During recent years much of the emphasis in groundwater investigations in industrialized countries has shifted from problems of groundwater supply to considerations of groundwater quality. As a result of our consumptive way of life, the groundwater environment is being assaulted with an ever-increasing number of soluble chemicals. Current data indicate that in the United States there are at least 17 million waste disposal facilities emplacing more than 6.5 billion cubic meters of liquid into the ground each year (US. Environmental Protection Agency, 1977). As time goes on, the vast subsurface reservoir of fresh water, which a few decades ago was relatively unblemished by man’s activities, is gradually becoming degraded.
The problem of water quality degradation of rivers and lakes has been evident for a long time. In general, solutions to this problem have been found in the implementation of effective legislation for discontinuing contaminant emissions. Already in some parts of the world, effective emission abatement measures have led to great improvements in surface-water quality. Unfortunately, problems of groundwater quality degradation are in many ways more difficult to overcome. Because of the heterogeneities inherent in subsurface systems, zones of degraded groundwater can be very difficult to detect. The US. Environmental Protection Agency (1977) has reported that almost every known instance of aquifer contamination has been discovered only after a water-supply well has been affected. Often by the time subsurface pollution is conclusively identified, it is too late to apply remedial measures that would be of much benefit. From a water quality viewpoint, degradation of groundwater often requires long periods of time before the true extent of the problem is readily detectable. Long periods of groundwater flow are often required for pollutants to be flushed from contaminated aquifers. Groundwater pollution often results in aquifers or parts of aquifers being damaged beyond repair.
Whereas the problem of achieving acceptable quality of surface waters focuses mainly on decreasing the known emissions of pollutants to these systems, the problem facing scientists and engineers involved in the protection of groundwater resources is to identify the areas and mechanisms by which pollutants can enter groundwater flow systems and to develop reliable predictions of the transport of contaminants within the flow systems. This is necessary as a basis for minimizing the impact of existing or proposed industrial, agricultural, or municipal activities on groundwater quality.
The purpose of this chapter is to provide some insight into the physical and chemical factors that influence the subsurface migration of dissolved contaminants. To this end the behavior of nonreactive solutes and of solutes that undergo reactions during subsurface migration will be considered. Following this, more specific contamination problems related to activities such as agriculture, mining, nuclear power development, and disposal of refuse, sewage, and industrial wastes will be briefly reviewed.
Throughout this chapter all solutes introduced into the hydrologic environment as a result of man’s activities are referred to as contaminants, regardless of whether or not the concentrations reach levels that cause significant degradation of water quality. The term pollution is reserved for situations where contaminant concentrations attain levels that are considered to be objectionable.
The emphasis in this chapter is on the occurrence and processes that control the migration of dissolved contaminants in groundwater. Groundwater can also be contaminated by oily substances that exist in a liquid state in contact with water in a manner that does not lead to mixing of the oils in a dissolved form. The oily liquid is said to be immiscible in the water. The physical processes that control the movement of immiscible fluids in subsurface systems are described by Bear (1972) and are introduced in Section 9.5.
9.1 Water Quality Standards
Before proceeding with discussions of the principles of contaminant behavior in groundwater flow systems and of sources of groundwater contamination, we will briefly examine some of the more important water quality standards. These standards serve as a basis for appraisal of the results of chemical analyses of water in terms of suitability of the water for various intended uses. The most important of these standards are those established for drinking water (Table 9.1). The recommended limits for concentrations of inorganic constituents in drinking water have existed for many years. Limits for organic constituents such as pesticide residues are a recent addition. There is considerable controversy with regard to the specific organic constituents that should be included in drinking water standards and the Concentration limits that should be established for them.
In Table 9.1 the major constituents for which recommended permissible limits are listed are total dissolved solids (TDS), sulfate, and chloride. Consumption by humans of waters with concentrations somewhat above these limits is generally not harmful. In many regions groundwater used for drinking-water supply exceeds the limits of one or more of these parameters. Several hundred milligrams per liter of chloride must be present in order for saltiness to be detected by taste.
|Constituent||Recommended concentration limit* (mg/ℓ)|
|Total dissolved solids||500|
|Hydrogen sulphide (H2S)||0.05|
|Maximum permissible concentration‡|
|Carbon chloroform extract||0.2|
|Radionuclides and radioactivity||Maximum permissible activity (pCi/ℓ)|
|Gross beta activity||30|
|Gross alpha activity||3|
|Total coliform bacteria||1 per 100m/ℓ|
Hardness of water is defined as its content of metallic ions which react with sodium soaps to produce solid soaps or scummy residue and which react with negative ions, when the water is evaporated in boilers, to produce solid boiler scale (Camp, 1963). Hardness is normally expressed as the total concentration of Ca2+ and Mg2+ as milligrams per liter equivalent CaCO3. It can be determined by substituting the concentration of Ca2+ and Mg2+, expressed in milligrams per liter, in the expression
Each concentration is multiplied by the ratio of the formula weight of CaCO3 to the atomic weight of the ion; hence the factors 2.5 and 4.1 are included in the hardness relation. Water with hardness values greater than 150 mg/ℓ is designated as being very hard. Soft water has values less than 60 mg/ℓ. Water softening is common practice in many communities where the water supply has a hardness greater than about 80–100 mg/ℓ. Water used for boiler feed will cause excessive scale formation (carbonate-mineral precipitation) if the hardness is above about 60–80 mg/ℓ.
Of the recommended limits specified for minor and trace inorganic constituents in drinking water, many have been established for reasons other than direct hazard to human health. For example iron and manganese are both essential to the human body. Their intake through drinking water is normally an insignificant part of the body requirement. The recommended limits placed on these metals in the Standards is for the purpose of avoiding, in household water use, problems associated with precipitates and stains that form because oxides of these metals are relatively insoluble (Camp, 1963). The recommended limit for zinc is set at 5 mg/ℓ to avoid taste produced by zinc at higher concentrations. Concentrations as high as 40 mg/ℓ can be tolerated with no apparent detriment to general health. Zinc concentrations as low as 0.02 mg/ℓ are, however, toxic to fish. Zinc contamination can be regarded as severe pollution in ecological systems where fish are of primary interest but may be only of minor significance if human consumption is the primary use of the water.
The most common identifiable contaminant in groundwater is nitrate (). The recommended limit for nitrate in drinking water is 45 mg/ℓ expressed as NO or 10 mg/ℓ expressed as N. In Europe the limit recommended by the World Health Organization is 50 mg/ℓ as and 11.3 mg/ℓ as N. Excessive concentrations of NO have potential to harm infant human beings and livestock if consumed on a regular basis. Adults can tolerate much higher concentrations. The extent to which NO in water is viewed as a serious pollutant therefore depends on the water use.
The constituents for which maximum permissible concentration limits have been set in drinking water standards (Table 9.1) are all considered to have significant potential for harm to human health at concentrations above the specified limits. The specified limits are not to be exceeded in public water supplies. If the limits for one or more of the constituents are exceeded, the water is considered to be unit for human consumption. The limits indicated in Table 9.1 are representative of the current standards in the United States and Canada. The limits are continually being appraised and modifications occur from time to time. As more is learned about the role of trace constituents in human health, the list of constituents for which maximum permissible limits exist may expand, particularly in the case of organic substances.
In many regions the most important uses of groundwater are for agriculture. In these situations it is appropriate to appraise the quality of groundwater relative to criteria or guidelines established for livestock or irrigation. Recommended concentration limits for these uses are listed in Table 9.2. The list of constituents and the concentration limits are not as stringent as for drinking water. These water quality criteria do serve to indicate, however, that concentration increases in a variety of constituents due to man’s activities can cause serious degradation of groundwater quality even if the water is not used for human consumption.
|Livestock: Recommended limits (mg/ℓ)||Irrigation crops: Recommended limits (mg/ℓ)|
|Total dissolved solids|
9.2 Transport Processes
The common starting point in the development of differential equations to describe the transport of solutes in porous materials is to consider the flux of solute into and out of a fixed elemental volume within the flow domain. A conservation of mass statement for this elemental volume is
The physical processes that control the flux into and out of the elemental volume are advection and hydrodynamic dispersion. Loss or gain of solute mass in the elemental volume can occur as a result of chemical or biochemical reactions or radioactive decay.
Advection is the component of solute movement attributed to transport by the flowing groundwater. The rate of transport is equal to the average linear groundwater velocity, , where , v being the specific discharge and n the porosity (Section 2.12). The advection process is sometimes called convection, a term that in this text is reserved for use in discussion of thermally driven groundwater flow as described in Chapter 11. The process of hydrodynamic dispersion, which is described in Section 2.13, occurs as a result of mechanical mixing and molecular diffusion.
Mathematical descriptions of dispersion are currently limited to materials that are isotropic with respect to dispersion properties of the medium. The principal differential equation that describes transport of dissolved reactive constituents in saturated isotropic porous media is derived in Appendix X. This equation is known as the advection-dispersion equation. Our purpose here is to examine the physical significance of the terms in this equation (advection, dispersion, and reaction). We will start with the physical processes and then turn our attention to the chemical processes.
The one-dimensional form of the advection-dispersion equation for nonreactive dissolved constituents in saturated, homogeneous, isotropic, materials under steady-state, uniform flow [Eq. (A10.11), Appendix X] is
where l is a curvilinear coordinate direction taken along the flowline, is the average linear groundwater velocity, Dl, is the coefficient of hydrodynamic dispersion in the longitudinal direction (i.e., along the flow path), and C is the solute concentration. The effects of chemical reactions, biological transformations, and radioactive decay are not included in this form of the transport equation.
The coefficient of hydrodynamic dispersion can be expressed in terms of two components,
where αl, is a characteristic property of the porous medium known as the dynamic dispersivity, or simply as dispersivity [L], and D* is the coefficient of molecular diffusion for the solute in the porous medium [L2/T]. The relation between D* and the coefficient of diffusion for the solute species in water is described in Section 3.4. Some authors have indicated that a more accurate form of the mechanical component of the dispersion coefficient is , where m is an empirically determined constant between 1 and 2. Laboratory studies indicate that for practical purposes m can generally be taken as unity for granular geologic materials.
The classical experiment shown in Figure 9.1(a) is one of the most direct ways of illustrating the physical meaning of the one-dimensional form of the advection dispersion equation. In this experiment, a nonreactive tracer at concentration C0 is continuously introduced into a steady-state flow regime at the upstream end of a column packed with a homogeneous granular medium. For illustrative purposes it is assumed that the tracer concentration in the column prior to the introduction of the tracer is zero. It is convenient to express the tracer concentration in the column as a relative concentration, defined as C/C0, where C is the concentration in the column or in the output.
The tracer input can therefore be represented as a step function, as shown in Figure 9.1(b). The concentration versus time relation of the column outflow, known as the breakthrough curve, is shown in Figure 9.1(c). If it is assumed that the tracer moves through the column with no mechanical dispersion or molecular diffusion, the tracer front will pass through as a plug and will exit from the column as a step function. This condition is shown as a vertical dashed line in Figure 9.1(c). In real situations, however, mechanical dispersion and molecular diffusion occur and the breakthrough curve spreads out causing the tracer to begin to appear in the outflow from the column (at time t1) before the arrival of water traveling at the velocity of if (time t2). This is represented in Figure 9.1(c).
Figure 9.1(d) shows instantaneous “pictures” of the dispersion interface inside the column at various times prior to breakthrough. The tracer front is spread out along the flow path. The spread of the profile increases with travel distance. The positions represented by points 1 and 2 in Figures 9.1(d) correspond to times t1 and t2 in Figure 9.1(c). Mechanical dispersion and molecular diffusion cause some of the tracer molecules to move faster than the average linear velocity of the water and some to move slower. The average linear velocity of the water in the column is determined by dividing the water input rate (Q) by nA, where A is the cross-sectional area of the column and n is the porosity [Eq. (2.82)].
The boundary conditions represented by the step-function input are described mathematically as
For these boundary conditions the solution to Eq. (9.3) for a saturated homogeneous porous medium is (Ogata, 1970)
where erfc represents the complementary error function, which is tabulated in Appendix V; l is the distance along the flow path; and is the average linear water velocity. For conditions in which the dispersivity of the porous medium is large or when l or t is large, the second term on the right-hand side of the equation is negligible. Equation (9.5) can be used to compute the shapes of the breakthrough curves and concentration profiles illustrated in Figure 9.1(c) and (d). Analytical solutions for Eq. (9.3) with other boundary conditions are described by Rifai et al. (1956), Ebach and White (1958), Ogata and Banks (1961), Ogata (1970), and others.
The spreading out of the concentration profile and breakthrough curve of tracers or contaminants migrating through porous materials is caused by both mechanical dispersion and molecular diffusion. Figure 9.2 shows a concentration profile for the experimental conditions represented in Figure 9.1(a).
In this graph the contribution of molecular diffusion to the spread of the curves is indicated schematically. At a low velocity, diffusion is the important contributor to the dispersion, and therefore the coefficient of hydrodynamic dispersion equals the diffusion coefficient (Dl = D*), At a high velocity, mechanical mixing is the dominant dispersive process, in which case . Larger dispersivity of the medium produces greater mixing of the solute front as it advances. Laboratory experiments on tracer migration in saturated homogeneous granular materials have established relations between the influence of diffusion and mechanical dispersion, as illustrated in Figure 9.3.
The dimensionless parameter is known as the Peclet number, where the average particle diameter is denoted by d. The exact shape of the relation between the Peclet number and Dl = D* depends on the nature of the porous medium and on the fluid used in the experiments. The general shape illustrated in Figure 9.3 has been established by various investigators on the basis of experiments using different media (Bear, 1972).
In situations where the boundary conditions specified for Eq. (9.5) are applicable and where the groundwater velocity is so small that mechanical dispersion is negligible relative to molecular diffusion, Eq. (9.5) reduces to the one-dimensional solution to Fick’s second law. This “law” is described in Section 3.4. The rate at which one-dimensional diffusion occurs is expressed graphically in Figure 9.4, which shows, for periods of diffusion of 100 and 10,000 years, diffusion distances as a function of relative concentration.
The diffusion distances were obtained using Eq. (3.47) with diffusion coefficient values of 1 × 10–10 and 1 × 10–11 m2/s. These values are representative of a range typical of nonreactive chemical species in clayey geologic deposits. Values for coarse-grained unconsolidated materials can be somewhat higher than 1 × 10–10 m2/s but are less than the coefficients for the chemical species in water (i.e., < 2 × 10–9 m2/s). Figure 9.4 indicates that over long periods of time, diffusion can cause contaminants to move considerable distances, even through low-permeability materials. Whether contaminant migration on this time scale is important depends on the nature of the problem. In the case of subsurface disposal of radioactive wastes or highly toxic inorganic or organic compounds, diffusion can be an important process.
One of the characteristic features of the dispersive process is that it causes spreading of the solute, if the opportunity is available, in directions transverse to the flow path as well as in the longitudinal flow direction. This is illustrated schematically for a two-dimensional horizontal flow field in Figure 9.5(a). In this experimental sand box, a nonreactive tracer is introduced as a continuous steady input to the uniform flow field. Dispersion in this two-dimensional flow domain is illustrated in a different manner by the experiment shown in Figure 9.5(b). In this case the tracer is introduced as an instantaneous point source (i.e., a slug of tracer) into the uniform flow regime. As the tracer is transported along the flow path, it spreads in all directions in the horizontal plane. The total mass of the tracer in the flow regime does not change, but the mass occupies an increasing volume of the porous medium. The process of mechanical dispersion is directionally dependent even though the porous medium is isotropic with respect to textural properties and hydraulic conductivity. Figure 9.5(b) shows that the tracer zone develops an elliptical shape as the tracer is transported through the system. This occurs because the process of mechanical dispersion is anisotropic. Dispersion is stronger in the direction of flow (the longitudinal dispersion) than in direction normal to the flow line (transverse dispersion).
One-dimensional expressions for the transport of dissolved constituents, such as Eq. (9.5), are useful in the interpretation of laboratory column experiments, but are of limited use in the analysis of field problems because dispersion occurs in the transverse directions as well as in the longitudinal direction. As an example of a solution to the advection-dispersion equation in three-dimensions [Eq. (A10.9), Appendix X], we will follow an approach described by Baetslé (1969). As in Figure 9.5(b), the contaminant is assumed to originate as an instantaneous slug at a point source at x = 0, y = 0, z = 0. The mass of contaminant is then carried away from the source by transport in a steady-state uniform flow held moving in the x-direction in a homogeneous isotropic medium. As the contaminant mass is transported through the flow system, the concentration distribution of the contaminant mass at time t is given by
where M is the mass of contaminant introduced at the point source, Dx, Dy, and Dz are the coefficients of dispersion in the x, y, z directions and X, Y, and Z are distances in the x, y, z directions from the center of gravity of the contaminant mass. The position of the center of gravity of the contaminant mass at time t will lie along the flow path in the x direction at coordinates (xt, yt, zt), where yt = zt = 0 and , where is the average linear velocity, v is the specific discharge, and n is the porosity. In Eq. (9.6), , Y = y, and Z = z. It is apparent from Eq. (9.6) that the maximum concentration is located at the center of gravity of the contaminant cloud, where X = 0, Y = 0, and Z = 0. The mass of the contaminant introduced at the source equals C0V0, Where C0 is the initial concentration and V0 is the initial volume. In the mathematical formulation of the initial conditions, the contaminant input occurs at a point and therefore has mass but no volume. In practice, however, this is expressed by the quantity C0V0.
From Eq. (9.6) it follows that the peak concentration that occurs at the center of gravity of the contaminant cloud is given by
The zone in which 99.7% of the contaminant mass occurs is described by the ellipsoid with dimensions, measured from the center of mass, of , , , where σ is the standard deviation of the concentration distribution. This is illustrated in the xy plane in Figure 9.5(b). At low velocities molecular diffusion is the dominant dispersive mechanism, in which case the migrating contaminant cloud is circular. Because these equations are based on idealized conditions, such as the instantaneous point source and uniform flow, they have limited use in the analysis of most field situations. In simple hydrogeologic settings, however, they can be used to obtain preliminary estimates of the migration patterns that may arise from small contaminant spills or from leaching of buried wastes (Baetslé, 1969). A variety of other analytical solutions describing the migration of contaminants in two- and three-dimensional space are described by Fried (1975) and Codell and Schreiber (in press).
Mechanical dispersion in the transverse direction is a much weaker process than dispersion in the longitudinal direction, but at low velocities where molecular diffusion is the dominant dispersive mechanism, the coefficients of longitudinal and transverse dispersion are nearly equal. This is illustrated by the experimental results shown in Figure 9.6, which indicates small dispersion coefficients over a range of low velocities. Because mechanical dispersion in the transverse direction is much weaker than in the longitudinal direction, the transverse dispersion coefficient remains diffusion-controlled until the flow velocity is quite high.
The forms of the transport equation described above are based on the assumption that there is no significant density contrast between the contaminant or tracer fluid and the groundwater in the surrounding flow domain. Equations that make allowance for density contrasts are more complex. As a qualitative example of the effect of density contrasts, consider the sinking contaminant plume in an initially uniform flow field, as illustrated in Figure 9.7.
If the contaminant solution entering this flow regime has the same density as the groundwater, the contaminant plume will spread in a shallow zone close to the water table. If the contaminant solution is considerably more dense than the groundwater, the plume will sink steeply downward into the groundwater flow system. Prediction of contaminant migration patterns requires accurate knowledge of the density of the contaminant solution as well as that of the groundwater.
If it were not for the effects of heterogeneities in natural geological materials, the problem of prediction and detection of contaminant behavior in groundwater flow systems would be easily solved. Advection is the process whereby solutes are transported by the bulk mass of the flowing fluid. Advection is normally considered on the macroscopic scale in terms of the patterns of groundwater flow. These patterns are defined by the spatial and temporal distributions of the average linear velocity of the fluid. Flow patterns and flow nets have been described extensively in Chapters 5 and 6. Our purpose here is to consider in more detail the effects on flow lines and velocities exerted by various types of heterogeneities.
To illustrate the effect of simple layered heterogeneities on transport patterns, the cross-section flow-domain illustrated in Figure 9.8(a) is used. It is assumed that steady-state groundwater flow occurs through the cross section and that the flow domain is isotropic with respect to hydraulic conductivity.
To illustrate the effect of stratigraphic variations on the transport pattern of contaminants entering the system in a recharge area, a contaminant input zone is located on the cross section. In field situations this could represent seepage from a waste lagoon, sanitary landfill, or some other surface or near-surface source. Figure 9.8(b), (c), (d), and (e) shows the patterns of contaminant transport that would occur with various hypothetical stratigraphic configurations. The contaminant is assumed to be nonreactive and the effect of dispersion is neglected. The flow lines that depict the limits of the contaminant migration patterns were obtained by solution of the two-dimensional form of the steady-state groundwater flow equation [Eq. (2.69)], using the finite-element method in the manner described by Pickens and Lennox (1976). Figure 9.8(b) indicates that in situations where the How domain is homogeneous, the contaminant migration pattern would be simple and relatively easy to monitor. The conditions for the flow system shown in Figure 9.8(c) are similar to the previous case, with the exception of the inclusion of a thin, higher-conductivity horizontal layer that extends across the How domain. This would cause the contaminants to move through the how system almost entirely in this thin layer. The total travel time would be one-fifth of the nonstratified case illustrated in Figure 9.8(b). The thin higher conductivity bed has a conductivity 100 times larger than the rest of the system and exerts a very strong influence on the migration patterns and velocity distribution. If the lower-K medium (K1) represents a very fine-grained sand, the higher-K bed (K2) could represent a medium- or coarse-grained sand. In stratigraphic studies of waste disposal sites, a thin medium-grained sand bed in an otherwise fine-grained sand deposit could easily be unnoticed unless careful drilling and sampling techniques are used.
In Figure 9.8(d) a discontinuous layer of low-conductivity material exists in the cross section. The contaminant migration zone moves over the first lense and under the second one. To reach the discharge area, it passes through the second lense near the end of its flow path.
Figure 9.8(e) shows the contaminant migration pattern that would exist if a thin higher-conductivity bed is discontinuous through the central part of the cross section. The discontinuity causes a large distortion in the contaminant migration pattern in the middle of the cross section. The contaminated zone spreads out in the central part of the flow system and extends to the water-table zone. In situations where contaminants can be transferred through the unsaturated zone by advection, diffusion, or vegetative uptake, this condition could lead to spread of the contaminants in the biosphere. Figure 9.8(e) also illustrates some of the difficulties that can arise in monitoring contaminated flow systems. If little information were available on the stratigraphy of the system, there would be no reason to suspect that the type of distortions shown in Figure 9.8(e) would occur. Lack of this information could result in inadequate monitoring of the system. In nature, geologic cross sections typically include many stratigraphic units with different hydraulic conductivities. Large conductivity contrasts across sharp discontinuities are common. Relative to real situations, the effects of stratification illustrated in Figure 9.8 are very simple.
In the discussion above, layered heterogeneities on the scale that could, if necessary, be identified and mapped by careful drilling, sampling, and geophysical logging were considered. Heterogeneities in another category also exist in most geologic settings. These are known as small-scale heterogeneities. They cannot be identified individually by conventional methods of field testing. Even if identification is possible using special coring techniques, these heterogeneities usually cannot be correlated from borehole to borehole. In granular aquifers, heterogeneities of this type are ubiquitous. Hydraulic conductivity contrasts as large as an order of magnitude or more can occur as a result of almost unrecognizable variations in grain-size characteristics. For example, a change of silt or clay content of only a few percent in a sandy zone can have a large effect on the hydraulic conductivity.
Figure 9.9 illustrates the effect of two types of small-scale heterogeneities on the pattern of migration of a tracer or contaminant in granular porous media. In Figure 9.9(a) the pattern of dispersion is regular and predictable using the methods described above. In Figure 9.9(b) the lense-type heterogeneities cause the tracer front to advance in a pattern commonly referred to as fingering. In this case the contaminant is transported more rapidly in the lenses or beds of higher hydraulic conductivity. Figure 9.9(c) illustrates results obtained by Skibitzke and Robertson (1963) using dye tracers in a box model packed with fine sand and long sinuous lenses of coarser sand. These authors observed that a large angle of refraction at the boundary between sand of contrasting permeability caused accelerated spreading of the tracer zone.
In one of the very few detailed three-dimensional studies of contaminant movement in sandy deposits, Childs et al. (1974) observed that “plumes migrate along zones … that, although they are texturally similar, show subtle differences in fabric that result in slight variations in permeability. Bifurcations indicate that detection of a shallow plume does not negate the existence of the other plumes of the same constituent at depth” (p. 369).
Nearly all studies of dispersion reported in the literature have involved relatively homogeneous sandy materials under controlled conditions in the laboratory. These studies have indicated that the dispersivity of these materials is small. Values of longitudinal dispersivity are typically in the range of 0.1 to 10 mm, with transverse dispersivity values normally lower by a factor of 5–20. Whether or not these values are at all indicative of dispersivities in field systems is subject to considerable controversy at the present time. Many investigators have concluded that values of longitudinal and transverse dispersivities in held systems are significantly larger than values obtained in laboratory experiments on homogeneous materials or on materials with simple heterogeneities. Values of longitudinal dispersivity as large as 100 m and lateral dispersivity values as large as 50 m have been used in mathematical simulation studies of the migration of large contaminant plumes in sandy aquifers (Pinder, 1973; Konikow and Bredehoeft, 1974; Robertson, 1974).
To illustrate the effect of large dispersivities on the migration of contaminants in a hypothetical groundwater flow system, a cross-sectional flow domain similar to that shown in Figure 9.8(a) and (b) will be used. Figure 9.10 shows the effect of dispersivity on the spreading of a contaminant plume that emanates from a source in the recharge area of the flow system. Although the cross sections shown in Figure 9.10 are homogeneous, dispersivities for the system are assumed to be large as a result of small-scale heterogeneities. With assigned values of dispersivity the patterns of contaminant distribution can be simulated using a finite-element approximation to the transport equation expressed in two-dimensional form for saturated heterogeneous isotropic media [Eq. (A10.13), Appendix X]:
where sl and st, are the directions of the groundwater flowlines and the normals to these lines, respectively. The finite-element model used to obtain the contaminant distributions shown in Figure 9.10 is described by Pickens and Lennox (1976).
Other numerical models have been developed by Reddell and Sunada (1970), Bredehoeft and Pinder (197 3), Pinder (1973), and Schwartz (1975). The simulations presented in Figure 9.10 indicate that if dispersivity is large, contaminants can spread to occupy a portion of the flow system many times larger than would be the case in the presence of advection alone. If the transverse dispersivity is very large as indicated in Figure 9.10, contaminants transported along relatively horizontal flow paths can migrate deep into the how system. The longitudinal and transverse dispersivities represented in the simulated contaminant transport patterns shown in Figure 9.10 indicate that if dispersivity values are orders of magnitude larger than the values obtained from laboratory experiments, dispersion will exert a strong influence on contaminant transport. Whether or not dispersivities in nonfractured geologic materials under field conditions have magnitudes that are this large remains to be established by detailed field experiments. This topic is discussed further in Section 9.4.
In this section we will consider the transport of solutes that behave as those described above, but with the added influence of chemical reactions. Changes in concentration can occur because of chemical reactions that take place entirely within the aqueous phase or because of the transfer of the solute to or from other phases such as the solid matrix of the porous medium or the gas phase in the unsaturated zone. The myriad of chemical and biochemical reactions that can alter contaminant concentrations in groundwater flow systems can be grouped in six categories: adsorption-desorption reactions, acid-base reactions, solution-precipitation reactions, oxidation-reduction reactions, ion pairing or complexation, and microbial cell synthesis. Radioactive contaminants are influenced by radioactive decay in addition to the nonradiogenic processes. In the following discussion we will focus on adsorption as a concentration-altering mechanism. In Section 9.3 other types of reactions are considered.
For homogeneous saturated media with steady-state flow, the one-dimensional form of the advection-dispersion equation expressed in a manner that includes the influence of adsorption [Eq. (A10.14), Appendix X] is
where ρb is the bulk mass density of the porous medium, n is the porosity, and S is the mass of the chemical constituent adsorbed on the solid part of the porous medium per unit mass of solids. ∂S/∂t represents the rate at which the constituent is adsorbed [M/MT], and (ρb/n)(∂S/∂t) represents the change in concentration in the fluid caused by adsorption or desorption
Adsorption reactions for contaminants in groundwater are normally viewed as being very rapid relative to the flow velocity. The amount of the contaminant that is adsorbed by the solids (i.e., the degree of adsorption) is commonly a function of the concentration in solution, S = f(C). It follows that
In which the term (∂S/∂C) represents the partitioning of the contaminant between the solution and the solids.
The partitioning of solutes between liquid and solid phases in a porous medium as determined by laboratory experiments is commonly expressed in two-ordinate graphical form where mass adsorbed per unit mass of dry solids is plotted against the concentration of the constituent in solution. These graphical relations of S versus C and their equivalent mathematical expressions are known as isotherms. This term derives from the fact that adsorption experiments are normally conducted at constant temperature.
Results of adsorption experiments are commonly plotted on double-logarithmic graph paper. For solute species at low or moderate concentrations, straight-line graphical relations are commonly obtained over large ranges of concentration. This condition can be expressed as
where S is the mass of the solute species adsorbed or precipitated on the solids per unit bulk dry mass of the porous medium, C is the solute concentration, and Kd, and b are coefficients that depend on the solute species, nature of the porous medium, and other conditions of the system. Equation (9.12) is known as the Freundlich isotherm. The slope of the log-log adsorption relation is represented by the term b in Eq. (9.12). If b = 1 (i.e., if the straight-line relationship between S and C on a log-log plot has a slope of 45°), then the S versus C data will also plot as a straight line on an arithmetic plot. Such an isotherm is termed linear, and from Eq. (9.12) with b = 1,
where Kd is known as the distribution coefficient. This parameter is widely used in studies of groundwater contamination. Kd is a valid representation of the partitioning between liquid and solids only if the reactions that cause the partitioning are fast and reversible and only if the isotherm is linear. Fortunately, many contaminants of interest in groundwater studies meet these requirements. A comprehensive treatment of adsorption isotherms is presented by Helfferich (1962), who provides detailed information on many important types of isotherms in addition to the Freundlich isotherm.
The transfer by adsorption or other chemical processes of contaminant mass from the pore water to the solid part of the porous medium, while flow occurs, causes the advance rate of the contaminant front to be retarded. To illustrate this concept, the classical column experiment shown in Figure 9.1(a) will again be considered. It is assumed that two tracers are added to the water passing through the column. One tracer is not adsorbed and therefore moves with the water. The other tracer undergoes adsorption, and as it travels through the column part of its mass is taken up by the porous medium. The two tracers are added instantaneously to the water at the column input [step-function input as shown in Figure 9.1(b)]. As transport occurs, the two tracers are distributed in the column in the manner represented schematically in Figure 9.11. The transporting water mass represented by the nonreactive tracer moves ahead of the reactive tracer. The concentration profile for the nonadsorbed tracer spreads out as a result of dispersion. The concentration profile of the front of the reactive tracer also spreads out but travels behind the front of the nonreactive tracer. The adsorbed tracer is therefore said to be retarded.
For cases where the partitioning of the contaminant can be adequately described by the distribution coefficient (i.e., fast reversible adsorption, with linear isotherm), the retardation of the front relative to the bulk mass of water is described by the relation
where is the average linear velocity of the groundwater and is the velocity of the C/C0 = 0.5 point on the concentration profile of the retarded constituent. Equation (9.14) is commonly known as the retardation equation. The term is referred to as the retardation factor. The reciprocal of the retardation factor is known as the relative velocity . Equation (9.14) was originally developed on an empirical basis for use in chemical engineering by Vermeulen and Hiester (1952). It was first applied to groundwater problems by Higgins (1959) and Baetslé (1967, 1969). Baetslé indicated that it can be used to determine the retardation of the center of mass of a contaminant moving from a point source while undergoing adsorption.
To gain a more quantitative appreciation for the effects of chemical retardation on contaminant migration, some representative parameter values will be used in conjunction with Eq. (9.14). For unconsolidated granular deposits, porosity, expressed as a fraction, is commonly in the range 0.2–0.4. The average mass density of minerals that constitute unconsolidated deposits is approximately 2.65. The range of bulk mass densities, ρb that correspond to the porosity range above is 1.6–2.1 g/cm3. For these ranges of porosity and bulk mass density, ρb/n values range from 4 to 10 g/cm3. An approximation to Eq. (9.14) is therefore
The only major unknown in Eq. (9.15) is the distribution coefficient Kd. The distribution coefficient can be expressed as
The dimensions for this expression reduce to L3/M. Measured Kd values are normally reported as milliliters per gram (mℓ/g).
Distribution coefficients for reactive solutes range from values near zero to 103 mℓ/g or greater. From Eq. (9.15) it is apparent that if Kd = 1 mℓ/g, the midconcentration point of the solute would be retarded relative to the bulk groundwater flow by a factor between 5 and 11. For Kd, values that are orders of magnitude larger than 1, the solute is essentially immobile.
To further illustrate the effect of liquid- to solid-phase partitioning, a cross-sectional flow domain similar to the one represented in Figures 9.8 and 9.10 is used. The pattern of contamination in this cross section caused by an influx of water with contaminant species of different distribution coefficients is shown in Figure 9.12. The patterns were obtained by Pickens and Lennox (1976) using a finite-element solution to the transport equation with the reaction term described by Eq. (9.11). The case in which Kd = 0 shows the zone occupied by a contaminant species that is not affected by chemical reactions. Under this condition the processes of advection and dispersion cause the contaminant to gradually occupy a large part of the flow domain. The transport pattern is controlled by the contaminant input history, by the velocity distribution, and by dispersion. Contaminant species with Kd, values greater than zero occupy a much smaller portion of the flow domain. If Kd = 10 mℓ/g, most of the contaminant mass migrates only a very short distance from the input zone during the specified migration period. This situation can be anticipated from consideration of the magnitude of this Kd value in Eq. (9.15). There is an extensive zone beyond the C/C0 = 0.1 contours shown in Figure 9.12 in which the contaminant occurs at very low concentrations. If the contaminant is harmful at low concentrations, this zone can be extremely important, even though it includes only a small portion of the total contaminant mass in the flow system.
When a mixture of reactive contaminants enters the groundwater zone, each species will travel at a rate depending on its relative velocity, . After a given time t, the original contaminant cloud will have segregated into different zones, each advancing in the same direction at different velocities. Considering the instantaneous point-source example described by Eqs. (9.6) and (9.7), the position of the center of mass of the migrating cloud is obtained from the relative velocity defined by the reciprocal of calculated from Eq. (9.14). Equation (9.6) can be used to calculate the concentration distribution of the dissolved reactive species, with substitution of τ for t, where . Since the total standard deviation of a given distribution is a function of time as well as distance traveled, both parameters influence the dispersion pattern of each retarded species (Baetslé, 1969).
The distribution coefficient approach to the representation of chemical partitioning of contaminants in groundwater flow systems is based on the assumption that the reactions that partition the contaminants between the liquid and solid phases are completely reversible. As a contaminant plume advances along flow paths, the front is retarded as a result of transfer of part of the contaminant mass to the solid phase. If the input of contaminant mass to the system is discontinued, contaminants will be transferred back to the liquid phase as lower-concentration water flushes through the previously contaminated zone. In this situation the contaminant moves as a cloud or enclave through the flow system. This is illustrated in Figure 9.13, which shows the migration of a contaminant enclave through the cross section illustrated in Figures 9.10 and 9.12.
Initially, the contaminated zone is localized beneath the input area. After the input of contaminated water is discontinued, the contaminant mass moves along the flow paths, leaving a zone of less« contaminated water beneath the input area. As time goes on, the contaminants are flushed out of the flow system. If the partitioning reactions are completely reversible, all evidence of contamination is eventually removed from the system as complete desorption occurs. Thus, if the reactions are reversible, contaminants cannot be permanently isolated in the subsurface zone, even though retardation of the concentration front may be strong. In some situations a portion of the contaminant mass transferred to the solid part of the porous material by adsorption or precipitation is irreversibly fixed relative to the time scale of interest. This portion is not transferred back to the pore water as new water passes through the system and is therefore isolated in the subsurface environment.
When the distribution coefficient is used to determine contaminant retardation, it is assumed that the partitioning reactions are very fast relative to the rate of groundwater movement. Many substances, however, do not react sufficiently fast with the porous medium for this assumption to be valid. When contaminants of this type move through porous media, they advance more rapidly than would be the case if the reactions produce Kd type partitioning relations. This is illustrated in Figure 9.14, which shows the nonequilibrium front in a position between the front of a nonretarded tracer and the front of a retarded tracer described by the Kd relation. Analysis of the movement of contaminants that undergo partitioning in a manner that cannot be described by equilibrium relations requires information on the rates of reaction between the contaminant and the porous medium. This information is difficult to obtain. In field studies the retardation equation described above is often used because of its simplicity or because there is a lack of information on reaction rates. This can lead to serious errors in prediction of rates of contaminant migration in systems where kinetic factors are important.
Although contaminant transport in fractured geologic materials is governed by the same processes as in granular media-namely, advection, mechanical dispersion, molecular diffusion, and chemical reactions—the effects in fractured media can be quite different. The effective fracture porosity of fractured rocks and of consolidated cohesive materials that are fractured, such as jointed till, silt, or clay, is normally very small. Values in the order of 1–0.001%, or 10–2–10–5 expressed as a fraction, are not unusual. Although the porosities are small, the groundwater velocities can be large. The reason for this can be deduced from the modified Darcy relation (Section 2.12)
where is the average linear velocity of water in the fractures, K the bulk hydraulic conductivity of the fractured medium, nf the bulk fracture porosity, and dh/dl the hydraulic gradient. This relation treats the fractured medium as an equivalent porous medium. The parameters in the equations relate to a volume segment of the medium that is sufficiently large to be described by hydraulic conductivity and porosity averaged over the bulk mass. In this approach each fracture opening is considered to be very small relative to the bulk volume of the domain over which K is measured. The number of fractures in this domain therefore must be large.
For illustrative purposes we will consider a medium that has a bulk hydraulic conductivity of 10–8 m/s and a fracture porosity of 10–4. These values could represent conditions in a slightly fractured granite. Using a hydraulic gradient of 10–2, which is within the range commonly observed in field situations, the groundwater velocity computed from Eq. (9.16) is 10 m/yr. Compared to velocities in fine-grained unfractured granular materials, this velocity is very large. For example, an unfractured granular medium, such as a silt deposit, with this hydraulic conductivity and gradient and an intergranular porosity of 0.3 would have a groundwater velocity of about 0.003 m/yr. The flux of water (volume of water per unit time passing through a specified cross-sectional area) in these two cases is the same and is extremely small. Although Eq. (9.16) can be used to compute average velocities in fractured media, it provides no indication of the velocities in individual fractures. Depending on the fracture aperture and wall roughness, the velocity of groundwater may deviate from the average by orders of magnitude.
It was indicated above that in the mathematical analysis of mechanical dispersion in granular media, the media are assumed to be isotropic with respect to dispersivity. That is, longitudinal dispersivity at a point in the medium has a single value regardless of the direction of the velocity vector. Each of the transverse dispersivities has a single value relative to the longitudinal dispersivity. The differences between longitudinal and transverse dispersivities are related to the mechanism of dispersion rather than to directional properties of the medium. Fractured geologic materials, however, are notoriously anisotropic with respect to the orientation and frequency of fractures. It can be expected that the dispersion of solutes during transport through many types of fractured rocks cannot be described by the equations developed for homogeneous granular materials. Little is known about dispersion in fractured media. A common approach in field investigations of contaminant migration in fractured rock is to treat the problem in the same manner mathematically as for granular porous media. The scale at which this approach becomes valid in the analysis of field situations is not known. As a concluding comment on this topic, the statement by Castillo et al. (1972) is appropriate:
Although the basic theoretical aspects of . . . (dispersion) . . . have been treated at length for the case where the permeable stratum is composed of granular materials, the classical concept of flow through a porous medium is generally inadequate to describe the flow behaviour in jointed rock, and it becomes increasingly unsuitable for the analysis of dispersion. Despite these limitations, little work has been directed toward extending these ideas to handle flow through jointed rock formations (p. 778).
A modification in approach is necessary for the distribution coefficient or isotherm concept to be applicable in the analysis of the migration of reactive contaminants through fractured media. For granular materials the amount of solute adsorbed on the solid part of the porous media is expressed per unit mass of the bulk medium in a dry state. For convenience the unit mass of the porous medium is used as a reference quantity. A more mechanistic but less convenient approach would be to use a unit surface area of the porous medium as the reference quantity. This would be a reasonable approach because adsorption reactions are much more closely related to the surface area of the solid medium than to the mass of the medium. Nevertheless, for granular materials such as sands, silts, and clays, the use of mass density in the definition of the distribution coefficient normally produces acceptable results. With this approach, measurements of effective surface area are not necessary.
In the case of contaminant migration through fractured materials, it is more appropriate, as suggested by Burkholder (1976), to express the distribution coefficient Ka on a per-unit-surface-area basis.
It is therefore defined as
The dimensions for this expression are [M/L2 · L3/M] or [L]. The units that are commonly used are milliliters per square centimeter.
The retardation equation therefore becomes
where A is the surface area to void-space (volume) ratio [1/L] for the fracture opening through which the solute is being transported. It is apparent from this relation that fractures with smaller apertures produce greater retardation of reactive solutes. The distribution coefficient in this retardation expression has the same inherent assumptions as Eq. (9.14), namely: the partitioning reactions are reversible and fast relative to the flow velocity.
Equation (9.17) is simple in conceptual terms, but it is difficult to apply to natural systems. If information can be obtained on the aperture of a fracture and f the fracture surface is assumed to be planar, A = 2/b, where b is the aperture width (Section 2.12). Fracture surfaces usually have small-scale irregularities and therefore can have much larger surface areas than planar surfaces. In the determination of the adsorption isotherm or distribution coefficient for the fracture, he partitioning of the contaminant between fluid in contact with the fracture and he fracture surface is measured. If the fracture surface is irregular or contains coating of weathered material or chemical precipitates, the actual surface area with which the contaminant reacts is unknown. Without an elaborate experimental effort, it is indeterminant. A practical approach is to express the Ka, relative to the area of an assumed planar fracture surface, in which case the retardation relation becomes
It should be kept in mind that Eq. (9.17) is only valid for fractured materials in which the porosity of the solid mass between fractures is insignificant. When contaminants occur in fractures, there is a gradient of contaminant concentration between the fracture fluid and the fluid in the unfractured material adjacent to the fracture. If the solid matrix is porous, a portion of the contaminant mass will move by molecular diffusion from the fracture into the matrix. This mass is therefore removed, at least temporarily, from the how regime in the open fracture.
Figure 9.15 illustrates the effect of matrix diffusion on the concentration distribution of nonreactive and reactive contaminants migrating through a fracture in a medium with a porous matrix. For illustrative purposes it is assumed that dispersion within the fracture is insignificant. Comparison of Figure 9.15(a) and (b) indicates that diffusion into the matrix causes the concentration in the fracture to diminish gradually toward the front of the advancing contaminant zone. The bulk mass of the advancing contaminant zone in the fracture appears to be retarded because part of the contaminant mass is transferred to the matrix.
The general shape of the longitudinal profile is somewhat similar to that produced by longitudinal dispersion in granular materials. If the contaminant undergoes adsorption, the effect of diffusion is to cause adsorption to occur on a much larger surface area than would be the case if the contaminant mass remained entirely within the fracture. A portion of the contaminant is adsorbed on the surface of the fracture and as diffusion occurs a portion is adsorbed in the matrix. The combined effect of adsorption on the fracture surface and adsorption in the matrix is to cause the contaminant mass in the fracture to be retarded relative to the advance that would occur in the absence of adsorption [Figure 9.15(c)].
The contaminant distribution in a porous fractured aquifer receiving waste from a surface source is illustrated schematically in Figure 9.16. As time goes on, the zone of contamination will diffuse farther into the porous matrix. If the source of contamination is discontinued, the contaminant mass in the porous matrix will eventually diffuse back to the fracture openings as fresh water flushes through the fracture network.
Molecular diffusion is a process that occurs at a sufficiently rapid rate to exert a strong influence on contaminant behavior in many types of fractured materials. Even granite has appreciable primary porosity and permeability, with porosity values commonly as large as 0.05–1.0% and hydraulic conductivity in the order of 10–12 m/s. In the main limestone aquifer in Britain, detailed studies show that subsurface distributions of tritium and nitrate in the limestone are strongly influenced by diffusion of these constituents from the fractures, where rapid flow occurs, into the porous rock matrix (Foster, 1975). In the Plains Region of North America, deposits of glacial till and glaciolacustrine clay are commonly fractured (Section 4.4). Grisak et al. (1976) indicate that although the fractures are generally a major avenue of groundwater flow, the chemical evolution of groundwater is controlled by diffusion of dissolved reaction products from the clayey matrix into the fracture network.
9.3 Hydrochemical Behavior of Contaminants
In this section, the hydrochemical behavior of groundwater contaminants will be discussed. It is not feasible for all the hydrochemical processes that affect contaminants in groundwater to be considered in this text. Our purpose is to illustrate some of the most important processes that control the behavior of several groups of contaminants with different hydrochemical properties. The origin and causes of groundwater contamination are discussed in Section 9.4.
The most common contaminant identified in groundwater is dissolved nitrogen in the form of nitrate (). This contaminant is becoming increasingly widespread because of agricultural activities and disposal of sewage on or beneath the land surface. Its presence in undesirable concentrations is threatening large aquifer systems in many parts of the world. Although is the main form in which nitrogen occurs in groundwater, dissolved nitrogen also occurs in the form of ammonium (), ammonia (NH3), nitrite (), nitrogen (N2), nitrous oxide (N2O), and organic nitrogen. Organic nitrogen is nitrogen that is incorporated in organic substances.
Nitrate in groundwater generally originates from nitrate sources on the land surface, in the soil zone, or in shallow subsoil zones where nitrogen-rich wastes are buried (Figure 9.17). In some situations that enters the groundwater system originates as ; in wastes or fertilizers applied to the land surface. These are designated as direct nitrate sources in Figure 9.18. In other cases, originates by conversion of organic nitrogen or , which occur naturally or are introduced to the soil zone by man’s activities. The processes of conversion of organic nitrogen to is known as ammonification. Through the process of nitrification, is converted to by oxidation. Ammonification and nitrification are processes that normally occur above the water table, generally in the toil zone, where organic matter and oxygen are abundant. Thus, in Figure 9.18 these processes are represented as producers outside the boundaries of the groundwater flow system.
Concentrations of in the range commonly reported for groundwater are not limited by solubility constraints. Because of this and because of its anionic form, is very mobile in groundwater. In groundwater that is strongly oxidizing, is the stable form of dissolved nitrogen. It moves with the groundwater with no transformation and little or no retardation. Very shallow groundwater in highly permeable sediment or fractured rock commonly contains considerable dissolved O2. It is in these hydrogeologic environments where commonly migrates large distances from input areas.
A decline in the redox potential of the groundwater can, in some situations, cause denitrification, a process in which is reduced to N2O or N2 (Figure 9.17). This process is represented chemically in Table 3.11. In an ideal system, which can be described by reversible thermodynamics, denitrification would occur at a redox potential of about 4.2 as pE (or +250 mv as Eh) in water at pH 7 and 25°C. At this redox potential, the water would be devoid of dissolved O2 (i.e., below the detection limit). The is reduced to N2O and then, if the redox potential declines further, the N2O is reduced to N2. These reaction products exist as dissolved species in the groundwater. If the water moves into the unsaturated zone, a portion of the N2O or N2 may be lost by off-gassing to the soil air (Figure 9.18).
Figure 9.18 indicates that in addition to the denitrification pathway for the reduction of NO3, there is a pathway that leads to . For biochemical reasons only a small fraction of the that undergoes reduction follows this reduction path. If is produced in groundwater by this process, most of it would eventually be adsorbed on clay or silt-sized particles in the geologic materials.
From a water-quality viewpoint, denitrification in groundwater is a desirable process. Increased concentrations of dissolved N2 and N2O are not detrimental to drinking water. In contrast, at concentrations above 45 mg/ℓ renders water unfit for consumption by human infants. If water has more than 450 mg/ℓ of , it is unsuitable for consumption by livestock.
Denitrification is a process that has been observed in numerous investigations of soil systems in the laboratory and in the field. Given a source of organic matter and abundant , bacterial systems in soil are capable of denitrifying large amounts of . Denitrification in the groundwater zone, however, is a process about which little is known. It appears that a lack of suitable types or amounts of organic matter in the groundwater zone commonly inhibit the growth of denitrifying bacteria in groundwater. This limits the rate of denitrification, even if the redox system has evolved toward reducing conditions. However, since groundwater commonly flows at low velocity, a slow rate of denitrification may nevertheless be significant with respect to the nitrate budget of the subsurface environment. For discussions of field situations in which evidence indicating denitrification in aquifers is presented, the reader is referred to Edmunds (1973) and Gillham and Cherry (1978).
In recent years the mobility of trace metals in groundwater has received considerable attention. Of special interest are the trace metals for which maximum permissible or recommended limits have been set in drinking water standards. These include Ag, Cd, Cr, Cu, Hg, Fe, Mn, and Zn (see Table 9.1). During the next decade this list may grow as more is learned about the role of trace metals in human health and ecology. Although these elements rarely occur in groundwater at concentrations large enough to comprise a significant percentage of the total dissolved solids, their concentrations can, depending on the source and hydrochemical environment, be above the limits specified in drinking water standards. Most of the elements listed above are in an elemental group referred to by chemists as the transition elements. Many of these elements are also known as heavy metals.
Trace metals in natural or contaminated groundwaters, with the exception of iron, almost invariably occur at concentrations well below 1 mg/ℓ. Concentrations are low because of constraints imposed by solubility of minerals or amorphous substances and adsorption on clay minerals or on hydrous oxides of iron and manganese or organic matter. Isomorphous substitution or coprecipitation with minerals or amorphous solids can also be important.
A characteristic feature of most trace metals in water is their tendency to form hydrolyzed species and to form complexed species by combining with inorganic anions such as , , , Cl–, F–, and . In groundwater environments contaminated with dissolved organic compounds, organic complexes may also be important. Expressed in terms of the products of hydrolysis, the total concentration of a trace metal Mτ that in the unhydrolyzed form exists as Mn+ is
If the total concentration, Mτ, is known, the concentrations of the other species can be computed using mass-action equations with equilibrium constants derived from thermodynamic data (Leckie and James, 1974). Using zinc as an example, the hydrolyzed species and inorganic complexes that would form would include ZnOH+, , Zn(OH)42-, ZnCl–, , and . The occurrence and mobility of zinc in groundwater requires consideration of these and other dissolved species. Chemical analyses of zinc in groundwater provide direct information only on the total zinc content of water. The percent of the total concentration existing as hydrolyzed species increases with increasing pH of the water. Complexes of zinc with Cl–, SO42–, and increase with increasing concentrations of these anions in solution. In Section 3.3, it was shown that dissolved species in groundwater resulting from the formation of complexes with major ions can be computed from analyses of total concentrations of major constituents. In much the same manner, the concentration of trace-metal complexes can be computed using cons centration data from laboratory analyses. Capability to predict the mobility of trace metals in groundwater can depend on the capability for prediction of the concentrations of the most important complexes formed by the element in the water. Although information on the free and complexed forms are often required for an understanding of the mobility of trace metals, the concentration values listed in water quality standards are total concentrations.
Nearly all the trace metals of interest in groundwater problems are influenced by redox conditions, as a result either of changes in the oxidation state of the trace metal or of nonmetallic elements with which it forms complexes. The redox environment may also indirectly influence trace-metal concentrations as a result of changes in solid phases in the porous medium that cause adsorption of the trace metal. In the following discussion, mercury is used for illustration of the influence of redox conditions and complexing. Diagrams of pE-pH for Hg, in water that contains Cl– and dissolved sulfur species, are shown in Figure 9.19. Figure 9.19(a) indicates the main solid compounds of mercury that occur in the various pH-pE stability fields and the field in which liquid mercury occurs.
The dominant aqueous Hg species in equilibrium with these solid phases containing appreciable concentrations of SO42– and Cl– are shown in Figure 9.19(b). In high-Cl– water, is the dominant dissolved species of Hg in the normal pH range of groundwater under oxidizing conditions. At low Cl– concentrations, HgO is the equilibrium solid phase and is the dominant dissolved species at high redox potential. The main equilibrium reaction in this pH-pE environment is
At 25°C, log K for this reaction is –3.7. The equilibrium concentration from this reaction is therefore 47 mg/ℓ. This concentration is 4 orders of magnitude above the maximum permissible level for drinking water. In most of the pH-pE domain below the HgO(s) stability field, solubility constraints produce equilibrium concentrations of total dissolved mercury considerably below this level. In much of the redox domain, the equilibrium concentrations are below the maximum levels permitted in drinking water.
Some of the other trace metals also have large equilibrium concentrations in waters with high redox potential. In anaerobic groundwaters, the relative insolubility of sulfide minerals can limit trace metals to extremely low concentrations. In nonacidic groundwaters with high concentrations of dissolved inorganic carbon, solubility of carbonate materials will, if equilibrium is achieved, maintain concentrations of metals such as cadmium, lead, and iron at very low levels. This is the case provided that excessive amounts of inorganic or organic complexing substances are not present in the water.
In addition to the constraints exerted by solubilities of solid substances and the effects thereon caused by the formation of dissolved complexes, the occurrence and mobility of trace metals in groundwater environments can be strongly influenced by adsorption processes. In some groundwaters, many of the trace metals are maintained by adsorption at concentrations far below those that would exist as a result only of solubility constraints. Trace-metal adsorption in subsurface systems occurs because of the presence of clay minerals, organic matter, and the other crystalline and amorphous substances that make up the porous media. In some geologic materials trace-metal adsorption is controlled by crystalline or amorphous substances that are present in only small quantities. For example, Jenne (1968) indicates that hydrous oxides of Fe and Mn furnish the principal control on the fixation of Co, Ni, Cu, and Zn in soils and freshwater sediments. In oxidizing environments, these oxides occur as coatings on grains and can enhance the adsorptive capability of the medium far out of proportion to their percent occurrence relative to the other solids. The hydrous oxide coatings can act as scavengers with respect to trace metals and other toxic constituents.
Hydrous iron and manganese oxide precipitates are usually denoted as Fe(OH)3(s) and MnO2(s). FeOOH(s) is sometimes used to designate the iron oxide precipitates. The oxides of iron and manganese may be X-ray amorphous (i.e., noncrystalline) or crystalline. In crystalline form hydrous iron oxide is known as the mineral goethite, or if the composition is Fe2O3, as hematite. Hydrous iron oxide precipitates are generally mixtures of different phases. Crystalline forms such as goethite and hematite form as a result of long-term aging of amorphous precipitates (Langmuir and Whittemore, 1971).
A pH-Eh diagram for iron in water that contains dissolved inorganic carbon and dissolved sulfur species is shown in Figure 9.20. Within the pH range typical of groundwater, Fe(OH)3(s) is thermodynamically stable at moderate to high pE values. In groundwaters with appreciable, dissolved inorganic carbon and sulfur, FeCO3(s) (siderite) and FeS2 (pyrite or marcasite), are stable at lower pE values. In Figure 9.20 the boundaries of the Fe(OH)3(s) field have considerable uncertainty because of uncertainty in free-energy data for Fe(OH)3(s). Nevertheless, the pH-pE diagram serves to illustrate that the existence of Fe(OH)3(s) is dependent on the redox conditions.
From this it follows that the trace-metal adsorption capability of a groundwater system may vary greatly from one zone to another. If man’s activities disturb the pH-pE regime, a zone that initially has a strong capability for trace-metal adsorption may lose this capability, or the reverse situation may occur.
In summary, it can be concluded that the environmental chemistry of trace metals is complex. It is difficult to predict their transport behavior within groundwater now systems. In many subsurface environments adsorption and precipitation reactions cause fronts of these elements to move very slowly relative to the velocity of the groundwater. It is not surprising, therefore, that relatively few instances of trace-metal pollution of groundwater have been reported (Kaufman, 1974). In situations where trace-metal contamination does occur, however, the consequences can be serious.
More comprehensive reviews of trace-metal behavior in aqueous systems are provided by Leckie and James (1974) and Leckie and Nelson (1977). The occurrence and controls of trace metals in natural and contaminated groundwaters have been reviewed by Matthess (1974).
Of the many nonmetals listed in the periodic table of the elements, only a few have received much attention in groundwater investigations. These include carbon, chlorine, sulfur, nitrogen, fluorine, arsenic, selenium, phosphorus, and boron. Dissolved forms of carbon (, CO32–, CO2, H2CO3), of chlorine (Cl–), and of sulfur (SO42–, HS–, H2S) occur in abundance in most natural and contaminated groundwaters. The geochemical origin and behavior of these constituents is described in Chapters 3 and 7 and need not be pursued here. Nitrogen in groundwater was discussed previously in this chapter. Our purpose here is to briefly review the hydrochemical behavior of the other important nonmetallic, inorganic constituents that occur as contaminants or as toxic natural constituents in groundwater. The following constituents will be considered: arsenic, fluoride, selenium, boron, and phosphate. These constituents are rarely present in natural or contaminated waters at concentrations above 1 mg/ℓ. Limits for the hrst four constituents on this list are included in the drinking water standards (Table 9.1).
Arsenic and its compounds have been widely used in pigments, as insecticides and herbicides, as an alloy in metals, and as chemical warfare agents (Ferguson and Gavis, 1972). Synthetic organic compounds have now replaced arsenic in most of these uses, but because of past usage and contributions from ore processing wastes and from natural sources, arsenic is still an element of interest in terms of environmental quality. Based on a review of arsenic data from water supply and surface water environments, Ferguson and Gavis (1972) concluded that arsenic concentrations in natural waters often approach or exceed the limits specified in drinking water standards.
The geochemistry of arsenic has been described by Onishi and Sandell (1955). Ferguson and Gavis (1972) have reviewed the arsenic cycle in natural waters. Arsenic occurs in four oxidation states, +V, +III, 0, and –III. The –III state is stable only at extremely low pE values. In the pH range typical of groundwater, the stable solid arsenic forms are As2O5(s) and As2O3(s). These solids are soluble enough for dissolved arsenic species to exist at concentrations well above the permissible concentration in drinking water. Under oxidizing conditions, the following species of dissolved arsenic are stable: , , HAsO42– and AsO43–. Under mildly reducing conditions, , , and HAsO32– are predominant. At low pE values in waters with moderate or large concentrations of dissolved sulfur species, the sulfides As2S3 and AsS are stable. Under these conditions, total dissolved arsenic is limited by solubility constraints to concentrations far below the limit for drinking water. At higher pE conditions, however, dissolved arsenic species can occur at equilibrium concentrations that are much above the permissible limit for drinking water. The fact that the dominant dissolved species are either uncharged or negatively charged suggests that adsorption and ion exchange will cause little retardation as these species are transported along groundwater flow paths.
Of the various nonmetals for which maximum permissible limits are set in drinking water standards, two of these, fluoride and selenium, are of interest primarily because of contributions from natural sources rather than from man-derived sources. Although within the strict usage of the term, these constituents derived from natural sources are not contaminants even if they do occur at toxic levels, their occurrence will be discussed in this section.
Selenium is a nonmetallic element that has some geochemical properties similar to sulfur. Selenium can exist in the +VI, +IV, and –II oxidation states, and occurs in appreciable concentrations in such rocks as shale, in coal, in uranium ores, and in some soils (Lewis, 1976). The aqueous solubilities of selenium salts are in general greater than those of sulfate salts. In dissolved form in groundwater, selenium is present primarily as SeO32– and SeO44– ions. Experimental studies by Moran (1976) indicate that selenium concentrations in groundwater can be controlled by adsorption on coatings or colloidal particles of hydrous iron oxide. In many groundwater systems, however, there is so little selenium present in the rocks or soils that availability is the main limiting factor. There are, however, exceptions to this generalization. For example, Moran (1976) has described an area in Colorado in which waters from many wells have selenium concentrations that exceed the permissible limits for drinking water.
Fluoride, because of the beneficial effects on dental health that have been claimed for it and consequently because of its use as a municipal water-supply additive in many cities, is a constituent that has received much attention in recent decades. Fluoride is a natural constituent of groundwater in concentrations varying from less than 0.1 mg/ℓ to values as high as 10–20 mg/ℓ. Maximum permissible limits specified for drinking water range from 1.2 to 2.4 mg/ℓ (Table 9.1), depending on the temperature of the region. Concentrations recommended for optimum dental health are close to 1 mg/ℓ, but also vary slightly depending on the temperature of the region. Natural concentrations of F– in groundwater depend on the availability of F– in the rocks or minerals encountered by the water as it moves along its flow paths and on solubility constraints imposed by fluorite (CaF2) or fluorapatite, Ca3(PO4)2 · CaF2. Equilibrium dissolution-precipitation relations for these minerals in water are
Because of the lack of PO43- in most groundwater environments, CaF2 is probably the mineral phase that exerts the solubility constraint in situations where F– is available from the host rock. However, as can be determined by substituting values in Eq. (9.20), Ca2+ concentrations of many hundreds of milligrams per liter are required for this solubility constraint to limit F– concentrations to levels below drinking water standards. The fact that nearly all groundwaters are undersaturated with respect to fluorite and fluorapatite suggests that the F– content of groundwater is generally limited by the availability of F– in the rocks and sediments through which the groundwater moves rather than by the solubility of these minerals. Groundwater with F– contents that exceed drinking water standards is common in the Great Plains region of North America and in parts of the southwestern United States. This suggests that F– is more readily available from the rocks of these regions than in most other areas of North America.
Although phosphorus is not a harmful constituent in drinking water, its presence in groundwater can be of considerable environmental significance. Phosphorus additions to surface-water bodies in even small amounts can, in some circumstances, produce accelerated growth of algae and aquatic vegetation, thereby causing eutrophication of the aquatic system. Because of this, phosphorus is regarded as a pollutant when it migrates into ponds, lakes, reservoirs, and streams. The occurrence and mobility of phosphorus in groundwater is important in situations where there is a potential for groundwater to feed phosphorus into surface-water environments. Through the widespread use of fertilizers and disposal of sewage on land, the potential for phosphorus influx to surface-water systems as a result of transport through the groundwater zone is increasing.
Dissolved inorganic phosphorus in water occurs primarily as H3PO4, , , and PO43–. Since H3PO4 is a polyprotic acid [see discussion in Section 3.3 and Figure 3.5(b)], the relative occurrence of each of these forms of dissolved phosphorus is pH-dependent. In the normal pH range of groundwater, and HPO42– are the dominant species. Because these species are negatively charged, the mobility of dissolved phosphorus in groundwater below the organic-rich horizons of the soil zone is not strongly limited by adsorption. The dominant control on phosphorus in the groundwater zone is the solubility of slightly soluble phosphate minerals.
Solubility control is usually attributed to one or more of the following minerals: hydroxylapatite, Ca5(OH)(PO4)3; strengite, FePO4 • 2H2O; and varisite, AlPO4 • 2H2O. From the law of mass action, equilibrium expressions for precipitation-dissolution reactions of these minerals in water can be expressed as
where Kn, Ks, and Kv are the equilibrium constants for hydroxylapatite, strengite, and varisite, respectively. The log K values are for 25°C and 1 bar. These solubility relations indicate that the concentrations of Ca2+, Fe3+, and Al3+ can control the equilibrium concentration of dissolved phosphorus in solution. Equilibrium concentrations of total dissolved phosphorus computed from the solubility relations above are shown in Figure 9.21.
Since the solubilities of hydroxylapatite, strengite, and varsite depend on the concentrations of Ca2+, Fe3+, and Al3+, respectively, each solubility line is valid only for a specified concentration of these ions. Two solubility lines for hydroxylapatite (lines c and d) are shown in order to illustrate the influence of Ca2+ on the equilibrium phosphate concentration. The varisite solubility line (line a) is based on the assumption that the Al3+ concentration is governed by the solubility of gibbsite, Al(OH)3(s). For strengite solubility (line b) it is assumed that Fe(OH)3(s) limits the Fe3+ concentration.
From Figure 9.21 it is apparent that equilibrium concentrations of total dissolved phosphate are large in waters that have low Ca2+ concentrations and pH values near or below 7. In anaerobic groundwater, Fe2+ rather than Fe3+ is the dominant form of dissolved iron. In this situation strengite solubility is not a limiting factor in phosphate occurrence. Groundwaters with these characteristics, namely, low Ca2+ concentrations and reducing redox conditions, occur in many regions that are underlain by crystalline igneous rocks or by deposits derived from these rocks. In regions such as the Precambrian Shield region of Michigan, Minnesota, and parts of Canada, migration of dissolved phosphorus from septic systems through shallow groundwater regimes into clear-water lakes poses a significant water-quality problem. Small increases in phosphorus influx to many of these lakes can cause extensive growth of algae and undesirable aquatic vegetation. Phosphorus mobility in groundwater can be a significant factor in the environmental impact of cottage and recreational developments near lakes.
For a more extensive review of the hydrochemical controls on phosphorus in aqueous systems and soils, the reader is referred to Stumm and Morgan (1970) and Beck and De Haan (1974).
In Chapter 3, it was indicated that all groundwater normally contains small amounts of dissolved organic substances of natural origin. These substances, which are referred to as humic and fulvic acids, are of little concern from a water quality viewpoint. Organic substances produced by man, however, are of great concern. The number of identified man-made organic compounds now totals near 2 million and is growing at a rate of about 250,000 new formulations annually, of which 300–500 reach commercial production (Giger and Roberts, 1977).
Increasing numbers of these substances are relatively resistant to biological degradation. Many resist removal in sewage treatment plants. It is estimated that up to one-third of the total production of today’s synthetic organic compounds eventually enters the biosphere (Iliff, 1973). More than 1200 individual man-made organic substances have been identified in drinking water supplies (Shackelford and Keith, 1976). This number is increasing rapidly as investigations of organic compounds in water supplies are intensified.
The question that should be addressed here is: To what extent and under what circumstances are organic compounds causing degradation of groundwater quality? Unfortunately, this question cannot be answered at present. Since there have been so few investigations of organic compounds in groundwater, it is not possible at the present time to draw any general conclusions. Our purpose here is to briefly review some of the factors that are expected to play a major role in the migration of organic compounds into groundwater systems.
Organic chemicals make their way to the land surface as a result of the use of pesticides, the use of land for sewage disposal, the use of sanitary landfills or refuse dumps for disposal of organic compounds, burial of containers with organic compounds at special burial sites, leakage from liquid waste storage ponds, and accidental spills along highways or other transportation routes. There are hundreds of thousands of locations in North America and Europe in which organic compounds may be a threat to groundwater quality.
Fortunately, there are several mechanisms that tend to prevent or retard the migration of most organic substances from the land surface or soil zone into deeper parts of the subsurface environment. These mechanisms include chemical precipitation, chemical degradation, volatilization, biological degradation, biological uptake, and adsorption.
Many organic substances have extremely low solubility in water. This generally limits the possibility for appreciable migration of large quantities in groundwater. However, because many of these substances are toxic at very low concentrations, solubility constraints are often not capable of totally preventing migration at significant concentration levels. For example, comparison of the solubilities and maximum permissible concentrations in drinking water of some of the common pesticides (Table 9.3) indicates that the solubilities generally exceed the permissible concentrations of these pesticides. A more comprehensive description of pesticide compositions and solubilities is presented by Oregon State University (1974).
|Compound||Maximum permissible concentration (mg/ℓ)||Solubility in water (mg/ℓ)|
Many organic substances are lost from the soil zone as a result of volatilization (i.e., conversion to the vapor state). When the substances transform from the solid phase or from the dissolved phase to the vapor phase, they are lost by diffusion to the atmosphere. This process can greatly reduce the concentrations available for transport in subsurface water. For volatilization to occur, however, a gas phase must be present. Therefore, this process cannot be effective if the compounds migrate below the water table, where species occur only in dissolved form.
Nearly all pesticides and many other organic substances that make their way to the land surface and hence into the soil zone undergo biochemical degradation. The soil zone contains a multitude of bacteria that can convert and consume countless numbers of organic compounds. If it were not for these organisms, the biosphere would long ago have become intolerably polluted with organic compounds. In terms of environmental contamination, the major concern is focused on those organic substances that are not readily degraded by bacteria, either in the soil zone or in sewage treatment facilities. These substances are known as refractory compounds. Their presence in the surface environment is becoming increasingly pervasive.
The organic substances that pose the greatest threat to the quality of groundwater resources are those that are relatively soluble, nonvolatile, and refractory, the main mechanism that prevents most of these compounds from readily migrating from the land surface into aquifer systems is adsorption. Minerals and amorphous inorganic and organic substances in the soil zone and in deeper geologic materials all provide surfaces for adsorption of organic compounds. Unfortunately, adsorption isotherms are available for only a small percentage of the existing organic chemicals that are entering the biosphere. These isotherms relate to only a small number of permeable geological materials under a limited range of hydrochemical conditions. Because of this paucity of adsorption data, it is not possible to draw general conclusions on the potential magnitude of the hazard to groundwater resources posed by increasing use and dependence on organic chemicals.
For readers interested in obtaining more information on the occurrence, classification, and movement of organic substances in groundwater and surface waters, the following introduction to the literature may be useful. Giger and Roberts (1977) describe the problems associated with characterizing refractory organic compounds in contaminated waters. A classification scheme for organic compounds in water is presented by Leenheer and Huffman (1977). The chemical, ecological, and adsorptive properties of a wide variety of insecticides and herbicides are described by the Oregon State University (1974). Malcolm and Leenheer (1973) indicate the usefulness of dissolved organic carbon measurement as a contamination indicator in groundwater and surface-water investigations. Based on an extensive literature review, Shackelford and Keith (1976) summarized the reported occurrences of organic compounds in groundwater and other waters used for drinking water supplies. The behavior of petroleum substances such as oil and gasoline in water is described by McKee (1956). Adsorption isotherms for several organic compounds in selected soils are described by Kay and Elrick (1967), Hamaker and Thompson (1972), Davidson et al. (1976), and Hague et al. (1974).
9.4 Measurements of Parameters
There are three groups of methods for determination of the velocity of groundwater. The first group includes all techniques that are directly dependent on use of the Darcy equation. The second group involves the use of artificial tracers. The third group consists of groundwater age-dating methods using environmental isotopes such as tritium and carbon 14. Darcy-based techniques require information on the hydraulic conductivity, hydraulic gradient, and porosity in the portion of the flow field in which velocity estimates are desired. From these data the average linear velocity can be computed using Eq. (2.82). Methods by which field values of hydraulic conductivity, hydraulic gradient, and porosity are determined are described in Chapters 2, 6, and 8. They need not be pursued here, other than to note that velocity estimates based on use of these parameters in Darcy-based equations have large inherent uncertainties that generally cannot be avoided. In field situations hydraulic conductivity determinations often have large uncertainties. Errors in hydraulic conductivity measurements combined with the errors associated with determination of the gradient and porosity result in considerable error being associated with the computed velocity. In some situations better accuracy can normally be achieved by use of artificial groundwater tracers, although this may involve greater expense.
The most direct method for groundwater velocity determination consists of introducing a tracer at one point in the How field and observing its arrival at other points. After making adjustments for the effect of dispersion, the groundwater velocity can be computed from the travel time and distance data. The literature is replete with descriptions of experiments of this type. Many types of nonradioactive and radioactive tracers have been used, ranging from such simple tracers as salt (NaCl or CaCl2), which can be conveniently monitored by measurements of electrical conductance, to radioisotopes such as 3H, 131I, 29Br, and 51Cr-EDTA (an organic complex with 51Cr), which can be accurately monitored using radioactivity detectors. Radioisotopes have the disadvantage of government licensing requirements for their use and of being hazardous when used by careless workers. Fluorescent dyes (fluoroscein and rhodamine compounds) have been used by many investigators. In field tests, visual detection of the dye can sometimes yield adequate results. Dye concentrations can be measured quantitatively to very low concentrations when necessary. Recent work suggests that Freon (Cl3CF) may be one of the best of the artificial tracers for use in groundwater velocity tests (Thompson et al., 1974). It is nonreactive with geologic materials and can be used in extremely small concentrations that are nonhazardous in public waters. For reviews of tracer techniques in groundwater investigations, the reader is referred to Knutson (1966), Brown et al. (1972), and Gaspar and Oncescu (1972).
The direct tracer method of groundwater velocity determination described above has four main disadvantages: (1) because groundwater velocities are rarely large under natural conditions, undesirably long periods of time are normally required for tracers to move significant distances through the flow system; (2) because geological materials are typically quite heterogeneous, numerous observation points (piezometers, wells, or other sampling devices) are usually required to adequately monitor the passage of the tracer through the portion of the How field under investigation; (3) because of (1), only a small and possibly nonrepresentative sample of the flow field is tested; and (4) because of (2), the flow field may be significantly distorted by the measuring devices. As a result of these four factors, tracer experiments of this type commonly require considerable effort over extended periods of time and are rarely performed.
A tracer technique that avoids these disadvantages was developed in the USSR in the late 1940’s. This technique, which has become known as the borehole dilution or point-dilution method, is now used extensively in Europe. Borehole dilution tests can be performed in relatively short periods of time in a single well or piezometer. The test provides an estimate of the horizontal average linear velocity of the groundwater in the formation near the well screen. A schematic representation of a borehole dilution test is shown in Figure 9.22(a). The test is performed in a segment of a well screen that is isolated by packers from overlying and underlying portions of the well. Into this isolated well segment a tracer is quickly introduced and is then subjected to continual mixing as lateral groundwater flow gradually removes the tracer from the well bore. The combined effect of groundwater through-flow and mixing within the isolated well segment produces a dilution versus time relation as illustrated in Figure 9.22(b). From this relation the average horizontal velocity of groundwater in the formation beyond the sand or gravel pack but close to the well screen is computed. (The theory on which the computational methods are based is described below.
Although adjustments in field technique and analytical methods can be made to take into account the effects of flow with a significant vertical component, the borehole dilution method is best suited for velocity determination in steady-state lateral flow regimes. We will proceed on this basis and with the additional stipulation that complete mixing of the tracer in the well-screen segment is maintained with no significant disturbance of the flow conditions in the formation. The effect of the well bore and sand pack in a lateral flow regime is shown in Figure 9.23.
The average linear velocity of the groundwater in the formation beyond the zone of disturbance is . The average bulk velocity across the center of the well bore is denoted by . It will be assumed that the tracer is nonreactive and that it is introduced instantaneously at concentration C0 into the isolated segment of the well screen. The vertical cross-sectional area through the center of the isolated segment is denoted as A. The volume of this well segment is W. At time t > 0, the concentration C in the well decreases at a rate
which, upon rearrangement, yields
Integration and use of the initial condition, C = C0 at t = 0, leads to
Thus, from concentration versus time data obtained during borehole dilution tests, values of can be computed. The objective of the test, however, is to obtain estimates of . This is accomplished using the relation
where n is the porosity and α is an adjustment factor that depends on the geometry of the well screen, and on the radius and hydraulic conductivity of the sand or gravel pack around the screen. The usual range of α for tests in sand or gravel aquifers is from 0.5 to 4 (Drost et al., 1968).
Borehole dilution tests performed at various intervals within a well screen can be used to identify zones of highest groundwater velocity. These zones are often of primary interest because contaminants can move through them at velocities much higher than in other parts of the system. Identification of the high-velocity zones, which may occur in only a thin segment of an aquifer system, can provide for efficient design of monitoring networks for groundwater quality.
The borehole dilution method is described in detail by Halevy et al. (1967) and Drost et al. (1968). In most borehole dilution tests described in the literature, radioactive tracers were used. The recent advent of commercially available electrodes for use with portable pH meters for rapid down-hole measurement of Cl– or F– has made it feasible to conduct borehole-dilution tests with these readily available tracers in a more convenient manner than was previously the case. An example is described by Grisak et al. (1977). An even simpler approach involves the use of salt as the tracer with down-hole measurement of electrical conductance as the salt is flushed from the well screen. Borehole dilution tests, like many other types of field tests used in groundwater studies, can be accomplished using simple, inexpensive equipment or more elaborate instrumentation. The choice of method depends on factors such as the hydrogeologic setting, availability of instrumentation, and the experimental precision and reproducibility that is desired.
From a measurement point of view, the most elusive of the solute transport parameters is dispersivity. Longitudinal dispersivity can be measured in the laboratory by passing a nonreactive tracer through cylindrical samples collected from boreholes or excavations. These experiments produce break-through curves as illustrated in Figure 9.1(c). The dispersivity of the sample can be computed by fitting solutions of the advection-dispersion equation to the experimentally determined breakthrough curve. If the breakthrough curve is obtained from a column test with step-function tracer input, Eq. (9.5) can be used in the analyses of the curve. The velocity is obtained by dividing the specific discharge of water through the column by the porosity. Dispersivity is then obtained as the remaining unknown in the equation. Dispersivity values obtained from column tests on disturbed or undisturbed samples of unconsolidated geological materials invariably yield values in the range 0.01–2 cm. Based on 2500 column dispersion tests of this type, Klotz and Moser (1974) observed that values of longitudinal dispersivity depend on grain size and grain-size distribution and are independent of grain shape, roughness, and angularity.
Longitudinal dispersivity values determined by column tests are generally viewed as providing little indication of the in situ dispersivity of the geologic materials. Dispersivity has the distinction of being a parameter for which values determined on borehole-size samples are commonly regarded as having little relevance in the analysis of problems at the field scale.
It is generally accepted that longitudinal and transverse dispersivities under field conditions are larger than those indicated by tests on borehole samples. In other words, tracer or contaminant spreading in the field as a result of dispersion is greater than is indicated by laboratory measurements. This difference is normally attributed to the effects of heterogeneities on the macroscopic flow field. Since most heterogeneities in geological materials occur at a larger scale than can be included in borehole samples, dispersivity values from tests on small samples can be viewed as representing a property of the medium but at a scale of insufficient size for general use in prediction of dispersion in the field.
Studies of contaminant migration under field conditions require dispersivity measurements in the field. Although this premise is generally accepted, there is little agreement on the types of field dispersivity tests or methods for test analysis that are most appropriate. This state of affairs may be the result of the fact that relatively few detailed field dispersivity tests have been conducted, rather than a result of excessive difficulties of the task. It has only been in recent years that dispersivity at the field scale has received much attention. In comparison to the many thousands of field hydraulic conductivity and transmissivity tests that have been conducted in the common types of geologic materials, only a few tens of field dispersivity tests are reported in the literature.
There are four main types of field dispersivity tests. These are (1) single-well withdrawal-injection tests, (2) natural-gradient tracer tests, (3) two-well recirculating withdrawal-injection tests, and (4) two-well pulse tests. In each of these tests a nonreactive tracer is introduced into the groundwater system. In the single-well test, the tracer is pumped in for a set time period followed by pumping from the well and monitoring the concentration levels. The dispersivity of the formation near the well screen is computed from the concentration response data (Percious, 1969; Fried et al., 1974). In the natural-gradient test the tracer is introduced into the system without much disturbance of the flow regime. Its migration is then monitored at one or more sampling points (Fried, 1976). In the two-well recirculating test, the tracer is injected into the flow regime at one well. It is pumped out of a second well and then recirculated through the withdrawal-injection system. The concentration versus time response at the withdrawal well serves as a basis for Computation of the dispersivity using analytical or numerical models (Grove and Beetem, 1971; Pickens et al., in press). In the two-well pulse test a tracer is introduced into a well situated within the drawdown cone caused by pumping of a second well. Concentration data from the pumping well are used for calculation of a dispersivity value for the segment of the formation between the two wells (Zuber, 1974).
Fried (1975) presents an outline of the test methods and the mathematical basis for analysis of data from the first three types of tests indicated above. In each case dispersivity values are obtained by fitting an analytical or numerical model to the experimental data. Zuber (1974) emphasizes that the dispersivity value obtained from a given field experiment depends, sometimes to a high degree, on the mathematical model used in the analysis, and on the scale of the experiment. Aquifers are commonly stratified and tracers travel at different rates through the different layers. Even though the differences in hydraulic conductivity between layers may be almost imperceptible, the design of wells used for sampling can exert a dominant influence on the dispersivity values computed from concentration response data. Tests in which monitoring wells with large screened intervals are used can yield large apparent dispersivities because of mixing in the well screen. Pickens et al. (in press) describe a multilevel point-sampling device that is well suited for use in dispersion tests in sandy aquifers. Castillo et al. (1972) show that the dispersive nature of fractured rocks can exhibit great complexities in comparison with that expected of granular materials.
Reactive contaminants transported by groundwater are distributed between the solution phase and other phases. Reactions between the dissolved species and the geological materials may cause a portion of the dissolved species to be transferred to the solids as a result of adsorption or ion exchange. Reactions primarily among the contaminant, other dissolved constituents, and the geological materials may cause a portion of the contaminant concentration to be incorporated into a solid form as a result of chemical precipitation. Above the capillary fringe, where a continuous gas phase normally exists in part of the void space, reactions may cause some of the contaminant mass to be transferred from the solution phase to the gas phase, such as occurs during denitrification in the unsaturated zone. In each of these processes the contaminant is partitioned between the solution and other phases. The ultimate fate of the contaminant in the subsurface zone can depend on the degree of irreversibility of the reactions. Prediction of the rate and concentrations at which a contaminant will be transported in groundwater requires knowledge of the rates and extent to which this partitioning will occur.
In this brief discussion of this broad topic, we will focus on the partitioning of contaminants between the liquid and solid phases. There are four main approaches to the determination of this type of partitioning. These include (1) use of computational models based primarily on thermodynamically derived constants or coefficients for equilibrium systems, (2) laboratory experiments in which the contaminant in solution is allowed to react under controlled conditions with samples of the geologic materials of interest, (3) field experiments in which the degree of partitioning is determined during passage of contaminant solutions through a small segment of the groundwater system and (4) studies of existing sites at which contamination has already occurred.
Insight into the computational approach based on equilibrium thermodynamics can be acquired from Chapters 3 and 7. If it is expected that the concentration of the contaminant in solution is controlled by precipitation-dissolution reactions and if the necessary thermodynamic data on the aqueous and solid components of the system are available, the equilibrium concentration of the contaminant in solution under specified conditions can be computed. Although the necessary computational techniques are well developed, this method has limited application for many types of contaminants because of uncertainties with regard to chemical composition and the free energies of the controlling solid phases or because of sluggish rates of the dominant reactions. In many cases the contaminant species of interest are transported in solutions that are chemically very complex. The presence of organic compounds can cause contaminant mobility to be enhanced considerably beyond that predicted based on inorganic considerations alone.
In the laboratory, the degree of contaminant partitioning is determined in column experiments and in what are known as batch experiments. In column experiments (Figure 9.1), prepared solutions or natural waters to which the contaminant is added are passed through cylindrical samples of the geologic materials of interest. If the flow rate and input water chemistry is regulated to approximate the held conditions and if disturbance of the sample prior to emplacement in the column has not caused the material to acquire properties that deviate significantly from held conditions, the degree of partitioning and retardation obtained from this type of experiment provides an indication of what will take place in the held. Column experiments, however, are rarely conducted with adherence to all of these requirements. There is considerable uncertainty, therefore, in application of the results to held situations. Column experiments are described by Rovers and Farquhar (1974) and Griffin et al. (1976) using sanitary landfill leachate, by Routson and Serne (1972) using trace concentrations of radionuclides, by Kay and Elrick (1967) and Huggenberger et al. (1972) using lindane (a pesticide), by Doner and McLaren (1976) using urea, and by many other investigators using various chemical constituents.
In batch experiments the contaminated solution and the geologic material in a disaggregated state are brought into contact in a-reaction vessel. After a period of time that normally ranges from hours to days, the degree of partitioning of the contaminant between the solution and the geologic materials is determined. For partitioning data from these experiments to be applied with confidence in the analysis of held situations, comparisons with results of column or held tests are necessary. Batch tests have the advantage of being relatively quick and inexpensive to conduct. For some contaminants, the batch test is a standard method for establishing adsorption isotherms or selectivity coefficients in ion exchange reactions. Sample disturbance and the lack of representation of field how conditions can detract from the validity of the results in the analysis of held situations. Samples used in batch tests are usually exposed to oxidizing conditions (i.e., to oxygen in the air) during sample preparation and during the tests. Since the adsorptive capabilities of oxidized materials can be much different than reduced materials, the test results can be invalid for analysis of contaminant behavior in field systems.
The most direct but rarely the most convenient method for determining the partitioning and retardation of the contaminant is to conduct field tests. Injection of a solution of appropriate composition into a small segment of the groundwater system followed by monitoring of its behavior can provide, in favorable circumstances, a basis for prediction of contaminant behavior elsewhere in the system. Field tests of this type can be time-consuming and expensive. In order to obtain adequate information, numerous tests may be required. In some situations the need to obtain reliable information on contaminant behavior is great enough to justify this effort.
Another approach for obtaining information on the partitioning and retardation of contaminants during transport in groundwater is to conduct investigations at existing sites where groundwater pollution has already occurred. For results of these investigations to have more than site-specific significance, not only must the distributions of the contaminants in the water and on the porous media be determined, but the factors that influence these distributions must also be investigated. During recent years an appreciable number of detailed studies of sites with subsurface contamination have been reported in the literature. Some of the more notable examples are those by McKee et al. (1972), Childs et al. (1974), Suarez (1974), Ku et al. (1978), Goodall and Quigley (1977), and Gillham and Cherry (1978).
9.5 Sources of Contamination
Land Disposal of Solid Wastes
In North America approximately 3 kg of refuse per capita is produced daily. More than 20,000 landfills across the continent accommodate more than 90% of the solid waste that is produced by municipal and industrial activities. According to Yen and Scanlon (1975), a city of 1 million people generates refuse with an annual volume equivalent to 80 ha covered 5 m deep. Although materials recovery and incineration may eventually decrease the amount of waste that is disposed of by landfilling, landfills will continue to be the primary method of disposal of these wastes during at least the next few decades.
The design, construction, and operational aspects of land disposal of refuse are described by Mantell (1975). For purposes of this discussion this information is not required, other than to recognize that much of the solid waste (refuse) that is now disposed of on land is emplaced in engineered disposal systems known as sanitary landfills. In sanitary landfills, solid waste is reduced in volume by compaction and then is covered with earth. Ideally, the earth cover is placed over the refuse at the conclusion of each day’s operation, but in practice less frequent cover application is common. The landfill, consisting of successive layers of compacted waste and earth, may be constructed on the ground surface or in excavations. In North America a large number of the older sites that receive municipal wastes are open dumps or poorly operated landfills. Newer sites are generally better situated and better operated. It is estimated that 90% of the industrial wastes that are considered to be hazardous are landfilled, primarily because it is the least expensive waste management option.
Our purpose here is to consider some of the effects that refuse disposal can have on the groundwater environment. With the exception of arid areas, buried refuse in sanitary landfills and dumps is subject to leaching by percolating water derived from rain or snowmelt. The liquid that is derived from this process is known as leachate. Table 9.4 indicates that leachate contains large numbers of inorganic contaminants and that the total dissolved solids can be very high. Leachate also contains many organic contaminants. For example, Robertson et al. (1974) identified more than 40 organic compounds in leachate-contaminated groundwater in a sandy aquifer in Oklahoma. These authors concluded that many of these compounds were produced by leaching of plastics and other discarded manufactured items within the refuse. Not only do the leachates emanating from landfills contain contaminants derived from solids, but many leachates contain toxic constituents from liquid industrial wastes placed in the landfill.
|Parameter||Representative range (mg/ℓ)|
|P as PO4||1–100|
|Total dissolved organic carbon||200–30,000|
|COD (chemical oxidation demand)||1000–90,000|
|Total dissolved solids||5000–40,000|
Concern has developed in recent years with regard to the effect of landfills on the quality of groundwater resources. Garland and Mosher (1975) cite several examples where groundwater pollution has been caused by landfills. A case where leachate migration caused serious pollution of a large aquifer used as a city’s water supply is described by Apgar and Satherthwaite (1975). It is expected that the cost of rectifying this situation will eventually total many millions of dollars.
Numerous investigations in North America and Europe have shown that in nonarid regions, infiltration of water through refuse causes water table mounding within or below the landfill. The mounding process is similar to that described in Section 8.11. Water-table mounding causes leachate to flow downward and outward from the landfill as illustrated in Figure 9.24. Downward flow of leachate may threaten groundwater resources. Outward flow normally causes leachate springs at the periphery of the landfill or seepage into streams or other surface-water bodies. If the paths of leachate migration do not lead to aquifers containing potable water, downward movement of leachate will not pose a threat to groundwater resources.
In situations where landfills are located in relatively permeable materials such as sand, gravel, or fractured rock, leachate migration may cause contamination over areas many times larger than the areas occupied by the landfills. An example of such a case is shown in Figure 9.25. At this landfill site on moderately permeable glaciodeltaic sand, a large plume of leachate-contaminated water, represented in Figure 9.25 by the Cl– distribution, has penetrated deep into the aquifer and has moved laterally several hundreds of meters along the paths of groundwater flow. This contamination developed over a period of 35 years. Infiltration of water through the landfill will continue to produce leachate for many decades. Transport by groundwater flow in the sand will cause the zone of contamination to greatly expand. In this particular case, however, the aquifer is not used for water supply.
The spreading contaminant plume is therefore not regarded as a significant problem. At a landfill on sand and gravel on Long Island, N.Y., Kimmel and Braids (1974) delineated a leachate plume that is more than 3000 m long and greater than 50 m in depth. These two examples and others described in the literature indicate that if leachate has access to active groundwater flow regimes, pollution can spread over very large subsurface zones. Physical and chemical processes are sometimes incapable of causing appreciable attenuation of many of the toxic substances contained within the leachate plume.
If landfills are situated in appropriate hydrogeologic settings, both groundwater and surface-water pollution can be avoided. It is commonly not possible, however, to choose sites with ideal hydrogeologic characteristics. In many regions land of this type is not available within acceptable transportation distances, or it may not be situated in an area that is publicly acceptable for land filling. For these and other reasons most landfills are located on terrain that has at least some unfavorable hydrogeologic features.
Although it is well established that landfills in nonarid regions produce leachate during at least the first few decades of their existence, little is known about the capabilities for leachate production over much longer periods of time. In some cases leachate production may continue for many decades or even hundreds of years. It has been observed, for example, that some landfills from the days of the Roman Empire are still producing leachate. Many investigators have concluded that at the present time there have been very few occurrences of leachate contamination of aquifers that are used for water supply. Whether or not it will be possible to draw similar conclusions many years from now remains to be established.
Farvolden and Hughes (1976) have concluded that solid waste can be buried at almost any site without creating an undue groundwater pollution hazard, provided that the site is properly designed and operated. A testing program to define the hydrogeological environment is essential. These authors indicate that if uncontrolled leachate migration is unacceptable, the leachate should be collected and treated as a liquid waste. One feasible way to ensure that no leachate leaves the site is to establish a hydraulic gradient toward the site, perhaps by pumping, Liners for emplacement beneath landfills are currently being evaluated as a control method but have not yet been established in practice. Some examples of controls on leachate migration using drains or wells are shown in Figure 9.26. These types of control measures require that the collected leachate be treated or otherwise managed in an appropriate manner.
In addition to the production of leachate, infiltration of water into refuse causes gases to be generated as biochemical decomposition of organic matter occurs. Gases such as CO2, CH4, H2S, H2, and N2, are commonly observed. CO2 and CH4 are almost invariably the most abundant of these gases. CH4 (methane) has a low solubility in water, is odorless, and generally is of little influence on groundwater quality. In the environmental impact of landfills, however, it can be of great importance because of its occurrence in gaseous form in the zone above the water table. It is not uncommon for CH4 to attain explosive levels in the refuse air. In some situations CH4 at dangerous levels can move by gaseous diffusion from the landfill through the unsaturated zone in adjacent terrain. Migration of CH4 at combustible levels from landfills through soils into residences has occurred in urban areas. In recent years, installation of gas vents in landfills to prevent buildup of methane in the zone above the water table has become a common practice.
In addition to hazards caused by the potential for methane explosion, gaseous migration from landfills can result in extensive damage to vegetation and odor problems. Case histories of gas migration from landfills have been described by Flower (1976). Mohsen (1975) has presented a theoretical analysis of subsurface gas migration from landfill sources. The interactions of the various factors that influence gas production in landfills have been described by Farquhar and Rovers (1973).
Sewage is placed on or below the land surface in a variety of ways. Widespread use of septic tanks and drains in rural, recreational, and suburban areas contributes filtered sewage effluent directly to the ground. Septic tanks and cesspools are the largest of all contributors of wastewater to the ground and are the most frequently reported sources of groundwater contamination in the United States (U.S. Environmental Protection Agency, 1977). Twenty-nine percent of the U.S. population disposes of its domestic waste through residential disposal systems. An increasing percentage of the municipal sewage in industrialized countries is being processed in primary and secondary sewage treatment plants. Although this decreases surfacewater pollution, it produces large volumes of solid residual materials known as sewage sludge. In many areas this sludge, which contains a large number of potential contaminants, is spread on agricultural or forested lands. In some regions liquid sewage that has not been treated or that has undergone partial treatment is sprayed on the land surface. Application of liquid sewage and sewage sludge to the land provides nutrients such as nitrogen, phosphorus, and heavy metals to the soil. This can stimulate growth of grasses, trees, and agricultural crops. Land that is infertile can be made fertile by this practice. One of the potential negative impacts of this type of sewage disposal is degradation of groundwater quality.
Primary- and secondary-treated sewage is being spread on forested land and crop land in an increasing number of areas in Europe and North America. For example, in Muskegon County, Michigan, more than 130 million liters per day of sewage effluent is sprayed on the land surface (Bauer, 1974). For many decades cities such as Berlin, Paris, Milan, Melbourne, Fresno, and many others have been using sewage for irrigation of crops. Not only are the nutrients in sewage effluent valuable, but the water itself is a valuable resource in many regions. In some situations intensely treated sewage effluent may be used as a source of artificial recharge for aquifers that serve for municipal water supply. Injection of treated sewage into coastal aquifers may serve as a means of controlling the intrusion of salt water.
Considering the many ways in which liquid and solid constituents from sewage reach the land surface and subsurface zones, it is reasonable to expect that over the long term the quality of groundwater resources in many areas will reflect the extent to which hydrogeologic factors are considered in the overall planning and operation of sewage management systems. In a textbook of this type it is not feasible to look specifically at the hydrogeologic and geochemical factors that are important in each of the land-application or disposal-of-sewage options that are in use. Before proceeding to other topics, however, we will provide a brief guide to some of the more important studies that have been conducted. For a detailed guide to the literature in this area, the reader is referred to the US. Environmental Protection Agency (1974a).
In the mid-1960’s the detergent industry replaced ABS with linear alykl sulfonate (LAS), a compound that is readily biodegradable in aerobic environments. Cases of LAS and ABS contamination of wells have been a rare occurrence since LAS gained widespread use, a somewhat surprising situation considering that many septic systems drain into anaerobic groundwater environments where the effects of biodegradation are probably minimal. LAS may undergo considerable retardation as a result of adsorption.
Effluent from septic systems includes many other types of contaminants. One of the most frequently reported of these contaminants in groundwater is nitrate. As indicated in Section 9.3, nitrate commonly does not undergo complete biochemical reduction to N2 even if the groundwater system is anaerobic. Nitrate emanating from septic systems into groundwater is transported along the groundwater flow paths. A detailed case history of the migration of nitrate and other contaminants in groundwater as a result of discharge from septic systems was presented by Childs et al. (1974).
In some areas the primary concern with regard to contaminant migration from septic systems is surface-water quality rather than groundwater quality. This is particularly the case in areas of recreational lakes where cottages and tourist facilities use septic systems located near lakes. Transport of nitrogen and phosphorus through the groundwater zone into lakes can cause lake eutrophication manifested by accelerated growth of algae and decrease in water clarity. Some examples of hydrogeologic investigations in recreational lake environments are described by Dudley and Stephenson (1973) and Lee (1976).
Another concern associated with the disposal of treated or untreated sewage on or below the land surface revolves around the question of how far and how fast pathogenic bacteria and viruses can move in subsurface flow systems. This problem is also crucial in the development of municipal water supplies by extraction of water from wells located adjacent to polluted rivers. The literature is replete with investigations of movement of bacteria through soils or granular geological materials. As bacteria are transported by water flowing through porous media, they are removed by straining (filtering), die-off, and adsorption. The migration of the bacterial front is greatly retarded relative to the velocity of the flowing water. Although bacteria can live in an adsorbed state or in clusters that clog parts of the porous medium, their lives are generally short compared to groundwater flow velocities. In medium-grained sand or finer materials, pathogenic and coliform organisms generally do not penetrate more than several meters (Krone et al., 1958). Field studies have shown, however, that in heterogeneous aquifers of sand or gravel, sewage-derived bacteria can be transported tens or hundreds of meters along the groundwater flow paths (Krone et al., 1957; Wesner and Baier, 1970).
Viruses are very small organic particles (0.07–0.7 μm in diameter) that have surface charge. There is considerable evidence from laboratory investigations indicating that viruses are relatively immobile in granular geological materials (Drewry and Eliassen, 1968; Robeck, 1969; Gerba et al., 1975; Lance et al., 1977). Adsorption is a more important retardation mechanism than filtering in highly permeable granular deposits. Problems associated with sampling and identification of viruses in groundwater systems have restricted the understanding of virus behavior under field conditions. Advances in sampling technology (Wallis et al., 1972; Sweet and Ellender, 1972) may lead to a greatly improved understanding of virus behavior in aquifers recharged with sewage effluent.
Although there is considerable evidence indicating that bacteria and viruses from sewage have small penetration distances when transported by groundwater through granular geologic materials, similar generalizations cannot be made for transport in fractured rock. It is known that these microorganisms can live for many days or even months below the water table. In fractured rocks, where groundwater velocities can be high, this is sufficient time to produce transport distances of many kilometers.
As man relies more heavily on land application as a means of disposal for municipal sewage effluent and sludge, perhaps the greatest concern with regard to groundwater contamination will be the mobility of dissolved organic matter. Sewage effluent contains many hundreds of dissolved organic compounds, of which very little is known about their toxicity and mobility. Some of these compounds may eventually be shown to be more significant in terms of degradation of groundwater quality than nitrate, trace metals, bacteria, or viruses.
Of all the activities of man that influence the quality of groundwater, agriculture is probably the most important. Among the main agricultural activities that can cause degradation of groundwater quality are the usage of fertilizers and pesticides and the storage or disposal of livestock or fowl wastes on land. The most widespread effects result from the use of fertilizer. In industrialized countries most fertilizer is manufactured chemically. This type of fertilizer is known as inorganic fertilizer. In less developed countries, animal or human wastes are widely used as organic fertilizer.
Fertilizers are categorized with respect to their content of nitrogen (N), phosphorus (P), and potassium (K). These are the three main nutrients required by crops. The annual application rates of fertilizers vary greatly from region to region and from crop to crop. Nitrogen applications, (expressed as N), generally vary from about 100 to 500 kg/ha • yr. Because fertilizer is used year after year, it is to be expected that in many areas some of the N, P, or K is carried by infiltrating water downward to the water table, where it can migrate in the groundwater flow regime. For reasons explained in Section 9.3, nitrogen in the form of is generally much more mobile in subsurface flow systems than dissolved species of phosphorus. Cation exchange causes K+ to have low mobility in most nonfractured geologic materials.
Of the three main nutrients in fertilizer, N in the form of is the one that most commonly causes contamination of groundwater beneath agricultural lands. High concentrations have been delineated in extensive areas in many parts of the world, including Israel (Saliternik, 1972), England (Foster and Crease, 1972), Germany (Groba and Hahn, 1972), California (Calif. Bureau Sanitary Eng, 1963; Nightingale, 1970; Ayers and Branson, 1973), Nebraska (Spalding et al., 1978), southern Ontario, and southern Alberta. Many wells in these areas have concentrations that exceed the recommended limit for drinking water. In areas where contamination is areally extensive, fertilizer rather than animal wastes from feedlots or lagoons or septic field seepage is usually identified as the primary nitrogen source. Nitrate is the principal dissolved nitrogen component, with ammonium and organic nitrogen present in much lower concentrations. Although in many aquifers that are contaminated by the concentrations are below the limits recommended for drinking water, it is disturbing to note that gradual increases in have been observed. The widespread use of inorganic fertilizers began after World War II. The major impact on groundwater quality resulting from this change in agricultural practice is probably not yet fully developed. Nitrate contamination is rarely reported at depths of more than about 10–100 m below the water table. As time goes on, however, contamination may extend to greater depth in areas where there are significant downward flow components. For example, in deep wells in California, ranging in depth from 240 to 400 m below ground surface, increased from approximately 1 mg/ℓ in 1950 to 10–17 mg/ℓ in 1962 (Broadbent, 1971). The extent to which denitrification occurs as water moves along regional flow paths is a major uncertainty inherent in predictions of long-term increases in aquifers.
In England, contamination of a large regional carbonate-rock aquifer is widespread. Analysis of the occurrence and movement of in this aquifer is complicated by the fact that is carried in groundwater flowing in a network of joints and solution channels, while some of the is lost from the active flow regime as a result of diffusion into the porous matrix of the limestone (Young et al., 1977). If at some time in the future the concentration in the flow network declines, will diffuse from the matrix back into the flow regime.
Although extensive contamination of shallow groundwater can often be attributed to leaching of fertilizer, in shallow groundwater in large areas in southern Alberta (Grisak, 1975), southern Saskatchewan, Montana (Custer, 1976), and Texas (Kreitler and Jones, 1975) is not caused by fertilizer use. In these areas it appears that most of the is derived by oxidation and leaching of natural organic nitrogen in the soil. The greater abundance and deeper penetration of oxygen into the soil has occurred as a result of cultivation. In some areas the initial turning of the sod as settlers moved on the land was probably a major factor. In other areas continual deep cultivation during the modern era of farming has been a major influence.
In many agricultural areas shallow groundwater has become contaminated locally as a result of leaching of from livestock and fowl wastes. The conversion of organic nitrogen in these wastes to takes place through biochemical processes. Relatively small source areas such as farm manure piles, fowl-waste lagoons, and feedlots contribute to groundwater, but if these contaminant sources are not directly underlain by aquifers, the contamination is rarely very significant. Specific cases of groundwater contamination from animal wastes are reported by Hedlin (1972) and by Gillham and Webber (1969). In agricultural areas contamination of shallow wells by and other constituents commonly occurs because of faulty well construction. If wells are not properly sealed by grout or clay along the well bore above the screen, contaminated runoff can easily make its way to the aquifer zone near the well screen.
Concurrent with the widespread increase in the use of chemical fertilizers since World War II has been the rapid development and use of a multitude of organic pesticides and herbicides. In a report on groundwater pollution in the southwestern United States, Fuhriman and Barton (1971) concluded that pollution by pesticides must be listed as an important potential hazard. However, they obtained no direct evidence indicating significant pesticide contamination of groundwater. Kaufman (1974), in a review of the status of groundwater contamination in the United States, indicates that this conclusion appears to characterize today’s situation—that of a potential but as-yet-unrealized problem. Based on a literature review and field studies in Kent, England, Croll (1972) arrived at a similar conclusion. It is well known from laboratory experiments that many pesticides and herbicides with appreciable solubility in water have significant mobility in some types of geologic materials, particularly clean sands and gravels (Burns and Mclaren, 1975; Adams, 1973). It is not unreasonable to expect that the use of these chemicals in agriculture will eventually cause parts of some aquifers to become contaminated. Davidson et al. (1976) have pointed out that because of the immense size to which the pesticide industry has grown, the problems associated with the disposal of surplus and waste pesticide materials and empty or partially empty pesticide containers has become acute. High concentrations of pesticides in groundwater can result in greater mobility than at lower concentrations. At higher concentrations, the exchange sites are more readily saturated with the pesticide, or the biodegradation capabilities of the medium may be exceeded.
In industrialized countries hundreds of thousands of steel gasoline storage tanks lie buried at filling stations. Many thousands of kilometers of underground pipelines carry petroleum products across continents. Tanker trucks with oil and gasoline are continually on the move. It is not surprising, therefore, that leakages and spills from these sources are an increasing threat to groundwater quality. Most of the buried storage tanks at filling stations were placed in the ground since World War II. Because stringent requirements for tank testing and replacement are only gradually being implemented in most countries, leakage problems caused by older tanks are common, particularly in regions of high water tables and frequent infiltration.
Contamination of groundwater by petroleum products from leaky tanks, from pipelines, or from spills is a much different type of problem than those described elsewhere in this chapter. The major difference lies in the fact that oils and gasoline are less dense than water and are immiscible in water. As a consequence of this, oil or gasoline from leakages or spills migrate almost exclusively in the unsaturated zone. The processes of petroleum movement in the unsaturated zone have been described in detail by Schwille (1967), van Dam (1967), and Dietz (1971). The following discussion is based primarily on these references.
Figure 9.27 illustrates the main subsurface migration stages that occur when oil seeps in the ground. In this case, the hydrogeologic conditions are simple. There is appreciable depth of unsaturated zone beneath the level of oil entry into the system. The term “oil” is used here to refer to both crude oil proper and its liquid derivatives, such as gasoline.
In the first migration stage, the oil movement is primarily downward under the influence of gravitational forces. During this seepage stage, capillary forces produce some lateral migration. This causes a zone, referred to as the oil wetting zone, around the core of the infiltration body. It is comparable in origin to the natural capillary fringe on the water table. In the oil wetting zone, the degree of oil saturation decreases outwardly and capillary forces (surface tension) are dominant. In the main seepage zone, only gravitational forces exist.
Downward seepage of oil ceases when the seepage front reaches the water table. Although it might be expected that the oil would spread laterally on top of the capillary fringe rather than along the water table, experimental and field evidence indicates that considerable migration occurs within the capillary fringe at or very near the water table. Since oil is immiscible in water and is less dense than water, it may slightly depress the water table. Except for small amounts of hydrocarbons that go into solution, the oil does not penetrate below the water table. As oil accumulates on the water table, the oil zone spreads laterally, initially under the influence of gradients caused by gravity and later in response mainly to capillary forces. Capillary spreading becomes very slow and eventually a relatively stable condition is attained. In theory, stability occurs when a condition known as residual oil saturation or immobile saturation is reached. The experience of oil production engineers is that below a certain degree of saturation, oil is held in a relatively immobile state in the pore spaces. If the percent oil saturation is reduced further, isolated islands or globules of oil become the dominant mode of oil occurrence. Over the range of pressure gradients that can occur, these islands are stable. As the mass of oil spreads laterally due to capillary forces, the residual oil saturation condition must eventually be attained, provided that the influx of oil from the source ceases. This is referred to as the stable stage.
When the oil spill or leakage volume is small relative to the surface available for contact as the oil moves through the zone above the water table, the oil migration zone may attain residual saturation and become immobile before penetrating to the water table. The volume of porous medium required to immobilize a given amount of oil depends on two factors: the porosity and the nature of the hydrocarbons that comprise the oil. The volume B of porous geologic materials that is required to immobilize a spill or leakage volume can be estimated from the relation
where B0 is the volume of oil entering the system, n the porosity, and S0 the residual oil saturation. If the depth to water table and values for n and S0 are known, this relation can be used to estimate the likelihood that spilled oil will penetrate to the water table (American Petroleum Institute, 1972). Van Dam (1967) presents equations that describe the shape of the stable layer of oil if penetration to the water table occurs. In practice, however, it is generally not possible in field situations to obtain sufficient data on the distribution of relative permeabilities for more than a qualitative analysis to be made (Dietz, 1971). Laboratory model experiments by Schwille (1967) have demonstrated that minor differences in permeabilities laterally or vertically can cause strong distortions in the shape of the oil migration zone.
Because oil leakages or spills usually do not involve large fluid volumes of oil, and because migration is limited by residual oil saturation, one might expect that oil is not a significant threat to groundwater quality. This is unfortunately not the case. Crude oil and its derivatives contain hydrocarbon components that have significant solubility in water. In general, the lighter the petroleum derivative, the greater is the solubility. Commercial gasoline, for example, has solubility of 20–80 mg/ℓ. It can be detected by taste and odor at concentrations of less than 0.005 mg/ℓ (Ineson and Packham, 1967). Because the solubility of the lighter hydrocarbons greatly exceeds the concentration levels at which water is considered to be seriously polluted, it is not difficult to envision situations where the effect of hydrocarbon dissolution is of much greater concern in terms of groundwater quality than the localized immobile zone of immiscible hydrocarbons on and above the water table. For example, in the situation represented in Figure 9.28, the lateral flow of groundwater beneath the zone of immobilized oil could cause soluble hydrocarbons to be transported large distances along the groundwater flow paths.
In situations where oil penetrates to the water-table zone and then spreads out and becomes relatively immobile, the effect of water table fluctuations can be important. If the water table falls, much of the oil remains in the newly created zone as a thin coating on the surface of the porous medium. This film is not removed by water flushing or air ventilation. The problem can be ameliorated by the effect of bacteria. There exist species of aerobic and anaerobic bacteria that grow rapidly in the presence of oil or gasoline if the other necessary nutrients are also available. In favorable circumstances, bacterial oxidation can consume much of the oil or gasoline that accumulates above the water table as a result of leakages or spills. A fluctuating groundwater table is believed to promote the processes of biodegradation.
Of the many examples of petroleum contamination of groundwater that have been reported, the case history described by McKee et al. (1972) is particularly illustrative of the problems and processes involved and remedial measures that can be used to minimize the damage to subsurface water quality. Control and remedial procedures are described by the American Petroleum Institute (1972).
It has been several decades since nuclear engineers and scientists at an Idaho research station watched four household light bulbs flicker to life as a result of man’s first generation of electricity from atomic energy. From this modest beginning, nuclear-power-generating facilities have grown to the point where they now produce more than 15% of the electricity in the United States and Canada and larger percentages in some European countries. By the turn of this century it is expected that in North America and Europe, the percentages will be much larger. Feared by many as a threat to mankind’s future and hailed by others as the answer to the world’s energy problems, nuclear generation of electricity has sparked controversy around the globe. At the present time there are several uncertainties inherent in activities associated with the generation of nuclear power. One of them is man’s capability to safely isolate radioactive wastes from the biosphere for long periods of time. Because of its hydrogeologic nature, this nuclear-power-related topic is worthy of discussion in this text.
The hydrogeologic aspects of the nuclear power industry will be considered within the framework of the nuclear fuel cycle. This expression refers to all the stages in the nuclear power industry in which nuclear fuel is developed and used and in which radioactive wastes are generated. This includes uranium mining, milling, refining, uranium enrichment, fuel fabrication, fuel consumption in reactors, fuel reprocessing, waste solidification, and burial of solidified waste or unreprocessed spent fuel in deep geological repositories. The “front end” of the nuclear fuel cycle involves mining and milling of uranium ore. An undesirable by-product of these activities is the production of large volumes (hundreds of millions of cubic meters per year in North America) of waste rock from mining and tailings from milling. Waste rock and tailings are usually placed in piles on the land surface or as fill material in topographic depressions confined by small earth embankments or dams. Because they contain isotopes of uranium, thorium, and radium, waste rock and tailings are a form of solid low-level radioactive waste. Radium 226 (226Ra), with a half-life of 1620 years, poses the greatest environmental hazard. Table 9.1 indicates that the maximum permissible concentration of 226Ra in drinking water is 3 pCi/ℓ, which is equivalent to 10–9 mg/ℓ. This concentration is so small that it is orders of magnitude below the maximum permissible concentrations for trace metals such as lead, mercury, or cadmium (Table 9.1). Extremely small amounts of 226Ra leached from waste rock or tailings into groundwater can therefore cause the water to be unfit for human consumption. Uranium mining in North America generally occurs in areas remote from population centers and from industrial or agricultural developments. In these areas, groundwater quality has, until recently, not been a subject of significant concern. The extent to which 226Ra and other hazardous constituents from waste rock or tailings are entering groundwater and their fate within groundwater flow systems is not known. We can expect, however, that in the next decade hydrogeological factors will play a much greater role in the design and evaluation of disposal sites for uranium mining and milling wastes than has previously been the case.
The next stage in the nuclear fuel cycle is uranium refining, a process in which the mill product is upgraded in preparation for uranium enrichment into nuclear fuel (the US. and European approach) or unenriched fuel fabrication (the Canadian approach). In the refining process, small quantities of solid or semisolid, low-level radioactive wastes are generated. The chemical nature of these wastes varies from refinery to refinery, but the wastes generally contain 226Ra, 230Th, and 238U in what are normally small but significant concentrations. As in the case of mining and milling wastes, 226Ra is the isotope of main concern. The refinery wastes are assigned to near-surface burial grounds that are located near the refineries. After more than 20 years of use, the burial ground at the principal uranium refinery in Canada (Port Hope, Ontario) was found in the mid-1970’s to be emitting leachate with 226Ra, as well as other nonradioactive contaminants. Although no aquifers were in jeopardy, remedial measures include excavation of the wastes with subsequent reburial at a site with hydrogeologic capability for longer-term isolation of the wastes from the biosphere.
The next major waste-generation stage in the nuclear fuel cycle is the operation of nuclear reactors for power production, weapons production, or research. In this stage, low-level solid radioactive wastes in the form of discarded equipment, assorted slightly radioactive refuse, and ion-exchange materials from decontamination facilities are produced. These wastes are known as reactor wastes. The term “low-level” is used here in a qualitative sense to distinguish these wastes from the very highly radioactive materials in used nuclear fuel or material derived directly from the used fuel.
Since the start of nuclear power production on a commercial basis in North America, the total volume of accumulated reactor wastes has amounted to about 40,000 m3 (as of 1975). These wastes have been emplaced in shallow burial sites at 11 major locations in the United States and four major locations in Canada. In the United States, the volume is expected to rise to 50,000 m3 by 1980 and then to more than 300,000 m3 by the year 2000 (Nuclear News, 1976). By use of existing and economically viable technology, the projected volumes can be reduced by a factor of 2 or 3. It is hoped, of course, that new waste processing methods will be developed to provide further volume reduction.
In spite of the improvements that may develop, the volumes of reactor waste are expected to be enormous in comparison with the volumes that have been handled in the past. Since the standard method of managing reactor wastes is to assign them to shallow burial sites, these mounting waste volumes can be viewed as a potential source of contamination to groundwater and other environments. The past history of shallow low-level waste burial in the United States is less than satisfactory. Of the 11 existing sites at which radioactive wastes resulting from commercial power production have been buried, three are leaking radioactive constituents to the environment (Ground Water Newsletter, 5, no. 3, 1976). Although at present this leakage to subsurface flow systems does not present a hazard to potable water supplies, it is striking evidence that undesirable consequences of inadequate hydrogeologic studies of waste management sites can become evident many years or decades after site usage begins. There is now little doubt that at the time most of these sites were established many years ago, more attention was given to the economics of handling the wastes, the ready availability of land for burial use, and proximity to transportation routes than was given to the ultimate fate of the wastes. With these lessons in hand, the problem facing hydrogeologists now is to ensure, through use of proper site search and evaluation methodologies, that future sites for shallow burial of solid low-level wastes have adequate radionuclide containment capabilities and that proper subsurface monitoring facilities are installed and operated.
Reactor wastes contain a variety of radionuclide species, with half-lives ranging from seconds to many decades or longer. Of these nuclides, 137Cs, 90Sr, and 60Co, with half-lives of 28, 33, and 6 years, respectively, are usually regarded as posing the most significant environmental hazard. Wastes with these radionuclides need several hundred years to decay to very low radioactivity levels.
Figure 9.29 illustrates several types of waste-burial options. Although other situations exist, these will serve as a basis for discussion of some of the main concepts used in the development of burial facilities. In Figure 9.29(a), the wastes are placed in strong engineered containers constructed of materials such as concrete and steel situated on the ground surface. In these containers they can remain in storage in areas from which the public is excluded. Deterioration of the containers can be readily monitored. If problems arise, the containers can be repaired or the Wastes can be placed in new containers, provided, of course, that a responsible organization remains in charge of the facilities. The facilities shown in Figure 9.29(b) are similar, except that the storage containers are covered with earth materials. If these materials are properly designed, they will protect the containers from weathering and thereby extend their life expectancy. The earth materials can be thought of as an engineered hydrogeological environment.
In Figure 9.29(c) and (d), the wastes are stored in containers situated a few meters below ground surface, either above or below the water table. In the case shown in Figure 9.29(c), earth material from the excavation is used as backfill around the containers. In Figure 9.29(d), part of the till in the excavation is designed to provide enhanced containment capabilities for the system. If the containers are located above the water table and if there is good reason to believe that during periods of water-table fluctuation, the water table will not rise into the burial zone, there will be little possibility of radionuclides escaping into the environment. If the water table fluctuates within the zone of burial, the containers are subjected to variable hydraulic and geochemical conditions. Their life expectancy and the fate of radionuclides in the event of container failure is much less certain than in the case of burial entirely above the water table.
In Figure 9.29(e) and (f), the containers are buried in large holes about 10 to 20 m deeper than in the previous examples. In nonarid regions the permanent water table in these situations would normally be above the containers. In the case depicted in Figure 9.29(e), the hole is simply backfilled with material originally removed from the site. In Figure 9.29(f) the excavated zone around the containers is packed with geological material such as bentonitic clay chosen to improve the long-term containment capabilities of the burial facility.
Nearly all of the burial sites for reactor wastes in the United States and Canada are in the category represented by Figure 9.29(e), with the water table within or just slightly below the burial zone. Some of the wastes have been placed directly in the ground without the protection of watertight containers. Most of the sites are located in poor hydrogeologic settings. It is not surprising, therefore, that subsurface migration of radionuclides from the burial zones is a common circumstance.
To avoid problems of subsurface radionuclide migration at future sites, numerous investigators have proposed that future sites be located in hydrogeologic environments that are shown to have long-term containment capability. To achieve this capability, the site should have the following characteristics: geomorphic and structural stability, isolation from fractured bedrock or other subsurface flow regimes that are too complex for development of reliable pathway analyses (i.e., the site should have a simple hydrogeologic framework), absence of subsurface flowlines that lead directly to the biosphere or to subsurface zones of potable Water, and low predicted radionuclide velocities resulting from favorable combinations of groundwater velocity and chemical retardation. In addition to hydrogeologic criteria such as these, various investigators have indicated that the water table should be deep enough to permit waste burial to occur entirely in the unsaturated zone (Cherry et al., 1974). The predicted upper level of the range of water table fluctuation over many centuries should be below the bottom of the burial zone. The criteria outlined above, if adhered to, would lead to the development of waste-burial facilities that would provide considerably better long-term containment than is the case for existing sites. Unfortunately, in most humid and semihumid regions of North America, this desired ideal combination of hydrogeologic properties is rare, or even nonexistent, within areas that also meet social and economic criteria. Water-table depth is generally shallow, which prevents establishment of burial zones at appreciable depth below ground surface. Wastes that are buried within a few meters, or even within 5–10 m of ground surface, are generally not considered to be in isolation from future generations during the hundreds of years or more that will be necessary for decay to reduce radioactivity in the waste to low levels.
As an alternative approach to the siting of burial grounds for low-level solid radioactive waste, Cherry et al. (in press) proposed that in humid and semihumid regions, burial zones be located in unfractured clayey aquitards in the manner represented in Figure 9.29(f). According to this scheme, burial would occur at the bottom of large (2–5 m in diameter) auger holes more than 15 m below ground surface. The burial zone would be at considerable depth below the water table and below the zone of active weathering. The wastes would therefore be isolated from the biosphere in a hydrogeologic environment in which groundwater velocity is extremely low and chemical retardation is great. Accidental unearthment of the waste by future generations would be much less probable than is the case for shallow, above-water-table waste burial. For this approach to be evaluated in detail it will be necessary to direct research toward the hydrogeologic properties of clayey aquitards such as clayey till, glaciolacustrine clay, and soft shale.
For additional information on the hydrogeologic aspects of the management of low-level solid radioactive waste (primarily reactor waste), the reader is referred to Peckham and Belter (1962), Richardson (1962a, 1962b), Mawson and Russell (1971), Cherry et al. (1973), and Pferd et al. (1977).
In the nuclear reactors that are currently used for power production, fuel rods composed of solid uranium oxide undergo fission reactions that release heat and decay particles. After a period of time in the reactor, the fuel rods are replaced. The spent fuel contains a wide variety of toxic radioactive isotopes produced from the uranium and from other elements. The ultimate fate of these man-made radionuclides is the core of what has become known as the high-level radioactive waste disposal problem. Numerous proposed solutions to this problem have been suggested, each with the objective of isolating the radionuclides from the biosphere for the life span of their radioactivity. This is the final stage of the nuclear fuel cycle.
Disposal options such as burial within the Antarctic ice cap, emplacement in the ocean floor at locations where natural burial beneath migrating continental plates will occur, and rocket transport beyond the earth’s gravity held have been excluded because of impracticalities during the next few decades. It is now generally hoped that a satisfactory solution can be obtained by emplacing the radioactive material in an engineered repository in geologic strata in which it will be isolated from zones of active groundwater flow. This approach is commonly referred to as terminal storage. This implies that for a generation or two the repository environment will be monitored, and that if all goes well the wastes will then be regarded as having been permanently disposed.
In the United States, four main hydrogeologic possibilities are being investigated for suitability for repository development. These are (1) deep salt beds, (2) deep crystalline igneous rocks, (3) deep shale strata, and (4) thick unsaturated zones in arid regions. Because of differences in climate, option (4) is not available in Canada or in European countries. The most critical question in the evaluation of these options is whether or not the wastes will be isolated from the biosphere for periods of time that are considered to be acceptable.
The waste will contain numerous radionuclide species, but by the year 2020, 99% of the projected accumulation of radioactivity will be due to the presence of 90Sr and 137Cs (Gera and Jacobs, 1972) which have half-lives of 28 and 33 years, respectively, and which will decay to very low levels within about 1000 years. Much longer periods of time, however, are required for decay to low levels of long-lived transuranic nuclides in the waste, namely, 238Pu, 239Pu, 240Pu, 241Am, and 243Am, with half-lives ranging from 89 years for 238Pu to 24,000 years for 239Pu. Radioactive decay of these elements in the waste produces other radionuclides, known as daughter products (237Np, 226Ra, 129I, 99Tc, and others). If these are taken into account, the material will remain hazardous for millions of years, although at much lower radioactivity levels than will occur during the first thousand years.
The radionuclides can be placed in the repository in their original form as spent fuel or they may be incorporated into other materials after the spent fuel has been reprocessed. Reprocessing is a chemical treatment in which spent fuel is dissolved in acid and plutonium is separated from the other radionuclides. Plutonium is viewed by the nuclear power industry as a valuable commodity because it can be used to produce power in fast-breeder reactors. After the extraction of plutonium, a hot, highly radioactive waste solution containing the remaining radionuclide species and some plutonium residue remains as waste. It is now generally agreed by the nuclear regulatory agencies in the various countries working on the problem that these wastes must be solidified and incorporated into solid relatively inert materials such as ceramics or glass. This must be done before proceeding with commitment to any scheme for long-term subsurface storage or disposal. From the chemical reprocessing plant the waste will proceed, after a period of interim storage, through a solidification plant. After being solidified and encapsulated, the waste will then be ready for emplacement in a subsurface geological repository. Although reprocessing removes most of the plutonium from the spent fuel, the waste that remains after reprocessing is still highly radioactive. From a hydrogeological viewpoint, isolation of spent fuel or isolation of solidified waste from reprocessing are fundamentally the same problem.
For a subsurface repository to be viewed as satisfactory within the abovementioned time framework, it must be capable of protecting the wastes from the effects of landscape erosion caused by wind, water, and even glaciers. It must be located in an area that does not have a significant seismic hazard or potential for volcanic activity. The protective materials within which the wastes are placed in the repository and the hydrogeologic environment outside the repository must be capable, within an exceptionally high degree of predictive confidence, of preventing migration of radionuclides in groundwater from the repository into the biosphere. It is this latter criterion that is the most difficult to establish at the level of confidence that is required. Never before in the history of mankind have engineers and scientists been asked to provide safety analyses relevant to such a long period of time. The feasibility of achieving long-term waste isolation in each of the four hydrogeologic repository options listed above is currently under evaluation. The deep-salt option is discussed by Bradshaw and McClain (1971) and by Blomeke et al. (1973). The potential for repository development in shale is described by Ferro et al. (1973). Winograd (1974) reviewed the hydrogeologic aspects of the arid region/unsaturated zone option. For a broader view of the high-level radioactive-waste-disposal problem, the reader is referred to Kubo and Rose (1973) and Cohen (1977).
Injection of liquid wastes, mainly of industrial origin, has been widely adopted as a waste disposal practice in North America. The purpose of this procedure is to isolate hazardous substances from the biosphere. As the discharge of pollutants to rivers and lakes has become increasingly objectionable, and as legislation for protection of surface water resources has become more stringent, the use of deep permeable zones for liquid waste disposal has become an increasingly attractive waste management option for many industries. Inventories of industrial liquid-waste injection wells in the United States were conducted in 1964, 1967, 1968, 1972, and 1973. During this period the number of waste injection wells increased from 30 in 1964 to at least 280 in 24 states in 1973 (Warner and Orcutt, 1973). In Canada in 1976 at least 80 injection wells were in use. Injection wells used for return of brines extracted during oil or gas field pumping are normally not categorized as waste injection wells. There are more than 100,000 of these wells in North America. Chemical, petrochemical, and pharmaceutical companies are the largest users of waste injection wells. Other important users are petroleum refineries, gas plants, steel mills, potash mines, uranium mills, and processing plants. In Florida, Hawaii, Louisiana, and Texas, injection of sewage effluent into saline-water aquifers occurs on a minor scale. Nearly all the waste injection wells are in the depth range of 200–4000 m. Most are between 300 and 2000 m deep. The injection zones are generally located in sandstones, carbonate rocks, and basalt.
Most injection wells are operated at injection pressures less than 7 × 103 kN/m2. The trend in recent years is toward lower injection pressures and injection rates in the range of 500–1400 ℓ/min. The effect of an injection well on the hydrodynamic conditions in a hypothetical horizontal aquifer in which there is regional flow, are shown in Figure 9.30. The injection well causes a mound in the potentiometric surface. The mound extends unsymmetrically in the direction of regional flow in the aquifer. As injection continues the areal extent of the mound spreads to occupy an ever-increasing area. This process can be viewed as the inverse of the effect of a pumping well in a confined aquifer, and in fact is described mathematically by the same equations, modified for the effect of injection rather than pumping (Warner, 1965).
If injection wells are located close together, the potentiometric mounds coalesce in a manner analogous to the drawdown interface in fields of pumping wells. The spread of the front of the potentiometric mound is very rapid in comparison to the spread of the zone of injected waste. The front of the potentiometric mound spreads by pressure translation. The front of the waste zone spreads as volume displacement occurs. The waste zone spreads in direct proportion to the cumulative volume of waste that is forced into the aquifer. The interface between the formation water and the waste will be gradational as a result of dispersion.
Based on reviews of status and impact of waste injection practices in the United States and Canada, Warner and Orcutt (1973) and Simpson (1976) concluded that documented cases of even minor disposal system failure and related contamination of surface and near-surface waters are rare. This may not continue to be the case, however, as waste injection becomes an increasingly common waste disposal practice and as the length of time that strata have received wastes increases. One of the few reported instances in which waste injection has caused surface contamination occurred in southern Ontario near Sarnia, where most of the Canadian petrochemical industry is located. Contamination was caused by live injection wells, the first of which was drilled in 1958 and the others in 1960. Refinery caustic and phenolic waters were injected down the wells at rates less than 400 ℓ/min into carbonate rock strata at depths between 200 and 260 m below surface. In the late 1960’s and early 1970’s, phenol occurrences in fluids reaching the surface in the Sarnia district were observed (Simpson, 1976). It is believed that the contamination occurred because of upward leakage of the wastes through abandoned unplugged wells. There may be as many as 30,000 unplugged wells in southwestern Ontario in the vicinity of Sarnia (van Everdingen and Freeze, 1971). Many of these were drilled decades ago before plugging of abandoned wells was required. The hazard represented by unplugged wells in areas of waste injection is a particularly insidious one because the location of many of the wells is unknown. Some no longer even exist at ground surface but provide vertical connections below ground surface. There may be more than 1 million unplugged, unlocated wells in North America. In regard to the long-term effect of deep-well injection of noxious wastes, van Everdingen and Freeze (1971) have suggested that vertical connections provided by unplugged wells may well be the most important hazard.
Another major hazard associated with the practice of waste injection is the inducement of earthquakes as a result of increasing pore-water pressures along faults. This topic is discussed in Chapter 11.
As a concluding statement on deep-well disposal, the comments by A. M. Piper (1970) of the United States Geological Survey seem appropriate:
In its predilection for grossly oversimplifying a problem and seeking to resolve all variants by a single massive attack, the United States appears to verge on accepting deep injection of wastes as a certain cure for all the ills of water pollution (p. 2).
Injection does not constitute permanent disposal. Rather, it detains in storage and commits to such storage-for all time in the case of the most intractable wastes-under-ground space of which little is attainable in some areas, and which definitely is exhaustible in most areas (p. 6).
Admittedly, injecting liquid wastes deep beneath the land surface is a potential means for alleviating pollution of rivers and lakes. But, by no stretch of the imagination is injection a panacea that can encompass all wastes and resolve all pollution even if economic limitations should be waived. Limitations on the potentials for practical injection are stringent indeed—physical, chemical, geologic, hydrologic, economic and institutional (including legal) limitations (p. 5).
There are many other sources that contribute contaminants to the groundwater zone. In the northern United States and in Canada large quantities of salts are applied to roads to combat adverse ice conditions during the winter months. Contamination of shallow aquifers along roads that receive salt is not uncommon in these regions. Since salts such as NaCl and CaCl2 are highly soluble and relatively mobile in groundwater, there is little that can be done to prevent this situation, other than to decrease the amount of salt usage.
Activities of the mining industry are another potential cause of groundwater contamination. The effects range from changes in groundwater chemistry caused by mining to infiltration of leachate from tailings and other wastes. The extensive occurrence of acid water drainage from abandoned coal mines in the Appalachian region of the United States is the most visible example of adverse effects of mining on groundwater and surface water.
Seepage from industrial waste lagoons is another cause of groundwater contamination. Across North America there are thousands of artificial ponds and lagoons that contain countless types of liquid wastes. In many cases the lagoons are not lined with impermeable barriers, thereby providing opportunity for seepage of wastes downward into the subsurface environment. In situations where potable water aquifers are located nearby, this can cause serious problems. Many years may pass, however, before the extent of the problem becomes evident.
In some regions urbanization is spreading into the recharge areas of major aquifers. Even if centralized sewage treatment facilities rather than septic systems are used, urban activities produce numerous sources of contamination to the ground. Prediction of the long-term effect of urbanization on groundwater quality is a difficult task but a necessary one if we are to develop methods of land use planning that will minimize adverse impacts on the groundwater environment.
As an introduction to the literature on groundwater contamination and related topics, the reader is referred to Hall (1972), Summers and Spiegal (1974), Todd and McNulty (1976), and Wilson et al. (1976). An indication of the nature and regional extent of groundwater contamination in the United States is presented in the reviews by Fuhriman and Barton (1971), Scalf et al. (1973), Miller et al. (1974), and Scalf (1977). A summary of waste disposal practices and their effects on groundwater in the United States has been described by U.S. Environmental Protection Agency (1977). A review of recent research activities related to chemical problems in hydrogeology has been presented by Back and Cherry (1976).
BAETSLÉ, L. H. 1969. Migration of radionuclides in porous media, Progress in Nuclear Energy, Series XII, Health Physics, ed. A. M. F. Duhamel. Pergamon Press, Elmsford, N.Y., pp. 707–730.
CHERRY, J. A., R. W. GILLHAM, and J. F. PICKENS. 1975. Contaminant hydrogeology: Part 1: Physical processes, Geosci. Can., 2, pp. 76–84.
FRIED, J. J. 1976. Ground Water Pollution. Elsevier, Amsterdam, pp. 1–47.
OGATA, AKIO. 1970. Theory of dispersion in a granular medium. U.S. Geol. Surv. Prof. Paper 411–1, p. 134.
U.S. Environmental Protection Agency. 1977. Waste disposal practices and their effects on ground water. The Report to Congress, pp. 81–107.