Flow

Because the range of water content in partly saturated soils can vary from relatively dry to relatively wet, the mechanism for water transfer for either situation will be different. The term water transfer is used in preference to water movement because the presence of a vapor phase and its movement adds or subtracts water from any one location. Where high temperature gradients exist in soils, vapor transfer is greater than liquid transfer in the relatively dry soils. However, in relatively wet but partly saturated soils, liquid transfer outweighs vapor transfer. Vapor movement occurs by convective ﬂow of the air in the soil and/or by diffusion of water molecules in the direction of decreasing vapor pressure. Vapor-pressure gradients can develop not only because of temperature differences but also because of salt concentration differences and differential suction in the soil. Vapor transfer in partly saturated soils can also occur under isothermal conditions. In natural soils, and especially in natural clay soils, it is not unusual to ﬁnd from 2 to 10% air content in so-called saturated soils. This is because of the presence of entrapped air in the soil voids ﬁlled with water. In the absence of external pressure gradients, most of the water movement in the liquid phase is due to the gradients of the ym and to physico-chemical forces associated with the interaction between the soil particles and water. Differences in concentrations of solutes between two points create ﬂow. Figure 3.11 shows an osmometer-type device where the righthand cell contains pure water with access to (1) a moist clay soil containing a speciﬁed concentration of solutes in the left-hand cell (top device) and (2) a solution containing the same solutes as in the moist clay soil and at the same concentration in the left-hand cell (bottom) replacing the moist clay soil. The water in the right-

86

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Pure water Moist clay soil with solutes

Shaded length shows suction needed to prevent water from entering left-hand cell

Suction device

S = SS + SM SM

Solution with same solute concentration as in moist clay soil

SS

Selective membrane permitting transport only of water molecules through membrane

Figure 3.11

Osmometer-type cells showing development of suction required to counter ﬂow of water into the left-hand side chambers because of the total potential yT in the moist clay soil (top) and because of the solute potential ys in the cell with the solution containing the same concentration of solutes (bottom). S, SS and SM refer to the total suction, solute suction and matric suction, respectively.

hand cell is connected to a suction measuring device. For simplicity in visualization, the length of the shaded horizontal column represents the equivalent suction needed to prevent transport of the water in the right-hand cell to the left-hand cells in response to the potential gradients established by the moist soil or the solutes in the solution. The components of the total potential y described in Section 3.3 produce gradients that provoke ﬂow (liquid transfer). Because of the concentration of solutes in the solution in the left-hand cell (bottom), there is a tendency for the pure water in the right-hand cell to diffuse into the left-hand cell in response to the gradient set up by the solute potential ys in each of the cells. The suction required to prevent the diffusion of water is shown in the diagram as Ss. When the left-hand cell contains a moist clay soil with the same kind and concentration of solutes, the suction required to prevent diffusion of water into the moist clay cell (top device) is much higher owing to the addition of the ym in the moist clay soil. The schematic illustration shown in Figure 3.11 demonstrates that even in the absence of external forces or pressures, liquid water moves in soils in response to internal gradients developed as a result of the differences in potentials between adjacent points. Water movement in partly saturated soils occurs along ﬁlm boundaries in soil pore spaces that are not completely ﬁlled with water and as pore channel ﬂow for those pore spaces that are completely ﬁlled with water. Figure 3.12 shows a sequence schematic that depicts water uptake into a dry soil. The representative volume

SOIL-WATER SYSTEMS AND INTERACTIONS

87

Representative volume “Exploded” 4-particle arrangement in mu

Microstructural unit (mu) Film boundary

“Full” saturation from uptake and unsaturated flow

Further uptake of water is by film boundary transport

Idealized edge view showing water uptake beginning with hydration layers

Entrapped air bubble

Figure 3.12

Water uptake into a dry soil. The schematic drawing shows an exploded view of a four-particle arrangement in a microstructural unit taking on water through hydration processes and further transport into the soil from ﬁlm boundary transport. At apparent full saturation, it is possible to entrap air bubbles in the structural units because of the inability of the air bubbles to escape.

element shown in the top of the diagram depicts three microstructural units interacting with each other. The top right of the diagram shows an “exploded” view of a four-particle arrangement in a microstructural unit taking on water. Initially, water uptake occurs through hydration processes. Further water transport into the soil occurs from ﬁlm boundary transport. As more water is drawn in, the air within the soil must escape to the surface. At “full” saturation, if trapped air in some of the voids cannot escape, it remains in the microstructural unit. Figure 3.13 shows the activation energy requirements for movement of ions in a boundary layer associated with wetted particles. It is important to realize that the processes associated with assimilation of contaminants and/or pollutants, and with degree of bonding between soil particles and contaminants and pollutants, are to some degree affected by the properties of this layer. Pore channel ﬂow has been modeled as saturated ﬂow and will be discussed in the next section. Obviously, we cannot evaluate or analyze unsaturated ﬂow using two separate and different analytical models; i.e., it is not practical to perform ﬁlm boundary ﬂow analysis in conjunction with saturated ﬂow analysis since we have no means to determine the proportion of ﬁlm boundary or saturated ﬂow contributing to the total ﬂow. Instead, we have to rely on a deterministic analysis of the combined unsaturated ﬂow phenomenon. Figure 3.14 shows that the computed hydraulic con-

88

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

We need to be very aware of the influence of the hydration (adsorbed) layer and surface charge distribution of clay particles on pollutant interaction and assimilative processes Higher activation energies for movement of exchangeable ions in the hydration or boundary layer due to bond breakage with neighbors and charge sites, and also in “shoving” aside of water molecules. Activation energy z A2 i = n 1

Ei =

D

5

i=1

ri

E i = electrical energy of ion; r = distance between ion and charge site i; n = number of charge sites on particle; z = valence; D = dielectric constant; A = electronic charge

Boundary layer movement in wetting process responds to forces associated with reactive surfaces, and is more an IHP phenomenon – distinctly different from movement in the regions outside the OHP Soil structure, size and distribution of units in macrostructure become important players in control of transport processes – e.g., proportion of film boundary to “regular” transport, and abundance of “menisci barriers” in units.

Figure 3.13

Activation energy requirements for movement of ions in boundary layer associated with wetted particles. Assimilation of contaminants and pollutants and degree of bonding between soil particles and contaminants are all part of the equation for water uptake and boundary layer transport.

ductivity coefﬁcient, k, obtained in typical fully saturated ﬂow experiments decreases as the water content of the soil is decreased. For comparison, the corresponding water diffusion coefﬁcients, D, for the same soils are given in respect to the volumetric water contents, q. It must be noted here that the use of the water diffusion coefﬁcient D does not mean that the associated water ﬂow is diffusive in nature. The analytical treatment of ﬂow in partly saturated soils is facilitated by using diffusion-type models to analyze unsaturated transport of water even though the mechanism of water transport includes both ﬁlm boundary and saturated ﬂow. For a proper evaluation of the attenuation performance and capability of soils in actual ﬁeld conditions, it is important to recognize that initial “degree of saturation” state of the ﬁeld soil is a signiﬁcant parameter. For soils located in the upper horizon where wetting and drying occurs, the ﬁeld capacity of the soil plays a prominent role. It is at this point that most (if not all) of the water in the partly saturated soil is distributed in the soil as ﬁlm boundaries; i.e., no pore spaces are completely ﬁlled with water. The rate of water movement is extremely slow. Removal or movement of the hydration water layer associated with the soil particles requires energy input higher than those described in Equations 3.1 and 3.2. With respect to movement of contaminants that require water as the transport agent, diffusion of the contaminants as solutes occurs along connecting ﬁlm boundaries. So long as con-

89

D – silty sand 102

104 k – silty sand

1

103

10-2

102

D – clay soil

10-4

10-6

10

1 k – clay soil

10-8

Water diffusivity coefficient , D, cm2/day

Hydraulic conductivity coefficient, k, cm/day

SOIL-WATER SYSTEMS AND INTERACTIONS

0 0

0.1

0.2

0.3

0.4

0.5

Volumetric water content, G

Figure 3.14

Variation of hydraulic conductivity coefﬁcient k and water diffusivity coefﬁcient D with volumetric water content for a silty sand and a medium-type clay.

nected ﬁlm boundaries exist, diffusive transport of contaminants along the ﬁlm boundaries or boundary layers can occur. In making measurements and computations of ﬂow in partly saturated soils, it is often easier to measure differences in volumetric water content q between neighboring points instead of rate of ﬂow between the same two points. Using the macroscopic ﬂow velocity approach, and assuming a no volume change condition, the equation of continuity, which states that the ﬂow of water into or out of a unit volume of soil is equal to the rate of change of the volumetric water content, is given as (xn/xx) = -(xq/xx), where v = macroscopic velocity and x = spatial coordinate. Although not totally appropriate for unsaturated ﬂow at low water contents, as shown in Chapter 2, the Darcy relationship n = -k(q)(xY/xx) is often used in conjunction with the continuity condition, where y is the total soil-water potential referred to in Section 3.3. This permits one to obtain the unsaturated ﬂow relationship in terms of a changing q with distance as follows: xq x Ê xy ˆ = Á k (q ) ˜ xx xx Ë xx ¯ x Ê xy xq ˆ = Á k (q ) ˜ xx Ë xq xx ¯

(3.6)

90

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Assuming that for any one soil, y is a single-valued function of q, we can introduce the water diffusivity coefﬁcient (Figure 3.11) as D(q) = k(q)((xy)(xq)) and rewrite Equation 3.6 as xq x Ê xq ˆ = Á D(q) ˜ xt xx Ë xx ¯

(3.7)

For vertical ﬂow, using the coordinate z to designate the vertical spatial coordinate, we have xq x Ê xq ˆ xk (q) = Á D(q) ˜ + xt xz Ë xz ¯ xz

(3.8)

The wetting front proﬁle obtained from a horizontal inﬁltration experiment conducted on a soil sample shows a typical form as in Figure 3.15. Much can be learned from the shape of the wetting front. Using the solutions provided by Yong and Wong (1973), speciﬁc cases governing q and D can be studied, using Equation 3.7 to determine their inﬂuence on the shape of the wetting front proﬁle. Figure 3.16 shows the various proﬁles for the case where D = constant to cases where D varies as q varies. As indicated in the diagram, we can learn a lot from studying the shape of the wetting front, especially in regard to how D varies as the volumetric water content in the soil varies as it continues to take in water.

3.8 MOVEMENT OF SOLUTES Movement of pore water solutes (i.e., solutes in the pore water) in soils occurs in saturated and partly saturated soils. The obvious necessary condition for movement of solutes is the presence of water within the pore spaces either as ﬁlm boundaries or as water-ﬁlled pores. If continuity in ﬁlm boundaries is established through contact between adjoining particle ﬁlm boundaries, movement of the solutes will progress as diffusive movement. This is to say that in partly saturated soils, the predominant mechanism of transport or movement of solutes in the pore water occurs by diffusive means. Diffusive ﬂow of solutes also occurs in the water-ﬁlled micropores of the microstructural units previously shown in Chapter 2 (Section 2.5.1) because of the forces of interaction between adjacent particles. In saturated soils, there is both diffusive and advective ﬂow, depending on whether the external hydraulic gradients are sufﬁciently large; i.e., diffusive ﬂow of solutes occurs in the micropores, and advective ﬂow of solutes likely occurs in the macropores if the hydraulic gradients are sufﬁciently large (see next section). 3.8.1

Diffusion of Solutes and Diffusion Coefﬁcient

Brownian activity of the solutes in the pore water results in a net diffusive ﬂow of the solutes. When the pore water is in a more-or-less immobile state, e.g.,

SOIL-WATER SYSTEMS AND INTERACTIONS

91

Transmission zone Wetting zone

Volumetric water content G

Wetting front

Gsat

Wetting front profile

Gini Soil column 0

Figure 3.15

Distance from water source input Water source

Characteristics of a wetting front proﬁle. Shaded area represents the volumetric water content in the zone behind the wetting front. qini and qsat are initial and saturated volumetric water contents, respectively.

D increases faster than q increases; D = becq D increases linearly with q; D = a q

Volumetric water content q

D increases slower than q increases; D = b(1-ecq)

D = constant D decreases as q increases; D = be-cq

Distance from water source Figure 3.16

Analytically obtained wetting front proﬁles for various cases of D varying with q. (a, b and c are constants.)

92

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

hydration water layer or water in the diffuse double layer, diffusion is the more likely mechanism of transport of the solutes. From a theoretical point of view, we can consider the diffusion coefﬁcient of a target solute Ds to be equal to the effective molecular diffusion coefﬁcient. A useful reference point is the inﬁnite solution coefﬁcient Do, which refers to a single ionic species in dilute solutions. From the studies of molecular diffusion of both Nernst (1888) and Einstein (1905) involving the movement of suspended particles controlled by the osmotic forces in the solution, we cite the following: Nernst-Einstein

Do =

uRT = ukT N

Einstein-Stokes

Do =

RT T = 7.166 ¥ 10 -21 6 pNhr hr

Nernst

Do =

RTl Tl = 8.928 ¥ 10 -10 F2 z z

(3.9)

(3.10)

(3.11)

where u = absolute mobility of solute, R = universal gas constant, T = absolute temperature, N = Avogadro’s number, k = Boltzmann’s constant, l = conductivity of the target solute, r = radius of hydrated solute, h = absolute viscosity of the ﬂuid, z = valence of the ion, and F = Faraday’s constant. Inﬁnite solution diffusion Do models incorporate such factors as ionic radius, absolute mobility of the ion, temperature, viscosity of the ﬂuid medium, valence of the ion, equivalent limiting conductivity of the ion, etc. Compiled values for Do for various conditions can be found for example in Li and Gregory (1974) and Lerman (1979), and experimental values for l for many major ions at various temperatures can be found in Robinson and Stokes (1959). Discussions of the effects of varied contaminant solutes on diffusion coefﬁcients can be found in these same references. The Peclet number Pe can be used as a screening tool to be ensure use of the appropriate transport in evaluation of transport of solutes in the pore wate. The Peclet number is deﬁned as Pe = vL d/Do, where Do = the diffusion coefﬁcient in an inﬁnite solution, d = the average soil particle diameter, and vL = the longitudinal ﬂow velocity (advective ﬂow). Figure 3.17 shows the transport diagram using information reported by Perkins and Johnston (1963). This indicates that it is appropriate to consider transport of solutes as being diffusive when Pe < 1. In this range, diffusive transport of solutes in the pore water dominates any advective transport. Partitioning of solutes during diffusive ﬂow in the pore water can be considered in respect to (1) association with the volumetric water content and/or (2) the ﬂuxes associated with the respective thermodynamic gradients. In the ﬁrst approach, using the volumetric water content association, we can specify xy s ˆ xq s x Ê = Á k (q, s) ˜ xt xx Ë xx ¯

(3.12)

SOIL-WATER SYSTEMS AND INTERACTIONS

93

where s = concentration of solute under consideration and ys = y(q,s) = solute potential. Assuming that the Darcy permeability coefﬁcient k is relatively insensitive to the presence of solutes in the ﬂuid phase, i.e., k is a function only of q, and further assuming that ys is also only a function of q, we obtain from Equation 3.12 the following relationship: xq s x Ê xs ˆ x(r * s*) = Á D (q ) ˜ xt xt xx Ë s xx ¯

(3.13)

where r* and s* represent the bulk density of soil divided by the density of water and the concentration of solutes sorbed by the soil particles. The ﬂuxes associated with the respective thermodynamic forces are as follows: Jq = Lqq

xy q xy s + Lq s xx xx

xy q xy s + Lss J s = Lsq xx xx

(3.14)

where Jq and Js = ﬂuid and solute ﬂuxes, respectively, and Lqq, Lqs, Lsq and Lss are the phenomenological coefﬁcients. The coupled relationships can be expressed in conjunction with the various diffusivity coefﬁcients as follows: xq x = xt xx

xq xs ˘ È ÍÎ Dqq xx + Dq s xx ˙˚

xs x È xq xs ˘ r xs * D = + Dss ˙ xt xx ÍÎ sq xx xx ˚ r w q xt

(3.15)

where Dqq = Lqq(xyq/xx), Dsq = Lsq(xyq/xq), Dss = Lss(RT/s) and Dqs = Lqs(RT/s) = moisture, solute-moisture, solute and moisture-solute diffusivity coefﬁcients, respectively (Elzahabi and Yong, 1997). The choice of the functionals for the phenomenological coefﬁcients are based on experimental information on the distribution of solutes along columns of test samples. Yong and Xu (1988) provide a useful identiﬁcation technique for evaluation of these phenomenological coefﬁcients. 3.8.2

Solute Movement in Saturated Soils

For Peclet numbers greater than 10–2, i.e., Pe > 10–2 (Figure 3.17), it is necessary to consider the effects of advective velocities on the transport of solutes in the pore water. At the range of Pe > 10, advection plays a dominant role. When advection needs to be considered in the transport of solutes, the diffusion coefﬁcient DL is used. This coefﬁcient is identiﬁed as the longitudinal diffusion-dispersion coefﬁcient and is meant to reﬂect the advective velocity modiﬁcation of the diffusive ﬂow of

94

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

102 D L = longitudinal diffusion coefficient d = average soil particle diameter vL =

101

longitudinal velocity

DL Do Advection dominant

100 Diffusion dominant

Transition zone

10-1 10 -3

10

-2

10 -1

100

101

102

Pe = v Ld D o Figure 3.17

Diffusion and advection dominant ﬂow regions for solutes in relation to Peclet number. (Adapted from Perkins, T.K., and Johnston, O.C., J. Soc. Pet. Eng., 17, 70–84, 1963.)

solutes. The term longitudinal is used in conjunction with this coefﬁcient to signify ﬂow in the direction of the advective velocity. For convenience in terminology, DL is often referred to as the longitudinal diffusion coefﬁcient, even though advection is known to be involved in the movement of the solutes under consideration. The common expression used for DL is: DL = Dm + avL. Dm represents molecular diffusion and is equal to Dot, where t = tortuosity factor, a = dispersivity parameter, and vL = advective velocity. A general one-dimensional transport relationship based on Fickian diffusion of solutes taking into account the effects of advective velocities can be obtained as xs x2s xs r * xs * = DL 2 - v xt xx xx n xt

(3.16)

If we assume a linear relationship between the concentration of solutes sorbed by the soil particles and the concentration of solutes remaining in equilibrium in the aqueous phase (i.e., constant adsorption isotherm as given in Figure 3.9), we can introduce the distribution coefﬁcient kd discussed in Section 3.6.1 into Equation 3.17 to obtain R

xs x2 s xs = DL 2 - v xt xx xx

(3.17)

SOIL-WATER SYSTEMS AND INTERACTIONS

95

r* ˘ where s* = kds and R = ÈÍ1 + k [1 + (r*/n)kd] = retardation factor. When a n d ˙˚ Î nonlinear adsorption isotherm is used, Equation 3.16 should be used in conjunction with the full functional form for s*. Transport processes become somewhat more complicated when the soil is not a uniform homogeneous soil, as is the case for natural soils. We discussed the structure of soil in Section 2.5 in relation to some physical characteristics and properties of soils. Most, if not all natural soils possess microstructural units (mu’s) that are the building blocks for the macrostructures that characterize such soils. The pore spaces in a natural soil are not uniform, not only because of the irregular shapes and sizes of the soil particles, but also because of the presence and distribution of the microstructural units. The microstructural units vary in size. The pore spaces in these units, which are classiﬁed as micropores, have different ﬂuid and solute conducting characteristics. Many researchers suggest strongly that we should pay more attention to the wide range of solute velocities within and between soil pores (Philip, 1968; Skopp and Warrick, 1974; Rao et al., 1980) and that we should recognize that some pore spaces could be nonconducting (stagnant). This is because continuity between the pore spaces may not exist and because many micropores are too small to permit easy transmission of solutes because of the prohibitive energy requirements. However, because of the disparity in sizes between the macro and micropores in the microstructural units, we can consider the microstructural units as sources/sinks for solutes (Figure 3.18). To account for this phenomenon, the D coefﬁcient could be deﬁned as follows (Paissioura, 1971; Rao et al., 1980; Wagenet, 1983): D = (p + Dh + Ds) where

(3.18)

Dp = Dot = effective molecular diffusion coefﬁcient, Do = inﬁnite solution diffusion coefﬁcient, Dh = avL, Ds = dispersion coefﬁcient accounting for dispersion effects caused by diffusion of solutes from stagnant to mobile regions and is given as: È l2 L ˘ Ds = f Íj, , ,....˙ Î Des v ˚

t = physico-chemical tortuosity factor = 2

Ê Lˆ qÁ ˜ wc Ë Le ¯ w = coefﬁcient relating to effect of charged soil particles on water viscosity, c = coefﬁcient accounting for effects of anion exclusion,

96

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Soil organic matter Hydroxyl

Carboxyl Amine

COOH

-

OH

-

+

NH x

Carbonyl CO

+

O Methoxyl O

OH

+

CH 3

-

Phenolic

O

+

Quinone

Figure 3.18

Schematic diagram showing sink-source phenomenon created by presence of microstructural units in diffusive ﬂow of water through the soil-water system.

Des = effective diffusion coefﬁcient in the stagnant region, and j = pore water fraction in the conducting region. Using the expression for Ds given by Paissioura (1971) and Rao et al. (1980), the ﬁnal relationship for the D coefﬁcient given as Equation 3.18 was obtained by Yong et al. (1992) as Ê v 2 r 2 (1 - j) ˆ D = Á Do t + av + 15 Des ˜¯ Ë

(3.19)

where r is the average equivalent diameter of the soil particles. The signiﬁcance of the expression given in Equation 3.19 lies in the recognition of the inﬂuence of the wide differences in pore sizes in a natural soil. Leaching column tests and diffusion cell tests with natural samples provide information that at best can be considered as representative of that particular sample in the test column or cell. Extrapolation or direct use of laboratory values for ﬁeld predictions of transport of solutes will necessarily be dependent on the degree to which the laboratory samples have been able to replicate the structure of the ﬁeld soils. 3.9 CONCLUDING REMARKS The primary focus of this chapter has been to develop a better appreciation of soil as a material composed of various soil fractions or soil solids in interaction with an aqueous phase. We have learned that:

SOIL-WATER SYSTEMS AND INTERACTIONS

97

• The basic structure of soil fractions and their surface functional groups form complexes between the functional groups and the contaminants or solutes in the pore water. • The functional groups at the planar and edge surfaces of inorganic soil fractions, together with isomorphous substitutions in the lattices of the layer-lattice clay minerals, result in the development of negative and positive charges distributed on the surfaces and edges of the soil particles. The nature and extent of surface complexation depend on the reactive properties of the soil particles and the contaminants themselves. Evaluation of the complexes formed can be performed using a variety of surface complexation models such as the single-layer, doublelayer and triple-layer models. • Charge reversal occurs when the net charges on a particle surface (i.e., charge density) change in sign from positive to negative or vice versa when the system pH progresses from below the pzc to above the pzc. • Interactions between soil solids, water and the various dissolved solutes in the pore water can be characterized in terms of energy relationships known as soilwater characteristics. These provide information about the water-holding capacity of a soil. • The pH of the soil-water system and the various other dissolved solutes in the pore water contribute to interaction mechanisms such as acid-base reactions, speciation, complexation, precipitation and ﬁxation. • Bonds are developed as a result of interactions between charged particles of the various soil fractions and the charged contaminants. These bonds, which constitute one of the principal means for sorption of contaminants, include interatomic bonds such as ionic, covalent and coordinate covalent bonds. • Determination of partitioning of contaminants or solutes is generally obtained through batch equilibrium adsorption isotherm tests. The procedure is conducted with soil solutions, and the results obtained are generally indicative of maximum sorption partitioning. • Movement of all nonliquid pollutants in soils can be analyzed in terms of their association with the volumetric water content of the soil. When the pore water is in a more or less immobile state (e.g., hydration water layer or water in the diffuse double layer) diffusion is the more likely mechanism of transport of the pollutant solutes. • Water movement in partly saturated soils occurs along ﬁlm boundaries in soil pore spaces that are not completely ﬁlled with water. Pore channel ﬂow occurs for those pore spaces that are completely ﬁlled with water. • For Peclet numbers Pe >> 10-2, the effects of advective velocities on the transport of solutes in the pore water cannot be ignored.

A knowledge of the basic interactions between soil solids and contaminants or dissolved solutes in the pore water is necessary if we are going to develop a better appreciation of the factors that contribute directly to the assimilative capacity of soils. The nature of the reactive surfaces in the soil-water system and how these surfaces are obtained will give us insight into how the soil conditions in the ﬁeld will impact directly on the transport and fate of the contaminants under consideration.

98

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

REFERENCES Allison, J.D., Brown, D.S., and Novo-gradac, K.J., MINTEQA2/PRODEFA2: A geochemical assessment model for environmental systems, EPA/600/3-91/021, U.S. Environmental Protection Agency, Athens, GA. 1991. Appelo, C.A.J. and Postma, D., Geochemistry, Groundwater and Pollution, Balkema, Rotterdam, The Netherlands, 1993, 536 pp. Bockris, J.O’M. and Reddy, A.K.N., Modern Electrochemistry, Vols. 1 and 2, Plenum Press, New York, 1970. Bolt, G.H. Adsorption of anions by soil, in Soil Chemistry; A: Basic Elements, Bolt, G.D. and Bruggenwert, M.G.M., Eds., Elsevier Scientiﬁc, New York. 1978, Chap. 5. Buckingham, E., Studies on the movement of soil moisture, U.S. Dep. Agric., Bur. Soils, Bull. 38, 61p. 1907. Einstein, A. Uber die von der Molekularkinetischen theorie der Warme Geforderte Bewegung von in Ruhenden Flussigkeiten Suspendierten Teilchen, Ann. Phys., 4, 549–660, 1905. Elzahabi, M. and Yong, R.N., Vadose zone transport of heavy metals, in Geoenvironmental Engineering: Contaminated Ground: Fate of Pollutants and Remediation, Yong, R.N. and Thomas, H.R., Eds., Thomas Telford, London, 1997, pp. 173–180. Grahame, D.C., The electrical double layer and the theory of electocapillarity, Chem. Rev., 41, 441–501, 1947. Greenland, D.J. and Mott, C.J.B., Surfaces of soil particles, in Greenland, D.J. and Hayes, M.H.B., Eds., The Chemistry of Soil Constituents, John Wiley & Sons, New York, 1985, pp. 321–354. Hayes, M.H.B. and Swift, R.S., The chemistry of soil organic colloids, in Greenland. D.J. and Hayes, M.H.B., Eds., The Chemistry of Soil Constituents, John Wiley & Sons,New York, 1985, pp. 179–320. Huang, P.M. and Jackson, M.L., Communication in Nature, 211, 779, 1966. Kruyt, H.R., Colloid Science, Vol. I, Elsevier, Amsterdam, 1952, 389 pp. Lerman, A., Geochemical Processes: Water and Sediment Environments, John Wiley & Sons, New York, 1979, 481 pp. Li, Y.H. and Gregory, S., Diffusion of ions in sea water and in deep-sea sediments, Geochem. Cosmochim. Acta, 38, 603–714, 1974. Nernst, W., Zur knetik der in losung beﬁnlichen korper, Z. Phys. Chem., 2, 613–637, 1888. Paissioura, J.B., Hydrodynamic dispersion in aggregated media: I. Theory, Soil Sci., 111, 339–344, 1971. Parkhurst, D.L., Thorstenson, D.C., and Plummer, L.N., PHREEQE: A Computer Program for Geochemical Calculations, U.S. Geol. Surv., Water Resources Investigation Report, pp. 80–96, 1980. Parks, G.A., The isoelectric points of solid oxides, solid hydroxides, and aqueous hydroxy complex systems, Chem. Rev., 65, 177–198, 1965. Perkins, T.K. and Johnston, O.C., A review of diffusion and dispersion in porous media, J. Soc. Petrol. Eng., 17, 70–84, 1963. Philip, J.R., Diffusion dead-end pores, and linearized adsorption in aggregated media, Aust. J. Soil Res., 6, 31–39, 1968. Rao, P.S., Rolston, D.E., Jessup, R.E., and Davidson, J.M., Solute transport in aggregated porous media: Theoretical and experimental evaluation, Soil Sci. Soc. Am. J., 44, 1139–1146, 1980. Ritchie, G.S.P. and Sposito, G., Speciation in soils, in Ure, A.M. and Davidson, C.M., Eeds., Chemical Speciation in the Environment, Blackwell Scientiﬁc, Oxford, UK, 2002.

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Robinson, R.A. and Stokes, R.H., Electrolyte Solutions, 2nd ed., Butterworths, London, 1959. Singh, U. and Uehara, G., Electrochemistry of the double layer principles and applications to soils, Soil Physical Chemistry, Sparks, D.L., Ed., CRC Press, Boca Raton, FL, 1986, Chap. 1, pp.1–38. Skopp, J. and Warrick, A.W., A two-phase model for the miscible displacement of reactive solutes through soils, Soil Sci. Soc. Am. Proc., 38, 545–550, 1974. Sposito, G., The Surface Chemistry of Soils, Oxford University Press, New York, 1984, 234 pp. Sposito, G. and Coves, J., SOILCHEM: A Computer Program for the Calculation of Chemical Speciation in Soils, Kearny Foundation of Soil Science, University of California, Riverside, 1988. Sposito, G. and Mattigod, S.V., GEOCHEM: A Computer Program for Calculating Chemical Equilibria in Soil Solutions and Other Natural Water Systems, University of California, Riverside. 1979,110 pp. Stern, O. Zur Theorie der elektrolytischen Doppelschicht, Z. Electrochem. 30, 508–516. 1924. van Olphen, H. An Introduction to Clay Colloid Chemistry, 2nd. ed., John Wiley & Sons, New York, 1977, 318 pp. Wagenet, R.J., Principles of Salt Movement in Soils, Chemical Mobility and Reactivity in Soil Systems, SSSA Spec. Publ. 11, 123–140, 1983. Wolery, T.J., EQ3NR: A Computer Program for Geochemical Aqueous Speciation-Solubility Calculations, University of California Lawrence Livermore Lab. 1983. Yong, R.N., Soil suction and soil-water potentials in swelling clays in engineered clay barriers, Eng. Geol., 84, 3–14, 1999. Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate, and Mitigation, CRC Press, Boca Raton, FL, 2000, 307 pp. Yong, R.N., Mohamed, A.M.O., and Warkentin, B.P., Principles of Contaminant Transport in Soils, Elsevier, Amsterdam, 1992, 327 pp. Yong, R.N. and Wong, H.Y., Water Movement in Unsaturated Swelling Soils, Soil Mechanics Laboratory, McGill University, Montreal, Research Report, 1973, 15 pp. Yong, R.N. and Xu, D.M., An identiﬁcation technique for evaluation of phenomenological coefﬁcients in unsaturated ﬂow in soils, Int. J. Num. Anal. Methods Geomech., 12, 283–299, 1988.

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Batch

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Soil leaching columns

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Analysis of soil solids in sections for exchangeable and extractable ions

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Pb (PEA3)

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14

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80

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60

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2

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6

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Pb concentration, ppm

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Abundance of 2,6 DMP dimers (mass 242) (Million)

6 Al-montmorillonite clay, Fe-montmorillonite clay, and Al-sand Mass of isomer 2,6-dimethylphenol = 122

5

0

0

0 0

4

0

Al-clay 0

0

0 0

3

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Fe-clay

0

0 0

0 0

2

0 0

0 0

1

0 0

0

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3

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5

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CHAPTER 7 Field Performance and Assessments 7.1 INTRODUCTION Before the 1990s, the only options available for remediation of contaminated sites were through engineered bioremediation. The more common ones involved addition of oxygen and nutrients or other practices such as pump and treat or excavation for treatment offsite. Prior to the 1970s, the subsurface was actually considered to be devoid of biological activity. We now know better. There is microbial activity throughout almost the entire subsurface. Since the 1990s, it has been determined that there are numerous subsurface processes that lead to the natural attenuation of contaminants. The processes (described in the previous chapters) include biodegradation, dilution, assimilation processes and partitioning, volatilization, and biological or chemical reactions. It is important at this stage to distinguish between the use of the term natural attenuation as a process and the term natural attenuation capacity as a property of a soil. When we refer to the use of natural attenuation as a process for remediation of contaminated soils, the abbreviation NA is often used to denote this process. When we refer to the property of natural attenuation, this term will be italicized. Contaminants can be sorbed to the soil, in the pore spaces in the vapor phase or dissolved in the groundwater. This is where most of the processes for NA of contaminants and pollutants occurs. In the U.S., remediation by NA increased at EPA Superfund sites from 6% in 1990 to 23% in 1997 (Renner, 2000). It is used extensively at sites where there are leaking underground storage tanks (U.S. EPA, 1996; ASTM, 1995) and in 1997 was used at 15,000 such sites (Renner, 2000). As there are, however, 358,000 releases or leaking underground storage tanks, there is the potential for extensive use of NA at these sites — attributable to the extensive cost of engineered remediation. According to NATO/CCMS (2003), NA is viewed in the U.S. as an alternative to pump and treat or for use after application of various other in situ technologies. What this suggests is that natural attenuation as a process could be usefully applied as a technique for remediation in support of a treatment train–type of application. Countries such as the United Kingdom have various

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initiatives with NA as the focus area. These include the Bioremediation LINK Programme and the Faraday Centre for Remediation of the Polluted Environment (NATO/CCMS, 2003). However, as the use of natural attenuation processes for application to remediation of contaminated sites has increased, we need to ask whether the protocols, procedures and capabilities for scientific evaluation of the effectiveness of these processes have been appropriate and satisfactory. This concern is being driven by awareness that the use of monitored natural attenuation (MNA) as a remediation tool is a knowledge-based technique. In this chapter, we will be looking at reported performances of MNA at various sites, with a view to determining whether the stakeholders consider the technique to be successful and effective. Assessment procedures, monitoring protocols, modeling capabilities used for the reported field application cases and field data in support of effectiveness (or otherwise) of natural attenuation as a process will be examined. The various protocols and requirements for MNA application are discussed in the next chapter.

7.2 ASSESSMENT OF POTENTIAL FOR NATURAL ATTENUATION The capacity for natural attenuation of contaminants should be determined to evaluate the rate of natural attenuation and to determine the level of contaminant removal to be achieved and the potential risks to receptors. Figure 7.1 shows some of the principal mechanisms involved in attenuating contaminants associated with the pollutant plume originally shown in Figure 1.2. Reliance on the many assimilative and transformation processes associated with the natural attenuation of contaminants to protect the biotic receptors from the threats posed by the contaminants requires confidence that the soil-water system will function as planned. Chapter 8 provides the protocols and requirements for determining whether natural attenuation processes can be effectively used to remediate a contaminated site. The lines of evidence and evidence of success protocols will be discussed in detail in that chapter. In this chapter, we will turn our attention to lessons we can learn from indirect and direct field applications of NA. It is well accepted that for successful application of natural attenuation as a tool for remediation of contaminated sites, the geological, biogeochemistry and hydrological conditions for each site must be evaluated on a case-by-case basis. A detailed and comprehensive understanding of the site-specific conditions, including availability of substrates and nutrients, will lead to higher success rates. Fate and transport models are used to provide predictions of distribution and concentrations of contaminants at various times and over specified distances from the source. These predictions provide guidance for initial evaluation and throughout the natural attenuation process. They are particularly useful in instances where biodegradation is a major mechanism contributing to the total attenuation process. That being said, decay rates that need to be determined for input into the models are problematic. These have been reviewed by Suarez and Rifai (1999) for petroleum and chlorinated hydrocarbons. Methods for improving our understanding of the processes involved and for assessing the potential for application of natural attenuation processes in remediation of contaminated sites are still under development.

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Figure 7.1

7.2.1

225

Schematic showing contaminated ground with a pollutant plume as the potential source of health and environmental threats to environmental and human receptors. Also shown are the processes associated with the transport of the pollutant plume.

Processes and Mechanisms Involved in NA

Future land use will also be a major consideration regarding the potential for application of NA as a remediation technology. If contaminants are continually released at an existing facility, they will have the potential to affect human health and the environment. Thus, application of NA as a remediation strategy could be a viable option if the pollutant source can be controlled. Various factors at a site need to be evaluated. For natural attention of metals and radionuclides to be realized, the primary attenuating mechanisms of sorption and dilution must be available. In the case of radioactive materials, daughter nuclides will be produced, and it must be determined whether they are more hazardous than the parent. The primary mechanisms contributing directly to the natural attenuation of contaminants for the three types of contaminants are shown in Figure 7.2. Various factors and issues need to be considered in detail to determine the potential for NA. Some of these include the following considerations: • At high pH values, net negative charges in the clay minerals are expected to be high, and the potential for increased cation sorption will be correspondingly increased. The likelihood of precipitation of metals is also expected to be high. • Anion sorption increases as the pH decreases as shown in Chapter 4. Low pH conditions also affect the solubility of iron hydroxides and carbonates. The results of these, in turn, affect sorption. • Calcium is known to provide an indication of calcium carbonate stability.

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Figure 7.2

Primary mechanisms and processes contributing directly to natural attenuation of contaminants and pollutants.

• Low redox values dissolve iron hydroxides by lead to sulfide formation. • The presence of chelating agents such as ethylenediaminetetraacetic acid (EDTA), nitrilotriacetic acid (NTA), diethylenetriamine pentaacetic acid (DTPA), citrate and oxalic acids can significantly influence the rate of transport of metals. • The biodegradation rates of the chelating agents mentioned above affect partitioning of the metals during transport through soils. Waters et al. (1998) have shown that the chelating agents degrade in the following order: citrate ~ oxalate >> NTA > EDTA > DTPA. These agents influence sorption by decreasing sorption or increasing complexes with the contaminant and the soil surface. • Excessive carbon dioxide produced can also increase porosity from calcite and dolomite dissolution under acidic conditions (Bennett et al., 2000). Changes in porosity can occur as a result of dissolution processes. • Other reactions under anoxic conditions such as carbonate and bicarbonate saturation with calcite can plug pore spaces and decrease permeability.

Since the pollutant source responsible for contaminating the soil is usually removed from the contaminated site, the sorbed metals that are left need to be evaluated for potential desorption or detachment since the desorbed metals will be mobile and will pose a threat to the well-being of the immediate environment. Performance assessment calculations usually assume reversible desorption, but although these rates can be very slow, they can be the determining factor in the mobility of these types of contaminants. 7.2.2

Bioattenuation and Bioavailability

Determination of the capacity for bioattenuation has not received a great deal of attention in assessment of the natural attenuation of organic chemical pollutants.

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Substrates can become less bioavailable via interaction with negatively charged clay particles and organic material (Alexander, 1994). Sorption and sequestration can be influenced by pH, organic matter content, temperature and pollutant characteristics. The biodegradation of polycyclic aromatic hydrocarbons (PAHs) is particularly affected by sorption. Bioavailability of electron acceptors can influence microbial activity. Solid Fe(III) must be available and must be in direct contact with the microorganisms or obtained from humic acids that can chelate iron. Their presence can significantly increase iron bioavailability (Lovley, 2001). Nutrients such as nitrogen and phosphorus can also be limiting. Microorganisms attached to the soil minerals are capable of degrading soils that release phosphorus. There have been good indications that the rate of degradation or weathering can correlate well with microbial activity (Bennett et al., 2000). This leads to the suggestion that determination of sediment or soil contents could be interpreted in terms of biodegradation potential. While this thesis is tenable, much work needs to be done to support it — given that many other chemical and physico-chemical factors contribute significantly to the weathering of rocks and soils. 7.2.3

Other Factors

The presence of other predatory organisms such as protozoa can also be detrimental. As the contaminant levels increase, so do the levels of microorganisms. This also increases the likelihood of grazing by protozoa. The rate of predation is influenced by the soil and pollutant characteristics, groundwater velocity and redox conditions (Kota et al., 1999). According to the U.S. EPA Office of Solid Waste and Emergency Response (OSWER) directive (U.S. EPA, 1999), the contaminated site needs to be characterized for the nature and concentrations of contaminants and potential impact to receptors (Figure 7.2). Contributions of sorption, dilution and dispersion of the contaminants should be evaluated in the groundwater regime in addition to the hydraulic regime. This includes recharge, discharge areas and volumes. For biodegradation, the presence of nutrients, electron donors and acceptors, metabolites and by-products, and presence of available microbial populations should be evaluated. These pieces of information need to be incorporated into a site-specific fate and transport model. Technological limitations for application of NA may include the presence of solution channels, fractures, joints or foliations in the rock aquifers. Materials with anisotropic properties can make it difficult to provide accurate predictions of groundwater flow and direction. The difficulties and errors can be minimized with recognition and proper accounting of the anisotropic properties. Evaluation of the potential and capability of NA at a site will be enhanced as information of bacterial species and their activities is cataloged and evaluated. It is necessary to bear in mind that the tools presently available for assessment of the rates of potential biodegradation are still not reliable. Determination of biodegradation rates in the laboratory can be one to two orders greater than those in the field (Kota, 1999). Surface heterogeneity is one of the complicating factors. No accepted methods exist for incorporating rate data into mathematical models. In the Kao and

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Wang (2001) mass flux method for assessment of natural attenuation of BTEX (benzene, toluene, ethylbenzene and xylene), corrections were made for the biodegradation rate using dispersion, dilution, volatilization and sorption. The basis for the corrections was obtained by using the recalcitrant tracer, 1,2,4-trimethylbenzene (TMB). Biodegradation was estimated using a first-order decay rate. Evidence for BTEX removal included the following: • • • • • •

Exhaustion of dissolved oxygen concentration Production of Fe(II), carbon dioxide and methane Increase in pH Decreases in BTEX carbon to total organic carbon (TOC) ratio as it is transported Decreases in BTEX mass in comparison to the tracer Faster removal of some components in the BTEX

Toluene degraded the fastest, followed by o-xylene, then m- and p-xylene, then xylene and finally ethylbenzene. The technique of using mass flux and an in situ tracer seems to be a more direct way for determination of the natural attenuation of the BTEX instead of relying on microcosm tests and fate and transport models. Approximately 87% of BTEX was removed by biodegradation. 7.3 ASSESSMENT OF SUSTAINABILITY One way to evaluate NA is to determine its sustainability (NRC, 2000). The mechanisms for natural attenuation of contaminants need to be sustainable, particularly if the rate of contaminant release is high. This can occur in the case of a pool of nonaqueous phase liquids (NAPLs) or a tailings pond. The various contaminant attenuating processes and mechanisms need to be evaluated in respect to long-term capabilities since groundwater and soil qualities must be assessed in view of the fate and transport of these contaminants. The availability of electron acceptors from the groundwater or soil as well as electron donors should be evaluated on a longterm basis. In the case of soil mineralogy, sites for sorption, electron acceptors and alkalinity are not in infinite supply. Renewable and nonrenewable mechanisms and processes need to be determined as shown in Figure 7.3. Mass budgeting can be used as a tool for evaluating sustainability. Mass budgeting determines the contaminant destruction or sequestration and compares it to production of components. This compares the importance of natural attenuation processes against each other. The advantage of determination of sustainability of natural attenuation processes on a long-term basis can be seen in the example of a nonsustainable NA application at a former copper mining site where acid drainage (pH 2 to 3) was released from an unlined pond. High concentrations of sulfate, calcium, iron, manganese, copper, aluminum and zinc were present in the groundwater plume 12 km from the source. Initial studies (Stollenwerk, 1994) showed that the soil carbonate raised the pH from 5 to 6 and led to precipitation or sorption of iron, copper, zinc and other contaminants. As the carbonate became depleted, the pH dropped and the metals were remobilized. This tells us that we need to evaluate the sensitivity of natural attenuation to changing environmental conditions. In addition, we need to develop methods to assess the

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Figure 7.3

229

Schematic showing transport of pollutant plume through a soil-water system. The concerns for long-term performance of the mechanisms and processes contributing to the natural attenuation of the pollutants are central to the problem at hand. Also shown are the issues regarding long-term supply and availability of electron acceptors and donors, sorption sites and long-term status of pH and Eh.

effects of changing environments and their effects on the long-term sustainability of NA and its long-term effectiveness. Changes in the immediate environment could be the result of interactions in the soil-water system with other contaminants, formation of toxic by-products, climate change and adverse effects from other remediation technologies. 7.4 PROCEDURES FOR MONITORING If NA is effective, the mass and toxicity of the contaminants will be reduced. Monitoring of the quality of both the groundwater (pore water) and soil material permits verification of the effectiveness of NA and deduction of the processes responsible for the attenuated results. As will be seen in Chapter 8, there are strict protocols and criteria that have to be met. If we are to be comfortable with application of NA as a treatment process, we need to be assured that the process will continue to operate in the long term. A thorough knowledge of the subsurface geology, hydrology, microbial populations and degradation and conversion processes is required (Chapter 8). 7.4.1

Importance of Monitoring

The U.S. EPA directive (U.S. EPA, 1999) says that when NA is used for remediation of soil and groundwater, it cannot be a no action approach. The application

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of NA is dependent on extensive monitoring. The designation of MNA is used to reflect this. Lines of evidence have been established and include: • Historical groundwater and soil chemistry data showing decreases in contaminant mass or concentration of time at monitoring or sampling points. • Indirect evidence from hydrologeological and geochemical data indicating that natural attenuation is occurring by sorption, dilution, volatilization or biodegradation and the required levels will be achieved. • Use of field or microcosm studies for direct demonstration of natural attenuation or degradation of the contaminant. The studies are conducted in situ or with actual contaminated site media.

If the first line of evidence is not conclusive, then the second will be required. If the second cannot be proven, then the third will be necessary. Site-specific factors such as size and nature of contaminants, availability of and risk to receptors, and environmental factors such as climatic conditions, hydrogeology, hydrology and subsurface geology can influence the amount of information required. The EPA states that MNA is most appropriate if a reasonable time frame for remediation can be demonstrated. To accomplish this, MNA can be combined with source control or groundwater extraction or other remediation processes. Therefore, the plume should be no longer increasing or should be decreasing in size. Monitoring is thus a substantial part of MNA. According to the U.S. EPA (1999), the main objectives include: • Showing that natural attenuation is performing as expected • Detecting hydrologic, geochemical, microbiological or other changes that could affect the natural attenuation process • Identifying by-products that are toxic and/or mobile • Determining if the plume is expanding in a downgradient, lateral or vertical direction • Ensuring that there is no impact on receptors and that controls are sufficient to protect them • Determining if new releases of contaminants are occurring that could affect the natural attenuation process • Verifying that the remediation objectives are achieved

Figure 7.4 shows the basic elements involved in a monitoring scheme. Note that the number of monitoring and sampling wells shown in the diagram is not representative of what will generally be required in the field. Experience from field studies has shown that there can be well over 100 wells and sampling points, depending on the size of the management zone as shown in Figure 7.4. Monitoring and sampling must continue until one is assured that there are no longer any threats to the health of humans or to the environment. This is typically from 1 to 3 years according to the U.S. EPA (1999), but this period can be substantially longer. As with the lines of evidence, monitoring will be dealt with in greater detail in Chapter 8. 7.4.2

Components of Monitoring

The term monitoring can have different meanings depending on the context in which it is used. In the context of NA, monitoring involves determination of the

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Figure 7.4

231

Monitoring network for performance, detection and ambient monitoring as required in application of natural attenuation as a remediation tool.

status of the pollutant plume and the effectiveness of the NA treatment. In that sense, there are several components that comprise the monitoring scheme and requirement. Monitoring for the natural attenuation of heavy metals in the subsurface soil is a good case in point. The components that constitute the monitoring (sampling and testing) requirements include the following: • Distribution and concentration of target heavy metal contaminants; precipitated metals as distinct species • Speciations • pH and Eh

Because heavy metals are more likely to be sorbed onto soil solids, the monitoring program must include sampling of the subsurface soils for determination of sorbed metal species and solutes in the pore water. The other components such as speciation, pH, Eh, etc. are necessary as monitoring components because they provide guidance about the status of the metal pollution plume. In respect to organic chemical pollutants, according to the National Research Council protocol (NRC, 2000), metabolism by microorganisms is documented by: • Loss of the contaminant • Laboratory assays showing that microorganisms have the potential to transform the contaminant • Evidence that biodegradation is occurring

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Modeling parameters are often obtained by the laboratory assays. These assays involve measuring the decrease in the contaminant concentration caused by introduction of laboratory microcosms. While all these criteria cannot always be demonstrated, many others have been shown by the Christensen group (Nielsen et al., 1995), as for example the decrease in concentrations of BTEX in the anaerobic landfill leachate plume. A chloride tracer showed sorption was not significant. Laboratory experiments indicated that although toluene and xylene degraded, benzene did not degrade. Several reasons exist that can explain why laboratory and field conditions do not give the same results (Roling and van Verseveld, 2002). Some of these include: • Data obtained from the site are not representative or are very small compared with the size of the site. • Site heterogeneity in terms of geology, hydrology and geochemistry can cause significant variations in the microbial populations. • Microbial assays can be misleading or inaccurate, resulting in erroneous determination of the microbial activities and population. • Changes in the chemical or physical properties of the sample can alter microbial diversity. While tests should be performed as soon as possible, it should be recognized that incubation tests take a long time. Caution should be exercised when interpreting lab tests concerning the microbial activities and community from in situ samples.

The U.S. Department of Energy (DOE) has provided technical guidance for natural attenuation at their sites (DOE, 1999). As shown in Figure 7.4, three types of monitoring are utilized: • Performance monitoring within and adjacent to the plume • Detection monitoring at the boundary of the management zone • Ambient monitoring (upgradient of the plume)

Performance monitoring is used to determine the progress or the effectiveness of the natural attenuation process. At the same time, checks on the parameters that can impact directly on the process can be made. The protocols, future directions and detailed requirements are discussed in detail in Chapter 8. In this section, we will look at present concerns in monitoring procedures. For proper capture of subsurface information, monitoring wells need to be located not only in the contaminant plume but also along the sides and ahead of the plume. Figure 7.4 shows the major details of such a scheme. The number of sampling locations will depend on several factors. These include: • • • •

The nature of the pollutants and source location The dimensions and features of the existent pollutant plume The hydrogeology and subsurface features of the site Model predictions of plume advance

The intervals between monitoring stations should be sufficient to show the progress of the natural attenuation process, frequent at the beginning and then

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decreasing in frequency as the progress slows. Data should be able to demonstrate lines of evidence including: • Reduction in mass concentrations of the contaminant over time • Presence of conditions to allow geochemical or biological attenuation processes to occur • Indicators that geochemical or biological processes such as degradation processes are occurring • Analysis from the field of soil samples indicating precipitation or adsorption onto aquifer materials.

It is important to anticipate what will be measured or detected in the monitoring system. In particular, it is important to know what types of contaminants and concentrations can be expected at the various locations and at various time periods. Taking the example of the six groups of contaminants found at DOE sites (fuel hydrocarbons, chlorinated organics, high explosives, metals, inorganic anions and tritium) procedures and techniques required to detect each group will be different. Each group requires different and specific considerations in implementation of monitoring and testing schemes. For metals, since sorption can be either reversible or irreversible, it is necessary to incorporate measures that would permit examination of surface processes in the monitoring process. Sequential soil extractions can be used during the development of the conceptual model stage to determine the bioavailability of the metals. Mulligan et al. (2001) have shown that remedial techniques can be monitored by application of selective sequential extraction techniques. Sequential leaching tests can also be used during long-term monitoring of cationic metals. Measurement of pH is required since (1) cationic metals sorb strongly to the soil solids above pH 5 and (2) precipitation and coprecipitation of these metals occurs at high pH levels. Redox potential should also be determined because of its potential effect on, for example, iron hydroxide stability for cations. Redox conditions also impact directly on the reduction of chromium and technetium anions by soil organic matter. Regarding analytical methods, the DOE (1999) suggests the use of EPA publication SW 846 (U.S. EPA, 1987) and the American Society for Testing and Materials (ASTM) standards, (1992) or their own documentation (Goheen et al., 1997). 7.4.3

Established Monitoring Techniques

Chemical methods are used to determine the changes in concentrations of contaminants and the production of by-products and coreactants. Tracers are used to determine transport or sorption. Decreases in concentration of organic chemical pollutants, however, do not mean that complete degradation is taking place. Other pieces of evidence are required. These include (1) carbon dioxide and methane production, and/or (2) changes in the concentration of electron acceptors. A good review of many of the problems associated with evaluation of electron acceptors can be found in Christensen et al. (2000). Sorption, precipitation and volatilization can complicate verification of subsurface bioattenuation. The pH is a significant factor in determining the influence of the factors on biodegradation.

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Table 7.1 Parameters for Monitoring in Soil Parameter

Reason for Monitoring

Volatile organic compounds Polycyclic aromatic hydrocarbons Chlorinated solvents Heavy metals Total organic carbon Dissolved oxygen Carbon dioxide Nitrate Iron (III) Iron (II) Sulfate Chloride Oxidation reduction potential Alkalinity Temperature pH Hydroxide and carbonate content

Degree of soil contamination Degree of fuel components Degree of soil contamination Degree of soil contamination Evaluation of bioactivity in vadose zone Evaluation of bioactivity in vadose zone Evaluation of bioactivity in vadose zone Potential electron acceptor for organic compounds Potential for iron reduction Electron donor Electron acceptor for organic compounds Monitoring of chlorinated solvent reduction Indication of the nature of degradation Buffering capacity Inﬂuences rate of biodegradation Inﬂuences rate of biodegradation Potential for metal sorption or precipitation

Determination of intermediate metabolites can provide significant information regarding the biodegradation of the parent compound of interest. For metal contamination, potential by-products such as sulphide should also be monitored. The compound should also be stable chemically and biochemically. Knowledge of the various microbial pathways is required. Energy requirements can also have a significant influence on determining the probability of the existence of a compound. For example, if a compound requires significant energy for production, it is unlikely to be produced (Zwolinski et al., 2000). Table 7.1 provides some of the more significant parameters that need to be determined in the soil-water system for a better appreciation of the effectiveness of the NA treatment option. In addition to those parameters, are other measurements such as methane, used to indicate anaerobic production, and also turbidity and conductivity. The frequency of sample procurement depends on the rate of groundwater movement and spacing of monitoring wells. The fundamental requirement is to be able to determine the rates of contaminant attenuation. One should be prepared to alter the frequency of monitoring and sampling dependent on the results forthcoming from the ongoing monitoring and testing program. The sampling process should continue until it is proven that there is no risk to the health of humans or the environment. 7.4.4

Development of Monitoring Techniques

For the determination of metals and radionuclides in soil-water systems and groundwater, the techniques used in the field must provide information that would permit determination of mobility and bioavailability. Some of these, according to the Sandia National Laboratory (Waters et al., 1998), include scanning electron microscopy (SEM), isotope exchange techniques and soil digestion.

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Newer techniques for detecting and evaluating populations of microorganisms that are culture independent are being developed and utilized to provide a better understanding of the natural attenuation process. Polymerase chain reaction (PCR) is one of these techniques. This is based on small quantities of gene fragments. The potential for treatment of specific contaminants can be determined by evaluating the presence of various genes responsible for degradation or transformation (Brockman, 1995; Stapleton et al., 1998). Probes can be developed to evaluate the presence of various microbial strains. Most of the current information available is related to the species responsible for the aerobic biodegradation of BTEX and PAHs. These are the Pseudomonas species. The database called Genbank (www.ncbi.nlm.nih.gov/entrez) contains the available sequence information. This method is complicated by the fact that a specific pollutant can be biodegraded via various pathways. For this method to be viable, a large amount of information on gene fingerprints and environmental data needs to be collected from many sites. Davis et al. (2002) examined 16S RNA gene sequences from a chlorinated ethane–contaminated aquifer to characterize the microorganisms in the sediment. The reductive and oxidative mechanisms occurring in the aquifer were supported by the types of organisms present. The results of studies on natural attenuation of uranium in a tailings disposal site reported by Abdelouas and coworkers (Abdelouas et al., 1998, 2000, 2002) indicate that uranium does not significantly adsorb on the surfaces of the sandstone minerals owing to the negative charges of the carbonate minerals. Biological reactions to precipitate U(VI) by conversion to U(IV) are also possible with sulfatereducing bacteria. Other contaminants such as technetium can potentially be removed from subsurface soils and groundwater via reduction of Te(VII) and precipitation of TeO2 and/or TeS2 by sulfate-reducing bacteria. Sulfate-reducing bacteria (SRB) are found in many types of environments including metal-contaminated media such as soil, mining residues and wastewaters, among others. They can function in both anaerobic and aerobic environments. The results of the study of the distribution of SRB at a uranium mill tailings disposal site (Chang et al., 2001) showed the presence of dissimilatory sulphite reductase (DSR) gene sequences, suggesting the dominance of the genus Desulfotomaculum in up to 1500 ppb of uranium in the groundwater. The role of these SRB in reducing U(VI) to U(IV) and Te(VI) to Te(IV) needs to be studied. Other techniques under development include: • Measurements of the isotopes 13C/12C to determine if hydrocarbon degradation occurs • RNA probes from specialized bacteria for chlorinated solvent degradation • Determination of oxygenase activity

Oxygenases are required for the cometabolic degradation of chlorinated solvents (Sinke, 2001). Pennington et al. (1999) performed an evaluation of the natural attenuation of explosives in soil. They determined that the stable isotope of soil organic matter was too small for detection and could not be related to changes in

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

2,4,6-trinitrotoluene (TNT) concentration. However, they found a correlation of stable isotope ratios of nitrogen with changes in TNT in groundwater. Most of the studies to evaluate the effectiveness of NA monitor groundwater to determine the presence of electron acceptors. Since, however, Fe(III) and SO42- result in the mineral forms of Fe and S, the mineral forms of iron sulphide and solid Fe(II) and Fe(III) need to be determined (Kennedy et al., 1998). This is because aqueous forms may not show noticeable changes during the natural attenuation process. A mild acid (0.05 N HCl) can be used to extract Fe(II) followed by spectrometric analysis. Fe(III) and sulphides can be analyzed after strong acid extraction (6 N HCl) and further chromium and zinc precipitation for the sulphides (acid volatile sulphides) (Kennedy et al., 1998). Background conditions should also be examined in noncontaminated sections of the aquifer. This analysis can lead to a much better estimate of whether sulfate reduction is a factor in the development of the processes contributing to the natural attenuation of contaminants. There is consensus that better sensor technology needs to be developed to meet the requirements for proper monitoring of NA. Some of the newer sensor tools include: • Direct push smart probes and cone penetrometer system (CPT) penetrometer. The CPT can be coupled with various detectors such as a membrane interface probe system (MIPS) for determining volatile organic compounds (VOCs) in the subsurface. • Laser-induced fluorescence (LIF) or discrete multilayer samples (DMLs) for vertical contaminant determination. LIF can be used to measure hydrocarbons in real time in undisturbed soil and groundwater. • CPT-laser-induced fluorescence delineation. This technique was used in a qualitative manner to show the soil areas affected by contamination from underground storage tanks and piping by BTEX at Keesler Air Force Base in Biloxi, Mississippi (Weidemeier et al., 1999). Information of a qualitative and semiquantitative nature can be obtained.

Various types of sensors can be integrated with CPT. Hydrocarbons such as fuel oil, motor oil, grease, coal tar, gasoline, jet fuel and diesel fuel can be screened by this method. The boundaries of a plume can be detected before and during treatment to monitor the progress of the remediation. Figure 7.5 shows a fiber-optic LIF sensor coupled to a CPT. A specific wavelength of light generated by a laser is conducted downward to a fiber-optic cable toward a sapphire window at the cone tip, which advances in the subsurface. Two- and three-ring aromatic compounds and PAHs are detected as the laser causes them to fluoresce. The induced fluorescence returns over the second fiber to the surface detector system. The intensity of the signal is a measure of the concentration of the contaminant. In general, sandy soils offer higher fluorescence responses than clay soils. Other components in the subsurface can also give fluorescent response signals including de-icing agents, antifreeze and other detergents. The availability of a response signal database is essential if proper differentiation between the various receptors is to be obtained. Detection limits are generally 50 to 1000 mg/kg, and responses are linear up to about 10,000 mg/kg. Calibration is performed by spiking known quantities into representative soil samples. The main advantage of this type of system is that real-time data can be obtained

FIELD PERFORMANCE AND ASSESSMENTS

Figure 7.5

237

Laser-induced ﬂuorescence sensor attached to a cone penetrometer for detection of polycyclic aromatic hydrocarbons.

in the field. CPT pushes (penetration into the subsurface) of 65 to 100 m can be done in a 10 hour period and can only be undertaken in unconsolidated materials such as sediments and soft-to-medium clays. Limitations in the optical fiber constrain the CPT push to a depth of 50 m. This type of analysis is most cost-effective for large sites. Geophysical methods traditionally used in the petroleum and mining industries have recently been adapted for hazardous waste sites and indirect detection of aqueous and nonaqueous phase contaminants such as light NAPLs (LNAPLs) and dense NAPLs (DNAPLs). These methods are useful for detection of buried drums and other structures at hazardous waste sites that are sources of contamination. Geophysical methods that can be used for monitoring natural attenuation phenomena include electromagnetometry, magnetometry and ground-penetrating radar (Figure 7.6). The main types of contaminants that can be characterized by these methods include petroleum compounds, chlorinated solvents, polyaromatic hydrocarbons and inorganic compounds. Electrical conductivity and resistivity (Figure 7.7), electromagnetometry and to some extent ground-penetrating radar can directly detect contaminants by determining changes in soil conductivities by the chemicals. Other techniques include magnetometry, seismic reflection and refraction, and gamma logging. These are able to identify subsurface characteristics that provide pathways for contaminant movement. Other techniques that are used increasingly for monitoring natural attenuation are summarized in Table 7.2. A technique for sampling soil vapors or gases that has been described by Tartre (2001) involves purging the soil with nitrogen or another inert gas to determine the

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Geophones

Source

Overburden

Seismic wave direction

Rock

Figure 7.6

Schematic of ground-penetrating radar.

Battery or current source

Current meter

Voltage meter

Current flow

S=electrode spacing Figure 7.7

Schematic of resistivity survey.

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Table 7.2 Direct Push Instruments Being Developed for Monitoring Natural Attenuation Processes, Particularly Related to Soil Properties Instrument Cone pressure and sleeve friction Laser-induced ﬂuorescence

Modeling software with global positioning system and autocad Moisture probe (time domain reﬂectometry) and conductivity Laser-induced breakdown spectroscopy Soil gas sampler

Video microscope

Application For identiﬁcation of ﬂow pathways and conﬁning layers Deﬁning hydrocarbon plumes and possible degradation (currently experimental) through emission spectra Conceptual models for geological and chemical mapping, chemical fate and transport Determination of unsaturated hydraulic conductivity and location of ionic fronts from saltwater intrusion, river recharge and inﬁltration Determination of metal contaminants in soil Detection of volatile organic compounds from contaminants or biological activity, detection of respiration rates Grain size determination, permeability and hydraulic conductivity estimation, cone pressure conﬁrmation and free-phase nonaqueous phase liquids (under development)

Source: Adapted from Meuzelaar, H.L.C., Ed., Field Analytical Chemistry and Technology, Vol. 2, No. 2, John Wiley & Sons, New York, 1998.

volatile phase in equilibrium with the dissolved, sorbed or free hydrocarbon phase. The purging technique can be applied to both the vadose and saturated zones. Mass transfer and mass production from a small volume of soil can be determined.

7.5 MODELS TO SIMULATE NATURAL ATTENUATION 7.5.1

Background on Modeling

Models are used to predict the fate and transport of contaminants for the vadose and saturated zones. Some of the more complex ones reviewed by Weidemeier et al. (1999) include SESOIL, VADSAT, Jury’s model and SAM. These models require from 19 to 40 site and chemical parameters for input, depending on the model. SESOIL is the only one that considers chemical processes such as cation exchange capacity, hydrolysis and metal complexation. However, groundwater dilution of the leachate is not taken into account. A comparison of these models by Connor et al. (1994) shows that the main differences are (1) volatilization is neglected by the SAM model and (2) better correlation of results between models is obtained for transport in clay than in sandy soil. Models have been developed for evaluation of NA and prediction of impacts on receptors. They can also be used to determine the effects of source removal or reduction, pump and treat, and the impact of barriers on NA. Models can predict plume length either without calibration or for fitting with observed data as a calibration procedure for the model. Risk-based corrective action program use models without calibration, whereas the Air Force Natural Attenuation Initiative

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

is based on the second approach and the modeling results are used as lines of evidence. Both analytical and numerical models can be used to simulate the conditions at a contaminated site. Analytical models provide exact solutions consistent with their mathematical formulations. This should by no means be construed to mean that they provide exact solutions to real field problems. This will only happen when the analytical model accurately represents all the processes and boundary conditions in the field problem. Numerical models provide approximate solutions and are generally favored because of their capability to handle a wide range of hydrogeological conditions. Many processes must be approximated for analytical models. The heterogeneity of the subsurface can lead to over- or underestimation of contaminant movement. Decisions concerning what model to use should be based on available data, the complexity of the site and the accuracy of the information required. Assumptions within each of the models must be known to understand their limitations. Analytical models are used to simulate advection, biodegradation, dispersion and sorption in one, two or three dimensions as shown in Table 7.3. They are more useful if limited data is available. Numerical models are more applicable for cases where there is heterogeneity in the aquifer. Therefore, distribution of the hydraulic and contaminant properties within the aquifer is required. 7.5.2

Available Models for Natural Attenuation

Various packages are available for predictive modeling of NA. The analytical model BIOSCREEN developed for the Air Force’s Center for Environmental Excellence by Groundwater Services Inc. (Houston, TX) is used to simulate the natural attenuation from petroleum fuel releases. The Domenico analytical transport model is the basis for the model that is in a Microsoft Excel format and includes the assumption that the source is infinite and concentrations from the source do not Table 7.3 Common Analytical Models for Fate and Transport Processes Simulated Advection, dispersion, linear sorption biodegradation with constant or decaying source Advection, dispersion, linear sorption biodegradation with constant source Advection, dispersion, linear sorption biodegradation with constant or decaying source

Description

References

Semi-inﬁnite solute transport with continuous or decaying source, ﬁrstorder biodegradation decay, concentration over time and distance in one dimension Semi-inﬁnite solute transport with continuous or decaying source, ﬁrstorder biodegradation decay, concentration over time and distance in two dimensions Semi-inﬁnite solute transport with continuous or decaying source, ﬁrst order biodegradation decay, concentration over time and distance in three dimensions

Bear (1972) Van Genuchten and Alves (1982) Wexler (1992) Wilson and Miller (1978)

Domenico (1987)

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241

change. The BIOSCREEN model assumes a declining source concentration. Attenuation processes for the soluble hydrocarbons include advection, dispersion and adsorption, in addition to aerobic and anaerobic degradation. Since only vertical zones are considered from the source, the program is more applicable for LNAPL than DNAPL sites (Wiedemeier et al., 1999). Regarding biodegradation, decreases in electron acceptors such as dissolved oxygen, nitrate and sulfate and metabolic by-product formation [iron(II) and methane] are also incorporated. These parameters are accepted as evidence that biodegradation is occurring, and are used to determine if NA is feasible or whether other lines of evidence are needed. The more sophisticated BIOPLUME model has been at Superfund and Resource Conservation and Recovery Act (RCRA) sites. Outputs from the model include graphs of the plume centerline, plume concentrations as a three-dimensional color plot, and mass balances showing removal by each electron acceptor. A mass flux calculator has been added to Version 1.4 to show the mass fluxes at various points in the plume. BIOCHLOR is a model that simulates the natural attenuation of chlorinated solvents. Some of the features included in the program include (1) determination of reductive dechlorination of chlorinated ethenes to ethenes and chlorinated ethanes to ethenes in sequential steps, (2) separation of higher dechlorination rates into a different zone from lower dechlorination rates, (3) development of a database to assist in the prediction of first-order decay rates for input into BIOCHLOR and (4) incorporation of advection, dispersion and adsorption in the program scheme. Both this software and BIOSCREEN are available from the U.S. EPA’s Center for Subsurface Modeling Support Internet Site. The database is found at www.gsinet.com. The BIOSCREEN program was used to estimate the time for remediation of BTEX by natural attenuation processes at a facility contaminated with fuel hydrocarbons (Suarez and Rifai, 2002). Field data for monitoring along a flow line were used to calibrate the model, and input information included field data such as hydraulic conductivity and gradient, porosity, retardation factor and biodegradation decay rate. Information on biodegradation rate was obtained from a well at the source, and benzene was used as the key compound to obtain the attenuation rate. With this calibration, the data agreed fairly closely with the measured values, and an estimate of a 250-year time period was made for effective biodegradation of the benzene in the contaminated site, with an initial concentration of approximately 900 mg/L at the source. One of the major drawbacks of using models such as BIOSCREEN is that a first-order decay coefficient is used to calibrate the models (Odencrantz et al., 2001). The Domenico solution for advection-dispersion-biodegradation is given by

È x C C( x, t ) = 0 exp Í 2 ÍÎ 2a x

Ê 1 + 4la x Á1 v Ë

È 4la x Í x - vt 1 + ˆ˘ v Í ˜ ˙erfc Í 2 a x vt ¯ ˙˚ Í Î

˘ ˙ È ˘ ˙erf Í Y ˙ ˙ Í2 ayx ˙ ˚ ˙ Î ˚ (7.1)

242

Figure 7.8

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Impact of attenuation processes at a site. The more the attenuation processes, the higher will be the overall attenuation rate. (Adapted from Newell, C.J., Rifai, H.S., Wilson, J.T., Connor, J.A., Aziz, J.A. and Suarez, M.P., Calculation and Use of First-Order Rate Constants for Monitored Natural Attenuation, Ofﬁce of Research and Development, National Risk Management Research Laboratory, EPA/540/S-02/500, Cincinnati, OH, 2002.)

where Co is the initial concentration, ax is the longitudinal dispersivity, ay is the transverse dispersivity, l is the rate of biodegradation, t is time, x is distance from source, v is the retarded velocity due (v = vs/R, R is the retardation factor and vs is seepage velocity) and Y is the width of the source. The effect of the various factors on the calculations is shown in Figure 7.8 and Figure 7.9. The first-order decay coefficient, i.e., the rate of biodegradation l, is adjusted until it coincides with field analyses. All information regarding dispersion, sorption and biodegradation is represented as one parameter. This can lead to significant inaccuracies. More details can be found at www.epa.gov/ada/bioscreen.html. The BIOPLUME III two-dimensional model (Rifai et al., 1997) simulates natural attenuation by taking into account advection, dispersion, ion exchange and biodegradation. Both aerobic and anaerobic electron acceptors are included. The U.S. Geological Survey (USGS) formed the basis of the model in July 1989 (Konikow and Bredehoeft, 1989). It is mainly used for evaluation of groundwater flow and transport. Biodegradation reactions can be simulated by instantaneous degradation, first-order decay and Monod kinetics. Electron acceptors include oxygen, nitrate, iron, sulfate and carbon dioxide. One of the main limitations is that components such as BTEX are treated as one component, not as separate ones. This is particularly problematic since benzene and toluene are more biodegradable than xylene. Other models are available such as BIOPLUME II (Rifai et al., 1988) and Bio1D (Srivinsan and Mercer, 1988). They are applied mainly for evaluation of the transport and biodegradation of petroleum compounds. BIOPLUME III can simulate the aerobic and anaerobic biodegradation of organic contaminants. The electron

FIELD PERFORMANCE AND ASSESSMENTS

Figure 7.9

243

Effect of the different mechanisms of contaminant concentration vs. time near the source. (Adapted from Newell, C.J., Rifai, H.S., Wilson, J.T., Connor, J.A., Aziz, J.A. and Suarez, M.P., Calculation and Use of First-Order Rate Constants for Monitored Natural Attenuation, Ofﬁce of Research and Development, National Risk Management Research Laboratory, EPA/540/S-02/500, Cincinnati, OH, 2002.)

acceptors, oxygen, nitrate, iron (III), sulfate and carbon dioxide are depleted sequentially. Transport and production of Fe(II) can also be simulated if it is used as an electron acceptor. First-order decay, instantaneous reactions and Monod kinetics can be simulated. BIOPLUME III can be used to determine the extent of the plume and its persistence with or without engineered controls. As an example, this model was applied to examine the feasibility of natural attenuation at Patrick Air Force Base (Rifai et al., 2000). It was determined that it would take 30 to 40 years to stabilize the plume at a distance of less than 170 m from the source area. The breakdown of aerobic and anaerobic biodegradation was 30 and 65% of the total losses, respectively. The models MODFLOW and RT3D (Reactive Transport in Three Dimensions; Sun et al., 1996) are also used extensively. Both sorbed (sorption and desorption processes) and aqueous species can be simulated, in addition to dissolution of NAPL. Contaminants include heavy metals, explosives, petroleum and chlorinated hydrocarbons. These are three-dimensional models for simulation of solution solute and microbial transport with applications for natural attenuation and enhanced bioremediation processes. BioRedox-MT3DMS is a three-dimensional multispecies fate and transport model that was developed to predict the performance of intrinsic or accelerated

244

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

bioremediation remedies for petroleum hydrocarbons and chlorinated solvents. BioRedox was developed from the MT3DMS fate and transport model and works in conjunction with MODFLOW. An oxidation-reduction reaction database is incorporated into the model. Iron hydroxide depletion in the soils can be modeled. Biological reactions can be first order or instantaneous. Other models are described in Table 7.4. 7.5.3

Application, Calibration and Veriﬁcation of Models

Utilization of analytical and computer models to a particular site requires contaminant transport data. In addition to usage of models for prediction and analysis of attenuation events at a site, models can be used to determine if more data are needed. These can be location specific or general. In the final analysis, models provide us with an appreciation of the effectiveness of NA at the site under investigation. An approach for a model methodology can be seen in Figure 7.10. A well thought out conceptual model (Figure 7.11) is the key to a successful natural attenuation model. It aids in the understanding of the site, assessment of available data and determination of other required inputs and further required sampling. The choice of model and the suitability of the particular model depend on the limitations, assumptions and site characteristics such as hydrogeology (Corapcioglu and Baehr, 1987; Carey et al., 1998). The scope, availability and quality of input data are the key to a successful model, particularly if the site is complex. Calibration, verification and prediction are all essential to improve the accuracy of the model for site conditions. Calibration is used to adjust the results so that they can accord better with measured data. Biodegradation rate constants are often modified when the results obtained are not close to field data. Use of laboratory data can lead to serious errors when these are extrapolated to field scale. To a large extent, this is because of a lack of proper scaling laws. The use of data from other sites for application to the site under investigation can lead to significant errors in predictions because of the differences between the native conditions in the two sites. At all times, sitespecific field data should be inputted into the model. The contribution of biotic factors should be determined so that the models can be used to accurately predict the movement and concentration of contaminants and the effectiveness of NA (Mercer, 1998; Rifai et al., 1995b). Models such as BIOSCREEN (Newell et al., 1996) and BIOCHLOR (Aziz et al., 2000) use first-order rates for attenuating contaminants as they leave the source zone. The rates and their applications are not clearly defined, and natural attenuation processes must be clearly understood for proper application of the rate constants. Three types of rate constants (the first two shown in Figure 7.12) are used (Newell et al., 2002): 1. Pollutant concentration vs. time constant (inverse time) to show the time for achievement of remediation goals. It is determined by plotting the natural log concentration at a specific location vs. time and determining the slope. 2. k concentration vs. distance constant (inverse time), which is a bulk attenuation rate to show if the plume is changing in size owing to dispersion, biodegradation

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245

or other attenuation processes. It is determined by plotting the natural log concentration at a specific location vs. distance, determining the slope and multiplying by the groundwater seepage velocity. 3. l-Biodegradation rate constant (inverse time) for soluble contaminants to determine the effect on solute transport. It is determined in various ways such as by comparison of contaminant and tracer transport or by calibrating a model (BIOSCREEN, BIOCHLOR, BIOPLUME, MT3D) with field data. Field studies can utilize a tracer or the steady-state contaminant plume method of Buscheck and Alcantar (1995). For the latter method, the number of wells must be sufficient and the plume must be at steady state to give appropriate results.

Source degradation and reduction owing to the various attenuation processes (Figure 7.8 and Figure 7.9) can have significant influences on contaminant concentrations. In addition, uncertainties in the calculations for the rate of biodegradation can originate from sources including the placement of monitoring wells, variations throughout the year, sampling and analytical procedures, as well as the heterogeneity at the site (Newell et al., 2002). The typical rate values shown in Table 7.5 contain rate constants for some lower molecular PAHs. High molecular-weight compounds were not determined, and it is not known how the presence of higher molecularweight compounds influences the degradation of the lower molecular-weight ones. The models typically incorporate linear or reversible performances. However, for PAHs, sorption is often nonlinear and hysteric and highly dependent on the type of media. In the future, models will need to determine the attenuation of PAHs by dilution, sorption and biodegradation in a more specific manner and less of a “black box” depletion approach (Rogers et al., 2002). Regarding the use of models, the NRC (2000) discussed several common problems with their application. We cite three categories as follows: 1. The first category is related to the model framework that includes conceptual models and data used as the basis for these models. The present and future site conditions must agree with the assumed boundary conditions, equations and parameters used. Data should not be taken from another site, and observations must be weighted appropriately. The most appropriate model for the site should be used. 2. The second category essentially states that a closed mind should not be maintained when evaluating the model and the field data. Models should not be forced if they do not fit the data. Other models or an alternate conceptual model should be considered if this problem arises. 3. The final category involves a model’s end results. The model should not be extrapolated beyond its ability, nor should its accuracy be overestimated. Uncertainties in the data, data assumptions, methods of verification and validation, parameters inputted and the complexity of the site can also influence the model’s results.

7. 6 PROTOCOLS DEVELOPED FOR NATURAL ATTENUATION The various protocols for use of NA as a remediation tool, together with the lines of evidence and evidence of success of NA plus enhanced natural attenuation

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Table 7.4 Summary of Available Natural Attenuation Model Soils and Groundwater Model Name

Applicability

BIO1D

1-D model for biodegradation and sorption of hydrocarbons BIOREDOX 3-D model for chlorinated solvents and petroleum hydrocarbons; couples biodegradation and reduction of oxygen, nitrate, sulfate and carbon dioxide BIOPLUME II 2-D model for transport of a single dissolved hydrocarbon species with oxygen-limited biodegradation, ﬁrst order decay, linear sorption, advection and dispersion BIOPLUME III 2-D model for multiple hydrocarbons

BioTracker

CHEMFLO

3DFATMIC

MODFLOW

MT3D

PESTAN

RITZ

RT3D

SESOIL SWIFT, SWIFT/486

1-D natural attenuation screening model with visualization tools for groundwater multispecies transport and transformation; used with Sequence and BioTrends; based on Bioredox To simulate water movement and chemical transport in unsaturated soils and the convection-dispersion equation (chemicals); public domain 3-D model to simulate subsurface ﬂow, transport and fate of contaminants that are undergoing chemical and/or biological transformations for both saturated and unsaturated zones Model for estimating the vertical migration of dissolved organic solutes through the vadose zone to groundwater; a closed-form analytical solution of the advective-dispersive-reactive transport equation For simulation of unsaturated zone ﬂow and transport of oily wastes; partitioning of pollutant between the liquid, soil, vapor and oil phases by linear equilibrium isotherms; degradation of pollutant and oil is described as ﬁrst-order process Finite difference model for transient and steady state groundwater ﬂow; used with transport models MT3D, Biotrans, RAND3D 3-D transport model for advection, linear and nonlinear sorption dispersion, ﬁrst-order decay of single species; coupled with MODFLOW Modiﬁcation of MT3D; for multispecies transport of chlorinated compounds, by-products and solid-phase species; instantaneous aerobic degradation, BTEX degradation with multiple electron acceptors, sequential anaerobic degradation of PCE/TCE, and combined aerobic/anaerobic degradation of PCE/TCE Transport through vadose zone; can be combined with MODFLOW 3-D ﬁnite difference model to simulate contaminant, ﬂuid and heat transport in porous and fractured media; linear and nonlinear desorption, dispersion, diffusion, dissolution, leaching and decay; public domain

Developer or Reference GeoTrans Inc. Carey et al. (1998)

Rifai et al. (1988)

Development commissioned by AFCEE Scisoftware

http://www.epa.gov/a da/csmos/models/ch emﬂo.html http://www.epa.gov/a da/csmos/models/3d fatmic.html McDonald and Harbaugh (1988)

S.S. Papadopoulos and Associates Inc.

http://www.epa.gov/a da/csmos/models/pe stan.html http://www.epa.gov/a da/csmos/models/rit z.html Washington State University and Paciﬁc Northwest National Laboratory

U.S. Salinity Lab Sandia National Laboratory

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247

Table 7.4 Summary of Available Natural Attenuation Model Soils and Groundwater (Continued) Model Name SWMS_2D

UTCHEM

VLEACH

Applicability

Developer or Reference

3-D Modeling transient and steady-state ﬂow and U.S. Salinity Lab mass transport in the groundwater (saturated) and vadose (unsaturated) zones of aquifers. Physical, chemical and biological processes. Includes multiple organic NAPL phases; the dissolution and/or mobilization of NAPLs by nondilute remedial ﬂuids, chemical and microbiological transformations and changes in ﬂuid properties; includes nonequilibrium interphase mass transfer, sorption, geochemical reactions and the temperature dependence of pertinent chemical and physical properties; model includes inhibition, sequential use of electron acceptors, and cometabolism for a general class of bioremediation processes 2-D model for transport of water and solutes in various http://www.epa.gov/a saturated media; linear sorption, zero-order da/csmos/models/ut production, ﬁrst-order decay, dispersion; public chem.html domain. 1-D ﬁnite difference model for evaluating effects on http://www.epa.gov/a groundwater from the leaching of volatile, sorbed da/csmos/models/vl contaminants through the vadose zone; includes each.html liquid-phase advection, solid-phase sorption, vaporphase diffusion, and three-phase equilibration in terms of soil properties, recharge rates, depth of water, or initial conditions; public domain.

Note: NAPL, nonaqueous phase liquids Source: Weidermeier, T.H., et al., Technical Protocol for Evaluation Natural Attenuation of Chlorinated Solvents in Groundwater, U.S. Air Force Center For Environmental Excellence, Technology Transfer Division, Brooks Air Force Base, San Antonio, TX, 1996.

(ENA) and engineered natural attenuation (EngNA) are developed in Chapter 8 in terms of where we should go from here. These recognize the need or usefulness in obtaining as much supporting information as possible, to support the knowledgebased requirements for application of natural attenuation processes in the treatment of contaminated ground. In this section, we will pay more attention to present usage of protocols and requirements as evidenced from field applications of NA. 7.6.1

Various Technical Protocols

Various technical protocols have been established such as the Designing Monitoring Programs to Effectively Evaluate the Performance of Natural Attenuation by the Air Force Center for Environmental Excellence (Weidemeier and Haas, 1999). Sampling type, frequency and location, and analyses required for NAPL contaminants are described. Concentrations of dissolved oxygen, nitrate, Fe(II), sulfate and methane can be determined with contaminant data obtained downgradient of the plume. Monitoring wells should be located to obtain this data, and sampling frequency should be determined on the basis of distribution and location of monitoring stations, evidence of attenuation, and hydrogeology. Computational models and

248

Figure 7.10

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Simple procedure for model development.

statistical methods can be used to determine the time of cleanup and plume behavior to aid in the planning of monitoring frequency and monitoring intensity. A summary of well locations and analytical parameters to be monitored is shown in Table 7.6 for a LNAPL plume. Background groundwater quality information can be obtained from up- and sidegradient wells (see Figure 7.4). Those monitoring stations within the plume should provide information that indicates the progress of the natural attenuation process. Downgradient wells can be employed to monitor the plume progress and contaminant concentrations. A full set of geochemical data for hydrocarbon contaminants includes dissolved oxygen, nitrate, Fe(II), sulfate and methane concentrations, pH, conductivity, alkalinity and redox potential. Additional parameters are required for chlorinated hydrocarbons including chloride, total organic carbon and hydrogen. The EPA 1998 Technical Protocol for Evaluation of Natural Attenuation of Chlorinated Solvents in Groundwater (Weidemeier et al., 1998) was established to demonstrate mechanisms of chlorinated solvent natural attenuation. The EPA (2001) points out, however, that field information can be substantially different from laboratory research, particularly regarding dechlorination rates and product concentrations. For chloromethane, chlorothanes, chlorinated benzenes and chlorinated ethers, the reductive dechlorination rates need to be compared in the laboratory and field. Field information is not available for many processes and their reaction rates. To complicate matters, there are other uncertainties such as those related to interactions

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Figure 7.11

249

Schematic showing conceptual model for nonaqueous phase liquid release from a leaking underground storage tank.

A

Natural Log Pollutant Concentration

Slope = kpoint

Time

B Natural Log Pollutant Concentration

Slope = k/v

Distance from source

Figure 7.12

Determination of (A) time rate constant (kpoint) and (B) distance rate constant (k).

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Table 7.5 Typical Rate Constants Obtained from Field Data Rate Constant Point decay rate constant (kpoint)

Bulk attenuation rate (k)

Biodegradation rate constant (l)

Rate Constant Values from Field Data For benzene, 0.46 per year and for MTBE, 0.44 per year (49 gas stations) For benzene, 0.14 per year and for MTBE, 0.04 per year For benzene, 0.33 per year (159 plumes at gas stations); for TCE, 0.15 per year (37 plumes) For BTEX, 0.001 to 0.01 per day For benzene, 0.0005 to 0.005 per day (14 sites) For BTEX, 0.001 to 0.01 per day; for chlorinated solvents (aerobic/anaerobic), 0 to 1.96 per day For TCE, 0.001 to 0.008 per day (10 sites) For cis DCE, 0.002 to 0.008 per day (9 sites) For VC, 0.002 to 0.15 per day (7 sites) For anaerobic BTEX plume, 0.003 to 0.03 per day for Elmendorf Air Force Base, Alaska, 0.004 per day used in Bioplume II model For total BTEX, 0.01 per day along center line and 0.02 per day along periphery at Hill Air Force Base For TCE, cis DCE and VC, near the source, 1.24, 0.75 and 0.29 per year, respectively, 300 m from source, 0.3, 0.07 and 0.47 per year, respectively, at Plattsburgh Air Force Base For PCE, 4.0 per year (4 sites) For TCE, 1.1 per year (18 sites) For cis-DCE, 1.6 per year (13 sites) For VC, 1.3 per year (6 sites) For naphthalene, 0.00057–0.0063 per day For acenaphthylene, 0.00027 per day For phenanthrene, 0.00027–0.063 per day For MTBE, 0.0001 to 0.001 per day

Reference Reid and Reisenger (1999) Peargin (2002) Newell et al. (2002)

Newell et al. (2002) Rifai et al. (1995a) Suarez and Rifai (1999) Aziz et al. (2000)

Weidemeier et al. (1999)

Wilson (1998)

Rogers et al. 2002

Day (2000)

Note: MTBE, Methyl tert-butyl ether; TCE, trichloroethylene; BTEX, benzene, toluene, ethylbenzene and xylene; DCE, dichloroethylene; VC, vinyl chloride; PCE, perchloroethylene

with other contaminants, high concentrations at the source area and mechanisms for degradation other than reductive dechlorination. The protocol by the ASTM applies to petroleum contamination from underground storage tank releases into the groundwater (ASTM, 1998). Included in the protocol are requirements and considerations such as: • Site characterization, evaluation of potential risks and the ability to meet remediation goals • Stable or shrinking size of pollutant plume

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Table 7.6 Well Location and Sampling Requirements for Monitoring of a Nonaqueous Phase Liquid (NAPL) Plume Well Location

Purpose

Long-Term Sampling

Up- or sidegradient

Backgroundwater quality monitoring

Source area

Determination of composition or source concentration Plume monitoring over time

No analysis or a limited set of parameters including dissolved oxygen, redox potential, temperature and pH unless changes are indicated or indicator contaminants are detected Analysis of indicator contaminants in the NAPL and groundwater below NAPL and limited set of geochemical parameters

Downgradient of source along plume center line Downgradient of plume Contigency well

Surface water

Detection of plume migration Determination of migration toward receptors Determination of impact on surface water

Analysis of indicator contaminants unless there is a change in plume behavior when full set required Analysis of indicator contaminants unless there is a change in plume behavior when full set required Analysis of indicator contaminants unless there is a change in plume behavior when full set required Analysis of indicator contaminants unless there is a change in plume behavior when full set required

Source: Adapted from Weidemeier, T.H., Rifai, H.S., Newell, C.J., and Wilson, J.T., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley and Sons, New York, 1999.

• Time frame for achieving the targets and goals to be determined by the regulatory agency • Integration of source control into the remedial decisions • The decision to remove or allow the source to remain, to be made by the regulatory agency • Monitoring requirements and implementation, which should be based on sitespecific conditions and maintained until objectives are met • Contingency plans to be structured and implemented if the objectives cannot be met by natural attenuation

To prove that natural attenuation is occurring, lines of evidence are established to indicate decreases in contaminant concentrations (NRC, 1993). They include: • Decreases in contaminant concentration and/or plume size over time. This is used to indicate that biodegradation is occurring faster than plume size is increasing. • Chemical indicators of microbiological activity in the groundwater chemistry such as consumption of oxygen, nitrate and sulfate and production of Fe(II), Mn(II) and methane. • Laboratory microcosm studies, which are used to simulate aquifer conditions to determine if bacteria at the site can biodegrade the contaminants and at what rate. This step is mainly used if neither of the first two clearly indicates significant trends.

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Of the three lines of evidence indicated by the NRC (1993), it is suggested that primary lines are usually required. Secondary or tertiary lines are required only if primary lines are insufficient. Site-specific attenuation rates can be used as a secondary line of evidence showing that attenuation is occurring and the rate at which it is occurring. The time frame establishment of goals and the requirement for source control are determined by the regulatory agency. The remediation goals must be established earlier, which differs from the U.S. Air Force Protocol (Weidermeier et al., 1995). Monitoring frequency is according to the potential impact on receptors, ability to meet remedial goals, plume behavior and institutional controls used. Contingency plans are necessary only if the goals cannot be met. The details of protocols for lines of evidence, evidence of success and monitoring provided in Chapter 8 differ somewhat from these. The requirements and protocols discussed in Chapter 8 pay attention to the same issues and factors discussed in this chapter but add the soil-contaminant interaction components to the assessment procedures. The EPA guidance document covers all contaminants in soil and groundwater at RCRA and Superfund sites and for underground storage tanks (U.S. EPA, 1999). For the OSWER policy on natural attenuation (Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites), if natural attenuation is chosen for remediation, the following must be demonstrated: • Natural attenuation is occurring as expected. • Toxic by-products are identified [trichloroethylene (TCE), dichloroethylene (DCE), vinyl chloride, methyl mercury, etc.]. • Extent of plume expansion is demonstrated. • New releases or other environmental conditions that could affect the progress of natural attenuation are identified. • Cleanup objectives can be obtained. • Contingency plans have been made in the event that natural attenuation is not effective.

Concerns that need to be addressed include (1) release of minerals such as arsenic and manganese into the aquifer (U.S. EPA, 1998) in the course of remediation of petroleum or chlorinated hydrocarbons and (2) transfer of other components across the various media — from groundwater to soil, soil to air or groundwater to surface water. The NRC (2000) has reviewed other protocols. It is useful to note, from the document, that the number of samples, parameters to be monitored and contaminant concentrations to be obtained can vary substantially among the protocols. For example, in addition to those mentioned above, 14 federal, state, professional and industry protocols were evaluated and reported on in the document, together with community concerns, scientific and technical issues and implementation issues. Table 7.7 reviews the protocols. Only seven have been peer reviewed. In general, the NRC (2000) noted that only the DOE guidance entitled Site Screening and Technical Guidance for Monitored Natural Attenuation at DOE Sites (Brady et al., 1998) and the EPA protocol address inorganic contaminants but in a very limited way. The DOE protocol gives some guidance regarding the sorption and sequestration of inorganics by giving

Inorganic and organic Chlorinated solvents Chlorinated solvents Fuels Fuels Chlorinated solvents Fuels Fuels Inorganic and organic

EPA (1999) RTDF (1997) EPA ORD (1998) Navy (1998) Air Force (1995) Air Force (1997) ASTM (1997) API (1997) DOE (1998)

D M ND M M D D ND ND

Contingency Plans ND D D D D D D D M

Science-Based Underpinning M ND D ND D D M ND ND

Source Characterization ND M M M M D M ND ND

Intrinsic Capacity

ND ND M ND ND ND D M ND

Peer Reviewed

ND D D D D D D D D

Usability

Note: D, discussed; M, mentioned; ND, not discussed or not applicable Source: Adapted from NRC, Natural Attenuation for Groundwater Remediation, Committee on Intrinsic Remediation, Water Science and Technology Board and Board on Radioactive Waste Management, Commission on Geosciences, Environment and Resources, National Academy Press, Washington, DC, 2000.

Contaminant Covered

Protocol

Table 7.7 Comparison of Available Natural Attenuation Protocols

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default values for sorption coefficients and suggesting that sequential extraction or isotopic pulsing could be used to evaluate irreversible uptake. The U.S. Air Force Technical Protocol for Implementing Intrinsic Remediation with Long-Term Monitoring for Natural Attenuation of Fuel Contaminants (Weidemeier et al., 1995) was evaluated as one of the most scientifically sound protocols, as it describes ways to estimate biological activity, dilution, sorption and dispersion. Methyl tert-butyl ether (MTBE) is not addressed, however. A more recent document by the EPA (2001) has made various recommendations concerning MTBE including: • Evaluating the biodegradability of MTBE and other oxygenates under field conditions • Determining with more certainty the rate of dissolution of MTBE and other fuel components • Establishing a database on the natural attenuation of MTBE

In the same document, the EPA (2001) has also tried to establish procedures for evaluating the natural attenuation of inorganics. For example, it is suggested that (1) the natural attenuation mechanisms for arsenic and other inorganics should be determined, (2) the effect of geochemical conditions on remobilization needs to be established, (3) the effect of organics on inorganic contaminant behavior and vice versa should be studied, (4) guidelines need to be developed to help understand immobilization processes based on laboratory and field data and the use of models, and (5) uncertainty analysis needs to be incorporated into decision making. The majority of the available protocols address only fuel hydrocarbons or chlorinated solvents. Other organic contaminants such as PAHs, polychlorinated biphenyls (PCBs), explosives and pesticides are not addressed, and metals, inorganics and radionuclides are infrequently discussed. There are major shortcomings in these protocols. The following are some of the recommendations made by the NRC (2000): • Agreement is required on the use of protocols for natural attenuation and how they are used to obtain regulatory approval for natural attenuation. • All protocols should be peer-reviewed by independent experts. • Conceptual models are preferred over scoring systems as decision guidelines. • Documents to support the protocols should be easy to use. • Training by neutral organizations should be provided for proper use of the protocol.

7.6.2

Inclusion of Soils and Sediments in Protocols

One aspect not considered by the NRC, since its focus was on groundwater quality, was that most protocols are designed for groundwater natural attenuation and not for natural attenuation of contaminants in soils or sediments. Few protocols exist for soil, with the exception of the EPA’s (1999) and those by the DOE. Chapter 8 addresses this problem. Sediments have not received much attention at all. Sediments differ from soils since they can be transported. In addition, organisms can transport contaminants, and there is considerable variability at sites. Technical protocols have not been developed for sediments. The EPA (2001) has recommended

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that research be expanded to determine natural attenuation mechanisms in sediments. Monitoring methods need to be developed for quantifying natural attenuation, contaminant transport and bioaccumulation for analysis and assessment. Research specific to freshwater, coastal and marine aquatic environments is also required. Petroleum contaminants that are of low solubility tend to remain in the soil. It is imperative that these contaminants do not pose threats in the soil through direct contact or leaching to the groundwater. Leaching and environmental mobility of sorbed contaminants are significant agents for release of these contaminants for transport to biotic receptors. As another example, chlorinated solvents have the potential to slowly leach into the groundwater over long periods of time. For inorganic contaminants in soils, it is particularly important to identify the mechanisms of interaction such as the type of sorption or redox reaction for retention. Some precipitation and absorption reactions are quite stable such as precipitation of cesium onto clay minerals, whereas others such as surface adsorption of uranium onto ironoxide minerals and organic complexation are less stable or more reversible. The presence of chelating agents such as EDTA can also increase the mobility of trivalent chromium. Other factors such as pH changes, redox, chemical speciation or concentration changes may also influence mobility. Radionuclides can also decay to other daughter products such as Pu-241 to Am241 and Np-237. These can be more mobile and toxic than the parent. Radionuclides are also a concern since direct contact is not necessary for harm to come to humans. Gamma radiation or x-rays from below can increase the risk to humans at the near surface. Natural attenuation of inorganic contaminants can be effectively applied if the process or mechanism for attenuation is demonstrated to be irreversible.

7.7 CASE STUDIES OF NATURAL ATTENUATION 7.7.1

Natural Attenuation of Chlorinated Solvents

Fuel components such as benzene, toluene, ethylbenzene and xylene, known as BTEX, are readily biodegraded in the subsurface. Others such as chlorinated hydrocarbons are not believed to degrade easily. Compounds such as tetrachloroethylene (PCE) are converted by reductive dechlorination to dichloroethylene (DCE) and vinyl chloride (VC). PCE Æ TCE Æ DCE Æ VC Æ ethylene or ethane Increasing water solubility Æ An analysis of 61 BTEX-contaminated sites showed no reductive chlorination occurred at 23 of them. Accumulation of DCE was detected at 18, and vinyl chloride production was at 20 sites (McNab et al., 2000). It had previously been expected that vinyl chloride would accumulate (Vogel et al., 1987). However, no accumulation of vinyl chloride, ethylene or ethane were detected since it was likely that bacteria in the subsurface could have converted these components to carbon dioxide with

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various electron acceptors including iron (III), manganese (IV) and oxygen. Anaerobic oxidation of vinyl chloride or cis-DCE could have occurred with the electron acceptors oxygen, iron (III), manganese (IV) or natural organic matter. Brigmon et al. (1998) determined that sorption of the TCE from the water onto the soil was a major mechanism, removing up to 90% of the TCE from the water in the U.S. Department of Energy Savannah River Site, SC, where contamination of the groundwater by TCE had occurred. Microcosm tests were designed for the evaluation of abiotic and biotic attenuation. Linear partitioning coefficients, kd, were determined to be from 0.83 to 7.4 mL/g, and organic carbon partitioning coefficients, koc, were from 72 to 180 mL/g of carbon. Diffusion was a significant factor in the attenuation process. Only small amounts (less than 5%) of TCE were biodegraded, and the main biological mechanism was anaerobic reductive dechlorination, as indicated by the presence of the by-products cis-1,2-dichloroethylene (c-DCE) and trans-1,2-dichloroethylene (t-DCE). Addition of methane, oxygen and methanol did not enhance biodegradation, leading to the conclusion that sorption was the principal attenuating mechanism. 7.7.2

Natural Attenuation of MTBE

The natural attenuation of MTBE, particularly in the groundwater, has been a matter of much discussion. Emphasis has been placed on its potential for biodegradation. In the Odencrantz (1998) evaluation of field studies of MTBE mobility across North America, the plumes were found to be very large despite the distance from the source. The results of three field studies are summarized in Table 7.8. These cases showed no real evidence of biodegradation. Plumes were up to 330 m long in the sandy aquifers, with aquitards 6 to 17 m below the water table. This is in contrast to hydrocarbon fuel sites in California (Rice et al., 1995) and 75% of benzene plumes in Texas (Mace et al., 1997), which were less than 80 m long. Jansen et al. (2002) have used NA as part of, or as a sole procedure, for treatment of MTBE-contaminated sites with concentrations in soil or groundwater less than 5 ppb. This treatment procedure has been used at 15 gas station sites — with NA being the lone remedial technique at 6 of them. The review by Seagren and Becker (2002) on the natural attenuation of BTEX and MTBE concluded that MTBE was biodegradable in shallow aquifers but at rates slower than BTEX. Dispersion and dilution are important mechanisms for MTBE natural attenuation (Landmeyer et al., 1998). The NRC (2000) rates the level of understanding for natural attenuation of MTBE as moderate with a low chance of success. Field studies have generated highly variable results. Present information shows that MTBE is much more mobile than BTEX. This could lead to restrictions in the number of sites where natural attenuation can be used. However, there are indications that the MTBE plumes stabilize eventually. The review of the status of natural attenuation of MTBE in groundwater at several sites (Fiorenza et al., 2002) (Base Borden in Ontario, Canada; Sampson County site and Elizabeth City site in North Carolina; 74 Amoco service stations in the U.S. and the Laurel Bay Exchange site in South Carolina) indicated that:

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Table 7.8 Results of Methyl Tert-Butyl Ether (MTBE) Field Studies at Three Sites Site

Plume Characteristics

Transport Mechanisms

Biodegradability

Santa Monica Charnock (42 lb per month)

1200 ft length High permeability Minimal 50 to 150 ft wide aquitard Minimum of 10 ft thick Cross-connection between monitoring wells Poorly abandoned water wells

North Carolina State University (Borden) (3 lbs in 700 days)

580 ft long (up to 700 to 800 ft) 250 ft wide 28 ft thick on a clay aquitard

Waterloo MTBE study, Borden aquifer, Ontario (1.7 lb immediately)

800 to 1000 ft from Travel along top of No formation of source after 3000 aquitard for 7 to 8 years intermediates days Loss of MTBE by TBA or TBF after Plume 165 ft long dispersion or 7 years 150 ft. wide heterogeneities in No real evidence 15 to 20 ft in thickness region

Some degradation at Minimal source No decay downgradient Possible leakage through aquitard

Source: Adapted from Odencrantz, J.E., Implications of MTBE for intrinsic remediation of underground fuel tank sites, Remediation, 9(3), 7–16, 1998.

• Mechanisms for natural attenuation seem to be highly uncertain. • At the Borden site, only 3% of the MTBE remained, but there did not seem to be evidence of biodegradation. • Degradation of MTBE occurred only close to the source at the Sampson County site. A first order decay coefficient of 0.0010 per day for MTBE was obtained. Further positions downgradient from the source gave a decay coefficient of zero for MTBE, indicating a potential risk distant from the source. • A plume of methane (7 mg/L) was concurrent with that of MTBE at the Elizabeth City site. • MTBE biodegradation rates decreased according to the groundwater seepage velocity, 5.0 per year for the highest velocity to 2.2 per year for the lowest velocity. At that rate, it was estimated that MTBE concentrations could decrease to 30 mg/L in 60 years, from an initial source of 1200 mg/L. • Biodegradation rates were determined for 6 of the 74 gasoline sites and shown to range from 0.3 to 10.9 per year. There also seemed to be some association with methanogenic sites, although there is little support for degradation by methanogenic bacteria. • At the Laurel exchange site, although there could have been biodegradation, dilution and dispersion seemed to be the major attenuation mechanisms.

7.7.3

Natural Attenuation as a Sole Remediation Technology

Most of the reported case studies of NA have focused on the groundwater aspects of the application. This is not surprising, since the initial thrust of NA has been directed toward management of groundwater quality. Reported case studies on NA

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

in soils and sediments are relatively scarce. However, the evidence shows that there will be considerably more application of NA to contaminated sites as knowledge and understanding of the abiotic and biotic processes contributing to the natural attenuation of contaminants in soils and sediments increases. Some of the reported case studies of NA in soils and sediments will be described in this section. At a U.S. Department of Energy site (NRC, 2000) at Hanford 216-B-5 (in the state of Washington), although various radionuclides (90Sr, 137Cs, 239Pu and 240 Pu) were found in a plume, most appeared to be associated with the sediment solids close to the source through adsorption and precipitation mechanisms. Sorption of cesium onto biotite, smectite and vermiculite occurs by ion exchange mechanisms following diffusion in the mineral interlayers. While ion exchange is also responsible for sorption of strontium, there is apparently some unconfirmed precipitation of strontium with phosphate or BiPO4. Plutonium (IV) is associated with hydrous oxide solids in the sediments close to the source and does not pose a risk further downstream (DOE-RL, 1996) unless it somehow subsequently becomes detached and mobile. The ASTM risk-based approach (Figure 7.13) used by Khan and Husain (2001) on a site contaminated by 150,000 barrels of oil spilled from a pipeline concluded that attenuation of xylene and ethylbenzene would take 80 years. The procedure used for the analysis included the following steps (Figure 7.14):

Figure 7.13

Risk-based approach for site assessment. (Adapted from Khan, G.I. and Husain, T., J. Hazard. Mat., B85, 243–272, 2001.)

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Figure 7.14

259

Methodology for risk-based monitored natural attenuation (MNA) process. (Adapted from Khan, G.I. and Husain, T., J. Hazard. Mat., B85, 243–272, 2001.)

• • • •

Review of available data and characterization Conceptual model development Refinement of the conceptual model Modeling by analytical and/or numerical fate/transport models for soil and groundwater • Analysis of the exposure pathway • Planning for long-term monitoring • Obtaining approval for MNA

The three models used in the Khan and Husain (2001) case study included SESOIL for contaminant transport in the unsaturated zone, AT123D for contaminant transport in the saturated zone and BIOSCREEN for natural attenuation. In the vadose zone, it was estimated that 76% of the benzene was volatilized and the remainder biodegraded, with small amounts leaching into groundwater in 80 years. While toluene showed similar behavior to benzene, only 10% of the ethylbenzene and xylene were degraded. These behaved as NAPLs in an area that contained sandy clay with porous gravel and clay loam. Although the proportions differed in the four regions studied, it was evident that volatilization was a major mechanism, followed by biodegradation and then transport for benzene and toluene. The study of the natural attenuation of polychlorinated biphenyls and chlorinated ethenes at three sites, reported by Kastanek et al. (1999), indicates that the individual PCB congeners’ composition seemed to vary over the 20-year contamination period. This would indicate that some biodegradation of the lighter congeners occurred in

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

the oxygenated superficial soil layers. Anaerobic dechlorination and volatilization or solubilization into the groundwater may also have occurred. These could be responsible for the natural attenuation of the lighter PCBs. PCBs with fewer chlorine atoms can completely biodegrade, as confirmed in this study where the PCBs with low numbers of chlorines decreased more than the PCBs with a high content of chlorines. Reductive dechlorination of PCBs has been shown to occur in sediments (Sokol et al., 1994). Mechanisms for natural attenuation for PCBs must be further investigated. Natural attenuation of PCBs has also been studied in the sediments of the Hudson River (NRC, 2000). Highly chlorinated PCBs had become lightly chlorinated over the years (Brown et al., 1987) by microbial reductive dechlorination processes (Bedard and Quensen, 1995). For complete biodegradation to occur, aerobic and anaerobic processes must work together. Anaerobic zones can become aerobic through stream-channel and bioturbation processes. Recent studies by McNulty (1997) show that dechlorination of PCBs in the highly contaminated sediments occurs within the first year but drops off substantially subsequently, and it remains to be seen if complete dechlorination will occur even after decades. It is important to be aware that although there are signs that natural attenuation is proceeding, reduction in concentration and toxicity of the contaminants may never reach the target levels specified by regulatory standards. A study was conducted at an alpine skiing area (2875 m above sea level) for the natural attenuation and biostimulation of diesel in soil contaminated with diesel from field-incubated lysometers (2612 mg diesel fuel/kg) (Margesin and Schinner, 2001). Hydrocarbon contents were monitored in both the soil and leachate for three summers (July to September). After three summers, hydrocarbon levels decreased by 50% and 70% in the nonfertilized and fertilized soils, respectively. Levels still had 1296 mg/kg and 774 mg/kg of diesel after this time. Although significant levels of diesel fuel removal at extreme conditions could be achieved in the first two summers, the results from the third summer showed that removal rates had decreased considerably. Leaching and natural attenuation of kerosene have been studied by Dror et al. (2001) in field experiments with irrigation water over a 39-day period to a depth of 100 cm. Volatilization of many components was immediate, while transport was minimal. When the soil moisture was high, volatilization of low-vapor-pressure compounds was substantial since infiltration to deeper soil layers was delayed. The heavier compounds remained in the upper layers of the soil. Few studies have been performed on the natural attenuation of contaminants in clay soils. Berry and Burton (1997) examined the natural attenuation of diesel fuel (5000 mg/kg) in a smectitic clay soil. As can be expected from the discussions in Chapters 2 and 3, the hydraulic conductivity of the smectite soil would be very low, and thus, anaerobic conditions would most likely prevail, even near the soil surface. This could inhibit biodegradation. However, the low hydraulic conductivity of the clay soil would limit migration of the diesel to the groundwater. Two methods were used to determine complete attenuation and hydrocarbon degradation: (1) the extractable organic method (EPA 3520B and EPA 8000A) for determination of time for complete attenuation and (2) the enzyme-linked immunosorbent assay (ELISA) to show that hydrocarbon degradation was nearly complete.

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While complete attenuation was shown to occur within 74 days, the mechanism of attenuation could not be determined, inasmuch as the only indicator for complete attenuation was the disappearance of the organic chemical. No biological transformations to other by-products were determined in the course of the investigation. The low levels of hydrocarbons recorded indicated insignificant amounts of contaminant migration. 7.7.4

Combination of Natural Attenuation with Other Remediation Processes

NA by itself may not be sufficient to achieve effective remediation of contaminated sites in a reasonable time frame. When this is judged to be the case, NA can be combined with engineered treatment processes to reduce risk to biotic receptors and to enhance remediation rates and efficiency. Source reduction and oxygen addition are methods of enhancing NA. Source reduction technologies include free product removal, soil vapor extraction, bioventing and bioslurping, as described by Mulligan (2002). Oxygen addition can be achieved by biosparging and the addition of oxygen-releasing compounds. An example of an attempt to enhance NA is the field tests presently being conducted using sodium lactate (an electron donor in many biochemical reactions) injected into the groundwater to enhance biodegradation of TCE. Initially, 1135 L of lactate was injected weekly. After 5 weeks, concentrations of 3800 mg/L of TCE dropped to 10 mg/L. The injection program began in 1998, and currently 5000 L of lactate is injected every two months (Strzelecki, 2002). A similar approach can be adopted by adding hydrogen release compound (HRC; Regenesis, San Clemente, CA). After injection of the compound into the groundwater, lactic acid is slowly released and hydrogen gas is produced by the bacteria. The hydrogen serves as an electron donor for chlorinated hydrocarbon and nitroaromatic compounds. An application of HRC as an aid to NA can be found in the treatment of the Army Chemical Depot in Texas. Heavy metal and VOC contamination of the groundwater and soil in the facility occurred from the 1940s to 1974 (Koenigsberg and Vigue, 2002). The explosives that were discharged contaminated the upper aquifer with TNT, 2,4-dinitrotoluene (2,4-DNT), hexahydro-1,3,5-triazine (RDX), 1,3,5-trinitrobenzene (1,3,5-TNB) and nitrate. While the TNT naturally degraded, the other products did not. HRC was injected into the soil at 30 locations to enhance biodegradation. Monitoring was undertaken at 15 well locations. The monitored results showed the presence of the TNT by-products 2,4-DNT and 1,35-TNB. Significant reduction of all by-products was observed as follows 105 days after injection of the HRC: 72 to 98% for 2,4-DNT, 45 to 99% for RDX, 86 to 100% for 1,3,5-TNB and 52 to 94% for nitrate. A combination of NA and phytoremediation has been used at a site contaminated from underground storage tanks (Nzengung and Ramaley, 2001). Groundwater samples taken at the site showed the presence of PCE, TCE and cis-1-2-DCE, and surface-water lake samples showed PCE, cis-1,2-DCE, 1,1-DCE and vinyl chloride, while sediment samples indicated PCE and TCE. This indicates that many different processes contributing to natural attenuation were ongoing. Field and laboratory

262

Figure 7.15

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Zones of perchloroethylene attenuation. (Adapted from Nzengung, V.A. and Ramaley, S., Coupling natural attenuation and phytoremediation to clean up a shallow chlorinated solvent plume at the former Naval Training Center in Orlando, Florida, Proceedings of 2001 International Containment and Remediation Technology Conference and Exhibition (10–13 June), Orlando, FL, 2001.).

analyses showed that halorespiration mechanisms were occurring in one part of the sediments in the source area (group I in Figure 7.15), while cometabolic processes occurred in the sediments in the nonsource area (group II). Sorption was also an important removal mechanism since the foc in group II sediments was greater than 2.5%. Group III sediments were deep, and test results indicated limited amounts of biodegradation (only PCE to DCE occurred) and sorption. As natural attenuation processes would not completely dehalogenate PCE in the nonsource area and deep sediments, it could not be used as the only technology for remediation. Therefore, amendments were used to accelerate reductive dechlorination. Laboratory experiments showed this could be accomplished through addition of carbon sources, particularly for the deep sediments. Phytoremediation was utilized since reductive dechlorination and mineralization of PCE and TCE occurs with willows and cottonwood trees. The trees would enable (1) removal of residual chlorinated solvents after source removal, (2) reduction of the high discharge from the aquifer into the lakes and (3) enhancement of natural attenuation by increasing the dissolved organic carbon in the aquifer. Therefore, willow trees were planted near the source area and a wetland was constructed near the lake. While most engineered remedial technologies are designed to have a positive effect on attenuation of contaminants, there can be negative effects (Weidemeier and Haas, 1999). This factor should be considered when making contingency plans. For example, if air is introduced by chemical oxidants, oxygen-releasing materials, air sparging, bioslurping or pump and treat to an aquifer contaminated with chlorinated solvents, it can disrupt the natural reductive dechlorination processes. Other processes may lead to the spread of contaminants into previously uncontaminated areas

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such as pump and treat and in-well circulation for petroleum hydrocarbons and soil vapor extraction for nondegradable VOCs. These factors should always be considered if natural attenuation is to be used with an engineering remedial technology.

7.8 CONCLUDING REMARKS In this chapter, we have examined the various aspects of field application of NA. We have looked at assessing the potential for natural attenuation, the methods for monitoring, the models and protocols used and some reported case studies. We have seen that there are various advantages to natural attenuation including: • The possibility of destroying the contaminant completely • A remediation method that is potentially more acceptable to the public than other remediation technologies • A technology that can be used with other remediation technologies as a pre- or posttreatment at a site • A technology that can potentially reduce remediation costs significantly

Potential disadvantages are: • Longer remediation times compared with other technologies • Lack of knowledge concerning mechanisms for remediation, particularly with regards to inorganic contaminants • Substantial and complex requirements for monitoring • By-products that can be more environmentally hazardous or mobile than the parent compounds • Possible desorption, resolubilization or transfer of contaminants to other media • Changes in the hydrological or geochemical conditions, which can influence stability and mobility of contaminants, such as naturally occurring metals • Slow public acceptance, requiring more education • The need for modeling data to be evaluated with caution and examined from where the data originated and the assumptions were made

Many similarities exist between existing protocols and guidelines for NA, particularly concerning lines of evidence and data requirements. Very few consider the soil and soil gas in their protocols, and most have been adapted for hydrocarbon and chlorinated solvent contamination. The natural attenuation of many other contaminants such as PAHs, PCBs, pesticides and inorganic contaminants has not been investigated extensively. Monitoring techniques will need to be substantially improved, particularly in the soil. The factors influencing the rate of NA of compounds that are not readily attenuated, including metals, chlorinated solvents and high molecular weight organics will need to be determined. For monitoring, direct indicators of bioremediation, particularly for NA (e.g., genetic markers, metabolic by-products and techniques for characterizing the fate of contaminants, including degradation products, during natural attenuation) will need to be further developed.

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Available computational models are mainly concerned with groundwater plumes. Integration of soil and groundwater models is complicated. Care should also be exercised regarding source and quality of input data. Interaction of organic and inorganic contaminants with the soil components are important factors in the processes that contribute to the natural attenuation of contaminants. Inorganic compounds such as manganese and iron oxides can actively promote catalytic activity for the remediation of soils containing organic chemical pollutants. REFERENCES Abdelouas, A., Fattahi, M., Grambow, B., Vichot, L., and Gautier, E., Precipitation of technetium by subsurface sulfate-reducing bacteria, Radiochim. Acta, 90, 773–777, 2002. Abdelouas, A., Lutze, W., Gong, W., Nuttall, E. H., Strietelmeier, B.A. and Travis, B.J., Biological reduction of uranium in groundwater and subsurface soil, Sci. Total Environ., 250, 21–35, 2000. Abdelouas, A., Lutze, W., and Nuttall, E., Chemical reactions of uranium in groundwater at a mill tailings site, J. Contamin. Hydrol., 34, 343–361, 1998. Alexander, M., Biodegradation and Bioremediation, Academic Press, London, 1994. American Public Health Association, Standard Methods for the Examination of Water and Wastewater, 18th ed., Published jointly by American Public Health Association, American Water Works Association and Water Pollution Control Federation, 1992. Anderson, M.P. and Woessner, W.W., Applied Groundwater Modeling, Academic Press, San Diego, CA, 1992. American Society for Testing and Materials (ASTM), Standard Guide for Remediation of Groundwater by Natural Attenuation at Petroleum Release Sites, ASTM Designation E1739-95, ASTM, West Conshohocken, PA, 1995. American Society for Testing and Materials (ASTM), Standard Guide for Remediation of Groundwater by Natural Attenuation at Petroleum Release Sites, ASTM Designation E1943-98, ASTM, 1998. Aziz, C.E., Newell, C.J., Gonzales, J.R., Haas, P.E., Clement, T.P., and Sun, Y., BIOCHLOR Natural Attenuation Decision Support System, User’s Manual Version 1.1, U.S. EPA, Office of Research and Development, EPA/600/R-00/008, Washington, DC, 2000, www.epa.gov/ada/csmos/models.html. Bear, J., Dynamics of Fluids in Porous Media, Dover, New York, 1972, 764 pp. Bedard, D.L. and Quensen, J.F., III,, Microbial reductive dechlorination of polychlorinated biphenyls, in Microbial Transformation and Degradation of Toxic Organic Chemicals, Young, L.Y. and Cerniglia, C.E., (Eds.), Wiley-Liss, New York, 1995, pp. 127–216. Bennett, P.C., Hiebert, F.K., and Roger, J.R., Microbial control of mineral-groundwater equilibria: Macroscale to microscale, Hydrogeology J., 8, 47–92, 2000. Berry, K.A. and Burton, D.L., Natural attenuation of diesel fuel in heavy clay soil, Can J. Soil. Sci., 77, 469–477, 1997. Brady, P.V., Spalding, B.P., Krupka, K.M., Waters, R.D., Zhang, P., Borns, D.J., and Brady, W.D., Site Screening and Technical Guidance for Monitored Natural Attenuation at DOE Sites, Sandia National Laboratory, Albuquerque, NM, 1998. Brigmon, R.L., Bell, N.C., Freedman, D.L., and Berry, C.J., Natural attenuation of trichloroethylene in rhizosphere soils at the Savannah River site, J. Soil Contam., 7, 433–453, 1998. Brockman, F.J., Nucleic-acid based methods for monitoring the performance of in situ bioremediation, Mol. Ecol., 4, 567–578, 1995.

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Brown, J.F., Bedard, D.L., Brennan, M.J., Carnahan, J.C., Feng, H., and Wagner, R.E., Polychlorinated biphenyl dechlorination in aquatic sediments, Science, 236,709–712, 1987. Brown, R.A., Hicks, R.J., and Hicks, P.M. , Use of air sparging for in situ bioremediation, in Air Sparging for Site Remediation, Hinchee, R.E., Ed., Lewis, Boca Raton, FL, 1994. Buscheck, T.E.G. and Alcantar, C.M., Regression techniques and analytical solutions to demonstrate intrinsic bioremediation, in Proceedings of the 1995 Battelle International Conference on In-Situ and On-Site Bioreclamation, Battelle Press, Columbus, OH, 1995, pp. 109–16. Carey, G.R., van Geel, P.J., Murphy, J.R., McBean, E.A., and Rover, F.A., Full-scale field application of a coupled biodegradation-redox model (BIOREDOX), in Natural Attenuation of Chlorinated Solvents, Wickramanayake, G.B. and Hinchee, R.H., eds., Batelle Press, Columbus, OH, 1998, pp. 213–218. Chang, Y.-J., Peacock, A.D., Long, P.E., Stephen, J.R., McKinley, J.P., MacNoughton, S.J., Hussain, A.K.M.A., Saxton, A.M., and White, D.C., Diversity and characterization of sulfate-reducing bacteria in groundwater at a uranium mill tailings site, Appl. Environ. Microbiol., 67, 3149–3160, 2001. Christensen, T.H., Bjerg, P.L., Banwart, S.A., Jakobsen, R., Heron, G., and Albrechtsen, HJ., Characterization of redox conditions in groundwater contaminant plumes. J. Contam. Hydrol. 45 (3), 2000, 165–242. Connor, J.A., Newell, C.J., Nevin, J.P., and Rifai, H.S., Guidelines for use of groundwater spreadsheet models in risk-based corrective action design, in Proceedings of the National Groundwater Association Petroleum Hydrocarbons and Organic Chemicals in Groundwater Conference, Houston, TX, Nov., 1994, pp. 43–55. Corapcioglu, M.Y. and Baehr, A.L., A compositional multiphase model for groundwater contamination by petroleum products. 1. Theoretical considerations, Water Resour. Res., 23 (1), 191–200, 1987. Davis, J.W., Odom, J.M., DeWeerd, K.A., Stahl, D.A., Fishbain, S.S., West, R.J., Klecka, G.M., and DeCarolis, J.G., Natural attenuation of chlorinated solvents at Area 6, Dove Air Force Base: Characterization of microbial community structure, J. Contam. Hydrol., 57, 41–59, 2002. Day, M.J., Fate and transport of fuel components below slightly leaking underground tanks, Soil Sediment and Groundwater MTBE, special issue, March, 21–24, 2000. DOE, Technical Guidance for the Long-Term Monitoring of Natural Attenuation Remedies at Department of Energy Sites, October 8, 1999. DOE-RL, 200-BP-5 Operable Unit Treatability Test Report, DOE/RL-96-59, U.S. Richland, WA, Department of Energy, Richlands Operations Office, 1996. Domenico, P.A., An analytical model for multidimensional transport of a decaying contaminant species, J. Hydrol., 91, 49–58, 1987. Dror, I., Gerstl, Z., and Yaron, B., Temporal changes in kerosene content and composition in field soil as a result of leaching, J. Contam. Hydrol., 48, 305–323, 2001. Fiorenza, S., Suarez, M.P., and Rifai, H.S., MTBE in groundwater: Status and remediation, J. Environ. Eng., 128(9), 773–781, 2002. Goheen, S., McCulloch, M., Riley, R., Sklarew, D., Sharma, A., and Fadeff, S., DOE Methods for Evaluating Environmental and Waste Management Samples, Batelle Press, Columbus, OH, 1997. Jansen, R., Moyer, E., Woodward, R., and Sloan, R., MTBE Remediation Seminar Series, Spring, 2002. Kao, C.M. and Wang, Y.S., Field investigation of the natural attenuation and intrinsic biodegradation rates at an underground storage tank site, Environ. Geol., 4, 622–631, 2001.

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Kastanek, F., Demnerova, K., Pazlarova, J., Burkhard, J., and Maleterova, Y., Biodegradation of polychlorinated biphenyls and volatile chlorinated hydrocarbons in contaminated soils and groundwater in field condition, Int. Biodeterior. Biodegrad., 44, 39–47, 1999. Kennedy, L.G., Everett, J.W., Ware, K.J., Parsons, R., and Green, V., Iron and sulfur mineral analysis methods for natural attenuation assessments, Bioremediation J., 2(3), 259–271, 1998. Khan, G.I. and Husain, T., Risk-based monitored natural attenuation — a case study, J. Hazard. Mat., B85, 243–272, 2001. Koenigsberg, S.S and Vigue, B.W., The next big thing: Cost-effective DOD sites using slowrelease compounds, Pollut. Eng., 34(3), 14–17, 2002. Konikow, L.F. and Bredehoeft, J.D., Computer model of two-dimensional solute transport and dispersion in groundwater, in Techniques of Water Resources Investigation of the United States Geological Survey, Book 7, U.S. Geological Survey, Reston, VA, 1978. Kota, S., Borden R.C., and Barlaz, M.A., Influence of protozoa grazing on contaminant biodegradation, FEMS Microbiol. Ecol., 29, 179–189, 1999. Landmeyer, J., Chapelle, F., Bradley, P., Pankow, J., Church, C., and Tratnyek, P., Fate of MTBE relative to benzene in a gasoline-contaminated aquifer (1993–1998), Groundwater Monit. Rev., 18(4), 93–102, 1998. Lovley, D.R., Reduction of iron and humics in subsurface environments, in Subsurface Microbiology and Biogeochemistry, Fredickson, J.K. and Fletcher, M., Eds., WileyLiss, New York, 2001. Mace, R.E., Fisher, R.S., Welch, D.M., and Parra, S.P., Extent, Mass and Duration of Hydrocarbon Plumes from Leaking Petroleum Storage Tank Sites in Texas, Bureau of Economic Geology Geological Circular 97–1, 1997, 52 pp. Margesin, R. and Schinner, F., Bioremediation (natural attenuation and biostimulation) of diesel-oil-contaminated soil in an alpine glacier skiing area, Appl. Environ. Microbiol., 67, 3127–3133, 2001. McDonald, G. and Harbaugh, A.W., A modular three-dimensional finite-difference ground water flow model, U.S. Geological Survey Techniques of Water Resources, Investigations, Technical Report, U.S. Geol. Survey, Reston, VA, 1988, 6, Chap. A1. McNab, W.W., Jr., Rice, D.W., and Tuckfield, C., Evaluating chlorinated hydrocarbon plume behavior using historical case population analyses, Bioremediation J., 4, 311–335, 2000. McNulty, A.K., In situ anaerobic dechlorination of PCBs in Hudson River sediments, Master’s thesis, Rensselaer Polytechnic Institute, Troy, NY, 1997. Mercer, J.W., Transport and remediation of chemicals in groundwater, in Encyclopedia of Environmental Analysis and Remediation, Meyers, R.A., Ed., John Wiley & Sons, New York, 1998. Meuzelaar, H.L.C., Ed., Field Analytical Chemistry and Technology, Vol. 2, No. 2, John Wiley and Sons, New York, 1998. Mulligan, C.N., Environmental Biotreatment, Government Institutes, Rockville, MD, 2002, 395 pp. Mulligan, C.N., Yong, R.N., and Gibbs, B.F., The use of selective extraction procedures for soil remediation, Proc. Int. Symp. on Suction, Swelling, Permeability and Structure of Clays, Balkema, Rotterdam, The Netherlands, 2001. NATO/CCMS, Pilot Study Evaluation of Demonstrated and Emerging Technologies for the Treatment and Clean Up of Contaminated Land and Groundwater (Phase III) 2002, Annual Report, Number 255, North Atlantic Treaty Organization, EPA-542-R-02010, 2003. Newell, C.J., McLeod, R.K., and Gonzales, J., BIOSCREEN Natural Attenuation Decision Support System, EPA/600/R-96/087, August, 1996, www.epa.gov/ada/csmos/models.html.

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Newell, C.J., Rifai, H.S., Wilson, J.T., Connor, J.A., Aziz, J.A. and Suarez, M.P., Calculation and Use of First-Order Rate Constants for Monitored Natural Attenuation, Office of Research and Development, National Risk Management Research Laboratory, EPA/540/S-02/500, Cincinnati, OH, 2002, 25 pp. Nielsen, P.H., Bjarnadottir, H., Winter, P.L., and Christensen. T.H., In situ and laboratory studies on the fate of specific organic compounds in an anaerobic landfill leachate plume. 2. Fate of aromatic and chlorinated aliphatic compounds, J. Contam. Hydrol., 20, 51–66, 1995. NRC (National Research Council), In Situ Bioremediation: When Does It Work? National Academy Press, Washington, DC, 1993. NRC (National Research Council), Natural Attenuation for Groundwater Remediation, Committee on Intrinsic Remediation, Water Science and Technology Board and Board on Radioactive Waste Management, Commission on Geosciences, Environment and Resources, National Academy Press, Washington, DC, 2000, 274 pp. Nzengung, V.A. and Ramaley, S., Coupling natural attenuation and phytoremediation to clean up a shallow chlorinated solvent plume at the former Naval Training Center in Orlando, Florida, Proceedings of 2001 International Containment and Remediation Technology Conference and Exhibition (10–13 June), Orlando, FL, 2001. Odencrantz, J.E., Implications of MTBE for intrinsic remediation of underground fuel tank sites, Remediation, 9(3), 7–16, 1998. Odencrantz, J.E., Varljen, M.D., and Vogel, R.A., Natural attenuation: Is dilution the solution?, Lust Bull., 40(March), 8–12, 2002. Peargin, T.R., Relative depletion rates of MTBE, benzene, and xylene from smear-zone nonaqueous phase liquid, in Bioremediation of MTBE, Alcohols and Ethers, Proc. Sixth Int. In Situ and On-Site Bioremediation Symposium, Magar, V.S., Gibbs, J.T., O’Reilly, K.T., Hyman, M.R., and Leeson, A., (Eds.), San Diego, CA, June 4–7, 2001, Batelle Press, Columbus, OH, 2002, pp. 67–74. Pennington, J.C., Miyares, P.H., Ringelberg, D.B., Zakikhani, M., Reynolds, C.M., Felt, D., Coffin, R.B., Gunnison, D., Cifuentes, L., Fredrickson, H.L., and Jenkins, T.F., Natural Attenuation of Explosives in Soil and Water Systems at Department of Defense Site, Prepared for Strategic Environmental Research and Development Program, U.S. Army Corps of Engineers, Waterways Experiment Station, Technical Report SERDP99-1, July, 1999. Prabhakar Clement, T., Truex, M.J., and Lee, P., A case study for demonstrating the application of U.S. EPA monitored natural attenuation screening protocol at a hazardous waste site, J. Contam. Hydrol., 59, 133–162, 2002. Reid, J.B. and Reisenger, H.J., Comparative MtBE versus Benzene Plume Length Behavior BP Oil Company Florida Facilities, prepared by Integrated Sciences and Technology, Marietta, Georgia, for BP Oil Company, Cleveland, OH, 1999. Renner, R., Natural attenuation’s popularity outpaces scientific support, NRC finds, Environ. Sci. Technol., 34, 203A–204A, 2000. Rice, D.W., Dooher, B.P., Cullen, S.J., Everett, L.G., Kastenberg, W.E., Grose, R.D., and Marino, M.A., Recommendation to Improve the Cleanup Process for California’s Leaking Underground Fuel Tanks (LUFTs), report submitted to the California State Water Resources Control Board and the Senate Bill 1764 Leaking Underground Fuel Tank Advisory Committee, California Environmental Protection Department, Sacramento, CA, 1995, 20 pp. Rifai, H.S., Bedient, P.B., Wilson, J.T., Miller, K.M., and Armstrong, J.M., Biodegradation modeling at aviation fuel spill site, J. Environ. Eng., 114, 1007–1029, 1988.

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Rifai, H.S., Borden, R.C., Wilson, J.T., and Ward, C.H., Intrinsic bioattenuation for subsurface restoration, in Intrinsic Bioremediation, Hinchee, R.E., Wilson, J.T., and Downey D.C., Eds., Batelle Memorial Institute, Columbus, OH, 1995a, pp 1–29. Rifai, H.S., Newell, C.J., Gonzales, J.R., Dendrou, S., Kennedy, L., and Wilson, J., BIOPLUME III Natural Attenuation Decision Support System, Version 1.0 User’s Manual, prepared for the U.S. Air Force Center for Environmental Excellence, Brooks Air Force Base, San Antonio, TX, 1997. Rifai, H.S., Newell, C.J., Gonzales, J.R., and Wilson, J.T., Modeling natural attenuation of fuels with BIOPLUME III, J. Environ. Eng., 126, 2000, 428–438. Rifai, H.S., Newell, C.J., Miller, R., Taffinder, S. and Roundsville, M., Simulation of natural attenuation with multiple electron acceptors, Bioremediation. 3, 53–58, 1995b. Rogers, S.W., Ong, S.K., Kjartanson, B.H., Golchin, J., and Stenback, G.A., Natural attenuation of polycyclic aromatic hydrocarbon-contaminated sites: Review, Practice Periodical Hazardous Toxic Radioactive Waste Manage., 6, 141–155, 2002. Roling, W.F.M. and van Verseveld, H.W., Natural attenuation: What does the subsurface have in store?, Biodegradation, 13, 53–64, 2002. Seagren, E.A. and Becker, J.G., Review of natural attenuation of BTEX and MTBE in groundwater, Practice Periodical Hazardous Toxic Radioactive Waste Manage., 6, 156–172, 2002. Sinke, A., Monitored natural attenuation: moving forward to consensus, Land Contam. Reclam., 9(1), 111–118, 2001. Seagren, E. A. and Becker, J.G., Review of Natural Attenuation of BTEX and MTBE in Groundwater Practice Period. Hazard. Toxic Radioact. Waste Manage. 6 (3), 2002, 156–162. Sokol, R.C., Kwon, O-Seob. and Bethoney, Ch.M., Reductive dechlorination of PCBs in St. Laurence River sediments and variations in dechlorination characteristics, Env. Sci. Technol., 28, 2054–2064, 1994. Srivinson, P. and Mercer, J.W., Simulation of biodegradation and sorption processes in groundwater, Groundwater, 26(4), 475–487, 1988. Stapleton, R.D., Ripp, S., Jimenez, L., Cheol-koh, S., Fleming, J.T., Gregory I.R., and Sayler, G.S., Nucleic acid approaches in bioremediation: Site assessment and characterization, J. Microbiol. Methods, 32, 165–178, 1998. Steollenwerk, K.G., Geochemical interactions between constituents in acidic groundwater and all uvum in an aquifer near Globe, Arizona, Appl. Geochem., 9, 353–369, 1994. Strzelecki, D., Low-tech remedies save millions at DOE site, Pollut. Eng., 34(3), 41, 2002. Suarez, M.P. and Rifai, H., Biodegradation rates for fuel hydrocarbons and chlorinated solvents in groundwater, Bioremediation J., 3, 337–362, 1999. Suarez, M.P. and Rifai, H.S., Evaluation of BTEX remediation by natural attenuation at a coastal facility, Ground Water Monitoring and Remediation, 22, 62–77, 2002. Sun, Y., Petersen, J.N., Clement, T.P., and Hooker, B.S., A monitoring computer model for simulating natural attenuation of chlorinated organics in saturated ground-water aquifers, in Proceedings of the Symposium on Natural Attenuation of Chlorinated Organic in Groundwater, Dallas, TX, Sept. 11– 13, EPA/540/R-96/509, U.S. EPA, Washington, DC, 1996. Surampalli, R. and Banerji, S., Long-term performance monitoring at natural attenuation site, Practice Periodical Haz. Toxic Radioactive Waste Manage., 6, 173–176, 2002. Tartre, A., Plume Delineation and Monitoring of Natural Attenuation Processes via in situ Flux Measurement, Presented at 2001 International Containment and Remediation Technology Conference and Exhibition, 10–13 June, Orlando, FL, 2001.

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U.S. EPA, Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvent in Groundwater, EPA/600/R-98/128, National Risk Management Laboratory, Ada, OK, 1998. U.S. EPA, Office of Solid Waste and Emergency Response (OSWER), Compendium of Superfund Field Operations Methods, OSWER Directive 9355.0-4, 1987. U.S. EPA, Office of Solid Waste and Emergency Response (OSWER), Use of Risk-Based Decision-Making in UST Corrective Action Programs, Directive 9610.17, 1996. U.S. EPA Office of Solid Waste and Emergency Response (OSWER), Use of Monitored Natural Attenuation at Superfund, RCRA Corrective Action and Underground Storage Tank Sites, Directive Number 9200.4-17P, April 21, 1999, 32 pp. U.S. EPA Science Advisory Board, Monitored Natural Attenuation; USEPA Research Program: An EPA Science Advisory Board Review, Science Advisory Board (1400A), Washington, DC, 2001. Van Genuchten, M.Th. and Alves, W.J., Analytical Solutions of the One-Dimensional Convective-Dispersive Solute Transport Equation, Technical Bulletin 1661, U.S. Department of Agriculture, Washington, DC, 1982, 151 pp. Vogel, T.M., Criddle, C.S., and McCarty, P.L., Transformations of halogenated aliphatic compounds, Environ. Sci. Technol., 21, 722–736, 1987. Waters, R.D., Brady, P.V., and Borns, D.J., Natural Attenuation of Metals and Radionuclides: An Overview of the Sandia/DOE Approach, Presented at Waste Management ’98, Tucson, AZ, Spring, 1998. Weidemeier, T.H., et al., Technical Protocol for Evaluation Natural Attenuation of Chlorinated Solvents in Groundwater, U.S. Air Force Center For Environmental Excellence, Technology Transfer Division, Brooks, Air Force Base, San Antonio, TX, 1996. Weidemeier, T.H. and Haas, P.E., Designing Monitoring Programs to Effectively Evaluate the Performance of Natural Attenuation, U.S. Air Force Center for Environmental Excellence, San Antonio, TX, 1999. Weidemeier, T.H., Rifai, H.S., Newell, C.J., and Wilson, J.T., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley and Sons, New York, 1999. Weidemeier, T., Swanson, M.A., Moutoux, D.E., Gordon, E.K., Wilson, J.T., Wilson, B.H., Kampbell, D.H., Haas, P.E., Miller, R.N., Hansen, J.E., and Chapelle, F.H., Technical Protocol for Evaluating Natural Attenuation of Chlorinated Solvents in Groundwater, EPA/600//R-98/128, September, EPA Office of Research and Development, Washington, DC, 1998. Weidemeier, T., Wilson, J.T., Kampbell, D.H., Miller, R.N., and Hansen, J.E., Technical Protocol for Implementing Intrinsic Natural Remediation with Long-Term Monitoring for Natural Attenuation of Fuel Contamination Dissolved in Groundwater, Vol. I and II, Air Force Center for Environmental Excellence Technology Transfer Division, Brooks, AFB, San Antonio, TX, 1995. Wexler, E.J., Analytical solution for one- two- and three-dimensional solute transport in groundwater systems with uniform flow, in Techniques of Water Resources Investigations of the United States Geological Survey, Book 3, U.S. Geological Survey, Washington, DC, 1992, Chap. B7, 190 pp. Wilson, J.L. and Miller, P.J., Two-dimensional plume in uniform ground-water flow, ASCE J. Hydraul. Div., 104, 503–514, 1978. Wilson, J.T., Attenuating biodegradation and attenuation rate constants, Seminar Series on Monitored Natural Attenuation for Groundwater, EPA/625/K-98/001, Office of Research and Development, Washington, D.C., 1998, pp. 5-3–5-5. Zwolinski, M.D., Harris, R.F., and Hickey W.J., Microbial consortia involved in the anaerobic degradation of hydrocarbons, Biodegradation, 11, 141–158, 2000.

CHAPTER 8 Present Application and Future Directions For Natural Attenuation 8.1 INTRODUCTION While there are many schemes and projects that utilize the natural attenuation capability of soils to one degree or another, the major kinds of projects presently exploiting the natural attenuation capability of soils, with respect to management of leachates and pollutant transport, can be broadly divided into two classes according to how this attenuation property is used: 1. Mitigation of contamination: Three good examples of this are (1) monitored natural attenuation (MNA) as a remediation option, (2) treatment zones as part of engineered treatment systems and (3) permeable reactive walls to capture and treat leachate or contaminant plumes. 2. Control of contaminant transport: Examples of this are found in the engineered liner and barrier systems using specified thicknesses of clay soil, singly or in combination with other materials as containment liner material.

In this chapter, we will examine some of these examples and discuss the present sets of protocols in the use of natural attenuation as a property for remediation of contaminated sites. This should not be confused with the protocols discussed in Chapter 7 that have been set forward by the various regulatory agencies. What we need to establish in this chapter are the protocols and requirements that take into account the interactions between soil fractions and contaminants in a soil-water system. As we have pointedly remarked throughout this book, natural attenuation of contaminants involves interactions of the contaminants and pollutants with both the fluid phase and the solid phase (soil solids). Accordingly, determination of lines of evidence, for example, requires analyses relating to sorption behavior of the contaminants in relation to soil type and composition. As in the previous chapter, the term natural attenuation is italicized when it is used in the context of a material property. This is to distinguish it from natural attenuation of contaminants as a process derived from the various mechanisms of contaminant assimilation and 271

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associated biological activities. In addition, we will continue to use NA (as in Chapter 7) to indicate the use of natural attenuation as a remediation treatment option. In the previous chapters, we have seen that the many variables in the interactions in the soil-water system between the various pollutants, aqueous phase, and soil fraction, can produce anywhere from minimal to significant pollutant attenuation in transport through the subsurface soil. Since the evidence of success or failure of the soil-water system to attenuate pollutants is not readily perceptible, monitoring, sampling and testing of the elements in the region of suspected pollutant plume presence and advance are necessary. We need to determine the necessary sets of protocols that would provide us with the assurance that the processes that occur in the ground are indeed producing the desired sets of results. This is perhaps one of the more significant requirements for future application of natural attenuation as a process and as a tool in pollutant plume management. In other words, the necessary sets of proof that will testify to the attenuation performance of the subsurface soil need to be properly developed. While these sets of proofs would be site and problem specific, there are some basic elements of physical, chemical and biological evidence that form the underpinning of the proofs. We need to be fully aware of these if we wish to exploit the natural attenuation property of soils.

8.2 MITIGATION OF CONTAMINATION It was pointed out in Chapter 1 that the costs for remediation of hazardous waste sites can reach proportions in the order of tens of billions of dollars in the U.S. If we project that to include other parts of the world, it is not difficult to imagine the degree of magnification of this cost. The attraction of exploiting the natural attenuation processes of soils as a remediation option for contaminated sites is self-evident. As we pointed out in the previous chapter, up until very recently, most, if not all of the reported uses of natural attenuation treatment procedures have been confined to remediation of sites contaminated by organic chemicals. Application of the procedure to sites contaminated solely by inorganic contaminants such as heavy metals has been very limited. This is unfortunate, since the various soil sorption mechanisms can play significant roles in the partitioning of inorganic contaminants. Ther e is general agreement among most researchers in the field of contaminant transport that natural attenuation of contaminants in soil is a process that “involves the biodegradation, dispersion, dilution, sorption, volatilization of contaminants, together with chemical and biochemical reactions and transformations of the contaminants to reduce contaminant toxicity, volume, mass, and concentrations to levels considered as non-threatening to biotic receptors and the environment” (Suthersan, 2002; Brady, 1998). By all accounts, although some application of natural attenuation processes has been directed toward remediation of inorganic pollutants in contaminated sites, its major use as a treatment option seems to have been concentrated on remediation of organic chemical pollutants to exploit the natural presence of microorganisms in soils. The oxidation-reduction reactions occurring in interactions between organic chemical pollutants and soil fractions under abiotic and biotic conditions are

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considered to be the major force in remediation of ground contaminated by organic chemicals. Redox reactions occur abiotically and under biotic conditions, and in general, there will be some microbial activity in both the classes of reactions. In the absence of significant microbial activity, and, in particular, for organic chemicals that are resistant to biodegradation, hydrolysis is a significant attenuating mechanism. Hydrolysis half-lives for such chemicals as halogenated aliphatics and chlorinated organics can range from months to a few years in the pH range commonly found in groundwater. 8.2.1

Monitored Natural Attenuation Requirements

MNA as a remediation process has been called passive remediation and intrinsic remediation. The implicit requirement for application of the natural processes in the soil to remediate contamination by natural attenuation is a strict monitoring program that would supply essential information according to the criteria set forward in the lines of evidence. Some of the major categories of items that need attention in the lines of evidence are shown in the schematic diagram in Figure 8.1. The lines of evidence provide the framework for gathering and evaluating the required markers in the various categories shown in Figure 8.1. The markers are perhaps one of the most critical elements of the MNA remediation program. They form not only the backbone of the protocols for lines of evidence, but also for the evidence of success (Section 8.3.1) and for positive evidence of engineered natural attenuation (EngNA) capability (Section 8.5.1). These markers are in essence sign posts that are needed to guide one through the sets of protocols. They consist of the following: • Proper site characterization to establish: – That the proper hydrogeology exists to promote transport of the pollutants in the soil-water system that possesses the qualities that will promote attenuation of the contaminants (pollutants) – The nature of the groundwater geochemistry and, in particular, the redox condition – That the essential site elements that can promote and sustain natural attenuation processes are present, e.g., composition and assimilative capacity of the soils, cation exchange capacities (CECs) and specific surface areas (SSAs) of the soils, soil organic matter, pH and Eh, microorganisms, electron acceptors and donors. • Determination of evidence of prior occurrence of natural attenuation at the site. The question that needs to be answered is, “Can we find evidence of the effects or results of natural attenuation of contaminants previously present in the site?” The pieces of evidence needed include determination of: – – – – – – –

Occurrence of natural bioremediation Occurrence of partitioning of pollutants Occurrence of transition products from abiotic reactions Biotransformation and biodegradation products Reduced and oxidized compounds, e.g., sulfate and iron reduction Speciation Groundwater geochemistry and hydrogeology

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Figure 8.1

Schematic showing pollution plume and the main categories of information required under the lines of evidence for proper consideration of monitored natural attenuation as a remediation process.

• Determination that the source of pollution can be controlled, limited or even removed. Ideally, one would want to remove or contain the source of the pollutants so that the natural attenuation process will be sufficient to achieve effective complete remediation of the pollutant plume; i.e., the pollutant plume will be completely attenuated. • Supporting laboratory research including microcosm studies, sorption characterization and tests for assimilative capacity of soils. – Microcosm studies, which provide the control environment for evaluation of the degradation and/or transformation of the target organic chemical pollutants under specified conditions, should be structured to provide information on the various factors and conditions that inhibit or promote mineralization and biotransformation of the target organic chemicals. – Partitioning tests for both inorganic and organic chemical pollutants should be conducted. These should include tests for determination of koc, kd, kow, adsorption isotherms, leaching column tests and desorption tests. • Modeling of transport and fate of pollutants. The models should provide the capability to predict transport and fate of the pollutants in the contaminated site. Uncertainties in prediction of the fate of the various pollutants can often be traced to (1) improper and incomplete accounting or understanding of the complex processes contributing to attenuation of the contaminants and (2) poor quality information on the various parameters, site conditions and boundary conditions. For proper prediction, the following items need to be considered:

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 275

– Accurate site-specific data – Proper quantitative descriptions of the various processes controlling mass transport and chemical mass transfer such as redox reactions, hydrolysis, precipitation, complexation and biological processes – Changes in reaction rates and constants because of transformation, speciation and complexation • Examination and evaluation of the potential advantages and disadvantages of application of MNA. The balance between length of time needed to effect acceptable attenuation of pollutants and costs needs to be examined in conjunction with long-term monitoring requirements and health and safety assessments. The list of some of the major items includes: – There should be less site disturbance and fewer chances of exposure of abovesurface biotic receptors to contaminants. – Generally speaking, the overall costs for application of MNA as a remediation process should be much less than other conventional methods. – A longer time period is needed to achieve remediation goals and standards. – Requirements for determining lines of evidence can lead to considerable expenditure of time, effort and costs. – Transformation and intermediate products resulting from natural attenuation processes may be more toxic than the original pollutants. The toxicity, persistence and mobility of the intermediary metabolites that result from incomplete biodegradation of the parent organic chemical compound are concerns that need to be fully addressed. – Requirements for long-term monitoring and sampling, together with associated tests to ensure attenuation occurs at the rate predicted by the transport models, can be a distinct disadvantage. – Changes in the hydrogeology of the site owing to circumstances occurring outside the boundaries of the contaminated site over the treatment life could have a negative impact on the effectiveness of the MNA process.

The sets of parameters that constitute the basic elements of the markers are shown assembled into three groups in Figure 8.2. The groups are distinct in that each group encompasses a set of tests and investigations that are needed to satisfy the technical protocols in the framework defined by the lines of evidence. For proper consideration of MNA as a remediation option, the markers and information requirements specified must be addressed. 8.2.2

Natural Attenuation and Lines of Evidence

To implement the protocols associated with lines of evidence, it is necessary to fully understand the various mechanisms and processes that combine to produce assimilation of pollutants in soils — in addition to the abiotic and biotic processes and reactions occurring in the soil. The previous chapters have discussed many of these as specific processes involving interactions between contaminants and soil fractions. These previous discussions will now be used to support and expand upon the material presented in the previous chapter regarding the lines of evidence used in the reported field studies or mandated by regulatory bodies. We take into account the role of the soil fractions in the attenuation processes and consider that a proper

276

Figure 8.2

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Parameters and markers for technical protocols in consideration of monitored natural attenuation as a remediation tool.

accounting of these is needed in the program of scrutiny required to establish the success of MNA. 8.2.2.1 Organic Chemical Pollutants To understand attenuation of organic chemical pollutants by soils, we need to go back to the previous chapters and recall some of the mechanisms of interactions between the organic chemicals and clay soils. As we have seen in the previous chapters, the chemical properties of the functional groups of the soil fractions contribute appreciably to the acidity of the soil particles. This is a significant property of the soil, since surface acidity is very important in the adsorption of ionizable organic molecules of clays. Surface acidity plays a major role in clay adsorption of amines, s-triazines, amides and substituted urea owing to protonation on the carbonyl group. A good example of this is the hydroxyl groups in organic chemical compounds, which consist of two broad classes of compounds [alcohols (ethyl, methyl, isopropyl, etc.) and phenols (monohydric and polyhydric)], and the two types of compound functional groups, i.e., those having a C-O bond (carboxyl, carbonyl, methoxyl, etc.) and the nitrogen-bonding group (amine and nitrile). Amine, alcohol and other organic chemicals possessing dominant carbonyl groups that are positively charged by protonation can be readily sorbed by clays. In amines, for example, the NH2 functional group of amines can protonate in soil, thereby replacing inorganic cations from the clay complex by ion exchange. As we saw in Chapter 5, the extent of sorption of these kinds of organic molecules depends on:

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 277

• The CEC of the clay minerals • The composition of the clay soil (soil organics and amorphous materials present in the soil) • The amount of reactive surfaces • The molecular weight of the organic cations

Large organic cations are adsorbed more strongly than inorganic cations because they are longer and have higher molecular weights, and polymeric hydroxyl cations are adsorbed in preference to monomeric species. This is because of the lower hydration energies and higher positive charges and stronger interactive electrostatic forces. Because of the unsymmetrically shared electrons in the double bond, carbonyl compounds possess dipole moments. This permits sorption onto clay minerals by hydrogen bonding between the OH group of the adsorbent and the carbonyl group of the ketone or through a water bridge. For the carbonyl group of organic acids such as benzoic and acetic acids, sorption onto clays occurs directly with the interlayer of cation or by formation of hydrogen bonds with the water molecules (water bridging) coordinated to the exchangeable cation of the clay complex. We will use an example of organic chemical pollutants in a contaminated site to demonstrate the importance of the protocols that attend the serious consideration of MNA as a remediation option. Organic chemicals that find their way into the land environment have origins in various chemical industrial processes and as commercial substances for use in various forms. These are generally classified as xenobiotic compounds. Chemical products such as organic solvents, paints, pesticides, oils, gasoline, creosotes, greases, etc. are responsible for many of the chemicals found in contaminated sites. Because there are thousands of organic chemical compounds in existence, it is not possible to categorize them all in respect to how they would interact in a soil-water system. The more common organic chemicals found in contaminated sites can be grouped into three convenient groups as follows: 1. Hydrocarbons including the PHCs (petroleum hydrocarbons), the various alkanes and alkenes, and aromatic hydrocarbons such as benzene, MAHs (multicyclic aromatic hydrocarbons) such as naphthalene, and PAHs (polycyclic aromatic hydrocarbons) such as benzo-pyrene 2. Organohalide compounds(the chlorinated hydrocarbons are perhaps the best known) including TCE (trichloroethylene), carbon tetrachloride, vinyl chloride, hexachlorobutadiene, PCBs (polychlorinated biphenyls) and PBBs (polybrominated biphenyls) 3. Miscellaneous compounds including oxygen-containing organic compounds such as phenol and methanol and nitrogen-containing organic compounds such as TNT (trinitrotoluene).

The density of these compounds in comparison to that of water is of particular interest since this will control the transport characteristics of the organic chemical. We classify nonaqueous phase liquids (NAPLs) into the light NAPLs (LNAPLs) and the dense ones (DNAPLS), as illustrated in Figure 5.5 in Chapter 5. Since the LNAPLs are lighter than water and the DNAPLs are heavier and denser than water, one would

278

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expect the NAPL to stay above the water table and the DNAPL to sink through the water table and come to rest at an impermeable bottom (bedrock). LNAPLs include gasoline, heating oil, kerosene and aviation gas. DNAPLs include the organohalideand oxygen-containing organic compounds such as 1,1,1-trichloroethane, chlorinated solvents such as tetrachloroethylene (PCE), trichloroethylene (TCE) and carbon tetrachloride (CT), PCBs, PCPs (pentachlorophenols) and TCPs (tetrachlorophenols). The results of transformations and biodegradation of organic chemicals were discussed in various forms in Chapter 6. The significant outcome of the application of MNA as a remediation process is the evidence of biodegradation and transformation of the target organic chemicals in the MNA process, as indicated in the requirements in lines of evidence. In the case of abiotic transformation processes, as pointed out in Chapter 5, these occur without the mediation of microorganisms. These processes include chemical reactions such as hydrolysis and oxidation-reduction. The markers used to satisfy lines of evidence requirements involve determination of both the decrease in concentration of the pollutants and the transformation of the organic chemical pollutants. This requires determination of the nature and composition of the transformed products of the original organic chemical pollutants. It is important to distinguish between the products obtained via abiotic and biotic processes. As noted in Chapter 6, biotic transformation processes are biologically mediated transformation reactions — with associated chemical reactions arising from microbial activities. The major difference between the transformation products from abiotic and biotic processes is that abiotic transformation products are generally other kinds of organic chemical compounds, whereas transformation products resulting from biotic processes are mostly seen as stages (intermediate products) toward mineralization of organic chemical compounds. To interpret the information, it is necessary to understand the evidence of occurrence of natural bioremediation. We recall from the previous chapters that biologically mediated transformation processes are the only types of processes that can lead to mineralization of the subject organic chemical compound. Complete conversion to CO2 and H2O (i.e., mineralization) does not always occur. However, intermediate products can be formed during the mineralization. Note that intermediate products obtained from abiotic and biotic transformation processes can themselves become greater environmental threats, as shown for example in Chapter 6 in the transformation of (C2Cl4) to TCE (C2HCl3), to 1,2-dichloroethylene (DCE; C2H2Cl2) to vinyl chloride (C2H3Cl) and to ethane (CH3CH3). Beginning with PCE, where the log koc value indicates good partitioning to the soil fractions, degradation of PCE to TCE and onward to vinyl chloride shows that the log koc values diminish considerably to a very low value for the vinyl chloride, as shown in Figure 6.9 in Chapter 6. As the PCE continues to degrade, more of the chemical substance is released into the aqueous phase (pore water). This is particularly true for vinyl chloride, for which the low values of log koc and high water solubility values suggest that this chemical can be environmentally mobile. 8.2.2.2 Heavy Metals For sites contaminated by heavy metals, we note from Chapter 4 that environmental mobility of these metals is dependent upon whether they are in the pore

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 279

water as free and complexed ions or sorbed onto the soil particles. Mobility of free ions and complexed ions in the pore water is governed by advection and diffusion mechanisms. So long as the full assimilative potential of the soil for heavy metal is not reached, attenuation of the heavy metals will continue. The mechanisms for heavy metal assimilation discussed in Chapter 4 show, for example, that the metals that are sorbed onto the soil particles are held by different sets of forces — determined to a large extent by the soil fractions and the pH of the soil-water system. The selective sequential extraction (SSE) procedure and test results shown in Figure 4.23 (Chapter 4) for Pb sorbed by an illite soil, and in Figure 4.24 and Figure 4.25 demonstrate that the various types of soils and their different soil fractions have different sorption capacities, dependent on the nature and distribution of the heavy metals and the pH of the system. Although precipitation of heavy metals is not, strictly speaking, a sorption phenomenon, precipitation of heavy metals as hydroxides, sulphides and carbonates generally are classified as part of the assimilative mechanism of soils because the precipitates form distinct solid material species. They are classified under the category of “removal of solutes from the pore water” and are thus most often considered as part of the attenuation process. Either as attached to soil particles or as void pluggers, precipitates of heavy metals can contribute significantly to attenuation of heavy metals in contaminant plumes. Hydroxide precipitation is favored in alkaline conditions as, for example, when Ca(OH)2 is in the groundwater in abundance. With available sulfur and in reducing conditions, sulphide precipitates can be obtained. Sulphide precipitates can also be obtained as a result of microbial activity, but this will not be a direct route. As shown in Chapter 6, sulfate reduction by anaerobic bacteria produces H2S and HCO3_, thus producing the conditions for formation of metal sulphides.

Figure 8.3

General protocol for considering monitored natural attenuation as a remediation tool.

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Figure 8.3 shows the general protocol for considering MNA as a remediation tool for treating contaminated ground, using Figure 8.1 and Figure 8.2 as reference guides. The data and information inputs shown on the left-hand side of the diagram, which are described in detail in Section 8.2.1, are taken directly from Figure 8.2. These tell us what is required to satisfy site-specific conditions and whether the markers that point toward evidence of natural bioremediation are sufficient to proceed with further examination to satisfy that the MNA is a viable treatment option. A “No” response from the first two decision steps will automatically trigger a technological and/or engineered solution to the remediation problem. The laboratory research input in the third step refers to the microcosm studies and the various laboratory procedures for determination of partitioning. These, together with the transport and fate predictions from the models developed for the specific situation, constitute the supporting studies shown in the third decision box. Analysis of the results from the supporting studies (laboratory research and predictions) should provide information about the ability of the site materials and conditions to attenuate the pollutants. A “No” response from the third decision step will allow incorporation of natural attenuation processes as part of a technological remediation solution. This will be discussed in Section 8.4 8.2.3

Desorption of Pollutants and Augmentation of Pollutants

Augmentation of pollutants, as opposed to attenuation of pollutants, refers to the increase in concentration of pollutants in the pollutant plume during transport. This is the result of the addition of pollutants desorbed from the soil solids. This can occur when the bonds holding the pollutants to the soil solids are disrupted or ruptured, causing the pollutants to be released into the pore water. The released pollutants will add to the concentration of the incoming leachate, thereby resulting in an increase in the concentration of the pollutants (augmentation). Dislocation and removal of inorganic pollutants such as heavy metals and organic chemicals sorbed or bound to soil solids arise because of the dynamics of the various biogeochemical actions or reactions occurring in the soil-water system. It is important to remember that the various soil mineral particles can undergo changes in the surface characteristics when the chemistry of the immediate environment (e.g., pore water) is changed. If the chemistry of the soil-water system changes such that low pH values result, solubility of exposed structural Al(III) on the clay particle surfaces will occur. The amount of Al(III) dissolved from the surface will be a function of time and pH level. Greater amounts of Al(III) will be dissolved with time and lower pH values (Wieland & Stumm, 1992). If the pH is raised to a high level, solubility of the Si(IV) could occur. The released Al(III) and Si(IV) could compete for the metal ions or could add to the pollutants in the contaminant plume. The report by Wieland and Stumm (1992) suggest that the preferential release of Si(IV) obtained in their study could have been due to the simultaneous resorption of Al(III) during the dissolution process. For disruption of bonds between the pollutants (inorganic and organic chemicals) and clay soil solids, it is necessary to either weaken or overcome the energies of interaction between the pollutants and the soil fractions. Another mechanism for

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 281

bond disruption or breakage is to alter or change the initial bonding condition. For example, de-ionization of ionized organic chemicals or changes to the surface acidity of clay mineral particles could trigger release of sorbed organics. Introduction of ligands could set up competition for sorption of inorganic contaminants. Biotransformation of sorbed organic chemicals could also alter the bonding relationships previously established between the organic chemical and the respective soil fraction, resulting in the release of the organic chemical from the soil solids. It is useful to remember that the soil-water or soil-aqueous system is a dynamic system subject to biogeochemical processes at all times. The results of such processes cannot be easily predicted. We need to be fully aware of the dynamics of the system and to anticipate that both augmentation and attenuation mechanisms exist since these are in essence two sides of the same coin. 8.2.4

Pollutant Release Studies for Augmentation Assessment

Studies designed to establish the potential for pollutant release from the soil fractions are necessary to determine if pollutant augmentation could (1) be responsible for the nonconforming or unexpected results from measurements of pollutant concentrations and performance characteristics, (2) pose a serious health and environmental threat downstream and (3) be a significant problem in itself. Since the biogeochemical environment in a soil-water system is never constant, we should investigate the possibilities for bond disruption and pollutant release because of the changing environment. 8.2.4.1 Batch Equilibrium Studies The simplest procedure to determine pollutant release in soils contaminated with inorganic contaminants is to conduct batch equilibrium tests (see Chapter 4). At least two types of soil suspension studies can be conducted: (1) determination of desorption characteristics of the candidate polluted soil and (2) soil washing–type studies designed to weaken the bonds. In the case of soils contaminated with heavy metals, desorption-extraction tests using extractants at various pH levels can be informative about the extractability of the metals in the metal-polluted soil. A typical set of desorption test results is shown in Figure 8.4 for a Pb-polluted illite soil. The results show that the number of washings needed to achieve metal release from the soil is dependent on the initial state of the soil. Neutral salts are generally used as the reagent solution of interest in the evaluation of desorption of heavy metals. The use of reagents in the desorption characterization studies are principally designed to seek extraction of the pollutants attached to the soil fractions. SSE studies such as those described in Chapter 4 are very useful for evaluation of the ease of removal of heavy metal pollutants in contaminated (polluted) soil samples. Figure 8.5 shows the sets of typical reagents used as extractants to remove heavy metal pollutants. Mulligan et al. (2001) have used this procedure to evaluate the effectiveness of various kinds of biosurfactants in the removal of sorbed heavy metals in a soil contaminated with both organic chemicals and heavy metals.

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

100 Numbers refer to Pb concentration in mg/kg of soil

Percent removed

80

50,000 mg/kg

60

5000 mg/kg

40

20 500 mg/kg 0 0

1

2

3

4

5

6

7

8

9

10

11

12

Number of washings with pH 2 solution

Figure 8.4

Desorption batch equilibrium test results for Pb-contaminated illite soil. Numbers refer to initial Pb concentration in the soil, and HNO3 is used as the washing solution. (Adapted from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000.)

Bonding between Pb and the different soil fractions (minerals, natural organic matter, oxides/hydroxides, carbonates), as shown for example in Figure 8.5, for the illitic soil is determined by mechanisms that range from ion exchange to precipitation and/or coprecipitation. As shown in the figure, the dominant mechanisms responsible for accumulation of heavy metal pollutants are sensitive to the pH of the immediate environment. This is because of the solubility of the hydroxide species of the heavy metals. When the pH in the pore water increases to a certain level (generally seen to be near the precipitation pH of the metal contaminant), Pb begins to form hydroxy species, resulting in the onset of Pb retention by the hydroxide fractions. The Pb precipitated or coprecipitated as natural carbonates can be released if the immediate environment is acidified. Thus, if the pH of the local environment changes and becomes acidified, augmentation of heavy metal concentration in the pore water can result from the release of the previously sorbed heavy metals. Bonding mechanisms established between Pb and amorphous or poorly crystallized Fe, Al and Mn oxides include exchangeable forms via surface complexation with functional groups (e.g., hydroxyls, carbonyls, carboxyls, amines, etc.) and interface solutes (electrolytes) and through precipitation and coprecipitation of the heavy metals. It is important to realize that metals attached to amorphous or poorly crystallized Fe, Al and Mn oxides can be removed by redox gradients. pH sensitivity in heavy metals bonding to soil solids is a very serious issue, as can be seen in the results in Figure 8.5. When low pH values exist in the soil-water system, neutral salts such as MgCl2, CaCl2 and NaNO3 become effective extractants. These can

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 283

Figure 8.5

Some representative reagents that will extract heavy metals sorbed by the various soil fractions shown in the diagram. Results shown in the diagram are for Pb sorbed by an illite soil. (From Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000.)

promote release of ions physically bound by electrostatic attraction to the negatively charged sites on the soil particle surfaces. Because of the affinity of groups II and III cations (valence of 2 or 3) for most surface sites, the cations in these extractant solutions have to be present in larger concentrations than the heavy metals bound by electrostatic attraction. 8.2.4.2 Leaching Column Tests Leaching columns are used not only to determine sorption of leachates in soil samples, but also to determine leachability or environmental mobility of sorbed pollutants in contaminated soils. To determine leachability, the contaminated samples are placed in the leaching cells or columns, and the fluids used as extractants are those that are expected to be present in the site. Common practice is to leach initially with water as a reference base. Most often, little evidence of leaching of sorbed pollutants will be seen. This is to be expected since the SSE experiments confirm that even with pollutants sorbed by exchange mechanisms, leaching with water does not result in significant exchange of the sorbed cations. Since the pore water in the soil contains dissolved solutes, leaching with representative solutions is required to determine leachability and environmental mobility of the sorbed pollutants. The results of two leaching tests are shown in Figure 8.6 and Figure 8.7. In Figure 8.6, the results show leaching of arsenic from an arsenic-contaminated soil in relation to the number of pore volumes of water used as the leaching fluid. Soil organic

284

Figure 8.6

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Leaching of arsenic from arsenic-contaminated soil in relation to number of pore volumes of water used as leaching solution. Values in the ordinate refer to concentration of As in the efﬂuent. (Data from Tan, B.K., Update on Arsenic Leaching Tests, unpublished report submitted to Geoenvironmental Research Centre, Cardiff University, Wales, 2003.)

matter content appears to influence the release concentrations. However, the nature of the influence has yet to be fully determined. In Figure 8.7, leaching of a DNAPLpolluted soil is shown. The results indicate that although leaching with water is not a major issue, some leaching of the DNAPL into the pore water occurs — in addition to a small fraction identified as desorbed. Mass balance calculations were made to determine the losses owing to volatilization and other degradative effects. Determination of the chemical compounds in the pore water together with the DNAPL product remaining sorbed onto the soil fractions will provide information of attenuation and augmentation of the DNAPL.

8.3 MONITORING AND EVIDENCE OF SUCCESS The term monitoring is used in many different ways. In the context of monitoring of a particular site to determine whether the events expected to occur in the site have indeed transpired, it is necessary to gather all pertinent pieces of information providing evidence that those events have occurred. We interpret the definition of monitoring in the previous chapter to mean a program of sampling, testing and evaluation of the status of the situation being monitored. In the case of the MNA program, the situation being monitored would be the management zone defined in Figure 7.4. To determine whether attenuation of pollutants in a contaminated site has been effective, it is necessary to obtain information pertaining to the nature,

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 285

Figure 8.7

Distribution of dense nonaqueous phase liquid (DNAPL) in a DNAPL-polluted soil column as a result of leaching with water after 1, 5 and 10 pv. (From Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000.)

concentrations, toxicity, characteristics and properties of the pollutants in the attenuation zone. The pollutants reside both in the pore water (or groundwater) and on the surfaces of the soil solids. Residence associated with the soil solids can take the form of sorbates and coprecipitates. In turn, the sorbates can be complexed with the soil solids and remain totally fixed within the structure of the soil solids. However, the sorbates can also be held by ionic forces that can be easily disrupted, thus releasing the sorbates. What the preceding discussion of the residence status of the pollutants tells us is that we need to monitor and sample not only the pore water or groundwater, but also the soil fractions in the contaminant attenuation zone. Two types of monitoringsampling systems are needed. For pore water or groundwater, monitoring wells are generally used. These wells are necessary to provide access to groundwater at various spots (vertically and spatially) in a chosen location. The choice of type of monitoring wells and distribution and location of wells depends on the purpose of the wells. In respect to determination of whether natural attenuation can be used as a treatment process, there are at least three separate and distinct monitoring schemes that need to be considered. These range from the initial site characterization studies to verification monitoring and long-term conformance monitoring. The term monitoring scheme is used deliberately to indicate the use of monitoring and sampling devices to obtain both soil and water samples. Figure 8.8 shows some typical devices used as monitoring wells to monitor groundwater at various levels. In the left-hand group are individual monitoring wells with sampling ports located at different depths but

286

Figure 8.8

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Some typical groundwater monitoring and sampling wells.

grouped together in a shared borehole. This is generally identified as a single borehole multilevel monitoring well system. The middle drawing shows a nest of single monitoring wells in their own separate boreholes, and the right-hand drawing shows a single tube system with monitoring portholes located at the desired depths. With present technological capabilities, monitoring wells and the manner of operation have reached levels of sophistication where downhole sample analysis of groundwater can be performed without the need for recovery of water samples. Some of these were described in Chapter 7. Site characterization monitoring is necessary to provide information on the hydrogeology of the site. It is necessary to properly characterize subsurface flow to fully delineate or anticipate the transport direction and extent of the pollutant plume. Determination of the direction and magnitude of groundwater flow is important. Obviously this means a judicious distribution of monitoring wells upgradient and downgradient. A proper siting of the monitoring wells and analysis of the results should provide knowledge of the source of the pollutants and the characteristics of the pollutant plume. Verification monitoring requires placement of monitoring wells and soil sampling devices within the heart of the pollutant plume and at positions beyond the plume. Figures 8.9 and 8.10, respectively, show the vertical and plan views of how the wells and sampling stations might be distributed. The more monitoring and sampling devices there are, the better one is able to properly characterize the nature of the pollutant plume — assuming that the monitoring wells and sampling devices are properly located. The monitoring wells and sampling devices placed outside the pollutant plume, shown in Figure 8.10, also serve as monitoring wells and sampling devices for long-term conformance assessment.

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 287

Figure 8.9

Simple monitoring and sampling scheme for monitored natural attenuation effectiveness. Note that the nest of multilevel sampling tubes is meant to provide access to soil sampling and specimen recovery for analysis of partitioning of pollutants. The nest of monitoring and sampling wells provides access to aqueous samples for determination of aqueous chemistry.

The tests required of samples retrieved from monitoring wells are designed to determine the nature of the pollutants in the pore or groundwater. These provide information about the concentration, composition and toxicity of the target pollutant. For prediction of further attenuation of the target pollutant, the partition coefficients and solubilities of the various contaminants are needed as input to transport and fate models. If biotransformation of the target pollutants occurs, supporting laboratory research will be needed to determine the likely fate of the transformed or intermediate products. Tests on recovered soil samples at various locations and depths in the contaminant attenuation zone are required as part of the monitoring program. These tests are designed to determine the environmental mobility of the pollutants and the nature and concentration of pollutants sorbed onto the soil particles (soil solids). These pollutants or sorbates can be lightly bonded or strongly bonded to the soil solids, and it is necessary to determine if changes in the pore water chemistry, redox changes, pH changes, etc. would dislodge the sorbates. Detailed discussions of many of the bonding mechanisms, their sensitivity to changes in the immediate environment and the methods for determining partitioning of the pollutants have been developed in Chapters 3, 4 and 5. Determination of environmental mobility of pollutants sorbed onto soil solids is generally conducted using leaching columns. In

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Figure 8.10

Plan view of distribution of monitoring wells and soil sampling boreholes for veriﬁcation monitoring and long-term conformance monitoring.

such tests, the contaminated soil samples recovered from the site are subject to leaching by deionized water as a calibration or standard case and by various fluids selected to mimic the chemistry of the groundwater in the site of interest. Such a calibration desorption test is shown in Figure 8.11 for a laboratory study on arsenic release from arsenic-contaminated soils. Changes in pH and Eh may also be factored into the leaching tests as environmental factors to determine their influence on the release of contaminants. 8.3.1

Evidence of Success

The long-term monitoring program required as part of the MNA implementation scheme provides the information necessary to satisfy the protocols associated with the evidence of success. The evidence of success is comparable in philosophy to the lines of evidence. Figure 8.12 shows the protocols for determination of evidence of success, beginning with requirements for proper siting of the monitoring wells and sampling devices, as shown in the top decision box. The lower decision boxes, which deal with the verification requirements, are the essence of the success testimony for MNA. Without positive verification, it is not likely that application of the MNA as a remediation process will satisfy regulatory requirements. The key elements that are essential to the verification monitoring scheme portion of the evidence of success come from the supporting studies and model predictions. The supporting studies and model predictions are the same as those described in Section 8.2.1 and shown

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 289

120

Arsenic (mg/L)

100

80

60

40

20

0 2

4

6

Pore volumes, pv Figure 8.11

Desorption of arsenic from an arsenic-contaminated soil with deionized water. Measurements of concentration of arsenic are in the efﬂuent. (Data from Tan, B.K., Update on Arsenic Leaching Tests, unpublished report submitted to Geoenvironmental Research Centre, Cardiff University, Wales, 2003.)

in Figure 8.3. The importance of verification of the markers for attenuation of pollutants using comparative leaching column studies and microcosm studies cannot be overstated. Exact or close corroboration is not expected between field and laboratory values because of size and time scaling problems. One looks for patterns and similarities in partitioning performance and degradation patterns and pathways. Further corroboration is sought between field performance and predictions from analytical-computer models of system performance. While it would be desirable to obtain matching between predicted and observed field performance, the more realistic view is to look for trends in field performance that closely match model predictions. If the trends are in the right direction, and if the degree of attenuation predicted by the fate and transport model accords with the observed field performance, MNA can work as a remediation option. The biggest challenge in evaluating the results from the verification monitoring program is to determine an acceptable level of risk that would satisfy the decision makers associated with the verification process. Risk evaluation, risk assessment and risk analysis are necessary components of the evaluation scheme, as seen in the fourth decision box in Figure 8.12. In essence, we can describe MNA as a deliberate and calculated risk management scheme, with levels of risk carefully managed by strict application of the lines of evidence and evidence of success protocols. It is useful to note that if all attenuation markers are not totally satisfied, it is possible

290

Figure 8.12

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Protocols for establishing evidence of success for monitored natural attenuation using monitoring information and supporting laboratory research and predictions from fate and transport models.

to use NA as part of a set of structured technological solutions, as shown in the right-hand side of the diagram. These aspects will be discussed in the next section together with other uses for natural attenuation. 8.4 ENHANCED NATURAL ATTENUATION Recognizing that bioremediation of soil contaminated by organic chemical pollutants exploits soil microorganisms to metabolize the organic chemical compounds, and further recognizing that many natural soil organics [soil organic matter (SOM)] are similar to synthetic organic chemical compounds, we can expect to have naturally occurring consortia of microorganisms present in a contaminated site containing soil organics. As an example, natural aromatic soil organics such as vanillin, lignin and tannin are similar to the synthetic aromatic organic compounds represented by benzene, toluene, PAHs, etc. The naturally occurring consortia of microorganisms, ranging from bacteria and fungi to viruses, described in Chapter 6 should successfully address the synthetic organic chemicals since they are well adapted to the specific habitat. The available energy sources and all the other microenvironmental factors such as pH, temperature, water content, etc. produce the suites of biomass that have adapted to the microenvironment. Previous discussions have shown that natural attenuation can be used directly as a remediation tool in the control and management of pollutant plumes. However,

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 291

Figure 8.13

Exploitation of natural attenuation potential of soils. The dashed arrow from ENA to EngNA indicates that one may or may not wish to take advantage of ENA properties in the construction of EngNA systems.

the natural attenuation capacity of a soil can also be used as part of a design or engineered scheme, as seen in Figure 8.13, and for management and control of pollutant plumes. This can take the form of augmentation or enhancement of the assimilative capacity of the soil or use of the soil as part of a barrier system. These aspects are discussed in this section. When the assimilative capacity of a soil is not capable of providing (1) complete pollutant attenuation, or (2) the desired level of attenuation of the pollutants in the pollutant plume or (3) effective contaminant attenuation within a specified time frame, then there are several ways to increase the assimilative capacity of the soil. While some may consider these procedures to be slightly intrusive, they are nevertheless useful since they are relatively inexpensive and simple to implement. Figure 8.13 shows some of the more popular methods for enhancing the natural attenuation capacity of the soil. The boxes in the top portion of the figure show two abioticand two biotic-related procedures. Strictly speaking, the integration of abiotic procedures into the treatment procedure cannot be called natural attenuation. While the term enhanced natural attenuation (ENA) has been used in the literature, and will be used herein, it must be noted that it concerns improvement of the assimilative capacity of the soil and/or the biological activities therein. 8.4.1

Biostimulation

Probably the simplest procedure for improving the intrinsic bioremediation capability of a soil is to provide a stimulus to the microorganisms that already exist in the site. This procedure is called biostimulation, i.e., adding nutrients and other

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growth substrates, together with electron donors and acceptors. The intent of biostimulation is to promote increased microbial activity with a set of stimuli to better degrade the organic chemical pollutants in the soil. With the addition of nitrates, Fe(III) oxides, Mn(IV) oxides, sulfates and CO2, for example, anaerobic degradation can proceed. This technique is used for sites contaminated with organic chemical pollutants and is perhaps one of the least intrusive of the methods of enhancement of natural attenuation. The other method of enhancement that falls in the same class of less-intrusive enhancement procedure is bioaugmentation. 8.4.2

Bioaugmentation

If the indigenous microbial population is not capable of degrading the organic chemicals in the soil (for whatever reason, e.g., concentrations, inappropriate consortia, etc.) other microorganisms can be introduced into the soil. These are called exogenous microorganisms. Their function is to augment the indigenous microbial population such that effective degradative capability can be obtained. If need be, biostimulation can also be added to the bioaugmentation to further increase the likelihood of effective degradative capability. It is important to be conscious of the risks that arise when unknown results are obtained from interactions between the genetically engineered microorganisms and the chemicals in the contaminated ground. The use of microorganisms grown in uncharacterized consortia which include bacteria, fungi and viruses can produce toxic metabolites (Strauss, 1991). In addition, the interaction of chemicals with microorganisms may result in mutations in the microorganisms themselves and microbial adaptations. 8.4.3

Geochemical and Biogeochemical Intervention

The techniques for introducing the nutrients, growth substrate and exogenous microorganisms to implement biostimulation and bioaugmentation can also be used to provide geochemical intervention. Generally, the simplest forms of geochemical and biogeochemical interventions are pH and pE or Eh manipulation. These types of intervention are also well suited for treatment of inorganic contaminants or pollutants. Changes in toxicity for some inorganics (e.g., Cr and As) can result because of changes in their oxidation state, resulting in either a decrease or increase in the toxicity level of the contaminant. Chromium (Cr) is an example. Cr(III) is an essential nutrient that helps the body use sugar, protein and fat, while Cr(VI) has been determined by the World Health Organization to be a human carcinogen. Cr(III) can be oxidized to Cr(VI) by dissolved oxygen and possibly with manganese dioxides. Since this is not a desirable situation, geochemical and biogeochemical means can be used to create a reducing environment in the subsurface. One useful method is to deplete the oxygen in the subsurface to create a reduced condition in the soil. In the case of As, As(III) is more toxic than As(V). What is needed, therefore, is to provide an oxygen source to ensure oxidation of the As(III). Manipulation of pH is quite common in agricultural practice. When applied to assist the assimilative capacity soils, pH manipulation, among other things, addresses the precipitation of heavy metals in solution (pore water) or dissolution of precipitated heavy metals. Changes in pH also result in changes in the sign of surface

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 293

electrostatic charges for those materials that have pH-dependency charge characterization. When this happens, bonding of metals or release of heavy metals from disruption of bond disruptions occurs. Both pH and Eh changes also have considerable influence on the various acid-base reactions and on abiotic and biotic electron transfer mechanisms. Transformations and degradation of organic chemical pollutants resulting from acid-base and oxidation-reduction reactions are less significant than those obtained via biotic processes. Reaction kinetics in relation to such processes, and those initiated by the catalytic action of soils resulting in abiotic transformation, are considered to be relatively slow. 8.4.4

Soil Improvement

The procedures for soil improvement generally include addition of soil amendments to increase the sorption or assimilative capacity of the soil. While soil improvement is perhaps the most intrusive of the enhancement procedures, it is not uncommon to find soil amendments used quite widely. Consider the problem of a site contaminated by heavy metals as an example. In the past, lime has been used as the additive of choice because of its ability to raise the pH of the soil-water system. This would precipitate the heavy metal pollutants and thus make them less bioavailable. However, it must be recognized that this is not a permanent solution because if the pH of the system is subsequently reduced by environmental forces or external events, the same heavy metals will become mobile again. To obtain a more permanent solution, not only a pH change, but also amendments that would bind the metals to the soil should be considered. This can take the form of addition of those soil fractions that have the greatest sorption capability (see Chapters 2 and 3) such as clay minerals, SOM, and amorphous materials. The aim is to increase the CEC and the SSA of the soil. These, in turn, increase the partition coefficient of the soil. Figure 8.14 shows the breakthrough curves for arsenic in relation to a soil in its original state and with additions of SOM. It would appear that changes in soil structure with the SOM have significant influence on the initial sorption of arsenic. In addition to the addition of beneficial soil fractions, inorganic and organic ligands can be introduced in the pore water to promote speciation and formation of various compounds. 8.4.5

Application of Enhanced Natural Attenuation

Figure 8.15 shows a direct application of ENA as an in situ remediation process. Enhanced treatment of a region (spatial and vertical) of the site downgradient from the contaminated site permits the ENA to function as planned. The treated region is called the in situ reactive region (IRR) or treatment zone and can be used in conjunction with other treatment procedures. Figure 8.15 illustrates the use of the IRR as a treatment procedure for the pollutant plume in the region in front of the permeable reactive barrier. Treatment procedures using treatment wells or boreholes and associated technology include: • Geochemical procedures such as pH and Eh manipulation • Soil improvement techniques such as introduction of inorganic and organic ligands and introduction of electron acceptors and donors • Various other biostimulation procedures and bioaugmentation

294

Figure 8.14

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Sorption of arsenic showing breakthrough performance and inﬂuence of soil organic matter content on performance. (Data from Tan, B.K., Update on Arsenic Leaching Tests, unpublished report submitted to Geoenvironmental Research Centre, Cardiff University, Wales, 2003.)

The choice of any of these, or a combination of these methods of augmentation depends on the type, distribution and concentration of pollutants in the contaminated site and on the results obtained from microcosm and treatability studies. 8.4.5.1 In Situ Reactive Regions Figure 8.15 shows an example of the IRR, i.e., the region immediately in front of the permeable reactive barrier. The purpose of an IRR is to provide not only pretreatment or preconditioning in support of another treatment procedure, but also as a posttreatment process for sites previously remediated by other technological procedures. Figure 8.15 shows the IRR used in support of the permeable reactive barrier (PRB) treatment procedure. Other treatment procedures can be used in place of the PRB. The presence of heavy metals in combination with organic chemicals in the pollutant plume is not an uncommon occurrence. One could, for example, envisage using IRR as a treatment procedure in combination with a subsequent procedure designed to fix or remove the metals. In application of IRR as a posttreatment process, IRR should be the final cap for some kind of design or technological process for remediation of a contaminated site. This is generally part of a multiple-treatment process — as opposed to the use of IRR in a pretreatment or preconditioning process. A good example of this is the use of pump-and-treat as the first phase of the remediation program, followed by

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 295

Figure 8.15

Enhancement of natural attenuation using treatment wells. Treatments for enhancement can be any or all of the following: geochemical intervention, biostimulation and bioaugmentation. Treatment occurs in the pollutant plume and downgradient from the plume.

the IRR as a posttreatment process in which the treated pollutant plume receives its final cleanup. The efficiency of cleanup using pump-and-treat methods rapidly decreases as greater pollutant extraction from the groundwater or pore water is required. It is not unusual to remove some large proportion of the pollutants from the groundwater or pore water and to leave the remaining proportion to be removed via natural attenuation processes in an IRR. 8.4.5.2 Permeable Reactive Barriers The intent of a PRB is to provide treatment as a remediation procedure to a pollutant plume as it is transported through the PRB so that the plume no longer poses a threat to biotic receptors when it exits the PRB. Figure 8.16 shows the crosssectional and plan views of a PRB application. The pollutant plume is seen to migrate into the PRB, where various assimilative mechanisms are brought to bear to attenuate the pollutants. The PRB needs to be strategically located downgradient to intercept the pollutant, and if needed, the pollutant plume can be channeled to flow through the PRB as shown in Figure 8.17. The funnel-gate technique shown in Figure 8.17 is one of the more common techniques used to channel the pollutant plume to flow through the reactive barrier. This funnel, which is constructed or placed in the contaminated ground, is composed basically of confining boundaries of impermeable

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Figure 8.16

Cross-section and plan view of permeable reactive barrier (PRB), which is sometimes called a treatment wall. Plan view shows leachate plume entering the PRB with pollutants and leaving the PRB without pollutants.

material (e.g., sheet pile walls) that narrow toward the funnel mouth where the reactive wall is located. Other variations of the funnel-gate technique exist — obviously in accord with site geometry and site specificities. PRBs are also known as treatment walls. The soil materials in these walls or barriers can include a range of oxidants and reductants, chelating agents, catalysts, microorganisms, zero-valent metals, zeolite, reactive clays, ferrous hydroxides, carbonates and sulfates, ferric oxides and oxyhydroxides, activated carbon and alumina, nutrients, phosphates, and soil organic materials. The choice of any of these treatment materials is made on the basis of site-specific knowledge of the interaction processes between the target pollutants and material in the PRB. Laboratory tests and treatability studies are essential elements of the design procedure for the treatment walls (PRBs). When designed properly, a PRB provides the capability for assimilation of the pollutants in the pollutant plume as it migrates through the barrier. In that sense, PRBs function in much the same manner as IRRs, except that the region is a constructed barrier. Some of the assimilative processes in the PRB include the following: • Inorganic pollutants: sorption, precipitation, substitution, transformation, complexation, oxidation and reduction • Organic pollutants: sorption, abiotic transformation, biotransformation, abiotic degradation and biodegradation

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 297

Figure 8.17

Funnel and gate arrangement of permeable reactive barrier (PRB) treatment of pollutant pl ume. Funnel effect is provided by the impermeable walls that channel pollutant plume transport to the PRB gate.

8.5 ENGINEERED NATURAL ATTENUATION There are some who would argue that engineered natural attenuation (EngNA) is perhaps a misnomer in that natural attenuation is no longer natural. To a certain extent this is true because the soil attenuation layer of an engineered barrier system is often composed of soil materials that are chosen for their attenuation capability — a designed soil-water system that is generally called an engineered clay barrier. Figure 8.18 shows a general view of an engineered barrier system used for containment of a waste pile. The details of the filter, membrane and leachate collection system and the nature and dimensions of the pollutant attenuation layer are specified by regulatory command and control requirements or by performance requirements. Figure 8.19 shows an example of these details and dimensions for the engineered barrier system used in containment of municipal solid wastes. The nature of material comprising the engineered clay barrier that underlies the synthetic membrane is determined on the basis of a maximum permissible hydraulic conductivity performance expressed in terms of the Darcy permeability coefficient k. There is an implied understanding (not always well founded) that the minimum specified hydraulic conductivity is somehow related to the attenuation capability of the engineered soil. The minimum dimensions specified in Figure 8.19 vary between different countries.

298

Figure 8.18

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Pollutant attenuation layer constructed as part of an engineered barrier system. The dimensions of the attenuation layer and the speciﬁcation of the various elements that constitute the ﬁlter, membrane and leachate collection system are generally determined by regulations or by performance criteria.

The basic idea in the design details of the engineered barriers shown in Figure 8.18 and Figure 8.19 is that if leachates inadvertently leak through the high-density polyethylene membrane and are not captured by the leachate collection system, the pollutants in the leachate plumes will be attenuated by the engineered clay barrier. The engineered clay barrier serves as the second line of defense or containment. Even though the specifications shown in Figure 8.19 refer only to a maximum permissible k value (generally in the order of 10–9 m/sec) for the engineered clay barrier material, it is prudent to conduct additional tests of the material. These tests, which determine the pollutant assimilation capability of the clay material, are part of the protocol in the evidence of EngNA capability, which assesses the capability of the engineered clay barrier material to attenuate the pollutants in the leachate plume. This will be discussed in the next section. The EngNA shown in Figure 8.13 as a direct link from NA is also used as the foundation base for the double-liner barrier system for landfill containment of hazardous waste, as shown for example in Figure 8.20. There are several options for the foundation base seen in the diagram. Since a fully compacted foundation base is a standard requirement — to provide support for the material contained above — there is the option of working with the native material if it has the proper assimilative potential or with imported fill material. Once again, the purpose of the foundation base is to provide attenuation of pollutants if leakage of pollutants through the double-liner system occurs. This in essence constitutes a third line of defense against pollutant transport into the subsurface soils.

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 299

Figure 8.19

8.5.1

General speciﬁcations for engineered barrier systems used as bottom-liner barriers to contain municipal solid wastes. (Adapted from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000.)

Evidence of EngNA Capability

At present, there are no absolute requirements for establishing positive evidence of the capability of the engineered barrier material to function in accord with design expectations of attenuation of pollutants — except for controls on maximum allowable hydraulic conductivity. Section 2.5.2 showed that to a certain extent, the transmission properties as expressed by the permeability coefficient k do take into account the surface areas of the soil particles in contact with the permeating fluid. By extension, one would anticipate that if the permeating fluid is a leachate containing pollutants, these would interact with the reactive forces associated with the surface areas of the soil particles. For assurance that the engineered barrier material, or the foundation base material and preparation technique, will provide for a capable barrier or foundation, it is necessary to obtain evidence of capability of the material to function according to specifications. At the present time, there are few mandatory specifications. If specifications are written, they are generally site and material specific and may not include tests that would provide evidence of the necessary assimilative capacity of the soil material used for the barrier. Prudent practice requires evidence of capability of the material to function as required and that the placement technique provide the proper barrier performance capability.

300

Figure 8.20

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Suggested minimum dimensions for double-membrane-liner system with an engineered clay layer as foundation base as a bottom-liner system for landﬁll containing hazardous wastes. The minimum thickness of the engineered clay layer in the foundation base should be at least 1 m. (Adapted from Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000.)

Evidence of capability of the engineered natural attenuation (evidence of EngNA capability) barrier to provide effective attenuation of pollutants requires attention to a set of tests and analyses as outlined in the diagram shown in Figure 8.21. The supporting laboratory studies are the same supporting laboratory studies shown in Figure 8.3 and listed in Section 8.2.1. As in the case of the evidence of success protocols, the procedures for determination of positive evidence of EngNA capability require almost the same sets of procedures used in the determination of evidence of success for MNA. One of the significant benefits of applying the protocols for evidence of EngNA capability is the determination of the required thickness of the pollutant attenuating layer shown in Figure 8.18 and the engineered clay barrier shown in Figure 8.19. The second decision box in Figure 8.21 requires specification of a thickness of the pollutant attenuating layer or engineered clay barrier needed to satisfy attenuation remediation requirements. This specification can be obtained from the calculations from fate and transport models using the results of supporting laboratory research on partitioning and other attenuating phenomena. Since the dimensions for the engineered clay barriers shown in Figure 8.19 specify greater than or equal to designations, application of the protocols for evidence of EngNA capability will likely provide dimensions that should satisfy regulatory requirements.

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 301

Figure 8.21

Protocol for determination of positive evidence of engineered natural attenua(EngNA) capability as required for use of engineered clay barriers or as foundation base material for double-liner systems.

tion

8.6 OBSERVATIONS ON FUTURE APPLICATIONS The existence of natural bioremediation assures us that, given an allowed time factor, exploitation of natural attenuation for remediation of organic chemical pollutants in the ground can be successful either by itself or, more likely, as an added element in the total treatment of a contaminated site. That is, the assimilative potential of soils can be exploited by use of this potential in concert with other technologies for remediation of contaminated sites. This exploitation can be enhanced with biostimulation and/or bioaugmentation plus geochemical and biogeochemical manipulation. For leachate streams containing heavy metals, natural attenuation works in the form of partitioning, precipitation and speciation of the metals in the soil. Natural bioremediation of soils contaminated with heavy metals occurs through processes such as bioaccumulation and biological oxidation and reduction. It is important to bear in mind that there are both advantages and disadvantages to the use of natural attenuation in pollutant plume management. Many of these have been discussed as part of the markers in Section 8.2.1. That being said, there is every expectation that exploitation of the various processes that contribute to natural attenuation will flourish and expand. A good example of this is the recent use of phytoremediation. Uptake of heavy metals can also be accomplished by some species of plants under processes classified as phytoremediation. Some of these aspects have already

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been mentioned in Chapter 6. There are some who might argue that phytoremediation should not be included in the group of processes contributing to the natural attenuation of pollutants in soils because they are not totally associated with actions of the soil-water system. However, since the processes classified as phytoremediation include phytoextraction (uptake) and phytodegradation, and since these directly interact with the soil-water system, phytoremediation can be accepted as a process contributing to the natural attenuation of pollutants in soils. Many plants have the ability to extract and concentrate certain kinds of elements in the soil. Their root system absorbs and accumulates the necessary nutrients (and water) to sustain their growth. While metal-tolerant plants have some tolerance for toxic metal ion uptake, generally, their tolerance level for such metals is very low. However, hyperaccumulating plants have higher levels of tolerance for toxic metal ions and can tolerate heavy metal ions up to several percent of their dry weight. Continued research into this area shows considerable promise in the use of specific plants for phytoextraction or phytodegradation. Mulligan (2001) explains phytodegradation as the metabolism of contaminants in the leaves, shoots and roots, resulting in the release of enzymes and other components for stimulation of bacterial activity or biochemical conversion. Examples cited from various sources by Mulligan (2001) show removal of TCE, PCE, DCE and xylene from groundwater by hybrid poplars. 8.6.1

Use of Clay Soils as Catalysts

The use of clay soils as a pollutant attenuating material has generally been considered in terms of physical interactions between sorbent and sorbate. Chapters 3 to 5 elaborated on the interactions between pollutants and soils in respect to reactive forces and mechanisms associated with the surfaces of the soil solids and their respective functional groups. The catalytic role of clay soils, and mineral soils in particular, has not been well exploited. In part, this is because soil catalysis is not a well-appreciated phenomenon and in part because soil improvement methods for pollutant assimilation have been confined to the previously stated sorbent-sorbate interacting relationship. In Section 3.4.1 we briefly discussed hydrolysis as an acid-base reaction. The importance of hydrolysis or hydrolytic reactions in respect to pollutants in a soilwater system cannot be overstated. Section 4.4.1 addressed hydrolysis reactions of metal ions and indicated that these reactions are influenced by the pH of the active system, temperature, redox and type and concentration of the metal ions. The hydrolysis reactions of metals (MX) can be expressed as MX + H2O Æ MOH + X– + H+

(8.1)

With respect to organic chemical pollutants, hydrolysis involves reactions between an organic molecule and water, resulting in the formation of a new covalent bond with OH. Designating the organic molecule as RX, cleavage of the covalent bond with X occurs. The net reaction is given by Mill and Mabey (1988) for displacement of X by OH– as

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 303

Figure 8.22

Surface acidity of kaolinite expressed as pka of Hammett indicators and also as percent of H2SO4 of equivalent pka. (Adapted from Solomon, D.H., and Murray, H.H., Clays Clay Miner., 20, 135–141, 1972.)

RX + H2O Æ ROH + X– + H+

(8.2)

Metal ion catalysis occurring at interfaces between two phases (pore water and mineral soil particle surface) is referred to as heterogeneous catalysis. Clay particle surfaces exhibit surface acidity that can be from two to four units lower than the acidity of the bulk pore water. Soil-catalyzed hydrolysis reactions associated with the surface acidity of clay minerals can be significant because they can affect the hydrolysis half-lives of the reacting organic chemicals; i.e., they affect the kinetics of hydrolysis. Figure 8.22 shows the surface acidity of kaolinite in relation to the moisture content of the soil. This surface acidity, which is derived from the surface hydroxyls on the octahedral layer of the mineral particles, is reduced significantly as the moisture content of the soil is increased. Surface acidity in the case of montmorillonites is due to isomorphous substitution and to interlamellar cations. The charge and nature of the cations affect the degree of catalytic activity since the cations impact directly on the polarizing power and the degree of dissociation of the water in the inner Helmholtz plane (adsorbed water). In addition, the surface acidity (of montmorillonite) increases as the valency of the exchangeable cations is increased; e.g., for Na- Mg- and Al-montmorillonites, surface acidity increases in the order given as Na, Mg and Al. For sites contaminated with heavy metals, metal-ion catalysis of hydrolysis occurs through the heavy metals sorbed by the soil fractions. Direct polarization results from coordination of hydrolyzable functional groups by the metals. According to Buckingham (1977), direct polarization mechanisms can accelerate hydrolysis rates by factors of 10,000 or more. It is also possible for metal ion–catalyzed

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hydrolysis to form a metal-coordinated nucleophile that is more reactive than a comparable free nucleophile (Plastourgou & Hoffman, 1984). M(H2O)mn+ ´ M(H2O)m–1OH(n +1)+ + H+

(8.3)

M(H2O)m–1OH(n + 1)+ + RCOX Æ (H2O)mM – O – CRn+ + X:

(8.4)

The increase in acidity of the water molecules, in situ generation of OH_, results from metal coordination. 8.6.2

Use of Soils to Promote Oxidation-Reduction Reactions

The promotion of oxidation-reduction reactions by soils is generally called soilcatalyzed oxidation and soil-catalyzed reduction. In essence, this is soil-mediated electron transfer. Soil-catalyzed oxidation depends on several factors, the most important of which is the presence of redox-active metals such as iron, manganese, copper, aluminum and trace metals. Free radical oxidation (homolytic oxidation) in soils is dependent on lower solubility limits and chemical structure of the organic chemicals (Dragun, 1988). The structural groups are classified according to the characteristics of electron withdrawal and donation, as for example, aromatic compounds with only electron-donating substituents and aromatic compounds with electron-withdrawing with (1) weak electron-donating fragments and (2) strong electrondonating groups. In the example of the oxidation of phenols by montmorillonite given in Chapter 5, the work reported by Yong et al. (1997) shows the greater oxidizing capability of the Fe (II)-clay in transforming the monomer 2,6-dimethylphenol of mass 122 to a 2,6-dimethylphenol dimer of mass 242 — in comparison to other clays. As noted in Chapter 5, the major groups of organic chemicals that undergo reductive transformation in reducing environments are the halogenated aliphatic and aromatic compounds. For abiotic processes, we look toward electron-mediated reductions to accomplish reductive transformations. SOM and clay minerals provide electron-mediated reductions through iron and sulphide components in the mineral structure. Kriegman-King and Reinhard (1991) reported on the transformation of hexachloroethane (HCA) to PCE by biotite and vermiculite according to the following mechanism: 2[Fe2+, nM+]mineral + HCA Æ 2[Fe3+, (n-1)M+] + PCE + 2Cl– + 2M+

(8.5)

8.7 CONCLUDING REMARKS It seems fairly clear that we will want to continue exploiting the natural assimilative and bioremediative potential of soils — with or without enhancements. It also seems clear that this property of natural attenuation will be a useful tool in a multiple-treatment process. The protocols set forward for lines of evidence, evidence

PRESENT APPLICATION AND FUTURE DIRECTIONS FOR NATURAL ATTENUATION 305

of success and positive evidence of EngNA capability in this chapter rely on the execution of several mandatory tests and analyses that we consider to be of utmost importance and significance. While we recognize and value the importance of monitoring and analyzing groundwater and pore water samples, we cannot over-emphasize the importance of analyses of soil samples from contaminated sites and from monitoring stations. We need to be fully aware of the nature of the pollutants sorbed onto the soil solids and to be conscious of the possibility that these can in themselves pose health threats to biotic receptors and that at some point in time they may become mobile. The material presented in this book has attempted to present the total picture of attenuation from the soils’ point of view, in the hope that the reader will begin to better appreciate the need for more attention to the role of soils in the determination of pollutant attenuation.

REFERENCES Brady, P.V., Brady, M.V. and Born, D.J., (1998), Natural Attenuation: CERCLA, RBCA’s, and the Future of Environmental Remediation, Lewis, Boca Raton, 245p. Buckingham, D.A., Metal-OH and its ability to hydrolyze (or hydrate) substrates of biological interest, in Addison, A.W., Cullen, W.R., Dolphin, D., and James, B.R., (Eds.), Biological Aspects of Inorganic Chemistry, Wiley, New York, 1997. Dragun, J., The Soil Chemistry of Hazardous Materials, Hazardous Materials Control Institute, Silver Spring, MD, 1988, 458 pp. Fogel, M.N., Taddeo, A.R., and Fogel, S., Biodegradation of chlorinated ethenes by a methaneutilizing mixed culture, Appl. Environ. Microbiol., 54, 720–724, 1986. Frenkel, M., Surface acidity of montmorillonites, Clays Clay Miner., 22, 435–441, 1974. Kriegman-King, M.R. and Reinhard, M., Reduction of hexachloroethane and carbon tetrachloride at surfaces of biotite, vermiculite, pyrite, and marcasite, in Baker, R.A., Ed., Organic Substances and Sediments in Water, Vol. 2, Lewis, Chelsea, MI, 1991, pp. 349–364. McCarty, P.L. and Semprini, L., Ground-water treatment for chlorinated solvents, in Bioremediation of Groundwater and Geologic Material: A review of In-Site Technologies, Government Institutes., Rockville, MD, 1994, Section 5. Mill, T. and Mabey, W., Hydrolysis of organic chemicals, in Hutzinger, O., (Ed.), Handbook of Environmental Chemistry, Vol. 2D, Reactions and Processes, Springer-Verlag, New York, 1988, pp. 71–111. Mortland, M.M., Clay-organics complexes and interactions, Adv. Agron., 22, 75–117, 1970. Mulligan, C.N., Environmental Biotreatment:Technologies for Air, Water, Soil, and Wastes, Government Institutes, Rockland, MD, 2001, 395 pp. Mulligan, C.N., Yong, R.N., and Gibbs, B.F., The use of selective extraction procedures for soil remediation, Proc. Int. Symp. on Suction, Swelling, Permeability and Structure of Clays, Balkema, Rotterdam, The Netherlands, 2001. Plastourgou, M. and Hoffman, M.R., Transformation and fate of organic esters in layeredflow systems: The role of trace metal catalysis, Environ. Sci. Technol., 18, 756–764, 1984. Schwarzenbach, R.P., Gschwend, P.M., and Imboden, D.M., Environmental Organic Chemistry, Wiley & Sons, Inc. New York, 1993, 681 pp.

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NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Solomon, D.H., and Murray, H.H., Acid-base interactions and properties of kaolinite in nonaqueous media, Clays Clay Miner., 20, 135–141, 1972. Strauss, H., Final Report: An Overview of Potential Health Concerns of Bioremediation, Env. Health Directorate, Health Canada, Ottawa, 1991, 54 pp. Suthersan, S.S., (2002), Natural and Enhanced Remediation Systems, Lewis, Boca Raton, 440p. Tan, B.K., Update on Arsenic Leaching Tests, unpublished report submitted to Geoenvironmental Research Centre, Cardiff University, Wales, 2003. Verschueren, K., Handbook of Environmental Organic Chemicals, 2nd ed., Van Nostrand Reinhold, New York, 1983, 1310 pp. Wieland, E. and Stumm, W., Dissolution kinetics of kaolinite in acidic aqueous solution at 25˚C, Geochim. Cosmochim. Acta, 56, 3339–3355, 1992. Wilson, J.T. and Wilson, B.H., Biotransformation of trichloroethylene in soil, Appl. Environ. Microbiol., 49, 242–243, 1985. Yong, R.N., Desjardins, S., Farant, J.P., and Simon, P., Influence of pH and exchangeable cation on oxidation of methylphenols by a montmorillonite clay, Appl. Clay Sci., 12, 93–110, 1997. Yong, R.N., Geoenvironmental Engineering: Contaminated Soils, Pollutant Fate and Mitigation, CRC Press, Boca Raton, FL, 2000, 307 pp.

Index A

Amphiprotic substance, 76 Anaerobic biodegradation, 207 Anthracene biodegradation, 209 Aquatic plants, 194 Aquifer, chlorinated ethane-contaminated, 235 Aromatic amines, 13 Aromatic compounds, degradation of, 205 Aromatic hydroxy compounds, 163 Arsenate adsorption, 118 Arsenic adsorption of in soils, 139 mobility of in environment, 104 Arthrobacter, 194 ASTM protocol, 250 ATP, see Adenosine triphosphate Attenuating processes, 8 Attenuation, see also Natural attenuation contaminant, 8, 10 degree of, 9 Autoprotolysis, 110 Avogadro’s number, 55

Abiotic reactions, 179 Acetic acid, 165, 179 Achromobacter, 194 Acid acetic, 165, 179 base reactions, 76, 109 digestion, 120 mine drainage, 7 rain, 6 strip solutions, 11 volatile sulphides, 236 Acinetobacter, 194 Acinetobacter calcoaceticus, 210 Actinomycetes, 198 Adenosine triphosphate (ATP), 198 Adsorption isotherm(s), 82, 114, 116 model, 177 nonlinear, 155 organic chemical pollutants, 170 tests, batch equilibrium, 145 Advection-diffusion relationship, 153 Advection-dispersion -biodegradation, 241 reactive equation, 177 Aerobic degradation products of, 217 Agricultural lands, depletion of, 5 Alcohols, 163, 164 Aldehydes, 165 Algae growth of, 193 treatment of contaminants by, 191 Aliphatic hydroxy compounds, 163 Alkaligenes eutrophus, 210 Alkali metals, 102 Alkaline cleaning agents, 11 Alkanes, 203 Alkenes, 203 Amino functional group, 165 Amorphous soil organics, 39, 40

B Bacillus, 194 Bacillus subtilis, 194, 200 Bacteria classiﬁcation of, 194 growth of, 196 nitrifying, 198 reproduction of, 194 sulfate-reducing, 203, 235 sulfur, 214 treatment of contaminants by, 191 Bacterial metabolism nitrogen, 213 sulfur, 214 Bacteriophages, 194 Batch equilibrium isotherm, BTEX, 173 tests, 116, 117, 122, 154

307

308

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Beggiatoa, 214 Benchmark tests, SSFA tests as, 143 Bentonites, 38 Benzene, toluene, ethyl benzene and xylene (BTEX), 18, 101, 159, 204–205 batch equilibrium isotherm, 173 components, aerobic degradation of, 205 degradation of, 217 mobility, 174 MTBE Henry’s law constant and, 168 natural attenuation of, 256 removal, 228 Benzoic acid, 165 BET adsorption isotherm model, see Brauer, Emmett and Teller adsorption isotherm model Bingham yield stress, 44, 45 Bioaugmentation, 292 Bioavailability, 107 Biochemical oxygen demand (BOD), 196 Biocide application, 11 Biodegradation anaerobic, 207 anthracene, 209 chlorinated ethene, 208 MTBE, 257 PAH, 217, 227 rate of, 242 resistance of MTBE to, 207 Biological magniﬁcation, 6 Biological transformation of contaminants, 191–222 biodegradation of organic chemicals, 203–212 alkanes, 203 alkenes, 203 BTEX, 204–206 cycloalkanes, 204 halogenated aliphatic compounds, 207–209 halogenated aromatic compounds, 209–211 MTBE, 207 nitroaromatics, 211–212 petroleum hydrocarbons, 203 polycyclic or polynuclear aromatic hydrocarbons, 206–207 biotransformation of metals, nonmetals and radionuclides, 212–216 bacterial metabolism of nitrogen, 213–214 bacterial metabolism of sulfur, 214–216 metals, 212–213 classiﬁcation of microorganisms, 198–199 physical and chemical environmental effects on microorganisms, 199–202 chemical factors, 200

cometabolism, 202 contaminant availability, 200 halorespiration, 201–202 heavy metals and other compounds, 201 molecular structure, 201 toxicity, 200–201 types of organisms, 191–197 algae, 193 bacteria, 194–197 fungi, 192–193 plants, 194 protozoa, 192 viruses, 193–194 worms, 194 Biomethylation, metals, 212 Bioremediation, natural, 278 Bioslurping, 261 Biostimulation, 291 Biotites, removal of potassium from, 38 Black box deletion approach, PAH, 245 BOD, see Biochemical oxygen demand Boltzmann constant, 92 Boltzmann distribution, 70 Brassica juncea, 194 Brauer, Emmett and Teller (BET) adsorption isotherm model, 177 Breakthrough curves, 123, 124, 125 Brevibacterium, 194 Brevibacterium ethrogenes, 203 Brownian activity, solute, 90 BTEX, see Benzene, toluene, ethyl benzene and xylene Buchner-type pressure-tension apparatus, 72 Buffering capacity, 77 Bulk density values, 58

C Capillaric velocity, 50 Carbonates, 39, 138 Carbonation, 27 Carbonyl compounds, 165 Carboxylates, formation of, 207 Cation(s) exchangeable, 80, 81 sorption, 81 Cation exchange capacity (CEC), 49, 53, 80, 273 clay minerals, 56 correlation of partition coefﬁcients with, 178 illite soil, 82 kaolinite, 108, 124 measurement of, 81 CEC, see Cation exchange capacity Cell leaching tests, 113

INDEX

Charge reversal, 66, 97 surface, pH-dependent, 68 Chemical buffering capacity, 81 oxygen demand (COD), 196 speciation, 71 waste categories, 152 weathering, 25 Chemisorption, 63, 77, 80, 161 Chemolithotrophs, 198 Chloride tracer, 232 Chlorinated benzenes, 174, 175 Chlorinated ethenes, biodegradation of, 208 Chlorinated hydrocarbons, 224 Chlorinated phenols, 9 Chlorinated solvents, 13 Chlorites, 37–38 Chlorobenzenes, desorption of, 176 Chlorophyll, 198 Chromium, sources of, 104 Clay(s) illite, 24 mineral(s), 28, 33 basic building blocks for, 34 CEC for, 56 charge characteristics for, 56 composition of, 34 fractions, interactions between heavy metal pollutants and, 127 heavy metal retention by, 107 particles, 34, 64 phenol oxidation by, 181 SSA for, 56 mixed-layer, 36 particle(s) electriﬁed interface, 66 surfaces, electrostatic charges on, 55 size, 28 soil(s) attribute of, 63 microstructures in, 43 particle sizes in, 31 use of as catalysts, 302 water uptake by dry, 84 surfaces, bridging to, 177 Climate change desertiﬁcation, 6 Coarse-grained soils, 30 COD, see Chemical oxygen demand Cohesionless soils, 30 Column leaching tests, 118, 122 Cometabolism, deﬁnition of, 201 Competitive adsorption model, 123 Composite species, 130 Cone penetrometer system (CPT), 236

309

Constant adsorption curve, 114 model, 115 Constant-type isotherms, 82 Contaminant(s) assimilation of, 72 attenuation, 8, 10, 285 availability, microorganisms, 200 biotic reactions between soils and, 58 concentration, 32 deﬁnition of, 9 heavy metal, 117 hindrance partitioning of, 52 inorganic, 102 movement of, 88 organic chemical, 151 partitioning of, 15, 59, 77, 112 petroleum, 255 physical adsorption of, 78 transport control of, 271 prediction of, 153 volatilization of, 272 Contamination dense nonaqueous phase liquid, 152 mitigation of, 271, 272 Corynebacterium, 194, 203 Coulombic interaction energy, 69 Covalent bonding, 79 CPT, see Cone penetrometer system Critical concentration(s) regulatory control standard, 16 speciﬁcation of, 20 Crop plants, 194 Crystalline water, 84 Cycloalkanes, 204, 206

D Darcy model, 46 Darcy permeability coefﬁcient, 93 Darcy’s law, 49 DCE, see Dichloroethylene DDL model, see Diffuse double-layer model DDLs, see Diffuse double layers DDT, octanol-water partition coefﬁcient, 172 Debye forces, 160 Deforestation, 6 Degree of attenuation, 9 Dehalogenation, 209 Dehalorespiration, 201 Dehydrohalogenation, 180, 184 Dense NAPLs (DNAPLs), 162, 163, 241, 277, 284

310

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Dense nonaqueous phase liquid contamination, sources of, 152 Deoxyribonucleic acid (DNA), 191, 193 Desertiﬁcation, 6 Desorption, prediction of, 183 Desulfotomaculum, 235 Desulfovibrio desulﬁcans, 215 Dichloroethylene (DCE), 18, 255 Diethylenetriamine pentaacetic acid (DTPA), 226 Diffuse double-layer (DDL) model, 78, 84 Diffuse double layers (DDLs), 159 Dihydrodiols, 205 Dilution, pollutant attenuation through, 16 Dioxin, cometabolism of, 202 Discrete multilayer samples (DMLs), 236 Dispersivity parameter, 94 Dissimilatory sulphite reductase (DSR), 235 Distribution coefﬁcients, 82, 83, 121, 156 DMLs, see Discrete multilayer samples DNA, see Deoxyribonucleic acid DNAPLs, see Dense NAPLs DSR, see Dissimilatory sulphite reductase DTPA, see Diethylenetriamine pentaacetic acid

E EDTA, see Ethylenediaminetetraacetic acid Eh-pH relationships, 111 Electron acceptors, 242 donors, 198 transfer, 109 ELISA, see Enzyme-linked immunosorbent assay ENA, see Enhanced natural attenuation Engineered clay barrier, 297 Engineered natural attenuation (EngNA), 3, 18, 247, 273 capability, evidence of, 298, 299, 300 misnaming of, 297 EngNA, see Engineered natural attenuation Enhanced natural attenuation (ENA), 18, 290, 291 application of, 293 use of treatment wells for, 295 Entamoeba histolytica, 192 Enzyme-linked immunosorbent assay (ELISA), 260 Equilibrium partition coefﬁcient, 170, 171 Escherichia, 197 Escherichia coli, 214 Ethylenediaminetetraacetic acid (EDTA), 226 Eukaryotes, 191 Exchangeable cations, 80, 81 Exchangeable metals, 137 Exogenous microorganisms, 292

Exploitation, ground contamination and, 10 Extraction procedures, 136

F Fate and transport models common analytical, 240 purpose of, 224 Ferns, 194 Fertilizer application, 11 phosphate, 103 Field capacity, soil, 88 Field performance and assessments, 223–269 assessment of potential for natural attenuation, 224–228 bioattenuation and bioavailability, 226–227 other factors, 227–228 processes and mechanisms involved in NA, 225–226 assessment of sustainability, 228–229 case studies of natural attenuation, 255–263 combination of natural attenuation with other remediation processes, 261–263 natural attenuation of chlorinated solvents, 255–256 natural attenuation of MTBE, 256–257 natural attenuation as sole remediation technology, 257–261 models to simulate natural attenuation, 239–245 application, calibration and veriﬁcation of models, 244–245 available models for natural attenuation, 240–244 background on modeling, 239–240 procedures for monitoring, 229–239 components of monitoring, 230–233 development of monitoring techniques, 234–239 established monitoring techniques, 233–234 importance of monitoring, 229–230 protocols developed for natural attenuation, 245–255 inclusion of soils and sediments in protocols, 254–255 technical protocols, 247–254 First-order decay coefﬁcient, 242 First wetting, 158 Flatworms, 194 Flavobacterium, 194, 210, 211 Flow

INDEX

311

computations of, 89 prediction of unsaturated, 50 unsaturated, 85 velocity, 50 vertical, 90 Fluid conductivity, 55 Forest ﬁres, 104 Frame of reference, soil particles as, 74 Freundlich isotherms, 82, 83, 114, 170 Fuel hydrocarbons, 233 Fulvic acid, 40 functional groups for, 54 sorption selectivity of, 129 Functional group(s) amino, 165 organic chemical compounds, 162 Fungi reproduction of, 192–193 treatment of contaminants by, 191 Fungicides, 103 Funnel-gate technique, 295, 297

G Gasoline, additive to, 207 Gaussian statistics, 58 GC-DDL model, see Gouy-Chapman diffuse double layer model Genbank, 235 Geologic origin, soil, 27 Giardia lambia, 192 Global positioning system, modeling software with, 239 Goethite, sorption selectivity of, 129 Gouy-Chapman diffuse double layer (GC-DDL) model, 68, 69 Gouy forces, 10 Grasses, 194 Greenhouse gases, 6 Ground contamination, natural attenuation and, 1–21 contaminated ground, 7–10 control of ground contamination through attenuation, 8–9 pollutants, contaminants, groundwater and pore water, 9–10 control and management of pollution plumes, 12–17 constructed and emplaced barriers, 14–15 dilution, retardation, retention and attenuation, 15–17 engineered barrier-liner systems, 13–14 land use, 4–7 land suitability and use, 4

land use, ground contamination and sustainable development, 5–7 natural attenuation, 2–3 natural attenuation and regulatory attitudes, 17–20 pollutants and ground contamination, 10–12 Ground-penetrating radar, 237, 238 Groundwater deﬁnition of, 9–10 ﬁlter socks added to, 199 monitoring wells, 286 samples, importance of monitoring, 305 velocity, 227

H Hagen-Poiseuille relationship, 50 Halloysite, 36 Halogenated aliphatic compound, examples of, 207 Halogenated alkanes, 179 Halogenated aromatic compounds, 209 Halorespiration, 201, 202 Hard acids, 110 Hard bases, 110 Heavy metals (HMs), 11 estuarine alluvium retention of, 140 hydroxy species and, 139 monitoring for natural attenuation of, 231 natural bioremediation of soils contaminated with, 300 pH sensitivity in, 282 precipitation of, 106 sorption selectivity, 129 Heavy metals, partitioning and mobility of, 101–149 inorganic contaminants, 102–107 interactions and assimilation of heavy metals, 108–113 acid-base reactions, 109–111 molecular interactions, 112–113 redox and Eh-pH relationships, 111–112 mobility, availability and SPR, 107–108 partitioning of heavy metals, 113–127 adsorption isotherms, 114–118 breakthrough curves and retardation factors, 123–127 column leaching tests, 118–121 distribution coefﬁcient, 121–123 preferential sorption and selectivity, 127–132 inﬂuence of ligands, 131–132 selectivity, 127–130 soil fractions and sorption, 132–143 selective sequential extraction, 134–141

312

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

selective sequential fraction addition or removal, 142–143 Height of capillary rise experiment, 73 Helmholtz plane, 67, 68, 78, 303 Henry’s law constant, 167, 168, 169 Herbicide(s) adsorption, 172 sorption and desorption of, 170 tests on, 178 High-afﬁnity speciﬁcally sorbed ions, 80 High-afﬁnity-type isotherms, 82 HMs, see Heavy metals Household wastes, 152 HRC, see Hydrogen release compound Humic acid(s), 40 functional groups for, 54 sorption selectivity of, 129 Humins, 40, 54 Hydration water, volume expansion associated with, 75 Hydraulic conductivity, 58 coefﬁcient, variation of, 89 factors affecting, 46 Hydrocarbon(s), 277 contaminated soils, 141 Henry’s law constants of, 169 Hydrogen bonding, 160, 177 release compound (HRC), 215, 262 Hydrolysis reactions, soil-catalyzed, 303 Hydrophobic bonding, 161 Hydrophobic substances, soil-organic chemical permeabilities of, 159 Hydrous oxides, 54 Hydroxides, 40 Hydroxy compounds aliphatic, 163 aromatic, 163 Hydroxylated surfaces, 52 Hydroxy species, heavy metals and, 139

I Illicit dumping, 1 Illite(s), 36 clay, 24 CEC value, 124 sorption selectivity of, 129 interlayered, 37 removal of potassium from, 38 soil Pb-polluted, 281 retention of heavy metals by, 128 Indifferent ions, 78

Industrialization, ground contamination and, 10 Inorganic contaminants, 102 Inorganic crystalline material, 30 Inorganic noncrystalline material, 30 In situ reactive region (IRR), 293, 294 Interlayered vermiculites, 37–38 Intrinsic remediation, 273 Ion(s) activation energy requirements for movement of, 88 exchange, 161 high-afﬁnity speciﬁcally sorbed, 80 indifferent, 78 -ion interactions, 79 potential determining, 80 speciﬁc, 78 Iron oxides, 41 IRR, see In situ reactive region Isoelectric point, 66 Isolation barriers, 14 Isotope exchange techniques, 234

K Kandite, 34 Kaolin, 34 Kaolinite(s), 27, 34, 35 CEC, 124 clay, sorption selectivity of, 129 edges of, 165 hydroxyl terminals of, 53 rheograms for, 44 soil adsorption isotherm for, 122 sorption of Cd by, 133 suspension, 43 surface acidity, 162 charges in, 81 K-C model, see Kozeny-Carman model Keesom forces, 160 Kozeny-Carman (K-C) model, 46, 47

L Land(s) agricultural, depletion of, 5 deﬁnition of, 4 environment basic component of, 4 quality of, 20–21 sustainable, 21 quality, maintenance of, 21 use

INDEX

313

sustainability of, 4 types of, 4 Landﬁlls construction of, 19 improperly managed, 152 waste disposal in, 14 Langmuir-type isotherms, 82, 114 Laser-induced ﬂuorescence (LIF), 236, 237, 239 Layered treatments, 3 Leaching column samples, sorption curves for, 119 tests, 96, 156, 283 tests, 82, 113, 233 Lead adsorption, 103 Leaking underground storage tanks, 207 Lewis acid(s) deﬁnition of, 110 sites, 53, 64, 108 Lewis bases, deﬁnition of, 110 LIF, see Laser-induced ﬂuorescence Ligand(s), 105 exchange, 161 heavy metal retention and, 131 Light NAPLs (LNAPLs), 162, 163, 241, 277 Lines of evidence, 19 MNA, 230 requirements for determining, 275 LNAPLs, see Light NAPLs London dispersion forces, 160 London-van der Waals forces, 160 Longitudinal diffusion coefﬁcient, 94

M Macrophytes, 194 MAHs, see Monocylcic aromatic hydrocarbons Manufacturing plants, abandoned, 151 Matric potential, 74 MBTE ﬁeld studies, results of, 257 Mesophiles, 199 Metal(s), see also Heavy metals alkali, 102 degreasing, 11 exchangeable, 137 microbial conversion of, 212 oxides, metals contaminants associated with, 138 plant species that accumulate, 194 species sorption, selectivity in, 130 transition, 102 Methanotrophs, 202 Methemoglobinemia, 213 Methyl tertiary-butyl (MTBE), 199

behavior of, 167 biodegradation rates, 257 chemical reactions, 179 mobility, 174 natural attenuation of, 256 recommendations concerning, 254 resistance of to biodegradation, 207 Micas, 36, 38 Micrococcus, 194 Microcosm studies, 274 Microenvironment, pH of, 105 Microorganisms chemical factors affecting, 200 classiﬁcation of, 198 contaminant availability, 200 environmental effects on, 199 exogenous, 292 molecular structure, 201 nutritional needs of, 198 Microsoft Excel, 240 Mineral, chlorite-type, 38 Mixed-layer clays, 36, 37 MNA, see Monitored natural attenuation Model(s) application, calibration and veriﬁcation of, 244 BET adsorption isotherm, 177 BIOCHLOR, 241, 244 BIOPLUME, 241, 242, 243 BioRedox, 243, 244 BIOSCREEN, 240, 241, 244, 259 competitive adsorption, 123 constant adsorption, 115 Darcy, 46, 50 development, procedure for, 248 diffuse double-layer, 78 EQ3, 71 fate and transport common analytical, 240 purpose of, 224 Freundlich, 178 GEOCHEM, 71 Gouy-Chapman diffuse double layer, 68, 69 Jury’s, 239 Kozeny-Carman, 46, 47 MINTEQA2, 71 MODFLOW, 243, 244 numerical, 240 PHREEQE 71 PHREEQM, 71 Poiseuille, 46 predictions, ﬁeld performance trends and, 289 RT3D, 243 SESOIL, 239, 259 SOILCHEM, 71

314

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

Stern layer, 67 summary of natural attenuation, 246–247 surface complexation, 69 transport and fate, 152 design of, 153 focus of, 153 requirements for, 154 VADSAT, 239 Monitored natural attenuation (MNA), 3, 273 deﬁnition of, 18 effectiveness, 287 implementation, 19, 288 key element in application of, 152 lines of evidence, 144, 230 process effectiveness of, 20 modeling of, 19 remediation scheme, 3 requirements, 273 use of as remediation tool, 224 Monitoring scheme, 285 stations, analysis of samples from, 20 wells distribution of, 288 placement of, 19 Monocylcic aromatic hydrocarbons (MAHs), 166 Monod equation, 196 Monod kinetics, 242 Monooxygenase enzymes, 202 Montmorillonite, 38, 56, 112 clay, sorption selectivity of, 129 distribution of Cd retained in, 140 surface acidity of, 162, 303 Mosses, 194 MTBE, see Methyl tert-butyl ether Mucor,193 Multi-component species, 130 Mushrooms, components of, 193 Mycobacterium fortuitum, 203 smegmatis, 203

N NA, see Natural attenuation Nacrites, 36 Naphthalenes, 9, 178 NAPLs, see Nonaqueous phase liquids Natural attenuation (NA), 223 advantages, 263 assessment of potential for, 224 available models for, 240 capability, soils, 1

capacity, 223 disadvantages, 263 effectiveness, 244 engineered, 3 enhanced, 18 heavy metals, 231 monitored, 3, 144 processes and mechanisms involved in, 225 protocols, 245–255 comparison of available, 253 inclusion of soils and sediments, 254 technical, 247 regulatory attitudes and, 17 role of in ground pollution management, 2 sustainability of, 228 Natural attenuation, present application and future directions for, 271–306 engineered natural attenuation, 297–301 enhanced natural attenuation, 290–297 application of enhanced natural attenuation, 293–297 bioaugmentation, 292 biostimulation, 291–292 geochemical and biogeochemical intervention, 292–293 soil improvement, 293 mitigation of contamination, 272–284 desorption of pollutants and augmentation of pollutants, 280–281 monitored natural attenuation requirements, 273–275 natural attenuation and lines of evidence, 275–280 pollutant release studies for augmentation assessment, 281–284 monitoring and evidence of success, 284–290 observations on future applications, 301–304 use of clay soils as catalysts, 302–304 use of soils to promote oxidationreduction reactions, 304 Natural resources, increased demand on, 6 Negative charges, kinds of, 66 Nematodes, 194 Net surface charges, 66 Neutralization, 110 Neutral salts, 281 Nitrilotriacetic acid (NTA), 226 Nitroaromatics, 211 Nitrobacter, 198 Nitrobacter agilis, 194 Nitrogen, bacterial metabolism of, 213 Nitrosomonas, 198 Nocardia sp., 203, 204 Nonaqueous phase liquids (NAPLs), 159, 211, 277

INDEX

315

contaminants, 247 natural attenuation evaluation and, 228 plume, well location for monitoring of, 251 Non–point source pollution, 7 NTA, see Nitrilotriacetic acid

O Octanol-water partition coefﬁcient, 157, 183 Oil sludges, 13 Organic(s) acids, 165, 201 adsorption, primary mechanisms of, 157 animal-derived, 40 matter, metals associated with, 138 plant-derived, 40 sludges, 152 transformed, 39 unaltered, 39 Organic contaminants, interactions between soil–water systems and, 151–189 adsorption and bonding mechanisms, 158–169 functional groups and bonding, 161–167 intermolecular interactions, 159–161 volatilization, 167–169 interactions and fate, 179–182 abiotic reactions, 179 dehydrohalogenation, 180 hydrolysis, 179–180 oxidation-reduction reactions, 180–182 mobility of organic chemical contaminants, 152–157 organic chemical contaminants, 151–152 partitioning of organic chemical pollutants, 169–178 adsorption isotherms, 170 equilibrium partition coefﬁcients, 170–176 studies on sorption of various organic chemicals, 176–178 Organisms, ﬁve-kingdom classiﬁcation, 191, 192 Organochloride compounds, 13 Organohalide compounds, 277 Osmotic potential, 74 OSWE, see U.S. Environmental Protection Agency Ofﬁce of Solid Waste and Emergency Response Oxidation-reduction reactions, 76, 77, 180, 304 Oxides, 40, 64 Oxyanions, 213 Oxyhydrates, extractant selected for, 138 Oxyhydroxides, 40

P PAHs, see Polycyclic aromatic hydrocarbons Parent rock, 26, 27 Particle surfaces, availability of, 33 Partitioning, 82 contaminants, 15, 32, 113 heavy metals, 113 organic chemical pollutants, 169 tests, types of, 145 Passive remediation, 273 PCBs, see Polychlorinated biphenyls PCE, see Perchlorethylene PCO, see Pentacholorphenol PCR, see Polymerase chain reaction Peclet number, 92 Penicillium, 193 Pentacholorphenol (PCP), 199 Perchlorethylene (PCE), 18 Permeability coefﬁcient, variability of, 48, 49 Permeable reactive barrier (PRB), 294, 295, 296 Pesticides, 9, 178 Petroleum contaminants, 255 hydrocarbons (PHCs), 11, 101, 203, 224, 277 desorbed, 167 partitioning of, 182 petroleum fractions in, 166 transport of in soils, 158 water solubility of, 166, 167, 176 pH, manipulation of, 292 Phaenaerochaete chrysosporium, 193 PHCs, see Petroleum hydrocarbons Phenols, 163, 181 Phenomenological coefﬁcients, 93 Phosphate fertilizers, 103 Photolithotrophs, 198 Photoorganotrophs, 198 Photosynthesis, 6 Physical hindrance, contaminant concentration and, 32 Physisorption, 63 Phytoremediation, 262, 300 Pipeline oil spill, 258 Plants metal-tolerant, 302 remediation technology and, 194 Plutonium, 258 Point of zero charge, 65 Poiseuille model, 46 Pollutant(s) augmentation of, 280 cost effectiveness of removal of, 1 deﬁnition of, 9 desorption of, 280

316

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

discharges, 7 ground contamination by, 5, 7 heavy metal, 101 inorganic, 296 leachate plume with, 12 mobility of, 12 movement of nonliquid, 84, 97 natural attenuation of, 302 organic chemical, 164, 276 adsorption isotherms, 170 partitioning of, 169 partitioning tests, 156 plume management, 272 schematic, 274 priority, 102 proﬁles, 119, 120 pulses, 17 release studies, augmentation assessment, 281 -soil bonds, destroying, 134 sorbed, 108 threats posed by, 1 tracking requirement for pollution plume associated with, 152 transport, prediction of, 153 types of in contaminated ground, 11 Pollution non–point source, 7 plumes control and management of, 12 remedy against spread of, 14 waste, 5 Polychlorinated biphenyls (PCBs), 9, 11, 156, 162 cometabolism of, 202 natural attenuation of, 260 reductive dechlorination of, 260 Polycyclic aromatic hydrocarbons (PAHs), 101, 157, 199, 203, 206 biodegradation of, 217, 227 black box deletion approach, 245 degradation, 207 hydrolysis reactions for, 179 molecular weight of various, 208 redistribution of, 176 sorption, 245 Polymerase chain reaction (PCR), 235 Polyvinyl chloride (PVC), 103 Populus, 194 Pore channel ﬂow, 87 cross-section area, typical value for, 47 ﬂuid analysis, 120 Pore water chemical reactions in, 75 deﬁnition of, 9–10

leaching of DNAPL into, 284 ligands present in, 131 removal of solutes from, 279 samples, importance of monitoring, 305 Positive charges, 66 Potential determining ions, 67, 80 PRB, see Permeable reactive barrier Precipitation mechanisms, retention by, 141 Predicted sorption curve, 131 Pressure-membrane apparatus, 73 Proton acceptor, 76, 109 donor, 76, 109 Protonation, 162 Protophillic solvent, 76 Protozoa size of, 192 treatment of contaminants by, 191 Pseudomonas, 194, 204, 210, 235 Pseudomonas aeruginosa, 158, 194, 200 Psychrophiles, 199 Public health, potential threats to, 182 Pump-and-treat methods, 295 PVC, see Polyvinyl chloride

R Radionuclides, 213, 255 Rate constants, ﬁeld data, 250 RCRA sites, 252 RDX, see Royal demolition explosive Reaction rates, quantiﬁcation of, 180 Reactive surfaces, 52, 59 Redox potential, 76 Reductive dehalogenation, 209, 211 Regional controls forces exercised as result of, 27 soil, 27 Regulatory command and control requirements, 297 Regulatory control standard, 16 Remediation process, MNA as, 273 technology, natural attenuation as sole, 257 Representative elementary volume (REV), 55 Retardation factors, 123, 126 Retention proﬁles, 120 REV, see Representative elementary volume Reynolds’ number, 51 Rhizopus, 193 Rhodococcus chlorophenolicus, 210 Ribonucleic acid (RNA), 193, 235 Rio Declaration, 5 Risk evaluation, 289

INDEX

317

RNA, see Ribonucleic acid Rock(s) chemical weathering of, 28 erosion, 104 physical weathering of, 25 Royal demolition explosive (RDX), 212

S Scanning electron microscopy (SEM), 234 Seed plants, 194 Selective sequential extraction (SSE), 132, 279 implementation of, 135, 137 reagents used in, 134, 137 underlying though behind, 134 Selective sequential fraction addition (SSFA), 132, 142 removal (SSFR), 132, 142 SEM, see Scanning electron microscopy Sequential soil extractions, 233 Sewage sludges, 103 Sink-source phenomenon, 96 Site characterization, 273, 286 Slime molds, 193 Smectites, 38 Soft acids, 110 Soft bases, 110 Soil(s) arsenic-contaminated, 283, 284 assimilative capacity of, 8, 291 breakthrough performance of, 124, 126 -catalyzed oxidation, 304 chemical buffering capacity, 81 classiﬁcation, 27, 28 clay, microstructures in, 43 coarse-grained, 30 cohesionless, 30 composition, 20, 30 control of pollutants in, 16 digestion, 234 environment, 183 ﬁeld capacity, 88 ﬂuid transmission properties of, 28 fractions, 29 gas sampler, 239 improvement, procedures for, 293 macrostructure, 33 major sources of, 25 mass, density of, 42 most signiﬁcant animals in, 194 movement of water in, 86 natural attenuation capacity of, 14, 15, 19 natural levels of zinc in, 103 organics, 30, 39

acidic properties associated with, 65 amorphous, 39, 40 important functional groups in, 53 natural aromatic, 290 paddy, 178 parameters for monitoring in, 234 permeability, expression of, 48 pH, 40 -PHC bonding, 158 -pollutant interaction mechanisms, 2 pores, air-water interfaces in, 73 saturated, 85 solids copper retention by, 26 energy relations between water and, 72 interactions between heavy metal pollutants and, 101 ion-ion interactions between, 79 lead retention by, 25, 26 pollutant residence associated with, 285 solute movement in saturated, 93 structure, 42 suction, 74 swelling, 75, 85 transmission property of, 45 uptake capability, 52 vapor(s) extraction, 261 technique for sampling, 237 water -holding relationships, characterization of, 73 movement in, 84 Soil composition and transmission properties, 23–61 clay minerals and soil fractions, 33–41 carbonates, 39 illites, micas and mixed-layer clays, 36–37 kaolin, kaolinite and kandite, 34–36 oxides, hydroxides and oxyhydroxides, 40–41 smectites, 38 soil organics, 39–40 vermiculites, interlayered vermiculites and chlorites, 37–38 nature of soil, 25–31 soil composition, 29–31 soil material and classiﬁcation, 27–29 physical characteristics and properties, 41–52 Darcy’s law, low water contents and unsaturated ﬂow, 49–52 soil composition and soil structure, 42–45 soil structure and transmission properties, 45–49

318

NATURAL ATTENUATION OF CONTAMINANTS IN SOILS

soil fractions, composition and attenuation, 31–33 soil properties and natural attenuation, 24–25 surface properties of soils, 52–58 reactive surfaces of soil fractions, 52–53 speciﬁc surface area, 55–58 surface charge density, 53–55 Soil organic matter (SOM), 30, 80, 117, 290 classiﬁcation of, 40 functional groups, 54, 163 hydrophobic chemical compounds and, 173 oil retention and, 158 organic carbon content in, 173 oxides and, 64 Soil-water systems and interactions, 63–99 chemical buffering and partitioning, 81–83 chemical reactions in pore water, 75–77 acid-base reactions and hydrolysis, 76 oxidation-reduction reactions, 76–77 functional groups and electric charges, 63–71 applications and chemical speciation, 71 clay mineral particles, 64 electriﬁed interface and interactions, 66–69 interactions and surface complexation models, 69–71 negative and positive charges, 66 oxides and soil organic matter, 64–65 interactions, exchanges and sorption, 77–81 bonding and sorption mechanisms, 79–80 cation exchange, 80–81 movement of solutes, 90–96 diffusion of solutes and diffusion coefﬁcient, 90–93 solute movement in saturated soils, 93–96 soil-water energy characteristics, 71–75 water movement in soils, 84–90 unsaturated ﬂow, 85–90 water uptake by dry clay soil, 84–85 Solute(s) Brownian activity of, 90 diffusion front, 126 electron gain by, 111 Fickian diffusion of, 94 mechanisms of transport of, 92 movement of, 90, 93 retardation front, 126 Solvent(s) chlorinated, 13 dry cleaning, 152 protophillic, 76 SOM, see Soil organic matter Sorption, 77 capacity limit, 119 cation, 81

curve, predicted, 131 deﬁnition of, 112, 113 herbicide, 170 mechanisms, 79 naphthalene, 178 organic chemicals, 176 preferential, 127 selectivity, heavy metal, 129 soil fractions and, 132 Source-receptor-path (SRP), 2, 12, 107 Speciation, 131, 145 Speciﬁc ions, 78 Speciﬁc surface area (SSA), 33, 49, 55, 121 clay minerals, 56 ﬁne-grained soils, 33 laboratory measurement of, 57 values of, 59 Spent acids, 13 SRB, see Sulfate-reducing bacteria SRP, see Source-receptor-path SSA, see Speciﬁc surface area SSE, see Selective sequential extraction SSFA, see Selective sequential fraction addition SSFR, see Selective sequential fraction removal Steel pickling, 11 Steric bond, 80 Stern layer, 78 charge, 68 model, 67 Storage tanks, leaking, 11 Subsoil material, interaction between organic chemicals and, 151 Sulfate-reducing bacteria (SRB), 203, 235 Sulfonylurea, 200 Sulfur, bacterial metabolism of, 214 Superfund sites, 11, 252 Surface charge density, 53, 70 complexation mechanisms of, 177 models, 69 functional groups, 52, 59 Sustainable development, land use, 5 Swelling soils, 85

T Target concentration,(s) regulatory control standard, 16 speciﬁcation of, 20 TCA, see Tetrachloroethane TCE, see Trichloroethylene TCLP, see Toxicity Characteristics Leaching Procedure

INDEX

319

Tetrachloroethane (TCA), 202 Thlaspi calaminare, 194 TOC, see Total organic carbon Total organic carbon (TOC), 196 Total petroleum hydrocarbons (TPHs), 196 Toxicity Characteristics Leaching Procedure (TCLP), 11 TPHs, see Total petroleum hydrocarbons Transformed organics, 39 Transition metals, 102 Transport and fate models, 152 design of, 153 focus of, 153 requirements for, 154 Transport processes, 95 Treatment trains, 3 wells, ENA using, 295 Trichloroethylene (TCE), 18, 177, 278 cometabolism of, 202 degradation of, 202 hydrolysis reactions, 179 Triclinic unit cell, 35

U Unaltered organics, 39 Underground storage tanks, leaking, 141, 207 Unsaturated ﬂow, 85 Urbanization, ground contamination and, 10 U.S. Air Force Protocol, 252 U.S. Department of Energy, technical guidance for natural attenuation by, 232 U.S. Environmental Protection Agency (U.S. EPA), 1 Center for Subsurface Modeling Support Internet Site, 241 deﬁnition of MNA by, 18 directive, natural attenuation use, 229 Ofﬁce of Solid Waste and Emergency Response (OSWE), 227 Priority Pollutants List, 9 U.S. EPA, see U.S. Environmental Protection Agency

Vermiculite, 37–38, 56 Vibrio, 194 Video microscope, 239 Viruses replication of, 193 treatment of contaminants by, 191 VOCs, see Volatile organic compounds Void ratio, 58 Volatile organic compounds (VOCs), 177 nondegradable, 263 sorption of, 177 Volatilization, 184 negligible, 168 rate of, 167 Volcanoes, 104

W Waste(s) disposal landﬁll systems, 14 generation and pollution, 5 household, 152 impoundment facilities, engineered containment barriers lining, 13 storage, insufﬁcient, 151 streams, 11, 13 Water, see also Pore water crystalline, 84 distribution and migration of, 84 energy relations between soil solids and, 72 hydration, volume expansion associated with, 75 movement, 84, 85 oxygen solubility in, 199 transfer, 85 uptake, dry soil, 87 Weathering processes, soils formation and, 27 Wetted surface ratio (WSR), 48 Wetting front proﬁle, 90, 91 World Health Organization, 292 Worms, signiﬁcance of in soil, 194 WSR, see Wetted surface ratio

Y

V Yeasts, 193 van der Waals forces, 79, 160 Vapor transfer, 85 Variable-charge surfaces, 66 Vegetation, decaying, 25

Z Zinc, natural levels of in soils, 103