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Ground Water Recharge Using Waters of Impaired Quality 3 Soil and Aquifer Processes The desired role of the unsaturated soil zone (vadose zone) in a recharge system is a straightforward one: remove or reduce chemical and biological constituents that pose a potential health risk before the recharge water enters the ground water. Unfortunately, the processes by which removal occurs are not completely efficient in a natural setting, and not all constituents are retained or degraded to the same extent. Moreover, management strategies that may enhance the removal of one chemical or pathogen may actually decrease the efficiency of removal of another. This chapter describes the major processes by which soils and aquifers can remove chemicals and pathogens. The processes that occur in soils and aquifers are chemical-, pathogen-, and soil-specific, depending on a number of conditions that can vary significantly from site to site and from one compound to another. Thus this chapter reviews the principal processes governing transport and fate in sod first in a generic fashion, and later with respect to the behavior of specific chemical or pathogen groups. The soil properties that are important in a properly functioning soil-aquifer treatment (SAT) system are reviewed, and properties or processes that can create difficulties for chemical and pathogen removal processes are identified. Three separate issues of concern are addressed: (1) the overall effectiveness of the SAT system and its ability to ensure that the quality of the underlying water resources will not be impaired, (2) the long-term sustainability of the system, and (3) the feasibility of monitoring to determine both the performance and the safety of the operation.
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Ground Water Recharge Using Waters of Impaired Quality CONDITIONS INFLUENCING PRETREATMENT The soil and aquifer properties, recharge method, type of wastewater, and ultimate destination and intended use of the recovered ground water collectively dictate the degree of pretreatment required before recharge. In addition, recharge rates, regardless of the method used, depend to some degree on the quality of the source water that is recharged. Cost-effective recharge operations are achieved through tradeoffs between maximizing recharge rates and minimizing treatment costs. The selection of a wastewater treatment process depends on the characteristics of the wastewater, the required effluent quality, and the cost of the selected treatment option. When considering a water source as a potential for recharge, the likely variety and concentration of contaminants also need to be considered. The use of waters of impaired quality as sources for recharge has raised questions about the level of pretreatment necessary prior to recharge. The most conservative approach is to assume that passage through the soil to the aquifer and through the aquifer to the withdrawal location provides no treatment and that pretreatment processes therefore should improve the source water to the quality level needed by the end user. This approach, however, can lead to expensive systems. Soil-aquifer processes can be counted on to provide treatment benefits. Soil Properties The ideal porous medium for an SAT operation is one that allows rapid infiltration and complete removal of all constituents of concern. Unfortunately, no such medium exists because the attributes required to achieve one goal hamper the achievement of the other. In surface soil, coarse-textured materials are desirable for infiltration because they transmit water readily; however, the large pores in these soils are inefficient at filtering out contaminants, and the solid surfaces adjacent to the main flow paths are relatively nonreactive. In contrast, fine-textured soils are efficient at contaminant adsorption and filtration, but they have low permeability and their small pores clog easily. Structured soils con-mining biological channels (e.g. worm holes or root holes) or cracks are permeable, but the large flow paths completely dominate the movement of material and much of the matrix is bypassed. The best choice for an SAT soil is therefore a compromise, such as a fine sand or a sandy loam with relatively little structure (Bouwer, 1985). Nature of Recharge Operation Of the two basic methods of artificial recharge—surface infiltration and well recharge—well recharge requires water of much higher quality. This is particularly true where an aquifer composed of granular rocks is to be recharged.
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Ground Water Recharge Using Waters of Impaired Quality The flux per unit area of rock surface at the point of recharge is generally much greater for injection wells than for surface infiltration systems. Consequently, equivalent amounts of clogging material and nutrients for biological growth result in more severe operational problems in recharge wells. Surface infiltration systems can function effectively over a broad range of water quality and can readily tolerate variations in quality, although, in general, higher quality results in higher infiltration rates. Injection well operations, on the other hand, are much more sensitive to quality variations, except in the case of conduit-flow rocks such as solution-riddled limestone or fractured rock. Recharge wells require water that is virtually free of suspended matter, especially where granular aquifers are to be used. Any injected suspended matter accumulates at and near the well-aquifer interface, and because this circumferential area is limited and the flux through it is large, rapid hydraulic head loss and reduction in injection capacity occur quickly. The clogging caused by the accumulation of the suspended material and biological growth must be remedied by pumping the well for backflushing, by surging or jetting, or at times by dosing the well with chemicals to loosen and/or dissolve the accumulated clogging materials. Removal of suspended solids (to very low levels, i.e., less than 1 milligram per liter) in the recharge water is required for successful operation of recharge wells, except where karstic or fractured rock aquifers are to be recharged. Wells recharging solution-riddled or fractured rock aquifers can tolerate water having higher levels of suspended solids without experiencing severe operational problems. Wastewater Composition Quality parameters of concern in the operation of surface infiltration systems are the suspended solids (SS) and total dissolved solids (TDS) content as well as the concentrations of nutrients that stimulate biological growth and of major cations such as calcium, magnesium, and sodium, which determine the sodium adsorption ratio (SAR). The suspended solids concentration of the recharge water is the most important factor. Suspended solids settle out or are filtered from the water and accumulate on the soil and/or at a short distance below the water-soil interface. The accumulation reduces the permeability of the soil and retards movement of water into the subsurface. Swelling and deflocculation of clay minerals contained in the aquifer can occur if the recharge water contains a higher ratio of monovalent to divalent cations (higher SAR) than does the native water. Clay dispersal can also occur if freshwater is recharged into a saline aquifer. The combination of low TDS with high SAR causes clays to disperse. If clays disperse, infiltration rates drop. Chemical reactions between the recharge water and the native ground water can cause precipitates to form and these can clog pores and reduce injection capac-
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Ground Water Recharge Using Waters of Impaired Quality ity. Compounds of carbonates, phosphates, or iron oxides are the most likely to cause such problems. To avoid problems caused by clay dispersal or chemical precipitates, an evaluation of the chemical compatibility of the recharge water with the aquifer materials and the native ground water should be conducted. The evaluation may indicate the need to chemically modify the recharge water to avoid clogging problems caused by chemical incompatibility. Water recharged by wells must be free of entrained or dissolved gases that may evolve when cold recharge water is injected into warmer ground water. Entrained air or gases that come out of solution will reduce the aquifer's hydraulic conductivity and, consequently, the injection capacity of the well. Bacterial growth is yet another concern with regard to clogging of recharge wells. Removal of nutrients and biodegradable matter from the recharge well and disinfection of the water minimize the potential for bacterial growth in the immediate vicinity of the recharge well. When the recharge water contains a chlorine residual, well clogging is slower (Vecchioli et al., 1980). An alternative to continuous chlorination of the recharge water is to backpump the well frequently (about once a day) or to dose the well heavily with chlorine periodically to destroy the bacterial growth and then backpump the well to remove the spent chlorine solution and organic residue. Recharge wells ending in the vadose zone (dry wells) cannot be redeveloped readily because it is not possible to pump water from them to remove accumulated suspended matter or other clogging materials. Therefore, water recharged to dry wells must be free of suspended solids and not cause clay dispersal or bacterial growth. On the other hand, because of the generally shallow depths of dry wells, replacing clogged wells is considerably less costly than for wells injecting water below the water table. GENERAL DESCRIPTION OF SUBSURFACE PROCESSES Vadose Zone Processes The ultimate goal of a ground water recharge project is to resupply the subsurface with water that does not impair the quality of the underlying resource. Thus, the role of the unsaturated or vadose zone in recharge systems is to help filter out or transform harmful constituents in the soil solution as recharge water moves through the soil matrix en route to the aquifer. The vadose zone is a much more complex transport medium than an aquifer, for several reasons. Because only part of the void space is failed with water, chemicals with a significant vapor pressure can move in the gas phase as well as in solution. The water flow rate can vary significantly. The resistance offered by the vadose zone to the flow of water through a given local soil volume is a nonlinear function of the water content, whereas in the saturated zone, it is a constant. The temperature varies in the surface regime in response to the cyclic
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Ground Water Recharge Using Waters of Impaired Quality inputs of radiant energy, and the composition of the air, solid, and solution phases of the soil is also dynamic, causing spatial and temporal variations in the chemical and biological reactions that transform chemicals in the vadose zone. Also, the amount of water retained against gravity varies significantly with soil texture. Coarse-textured, sandy soils may hold as little as 10 to 20 percent of water-saturation after drainage becomes insignificant, while fine-textured silts or clays may hold as much as 90 percent. Restricting layers comprised of clay lenses or cementing agents can retard drainage greatly, even in otherwise permeable media. Soils that retain extensive water are prone to aeration problems. There are many different processes that can remove chemicals or pathogens from the recharge water as it flows through the vadose zone. Some chemicals volatilize and escape to the atmosphere. They can be chemically or biologically transformed to a new form that may or may not be toxic. They can attach to stationary soil mineral or organic surfaces or precipitate out of solution. They can form complexes with dissolved constituents or particulate matter in solution, thereby reducing their attraction to the soil solid phase and enhancing their mobility in solution. Large pathogens such as parasites, some bacteria, and colloidal material containing contaminants can be filtered out of solution by narrow soil pores, a process that slowly clogs the medium and eventually reduces its permeability if the contaminants are not biodegraded. Viruses can be retained by soil solid phases and inactivated by reactions occurring in the soil. Aquifer Processes Chemicals or pathogens that are still present in solution when the recharge water reaches the aquifer are subject to many of the same processes that occur in the vadose zone, with several exceptions. The biological activity in ground water is much slower than in the near-surface zone, so degradation is greatly reduced. Water fills all of the pore spaces within the ground water zone, so the only place where volatilization can occur is in the capillary fringe above the water table interface. Aquifers used for ground water recharge projects are generally much more coarse textured than soils, so colloids and large pathogens, should they still be present in ground water, are not as easily filtered out of solution as they axe in surface soil. Volatilization Volatilization refers to the evaporation of chemical vapor from soil or water bodies and its subsequent loss to the atmosphere. Many organic compounds are volatile in water, as are some nitrogen compounds (e.g., ammonia, nitrous oxide) generated from biological transformations. In addition, some inorganic chemicals (e.g., selenium compounds) may be rendered volatile through biological reactions.
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Ground Water Recharge Using Waters of Impaired Quality The chemical characteristic that is most indicative of a compound's volatility is its Henry's constant, which is the ratio between the vapor pressure and the solubility of the pure chemical in water. It may also be expressed in dimensionless form as the ratio between the chemical concentration in the gaseous and aqueous phases, where Cg is the mass of chemical vapor per unit volume of air and Cw is the mass of chemical dissolved per unit volume of aqueous solution. For a given chemical, the extent of volatilization loss is very dependent on the soil and atmospheric conditions. The primary factor determining loss is the air phase concentration maintained at the interface with the atmosphere. In general, volatilization is greatly reduced in soil compared to water because the soil solid phase retains the chemical mass, thereby reducing its vapor pressure. In addition, the soil can offer substantial resistance to the transport of chemical from the soil profile to the surface, particularly if the soil is wet and little upward flow of water is occurring. Therefore, in a typical SAT process where recharge water is ponded over the surface for prolonged periods of time, the primary route for volatilization loss will be from the surface of the standing water. During the drainage cycle when the soft becomes unsaturated, volatile constituents in solution near the surface can also evaporate and escape to the atmosphere. Volatilization from Surface Water Volatilization from standing water can be represented conceptually as a two-film resistance model, in which the dissolved compound moves from the bulk fluid through a liquid film to the evaporating surface, and then diffuses through a stagnant air film to the well mixed atmosphere above (Liss and Slater, 1974). The two-film model assumes that the chemical is well mixed in the bulk solution below the liquid film and that mass transfer across each film is proportional to the concentration difference. With these assumptions, a chemical in the water body volatilizes at a rate proportional to the bulk concentration, so that the entire loss process can be characterized by an effective ''half-life," defining the mount of time required to reduce the mass in solution by 50 percent. Thomas (1982) reviewed volatilization loss models for chemicals present in water bodies and performed model calculations for a number of compounds. The film thicknesses depend on specific conditions within the water and air, but may be crudely estimated from default values given in Thomas (1982) when no actual data are available. Table 3.1 summarizes effective volatilization half lives calculated for a range of Henry's constant values. As seen from Table 3.1, the effective half-life varies widely, depending on
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Ground Water Recharge Using Waters of Impaired Quality TABLE 3.1 Effective Volatilization Half-Life Ranges as a Function of Dimensionless Henry's Constant Values for a Stagnant Water Body of 1-m Depth Henry's Constant KH Example Half life (days) 10-8 - 10-7 Bromacil 104 - 105 10-7 - 10-6 Amine 103 - 104 10-6 - 10-5 Phenol 102 - 103 10-5 - 10-4 Diazinon 10-100 10-4 - 10-3 EPTC 1-10 10-3 - 10-2 Bromobenzene ˜ 1 10-2 - 10-1 Benzene ˜ 1 10-1 - 1 Methyl bromide ˜ 1 1-∞ Vinyl chloride ˜ 1 Source: Default values for film transfer coefficients are taken from Thomas (1982), and KH values are taken from Jury et al. (1984b). the value of the Henry's constant. Clearly, compounds with half lives that are considerably less than the detention time of the water on the surface will not enter the soil. For instance, volatilization losses of 22 to 73 percent were found for a wide spectrum of hydrocarbons in sewage effluent infiltration basins in Phoenix, Arizona (Bouwer et al., 1986). Volatilization from Soil The volatilization loss rates of chemicals from soil are generally smaller than those from standing water for several reasons. First, the diffusion resistance of soil is greater than that of flee air because of the solid and liquid barriers to gas movement. Second, adsorption of chemical to soil solids reduces the vapor pressure of the compound by removing mass from solution. Because the transport pathways from the soil to the surface are much more complex than in free water, the two-film resistance model of volatilization is not applicable to soil, and more sophisticated estimation methods must be used. Jury et al. (1983, 1984a,b,c, 1990) developed a comprehensive screening
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Ground Water Recharge Using Waters of Impaired Quality model for evaluating volatilization losses of chemicals after their incorporation into a soil layer of arbitrary thickness. They performed calculations covering a range of initial conditions on a large group of organic compounds, allowing the volatilization losses of different chemicals to be compared and grouped. In general, they found that compounds with dimensionless Henry's constant values less than 10-4 were not prone to significant volatilization after deposition in the soil, but that extremely volatile compounds could move upward to the surface from substantial distances if they were not rapidly degraded. Under certain conditions, water evaporation can greatly enhance volatilization from soil by concentrating the chemical mass at the surface and raising its vapor pressure. There are several good references (see Howard, 1990; Howard et at., 1991) containing compendia of chemical properties and soil-chemical interaction coefficients (such as sorption coefficients and degradation rate coefficients). These are more reliable Sources of information on chemical properties than earlier references, which frequently contain outdated or inaccurate data. Volatilization During Ground Water Recharge Chemicals likely to volatilize during ground water recharge operations include nitrogen compounds and organics with high Henry's constant values. Operations conducted under high water saturation, such as ponding, will minimize the loss of organics from soil by volatilization because these will be blocked from entering the gas phase and reaching the soil surface if the soil air space is filled with water. However, the presence of standing water on the surface enhances the loss from water, so that any chemicals with volatilization half-lives that are significantly less than the total detention time in surface water Will probably evaporate out of solution before they ever reach the soil. Recharge water containing fertilizer or sewage contributions may contain nitrogen compounds that can be transformed to volatile species. Ammonia is very volatile and vaporizes from anhydrous form immediately upon exposure to the air. The ammonium ion NH4+, which is a constituent of many fertilizers, partitions into a volatile gas when exposed to air; therefore, any dissolved NH4+ near the atmospheric interface will lose nitrogen to the atmosphere. NH4+ is readily transformed by soil bacteria to nitrate (NO3-), which is very mobile in soil and is nonvolatile. However, NO3- may be transformed anaerobically by biological denitrification to several gaseous species (primarily N2O and N2) when soil water content is high and a source of organic carbon is present. Ground water recharge by ponding may therefore enhance removal of residual nitrogen through this process. There is evidence that potentially harmful inorganic compounds of selenium and other trace metals may be removed from contaminated soil by volatilization after the compounds have been methylated through biological transformation (Karlson and Frankenberger, 1989).
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Ground Water Recharge Using Waters of Impaired Quality Transport of Dissolved Chemicals Unless a compound is very volatile, it moves primarily in the aqueous solution phase. At the scale of the soil pore, there are two transport mechanisms that can move solutes through the medium: convection and diffusion. Convection is the transport of a dissolved chemical by virtue of bulk movement of the host water phase, while diffusion is the random mixing caused by collisions at the molecular scale. The local water flux describes three-dimensional flow around the solid and gaseous portions of the medium and is not measurable. Instead, the local quantifies are volume averaged to produce a larger-scale representation of the system properties. The averaging volume must be large enough that the statistical distribution of geometric obstacles is the same from place to place. If the porous medium contains the same material and density throughout, then the mean value produced by this averaging is macroscopically homogeneous over the new transport volume containing the averaged elements. The large-scale convective solute flux is then the product of the average water flux and the average solution concentration Cw. However, in the process of averaging, some of the solute motion is lost. The small-scale migration of solution through tortuous pathways within the porous medium no longer appears in the water flux expression after volume averaging, because the fluctuations about the mean motion do not contribute to the net water transport. Dissolved chemicals convected along these tortuous flow paths do contribute to solute transport and cannot be neglected. The motion of solute due to small-scale convective fluctuations about the mean motion is called mechanical dispersion (Bear, 1972). The combined mixing associated with diffusion and mechanical dispersion is called hydrodynamic dispersion. Development of models to describe the hydrodynamic dispersion flux is one of the most active research areas in soil physics and hydrology today (Dagan, 1986). The most widely used representation of dissolved chemical movement is the convection-dispersion model, which assumes that the dispersive mixing is random or diffusionlike within the moving fluid. Solute Velocity The water flux (or Darcy velocity) q is the volume of water flowing per unit time per unit area of soil. Because the water flows only through the water-filled regions of the porous medium, the actual average velocity of the water is equal to where θ is the volumetric water content (volume of water per volume of soil). Equation (2) describes the average linear velocity of the water and therefore also represents the movement of a chemical that travels freely with water and
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Ground Water Recharge Using Waters of Impaired Quality does not interact with solid surfaces in the soil. At normal pH levels, negatively charged ionic species such as NO3-are not attracted to soil mineral or organic surfaces and move relatively freely with flowing solution. In fact, the negatively charged clay lattice actually has a repulsive effect on the unions in solution, causing them to avoid the solution region that is closest to the surfaces, in which case (2) will somewhat underestimate the solute velocity. This effect is minor unless the soil is high in clay. Therefore, a chemical moving in solution at a water flux q that flows through a water-filled volume fraction θ without interacting with the solid phase will move with a velocity given by (2) and hence will require a time to move a distance L through the soil or aquifer. Because of diffusion and dispersion, not all of the solute molecules will move at the same speed, but will spread out around the average arrival time given in (3). Solute Dispersion During Transport The effect of dispersion is to produce chemical spreading during transport. It is greatly affected by soil structure and depends also on the scale over which the water flux is volume averaged. Soils with a pronounced macrostructure can transmit chemicals rapidly, leading to early arrival that greatly precedes the main pulse or front. In addition, aggregated regions of the medium containing water that is not flowing can act as repositories for diffusing chemicals, causing portions of a front or pulse to lag far behind the average flow. Dispersion modeling is poorly developed in unsaturated field soils. Near the soil surface of an alluvial soil, it is likely that dispersion will be dominated by differences in water velocity at different points in the soil. In this case, the common convection-dispersion model assuming random solute spreading is not likely to be accurate, and a stochastic convective model assuming parallel flow of solute in isolated stream tubes may represent the solute mixing process better (Butters and Jury, 1989). Solute Sorption During Transport Chemicals that are hydrophobic or positively charged do not travel at the speed of the flowing water, but rather are slowed by their attraction to stationary solid sorption sites. Although the sorption process is very dynamic at the molecular scale, it is useful to conceptualize an adsorbed molecule as existing in a distinct phase that is temporarily immobilized by virtue of its attachment to stationary solid matter.
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Ground Water Recharge Using Waters of Impaired Quality There are several different types of sorption reactions, distinguished primarily by the nature of the sorbing surface and the charge characteristics of the sorbing molecule. Positively charged ions in solution are attracted to negatively charged clay mineral surfaces and are temporarily immobilized by the process known as cation exchange. This is a partitioning reaction that divides the chemical mass between solution and adsorbed phases; it does not completely strip a compound from solution, nor is it permanent. The reaction depends on the nature of the molecule and also on the composition of the soil solution (Sposito, 1981). In addition, some positively charged species, notably the trace metals, appear to be specifically sorbed strongly to certain oxide surfaces (Chang and Page, 1985). Anions are repelled from clay mineral surfaces that are negatively charged, but are attracted to positively charged broken end faces of minerals and also to free oxides in the soil. These surfaces have charges that are strongly pH dependent, and attract anions most strongly under acidic conditions. Neutral organic molecules such as nonionic pesticides sorb primarily to organic matter surfaces in a reaction that can be approximated by a partition coefficient. The form of this relation is denoted through the linear sorption model where Cs is the sorbed chemical concentration (mass of chemical sorbed per mass of soil), Cw is the chemical concentration in the soil solution (mass of chemical per mass of water), and the proportionality coefficient Kd is called the distribution coefficient (Hamaker and Thompson, 1972). The distribution coefficient has units of volume per mass. Because most of the sorption occurs on organic matter surfaces, the distribution coefficient may be subdivided into where ƒoc is the fraction of soil organic carbon content (mass per mass of soil) and Koc is the distribution coefficient per unit organic carbon, called the organic carbon partition coefficient. Hamaker and Thompson (1972) showed that the Koc of a given compound varied significantly less between soils than the Kd. In that sense, it represents an intrinsic sorption affinity for a chemical that is soil independent. The organic carbon partition coefficient for chemicals generally decreases with increasing water solubility. The retardation factor R is another measure of the relative partitioning of a chemical between the soil and the aqueous phases: where ρb is the soil bulk density. R is the ratio between the total mass density and the mass per soil volume θCw in the dissolved phase. A simple derivation of
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Ground Water Recharge Using Waters of Impaired Quality with ultraviolet radiation, the question of the stability of DBPs in ground water systems becomes important. Changes in the DBP chemical composition of recharge water will occur in the infiltration basin and in the soil vadose zone for surface recharge operations, and in the aquifer during transport and storage, regardless of the input procedure. In surface recharge operations, the water to be recharged will undergo some chemical change due to microbiological activity in the infiltration basin. As organisms respond to the DOC and begin growth cycles, total organic carbon (TOC) may increase or decrease slightly, depending on retention time and the acclimation phase (Bouwer, 1984). In the vadose zone, hydrophobic compounds will be retarded by the soil organic material and subjected to microbial decomposition. Halogenated organic compounds, including the nonvolatile DBPs, will have greater hydrophilicity and correspondingly less retention on soft organic matter and more rapid transport to the water table. Some DBPs, such as the halogenated acetic acids, will behave as fully ionized organic anions at the pH values of water in the vadose zone and may be expected to move at the same speed as the subsurface water. Thus, it might be expected that the smaller halogenated materials will persist longer and penetrate more deeply into the soil horizons than the larger and more hydrophobic molecules. The degree of microbial utilization is most difficult to predict and will probably vary from site to site. The bioremediation literature notes that trichloroacetic acid is an intermediate in the oxidative degradation of vinyl chloride produced from anaerobic degradation of trichloroethylene and that it does not build up in the soil. This suggests that complete mineralization is possible for this DBP, at least in systems with fully adapted microorganisms. The same is probably true for the other haloacetic acids. Thus, control of the redox environment in the vadose zone through optimizing application rates is probably an important management criterion for these DBPs. Of the list of DBPs in Table 3.5 one might expect significant losses of formaldehyde and chloroform in the infiltration basin due to volatilization, with somewhat less significant losses of any bromodichloromethane and 1,1,1-trichloropropanone, owing to lower volatilities. Losses of trichloroacetic acid may be expected in the vadose zone. MX presents an unusual case because its polarity depends on whether the molecule is in the open form, which is a substituted botanic acid, or in the closed form, which is a substituted hydroxyfuranone. A pH value of less than 5 would give the closed form, which is substantially less polar and therefore more interactive with soil organic matter and more likely to be retained for microbial utilization. In the open acid form, it might be expected to behave like trichloroacetic acid. In addition, MX and its congeners are probably reactive with reduced forms of sulfur in the subsurface environment, but their reaction products are probably of less toxicological significance. Singer et al. (1993) have studied the fate of chlorinated DBPs when treated
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Ground Water Recharge Using Waters of Impaired Quality and disinfected drinking water was injected and stored in an aquifer at five sites. These authors concluded that THMs and haloacetic acids (HAAs) are removed from chlorinated drinking water during aquifer storage. HAA removal apparently precedes THM removal, and the more highly brominated species tend to be eliminated earliest. Removal of DBPs during aquifer storage was accompanied by a decrease in dissolved oxygen, implicating anaerobic biological mechanisms. In addition to DBP removal, reductions in the concentrations of organic precursor material were observed and correlated with measured reductions in THM formation potential and HAA formation potential. No effort was made to establish the identities of THM or HAA degradation products. The results from Singer et al. (1993) concur with those from Roberts (1985) who evaluated organic contaminant behavior during a recharge project in Palo Alto, California. He also found that THM losses were correlated with losses of dissolved oxygen. In addition, he found that the loss rate for THMs was nearly 10 times as rapid as the rate of concentration decrease for compounds containing two carbon atoms, represented by trichloroethylene, tetrachloroethylene, and 1,1,1-trichloroethane. For these reasons, disinfection to control viruses or other pathogens prior to infiltration may not necessarily result in an increase of DBPs in the ground water aquifer. However, it is possible to overload the removal mechanisms, and this is driven by TOC levels in the source water primarily. To provide disinfection and prevent DBP buildup in the ground water it would be necessary to pretreat source water in order to reduce levels of TOC. On the other hand, if the pathogen content of the ground water is not of primary concern, then disinfection (and DBP formation) may be avoided prior to recharge. SUSTAINABILITY OF THE SAT SYSTEM Sustainability refers to the long-term viability of an SAT operation, and specifically to the ability of the soil and aquifer to receive recharge water indefinitely without suffering deleterious side effects. Within the vadose zone, clogging is an inevitable side effect of SAT procedures, but is largely manageable with periodic drying to oxidize the accumulated organic material and restore the infiltration rates, along with periodic physical removal of the clogging layer by scraping, raking, or other techniques (Bouwer, 1985). Of the material that is not completely removed from the recharge water before reaching the aquifer, viruses are of special concern because they may cause disease or pose an unacceptably high risk even when present in low concentrations. Sustainability must therefore include considerable travel distance between the receiving basin of the aquifer and any route to drinking water supplies to allow sufficient vital inactivation to occur (Yates et al., 1986). Nitrates that enter the aquifer should be regarded as nondegrading, and therefore the recharge operation must operate very efficiently with respect to nitrogen
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Ground Water Recharge Using Waters of Impaired Quality RAPID INFILTRATION-EXTRACTION PROJECT IN COLTON, CALIFORNIA The cites of San Bernardino and Colton, California, are required to filter and disinfect their secondary effluent prior to discharge to the Santa Ana River, which is a source of drinking water and is used for body contact recreation. The two cities joined under the auspices of the Santa Ana Watershed Project Authority (SAWPA) to seek a regional solution to their wastewater treatment requirements. They hoped to develop a cost-effective alternative to conventional tertiary treatment (chemical coagulation, filtration, and disinfection) that would still result in an effluent that was essentially free of measurable levels of pathogens. SAWPA conducted a one-year demonstration project to examine the feasibility of a rapid infiltration-extraction (RIX) process to treat unchlorinated secondary effluent prior to discharge to the river and determine whether or not the RIX process is equivalent—in terms of treatment reliability and quality of the water produced—to the conventional tertiary treatment processes specified in the California Department of Health Services Wastewater Reclamation Criteria. The demonstration was conducted on a site in Colton, California, adjacent to the Santa Ana River bed. Infiltration basins were built at two sites on the property to allow testing the RIX system under a variety of operating conditions. The soils were coarse sands with clean water infiltration rates of about 15 m/day (50 ft/day). Forty four monitoring wells were sampled for a number of organic, inorganic, microbiological, and physical measurements. The study indicated that the optical filtration rate was 2 m/day (6.6 ft/day) with a wet to day ratio of 1:1 (1 day of flooding to 1 day of drying). Mounding beneath the infiltration basins ranged from 0.6 to 0.9 m (2 to 3 ft), and infiltrated wastewater migrated up to 24.4 m (80 ft), vertically and over 39.5 m (100 ft) laterally before it was extracted and had an aquifer residence time of 20 to 45 days, depending on the recharge site. Extraction of 110 percent of the volume of infiltrated effluent by downgradient extraction wells effectively contained the effluent on the RIX site and minimized mixing with regional ground water outside the project area. The soil-aquifer treatment reduced the concentration of total coliform organisms 99 to 99.9 percent in samples collected approximately 7.6 m (25 ft) below the infiltration basins, and, generally, water from the extraction wells prior to disinfection contained less than 2.2 total coliforms per 100 ml. While viral levels were as high as 316 viruses per 2001 in the unchlorinated secondary effluent applied to the infiltration basins, only one sample of extracted water prior to disinfection was found to contain detectable levels of viruses. In addition, the turbidity of the extracted water generally averaged less than 0.04 NTU (Foreman et al. 1993). Although the RIX process greatly reduced microorganism levels in the wastewater, disinfection of the extracted wastewater proved to still be necessary prior to its discharge to the Santa Ana River. Ultraviolet radiation was evaluated as an alternative to chlorination for disinfection. Due, in part, to the high quality of the extracted water, ultraviolet radiation was shown to be effective for the destruction of both bacteria and viruses and will likely be used instead of chlorine in the fullscale project (CH2M Hill, 1992).
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Ground Water Recharge Using Waters of Impaired Quality removal in the vadose zone where it is transformed. Because proper management of an SAT system has been able to achieve as high as 80 percent removal of nitrogen from the recharge water, control of this chemical may be manageable in long-term operations. Trace element accumulation in the vadose zone during operation may occur so slowly that exhaustion of the assimilative capacity of the surface zone may not occur in a time that will limit the economic viability of the operation (Chang and Page, 1985); however, the capacity of the soil to remove metal cations is not infinite, and their buildup should be monitored. Site closure at the termination of operation of an SAT system should be regulated to ensure that the metals residing in the surface soil are not disturbed by operations that might change the chemical characteristics of the soil solution and mobilize the contaminants. Pesticides and other organic chemicals present in recharge water vary significantly in their mobility, persistence, and suspected health effects. Water that contains pesticides known to be resistant to microbial or chemical degradation in soil may pose problems for SAT systems in the long run, because refractory organics are only slowed by solid-phase sorption and not permanently removed from solution. Ground water recharge operations, either from surface spreading or injection wells, introduce microorganisms to the subsurface environment in addition to organic and inorganic chemicals. Microbial activity associated with artificial recharge may have three principal effects: (1) bacteria may grow near the recharge facility and cause clogging, with a gradual reduction of soil and aquifer hydraulic conductivity; (2) microbial activity may produce substances that adversely affect the taste and odor of recovered water; and (3) pathogenic organisms in the recharge water may travel through the aquifer and cause illness when the water is later recovered for use without disinfection (Ehrlich et al., 1979a). During tests with injection of highly treated sewage into a sand aquifer on Long Island, chlorination of the injectant to 2.5 mg/1 suppressed bacterial growth and clogging (Ehrlich et al., 1979b). Similar tests at another injection site showed that clogging could be controlled with a chlorine residual maintained between 0.2 and 1.5 mg/l at the treatment plant (Schneider et al., 1987). The tests of Ehrlich et al. (1979b) also showed that movement of bacteria from the injection well into the aquifer was not extensive. Pathogenic bacteria are generally incapable of competing with soil bacteria for nutrients, while viruses are incapable of reproducing outside of a living cell. They do not multiply and eventually they die and decompose (Bouwer, 1978). PERFORMANCE AND COMPLIANCE MONITORING Vadose Zone Monitoring Ideally, monitoring should be instituted near the point of application in the
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Ground Water Recharge Using Waters of Impaired Quality soil to determine whether the SAT system is functioning in the desired manner. Vadose zone monitoring is difficult, however, for a number of reasons. First, there are essentially only two monitoring methods (soil sampling and vacuum extraction of solution), both of which have drawbacks. Soil sampling is destructive and very localized, so many samples must be taken to give an accurate picture of the average and extreme behavior of the SAT system. This causes substantial land disturbance. Vacuum extraction through porous soil solution samplers permanently installed in the soil is less destructive than soil sampling, but also can produce preferential flow pathways along the sides of the sampling tube if special care is not taken to produce and maintain a tight contact with the surrounding soil. Vacuum extractors also sample an unknown volume of soil and may be very inefficient at intercepting rapidly flowing solution such as might be present in preferential flow channels (Roth et al., 1991). They may also ''filter out" microorganisms and chemicals. For these reasons, the contaminant transport characteristics revealed by solution samplers, particularly those located near the surface, might not be representative of the true behavior in the field (Roth et al., 1991). A second limitation to monitoring near the surface is that the transformation processes occurring within this zone may not have reached completion at the point of monitoring, so interpretation of the information recorded is problematic. Spatial variations in water velocity, chemical concentration, and soil structure are most pronounced near the surface, so the monitoring density required for accurate assessment of average and extreme behavior is greatest in this zone. For these reasons, the best zone of monitoring is generally farther from the surface, such as in the top of the saturated zone. Vadose zone monitoring is generally confined to periodic sampling of the soil and solution to determine general trends. Tracing of Recharge Water To follow the migration and fate of recharge water, water quality changes may be monitored by using water samples from wells and lysimeters. Chemical analyses usually focus on indicator chemicals rather than attempting to identify all recharge water chemical species. If a sufficient contrast exists between recharge and ambient waters, then TDS or conductivity may serve as indicators of general migration. Other mobile chemical species that may be used as indicators of recharge waters include chlorides, sulfates, and nitrates. Migration of organic species may be followed by grouping chemicals based on their potential mobility through R values. Lumped parameters such as TOC may serve as a surrogate for wastewater, with values less than 1 mg/1 TOC at the wellhead showing adequate dilution and TOC losses for potable use. Tracing of recharge waters is more difficult if the recharge water quality is similar to that of native formation waters. At the Fred Hervey water reclamation
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Ground Water Recharge Using Waters of Impaired Quality plant in E1 Paso, Texas, the recharge and formation water are very similar, and a, combination of chlorine, nitrogen, and 18O concentrations are used to determine breakthrough of recharge waters at monitoring and recovery wells. Mixing of Recharge and Ambient Ground Water The extent of mixing between recharge and ambient ground waters is important in determining whether potentially toxic chemicals in the recharge source waters exceed concentrations that may adversely affect human health when extracted waters are used for potable purposes. The degree to which recharge water has mixed with ambient ground water at a site downgradient from the recharge facility may often be estimated from water quality parameters. If chemical constituents behave conservatively during subsurface migration and mixing, then a linear weighting of concentrations may be used to estimate the degree of mixing (Reeder et al., 1966). For example, chloride and sodium concentrations were used to determine mixing of reclaimed water recharged at East Meadow, Long Island, with the ambient ground water (Schneider et al., 1987). The median sodium and chloride concentrations in the recharge water were approximately 115 and 157 mg/l, respectively, whereas those in ambient ground water were 22 and 27 mg/l, respectively. If the chloride concentration in the ambient ground water is approximately 30 mg/l, in the recharged water is 160 mg/l, and in the observation well sample is Co mg/l, then the fraction F of recharge water in samples from an observation well can be calculated from the following equation: An observation well sample concentration of Co = 100 mg/l would correspond to a recharge water fraction of 0.54, or 54 percent in the water sampled. California is developing regulations that place restrictions on the percentage of recharged and reclaimed water to be included in water produced by supply wells for potable use. A supply well can produce water from strata above and/or below as well as from the zone in which subsurface migration of the recharge water is occurring. Also, as a result of radial flow toward the supply well during pumping episodes, part of the water produced by the well may be from parts of the aquifer containing ambient ground water. Under either of these conditions, the water produced by the supply well may be a mixture of recharge water and ambient ground water. Equation (8) can be used to estimate the recharge water fraction in the produced supply water. Not all chemical constituents allow a simple linear weighting of concentrations to be used to estimate the degree of mixing, especially reactive chemicals and those contributing to the buffering capacity of the system. As an example,
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Ground Water Recharge Using Waters of Impaired Quality calcium may not serve as a good indicator of mixing because Of its affinity for clay minerals and because of its role in the calcium-carbon dioxide system. If ambient and recharge waters are mixed within the formation, then calcium carbonate may precipitate, remain in equilibrium with the resulting water, or dissolve, depending on the ratios of carbonate and bicarbonate in the two source waters (Huisman and Olsthoom, 1983). Posttreatment of Recharge Waters For many recharge systems the recharge water tends to lose its chemical identity while in the subsurface, and recovered waters have traditionally required little treatment before use. Even where potable reuse is contemplated, if pretreatment of recharge waters is sufficient, little posttreatment is required. For example, recharge waters in E1 Paso, Texas, are planned for potable reuse. The moderately weak domestic sewage receives advanced wastewater treatment before injection, and the recovered ground water receives only disinfection before use. Where treatment is required, posttreatment of recharge waters is no different from ground water treatment. Tertiary treatment can remove taste and odor problems. For removal of organics, aeration (air stripping) and sorption using activated carbon are particularly effective, and neither is inordinately expensive nor difficult to install in a water treatment system. Activated carbon may not remove DOC to sufficiently low values. In that case, membrane filtration or reverse osmosis can be used. Point-of-use water treatment systems may include chlorination, ozonation, or ultraviolet disinfection, ion exchange for water softening, and filtration with activated carbon or membrane filtration for control of organics as well as taste and odor. SUMMARY The soil and underlying aquifer have a great capacity to remove chemical contaminants and pathogens from recharge water. The assumption that passage through the soil to the aquifer and through the aquifer to the point of withdrawal provides no treatment and that an treatment must be provided before recharge or after extraction is overly conservative when applied to most chemicals and microorganisms. Ground water recharge and recovery systems can provide significant treatment benefits at relatively low costs and in appropriate circumstances can make use of water of impaired quality attractive. The ideal soil for an SAT system balances the need for a high recharge rate, which occurs in come-textured soils, with the need for efficient contaminant adsorption and removal, which are better in free textured soils. Because structured soils are undesirable for obvious reasons, the best choice of soil texture for SAT is a structureless fine sand or sandy loam.
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Ground Water Recharge Using Waters of Impaired Quality The vadose zone has the capacity to remove many constituents of concern during passage of recharge water toward the underlying aquifer. Nitrogen, for example, quickly transforms to nitrate, which is very mobile under normal conditions in the soil but can be removed only by denitrification under anaerobic conditions. This reaction can be enhanced by the proper combination of alternate wetting and drying cycles. Phosphorus levels are reduced by sorption and precipitation, but not completely, Trace metals, with the exception of boron and arsenic, are strongly attenuated and precipitated in the soil. There is some concern about their eventual passage through a soil that has been under SAT for many years or after closure of a site. Organic chemicals are removed to varying extents by volatilization or chemical or biological degradation during passage through the vadose zone. Some pathogen removal in soil occurs by filtration in the surface clogging mat for the largest organisms and by sorption for bacteria and viruses. Viruses are considerably more mobile in soil than the larger pathogens, although they inactivate in soil eventually. Traditional disinfection by chlorination produces DBPs, which are mobile in soil and persistent to varying degrees. With proper management and pretreatment, an SAT operation employing surface wastewater spreading with periodic drying to reduce clogging should be sustainable indefinitely. Slow trace element migration remains a concern, however, that requires monitoring during SAT use and regulation after closure. Although near-surface monitoring is desirable for proper vigilance, soil variability makes it difficult to achieve complete coverage with existing devices. Therefore, a combination of periodic near-surface and distant monitoring is important. With adequate management and monitoring, SAT systems can reduce pretreatment costs. REFERENCES American Society of Civil Engineers. 1961. Recharge and withdrawal. Committee on Ground Water. ASCE Man. Eng. Pract. 40:72-92. Amy, G. L., P. A. Chadki, and P. H. King. 1984. Chlorine Utilization during formation of THM in presence of ammonia and bromide. Environ. Sci. Technol. 18:781-786. Anderson, L. J., J. D. Johnson, and R. F. Christman. 1985. Reaction of ozone with isolated aquatic fulvic acid. Organic Geochemistry. 8:65-69. Armstrong, D. E., and J. Konrad. 1974. Nonbiological degradation of pesticides. In Pesticides in Soil and Water, W. Guenzi, ed. Madison, Wisc: Soft Science Society of America. Asano, T. 1985. Artificial Recharge of Groundwater. Boston, Mass: Butterworth. Bear, J. 1972. Dynamics of Fluids in Porous Media. New York: Elsevier. Beven, K., and P. Germann. 1982. Macropores and water flow in soils. Water Res. Res. 18:1311-1325. Bouwer, E. J., P. L. McCarty, H. Bouwer, and R. C. Rice. 1984. Organic contaminant behavior during rapid infiltration on secondary wastewater at the Phoenix 23rd Avenue Project. Water Res. 18(4):463-472. Bouwer, H. 1978. Groundwater Hydrology. New York: McGraw-Hill. Bouwer, H. 1984. Wastewater renovation in rapid infiltration systems. Ground Water 22:696-705.
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Ground Water Recharge Using Waters of Impaired Quality Bouwer, H. 1985. Renovation of wastewater with rapid-infiltration land treatment systems. Pp. 249-282 in Artificial Recharge of Groundwater, T. Asano, ed. Boston, Mass.: Butterworth. Bouwer, H. 1991. Simple derivation of the retardation equation and application to preferential flow and macrodispersion. Ground Water 29(1):41-46. Bull R. J., and F. C. Kopfler. 1991. Health effects of Disinfectants and Disinfection By-products. Denver: Am. Water Works Assoc. Res. Found. Butters, G. L., and W. A. Jury. 1989. Field scale transport of bromide in an unsaturated soil. II. Dispersion modeling. Water Res. Res. 25:1582-1589. Chang, A. C., and A. L. Page. 1985. Soil deposition of trace metals during groundwater recharge using surface spreading. Pp. 609-626 in Artificial Recharge of Groundwater, T. Asano, ed. Boston, Mass.: Butterworth. CH2M Hill. 1992. UV Disinfection Pilot Study: Rapid Infiltration/Extraction (RIX) Demonstration Study. Prepared for the Santa Ana Watershed Project Authority, City of San Bernardino, and City of Colton. Santa Am, Calif. Crites, R. W. 1985. Micropollutant removal in rapid infiltration. Pp. 597-608 in Artificial Recharge of Groundwater, T. Asano, ed. Boston, Mass: Butterworth. Dagan, G. 1986. Statistical theory of groundwater flow and transport: Pore to laboratory, laboratory to formation, and formation to regional scale. Water Resour. Res. 22:120S-134S. Ehrlich, G. G., E. M. Godsy, C. A. Pascale, and J. Vecchioli. 1979a. Chemical changes in an industrial waste liquid during post-'rejection movement in a limestone aquifer, Pensacola, Florida. Ground Water 17(6):562-573. Ehrlich, G. G., H. Ku, J. Vecchioli, and T. Ehlke. 1979b. Microbiological effects of recharging the Magothy Aquifer, Bay Park, New York, with tertiary-treated sewage . U.S. Geol. Sun,. Prof. Paper 751-E. Foreman, T. L., G. Nuss, J. Bloomquist, and G. Magnuson. 1993. Results of a 1-year rapid infiltration/extraction (RIX) demonstration project for tertiary filtration. Pp. 21-36 in proceedings of the Water Environment Federation 66th Annual Conference and Exposition, Oct. 3-7, 1993, Anaheim, Calif. Alexandria, Va.: Water Environment Federation. Gerba, C., and S. Goyal. 1985. Pathogen removal from wastewater during groundwater recharge. Pp. 283-318 in Artificial Recharge of Groundwater, T. Asano, ed. Boston, Mass.: Butterworth. Ghodrati, M., and W. A. Jury. 1990. A field study using dyes to characterize preferential flow of water. Soil Sci. Soc. Am. J. 54:1558-1563. Hamaker, J. W., and J. M. Thompson. 1972. Adsorption. Pp. 49-144 in Organic Chemicals in the Soil Environment, C. A. I. Goring and J. W. Hamaker eds. New York: Marcel Dekker. Hillel, D. 1987. Unstable flow in layered soils; a review. Hydrological Processes 1(2):143-147. Howard, P. H. 1990. Handbook of Environmental Fate and Exposure Data for Organic Chemicals Vol I and II. Chelsea, Mich: Lewis. Howard, P. H., R. Boethling, W. Jarvis, W. Meylan, and E. Michalenko. 1991. Handbook of Environmental Degradation Rates. Chelsea, Mich: Lewis. Huisman, L., and T. N. Olsthoorn. 1983. Artificial groundwater recharge. Boston, Mass.: Pitman. Jenne, E. A. 1968. Controls on Mn, Fe, Co, Ni, Cu, and Zn concentrations in soils and water. Pp. 337-387 in Trace Organics in Water, R. F. Gould ed. Advances in Chemistry Series, Vol. 73. Washington, D.C.: American Chemical Society Jury, W. A. 1985. Spatial Variability of Soil Physical Parameters in Solute Migration: A Critical Literature Review. EPRI Topical Rep. E4228, Electric Power Research Institute. Palo Alto. Calif. Jury, W. A., and H. Fluhler. 1992. Transport of chemicals through soil: Mechanisms, models, and field applications. Adv. Agron. 27:141-201. Jury, W. A., and K. Roth. 1990. Evaluating the role of preferential flow on solute transport through unsaturated field soils. Think Tank Workshop Report. Pp. 23-30 In Field Scale Water and Solute Flux in Soils, K. Roth et al., eds. Basel, Switzerland: Birkhaeuser.
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Ground Water Recharge Using Waters of Impaired Quality Jury, W. A., W. F. Spencer, and W. J. Farmer. 1983. Model for assessing behavior of pesticides and other trace organics using benchmark properties. I. Description of model. J. Environ. Qual. 12:558-564. Jury, W. A., W. J. Farmer, and W. F. Spencer. 1984a. Model for assessing behavior of pesticides and other trace organics using benchmark properties. II. Chemical classification and parameter sensitivity. J. Environ. Qual. 13:567-572. Jury, W. A., W. F. Spencer, and W. J. Farmer. 1984b. Model for assessing behavior of pesticides and other trace organics using benchmark properties. III. Application of screening model. J. Environ. Qual. 13:573-579. Jury, W. A., W. F. Spencer, and W. J. Farmer. 1984c. Model for assessing behavior of pesticides and other trace organics using benchmark properties. IV. Review of experimental evidence. J. Environ. Qual. 13:580-585. Jury, W. A., D. D. Focht, and W. J. Farmer. 1987. Evaluation of pesticide groundwater pollution potential from standard indices of soil-chemical adsorption and biodegradation. J. Environ. Qual. 16:422-428. Jury, W. A., D. Russo, G. Streile, and H. Elabd. 1990. Evaluation of volatilization by organic chemicals residing below the soil surface. Water Resour. Res. 26:13-20. Jury, W. A., W. R. Gardner, and W. H. Gardner. 1991. Soil Physics. New York: John Wiley. 239 PP. Karlson, U., and W. T. Frankenberger. 1989. Removal of selenium from evaporation pond sediments by biological transformation. Sci. Total Environ. 92:41-54. Kung, K.-J. S. 1990. Preferential flow in a sandy vadose zone: 1. Field observation. 2. Mechanism and implications. Geoderma 46:51-72. Lance, J. C., and C. P. Gerba. 1984. Virus movement in soil during saturated and unsaturated flow. Applied Environmental Microbiology. 47:335-337. Liao, W. T., R. F. Christman, J. D. Millington, and J. R. Hass. 1982. Structural characterization of aquatic humic material. Environ. Sci. Technol. 16:403-410. Liss, P. S., and P. G. Slater. 1974. Fluxes of gases across the air-sea interface. Nature 247:181-184. Nellor, M. A. 1980. Health effects of water reuse by ground water recharge. Tech. Rep. CRWR-175. Center for Research in Water Resources, University of Texas, Austin. Omoti, U., and A. Wild. 1979. Use of fluorescent dyes to mark the pathways of solute movement through soils under leaching conditions. 2. Field experiments. Soil Sci. 128:93-104. Powelson, D. K., and C. P. Gerba. 1993. Virus transport and removal in wastewater during aquifer recharge. Water Research. 27:583-590. Powelson, D. K., J. R. Simpson, and C. P. Gerba. 1990. Virus transport and survival in saturated and unsaturated flow through soil columns. J. Environ. Qual. 19:396-401. Pratt, P. F., W. W. Jones, and V. E. Hunsaker. 1972. Nitrate in deep soil profiles in relation to fertilizer rates and leaching volume. J. Environ. Qual. 1:97-102. Rao, P. S. C., A. G. Hornsby, and R. E. Jessup. 1985. Indices for ranking the potential for pesticide contamination of groundwater. Proc. Soil Crop Soc. Fla. 44:1-8. Reeder, H., W. Wood, G. Ehrlich, and R. Sun. 1966. Artificial recharge through a well in fissured carbonate rock, West St. Paul, Minnesota. U.S. Geological Survey Water Supply Paper 2004. Ritchie, J. T., D. E. Kissel, and E. Burnett. 1972. Water movement in undisturbed swelling clay soil. Soil Sci. Soc. Am. Proc. 36:874-879. Roberts, P. V. 1985. Field observations of organic contaminant behavior in the Palo Alto baylands. Pp. 647-679 in Artificial Recharge of Groundwater, T. Asano, ed. Boston, Mass: Butterworth. Roth, K., W. A. Jury, H. Fluhler, and W. Attinger. 1991. Transport of chloride through an unsaturated field soil. Water Resour. Res. 27:2533-2541. Schneider, B. J., H. Ku, and E. Oaksford. 1987. Hydrologic effects of artificial recharge experiments with reclaimed water at East Meadow, Long Island, NY. U.S. Geol. Surv. Water Resour. Inv. Rep. 85-4323.
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Ground Water Recharge Using Waters of Impaired Quality Scotter, D. R., and P. Kanchanasut. 1981. Anion movement in a soil under pasture. Austr. J. Soil Res. 19:299-307. Sharpless, R., E. Wallihan, and F. Peterson. 1969. Retention of zinc by some add zone soils treated by zinc sulfate. Soil Sci. Soc. Am. Proc. 33:901-904. Singer, P. C., R. D. G. Pyne, M. AVS, C. T. Miller, and C. Mojonnier. 1993. Examining the impacts of aquifer storage and recovery on DBPs. J. American Water Works Assn. 85(11):85-94. Sposito, G. 1981. The Thermodynamics of Soil Solutions. Oxford: Oxford University Press. Thomas, R. G. 1982. Volatilization from water. Chap. 15 in Handbook of Chemical Property Estimation Methods, W. J. Lyman et al., eds. New York: McGraw-Hill. Vecchioli, J., H. F. H. Ku, and D. J. Sulam. 1980. Hydraulic Effects of Recharging the Magothy Aquifer, Bay Park, New York, With Tertiary-Treated Sewage. Geol. Surv. Prof. Paper 751-F. U.S. Department of the Interior, U.S. GPO, Washington, D.C. Yates, M. V., and S. R. Yates. 1987. Modeling virus survival and transport in the subsurface. J. Contam. Hydrol. 1:329-345. Yates, M. V., S. R. Yates, A. Warrick, and C. Gerba. 1986. Predicting virus fate to determine septic tank setback distances using geostatistics. Appl. Environ. Microbiol. 52:479-483.
Representative terms from entire chapter: