Land Treatment of Wastewater

LAND TREATMENT OF WASTEWATER Herman Bouwer and R. L. Chaney US. Department of Agriculture, Agricultural Research Service, US. Water Conservation Laboratory, Phoenix, Arizona, and US. Department of Agriculture, Agricultural Research Service, Biological Waste Management Laboratory, Beltsville Agricultural Research Cenfer, Beltsville, Maryland

I. Introduction .................................................... 11. Fate of Wastewater Constituents in Soil . . . . . . . . . . . . . . . . . . . . . . . . . . A. Suspended Solids and Clogging . . . . . . . . . . . . . . . . .............. B. Organic Carbon and Oxygen Demand ............................ C. Bacteria and Viruses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D. Nitrogen . . . . . . . . . . . . . . . . . . . ........................ E. Phosphorus . . . . . . . . . . . ................................. F. Fluorine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . G. Boron ...................................................... H. Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . I. Dissolved Salts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . J. pH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 111. Crop Response . . . . . . . . . . . . . . . ............................. A. Effects on Yield and Quality . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B. Uptake of Pollutants and Location in Plant ........................ IV. Selection and Design of System . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References ....................... ........................... I.

133 135 135 137 141 146 151 155 155

157 162 163 164 164 165 167 169

Introduction

Public awareness of the need for preserving the quality of our surface water and increasingly severe legal restrictions on the discharge of pollutants into streams and lakes have revived interest in the use of land for disposal, treatment, and utilization of sewage effluent and other liquid wastes. Such systems have great public appeal. Wastewater is not only kept out of surface water, but land treatment also implies recycling, where “pollutants” become nutrients for plant growth. The simplicity, reliability, and low energy requirements of land treatment, as contrasted with the complex technology and high energy requirements of advancedtreatment plants, are other favorable aspects. Expressions, such as cleaning waste in nature’s way, living filters, plant-soil filters, soil mantle as sewage treatment plant, green-land clean-streams, etc., abound in the literature on land treatment of waste (Kardos, 1967; McGauhey and Krone, 1967; Stevens, 1972). Conversely, proponents of in-plant treatment have labeled 133

134

HERMAN BOUWER AND R. L. CHANEY

land treatment “a giant step backward” (Egeland, 1973). Selection of a certain system for treatment of wastewater should be free fron emotionalism. The economics and environmental aspects of various alternatives should be carefully considered so that the best system can be rationally selected. Liquid wastes commonly applied to land include conventionally treated sewage; wet sewage sludge (about 95% water), liquid animal waste (including feedlot runoff and lagoon or oxidation ditch effluents); and effluents from fruit or vegetable processing plants, animal processing plants, dairies, and fiber products industries. While these wastes vary widely in their composition, they all generally contain organic material, nitrogen, phosphorus, dissolved salts, trace elements, and microorganisms. Land treatment systems can generally be divided into three types: overland flow systems, low-rate application systems, and high-rate application systems (Bouwer, 1968; Thomas, 1973a). Overland flow systems are used where the soil is too impermeable or the suspended solids content of the wastewater too high to allow significant infiltration rates, causing most of the wastewater to run off. These systems are sometimes also called grass or vegetation filtration systems, or spray-runoff systems. With low-rate application systems, all wastewater applied infiltrates into the soil, but the dosages are rather small and of the same order as the water requirements of the crop or vegetation. Typically, the amounts are 2-10 cm per week, which may be given in one or several applications. Low-rate systems include all systems where wastewater is used for crop irrigation. Other uses of wastewater in this category are for revegetation of mine spoils, greenbelts, recreation areas, etc. With high-rate application systems, all wastewater again infiltrates into the soil, but the dosage is much greater than that necessary for crop growth. Amounts may range from about 0.5 m per week to several meters per week. Infiltration periods are rotated with drying or resting periods, to allow recovery of infiltration rates (infiltration rates generally decrease during application of wastewater) and to oxygenate the upper portion of the soil profile. High-rate systems require permeable soil. Often, the main function of the land with these systems is to receive and treat wastewater for groundwater recharge and reuse for irrigation, recreation, or industrial-municipal purposes. Agricultual utilization of the infitration system is of no or secondary importance. For both low-rate and high-rate systems, wastewater may be applied with sprinklers or, if the topography permits, with furrows, borders, or basins. Plow-in systems are sometimes used for essentially one-time application of thick liquids, such as sewage sludge or slurries from processing plants. The waste is injected into the plow furrow and covered with soil,

LAND TREATMENT OF WASTEWATER

II.

Fate of Wastewater Constituents in

135

Soil

A. SUSPENDED, SOLIDSAND CLOGGING Wastewater is usually screened, settled, or comminuted before it is applied to land. Thus, suspended solids received by the soil are usually rather fine and mainly in the organic form (sewage sludge, bacteria flocs, fibrous materials, fruit and vegetable peelings, straw or other roughage, algae cells, etc.) . These solids accumulate on the soil, forming a layer of high hydraulic impedance. This layer reduces the infiltration rate and, because it consists of biodegradable organic material, also constitutes an oxygen sink. This sink can cause small plants and seedlings to die, and it may diminish the movement of oxygen in the soil during drying. When worked into the soil, the solids initially could immobilize nitrogen if the nitrogen content is less than 1.3% on a dry-weight basis (Viets, 1973). Fine suspended material, such as colloidal clay particles, may move deeper into the soil (Goss and Jones, 1973). Movement of algal cells into dune sand was reported by Folkman and Wachs (1970). The soil, however, is a very effective filter, and suspended solids will be essentially completely removed from the wastewater after about 1 m of percolation. Since clogging at or near the surface of the soil is much easier to control and rectify than when it occurs at greater depth, it is important to know where the clogging is concentrated. The “symptoms” of clogging at the surface are decreasing water pressures (increasing tensions) and decreasing water contents in the upper portion of the soil profile, and increased effect of the water depth above the surface on the infiltration rate (Bouwer et al., 1974a). Clogging at greater depths is accompanied by increasing water pressures (decreasing tensions) and increasing water contents in the upper portion of the soil, and a decrease of the effect of depth of ponding on the infiltration rates. Clogging of the surface soil in a rapid-infiltration system receiving secondary sewage effluent was mainly a physical process due to the accumulation of suspended solids (Rice, 1974). The hydraulic impedance of the clogged layer was directly proportional to the total solids load. For a given solids load, high hydraulic gradients in the surface layer of the soil produced more compact layers of solids than did low hydraulic gradients. The more compact layers had a greater hydraulic impedance than the less compact layers for the same total solids load. Drying effectively restored the infiltration rate (Rice, 1974; Bouwer, et al., 1974a), as a result of the clogged layer decomposing. If the effluent contained suspended solids much in excess of 10 mg/liter, periodic removal of the sludge layer was

136

HERMAN BOUWER AND R. L. CHANEY

required to avoid a build-up of solids and, hence, a general decline in the infiltration rates in the basins. Biological clogging of the surface soil may also be caused by bacterial action, including production of polysaccharides and other organic compounds, if the wastewater contains a high dissolved organic matter content. Human urine, for example, with a chemical oxygen demand of 4000 mg/liter caused such clogging in fine-sand filters that the resulting infiltration rates were too low for practical application, and coarser sand had to be used (California Institute of Technology, 1969). Thomas et al. (1966) observed accelerated clogging of soil columns flooded with septic-tank effluent when the soil became anaerobic. Clogging was concentrated in the top centimeter. While sulfide accumulation could be used as an indicator of anaerobiosis, it was not a direct cause of clogging. Drying the soil caused infiltration recovery equivalent to the decrease in infiltration during anaerobic conditions. Since organic matter was the only material that declined during drying, clogging was attributed to the accumulation of polysaccharides, polyuronides, and other organic compounds during flooding. Nevo and Mitchell (1967) found that low redox potentials inhibited degradation of polysaccharides in laboratory experiments, but had little effect on the production of polysaccharides, indicating the need for regular drying or resting periods of treatment fields to avoid declines in infiltration rates. These workers also found that at temperatures below 20°C decomposition of polysaccharides was inhibited but synthesis slowly continued. Between 20 and 30"C, production and degradation rates of polysaccharides were approximately equal, and both rates increased with temperature. At 37"C, little polysaccharide was produced, but the decomposition rate continued to increase. This indicates that soil clogging caused by formation of polysaccharides may be of greater concern in cool climates than in warm climates. Regardless of climate, the optimum schedule of wastewater application and drying or resting of the soil must be evaluated by local experimentation. For the Flushing Meadows Project (Bouwer, 1973a; Bouwer et al., 1974a), maximum long-term infiltration rates were obtained with flooding periods of about 18 days, rotated with drying periods of about 10 days in the summer and 20 days in the winter. At the Whittier Narrows spreading grounds (McMichael and McKee, 1965), basins are flooded for about 9 hours and then dried for about 15 hours. With this schedule, about 2 feet per day infiltrated into the soil. Wastes containing very high solids contents may be applied only a few hours each week to allow drying and decomposition of the solids layer. Bendixen el al. (1968) reported satisfactory performance of a ridge-and-furrow system in northern latitudes where sec-

LAND TREATMENT OF WASTEWATER

137

ondary sewage effluent was applied on a 2 weeks on and 2 weeks off schedule. Clogging from excessive accumulation of suspended solids on the surface of the soil can be a problem in disposal fields with poor surface drainage. Because of reduced infiltration rates, surface runoff will develop and water will collect in the low places of the field, causing anaerobic conditions in the solids layer and the underlying soil. This will reduce the rate of decomposition of the solids, and odor and insect problems may develop. It is generally desirable to remove as much suspended material from the wastewater as possible before the water is applied to land. Overland-flow systems can effectively remove suspended solids of wastewater. Thomas (1973b) reported a suspended solids reduction from an average of 160 mg/liter (range 52-420) to 6-12 mg/liter for comminuted raw sewage applied to vegetated plots that were 36 m long and had a slope of 2-4%. The loading rates were from 7.4 to 9.8 cm/week, applied daily (except Sundays) in 8-9 hours. Law et al. (1970) found that the suspended solids content of screened cannery waste was reduced from 245 to 16 mg/liter by vegetation filtration over a distance of 45-100 m at loading rates of 0.9 cm/day applied in 6-8 hours. Other solids removal percentages are 95% for primary sewage effluent after 365 m of overland flow at the Melbourne system (Kirby, 1971), a reduction of 56.4 to 15.0 mg/liter for humus tank effluent at the high loading rate of 85 cm/day in an English study (Truesdale et al., 1964), and from 5215 to 63 mg/liter for sugar beet waste in a Nebraska study (Porges and Hopkins, 1955; Hopkins et al., 1956). CARBON AND OXYGEN DEMAND B. ORGANIC

Wastewater contains a variety of natural and synthetic organic compounds, usually not individually identified, but collectively expressed in terms of the biochemical oxygen demand (BOD, determined normally after 5 days incubation), the chemical oxygen demand (COD, usually determined with the dichromate technique), or the total organic carbon content (TOC, determined as the difference between total and inorganic carbon). The BOD and COD tests were developed primarily for oxygen regimes in aquatic environments. For land treatment, however, TOC content may be the most appropriate parameter. In addition to the carbonaceous oxygen demand, wastewater contains a nitrogenous oxygen demand for oxidation of organic or ammonia nitrogen to nitrate. The oxygen demands for other constituents are negligible, except perhaps for certain special wastes containing large amounts of sulfide, reduced iron, or other reduced compounds.

138

HERMAN BOUWER AND R. L. CHANEY

Good-quality secondary sewage effluent may have a BOD of around 10-20 mg/liter, a COD of 30-60 mg/liter, and a TOC content of 10-30 mg/liter. The relation between TOC and COD was evaluated as

TOC = 0.25 COD

+ 1.30

for secondary effluent (domestic and light industry) from the Phoenix area (Bouwer et al., 1974b). This relationship also includes measurements on renovated sewage water obtained by high-rate land treatment. The nitrogenous oxygen demand of secondary sewage effluent where most of the nitrogen is in the ammonium form, may be in the range of 100-200 mg/liter. Wastes from vegetable or fruit processing plants may have a BOD of several hundred to several tens of thousands of milligrams per liter (W. G. Knibbe, personal communication, 1973; California State Water Resources Control Board, 1968; Splittstoesser and Downing, 1969; Rose et al., 1971; Colston and Smallwood, 1973). Splittstoesser and Downing ( 1969) reported a COD/BOD ratio of 1.4-2 for vegetable processing effluents. Incompletely digested sewage sludge and liquid animal wastes have BOD’S of several hundred to several tens of thousands of milligrams per liter, depending on the density of the slurry or effluent (Loehr, 1968; Erickson et al., 1972). The COD of animal wastes may be 2 to 3 times as high as the BOD (Erickson et al., 1972). The soil with its biomass is extremely versatile and effective in decomposing natural and synthetic organic compounds. The processes can be divided into aerobic metabolisms where CO,, H,O, microbial cells, and NO,- and SO,*- are the main end products, and anaerobic metabolisms. The latter occur at a slower rate and are less complete, organic intermediates being formed. These include acids, alcohols, amines, and mercaptans. The end products of anaerobic decomposition consist of CH4,H,,NH,+, and H,S in addition to CO, and H,O (Miller, 1973). Organic carbon, whether supplied to the soil by the wastewater or produced in the soil by autotrophic bacteria, is a main factor in denitrification, since it supplies the energy for the denitrifying bacteria. Theoretically, aerobic conditions in the soil should prevail so that the total oxygen demand (sum of carbonaceous, nitrogenous, and other oxygen demands) of the waste load is balanced against the amount of oxygen entering the soil. Oxygen enters the soil (1 ) as dissolved oxygen in the wastewater applied (usually negligible), (2) as mass flow after the start of a drying or resting period, when the soil drains and air replaces the draining water in the soil, and (3) by diffusion from the atmosphere after the soil has drained. The deeper the water table and the higher the drainable pore space fraction of the soil, the more oxygen enters the soil as

LAND TREATMENT OF WASTEWATER

139

mass flow after infiltration stops. The longer the drying period, the more oxygen will enter by diffusion in relation to that which has entered the soil by mass flow. The depth to which oxygen can penetrate the soil by diffusion is limited and does not exceed a distance of about 1 meter in all but the most porous soils (Pincince and McKee, 1968; Lance et al., 1973). Lance et al. (1973) also found that the amount of oxygen entering by diffusion was 1.5 times greater than the amount entering by mass flow when laboratory soil columns were flooded with secondary sewage effluent on a 2-day wet, 5-day dry cycle, but twice that amount with a 9-day wet, 5-day dry cycle. Most of the oxygen was used to convert ammonium to nitrate and only a relatively small fraction was used to reduce COD. If wastewater is applied with sprinklers, considerable amounts of oxygen may enter the soil during the short periods between sprinkler revolutions, particularly on fast-draining soils. Some organic compounds are easier to degrade and exert a higher initial oxygen demand on the soil than others. The oxygen demand of secondary sewage effluent is sufficiently small and mostly due to readily degradable material. Thus, BOD is essentially completely removed as the effluent moves through the soil, even for high rate systems. In laboratory and field studies, prolonged flooding and obvious depletion of oxygen did not seem to affect the removal of BOD or COD (Bouwer et al., 1974b; Lance et al., 1973). Thus, anaerobic processes were also effective for BOD removal. This agrees with studies by Thomas and Bendixen (1969), who detected little or no effect of loading rate, duration of dosing, and temperature, on the organic carbon removal from septic-tank effluent passing through soil columns. Small, frequent applications, such as the 3 to 6 times per day rate recommended by Robeck et al. (1964) for best removal of COD, may be necessary if the wastewater contains high concentrations of organic compounds. Such schedules may increase the rate of biodegradation of these compounds in the soil, as was demonstrated by HaIIam and Bartholomew (1953) for plant residue. The BOD loading and removal at the Flushing Meadows Project was 100 kg/ha per day during flooding (Bouwer ef al., 1974b). At the Whittier Narrows Project, complete BOD removal was obtained from secondary sewage effluent at infiltration rates of about 0.6 m/day, or a BOD load also of about 100 kg/ha per day. In this rapid-infiltration system, 9-hour flooding periods were rotated with 15-hour drying periods. The sum of the carbonaceous and nitrogenous oxygen demands was about 750-1 000 kg/ha per day. Of this, about four-fifths was for nitrification of ammonium (McMichael and McKee, 1965). About three-fourths of the carbonaceous oxygen demand was removed in about 1.2 m of percolation of the effluent

140

HERMAN BOUWER AND R. L. CHANEY

through the soil. Erickson et al. (1972) reported BOD reductions from about 1200 mg/liter to 5 mg/liter when dairy waste was applied to the Barriered Landscape Wastewater Renovation System (BLWRS) at rates of about 2 cm/day, or a BOD load of about 240 kg/ha per day. Higher oxygen demands on the soil system and less complete removal of BOD are possible with effluents from vegetable or fruit processing plants, concentrated animal-waste slurries, or incompletely digested sewage sludges, where the BOD levels may be in the tens of thousands of milligrams per liter and the organic compounds readily biodegradable. D. M. Parmelee (personal communication, 1973) recommended that BOD loading rates not exceed 450 kg/ha per day for food processing plants. At these rates, W. G. Knibbe (personal communication, 1973) found that the COD of vegetable processing plant effluent was reduced from a range of about 500 to 2000 mg/liter to about 25 mg/liter in the first 50 cm of movement through soil (the COD of these effluents was about 1.7 times as high as the BOD). Higher loadings produced higher COD levels in the renovated water. Where soils are heavily overloaded with organic compounds in liquid wastes, solids in the wastewater and solids formed by bacterial activity in the soil may build up under the anaerobic conditions caused by the high oxygen demand. This will in turn cause a decrease in the infiltration rate, and hence in the oxygen demand exerted on the soil. Thus, soil may have some form of “self-defense” against excessive loadings of oxygen demand. Overland flow systems can also be effective in removing oxygen demand. provided the loading rate is sufficiently small and land has been sufficiently prepared to avoid channeling or short-circuiting. Thomas ( 1973b) reports a BOD reduction from an average of 150 mg/liter to a range of 8 to 12 mg/liter by flowing comminuted raw sewage over vegetated soil. Truesdale el al. (1964), using a much higher loading rate, found that BOD of humus tank effluent was reduced from a 16 to 24 mg/liter range to a 7 to 10 mg/liter range by overland flow in grassed plots. Wilson and Lehman (1967) obtained a reduction of only about 20% in the COD of primary effluent by flowing it through bermudagrass irrigation borders. For cannery wastes, the BOD was reduced from 580 mg/liter to 9 mg/liter in a Texas project (Law et al., 1970). A BOD reduction from 483 to 158 mg/liter was obtained for sugarbeet wastes in a field not very well graded and showing considerable channeling (Porges and Hopkins, 1955). Vela and Eubanks (1973) demonstrated that for land treatment of cannery wastes, soil bacteria, rather than enzymes or bacteria already present in the plant effluent, were responsible for the decomposition of organic matter. Thus, soil treatment can be expected to be more effective in reduc-

LAND TREATMENT OF WASTEWATER

141

ing the BOD of wastewater than, for example, lagooning or other treatment where the plant effluent will not be in contact with the soil. Only a small fraction of the bacteria population in the soil (16 out of 100 species) contributed directly to the decomposition of organic matter, which consisted of hydrolysis of the polymers followed by oxidation of the monomers. The other bacterial species probably contributed indirectly to the mineralization process. Because of this, bacteria in the soil did not correlate with the oxidative capacity of the soil. Shuval and Gruener’s (1973) statement that “. . . advanced wastewater renovation technology still cannot reduce COD or TOC to an absolute zero concentration . . .”, also applies to land treatment of wastewater. For example, while BOD was completely removed and COD reduced to the same level as that of the native groundwater at the Flushing Meadows Project, TOC values of the renovated water averaged 5 mg/liter after 9 m soil precolation (Bouwer et al., 1974b). The identity of this organic carbon is not very well known. Thus, it is subject to speculation regarding toxicants, teratogens, mutagens, and carcinogens. Perhaps this TOC can be reduced by treatment with a strong oxidant, such as ozone. Wastewaters, and particularly sewage effluent from industrialized communities, may contain hydrocarbons, detergents, pesticides, phenolic compounds, and other undesirable constituents. Usually, however, their concentrations are so low that with adsorption and gradual biodegradation generally occurring in the soil, few or no adverse effects are expected (Miller, 1973). Special precautions need to be taken, however, with land treatment of wastewaters containing unusually large concentrations of these compounds, or where porous soils, fissured rock, or cavernous limestones in the treatment fields offer little opportunity for appreciable renovation of the wastewater. Until further research has demonstrated that the refractory organics and other substances in renovated wastewater are harmless, direct use of such water (particularly sewage water) for domestic purposes is not recommended as a general practice (Long and Bell, 1972; American Water Works Association, Board of Directors, 1973; Ongerth et al., 1973). c.

BACTERIAAND VIRUSES

Of the numerous microorganisms possibly present in the wastewater, particularly in sewage effluents and sludges, the fate of pathogenic bacteria and viruses when the water moves through the soil is of utmost concern. The fecal coliform test is useful for indicating fecal pollution and, hence, possible presence of pathogens in surface water. For land treatment systems, low fecal coliform densities in the percolate or renovated water

142

HERMAN BOUWER AND R. L. CHANEY

probably mean absence or low levels of pathogenic bacteria or viruses. However, the absence of such organisms can be determined only by testing for specific microbial pathogens. The pathogenic bacteria commonly found in sewage effluent include Salmonella, Shigella, Mycobacterium, and Vibrio comma (Foster and Engelbrecht, 1973). Viruses include the enteroviruses and adenoviruses. The hepatitis virus is of great concern, but tests to detect its presence have not yet been developed. Other pathogens include the protozoa, such as Endamoeba histolytica, and helminth parasites, for example, ascaris and tapeworm ova. Fortunately, the soil is an effective filter and many reports indicate absence or very low levels of fecal coliforms or other organisms after water has moved one to several meters through soil (Stone and Garber, 1952; California State Water Pollution Control Board, 1953; Baars, 1964; McMichael and McKee, 1965; Drewry and Eliassen, 1968; Merrel and Ward, 1968; Romero, 1970; Young and Burbank, 1973; Bouwer et al., 1974b). On the other hand, situations have also been reported where appreciable numbers of microorganisms were detected in the renovated water after considerable distance of underground movement (Romero, 1970; Randall, 1970; Allen and Morrison, 1973). Such long underground travel distances of microorganisms are usually associated with macropores, as may be found in gravels, coarse-textured soils, structured clay soils, fractured rock, cavernous limestones, etc. The retention of microorganisms in the soil is largely due to physical entrapment for the larger organisms and to adsorption to clay and organic matter for viruses and other amphoteric organisms (McGauhey and Krone, 1967; Krone, 1968). Drewry and Eliassen (1968) found that virus adsorption was more rapid when the pH was below 7-7.5 than when the pH was higher. An increase in the cation concentration of the liquid phase in the soil also increased the adsorption of viruses. Young and Burbank (1973) reported virus removal in soil as a pH-dependent adsorption process. Cookson (1967) found that the adsorption of viruses by activated carbon could be described by a diffusion equation with a Langmuir adsorption boundary condition. Virus removal due to adsorption during phosphate precipitation was described by a pH-dependent Freundlich isotherm by Brunner and Sproul ( 1970). Microorganisms retained in the soil are subject to normal die-off, which usually takes several weeks to several months (Van Donsel et al., 1967). This is about the same as the die-off times in surface waters (Andre et al., 1967). Much longer survival times in soil have also been reported, however, such as 6 months to l year for salmonella (Rudolfs et al., 1950) and up to 4 years for Escherichia coli (Mallman and Mack, 1961). Miller

LAND TREATMENT OF WASTEWATER

143

(1973) found that fecal streptococci from sewage sludge survived up to 6 months in a clay soil, but not as long in coarser soils. The die-off of pathogens and other foreign microorganisms brought into the soil with the wastewater is due to the “homeostatic” reaction of the existing microbiological community in the soil (Alexander, 1971). This rejection of foreign organisms may result from production of toxins, lysis by enzymes, consumption by predatory protozoa, parasitic organisms, competition, and the general hostility of the soil environment to pathogenic organisms that are more at home in men and other warm-blooded creatures. Normally, fecal coliform bacteria are essentially completely removed after the water has traveled 1 m or at most 2 or 3 m through the soil. However, Bouwer et al. (1974b) found much deeper penetration of fecal coliforms below rapid-infiltration sewage basins after the basins were flooded following an extended drying or resting period. This was probably due to reduced entrapment of E. coli on the surface of the soil. The clogging layer of organic fines that had accumulated on the soil during flooding, forming an effective filter, was dry and partially decomposed after drying, thus yielding a more open surface of the soil and a less effective filter when flooding was resumed. Also, the bacteria population in the soil undoubtedly declined during drying because the nutrient supply was discontinued. Consequently, there was less competition from the native soil bacteria, and hence greater survival of the fecal coliforms when flooding was resumed. As flooding continued, however, fine suspended solids accumulated again on the surface of the soil and the bacteria population also increased, both resulting in increased retention of E. coli and return of the fecal coliform levels to essentially zero in renovated water sampled from a depth of 9 m. Almost all the removal of the fecal coliforms took place in the first 1 m of soil. Pathogenic and other foreign microorganisms may survive for some time in the soil, but they do not multiply (Benarde, 1973). The same has been observed for surface water (Deaner and Kerri, 1969). McMichael and McKee (1965) observed increased coliform counts in the soil with depth below spreading basins. They attributed this to a growth in Aerobacter aerogenes, which is a common soil bacterium of the coliform group, rather than to E. coEi. However, Masinova and Cledova (1957) reported that E. coli can sufficiently change in soil or water to give the biochemical tests more typical of the intermediate coliform types, including A . aerogenes. Cohen and Shuval (1973) studied the survival of coliforms, fecal coliforms, and fecal streptococci in surface water and sewage treatment plants. Fecal streptococci were generally more resistant than the other indicator organisms. In two systems, the survival of fecal streptococci paralleled the survival of enteric viruses better than the survival of coliforms.

144

HERMAN BOUWER AND R. L. CHANEY

The best insurance against contamination of groundwater by pathogenic microorganisms due to land treatment of wastewater is to allow sufficient distance between the land treatment facility and the point where groundwater leaves the aquifer for human consumption. Recommendations for this distance vary from about 10 m to 100 m (Romero, 1970; Drewry and Eliassen, 1968), depending on the soil type. Very coarse soils, wellstructured soils, and fractured or cavernous rocks cannot be expected to effectively retain microorganisms, and they should be avoided. In addition to moving underground, pathogenic organisms can spread from a land treatment site through the air, particularly if the wastewater is applied by sprinklers. Adams and Spendlove (1970) found that trickling filters of sewage plants emitted coliform bacteria into the air, and that E. coli could be sampled from the air as far as 1.2 km downwind. No matter what precautions are taken and how failsafe a land treatment system may be, some contamination and some survival of microorganisms may still take place. The simplest precaution against the possibility of infectious disease may be to chlorinate or otherwise disinfect all water for human consumption that is pumped from wells within underground traveling distance from land treatment sites or other possible sources of groundwater contamination. Most waterborne disease outbreaks are due to consumption of undisinfected groundwater (Craun and McCabe, 1973). These authors also recommend disinfection of groundwater as an easy and simple means to reduce the incidence of water-borne disease. Chlorination for virus control in wastewater is not effective if the water has a high suspended solids content. Thus, virus survival in chlorinated secondary sewage effluent is often observed (Mack, 1973). Culp et al. (1973) reported that disinfection for virus removal is most effective in water having a turbidity below 1 JTU (Jackson Turbidity Units) and as near as 0.1 JTU as possible. Chlorination to a free residual of 1 mg/liter with a contact time of 30 minutes is normally adequate to completely remove or inactivate all viruses. Since soil filtration of wastewater removes essentially all suspended solids, chlorination of the percolate or renovated water for virus and bacteria removal should be much more effective than chlorination of the wastewater prior to land treatment. In overland flow systems bacteria and viruses are removed primarily by settling and entrapment of suspended solids harboring the microorganisms. Detention times in overland flow systems normally are too short to reduce bacteria and viruses substantially by normal die-back, as usually happens in ponds or streams (Andre et al., 1967; Cohen and Shuval, 1973). Seidel (1966) reports much faster die-back of fecal coliforms in shallow impoundments where rushes (Scirpus lucustris and Spartina Townsendii) were growing than in impoundments without such vegetation.

LAND TREATMENT OF WASTEWATER

145

The removal of microorganisms in overland flow systems may possibly be improved if a flocculant such as alum or lime is added to the wastewater prior to land application. Viruses and other microorganisms may then become attached to the flocs and be detained on the treatment field. Excellent virus reductions, for example, have been obtained by flocculation and sand filtration of secondary sewage effluent (Berg et al., 1968). The addition of flocculants also helps to precipitate phosphates (Brunner and Sproul, 1970), and hence, may increase the phosphate removal in overland flow systems. Bacteria and viruses in the wastewater restrict the type of crop that can be grown on the land treatment fields. While entry of certain viruses into the plant through the root system has been observed (Murphy et al., 1958; Murphy and Syverton, 1958), normally the main concern is with pathogenic organisms that could collect on the surfaces of fruits and vegetables consumed raw (National Technical Advisory Committee, 1968). This committee suggests an interim guideline of not more than 5000 total coliform bacteria per 100 ml and not more than 1000 fecal coliforms per 100 ml, for irrigation water of crops where tops or roots are directly consumed by man or livestock. More conservative health guidelines were presented by Krishnaswami ( 1971 ) . A number of states have adopted quality criteria for irrigation with sewage effluent, sometimes based on what is theoretically desirable and practically achievable while avoiding criteria that are so stringent that they could not be met by normal irrigation water. As an example, the Arizona State Health Department ( 1972) requires secondary treatment, or its equivalent, if the sewage is used for irrigation of fibrous or forage crops not intended for human consumption, or orchard crops where the water does not come in contact with fruit or foliage. Secondary treatment and disinfection or equivalent treatment to reduce the total coliform density to 5000 per 100 m1 and the fecal coliform density to 1000 per 100 ml are required for higation of food crops that are sufficiently processed to destroy pathogens, or for orchard crops where the irrigation water does come in contact with fruit and foliage, or golf courses, cemetaries, etc. Tertiary treatment to produce a BOD and suspended solids content both of less than 10 mg/liter and disinfection or equivalent treatment to reduce the fecal coliform count to less than 200 per 100 ml are required if the effluent is to be used for irrigation of food crops that are consumed raw by man, or of play grounds, lawns, parks, etc., where children can be expected to play. One of the biggest questions with respect to the health hazards of land treatment of sewage effluent and other wastewaters is: What are acceptable levels of microorganisms, and particularly pathogens, in the renovated water or crops produced by such systems? Some persons may advocate

146

HERMAN BOUWER AND R. L. CHANEY

complete sterility, but this may not be necessary even if it were achievable. The environment as a whole is not sterile. Bacterial pathogens have been recovered from pristine mountain streams (Fair and Morrison, 1967). While some people may be alarmed to hear that fecal coliforms and, hence, possibly pathogenic bacteria, can travel through the air for long distances around sewage treatment plants (Adams and Spendlove, 1970), sewage treatment plant workers apparently do not have poorer health than people in other occupation groups. As a matter of fact, sewage plant workers were found to have the lowest absenteeism rate among a group of occupations studied, and this was attributed to the fact that “sewage workers were regularly immunized by their exposure to small amounts of infected material” (J. L. Melnick, as quoted by Benarde, 1973). Benarde (1973) also states that “one must be chary of the type of microbiological thinking that equates the presence of microbes with the potential for illness. The fact is that illness is an unusually complex phenomenon that does not have a 1 to 1 relationship to microbes.” Little is known about minimum infecting doses of pathogenic organisms and the combination of factors necessary to produce illness (Dunlop, 1968; Benarde, 1973). From a communicable disease standpoint, however, land treatment is far less hazardous than disposal of sewage effluent and other liquid wastes into rivers and streams (Benarde, 1973 ) . D.

NITROGEN

The nitrogen content of liquid waste may be as low as essentially zero for some cannery wastes and as high as 700 mg/liter for slurries of fresh swine waste (Erickson et al., 1972). Secondary sewage effluent generally contains 20-40 mg of nitrogen per liter (California Department of Water Resources, 1961) and sewage sludge 3-5% nitrogen (on a dry weight basis). Winery wastewaters may have 4-10 times as much nitrogen as domestic sewage (Schmidt, 1972). Wet sewage sludge (95% water) generally contains 1500-2500 mg of nitrogen per liter (Hinesly, 1973; Peterson et al., 1973). For cannery wastes, where the organic material consists essentially of cellulose and other carbonaceous materials, nonleguminous crops may actually become nitrogen deficient at high waste loadings in the same way that nitrogen deficiency may occur after application of crop residue containing less than about 1.3% nitrogen. The C/N ratio of these materials is usually about 35. For such wastes, release of significant amounts of nitrogen cannot, be expected unless the nitrogen content exceeds about 1.8 % on a dry weight basis (Wets, 1973). For secondary sewage effluent and similar liquid wastes, a significant

LAND TREATMENT OF WASTEWATER

147

amount of nitrogen can be removed by crop uptake if the nitrogen loading rates are not much more than the fertilizer requirement of the crop, which generally ranges from 50 to 600 kg/ha. For example, wheat removed 92 and 60% of the nitrogen applied with sewage effluent at the Pennsylvania State Project, using applications of 2.5 and 5 cm per week, respectively (Kardos, 1967). Sewage sludge should be applied so that the nitrogen load is about the same as the nitrogen requirements of the crop (Hinesly, 1973). When sewage effluent is applied in small amounts, the soil is predominantly aerobic and the nitrogen in the effluent (which is mostly in the ammonium form) will be converted to nitrate. The fate of this nitrogen will probably be about the same as that of fertilizer nitrogen; i.e., about 50% will be used by the plants, 25% will be lost by denitrification, and the remaining 25% will be lost by other processes, such as ammonia volatilization (Woldendorp, 1963). Since soils where wastewater is frequently applied may have high water contents, denitrification losses may be higher in land treatment fields than in normal agricultural fields, particularly if the wastewater contains organic carbon that can be used as an energy source by the denitrifying bacteria. However, as long as the wastewater is applied in normal irrigation schedules (for example, once every 1 to 3 weeks), nitrogen entering the soil in excess of fertilizer requirements tends to be converted to nitrate and moved down to the groundwater. If sewage effluent is used as the sole water source for irrigation in warm, arid regions, nitrogen loading may exceed crop uptake and normal denitrification and other losses. The excess nitrogen will then move down as nitrate to the groundwater. Thus, increases in the nitrate content of the groundwater below sewage irrigated fields are frequently observed (Matlock et al., 1972; Schmidt, 1972; Wells and Sweazy, 1973). Because the salt concentration of the Ieachate from the root zone of an irrigated crop may be 3 to 10 times as high as that of the irrigation water (Bouwer, 1969 ) , nitrate levels in the groundwater below these fields could exceed those in the sewage effluent. For high-rate systems, complete conversion of the nitrogen to the nitrate form is commonly observed if the wastewater applications are relatively short and frequent. This frequency may range from 3 to 6 short applications per day (Robeck et al., 1964), or about 8 hours flooding per day and 16 hours drying (McMichael and McKee, 1965) to 2 or 3 days flooding alternated with about 5 days drying (Bouwer et al., 1974b). For secondary sewage effluent or similar wastes with a relatively low organic carbon content, most of the organic carbon will also be oxidized under the predominantly aerobic soil conditions with these frequencies, leaving insufficient carbon for subsequent denitrification. As shown in Fig. 1 for July and August, the nitrate nitrogen concentrations in the renovated

148

HERMAN BOUWER AND R. L. CHANEY

30

20

10

0

FIG.1. 'Total nitrogen in effluent (0-0)

and nitrate ( Q - - - A ) and ammonium content of renovated water samples at 9.1 meters below the basins of the Flushing Meadows Project (Bouwer et al., 1974b).

(0-0)

wastewater will then be about the same as the total nitrogen concentrations in the wastewater (Lance and Whisler, 1972; Bouwer et al., 1974b). Denitrification is the most important process whereby nitrogen applied with wastewater in excess of crop requirements can be removed from the soil-water system (Lance, 1972). This requires the presence of nitrates and organic carbon under anaerobic conditions (Broadbent and Clark, 1965; Lance, 1972; Bouwer, 1973b). About 1 mg of organic carbon is required for each milligram of nitrate nitrogen to be denitrified. Denitrification in land treatment systems should be easiest to accomplish if the nitrogen in the wastewater is already in the nitrate form, and the wastewater contains sufficient organic carbon. Then all that is necessary to stimulate denitrification is to maintain anaerobic conditions in the soil by flooding for long periods (assuming that other factors, such as pH and temperature, are favorable for denitrifying bacteria). If organic carbon is limiting, it may be added by incorporating crop residues into the soil or by adding carbon sources to the wastewater. If the nitrogen in the wastewater is predominantly in the organic or ammonium form, as is usually the case with sewage water, an aerobic phase in the soil is necessary first, to convert the nitrogen to nitrate, before denitrification can take place. During this aerobic phase, organic carbon in the wastewater also will be oxidized by the numerous heterotrophic aerobic bacteria in the soil, leaving less organic carbon for denitrification

LAND TREATMENT OF WASTEWATER

149

when the wastewater moves into anaerobic zones. This could limit subsequent denitrification for secondary effluent and similar wastewaters which already contain relatively low organic carbon levels. The C/N ratio for secondary sewage effluent, for example, is of the order of 0.7. Denitrification following nitrification was successfully achieved by Erickson et al. (1972) for fresh swine and dairy waste slurries in the Barriered Landscape Wastewater Renovation System (BLWRS) . This is a specially constructed soil filter with an artificial barrier at a depth of about 2 m to create a perched groundwater table below which anaerobic conditions can prevail. Drains along both sides of the barrier collect the wastewater in renovated form. By applying the wastewater in frequent, small amounts (for example less than 2 cm per day), the upper portion of the soil is sufficiently aerobic to convert the nitrogen in the wastewater (concentration 3 10-660 mg/liter, mostly as organic nitrogen and ammonium) to nitrate. Because the organic carbon of the wastewater is high ( a COD of 2000-3000 mg/liter) , sufficient organic carbon is left for denitrification when the waste liquid moves from the upper aerobic zone into the lower anaerobic zone below the perched water table. This system removed 96-99% of the nitrogen from the wastewater. Additional nitrogen removal was obtained by mixing organic carbon as corn cobs, molasses, etc. in the soil above the barrier during construction. For the summer period, denitrification removed about 700 kg of N per ha per month. This is much higher than denitrification rates in normal agricultural fields, which may be about 25 kg/ha per growing season. Denitrification in secondary sewage effluent was achieved below the high-rate infiltration basins of the Flushing Meadows Project when relatively long flooding and drying periods were used; for example, 2 weeks flooding alternated with 10 days drying in summer and 20 days drying in winter (Bouwer et al., 1974b). With these schedules, oxygen became depleted in the soil below the basins shortly after flooding was started, so that nitrification could no longer occur. This left the nitrogen in the ammonium form, which was then adsorbed by the cation exchange complex of the soil, yielding both low nitrate and ammonium levels in the renovated water (Fig. 1). Flooding had to be stopped before the cation exchange complex became saturated with ammonium; otherwise, the ammonium content of the renovated water increased (Lance and Whisler, 1972). The oxygen entering the soil during subsequent drying caused bio-oxidation of the adsorbed ammonium to nitrate, part of which was then denitrified in anaerobic microenvironments. Such microenvironments could exist even in predominantly aerobic zones due to locally low oxygen diffusion rates and oxygen sinks caused by nitrification or decomposition of organic material. Nitrate not denitrified in this way was then

150

HERMAN BOUWER AND R. L. CHANEY

leached out by the newly infiltrating effluent when flooding was resumed. Some of this nitrate could be denitrified as it moved to deeper anaerobic zones. The rest of the nitrates stayed in the water and caused a nitrate peak in the renovated water collected from wells in the area upon arrival of the newly infiltrated water (Fig. 1 ). Flooding and drying should be scheduled so that the amount of ammonium adsorbed during flooding is not more than can be nitrified during drying (Lance er al., 1973). Otherwise, some adsorbed ammonium will not be oxidized, causing less ammonium to be adsorbed during subsequent flooding and hence an increase in the ammonium content of the renovated water. When this is observed, a sequence of short, frequent flooding periods or several long drying periods should be used to nitrify the adsorbed ammonium (Bouwer et al., 1974b). The total nitrogen concentration in the renovated water between NO, peaks was sometimes 80% less than that of the secondary effluent (Fig. 1) , During NO, peaks, the renovated water often contained as much total nitrogen as the effluent, and sometimes even more. The total nitrogen removal for sequences of sufficiently long flooding and drying periods to yield NO, peaks in the renovated water was about 30%. This figure was obtained by combining nitrogen relations in effluent and renovated water with infiltration rates in the basins (Bouwer et al., 1974b). The 30% removal agreed with the percentage obtained from the average total nitrogen concentration in the renovated water from the more distant wells, where the NO, peaks were attenuated by mixing and dispersion (Bouwer er al., 1974b). It also agreed with results from laboratory studies (Lance and Whisler, 1972). Since the annual nitrogen load was about 25,000 kg/ha, the 30% removal rate corresponded to a nitrogen loss of 7500 kg/ha per year, or about 625 kg/ha per month. This is close to the 700 kg/ha per month removed by denitrification in the BLWRS (Erickson et al., 1972). Most of the 70% of the nitrogen not removed in the Flushing Meadows Project is concentrated in the NOs peaks (Fig. 1) . Laboratory studies have indicated that if the portions of the renovated water containing the NOa peaks are pumped back into the basins, where they can mix with the effluent and pass once more through the soil, the total nitrogen removal can be increased to almost 80% (Lance and Whisler, 1973). These authors also increased nitrogen removal by adding organic carbon to the effluent prior to infiltration, or by reducing the infiltration rate. The latter was accomplished by decreasing the depth of ponding above the soil. At nitrogen loadings of 25,000 kg/ha per year, crop uptake of nitrogen is insignificant. However, crops may increase the nitrogen removal by stimulating denitrification in the root zone due to exudation of organic carbon and the creation of low oxygen levels, as reported by Woldendorp (1963) and Stefanson (1973). Some evidence of lower nitrate contents in the reno-

LAND TREAWENT OF WASTEWATER

151

vated water below grass-covered basins as compared to that below nonvegetated basins, was obtained at the Flushing Meadows Project (Bouwer et al., 1974b). However, these lower nitrate contents could also be the result of inhibitatory effects of roots on nitrification, as reported by Moore and Waid (1971). The slower release of nitrate resulting from this action could lead to reduced nitrate leaching during the initial stages of a new flooding period, and to more denitrification in the biologically active upper soil layers. Nitrogen from wastewater treated by overland flow or spray runoff systems can be removed by adsorption of ammonium to the soil and by denitrification in the biologically active surface layer of the soil. Organic or ammonium nitrogen in the wastewater can be converted to nitrate in the overland flow sheet, which is in direct contact with atmospheric oxygen. Shallow flow and relatively long detention times are required for significant nitrogen removal. Thus, while high loading rates yielded little or no nitrogen removal in overland flow systems (Truesdale et al., 1964; Wilson and Lehman, 1966), lower rates showed nitrogen reductions from an average of 23.6 mg/liter in the raw sewage to a range of 2.2 to 7.2 mg/liter in the runoff, depending on loading rate and age of the system (Thomas, 1973b). Law et al. (1970) reported nitrogen reductions from 17.2 to 2.8 mg/liter in an overland flow system for treatment of cannery waste.

E. PHOSPHORUS The phosphate content of secondary effluent varies widely among municipalities (Pound and Crites, 1973a,b). The observed range is about 0.5 to 40 mg of phosphorus per liter. The EPA “theoretical effluent” contained . waste dis10 mg of phosphorus per liter (Thomas, 1 9 7 3 ~ ) Industrial charges can reduce the phosphate concentration in municipal sewers, or greatly increase it. The phosphate content of a municipality’s wastewater may vary with time. The phosphate in the wastewater used at the Pennsylvania State University project fell steadily from 9.7 mg of phosphorus per liter in 1963 to 4.2 mg in 1970 (Sopper and Kardos). A similar decrease was reported by Bouwer et al. ( 1974b). Technology has been developed to minimize effluent phosphate by additions of phosphate precipitant chemicals (lime, aluminum sulfate, ferric chloride) during sewage treatment (Barth and Ettinger, 1967). The effluents from these processes contain low levels of phosphate. Also, alternative biological technology has been developed to reduce effluent phosphate to as low as 0.55 mg of phosphorus per liter (Levin et al., 1972). Total sewage phosphate removed by conventional treatment processes ranges from 20 to 90%. This variation led to the search for the improved biological technology to remove phosphate (Levin et al., 1972). The treat-

152

HERMAN BOUWER AND R. L. CHANEY

ment processes generally lead to hydrolysis of sewage polyphosphates to orthophosphate (Bunch et al., 1961). Polyphosphates are also rapidly hydrolyzed in soil (Gilliam and Sample, 1968). The reactions of wastewater phosphate in soils recently have been described by Ellis (1973), Ellis and Erickson (1969), Lindsay (1973), and Thomas ( 1 9 7 3 ~ ) The . Langmuir adsorption isotherm has been applied to the adsorption of phosphate in soils by numerous authors (Griffin and Jurinak, 1973 ) . Schneider and Erickson ( 1972), Ellis ( 1972), and Ellis and Erickson (1969) described the use of Langmuir constants to estimate the phosphate adsorption capacity of particular soils from a solution containing 10 mg of phosphorus per liter, The adsorbing capacity of the soils seemed to be related to the iron and aluminum contents. For example, the phosphate absorbing capacity of some highly weathered soils was much higher in the B-horizon than in the A-horizon, presumably because iron and aluminum oxides had accumulated in the B-horizon. The calcareous soil used in the study had a low phosphate adsorbing capacity. Such soils contain little iron and aluminum oxides, and phosphate removal may be due to precipitation of calcium phosphates. Ellis ( 1973) noted that the adsorption capacity of phosphorus-saturated soil was regenerated during 3 months’ incubation. The regeneration was probably due to crystallization of adsorbed phosphate into less soluble compounds and to the production of more iron and aluminum oxides by weathering. Thus, use of Langmuir constants to calculate the potential life of a land treatment site can lead to serious underestimation. On the other hand, presumption that all the hydrous oxides of iron and aluminum will be available to adsorb phosphate (Bauer and Matsche, 1973) can lead to overestimation of the life of a site. Schneider and Erickson (1972) compiled phosphate adsorption capacities for Michigan soils based on Langmuir constants. Griffin and Jurinak (1973) modified Langmuir adsorption isotherms to account for two simultaneous adsorption reactions. A convenient one-point method has been developed by Bache and Williams (1971) to determine Langmuir constants. Various phosphorus compounds also precipitate in soils depending on concentrations of phosphate, Fe3+, Al”’, Ca’+, F-, CO,“, and on pH. Lindsay and Moreno (1960) developed a solubility vs pH diagram for variscite, strengite, fluoroapatite, hydroxyapatite, octacalcium phosphate, and dicalcium phosphate dihydrate as end products of the adsorption-precipitation sequence. The kinetics of some of these phosphate precipitation reactions are relatively slow, and equilibrium with the predicted crystalline precipitates should not be expected. However, phosphate precipitation may be the main mechanism for phosphate removal from wastewater in calcareous soils.

LAND TREATMENT OF WASTEWATER

153

With low-rate systems, so little phosphate can be applied that crop removal balances phosphorus additions with wastewater. Sometimes, phosphorus fertilizer may have to be added to maintain fertility. Kardos and Sopper (1973) described renovation of secondary sewage effluent by sampling from porous cups installed 15, 60, and 120 cm deep in a soil cropped to corn and Reed canarygrass, and in two forested soils. Although the phosphorus level in the soil water at the 15-cm depth was increased by wastewater application, the phosphorus concentration at 120 cm was only slightly affected. Areas covered by Reed canarygrass received about twice as much phosphorus as areas in corn, but the phosphorus concentration in the soil water at the 120-cm depth was lower in the Reed canarygrass areas than in the corn areas. The phosphorus concentration in the soil solution was higher where effluent was applied at 5 cm/week than at 2.5 cm/week. Through 1970, the removal of phosphorus from the wastewater was about equal on Hubersburg silt loam and Morrison sandy loam. Sopper and Kardos (1973) reported the crop responses to wastewater application, and crop removal of phosphorus. In the early years of their project, wastewater phosphorus at 5 cm/week application supplied as much as 134 kg of phosphorus per hectare per year, clearly in excess of crop removal. However, by 1971, the wastewater phosphorus application had dropped considerably, and corn silage or Reed canarygrass removed more phosphorus than was added with wastewater. Forest crops did not remove nearly as much phosphorus. In 1970, only 19% of the applied phosphorus was removed. Hook et al. (1973) reported the soil phosphorus relations for these same plots. The Hubersburg silt loam showed considerable increase in Bray-extractable phosphorus in the surface 30 cm of soil, but little change in the second 30 cm. Morrison sandy loam soil showed increased extractable phosphorus as deep at 120 cm. They considered three bases for the deeper penetration of phosphorus in the Morrison soil: (1 ) crops had not been removed; ( 2 ) the sandy loam has a greater hydraulic conductivity, thus phosphorus in percolating solution has less time to react with particle surfaces; and ( 3 ) the concentrations of free iron and aluminum oxides are much lower in the Morrison soil. Day et al. (1972) found that the available phosphorus was increased in the Ap-horizon after 14 years of irrigation of calcareous soils with sewage effluent; available phosphorus was not significantly increased in the C-horizon. Relatively low application rates were used, and no data have been reported on soil solution levels of phosphate at different depths. Kirby (1971) reported that the moderately acid soils of the infiltration system at Werribee, Australia, removed about 80% of the phosphorus in the (settled) sewage. By 1958 (after 70 years of operation) as much as

154

HERMAN BOUWER AND R. L. CHANEY

3200 mg of phosphorous per kilogram of soil (surface 10 cm) had accumulated in the irrigated areas, with considerable movement below 30 cm (Khin and Leeper, 1960). The 3200 mg of phosphorus per kilogram of soil is about 10 times the adsorption maximum for high adsorbing soils determined by Ellis and Erickson (1969) using Langmuir constants. More recently, R. D. Johnson, R. L. Jones, T. D. Hinesly, and D. J. David (personal communication, 1974) found greater accumulation and deeper penetration of phosphorus than did Khin and Leeper. In characterizing the soil phosphate in irrigated and control areas, .Khin and Leeper (1960) found that one-third of the phosphorus was organic bound. Crop removal accounted for little phosphorus removal because grazing cattle and sheep returned about 85% of dietary phosphorus to the soil. With high-rate systems, the phosphate applied to the soil greatly exceeds crop uptake. At the Flushing Meadows Project (Bouwer et al., 1974b), the annual application was about 10,000 kg of phosphorus per hectare. The PO,-phosphorus concentration of the renovated water 9 m below the basins was 30-70% less than in the sewage effluent, depending on hydraulic loading and PO,-phosphorus content of the effluent. Further underground travel through the predominantly sandy and gravely materials resulted in additional PO, reduction. Wells 6 m deep and 30 m away from the basins yielded renovated water with phosphorus concentrations of 1-3 ppm, or removal percentages of 70-90%. After 5 years of operation of the project, during which a total of almost 50,000 kg of PO,-phosphorus was applied per hectare, the phosphorus removal efficiency of the system was still stable. Since the soils were calcareous sands and gravels, which contained little or no iron and aluminum oxides and less than 2% clay, the phosphorus was probably removed by precipitation of calcium phosphates. Larson (1960) found that 75% of the 2700 kg of phosphorus per hectare per year applied with wastewater was removed after 9 m of movement through coarse soil. Significant reductions in phosphorus concentrations of wastewater have also been observed in overland-flow systems. Kirby ( 197 1 ) observed 35 % removal from sewage effluent at the Werribee, Australia, system. Law et al. ( 1970) reported reductions in phosphorus-concentrations of cannery waste from 7.4 to 4.3 mg/liter due to overland flow, with daily applications. The phosphorus removal was essentially doubled when the frequency of application was reduced to three times per week. Thomas (1973b) reported phosphorus reduction in raw sewage from an average of 10 mg/liter to an average of 4.0 to 5.4 mg/liter, depending on age of the overland-flow system and loading rate. Most work on phosphorus removal from wastewater applied to soil has consisted of determining phosphorus concentrations in the renovated water.

LAND TREATMENT O F WASTEWATER

155

More knowledge of the reaction kinetics of phosphorus precipitation and adsorption is needed before the phosphorus removal capacity, and hence the useful life, of a land treatment system can be accurately assessed.

F. FLUORINE Wastewater is enriched in fluoride by industrial and domestic additions. Many cities now add fluoride to the drinking water so that it contains about 1 mg of fluoride per liter. Fluoride is adsorbed by various soil components, especially hydrous aluminum oxides, according to the Langmuir adsorption equation (Bower and Hatcher, 1967). The adsorption of wastewater F by soil and its subsequent equilibration with fluorite (CaF,) and fluoroapatite leads to both retention of fluoride in the soil and control of injury to plants and food chain. The Ca2+added with wastewater maintains the soil Ca level high enough to prevent fluoride injury. Injury from added NaF has been demonstrated in acidic soils low in Ca, but not in well-limed soils (Prince et al., 1949). Crops raised on fluoride-enriched soils show little increased F uptake as long as the soil is near neutral pH. A recent review by Brewer (1966) summarizes plant and soil relationships of fluoride. Larsen and Widdowson ( 1971 ) have examined “labile” fluoride in soils. The maximum limit of fluoride in irrigation water for continuous use on all soils is 2 mg of fluorine per liter (National Academy of Science-National Academy of Engineering, 1972). Few surface waters exceed 1 mg of fluorine per liter. Very little study of the fate of fluoride during wastewater irrigation has been reported. Bouwer et d.(1974b) reported that the fluorine content of secondary effluent was reduced from 4.1 to 2.6 mg/liter after 9 m of movement through sandy material in a high-rate system, and reduction continued with further movement through the coarse textured soil. The fluoride removal somewhat paralleled the phosphate removal, suggesting precipitation of fluorapatite and fluorite. Soil retention of fluorine should be related somewhat to kinetics of water movement and length of path (Bower and Hatcher, 1967). Possibly, irrigation water containing higher levels of fluorine will lead to slightly higher fluorine content of plants (Rand and Schmidt, 1952), perhaps because of the temporary surface adsorption of fluorine in forms more available to plants than fluorite and fluorapatite. G . BORON

As borates are substituted for phosphate in household detergents the boron content of sewage effluent may increase. Bouwer et al. (1974b)

156

HERMAN BOUWER AND R. L. CHANEY

found that boron in the secondary effluent from the city of Phoenix, Arizona, had risen from 0.4 mg/liter in 1969 to 0.9 mg/liter in 1971. The effects of boron on crops and soils have been studied for many years because they are a hazard in some natural irrigation waters in arid areas. Differences in crop sensitivity to boron in irrigation water have been identified (Eaton, 1944; Richards, 1954). Sensitive crops showed toxicity at 0.5-1 mg of boron per liter, semitolerant crops at 1-2 mg of boron per liter, and tolerant crops at 2-4 mg of boron per liter. The maximum level of boron in irrigation water for continuous use on all soils is 0.75 mg of boron per liter (National Academy of Science-National Academy of Engineering, 1972); this level is based on studies of the boron-sensitive citrus crops. Ellis and Knezek (1972) summarized the reactions of boron with soils. Boron adsorption appears to occur on: (1 ) iron and aluminum-hydrous oxide coatings on clay minerals; ( 2 ) iron and aluminum oxides; ( 3 ) clay minerals, particularly micaceous-type clay minerals; and (4) magnesiumhydroxy clusters or coatings that exist on the weathering surface of ferromagnesian minerals. Several authors have found that boron adsorption can be described by the Langmuir adsorption equation, at least over a limited range of concentration. Studies of soil adsorption of boron particularly relevant to irrigation have been made using soil columns (Biggar and Fireman, 1960; Hatcher and Bower, 1958; Okazaki and Chao, 1968; Rhoades et al., 1970; Tanji, 1970). Rhoades et al. (1970), studying leaching of soils to remove naturally occurring excess boron, found that weatherable boron can be released during incubation after the leachable adsorbed boron has been removed. Thus, wastewater boron will be retained until its concentration reaches equilibrium with the soil solution boron. Wastewater irrigation effects on plant, soil, and percolating water boron have been reported in only a few studies. Bouwer et al. (1974b) found essentially no boron removal in the sandy and gravelly soils below their infiltration basins. On the other hand, Sopper and Kardos (1973) found that the heavier soils were still retaining up to 90% of the added boron (as measured by soil solution extracted at 1.2 m) . The soil solution boron was higher where 5 cm of wastewater were added per week (ca. 0.09 mg/liter) than where 2.5 cm was added per week (ca. 0.06 mg/liter) (control was ca. 0.03 mg/liter) . The added wastewater contained about 0.29 mg of boron per liter. Analysis of several crops growing in the different experiments showed only a slight increase in foliar boron content. In humid areas rainfall will leach some of the boron adsorbed from wastewater. Clearly, however, questions remain about the safety of boron additions with wastewater added in excess of crop requirements, especially when some wastewaters already contain boron in excess of the recommended limit for irrigation water (Pound and Crites, 1973a).

LAND TREATMENT OF WASTEWATER

157

H. METALS The concentration of heavy metals in various wastewaters is generally quite low if there is no specific metal pollution. Most of the metals in sewage end up in the sludge. Brown et al. ( 1973), found that the higher the influent metal level or the higher the suspended solids removal, the higher the observed metal removal efficiency. Cadmium removal was poor (averaging 16% ), apparently because of its low concentration. Argo and Culp (1972) and Nilsson (1971 ) summarized metal removal by different sewage treatment practices. Mytelka et al. (1973) reported the contents of silver, cadmium, cobalt, chromium, copper, iron, mercury, manganese, nickel, lead, and zinc in raw and treated sewage collected from treatment plants in the Interstate Sanitation District (New York, New Jersey, and Connecticut). The range and median values for selected elements are presented in Table I. Blakeslee (1973) reported the total and dissolved cadmium, chromium, copper, mercury, nickel, lead, and zinc of wastewater effluents from 5 8 treatment plants in Michigan. The range and median values are shown in Table 11. The amount oi metals that would enter the soil with the wastewater could be considerably lower than permitted under the 1972 Irrigation Water Standard (National Academy of Science-National Academy of Engineering, 1972), as shown in Table 111. The reactions of heavy metals with soils and uptake by plants have recently been reviewed by several authors (Allaway, 1968; Chaney, 1973; Ellis and Knezek, 1972; Hodgson, 1963; Jenne, 1968; Knezek, 1972; TABLE I Range and Median Heavy Metal Contents of Wnstewater Treatment Plant Effluents in t h e Tnterstate Sanitation Districtn Range

Elemerit

rdow

(Ing/liter)

High (mg/liter)

Median (ing/liter)’


6.4


05 <0.05

0.05

<0.05 < 0 , 0001

5 , !I 0.1%

<0.05 <0.05 0 .1 0 n . QOOB

<0.1

1.5 G.O



6.8

50.0

From Mytelka el al. (1973).

<0.1

< 0 . 05 Q,l5

158

HERMAN BOUWER AND R. L. CHANEY

TABLE I1 Range and Median Heavy Metal Contents of 58 Wastewater Treatment Plant Effluents in Michigan" Range Element Cd Cr cu Hg Ni Pb Zn a

Low (mg/liter)

High (mg/liter)

Median (mg/liter)

10.005

<0.005

0.01
0.15 1.46 1.3 0,001


5.4

0.03

4.7


0.025 0.04 0.2 0.02 0.05 0.19

1.3

From Blakeslee (1973).

TABLE I11 Comparison of the Amount of Toxic Metals That Would Be Applied with a Typical Effluent vs That Sdded under the 1972 Irrigation Water Standards@ 1972 Irrigation Water

Standard

Element Cd cu Ni

Pb Zn a

Concentration Amountb (mg/liter) (kg/ha/year) 0.01 0.20

0.20 5.0 2.0

0.2 4.0 4.0 100 40

National Academy of Science-National

Typical effluent Concentration (mg/liter)

Amountb (kg/ha/year)

< O . 005

0.1

0.10 0.02 0.05 0.15

2.0 0.4

1 .o

3.0

Academy of Engineering

(1974).

Applied at 5 cm/week for 40 weeks/year.

Leeper, 1972; Lindsay, 1973; Lisk, 1972; Mitchell, 1964; Murrmann and Koutz, 1972; Page, 1973). The metal ions are bound by clay, organic matter, and hydrous oxides components of the soil. A high pH favors immobilization of some metals in the soil. Jenne (1968) suggested that the hydrous oxides of manganese and iron were very important in adsorption of heavy

LAND TREATMENT OF WASTEWATER

159

metals. Stable organic matter in the soil may significantly contribute to the binding of metals (Ellis and Knezek, 1972); the chelation of copper by organic matter is especially important. Bondietti and Sweeton (1973) observed that the apparent stability constant of cadmium with soil organic matter is dependent on the percentage of the cadmium binding capacity filled. Apparently, small amounts are bound to relatively specific sites of the organic matter which form higher stability chelates with the metal ion. Large amounts of metals (>50% saturation) are held much less firmly, more nearly like cation exchange than chelation. The heavy metals do enter into the general cation exchange reactions with clays and organic matter, in addition to the chelation reactions with organic matter. The metal ions in wastewater should occur largely as low molecular weight soluble chelates which could affect the kinetics and extent of metal reactions and movement in the soil. Norvell ( 1972) recently summarized the reactions of metal chelates in soils. Volk (1970) found that zinc EDTA moved rapidly through soil. Thus the presence of strong chelating agents in wastewater could lead to deeper penetration of heavy metals into the soil. On the other hand, the physical filtering activity of soils could remove high molecular weight metal complexes from the wastewater. The ability of soils and plants to remove heavy metals from wastewater is generally considered to be a benefit of wastewater irrigation. In the past, raw sewage was often applied and industrial release of heavy metals to the sewers was not controlled. Thus, irrigation with sewage sometimes has led to substantial accumulations of metals in soils. Phytotoxicity to metalsensitive crops has sometimes been observed. Blood (1963) reported that “on a sand land farm with a history of annual applications of [raw] sewage effluent over 80 years, the soil now contains 500 ppm of Zn down to 25 cm, whilst soil from adjacent land receiving no effluent contains less than 20 ppm.” Sugar beets failed where the soil pH fell below 6.4. Rohde (1962) reported on heavy metals accumulated in the soils of the (raw) sewage farms of Berlin and Paris. He found that copper and zinc were significantly increased in the irrigated soil, and that local areas where “exhaustion” was observed contained higher levels of metals than nearby areas where plants looked healthy. The problems appear to have been corrected by more careful management of soil pH and use of crops less sensitive to toxic metals. The Paris farm may have experienced zinc-induced manganese deficiency; foliar sprays of manganese can be used to correct this deficiency (Trocme ef al., 1950). A more recent study of the Werribee Sewage Farm at Melbourne, Australia reported by R. D. Johnson, R. L, Jones, T. D. Hinesly, and D. J. David (personal communication, 1974) has revealed that substantial levels

160

HERMAN BOUWER AND R. L. CHANEY

of metals have accumulated during 70 years of application of raw and settled sewage. Six irrigated locations and one unirrigated location were sampled at three depths. The samples were taken from the center of the irrigation paddocks; the accumulation of metals, etc., would have been much higher nearer to the sewage outfall because flood irrigation would lead to grass filtration near the outfall. The samples were analyzed for total and 0.1 N HCI extractable metals (Table IV). It is clear that these metals have accumulated in the soil, and that considerable leaching of the metals has occurred. The extent of leaching was higher at locations with higher total metal accumulation. The zinc content of the largely perennial rye grass crop was 50 mg/kg on unirrigated soil and 104 mg/kg on irrigated soil (mean of four irrigated sites). The inactivation or reversion of non-wastewater-related metals observed by other workers (Follett and Lindsay, 1971) was not observed at Werribee. Nearly 80% of all zinc added remained in a form extractable by 0.1 N HC1. It is not clear whether the breakdown of organic matter, cycling of redox potential, or other factors related to the addition of sewage effluent led to this result. Perhaps, the observations of Follett and Lindsay have limited application when the metals are applied with sewage effluent. The existence of some of these older wastewater irrigation sites with accumulated metals, phosphate, etc., provide a valuable opportunity to determine the environmental impact of land disposal of wastewater. Several TABLE IV Total Metals and 0.1 N HC1 Extractable Heavy Metals in Soils at Werribee Soil Total metals Irrigated (6) Unirrigated (1)

320

0-2.5 2.5-18 15-45 0-2.5 2.5-18 25-45

82 60 51

144 91 70 60 84

47 42 56

23

30

21 47

30 58

-

-

-

0.1 N HCl extractable metals ~~

Irrigated (6) Unirrigated (1)

0-2.5 2.5-15 25-45 0-2.5 2.5-15 15-45

~~~

210 70 17 16 1.8 0.9

~

2.0 4.0 2.9 1.1 0.8 2.2

13.0 4.8 4.7 1.9 1.0 3.0

1.78 0.57

0.%3 0.17 <0.13 <0.13

LAND TREATMENT OF WASTEWATER

161

recent surveys have been conducted to identify these sites (Sullivan et al., 19?3; Pound and Crites, 1973b). Wheatland and Borne (1961) found considerable removal of chromium, copper, manganese, nickel, lead, and zinc from river water but did not examine soil parameters involved. Lehman and Wilson (1971 ) studied removal of metals from wastewater by filtration through mostly sandy calcareous soils. The concentrations of iron, manganese, nickel, copper, zinc, lead, and cadmium were effectively reduced. Aerobic conditions in the soil resulted in greater immobilization of the metals than anaerobic conditions. Metal concentrations in the secondary effluent and in the renovated water from a well 27 m from the basins of the Flushing Meadows project showed considerable removal of copper and zinc, but not of cadmium and lead (Table V ) . The maximum permissible limits for these metals in raw public water supplies, listed by the National Technical Advisory Committee (1968), are shown for comparison. Metals did not accumulate in the surface 1.5 m of the soil (R. L. Chaney, R. C. Rice, and H. Bouwer, unpublished, 1972) probably because of the low organic matter and clay content of the basin soils, the low retention times of the water in the surface soils, and the low metal concentrations in the effluent. A study of metal accumulation in infiltration basins at Ft. Devens, Massachusetts, by E. P. Meier and S . A. Schaub (personal communication, 1973) revealed a peak of heavy metals which coincided with an organic matter accumulation zone at 45 cm. The organic matter in this zone and its metal content appeared to increase during winter and decrease during summer. K. W. Brown, C. E. Woods, and J. F. Slowey (personal communication, 1973) at Texas A&M are studying soil retention of metals from wastewater enriched to 1 mg/liter each of cadmium, copper, nickel, lead, and TABLE V Metal Concentrations in Secondary Effluent and Renovated Water at Flushing Meadows Project" and Maximum Limits in Raw Municipal Water Suppliesb

Element Zn cu Cd

Pb Hg a

Secondary effluent Renovated water Maximum limits (mg/iiter) (mg/liter) (mg/liter) 0.193 0.143 0.008 0.084 0,002

0.037 0.017 0.007 0.066 0.001

5 1 0.01 0.05 Not given

From Bouwer d al. (1974b). From the National Technical Advisory Committee (1968).

162

HERMAN BOUWER AND R. L. CHANEY

zinc. This retention is studied in relation to depth, variations in soil pH, and clay and organic matter contents. In general, most studies of metal leaching have found that the pH, organic matter, and ion exchange capacity are the dominant factors in metal movement through soils. Jones and Belling ( 1967) and Jones et al. (1957) examined the effect of water, C0,-saturated water, superphosphate, and an aqueous extract of lucerne on leaching zinc, copper, and cobalt in several soils. They found appreciable movement only in a sand with low organic matter. Smith et al. (1962) also observed movement of copper and zinc only in soils of low cation exchange capacity and low organic matter; low pH promoted leaching. Peterson and Geschwind (1972) found leaching of heavy metals from acidic strip mine spoils amended with sewage sludge only at low pH. Ng and Bloomfield (1962) suggest that under reducing conditions (unstable organic amendments and waterlogging) heavy metals can be mobilized and subsequently leached. In contrast, when a method to obtain zinc leaching through soils was desired in order to fertilize apple trees, Benson (1966) found that this could be achieved by applying an excess of potassium, calcium, or magnesium salts with the zinc. A comparison of amounts of nitrogen and heavy metals added to a given soil with sludge vs effluent shows that, at application rates that avoid nitrate pollution, more heavy metals can be added with sludge in one year than are added in a century of effluent irrigation. It is generally considered that the amount of plant-available nitrogen limits the yearly application of sludge. However, some sludges contain large amounts of metals (Chaney, 1973; Page, 1973), and metal toxicity to sensitive crops could result from sludge applications that do not contain more nitrogen than can be removed by a crop. The level of metals in municipal wastewater should soon be sufficiently low that heavy metals will not be a limiting factor in long-term wastewater irrigation practices. New effluent and pretreatment guidelines (Environmental Protection Agency, 1973) should lead to control of metal release from industrial sources. Although other pretreatment technologies to remove metals from industrial wastewaters are available (precipitation, ion exchange resins, and reverse osmosis), Wentink and Etzel (1972) found that soil filtration was an adequate pretreatment to remove metals. I. DISSOLVED SALTS

Sewage effluent usually contains about 100-300 mg more salt (predominantly NaCl) per liter than the water going into the municipal water supply system (California Department of Water Resources, 1961). Cattle manures may contain high NaCl levels, whereas effluent from vegetable processing

LAND TREATMENT OF WASTEWATER

163

plants may contain considerable amounts of NaOH if the lye-peeling process is used. Dissolved salts in the wastewater react with the soil through ion exchange and dispersion or flocculation of clay. While ion exchange may initially affect the quality of the percolate or renovated water, the ionic composition of the renovated water will eventually be the same as that of the wastewater as equilibrium between the cation exchange complex and the soil solution is reached. Much higher salt concentrations in the percolate than in the original wastewater can be expected if the amount of wastewater applied (plus rainfall) is not much larger than the evapotranspiration of the crop. For normally irrigated fields in arid regions, the salt content has increased 3- to 10-fold as the irrigation water moved through the root zone (Bouwer, 1969). The salt balance equation (Bouwer, 1969) shows that the ratio of the salt concentration in the percolate, Cd, to that in the wastewater, Ci,can be calculated as

C d C i = Di/(Dt - DJ where Di is the amount of water applied and D, is the evapotranspiration. Where rainfall or precipitation of salts in the soil are significant, appropriate corrections should be made in this equation to take these effects into account. The equation shows that even when three times as much water is applied as is needed by the crop, the salt concentration in the percolate will be 1.5 times that in the wastewater. Thus, overirrigation with wastewater in arid regions tends to produce renovated water with too high salt concentration. Only when D iis much greater than D, (high-rate systems) can renovated water with essentially the same salt content as the wastewater be obtained. At the Flushing Meadows Project, for example, D , = 1.8 m/year and Di is about 100 m (Bouwer et al., 1974b). This gives renovated water with only about 2 % more salt than in the sewage effluent.

J. PH The pH of soil and the pH of wastewater can be modified by the chemical reactions in the soil. Wastewater containing organic acids may show an increase in pH as it moves through the soil because of biodegradation of the acids. For biodegradable material with a near neutral pH, such as sewage effluent, the pH may decrease because soil microbial activity produces CO, and organic acids. At the Flushing Meadows project, for example, the p H of sewage effluent was about 8, whereas that of the renovated water was about 7 (Bouwer et al., 1974b). Such a decrease in pH could result in increased solubility of CaCO, which may occur in the soil as a precipitate if the original p H was in the 7.5 to 8.2 range.

164

HERMAN BOUWER AND R. L. CHANEY

Ill.

A.

Crop Response

EFFECTSON YIELDAND QUALITY

Experimental examination of the effect of wastewater irrigation on crop yield is made difficult by the need to compare unirrigated unfertilized control, unirrigated fertilized control, water irrigated control, and water plus fertilizers equal to wastewater treatment control with the wastewater treatment. Where rate of wastewater application is varied, comparable treatments for each rate would also be needed. Without this group of treatments, it is difficult to evaluate the role of water, plant macronutrients and micronutrients, etc., on crop yield and quality. Major studies to date, while finding a few negative and many positive yield responses of crops to wastewater, have not included this range of treatments. For arid regions, the unirrigated controls are not necessary. Research at Pennsylvania State University has compared unirrigated, unfertilized control (for forest crops) or unirrigated, fertilized control (for field crops) with one or several rates of wastewater application. Sopper and Kardos (1973) report 9 years’ experience: “Effluent irrigation at 2 inches per week resulted in annual yield increases ranging from -8 to +346% for corn grain, 5 to 130% for corn silage, 85 to 191% for red clover, and 79 to 139% for alfalfa.” Precipitation during the growing season greatly influenced yield differences, but the experimental design prevented study of this factor. Since the major theme of the studies was the ability of crops and soils to remove nutrients from percolating wastewater, wastewater application rates exceeded the water requirements of the crops. In the University of Arizona studies, wastewater is considered an alternative water source for otherwise necessary irrigation; thus, one wastewater application rate was used: the amount of water required to meet recommended irrigation practices. Day and co-workers (1962, 1963; Day and Tucker, 1959, 1960; Day and Kirkpatrick, 1973; Day, 1973) in Arizona studied growth of several crops (barley, oats, and wheat as forage and grain) irrigated with water; water plus recommended N, P, and K; water plus N, P, and K equal to contents of these elements in secondary sewage effluent; and secondary effluent. The application rates were governed by the irrigation requirements of the crops. Wastewater generally produced equal or somewhat higher yields of grain or forage than well water with N, P, and K added equal to the N, P, and K of the wastewater, although barley was more easily injured by wastewater than were oats and wheat. Effects of wastewater irrigation on crop quality were included in both

LAND TREATMENT OF WASTEWATER

165

the Arizona and Pennsylvania studies. Day and co-workers ( 1962, 1963; Day and Tucker, 1959, 1960; Day and Kirkpatrick, 1973; Day, 1973) found that small-grain forage contained similar amounts of protein and digestible laboratory nutrients when irrigated with wastewater or water plus N, P, and K equal to wastewater. The danger of nitrate poisoning from forage grown with wastewater was no greater than the danger from forage grown with well water plus N, P, and K (Day et al., 1961). Although the yields and quality of wheat grain for livestock feed were not impaired by wastewater irrigation, the milling and baking qualities of wheat grain produced with wastewater were lower than those of grains grown with well water plus N, P, and K (Day, 1965). No explanation is available for this observation. The quality of other crops as food or fiber could be modified by wastewater irrigation. Slightly higher N, P, and K in wastewater irrigated forage crops than in chemically fertilized, unirrigated crops were found in the Pennsylvania State University studies. No reduction in feed quality of these forages was observed. Although harvest of corn silage and Reed canarygrass forages can remove as much N, P, and K as is added with some wastewaters, cycling the forages through animals will generate manures that will contain much of the N, P, and K removed from the wastewater irrigation site. The properties of red pine and red oak necessary for pulpwood were enhanced by wastewater irrigation (Murphey et al., 1973). Sopper and Kardos (1973) report considerable variation in response of different tree species to wasterwater irrigation; white spruce was quite responsive to wastewater, whereas red pine showed yield response for 2.5 cm per week application but none to 5 cm per week. In one of their experiments, unusual weather conditions led to windthrow of all trees on the irrigated treatment apparently because irrigation decreased the rooting depth of red pine. While nitrogen in sewage effluent or other wastewater has fertilizer value, nitrogen in irrigation water is not always desirable because it may unfavorably affect yield and quality of some crops. Baier and Fryer (1973) report reduced yield, fruit size, and/or fruit quality of certain fruit crops. Too much nitrogen at the end of the growing season may delay the maturity of the crop (cotton, for example). It can also reduce the sugar content of sugar beets and the starch content of potatoes. Lodging may be a problem in grain crops.

B. UPTAKEOF POLLUTANTS AND LOCATION IN PLANT It is important to assess the absorption by plants of pollutants in wastewater even if they do not appear to affect the plants themselves, because

166

HERMAN BOUWER AND R. L. CHANEY

of the potential for injury to organisms higher in the food chain. The pollutants of major interest include toxic elements, chlorinated hydrocarbons, and other pesticides. Many ill-defined organic chemicals enter the wastewater of municipalities because of the complexity of our industrial and domestic effluents. Little information has been obtained on the reactions of the organic chemicals in wastewater with soils, or their uptake by plants. “Only a small number of pesticides have been investigated for uptake by plants,” is the way Nash (1974) summarized the state of knowledge in this area. In most cases, the amounts of persistent chlorinated hydrocarbons added with wastewater will be considerably lower than those added during normal agricultural operations. Some specific industrial compounds such as polychlorinated biphenyls (PCB’s) have no agricultural use. Several authors have recently reviewed plant accumulation of pesticides (Caro, 1969; Foy et al., 1971; Nash; 1974). Considerable variation in uptake was observed among plant species. External surfaces of root crops were heavily laden with organochlorine compounds when grown in soils containing such compounds, but leaves and root interiors had only very low amounts. Nash and Harris ( 1973) found that soybeans transported several organochlorine pesticides to the grain; other crops differed sharply in uptake and translocation (corn, oats, wheat). Little is known about the uptake of PCB’s by plants. G. B. Jones and T. D. Hinesly (personal communication, 1974) observed no increase in the PCB content of corn grain due to application of sewage sludge containing these compounds. The impact of land application of wastewater on soil and plant levels of pesticides has remained essentially unassessed. The toxic trace elements of major interest include arsenic, cadmium, copper, mercury, molybdenum, lead, selenium, and zinc. Plant absorption of these metals has been reviewed by Allaway (1968), Chaney (1973), Leeper (1972), Lisk (1972), Page (1973), and Tiffin et al. (1973). Again, plant species differ markedly in accumulation of these elements. The beet family accumulates large amounts of many of these elements in its leaves. Most plants exclude toxic trace elements from their seeds and fruits; most root crops exclude these elements from the edible root, although carrot roots accumulate considerable amounts of cadmium, lead, etc., from rnetal-enriched soils. The data shown in Table III suggest that the accumulation of toxic trace elements need not be a food chain hazard where “normal” domestic effluents are applied to land. Although foliar uptake or fixation of toxic elements from spray irrigated effluents could occur, this topic is similarly unresearched. In most cases the hazard will build up over decades in contrast to sludge application where a single application of some sludges can lead to potential food chain hazards.

LAND TREATMENT OF WASTEWATER

IV.

167

Selection and Design of System

Minimum impact on the environment and minimum total cost of operation are the two main design criteria for land treatment of liquid waste. The choice of system is largely controlled by soil and hydrogeologic conditions, and by the availability of land. If the available soils in a certain region have very low infiltration rates, about the only choice is an overland flow system (this term is preferred over spray-runoff systems because the wastewater does not necessarily have to be sprayed on the field but can also be applied with ditches, hydrants, or other techniques normally used in surface irrigation). Low-rate systems are the only possibility for soils that are not permeable enough for high-rate systems but too permeable for overland flow systems. Sandy loams and coarser-textured soils generally have sufficient permeability to permit high-rate systems, so that a choice between low-rate and high-rate systems must be made. Coarse sands, gravels, or shallow soils underlain by coarse material or by fractured or cavernous rock are usually not suitable for land treatment because of the hazards of groundwater contamination. While a deep water table ( 1 m or more) is desirable for land treatment fields (other than overland flow systems) , temporarily higher water tables can be tolerated, especially during infiltration, provided that the soil drains rapidly after wastewater infiltration stops. Even with very deep water tables, aerobic conditions may be restricted to the top meter or so of the soil profile (Pincince and McKee, 1968; Lance et al., 1973). Thus, water table depths of several meters as have sometimes been recommended (Robeck et al., 1964) may not be required under those conditions. The greater the depth to which the water table drops after the start of a resting period, the greater will be the amount of oxygen which enters the soil as air replaces the water draining from the soil. Entry of greater amounts of oxygen into the soil during drying would permit higher loadings of oxygen-demanding materials during wastewater application. The same would be true for porous soils, which have a higher oxygen diffusion rate than finer soils. Thus, the optimum schedule of infiltration and resting periods of a treatment field can be influenced by the drainage and diffusion parameters of the soil. In soils with restricted internal drainage, subsurface drains may be necessary to assure sufficiently rapid drainage of the soil profile after the start of a resting period. Overland flow systems should be designed to yield uniform, shallow depths of the flow above the soil and sufficient detention times. The slope of the fields should be high enough to give small depths of water on the soil but low enough to prevent channeling. The slopes in the system for

168

HERMAN BOUWER AND R. L. CHANEY

cannery wastes at Paris, Texas, were 2-6% and the overland flow distances were 60-90 m (Law et al., 1970; Gilde, 1973). The average application was about 1.3 cm/day, which was applied in 6-8 hours each day. Thomas (1973b) used slopes of 2-4%, overland flow distances of 36 m, and applications of 1-1.4 cm/day applied in about 8 hours, in his study with raw sewage. The loading rates and hence the land requirements for overland flow systems are approximately in the same category as those for low-rate infiltration systems, Overland flow systems require smooth topographies and uniform application of the wastewater at the upper end of the reach to prevent channeling. The vegetation may consist of Reed canarygrass in temperate climates and of bermudagrass in warm climates. Other grasses have also been used. Certain rushes (Scirpus Zacustris) have been reported to be very effective in removing pollutants from wastewater (Seidel, 1966). Low-rate systems are best suited for humid climates with relatively low evapotranspiration rates. Here, the salt content of the percolate will not be much higher than that of the original wastewater. Low-rate systems often permit normal agricultural use of the receiving fields. Compared to other land treatment systems, low rate systems normally will yield the best quality renovated water. If used on a large scale, however, native groundwater supplies may be affected over a large area. The spread of contamination may be difficult to control, raising concern over the long-term effects of diffuse sources (Walker, 1973). Wastewater may be applied with sprinklers, basins, or furrows, depending on the topography. Bendixen et al. (1968) reported no significant differences between these application techniques with respect to the performance of the system. Sands and other permeable soil may drain so rapidly that, when the wastewater is applied with sprinklers, sufficient air may enter the soil between sprinkler-head rotations to maintain aerobic conditions in the upper .,ortion of the soil. This will restrict denitrification and complete nitrificaion of the nitrogen in the wastewater can be expected, even after prolonbed application (Smith, 1971 ) . If wastewater is used for irrigation or similar low-rate systems in arid regions with high evapotranspiration rates, the salt content of the percolate will be much higher than that of the wastewater and the percolate will be unsuitable for groundwater recharge (Bouwer, 1974). Thus, artificial drainage may be required to remove the salty deep percolation water from the soil, as commonly practiced in irrigated agriculture. The large land requirements of low-rate systems may pose a problem when large volumes of effluent are to be applied. At an application rate of 2.5 cm/week for example, a city of 100,000 people would require some 1200 ha to dispose of its effluent. Muskegon County in Michigan has acquired 4000 ha of land northeast of the city of Muskegon to handle

LAND TREATMENT OF WASTEWATER

169

a domestic and industrial wastewater flow of 164,000 m3/day (Chaiken et al., 1973). Where permeable soils (loams, sandy loams, and fine sands) are available, high-rate systems are possible. The land requirements of high-rate systems may only be a few percent of those for low-rate systems. However, normal agricultural utilization of the land is often not feasible, requiring that the land be dedicated to wastewater renovation. High-rate systems are the only land treatment systems that can be used in warm, dry climates to yield renovated water that has about the same salt content as the original wastewater (Bouwer, 1974). This is important if the renovated water is to be reused. The spread of renovated water into the groundwater basin below highrate systems can be controlled by collecting the renovated water with drains if the aquifer is shallow, or with wells if the aquifer is deep (Bouwer, 1970, 1973c, 1974). The drains or wells must be located far enough from the infiltration system to allow sufficient time and distance of underground travel for the wastewater. The system can be so designed and operated that theoretically all the wastewater infiltrating into the soil can be collected by the wells or drains, without any renovated water moving into the groundwater outside the system of infiltration fields and collection facilities. This means that the portion of the aquifer between the infiltration and collection facilities is dedicated to renovation of wastewater. After collection, the renovated water can be used for unrestricted irrigation, recreation, industrial purposes, or it can be discharged into surface water. Domestic use of this water is not recommended until it is proven safe (Long and Bell, 1972; American Water Works Association Board of Directors, 1973). High-rate systems lend themselves for pre- or posttreatment of water in connection with advanced in-plant treatment of wastewater. This is done in various countries where low-quality surface water is used for municipal water supplies. Since the performance of a land-treatment system depends so much on the local soil, hydrogeology, and climate, as well as on the waste characteristics themselves, local experimentation and pilot projects are usually needed if land treatment is considered and local experience with such systems is not available. After installation of the full-scale project, good management, and monitoring of the system so that undesirable performance can be corrected before too much damage is done, are essential. REFERENCES Adams, A. P., and Spendlove, J. C. 1970. Science 169, 1218-1220. Alexander, M. 1971. “Microbial Ecology.” Wiley, New York. Allaway, W. H. 1968. Advan. Agron. 20, 235-274.

170

HERMAN BOUWER AND R . L. CHANEY

Allen, M. J., and Morrison, S. M. 1973. Ground Water 11, 6-10. American Water Works Association Board of Directors. 1973. J . Amer. Wafer Works Ass. 65, Part 11, 64. Andre, D. A., Weiser, H. H., and Maloney, G. W. 1967. J. Amer. Water Works Ass. 59, 503-508. Argo, D. G., and Culp, G. L. 1972. Water Sewage Works 119 (8), 62-65; (9), 128-132. Arizona State Health Department. 1972. “Proposed Rules and Regulations for Reclaimed Wastes.” Phoenix, Arizona. Baars, J. K. 1964. “Principles and Applications in Aquatic Microbiology,” Rudolfs Res. Conf. Proc., 1963, Chapter 17. Wiley, New York. Bache, B. W., and Williams, E. G. 1971. J . Soil Sci. 22, 289-301. Baier, D. C., and Fryer, W. B. 1973. J . Irrig. Drain. Div., Amcr. SOC. Civil Eng. 99(IR2), 133-141. Barth, E. F., and Ettinger, M. B. 1967. J . Water Pollut. Contr. Fed. 39, 1362-1368. Bauer, W. J., and Matsche, D. E. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 345-363. Penn. State Univ. Press, University Park, Pennsylvania. Benarde, M. A. 1973. J . Amer. Warer Works Ass. 65, 432-439. Bendixen, T. W., Hill, R. D., Schwartz, W. A., and Robeck, G. G. 1968. J . Sanit. Eng. Div., Amer. SOC. Cicil Eng. 94, 147-157. Benson, N. R. 1966. Soil Sci. 101, 171-179. Berg, G., Dean, R. B., and Dahling, D. R. 1968. J . Amer. Water Works Ass. 160, 193-198. Biggar, J. W., and Fireman, M. 1960. Soil Sci. SOC.Amer., Proc. 24, 115-120. Blakeslee, P. A. 1973. I n “Recycling Municipal Sludges and Effluents on Land,” pp. 183-198. Nat. Ass. State Univ. and Land-Grant Colleges, Washington, D.C. Blood, J. W. 1963. N.A.A.S. Quart. Rev. 14, 97-100. Bondietti, E. A., and Sweeton, F. H. 1973. Agron. Abstr. p. 89. Bouwer, H. 1968. J . Soil Water Conserv. 23, 164-168. Bouwer, H. 1969. J. Irrig. Drain Div., Amer. SOC. Civil Eng. 95(IR1), 153-170. Bouwer, H. 1970. J . Sanit. Eng. Div., Amer. SOC. Civil Eng. 96(SAl), 59-74. Bouwer, H. 1973a. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 164-175. Penn. State Univ. Press, University Park, Pennsylvania. Bouwer, H. 1973b. “Nitrification-denitrscation in the Soil.” Water Res. Center, University of Massachusetts, Boston. Bouwer, H. 1973c. In “Underground Waste Management and Artificial Recharge” (J. Braunstein, ed.), Vol. I, pp. 23-33. Amer. Assoc. Petrol. Geol., Tulsa, Oklahoma. Bouwer, H. 1974. Ground Water 12, 140-147. Bouwer, H., Rice, R. C., and Escarcega, E. D. 1974a. J . Water Pollitf. Contr. Fed. 46, 834-843. Bouwer, H., Lance, J. C., and Riggs, M. S . 1974b. J . Wuter Pollut. Contr. Fed. 46, 844-859. Bower, C. A., and Hatcher, J. T. 1967. Soil Sci. 103, 151-154. Brewer, R. F. 1966. In “Diagnostic Criteria for Plants and Soils” (H. D. Chapman, ed.), pp. 180-196. Div. Agr. Sci., University of California, Berkeley. Broadbent, F. E., and Clark, F. E. 1965. I n “Soil Nitrogen” (W. V. Bartholomew and F. E. Clark, eds.), Agron. Monogr. No. 10, pp. 344-359. Amer. SOC.Agr., Madison, Wisconsin.

LAND TREATMENT OF WASTEWATER

171

Brown, H. G., Hensley, C. P., McKinney, G. L., and Robinson, J. L. 1973. Environ. Lett. 5, 103-114. Brunner, D. R., and Sproul, 0. J. 1970. J . Sanit. Eng. Div., Amer. SOC. Civil Eng. 96(SA2), 365-379. Bunch, R., Barth, E., and Ettinger, M. 1961. J. Wafer Pollut. Contr. Fed. 33, 122-126. California Department of Water Resources. 196 1. “Feasibility of Reclamation of Water from Wastes in the Los Angeles Metropolitan Area,” Bull. No. 80. Calif. Dept. Water Resour., Sacramento. California Institute of Technology. 1969. “Investigation of Biochemical Stabilization of Aqueous Solutions of Organic Compounds by Unsaturated Flow through Porous Media.” J. E. McKee, Principal investigator. W. M. Keck Laboratory, CIT, Pasadena. California State Water Pollution Control Board. 1953. “Field Investigation of Water Reclamation in Relation to Groundwater Pollution,” Publ. No. 6. California State Water Resources Control Board. 1968. “Cannery Waste Treatment, Utilization and Disposal,” Publ. No. 39. Caro, J. H. 1969. Phyropathology 59, 1191-1197. Chaiken, E. I., Polonesik, S., and Wilson, C. D. 1973. Civil Eng. 43, 49-53. Chaney, R. L. 1973. In “Recycling Municipal Sludges and Effluents on Land,” pp. 129-141. Nat. Ass. State Univ. and Land-Grant Colleges, Washington, D.C. Cohen, J., and Shuval, H. I. 1973. Water, Air, Soil Pollut. 2, 85-95. Colston, N. V., and Smallwood, C., Jr. 1973. J. Water Pollut. Contr. Fed. 45, 1972-1977. Cookson, J. T., Jr. 1967. Sci. Technol. 1, 157-160. Cosgrove, D. T. 1973. J. Environ. Quai. 2,483-485. Craun, G. F., and McCabe, L. I. 1973. J . Amer. Water Works Ass. 65, 74-84. Culp, G. L., Culp, R. L., and Homann, C. L. 1973. J. Amer. Water Works Ass. 65, 641-646. Day, A. D. 1965. 1964 Wheat Newsleft. 11, 39-40. Day, A. D. 1973. Proc. Nut. Workshop Land Appl. Municipal Sludge Efluent, pp. 155-160. Day, A. D., and Kirkpatrick, R. M. 1973. J . Environ. Qual. 2, 282284. Day, A. D., and Tucker, T. C. 1959. Agron. J. 51, 569-572. Day, A. D., and Tucker, T. C. 1960. Agron. J . 52, 238-239. Day, A. D., Tucker, T. C., and Vavich, M. G. 1961. 1960 Oat Newsleft. 11, 29-3 1. Day, A. D., Tucker, T. C., and Vavich, M. G. 1962. Agron. J . 54, 133-135. Day, A. D., Dickson, A. D., and Tucker, T. C. 1963. Agron. J . 55, 317-318. Day, A. D., Stroehlein, J. L., and Tucker, T. C. 1972. J . Water Pollut. Contr. Fed. 44, 372-375. Deaner, D. G., and Kerri, K. D. 1969. J . Amer. Wafer Works Ass. 61, 465-468. Drewry, W. A., and Eliassen, R. 1968. J . Water Pollut. Contr. Fed. 40, R257-R271. Dunlop, S . G. 1968. In “Municipal Sewage Effluent for Irrigation” ( C . W. Wilson and F. E. Beckett, eds.), pp. 107-121. Louisiana Polytech. Inst., Ruston. Eaton, F. M. 1944. J. Agr. Res. 69, 237-277. Egeland, D. R. 1973. J. Water Polluf. Contr. Fed. 45, 1465-1475. Ellis, B. G. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through

172

HERMAN BOUWER AND R. L. CHANEY

Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 46-70. Penn. State Univ. Press, University Park, Pennsylvania. Ellis, B. G., and Erickson, A. E. 1969. “Movement of and Transformation of Various Phosphorus Compounds in Soil” (mimeo.). Michigan State University, East Lansing. Ellis, B. G., and Knezek, B. D. 1972. In “Micronutrients in Agriculture” (J. J. Mortvedt et al., eds.), pp. 59-78. Soil Sci. SOC.Amer., Madison, Wisconsin. Environmental Protection Agency. 1973. “Pretreatment of Discharges to Publicly Owned Treatment Works (40 CFR 128).” U.S. Environ. Protect. Agency, Off. Water Programs Operations, Washington, D.C. Erickson, A. E., Tiedje, J. M., Ellis, B. G., and Hansen, C. M. 1972. Proc. Cornell Agr. Waste Manage. Conf., pp. 405-410. Fair, J. F., and Morrison, S . M. 1967. Water Resour. Res. 3, 799-803. Folkman, Y.,and Wachs, A. M. 1970. Proc. Amer. SOC.Civil Eng. 96, 675-690. Follett, R. H., and Lindsay, W. L. 1971. Soil Sci. SOC.Amer., Proc. 35, 600-602. Foster, D. H., and Engelbrecht, R. S. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 247-270. Penn. State Univ. Press, University Park, Pennsylvania. Foy, C. L., Coats, G. E., and Jones, D. W. 1971. In “Respiration and Circulation” (P. L. Altman and D. S. Dittmer, eds.), Biol. Handb., pp. 743-791. Fed. Amer. SOC.Exp. Biol., Bethesda, Maryland. Gilde, L. C. 1973. J . lrrig. Drain. Div., Amer. SOC.Civil Eng. 99(IR3), 339-352. Gilliam, J. W., and Sample, E. C. 1968. Soil Sci. 106, 352-357. Goss, D. W., and Jones, 0. R. 1973. In “Underground Waste Management and Artificial Recharge” (J. Braunstein, ed.), Vol. I, pp. 468-480. Amer. Ass. Petrol. Geol., Tulsa, Oklahoma. Griffin, R. A., and Jurinak, J. J. 1973. Soil Sci. SOC. Amer., Proc. 37, 847-850. Hallam, M. J., and Bartholomew, W. V. 1953. Soil Sci. SOC. Amer., Proc. 17, 365-368. Hatcher, J. T.,and Bower, C. A. 1958. Soil Sci. 85, 319-323. Hinesly, T. D. 1973. Water Spectrum 5, 1-8. Hodgson, J. F. 1963. Advan. Agron. 15, 119-159. Hook, J. E., Kardos, L. T., and Sopper, W. E. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 200-219. Penn. State Univ. Press, University Park, Pennsylvania. Hopkins, G. J., Neel, J. K., and Nelson, F. L. 1956. Sewage Inn. Wastes 28, 1466-1474. Jenne, E. A. 1968. Advan. Chem. Ser. 73, 337-387. Jones, G.B., and Belling, G. B. 1967. Aust. J . Agr. Res. 18, 733-740. Jones, G. B., Riceman, D. S., and McKenzie, J. 0. 1957. Aust. J . Agr. Res. 8, 190-20 1. Kardos, L. T. 1967. In “Agriculture and the Quality of our Environment,” Publ. NO. 85, pp. 241-250. Amer. Ass. Advan. Sci., Washington, D.C. Kardos, L. T., and Sopper, W. E. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 148-163. Penn. State Univ. Press, University Park, Pennsylvania. Khin, A., and Leeper, G. W. 1960. Agrochimica 4, 246-254. Kirby, C. F. 1971. “Sewage Treatment Farms” (Post graduate course in Public

LAND TREATMENT OF WASTEWATER

173

Health Engineering), Sess. No. 12. Dept. Civil Eng., University of Melbourne, Australia. Knezek, B. D. 1972. Mich., Univ., Inst. Water Res. [Tech. Rep.] 30, 27-43. Krishnaswami, S. K. 1971. Can. J. Pub. Health 62, 36-41. Krone, R. B. 1968. In “Municipal Sewage Effluent for Irrigation” (C. W. Wilson and F. E. Beckett, eds.), pp. 75-106. Louisiana Polytech. Inst., Ruston. Lance, J. C. 1972. J. Water Pollut. Contr. Fed. 44, 1352-1361. Lance, J. C., and Whisler, F. D. 1972. J . Environ. Qual. 1, 180-186. Lance, J. C., and Whisler, F. D. 1973. Proc. Int. Conf. Land Waste Manage., 1973 (in press), Lance, J. C., Whisler, F. D., and Bouwer, H. 1973. J. Environ. Qual. 2, 345-350. Larsen, S., and Widdowson, A. E. 1971. J. Soil Sci. 22, 210-221. Larson, W. E. 1960. I n “Algae and Metropolitan Wastes, Transactions 1960 Seminar,” pp. 125-129. U.S. Dept. of Health, Education and Welfare, Washington, D.C. Law, J. P., Jr. 1968. “Agricultural Utilization of Sewage Effluent and Sludge,” An annotated bibliography. Fed. Water Pollut. Contr. Admin., U.S. Dept. of Interior. Law, J. P., Jr., Thomas, R. E., and Myers, L. H. 1970. J. Water Pollrit. Contr. Fed. 42, 1621-1631. Leeper, G. W. 1972. “Reactions of Heavy Metals with Soils with Special Regard to their Application in Sewage Wastes,” Contract No. DACW73-73-C-0026. Dept. of Army, Corps Engineers, Washington, D.C. Lehman, G. S., and Wilson, L. G. 1971. Water Resour. Res. 7 , 90-99, Levin, G . V., Topol, G. .I. Tarnay, , A. G., and Samworth, R. B. 1972. 1. Water Pollut. Contr. Fed. 44, 1940-1954. Lindsay, W. L. 1973. I n “Recycling Municipal Sludges and Effluents on Land,” pp. 91-96. Nat. Ass. State Univ. and Land-Grant Colleges, Washington, D.C. Lindsay, W. L., and Moreno, E. C. 1960. Soil Sci. SOC. Amer., Proc. 24, 177-182. Lisk, D. J. 1972. Advan. Agron. 24, 267-325. Loehr, R. C. 1968. “Pollution Implications of Animal Wastes-A Forward Oriented Review.” Robert S. Kerr Water Res. Cent., Environmental Protection Agency, Ada, Oklahoma. Long, W. H., and Bell, F. A., Jr. 1972. J . Amer. Wafer Works Ass. 64, 220-225. McGauhey, P. H., and Krone, R. B. 1967. “Soil Mantle as a Wastewater Treatment System,” SERL Rep. No. 67-1 1. University of California, Berkeley. Mack, W. H. 1973. J. Amer. Water Works Ass. 65, 347-348. McMichael, F. C., and McKee, J. E. 1965. “Research on Waste Water Reclamation at Whittier Narrows.” W. M. Keck Lab. Environ. Health Eng., California Institute of Technology, Pasadena. Mallman, W. L., and Mack, W. H. 1961. “Groundwater Contamination,” Tech. Rep. W61-5, pp. 35-43. U S . Pub. Health Serv., Washington, D.C. Masinova, L., and Cledova, V. 1957. “Evaluation of the Positive Test for Coliform Bacteria in Sanitary Tests of Water and Soil,” Rep. No. 13-14. Inst. Hyg., Prague. Matlock, W. G. et a/. 1972. “Sewage Effluent Pollution of a Groundwater Aquifier.” Amer. SOC.Agr. Eng., Chicago, Illinois. Merrell, J. C., Jr., and Ward, P. C. 1968. J . Ainer. Water Works Ass. 60, 145-153. Miller, R. H. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 71-94. Penn. State Univ. Press, University Park, Pennsylvania.

174

HERMAN BOUWER AND R. L. CHANEY

Mitchell, R. L. 1964. In “Chemistry of the Soil” (F. E. Bears, ed.), 2nd ed., pp. 320-368. Van Nostrand-Reinhold, New York. Moore, D. R. E., and Waid, J. S. 1971. Soil Biol. & Biochem. 3, 69-83. Murphey, W. K., Brisbin, R. L., Young, W. J., and Cutter, B. E. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” ( W. E. Sopper and L. T. Kardos, eds.), pp. 295-310. Penn. State Univ. Press, University Park, Pennsylvania. Murphy, W. H., and Syverton, J. T. 1958. Virology 6, 623-636. Murphy, W. H., Eylar, 0. R., Schmidt, E. L., and Syverton, J. T. 1958. Virology 6, 612-622. Murrmann, R. P., and Koutz, F. R. 1972. In “Wastewater Management by Disposal on the Land” (S.C. Reed, ed.), CRREL Spec. Rep. No. 171, pp. 48-76. Mytelka, A. I., Czachor, J. S., Guggino, W. R., and Golub, H. 1973. J . Water Pollut. Contr. Fed. 45, 1859-1864. Nash, R. G. 1974. In “Pesticides and their Effects on Soils and Water,” SpecPubl. No. 8 (revised), Chapter 11. Soil Sci. SOC. Amer., Madison, Wisconsin (in press). Nash, R. G., and Harris, W. G. 1973. J. Environ. Qual. 2, 269-273. National Academy of Science-National Academy of Engineering. 1974. “Water Quality Criteria 1972” (in press). National Technical Advisory Committee to the Secretary of Interior. 1968. “Water Quality Criteria.” Environmental Protection Agency, Washington, D.C. Nevo, Z., and Mitchell, K. 1967. Water Res. 1, 231-236. Ng, S. K., and Bloomfield, C. 1962. Plant Soil 16, 108-135. Nilsson, R. 1971. Water Res. 5, 51-60. Norvell, W. A. 1972. In “Micronutrients in Agriculture” (J. J. Mortvedt et al., eds.), pp. 115-138. Soil Sci. SOC.Amer., Madison, Wisconsin. Okazaki, E., and Chao, T. T. 1968. Soil Sci. 105, 255-259. Ongerth, H. J., Spath, D. P., Crook, J., and Greenberg, A. E. 1973. J. Amer. Water Works Ass. 65, 495-498. Page, A. L. 1973. “Fate and Effects of Trace Elements in Sewage Sludge when Applied to Agricultural Lands. Literature Review Study for Ultimate Disposal Research Program.” Environmental Protection Agency, Cincinnati, Ohio. Peterson, J. R., and Geschwind, J. 1972. J . Environ. Qual. 1, 410-412. Peterson, J. R., Lue-Hing, C., and Zenz, D. R. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds.), pp. 26-37. Penn. State Univ. Press, University Park, Pennsylvania. Pincince, A. B., and McKee, J. E. 1968. J . Sanif. Eng. Div., Amer. SOC. Civil Eng. 94, 1093-1 119. Porges, R., and Hopkins, G. J. 1955. Sewage Znd. Wastes 27, 1160-1170. Pound, C. E., and Crites, R. W. 1973a. In “Recycling Municipal Sludges and EWuents on Land,” pp. 49-61. Nat. Ass. State Univ. and Land-Grant Colleges, Washington, D.C. Pound, C. E., and Crites, R. W. 1973b. “Wastewater Treatment and Reuse by Land Application,” Vols. I and 11, EPA-660/2-73-006a and b. U.S. Govt. Printing Office, Washington, D.C. Prince, A. L., Bear, F. E., Brennan, E. G., Leone, I. A., and Daines, R. H. 1949. Soil Sci. 67, 269-277. Rand, W. E., and Schmidt, H. J. 1952. Amer. J . Vet. Res. 13, 50-61.

LAND TREATMENT OF WASTEWATER

175

Randall, A. D. 1970. J. Amer. Water Works Ass. 62, 716-720. Rhoades, E. D. 1964. Trans. ASAE (Amer. SOC.Agr. Eng.) 7, 165-166. Rhoades, J. D., Ingvalson, R. D., and Hatcher, J. T. 1970. Soil Sci. Sco. Amer., Proc. 34, 871-875. Rice, R. C. 1974. J . Water Pollut. Contr. Fed. 46, 708-716. Richards, L. A., ed. 1954. “Diagnosis and Improvement of Saline and Alkali Soils,” Handb. No. 60. U.S.Dept. of Agriculture, Washington, D.C. Robeck, G. G., Bendixen, T. W., Schwartz, W. A., and Woodward, R. L. 1964. J . Water Pollut. Contr. Fed. 36, 971-983. Rohde, G. 1962. J . Inst. Sewage Purification 1962 pp. 581-585. Romero, J. C. 1970. Ground Water 8, 37-49. Rose, W. W., Merier, W. A., Katsuyuma, A., Sternberg, R. W., Brauner, G. V., Olsen, N. A., and Weckel, K. G. 1971. Proc. Nut. Symp. Food Process. Wastes, Pac. Northwest Water Lab., EPA Nat. Canners’ Ass., 2nd. pp. 109-127. Rudolfs, W., Falk, L. L., and Ragotzkie, R. A. 1950. Sewage Ind. Waste 22, 1261-1 28 1. Schmidt, K. D. 1972. Ground Water 10, 50-61. Schneider, I. F., and Erickson, A. E. 1972. Mich., Agr. Exp. Sta., Res. Rep. 195. Seidel, K. 1966. Naturwissenschaften 53, 289-297. Shuval, H. I., and Gruener, N. 1973. Environ. Sci. Technol. 7 , 600-604. Smith, P. F., Rasmussen, G. K., and Hrnciar, G. 1962. Soil Sci. 94, 235-238. Smith, T. P. 1971. “Actual Spray Field Operations,” Proc. Land Spreading Conf., Pap. No. 8. East Central Florida Regional Planning Council, Orlando, Florida. Sopper, W. E., and Kardos, L. T. 1973. In “Recycling Treated Municipal Wastewater and Sludge Through Forest and Cropland” (W. E. Sopper and L. T. Kardos, eds), pp. 271-294. Penn. State Univ. Press, University Park, Pennsylvania. Splittstoesser, D. F., Downing, D. L. 1969. N. Y., Agr. Exp. Sta., Geneva, Res. Circ. 17, 1-15. Stefanson, R. C. 1973. Aust. J. Soil Res. 10, 183-195. Stevens, R. M. 1972. “Green Land-Clean Streams,” Report by Center for the Study of Federalism. Temple University, Philadelphia, Pennsylvania. Stone, R., and Garber, W. F. 1952. Trans. Amer. SOC.Civil Eng. 117, 1189. Sullivan, R. H., Cohn, M. M., and Baxter, S. S. 1973. “Survey of Facilities Using Land Application of Wastewater,” EPA-430/9-73-006. U.S. Govt. Printing Office, Washington, D.C. Tanji, K. K. 1970. Soil Sci. 110, 44-51. Thomas, R. E. 1973a. I . Water Pollut. Contr. Fed. 45, 1476-1484. Thomas, R. E. 1973b. Proc. Int. Conf. Land Waste Manage., 1973 (in press). Thomas, R. E. 1973c. Proc. Conf. Land Disposal Municipal EfPuents Sludges EPA-902/9-73-001, pp. 181-200. Thomas, R. E., and Bendixen, T. W. 1969. J. Water Pollut. Contr. Fed. 41, 808-813. Thomas, R. E., Schwartz, W. A., and Bendixen, T. W. 1966. Soil Sci. SOC. Amer., Proc. 30, 641-646. Tiffin, L. O., Lagerwerff, J. V., and Taylor, A. W. 1973. “Heavy Metal and Radionuclide Behavior in Soils and Crops. A Review,” AEC Res. Contract AT(49-7)-1, 182 pp. Trocme, S., Barbier, G., and Chabannes, J. 1950. Ann. Agron., Ser. A 1, 663-685. Truesdale, G. A., Birkbeck, A. E., and Shaw. D. 1964. In “Proceedings of The Institute of Sewage Purification,” Part I, pp. 3-23. Water Pollut. Res. Lab., Stevenage, U. K.

176

HERMAN BOUWER AND R. L. CHANEY

Van Donsel, D. J., Geldreich, E. E., and Clarke, N. A. 1967. Appl. Microbiol 15, 1362-1370. Vela, G. R., and Eubanks, E. R. 1973. J . Water Pollut. Contr. Fed. 45, 1789-1794. Viets, F. G., Jr. 1973. “Nitrogen Transformations in Organic Waste Applied to Land,” Int. Rep., Agr. Res. Serv., US. Department of Agriculture, Washington, D.C. Volk, B. G. 1970. Ph.D. Thesis, Michigan State University, Ann Arbor (Diss. Abstr. B 31, 63-79). Walker, W. H . 1973. Ground Water 11, 11-20. Wells, D. M., and Sweazy, R. M. 1973. Proc. Znt. Conf. Land Waste Manage., 1973 (in press). Wentink, G. R., and Etzel, J. E. 1972. J. Water Pollut. Contr. Fed. 44, 1561-1574. Wheatland, A. B., and Borne, B. J. 1961. Water Waste Treat. I . 8, 330-335. Wilson, L. A., and Lehman, G. S. 1966. “Grass Filtration of Sewage Effluent for Quality Improvement Prior to Artificial Recharge.” Amer. SOC.Agr. Eng., Chicago, Illinois. Wilson, L. G., and Lehrnan, G. S. 1967. “Progressive Agriculture in Arizona,” July-August issue, pp. 22-24. College of Agriculture, University of Arizona, Tucson. Woldendorp, J. W. 1963. “The Influence of Living Plants on Denitriftcation,” Bull. No. 63( 13). Agr. Univ., Wageningen, The Netherlands. Young, R. H. F., and Burbank, N. C., Jr. 1973. J . Amer. Water Works Ass. 65, 598-603.