treatment of wastewater with slow rate systems: a review of

Critical Reviews in Environmental Science and Technology, 36:187–259, 2006Copyright © Taylor & Francis Group, LLCISSN: 1064-3389 print / 1547-6537 onlineDOI: 10.1080/10643380500542756

Treatment of Wastewater With Slow RateSystems: A Review of Treatment Processes

and Plant Functions

NIKOLAOS V. PARANYCHIANAKIS and ANDREAS N. ANGELAKISNational Agricultural Research Foundation (NAGREF), Institute of Iraklio, Iraklio, Greece

HAROLD LEVERENZ and GEORGE TCHOBANOGLOUSDepartment of Civil and Environmental Engineering, University of California, Davis,

California, USA

Land treatment systems constitute a viable alternative solution forwastewater management in cases where the construction of con-ventional (mechanical) wastewater treatment plants (WWTPs) arenot affordable or other disposal options are not available. Theyhave proven to be an ideal technology for small rural communi-ties, clusters of homes, and small industrial units due to low energydemands and low operation and maintenance costs. In addition,slow rate systems (SRS) may be designed using the “zero discharge”concept. The purpose of this article is to review the current trendsand developments in the field of SRS, focusing on those systems inwhich effluent application is based on plant water requirements.Vegetation has an important role in treatment efficiency throughits effects on hydraulic loading rate, nutrient removal, and biomassproduction. In addition, vegetation may affect the fate of trace ele-ments and the degradation/detoxification of recalcitrant organics.Detailed knowledge of the basic processes involved in wastewatertreatment and the factors governing the performance of SRS is fun-damental for enhancing treatment efficiency and eliminating po-tential environmental and health risks. Finally, monitoring perfor-mance of SRS and adopting the appropriate management strategiesare of paramount importance to maintain treatment efficiency overthe a long term.

Address correspondence to Nikolaos V. Paranychianakis, National AgriculturalResearch Foundation (NAGREF), Institute of Iraklio, PO Box 2229, 71307 Iraklio, Greece.E-mail: [email protected]

187

188 N. V. Paranychianakis et al.

KEY WORDS: land application, nutrient removal, organic pollu-tants, pathogens, vegetation

I. INTRODUCTION

Land application of pretreated wastewater is an old technology that has gonethrough different stages of development with time. A brief historical overviewalong with an introduction to the fundamental processes involved in theoperation of these systems is provided in this section.

A. Historical Overview

Wastewater spreading to the soil, also known as “land treatment,” has along history, as demonstrated by the elaborate sewerage systems associatedwith ancient palaces and cities of Minoan Civilization since cirea 2600 BC(Angelakis and Spyridakis, 1996). Indications of wastewater being applied inthe irrigation of agricultural land extend back to the ancient Greek civiliza-tions in Crete, Athens, Dion, Cyprus, and Sparti, approximately 4000 yearsago (Angelakis et al., 2005). In more recent history, wastewater effluentswere being used in Bunzlau, Germany, for beneficial crop production begin-ning in 1531 (Gerhard, 1909). In 1650 the “Crargentinny Meadows” projectwas developed Edinburg (Scotland), where the city’s sewage was directedto adjacent fields for crop irrigation (Stanbridge, 1976). However, the expan-sion of land treatment did not occur until in the middle of the 1800s. At thistime, great epidemics of cholera that occurred in England made urgent theneed for sanitation of wastes (Gerhard, 1909). “Sewage farming” (the oldestterm used in the literature) became relatively common as a first attempt toprotect public health and to control water pollution. This technology wasmainly developed from 1850 to 1890 in England (Folsom, 1876; Stanbridge,1976), while during the same period the first slow rate systems (SRS) ap-peared in the United States, France, and Germany (Reed et al., 1995; U.S.EPA, 1979). Thereafter, the application of SRS for wastewater managementdeclined due to the development and expansion of mechanical wastewatertreatment plants. A partial return to land treatment systems in the UnitedStates was observed after passage of the Clean Water Act (CWA) of 1972,when it was realized that land treatment systems could be used to meet therequirements of the CWA (Reed et al., 1995).

During recent decades, there has been renewed interest in the devel-opment of SRS for the treatment and reuse of wastewater. This interest hasmainly been caused by (a) the inability of centralized wastewater treatmentplants (WWTPs) to serve clusters of homes, isolated rural communities, orinstitutions; (b) the high construction, operation, and maintenance costs ofconventional plants, particularly for small communities (<10,000 e.p.); (c) the

Wastewater Treatment With Slow Rate Systems 189

need for further treatment for reuse of effluents that have been treated pre-viously in conventional WWTPs; and (d) the production of biomass thatmay provide an economic return to municipalities. Moreover, the use ofSRS has been expanded to treat various types of wastewater, includinglandfill leachates (Hasselgren, 1998), dairy effluents (Sparling et al., 2001),meat processing wastewater (Guo and Sims, 2003), olive oil mill wastew-ater (Cabrera et al., 1996), agricultural drainage (Rhoades, 1989), and con-taminated groundwater (Negri et al., 2003). Recognizing the importance ofwastewater management in meeting future water demands, preventing envi-ronmental degradation, and ensuring sustainable growth, the use of SRS inwastewater management is expected to increase.

B. Fundamental Processes

Slow rate systems purify the applied wastewater through physical,chemical, and biological mechanisms that occur concurrently in thesoil–water–atmosphere environment. These mechanisms include filtration,transformation, degradation, predation, natural die-off, soil adsorption, chem-ical precipitation, denitrification, volatilization, and plant uptake (Figure 1).Detailed knowledge of the factors regulating these fundamental processes,

FIGURE 1. Principal processes involved in the removal of carbon, nutrients, trace elements,and pathogens during effluent application to land.


as well as the complex interactions among them, is prerequisite for achiev-ing a reliable treatment, particularly in terms of organic matter degradation,pathogen elimination, and nutrient removal. Moreover, the long-term treat-ment performance and the sustainability of the land remain equally importantissues. Plant selection is among the most critical components mediating thesuccessful performance of SRS (Crites et al., 2000). Vegetation may affectdramatically the performance of land treatment systems through its effectson hydraulic loading, nutrient uptake, biomass production, microbial com-munity (structure and activity), and other specific functions such as traceelements uptake/inactivation and toxic organics degradation/inactivation.

Despite the important advantages of SRS in wastewater management,the ecological and health risks need to be considered, including the potentialfor contamination of surfacewater and groundwater, and impairment of landsustainability (Bond, 1998; Magesan et al., 1998; Correl, 1998). These risks aremainly due to improper estimation of plant water requirements and nutrientloads, reductions in soil hydraulic conductivity, and accumulation of salts andtrace elements. Application of an appropriate monitoring plan will eliminatethese risks and maintain system performance in the long run.

This article reviews the treatment processes that occur during land ap-plication of wastewater effluent, the role of vegetation, and the appropriatemanagement practices to ensure long-term performance of SRS, environmen-tal protection, and the sustainability of the land.

II. SLOW RATE SYSTEMS: DEFINITIONS—CATEGORIES

Slow rate systems (the predominant natural treatment system currently in use)involve the application of pretreated wastewater to vegetated soil to providetreatment and meet growth needs of the existing vegetation. Depending ontheir design objectives, SRS can be classified as follows:

� Type 1: Systems designed with the objective of wastewater reuse throughcrop or vegetation (including forest trees or forage grasses) growth. Hy-draulic loads applied on those systems depend on climatic conditions,plant species, and leaching requirements. The application of wastewatereffluent to the planted soil can be accomplished by using a variety ofmethods, including sprinklers, surface methods (flooding, furrows, andlocalized methods), or subsurface techniques.

� Type 2: Systems designed on the basis of the limiting design factor. Thecritical parameters mediating hydraulic loading rates of those systems arethe soil hydraulic conductivity, nitrogen concentration, and the pollutantconcentration. Type 2 systems are mainly used in humid areas wherethe disposal of wastewater effluents, rather than reuse, is the principalconcern.


� Filter bed: Systems designed for the application of effluent to filter bedsby flood methods and collection of the treated filtrate by an extensivesubsurface drainage system. These systems are effective in nutrient re-moval from percolating wastewater, which may be discharged to surfacewaters or used for other purposes. These systems are suitable in areaswith high-value land, for example, around urban centers, and can be usedfor wastewater treatment during rainy periods. Finally, the filter bed, withappropriate modifications, can also be used for various types of efflu-ents, such as saline drainage, industrial wastewaters, and effluents frompiggeries and dairies (Jayawardane et al., 1997).

Most of the treatment occurs in the upper soil horizons and vadose zonebecause the soil has a great capacity for carbon and nutrient assimilation. Inaddition, the applied organic matter and nutrients may have a beneficial ef-fect in the treatment ecosystem if they are applied at balanced rates. Thesesystems combine both treatment and reuse of wastewater and may be de-signed as “zero discharge” systems. Most of these systems are found in UnitedStates (U.S. EPA, 1981), but an increasing number may be found in Australia(Duncan et al., 1998) and Europe, especially in central and northern coun-tries, where these systems are used for biomass production (Perttu, 1999).SRS are capable of efficient wastewater treatment in a wide range of climaticconditions, varying from extremely hot and dry climates (Al-Jamal et al.,2002) to cool and wet (Perttu, 1999). This review is concentrated on Type 1SRS; however, information regarding the treatment processes and vegetationfunction is also applicable to the other types of land treatment systems.

III. PREAPPLICATION TREATMENT

A minimum level of treatment before the application of wastewater on theland is essential to avoid problems in transfer and distribution systems, re-duce the potential for odor development, increase the treatment efficiency ofSRS, and minimize environmental and health risks. The level of preapplica-tion treatment should be the simplest and the most cost-effective. Commonpractices include screening and primary treatment to remove coarse solidsand grass, and reduce organic matter. Lagoons or ponds have been usedeffectively for the treatment of wastewater effluents before their applicationin SRS (Al-Jamal et al., 2002; Reed et al., 1995). Lagoons and ponds providean increased level of treatment in terms of carbon, nitrogen and pathogenremoval, and also serve as storage units when climatic conditions are un-suitable for the operation of SRS or effluent application is interrupted forthe system maintenance. Septic tanks have also been used as an effectivepretreatment process. Tzanakakis et al. (2003) found that retention timesvarying from 12 to 24 h for municipal wastewater in a septic tank reduced


biochemical oxygen demand (BOD5), total suspended solids (TSS), and totalcoliform (TC) by an average of 40, 53, and 83%, respectively.

A. Factors Affecting Preapplication Treatment

Principal factors that affect the preapplication treatment requirements are(a) the degree of public access to the site; (b) the degree of process con-trol the application area; (c) the end use of the irrigated crop; and (d) thetreatment object (e.g., removal of organic carbon, nitrogen, or pathogens).Thus, primary treatment should be acceptable for isolated sites with restrictedpublic access when irrigating crops are not intended for direct human con-sumption or when effluent application is implemented by subsurface tech-niques and the underground part of the irrigated crops is not consumed raw.Biological treatment using lagoons or other processes, and strict control ofpathogens should be practiced in locations with public access or for crops tobe eaten raw. In such cases, pathogen concentration in wastewater effluentmust not exceed the limits suggested by the existing regulations or guidelinesfor effluent reuse.

The situation becomes more complicated in the case of industrial efflu-ents. In such cases, advanced treatment processes may be required beforeapplication to land. Depending on the specific wastewater characteristics,common preapplication processes include chemical precipitation, pH adjust-ment, sedimentation, and oil or trace elements removal. Limits for hazardouschemical concentrations in wastewater or the maximum permitted loads ap-plied to the soil have been established by many countries and internationalorganizations.

B. Considerations for Nutrient Removal

For effluents containing increased concentrations of nutrients, for example,food processing wastewater and livestock feedlot runoff, appropriate preap-plication treatment should be used to reduce nutrients at acceptable levelseliminating the environmental impacts. When enhanced removal of nitrogenis the objective of the treatment, an adequate source of carbon is needed inthe water applied to the SRS. Primary effluent with a BOD/N ratio >3 shouldbe used to support nitrification and denitrification processes. To maintainthe BOD/N ratio within the optimum range, only primary sedimentation orequivalent treatment is required. Ponds may be used efficiently to reduce thenitrogen concentration, in addition to other nutrients. Removal of phosphorusmay be also a critical issue before effluent application onto land, especiallywhen land treatment systems are located in sites adjacent to surface waterbodies.


IV. PLANT SELECTION

Vegetation plays a significant role in the treatment efficiency of SRS, affectingthe hydraulic load, nutrient removal, and biomass production. Thus, the pri-mary criteria when selecting vegetation are (a) water requirements, (b) thepotential for nutrient uptake, (c) salt tolerance, (d) trace elements uptakeand/or tolerance, and (e) biomass production. Additional features that shouldbe taken into consideration include climate (frosts, temperature, photoperiod,length of growing season), soil properties (pH, salt, and nutrient concentra-tion), plant availability, length of vegetation cycle, and production direction(e.g., pulp, wood, biofuels and other products).

Currently, a variety of annual crops, perennial grasses, and forest treesare used in SRS worldwide (Table 1). The primary focus is on trees suitable

TABLE 1. Selected Plant Species Used for the Treatment of Various Types of WastewaterEffluents in Slow Rate Systems

Plant species Wastewater type Reference

Arundo donax (reeds) Primary effluent Tzanakakis et al. (2003)Eucalyptus botryoides

(Southern mahogany)Meat processing effluent Guo et al. (2002)

Eucalyptus camaldulensis (redgum)

Primary effluent,stormwater pond

Tzanakakis et al. (2003),Rockwood et al. (1995)

Eucalyptus ovata (swampgum)

Meat processing effluent Guo et al. (2002)

Eucalyptus grandis (rose gum) Secondary effluent,stormwater pond

Falkiner and Smith (1996),Rockwood et al. (1995),Duncan et al. (1998)

Eucalyptus globulus(Tasmanian bluegum)

Secondary effluent, meatprocessing effluent

Guo et al. (2002), Duncanet al. (1998)

Eucalyptus cyanophylla(blue-leaved mallee)

Primary effluent Tzanakakis et al. (2003)

Chloris gayana (Rhodes grass) Secondary effluentenriched with nitrogen

Edraki et al. (2004)

Eucalyptus robusta (swampmahogany)

Secondary effluentenriched with nitrogen

Edraki et al. (2004)

Pinus radiata (Monterey pine) Secondary effluent,primary effluent

Falkiner and Smith (1996),Tomer et al. (1997)

Pasture grass Dairy processingeffluent

Sparling et al. (2001)

Populus × euramericana(hybrid poplar)

Primary effluent, foodprocessing wastewater

Vermes (1996)

Populus robusta (black poplarhybrid)

Primary effluent, foodprocessing wastewater

Vermes (1996)

Populus sp. (hybrid poplar) Domestic effluent andsludge

Moffat et al. (2001)

Salix schwerinii (willow) Secondary effluent Hasselgren (1998)Salix dasyclados (willow) Landfill leachate,

secondary effluentHasselgren (1998), Elowson

(1999)Salix vinimalis (willow) Landfill leachate,

agricultural drainageHasselgren (1998), Elowson

(1999)Acer rubrum (red maple) Landfill leachate Shrive and McBridge (1995)


for short-rotation coppice or perennial grasses regularly harvested during thegrowing season to maximize nutrient removal. A significant number of plantspecies, products of clonal selection, traditional breeding, and genetic engi-neering are currently under evaluation in terms of biomass production, wateruse efficiency, nutrient uptake, and salinity tolerance for future use in SRS.Furthermore, current progress in the field of phytoremediation reveals thatvegetation may contribute in the removal of specific pollutants either directlythrough uptake, stabilization, volatilization, transformation, and degradationor indirectly by stimulating the growth of microorganisms that degrade and/orimmobilize such constituents (Salt et al., 1998; Meagher, 2000; Davis et al.,2002; Siciliano and Germida, 1998). Detailed knowledge of the ability ofvegetation to remove a wide spectrum of constituents, both inorganic andorganic, is important when selecting the most suitable plant species based onthe effluent composition. Slow rate systems can be expanded for the treat-ment of toxic and hazardous wastes under the condition that the applicationrate of wastewater is adjusted based on the vegetation capacity of the systemfor remediation.

The following sections deal with the ability of plants to withstand en-hanced concentrations of salts, trace elements, and conditions of flooding.The effects of vegetation on hydraulic loading, nutrient uptake and removalof persistent organics, are discussed in subsequent sections.

A. Salinity Tolerance

Wastewater effluents contain increased concentrations of salts that must betaken into account when selecting vegetation for SRS since they may ad-versely affect plant growth and survival and hence impair treatment efficiencyof the SRS. Salinity affects plant performance through the development of anosmotic stress and the disruption of ion homeostasis (Paranychianakis andChartzoulakis, 2005). At a whole-plant level, salinity results in growth inhi-bition and yield reductions; in more severe cases, leaf injuries develop thatcan lead to the complete defoliation of plants and subsequent desiccation.

To cope with salinity, plants employ a variety of mechanisms that oper-ate at a cellular and whole-plant level, allowing their adaptation and survivalin saline environments. Mechanisms conferring salt tolerance in plants canbe grouped in three distinct categories: (a) minimize salt accumulation inshoots (salt restriction and/or exclusion, translocation to older tissues, com-partmentation), (b) optimize water use (osmoregulation, increase of wateruse efficiency, morphological changes), and (c) scavenge reactive oxygenspecies (Paranychianakis and Chartzoulakis, 2005).

A number of guidelines have been developed to predict and manage theimpacts of salts contained in irrigation water on plant performance. Theseguidelines were integrated and modified by Ayers and Westcot (1985) to


expand their usefulness. The most common parameter used to predict theimpacts of water quality on crops is the total content of salts expressedas electrical conductivity (ECw) or as concentration of total dissolved solids(TDS). However, these parameters do not take into consideration variationsin the salt tolerance amongst plant species, cultivars, or clones. Maas (1986)provides salt tolerance data for a wide range of agricultural crops includingannual and perennial species used in SRS. However, information on the salttolerance of forest trees is limited. A rating of plant species used in SRS ac-cording to their ability to withstand enhanced levels of salts is shown in Table2. It should be stated that the salt tolerance of a given genotype may vary fromone area to another. Genotypic variations in salt tolerance are attributed to dif-ferences in environmental factors (soil fertility, soil physical conditions, andclimatic factors) between regions and genetic diversity among populations.

TABLE 2. Salt Tolerance Rating of Selected Plant Species With Potential Use in SRS

Tolerance to salt Scientific name Common name

High Pinus radiata Monterey pineEucalyptus spathulata Swamp malletEucalyptus occidentalis Swamp yateEucalyptus camaldulensis Red gumAcacia stenophylla Shoestring acaciaAcacia saligna Orange wattlePopulus euphratica Euphrates poplarMelaleuca lanceolata MoonahMelaleuca bracteata Black teatreeCynodon spp. Bermuda grassBeta vulgaris Sugar beet

Moderate Acacia auriculiformis Darwin black wattleAcacia salicina Willow wattleEucalyptus grandis Rose gumEucalyptus robusta Swamp mahoganyEucalyptus astringens Brown malletEucalyptus botryoides Southern mahoganyMelaleuca leucadendra CajuputMelaleuca armillaris Green globeSorghum verticilliflorum Sudan grassChloris gayana Rhodes grassFestuca spp. Tall fescueMedicago sativa AlfalfaZea mays Corn

Low Acacia pendula Weeping myallAcacia victoriae Bramble wattleEucalyptus globulus Tasmanian bluegumEucalyptus intermedia BloodwoodPopulus x euramericana Hybrid poplarPlatanus spp. SycamorePopulus spp. PoplarSorgum spp. SorgumTrifolium spp. Clover


B. Trace Elements

Trace elements in wastewater effluents exhibit great variations depending ontheir concentration in the corresponding water supply, the type of effluent(agricultural, industrial, municipal), and the level of preapplication treatment(Page and Chang, 1988). Trace elements at concentrations potentially harmfulto environment and public health are found in industrial effluents, agricul-tural drainage, landfill leachates, and in some cases in municipal wastewater,especially in developing countries where industrial discharges are not con-trolled efficiently (Lin et al., 2002; Tanji and Kielen, 2002; Mapanda et al.,2005; Scott et al., 2005).

Vegetation can play an important role in the management of trace ele-ments. Criteria for the successful selection of vegetation will depend on thetype and concentration of trace elements in the wastewater effluent and theend use of irrigated crop.

When wastewater contains relatively high concentrations of trace ele-ments, plant species able to withstand these levels should be used to main-tain the treatment efficiency of the system. The physiological mechanismsinducing tolerance to trace elements are (a) exclusion from entering roots;(b) preferential accumulation in roots instead of shoots; (c) rendering oftrace elements to non-toxic forms through chemical binding; and (d) com-partmentation in the vacuoles. It should be stated, however, that the regu-lations/guidelines regarding the maximum allowable loads of heavy metalson land tend to become stricter, thus limiting the possibility for impacts onplant performance.

In the case of wastewater effluents used for irrigation of crops that are tobe consumed by humans or livestock animals, plant species or genotypes thathave increased ability to exclude heavy metals or sequester them in noned-ible tissues should be used. This characteristic is particularly important forelements that can be harmful to animals at concentrations too low to affectplants, such as cadmium(III), copper(II), and molybdenum(II) (Crook, 1998).When vegetation is used for wood, energy, or pulp production, higher ratesof trace element uptake and partitioning in the hypergeous organs are a desir-able characteristic to maintain soil quality. Although there are not known treespecies that can be characterized as accumulators, the greater biomass pro-duction can counteract this limitation. Greger (1999) evaluated various Salixspp. clones for cadmium uptake and found that some removed five timesmore cadmium than the known phytoaccumulators Thlaspi caerulescens andAlyssun murale. Poplar trees irrigated with municipal wastewater effluent andamended with biosolids displayed an ability to remove cadmium at greaterrates than it was applied on soil (Moffat et al., 2001). It can be inferred thatideal plant species for SRS used for the treatment of effluents with relativelyhigh concentration of trace elements are those that can accumulate largeamounts of trace elements in their tissues, without posing a risk for wildlifeor livestock animals, and produce a high amount of biomass.


In addition, vegetation may affect the risks of trace elements throughits effects on the availability and (phyto)volatilization rate. Plants can mod-ify the chemical composition of the rhizosphere by excreting various sub-stances known as “root exudates.” For example, Eucalyptus spp. reduces soilpH, thus increasing the availability of trace elements that may prove detri-mental to plant growth and the activity of soil microbial community. Also,complex interrelationships between plant species and the structure of mi-crobial community may strongly affect their availability. Plants and microbialbiomass release chelators that increase the availability of metals for plant up-take. Additionally, the release of enzymes in the rhizosphere may reduce theavailability of elements such as copper(II), selenium(II), and chromium(VI),by converting them into non-available forms.

Volatilization by plants can substantially contribute to the removal ofsome trace elements. Adequate information is available for selenium, an el-ement that occurs in relatively high concentrations in agricultural drainage.Terry and Lin (1999) found that the volatilization rate of selenium amongplant species may display a 100-fold variation. The highest volatilizationrates, which approach 350 µg/m2 · soil surface · d, were measured for broc-coli and cabbage, and the lowest rates in sugar beet, lettuce, onion, andbean (15 µg/m2 · soil surface · d) (Terry et al., 1992). Complex interac-tions between microorganisms and vegetation have been found to influ-ence the volatilization rate of selenium. Axenically grown plants or plantstreated with antibiotics showed lower rates of selenium volatilization com-pared to control plants (Terry et al., 2000). Volatilization due to the action ofsoil microorganisms has been also reported for arsenic and mercury. Rughet al. (1996) reported volatilization of mercury from transgenic Arabidopsisthaliana plants expressing a modified bacterial mercury reductase (merA)gene.

C. Flooding Tolerance

Waterlogging constitutes a common situation for Type 2 SRS, in which hy-draulic loadings are based on soil hydraulic conductivity. Such conditionsmay also prevail in Type 1 systems temporarily after effluent application orthe occurrence of a heavy rainfall in conjunction with inadequate drainageor shallow soils. In addition, reductions in infiltration rate, due to sodicity orbiological clogging, could also favor the prevalence of such conditions. Wa-terlogging may be detrimental to plant growth and development and henceon the treatment efficiency of the system. High rates of effluent applicationin a shallow “podzolic” soil at Queensland resulted in the prevalence of sat-urated conditions for prolonged periods causing the death of sensitive plantspecies (ISTRAM, 1993).

Under saturated or flooded conditions, available oxygen in root zone isdepleted, resulting in metabolic disruptions in vegetation due to anoxic con-ditions, reduction of nutrient availability, growth inhibition, root decay, and


TABLE 3. Flooding Tolerance of Selected Plant Species with Potential Usein SRS

Tolerance to flooding Scientific name Common name

High Eucalyptus ovata Swamp gumEucalyptus robusta Swamp mahoganyEucalyptus occidentalis Swamp yatePopulus deltoides Eastern cottonwoodSalix nigra Black willowPopulus balsamifera Balsam poplar

Moderate Eucalyptus camaldulensis Red gumPlantanus occidentalis American sycamore

Low Eucalyptus grandis Rose gumEucalyptus globulus Tasmanian bluegumPinus radiata Monterey pinePopulus grandidentata Big tooth aspenPopulus tremuloides Quaking aspen

early senescence, and after prolonged periods plant death may occur. Theextent of damage is a function of plant age, stage of development, compo-sition of floodwater, and duration of flooding (Kozlowski, 1997). In general,crops are more susceptible to flooding at early stages of development whileenhanced levels of salts exacerbate the impacts of flooding to plant perfor-mance. Plant tolerance to flooding varies appreciably among plant speciesand genotypes (Table 3) due to morphological (adventitious roots, oxygenabsorption by aerial tissues and transport to root, changes in aerenhyma) andmetabolic adaptations (control of energy metabolism, availability of energyresources, and protection against postanoxic injury) (Armstrong et al., 1994;Yamamoto et al., 1995).

V. HYDRAULIC LOADING RATES

As already mentioned, application of wastewater effluent on Type 1 SRS isbased on plant water requirements. Accurate estimation of plant water de-mands, defined as evapotranspiration (ET), is important for achieving a highdegree of treatment. Overestimation of ET may cause environmental risksresulting from nutrient and/or pathogen leaching or runoff. Application oflower hydraulic loads would reduce treatment efficiency of the system andincrease land requirements. Because land acquisition represents a major costfor land treatment systems, it is often a limiting factor for the developmentof such systems (Reed et al., 1995). As a consequence, there is interest inreducing the land requirements by adopting vegetation with high ET rates.However, ET rate should not be the only parameter considered when se-lecting plant species for SRS, because the application of hydraulic loads tomeet plant water demands generally exceeds vegetation capacity for nutrient


uptake and removal. Nutrient overloading becomes more pronounced in thecase of nutrient-rich effluents such as dairy, meat processing, piggery, oliveoil, and food-processing effluents. Thus, a major issue in SRS is the need tomatch ET with the capacity of vegetation to remove nutrients. Plant speciesexhibiting high water use efficiency (WUE) and low nutrient use efficiency(defined here as the dry matter production per nutrient content) should beused when wastewater effluent contains enhanced levels of nutrients. Incontrast, when relatively low concentrations of nutrients are contained, as inthe case of municipal wastewater after advanced preapplication treatment,species with low WUE can be used.

Plant species possessing the C4 metabolic pathway, such as maize, sugarcane, miscanthus, and switchgrass, have nearly double the WUE comparedto C3 plants due to their higher efficiency of photosynthesis (Stanhill, 1986).Great variations in WUE have also been reported among plant species pos-sessing the same metabolic pathway and among clones of the same species.For instance, significant differences in WUE were found among clones ofEucalyptus spp. that further changed with soil water status and nutrient avail-ability (Weih and Nordh, 2002). Typical values of WUE for species used inSRS are shown in Table 4.

TABLE 4. Indicative Values of Water Use Efficiency (WUE) of Selected Agri-cultural Crops and Forest Trees Commonly Used in SRS

Plant species WUEa (g dry matter/kg water) Reference

Maize 3.00 Howell et al. (1988)Miscanthus 7.80 to 9.50 Beale et al. (1999)Beets 13.5b Fabeiro et al. (2003)Sweet sorghum 5.7 Mastrorilli et al. (1999)Willow 3.0 to 3.7 Lindroth et al. (1994)c

Eucalyptus spp. Tzanakakis et al. (2003)c

1st year 0.712nd year 1.713rd year 2.6

Acacia spp. Tzanakakis et al. (2003)c


Reeds Tzanakakis et al. (2003)c


Populus spp. Tzanakakis et al. (2003)c


aEstimation of WUE is based on evapotranspiration losses and aboveground biomass.bDuring 1 September to 31 October, beets were irrigated at 0.85ET.cCalculated using data provided in specified reference.


A. Estimation of Evapotranspiration

Crop evapotranspiration rate is mainly governed by climatic conditions thatprevail in a given area and crop characteristics. Various methodologies havebeen developed to estimate the influence of climate (defined as potentialevapotranspiration, ETo) on ET, including the Blaney–Criddle, radiation, Panevaporation, and Penman–Monteith methods (reviewed in Doorendos andPruit, 1977). These methods differ in the accuracy of the estimate of ETo andthe climatic parameters required. When application of wastewater effluent ispracticed, increased accuracy is needed to ensure that deep percolation orrunoff does not occur; thus, pan evaporation or Penman–Monteith methodsare mostly recommended. Allen et al. (1998) updated the classical Penman–Monteith method, which provides a greater accuracy for the calculation ofETo for different intervals of time and under nonstandard conditions, suchas for crops suffering from water and salinity stress, nutrients deficiency,and waterlogging, for isolated stands and for crops grown together in thesame field. This method is currently accepted as the most precise availableto estimate reference evapotranspiration (ETo).

B. Affect of Plant Species

Crop evapotranspiration may differentiate substantially from the ETo due todifferences in physiological mechanisms and morphological factors amongplant species (water uptake, transpiration rate, foliage development, plantheight, and root distribution and density). Differentiation in ET among plantspecies, resulting from differences in leaf area development, will diminishafter canopy closure, when water consumption reaches its maximum value.In contrast, differences in ET resulting from transpiration rate will remainand after full canopy development in plantations. Moreover, differences intranspiration rate among plant species appear to be also a function of climaticconditions. Benyon et al. (2001) reported a 22% decline in transpiration rateper unit of leaf area of Eucalyptus leucoxylon trees when the vapor pressuredeficit reached values of 6 to 7.5 kPa compared with Eucalyptus spathulata,Eucalyptus occidentalis, and Eucalyptus cladocalyx, while at lower vaporpressure deficits no differences were observed in transpiration rate per unitof leaf area among plant species.

Differences in water consumption among plant species are reflectedby the crop coefficient (Kc), which varies predominately with the specificcrop characteristics and to a lesser extent by climate. The value of Kc fora given crop varies during the growing season due to changes in canopydevelopment and plant height. Overall, Kc increases as the climate becomesdrier and the wind speed increases until a threshold value depending on theplant species.

Representative values of Kc for a wide variety of plant species are pre-sented by Allen et al. (1998). However, limited data are available regarding


the Kc values for forest trees used in SRS. The crop coefficients for thesespecies are expected to be equivalent to those of fruit and nut trees, ap-proximately 0.9 (Allen et al., 1998). Based on the data of Myers et al. (1994),Al-Jamal et al. (2002) calculated the Kc for an Eucalyptus grandis plantationlocated in Australia and irrigated with treated wastewater. However, whenthat Kc was used to estimate the ET of Eucalyptus camaldulensis trees inMexico resulted in an underestimation of crop water needs and the treessuffered from water stress. Although the reasons for the differences in theKc value between the two locations were not discussed in that work, likelydifferences between the two species in transpiration rate and/or the differentclimatic conditions may have resulted in the altered water needs. The greaterwater requirements are somehow surprising because wastewater effluent inMexico was characterized as having higher salt concentrations compared tothat of Australia, and because leaching of the accumulated salts was notpracticed, plant water use should have been reduced. These results questionthe validity of transfer of Kc values among regions with distinct differencesin climatic conditions even for plant species of the same genera. Moreover,differences in water consumption have been found not only between plantspecies but also among clones or provenances. Values of Kc for a wide varietyof plant species used in SRS are shown in Table 5.

Water consumption by plant species appears to have an upper threshold.Values of ET up to 15 to 17 mm/d have been measured, but wilting may occurunder such conditions, which decreases the ET (Doorendos and Pruitt, 1977).In general, few commercial crops show seasonal water use higher than 1.5 m,

TABLE 5. Values of Crop Coefficient (kc) for Selected Crops with PotentialUse in SRS at Different Stages of Growth Cycle

Crop coefficient (kc)

Plant species Initial growth Intermediate growth End of growth

Maize — 1.20 0.35–0.60Beets 0.35a 1.20a 0.50a

Alfalfa 0.40a 1.20a 1.15a

Sorghum — 1.10a 0.55a

Bermuda grass 0.55a 1.00a 0.85a

Rye grass 0.95a 1.05a, 0.79b 1.00a

Sudan grass 0.50a 1.15a 1.10a

Eucalyptus 0.50c 1.20c, 0,85b 0.50c

Poplars 0.30c 0.70d, 1.00c 0.30c

Willow — 1.1–1.5e —Pine 0.10c 0.80c 0.10c

aFrom Allen et al. (1998).bEdraki et al. (2004).cMyers et al. (1999).dGochis and Cuenca (2000).eDepending on clone and growing season (Jørgensen and Schelde, 2001).


but values up to 2.5 m have been reported (Al-Jamal et al., 2002; Jimenezand Chavez, 2004).

C. Effect of Cultural Practices

Cultural practices and management strategies such as planting density, lengthof harvesting cycle, multiple cropping systems, application intervals, andsalinity management may have a strong effect on crop water demands. Waterconsumption increases with planting density due to the greater leaf areadevelopment, but that effect disappears after canopy closure. Thus increasingof planting density in short rotation coppice (2 to 3 years) will significantlyincrease water consumption.

Weeds or crop cover may significantly affect water use in SRS plantedwith trees. The contribution of cover crop to the system water balance de-clines as trees grow, due to shading, and nearly disappears when canopyclosure is achieved. Water use by the understory grass in a Eucalyptus spp.plantation was estimated at 97 and 86 cm during the first and the secondyear after the establishment of plantation, respectively, but it was negligiblethe third year (Al-Jamal et al., 2002). Plantations of Eucalyptus robusta withgrass understory showed only a slight increase in ET (15%) compared tograss plots (Edraki et al., 2004). In the case of annual crops, the adoptionof multiple cropping systems in SRS would increase water use due to thegreater length growing period.

D. Models for Estimating Water Balances

Currently a wide range of models such as HYDRUS, UNSATH, LEACHM,HELP, and tipping-bucket models have become available to estimate waterbalances under field conditions (Allen et al., 1998). Selection of the suitablemodel is a function of project objectives, site characteristics, and environ-mental considerations. Mahmood et al. (2003) developed a simple and prac-tical model for effluent application in SRS that requires data of soil moisturecontent, rainfall, and groundwater quality. In addition, Snow et al. (1999)developed a water balance model (APSIM) for effluent-irrigated plantations.The APSIM model includes modules for (a) tree growth, nutrient uptake,and water demand (GRANDIS); (b) weed growth, nutrient uptake, and waterdemand (WEED); (c) interception of rainfall and irrigation, evaporation ofirrigation (INTERCEPT); (d) water and solute movement (SWIM); (e) inter-ception of radiation by trees and weeds (CANOPY); and (f) salt and chloridetracking (SOLUTE). During 5 years of monitoring, of the 3370 mm of rainfalland 4480 mm of effluent application, 6710 mm were lost as evapotranspira-tion and 1080 mm as deep drainage. The APSIM model therefore providespromising results for application in effluent-irrigated sites, although somecalibration is further needed to reduce deep drainage.


VI. ORGANIC MATTER REMOVAL

Despite the beneficial effects of organic matter on soil fertility, its presencemay favor the survival of some pathogens and the transport of both pathogensand pollutants (trace elements and toxic organics) downward. Furthermore,there has been concern regarding the fate of a wide spectrum of persistentand toxic organic substances that are found in the various types of wastew-ater effluents and particularly in agricultural drainage (pesticides), olive milleffluents (phenols), industrial wastewaters, and landfill leachates (Jones et al.,2005a; Scott et al., 2005; Azbar et al., 2004). Even for municipal wastewaters,where such contaminants occur at relatively low concentrations, long-termaccumulation in the soil may result in the risk of groundwater contamina-tion (Bouwer, 2000). Moreover, the application of effluent to land has beenassociated with the deep percolation of organic pollutants, such as pesti-cides, that already exist in the soil (Downs et al., 1999). Traces of organicscontaminants were also reported in the runoff from an effluent-applied site(Pedersen et al., 2003).

A. Biodegradable Organic Matter

Biodegradable organics, such as carbohydrates, amino acids, and proteins,do not limit the performance of SRS. Applied organic loads may vary greatly,depending on the level of preapplication treatment, type of wastewater, andhydraulic loading rates. Typical organic loads applied to SRS receiving do-mestic wastewater fluctuate from 50 to 100 kg BOD/ha · d, but can be ashigh as 330 kg/ha · d in the case of food processing effluents or other indus-trial effluents. It is recommended that organic loads do not exceed 500 kgBOD/ha · d (Reed et al., 1995). Indeed, successful treatment of dairy wastew-ater in an SRS receiving at rates of 36,300 kg C/ha annually (about 202 kgBOD/ha · d) for 22 years has been reported (Sparling et al., 2001), demonstrat-ing the outstanding ability of these systems to degrade organic matter whenmanaged properly. Despite these high organic loads, system performancehad not been adversely affected and soil organic matter did not increase(Degens et al., 2000).

Organic matter in wastewater effluents is present in the form of par-ticulate matter, colloidal, and dissolved carbon. The relative proportionsof these fractions may vary appreciably depending on the effluent originand the preapplication treatment. Initially, during the application of effluentto the land, physicochemical processes such as straining, sedimentation, in-terception, and adsorption may contribute in the removal of organic matter.Thereafter, biological degradation and oxidation dominate in the removal oforganic matter. Generally, soil microbial biomass and the activity of enzymesinvolved in carbon cycling (xylanase, invertase, β-glucosidase, and cellobio-hydrolase) and nutrients cycling (urease, deaminase, N -acetylglucosamidase,


phosphomonoesterase) increase after the application of wastewater effluentto the land, due to the enrichment of the soil with energy substrates and nutri-ents. Application of wastewater effluent to the soil increased the numbers ofbacteria, actinomycetes, and fungi compared to soil that had never receivedeffluent. Similarly, the ATP content and the activities of β-glucosidase, β-acetylglucosaminidase, and proteinase were higher in effluent-irrigated soil(Filip et al., 2000). Application of dairy effluent to the soil was found to in-crease protease and deaminase activity (Zaman et al., 2002). Likewise, appli-cation of olive mill wastewater resulted in a rapid increase in respiration, de-hydrogenase and urease activities, and microbial biomass (Piotrowska et al.,in press).

The major portion of organic matter mineralization occurs in the first fewcentimeters of soil. Zaman et al. (2002) found that during the application ofdairy effluent, the increase in microbial biomass and the activity of enzymesresponsible for carbon and nitrogen mineralization occurred mainly in thefirst 10 cm of soil depth. Tzanakakis et al. (2003) reported 91% removal ofCOD at a depth of 15 cm, but no further reduction was observed below thisdepth in a clay-loam soil. It is likely that the anoxic conditions that prevailat higher depths reduce the activity of microbial biomass (Angelakis andRolston, 1985). Under such conditions, microbial activity is reduced to lowlevels, which may result in the accumulation of easily degradable organicsubstances (Kuzel and Drake, 1999).

Since biodegradation is mainly accomplished by aerobic and facultativemicroorganisms, unsaturated flow is essential for the maintenance of high re-moval rates. Hydraulic conditions should be controlled in order to increasethe contact time of percolating wastewater with the soil matrix and hencewith the active biofilm. Cyclic application of wastewater effluent is neces-sary to allow for the restoration of aerobic conditions and to control biofilmdevelopment.

1. AFFECT OF SOIL FACTORS

Soil texture, structure, and chemical attributes may exert a strong influenceon the degradation rate of organic matter, through their effects on soil aer-ation, organic matter availability, and the structure and activity of the mi-crobial population. Lower mineralization rates have been reported in clayeysoils compared to coarse, sandy soils, which were attributed to (a) the re-duced availability of carbon to microorganisms in clayey soils and (b) en-hanced microbial growth efficiency (van Veen and Kuikman, 1990). Clayeysoils are characterized by small pores, with more than 50% correspondingto sizes smaller than 0.2 µm. In these small pores, dissolved organic matter(DOM) is not easily accessible by microorganisms and the degradation ratedeclines (Marschner and Kalbitz, 2003). Another factor that should be takeninto consideration with clayey soils is the prevalence of anaerobic conditions


for a longer interval after effluent application, which may adversely affectmineralization rate due to the lower availability of oxygen. Stenger et al.(2001) did not find a clear relationship between soil texture and the mineral-ization of organic matter when dairy farm wastewater was applied to intactsoil cores. Soil pH can affect organic matter biodegradability by affecting mi-crobial community structure and activity, organic matter solubility, or nutrientand heavy metal availability (Marschner and Kalbitz, 2003).

2. AFFECT OF ORGANIC MATTER CHARACTERISTICS

The composition of organic matter, which may vary appreciably among differ-ent types of effluents, exerts a significant influence on carbon mineralization.Generally, proteins, amino acids, starch, and sugars are easily biodegradablesubstrates compared to lignin, cellulose, phenols, and aromatic substances,which are characterized by slow decomposition rates (Rowell et al., 2001;Marschner and Kalbitz, 2003; Jones et al., 2005b). However, the chemistryof organic matter cannot fully explain mineralization potential of a given or-ganic substrate since bonding with other organics and differences in the sizeof organic matter may substantially affect mineralization rate (Marschner andKalbitz, 2003). In a recent study, 23 wastewaters from various food processingindustries were characterized according to their carbon and nitrogen concen-trations, dissolved and solid organic phases, and acid solubility. Their effectson mineralization during soil incubation were then assessed (Parnaudeauet al., in press). Carbon mineralization was found to correlate with the C:Nratio of the acid-soluble fraction and the carbon present in the liquid fraction,but these relationships could explain only a small proportion of the varianceobserved in mineralization rates among wastewater types. Furthermore, greatdifferences in mineralization rates among wastewaters of the same origin andeven of the same wastewater at different sampling dates were observed, sug-gesting that additional physical and chemical factors should be consideredwhile predicting mineralization rate of organic matter during land application.

3. AFFECT OF NUTRIENTS

Generally, nutrients are abundant in most wastewater effluents and do notappear to limit the growth of heterotrophic microorganisms, which are re-sponsible for organic matter mineralization. It is important to note, however,that it is not only the absolute concentration of nutrients but also their rel-ative availability compared to organic matter content and form that affectsthe mineralization rate. The ratio of C:N has been extensively used to char-acterize the decomposition potential of various organic substrates (Magesanet al., 1999; Trinsoutrot et al., 2000). Increasing the C:N ratio in wastewatereffluent resulted in the accumulation in extracellular organic compounds andsoil blockage (Magesan et al., 1999). Parnaudeau et al. (2005) reported a pos-itive correlation between carbon mineralization and the C:N ratio of the acid


soluble fraction of various food processing effluents. Nutrients may also affectorganic matter mineralization through their effects on carbon bioavailability.Both positive and adverse effects have been reported for calcium, depend-ing on the nature of the DOM. The degradation of hydrophobic compoundswas stimulated when calcium was present at similar molar concentrationsto DOM (Jandl and Sletten, 1999). When calcium availability was increased,the degradation of hydrophilic compounds was inhibited, an effect that wasattributed to the formation of stable complexes with the compounds or theirmetabolites. Enhanced concentrations of potassium have been reported toincrease the solubility of organic matter (Marschner and Kalbitz, 2003).

4. AFFECT OF HEAVY METALS

Heavy metals may influence the degradation rate of organic matter eitherby reacting with the DOM to form stable complexes (Alberts et al., 2001) orthrough effects on the microbial population and activity (Giller et al., 1998;Hayat et al., 2002). However, it is difficult to establish threshold levels forheavy metals that are toxic to the soil microbial community (Giller et al.,1998). This uncertainty is due to the inability of available methodologies todetermine with accuracy the bioavailability of heavy metals and differences inmicroorganism sensitivity. In addition, toxic effects are dependent on the typeand combination of heavy metals in the soil (Renella et al., 2002). The impactsof trace elements on microbial biomass determined in laboratory studies maynot accurately reflect that under field conditions. The gradual accumulation ofheavy metals under field conditions allows microbial communities to developtolerance mechanisms; thus, the more resistant traits dominate and the effectsof metal toxicity are reduced. In addition, a portion of the accumulated metalsis converted into nonavailable forms due to sorption to solid phases or toprecipitation reactions.

5. AFFECT OF SALTS

Although there is no published work on the effect of dissolved salts inwastewater effluent on the mineralization of organic matter in SRS, it canbe inferred that salts decrease the mineralization rate, as this has been ob-served for natural organic matter in soils with increased salinity. The adverseeffects of salinity on mineralization rate can be attributed to both the toxic ef-fect on soil microbial biomass caused by enhanced levels of chloride(−I) andsodium(I), and to the reduction of osmotic potential. Contradictions found inthe literature with regard to the effects of salinity on soil microbial biomassprobably result from differences in the levels of salinity applied and theexperimental conditions and methods used. The effects of salinity are moredifficult to characterize in sodic soils, where the physical effects of sodium onsoil texture (clay dispersion) may result in conditions that limit the availabilityof organic matter or form protective coatings of clay over particulate matter


(Nelson et al., 1996). Rietz and Haynes (2003) reported a negative exponen-tial relationship between EC of saturated past extract and microbial biomass,the percentage of organic carbon present as microbial biomass, and the ac-tivities of extracellular enzymes. Negative correlations were also obtained forsodium adsorption ratio (SAR) and exchangeable sodium percentage (ESP),but they were best described with linear relationships. Although it is not pos-sible to separate the effects of salinity from those of sodicity, the exponentialdecrease in the amount of microbial biomass indicates that slight increases insoil salinity may have a significant impact on the degradation rate of organicmatter in the land. Increasing soil EC from 0 to 50 dS/m reduced the amountof microbial biomass from 600 to approximately 200 mg C/kg soil, but afterthis threshold value no further reduction of microbial biomass was observed.Pathak and Rao (1998) reported that carbon mineralization decreased from39 to 16% of added organic matter, when the soil EC increased to 97 dS/m.In contrast, increasing sodicity from 2 to 88.8% did not have any effect oncarbon mineralization. Solubilization of organic matter due to high pH mayhave counteracted the adverse effects of sodium on soil microorganisms andresulted in the discrepancy in the effects of sodium on soil microbial biomass.A significant difference between saline soils and SRS is that in the former casedecreased organic matter decomposition is accompanied by a reduction inorganic matter input because of reduced plant growth and the fact that plantorganic matter does not accumulate. In contrast, in SRS where the organicmatter is mainly exogenous, reduced rates of decomposition due to salin-ity may result in organic matter accumulation, odor development, and otherenvironmental impacts.

B. Toxic Organic Pollutants

In general, SRS are considered to be more efficient in the removal of toxicand persistent organic compounds than other types of natural systems orconventional WWTPs (Reed et al., 1995). The principal processes that deter-mine the fate of persistent organics during land application are sorption onsoil colloids and organic matter, volatilization, microbial degradation, plantuptake, transformation, and/or subsequent degradation. The contribution ofeach mechanism to the remediation of xenobiotics varies greatly with thephysicochemical properties, of the contaminant, soil properties and environ-mental conditions.

1. SORPTION

The term sorption is used to denote the uptake of a solute by the soil matrixwithout reference to a specific mechanism (Chiou, 1989). Sorption includesboth adsorption and absorption but it is often impossible to separate theseprocesses (Pignatello, 1989). Here the term sorption is adopted to describeboth processes.


Sorption plays a significant role in the fate and transport of environ-mental contaminants as it reduces their bioavailability and thus minimizesecological risks (Alexander, 2000). However, sorption reduces the biodegra-dation of xenobiotics as they become less accessible to microorganisms fortransformation or degradation. Under certain circumstances, desorption oforganic pollutants can occur and contaminated soils may act as a source oforganic pollutants. Release of organic contaminants was observed from soilcolumns of two former effluent infiltration sites when microbial respirationincreased as a result of alterations in solution pH, temperature, and drying(Reemtsma et al., 2003).

Sorption is due to a range of physical and chemical processes suchas ion trapping in micropores, exchange, cation bridging, ion-dipole inter-actions, ligand exchange, charge transfer, hydrogen bonding, and van derWaals forces. It is reported that sorption takes place in two distinct phases:a rapid phase, which may last from a few hours to a few days, and a secondslow phase, which may last several days or months (Brusseau et al., 1991).The rapid phase is attributed to the fast movement of contaminants intomacropores and their adsorption on soil particles or organic matter. With thepassage of time, contaminants may enter micropores or the condensed matrixstructure of humic acids, which makes their desorption difficult (Pignatelloand Xing, 1996).

Physicochemical properties of the contaminants, such as solubility, po-larity, charge, pH, redox potential, and hydrophobicity, have a great influenceon the sorption of organic contaminants to the soil matrix. Hydrophobicityis the most important parameter affecting the sorption of nonanionic sub-stances, and it is expressed by the octanol/water partition coefficient (Kow),or log Kow. Sorption of a pollutant in the soil matrix increases with increasinglog Kow. Thus, hydrophobic molecules, such as polychlorinated biphenyls(PCBs), and polychlorinated aromatic hydrocarbons (PAHs), with values oflog Kow > 3, are tightly bound to the soil organic matter. Values of Kow for avariety of organic compounds of interest are presented in Table 6.

Soil properties, particularly clay content, organic matter content andform, and pH, have been found to influence the sorption organic contami-nants (Karthikeyan and Kulakow, 2003; Sun et al., 2003). Clay content has asignificant effect in the case of charged and polar organic compounds. Soilorganic matter plays a critical role in the sorption of organic pollutants, andit has been suggested as the most important process affecting the removal ofnon-ionic hydrophobic pollutants (Chiou et al., 1983; Schwarzenbach et al.,1993; Sun et al., 2003). Knowledge of the distribution coefficient for partition-ing of pollutants into organic matter (Koc) and the mass of soil organic matteris required for estimating the sorption of pollutants to soil matrix. The valueof Koc for a given chemical has been suggested to be fairly constant (Changand Page, 1988). However, differences in the composition of soil organicmatter among areas have been reported to affect Koc value by up to one


TABLE 6. Soil Distribution Coefficient for Partitioning Into Or-ganic Matter (Koc), Water–Air Partitioning Coefficient (Hi), andOctanol–Water Partitioning Coefficient (Kow) of Selected Or-ganic Substances

Compound Koc Hi Kow

Chloroform 81 8.3 93Bromodichloromethane 70 — 76Dibromochloromethane 99 — 123Bromoform 140 — 200Carbon tetrachloride 246 1 437Methyl chloride 129 — 1781,1,1-Trichloroethane 113 0.7 to 6.4 148Tetrachloroethylene 230 — 398Chlorobenzene 343 6.9 to 9.2 6921,2-Dichlorobenzene 839 12.2 23991,3-Dichlorobenzene 838 6.8 23981,4-Dichlorobenzene 853 7.5 24551,2,4-Trichlorobenzene 3607 7.3 18,197Ethylbenzene 573 3.74 1413m-Xylene 622 — 1585p-Xylene 573 — 1413Naphthalane 825 — 2344Dimethyl phthalate 104 47,076 132Diethyl phthalate 643 — 1660Di-n-butyl phthalate 17,140 — 158,489Bis(2-ethyl hexyl) phthalate 20,230 — 199,526PCB (Arochor 1242) 32,181 40.5 380,189Lindane 1474 1047 5248

Note. Adapted from Chang and Page (1988).

order of magnitude, with humic acid to be characterized by higher values ofKoc compared to fulvic acid (Kimani-Njoroge et al., 1998; Kille et al., 1995).Moreover, soil pH may affect sorption through its effects on the charge ofsoil organic matter, clay minerals, and organic pollutants. Increases in soil pHresult in the ionization of weakly organic acids, making them more anionicand therefore increasing the electrostatic repulsion with negatively chargedsoil colloids; an opposite effect occurs for cationic pollutants. In general,sorption reaches its maximum potential when soil pH approaches the pKa

value of the pollutant.In terms of effluent composition, pH and organic matter present in col-

loid or dissolved form are expected to have the greatest influence. Organicmatter in wastewater effluent may have dual effects on the sorption by form-ing organic complexes with pollutants or by competing with them for theavailable sorption sites. The effects of effluent DOM on the sorption of pol-lutants are not clear, due to differences in its structural fractions among var-ious wastewater types, the properties of the pollutants, soil properties, andthe complex interactions among these factors. As a consequence, contrast-ing effects are found in the literature. Seol and Lee (2000), investigating the


effect of DOM from different sources (municipal effluent, swine effluent, andwater enriched with humic acids) on triazine sorption, did not find a signif-icant influence. In contrast, the desorption of phenathrene and pyrene wasfavored by the application of DOM derived from pig manure effluent andcompost (Cheng and Wong, in press). Ilani et al. (2005) found that the hy-drophobic neutral fraction of effluent DOM had a greater sorption capacityfor s-triazines and PAH than the hydrophobic acid fraction. Moreover, en-hanced sorption of the compounds tested was observed in the effluent DOMof the Netanya plant, which contained a higher level of hydrophobic frac-tions, suggesting that hydrophobic fraction of DOM in wastewater effluentexerts a strong effect on the sorption of pollutants.

2. VOLATILIZATION

Volatilization is the evaporation of compounds found in a liquid or solid statein the soil matrix to the atmosphere. Volatilization is an important pathwayfor the removal of nonionic compounds for which sorption does not con-stitute a significant removal pathway. In municipal effluents, trace organicsconsist primarily of low-molecular-weight and nonpolar organic substanceswhose high vapor pressure and low solubility in water make them especiallyconductive to volatilization (Chang and Page, 1988). Losses of diesel fuelthrough volatilization were found to be as high as 58% over 360 days, sug-gesting that volatilization is an important pathway in the removal of PAH, afact that is often overlooked in short-term studies (Kroening et al., 2001). Incontrast, volatilization of phenols from a soil receiving olive mill wastewaterwas estimated at only 0.1% (Rana et al., 2003).

Volatilization rates are strongly dependent on the physicochemical prop-erties of the compound, the soil characteristics, and the climatic conditions.The potential of a compound for volatilization can be estimated quantitativelyusing Henry’s law. The constant of Henry’s law (Hi) provides a measure ofthe compound’s tendency to partition to air relative to water. Organic com-pounds with Hi > 10−4 tend to move mainly through the air spaces in thesoil matrix and hence display enhanced volatilization rates, whereas organ-ics with Hi < 10−6 tend to move predominantly in the liquid phase of thesoil. Contaminants with intermediate Hi values are mobile in both air andwater phases (Ryan et al., 1988). However, it appears that the ratio of Hi tothe octanol–water partition coefficient (Kow) (Wang and Jones, 1994) or theoctanol–air partition coefficient (Koa) (Harner and Mackay, 1995) providesa better estimation of volatilization, because both constants account for thesorption of compounds into organic matter.

Soil texture and properties may also have a strong effect on the volatiliza-tion of toxic and persistent organics. Organic content favors the sorption oforganic compounds, particularly nonionic ones, by increasing the soil parti-tion (Cousins et al., 1999). Soil porosity, including both pore distribution and


size, is also important. The higher rate of triallate (herbicide) volatilizationin a loamy soil compared to a clay one was attributed to its larger size ofpores. Soil texture in general exerts minor effects except in the case of polarcompounds whose the volatilization rate may be lower in clayey soils. Withregard to climatic factors, temperature effects include the increases in thevapor pressure of the contaminants and the upward movement of water dueto evaporation. Wind may also affect volatilization through its effect on theresistance of gas-phase transport. Increasing wind velocity from 3 to 9 m/sraised volatilization of triallate from 40 to 53% in a loam soil and from 60 to73% in a sandy soil (Atienza et al., 2001).

The effect of vegetation on volatilization rate is less defined becauseboth stimulative and inhibiting effects can occur concurrently. Watkins et al.(1994) reported greater losses of naphthalene from a planted than from anonplanted soil. They attributed these increased losses to plant uptake andsubsequent volatilization; however, it is more likely that other factors, inparticular evapotranspiration, which is greater in vegetated soils, inducednaphthalene transport to the soil surface (Cousins et al., 1999). Vegetationmay also alter the physical, chemical, and biological properties of the soiland hence may have indirect effects on volatilization losses. Root growthincreases soil bulk density, which in turn leads to a reduction in soil poros-ity, limiting the transfer of contaminants to the gaseous phase (Karthikeyanand Kulakov, 2003). The increase in soil organic matter in planted soils, dueto root exudates and litter decomposition, enhances the fraction of vola-tized compounds that partition onto the organic matter. Finally, vegetationcan modify the local microclimate by decreasing soil and air temperatureand wind velocity, thus resulting in lower volatilization losses for plantedsoils than for bare soils. To develop valid models for estimating losses ofvolatile organic compounds (VOC) from SRS, more research is required inorder to understand the effect of vegetation on the volatilization rate of VOC.Moreover, management practices may influence volatilization rate of organicpollutants. Subsurface application methods will result in significantly lowervolatilization rates than surface methods, especially sprinkler application.

The estimation of volatilization losses should be a major concern for SRS,not only for the treatment efficiency, but also for the potential health risksthat originate from the inhalation of contaminants. The emission of phenolsfrom soils after olive mill wastewater has been spread was found to constitutea serious risk to humans, as levels have been shown to exceed the allowablelimits suggested by European Union (Rana et al., 2003).

3. MICROBIAL DEGRADATION

Microbial degradation constitutes an important pathway for the removal oftoxic and persistent organics. Contaminants can be utilized by soil microor-ganisms as a source of carbon and energy, or they can be cometabolized


in the presence of an organic substrate that is used as a carbon and energysource. Progress in the field of bioremediation over the last two decades hasrevealed a wide range of microorganisms capable of degrading toxic contam-inants and strains able to adapt their metabolism to utilize available carbonsources. Such microorganisms include bacteria of the Acinetobacter spp., An-throbacter spp., Corynebacterium spp., Nocardia spp., Pseudomonas spp.,Xanthomonas spp. and fungi of the Aspergillus spp., Candida spp., Peni-cillium spp., and Trichosporon spp., while the list of such microorganismscontinues to increase. Web-based databases are currently available that pro-vide detailed information on degradative microorganisms and the hazardoussubstances that they degrade (Urbance et al., 2003).

The extent to which biodegradation contributes to the removal of organicpollutants depends on the chemical properties of pollutant, soil properties,environmental conditions, and effluent composition through their effects oncontaminant bioavailability, the composition and the activity of microbialbiomass, and the potential of the compound for decomposition. It is im-portant to note that the soil microbial community may influence pollutantbioavailability through the release of enzymes, surfactants, and emulsifiers,which promote the desorption of pollutants (Davis et al., 2002; Makkar andRochne, 2003; Fava et al., 2004). Moreover, recent findings suggest that soilmicroorganisms are able to degrade a portion of organics that are sorbedonto the soil matrix and that are not considered as bioavailable (Guerin andBoyd, 1997; Park et al., 2001). Guerin (2000) reported that land applicationwas efficient in the treatment of oily wastewater (2% w/v petroleum hydro-carbon to soil), an effect that was attributed to the increase of hydrocarbon-degrading microbial populations. The maximum rates of degradation reached250 mg/kg · d, while the average degradation rates varied from 10 to35 mg/kg · d. The adoption of certain management practices, such as bioaug-mentation, optimization of ecological conditions that stimulate the growthand the activity of such microorganisms, and the selection of vegetation, maysubstantially improve the potential of the soil for organic pollutants removal.

It is important to note that most findings regarding the removal of or-ganic pollutants in the soil come from laboratory studies when pure organicsare added to the soil. When wastewater effluent is applied to the soil, theorganic matter present may have dual effects on the degradation of xenobi-otics. It may promote the sorption of organic contaminants onto DOM, thuslowering their degradation, or may induce it by stimulating the populationand the activity of the microbial biomass. Overcash et al. (2005) compiled thefindings of studies investigating the fate of xenobiotics applied to the soil inpure form or in conjuction with effluent or biosolids. They found that degra-dation of xenobiotics was stimulated in half the studies but in the other halfit was reduced. Huang et al. (2000) found that effluent application to the soilenhanced biodegradation of chlorpyrifos. In contrast, attrazine degradationdecreased as a consequence of long-term application of effluent (Masaphy


and Mandelbaum, 1997). The contrasting effects observed in organic con-taminant biodegradation for effluent application to land probably result fromdifferences in effluent matrix, soil texture and properties, the presence orabsence of vegetation, and the plant species used, via their effects on thecomposition and activity of microbial biomass and the bioavailability of pol-lutants. This makes it difficult therefore to draw general conclusions abouttheir transport from one site to another without detailed consideration of theprevailing conditions.

Vegetation may affect the biodegradation of toxic organic compoundsthrough specific interactions occurring between microbes and vegetation(Siciliano and Germida, 1998, and references there in). Vegetation may favorthe desorption of pollutants or by exerting a mechanistic action due to rootpenetration in the soil profile and the exposure of contaminants to microbialbiomass previously entrapped in soil micropores. Moreover, it may stimulatethe growth of soil microorganisms through the secretion of roots exudates.Root exudates can affect the degradation of pollutants by increasing theirbioavailability, by serving as a growth substrate that increases the popula-tion of certain microorganisms and their activity, and by secreting substancessuch as phenolic compounds that induce bacterial enzymes involved in PAHsand PCBs degradation (Davis et al., 2002; Reilley et al., 1996). Variations inthe secretion rates of root exudates and/or in their chemical compositioncan be considered responsible for differences in the removal of xenobioticsamong systems planted with different plant species. The release of salicylatefrom plant roots promotes microbial degradation of naphthalene, pyrene, andPCBs (Chen and Aitken, 1999; Singer et al., 2003). The presence of terpenesin some plant species has also been associated with a higher rate of PCBsdegradation. Soils amended with leaves of plant species, rich in terpenes,resulted in the presence of 105-fold more biphenyl utilizers (Hernadez et al.,1997). Common SRS plant species, such as Eucalyptus spp. and Pinus spp.,were included in the study, implying that SRS planted with these species willdisplay greater removal of PAHs and PCBs.

4. PLANT UPTAKE AND TRANSFORMATION

Another pathway for xenobiotics removal is plant uptake and their subse-quent transformation and/or degradation. This pathway, however, is in gen-eral less important than the processes reported previously. Organic contami-nants mainly enter plants roots via diffusion. Thus, principal factors affectingthe uptake of organics are pH, hydrophobicity, polarity, charge, and molec-ular weight (Briggs et al., 1982; Korte et al., 2000). The best correlations oforganics pollutants uptake have been obtained with the logarithm of octanol–water partition coefficient (log Kow). Burgen (2003) compiled the findings ofearlier studies (Briggs et al., 1982; Burken and Schnoor, 1998; Hsu et al., 1990)and suggests that the optimum values of log Kow for the uptake of nonionic


pollutants range from 1.8 to 3.1. Organic compounds with log Kow > 4 aregenerally not available for uptake since they are bound to cell membranes,whereas compounds with log Kow < 1 are not hydrophobic enough to passplant membranes. However, a number of studies indicate a poor correlationbetween log Kow and uptake (Aitchison et al., 2000; Davis et al., 1998), sug-gesting that additional factors may have also a strong effect on the uptake ofxenobiotics.

Entering plant root, organic contaminants can be bound in lignin or cel-lulose, sequestered to vacuoles, volatilized, or metabolized, depending ontheir chemical properties (Davis et al., 2002). Degradation of xenobiotics inplants occurs in three stages, including transformation, conjugation, and com-partmentation, depending on the properties of the pollutant (Sandermann,1992). A variety of enzymes have been characterized, such as nitroreductases,glycosyl and glutathione transferases, oxidases, phosphatases, nitrilases, anddehalogenases, which catalyze transformation and/or degradation reactions,including dehalogenation, hydroxylation, hydrolysis, and oxidation. Thesemetabolic processes have been reviewed in detailed by Wolfe and Hoehamer(2003).

Some of the plant species used in SRS, such as Populus spp., Eucalyptusspp., and Salix spp., have been found to have a high potential for uptakeand degradation of common xenobiotics found in wastewater and other pol-luted waters (Corseuil and Moreno, 2001; Gordon et al., 1998). It shouldbe stressed that most of the results regarding the phytoremediation of thereported compounds come from laboratory studies (pots and cell cultures),which may limit their extrapolation to field conditions.

VII. NUTRIENT REMOVAL

The fate of nutrients, in particular nitrogen and phosphorus, in SRS is ofparticular importance because of the environmental impact on surface andground water associated with these elements. Therefore, in contrast to agri-cultural land, in SRS interest is focused on factors limiting nutrient availabilityrather than increasing it for plant use. The principal mechanisms for nutrientremoval include sorption on soil particles or organic matter, plant uptake,chemical precipitation, denitrification, and volatilization.

A. Phosphorus Sorption

Sorption is considered a significant mechanism of phosphorus removal dur-ing effluent application onto land. It can last for long periods, depending onsoil properties, effluent ionic strength, and hydraulic loading rates.

The extent to which sorption contributes in phosphorus removal variesgreatly among soils with different properties. Phosphorus sorption is greaterin soils with a high proportion of clay than it is in sandy soils, due to the


higher reactive surface area. High organic matter content, often observed insoils receiving wastewater effluent, improves the capacity of soils for phos-phorus retention as it provides additional sorption sites (Eghball et al., 1996).Iron and aluminum oxides also improve the potential of soils for phospho-rus sorption due to their large surface area. The greater capacity to adsorbphosphorus of a “vertisol” soil receiving piggery wastewater compared to a“sodosol” was attributed to its higher content of iron and aluminum oxides(Phillips, 2002).

Phosphorus can be present in soil solution (dissolved), absorbed on thesurface of soil colloids, or sorbed on the soil matrix. It has been reportedthat both fast, reversible sorption processes and slower, time-dependent pro-cesses affect the distribution of phosphorus in the soil matrix (McGechan andLewis, 2002). The initial rapid phase of phosphorus sorption is generally as-sumed to last approximately 24 h and has been attributed to the presenceof high-affinity sites (Sanyal and de Datta, 1991; Phillips, 2002). The slowerphase is attributed to phosphorus diffusion in poorly accessible sites and/orto chemical precipitation. In acid soils, phosphorus deposition in a solid,crystalline form below the surface sorption sites is the dominant slow sorp-tion process. However, in calcareous soils, precipitation reactions dominateand result in the formation of phosphorus carbonates (McGechan, 2002).More than 99% of the phosphorus applied to a “vertisol” was sorbed 10 daysafter the application. In contrast, in a “sodosol” soil only 80% of the addedphosphorus had been sorbed within 3 days after effluent application, and 18days later almost 90% was sorbed (Phillips, 2002).

Continuous applications of phosphorus to the soil may reduce its ca-pacity for phosphorus retention in the long term. Decreases in the capacityof soils for phosphorus sorption have been found in wastewater-irrigated(Menzies et al., 1999; Falkiner and Polglase, 1999) or sludge-amended soils(Sui and Thompson, 2000). This has been attributed to the loss of high-affinity and readily accessible sites. In addition, Menzies et al. (1999) suggestthat part of the decreased capacity of effluent-received soils to adsorbphosphorus is attributed to competition with organic ligands. However, suchan effect was not observed by Phillips (2002), who used both wastewaterand an inorganic solution with the same phosphorus concentration. Variousmodels are currently available that can be used to predict the capacity ofsoils for phosphorus retention. A detailed description of these models, aswell as specific information on their advantages and disadvantages, can befound in McGechan (2002).

B. Nitrogen Sorption—Nitrification

Sorption can also contribute to the removal of nitrogen present in the formof ammonia. After the application of dairy effluent to a soil, ammonium ionswere found to retain predominantly in the surface 5 cm of the soil depth


(Zaman et al., 2002). With the progress of time, however, the concentrationof ammonium decreased, but this decrease was accompanied by an increasein the nitrates concentration. Likewise, ammonia in the soil solution wasfound to decrease continuously over a 21-day period, after the application ofpiggery wastewater (Phillips, 2002). Because the adsorption of ammonia isconsidered to be instantaneous, the decline in the concentration of ammoniawas mainly attributed to its conversion to nitrates. In fact, an increase in thenitrate concentration in the soil solution was noted in the study. The ratesof nitrification measured ranged from 2.0 to 3.6 mg/kg · d, higher than theaverage values reported in the literature. Sparling et al. (2001) also found agreater nitrification potential in soils receiving dairy effluent (3.03 to 4.21 µgN/cm3 · h) than in nontreated soils (0.09 to 0.28 µg N/cm3 · h). Furthermore,Tzanakakis et al. (2003) observed a rapid transformation of ammonia to ni-trate immediately after the application of wastewater to the soil. It is possiblethat the hydrological conditions (unsaturated flow) that ensure increased oxy-gen availability, the high soil moisture, and the availability of organic carbonare the main factors responsible for the fast nitrification reactions in landtreatment systems.

Nitrates have a limited chance for sorption to negatively charged claysurfaces due to their negative charge. However, the presence of positivelycharged metal oxides in the soil favors the sorption of nitrates. Sparling et al.(2001) reported an increase in the ability of a sandy loam soil to adsorbnitrates when it was amended with volcanic ash. The presence of Fe ox-ides and Al oxides, which are characterized by variable charge at the edgesof ferro-/aluminosilicate material, favored the sorption of nitrates (McLarenand Cameron, 1996). However, their sorption was found to be a completelyreversible reaction, leading to the conclusion that plant uptake and denitrifi-cation are the principal pathways for nitrogen removal in SRS.

C. Gaseous Losses of Nitrogen

Considerable amounts, up to 50% of the applied nitrogen in SRS, can be lostas gas emissions via ammonia volatilization and via the losses of nitrogengas and nitrous oxide through denitrification. The amount of nitrogen emit-ted to the atmosphere when applying wastewater to land depends on thewastewater characteristics (concentration and nitrogen form, redox potential,pH), carbon availability, environmental conditions, soil texture and proper-ties, and effluent application methods. Although gaseous losses improve thepotential of SRS for nitrogen removal, emissions of nitrous oxide and ammo-nia contribute to global warming, the destruction of the ozone layer (whenconverted to nitric oxide), and ecosystem eutrophication.

1. VOLATILIZATION

The volatilization of ammonia contributes to nitrogen removal during landapplication of wastewater only in the case when the nitrogen is present


mainly in the form of ammonia and the wastewater effluent and soil is char-acterized by high pH values. Volatilization of ammonia was estimated to beonly a few percent of the loaded nitrogen when municipal wastewater wasapplied to land (Smith et al., 1996). However, the amount of nitrogen lostthrough this pathway reached up to 35% of the ammonia when swine efflu-ent was applied to the land via sprinklers (Sharpe and Harper, 2002); 12%was lost during the application event, while the remainder was lost from thesoil surface within 48 h of effluent application. It was also found that theamount of ammonia volatilized was positively correlated with wind speedand temperature. Similar levels of ammonia volatilization before the effluentreaches the soil surface during application events have also been reportedin earlier studies, but the amount of ammonia removed from the soil sur-face was found to be quite variable (10 to 70%) (Sharpe and Harper, 1997;Schilke-Gartley and Sims, 1993). Al-Kaisi and Waskom (2002) found that theamount of ammonia volatilized, occurring during and after the sprinkler ap-plication of swine effluent, ranged from 32 to 83% of the ammonium in theeffluent, with an average of 58%. In this study, ammonium losses from thesoil ranged from 24 to 56% of the applied ammonium. Climatic conditionswere also found to have a strong effect on ammonium losses. During coldweather, sprinkler application losses were estimated at 10% and soil losses atan additional 25% of the applied ammonium. It should be stressed, however,that the mass balance method for nitrogen estimation used in this study mayhave resulted in an overestimation of the amounts of ammonia lost throughvolatilization, since it doesn’t take plant uptake and denitrification losses intoconsideration.

No estimates are available for ammonia losses occurring when wastew-ater effluent is applied through localized application methods. It can be as-sumed that they are reduced proportionally to the wetted soil surface. Fur-thermore, a lack of information exists about ammonia volatilization in SRSplanted with trees. The canopy of trees may exert a greater effect on ammo-nia volatilization rate than forage crops because of its effects on wind speed,irradiation, and temperature.

2. DENITRIFICATION

The removal of nitrates by denitrification process from land treatment systemsis considered a beneficial process that prevents the pollution of ground andsurface waters. Denitrification has been estimated to constitute approximately25% of the removal of nitrogen applied in SRS (U.S. EPA, 1981). However,wide variations (between 2 and 239 kg N/ha · yr) have been reported indenitrification rate when applying wastewater to the land (Lowrance et al.,1998; Smith and Bond, 1999; Meding et al., 2001). Differences in soil texture,climatic conditions, pH, nitrate concentration, and carbon availability aremost likely responsible for this variation. Soil texture influences denitrification


rate through its effects on soil moisture content and the availability of oxygen.The low denitrification rates observed in a Pinus radiata plantation wereattributed to the texture of the soil (sandy loam), which favored drainageand soil aeration (Smith and Bond, 1999). In contrast, a denitrification rateof 77 kg N/ha · yr was measured in a clay soil receiving wastewater effluentand planted with Eucalyptus spp. (Hooda et al., 2003).

Denitrification rate exhibits great temporal and spatial variations withina given site, making the accurate estimation of nitrogen losses difficult. Sig-nificant differences observed in the denitrification rate on the hilltop, themidslope, and the riparian zone of a forested land treatment system werefound to be poorly correlated with soil physical properties (Meding et al.,2001). Nitrate availability in the soil has been correlated positively with thedenitrification rate (Barton et al., 1999), but Hooda et al. (2003) found thatsuch a relationship can be established only on the first day after effluentapplication. The limiting effects of soil moisture or oxygen availability weremost likely responsible for the absence of a strong relationship during the fol-lowing days. In general, the denitrification rate increases with the availabilityof organic carbon. However, there are no threshold values for organic car-bon to sustain or increase the denitrification process. The lack of a thresholdvalue may arise from the dual effects of carbon on denitrification (electrondonor and effects on soil oxygen status) and the techniques used to assesssoil organic carbon (Barton et al., 1999). Typically, a carbon to nitrogen ratiohigher than 3:1 in wastewater is considered adequate to sustain denitrification(Reed et al., 1995). Such a ratio can be found in most municipal wastewatersthat have received primary treatment. However, organic carbon may be alimiting factor for denitrification in the case of effluents having received asecondary or higher treatment, before their application to the land.

All these factors have meant that the models that are currently avail-able to predict denitrification rates show great variations (50 to 150%) inthe measured rates in agricultural soils (Barton et al., 1999). Furthermore,denitrification rates may change over the long term due to alterations in thesoil physical properties, nitrogen availability, and the accumulation of po-tential inhibitors, such as heavy metals or salts. Thus, additional research isrequired before it is possible to predict accurately the losses of N throughdenitrification.

D. Plant Uptake

Uptake of nutrients by plants and their sequestration in hypergeous biomassconstitute an important pathway for nutrient removal in SRS when biomassharvesting is implemented. The capacity of vegetation to remove nutrientsvaries greatly among plant species or genotypes, and from one area to an-other, indicating that genetic factors, environmental conditions, and culturalpractices can strongly influence the removal of nutrients.


Initially, most SRS were implemented in existing forests (Crites andFehrmann, 1980). However, as tree harvesting was not practiced in thosesystems, nutrient removal by forest trees can be extremely low because oflow growth rates and the return of nutrients into soils as plant litter. Partic-ularly in mature forests, when stand biomass remains nearly stable, nutrientremoval through uptake is negligible and could probably explain the failureof natural forests in treating applied wastewater in terms of nutrient removal.In contrast, the use of forage vegetation or trees regularly harvested (short-rotation coppice) improves substantially the potential of SRS for nutrient re-moval. The amount of nutrients that can be removed is a function of theamount of produced biomass and nutrient content of the plant tissues. Thus,vegetation with a high capacity for biomass production and/or managementpractices that induce biomass production and the uptake of nutrients wouldresult in greater rates of nutrient removal.

1. EFFECT OF PLANT SPECIES

The uptake rate of nutrients and their accumulation in aboveground biomassmay differ greatly among plant species and can often compensate for differ-ences in biomass production. Woodard et al. (2002) evaluating the capacityfor N removal of a bermuda grass–rye and a corn–forage sorghum–rye sys-tem for four consecutive growing seasons, found a clear superiority of theformer for nitrogen removal, which was attributed to higher nitrogen concen-tration in bermuda grass (18.1–24.2 g kg−1) than in corn and forage sorghum(10.3–14.7 g kg−1). Significant differences in the nutrient content of roots,leaves, and stems among annual ryegrass and different species of cereals andlegumes receiving poultry manure were reported by Pederson et al. (2002).In contrast, the absence of differences in plant tissue content of nitrogen andphosphorus and in biomass production of various grasses (reed canary grass,meadow foxtail, and smooth brome grass) receiving wastewater resulted insimilar removal of nutrients (Geber, 2000).

Distribution of biomass among different plant organs can also mediatethe amounts of removed nutrients. Annual tissues, in particular leaves, exhibita greater content of nutrients than old wood. Thus, species that distribute ahigher proportion of their total biomass to leaves will have a higher poten-tial for nutrient removal. Indeed, trees of Eucalyptus botryoides displayed agreater removal of nitrogen and phosphorous than Eucalyptus ovata, despitetheir lower production of biomass. This effect was due to the larger distri-bution of produced biomass to leaves (Guo et al., 2002). Estimates of thecapacity of different plant species or clones receiving different managementpractices to remove nutrients are shown in Table 7.

2. EFFECT OF CULTURAL PRACTICES

Despite the inherent ability of plant species to remove nutrients, as de-termined by the potential for biomass production and distribution, tissue

TAB

LE7

.Pote

ntia

lofB

iom

ass

Pro

duct

ion

and

Nutrie

nt(N

itroge

nan

dPhosp

horu

s)Rem

ova

lofPla

ntSp

ecie

sor

Cro

ppin

gSy

stem

s

Age

of

Bio

mas

sN

rem

ova

lP

rem

ova

lPla

ntsp

ecie

sor

crop

crop

(t/h

a)(k

g/ha)

(kg/

ha)

Ref

eren

ce

Aca

cia

cya

nop

hyl

ous

213

.9—

—Tza

nak

akis

etal

.(2

003)

Eu

caly

ptu

sbo

tryo

ides

(South

ern

mah

oga

ny)

339

.742

542

Guo

etal

.(2

002)

Eu

caly

ptu

sgl

obu

lus(T

asm

ania

nblu

egum

)3

64.7

to80

651

55G

uo

etal

.(2

002)

Dunca

net

al.(1

998)

Eu

caly

ptu

sov

ata

(sw

amp

gum

)3

45.5

401

37G

uo

etal

.(2

002)

Eu

caly

ptu

sca

ma

ldu

len

sis

(red

gum

)2

4.5

——

Tza

nak

akis

etal

.(2

003)

Eu

caly

ptu

sgr

an

dis

(rose

gum

)3

36.1

——

Dunca

net

al.(1

998)

Pop

ulu

ssp

.(h

ybrid

popla

r)4

44to

111

241

to42

041

to10

5H

eilm

anan

dN

orb

y(1

998)

Pop

ulu

ssp

.(h

ybrid

popla

r)3

4.82

to8.

0811

0.4

18M

offat

etal

.(2

001)

Pop

ulu

str

ich

oca

rpa

(bla

ckco

ttonw

ood)

32.

19to

5.45

a72

.911

.2M

offat

etal

.(2

001)

Sali

xSp

.(v

ario

us

will

ow

clones

)1

15to

2275

to86

10to

11A

deg

bid

iet

al.(2

001)

Sali

xvi

nim

ali

s(h

ybrid

will

ow

)La

bre

cque

and

Teodore

scu

(200

3)1s

tRota

tion

335

.08

——

2nd

Rota

tion

358

.80

389

46.1

Cyn

odon

sp.(B

erm

uda

gras

s)—

6.7

to13

129

to30

212

to34

Adel

iet

al.(2

003)

Ph

ala

ris

aru

nd

ina

cea

(Can

ary

gras

s)—

9to

1110

2to

202

27to

31G

eber

(200

0)

Mis

can

thu

ssp

.—

7.2

to10

.440

.3to

71.7

Jørg

ense

n(1

997)

Pa

nic

um

virg

atu

m(s

witc

hgr

ass)

—3

to17

56to

189

19to

42Sa

nder

son

etal

.(2

001)

Aru

nd

od

ona

x(r

eeds)

—8.

13—

—Tza

nak

akis

etal

.(2

003)

Ber

muda

gras

s,ry

esy

stem

—16

.1to

21.1

418

to60

3—

Woodar

det

al.(2

002)

Corn

,so

rgum

,ry

esy

stem

—25

.5to

26.9

328

to35

6—

Woodar

det

al.(2

002)

aBio

mas

sas

sess

edaf

ter

leaf

fall.

220


nutrient content, and environmental factors, management practices may alsohave a strong effect on nutrient removal. Such practices include plantingdensity, length of harvesting cycle, and the use of multiple cropping systems.

Planting density in a short-rotation coppice may vary from about 1000to more than 100,000 trees per hectare (Duncan et al., 1998; Kopp et al.,1996). Short harvesting cycles should be adopted for energy production plan-tations, while longer cycles (6 and 12 years) should be adopted for woodproduction. In general, biomass would increase with tree density, but con-flicting data have been reported about the optimum plant density. Increasingplanting density in a Eucalyptus spp. plantation from 1333 trees/ha to 2667trees/ha resulted in 1.4-fold more biomass production (Duncan et al., 1998).In a stand where different clones of Salix spp. were investigated, the in-crease of planting density from 15,000 to 107,600 trees/ha was found to havedifferential effects on biomass production (Adegbidi et al., 2001). For Salixalba (SA22 clone) biomass production increased, whereas for Salix dasycla-dos (SV1 clone) the biomass production remained stable. In another study,spacings of 0.30 × 0.30, 0.30 × 0.90, and 0.60 × 1.1 m were not found to af-fect biomass production of Salix dasyclados trees (Kopp et al., 1997). Higherproduction of biomass was reported at plant spacings of 0.15 × 0.15 m whenthe harvesting was performed annually, but at larger rotation cycles, spacingsof 0.46 × 0.46 m were more productive (Kopp et al., 1996). However, poplartrees planted at 0.23 × 0.23 m spacing were twice as productive as treesplanted at 0.61 × 0.61 m spacing when harvested at 3-year cycles (Dawsonet al., 1976). Differences among plant species or cultivars in root density anddistribution, juvenile growth, the potential for assimilates production, andability to cope with severe competition for nutrients and water may explainthese complex interactions among plant species, density, and the length ofharvesting cycle. It can be concluded that dense plantings (<0.60 × 0.60 m)stimulate biomass production only at very short rotations (2–4 years) be-cause denser canopies are developed earlier; thereafter, the severe compe-tition among trees and the shading effects counteract this effect and mayeventually result in a decline in aboveground biomass and hence in nutrientremoval. Furthermore, dense spacings should be adopted in short-rotationcoppice only when the increase in biomass production compensates for thecosts of the additional trees needed for planting and the costs associated withshorter rotation cycles.

The optimum length of the harvesting cycle can vary substantially, de-pending on species-/clone-specific growth characteristics and planting den-sity. Because the goal in most SRS is to maximize nutrient removal andbiomass production, harvesting should be accomplished when the produc-tivity reaches its peak. Peak productivity generally occurs when the mean an-nual increment (MAI), the ratio of total biomass to age of plantation, reachesits highest values (Kauter et al., 2003). The time required for balsam poplarsto reach the maximum MAI varies from 4 to 10 years, depending on the clone


used (Schirmer, 1996), but for aspen it generally exceeds 12 years (Kauteret al., 2003). Kopp et al. (1997) found that mean annual productivity at spac-ings of 0.30 × 0.30, 0.30 × 0.90, and 0.60 × 1.1 m was considerably higherwhen triennial rotation cycles were applied than with biennial or annualharvesting, irrespective of stand density. Schoenborn and Duncan (2001)reported significant interactions between plant species and rotation lengthin an effluent-irrigated Eucalyptus spp. plantation. For Eucalyptus grandis,biomass accumulation in the 3 + 3-year rotation was similar to that for the6-year, but for Eucalyptus globulus it was significantly reduced in the 3 +3-year rotation compared to that of 6 years. In general, nutrient uptake fromthe soil declines sharply after canopy closure is achieved, due to the fact thatthe increase in biomass is attributed mainly to wood rather than leaves, inwhich is accumulated the major portion of the nutrients.

Nutrients availability also affects their uptake and the accumulation inplant tissues. However, the increase in tissue nutrient content and biomassproduction observed under conditions of high nutrient availability cannot ac-commodate the increased application of nutrients. Doubling the applicationof wastewater resulted in the removal of 103 and 161 kg N/ha, respectively,in 1995, and 123 and 201 kg N/ha in 1996; the total amount of nitrogen ap-plied was 150 and 300 kg in both seasons (Geber, 2000). Nitrogen removal ina Bermuda grass–rye system receiving wastewater was estimated to be 465,528, and 585 kg/ha when the applied loads of nitrogen were 500, 600, and900 kg N/ha, respectively, compared to 320, 327, and 378 kg N/ha for thecorn–forage sorghum–rye system (Woodard et al., 2002). Adeli et al. (2003)investigated the effects on nutrient removal in Bermuda grass of applicationrate and timing of swine effluent. Although increasing the application rateof the swine effluent above 10 cm/ha had little effect on biomass produc-tion, nitrogen and phosphorus tissue content continued to increase. Despitethe increase, it was not possible to compensate for the increased loads ofnutrients. At the lower application rate (5 cm/ha) the recovery of nitrogenand phosphorus was 64 and 40%, respectively, while at the highest rate (20cm/ha), the corresponding recovery was 40 and 18%, respectively. Applica-tion time of effluent was also found to influence both biomass production andnutrient removal. Application of an additional amount of effluent (2.5 cm)to the treatment that had received 10 cm of effluent on October 1 resultedin reduced yield and nutrient uptake than for the treatment that receivedeffluent only on September 1. These findings suggest that coupling of efflu-ent applications with the active period of growth will substantially eliminateenvironmental risks.

VIII. PATHOGEN REMOVAL

The elimination of pathogenic organisms is one of the most critical is-sues with land application of wastewater. Most studies investigating the


fate and transport of pathogens in land-based systems focus on infiltration–percolation studies (under both field and laboratory conditions), because ofthe elevated risks for groundwater contamination. However, differences inhydraulic conditions between infiltration–percolation studies and Type 1 SRSmay substantially affect the physical, chemical, and biological processes oc-curring in the soil matrix and determine pathogens’ removal. The risk forgroundwater contamination is lower when wastewater is applied at rates tosatisfy crop water requirements, as the wetted volume of the soil is restrictedto the active root zone, and unsaturated flow conditions prevail. However,overapplication of the effluent, need for salt leaching, and rainfall events, inconjunction with topography, soil texture, and aquifer depth, may also resultin pathogen percolation to groundwater or in runoff of applied effluent andpathogen transport to surface water bodies. Moreover, the survival and re-growth of pathogens in the soil remains a major concern when wastewatereffluent is applied to the land, to prevent contamination of grazing livestock,wildlife, or workers/consumers.

A. Physicochemical Processes

The contribution of physicochemical processes, such as filtration and ad-sorption to soil particles, in pathogen removal is mainly determined by soilproperties, effluent composition (pH, ionic strength), hydraulic conditions,and the applied management practices.

1. EFFECT OF SOIL TYPE

Soil texture and properties may strongly affect the efficiency of land-basedwastewater treatment systems to remove pathogens. Coarse-textured soilsfavor microbial movement downward (Huysman and Verstraete, 1993). Incontrast, soils forming small pores, like those with a high proportion of clay,are considered to be the most effective in pathogen purification as they re-strict pathogen leaching, in particular microorganisms of large size (bacteria,helminthes, and protozoa). Mawdsley et al. (1996) found that the majority ofoocysts (73%) applied to the soil was retained in the top 2 cm, with num-bers decreasing with depth. By increasing soil bulk density from 0.75 to1.15 g/cm3, recovery of E. coli O157:H7 decreased from approximately 90%of the applied cells to 0.1% (Artz et al., 2005).

Soils with a high proportion of clay promote the attachment ofpathogens, in particular bacteria and viruses, due to the greater surface areaand hence the larger number of available sorption sites compared to sandyor silty soils. Quanrud et al. (2003) found greater removal of coliphage insandy-loam columns (93%) than in a coarse-sand soil (73%). Because flowrates did not differ between two soil types, differences in grain size, specificsurface area, and cation exchange capacity were considered responsible forthe differences observed in virus removal. Similarly, in an earlier study, the


removal of reovirus was positively correlated to soil surface area (Mooreet al., 1982). Tzanakakis et al. (2003) did not find leaching of fecal coliformsbelow a depth of 15 cm during land application of wastewater. The increasedclay content (35.5%) of that soil probably promoted bacterial cell retention.However, in another study the type of medium (soil versus sand columns)did not affect bacteria and bacteriophage removal (van Guyk et al., 2001).

Another important factor that affects the extent to which pathogens aresorbed on soil particles is pH. In general, sorption of pathogens on soilmatrix increases with decreasing pH (Goyal and Gerba, 1979; Frankenbergeret al., 1988). Enhanced sorption occurs because most bacteria and viruseshave isoelectric points below 7, and thus they are negatively charged atthe pH ranges usually encountered in the field (Harden and Harris, 1953;Vilker, 1981). Decreases in pH result in less negatively charged pathogensand soil particles reducing the electrostatic repulsion between them favoringthe sorption of pathogens. The presence of oxides in the soil significantlyimproves its capacity for virus sorption and inactivation due to their positivelycharged binding sites (Chu et al., 2000, 2001).

With regard to protozoans, Cryptosporidium parvum removal was foundto increase with decreasing grain size in intermittent unsaturated sand filters(Logan et al., 2001). However, the authors did not attribute the higher effec-tiveness to the smaller sized filter media, since the size of pores was adequatefor oocysts leaching, suggesting that additional mechanisms must contributein their removal. Oocyst attachment to soil particles in general is not con-sidered as a mechanism of removal, since they are negatively charged at pHranges from 5.0 to 8.5 (Dai and Boll, 2003; Drozd and Schwartzbrod, 1996).However, Medema et al. (1998) reported an apparent attachment of oocyst toparticles of secondary treated effluent. The particles found in the effluents aremainly of organic origin and may have different sorption characteristics com-pared to negatively charged soil particles. Thus, organic matter contained inwastewater effluent may act as an absorption medium contributing to oocystremoval. Moreover, the development of a biofilm, which surrounds soil par-ticles under wastewater application, may have a positive effect on oocystsremoval by reducing medium porosity and/or by favoring their absorptionon organic matter.

Despite the higher inherent capacity of clayey soils to remove pathogens,increased earthworm populations and activity, factors often observed in suchsoils, and/or surface cracking may induce conditions of preferential flow andthus enhance the downward movement of pathogens. Greater numbers offecal coliforms, fecal enterococci, Escherichia coli, and bacteriophages wereleached from clayey soils than soils with a higher proportion of sand; thiswas attributed to preferential flow (Aislabie et al., 2001; McLeod et al., 2001).Artz et al. (2005) reported that relatively slight changes in the soil struc-ture of intact cores were associated with large variations in the percentageof E. coli O157:H7 leached (0.01% to 24%), suggesting that factors such as


compaction and the prevalence of macropores can strongly affect the trans-port of pathogenic bacteria downward. Likewise, Cryptosporidium parvumoocysts, applied to the surface of soil cores, were detected in the leachatefrom a clay-loam and a silty-loam soil as a consequence of bypass flow, butnot in that from a loamy-sand soil, following irrigation over a 21-day pe-riod (Mawdsley et al., 1996). Pathogen removal in various porous media andexperimental conditions (field and column studies) is shown in Table 8.

The adoption of management practices such as tillage and appropriateirrigation methods may prevent or at least limit preferential flow. Tillagedecreases the bulk density of the soil, causes the existing macropores andcrackings to disappear, and reduces earthworm activity. In cracking clayeysoils, low-rate effluent application methods such as localized systems thatensure the prevalence of unsaturated flow conditions will prevent effluentflow through soil cracks. In addition, increasing the frequency of effluentapplication will also contribute in reduced occurrence of soil cracking. Incoarse soils, however, in which fingered flow can be observed, the adoptionof high-rate effluent application systems is suggested to prevent conditionsfavoring the fingered flow.

2. EFFECT OF IONIC STRENGTH

The ionic strength of the applied effluent exerts a significant effect onpathogen sorption on the soil colloids. Pathogen sorption on the surfaceof negatively charged soil particles has been found to increase with the ionicconcentration of the applied solution due to the reduction of the electro-static repulsion between the negatively charged soil particles and pathogens(Jewett et al., 1995; Johnson et al., 1996; Redman et al., 1999; Chu et al., 2000).The extent to which ionic strength affects the sorption of pathogens on soilcolloids appears to be a function of the surface properties of the pathogen(Wan et al., 1994; Redman et al., 1999). Moreover, the type of cations also af-fects the potential of pathogens for sorption (Lance and Gerba, 1984; Zhuangand Jin, 2003). In general, divalent cations are more efficient than monovalentat forming pathogen-cation-clay bridges. Sorption of SJC3 bacteriophage wasstimulated when the solution ionic strength was increased by using divalentcations (Redman et al., 1999). In contrast, when pathogens and sorption sur-faces are oppositely charged, increases in the ionic strength of the appliedsolution may result in decreased sorption. Increasing the ionic strength ofsolution from 0.002 to 0.16 M resulted in reduced sorption of viruses in sandcoated with Al2O3, an effect attributed to reduced electrostatic attraction be-tween the oppositely charged viruses and Al oxide-coated surfaces (Zhuangand Jin, 2003). In general the high ionic strength of various wastewater efflu-ents favors the retention of pathogens during their passage through the soilprofile. However, an important issue that arises when effluent is applied toland is the possible mobilization of pathogens and their transport downwardafter rainfall events or when changing the effluent supply with fresh water.

TAB

LE8

.Rem

ova

lofPat

hoge

ns

Thro

ugh

Var

ious

Poro

us

Med

iaan

dExp

erim

enta

lConditi

ons

Syst

emdes

crip

tion

Typ

eofpat

hoge

ns

Sam

plin

gdep

th(c

m)

Rem

ova

l(%

)Ref

eren

ce

Septic

tank

effluen

tap

plie

dto

sand

lysi

met

ers

Feca

lco

liform

s96

–100

van

Guyk

etal

.(2

001)

INA

a

Applic

atio

nra

te:5

and

8cm

/dM

S-2

60an

d90

100

PRD

-1>

99>

95Effl

uen

tap

plic

atio

nin

sim

ula

ted

soil

aquifer

trea

tmen

tsy

stem

:Colip

hag

e10

0Q

uan

rud

etal

.(2

003)

Riv

ersa

nd

76Sa

ndy

loam

93A

rtifi

cial

was

tew

ater

applie

din

inte

rmitt

entfilte

rsco

nsi

sted

of:

Loga

net

al.(2

001)

100

Fine

sand

C.pa

rvu

m65

99.9

Coar

sesa

nd

Applic

atio

nra

te:4

cm/d

Soil

colu

mns

sandy

loam

:A

rtz

etal

.(2

005)

Inta

ct(h

igh

leac

hin

gra

tes)

80–9

9.9c

Inta

ct(low

leac

hin

gra

tes)

E.co

liO

157:

H7

4599

.6–9

9.9c

Rep

acke

db

20–9

9.6

Soil

core

sre

ceiv

ed5.

5×

107

cfu

inan

auto

clav

edca

ttle

slurr

yso

lutio

nan

dw

ere

leac

hed

by

sim

ula

ted

rain

(14.

2m

m/d

ay)

Applic

atio

nofdai

ryef

fluen

tin

various

soil

colu

mns:

Ais

labie

etal

.(2

001)

Cla

yey

Feca

lco

liform

s<

85c

Silty

-loam

yE.co

li70

–75

<99

.5Sa

ndy

Feca

len

tero

cocc

i10

0A

pplic

atio

nra

te:50

mm

follo

wed

by

sim

ula

ted

rain

fall

until

1PV

of

leac

hat

ew

asco

llect

ed.

226

Applic

atio

nofdai

ryef

fluen

tin

various

soil

colu

mns:

McL

eod

etal

.(2

001)

Cla

yey

(coar

sest

ruct

ure

)62

–94c

Sandy

(sin

gle

grai

nst

ruct

ure

)B

acte

riophag

e91

–99c

Silty

(fine

text

ure

)S.

typh

imu

riu

m70

100

Coar

sesa

nd

(fine

tosi

ngl

egr

ain

stru

cture

)99

.85

Applic

atio

nra

te:30

mm

ofw

ater

conta

inin

gth

ebac

teriophag

eat

ara

teof5

mm

/hfo

llow

edby

up

toab

out1.

8PV

ofsi

mula

ted

rain

fall.

Land

applic

atio

nto

asa

ndy

loam

soil

�X

174

6010

0O

akle

yet

al.(1

999)

Septic

tank

effluen

tA

pplic

atio

nra

te:0.

81–6

.5cm

/day

Land

applic

atio

nofse

ptic

tank

effluen

tM

S-2

6099

.9va

nG

uyk

etal

.(2

004)

Applic

atio

nra

te:0.

5–2.

7cm

/dPRD

-1La

nd

applic

atio

nofm

unic

ipal

was

tew

ater

under

fiel

dco

nditi

ons

Feca

lco

liform

s15

100

Tza

nak

akis

etal

.(2

003)

Tota

lco

liform

sA

pplic

atio

nra

te:Cro

pET

Land

applic

atio

nofm

unic

ipal

was

tew

ater

under

fiel

dco

nditi

onsd

Feca

lco

liform

s15

010

0Ree

det

al.(1

995)

Land

applic

atio

nofm

unic

ipal

was

tew

ater

under

fiel

dco

nditi

onsd

Feca

lco

liform

s8

100

Bel

lan

dB

ole

(197

8)

aIc

enucl

eotid

ebac

teria.

bD

epen

din

gon

the

bulk

den

sity

ofth

eso

ilco

lum

n.

cPre

fere

ntia

lflow

has

bee

ndocu

men

ted.

dCite

din

Ree

det

al.(1

995)

.

227


Redman et al. (1999) found that a change in the electrolyte composition ofpore solution from 10 mM NaCl to 10 mM CaCl2, which corresponds toa change in water hardness from 100 mg/L as CaCO3 to approximately 0mg/L, can mobilize viruses sorbed in quartz sand columns. Remobilizationof viruses was also observed in soil columns receiving artificial effluent aftersimulated rainfall (Quanrud et al., 2003). From a practical point of view itis therefore important to take into account the potential effects of rainfall inpathogens transport during land application of wastewater. Interruption ofeffluent application when rainfall is expected to occur or selection of soilswith a high ability to release ions in the soil solution will significantly reducethe potential risks.

3. EFFECT OF HYDRAULIC CONDITIONS

Hydraulic conditions may strongly influence the fate and transport of thepathogenic organisms present in wastewater effluent. Unsaturated flow re-sults in less transport of pathogen downward than does saturated flow(Gerba, 1984; Powelson and Mills, 2001). Unsaturated flow has been sug-gested to promote pathogen removal by increasing the contact time ofapplied wastewater with the solid phase and/or the surrounding biofilm,hence favoring the sorption of pathogens and their biological degradation.Moreover, air–water interfaces developed in unsaturated porous media havebeen reported to act as important sorption sites (Mills and Powelson, 1997;Thompson et al., 1999; Powelson and Mills, 2001). Decreasing the saturationof quartz columns from 100 to 46% was found to nearly double the reten-tion of bacterial cells (Jewett et al., 1999). Greater than 50% of the appliedbacterial cells were transported under saturated conditions, while the cor-responding fraction fell to 5% in columns saturated at 46%. The increasedefficiency of unsaturated quartz columns to remove bacteria was mainly at-tributed to the increase of air–water interfaces, as the bacterial numbers at-tached in the solid–water interface were found to be similar. Hydrophobicbacterial strains show a higher removal potential compared to hydrophilicstrains under conditions of unsaturated flow. Wan et al. (1994) reported thatthe relative recovery of bacterial cells from 30-cm columns saturated at 44%compared to saturated columns was 0.77 and 0.24 for the hydrophilic andthe hydrophobic strain, respectively. Similarly, the removal of both MS2 and�X174 increased under unsaturated flow conditions (Jin et al., 2000). How-ever, the extent of MS2 removal was greater than that of �X174. Mass balancestudies revealed that increased removal of �X174 was due to increased sorp-tion, whereas increased removal of MS2 was probably caused by inactivation.These findings suggest that the extent to which viruses are susceptible toair–water interfaces differs among various types of viruses, an influence thatprobably results from differences in structural characteristics of the capsidprotein.


B. Microbial Activity

The soil microbial community may exert a strong effect on pathogen inac-tivation in land-based wastewater treatment systems. Microbial action hasbeen attributed mainly to aerobic microorganisms, which have been foundto increase virus inactivation in the soil up to threefold, while anaerobicmicroorganisms were not found to have any contribution to virus removal(Hurst et al., 1988). A recent study, however, revealed that soil microbialactivity may have a significant influence on virus inactivation under satu-rated conditions, either by releasing extracellular enzymes or by promotingthe irreversible sorption of viruses on soil particles (Nasser et al., 2002). Theeffect of microbial activity increased with temperature and was found to bevirus dependent. Relatively low concentrations of protease pronase resultedin 90% inactivation of Cox-A9 virus, but did not affect poliovirus type 1, hep-atitis A, and MS2. The exposure of Cox-A9 and hepatitis A to P. aeruginosaextracellular enzymes resulted in a 99% inactivation in contrast to the otherinsensitive types (Nasser et al., 2002).

Lipson and Stotsky (1985) found that soil aerobic bacteria contributedto virus degradation by excreting substances that degrade virions and byutilizing viruses as growth substrate. Application of azide, an inhibitor ofaerobic respiration, reduced the cumulative removal of coliphage from ap-proximately 70 to 30% (Quanrud et al., 2003). Similarly, the survival time ofenteric bacteria was found to be longer in sterilized than in nonsterile soils(Gerba et al., 1975). Predation by protozoa and other bacteria has been alsosuggested to contribute in pathogens die-off (Frankenberger, 1988). Bomoet al. (2004) found that bacterial numbers increased in the effluent leachedfrom sand columns when columns were treated with the protozoan inhibitorcycloheximide. The increases in the number of protozoan species (Acantho-moeba spp. and Naegleria spp.) found in a soil receiving wastewater effluent(Gupta et al., 1998) may also provide evidence of the significance of proto-zoa on pathogen removal. The importance of microbial activity, which occursthrough biofilm formation, on wastewater purification is also outlined by vanGuyk et al. (2001). During the early operation of a laboratory-established infil-tration system, the number of fecal coliforms in leachate was 200 cfu/100 ml,but the number had decreased to 10 cfu/100 ml by week 20 and reachednondetectable levels at week 48 of operation.

C. Vegetation Effects

Root systems of aquatic plants (e.g., bulrush) have been reported to releasesubstances with antibiotic activity (Sundaravadivel and Vigneswaran, 2001).There is no available information on whether plant species used in SRS exudesuch substances and whether these substances have a significant influence onpathogen inactivation. It is well known that vegetation excretes a wide rangeof organic substances, including organic acids and enzymes, from the roots.


These substances can modify chemical conditions close to plant roots, thusindirectly affecting the fate of pathogen populations. Root exudates fromparticular plant species contain certain groups of substances, such as phe-nolic acids or isothiocyanates, that have biocidal effects on soil pathogens.For example, canola plants release glucosinolates, which can be hydrolyzedinto various substances, including isothiocyanates, in the presence of the en-zyme myrosinase (Rumberger and Marschner, 2003). These hydrolyzed com-pounds are toxic to a variety of soil pests and parasites. In a more specificstudy, Ocinum basilium plants displayed an ability to produce rosmarinicacid, which, when exposed to elicitors from Phytopthora cinnamoni, hadan antimicrobial effect on Pseudomonas aeruginosa, an opportunistic soilbacterium and human pathogen (Bais et al., 2002). More research shouldbe directed toward the investigation of the composition of root exudates ofplant species with potential use for land treatment systems. Plant specieswith extended and dense root systems are ideal candidates for increasingthe efficiency of SR systems to remove pathogens.

D. Factors Affecting Pathogen Survival in the Soil

Despite the outstanding ability of land treatment systems to removepathogens, significant populations may remain viable within the soil matrix.Factors that affect the survival time in the soil include the type of pathogens,soil factors, climatic conditions, and management practices.

1. TYPE OF PATHOGENS

Typically, protozoa and helminthes remain active in the soil for periods ofup to several months (Table 9), due to their ability to form (oo)cysts that areespecially tolerant to the prevailing environmental stresses. In general bac-teria survive for shorter periods of time in soil than viruses do, they usually

TABLE 9. Typical Intervals for Pathogen Survival in the Soil

Pathogens Survival (days)a

VirusesEnteroviruses <100, but typically <20

BacteriaThermotolerant coliforms <120, but typically <50Salmonella spp. <120, but typically <50Shigella spp. <120, but typically <50Vibrio cholerae <120, but typically <50

ProtozoaGiardia spp. Several monthsEntamoeda histolyca cysts <20, but typically <10

HelminthsAscaris lubricoides eggs Several months

aAdapted from Jimenez (2003) and Feachem et al. (1983).


do not exceed 70 days, while viruses can maintain their viability for slightlylonger, up to 100 days (Jimenez, 2003). Significant variations in the survivalperiod can be found in literature not only for different pathogen groups,but also within the same species. These variations are due to differences inenvironmental conditions prevailing from one area to another and to geneticdiversity between microorganisms of the same species. However, it is the dif-ferences among microorganisms of the same group which are of paramountimportance during land application of wastewater. Such differences makeit possible to use specific types of pathogenic microorganisms as conserva-tive indicators to protect public health. Evidence from studies investigatingthe fate of pathogens in the land reveals complex interactions among envi-ronmental conditions (climate, soil properties), management practices, andmicroorganisms, which bring into question the suitability of certain microor-ganisms as indicators under all environmental conditions.

2. CLIMATIC CONDITIONS

Radiation, temperature, and rainfall are considered to be the principal cli-matic variables that affect the survival period of pathogens in the soil. Radia-tion affects pathogen survival through desiccation and ultraviolet (UV) action(Frankenberger, 1988). However, its contribution to pathogen removal dur-ing land application of wastewater effluent is thought to be minor due to theshading effects of vegetation. Furthermore, the use of local irrigation meth-ods instead of flooding or sprinklers further eliminates the positive effects ofirradiation on pathogen inactivation.

Temperature appears to be the principal climatic parameter determiningpathogen survival in the soil. In general, survival rates of pathogens decreasewith increasing temperature. Poliovirus was found to survive in the soil forlonger than 6 months at 4◦C, but when the temperature was increased to37◦C no polioviruses were detected in the soil after 12 days (Yeager andO’Brian, 1979). The relatively high soil temperature (10◦C at 15 cm) that pre-vailed until the end of November was considered to be the principal factorhindering the survival of viruses in a soil receiving primary treated effluent(Oron et al., 1995). Thereafter, a decline in soil temperature was associ-ated with an increase in the number of viruses. Reddy et al. (1981) foundthat the decay of pathogenic bacteria and indicator organisms was nearlydoubled by increasing temperature in steps of 10◦C in the range from 5 to30◦C. In findings reported by Natvig et al. (2002) it was noted that repeatedfreeze-drying cycles were detrimental to the survival of S. enterica serovarTyphimurium and E. coli in manure-amendend soil. John and Rose (2005),reviewing the effects of temperature on pathogens inactivation in groundwa-ter, found that viruses show a temperature dependency with greater inacti-vation at greater temperatures; however, this occurs mainly at temperaturesgreater than 20◦C. In contrast, coliforms die-off in groundwater did not show


temperature dependency as viruses, indicating a complex interplay of in-activation and reproduction subject to influences from native groundwaterorganisms, temperature, and water chemistry. The survival period of proto-zoan oocysts was significantly reduced by increasing the temperature from4 to 30◦C (Fayer et al., 1998; Jenkins et al., 2002). The temperature effect onoocysts was found to be dependent on soil type, with inactivation higher insilty-clay-loam and loamy-sand soils than in a silt-loam one (Jenkins et al.,2002). The mechanisms by which temperature affects pathogen survival inthe soil are not fully understood. It has been suggested that temperature mayinduce damage of RNA or structural changes in virus capsid, or may renderpathogens more susceptible to other abiotic and or biotic factors (Yates andGerba, 1998).

3. SOIL FACTORS

Soil characteristics, such as pH, moisture content, organic matter, nutrientavailability, salts, and trace elements, also affect pathogen survival. Soil watermoisture strongly affects the survival in soil, but its effect varies greatly withthe specific type of pathogens (Mawdlsey et al., 1995). Oron et al. (1995)found minimal soil contamination with viruses when irrigation intervals wereextended. Investigating the survival of E. coli in different soil types, Mubiruet al. (2000) found lower mortality rates in a soil with higher water content,organic matter, and nutrients. Differences in available soil water between thetwo soils were the overriding factors in the survival of E. coli. In contrast, thesurvival of Cryptosporidium parvum oocysts was not affected by soil waterpotentials between −0.033 and −1.5 MPa (Jenkins et al., 2002). A significantinactivation of oocysts was observed, however, at lower soil water potentialvalues (Jenkins et al., 1999), implying that soil moisture availability may havean adverse effect on oocysts survival only when extreme soil water deficitprevails. However, such conditions are unusual in SRS with conventionalirrigation practices.

Organic matter extends the survival of pathogens in the soil (Yates andGerba, 1998). However, the mechanisms responsible for the enhanced sur-vival are not well understood (Jimenez, 2003). The organic matter may serveas a substrate for pathogen growth or it may have a protective effect. Evi-dence for the former is provided by the study of Artz and Killham (2002),where the presence of protozoa in well-water samples was associated withlonger survival of E. coli O157:H7. Furthermore, E. coli strain O157:H7 wasobserved to have a greater ability to survive and multiply in soil amendedwith manure (Gagliardi and Karnss, 2000). In addition, an indirect effect oforganic matter in pathogen survival may come from the increase in soil waterholding capacity and the greater availability of nutrients.

The effects of soil pH on pathogens can be either direct, by influencingtheir survival, or indirect, by altering the biological and physicochemical


conditions of the soil (Mawdsley et al., 1995). It has been reported to favorsurvival by affecting the adherence of viruses to soil particles (Gerba andBitton, 1984). Hurst et al. (1980) reported longer survival times for viruses inthe soil when the pH was decreased from 8.2 to 4.5. Oocysts of C. parvumdisplay a great ability to withstand acidic conditions. Even pH values of 2.75did not result in oocyst inactivation (Campbell et al., 1992). Extreme alkalineor acidic conditions decrease the survival of bacteria in the soil, while theirsurvival is favored in neutral conditions (Frankenberger, 1988).

The availability of certain elements in the soil may exert a profound effecton pathogen survival. Low concentrations of copper in water samples takenfrom wells, even lower than those having adverse effects on humans, werecorrelated with decreased numbers of E. coli (Artz and Killham, 2002). Otherstudies have also reported the adverse effects of the presence of copper andzinc in the soil on bacteria survival (Chaudri et al., 1999; Vulkan et al., 2000).Additions of aluminum, magnesium chloride, and magnesium peroxide werefound to increase the inactivation rate of MS2 virus (Frankenberger, 1988).

IX. LONG-TERM IMPACTS AND SYSTEM MAINTENANCE

Land application of effluent may adversely affect physical and/or chemicalproperties of the soil as well as its biological activities, all of which in turncould impair the basic treatment processes, resulting in reduced treatment ef-ficiency. The most common effects associated with wastewater application tothe soil include changes in pH, salt accumulation, sodicity problems, biolog-ical clogging, and nutrient accumulation. Trace element accumulation mayoccur in some cases, depending on the type of wastewater applied. With anappropriate monitoring schedule for these parameters it is possible to adoptsuitable management practices to maintain the long-term sustainability of thesystem and limit environmental impacts.

A. pH

In general, the application of wastewater to the soil would be expected todecrease soil pH due the oxidation of ammonia to nitrates and the min-eralization of organic matter resulting in CO2 and organic acid production.However, the effects of wastewater on soil pH are quite variable, dependingon effluent composition, application rates, soil chemical attributes, and veg-etation type. Soil pH has been found to increase in some studies (Sparlinget al., 2001; Falkiner and Smith, 1997; Schipper et al., 1996), or to decreasein others (Guo and Sims, 2003; Waly et al., 1987). Even slight reductionsin soil pH may adversely influence the structure and the activity of soil mi-croorganisms involved in carbon and nutrient cycling and impair the capacityof SRS for organic matter removal. Moreover, changes in soil pH may have


significant effects on the processes affecting the fate of pathogens to the soil,but these effects will vary depending on the specific type of pathogen.

B. Salinity

The residential use of fresh water supplies increases the concentration ofsalts in wastewater from 150 to 500 mg/L (Crook, 1998). The concentrationof salts increases even more in agricultural drainage or livestock wastes, inwhich ECw values often exceed 3 dS/m. Such levels of salts may accumulatein the soil and impair the performance of SRS. The buildup of salts has beenreported for effluent with relatively low concentrations of salts under unfa-vorable climatic conditions (arid or semiarid climates) and when appropriatemanagement practices are lacking. For example, the application of wastewa-ter effluent with an ECw of 0.79 dS/m to land caused accumulation of salts inthe soil profile (Falkiner and Smith, 1997). Moreover, Paranychianakis (2001)reported leaf burns in grapevines irrigated with effluent with an ECw valueof 1.8 dS/m.

Salinity could impair the performance of SRS mainly through its effectson nutrient uptake, crop water use, and the activity of the microbial com-munity. Nutrients (potassium and nitrate) uptake is directly inhibited by an-tagonism with sodium, and chloride and sulfate, respectively (Marschner,1995), and indirectly by reduced plant growth. Furthermore, because plantgrowth is more sensitive to salt accumulation than transpiration (Paranychi-anakis et al., 2004a, 2004b), salinity inhibits biomass production and hencenutrient removal to a greater extent than water consumption, increasing theenvironmental risks arising from the increased nutrient loads. Salinity alsoaffects the water consumption by the crops through its effects on leaf areadevelopment and the osmotic effect of salt accumulation in the rootzone.Because most classical methods used to estimate plant water demands donot take into account the effect of salinity on ET, overloading of the SRS mayoccur. It is therefore important that application rates of wastewater be read-justed. Allen et al. (1998) provide methodology for ET estimation of plantsgrown in saline environments. Recently, Theiveyanathan et al. (2004) devel-oped a simple spreadsheet model (WATSKED) that uses a water-budgetingapproach to enable scheduling of irrigation with saline water of eucalyptusplantations. Input data include pan evaporation, rainfall and applied irri-gation, the basic soil physical characteristics, the initial average soil salin-ity, salinity of the irrigation water, and threshold and critical soil salinity. Inaddition to water requirements, the model can be used to predict the tim-ing of leaching and the minimum amount of irrigation water required forleaching salts from soil to acceptable levels, but further model verificationis needed. The leaching requirements are a function of salt concentrationin irrigation water and irrigated crop salt tolerance. Leaching requirementsshould be calculated with the maximum possible accuracy to eliminate the


risk of groundwater contamination. The insurance of unsaturated flow con-ditions will limit the risk of percolation of pathogens and other contaminantsto underlying aquifers.

Other management practices include the selection and use of salt toler-ant plant species or genotypes to alleviate the impacts of salinity on plantgrowth, and application of water-saving methods such as microirrigation sys-tems to decrease salt accumulation in the root zone (Paranychianakis andChartzoulakis, 2005).

C. Sodicity

Increased sodium concentrations found in municipal effluents and otherwastewaters, such as tannery effluents, effluents from hide and skin pro-cessing plants, or meatworks (Menneer et al., 2001; Paranychianakis et al.,2004b), may result in sodium accumulation in the soil profile. Menneer et al.(2001) found an increase in soil exchangeable sodium percentage of up to31% in the first 30 cm of soil depth in soils receiving a mixture of effluentsfrom a rendering plant, hide and skin processing plants, and meatworks, for4 years. In a Eucalyptus spp. plantation at Waga Waga, the exchangeablesodium percentage increased from 0.2 to 25% in the 0.3–0.4 m soil layer after4 years of effluent application (Falkiner and Smith, 1997). Sodium accumu-lation in the soil may cause breakdown of soil aggregates (clay dispersion)and damage to soil structure. Clay dispersion reduces soil hydraulic conduc-tivity and may result in surface ponding of wastewater and runoff, inhibitionof plant growth due to toxicity or to prevailing waterlogged conditions, andodor problems. Therefore the maintenance of soil structure is of paramountimportance.

To predict the impacts of sodium accumulation on soil texture, soil ex-changeable sodium percentage (ESP), and soil adsorption ratio (SAR) can beused. The ESP and SAR can be calculated using Eqs. (1) and (2):

ESP = (100 × exchangeable Na)/∑

(exchangeable Ca2+

+ Mg2+ + Na+ + K+ + Al3+) (1)

where the cation concentrations are expressed in meq/100 g soil, and

SAR = Na+/[(Ca2++ Mg2+)/2]1/2 (2)

where the cation concentrations are expressed in meq/L.Based on previous findings, it can be concluded that there are no fixed

threshold values for both ESP and SAR above which the soil structure will bemodified (Halliwell et al., 2001). This is due to the complex interactions be-tween clay content and type, organic matter, pH, bicarbonates, and electrolyteconcentration, as well as differences in methodologies used for estimating


TABLE 10. Guidelines for Evaluation of Water Quality for Irrigation (Adapted FromAyers and Tanji, 1981)

EC (dS/m)

SAR No problem Slight to moderate problem Severe problem

0 to 3 <0.2 0.9 to 0.2 >0.93 to 6 <0.25 1.3 to 0.25 >1.36 to 12 <0.35 2.0 to 0.35 >2.012 to 20 <0.9 3.1 to 0.9 >3.1>20 <1.8 5.6 to 1.8 >5.6

SAR and ESP (Summer, 1993). It has been reported that ESP values lowerthan 5% can affect soil structure if low electrolyte concentrations are presentin the soil solution (Crescimanno et al., 1995; Summer, 1993). Overall, therisk of sodicity is lower in wastewater than fresh-water-irrigated soils, as ECw

values are, in most cases, higher than the critical value of clay coagulation(see Table 10). However, variations in EC of applied wastewater, stopping ofeffluent application, or winter rainfall may be detrimental for soil texture.

Carbonates and bicarbonates in irrigation waters precipitate calcium andmagnesium cations and therefore increase the impact of sodium on soil hy-draulic conductivity. It has been stated that it is the concentration of availablecations for adsorption onto soil matrix rather than the absolute concentrationof cations in irrigation water that determines ESP and SAR of irrigated soils,and an adjusted SAR must be used to estimate the real effects of sodium onsoil texture:

SAR = Na+/[(Ca2+

x + Mg2+)/2]1/2 (3)

where the concentrations of cations are expressed in meq/L and Ca2+x is

obtained from Table 11.However, the estimation of the effective SAR is more complicated in

wastewater irrigation than in fresh-water supplies (Halliwell et al., 2001).Binding of calcium and magnesium cations on organic substances may in-crease SAR and the potential risks for soil structure problems. For example,Metzger et al. (1983) found that 8 to 26% of the soluble calcium and mag-nesium was bound on organic substances. In addition, organic matter mayhave differential effects on clay dispersion either by protecting soil aggre-gates from dispersion or stimulating it, depending on the form of organicmatter (Nelson et al., 1999). Application of biosolids to a salinized soil im-proved aggregate stability, which was related to an increase in carbohydratesand a decrease in ESP (Garcıa-Orenes et al., 2005). Aggregate stability wasalso found to increase after long term irrigation with dairy factory effluent,an effect attributed to the presence of lactose in the effluents (Cameron et al.,2003).


TABLE 11. Cax Values for the Estimation of Adjusted SAR (Adapted From Victorian EPA, 1993)

Electrical conductivity (ECw) of irrigation water (dS/m)Ratio

HCO3/Ca 0.1 0.2 0.3 0.5 0.7 1.0 1.5 2.0 3.0 4.0 6.0 8.0

0.05 13.2 13.6 13.92 14.4 14.8 15.26 15.91 16.43 17.28 17.97 19.07 19.940.10 8.31 5.57 8.77 9.07 9.31 9.62 10.02 10.35 10.89 11.32 12.01 12.560.15 6.34 6.54 6.69 6.92 7.11 7.34 7.65 7.9 8.31 8.64 9.17 9.580.20 5.24 5.40 5.52 5.71 5.87 6.06 6.31 6.52 6.86 7.13 7.57 7.910.25 4.51 4.65 4.76 4.92 5.06 5.22 5.44 5.62 5.91 6.15 6.52 6.820.30 4.00 4.12 4.21 4.36 4.48 4.62 4.82 4.98 5.24 5.44 5.77 6.040.35 3.61 3.72 3.80 3.94 4.04 4.17 4.35 4.49 4.72 4.91 5.21 5.450.40 3.30 3.40 3.48 3.60 3.70 3.82 3.98 4.11 4.32 4.49 4.77 4.980.45 3.05 3.14 3.22 3.33 3.42 3.53 3.68 3.80 4.00 4.15 4.41 4.610.50 2.84 2.93 3.00 3.10 3.19 3.29 3.43 3.54 3.72 3.87 4.11 4.300.75 2.17 2.24 2.29 2.37 2.43 2.51 2.62 2.70 2.84 2.95 3.14 3.281.00 1.79 1.85 1.89 1.96 2.01 2.09 2.16 2.23 2.35 2.44 2.59 2.711.25 1.54 1.59 1.63 1.68 1.73 1.78 1.86 1.92 2.02 2.10 2.23 2.331.50 1.37 1.41 1.44 1.49 1.53 1.58 1.65 1.70 1.79 1.86 1.97 2.071.75 1.23 1.27 1.30 1.35 1.38 1.43 1.49 1.54 1.62 1.68 1.78 1.862.00 1.13 1.16 1.19 1.23 1.26 1.31 1.36 1.40 1.48 1.54 1.63 1.702.25 1.04 1.08 1.10 1.14 1.17 1.21 1.26 1.30 1.37 1.42 1.51 1.582.50 0.97 1.00 1.02 1.06 1.09 1.12 1.17 1.21 1.27 1.32 1.40 1.473.00 0.85 0.89 0.91 0.94 0.96 1.0 1.04 1.07 1.13 1.17 1.24 1.303.50 0.78 0.80 0.82 0.85 0.87 0.9 0.94 0.97 1.02 1.06 1.12 1.174.00 0.71 0.73 0.75 0.78 0.80 0.82 0.86 0.88 0.93 0.97 1.03 1.074.50 0.66 0.68 0.69 0.72 0.74 0.76 0.79 0.82 0.86 0.90 0.95 0.995.00 0.61 0.63 0.65 0.67 0.69 0.71 0.74 0.76 0.80 0.83 0.88 0.937.00 0.49 0.50 0.52 0.53 0.55 0.57 0.59 .61 0.64 0.67 0.71 0.74

10.00 0.39 0.40 0.41 0.42 0.43 0.45 0.47 0.48 0.51 0.53 0.56 0.5820.00 0.24 0.25 0.26 0.26 0.27 0.28 0.29 0.30 0.32 0.33 0.35 0.37

Increased sodium concentration in applied wastewater may also affectsoil hydraulic conductivity through its effects on organic matter dissolution.Land application of dairy effluent increased soil hydraulic conductivity de-spite its high sodium concentration (Sparling et al., 2001). This increase inhydraulic conductivity was attributed to the increase in soil porosity arisingfrom the dissolution of organic matter.

D. Biological Clogging

Decreases in soil hydraulic conductivity in soils receiving wastewater effluentmay also be due to biological clogging (Balks et al., 1997; Magesan et al.,1999). Biological clogging may be brought about by the accumulation ofmicrobial cells in the soil pores, the secretion of extracellular polymers, therelease of gases, and the accumulation of insoluble precipitates due to mi-crobial activity (Baveye et al., 1998). Magesan et al. (1999) found decreasesin hydraulic conductivity when a high C:N ratio (50:1) wastewater was ap-plied to soil, which was associated with increases in soil microbial biomassand extracellular C deposition. Similarly, application of meatworks to the


soil caused blocking of pores, this was attributed to biofilm formation. Thebiofilm was formed by extracellular polysaccharides and bacterial cells. Soilpermeability in that study was reduced by 70% within 4 days from effluentapplication, but recovered to its initial value 23 days after irrigation ceased.Magesan et al. (2000) found that the decrease of soil hydraulic conductiv-ity became more severe with increasing C:N ratio. Increasing C:N ratio from2.5:1 to 66:1 resulted in an 80% reduction of soil hydraulic conductivity. Thiseffect was attributed to the stimulative effect of available organic carbon onthe activity of soil microbial biomass. Based on these findings it can be con-cluded that the application of municipal wastewater to the soil is not thoughtto cause biological clogging, as most wastewaters are not characterized bysuch high C:N ratios.

Various practices to control biological clogging have been tested in lab-oratory and field studies, such as the application of sodium hypochlorite,sodium azide, and formaldehyde. Although some of these treatments maybe useful in aquifers or wells (Baveye et al., 1998), they cannot be applied atSRS because treatment efficiency is based on the biochemical function of thesoil microbial community. The most efficient way to control biological clog-ging is to apply wastewater at intervals sufficient to allow for the degradationof polysaccharides and the reduction of microbial biomass. These intervalsare a function of the organic loading and the climatic conditions. Balks et al.(1997) found that the required time for the recovery of soil permeability was23 days at 25◦C but increased to 50 days when the temperature droppedto 13◦C. The significance of the drying–wetting period to control biologi-cal clogging is also outlined in the study of Sparling et al. (2001). In thatstudy, despite the high loads of applied dairy effluent and the relatively highC:N ratio, soil hydraulic conductivity increased from 8.5 to 49.5 mm/h in asilty-loam soil and from 6.5 to 29.0 mm/h in a sandy soil (Sparling et al.,2001).

E. Nutrients

Numerous studies show that effluent application to land results in the ac-cumulation of nutrients above the levels required for optimum crop growth(Adeli et al., 2003; Geber, 2000; Smith and Bond, 1999; Redding et al., 2002;Woodard et al., 2002; Barton et al., 2005; etc.). Nutrients can be releasedto the environment by leaching, runoff, volatilization, or denitrification andcause adverse effects on surface-water bodies, groundwater, soil, and theatmosphere.

The nitrogen applied to the soil can be readily transformed to nitratesand leached to groundwater. Where the underground aquifers are used asdrinking-water sources or discharge to surface waters, there is concern aboutthe potential risks for both humans and wildlife. Health risks associated withincreased levels of nitrates in drinking water include methemoglobinemia


(blue baby syndrome), cancer, thyroid disease, and diabetes. As a conse-quence, limits for nitrates have been established for drinking water by in-ternational organizations and/or governmental agencies. However, the im-pacts of nitrates on human health has been questioned recently (L’hirondeland L’hirondel, 2001; Avery and L’hirondel, 2003). Fewtrell (2004) statedthat nitrates may be only one of the factors causing methaemoglobinaemia.However, when groundwater containing enhanced levels of nitrates dis-charges into surface water, it may induce algal development, deplete oxygen,and change pH, causing severe impacts on aquatic life. Additionally, toxinsproduced by algae have been associated with cancer, gastroenteritis, andnervous-system disorders (Ling, 2000; WHO, 1986). Predicting the impact ofexcessive nitrogen application on groundwater is a complex procedure, andoften it takes a considerable amount of time before these impacts becomeevident. For example, Tomer and Burkart (2003) found that excessive appli-cation of nitrogen during the period 1969–1974 continues to affect ground-water quality today. Principal factors determining the final concentration ofnitrates in groundwater are the texture of the subsoil through its effects ondenitrification rate, the groundwater depth, and the groundwater suppliesand flow rate because of dilution effect.

With respect to phosphorus, the concern that it will leach into thegroundwater is not supported by the existing evidence; however, leachingto small depths may occur under certain circumstances, depending on soilproperties and applied load of phosphorus (Johnson et al., 2004). Leachingof phosphorus at depths of 0.6 and 1.5 m was reported in two sites irrigatedwith piggery wastewater for 19 and 30 years, respectively (Redding, 2001).Likewise, increased concentrations of phosphorus were observed in drainagewater at a depth of 0.65 m in a soil having received livestock wastes for aperiod of 150 years (Heckrath et al., 1995).

Phosphorus accumulates close to the soil surface mainly increasing therisk of surface-water pollution through runoff. Redding et al. (2002), investi-gating the effects of long-term application of piggery wastewater to soils ofdifferent areas, reported an apparent increase in the total phosphorus in thesurface soil (0 to 5 cm). In this study, significant correlations were found in allsites between the applied loads of phosphorus and bicarbonate-extractablephosphorus, indicating an increase in the risk for surface runoff of phos-phorus. Application of piggery effluent for 11 years to a field planted withbermuda grass increased soil P availability 10 times above the optimum levelfor crop growth (King et al., 1990). Furthermore, excessive accumulation ofphosphorus in the soil may affect its future use in the long term becausesome plant species are sensitive to high P concentrations (Bond, 1998).

Adoption of an appropriate nutrient management plan that would takeinto account the site-specific factors is important to eliminate nutrient-induced environmental impacts. Successful management practices include(1) selection of vegetation with a high potential for biomass production and


nutrient uptake; (2) suitable vegetation management (length of harvestingcycle, multiple cropping systems, etc.); (3) practices that stimulate gaseouslosses of N (mainly as N2) and reduce runoff (berms, application methodand rate); and (4) higher levels of pretreatment especially for nutrient-richeffluents. To reduce environmental impacts arising from the excessive appli-cation on nutrients at SRS, several environmental agencies have developedguidelines/regulations that limit the amount of nitrogen and/or phosphorusapplied to the land according to the needs of the given crop. However, ad-ditional research is needed to estimate the nutrient demands of crops usedin SRS, as well as the principal processes and factors affecting the availabilityand the fate of nutrients in the soil.

F. Trace Elements

Trace elements in various types of effluents are of great importance becauseof their potential impact on human or animal health, crop performance, andon the population and activity of soil microfauna. Of particular concern is theoccurrence and the accumulation of cadmium, copper, and molybdenum, asthey can be harmful to human and animals at concentrations too low to affectplants (Crook, 1998).

Overall, trace elements in municipal wastewater that have received sec-ondary treatment do not pose a risk to plant growth or to the food chain,providing that industrial wastewaters are not discharged into sewerage sys-tems. During the past few years a significant decrease has been observedin the concentrations of trace elements in developed countries as a result ofthe stringent disposal standards for industrial effluents. This has tended toeliminate the risks arising from heavy metals during land application of mu-nicipal effluents. In developing countries, however, application of wastew-ater to the land often has been found to cause accumulation of trace ele-ments in the soil to levels above those suggested as potentially harmful byinternational organizations or environmental agencies. In Mexico the appli-cation of raw wastewater to land for 90 years has increased trace elementconcentrations (Ramirez-Fuentes et al., 2002). Similarly, enhanced levels ofcadmium, chromium, and lead were assessed in another effluent applicationarea in Mexico, which in the case of lead exceeded the limits established byEuropean Union regulations (Lucho-Constantino, 2005). In India, soil contentof zinc, copper, ferrum, nickel, and lead increased significantly after 20 yearsof effluent application compared to those received fresh water. Despite theincreased levels of trace elements in the soil, their concentrations in vege-tation were not found to pose a risk for human health (Rattan et al., 2005).Effluent irrigation of vegetables in Harare in Zimbabwe resulted in enhancedlevels of copper, zinc, cadmium, nickel, chromium, and lead in the soil. Theannual rates of heavy metal loading showed that heavy metals would haveexceeded the permitted limits within 5–60 years (depending on the metal


and the site), consequently posing serious environmental and health risks(Mapanda et al., 2005).

The concern about the accumulation of trace elements in SRS is mainlyfocused on those receiving industrial wastewaters, landfill leachates, and agri-cultural drainage. In order to deal with the environmental and health risks,several environmental agencies have adopted regulations that determine themaximum concentrations or loads of trace elements that can be applied toland (U.S. EPA, 1993, 2004). It should be stated, however, that the total loadof trace elements in the soil may overestimate the potential risks. It is the la-bile pool of trace elements that determines the potential risks. Trace elementsapplied to the soil are subjected to various physical, chemical, and biologicalprocesses, such as sorption to soil colloids, chemical precipitation, or theformation of complex compounds with organic matter or other elements,which may result in a significant decrease of their bioavailability. Cadmiumconcentration in grass was only slightly increased after 76 years of appli-cation of untreated effluent; similar findings were also obtained at differentsites in California (Reed et al., 1995). Plants may also play a complex rolein heavy metal management through their effects on the availability, uptake,and distribution to plant tissues in various plant species, as discussed in aprevious section.

X. CONCLUSIONS AND FUTURE RESEARCH NEEDS

The application of wastewater to the land, known as “sewage farming,” hasbeen practiced for centuries as a means to manage wastewater. It is an oldtechnique that with prudent management can be compatible with the currenthigh public health standards adopted by several countries and internationalorganizations and with the sustainable use of land. The use of land-basedsystems and especially SRS for treatment of municipal and other types ofwastewater is expected to further expand in the future. This expanded useis in response to the high construction and maintenance costs of complextertiary treatment processes, as well as the need to eliminate disposal ofeffluents into streams and lakes. Land-based wastewater treatment systemsappear to be an ideal practice, particularly in arid or semi-arid regions whereeffluents can be used efficiently for increasing irrigated areas.

Despite the important advantages of SRS in the treatment and reuse ofvarious types of effluents, critical health and environmental issues can arise ifappropriate management practices are not adopted. A vast amount of knowl-edge has been accumulated with regard to the principal processes affectingpathogen removal, inactivation, transport, and survival in the soil. Overall,land application of effluent serves as an effective way to purify effluents frompathogenic organisms (Table 8), with soil properties, hydraulic conditions,environmental factors, and effluent composition being the dominant factors.


Unsaturated flow conditions and the application of appropriate wetting–drying periods can appreciably decrease pathogen transport to underlyingaquifers and their survival in the land, eliminating the risks to public health.However, preferential flow, often observed in soils, appears to be the mostimportant factor favoring pathogen leaching; thus, emphasis should be givento the development of appropriate management practices to prevent macro-pore flow. Relatively little information is available regarding the effects ofvegetation (direct and indirect) on pathogens’ survival and transport. Vege-tation can alter (bio)chemical conditions of the rhizosphere, bulk density ofthe soil, microclimate, and organic matter, exerting direct and indirect effectson pathogen fate. Evidently more research efforts should be directed towardsthis subject.

In general, organic matter is not a limiting factor for the performance ofSRS as they are characterized by a high degradation potential. The principalconcern is the application of appropriate wetting–drying cycles to prevent bi-ological clogging of the soil and so maintain treatment efficiency and systemsustainability and prevent odor development. Relatively little information,however, is available about the fate of toxic organic substances contained ineffluents. In future, more emphasis should be given to the study of the fate ofspecific organic compounds, such as pharmaceutical agents, polycyclic aro-matic hydrocarbons, pesticides, and personal care products. Moreover, basicresearch is needed on the structure of the microbial community responsiblefor the biochemical functions in SRS, the factors affecting its performance,and its interaction with vegetation. Currently, molecular methods have beendeveloped that can be successfully applied to identify and study microbialcommunities in the soil. The application of such techniques is expected tobe a valuable tool for increasing our understanding of the role of specific mi-croorganisms in wastewater treatment and for improving treatment efficiencyof land-based wastewater treatment systems. Moreover, detailed study of theparameters affecting the movement of organic pollutants through ecosystemsand development of more accurate methodologies to estimate the environ-mental and health risks resulting from complex mixtures of such contami-nants is urgent.

With regard to the environmental impacts associated with nutrient loads,a major challenge is to match plant water requirements with the capacityof vegetation for nutrient uptake and particularly for nitrogen and phos-phorus. In regions with semi-arid and arid climate, the capacity of vegeta-tion to remove nutrients is exceeded when wastewater is applied at ratesneeded to satisfy evapotranspiration needs, even for wastewater contain-ing relatively low concentrations of nutrients. Therefore, application rates ofwastewater must be balanced with the capacity of the soil–plant system toremove/assimilate nutrients, or practices that induce nutrients uptake andremoval by vegetation should be applied. The selection and use of plantspecies with a high potential for biomass production, high rates of nutrient


uptake, and increased water use efficiency can substantially improve the ca-pacity of SRS to manage nutrients. Furthermore, attention should be given tothe application of management practices that enhance the removal of nutri-ents, such as applied optimum harvesting cycle, planting densities, and thosepractices that favor gaseous losses of nitrogen as N2.

The management of salts is a critical issue for the successful long-termperformance of SRS. Salt accumulation in the rhizosphere may reduce theactivity of the microbial community, affect the sustainability of plantation, anddamage soil structure and therefore the potential of soil to treat wastewater.Selection of salt-tolerant genotypes, accurate modeling of water consumptionunder such conditions, and applying sufficient leaching requirements willalleviate these impacts.

It is evident from the previous paragraphs that vegetation has an im-portant role in the treatment efficiency of SRS mainly through its effects onnutrient uptake and water use. More research should be directed to the eval-uation of plant species or clones for use in such systems by assessing theirpotential for nutrient and/or trace elements removal, biomass production,stimulation of the organic contaminants degradation, and water use. In ad-dition, transgenic plants that have been engineered with certain biochemicalpathways or proteins may also improve the capacity of SRS for wastewa-ter treatment. Currently, advances in the field of molecular plant physiologyhave revealed important information regarding the genes regulating the up-take and the metabolism of nutrients, growth, uptake and tolerance of saltsand heavy metals, water use and uptake, and metabolism of recalcitrant or-ganic compounds.

The development of guidelines/regulations that integrate the existingknowledge and adequately address the ecological, environmental and publichealth risks resulting from land application of wastewater effluents is impor-tant. These guidelines should take into consideration the type of wastewater,the required pretreatment level, the degree of public access to the site, sitecharacteristics (soil properties, topography, hydrology, aquifer depth andquality), effluent application method, and the end use of the irrigated crop.

In addition, of particular concern is how climate change will affect thetreatment efficiency of SRS. Based on the findings of previous studies de-signed to investigate the effects of climate change on plant performance,doubling the atmospheric CO2 concentration will result in reduced transpi-ration rate and increased photosynthesis, and hence on higher water useefficiency. Furthermore, as in SRS neither soil moisture nor nutrients are alimiting factor for growth, biomass production will increase favoring nutrientuptake. Higher temperatures will stimulate organic matter degradation andwill reduce pathogens survival in the soil. However, more detailed estima-tions of climate change in a spatial and temporal basis are required to predictwith accuracy the impacts of climate change on the processes affecting theperformance of SRS as well as the complex interactions amongst them.


ACKNOWLEDGEMENTS

Funding in support of this review was provided by the Greek National Schol-arships Institute. The authors thank Prof. Roubelakis-Angelakis, University ofCrete, Prof. Salgot, University of Catalonia, Dr. Tsagarakis, Technical Uni-versity of Crete, and Prof. Rose, Michigan State University, for their criticalreview and comments.

REFERENCES

Adegbidi, H.G., Volk, A.T., White, E.H., Abrahamson, L.P., Briggs, R.D., andBickelhaupt, D.H. Biomass and nutrient removal by willow clones in experi-mental bioenergy plantations in New York State, Biomass Bioenerg. 20, 399,2001.

Adeli, A., Varco, J.J., and Rowe, D.E. Swine effluent irrigation rate and timing effectson bermudagrass growth, nitrogen and phosphorous utilization, and residualsoil nitrogen, J. Environ. Qual. 32, 681, 2003.

Aislabie, J., Smith, J.J., Fraser, R., and McLeod, M. Leaching of bacterial indicators ofcontamination through four New Zealand soils, Aust. J. Soil Res. 39, 1397, 2001.

Aitchison, E.W., Kelley, S.L., Alvarez, P.J.J., and Schnoor, J.L. Phytoremediation of1,4-dioxane by hybrid poplar trees, Water Environ. Res. 72, 313, 2000.

Alberts, J.J., Takacs, M., and Pattanayek, M. Influence of IHSS standard and referencematerials on copper and mercury toxicity to Vibrio fiecheri, Acta HydrochimHydrobiol. 28, 428, 2001.

Alexander, M. Aging, bioavailability, and overestimation of risk from environmentalpollutants, Eniron. Sci. Technol. 34, 4259, 2000.

Al-Jamal, M.S., Sammis, T.W., Mexal, J.C., Picchioni, G.A., and Zachritz, W.H. Agrowth irrigation-scheduling model for wastewater use in forest production,Agric. Water Manage. 56, 57, 2002.

Al-Kaisi, M.M., and Waskom, R.M. Estimating ammonia loss from sprinkler-applicatedswine effluent, Agron. J. 94, 1156, 2002.

Allen, R., Pereira, G., Raes, L.S., and Smith, D.M. Crop evapotranspiration: Guidelinesfor computing crop water requirements, Irrigation and Drainage Paper 56, FAORome, 1998.

Angelakis, A.N., and Rolston, D.E. Transient movement and transformation of carbonspecies in soil during wastewater application, Water Resource Res. 21, 1141, 1985.

Angelakis, A.N., and Spyridakis, S.V. The status of water resources in Minoan times—A preliminary study, in Diachronic Climatic Impacts on Water Resources withEmphasis on Mediterranean Region, Angelakis, A. and Issar, A., Eds., Springer-Verlag, Heidelberg, Germany, 161, 1996.

Angelakis, A.N., Koutsoyannis, D., and Tchobanoglous, G. Urban wastewater andstormwater technologies in the ancient Greece, Water Res. 39, 210, 2005.

Armstrong, W., Brandle, R., and Jackson, M.B. Mechanisms of flood tolerance inplants, Acta Bot. Neerl. 43, 307, 1994.

Artz, R.R.E., and Killham, M. Survival of Eschericia coli O157:H7 in private drinkingwater influences of protozoan grazing and elevated copper concentration, FEMSMicrobiol. Lett. 216, 117, 2002.


Artz, R.R.E., Townend, J., Brown, K., Towers, W., and Killham, K. Soil macroporesand compaction control the leaching potential of Escherichia coli O157:H7,Environ. Microbiol. 7(2), 241, 2005.

Atienza, J., Tabernero, M.T., Alvarez-Bened, J., and Sanz, M. Volatilisation of triallateas affected by soil texture and air velocity, Chemosphere, 42, 257, 2001.

Avery, A., and L’hirondel, J.L. Nitrate and Methemoglobinemia, Environ. Health Per-spect. 111, 142, 2003.

Ayers, R.S., and Tanji, K.K. Agronomic aspects of crop with wastewater, in Proc. ofthe Water Forum Conference, ASCE, 1981.

Ayers, R.S., and Westcot, D.W. Water Quality for Agriculture, FAO Irrigation andDrainage Paper 21, Rev. 1, United Nations Food and Agriculture Organization,Rome, Italy, 1985.

Azbar, N., Bayram, A., Filibeli, A., Muezzinoglu, A., Sengul, F., and Ozer, A. A reviewof waste management options in olive oil production, Crit. Rev. Environ. Sci.Technol., 34, 209, 2004.

Bais, H.P., Walker, T.S., Schweizer, H.P., and Vivanco, J.M. Root specific elicitationand antimicrobial activity of rosmarinic acid in hairy root cultures of Ocomumbasilicum, Plant Physiol. Biochem. 40, 983, 2002

Balks, M.R., McLay, C.D.A. and Harfoot, C.G. Determination of the progression insoil microbial response, and changes in soil permeability following applicationof meat processing effluent to soil, Appl. Soil Ecol. 6, 109, 1997.

Barton, L., Schipper, L.A., Barkle, G.F., McLeod, M., Speir, T.W., Taylor, M.D., McGill,A.C., van Schaik, A.P., Fitzgerald, N.B., and Pandey, S.P. Land application ofdomestic effluent onto four soil types: Plant uptake and nutrient leaching,J. Environ. Qual. 34, 635, 2005.

Barton, L., McLay, C.D., Schipper, L.A., and Smith, C.T. Annual denitrification ratesin agricultural and forest soils: A review, Aust. J. Soil Res. 37, 1073, 1999.

Baveye, P., Vandevivere, P., Hoyle, B.L., DeLeo, P.C., and de Lozada, D.S. Environ-mental impact and mechanisms of the biological clogging of saturated soils andaquifer materials, Crit. Rev. Environ. Sci. Technol. 28, 123, 1998.

Beale, C.V., M.I.L., and Long, S.P. Water use efficiency of C4 perennial grasses in atemperate climate, Agric. For. Meteor. 96, 103, 1999.

Bell, R.G., and Bole, J.B. Elimination of fecal coliform bacteria from soil ir-rigated with municipal sewage lagoon effluent, J. Environ., Qual. 7, 193,1978.

Benyon, R.G., Marcar, N.E., Theieveyanathan, S., Tunningley, W.M., and Nicolson,A.T. Species differences in transpiration on a saline discharge site, Agric. WaterManage. 50, 65, 2001.

Bogan, B.W., and Sullivan, W.R. Physicochemical soil parameters affecting seques-tration and mycobacterial biodegradation of polycyclic aromatic hydrocarbonsin soil, Chemosphere 52, 1717, 2003.

Bomo, A.M, Stevic, T.K., Hovi, I., and Hanssen, J.F. Bacterial removal in biologicalsand filters, J. Environ. Qual. 33, 1041, 2004.

Bond, W.J. Effluent irrigation—An environmental challenge to soil science, Aust. J.Soil Res. 36, 543, 1998.

Bouwer, H. Integrated water management: Emerging issues and challenges, Agric.Water Manage. 45, 217, 2000.


Briggs, G.G., Bromilow, R.H., and Evans, A.A. Relationships between lipophilicityand root uptake and translocation of non-ionized chemicals by barley, Pestic.Sci. 13, 495, 1982.

Brusseau, M.L., Jessup, R.E., and Rao, P.S.C. Nonequilibrium sorption of organicchemicals: Elucidation of rate-limiting processes, Environ. Sci. Technol. 25, 134,1991.

Burgen, J.G. Uptake and Metabolism of organic compounds: Green liver model, inPhytoremediation: Transformation and Control of Contaminants, McCutcheon,S.C., and Schnoor, J.L., Eds. 59, 2003.

Burken, J.G., and Schnoor, J.L. Predictive relationships for uptake of organic con-taminants by hybrid poplar trees, Environ. Sci. Technol. 32, 3379, 1998.

Cabrera, F., Lopez, R., Martnez-Bordiu, A., Dupuy de Lome, E., and Murrilo, J.M.Land treatment of olive oil mill wastewater, Int. Biodeter. Biodegr. 38, 215, 1996.

Cameron, K.C., Di, H.J., Anwar, M.R., Russell, J.M. and Barnett, J.W., The “critical”ESP value: does it change with land application of dairy factory effluent?, N. Z.J. Agric. Res. 46, 147, 2003.

Campbell, A.T., Robertson, L.J., and Smith, H.V. Viability of Cryptosporidium parvumoocysts: correlation of in vitro excystation with inclusion or exclusion of fluo-rogenic vital dyes, Appl. Environ. Microbiol. 58, 3488, 1992.

Chang, A.C., and Page, A.L. Fate of wastewater constituents in soil and groundwater:Trace organics, in Irrigation with reclaimed wastewater: A guidance manual,Pettygrove, G.S., and Asano,T., Eds., Lewis Publishers, Chelsea, MI, 1988.

Chaudri, A.M., Knight, B.P., Bardosa-Jefferson, V.L., Preston, S., Raton, G.I., Killham,K., Coad, N., Nicholso, F.A., Chambers, B.J., and McGrath, S.P. Determination ofacute Zn toxicity in pore water from soils previously treated with sewage sludgeusing bioluminescence assays, Environ. Sci. Technol. 33, 1880, 1999.

Chen, S.H., and Aitken, M.D. Salicylate stimulates the degradation of high-molecularweight polycyclic aromatic hydrocarbons by Pseudomonas sacharofilia P15,Environ. Sci. Technol. 33, 435, 1999.

Cheng, K.Y., and Wong, J.W.C. Combined effect of nonionic surfactant Tween 80 andDOM on the behaviors of PAHs in soil–water system, Chemosphere, in press.

Chiou, C.T., Porte, P.E., and Schmedding, D.W. Partition equillibria of nonionic or-ganic compounds between soil organic matter and water, Environ. Sci. Technol.17, 227, 1983.

Chiou, C.T. Theoretical considerations of the partition uptake of nonionic organiccompounds by soil organic matter, in Reactions and Movement of Organic Chem-icals in Soils, SSSA Special Publication no. 22, 1989, 1.

Chu, Y., Jin, Y., and Yeates, M.V. Virus transport through saturated sand columns asaffected by different buffer solutions, J. Environ. Qual. 29, 2000.

Chu, Y., Jin, Y., Flury, M., and Yeates, M.V. Mechanisms of virus removal duringtransport in unsaturated porous media, Water Resour. Res. 37, 253, 2001.

Correll, D.L. The role of phosphorus in the eutrophication of receiving waters: Areview, J. Environ. Qual. 27, 261, 1998.

Corseuil, H.X., and Moreno, F.N. Phytoremediation potential of willow trees foraquifers contaminated with ethanol-blended casoline, Water Res. 35, 3013, 2001.

Cousins, I.T., Beck, A.J., and Jones, K.C. A review of the processes involved inthe exchange of semi-volatile organic compounds (SVOC) across the air–soilinterface, Sci. Total Environ. 228, 5, 1999.


Crescimanno, G., Iovino, M., and Provenanzano, M. Influence of salinity and sodicityon soil structural and hydraulic characteristics, Soil Sci. Soc. Amer. J. 59, 1701,1995.

Crites, R.W., and Fehrmann, R.C. Land application of winery stillage wastes, in Pro-ceedings Third Annual Madison Conference on Applied Research in MunicipalIndustrial Waste, Sept. p12, 1980.

Crites, R.W., Reed, S.C., and Bastian, R.K. Land Treatment Systems for Domestic andIndustrial Wastes. McGraw-Hill, Inc., New York, 2000.

Crook, J. Water reclamation and reuse criteria, in Wastewater Reclamation and Reuse,Asano, T., Ed., Technomic, Lancaster, PA, 1998.

Dai, X., and Boll, J. Evaluation of attachment of Cryptosporidium parvum and Giardialamblia to soil particles, J. Environ. Qual. 32, 296, 2003.

Davis, L.C., Castro-Diaz, S., Zhang, Q., and Ericson, L.E. Benefits of vegetation forsoils with organic contaminants, Crit. Rev. Plant. Sci. 21, 457, 2002.

Davis, L.C., Vnderhoof, S., Dana, J., Selk, K., Smith, K., Golpen, B., and Erickson,L.E. Movement of chlorinated solvents and other volatile organics through plantsmonitored by fourier transform infrared (FT-IR) spectometry, J. Hazard. WasteRes. 1, 1, 1998.

Dawson, D.H., Isebrands, J.G., and Gordon, J.C. Growth, dry weight yields, and spe-cific gravity of 3-year-old Populus grown under intensive culture, USDA ForestService N.C. Forest Exp. Station Res. Paper NC-122, 7, 1976.

Degens, B.P., Schipper, L.A., Claydon, J.J., Russell, J.M., and Yates, G.W. Irrigation ofan allophonic soil with dairy factory effluent for 22 years: Responses of nutrientstorage and soil biota, Aust. J. Soil Res. 38, 25, 2000.

Doorendos, J., and Pruitt, W.O. Crop Water Requirements, Food and AgricultureOrganization of the United Nations, Rome, 1977.

Downs, T.J., Cifuentes-Garcia, E., and Suffet, I.M. Risk screening for exposure togroundwater pollution in a wastewater irrigation district of the Mexico Cityregion, Environ. Health Perspect. 107, 553, 1999.

Drozd, C., and Schwartzbrod, J. Hydrophobic and electrostatic cell surfaceproperties of Cryptosporidium parvum, Appl. Environ. Microbiol. 62, 1227,1996.

Duncan, M.J., Baker, G., and Wall, G.C. Wastewater irrigated tree plantations: Produc-tivity and sustainability, 61st Annual Water Industry Engineers and Operators’Conference Civil Centre-Shepparton, 18, 1998.

Edraki, M., So, H.B., and Gardner, E.A. Water balance of swamp mahogany andRhodes grass irrigated with treated sewage effluent, Agric. Water Manage. 67,157, 2004.

Eghball, B., Binford, G.D., and Baltensperger, D.D. Phosphorus movement and ad-sorption in a soil receiving long-term manure and fertilizer application, J. Envi-ron. Qual. 25, 1339, 1996.

Elowson, S. Willow as a vegetation filter for cleaning of polluted drainage water fromagricultural land, Biomass Bioenerg. 16, 281, 1999.

Fabeiro, C., de Santa Olalla, F.M., Lopez, R., and Domınguez, A. Production andquality of the sugar beet (Beta vulgaris L.) cultivated under controlled deficitirrigation conditions in a semi-arid climate, Agric. Water Manage. 62, 215, 2003.

Falkiner, R.A., and Polglase, P.J. Fate of applied phosphorous in an effluent-irrigatedPinus radiata plantation, Aust. J. Soil Res. 37, 1995, 1999.


Falkiner, R.A., and Smith, C.J. Changes in soil chemistry in effluent irrigated Pinusradiata and Eucalyptus grandis plantations, Aust. J. Soil. Res. 35, 131, 1997.

Fava, F., Berselli, S., Conte, P., Piccolo, A., and Marchetti, L. Effects of humic sub-stances and soya lecithin on the aerobic bioremediation of a soil historicallycontaminated by polycyclic aromatic hydrocarbons (PAHs), Biotechnol. Bioeng.88, 214, 2004.

Fayer, R., Trout J.M., and Jenkins M.C. Infectivity of Cryptosporidium parvum oocystsstored in water at environmental temperatures, J. Parasitol. 84, 1165, 1998.

Fewtrell, L. Drinking-water nitrate and methemoglobinemia, Global burden of dis-ease: A discussion, Environ. Health Perspect. 112(14), 1371, 2004.

Filip, Z., Kanazawa, S., and Berthelin, J. Distribution of microorganisms, biomass ATP,and enzyme activities in organic and mineral particles of a long-term wastewaterirrigated soil, J. Plant Nutr. Soil Sci. 163, 143, 2000.

Folsom, C.F. Seventh Annual Report of the State Board of the Health of Massachusetts,Wright and Potter, Boston, 1876.

Frankenberger, W.T. Fate of wastewater constituents in soil and groundwater:Pathogens, in Irrigation with Recycled Water: A Guidance Manual, Asano, T.,Ed., Lewis Publishers, Cheslea, MI, 1988.

Gagliardi, J.V., and Karnss, J.S. Leaching of Eshericia coli 0157:H7 in diverse soilsunder various agricultural management practices, Appl. Environ. Microbiol. 66,877, 2000.

Garcıa-Orenes, F., Guerrero, C., Mataix-Solera, J., Navarro-Pedreno, J., Gomez, I., andMataix-Beneyto, J. Factors controlling the aggregate stability and bulk densityin two different degraded soils amended with biosolids, Soil Till. Res. 82, 65,2005.

Geber, U. Nutrient removal by grasses irrigated with wastewater and nitrogen balancefor reed canarygrass, J. Environ. Qual. 29, 398, 2000.

Gerba, C., and Bitton, G. Microbial pollutants, in Groundwater Pollution Microbiol-ogy, Bitton, G. and Gerba, C., Eds, John Wiley, New York, 1984.

Gerba, C.P. Applied and theoretical aspects of virus adsorption to surfaces, Adv. Appl.Microbiol. 30, 133, 1984.

Gerba, C.P., Wallis, C., and Melnick, J.L. Fate of wastewater bacteria and viruses inthe soil. ASCE Irrig. Drain. Div. J. IR3, 157, 1975.

Gerhard, W.P. Sanitation and Sanitary Engineering, New York, 1909.Giller, K.E., Witter, E., and McGrath, S.P. Toxicity of heavy metals to microorganisms

and microbial processes in agricultural soils: A review, Soil Biol. Biochem. 30,1389, 1998.

Gochis, D.J., and Cuenca, R.H. Plant water use and crop curves for hybrid poplars,J. Irrig. Drain. Eng. 126, 206, 2000.

Gordon, M., Choe, N., Duffy, J., Ekuan, G., Heilman, P., Muiznieks, I., Ruszai, M.,Shurtleff, B.B., Strand, S., and Wilmoth, L.A. Phytoremediation of trichloroethy-lene with hybrid poplars, Environ. Health Perspect. 106, 1001, 1998.

Goyal, S.M., and Gerba, C.P. Comparative adsorption of human enterovirus, simianrotavirus, and selected bacteriophages in the soil, Appl. Environ. Microbiol. 38,241, 1979.

Greger, M. Salix as phytoextractor, in Proceedings of 5th International Conference onthe Biogeochemistry of Trace elements, Wenzel, W.W., et al., Eds., Viena, Boku,872, 1999.


Guerin, T.F. Long term performance of a land treatment facility for the bioremediationof non-volatile oily wastes, Resource Conserv. Recyc. 28, 105, 2000.

Guerin, W.F., and Boyd, S.A. Bioavailability of naphthalene associated with naturaland synthetic sorbents, Water Res. 31, 504, 1997.

Guo, L.B., and Sims, R.E.H. Soil response to eucalypt tree planting and meatworkseffluent irrigation in a short rotation forest regime in New Zealand, Bioresource.Technol. 87, 341, 2003.

Guo, L.B., Sims, R.E.H., and Horne, D.J. Biomass production and nutrient cycling inEucalyptus short rotation energy forests in New Zealand I: Biomass and nutrientaccumulation, Bioresource Technol. 85, 273, 2002.

Gupta, V.V.S.R., Rogers, S., and Naidu R. Effects of secondary treated sewage effluentapplication on the populations of microfauna in a hardwood plantation soil:Bolivar HIAT trial, Geoderma, 84, 249, 1998.

Halliwell, D.J., Barlow, K.M., and Nash, D.M. A review of the effects of wastewatersodium on soil physical properties and their implications for irrigation systems,Aust. J. Soil Res. 39, 1259, 2001.

Harden, V., and Harris, J. The isoelectric point of bacterial cells, J. Bacteriol. 65, 1998,1953.

Harner, T., and Mackay, D. Measurements of octanol–air partitioning coefficients forchlorobenzenes, PCB and DDT, Environ. Sci. Technol. 29, 1599, 1995.

Hasselgren, K. Use of municipal waste products in energy forestry: Highlights from15 years of experience, Biomass Bioenerg. 15, 71, 1998.

Hayat, S., Ahmad, I., Azam, Z.M., Ahmad, A., and Inam, A. Effect of long termapplication of oil refinery wastewater on soil health with special reference tomicrobiological characteristics, Bioresource Technol. 84, 159, 2002.

Heckrath, G., Brooks, J.W., Poulton, P.R., and Goulding, K.W.T. Phosphorus leachingfrom soils containing different soil concentration in the Broadbalk Experiment,J. Environ. Qual. 24, 904, 1995.

Heilman, P., and Norby, R.J. Nutrient cycling and fertility management in temperateshort rotation forest systems, Biomass Bioenerg. 4, 361, 1998.

Hernandez, B.S., Koh, S.C., Chial, M., and Focht, D.D. Terpene-utilizing isolatesand their relevance to enhanced degradation of polychlorinated biphenyls,Biodegradation 8, 153, 1997.

Hooda, A.K., Weston, C.J., and Chen, D. Denitrification in effluent-irrigated claysoil under Eucalyptus globules plantation in south-eastern Australia, For. Ecol.Manage. 1, 61–62, 2003.

Howell, T.A., Tolk, J.A., Schneider, A.D., and Evett, S.R. Evapotranspiration, yieldand water use efficiency of corn hybrids differing in maturity, Agron. J. 90, 3,1988.

Hsu, F.C., Mrxmiller, R.L., and Yang, A.Y. Study of root uptake and xylem transloca-tion of cinmethylin and related compounds in detopped soybean roots, PlantPhysiol. 93, 1573, 1990.

Huang, X., Lee, L.S., and Nakatsu, C. Impact of animal waste lagoon effluents onchlopyrifos degradation in soils, Environ. Toxicol. Chem. 19, 2864, 2000.

Hurst, C., Gerba, C., and Cech, I. Effects of environmental variables and soil charac-teristics on virus survival in soil, Appl. Environ. Microb. 40, 1067, 1980.

Hurst, C.J. Influence of aerobic microorganisms upon virus survival in soil, Can. J.Microbiol. 34, 696, 1988.


Huysman, F., and Verstraete, W. Water-facilitated transport of bacteria in unsaturatedcolumns: Influence of cell surface hydrophobicity and soil properties, Soil Biol.Biochem. 25, 83, 1993.

Ilani, T., Schulz, E., and Chefetz, B. Interactions of organic compounds with wastew-ater dissolved organic matter: Role of hydrophobic fractions, J. Environ. Qual.34, 552, 2005.

ISTRAM, Soil and forestry study for the proposed augmentation of the Cleveland WaterPollution Control Works Effluent Irrigation Scheme, Prepared for Redland ShireCouncil, ISTRAM, University of Queensland, 1993.

Jandl, R., and Sletten, R.S. Mineralization of forest soil carbon: interactions withmetals, J. Plant Nutr. Soil Sci. 162, 623, 1999.

Jayawardane, N.S., Blackwell, J., Cook, F.J., Nicoll, G., and Wallet, D. Final reporton pollutant removal by the FILTER system during the period November 1994 toNovember 1996, Prepared for Griffith City Council, CSIRO Div. Land and WaterConsultancy Report No. 97, 1997.

Jenkins, M.B., Bowmam, D.D., Fogarty, E.A., and Ghiose, W.C. Cryptosporidiumparvum oocysts inactivation in three soil types at various temperatures andwater potentials, Soil Biol. Biochem. 34, 1101, 2002.

Jenkins, M.B., Walker, M.J., Bowman, D.D., Anthony, L.C., and Ghiorse, W.C.Use of a sentinel system for field measurements of Cryptosporidium parvumoocyst inactivation in soil and animal waste, Appl. Environ. Microbiol. 65,1999.

Jewett, D.G., Hilbert, T.A., Logan, B.E., Arnold, R.G., and Bales, R.C. Bacterial trans-port in laboratory columns and filters: Influence of ionic strength and pH oncollision efficiency, Water Res. 29, 1673, 1995.

Jewett, D.G., Logan, B.E., Arnold, R.G., and Bales, R.C. Transport of Pseudomonasfluorescens strain P17 through quartz sand columns as a function of water con-tent, J. Contam. Hydrol. 36, 73, 1999.

Jimenez, B., and Chavez, A. Quality assessment of an aquifer recharged with wastew-ater for its potential use as drinking source: “El Mezquital Valley” case, WaterSci. Technol. 50(2), 269, 2004.

Jimenez, B. Health risk in aquifer recharge with recycled water, in State of the ArtReport: Health Risks in Groundwater Recharge with Recycled Water, Aertgeerts,R., and Angelakis, A.N., Eds., WHO, WHO Regional Office for Europe, Copen-hagen, Denmark, 2003.

Jin, Y., Chu, Y.J., and Li, Y.S. Virus removal and transport in saturated and unsaturatedsand columns, J. Contam. Hydrol. 43, 111, 2000.

Jjemba, P.K. The potential impact of veterinary and human therapeutic agents inmanure and biosolids on plants grown on arable land: A review, Agric. Ecosys.Environ. 93, 267, 2002.

John, D.E., and Rose, J.B. Review of factors affecting microbial survival in ground-water, Environ. Sci. Technol. 39, 7345, 2005.

Johnson, A.F., Vietor, D.M., Rouquette, Jr., and Haby, V.A. Fate of phosphorus indairy wastewater and poultry litter applied on grassland, J. Environ. Qual. 33,735, 2004.

Johnson, W.P., Martin, M.J., Gross, M.J., and Logan, B.E. Facilitation of bacterialtransport through porous media by changes in solution and surface properties,Colloids Surfaces A Physicochem. Eng. Aspects, 107, 263, 1996.


Jones, A.H., Voulvoulis, N., and Lester J.N. Human pharmaceuticals in wastewatertreatment processes, Crit. Rev. Environ. Sci. Technol. 35, 401, 2005a.

Jones, D.L., Kemmitt, S.J., Wright, D., Cuttle, S.P., Bol, R., and Edwards, A.C. Rapidintrinsic rates of amino acid biodegradation in soils are unaffected by agriculturalmanagement strategy, Soil Biol. Biochem. 37, 1267, 2005b.

Jørgensen, U., and Schelde, K. Energy crop water and nutrient use efficiency, In-ternational Energy Agency, IEA Bioenergy Task 17, Short Rotation Crops, 36,2001.

Jørgensen, U. Genotypic variation in dry matter accumulation and content of N, Kand Cl in Miscanthus in Denmark, Biomass Bioenerg. 12, 155, 1997.

Karthikeyan, R., and Kulakow, P.A. Soil plant microbe interactions in phytoremedi-ation, Advances in Biochemical Engineering/Biotechnology, 78, 51, 2003.

Kauter, D., Lewandowski, I., and Claupein, W. Quantity and quality of harvestablebiomass from Populus short rotation coppice for solid fuel use—A review ofthe physiological basis and management influences, Biomass Bioenerg. 24, 411,2003.

Kille, D.E., Chiou, C.T., Zhu, H., Li, H., and Xu, O. Partition of nonpolar organic pol-lutants from water to soil and sediment organic matters, Environ. Sci. Technol.29, 1401, 1995.

Kimani-Njoroge, B.N., Ball, W.P., and Cherry, R.S. Sorption of 1,2,4,-trichlorobenzeneand tetrachloroethene within an authigenic profile: Changes in Koc with soildepth, J. Contam. Hydrol. 29, 347, 1998.

King, L.D., Burns, J.C., and Westerman, P.W. Long term swine effluent lagoon effluentapplications on “coastal” bermudagrass: II. Effect on nutrient accumulation inthe soil, J. Environ. Qual. 19, 756, 1990.

Kopp, R.F., Abrahamson, L.P., White, E.H., Burns, K.F., and Nowak, C.A. Cuttingcycle and spacing effects on biomass production by a willow clone in NewYork, Biomass Bioenerg. 12, 313, 1997.

Kopp, R.F., Abrahamson, L.P., White, E.H., Nowak, C.A., Zsuffa, L., and Burns, K.F.Woodgrass spacing and fertilization effects on wood biomass production by awillow clone, Biomass Bioenerg. 11, 451, 1996.

Korte, E., Kvesitadze, G., Ugrekhelidze, D., Gordeziani, M., Khatisashvili, G., Buadze,O., Zaalishvili, G., and Goulston, F. Organic toxicants and Plants, Ecotox. Envi-ron. Safety 47, 1, 2000.

Kozlowski, T.T. Responses of woody plants to flooding and salinity, Tree Physiol.Monogr. 1, 1, 1997.

Kroening, S.J., Leung, D.W., Greenfield, L.G., and Galilee, C. Losses of diesel oil byvolatilisation and effects of diesel oil on seed germination and seedling growth,Environ. Technol. 22, 1113, 2001.

Kuzel, K., and Drake, H.L. Microbial turnover of low molecular weight organic acidsduring leaf litter decomposition, Soil Biol. Biochem. 31, 107, 1999.

Labrecque, M., and Teodorescu, T.I. High biomass yield achieved by Salix clones inSRIC following two 3-year coppice rotations on abandoned farmland in southernQuebec, Canada, Biomass Bioenerg. 25, 135, 2003.

Lance, J.C., and Gerba, C.P. Effect of ionic composition of suspending solution onvirus adsorption by a soil column, Appl. Environ. Microbiol. 47, 484, 1984.

L’hirondel, J., and L’hirondel, J.L., Nitrate and Man: Harmless or Beneficial, CABI,Wallingford, UK, 2001.


Lin, Z.Q., de Souza, M. Pickering, I.J., and Terry, N. Evaluation of the macroalga,muskgrass, for the phytoremediation of selenium-contaminated agriculturaldrainage water by microcosms, J. Environ. Qual. 31, 2104, 2002.

Lindroth, A., Verwijst, T., and Halldin, S. Water-use efficiency of willow: Variationwith season, humidity and biomass allocation, J. Hydrol. 156, 1, 1994.

Ling, B. Health impairments arising from drinking water polluted with domesticsewage and excreta in China, in Water Sanitation & Health, Chorus et al., Eds.,IWA Publishing, London, 43, 2000.

Lipson, S.M., and Stotsky, G. Effect of bacteria on the inactivation and adsorption onclay minerals of reovirus, Can. J. Microbiol. 31, 730, 1985.

Logan, A.J., Stevic, T.K., Siegrist, R.L., and Ronn, R.M. Transport and fate of Cryp-tosporidium parvum in intermittent sand filters, Water Res. 35, 4359, 2001.

Lowrance, R., Johnson, J.C. Jr., Newton, G.L., and Williams, R.G. Denitrification fromsoils of a year-round forage production system fertilized with liquid dairy ma-nure, J. Environ. Qual. 27, 1504, 1998.

Lucho-Constantino, C.A., Alvarez-Suarez, M., Beltran-Hernandez, R.I., Prieto-Garcıa,F., and Poggi-Varaldo, H.M. A multivariate analysis of the accumulation andfractionation of major and trace elements in agricultural soils in Hidalgo State,Mexico irrigated with raw wastewater. Environment International 31, 313, 2005.

Maas, E.V. Salt tolerance of plants, Appl. Agric. Res. 1, 12, 1986.Magesan, G.N., McLay, C.D.A., and Lal, V.V. Nitrate leaching from a free-draining

volcanic soil irrigated with municipal sewage effluent in New Zealand, Agric.Ecosyst. Environ. 70, 181, 1998.

Magesan, G.N., Williamson, J.C., Sparling, G.P., Schipper, L.A., and Lloyd-Jones, A.R.Decreased hydraulic conductivity in soils irrigated with wastewater field andlaboratory studies, Aust. J. Soil Res. 37, 391, 1999.

Magesan, G.N., Williamson, Yeates, G.W., and Lloyd-Jones, A.R. Wastewater C:N ratioeffects on soil hydraulic conductivity and potential mechanisms for recovery,Bioresource Technol. 71, 21, 2000.

Mahmood, B., Wall, G.L., and Russell J.M. A physical model to make short-termmanagement decisions at effluent-irrigated land treatment system, Agric. WaterManage. 58, 55, 2003.

Makkar, R.S., and Rochne, K.J. Comparison of synthetic surfactants and biosurfac-tants in enhancing biodegradation of polycyclic aromatic hydrocarbons. Envi-ron. Toxicol. 22, 2280, 2003.

Mapanda, F., Mangwayana, E.N., Nyamangara, J., and Giller, K.E. The effect of long-term irrigation using wastewater on heavy metal contents of soils under vegeta-bles in Harare, Zimbabwe, Agric. Ecosyst. Environ. 107, 151, 2005.

Marschner, B., and Kalbitz, K. Controls of bioavailability and biodegradability ofdissolved organic matter in soils, Geoderma, 113, 211, 2003.

Marschner, H. Mineral Nutrition of Higher Plants, 2nd ed., Academic Press, London,1995.

Masaphy, S., and Mandelbaum, R.T. Atrazine mineralization in slurries from soilsirrigated with treated waste water, Appl. Soil Ecol. 6, 283, 1997.

Mastrorilli, M., Katerji, N., and Rana, G. Productivity and water use efficiency ofsweet sorghum as affected by soil water deficit occurring at different vegetativegrowth stages, Eur. J. Agron. 11, 207, 1999.


Mawdsley, J.L., Bardgett, R.D., Merry, R.J., and Theodorou, M.K. Pathogens in live-stock waste, their potential for movement through soil and environmental pol-lution, Appl. Soil Ecol. 2, 1, 1995.

Mawdsley, J.L., Brooks, A.E., and Merry, R.J. Movement of protozoan pathogen Cryp-tosporidium parvum through three contrasting soil types, Biol. Fertil. Soils, 21,30, 1996.

McGechan, M., and Lewis, D. Sorption of phosphorus by soil, Part 1: Principles,equations and models, Biosyst. Eng. 82, 1, 2002.

McGechan, M. Sorption of phosphorus by soil, Part 2: Measurement methods, resultsand model parameter values, Biosyst. Eng. 82, 115, 2002.

McLaren, R.G., and Cameron, K.C. Soil, plant and fertilizer nitrogen/phosphorus, inSoil Science, McLaren, R.G. and Cameron, K.C., Eds., Oxford University Press,Aukland, New Zealand, 192, 1996.

McLeod, M., Aislabie, J., Smith, J., Fraser, R., Roberts, A., and Taylor, M. Viraland tracer movement through contrasting soils, J. Environ. Qual. 30, 2134,2001.

Meagher, R.B. Phytoremediation of toxic elemental and organic pollutants, Curr.Opin. Plant Biol. 3, 153, 2000.

Medema, G.J., Schets, F.M., Teunis, P.F.M., and Havelaar, A.H. Sedimentation of freeand attached Cryptosporidium oocysts and Giardia cysts in water, Appl. Environ.Microbiol. 64, 4460, 1998.

Meding, S.M., Morris, L.A., Hoover, C.M., Nutter, W.L., and Carbera, M.L. Denitrifica-tion at a long-term forested land treatment system in the Piedmont of Georgia,J. Environ. Qual. 30, 1411, 2001.

Menneer, J.C., McLay, C.D.A., and Lee, R. Effects of sodium-contaminated wastewateron soil permeability of two New Zealand soils, Aust. J. Soil Res. 39, 877, 2001.

Menzies, N.W., Skilton, J.A., and Guppy, C.N. Phosphorous storage on effluent irri-gated land, J. Environ. Qual. 28, 750, 1999.

Metzger, L., Yaron, B., and Mingelgrin, U. Soil hydraulic conductivity as affected byphysical and chemical properties of the effluents, Agronomie, 8, 771, 1983.

Moffat, A.J., Armstrong, A.T., and Ockleston, J. The optimization of sewage sludgeand effluent disposal on energy crops of short rotation hybrid poplar, BiomassBioenerg. 20, 161, 2001

Moore, R.S., Taylor, D.H., and Sturman, L.S. Adsorption of reovirus by minerals andsoils, Appl. Environ. Microbiol. 44, 852, 1982.

Mubiru, D.N., Coyne, M.S, and Grove, J.H. Mortality of Eshericia coli O157:H7 intwo soils with different physical and chemical properties, J. Environ. Qual. 29,1821, 2000.

Myers, B.J., Bond, W.J., Falkiner, R.A., O’Brien, N.D., Polglase, P.J., Smith, C.J., andTheiveyanathan S. Waga Waga Effluent Plantation Project—Technical Report,September 1994. CSIRO Division of Forestry, User Series No. 17, 123, 1994.

Nasser, A.M., Glozman, R., and Nitzan, Y. Contribution of microbial activity to virusreduction in saturated soil, Water Res. 36, 2589, 2002.

Natvig, E.E., Ingham, S.C., Ingham, B.H., Cooperband, L.R., and Roper, T.R.Salmonella enterica serovar Typhimurium and Escherichia coli contaminationof root and root and leaf vegetables grown in soil with incorporated bovinemanure, Appl. Environ. Microbiol. 68, 2737, 2002.


Negri, M.C., Gatliff, E.G., Quinn, J.J., and Hincham, R.R. Root development androoting at depths, in Phytoremediation Transformation and Control of Contam-inants, McCutheon, S.C., and Schnoor, J.L., Eds., John Wiley and Sons, Hoboken,NJ, 2003.

Nelson, P.N., Baldock, J.A., Clarke, P., Oades, J.M., and Churcham, G.J. Dispersedclay and organic mater in soil: their nature and associations, Aust. J. Soil Res. 37,298, 1999.

Nelson, P.N., Ladd, J.N., and Oades, J.M. Decomposition of 14C-labeled plant materialin a salt-affected soil, Soil Biol. Biochem. 28, 433, 1996.

Oakley, S., Greenwood, W.P., and Lee, M. Monitoring nitrogen and virus removal inthe vadose zone with suction lysimeters, in Proc. 10th Nortwest On-Site Wastewa-ter Treatment Short Course and Equipment Exhibition, Seattle, 20–21 September1999, University of Washington, Seattle, 211, 1999.

Oron, G., Goemans, M., Mannor, Y., and Feyen, J. Poliovirus distribution in the soil–plant system under reuse of secondary wastewater, Water Res. 29, 1069, 1995.

Overcash, M., Sims, R.C., Sims, J.L., and Nieman J.K.C. Beneficial reuse and sustain-ability: The fate of organic compounds in land-applied waste, J. Environ. Qual.34, 29, 2005.

Page, A.L., and Chang, C. Fate of wastewater constituents in soil and groundwater:Trace elements, in Irrigation with Recycled Water: A Guidance Manual, Asano,T., Ed., Lewis Publishers, Chelsea, MI, 1988.

Paranychianakis, N.V., and Chartzoulakis, K.S. Irrigation with saline water in Mediter-ranean region: Tolerance mechanisms, photosynthesis and management prac-tices, Agric. Ecosyst. Environ. 106, 171, 2005.

Paranychianakis, N.V., Chartzoulakis, K.S., and Angelakis, A.N. Influence of root-stock, irrigation, and recycled water on water relations and leaf gas exchangeof Soultanina grapevines, Exp. Environ. Bot. 52, 185, 2004a.

Paranychianakis, N.V., Aggelides, S., and Angelakis A.N. Influence of rootstock, irri-gation, and recycled water on growth and yield of Soultanina grapevines, Agric.Water Manage. 69, 13, 2004b.

Paranychianakis, N.V. Influence of rootstock, irrigation level and recycled water onthe growth, nutrition, and physiology of Soultanina grapevines, PhD thesis, Agri-cultural University of Athens, 2001.

Park, J., Zhao, X., and Voice, T.C. Biodegradation of nondesorbable naphthalene insoils, Environ. Sci. Technol. 35, 2734, 2001.

Parnaudeau, V., Nicolardot, B., Robert, P., Alavoine, G., Pages, J., and Duchiron, F.Organic matter characteristics of food processing industry wastewaters affectingtheir C and N mineralization in soil incubation, Biores. Techn. in press.

Pathak, H., and Rao, D.L.N. Carbon and nitrogen mineralization from added organicmatter in saline and alkaline soils, Soil Biol. Biochem. 30, 695, 1998.

Pedersen, J.A., Yeager, M.A., and Suffet, I.H. Xenobiotic organic compounds in runofffrom fields irrigated with treated wastewater, J. Agric. Food Chem. 51, 1360, 2003.

Pederson, G.A., Brink, G.E., and Fairbrother, T.E. Nutrient uptake in plant parts ofsixteen forages fertilized with poultry litter: Nitrogen, phosphorus, potassium,copper, and zinc, Agron. J. 94, 895, 2002.

Perttu, K.L. Environmental and hygienic aspects of willow coppice in Sweden,Biomass Bioenerg. 16, 291, 1999.


Phillips, I.R. Phosphorus sorption and nitrogen transformation in two soils treatedwith piggery wastewater, Aust. J. Soil Res. 40, 335, 2002.

Pignatello, J.J., and Xing, B.S. Mechanisms of slow sorption of organic chemicals tonatural particles, Environ. Sci. Technol. 30, 1, 1996.

Pignatello, J.J. Sorption dynamics of organics compounds in soils and sedimends, inReactions and Movement of Organic Chemicals in Soils, SSSA Special Publicationno. 22, 1989, 45.

Piotrowska, A., Iamarino, G., Rao, A.M., and Gianfreda, L. Short-term effects of olivemill waste water (OMW) on chemical and biochemical properties of a semiaridMediterranean soil, Soil Biol. Biochem. in press.

Powelson, D.K., and Mills, A.L. Bacterial enrichment at the gas-water interface of alaboratory apparatus, Appl. Environ. Microb. 62, 2593, 1996.

Powelson, D.K., and Mills, A.L. Transport of Esherichia coli in sand columns withconstant and changing water contents, J. Environ. Qual. 30, 238, 2001.

Quanrud, D.M., Carroll, S.M., Gerba, C.P., and Arnold, R.G. Virus removal duringsimulated soil-aquifer treatment, Water Res. 37, 753, 2003.

Ramirez-Fuentes, E.C., Luchlo-Constantino, E. Escamilla-Silva, and Dendooven, L.Characteristics, and carbon and nitrogen dynamics in soil irrigated with wastew-ater for different lengths of time, Bioresource Technol. 85, 179, 2002.

Rana, G., Rinaldi, M., and Introna, M. Volatilisation of substances after spreadingolive oil wastewater on the soil in a Mediterranean environment, Agric. Ecosyst.Environ. 96, 49, 2003.

Rattan, R.K., Datta, S.P., Chhonkar, P.K., Suribabu, K., and Singh, A.K. Long-termimpact of irrigation with sewage effluents on heavy metal content in soils, cropsand groundwater—A case study, Agric. Ecosys. Environ. 109, 310, 2005.

Redding, M.R., Biggs, A., Gardner, T., and Duperouzel, D. An overview of land appli-cation of piggery effluent-P using soil P chemistry and mass balance calculations,Aust. J. Soil Res. 40, 81, 2002.

Redding, M. Pig-effluent application can increase risk of P transport: Two case stud-ies, Aust. J. Soil Res. 39, 161, 2001.

Reddy, K.R., Khaleel, R., and Overcash, M.R. Behaviour and transport of microbialpathogens and indicator organisms in soils treated with organic wastes, J. Envi-ron. Qual. 10, 256, 1981.

Redman, J.A., Grant, S.B., Olson, T.M., Adkins, J.M., Jacksom, J.L., Castillo, M.S.,and Yanko, W.A. Physicochemical mechanisms responsible for the filtration andmobilisation of a filamentous bacteriophage in quartz sand, Water Res. 33, 43,1999.

Reed, S.C., Crites, R.W., and Middlebrooks, E.J. Natural Systems for Waste Manage-ment and Treatment. 2nd Ed., McGraw-Hill, New York, 1995, 433.

Reemtsma, T., Savric, I., and Jekel, M. A potential link between the turnover of soilorganic matter and the release of aged organic contaminants, Environ. Toxicol.Chem. 22, 760, 2003.

Reilley, K.A., Banks, M.K., and Schwab, A.P. Dissipation of polycyclic aromatic hy-drocarbons in the rhizosphere, J. Environ. Qual. 25, 212, 1996.

Renella, G., Chaudri, A.M., and Brookes, P.C. Fresh additions of heavy metals do notmodel long-term effects on microbial biomass and activity, Soil Biol. Biochem.34, 121, 2002.


Rhoades, J.D. Intercepting, isolating, and reusing drainage waters for irrigationto conserve water and protect water quality, Agric. Water Manage. 16, 37,1989.

Rietz, D.N., and Haynes, R.J. Effects of irrigation-induced salinity on soil microbialfauna, Soil Biol. Biochem. 35, 845, 2003.

Rockwood, D.L., Pisano, S.M., Sprinkle, R.R., and Leavell, A. Stormwater remediationby tree crops in Florida, in Proc. 4th Biennial Stormwater Research Conference,October 18–20, Clearwater, FL, Southwest Florida Water Management District,Brooksville, FL, 1995.

Rowell, D.M., Prescott, C.E., and Preston, C.M. Decomposition and nitrogen min-eralization from biosolids and other organic materials: Relationship with initialchemistry, J. Environ. Qual. 30, 1401, 2001.

Rugh, C.L., Wilde, H.D., Stack, N.M., Thompson, D.M., Summers, A.O., and Meagher,R.B. Mercuric ion reduction and resistance in transgenic Arabidopsis thalianaplants expressing a modified bacterial merA gene, Proc. Natl. Acad. Sci. USA 93,3182, 1996.

Rumberger, A., and Marschner, P. 2-Phenylethylisothiocyanate concentration and mi-crobial community composition in the rhizosphere of canola, Soil Biol. Biochem.35, 445, 2003.

Ryan, J.A., Bell, R.M., Davidson, J.M., and Connor, G.A. Plant uptake of non-ionicorganic chemicals from soils, Chemosphere 17, 2299, 1988.

Salt, D.E., Smith, R.D., and Raskin, I. Phytoremediation, Annu. Rev. Plant Physiol.Mol. Biol. 49, 646, 1998.

Sandermann, H., Jr. Plant metabolism of xenobiotics, Trends Biol. Sci. 17, 82, 1992.Sanderson, M.A., Jones, R.M., McFarland, M.J., Stroup, J., Reed, R.L., and Muir, J.P.

Nutrient movement and removal in a switchgrass biomass-filter strip systemtreated with dairy manure, J. Environ. Qual. 30, 210, 2001.

Sanyal, S.K., and de Datta, S.K. Chemistry of phosphorus transformations in soil,Adv. Soil Sci. 16, 1, 1991.

Schilke-Gartley, K.L., and Sims, J.T. Ammonia volatilization from poultry manureamended soil, Biol. Fertil. Soils, 16, 5, 1993.

Schipper, L.A., Williamson, J.C., Kettles, H.A., and Speir, T.W. Impact of land-appliedtertiary-treated effluent on soil biochemical properties, J. Environ. Qual. 25,1073, 1996.

Schirmer, R. Aspekte der Pflanzenzuchtung schnellwachsender, in Baumarten furEnergiewalder, Burger et al., Eds, Schnellwachsende Baumarten, ihr Anbau undihre Verwertung, 6, 1996.

Schoenborn, C., and Duncan, M. Growth and nutrient uptake by effluent-irrigatedeucalypts, Agriculture Notes, AG0905, 1, November 2001.

Schwarzenbach, R.P, Gschwend, P.M., and Imboden, D.M. Environmental OrganicChemistry, Wiley Interscience, New York, 1993.

Scott, J., Beydoun, D., Amal, R., Low, G., and Camtlie, J. Landfill management,leachate generation, and leach testing of solid wastes in Australia and overseas,Crit. Rev. Environ. Sci. Technol. 35, 239, 2005.

Seol, Y., and Lee, L.S. Effect of dissolved organic matter in treated effluents onsorption of atrazine and prometryn by soils, Soil Sci. Soc. Am. J. 64, 1976, 2000.

Sharpe, R.R., and Harper L.A. Ammonia and nitrous oxide emissions from sprinklerirrigation applications of swine effluent, J. Environ. Qual. 26, 1703, 1997.


Sharpe, R.R., and Harper, L.A. Nitrous oxide and ammonia fluxes in a soybean fieldirrigated with swine effluent, J. Environ. Qual. 31, 524, 2002.

Shrive, S.C., and McBridge, R.A. Physiologic responses of red maple saplings to sub-irrigation with an untreated municipal landfill leachate, Waste Manage. Res. 13,219, 1995.

Siciliano, S.D., and Germida, J.J. Mechanisms of phytoremediation: Biochemical andecological interactions between plants and bacteria, Environ. Rev. 6, 65, 1998.

Singer, A.C., Crowley, D.E., and Thompson, I.P. Secondary plant metabolites in phy-toremediation and biotransformation, Trends in Biotechnology, 21, 123, 2003.

Smith, C.J., and Bond, W.J. Losses of nitrogen from an effluent irrigation plantation,Aust. J. Soil Res. 37, 371, 1999.

Smith, C.J., Freney, J.R., and Bond, W.J. Ammonia volatilization from soil irrigatedwith urban sewage effluent, Aust. J. Soil Res. 34, 789, 1996.

Snow, V.O., Bond, W.J., Myers, B.J., Theiveyanathan, S., Smith, C.J., and Benyon, R.G.Modelling the water balance of effluent-irrigated trees, Agric. Water Manage. 39,47, 1999.

Sparling, G.P., Schipper, L.A., and Russel, J. Changes in soil properties after appli-cation of dairy factory effluent to New Zealand volcanic ash and pumice soils,Aust. J. Soil Res. 39, 505, 2001.

Stanbridge, H.H. History of Sewage Treatment in Britain, Part 5, Land Treatment,Institute of Water Pollution Control, Maidstone, Kent, England, 1976.

Stanhill, G., Water use efficiency, Adv. Agron. 39, 53, 1986.Stenger, R., Barkle, G.F., and Burgess, C.P. Mineralization and immobilization of C

and N from dairy farm effluent (DFE) and glucose plus ammonium chloridesolution in three grassland topsoils, Soil Biol. Biochem. 33, 1037, 2001.

Sui, Y.B., and Thompson, M.L. Phosphorus sorption, desorption, and buffering ca-pacity in a biosolids-amended mollisol, Soil Sci. Soc. Am. J. 64, 164, 2000.

Summer, M.E. Sodic soils: New perspectives, Aust. J. Soil Res. 31, 683, 1993.Sun, H., Tateda, M., Michihiko, I., and Fujita, M. Short- and long-term sorp-

tion/desorption of polycyclic aromatic hydrocarbons onto artificial solids: Effectsof particle and pores size and organic matters, Water Res. 37, 2960, 2003.

Sundaravadivel, M., and VIgneswaran S. Constructed wetlands for wastewater treat-ment, Crit. Rev. Environ. Sci. Technol. 31, 351, 2001.

Tanji, K.K., and Kielen N.C. Agricultural Drainage Water Management in Arid andSemi-Arid Areas, FAO Irrigation and Drainage Paper 61, Rome, 2002.

Terry, N., and Lin, Z.Q. Managing High Selenium in Agricultural Drainage Waterby Agroforestry Systems: Role of Selenium Volatilization, Rep., California StateDepartment of Water Resources, Sacramento, 1999.

Terry, N., Carlson, C., Raab, T.K., and Zayed, A.M. Rates of selenium volatilizationamong crop species, J. Environ. Qual. 21, 341, 1992.

Terry, N., Zayed, A.M., de Souza, M.P., and Tarum, A.S. Selenium of higher plants,Annu. Rev. Plant Physiol. Plant Mol. Biol. 51, 401, 2000.

Theiveyanathan, S., Benyon, R.G., Marcar, N.E., Myers, B.J., Polglase, P.J., andFalkiner, R.A. An irrigation-scheduling model for application of saline waterto tree plantations, For. Ecol. Manage. 193, 97, 2004.

Thompson, S.S., Flury, M., Yates, M.V., and Jury, W.A. Role of the air-water-solidinterface in bacteriophage sorption experiments, Appl. Environ. Microbiol. 64,304, 1999.


Tomer, M.D., and Burkart, M.R. Long-term effects of nitrogen fertilizer use on groundwater nitrate in two small watersheds, J. Environ. Qual. 32, 2158, 2003.

Tomer, M.D., Schipper, L.A., Knowles, S.F., Rijkse, W.C., McMahon, S.D., Smith, C.T.,Thorn, A., and Charleston, T. A land based system for treatment of municipalwastewater at Whakarewarewa forest, New Zealand. N. Z. For. Res. Inst. Bull.199, 28, 1997.

Trinsoutrot, I., Recous, S., Bentz, B., Lineres, M., Cheneby, D., and Nicolardot, B.Biochemical quality of crop residues and carbon and nitrogen mineralizationkinetics under nonlimiting nitrogen conditions, Soil Sci. Soc. Am. J. 64, 918,2000.

Tzanakakis V.A., Paranychianakis, N.V., Kyritsis, S., and Angelakis, A.N. Evaluationof wastewater treatment and biomass production by SRS using different plantspecies, Water Sci. Technol. Water Supply, 3, 185, 2003.

Urbance, J.W., Cole, J., Saxman, P., and Tiedje, J.M. BSD: The biodegradative straindatabase, Nucleic Acids Res. 31, 152, 2003.

U.S. EPA, A History of Land Application as a Treatment Alternative, U.S. Environ-mental Protection Agency, Office of Water Program Operations, Wanshington,DC, 1979.

U.S. EPA, Land Treatment of Municipal Wastewater: Process Design Manual, U.S. En-vironmental Protection Agency, Center for Environmental Research Information,Cincinati, OH 1981.

U.S. EPA, Clean Water Act, U.S. Environmental Protection Agency, Washington, DC,1993.

U.S. EPA, Guidelines for Water Reuse, EPA/625/R-04/108, Camp Dresser & McKee,Inc., for the U.S. Environmental Protection Agency, Washington, DC, 2004.

van Guyk, S., Siegrist, R., Logan, A., Masson, S., Fischer, E., and Figueroa, E. Hydraulicand purification behaviors and their interactions during wastewater treatmentin soil infiltration systems, Water Res. 35, 953, 2001.

van Guyk, S., Siegrist, R.L., Lowe, K., and Harvey, R.W. Evaluating microbial pu-rification during soil treatment of wastewater with multicomponent tracer andsurrogate tests, J. Environ. Qual. 33, 316, 2004.

van Veen, J.A., and Kuikman, P.J. Soil structure aspects of decomposition of organicmatter by microorganisms, Biogeochemistry 11, 213, 1990.

Vermes, L. Special poplar plantation for water pollution control in agricultural areas,Hrvaske Vode, 4, 143, 1996.

Victorian EPA, Guidelines for Wastewater Irrigation, Publication 168, Victorian Envi-ronment Protection Authority, Australia, 1993.

Vilker, V.L. Simulating virus movement in soils, in Modeling wastewater RenovationLand Treatment, Iskandar, I.K., Ed., Wiley and Sons, New York, 223, 1981.

Vulkan, R., Zhao, F.J., Bardosa-Jefferson, V.L., Preston, S., Paton, G.I., Tipping, E.,and McGrath, S.P. Copper speciation and impacts on bacterial biosensors in thepore water of copper-contaminated soils, Environ. Sci. Technol. 34, 5115, 2000.

Waly, T.M., Abd Elnaim, E.M., Omran, M.S., and Nashar, M.B. Effect of sewage wateron chemical properties and heavy metals content of El Gadal El Asfar sandysoils, Biol. Wastes 22, 275, 1987.

Wan, J.M., Wilson, J.L., and Kieft, T.L. Influence of the gas–water interface on trans-port of microorganisms through unsaturated porous media, J. Appl. Environ.Microbiol. 60, 509, 1994.


Wang, M.J., and Jones, K.C. Behavior and fate of chlorobenzenes in spiked andsewage sludge-amended soils, Environ. Sci. Technol. 28, 1843, 1994.

Watkins, J.W., Sorenson, D.L., and Sims, R.C. Volatilization and mineralization ofnaphthalene in soil-grass microcosms, ACS Sympos. Ser. 563, 123, 1994.

Weih, M., and Nordh, N.E. Characterizing willows for biomass and phytoremediation:Growth, nitrogen and water use of 14 willow clones under different irrigationand fertilization regimes, Biomass Bioenerg. 23, 397, 2002.

WHO, Environmental Health Criteria 37: Aquatic (Marine and Freshwater) Biotox-ins, World Health Organization, Geneva, 1986.

Wolfe, N.L., and Hoehammer, C.F. Enzymes used by plants and microorganisms todetoxify organic compounds, in Phytoremediation Transformation and Controlof Contaminants, McCutheon S.C. and Schnoor, J.L., Eds., John Wiley and Sons,Hoboken, NJ, 159, 2003.

Woodard, K.R., French, E.C., Sweat, L.A., Graetz, D.A., Sollenderger, L.E., Macoon,B., Portier, K.M., Wade, B.L., Rymph, S.J., Prine G.M., and van Horn, H.H. Ni-trogen removal and nitrate leaching for forage systems receiving dairy effluent,J. Environ. Qual. 31, 1980, 2002.

Yamamoto, F., Sakata, S., and Tenazawa, K. Physiological, morphological andanatomical responses of Fraxinus mandshurica seedlings to flooding, TreePhysiol. 15, 713, 1995.

Yates, M.V., and Gerba, C.P. Microbial considerations in wastewater reclamation andreuse, in Wastewater Reclamation and Reuse, Asano, T., Ed., Technomic, Lan-caster, Pa, 1998.

Yeager, J.G., and O’Brien, R.T. Enterovirus inactivation in the soil, Appl. Environ.Microbiol. 38, 694, 1979.

Zaman, M., Cameron, K.C., Di, H.J., and Inubushi, K. Changes in mineral N, microbialbiomass and enzyme activities in different soil depths after surface applicationsof dairy shed effluent and chemical fertilizer, Nutr. Cycl. Agroecosysto. 63, 275,2002.

Zhuang, J., and Jin, Y. Virus retention and transport through Al-oxide coated sandcolumns: Effects of ionic strength and composition, J. Contam. Hydrol. 60, 193,2003.

treatment of wastewater with slow rate systems: a review of

Documents