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Critical Reviews in Environmental Science and Technology, 40:561–661, 2010Copyright © Taylor & Francis Group, LLCISSN: 1064-3389 print / 1547-6537 onlineDOI: 10.1080/10643380802471076

Contaminant Removal Processes inSubsurface-Flow Constructed Wetlands:

A Review

JOAN GARCIA,1 DIEDERIK P. L. ROUSSEAU,2 JORDI MORATO,3

ELS LESAGE,4 VICTOR MATAMOROS,5 and JOSEP M. BAYONA5

1Environmental Engineering Division, Hydraulics, Maritime and Environmental EngineeringDepartment, Universitat Politecnica de Catalunya, Barcelona, Spain

2Department of Environmental Resources, UNESCO-IHE Institute for Water Education,Delft, The Netherlands

3Laboratory of Health and Environmental Microbiology, Universitat Politecnica deCatalunya, Terrassa, Spain

4Laboratory of Analytical Chemistry and Applied Ecochemistry, Ghent University,Ghent, Belgium

5Environmental Chemistry Department, IIQAB-CSIC, Barcelona, Spain

The main contaminant removal processes occurring in subsurface-flow constructed wetlands treating wastewater are reviewed. Redoxconditions prevailing in the wetlands are analyzed and linked tocontaminant removal mechanisms. The removal of organic mat-ter and its accumulation in the granular medium of the wetlandsare evaluated with regard to particulate and dissolved componentsand clogging processes. The main biological processes linked to or-ganic matter transformation—aerobic respiration, denitrification,acid fermentation, sulfate reduction, and methanogenesis—are re-viewed separately. The processes of removal of surfactants, pes-ticides and herbicides, emergent contaminants, nutrients, heavymetals and faecal organisms are analyzed. Advances in wetlandmodeling are presented as a powerful tool for understanding multi-ple interactions occurring in subsurface-flow constructed wetlandsduring the removal of contaminants.

Els Lesage is currently affiliated with the Flemish Land Agency, Manure Bank, Ganzen-dries 149, 9000 Ghent, Belgium.

Address correspondence to Joan Garcıa, Environmental Engineering Division, Hydraulics,Maritime and Environmental Engineering Department, Universitat Politecnica de Catalunya,c/ Jordi Girona 1-3, Modul D-1, 08034 Barcelona, Spain; E-mail: [email protected]

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562 J. Garcıa et al.

KEY WORDS: reed beds, granular medium, hydraulic loadingrate, organic loading rate

INTRODUCTION

In recent years, extensive (or “natural”) wastewater treatment systems havebeen set up all over the world as an alternative to conventional intensivesystems for the sanitation of small communities. These systems have lowenergy requirements and are easy to operate and maintain, which makesthem suitable for wastewater treatment where land availability and landprices are not limiting factors (Garcıa et al., 2001; Puigagut et al., 2007a).

Subsurface-flow (SSF) constructed wetlands (CWs) are one of the mostcommon types of extensive wastewater systems used throughout the world.SSF CWs consist of beds that are usually dug into the ground, lined, filled witha granular medium, and planted with emergent macrophytes. Wastewaterflows through the granular medium and comes into contact with biofilmsand plant roots and rhizomes. Contaminants are removed by a wide rangeof processes. SSF CWs are therefore designed to simulate the processes thatoccur in natural wetlands but in a more controlled environment.

SSF CWs can be classified as either horizontal flow or vertical flowsystems. In a typical horizontal flow system, wastewater is maintained ata constant depth and flows horizontally below the surface of the granularmedium (Vymazal, 2005a). In vertical flow systems, wastewater is distributedover the surface of the wetland and trickles downward through the granularmedium (Brix & Arias, 2005). Vertical systems can be sorted in at least fourtypes depending on the hydraulic regimes: unsaturated flow (like conven-tional trickling filters), permanently saturated flow, intermittent unsaturatedflow, and flood and drain wetlands. With the exception of permanently sat-urated flow systems, the mode of functioning of the other vertical systemsimproves the aeration of the bed in comparison to horizontal flow wetlands.Consequently, vertical beds operate generally under more oxidized condi-tions than horizontal beds and are far more efficient, as they can treat highercontaminant loads. In addition, vertical systems produce nitrified effluents,whereas horizontal systems often have very limited nitrification capabilities(Brix & Arias, 2005). In practice, these two types of systems are often com-bined to form hybrid wetlands, which provide higher removal efficiency. SSFCWs are mainly designed to treat primary settled wastewater, although theyare also commonly used to improve the quality of secondary effluents.

During the last two decades, a large number of books have been pub-lished on technical and scientific aspects of constructed wetlands, includ-ing USEPA (1988, 1993, 2000), WPCF (1990), Reed et al. (1995), Wissingand Hofmann (1995), Cooper et al. (1996), Kadlec and Knight (1996),Vymazal et al. (1998), Campbell and Ogden (1999), Kadlec et al. (2000),Dias and Vymazal (2003), WERF (2006), Kadlec and Wallace (2008), and

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Processes in Constructed Wetlands 563

Vymazal and Kropfelova (2008). The book by Kadlec and Knight (1996)has been used as a key reference for many university courses and otherspecialized courses, and the recently published second edition (Kadlec andWallace, 2008) will be a key reference now and in the coming years. Thedescriptions of contaminant removal processes in SSF CWs given in thesebooks are often based on conclusions drawn from experiments in naturalwetlands or other wastewater treatment processes. This is because SSF CWsare very complex reactors in which many processes take place simultane-ously and are therefore difficult to study. Contaminants are removed fromwastewater in SSF CWs by physical, chemical, and biological processes, andthere is no single pathway that describes the complete range of processesinvolved in the removal of a given contaminant. Due to this complexity,constructed wetlands have been considered “black box” systems since theirintroduction.

However, recent years have seen a dramatic increase in the numberof research groups studying the processes involved in contaminant removalin SSF CWs. Similarly, the volume of knowledge acquired and informationpublished in SCI journals has increased considerably since 2000. Therefore,the main objective of this paper is to review the most recent advances incontaminant removal processes in SSF CWs considering a wide range ofpollutants. This is the main difference between the current review and otherprevious reviews on this topic.

In this paper, removal processes are reviewed for the following groupsof contaminants: organic matter, specific organic contaminants, nutrients,heavy metals and faecal microorganisms. The prevailing redox conditionsin SSF CWs have a strong effect on removal mechanisms and will be an-alyzed first, followed by a consideration of the removal of organic matter,its accumulation in the granular medium, and its relationship with cloggingprocesses. The main biological processes associated with organic mattertransformations are reviewed separately. We then review the removal pro-cesses of specific organic contaminants, such as surfactants, pesticides andherbicides, and emergent contaminants. Nitrogen and phosphorus transfor-mation reactions are then examined. We then look at the physicochemicaland biological processes involved in the removal of heavy metals, and con-sider the removal of faecal microorganisms and conduct a detailed review ofthe processes associated with faecal indicators, bacterial pathogens, viruses,protozoans and parasites. Finally, we review advances in wetland modeling,which is a powerful tool for understanding the multiple interactions thatoccur in SSF CWs during contaminant removal processes.

REDOX POTENTIAL BEHAVIOR

SSF CWs are distinguished from other wastewater treatment processes bythe simultaneous co-existence of areas with different redox status at the

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macro- and micro-scales. This intrinsic property allows different physico-chemical and biochemical processes to occur at the same time (Ojeda et al.,2008). The heterogeneous distribution of redox conditions in SSF CWs iscaused by several factors, including the presence of plant rhizospheres andfluctuations in the water level due to evapotranspiration and tidal flow. Thebiogeochemical cycles of elements such as C, N and S can be almost closedin the reactor itself thanks to the different redox conditions present. In ad-dition to spatial changes, short-term and long-term temporal variability ofredox conditions also occurs in SSF CWs. For example, Wießner et al. (2005)reported changes in redox status closely related to the daily light cycle.

It is difficult to study the internal processes involved in contaminantremoval in SSF CWs due to the heterogeneous oxidation-reduction condi-tions in these systems. Therefore, research into changes in redox, oxygenand other related variables at the macro- and micro-scales and associatedtemporal variations is vital to understanding the processes that occur in wet-land systems. In this section, spatial changes of redox at the macro- andmicro-scales are separately reviewed. Also the temporal variability of thisparameter is described.

Macro-scale Redox Potential

Macro-scale redox potential changes along the length of horizontal SSF CWsare characterized by a general increase, so stronger reducing conditions arefound near the inlet than at the outlet (Garcıa et al., 2003a; Headley et al.,2005). In theory, there should be no significant changes in redox conditionsacross the width of wetlands at the macro-scale, at least during the start-upof a given system; however, local differences in plant distribution, preferen-tial flow, short-circuiting and heterogeneous solids accumulation over timecan lead to changes in redox potential across wetland systems. Garcıa et al.(2003a) produced redox profiles for two 187.5 m2 horizontal SSF CWs treat-ing domestic wastewater from a hotel and recorded values between −200and +150 mV, depending on the location of the measurement point (seeFigure 1). The same authors found no significant differences across the widthof the wetlands and therefore suggested that redox potential measurementscould be used to detect lateral preferential flow. They also observed that re-dox potential decreased considerably with depth, because the processes thatprovide oxygen (surface reaeration and plant release) occur mainly in theupper layers of the wetland media. In contrast, Headley et al. (2005) studiedpilot horizontal SSF CWs of 8.8 m2 fed with primary treated municipal efflu-ent and found that the vertical redox potential gradients were insignificant.These authors measured the redox potential at three water depths in five lo-cations along the length of the wetland and recorded values between -88 and+222 mV. The absence of vertical redox gradients was associated with thesubstantial degree of vertical mixing detected by a tracer test. Garcıa et al.

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Bed 2EH, mV

-200 -100 0 100 200

ABC

Bed 1

-200 -100 0 100 2000

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Wat

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, cm 0

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March March

May May

June June

FIGURE 1. Redox potential variations with water depth in two horizontal SSF CWs (bed 1:unplanted; bed 2: planted) in Vilagrassa, Lleida (Spain) for three different periods (March,May, and June). The letters represent the location of sampling points: close to the inlet (A),in the middle (B), and close to the outlet (B). From Garcıa et al. (2003a), with permission.

(2003a) also conducted a tracer test and found that the degree of verticalmixing was very low, which was consistent with the vertical redox potentialgradients observed.

The results of the two studies cited above suggest that vertical macro-scale redox gradients are related to internal mixing, which is in turn a func-tion of the turbulence inside the reactor, depending on the flow rate perunit of cross-sectional area and the size of the granular medium. Headleyet al. (2005) suggested that other driving forces such as the transpiration-pump effect may also cause vertical mixing. All of these properties high-light the intra-technological complexity of horizontal wetlands, and indicatethat caution must be observed with categorical statements about any givenmechanism.

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Kayser et al. (2003) carried out a detailed study of the evolution ofeffluent redox potential and related parameters over time in a full-scalevertical SSF CW. The wetland treated lagoon effluent and had a surfacearea of 2,500 m2 and an average depth of 0.8 m. It was clearly observedover a period of two years that low redox potential (from −150 to 0 mV)coincided with low nitrification rates, whereas high redox potential (from 0to +200 mV) coincided with a greater degree of nitrification. These authorsrecommended the continuous measurement of redox potential in verticalSSF CW effluents because it is closely related to oxygen supply and removalefficiency, and is therefore a very reliable indicator of the state of the wetland.Kayser et al. (2003) also evaluated the daily evolution of redox potential.The redox potential and the oxygen concentration in the effluent decreasedinstantly after each feeding event. Minimum redox potential values alwayscoincided with maximum ammonium concentrations. On days when higherhydraulic loading rates (HLR) were recorded, the redox potential maximabefore each new feeding decreased continuously throughout the day, andthe ammonium concentration maxima increased continuously.

Micro-scale Redox Potential

Bezbaruah and Zhang (2004) studied changes in redox potential and dis-solved oxygen (DO) at the micro-scale in an experimental vertical SSF CW.The wetland was planted with Scirpus validus, had a surface area of 0.028 m2,and was fed with primary settled wastewater. Measurements were taken us-ing homemade microelectrodes attached to a micromanipulator. The redoxpotential at the surface of lateral roots ranged from +250 to +317 mV anddecreased to the values of the bulk solution (−54 to +14 mV) at distancesof more than 4.5 mm. Redox values at the surface of the main roots rangedfrom +24 to +124 mV and reached the values of the bulk solution at dis-tances of almost 3 mm. Increased redox potential close to the surface of thelateral roots correlated with the presence of oxygen, which was found toform a layer with a thickness of approximately 1 mm around the roots. Thisoxygen-rich layer was not found around the main roots. The authors suggestthat there may have been a small amount of oxygen close to the surfaceof the main roots but at a level below the detection limit of the oxygenmicroelectrode used. Although the measurements were taken in a samplingzone devoid of media (which might have affected the flow pattern and thebiofilm growth), there are clearly strong redox gradients at the micro-scalethat allow biochemical processes with very different environmental require-ments to occur simultaneously in very small areas. Bezbaruah and Zhang(2004) also found that feeding wastewater with different BOD concentra-tions produced the same DO concentration in the bulk water but a differentconcentration in the root-induced DO layers. DO content at the root surface

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increased nominally with oxygen demand, although the concentration didnot rise indefinitely.

Munch et al. (2005) considered redox changes at the micro-scale andstudied the distance from the roots of Phragmites australis at which nitrifica-tion or denitrification was detectable. Tests were conducted in experimental12 L SSF CWs fed with artificial wastewater in batch mode. They found thatoxygen saturation in an unplanted control wetland was comparable to thelevels recorded in a planted reactor at a root distance of 40 mm. There-fore, the oxygen layer around the roots was clearly thicker than the 1 mmreported by Bezbaruah and Zhang (2004). This suggests that the microgradi-ents induced by plants can vary dramatically between systems according toa wide range of properties, such as the load, the macrophyte species, andthe environmental conditions.

Wießner et al. (2005) observed daily changes in redox potential andpH in a laboratory-scale 0.06 m2 SSF CW fed with synthetic wastewater andoperated with complete mixing of the pore water to balance the macro-scale gradients. The redox values began to increase at about the same timeas it grew light in the early morning and decreased when the light faded.For example, on one day in August, the redox changed from −240 mV to+209 mV in a period of three and a half hours. The pH evolved accordingto the same pattern but the change was less dramatic. Interestingly, the levelof DO in the pore water throughout the day did not change in responseto variations in the redox potential, which the authors attributed to rapidoxygen consumption in the system.

MECHANISMS INVOLVED IN ORGANIC MATTER REMOVAL

Organic matter includes dissolved as well as particulate components, com-monly denominated as DOM (dissolved organic matter) and POM (particulateorganic matter). The mechanisms involved in the removal of DOM and POMare different, and for this reason in this paper are considered separately.

SSF CWs are essentially fixed-biofilm reactors in which organic matter isremoved through interactions between complex physical, chemical, and bio-chemical processes. Many studies have shown that organic matter removalrates are not clearly related to changes in water temperature, which suggeststhat the principal removal mechanisms are physicochemical and then bio-logical (McNevin et al., 2000). Influent particulate organic matter (POM) ismainly retained by purely physical processes such as filtration and sedimen-tation. Retained POM accumulates or disintegrates and undergoes hydrolysis,which generates dissolved organic compounds that can be degraded by dif-ferent pathways that occur simultaneously in a given wetland. In a wetlandwith prevailing aerobic conditions, organic matter is oxidized by organismsthat use oxygen as an electron acceptor (aerobic respiration). In systems

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with anoxic conditions, organic matter is oxidized by aerobic organisms thatcan use nitrate as an electron acceptor (denitrification) in addition to oxygen(Metcalf and Eddy, 2003). There is very limited potential for aerobic andanoxic conditions in horizontal SSF CWs because strong reducing conditionsoften prevail. Therefore, horizontal systems are generally considered to beanaerobic treatment systems (Garcıa et al., 2004a, 2005a; Marahatta, 2004). Incontrast, vertical SSF CWs often show prevalent oxidized conditions and aretherefore considered aerobic/anoxic treatment systems (Kayser et al., 2003).

Anaerobic conditions present in horizontal SSF CWs permit the develop-ment of many groups of bacteria, several intermediate steps, and alternativebiochemical pathways (Marahatta, 2004). Soluble monomers formed by hy-drolysis are converted into volatile fatty acids (VFAs) (i.e., acetate) through acomplicated multistage process (fermentation). Acetate and other intermedi-ates can be degraded by organisms that use sulfate as a final electron acceptor(sulfate reduction). Acetate can also be degraded by methanogenic organ-isms to produce methane and carbon dioxide (methanogenesis). The relativeimportance of the different biochemical pathways for removing organic mat-ter depends primarily on the redox conditions (Garcıa et al., 2004a, 2005a).Heterotrophic microorganisms using different metabolic pathways competefor electron donors (organic matter), and their success is largely dictated bythe amount of energy released when different electron acceptors are used.

In this section, we discuss in detail the processes related to organicmatter removal, focusing particularly on biochemical pathways. The firstpart of the section deals with internal and external organic matter loads,their sources, and relative importance. As shall be seen, POM retention andaccumulation are very important processes linked with clogging, which is theworst operational problem observed in SSF CWs. Aerobic respiration of DOMis a biochemical pathway strongly dependent on oxygen availability andfluxes, and will also be discussed. Because both physical and plant-mediatedoxygen fluxes seem to be not enough to allow the removal of all organicmatter present in a typical primary effluent through aerobic repiration, thethis section also deals extensively with anaerobic processes that are veryimportant in horizontal SSF CWs.

Internal and External Organic Matter Loads

SSF CWs treating wastewater receive organic matter from two differentsources: the organic matter contained in the wastewater itself, which canbe referred to as external loading, and the organic matter from the primaryproduction in the wetland itself, which constitutes internal loading.

Constructed wetlands are designed to remove the externally loaded or-ganic matter, but the internal loading must also be taken into account becauseit produces a background concentration that determines the maximum effi-ciency of the systems. The background concentration is usually represented

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as C∗. In terms of BOD, C∗ is often assumed to be around 3 mg L−1 (Rousseauet al., 2004a). Plants seem to be the largest internal source of organic matter,and below-ground production is thought to generate more organic matterthan above-ground production because the roots, rhizomes, leachates andexudates release matter directly into the granular medium (Tanner et al.,1998). Internal microbial autotrophic production of organic matter has notbeen studied, but it is likely to be a lesser source than plant production.

As plants grow, die, and decay, dissolved organic matter can leachinto the water. Bacterial decomposition of plant detritus converts particulateorganic matter into humic substances that increase the bulk dissolved organicmatter pool (Pinney et al., 2000; Quanrud et al., 2001). Fungi may also havean effect on plant decomposition, but we are not aware of any study thatdeals with this topic in constructed wetlands. Abiotic dissolution of plantmaterial also increases the concentration of dissolved organic matter. Pinneyet al. (2000) used a laboratory apparatus to simulate organic carbon leachingfrom dried cattails and observed that higher hydraulic retention time (HRT)increased the concentration of effluent dissolved organic carbon (DOC) (ofplant origin). However, mass-balance analysis revealed that only 5–8% of theC released by plants was leached and measured as DOC. The amount of Cretained in the plant biomass ranged from 45 to 60%, and the unaccountedorganic carbon (35–60%) exited the system as particulated organic C orwas converted into CO2 and CH4 through microbial activity. Therefore, theamount of plant material leached as residual DOC was smaller than the othercarbon pools in this study. The internal loading of organic matter only hasa significant effect on mass balance in SSF CWs used to treat wastewaterswith low organic carbon loading, such as secondary effluents. Alvarez andBecares (2006) speculated on the role of plant litter decomposition on theclogging rates of SSF CWs from results obtained in a study conducted ina surface flow wetland. They estimated the plant decomposition rates bymeans of the litter bag technique in a 44 m2 wetland planted with Typhalatifolia and estimated that more than 30% of the initial mass of plant detritusremained in the system after one year.

Most full-scale SSF CWs are designed to remove organic matter fromwastewater (external loads). The organic matter in municipal wastewatercomprises mixtures of components with sizes that range from less than0.001 µm (dissolved) to more than 100 µm (settleable) (Levine et al., 1991).The individual components of these mixtures can be degraded or removedfrom wetlands at different rates, which creates a distribution of the removalrate constant values across the various mass fractions of the mixture (Kadlec,2003). This is illustrated clearly by the different levels of wastewater partic-ulate and dissolved organic matter (detected by filtration through a standardfibre filter) along the length of horizontal SSF CWs. Most of the particulateorganic matter is removed close to the inlet, and the remaining dissolvedorganic matter is removed more slowly along the entire length of the beds

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(Garcıa et al., 2005a). Garcıa et al. (2005a) studied the variations in total or-ganic carbon (TOC) levels along the length of two 55 m2 pilot horizontal SSFCWs and used a Coulter counter to take simultaneous water particle counts.In general, greater than 50% of the TOC removal took place in the first quar-ter of the wetland length; the simultaneous Coulter counter measurementshowed that the number of particles larger than 0.7 µm (the smallest sizethat the Coulter was able to measure) in the same location was between 20and 30% of the initial figure (see Figure 2). In the area halfway along thelength of the beds, the number of particles had already fallen to less than5% of the initial figure. From the three-quarter point to the effluent, the TOCremoval efficiency was between 10 and 30%, most of which correspondedto the removal of dissolved organic matter.

The parameters used to quantify organic matter in wastewater treatmentsystems have traditionally been based on analytical procedures that lumpindividual chemical compounds into overall concentrations, such as totalsuspended solids, chemical oxygen demand, and total organic carbon. How-ever, since the end of the 1980s, organic matter fractionation has been usedincreasingly in studies of conventional systems because it is a requisite forsimulations such as activated sludge models (Henze et al., 2000). We are notaware of any research in which the different organic matter fractions con-sidered in activated sludge models (inert, slowly biodegradable, and readilybiodegradable) have been studied experimentally in constructed wetlands.

D2-June

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FIGURE 2. Changes in the number of particles remaining along the length of two horizontalSSF CWs (C2, with a water depth of 0.5 m; D2, with a water depth of 0.27 m) during asampling campaign carried out in June. The y-axis data has been standardized by dividing bythe number of particles in the influent. The line connects the averages from the data measuredon different days. The numbers refer to the sampling location: 1, close to the inlet; 2, in themiddle; and 3, close to the outlet. From Garcıa et al. (2005a), with permission.

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This probably reflects the fact that research into the modeling of constructedwetlands is far less advanced than equivalent research for conventional sys-tems. In relation to organic matter fractionation, Caselles-Osorio and Garcıa(2006) evaluated the removal efficiency of two experimental horizontal SSFCWs (surface area 0.54 m2, water depth 0.3 m, gravel size 3.5 mm, plantedwith Phragmites australis) operated under the same conditions and fed withsynthetic wastewater. Starch was used to simulate organic matter (as a slowlybiodegradable substrate) in one wetland and glucose (as a readily biodegrad-able substrate) was used in the other. After an experimentation period ofeight months, during which several HRTs were used, both wetlands showedvery similar average removal efficiencies for COD, which ranged from 85to 95%. In a companion study, Caselles-Osorio et al. (2007a) analyzed theefficiency of the same experimental wetlands as described above but applieda considerably higher organic loading rate (23 g COD m−2 d−1 vs. 6.0 g CODm−2 d−1 in the first study). Again, both wetlands were found to have verysimilar removal efficiencies. Therefore, SSF CWs do not appear to be sen-sitive to the type of organic matter in the influents, irrespective of whetherit is readily (like glucose) or slowly (like starch) biodegradable. Studies thattake organic matter fractionation into account are needed, particularly forthe development of mechanistic simulation models (Dittmer et al., 2005).

Particulate Organic Matter Accumulation

Particulate organic matter (POM) accumulation in the granular medium is atypical feature of SSF CWs and is one of the main factors behind clogging(Nguyen, 2001), which is the most serious problem in SSF CWs. POM accu-mulates progressively within the surface layers and on top of the mediumas sludge (or sediment), which is a mixture of wastewater solids, activebiofilms, and plant and microbial detritus. The net accumulation of POM inSSF CWs is determined by the balance between external and internal POMloadings and losses due to export and decomposition (Tanner et al., 1998).The increase in POM accumulation over time has been associated with thegradual decrease of HRTs in horizontal systems. However, the relationshipbetween these two parameters is indirect, and other factors such as the bulkdensity characteristics (related to the proportions of labile and stable POMfractions) and spatial patterns of the accumulations condition their effects onthe hydrodynamic behavior of the wetlands (Nguyen, 2001; Tanner et al.,1998).

In theory, the relative contributions of plant material, microbial material,and wastewater solids to POM accumulation should depend on the appliedload and the properties of the plants and biofilms growing in the system.Tanner et al. (1998) carried out a study in a five-year-old pilot horizontalSSF CWs (treating solids loads of between 2.2 and 7.3 g TSS m−2 d−1) andfound that POM accumulation rates (1.3 to 3.0 kg VSS/m2.year) were much

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higher than the rates potentially contributed by applied wastewaters (0.4to 1.6 VSS/m2.year). In a previous study of the same pilot system, Tannerand Sukias (1995) demonstrated the importance of plant-derived loadingand found that POM accumulation was 1.2–2.0 kg/m2 higher in plantedSSF CWs than in equivalent unplanted beds. Therefore, it was shown thatparticulate plant loading was greater than the external wastewater loading.Note that these studies did not evaluate POM accumulation caused by biofilmgrowth and decay. Nguyen (2001) carried out research in the pilot systemanalyzed by Tanner and Sukias (1995) and Tanner et al. (1998) and foundthat up to 90% of the POM was composed of recalcitrant fractions, probablyoriginating from lignocellulose. As lignocellulose compounds are not readilydecomposed under the anaerobic conditions that prevail in horizontal SSFCWs, Nguyen (2001) suggested studying the effect of periodic draining onthe decomposition of lignocellulose and its subsequent influence on mediumporosity.

Caselles-Osorio et al. (2007b) studied the loading of COD and solidsand the accumulation of solids in six full-scale horizontal SSF CWs used totreat urban wastewater in Spain and found a positive correlation betweensolids accumulation rates and loading rates (see Table 1). The authors alsomeasured the hydraulic conductivity of the granular medium and foundthat it was inversely related to the amount of accumulated solids. However,this relationship was indirect, as Tanner et al. (1998) also found for POMaccumulation and HRT.

POM accumulation rates seem to decrease with wetland maturation time.Tanner et al. (1998) found that the rates were approximately two times higherduring the first two years of operation than in the three subsequent years.Higher rates at the start of operation may well be related to the establishmentof the plants, but this correlation has not been researched. Chazarenc andMerlin (2005) measured seasonal variations in accumulated POM in three

TABLE 1. COD and TSS surface loading rates, accumulated solids, percentage of volatilesolids, and solids accumulation rates in six full-scale horizontal SSF CWs in northeast Spain

COD surface TSS surface Solidsloading rate loading rate Accumulated accumulation

(gCOD (gTSS solids POM/DM rate (kg DMSystem m−2 d−1) m−2 d−1) (kg DM m−2) (%) m−2 year−1)

Verdu 1 3.8–10.4 2.8–4.5 2.8–12.8 10–39 0.7–3.2Verdu 2 3.1–8.5 — 2.3–11.9 11–89 0.6–2.9Alfes 5.3–9.2 3.2–4.9 2.6–35.1 7–13 0.6–8.8Corbins 10.9–17.5 6.5–10.0 6.0–57.3 3–19 1.5–14.3Almatret north 4.5–13.8 4.5–6.2 2.3–9.6 5–30 0.8–3.2Almatret south 6.2–12.9 2.6–8.0 2.8–20.3 5–28 0.9–6.8

From Caselles-Osorio et al. (2007b). In this study, dry matter (DM) and particulate organic matter (POM)were measured as total suspended solids (TSS) and volatile suspended solids (VSS), respectively. Allsystems treat primary effluent except Verdu 2, which treats secondary (pond) effluent.

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full-scale vertical SSF CWs over a period of one year and found no clearseasonal trend. The POM/DM ratio was generally low and ranged from 25to 50%, possibly because the systems were fed with screened (not settled)raw sewage. Fine material released from the granular medium may also havecontributed to the low ratio values. Caselles-Osorio et al. (2007b) evaluatedthe POM/DM ratios of accumulated solids in six full-scale horizontal SSF CWsand found that the average values were below 20% in five systems (used totreat primary effluent) and above 75% in only one facility, which was used totreat maturation pond effluent (see Table 1). The low organic matter contentof most of the systems was related to the high degree of mineralization ofthe organic matter and the presence of mineral particles derived from thegranular media.

The main mechanisms for influent POM (and in general TSS) retentionin SSF CWs are those of physico-chemical nature, and include impact andretention encouraged by path variations of water flow owing to the grainsof the medium, settling due to low speed movement, and adhesion owingto superficial interaction forces.

There is no comprehensive study in the literature that evaluates therelative importance of these physicochemical processes in SSF CWs. Theremoval rates of influent TSS in SSF CWs are usually very high (>90%).

Disintegration and Hydrolysis of Particulate Organic Matter

In conventional wastewater treatment systems, disintegration is consideredas a non-biological process that mediates the breakdown of influent POMinto complex composite particulates, particulate carbohydrates, particulateproteins, particulate lipids, inert particulate material, and inert soluble mate-rial (Vavilin et al., 2008). In constructed wetlands, the disintegration step alsoincludes the lysis of microbial cells and plant material. Invertebrates livingin constructed wetlands may also influence disintegration processes, but thishypothesis has not yet been tested experimentally. During the hydrolysisstep, which occurs after disintegration a defined particulate or macromolec-ular substrate is degraded into its soluble monomers. Hydrolysis is a biolog-ical process mediated by the exoenzymatic activity of two types of enzymes:those excreted by microorganisms as a part of extracellular metabolism, andthose immobilized on medium colloids and humic materials (Shackle et al.,2000). Disintegration and hydrolysis are processes that occur either underaerobic, anoxic, or anaerobic conditions.

The hydrolysis reaction is one of the processes that most restricts theremoval of organic matter in wastewater treatment plants (Mino et al., 1995;Sanders et al., 2000). However, there are very few studies of (or referencesto) the organic matter hydrolysis reactions that occur in SSF CWs. McHenryand Werker (2005) monitored the hydrolysis kinetics of fluorescein diacetateinto easily measurable fluorescein and considered them to be an indirect

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indicator of the microbial biomass, as it is known that hydrolysis activ-ity is directly proportional to biomass. Experiments were conducted with12 L experimental SSF CWs batch fed with synthetic wastewater and plantedwith Phragmites australis. Comparison with unplanted systems showed thatplants did not increase the intensity of hydrolytic activity. Caselles-Osorioand Garcıa (2006) fed one experimental horizontal SSF CW with starch (asa carbon source in synthetic wastewater) at a rate of 6.0 g COD m−2 d−1

and observed that starch granules were present continuously in the inter-stitial water near the inlet zone. According to the starch hydrolysis ratesreported in other studies, the authors estimated that it was unlikely that thestarch would have accumulated in the system. These authors explained thatthe experiments were performed with a lower water temperature than thatused by other researchers (mostly on anaerobic reactors) and this could bethe reason for starch accumulation. Note that temperature influences bothdisintegration and hydrolysis (Vavilin et al., 2008).

Aerobic Respiration and Oxygen Fluxes

Influent dissolved organic matter (DOM) and that produced after disintegra-tion and hydrolysis processes can be removed by aerobic respiration throughthe metabolism of a large number of heterotrophic bacteria that use oxygenas a final electron acceptor. Most of these bacteria are facultative with respectto the electron acceptor and can use nitrate or nitrite when oxygen has beendepleted (denitrification). Aerobic respiration therefore requires oxygen,which can be the most limiting substrate for this reaction in SSF CWs. How-ever, oxygen availability varies greatly between different types of wetlands.Vertical systems are operated normally using intermittent loading alternatedwith resting periods (passive air pump effect), which favors the presenceof oxygen in the bulk water, whereas horizontal systems are permanentlyflooded, so the oxygen concentration is usually very low or undetectable.

The three sources of oxygen in SSF CWs are inputs by the influent,physical surface reaeration, and plant release. Oxygen is usually undetectedor present in very low concentrations in wastewater, so oxygen input bythe influent is generally negligible. Even if the wastewater contained oxygenup to the saturation point (7–11 mg L−1, depending mainly on the tempera-ture), the concentration would not have a noticeable effect on typical urbanwastewater, for which BOD5 ranges from 200 to 300 mg L−1.

Rousseau and Santa (2007) studied the physical oxygen transport ratesfrom air to water moving in an experimental unplanted horizontal SSF CW(surface area 0.63 m2 and water depth 0.3 m). The results clearly showed thatthe higher turbulence at lower hydraulic retention times (HRTs) increased theoxygen transfer rates. At the higher HRT tested (two days, which is actuallya little low for a horizontal SSF CW), the oxygen transfer rate was very low(0.7 g O2 m−2 d−1) and in the same order of magnitude as the typical rate

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in stagnant open water. Therefore, it can be seen that the amount of oxygentransported from air to water in a horizontal SSF CW is insignificant in com-parison to the oxygen demand of standard urban wastewater. The transferrate from the air to the bulk water in horizontal SSF CWs is related to theoxygen deficit as it is usually considered in running waters. The magnitudeof the oxygen deficit depends on the oxygen mass transfer coefficient, whichis directly proportional to the water velocity and inversely proportional tothe water depth. Empirical functions for calculating mass transfer coefficientsare available for rivers, but they cannot simply be transferred to SSF CWsbecause rivers have very different turbulence regimes.

Vertical SSF CWs operate under much more aerobic conditions thanhorizontal SSF CWs because the combination of intermittent loads and restingperiods promote the physical reaeration of the intergranular spaces in themedium (Kayser & Kunst, 2005; Kayser et al., 2003; Langergraber et al., 2007).Thus, when large volumes of wastewater are pumped into vertical SSF CWsin a short period of time, the HLRs exceed the hydraulic conductivity of themedium and a water layer is formed on top of the wetland. When this waterlayer migrates downward, the air in the pore spaces below is compressedand dissolves more easily in the water layer. In addition, an underpressureis created above the water layer that causes new air to be sucked in, whichis then available to microorganisms during the next pump phase (Rousseau,2005). According to Cooper (2005), oxygen transfer rates in vertical SSF CWsare at least 28 g O2 m−2 d−1.

The results of research to quantify rates of oxygen release by plantshave led to the general consensus that aerobic microsites are present inthe granular medium of horizontal SSF CWs (Bezbaruah & Zhang, 2005).However, it is widely accepted that wetland plants do release enough oxygeninto their immediate root environment to remove all of the OM in a typicalprimary effluent in either horizontal or vertical SSF CWs. In fact, the results ofseveral studies have shown that plant oxygen release rates are much lowerthan the oxygen demand of standard urban wastewater (Brix, 1994a, 1997;Tanner, 2001a).

Table 2 shows the plant oxygen release rates reported in the literature,ordered according to the estimation method used. As can be seen, the ratesvary considerably according to the different experimental approaches used,the seasonal variations in system conditions, and the different assumptionsused in the calculations. In general, the rates taken from laboratory studiesare much lower than those obtained by theoretical stoichiometric calcula-tions. From the information currently available in the literature, it seemsobvious that the rates based on theoretical calculations overestimated theoxygen release because they were based on the assumption that all BOD(particulate and dissolved) was removed by aerobic respiration (Stein &Hook, 2005). Therefore, other biochemical pathways, such as denitrification,fermentation, sulfate reduction, and methanogenesis, were disregarded, and

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it was assumed that a very large amount of oxygen was needed to OM re-moval. Recent studies have shown that aerobic respiration is not the onlyreaction to exert a significant influence on organic matter removal in hori-zontal SSF CWs (Aguirre et al., 2005; Baptista et al., 2003). In addition, thepresence of endogenous organic matter mobilized by root exudates couldcontribute to the overestimation of oxygen release (Brix & Headley, 2007).

Denitrification

Denitrification is the biochemical reduction of nitrate and nitrite to nitric ox-ide, nitrous oxide, and nitrogen gas. This process links the C and N cycles inCWs because it enables denitrifying bacteria to obtain energy from organiccompounds at the same time, as nitrate (and nitrite) is used as an electronacceptor. Denitrification is conducted by a wide range of heterotrophic aero-bic facultative bacteria that are able to use nitrate (and nitrite) as an electronacceptor under anoxic conditions. These bacteria species use oxygen pref-erentially over nitrate as an electron acceptor when it is available in thesurrounding environment. Consequently, significant denitrification rates areonly observed in depleted oxygen environments.

Garcıa et al. (2004a) used theoretical mass balances to estimate the rela-tive contribution of several biochemical pathways (including denitrification)to organic matter removal in eight pilot horizontal SSF CWs with differentconfigurations (55 m2 of surface area) that treated municipal wastewater.These pilot wetlands had different aspect ratios, contained granular media ofdifferent sizes, and had different water depths. The water depth was found tohave the strongest effect on the performance and efficiency of the wetlands.Denitrification was the most important biochemical reaction in removingdissolved organic matter in wetlands with a depth of 0.27 m, whereas sulfatereduction had a greater effect in wetlands with a depth of 0.5 m. Baptistaet al. (2003) used two experimental horizontal SSF CWs with a surface area of0.85 m2 fed with beer diluted in tap water and found that greater than 70% ofthe influent TOC was removed by denitrification. Although the denitrificationrates estimated by Garcıa et al. (2004a) and Baptista et al. (2003) were cal-culated under several assumptions, they showed clearly that denitrificationhas a strong effect on organic matter removal under certain conditions.

Munch et al. (2005) conducted denitrification activity assays in an ex-perimental batch-fed wetland and found that this reaction is stimulated bymacrophyte roots. Increasing denitrification activities were reported near theroots in relation with increasing DOC concentrations because of organic mat-ter leaching. Brix and Headley (2007) analyzed the effect of hydroponicallygrown Phragmites australis and estimated that the root release of DOC wasapproximately 0.25 t DOC/ha.year, which could produce a denitrificationrate of up to 1 t N/ha.year.

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Acid Fermentation

Fermentation is a multistage biochemical process in which soluble organicmonomers present in wastewater and those generated through hydrolysisare converted into volatile short-chain fatty acids (VFAs). A large numberof heterotrophic bacteria groups are involved in fermentation reactions. Fer-mentation occurs under anaerobic conditions and is therefore an importantreaction in horizontal SSF CWs, whereas its effect is apparently negligible invertical systems. There is little information in the literature about fermentationin SSF CWs identified by the presence of intermediates such as VFAs.

Huang et al. (2005) studied the influence of wetland design factors onthe removal or formation of VFAs. Experiments were conducted in eightpilot horizontal SSF CWs with surface areas of 55 m2 and fed with the sameurban wastewater flow. These pilot wetlands had different aspect ratios, con-tained granular media of different sizes, and had different water depths. Thewetlands were operated during several periods in which different hydraulicloading rates (HLRs) were applied (20, 27, 36, and 45 mm d−1). Analysisof the influent and the effluents showed that acetic acid was present in thehighest concentration (in the range of mg L−1) of all the VFAs detected.Propionic, butyric, valeric, and hexanoic acids were found in considerablylower concentrations (in the range of µg L−1). The average acetic acid con-centration was 25 mg L−1 in the influent and 7.9–14.4 mg L−1 in the effluents(the effluent concentration changed depending on the wetland). These av-erage concentrations accounted for 22% of the BOD5 in the influent and20–40% in the effluents. HLR and particularly water depth were found tobe the two major factors that controlled the degree of acetic acid removal.Higher removal rates were recorded in shallow wetlands (with water depthsof 0.27 m, as opposed to 0.5 m) and were consistent with the more oxidizedconditions found in these systems. Interestingly, the authors found that evenwhen the load was similar, acetic acid removal was considerably higher incold months than in warm months, which suggests that the acetic acid wasproduced within the system during warm months.

Garcıa et al. (2005a) studied the variability of VFAs and other relatedparameters along the length of two of the eight pilot horizontal SSF CWsstudied by Huang et al. (2005). Acetic acid was again found to be the mostabundant VFA in all of the samples analyzed (in the range of mg L−1). TheVFAs with the next-highest concentrations were isovaleric acid in the influ-ent and propionic acid in the effluent (in the range of µg L−1). The effluentconcentrations of acetic and propionic acids were higher than the influentconcentrations in many cases. In contrast, the concentration of longer-chainVFA was lower in the effluents of both SSF CWs in all cases. The overall be-havior of the acetic acid concentration varied considerably along the lengthof the SSF CWs. In some cases, the concentration remained relatively con-stant, whereas in others it decreased or even increased substantially between

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a point in the SSF CW and the outlet. The propionic acid concentration wasgenerally fairly stable along the length of the system. In contrast to the be-havior of acetic and propionic acids, the concentration of longer-chain VFAfell sharply close to the inlet of the wetlands and was fairly stable in the restof the system.

Sulfate Reduction

Interest in sulfate reduction in SSF CWs has grown in recent years becauseresearch has shown that it can contribute significantly to the removal oforganic matter in horizontal SSF CWs (Aguirre et al., 2005; Garcıa et al.,2004a, 2005a). Sulfate is a normal constituent of many types of wastewaterand can be used as an electron acceptor in the absence of oxygen by alarge group of strictly anaerobic heterotrophic microorganisms called sulfate-reducing bacteria. These microorganisms can grow by using a wide rangeof fermentation products as electron donors (e.g., acetate, lactate, butyrate,acetone, formate, isopropanol, hydrogen and butanediol) (Lloyd et al., 2004;Stein et al., 2007). Reduced sulfur compounds such as sulfide are releasedby the activity of sulfate-reducing bacteria and are known to be potentinhibitors of plant growth and certain microbial activities (Wießner et al.,2005; Gonzalias et al., 2007). These reduced compounds can also producenuisance odors (Huang et al., 2004a). Consequently, the sulfur cycle is a keyfactor that must be considered when the practical application of SSF CWs isbeing planned.

Garcıa et al. (2004a) examined the relative contribution of sulfate re-duction to dissolved organic matter removal with respect to the influenceof other biochemical pathways in eight pilot horizontal SSF CWs with sur-face areas of 55 m2 that were fed with the same urban wastewater flow.They compared theoretical mass balances and found that sulfate reductionwas the most effective biochemical reaction in removing dissolved organicmatter measured as COD (36–100% of the total removal) in wetlands with adepth of 0.5 m, while in wetlands with a depth of 0.27 m, denitrification wasthe most important and sulfate reduction the second (17–40% of removal).In this study, an inverse trend was furthermore observed between the ratesof sulfate reduction and the organic matter removal efficiency because shal-lower wetlands were the more efficient ones. In fact, the authors concludedthat the lower efficiency observed in the wetlands with a depth of 0.5 mwas related to the fact that less energetically favorable biochemical reactionssuch as sulfate reduction were predominant (as opposed to denitrification,which was more common in the shallow systems).

Garcıa et al. (2005a) and Aguirre et al. (2005) measured the changes insulfate concentration along the length of two of the eight horizontal SSF CWsused by Garcıa et al. (2004a), one with a water depth of 0.5 m and the other

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0.27 m, in different months. The sulfate concentrations varied in the differentperiods and even in the same SSF CW, and were found to increase, decrease,or remain stable in different cases. Therefore, the relative contribution ofsulfate reduction to the removal of organic matter can vary in the samewetland system depending on the organic load and other environmentalfactors, such as temperature. In these studies, BOD5 and ammonium removalrates were found to decrease with increasing sulfate reduction activity. TOCand ammonium concentrations decreased along the length of the wetlandswith a depth of 0.27 m, whereas sulfate concentrations remained largelyconstant (which indicates low sulfate reduction activity). In the wetlandswith a depth of 0.5 m, sulfate was removed close to the inlet, and TOC andammonium concentrations were higher (see Figure 3).

Wießner et al. (2005) found that BOD5 and ammonium removal rates de-creased with increasing sulfate reduction activity in laboratory-scale 0.06 m2

SSF CWs, operated with complete mixing of the pore water through per-manent water recycling. This configuration was designed by the authors toequalize the macro-scale gradients. Ammonium removal seemed to decreaseexponentially when sulfate removal was around 50% (influent 150 mg L−1

and effluent 75 mg L−1), whereas BOD5 removal decreased only slightly.Thus, these results suggest that ammonium removal is more sensitive to sul-fate reduction than BOD5 removal. Caselles-Osorio and Garcıa (2006) alsofound that the COD removal efficiency of two experimental 0.55 m2 hori-zontal SSF CWs was greater when no sulfates were present in the influent.However, they did not observe as clear a relationship between ammoniumremoval and sulfate reduction as that found by Wießner et al. (2005).

Stein et al. (2007) examined the effects of season, temperature, plantspecies, and organic loading on sulfate reduction in 16 experimental SSFCWs with surface areas of 0.031 m2 fed in batch mode. Eight wetlands wereplanted with Schoenoplectus acutus (hardstem bulrush), four with Typha lat-ifolia (cattail), and four were left unplanted as controls. Sulfate removal waslower in winter, higher in summer, and moderate in spring and fall. Seasonalvariations were more pronounced in planted wetlands. Redox measurementsshowed that plant-mediated oxygen transfer inhibited the activity of sulfate-reducing bacteria in the winter period, which reduced the level of sulfateremoval due to temperature. Consequently, the interactive effects of organicloading rate, temperature, season, and plant species controlled not only thedegree of sulfate reduction but also the balance of competition between thedifferent microbial consortia responsible for water treatment in constructedwetlands.

The results presented above confirm the importance of sulfate reductionand, in general, of all sulfur transformation processes in the behavior andeffectiveness of horizontal SSF CWs. Wießner el al. (2005) suggested thatfuture research should focus on the toxicity of reduced sulfur compounds toother processes, the competition for oxygen due to the oxidation of reduced

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COT

TO

C, m

g L

-10

10

20

30

40

50

60

C2D2

Sulphate

SO

42-, m

g L-1

0

20

40

60

80

100

120

140

Ammonia

Location

Influent

1 2 3 Effluent

NH

3, m

g N

L-1

0

10

20

30

40

50

60

FIGURE 3. Changes in TOC, sulfate, and ammonia concentration along the length of twohorizontal SSF CWs (C2, with a water depth of 0.5 m; D2, with a water depth of 0.27 m)during a sampling campaign carried out in July. The values displayed are the averages andthe standard deviations obtained from samples taken over five consecutive days. The numbersrefer to the sampling location: 1, close to the inlet; 2, in the middle; and 3, close to the outlet.From Aguirre et al. (2005), with permission.

species, the changes in micro-environmental conditions in the rhizospheredue to redox potential and sulfur deposits, the mobilization or immobilizationof nutrients, and the formation of biofilms.

Methanogenesis and Methane Fluxes

Methanogens are strictly anaerobic bacteria that produce methane as an endproduct of metabolism. Methanogens and sulfate-reducing bacteria requireenvironments with similar redox potential levels and use the same types of

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electron donors (i.e., hydrogen, methanol, and acetic acid). When these twogroups of bacteria grow together and the COD:sulfate ratio (expressed asCOD:S) is lower than 1.5, sulfate-reducing bacteria are able to outcompetemethanogens. When the ratio is greater than 6, methanogens predominateover sulfate-reducing bacteria (Stein et al., 2007). Methanogenesis can re-move significant quantities of organic matter from wastewater in SSF CWsand has been studied more extensively in recent years (Garcıa et al., 2007a).In addition, methane is a potent greenhouse gas that makes a substantialcontribution to global climate change when it is emitted. Methane emissionmeasurements can be used to determine the influence of methanogenesison organic matter removal in SSF CWs.

According to Brix et al. (2001a), the total flux of methane emitted intothe atmosphere in vascular plant-dominated natural wetlands (and thereforein SSF CWs) is mediated by three processes: gas diffusion, ebullition, andinternal plant-mediated transport. Brix et al. (2001a) estimated that the to-tal release by plant transport accounted for approximately 70% of methaneemissions in a natural wetland in the Czech Republic. The remaining pro-portion could be released by diffusion and ebullition, as well as by methan-otrophic oxidation (van der Nat & Middelburg, 1998). In the case of SSF CWs,the relative release by these different processes is unknown. However, theproportions could be quite different to those observed in natural wetlandsbecause wastewater contains high concentrations of organic matter.

Spatial variations of methane emissions in horizontal SSF CWs havebeen reported in several studies (Mander et al., 2003a; Picek et al., 2007;Teiter & Mander, 2005). Methane emission rates have a remarkable variation;however, the rates are usually grater near the inlet than near the outlet.This seems to be a logical pattern because methanogenic activity is directlyrelated to the organic load in the system, which is higher around the inlet of ahorizontal SSF CW than at the outlet. The data reported by Teiter and Mander(2005) indicate that methane emissions were between 10 and 20 times higherclose to the inlet than at the outlet in a full-scale horizontal SSF CW treatingwastewater from a hospital and designed for 40-person-equivalent (PE).

Garcıa et al. (2007a) studied short-term methane emissions (daily varia-tions in diffusive and ebullition release) in two pilot horizontal SSF CWs withsurface areas of 55 m2 and water depths of 0.5 m treating urban wastewater.They found that emission rates were constant during the measurement peri-ods (from 4 to 7 hours). Picek et al. (2007) analyzed diffusion and ebullitionmethane emissions in a full-scale horizontal SSF CW system in Slavosovice,Czech Republic (150 PE, surface area of each wetland of 374 m2, bed depthof 0.8–0.9 m) and also observed largely constant emissions throughout theday. This linear trend had already been reported in previous studies (Brix,1990; Tanner et al., 1997). Picek et al. (2007) also studied long-term methaneemissions and observed that emissions decreased gradually toward the endof the growing season of the plants in the system. It is thought that long-term

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TABLE 3. Field methane emission rates from horizontal SSF CWs measured in different studies

Average surface organicloading rate (g BOD m−2 d−1) CH4 (mg m−2 d−1) Reference

3.7 304 Brix (1990)0.95 0.05–2800 Mander et al. (2003a)0.95 0.93–210 Teiter and Mander (2005)1.7 0–2970 Picek et al. (2007)5.4–10.8 15–42 Garcıa et al. (2007a)

methane emissions are sensitive to temperature variations because they werefound to be much higher during summer than during winter (Søvik et al.,2006).

Methane emission rates show a very wide range of variation whencomparing values derived from different studies (see Table 3). In additionto environmental factors, such as temperature, organic loading rate shouldtheoretically be an important factor that affects methane emissions too. How-ever, the data in Table 3 show that those studies in which the lowest methaneemission rates were found also reported the highest organic loading rates.This illustrates the difficulties of comparing emission data from different SSFCWs (measured by non-standardized methods). Emissions also depend onthe quality of the organic matter in the wetland (such as the proportionsof particulate and dissolved matter or the amount of short-chain fatty acids)and may be affected by the rate of methane oxidation close to the plantrhizosphere and the surface of the water due to methanotrophic activity.Van der Nat and Middelburg (1998) found that reeds reduced methane emis-sions by 16% during the growing season in planted containers with tap andde-ionized water.

Teiter and Mander (2005) and Søvik et al. (2006) studied the potentialimpact of greenhouse gas emissions from SSF CWs (e.g., methane and ni-trous oxide) on global warming. In both studies, the authors conclude thatemissions from SSF CWs are globally insignificant. In fact, Teiter and Mander(2005) estimated that if wetlands were used to treat all global wastewater,they would only contribute 1% of the trace gas emission budget.

REMOVAL OF SPECIFIC ORGANIC CONTAMINANTS

Although the behavior of specific organic contaminants in conventionalwastewater treatment plants is well-documented, less information is avail-able about CWs, particularly SSF CWs (Haberl et al., 2003).

The main concern related to the fate of specific contaminants is linkedto their toxicity, persistence, and potential bioaccumulation and biomagnifi-cation when the treated effluent is discharged into the aquatic environment

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or intended for reuse. In this section, we first focus on surfactants that arewidespread specific contaminants in surface waters. These compounds arereviewed depending on the analytical methods used for their determination.As will be seen, the available information on the removal of pesticides andherbicides is scarce in the case of SSF CWs (which contrasts with surfaceflow constructed wetlands). We will see that the removal of pharmaceuticaland personal care products is dependent on the redox status of the wet-land, and that vertical systems seem to have higher removal efficiencies thanhorizontal flow wetlands.

Surfactants

Approximately 3–4 million tons of synthetic surfactants are produced in west-ern Europe, Japan, and the United States every year, making them the mostabundant synthetic compounds intentionally released into aquatic environ-ments. Various studies in the literature have examined the behavior of themost relevant classes of surfactants—non-ionic and anionic—in SSF CWs. Inthis review, articles will be presented separately according to the approachused to analyze the surfactant: chromatographic methods at the molecularlevel first (i.e., gas chromatography (GC) or liquid chromatography (LC)),and non-specific colorimetric methods second. This distinction is requiredbecause the two approaches provide different types of information. Colori-metric methods are class-specific but subject to many types of interferences,which lead to overestimations of surfactant concentration (e.g., humic acidsare detected as surfactants).

ANIONIC SURFACTANTS

Anionic surfactants are the most widely used type of surfactant for domes-tic applications and account for nearly 60% of the market share, followedby non-ionic surfactants. Linear alkylbenzene sulfonates (LAS) are the mostwidely used synthetic anionic surfactants for domestic applications; they re-placed the branched alkylbenzene sulfonates used until the mid-1960s (seeTable 4). LAS consist of a suite of linear alkylbenzene chains of differentlengths (ranging from 10 to 13 carbon atoms) and different positional iso-mers. The LAS degradation pathway begins with the sequential ω- and β-oxidation of the alkyl chain, which leads to the formation of low carbonnumber sulfophenyl carboxylates (SPC). The next steps consist in aromaticring cleavage and desulfonation (Schoberl, 1989) (see Figure 4).

Huang et al. (2004b) conducted a study in northeast Spain to exam-ine the removal of LAS from urban wastewater in a pilot plant with eighthorizontal SSF CWs (surface area of 55 m2 each planted with Phragmites aus-tralis). The pilot plant was fed with wastewater from an urban developmentwith nearly 200 inhabitants. The authors evaluated the effect of the aspectratio, the water depth (i.e., 0.3 and 0.5 m), the medium gravel size (i.e.,

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TABLE 4. Physicochemical properties of specific contaminants considered in the presentwork

CAS Water solubility LogName registry no. Structure Structure (mg L−1) 25◦Ca Kow∗

SurfactantsLAS 68411-30-3 C16H25NaO3S Anionic 177 2.02SPC n.a. — Biointermediate n.a. n.a.NP 25154-52-3 C15H24O Biointermediate 6.35 4.48NPE1–3 EO 9016-45-9 — Non-ionic n.a. 4.2

PesticidesSimazine 122-34-9 C7H12ClN5 Triazinic 6.2 2.18Parathion 56-38-2 C10H14NO5PS Organophosphorus 11 3.83Omethoate 1113-02-6 C5H12NO4PS Organophosphorus 1exp+6 -0.75Alachlor 15972-60-8 C14H20ClNO2 Carbamate 240 3.52Chlorpiryphos 2921-88-2 C9H11Cl3NO3PS Organophosphorus 1.12 4.96Endosulfan 115-29-7 C9H6Cl6O3S Organochlorine 0.325 3.83Lindane 58-89-9 C6H6Cl6 Organochlorine 7.3 3.72

HerbicidesMetolachlor 51218-45-2 C15H22ClNO2 Phenoxyacid 530 3.13MCPA 94-74-6 C9H9ClO3 Phenoxyacid 630 3.25Dicamba 1918-00-9 C8H6Cl2O3 Phenoxyacid 8310 2.21Mecoprop 93-65-2 C10H11ClO3 Phenoxyacid 620 3.13Diuron 330-54-1 C9H10Cl2N2O Phenylurea 42 2.68

∗Calculated from the Interactive PhysProp Database.Abbreviation: n.a. = not available

(CH2)R2

SO3-

(CH2)R1 HO

O

(CH2)R1 (CH2)R2

SO3-

ω-oxidation

S-CoA

O

(CH2)R2(CH2)R1-n

SO3-

β-oxidation

OH

SO3-

COO-COO-

SO3-

H2SO3-

O

COO-

COO-

OH

FIGURE 4. Degradation pathway of linear alkylbenzene sulfonates. Adapted from Scott andJones (2000).

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D60 = 3.5 mm and 10 mm), and the hydraulic loading rate (HLR) (i.e., 20–45 mm d−1) on LAS removal. LAS were analyzed at the homolog level(C10–C13) and sulfophenyl carboxylate (SPC9-SPC11) biointermediates wereanalyzed using liquid chromatography and mass spectrometry (LC-MS). Inthe most favorable conditions, LAS and SPC were removed by up to 72% and11%, respectively (see Table 5). SPC was removed less effectively than LASdue to de novo formation in the wetlands from related LAS precursors. Thehighest removal efficiencies were obtained in shallower wetlands (i.e., witha water depth of 0.3 m), and all other variables except the hydraulic loadingrate (HLR) were non-significant (on removal efficiency). The results of thisstudy suggest that LAS were removed more efficiently in shallow wetlandsbecause of the more oxidized conditions that prevailed in these systems(which determine the relative importance of biochemical pathways for con-taminant removal). In fact, a mass balance showed that organic matter wasmainly removed by denitrification and sulfate reduction in the shallow wet-lands, whereas sulfate reduction and methanogenesis were the predominantremoval mechanisms in the deeper wetlands (0.5 m water depth). Further-more, it has been demonstrated that the occurrence of LAS in wastewatersinhibits methanogenesis, especially when the concentrations are greater than15 mg L−1 (Krueger et al., 1998; Mosche & Meyer, 2002). However, the pre-cise biodegradation pathway of LAS in horizontal SSF CWs is unknown sinceanaerobic biodegradation of LAS is under scrutiny (Garcıa et al., 2005b;Mosche & Meyer, 2002). In a recent study, it was demonstrated that LASbiodegradation can occur at a very slow degradation rate (a half-life of sev-eral months) in strictly anaerobic sediments (Lara-Martin et al., 2006). Thereis no information in the literature about LAS adsorption on granular mediumof constructed wetlands, but the higher removal of the most hydrophobichomologs (C13-LAS) suggests that adsorption to the organic matter retainedin the gravel bed could account for LAS removal to a certain extent (Huanget al., 2004b). Further research is required to corroborate this hypothesis.

Nielsen (2005) evaluated the removal of LAS from digested sludge in aCW (sludge drying bed) and compared the results to those obtained whenother sludge treatment methods. Wetland treatment was more effective (98%removal) than anaerobic treatment in a container or treatment in a sludgepile turned over mechanically (90%). A similar removal pattern was obtainedfor ethoxylated nonylphenols.

Although colorimetric methods are generally used to detect and quan-tify total anionic surfactants in wastewaters (i.e., methylene-blue active sub-stances, MBAS), these methods are susceptible to interferences that leadto overestimations of the surfactant concentration (e.g., humic acids are de-tected as surfactants) (Schmitt, 2001). This may be why the surfactant removalrate was found to be negative in a SSF CW used to treat dairy parlor effluent(see Table 5) (Mantovi et al., 2003). Conte et al. (2001) conducted a studyin central Italy to measure MBAS removal in four horizontal SSF CWs of

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different sizes and used for different treatment stages (secondary or tertiary).They reported removal percentages of between 42 and 88%.

NON-IONIC SURFACTANTS

Non-ionic surfactants are used in cleaning products, pesticide formulations,inks and industrial processes, but have a lower market share than anionic sur-factants. Alkylphenol ethoxylates—namely, nonyl (NPEO)—and to a lesserextent octylethoxylates are several of the non-ionic surfactants of majorenvironmental concern due to the estrogenic properties of its degradationintermediates—namely, nonylphenol and octylphenol (White et al., 1994).

Mantovi et al. (2003) used colorimetric methods to evaluate the removalof total non-ionic surfactants measured as cobalt thiocyanate active sub-stances (CTAS) from milk parlor effluent treated in a SSF CW (see Table 5).CTAS removal of almost 69% was recorded, although this figure should onlybe considered as indicative, as the analytical process is susceptible to thesame types of interferences as described in the previous section.

Belmont and Metcalfe (2003) examined the feasibility of using the or-namental plant calla lily (Zantedeschia aethiopica) to remove a variety ofcontaminants including alkyl phenols in experimental SSF CWs. They useda setup consisting of six wetlands (each of which was 0.38 m wide and2.4 m long) located inside a greenhouse. The substrate was a 0.3 m layerof crushed rocks with diameters of between 3 and 5 cm. The water levelwas maintained 5 cm below the gravel surface, and the effective volumewas 80 L. The hydraulic retention times were between one and two days,and the simulated wastewater was river water spiked with fertilizers andtannic acid. Four wetlands were planted with calla lily and two wetlandswere left unplanted as control systems. High removal rates of NPEOs wereobtained (96.6%), but no statistical differences were observed between theplanted and unplanted wetlands. Therefore, the presence of plants did notinfluence the removal of the selected contaminants to a significant degree.The authors suggested that the main removal mechanism was adsorption tothe organic matter present in the granular medium because the short-chainalkyl NPEOs are relatively hydrophobic. Moreover, the removal of ethoxy-lated (1–3 ethoxy groups) nonylphenols was higher (96–98%) than that ofnonylphenol without ethoxy groups (54–57%). This trend was attributed tothe conversion of ethoxylated derivatives to nonylphenol, as has been pre-viously reported in conventional wastewater treatment plants (Giger et al.,1984). Full-scale studies are needed to determine the behavior of the NPEOsfound in these laboratory-scale SSF CWs.

Pesticides and Herbicides

Free water surface (also known as suface-flow) constructed wetlands havebeen used much more widely than SSF CWs for agricultural pesticide runoff

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attenuation (Braskerud & Haarstad, 2003; Moore et al., 2001, 2002; Schulzet al., 2003a, 2003b), at least in part because the former type has lowerconstruction costs. Surface flow systems remove pesticides mainly throughphotochemical reactions (Blough & Sulzberger, 2003), whereas SSF CWsremove pesticides primarily through adsorption (hydrophobic) and, to alesser extent, biodegradation.

Pesticides form a large class of contaminants with different physico-chemical properties, ranging from highly hydrophobic compounds, whichare no longer commonly used, to highly hydrophilic compounds, includingionic species (see Table 4). There is little information about the effective-ness of SSF CWs in removing pesticides (see Table 5). To the best of ourknowledge, only a few studies have been published on the use of SSF CWsto treat agricultural runoff (McKinlay & Kasperek, 1999; Stearmann et al.,2003). Stearmann et al. (2003) evaluated 14 wetlands (1.2 × 4.9 m or 2.4 ×4.9 m, with water depths of 0.30 or 0.45 m) used to collect the runoff froma 465 m2 container plant nursery. Half of the wetlands were planted withbulrush (Scirpus validus) and half were left unplanted as controls. The wet-lands were operated at four hydraulic loading rates (5, 10, 20 and 40 mmd−1) and contained quartz gravel of two different sizes (large: ≤ 38 mm, withan effective size of d10 = 19.1 mm; and small: ≤ 22 mm, with an effectivesize of d10 = 11.1 mm). Herbicides (i.e., simazine and metolachlor) werespread over the container nursery in one application at doses ranging from1.19 kg ha−1 (absolute amount: 110 g) to 4.78 kg ha−1 (absolute amount:220 g). The hydraulic retention time (HRT) of the wetlands ranged from2.1 to 20.6 d (data corrected for evapotranspiration), and it was observedthat higher HRTs correlated with higher removal efficiencies. In addition, theplanted wetlands showed higher removal rates (82.4% for metalochlor and77.1% for simazine) than the non-planted ones (63.2% for metolachlor and64.3% for simazine). Herbicides could be removed in non-planted wetlandsby sorption onto the organic matter accumulated in the granular medium.Sorption of simazine and metolachlor is a relatively rapid process that can becompleted in a matter of minutes or hours (Kookama et al., 1992), whereasbiodegradation takes a few days (Stearman et al., 2003). Moreover, the her-bicides may undergo desorption once they have been adsorbed, but the rateof this process in SSF CWs is unknown. It is thought that the rate of desorp-tion will increase with the time elapsed time after sorption, as occurs in soils(Pignatello & Huang, 1991).

Slemp et al. (2004) conducted a microcosm study in which they de-termined the preferred electron acceptor for the microbial degradation ofsimazine in horizontal SSF CWs. Denitrification and sulfate reduction did nothave a significant effect on simazine removal. According to the identifiedsimazine metabolites, aerobic microbial degradation in the plant rhizosphereappeared to be the predominant biodegradation mechanism. These resultsare consistent with those obtained by Stearmann et al. (2003), who observed

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higher simazine removal in planted wetlands and assumed that aerobic pro-cesses had a greater effect in these systems than in unplanted wetlands.

In another laboratory-scale study using a vertical SSF CW and waterwith high levels of atrazine, the effluent concentration of atrazine was foundto be below the detection limit after a period of 25–32 days (McKinlay &Kasperek, 1999). Based on the results obtained after the granular medium un-derwent a sterilization treatment (with hydrogen peroxide and peroxyaceticacid mixture), the authors suggested that atrazine removal was mediated bymicrobial activity. However, neither atrazine intermediates nor microorgan-isms involved in removal processes were identified. In another laboratory-scale study, twin-shaped vertical SSF CWs fed with upward and downwardflow were tested for the removal of pesticides (parathion (O,O-diethyl-O-(4-nitrophenyl)-thiophosphorylether) and omethoate (O,O-dimethyl-O-(4-nitro-phenyl)-thiolphosphorylether)), as well as herbicides (MCPA (4-chloro-2-methylphenoxyacetic acid) and dicamba (3,6-dichloro-o-anisic acid)) (Chenget al., 2002a). The plot was fed intermittently with artificial wastewater (seeTable 5) and planted with crop plants (Colocasia esculenta and Ischaemumaristatum). After 120 days of wastewater application, the removal efficienciesof the applied pesticides were almost 100%. The system was less efficientfor herbicide removal and showed an efficiency of only 36% for MCPA andno removal of dicamba. These results show that the organophosphorus pes-ticides were removed easily in the vertical SSF CW.

Matamoros et al. (2006) carried out an injection experiment of pesticidesand herbicides in pilot horizontal SSF CWs and also obtained different re-moval results according to the chemical species, which ranged from no netremoval of diuron to complete removal of lindane and endosulfan (see Table5). The fact that these pesticides were not detected in the granular mediumsuggests that they were biodegraded or phytosorbed. Studies on the up-take of lindane by ryegrass (Li et al., 2002, 2005) have shown a minimummetabolism, but it forms bound residues, and roots exhibited a higher sorp-tion than predicted from their lipidic composition. On the other hand, theremoval of endosulfan was linked to its biodegradation by indigenous bac-teria in anoxic soils (Guerin, 1999). Further research is needed to determinethe removal mechanism of organochlorine pesticides in subsurface-flow con-structed wetlands.

Independent studies have shown that herbicides belonging to thephenoxyacid class, such as MCPA, dicamba, and MCCP (mecoprop, 2-(4-chloro-o-tolyloxy)propionic acid), are poorly removed in SSF CWs (i.e.,0–22% removal) (Cheng et al., 2002a; Matamoros et al., 2006). In addi-tion, these herbicides are acidic and therefore interact poorly with thegranular medium and the rhizosphere. Diuron (3-(3,4-dichlorophenyl)-1,1-dimethylurea), which belongs to the phenylurea class, behaves as refractory(Matamoros et al., 2006). In fact, all of these herbicides are highly hydrophilic(see Table 4) and are not likely to interact with organic matter, biofilm or

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rhizosphere. Consequently, they are removed in very small quantities inSSF CWs.

Comparison of free water surface flow and subsurface-flow constructedwetlands shows that the two types of systems achieve similar removal ef-ficiencies for organophosphorus pesticides (i.e., chlorpyriphos, omethoate,parathion, parathion-methyl and azinphos-methyl) (see Table 5).

Emergent Contaminants

Pharmaceutical products and fragrances present in wastewater belong toa group called emergent contaminants (Daughton, 2004; Matamoros et al.,2008). They form a broad class of contaminants and are characterized by con-stant input into the aquatic environment through human activity. Althoughmost of these contaminants are readily degradable, some compounds are re-fractory to biodegradation and cannot be removed in conventional activatedsludge wastewater treatment plants (Ternes et al., 2004).

The behavior of pharmaceutical and personal care products (PPCPs)in SSF CWs depends greatly on the physical characteristics of the specificcompound, and can range from no to complete removal (see Table 4). Hy-drophobic compounds like polycyclic musks (see Table 4) are eliminated byadsorption to the organic matter present in the granular medium, whereas hy-drophilic contaminants are removed by different processes according to thespecific compound. All of these compounds can be classified as refractory,moderate, and efficiently removed. Figure 5 shows the different behavior ofpharmaceutical compounds injected into two horizontal SSF CWs with dif-ferent depth. The results indicate that the shallow horizontal SSF CW (waterdepth of 0.30 m) achieved a higher PPCP-removal efficiency than the deepersystem (water depth of 0.50 m). This trend of higher removal efficiency withlower depths was also described for ibuprofen and LAS (Huang et al., 2004b;Matamoros et al., 2005). Higher removal rates in shallower wetlands wererelated to the higher potential redox values found in these systems.

The results in Table 5 show that vertical SSF CWs achieve higher PPCP-removal efficiencies than horizontal wetlands, in relation to the fact thatvertical systems exhibit more oxidized conditions. As reported for biofilm re-actors (Zwiener et al., 2002), oxygen is a key factor affecting the removal ofsome pharmaceuticals. Aerobic biodegradation prevails in vertical SSF CWs,whereas anaerobic biodegradation pathways are predominant in horizontalSSF CWs. The removal rates of most of the PPCPs studied fit zero-order kinet-ics. Hydrophobic contaminants such as the polycyclic musks did not showdifferences in their removal rates between the vertical and horizontal wet-lands due to their non-oxygen dependence. When these contaminants areadsorbed onto organic matter, they become more recalcitrant to biodegra-dation (Alexander, 2000) and can accumulate on the medium at very highconcentrations (450–825 mg/kg) (Matamoros & Bayona, 2006).

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FIGURE 5. Cumulative percentage mass recovery for pharmaceuticals and bromide in twohorizontal SSF CWs (C2, with a water depth of 0.5 m; D2, with a water depth of 0.27 m). Atotal of 250 mg per pharmaceutical and 200 mg of NaBr were injected continuously over 200h. The following compounds were identified: carbamazepine (CBZ), ibuprofene (IB), clofibricacid (CLF). From Matamoros et al. (2005), with permission.

Other Specific Organic Contaminants

In this subsection, we discuss the removal of a variety of organic con-taminants with different functionalities that have been tested in SSF CWs.Specific organic contaminants are a broad class of compounds that can befound in many types of contaminated waters (i.e., groundwater, industrial

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effluents, and landfill leachates). As shown in the previous sections, the be-havior of these contaminants is strongly dependent on their physicochemicalproperties.

Explosives are a specific group of contaminants found not only in pro-duction plants but also in disposal and destruction areas. Behrends et al.(2001) studied the removal of two primary explosives (TNT and RDX) in a38 L microcosmos experimental setup planted with canary grass (Phalarisarundinacea). The system consisted of four SSF CWs, two with horizontalflow and two with vertical flow, working in full-depth reciprocation (se-quential resting periods) (Behrends, 1999). The vertical wetlands operatedin aerobic conditions. A significantly high degree of TNT and RDX removalwas obtained (>90%) under the combined anaerobic-aerobic conditions pre-vailing in the experimental setup. This system was also effective at removingnutrients and non-specific organic pollutants such as BOD5 (Behrends et al.,2001). Haberl et al. (2003) conducted a laboratory-scale study to examinethe removal of TNT and 2,6-DNT in a 1 m3 container horizontal SSF CWand found that the efficiency increased from 50 to greater than 95% whenthe system was fed with molasses. The authors suggested that anaerobicbiodegradation is a more effective removal pathway for these compounds(Haberl et al., 2003). However, other removal mechanisms may also occur,such as higher bed adsorption in the presence of molasses.

A pilot study to evaluate treatment of hydrocarbon-contaminated groundwater was performed in Casper, Wyoming, with horizontal SSF CWs (Haberlet al., 2003). Four SSF CWs filled with sand were fed with the slipstreamfrom a large-scale oil-water separator, and the effluent from the wetlandswas pumped to a large-scale air stripper. The wetlands were operated in anupward vertical flow mode at a flow rate of 5.5 m3 d−1. Two of the beds weresubjected to forced subsurface aeration (14 L min−1), and the effluent met thenational standards for benzene (0.005 mg l−1) from August to January, whichwas attributed to volatilization and biodegradation. The lowest removal rates(15–31%) were recorded for methyl-tert-butylether (MTBE), which is con-sistent with the fact that this contaminant exhibits more refractory behaviorthan BTEX in the studied environment (Florence & Rifai, 2003).

MECHANISMS INVOLVED IN NUTRIENT REMOVAL

This section deals with transformations of nitrogen and phosphorus. As shallbe seen, nitrogen can be successfully removed in SSF CWs, while removalof phosphorus is often poor. Processes linked with nitrogen removal arenumerous and have been intensely investigated in the last years. Perhapsthe removal of nitrogen is the issue that has received most attention byresearchers and practitioners. We will see that nitrification followed by den-itrification seems to be the most important pathway for nitrogen removal. At

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the end of the section, we will present the factors that lead to low phospho-rus removal rates, and the technical developments that are being investigatedin order to increase these rates.

Nitrogen

Nitrogen removal is a key point in CW technology, as many full-scale systemshave been designed for nitrogen removal as well as organic matter removal.Full-scale facilities often combine vertical and horizontal SSF CWs to achievemore effective treatment, and these hybrid systems are particularly effectivefor removing nitrogen. Vertical SSF CWs successfully remove ammonium,but the presence of very low denitrification rates means that they often donot remove sufficient quantities of nitrate. In contrast, horizontal SSF CWsprovide conditions that are more conducive to denitrification but are lesseffective at nitrifying ammonium (Vymazal, 2005a, 2007).

Vymazal (2007) noticed that the processes involved in nitrogen transfor-mation and removal during wastewater treatment in SSF CWs vary accordingto the prevalent N chemical species. Organic N, ammoniacal nitrogen (NH4

+

and NH3), nitrate (NO3−), and nitrite (NO2

−) are thought to be the principalforms of nitrogen involved in the nitrogen cycle in CWs (Borin & Tocchetto,2007; Green et al., 1998; Mayo & Mutamba, 2004). Removal of total nitrogenin SSF CWs generally ranges from 250 to 630 g N/m2.yr in terms of surfacemass loading and from 40 to 55% in terms of percentage removal efficiency.The removed load depends on the type of SSF CW and the inflow loading(Vymazal, 2007).

Nitrogen transformation and removal mechanisms in CWs include min-eralization (ammonification), ammonia volatilization, nitrification, denitrifica-tion, plant and microbial uptake, nitrogen fixation, nitrate reduction, anaer-obic ammonia oxidation (ANAMMOX), adsorption, desorption, burial, andleaching (Vymazal, 2007). Denitrification is generally the dominant N removalprocess in mature SSF CWs. Alternative microbial N removal processes in-cluding ANAMMOX may have a great effect in SSF CWs designed to treatammonium-rich wastewaters (Tanner, 2004). Horizontal SSF CWs show ahigh potential for nitrate reduction due to the presence of anaerobic condi-tions. Simultaneous occurrence of all or several of the mechanisms quotedabove give place to what has been described as N spiraling by Kadlec et al.(2005). This behavior increases the retention time of the inflowing N withrespect to the hydraulic retention time, produces long delays in the treat-ment system response to changes in N loading, and reduces the short-termfluctuations in N loading.

Nitrogen removal in SSF CWs is affected by the hydraulic retentiontime (HRT), the temperature, the vegetation type and the properties of themedium (Akratos & Tsihrintzis, 2007; Kuschk et al., 2003). HRT and tem-perature usually have a strong effect on the removal efficiencies for total

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Kjeldahl nitrogen (TKN) and ammonium. Nitrogen removal rates at a watertemperature greater than 15◦C are significantly higher than those observedat lower temperatures (Akratos & Tsihrintzis, 2007; Caselles-Osorio & Garcıa,2007). TKN and ammonium removal is also affected by the levels of availableoxygen (Wallace et al., 2000). In this subsection, we describe the principalmechanisms involved in nitrogen transformation and removal in SSF CWs.

AMMONIFICATION

Ammonification in CWs has not been studied as extensively as other nitro-gen transformation processes. It consists in the conversion of organic N inammonia through exoenzymatic activity from enzymes excreted by microor-ganisms as a part of extracellular metabolism (Vymazal, 2007). A wide rangeof ammonification rates have been reported for CWs, with values rangingfrom 0.004 to 0.53 g N m−2 d−1 (Tanner et al., 2002; Vymazal, 2006).

AMMONIA VOLATILIZATION

Ammonia volatilization is a physicochemical process in which ammonia de-rived from ammonification reactions or from wastewater is transferred fromwater to the atmosphere. This process is controlled by pH and can only bea significant mechanism for ammonia removal when the pH is above 10.Consequently, ammonia volatilization is not thought to be a major ammo-nia removal mechanism in SSF CWs treating urban wastewaters because thepH usually ranges from 7 to 8.5 (Vymazal, 2007). This may not be the casein systems used to treat wastewaters or wastes that are rich in ammoniacalnitrogen, such as pig manure (Poach et al., 2003).

NITRIFICATION

Nitrification is defined as the biological oxidation of ammonium to nitratewith nitrite as an intermediate in the reaction sequence. Nitrification itselfcannot remove nitrogen from wastewaters, but nitrification coupled withdenitrification seems to be the major removal pathway in many CWs (Mayo& Bigambo, 2005; Tanner et al., 2002; Vymazal, 2007). Nitrification limitsnitrogen removal in many CWs because ammonia is the dominant species ofnitrogen in urban and other industrial wastewaters. The degree of nitrificationis determined by oxygen availability (Cottingham et al., 1999). Therefore,vertical SSF CWs are more conducive to nitrification than horizontal SSFCWs.

Tanner et al. (2002) carried out detailed research into nitrogen removaland transformations in experimental horizontal SSF CWs operating in se-ries and planted with Schoenoplectus tabernaemontani. They found thatnitrification occurred concurrently with organic matter removal, even in theupstream stages of wetlands that received higher-strength wastewater inflow.The average net rates of nitrification ranged from 0.56 to 2.15 g m−2 d−1, andwere correlated with the average ammonium concentrations. The estimated

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oxygen demand required to support full nitrification of ammonium and min-eralized organic N was in the upper range of expected levels supplied bysurface reaeration and plant-mediated oxygen transport. Tanner and Kadlec(2003) inferred that other microbial processes such as ANAMMOX might alsoincrease the degree of N removal in SSF CWs.

Kuschk et al. (2003) used a large pool of data from an operating pe-riod of four years to examine the annual course of nitrogen removal in ahorizontal SSF CW with a surface area of 125 m2 in a moderate climate. Thenitrification rates differed dramatically between winter (0.15 g m−2 d−1) andsummer (0.70 g m−2 d−1). Nitrification was limited by temperature duringall seasons and by additional seasonal factors in midsummer (these weremainly linked to plant physiology, which affected root oxygen release). Thecorrelation between temperature and nitrification rates was also pointed outby Gerke et al. (2001).

Passive air pumps can be used to increase nitrification rates in SSF CWsand are commonly applied in vertical systems. Passive air pumps use fill anddraw cycles to remove oxygen-depleted air from the system and introducefresh air. Each volume of drained effluent is displaced by an equal volumeof fresh air. Green et al. (1998) set up a passive air pump in a variant of avertical SSF CW and obtained a nitrification efficiency of 96%.

Artificial aeration can also be used to increase nitrification rates. Wallace(2001) developed and patented an integral aeration system (Forced BedAerationTM) that consists of an air blower, a PVC distribution tube header,and perforated flexible HDPE tubing. The tubing is placed on top of theimpermeable liner. Air is delivered to the wetland bed to enhance oxygen-limited microbial processes. This aeration system has proven to providehigh removal efficiency at low energy costs in locations with cold weatherconditions.

DENITRIFICATION

Denitrification is the biochemical reduction of nitrate and nitrite to nitricoxide, nitrous oxide, and nitrogen gas, and is considered a major removalmechanism for nitrogen and organic matter in most types of CWs. However,nitrate concentrations are usually very low in wastewaters (with the excep-tion of drainage water from agriculture and some industrial wastewaters), sodenitrification should generally be coupled with nitrification to attain satisfac-tory net N removal. Environmental factors known to influence denitrificationrates include the absence of O2, redox potential, the moisture content of themedium, temperature, pH value, the presence of denitrifiying bacteria, themedium type, the presence and type of organic matter, the nitrate concen-tration, and the presence of overlying water. The presence of oxygen fornitrification and the absence of oxygen for denitrification are often the prin-cipal factors that limit overall nitrogen removal rates in CWs (Vymazal, 2007).

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Denitrification has a lesser environmental impact than ammonia volatiliza-tion and is therefore a more desirable nitrogen removal mechanism (Poachet al., 2003).

The optimum pH level for denitrification is between pH 6 and 8. Deni-trification occurs more slowly but still to a significant degree below pH 5 andis negligible or absent below pH 4. Denitrification is also strongly dependenton temperature. Denitrification proceeds at very slow but measurable ratesat temperatures below 5◦C. Incomplete denitrification leads to the release ofnitrous oxide, which is a strong greenhouse gas. Teiter and Mander (2005)found that vertical SSF CWs emitted a considerably higher volume of nitrousoxide than horizontal SSF CWs. Nevertheless, although nitrous oxide emis-sions were found to be relatively high, their overall influence on climatechange was estimated to be insignificant.

Tanner et al. (2002) conducted research with experimental horizontalSSF CWs operating in series and planted with Schoenoplectus tabernaemon-tani. The average net rates of denitrification ranged from 0.47 to 1.99 g m−2

d−1, and the degree of denitrification clearly correlated with organic mat-ter removal. Mayo and Bigambo (2005) operated two horizontal SSF CWsplanted with Phragmites mauritianus and estimated an average net rate ofdenitrification of 0.44 g N m2 d−1 by means of a mathematical model. Theyalso concluded that denitrification was the major pathway leading to a per-manent removal of nitrogen. Senzia et al. (2003) obtained slightly higherdenitrification rates in a wetland planted with Typha domingensis (19.3%)than in a wetland planted with Phragmites mauritianus (15.3%).

Progressive accumulation of organic matter in the granular medium ofSSF CWs increases denitrification rates. Organic wastes such as bark or strawcan be mixed with the granular medium in wetlands to increase denitrifica-tion rates, particularly at the beginning of the operating period, when thesystems contain only a small amount of accumulated organic matter (Søvik& Mørkved, 2008).

PLANT UPTAKE

Nitrogen can be removed in SSF CWs by harvesting the aboveground biomassof emergent plants, although this technique is not particularly suitable forurban wastewater treatment systems. In tropical regions, seasonal transloca-tion activity is very low, and several harvests can be made during the year,so plant uptake could play a significant role in nitrogen removal, especiallyin lightly loaded systems (Vymazal, 2007). The potential rate of nutrient up-take by plants is limited by their net productivity and the concentration ofnutrients in plant tissues.

The nitrogen standing stock in emergent wetland plants is in the range14–156 g N/m2, but more than half of this amount may be stored below

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ground (Vymazal, 2007). Vymazal (2007) found that aboveground N standingstock values reported in previous studies were in the range 0.6–88 g N/m2.

Lim et al. (2001) studied N content in the roots, rhizomes and leavesof Typha latifolia and recorded nitrogen uptake rates of between 0.45 and0.49 g m−2 d−1. Bachand and Horne (2000) studied the nitrate removal effi-ciency of three plant groups: bulrush (Scirpus spp.), cattail (Typha spp.), anda mixed stand of macrophytes and grasses including smartweed (Polygonumlapathifolium), duckweed (Lemna spp.) and barnyard grass (Echinochloacrusgalli). Significant differences were observed between the three groups:cattail removed 0.565 g N m−2 d−1, bulrush 0.261 g N m−2 d−1, and the mixedstand 0.835 g N m−2 d−1.

Tanner et al. (2002) conducted tests with experimental horizontal SSFCWs operating in series and planted with Schoenoplectus tabernaemontaniand obtained average net rates of plant uptake of between 0.28 and 0.47 gm−2 d−1. Mayo and Bigambo (2005) carried out experiments in two horizontalSSF CWs planted with Phragmites mauritianus and found that total N uptakewas 0.297 g N m−2 d−1, of which 0.140 g N m−2 d−1 was returned to thewater body as plant litter. The results of these two studies show that plantuptake is a responsible for less nitrogen removal than denitrification.

Phosphorus

Phosphorus removal rates are rather low in CWs. Vymazal (2005a) reviewedphosphorus removal rates in horizontal SSF CWs throughout the world andcalculated an average mass removal rate of 45 g P m−2 year−1 and an averagemass-based efficiency of 32%. Average effluent concentrations (mostly in sec-ondary treatment) was 5.15 mg L−1, which is above the permitted dischargelimit in most countries.

P-removal rates in vertical systems seem to be more variable than N-removal rates, and different authors have reported average concentration-based total phosphorus (TP) removal efficiencies of 25–40% (Verhoeven &Meuleman, 1999), 26–70% (Rousseau et al., 2004b), 0–40% (Esser et al., 2004),and 45–91% (Weedon, 2002). In all of these studies, TP effluent concentra-tions (mostly in secondary treatment) were above 2 mg L−1 and increasedover time.

There are three main factors that lead to low P-removal rates. First,microbial phosphorus removal is only a temporary sink. Second, as in thecase of nitrogen, plant uptake tends to have a relatively small effect on re-moval. Third, most substrates used as granular media have low P-sorptionand P-complexation capacities: in the case of gravel, from the start of theoperating period; in the case of sandy media, after a limited period oftime due to the depletion of sorption sites and complexing agents. De-tailed descriptions of these three mechanisms are given in the followingsections.

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MICROBIAL REMOVAL

Unlike denitrification, in which nitrite and nitrate are mainly converted intoharmless nitrogen gas that escapes from the wetland, microbial P-removalexhibits no similar sink. Bacteria can only take up and store P, which isa partly reversible removal mechanism. The continuous cycle of growth,die-off, and decay releases most of the initially assimilated phosphorus, andonly some refractory fractions become a permanent sink for P. Consequently,once the CW start-up phase has been completed, net microbial P-removal isgenerally very low. Mander et al. (2003b) recorded a cumulative P-retentionof 52.8 kg in a 40-person-equivalent (PE) horizontal SSF CW after an oper-ating period of five years, of which only 4.4% (2.3 kg) was due to microbialimmobilization. To the best of our knowledge, there is only one preliminarystudy on the potential for enhanced biological phosphorus removal (EBPR)in constructed wetlands due to the presence of phosphate-accumulating or-ganisms (PAOs). In this study, Alas et al. (2003) showed that different waterregimes and therefore different oxygenation conditions could stimulate mi-crobial P-removal. Recent findings made by Edwards et al. (2006) confirmthat microbial P-consumption is higher in aerobic regions of the rhizosphere,whereas P-mineralization is predominant in the anaerobic regions.

EXPORT THROUGH PLANT UPTAKE AND HARVESTING

According to Davies and Cottingham (1993), the P-uptake capacity of aquaticmacrophytes in CWs is very limited—about 6% of the influent load. Studiesof mature wetlands confirm this observation and show that the capacityvaries according to the plant species in the system, the climate, and theP-loading rates. Edwards et al. (2006), for example, found that in a five-year-old horizontal SSF CW planted with Phalaris arundinacea, plant uptake onlyaccounted for 1.5% of the annual P input, whereas Tanner (2001b) analyzeda system planted with Schoenoplectus tabernaemontani and recorded valuesof 6–13%.

The P-uptake capacities reported by different authors are fairly similar:50–150 kg P ha−1 year−1 (Brix, 1994b); 162 kg P ha−1 year−1 (Wood, 1995);55 kg P ha−1 year−1 (Drizo et al., 1999); and 80 kg P ha−1 year−1 (Meule-man, 1999). Radoux and Kemp (1982) harvested the aboveground biomassof an experimental CW in Viville, Belgium, and were able to export approx-imately 37 kg P ha−1 year−1, whereas Mandi et al. (1996) conducted tests ina warmer climate in Morocco and were able to export 62 kg P ha−1 year−1.Vymazal (2005a) indicates that plant uptake may have a stronger effect onP-removal in tropical countries because the standing crops are taller and notranslocation occurs in the autumn.

However, even low P-uptake capacities can have a considerable effecton P-removal in low-loaded tertiary treatment wetlands. Huett et al. (2005)

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studied the treatment of plant nursery run-off (average influent concentra-tion of 0.58 mg P L−1) in horizontal SSF CWs and found that plant uptakeaccounted for approximately 86% of the P-removal; the overall P-removalefficiency was higher than 96%. Unplanted systems performed far worse atefficiencies below 45%, and in some cases no net removal was observed.Interestingly, the shoots showed signs of nutrient deficiency, which suggeststhat the potential P-removal efficiency was even higher than that recorded.However, only 13–19% of the influent P load was stored in the shoots,whereas 86–96% was stored in the belowground biomass. Harvesting hastherefore only a low potential to export P from the system. In his reviewof horizontal SSF CWs in the Czech Republic, Vymazal (2002) confirms thatharvesting removed less than 10% or even 5% of P. However, when theplants are not harvested, phosphorus is partially re-released as the senescentplants decay and is stored partly in the detritus layer on the bed surface.

ADSORPTION/DESORPTION AND CHEMICAL PRECIPITATION

The degree of phosphate adsorption by the granular medium depends mainlyon its texture and grain size distribution, the Fe content and, to a lesserextent, the Al and Ca content. Gravel is the most widespread medium inhorizontal SSF CWs because it does not clog easily. It only adsorbs smallamounts of phosphorus because it has a coarse texture (small surface-to-volume ratio) and generally contains low levels of Fe and Al. In addition,the binding sites on the gravel usually become saturated within several weeksor months after the start-up, which further reduces the potential effect of theadsorption pathway on P-removal. Even net P export has been reported as aresult of P desorption during rainweather flows (Huett et al., 2005; Korkusuzet al., 2005). Luderitz and Gerlach (2002) describe that phosphates can alsobind to humic substances produced by the breakdown of dead wetlandvegetation. They showed that P-removal decreased by 50% when the plantswere removed from the horizontal SSF CW, and concluded that adsorptionto organic matter might be an intermediate step before chemical binding.

Another important process is chemical precipitation, during which phos-phates react with certain metals and form insoluble compounds. In oxi-dized and neutral-to-acidic conditions, Fe3+ and Al3+ compounds such asaluminium and ferric (hydroxy) phosphates are formed. However, underanaerobic conditions, ferric and aluminium phosphate compounds are re-dissolved, and ortho-phosphate ions are released into the water column.Under alkaline conditions, various Ca2+ and Mg2+ complexes prevail.

Recent research has focused mainly on identifying substrates for SSFCWs that combine high hydraulic conductivity with high P-sorption capacity,and on screening different chemical substances that can be used to amendthe medium to stimulate chemical precipitation. As stated by Del Bubbaet al. (2003), most of these studies do not differentiate between the sorption

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and precipitation reactions and consider them as a lumped process. In thefollowing paragraphs, we give an overview of the latest research in this area.

Drizo et al. (1999) studied bauxite, shale, burnt oil shale, limestone,zeolite, light expanded clay, and fly ash, and found that fly ash had thehighest P-adsorption values and maintained a sufficient level of hydraulicconductivity. Total P-removal was 650 mg kg−1 and 730 mg kg−1 duringshort-term batch and long-term column experiments, respectively. Arias et al.(2001) screened 13 Danish sands and found that their P-removal capacitieswere influenced more strongly by Ca-content than by Fe- and Al-contentbecause wastewater is normally slightly alkaline. P-removal ranged from20 to 165 mg kg−1 in the different types of sands. Wollastonite, a calciummetasilicate mineral mined in upstate New York, was proven to be a goodmedium for P-sorption, provided the contact time was long enough (i.e.,40 hours and above (Brooks et al., 2000). Korkusuz et al. (2005) comparedP-removal in wetlands filled with either gravel or blast furnace slag andconcluded that the latter substrate exhibited far better P-sorption capacities(up to 9150 mg kg−1). Drizo et al. (2006) also obtained reasonably goodresults with steel slag (up to 2200 mg kg−1). Field results showed that a full-scale steel slag filter treating effluent from waste stabilization ponds (averageflow 2000 m3 d−1) yielded sustained P-removal for five years at averageeffluent concentrations down to 2.3 mg P L−1, which produced a P-retentionratio of 1230 mg kg−1.

Pant et al. (2001) compared Lockport dolomite, Fonthill sand, andQueenston shale, and concluded that Fonthill sand not only had the highestP-sorption capacity (417 mg kg−1 for a new batch of sand) but also seemedto yield lower equilibrium P-concentrations and therefore lower P-effluentconcentrations. This study also showed that sorption characteristics appearto change over time due to the weathering of the material. Pant et al. (2001)therefore concluded that analyzing the maximum P-sorption capacity of amedium during the design or start-up phases may not produce reliable re-sults, and that repeated measurements are needed. Del Bubba et al. (2003)also compared different types of sand and concluded that P-sorption wasdetermined by Ca and Mg content, grain size, porosity, bulk density, andhydraulic conductivity. However, these authors also made a clear distinctionbetween sorption and precipitation. They showed that sorption only lastedfor about one year and reached maximum values in the range 35–79 mgkg−1, even for sands with high sorption capacities. These results prove thatP-precipitation must be stimulated to ensure sustained P-removal over longerperiods of time.

Søvik and Kløve (2005) studied P-retention in shell sand filter systems.They observed that sorption and precipitation occurred simultaneously butfound it difficult to distinguish between the two processes. Interestingly, thelaboratory-scale experiments in this study revealed that the soil-water ratiohad a very strong influence (up to a factor 10) on the maximum P-sorption

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capacity. A more realistic meso-scale experiment was therefore established,indicating a P-sorption capacity of shell sand of 285 mg kg−1. Prochaska andZouboulis (2006) issue similar warnings about experimental conditions. Theyremark that the use of synthetic influents and new filter material containingno plant material may produce results that deviate a lot from what willhappen in full-scale CWs.

Seo et al. (2005) studied the P-retention capacity of various substrates toestimate the potential longevity of a horizontal SSF CW for P-removal. Theyfound that P-removal increased considerably when oyster shell was added.Organic matter also appeared to adsorb large quantities of P, but there is arisk that P may be released again during degradation. The authors also foundthat P-sorption increased with decreasing particle size of the substrate.

Westholm (2006) published an extensive review of P-sorbing substrates.Although it is difficult to standardize and compare the results of differentstudies, Westholm (2006) concluded that Wollastonite, slag material, andLECA are the most suitable materials for this process. However, LECA hasfairly high production costs and might not be a sustainable solution, partic-ularly if we consider that LECA would need to be replaced approximatelyevery five years (Heistad et al., 2006) and the useful lifetime of a CW isusually in the order of 25 years.

As stated above, chemical substances can be amended to stimulate pre-cipitation reactions. In order of effectiveness, FeCl3, alum, Ca(OH)2, calcite,and dolomite were found to reduce soluble P-content substantially (Annet al., 2000; Reddy & D’Angelo, 1997). Luderitz and Gerlach (2002) alsoproved that adding iron filings was more effective than using Ca-rich soils.

Recent research has addressed the issue of how to manage saturatedfilter material after the SSF CW has been taken out of operation. Westholm(2006) suggests that certain substrates could be reused as fertilizers if thelevels of simultaneously sorbed and/or precipitated heavy metals are withinsafe limits. Kvarnstrom et al. (2004) tested the plant-availability of P adsorbedonto sand and LECA. The accumulated P occurred mainly as an inorganicand mobile fraction and was therefore freely available to rye grass. In theirextensive review of P-removal technologies, De-Bashan and Bashan (2004)conducted an extensive review of P-removal technologies and concludedthat P-recycling is preferable to P-removal sensu stricto.

EFFECT OF DESIGN PARAMETERS AND SEASONAL INFLUENCES

Although the composition of the granular medium material is the designparameter that has the greatest influence on P-removal, Garcıa et al. (2005a)also examined the effect of other key parameters, such as the hydraulic load-ing rate (HLR), the water depth, and the medium size. Shallow horizontalSSF CWs with a depth of 0.27 m removed greater quantities of dissolvedreactive phosphorus (DRP) than wetlands with depths of 0.5 m, but only

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during the first two growing seasons. This was due to several factors, includ-ing the higher redox potentials and oxygen concentrations in the shallowbeds and the greater degree of contact with plant roots. In addition, duringthe first growing season, the beds filled with finer gravel produced lowereffluent DRP concentrations than the beds filled with coarser material, pos-sibly because the finer gravel has a higher specific surface area. However,by the third growing season, the sorption sites had become saturated, andno significant differences were observed between beds with different depthsand medium sizes. The HLR appeared to be the most influential parame-ter throughout the experiment. Higher HLRs consistently produced higherP-effluent concentrations, which was confirmed by Drizo et al. (2006) in aseries of column studies. One should, however, note that in the experimentsof Garcıa et al. (2005a), DRP removal was below 22% and mostly even be-low 10%, indicating that significant percent differences had a low impact inabsolute terms.

P-removal in SSF CWs occurs mainly through non-biological processes,so Vymazal (2002) suggested that temperature only has a noticeable effect inlow-loaded systems with high levels of plant uptake, because P-removal istherefore related to the plant growth cycle. McCarey et al. (2004) monitoreda cold-climate horizontal SSF CW and reported net P-removal during allseasons except spring, when P from decaying plants appeared to be flushedout of the system due to the effects of snow melting and intensive rainstorms.

FUTURE DEVELOPMENTS

It is clear from the findings described above that when influent P-loadsare high and effluent-P standards are fairly strict, additional treatment isneeded to guarantee long-term compliance. Vohla et al. (2005) obtainedpositive results by adding a third-step oil-shale filter after treatment in a hy-brid vertical-horizontal SSF CW. Brix et al. (2001b) and Arias et al. (2003a)conducted similar trials with a polishing calcite filter but obtained less pos-itive results because saturation occurred after only a few months and thecalcite filter clogged easily. Based on these latter studies, the new Danishguidelines for on-site treatment of domestic wastewater with vertical SSF CWsnow recommend that the systems should be extended to include chemicalprecipitation of P with aluminium polychloride in the sedimentation tankwhenever P-removal of 90% or more is required (Brix & Arias, 2005). Like-wise, physicochemical P-removal by means of FeCl3 injection in the influentdistribution system has been attempted by Esser et al. (2004). Though theirlab-scale results were very good, field results were less conclusive because ofleaching of previously accumulated P. Finally, Meers et al. (2006) suggestedusing flocculants for physicochemical P-removal from the liquid fraction ofpig manure prior to treatment in a constructed wetland. Though these latter

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methods enhance P-removal considerably, it is important to note that theyalso produce higher sludge volumes and entail higher exploitation costs.

HEAVY METALS AND METALLOIDS

Kadlec and Knight (1996) reported a wide scatter in the literature informationabout heavy metals in CWs. Whereas many authors have studied the behaviorof heavy metals in CWs used to treat mine drainage (Dunbabin & Bowmer,1992; Wieder, 1989), very little research has been carried out with CWs usedto treat urban wastewater, particularly subsurface-flow systems (Vymazal,2003). However, growing awareness of the consequences of long-term metalaccumulation in CWs for domestic wastewater treatment has led to increasedresearch in this area in recent years.

Two approaches have been used to study metal removal processes:analysis of metal accumulation in the compartments of full-scale systems, andanalysis of removal processes, influencing factors, and distribution of metalsin more controlled environments in laboratory- and pilot-scale experiments.Experiments in controlled environments provide reliable information on thedifferent metal removal processes and can be used to enhance the designcriteria for new systems (Song et al., 2001). This information is particularlyimportant for systems that process influents with high metal loads, so mostof the laboratory- and pilot-scale experiments reported in the literature havebeen carried out with wastewaters with high metal concentrations, such asmine drainage.

When heavy metals enter a SSF CW, they are distributed between thecompartments. The main pools are the granular medium, the water column,and the vegetation (Dunbabin & Bowmer, 1992; Sheoran & Sheoran, 2006).The major processes responsible for the removal of heavy metals in CWs arebinding to the granular medium, sediments, particulates, and soluble organ-ics; precipitation as insoluble salts, mainly sulfides and (oxy-)hydroxides;and uptake by plants and bacteria. Gaseous removal mechanisms have avery limited effect, except in the cases of As, Hg, and Se. Sheoran and She-oran (2006) reviewed heavy metal removal in CWs used to treat acid minedrainage and subdivided the mechanisms into physical, chemical, and bio-logical processes, although they stressed that the different processes are notindependent of one another.

In this section, we present the most important physicochemical pro-cesses for removing heavy metals in the granular medium, discuss the rel-ative importance of heavy metal uptake by wetland plants and the indirecteffects of plants, and describe the effects of design parameters on the differ-ent removal mechanisms. As shall be seen, metals accumulate near the inletzone of the wetlands, with the exception of Mn and Fe, which have dynamics

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strongly affected by redox conditions. We will also show that metal removalby accumulation in macrophytes is low in comparison to other mechanisms.

Physicochemical Removal Processes in the Granular Medium

Major physicochemical removal processes of heavy metals occurring in thegranular medium of SSF CWs include sedimentation and filtration; sorption;and precipitation and co-precipitation. Sedimentation and filtration are im-portant physical processes that remove heavy metals associated with partic-ulate matter. Sorption is a broad term that can refer to adsorption and/orprecipitation. Heavy metals are adsorbed by surfaces either electrostati-cally by physical adsorption, which produces relatively weak complexes,or chemically by chemisorption, which produces strong complexes (Evan-gelou, 1998). Metals adsorbed on the surface of the granular medium and/orsediment can be substituted by other cations via cation exchange process.Fine-textured media containing appreciable amounts of organic matter accu-mulate heavy metals, whereas coarse-textured media generally have a lowaffinity for these elements (Gambrell, 1994).

In addition to the sorption and sedimentation processes describedabove, metals can also precipitate with (oxy-)hydroxides, sulfides, carbon-ates and other anions when solubility products are exceeded. The stability ofmetals precipitated as inorganic compounds is controlled primarily by pH.At near-neutral or alkaline pH, metals are effectively immobilized (Gambrell,1994). Bacterial production of bicarbonate by sulfate reduction, or the pres-ence of limestone in the medium, can lead to sufficiently high bicarbonatelevels to form precipitates with metals. Metal carbonates are less stable thanmetal sulfides but can influence the initial trapping of metals (Sheoran &Sheoran, 2006).

Hydrolysis and/or oxidation of metals leads to the formation of (oxy-)hydroxides. These oxides show very low solubility in the neutral-alkalinepH range found in many environments (Evangelou, 1998). Co-precipitationof metals with Fe- and/or Mn- (oxy-)hydroxides is another influential re-moval pathway in oxidized environments. However, co-precipitation withFe- and/or Mn- (oxy-)hydroxides is redox-sensitive and is not consideredto be a long-term removal mechanism (Sheoran & Sheoran, 2006). Theseoxides are unstable in reducing conditions (Eh < 100 mV), which leads tothe release of Fe, Mn and co-precipitated metals.

In anaerobic conditions and in the presence of an organic carbon source,sulfate is reduced by sulfate-reducing bacteria to sulfides. The sulfides thenreact with divalent metals to form insoluble metal sulfide precipitates (Gam-brell, 1994). Machemer and Wildeman (1992) suggested that metal sulfideprecipitation is the main long-term removal process in CWs used to treatacid mine drainage once the adsorption sites have become saturated.

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The type of flow (horizontal or vertical) in SSF CWs affects the pre-vailing oxidation-reduction conditions and, in turn, the metal removal pro-cesses. The prevalent conditions in vertical SSF CWs are usually aerobic andanoxic because the granular medium is exposed to the air during intermit-tent feeding. In contrast, the conditions in the bulk water of horizontal SSFCWs are typically anaerobic, although aerobic micro-sites are present in theplant rhizosphere due to radial oxygen loss (Stottmeister et al., 2003). Nextto sedimentation, filtration, and sorption processes, which are common inboth types of SSF CWs, it therefore seems logical that precipitation and/orco-precipitation with (oxy-)hydroxides is a removal process more typical ofvertical systems, whereas precipitation with sulfides can be an importantremoval processes in horizontal SSF CWs.

Most researchers who studied the distribution of heavy metals in full-scale SSF CWs found that heavy metals accumulate mainly in the granularmedium located closest to the inlet zone. Table 6 shows the range of metalconcentrations in the sediments or granular media of different horizontalSSF CWs and a hybrid system, as well as some of the principal designparameters. Gschloβl and Stuible (2000) found that Cu and Zn accumulatedto a considerable degree in the granular medium of two horizontal SSF CWsused to treat domestic wastewater after ten years of operation, particularlyclose to the inlet of the beds and in the organic layer on the medium surface.Maximum accumulation levels of 143 and 490 mg kg−1 were reported for Cuand Zn, respectively, in the horizontal SSF CW containing the finest medium.Vymazal and Krasa (2003), Vymazal (2003), and Lesage et al. (2007c) studiedthe accumulation of metals in the sediments of horizontal SSF CWs used totreat urban wastewater after three years of operation; all of these authorsfound that the metal concentrations were highest in the sediment close tothe inlet and decreased along the longitudinal profile of the bed. Obarska-Pempkowiak and Klimkowska (1999) analyzed a hybrid CW system usedto treat the effluent from a conventional wastewater treatment plant andalso observed that heavy metal accumulation was highest in the first sectionof the horizontal SSF CW. Although full-scale studies cannot be used toquantify the relative influence of different metal removal processes, it canbe clearly seen that the highest metal accumulation in SSF CWs is found inthe area closest to the inlet. This is thought to be the result of the followingprocesses:

• sedimentation and filtration of influent particulate matter and associatedmetals;

• sorption of metals onto organic matter accumulated close to the inlet area;and

• metal sulfide precipitation due to the input of organic matter and thecreation of reducing conditions.

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TA

BLE

6.

Met

alco

nce

ntrat

ions

inth

ese

dim

ent

or

gran

ula

rm

ediu

m(m

gkg

−1D

W)

of

SSF

CW

suse

dto

trea

turb

anw

aste

wat

er,

with

des

ign

par

amet

ers

and

refe

rence

s

Gsc

hlo

ßlan

dG

schlo

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dV

ymaz

al(2

003)

and

Lesa

geet

al.

Obar

ska-

Pem

pko

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kan

dSt

uib

le(2

000)

Stuib

le(2

000)

Vym

azal

and

Kra

sa(2

003)

(200

7c)

Klim

kow

ska

(199

9)

Country

Ger

man

yG

erm

any

Cze

chRep

ublic

Bel

gium

Pola

nd

Typ

eofSS

FCW

Horizo

nta

lH

orizo

nta

lH

orizo

nta

lH

orizo

nta

lH

ybrid

Per

sons-

equiv

alen

t67

510

065

035

015

0Su

rfac

ear

ea(m

2 )50

094

032

2413

0087

0Y

ears

ofoper

atio

n10

103

35

Gra

nula

rm

ediu

mSa

nd/g

rave

l∗Sa

nd

∗Cru

shed

rock

(8–1

6m

m)

Gra

vel(5

–10

mm

)So

il†

Pla

ntsp

ecie

sP

hra

gmit

essp

.,Typ

ha

sp.

Ph

ragm

ites

sp.,

Typ

ha

sp.

Ph

ragm

ites

au

stra

lis,

Ph

ala

ris

aru

nd

ina

cea

Ph

ragm

ites

au

stra

lis

Ph

ragm

ites

sp.

Met

alco

nce

ntrat

ions

(mg

kg−1

DW

)Cu

4–12

717

–143

12–9

7615

–288

18–6

5Zn

13–4

0252

–490

73–1

986

65–9

3423

–130

Cd

0.5–

3.4

0.3–

44.

5–16

60.

2–2.

51–

5.2

Ni

7–30

28–7

115

.1–3

4332

–42

Cr

10–3

538

–259

26–4

111

–41

Pb

4–25

7.3–

9052

–409

14–1

6210

–42

Mn

114–

1769

261–

947

60–2

18Fe

100–

1300

2345

8–35

351

∗ Gra

nula

rm

ediu

msi

zenotsp

ecifi

ed.

† Soil

med

ium

consi

sted

ofa

mix

ture

offine-

grai

ned

frac

tions

com

prisi

ng

clay

and

sand

inth

ehorizo

nta

lSS

FCW

,an

da

more

sandy

soil

inth

editc

hes

ofth

eca

scad

efilte

r.

608

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Several studies of full-scale SSF CWs highlight the importance of redoxconditions in the removal of Mn and Fe. Vymazal (2005b) conducted testsin a horizontal SSF CW (design capacity of 700 PE) with a total area of3520 m2 planted with Phragmites australis and Phalaris arundinacea andfound that Fe, Mn, and Ni concentrations were higher in the effluent than inthe influent. The significant reduction of the organic matter and the lack ofnitrates suggested that anaerobic conditions prevailed in the reed beds andthat Mn(IV) and then Fe(III) were reduced to Mn(II) and Fe(II), which aremore soluble and could therefore be partly washed out of the system. Ni wasco-precipitated with Mn- and/or Fe- (oxy-)hydroxides and showed a similarremoval pattern, which suggests that it was released when these compoundswere reduced. However, Obarska-Pempkowiak and Klimkowska (1999) andLesage et al. (2007a, 2007c) reported higher Mn concentrations in the sedi-ment toward the outlet of SSF CWs. Figure 6 shows the Mn concentrations inthe sediment of a horizontal SSF CW planted with Phragmites australis andused to treat urban wastewater of 350 PE in Flanders, Belgium, after threeyears of operation. The increasing trend of Mn concentrations in the sedi-ment could be attributed to the presence of reducing conditions close to theinlet area, which mobilizes the Mn(II), and the oxidizing conditions towardthe end of the wetlands, which causes oxidation to the insoluble Mn(IV).Garcıa et al. (2003a) found that the redox potential increased with the dis-tance from the inlet in horizontal SSF CWs used to treat urban wastewater,which supports the above findings.

Lim et al. (2001) demonstrated the importance of sedimentation andfiltration. They studied Cu removal in laboratory-scale horizontal SSF CWs(1.20 m long, 0.59 m wide) planted with Typha augustifolia or left unplanted.All wetlands were filled with layers of sand (<4 mm), pebbles (4–25 mm),

0

0.5

1

1.5

2

2.5

1 5 25 50Distance from inlet (m)

Mn

conc

entr

atio

n (g

kg-

1 D

W) left middle right

FIGURE 6. Mn concentration in the sediment of a horizontal SSF CW as a function of distancefrom the inlet and position with respect to the water current direction (left, middle, and rightside of the wetland, with a total width of 13 m). From Lesage et al. (2007c), with permission.

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and stones (>25 mm) to a depth of 0.2 m. The systems received primary-treated sewage spiked with increasing Cu levels (in mg L−1): 5.52, 10.06,23.75, 36.32, and 72.31. The retention time varied from 3.5 to 4.9 days.Cu was completely removed from the wastewater during the experimentalperiod of 141 days. As the dissolved Cu fractions accounted for only 3.0and 5.9% of the influent Cu load, because of the high organic content ofthe influent water, sedimentation and filtration appeared to be the majorremoval mechanisms.

Cheng et al. (2002b) demonstrated that heavy metals are re-moved efficiently from artificial wastewater in laboratory-scale twin-shapedvertical/reverse-vertical SSF CWs planted with Cyperus alterniflorus and Vil-larsia exaltata. Both wetlands in the system had surface areas of 1 m2 andwere filled with sand/gravel (0–8 mm) to depths of 0.55–0.45 m. The wet-lands had a shared drainage layer at the bottom that was filled with gravel(16–32 mm). After a preparation period of about 14 months to encourage thegeneration of bio-films on the surface of the sand, the systems were intermit-tently fed with 40 L of artificial wastewater six times a day with the followingcomposition (in mg L−1): 0.79 Al, 0.009 Cd, 1.04 Cu, 0.302 Mn, 0.0103 Pb,and 4.25 Zn. All of the metals except Mn were removed completely in thefirst wetland during the 150-day experimental period. The sand in the toplayer generally adsorbed more metals than that in the bottom layer.

Biological Removal Processes: Plant Uptake and Indirect Effects

Salt et al. (1995) compared the effect of plants with that of solar-drivenpumps, which can extract and concentrate certain elements from their en-vironment. All plants can extract essential metals, such as Fe, Mn, Zn, Cu,Mg, Mo, and possibly Ni from soil and water. Certain plants can also ac-cumulate metals or metalloids, such as As, Cd, Cr, Pb, Co, Ag, Se, and Hg.There are many reasons for investigating the uptake and storage of metals inwetland plant species. One of the most important ones is the potential bio-concentration of metals that can pose a health risk to animals and humans(Karpiscak et al., 2001).

Metal accumulation in plants usually accounts for only a negligible pro-portion of total metal removal in CWs (Stottmeister et al., 2003), and thistrend has been observed in several full-scale studies of SSF CWs used totreat urban wastewater (Lesage et al., 2007c; Vymazal, 2003; Vymazal &Krasa, 2003) and a mixture of domestic wastewater and dairy parlor effluent(Mantovi et al., 2003). Lim et al. (2001, 2003) also found that Typha angus-tifolia made a negligible contribution to the removal of Cu, Cd, Pb, and Znin laboratory-scale horizontal SSF CWs. Accumulation in the biomass of Ty-pha angustifolia accounted for less than 1, 2.5, 2.6, and 0.8% of the influentloads of Cu, Cd, Pb, and Zn, respectively. However, in other laboratory-scale

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experiments plants (mainly the fine roots) were found to make higher con-tributions to the overall metal removal. Sinicrope et al. (1992) found thatapproximately 35% and 13% of the mass loading of Cd and Zn, respectively,was accumulated in the fine roots of Scirpus lacustris growing in SSF CWmesocosms subjected to different hydroperiods, which varied between con-tinuous inundation and discharge, with two pulsed discharges per day. Theshoots, rhizomes, and coarse roots each accumulated less than 1%. Plant up-take was also found to be an influential removal process in a study by Chenget al. (2002b), who observed that more than 30% of Cu and Mn and 5–15%of the other influent metals were accumulated by Cyperus alterniflorus ina laboratory-scale twin-shaped vertical/reverse-vertical subsurface-flow con-structed wetland. The largest proportion of the metals was stored in thelateral roots, which form a layer at the (oxygenated) top of the vertical SSFCW. Higher metal concentrations are commonly encountered in roots thanin aboveground parts of plants (Weis & Weis, 2004). Metals sequestered inthe belowground plant parts are eventually buried in the sediment.

It is generally thought that the most effective removal processes of heavymetals in SSF CWs are physical and/or microbial. Although plants take uponly small quantities of metals, they are integral parts of CWs, and theirinteraction with the granular medium and microorganisms is thought to bean important process (Dunbabin & Bowmer, 1992; Williams, 2002). Plantscan indirectly affect metal removal mechanisms in SSF CWs by excretingprotons and exudates into the rhizosphere, by introducing organic matterinto the system when they decay, and by causing radial oxygen loss in therhizosphere.

Plant roots can excrete protons and exudates into the rhizosphere,which acidify and mobilize metals (Salt et al., 1995; Weis & Weis, 2004).Salt et al. (1995) described the excretion of metal-chelating molecules (phy-tosiderophores) that chelate and mobilize metals and favor their uptake byplant roots. Root exudates also increase microbial activity, buffering capacityand reducing power, and stabilize pH at approximately neutral levels (Dun-babin & Bowmer, 1992). Stottmeister et al. (2003) state that rhizodepositionhas a significant effect on metal removal in CWs that are fed with wastewatercontaining a low carbon load, which is the case of some industrial effluents.In contrast, they state that rhizodeposition is less influential in CWs used totreat urban wastewater.

If the aboveground biomass is not harvested, leaves and stems areeventually returned to the surface of the granular medium in the wetlands.Decaying plant biomass may introduce metals into the system through leach-ing and mineralization, but it can also act as a sink for metals. A number ofstudies in natural wetland systems have demonstrated that metal concentra-tions in plant litter generally increase with time. Different mechanisms forthis enrichment have been reported, such as adsorption of metals from the

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sediment, accumulation of fine particulates in the litter, and active uptake bythe microbial community (Weis & Weis, 2004).

Wetland plants are adapted to waterlogged conditions and can transportoxygen to the roots through the aerenchyma tissue. This oxygen transport isessential for respiration and causes an oxidative protective film—also calledplaque—to form on the root surface. Rhizosphere oxidation is particularly im-portant in horizontal SSF CWs because it creates aerobic microsites within theotherwise predominantly anaerobic conditions in the bulk water (Stottmeis-ter et al., 2003). By oxidizing the rhizosphere, plants can also remobilizemetals. The plaques are composed mainly of iron hydroxides and other met-als that are co-precipitated on the surface. There are contradicting reportsas to whether the presence of plaque reduces or increases the uptake ofmetals by the plants (Weis & Weis, 2004). Peverly et al. (1995) and Mantoviet al. (2003) identified an iron plaque on the roots of Phragmites australisplants in full-scale horizontal SSF CWs by analyzing root cross-sections witha scanning electron microscope.

Stein and Hook (2005) found that Scirpus acutus or Typha latifolia re-duced total removal of Zn in laboratory-scale SSF CWs in winter conditions.The wetlands (0.20 cm diameter) were filled with gravel (3–13 mm) andbatch-fed with synthetic mine-impacted water. Sucrose was added as a car-bon source to enhance bacterial sulfate reduction. Stein and Hook concludedthat the seasonal change in Zn removal was related to the fact that plantscreate more oxidized conditions in wetland systems during certain periods ofthe year. This effect is illustrated by higher redox potential, which limits sul-fate reduction and the precipitation of Zn with sulfides. Lesage et al. (2007b)also demonstrated that Phragmites australis plants reduced sulfate reductionand metal removal. They studied the removal of Co, Ni, Cu, and Zn fromsynthetic industrial wastewater in SSF CW microcosms (0.1 m diameter, 0.4m height) filled with gravel (3–8 mm) or a gravel/straw mixture (15% volu-metric straw content). Half of the microcosms were planted with Phragmitesaustralis and half were left unplanted. The systems were batch-fed every twoweeks, and the pore water was regularly monitored at two depths for heavymetals, sulfate, organic carbon, and redox potential. Glucose was added tothe microcosms to stimulate sulfate reduction and metal removal. The val-ues of all parameters indicated that sulfate reduction is a major pathwayfor heavy metal removal, and that plants limit sulfate reduction and metalremoval in the top layer of all microcosms.

Different researchers have studied the effects of plants on metal removalin laboratory-scale SSF CWs by using both unplanted and planted systemsin their experimental designs. Scholz et al. (2001), Scholz and Xu (2002),and Scholz (2003) found that the presence of Phragmites australis, Typhalatifolia, or both had no effect on the efficiency of Cu and Pb removal inlaboratory-scale vertical SSF CWs (0.58 m high, 0.32–0.40 m diameter) packedwith different filter media and used to treat simulated pre-treated minewater.

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The systems were operated in batch mode three times a week. Lee et al.(2006) obtained similar results in their study of Cu and Ni removal from gullypot liquor in vertical SSF CWs (0.83 m high, 0.10 m wide) filled with differentfilter media and either planted with Phragmites australis or left unplanted.The wetlands were operated in batch mode and periodically inundated anddrained to encourage air penetration. Lim et al. (2001, 2003) studied theremoval of Cu, Pb, Cd, and Zn in horizontal SSF CWs that were eitherplanted with Typha angustifolia or left unplanted. High removal efficiencieswere recorded in all systems and no plant effect was observed.

Buddhawong et al. (2005) related the removal of As and Zn inlaboratory-scale SSF CWs (0.5 m long, 0.3 m wide) to the presence of Juncuseffusus. The systems were operated discontinuously and were fed syntheticmining effluents contaminated with 5 mg L−1 Zn and 0.5 mg L−1 As, at pH4. As the gravel had a low adsorption capacity for As and the plants inhydroponic systems did not absorb large amounts of As, another removalmechanism must have been present. The authors suggest that organic com-pounds released by the plant roots favor a decrease in the redox potential,which leads to the dissolution of crystalline iron. In oxic zones, such asthose found in the rhizosphere, this dissolved iron precipitates again andcan co-precipitate other metals. The presence of this removal pathway wascorroborated by the patterns of Fe concentration and redox potential in thepore water, and by the higher As content in the roots of Juncus effususgrown in gravel beds than in hydroponic systems, due to the formation ofiron plaques. However, these observations only apply to mining and indus-trial effluents. In systems used to treat urban influents with higher organicloading, the contribution of organic compounds released by plant roots isassumed to be negligible. In addition, the organic loading would create morereducing conditions, at least in the section close to the inlet, and this wouldfavor other removal processes such as metal sulfide precipitation.

A general conclusion drawn from the abovementioned studies is thatin the case of the dominant processes being sedimentation, filtration, andsorption, and when the prevailing conditions are not strongly reducing, thepresence of helophytes either positively affected metal removal or did notaffect it at all. However, in SSF CWs in which metal sulfide precipitation isan important removal mechanism, the presence of helophytes can reducethe metal removal efficiency due to the radial oxygen loss by the plant roots.

More research is needed to assess the influence of indirect plant effectsin SSF CWs and constructed wetlands in general.

Design Factors That Influence Heavy Metal Removal Mechanisms

In this section, we review the two main design factors that affect heavy metalremoval: the granular medium and the hydraulic properties of the system.Many authors have conducted laboratory-scale and pilot-scale experiments

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to compare the effects of different types of granular media on the removalof heavy metals. Scholz et al. (2001), Scholz and Xu (2002), and Scholz(2003) studied Cu and Pb removal from simulated pre-treated minewater inlaboratory-scale batch-fed vertical SSF CWs packed with different media ofvarying adsorption capacities and costs. Different types of cheap washedgravel and sand media were tested, including large beach gravel (<40 mm),medium gravel (8–20 mm), small pea gravel (4–10 mm), coarse sand (0.3–4mm), and medium sand (0.15–2.36 mm). Filter media with a high adsorptioncapacity included Filtralite (a light expanded clay aggregate, 1.5–2.5 mm),granular activated carbon (12 × 40 US mesh size) and charcoal (a carbona-ceous adsorbent, 20 × 60 US mesh size). The authors suggested that sorptiononto the mature biofilm was responsible for the high removal efficienciesobserved in all systems. Removal efficiencies were higher than 94% for Cu,and Pb was completely removed in systems with a mature biofilm after 13months of operation (Scholz, 2003). Expensive adsorptive media such asFiltralite, granular activated carbon, and charcoal did not enhance metal re-moval. Similar results were obtained by Lee et al. (2006), who studied theremoval of Cu and Ni from gully pot liquor in batch-fed laboratory-scalevertical SSF CWs filled with different packing orders of different filter media.

Stark et al. (1996) studied the removal of Mn in unplanted laboratory-scale horizontal SSF CWs (1.60 m long, 0.25 m wide) that were filled toa height of 0.15 m with either spent mushroom substrate or limestone. Mnremoval was enhanced by the limestone in relation to the prevailing oxidizingenvironment. Sequential extraction of the medium was performed, revealingthat 72% of the Mn was present as oxide and about 16% as carbonates inthe limestone substrate. The authors made the general conclusion that Mnremoval is promoted in oxidizing conditions.

When metal sulfide precipitation is required, it is important to selecta suitable mixture of a granular medium and a carbon source. Chang et al.(2000) reported that raw biomass (e.g., oak chips) was initially a less efficientsubstrate type than biologically (e.g., compost) or chemically (e.g., sludgefrom a wastepaper recycling plant) treated biomass in promoting sulfatereduction and metal removal. Gibert et al. (2004) demonstrated that lowerlignin content in the substrate increased biodegradability and the capacityfor developing bacterial activity. Sheep manure was the most effective suc-cessful organic substrate in promoting sulfate reduction, whereas compostwas too poor in carbon. The results of the abovementioned laboratory-scalestudies, which test the effectiveness of different substrates in the passivetreatment of acid mine drainage (e.g., in permeable reactive barriers), couldalso be applied to the enhancement of sulfidogenesis in CWs. Song et al.(2001) studied the removal of Pb and Zn in unplanted laboratory-scale hori-zontal SSF CWs (0.9 m long, 0.35 m wide) filled with different mixtures of thefollowing substrate types: aged manure, hay, bark, gravel, peat moss, sand,and aged sludge. The systems were fed continuously with synthetic mine

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water at varying flow rates and had a hydraulic retention time of 2–7 days.The sludge and manure were added to stimulate bacterial growth; the hay,bark, and peat moss were added as main carbon sources; and the gravel andsand were used to improve hydraulic conductivity. The authors recorded av-erage removal percentages of 90% for Pb and 72% for Zn, independently ofthe substrate composition. The influent and effluent data could not be usedto distinguish between removal pathways. Therefore, processes other thanmetal sulfide precipitation may also have contributed to the metal removal.However, the reduction of sulfates and the detection of sulfides in the ef-fluent suggest that bacterial sulfate reduction accounted for a considerableproportion of the metal removal in this system.

Hydraulic design is considered by some authors to be the most impor-tant factor in controlling metal removal processes in constructed wetlands(Sheoran & Sheoran, 2006; Sinicrope et al., 1992). The major hydraulic pa-rameters that influence the effectiveness of metal removal processes includethe loading application (continuous or intermittent) and the hydraulic reten-tion time (HRT).

Metal sulfide precipitation is generally reduced at higher loading ratesand with shorter retention times. This hypothesis has been corroboratedby laboratory-scale experiments with acid mine drainage in which bacterialsulfate reduction was found to induce metal removal (Chang et al., 2000;Gibert et al., 2004). On the contrary, Song et al. (2001) did not observe aclear effect of HRT (when ranged from two and seven days) on the removalof Pb and Zn from synthetic mine water in laboratory-scale horizontal SSFCWs. This could be caused by the range of HRTs tested and the prevalenceof removal mechanisms other than sulfide precipitation. Stottmeister et al.(2003) suggested that the HRT affects the extent to which plants influencepollutant removal. They assume that plants have a greater effect on metalremoval in horizontal SSF CWs, which typically have long retention times,than in vertical SSF CWs, which usually have much shorter retention times.

Sinicrope et al. (1992) demonstrated the importance of the loading ap-plication strategy in their study of metal removal in mesocosms plantedwith Scirpus californicus subjected to different hydroperiods. The authorsrecorded the highest removal efficiencies in the mesocosms that wereflooded and drained twice a day. The metals were presumably removedthrough the formation of complexes with oxides and hydroxides, and lowertreatment efficiencies were generally observed in the continuously inun-dated treatment. The redox potential (mean value of ± 0 mV) and the lowsulfide levels in the effluent (± 0.02 mM S2−) indicate that significant sulfatereduction did not occur in this treatment, which may be due to the shortexperimental residence time of only 1–1.3 days.

The importance of the hydraulic design was emphasized by Munguret al. (1997). They studied the removal of Cu, Pb, and Zn from simulatedsurface run-off in a laboratory-scale horizontal SSF CW (2.24 m long, 0.6 m

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wide) filled with gravel to a depth of 0.25 m and planted with Typha lati-folia, Phragmites australis, Schoenoplectus lacustris, and Iris pseudacorus. Avery short HRT of four hours was used. Notably, the plants in this experi-ment were rooted in permeable buckets filled with peat. Short-circuiting andpreferential flow paths were present, as evidenced from the non-uniformlydistributed patterns of metal concentrations in the peat surrounding the rootsand the metal concentrations in the plants. Poor hydraulic design could leadto non-optimal use of the filter substrate and helophytes and thereby de-crease performance with time.

MICROBIAL PATHOGENS

Constructed wetlands have been found to reduce microbial pathogens withvarying but significant degrees of effectiveness. Since the introduction ofconstructed wetland technology, a number of studies have focused on howSSF CWs can be used to improve microbial water quality (Falabi et al., 2002;Gersberg et al., 1987a, 1987b, 1989a, 1989b; Green et al., 1997; John, 1984;Karpiscak et al., 1996; Mandi et al., 1996; Williams et al., 1995).

In this section, we first evaluate the effectiveness of SSF CWs in re-moving different types of microbes and then focus on processes related tomicrobial removal and the effects of design factors on removal efficiency. Wewill see that literature provides extensive information on removal efficienciesof faecal bacteria indicators, but scarce data are available for specific bac-teria, viruses, protozoa, and helminths. In that sense, molecular techniquesseem to be a promising tool for evaluating the effectiveness of SSF CWsin the removal of microbial pathogens. As shall be seen, abiotic and bioticmechanisms involved in microbial removal are far from understood. Thereare mechanisms frequently cited in the literature that have not been provenat all.

Removal CapabilitiesFAECAL BACTERIAL INDICATORS

Many studies have used colony-forming unit (CFU) methods such as totalplate counts to evaluate the removal of faecal bacteria from wastewater inCWs. However, although these procedures for monitoring microbial con-taminants (APHA, 1998) have been established for many years, they do notnecessarily detect the real numbers of faecal bacteria and therefore do notensure optimum protection of human health. These conventional methodsrequire long incubation and detection times and cannot provide the predic-tive capabilities of more rapid protocols.

Most studies of faecal microorganism removal in CWs only describetotal and faecal coliform removal (Leonard, 2000; Perkins & Hunter, 2000;

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Reed et al., 1995; Tanaka et al., 2006; Tuncsiper, 2007; Wang et al., 2005).Research using experimental, pilot, and full-scale SSF CWs has shown thatfaecal coliform bacteria inactivation usually ranges between 1.25 and 2.5log units (Arias et al., 2003b; Decamp et al., 1999; Gersberg et al., 1989a,1989b; Green et al., 1997; Hiley, 1990; Ottova et al., 1997; Rivera et al.,1995; Tanner et al., 1995; Vacca et al., 2005; Vymazal, 2005c; Williamset al., 1995). However, faecal coliform inactivation rates of 3 log units andhigher have been recorded in tertiary horizontal SSF CWs treating slaugh-terhouse wastewater with extremely high influent concentrations of fae-cal bacteria (6–11 log units/100 mL). The high degree of inactivation ob-served in these wetlands was related to the high influent microbial con-centration (Rivera et al., 1995). Gross et al. (2007) reported faecal coliforminactivation ratios of 3–4 log-units in vertical SSF CWs used to treat greywastewater.

Several published papers on the removal of faecal bacteria do not clearlydefine the period of study (Wood & McAtamney, 1996) or use low samplingfrequencies and only make qualitative assessments (Ottova et al., 1997). Itmust also be taken into account that the concentration of coliform bacteriain wastewaters is subject to significant daily fluctuations, so the highest con-centration in the influent of a given CW system will not necessarily coincidewith the highest concentration in the effluent (Cooper et al., 1996; Greenet al., 1997).

Garcıa et al. (2006) found that SSF CWs generally produced highersurface removal rates of faecal bacteria (CFU removed m−2 d−1) and higherpercentage removal efficiencies (in %) than algae-based systems. In addition,SSF CWs were more efficient on faecal bacteria removal when consideringsurface removal rates than free water surface flow constructed wetlands.Considering percentage removal efficiencies, the results showed that higherremoval rates were observed in free water surface flow systems for mostgroups, except Clostridia and Staphylococci (Garcıa et al., 2006).

Although reductions of faecal coliforms and Enterococci sometimes arevery encouraging when viewed on a percentage basis (i.e., >99%), even forhigh bacterial content effluents, such as municipal landfill leachates (Sawait-tayothin & Polprasert, 2006, 2007), caution is warranted given that averageoutflow concentrations of these bacterial indicators are usually greater than104–105 and 103 CFU 100 mL−1, respectively. For total coliforms and fae-cal coliforms, the outflow concentrations are usually in the range of 102 to105 CFU 100 mL−1, while for faecal streptoccoci the range is between 102 and104 CFU 100 mL−1 (Vymazal, 2005c). Accordingly, the use of SSF CWs as asingle treatment process for wastewater containing initially high levels ofmicroorganisms may not be an adequate alternative when considering themicrobial quality. In order to further reduce microbial populations, post-treatment of the wetland-processed effluents (e.g., chlorine, UV disinfection,maturation ponds) appears necessary.

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Williams et al. (1995) expressed the performance of CWs in removingfaecal microorganisms in terms of decimal reduction distance (DRD), whichis the length of wetland that produces a 90% reduction in indicator concen-tration. A negative linear correlation was observed between the reduction ofcoliforms and the DRD in a horizontal SSF CW fed with primary and sec-ondary effluents. Decamp and Warren (2000) found that the concentration ofE. coli decreased exponentially with distance along the length of a SSF CW(r2 ranged from 0.708 to 0.906). The concentration of bacteria decreased byone order of magnitude in the first two-thirds of the wetland and then stabi-lized. This would indicate that an increase in hydraulic retention time (HRT)would not necessarily lead to a significantly better removal rate of E. coli. Thesame conclusion was reached by Garcıa et al. (2003b). These observationsare consistent with results obtained in trickling filters and activated-sludgeplants, in which adsorption and other removal processes are not necessarilyrelated to hydraulic retention time (Gray, 1989).

BACTERIAL PATHOGENS

Little research has been carried out into the effect of CWs on the removalof specific bacterial pathogens. Although removal efficiencies can vary con-siderably (Barrett et al., 2001), it seems that bacterial pathogens (such asSalmonella, Shigella, and Yersinia) are removed to a lesser extent than fae-cal bacteria indicators. Removal rates in the range of 1.5–5.3 log units havebeen reported, depending on the particular pathogen (Gearheart, 1999).

Different authors have examined the reduction of microbial pathogensunder variable wetland conditions (Gersberg et al., 1987a, 1987b; Hill &Sobsey, 2001; Neralla et al., 2000). Gersberg et al. (1989a) studied the sur-vival of Salmonella by artificially seeding the inflow water of a SSF CWand monitoring the effluents over a period of several days. The test cul-ture of Salmonella was reduced by 93–96% over HRTs ranging from 23 to52 h. Similar reductions (96%) in Salmonella populations have been reportedin swine wastewater passed through a full-scale SSF CW (Hill & Sobsey,2001).

VIRUSES

Very few studies have examined the fate of viral indicators of wastewaterorigin in CWs, and little is known about the ability of CWs to remove disease-causing viruses. The scarcity of data on disease-causing virus survival inCWs is due primarily to the lack of reliable, widely accepted, simple-to-use,and cost-effective virus indicators. The detection and enumeration of humanviruses in wastewaters is a costly and time-consuming process for whichhighly specialized equipment is required.

Bacteriophages have been proposed as ideal indicators of viral pol-lution (Kott et al., 1974; Stetler, 1984) and as models for the behavior of

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enteroviruses in natural waters (Wentsel et al., 1982) and during wastew-ater treatment (Funderburg & Sorber, 1985). The assay for detecting thesebacteriophages is relatively simple, fast, and inexpensive. F-specific RNA(F-RNA) bacteriophages can be used as indicators of the fate of human en-teroviruses (Havelaar et al., 1984), as they are very similar in physical sizeand structure (Cramer et al., 1976) and capable of surviving many sewagetreatment processes (Grabow et al., 1978). They are thought to be more re-sistant to inactivation by disinfection than enteroviruses (Cramer et al., 1976)and are relatively resistant to environmental inactivation by heat (Burgeet al., 1981) and sunlight (Kapuscinski & Mitchell, 1983). Polprasert andHoang (1983) found that the removal of indigenous bacteriophages (E. coliB host) in an anaerobic rock filter follows first-order kinetics, with a meanremoval of 85–90% with a four-day HRT and removal efficiencies of 95%for indigenous F-specific phages in an unplanted SSF CW with a 5.5-dayHRT.

Bacteriophages are removed by wastewater treatment processes (Fun-derburg & Sorber, 1985; Grabow et al., 1978) and from groundwaters (Yateset al., 1985) at rates that are comparable to or lower than those reported forenteroviruses. Bacteriophage removal in CWs usually ranges from 0.5 to 2 logunits (Barret et al., 2001; Gerba et al., 1999; Hagendorf et al., 2000; Williamset al., 1995). The average removal rates in horizontal SSF CWs used for sec-ondary treatment are slightly below 95% (Gersberg et al., 1987b; Thurstonet al., 2001). If we consider that CWs still provided at least 90% reductionin these viral indicators of faecal contamination, we can see that coliphageremoval never yielded the reductions of two or more log units observed formost of the bacterial groups.

Some studies focused on the survival of seeded viruses, even thoughthese may not be associated with the suspended particulate matter in thewastewaters, unlike native viruses (Dowd & Pillai, 1997; Dowd et al., 1998;Funderburg & Sorber, 1985; Wentsel et al., 1982). Gersberg et al. (1987b) ob-served 99% removal of the virus MS2 seeded into the influent of a horizontalSSF CW. In another SSF CW study, a 98% inactivation for MS2, PRD1, andindigenous bacteriophages was observed after 3.67, 13.38, and 2.05 days, re-spectively (Vinluan, 1996). The survival of indigenous F-RNA bacteriophagesin the CWs was greater than that of seeded MS2 phage and seeded poliovirustype 1, suggesting that the behavior of animal viruses in the wetland envi-ronment may be conservatively predicted from the fate of the indigenousF-RNA phages (Havelaar et al., 1984).

Vidales-Contreras et al. (2006) evaluated the removal of the virus PRD1in surface-flow and SSF CWs. Both types of wetlands reduced the concen-tration of PRD1, but the SSF CW showed greater removal efficiency than thesurface-flow system. There was no clear explanation for this discrepancy.

The apparent low efficiency of bacteriophage removal in SSF CWsmay indicate that these types of treatment systems cannot easily remove

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pathogenic viruses. Consequently, post-treatment disinfection process unitsshould be considered if virus removal is required.

PROTOZOANS AND PARASITES

Very little data are available on the removal of pathogenic microorganismsin CWs, including protozoans such as Giardia and Cryptosporidium and par-asites such as helminthes. SSF CWs have been found to remove substantialproportions of protozoans and parasites (Stott et al., 2001a, 2001b; Thurstonet al., 2001), but the removal efficiency is lower than the common valuesreported for total and faecal coliforms. In general, horizontal SSF CWs canachieve between 0.4 and 3 log units of removal (Stott, 2003). To our knowl-edge, there is no information available on parasite removal in vertical SSFCWs.

Karpiscak et al. (1996) found that total and faecal coliforms were re-duced by average proportions of 98% and 93%, respectively, whereas Gi-ardia and Cryptosporidium were reduced by 73% and 58%, respectively.Thurston et al. (2001) studied the fate of pathogenic protozoa in secondarysewage effluent passed through a horizontal SSF CW and observed that Gi-ardia cysts and Cryptosporidium oocysts were reduced on average by 88%and 69%, respectively. Giardia cysts are larger and can therefore be removedmore efficiently (Thurston et al., 2001).

Although parasite eggs and oocysts are removed by entrapment andsedimentation within the medium as wastewater flows through the SSF CW(Stott, 2003), the exact role of sedimentation, filtration, and other physicalprocesses is still unclear. Williams et al. (1995) suggested that parasite eggs(and cysts) may be removed close to the inlet of horizontal SSF CWs, asmost suspended solids are retained in this part of the system. Retention andentrapment within wetland biofilms and subsequent grazing by predatoryfauna may also improve removal and inactivation, particularly of protozoanoocysts (Stott et al., 2001a).

Gerba et al. (1999) found that protozoan removal efficiency was greaterin a horizontal SSF CW (69–88%) than a surface-flow system (58–73%), with aretention time of less than four days in each case. They attributed these resultsto the greater surface specific area of the SSF CW, which is conducive togreater adsorption and filtration. In this case, the SSF CW effluents contained0.01–0.07 oocysts L−1, and the surface-flow CW effluents contained 4.9–16 oocysts L−1.

Helminth removal in horizontal SSF CWs in Morocco (1–4 d of HRT)was related to the length of the wetlands (Mandi et al., 1996) and the seasonduring which measurements were taken (Mandi et al., 1998). Removal rateswere higher in 50 m length wetlands (95%) than in 20 m wetlands (71%),and were also higher during the hot season (90%) than during the cold

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season (78%) in all studied systems. In Egypt, helminth ova (Ascaris lumbri-coides) removal increased with wetland length and a concomitant reductionin hydraulic loading rates (Stott et al., 2003a).

Evaluation of Removal Efficiencies Using Molecular Techniques

Different intrinsic limitations of cultivation-based methods may hamper thedetection of pathogens: the interference of background flora, the impossi-bility of distinguishing between pathogenic and non-pathogenic strains, andthe very long cultivation steps (Rompre et al., 2002). In water and wastew-ater quality assessment, these limitations can ultimately lead to outbreaks ofdisease, either because pathogenic bacteria remain undetected or becausethe delay in obtaining results makes it difficult to plan adequate preventionstrategies (Haugland et al., 2005).

The scientific community and health authorities are becoming increas-ingly interested in adding molecular techniques to the standard methods fordetecting pathogens in water and wastewater. Although routine examinationfor pathogenic organisms is not recommended due to the high cost and thelow numbers of specific pathogens present at any given time, the devel-opment of fast new molecular techniques capable of detecting pathogensin environmental samples will undoubtedly have an effect on health andenvironmental microbiology in the coming years.

Quantitative polymerase chain reaction (PCR) (Heid et al., 1996) iswidely used in molecular biology to amplify sections of DNA by in vitroenzymatic replication and has become a popular method for detecting bac-terial pathogens rapidly and with high sensitivity and specificity. A detectionlimit of 7.9 × 10−5 pg of E. coli O157:H7 DNA 100 mL−1 equivalent to ap-proximately 6.4 × 103 CFU of E. coli O157:H7 DNA 100 mL−1 based on platecounts was determined for CWs, whereas quantification was possible whencell numbers were >3.5 × 104 CFU 100 mL−1 (Ibekwe et al., 2003).

Vacca et al. (2005) developed a taxon-specific primer set for detectingEnterobacteriaceae as a means of monitoring microbial water quality. Thistaxon is the predominant bacteria in wastewater, and most of the pathogenicbacteria or facultative pathogens belong to this group. Therefore, the taxon-specific primers were used to evaluate the fate of the Enterobacteriaceaeentering CWs and localize harbouring in the rhizosphere (Vacca et al., 2005).The results show that bacteria are strongly influenced by filtration process inwetlands, and that diversity decreases substantially from the inlet to the out-let. Different microbial communities exist depending on the granular mediumtype and the presence of plants (Vacca et al., 2005).

Szewzyk et al. (2006) used culture-independent methods to determinethe reduction of wastewater bacteria in vertical and horizontal SSF CWs andcompared their results with data derived from standard cultivation proce-dures. The bacterial concentrations determined by direct microscopic counts

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after DNA staining in the influent and effluent indicated that the bacterialremoval efficiency of the horizontal beds were generally low (32% and 35%)and variable, whereas the vertical beds displayed higher removal rates of93% and 99.9%. In contrast, the bacterial concentrations determined by con-ventional heterotrophic plate counts for the same samples revealed a similarremoval pattern for horizontal and vertical SSF CWs: both wetland types dis-played a similar and constant bacteria removal efficiency of 1–2 orders ofmagnitude. Discrepancies between direct enumerations and cultivation werecaused by a decrease in the fraction of culturable bacteria during treatment inthe horizontal wetlands, which led to overestimation of their actual removalefficiency. The reduction in culturability during treatment in the wetlandswas probably due to a shift in the physiological state of the wastewaterbacteria in response to nutrient depletion (Szewzyk et al., 2006). If bacteriaadapt to rather oligotrophic conditions, they may grow less well or not at allin the standard culture media for selection of heterotrophic plate counts.

Mechanisms of Microbial Removal

The ecology of microorganisms in CWs is extremely complex. Pathogenicbacteria and viruses are the most important organisms in terms of publichealth, although protozoan pathogens and helminth worms must also betaken into account in tropical and subtropical countries (Rivera et al., 1995).Although many authors have discussed possible mechanisms of bacterialremoval (Armstrong et al., 1990; Burger & Weise, 1984; Decamp & Warren,2000; Morales et al., 1996), no systematic analyses on the removal processesand the fate of potential pathogenic bacteria in CWs have been conducted.

SSF CWs offer a suitable combination of the physical, chemical, and bio-logical mechanisms required to remove pathogenic organisms. The physicalfactors include filtration and sedimentation (Gersberg et al., 1989a; Pundsacket al., 2001), and the chemical factors include oxidation and adsorption toorganic matter (Gersberg et al., 1989a). The biological removal mechanismsinclude oxygen release and bacterial activity in the rhizosphere, aggrega-tion and retention in biofilms (Brix, 1997; Hiley, 1995), potential productionof bactericidal compounds or antimicrobial activity of root exudates (Axel-rood et al., 1996; Kickuth & Kaitzis, 1975; Seidel, 1976), predation by nema-todes and protists (Decamp & Warren, 1998; Decamp et al., 1999), attack bylytic bacteria and viruses (Axelrood et al., 1996), natural die-off (Gersberget al., 1989a, 1989b), and competition for limiting nutrients or trace elements(Gersberg et al., 1987a, 1987b; Green et al., 1997). In this review, we classifythe mechanisms for microbial removal as either abiotic or biotic.

ABIOTIC MECHANISMS

As water passes through a CW system, pathogens are removed by physicaland chemical mechanisms such as sedimentation, filtration, adsorption, and

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oxidation. The sediments of SSF CWs have been found to accumulate highconcentrations of pathogens, which suggests that sedimentation and filtrationare the main abiotic mechanisms that reduce the pathogen content of thewater phase in these systems (Decamp & Warren, 1998; Gersberg et al.,1987b; Green et al., 1997). In the study of Karim et al. (2004), it was foundthat on a volume/dry weight basis, the numbers of faecal coliforms andbacteriophages were one to two orders of magnitude higher in CW sediment.Therefore, the bottom sediments in SSF CWs promote survival of bacteria andother microbial pathogens and could potentially act as a reservoir for humanpathogens, which could be released into the water column by storm or man-made events. Laboratory studies demonstrated that sediment was capableof protecting poliovirus type 1 from the inactivating mechanisms driven bymicroorganisms, heat, salt, and temperature (Labelle & Gerba, 1982; Liew &Gerba, 1980; Smith et al., 1978).

Sedimentation and filtration of pathogens in SSF CWs appears to be size-dependent as higher removal rates are recorded for large particles. Virusesassociated with large particles soon leave the water column and settle intothe bottom sediments, while viruses adsorbed on colloidal particles tend tostay suspended in the water for a longer time. Giardia and Cryptosporid-ium are much larger compared to bacteria and viruses, and therefore cansettle in a greater degree. In fact, several authors reported that Giardiaand Cryptosporidium concentrations were 2–3 orders of magnitude greaterin sediment than in the water column, which suggests that sedimentationmay be the primary reduction mechanism in CWs (Falabi et al., 2002; Karimet al., 2004; Karpiscak et al., 1996). However, it also appears that Giardiadoes not survive for long periods in the bottom sediments of CWs (Karpiscaket al., 1996) and that the degree of survival depends strongly on temperature(Deregnier et al., 1989; Johnson et al., 1997). Karim et al. (2004) observedthat Giardia die-off was greater in CW sediment (0.37 log10 d−1) than in thewater column (0.029 log10 d−1). Therefore, Giardia survival was greater inthe water column and worse in the sediment, which may be due to biolog-ical antagonism or the presence of organic substances that enhance die-offof protozoan parasites (Chauret et al., 1998).

Adsorptive processes may play a major role in the removal of mi-croorganisms in SSF CWs and are particularly effective at removing viruses(McConell et al., 1984; Moore et al., 1981; Vega et al., 2003). Meschke andSobsey (1998) conducted research with Norwalk virus, poliovirus, and MS2bacteriophages and found that viruses exhibit different adsorption charac-teristics according to soil textures. They observed that bacteriophages wereremoved to a greater degree than the bacterial pathogen Salmonella in testswith a sand filter. pH is a key factor in controlling viral adsorption ontosediments (Dowd & Pillai, 1997; Schulze-Makuch et al., 2003). Oxidationis also thought to be one of the abiotic mechanisms for microbial removalin SSF CWs (Batchelor et al., 1990; Decamp, 1996; Gray, 1989), although

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there is no clear evidence of its relative importance with respect to othermechanisms.

BIOTIC MECHANISMS

There is very little information about the biotic mechanisms responsible formicrobial removal, and available information is limited to empirical obser-vations and assumptions. In this review, we discuss three of the suggestedmechanisms: aggregation and retention in biofilms, predation, and plant-mediated effects. Microbes attach to plant root surfaces and granular mediaand subsequently become trapped within biofilms (Brown & Reed, 1994;Okabe et al., 1998; Vymazal 2001a, 2001b). Biofilms act as filters that trap orsequester suspended solids (Larsen & Greenway, 2004). Therefore, the activ-ity of microbial communities—particularly those associated with biofilms—isthought to play a role in pathogen removal in SSF CWs. Biofilms are thoughtto be responsible for most of the microbial processing that occurs in CWs(Bodelier et al., 1996; Brix, 1997; Flood & Ashbolt, 2000).

Biofilms provide a habitat for microfauna that graze on associated de-tritus and bacteria. Many authors have stated that predation (including thebacterivory by nematodes, rotifers, and protozoa) is a key mechanism in re-moving bacteria from wastewaters in CWs and that microfauna densities andpredatory activity are greater in the presence of plants (Decamp & Warren,1998; Green et al., 1997; Rivera et al., 1995). Davies and Bavor (2000) showedthat bacterial numbers tend to fall more rapidly in CWs than in ponds, andthat bacterial predation may cause this decrease.

Free-living protozoa account for a substantial proportion of grazing mi-crofauna, particularly ciliates (Curds, 1992; Puigagut et al., 2007b, 2007c),which can ingest a variety of prey, including microbial pathogens (Stott et al.,2001a, 2003b). Some authors concluded that ciliate bacterivores may in factaccount for the entire removal rates of faecal coliforms in CWs (Decamp &Warren, 1998; Green et al., 1997; Rivera et al., 1995). Many of these ciliatesare filter-feeders that graze primarily on unattached bacteria, although crawl-ing forms such as oxytrichids may also feed on surface-associated bacteria.Further examination showed that predation by free-living ciliated protozoamay also be an important factor in the removal of Cryptosporidium oocystsfrom wastewater in CWs (Stott et al., 2001b).

It is often assumed that planted SSF CWs provide greater microbialremoval than unplanted systems, mainly because plants stimulate below-ground microbial populations. Although some authors found no differencesbetween planted and unplanted systems (Vacca et al., 2005), microbial re-moval is generally higher in the presence of plants. However, the resultsvary considerably, depending on the type of plant, the season, the hydraulicretention time and the wastewater characteristics (Tanner et al., 1995).

Removal efficiency was slightly enhanced (approximately 0.5 log) formost microbial groups, including Escherichia coli and other indicators, inthe presence of vegetation (Decamp et al., 1999; Garcıa & Becares, 1997;

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Hench et al., 2003; Leonard, 2000; Ottova et al., 1997; Perkins & Hunter,2000; Rivera et al., 1995; Soto et al., 1999; Thurston et al., 2001). SSF CWswere found to reduce E. coli by 31–91% in the rhizosphere but by only0–35% in unvegetated controls (Rivera et al., 1995). Decamp and Warren(2000) observed a clear difference in removal kinetics between a SSF CWand an unplanted system: the average decimal reduction distances (DRD)were 1.55 m for the SSF CW and 6.4 m for the unplanted system (Decamp& Warren, 2000).

Enteric bacteria appear to form part of the rhizosphere communitiesin SSF CWs and therefore may not be removed to a great extent by theexudates of macrophytes or other competing bacteria in the rhizosphere(Axelrood et al., 1996; Pierson & Pierson, 1996; Smalla et al., 2001). How-ever, several authors have observed that some plants produce secondarymetabolites with antibacterial properties (Vincent et al., 1994), even thoughtheir role in wastewater treatment has yet to be proved. Plant excretion of an-tibacterial compounds is a potential removal mechanism that is mentionedfrequently in the literature, but it has not yet been clearly demonstrated.Commonly cited papers by Seidel (1976), Gopal and Goel (1993), Ottovaet al. (1997), and other authors did not prove that plants have a direct effecton bacterial inhibition, but they found that the presence of plants was asso-ciated with higher bacterial reduction. The roots of many species of aquaticmacrophytes responsible for this possible antimicrobial mechanism excretetannic and gallic acids.

The specific effect of macrophytes on virus survival in SSF CWs hasnot been assessed quantitatively. However, Gersberg et al. (1987b) observedthat the survival of several indicators of viral pollution (indigenous F-specificbacteriophages, seeded MS2 bacteriophage, and seeded human poliovirustype 1) in primary treated municipal wastewater was significantly (p < 0.01)higher in planted SSF CWs than in unplanted systems. The greater removalcapabilities of planted systems suggest that macrophytes play a role in theremoval process.

The exact influence of macrophytes on parasite removal is unclear.While the presence of macrophytes in SSF CWs may not be a significantfactor in parasite removal, planted systems generally perform better than un-planted systems (Gersberg et al., 1987a). For example, experimental plantedSSF CWs were found to be 50% more effective in removing helminth eggsthan unplanted systems (Rivera et al., 1995). However, in another study,macrophytes had no apparent effect on helminth egg removal in a full-scaleSSF CW (Stott et al., 1996).

Influence of Design Factors on Microbial Removal

A number of design factors influence the apparent microbial removal effi-ciency of SSF CWs. They include operational conditions, properties of the

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granular medium, and physical characteristics such as water depth (Perkinsand Hunter, 1999).

A strongly negative trend of bacterial removal with transient high flowrates as a result of erosive conditions has been observed in different wetlandsystems (Green et al., 1997; Perkins & Hunter, 2000). Perkins and Hunter(1999) reported a faecal coliform concentration of 5 log units/100 mL (whichis a high concentration) in the effluent of a CW under high flow rates. Inanother study, Arias et al. (2003b) observed low total coliforms removal rates(ranging from 0.7 to 1.5 orders of magnitude) at high HLRs (from 520 to 1300mm d−1) in a vertical SSF CW.

The results of several studies of faecal bacteria removal suggest thatthere is a positive correlation between bacterial inactivation and HRT (Cooperet al., 1996; Curtis, 2003; Green et al., 1997; Mashauri et al., 2000; Tanneret al., 1995; Wright et al., 1995). Nevertheless, the inordinate variation infaecal bacteria concentration during the experimentation has not allowedfinding a clear relationship between bacterial inactivation and HRT. Garcıaet al. (2003b) observed that increases in HRT over a period of three daysdid not substantially increase the microbial inactivation ratio or by extensionreduce the microbial concentration in the effluents.

The granular medium is a key factor in the microbial removal process(Polprasert & Hoang, 1983). On average, smaller granular media improvethe microbial inactivation ratio by between 1 and 2 log units for both faecalcoliforms and somatic coliphages (Garcıa et al., 2003b; Ottova et al., 1997).The medium size can also affect the structure of the biofilms that develop inthe system. Although bacterial cell counts and total protein have not beenobserved to be significantly diffferent, Larsen and Greenway (2004) havefound that exopolysaccharides (EPS) development is higher on a gravel of 5mm compared with a gravel of 20 mm per volume basis.

Information on the relationship between water depth and microbialremoval rates in SSF CWs is scarce and contradictory (Coleman et al., 2001;Cooper et al., 1996; USEPA, 2000). However, Garcıa et al. (2004b) foundthat shallow horizontal SSF CWs with a water depth of 0.27 m were moreeffective in removing total coliforms and E. coli than wetlands with a depthof 0.50 m. There was no clear explanation for this pattern.

WETLAND MODELING: PREDICTIVE TOOLS FOR CONTAMINANTREMOVAL ESTIMATION

Extensive work has been done to develop mechanistic models as predictiveand interpretive tools for improving conventional wastewater treatment tech-niques, such as activated sludge plants or anaerobic reactors. In contrast, thestate of the art in CW modeling still consists of black-box approaches, whichfail to account for the inherent complexity of these artificial ecosystems. More

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comprehensive models have only been developed in recent years, and theyare not yet based on a common framework.

The first part of this section will focus on the first-order k-C∗ modelby discussing some recent findings on its limitations, as well as two newmethodologies for dealing with the inherent parameter uncertainty. In thesecond part, we introduce recently developed mechanistic models intendedto open up the black box.

First-Order k-C∗ Model

The first-order model is based on the common observation that contaminantremoval has an exponential relationship with distance traveled or time inhorizontal SSF CWs. When a number of simplifying assumptions are met,this can be represented as follows (Kadlec & Knight, 1996):

dC

dt= −kVC

[1]−→(

Cout − C∗

Cin − C∗

)= e(−kVτ) [2],[3],[4]−−−−→

(Cout − C∗

Cin − C∗

)= e(−kA/q)

Transformation equations:

[1] Cin = C(t = 0) and Cout= C(t =τ ), initial conditions

[2] kA = kVεd

[3] q = Q/A

[4] V = Qτ= Adε

Cin and Cout are influent and effluent concentrations, and C∗ is the back-ground concentration resulting from processes such as autochthonous pro-duction and/or sediment release (mg L−1 or CFU 100 mL−1). The first-orderrate constants can be expressed on a volumetric base, kV (d−1), or areal base,kA (m d−1). The other related variables are t (time, d), τ (mean hydraulicretention time, HRT, d), Q (flow rate, m3 d−1), q (hydraulic loading rate,HLR, m d−1), d (water depth, m), A (surface area, m2), V (pore volume, m3),and ε (porosity, dimensionless).

Assumptions needed for the development of the model are as follows:

• The wetland is in a stationary state, so there are no adaptation trends.• The data used should be averages of at least three hydraulic retention

times (HRTs).• The wetland has spatially invariant averaged flow, or precipitation = evap-

otranspiration.• There is no infiltration.• The time-averaged concentrations should be equal to the flow-weighted

concentrations.

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• There are no inputs into the wetland via atmospheric deposition.• The wetland is rectangular.• The wetland exhibits plug flow conditions (i.e., no backmixing or

bypassing).• There is no variation in the cross-flow direction.

Precipitation and evapotranspiration can have a considerable impact on con-centrations (dilution of concentration effects) and the HRT. Kadlec (1997)therefore proposes the following modified model, which expresses a power-law profile between the inlet and outlet for stationary conditions:

Cout − C′

Cin − C′ = (1 + [α/q])−(1+kA/α) with C′ = C∗[

kA

kA + α

]

The parameter α in the above equation is calculated as the difference be-tween precipitation and evapotranspiration (m d−1). The reader is referredto Kadlec and Knight (1996) for the theoretical background to this equation.

The influence of temperature on the removal rate is usually modeled bythe Arrhenius equation:

kA,T = kA,20θ(T−20) and kV,T = kV,20θ

(T−20)

where ki,T represents the k value at T◦C, k20 is the k value at 20◦C, and θ isthe temperature coefficient (dimensionless).

Finally, the reader should be aware that the areal-based version of theabove model has also been used for VSSF CWs (e.g., Arias et al., 2003a), al-though only due to a lack of suitable alternatives and as a means of compar-ing performance under different conditions, and not because experimentalor field observations had proved its validity for these systems.

The mathematical descriptions described above provide a maximumof four parameters (kA or kV, C∗, θ , and α) that must account for boththe complex removal processes in SSF CWs and external influences suchas climate conditions. Consequently, the parameter values reported varyconsiderably. Rousseau et al. (2004a) published an extensive review of theseparameter values for horizontal SSF CWs, and we provide a general overviewof kA values in Table 7. Values of kA vary from 0.06 to 1.00 m d−1 for BODremoval. The temperature coefficients θ also seem to be case-dependent.Kadlec and Knight (1996) reported a θ value of 1.00 for BOD, which indicatesthat organic matter removal is independent of temperature. However, otherresearchers claim that BOD removal increases with temperature. Reed andBrown (1995) and Tanner et al. (1995) obtained a θ value of 1.06 for BODremoval, and Stein et al. (2006) even claimed that k values decrease withtemperature or that θ is less than 1.

Many researchers have demonstrated that the first-order model cannotcapture the total variability encountered in SSF CW. For example, Kadlec

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TABLE 7. First-order areal rate constants kA (m d−1) for horizontal SSF CWs according todifferent authors

Contaminant Reference kA (m d1) Remarks

BOD Kadlec and Knight(1996)

0.085–1

Kadlec (1997) 0.49 C∗ > 3 mg L−1 and θ = 1.00(20◦C)

Vymazal et al. (1998) 0.19 Proposed by KickuthBrix (1998) 0.118 ± 0.022 Mean ± 95% limits; depends on

loadSchierup et al. (1990a) 0.083 Danish systemsCooper (1990) 0.067–0.1 UK systemsBrix (1994b) 0.16 C∗ = 3.0 mg L−1; soil-basedBrix (1994b) 0.068 C∗ = 0 mg L−1; soil-basedKadlec et al. (2000) 0.133 Czech Republic wetlandsCooper et al. (1996) 0.06 C∗ = 0 mg L−1; secondary

wetlandsCooper et al. (1996) 0.31 C∗ = 0 mg L−1; tertiary wetlandsKadlec et al. (2000) 0.17 C∗ = 0 mg L−1; tertiary wetlands

USATSS Kadlec and Knight

(1996)2.74 k20 with θ = 1 and C∗ > 7 mg L−1

Kadlec (1997) 8.22 k20 with θ = 1 and C∗ > 7 mg L−1

Kadlec et al. (2000) 23.1 Laboratory columnsKadlec et al. (2000) 31.6 Large-scale pilot wetlandKadlec et al. (2000) 0.119 Data from Czech Republic

TN Kadlec and Knight(1996)

0.074 k20 with θ = 1.05 and C∗ = 1.5mg L−1

Kadlec and Knight(1996)

0.007–0.1 kT with C∗ = 1.5 mg L−1

Kadlec et al. (2000) 0.028 Czech Republic systemsTP Kadlec and Knight

(1996)0.033 k20 with θ = 1.00 and C∗ = 0.02

mg L−1

(1997, 2000) and Headley et al. (2005) showed that the rate constants arein fact not constant at all and depend on influent concentrations, HLRs andwater depth. In addition, Rousseau et al. (2004a) argued that k values seemto correlate with void fractions, wetland age, and the chosen backgroundconcentration C∗. Finally, Stein et al. (2006) recently proved that k values areeven correlated with plant species; to increase the predictive power of theirmodel, the value of k had to be made dependent not only on temperatureand plant species but also on C∗. Stein et al. (2006) therefore added weightto the argument that macrophytes play an important role in CWs.

In addition to parameter variability and uncertainty, the first-order modelpresents one other major drawback. The equations are based on the assump-tions of prevailing plug flow and stationary conditions. However, small-scalewastewater treatment plants, which include most SSF CWs, often exhibitsubstantial influent variations (Boller, 1997) and are therefore under non-steady-state conditions. Physical dispersion, short-circuiting, and dead zones

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are also common in SSF CWs and cause non-ideal plug flow conditions,which renders the first-order model unsuitable (Kadlec, 2000).

Kadlec (2003) suggested that the parameter variability described abovecould be taken into account by applying discrete, linear, or gamma distri-butions to represent the HRT and the k value distribution. Retention timedistributions are relatively easy to measure using a tracer test and a tanks-in-series model. The basic premise behind k value distributions is easilyexplained by the example of BOD. This water quality parameter representsa range of organic materials, some of which can be degraded within a matterof hours, while others require several days. Each compound has its own kvalue, so a distribution of k values is produced. Kadlec (2003) defines theoverall k value for BOD removal as the “apparent k.” In fact, a simpler ver-sion of this idea had already been proposed by Shepherd et al. (2001) torepresent the degradation of the complex COD mixture in winery wastew-ater. Kadlec (2003) recommended using the more general form called the“relaxed tanks-in-series model,” which can be expressed mathematically asfollows:

Cout

Cin= 1

(1 + kappt/P)P

where P is the apparent number of equal-volume tanks in series and kapp isthe apparent tanks-in-series rate constant.

Given the numerous drawbacks and difficulties described above, itwould seem logical to develop more sophisticated and comprehensive mod-els. However, Marsilli-Libelli and Checchi (2005) advocate the use of a simplehydraulic model consisting of tanks in series and tanks in parallel in combi-nation with first-order kinetics. They claim that “the current level of modelcomplexity is not producing a proportional insight into the factors affect-ing pollution removal vis-a-vis the difficulty of estimating a large number ofparameters” (p. 217)—hence their proposal for a simple but robust model,which produced satisfying simulation results. Nevertheless, their model is ofblack-box type and therefore provides no insight into the complex internalprocesses of wetland systems.

Mechanistic Models

McGechan et al. (2005a, 2005b) published a spatially explicit model thatcan be used to simulate oxygen transport and consumption and nitrogentransport and transformations along a horizontal SSF CW. Transport occursboth by convection and diffusion and for oxygen also by plant root oxy-gen release. Nitrogen and carbon transformations are given as first-orderequations with fixed time constants for mineralization, volatilization, den-itrification, and anaerobic carbon degradation, whereas the rate constants

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for nitrification and aerobic carbon degradation are dependent upon oxygenconcentrations. In a second version (McGechan, 2005b), all rate constantswere also made temperature-dependent. Note that growth and decay andtherefore concentrations of bacterial biomass are not considered in theseequations. The model uses a 2D approach with horizontal and vertical lay-ers, which makes it suitable for handling the simultaneous aerobic, anoxic,and anaerobic processes in different wetland locations. As the model isnot overly complex, it was initially implemented in MS ExcelTM but waslater recoded in FORTRAN to optimize the simulation speed. Following atrial-and-error parameter-fitting procedure, the model outputs were found tocorrespond reasonably well with measured effluent concentrations. Interest-ingly, the root oxygen loss parameter had to be adjusted to 5 g m−2 d−1 toobtain optimal model predictions, which demonstrates that the plants havea significant indirect role in the nitrogen and carbon cycles of the systemstudied.

Mayo and Bigambo (2005) and Mayo and Mutambe (2005) took a fur-ther step toward capturing the complexity of N-transformations in horizontalSSF CWs. Their model considers sedimentation/filtration, mineralization, ni-trification, denitrification, plant uptake and decay, and microbial uptake anddecay, and is also capable of distinguishing between the activities of sus-pended and attached bacterial biomass. However, as far as can be seenfrom the published equations, some essential processes and relations arenot accounted for. Specifically, in the case of oxygen, it is unclear whetherthe model covers processes such as diffusion from air, root oxygen release,and consumption during BOD removal and influences such as the inhibitionof denitrification by high oxygen concentrations. The model implementa-tion and simulations were performed in STELLATM. Once the model hadbeen calibrated, it produced reasonably accurate predictions of the effluentconcentrations, although the simulated curve was smoother than the curveobtained from the experimental data. This may indicate that there was aproblem with the hydraulic sub-model, which was not clearly developed inthe paper. The model also confirmed the common finding that denitrifica-tion was the major sink for nitrogen. Finally, it was shown that plants have aphysical role in SSF CWs because they provide a bacteria attachment surface:when this biofilm was omitted, the predicted treatment efficiency decreasedsubstantially.

The model developed by Wynn and Liehr (2001) consists of six inter-linked sub-models that represent the carbon and nitrogen cycles, the wa-ter and oxygen balances, and the growth, decay and metabolism of het-erotrophic and autotrophic bacteria. The latter processes are modeled astypical Monod equations with switching functions for substrate and elec-tron acceptors. The hydraulic behavior is modeled using a tanks-in-seriesapproach to simulate the mixing regime and the Darcy equation to simu-late flow in a porous medium. The active biomass is split into aerobic and

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anaerobic fractions according to the bulk oxygen concentration in the tank,which simulates the concurrent existence of both aerobic and anoxic zonesin a completely mixed tank. The model design also assumes that suspendedsolids are removed completely, based on the common observation that ef-fluent suspended solids levels in SSF CWs are very low. The influence ofprecipitation and evapotranspiration on the water balance is also implicitlyincorporated. All model equations were implemented in STELLA II (HighPerformance Systems Inc.).

Wynn and Liehr (2001) calibrated the 42 model parameters graphicallyby visually comparing measured and simulated effluent concentrations forvarious parameter values. However, this procedure yielded values for sev-eral microbial parameters that were one or more orders of magnitude lowerthan those typically mentioned in the literature. The largest adaptations weremade to the aerobic and anaerobic heterotrophic growth rates, which had tobe lowered from 6.0 to 0.015 d−1 and from 4.0 to 0.014 d−1, respectively. Theauthors suggest that this problem may be due to limitations in the biofilm. De-spite the model uncertainty and the lack of high-quality data, the calibratedmodel reproduced most of the seasonal trends of oxygen, nitrogen, andcarbon and clearly identified the interactions between the different cycles,although it failed to detect most of the short-term variability. For a furtherStrengths, Weaknesses, Opportunities, and Limitations (SWOT) analysis ofthe Wynn and Liehr (2001) model, the reader is referred to Rousseau (2005).

Rousseau et al. (2005) applied a modified version of the Wynn andLiehr (2001) model to two full-scale horizontal SSF CWs in series used forsecondary treatment. First, to reduce the model uncertainty, influent and ef-fluent TOC and DOC were measured directly so that fixed conversion factorswere not required to translate BOD into dissolved and particulate organiccarbon and vice versa. Indeed, in reality, the BOD/TOC and BOD/DOCratios are highly variable. Second, some errors and inconsistencies in thecode were corrected. Third, the hydraulic sub-model was adapted to han-dle aboveground flow in the case of extreme rain events and zero outflowsduring warm periods with excessive evapotranspiration. Finally, the nitrogensub-model was enhanced with ammonium adsorption and desorption pro-cesses. A sensitivity analysis showed that wetland dimensions were the mostsensitive parameters because of their influence on the hydraulic retentiontime and/or the hydraulic resistance. It was also shown that the low-loadedsecond-stage horizontal SSF CW contained more sensitive parameters thanthe high-loaded first-stage wetland; at lower loading rates, there is no longerone or more dominant processes, and most processes have similar rates,which renders their parameters equally important. Finally, root oxygen lossalso proved to be a highly sensitive parameter that should be researched fur-ther so that the process can be quantified adequately. Figure 7 illustrates thecapabilities of the model for predicting effluent DOC and NH4 concentrationsafter optimal parameter calibration.

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FIGURE 7. Model predictions of the adapted Wynn and Liehr (2001) model for the first-stagehorizontal SSF CW of a 47 PE secondary treatment wetland in Saxby, UK. From Rousseauet al. (2005), with permission.

The multi-component reactive transport model CW2D (Constructed Wet-lands 2-Dimensional) was developed by Langergraber (2001, 2003) to modelthe transport and reactions of the principal wastewater constituents in verticaland horizontal SSF CWs. The model was implemented into the source codeof the simulation program HYDRUS-2D (Langergraber & Simunek, 2005).Water flow through the variably saturated porous media is represented byRichard’s equation. The transport model considers dispersion, diffusion, con-vection, and several sources and sinks such as adsorption/desorption andwater uptake by plant roots. HYDRUS-2D can also incorporate the concept oftwo-region, dual-porosity transport, which divides the liquid phase into mo-bile (flowing) and immobile (stagnant) regions. Unfortunately, CW2D onlyconsiders dissolved wastewater compounds and is currently unsuitable forinvestigating clogging phenomena. Biochemical transformations in CW2Dare represented according to the activated sludge models (Henze et al.,2000), which can describe the elimination of organic matter, nitrogen, andphosphorus. The model incorporates 12 components, 9 processes and, mostimportantly, 46 parameters, excluding the parameters for the hydraulic sub-model. A sensitivity analysis with 10% relative parameter changes revealedthat the hydraulic parameters were the most influential, followed by the oxy-gen release rate, yield coefficients, and lysis rates for the bacteria. A furtherSWOT analysis of the CW2D model can be found in Rousseau (2005).

The simulation results produced by CW2D showed very good fits withdata from an indoor experimental vertical SSF CW wetland for wastewatertreatment (1m2 surface area, 40 L d−1), which Langergraber (2001, 2003)attributed partly to the fact that extensive data were available for calibratingthe system. The simulation results from a second experimental vertical SSFCW for treating heavily polluted surface water (2 m2, one downflow andone upflow chamber) also showed a good match with the measured data.However, simulation of a full-scale two-stage (vertical SSF CW + horizontal

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SSF CW) system proved difficult due to hydraulic irregularities such as short-circuiting and preferential flow, which could not be reproduced by the 2Dmodel.

CW2D is used increasingly in case studies (e.g., Henrichs et al., 2007;Langergraber, 2007), which led to continuous refinements of the model struc-ture and more comprehensive knowledge of the parameter values, therebyreducing the output uncertainty.

Rousseau (2005) developed another horizontal SSF CW model basedon the activated sludge models. It is a mechanistic and dynamic modelof C and N transformations that reflects the competition between bacteriaand plants for nutrient uptake, as well as the competition between differ-ent microbial groups for substrates and electron acceptors. The processesresponsible for biochemical transformations of C and N compounds con-sidered in this model are aerobic and anoxic reactions (mainly followingthe recommendations given in the activated sludge models; Henze et al.,2000) and anaerobic reactions (following the model proposed by Kalyuzh-nyi and Federovich, 1998) for the competition between sulfate-reducing andmethanogenic bacteria. The bacterial groups considered include aerobic het-erotrophs, fermenters, sulfate reducers, methanogens, nitrifiers, and sulfideoxidizers. Microbial reactions are represented by standard Monod equationswith switching functions. Diffusion limitations in the biofilm are omitted be-cause biofilms growing in SSF CWs are typically very thin (Shipin et al., 2005).Underground flow of the wastewater is approached via the Darcy equation,and to mimic the dispersive characteristics, a combination of tanks-in-seriesand tanks-in-parallel is proposed. The model requires 27 input variables: 22for the influent (flow, COD, ammonia, etc.) and 5 for environmental factors(i.e., day length, air temperature, water temperature, precipitation, and sea-son). A total of 110 stoichiometric, kinetic, and other parameters are requiredto describe all of the processes that occur in the wetlands.

Data from an experimental horizontal SSF CW (surface area 0.55 m2) anda pilot-scale horizontal SSF CW (surface area 55 m2) and default parametervalues were used to verify the model. If we take into account the uncer-tainties of the COD, N, and S fractionations and the low-sampling frequencywith respect to the time-step used during the simulation, the model success-fully described the general trends for different seasons and different loadingrates. The results also confirmed that the oxygen balance has a strong impacton the microbial populations in horizontal SSF CWs and on the removal oforganic matter and nitrogen.

The Rousseau (2005) model also considers particulate substances andis therefore able to evaluate clogging process by simulating the pore vol-ume reductions. Garcıa et al. (2007b) used the model to verify the resultsfrom two experimental wetlands treating settled and physicochemically pre-treated wastewater and found that they were in good agreement with theexperimental results.

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CONCLUSIONS AND RECOMMENDATIONS

Subsurface-flow constructed wetlands (SSF CWs) have been used increas-ingly for wastewater treatment in recent years. At the same time, two principalresearch techniques have greatly increased our understanding of contami-nant removal processes: monitoring removal efficiencies and accumulationin the compartments of full-scale systems; and studying removal processesand influencing factors through laboratory experiments. Advances in thisfield have enabled specialists to design more effective full-scale SSF CWs.However, a number of processes are still poorly understood, and furtherresearch and development is necessary. In summary:

• It has been observed that SSF CWs have a heterogeneous redox structureat the micro- and macro-scale levels. The prevailing redox conditions havea strong influence on processes linked to contaminant transformation andremoval. Redox changes at the micro-scale have been poorly investigated.However, microelectrodes coupled to micromanipulators are promisingtools for studying these changes. Preliminary research suggests that redoxmicrogradients caused by the presence of plants can change dramaticallybetween systems, depending on a range of properties such as the load,macrophyte species, and environmental conditions.

• For a standard urban wastewater, the primary processes involved in or-ganic matter removal in a SSF CW are physicochemical followed by bio-logical ones. Physicochemical processes allow retention of organic matter(especially particulate fraction) in the wetland, while biological pathwaystransform, convert, and finally allow the removal of organic matter. Sev-eral studies have proven that internal organic matter loadings comingfrom plant decomposition are only significant in the carbon balance insystems treating influents having a low organic matter concentration (i.e.,secondary effluents).

• Particulate organic matter accumulation over time is a typical feature ofSSF CWs and occurs by three main processes: the retention of suspendedsolids, the growth of biofilms, and the accumulation of plant and microbialdetritus. The relative contribution of these processes to organic matteraccumulation is not known, although several authors have demonstratedor implied that all three can be significant. This is an important topic thatwarrants extensive research because it is linked with clogging, which isthe most serious operational problem in SSF CWs. There has been nocomprehensive study of the physicochemical processes involved in theretention of suspended solids. In addition, little information is available onthe disintegration and hydrolysis of particulate organic matter in SSF CWs.

• Many recent studies have highlighted the importance of biochemical path-ways other than aerobic respiration (i.e., denitrification, fermentation, sul-fate reduction, and methanogenesis) in the removal of organic matter in

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horizontal SSF CWs. The three main oxygen sources in horizontal SSFCWs (inputs by the influent, physical surface reaeration and plant release)provide considerably less oxygen than is needed to remove the organicmatter contained in a typical primary effluent. In contrast, in vertical SSFCWs, aerobic respiration is the most important reaction in the removal oforganic matter. In vertical systems, intermittent loads coupled with restingperiods produce high reaeration rates of the medium.

• Early studies using mass balances to estimate plant oxygen release ratesoverpredicted the rate values and led to the general view that macrophytesplayed a major indirect role in the removal of organic matter. However,several recent laboratory studies have shown that plant oxygen releaserates are much lower than the oxygen demand in standard wastewater.This remains a subject for debate, but it is becoming more widely acceptedthat root release of oxygen does not make a significant contribution to theoxygenation of wastewater. A standard procedure should be developedto evaluate plant oxygen release.

• Recent research has shown that design properties (i.e., water depth), op-erational conditions (i.e., organic loading rate (HLR)), and environmentalfactors (i.e., temperature) affect the predominant biochemical pathwaysresponsible for organic matter removal in SSF CWs and, by extension,the overall effectiveness of these systems. It seems that the removal effi-ciency decreases as more anaerobic reactions prevail and reactions suchas methanogenesis and sulfate reduction are promoted. In fact, severalstudies reported lower organic matter and ammonium removal rates withincreasing sulfate reduction activity. Research into biochemical reactionsinvolving substrates, intermediates, products, and microbial populationsmust expand our knowledge of the processes involved in organic matterremoval in the coming years.

• Interest in methanogenesis and methane emissions from SSF CWs has in-creased in recent years due to the significant contribution that methanemakes to global warming. Methane emission rates vary considerably be-tween the different studies and even at different points within the samewetland. Standard methods for estimating methane emission rates shouldbe developed. In addition, there is no information in the literature onthe relative importance of the three processes involved in emissions—gasdiffusion, ebullition and internal plant-mediated transport—or the impor-tance of methane oxidation.

• It has been proven that the removal efficiency of specific organic contami-nants in SSF CWs is comparable to or greater than efficiencies recorded inconventional wastewater treatment plants (activated sludge). The varietyof structures and physicochemical properties of specific organic contam-inants make it very difficult to predict and model their behavior in SSFCWs.

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• Specific organic hydrophobic contaminants (log Kow > 4–5) are mainlyremoved via sorption into the organic matter and biofilm present in thegranular medium, where they are slowly degraded. In contrast, the re-moval of hydrophilic contaminants (log Kow < 4) depends on their struc-ture and ranges from 0% to 99% (depending on the contaminant) withlow accumulation into the gravel medium. Further research is needed tomodel the behavior of hydrophilic contaminants, in particular pesticidesand pharmaceutical products. The role of macrophytes (direct or indirect)in the removal of specific organic contaminants is currently unknown.

• Vertical SSF CWs have been postulated as the best option to achievehigh specific organic contaminants removal with low hydraulic retentiontime—a few hours instead of the days or weeks typically used for horizon-tal CWs. Nevertheless, the performance in terms of efficiency is believedto be better by using hybrid systems.

• Many researchers have studied nutrient removal processes in SSF CWsin the last two decades, and these processes are better understood thanthose for the removal of other contaminants. The most important pathwayfor nitrogen removal is nitrification coupled with denitrification, wherenitrification is often the limiting step. A recent isotope study showed thatnitrogen spiralling occurs in SSF CWs, which increases the retention time ofthe influent nitrogen with respect to the hydraulic retention time, producesa long delay in the treatment system response to changes in nitrogenloading, and reduces short-term fluctuations in nitrogen loading. Anotherrecent report identified ANAMMOX bacteria in SSF CWs.

• Phosphorus removal rates in SSF CWs are always lower than nitrogenremoval rates because there are no microbial removal pathways, plantuptake is relatively small, and most media have low P-sorption and P-complexation capacities. SSF CWs require additional treatment units inorder to guarantee long-term compliance with strict effluent phosphorusstandards for standard urban wastewaters. Different studies conducted inrecent years have demonstrated that higher phosphorus removal (>90%)can be achieved reliably by extending the systems with chemical pre-cipitation of phosphorus with aluminium polychloride, iron chloride, ororganic flocculants.

• Although many recent studies have examined the distribution of metalsacross different compartments of SSF CWs, we still do not fully under-stand the processes through which they are removed, due to the interac-tion of many different components. Relatively few researchers have stud-ied metal speciation in the granular media, mainly because of a lack ofspeciation methods for coarse-textured media such as gravel. Speciationstudies of metals in SSF CW media could provide very useful informationabout the dominant removal processes and should be carried out in thefuture.

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• It has been demonstrated that heavy metals accumulate in the mediumsituated closest to the inlet zone of horizontal SSF CWs, with the ex-ception of Mn and Fe, the dynamics of which are strongly affected bythe redox conditions. Anaerobic conditions prevailing in the first sectionof horizontal SSF CWs can mobilize Mn(II) and Fe(II), which will eitherbe deposited further down the wetland, when conditions change fromreducing to oxidizing, or partly washed out of the system.

• The main removal processes of heavy metals are physical and micro-bial. Metal removal by accumulation in macrophytes usually accountsfor only a very small proportion of the total removal. However, macro-phytes can indirectly affect metal removal by acidifying and/or oxidizingthe rhizosphere. Different laboratory-scale experiments have shown thatmacrophytes may increase, decrease, or have no effect on metal removal.The dominant redox conditions of the medium in these experiments of-ten appeared to determine the effects of the macrophytes. The presenceof helophytes generally decreased or had no effect on metal removal ifthe medium was under mainly aerobic conditions, whereas macrophyteshave been reported to decrease metal removal in strongly reduced me-dia. However, more research is needed in this area to substantiate theseobservations.

• The hydraulic retention time (HRT) is an important hydraulic parameterthat affects metal removal processes, and higher HRTs generally lead togreater metal removal, particularly when metal sulfide precipitation is thedominant removal process.

• Although the literature provides extensive information on the removalefficiencies of faecal bacteria indicators in SSF CWs, complete data areunavailable for the removal effectiveness of specific bacterial pathogens,viruses, protozoa, and helminths. In addition, most studies used colony-forming unit methods to evaluate the removal of faecal bacteria indicators,which do not necessarily provide actual numbers of faecal bacteria. In fact,recent research comparing standard cultivation procedures and molecu-lar techniques has shown that plate counts overestimate the efficiencyof microbial removal in wetlands. Therefore, more studies based on thecomparison of the effectiveness considering molecular as well as standardtechniques are needed.

• Although many authors have discussed the possible processes involved inmicrobial removal in SSF CWs, none have conducted systematic analysesof these processes and the fate of microbes. Sedimentation and filtrationseem to be the most important abiotic mechanisms in the removal ofmany microbes and parasites. Specifically, adsorption is thought to playan important role in virus removal.

• Very little information is available on biotic mechanisms of microbial andparasite removal. Some empirical observations and assumptions have been

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made, but further research is clearly required. The most commonly sug-gested mechanisms are aggregation and retention in biofilms, predation,and plant-mediated effects. It has also been observed that ciliated proto-zoa prey on bacteria and constitute an important factor in the removal ofCryptosporidium oocysts. Comparisons between planted and unplantedsystems generally show more extensive microbial removal in the pres-ence of macrophytes. Plants are thought to stimulate biofilm growth andmicrofauna populations, as well as excrete antibacterial compounds. Thelatter process is frequently cited in the literature but has yet to be provenconclusively.

• The results of several studies suggest that there is a positive trend be-tween bacterial inactivation and HRT, and that media with smaller grainsize yield higher removal rates. However, the large variation in the influentmicrobial concentration makes it very difficult to determine these relation-ships, which only appear clearly under strictly controlled experimentalconditions.

• Modeling is a powerful technique for understanding contaminant removalprocesses in SSF CWs. The state of the art in wetland modeling still consistsof black-box approaches, which do not capture the inherent complexity ofwetland systems. However, several mechanistic models have been devel-oped in recent years that should help to define the processes and multipleinteractions that occur in wetlands.

• The most widely used black-box approach is the first-order model, whichis particularly suitable for sizing the systems. Many researchers have shownthat this model is not capable of capturing the full range of contaminantvariability. In fact, recent studies have demonstrated that first-order rateconstants are not “constant” at all and depend on several factors, includinginfluent concentrations, hydraulic loading rates, water depth, and plantspecies. A way to take into account this parameter variability consists inapplying discrete, linear, or gamma distributions to the first-order constantrates. Also, the use of more accurate hydraulic models consisting of tanks-in-series and tanks-in-parallel, in combination with first-order kinetics,allows one to account for the variability.

• Mechanistic models are at this moment useful tools to gain understandingof certain processes and are capable of demonstrating several interactionswithin the wetland ecosystem. However, none of the models developedto date successfully captures all of the associated hydraulic and biogeo-chemical complexities. The large number of parameters and a lack of in-depth knowledge currently limit the value of mechanistic wetland modelsas design tools. When considering the evolution of the Activated SludgeModels, it would be recommendable to agree on a common model frame-work, which could then be applied in numerous case studies to reduceboth model and parameter uncertainty.

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ACKNOWLEDGMENTS

This paper constitutes a result of the research projects and activities relatedto constructed wetlands that the authors have been conducted in the lastseveral years. We very much appreciate the cooperation of undergraduatestudents, technicians, and collegagues involved in these projects. Also, thesupport of public institutions is greatly acknoledged. Joan Garcıa thanks inparticular the cooperation of Eduardo Alvarez and Elif Bozdogan. Els Lesageacknowledges financial support from the Special Research Fund (BOF) ofGhent University. Joan Garcıa, Jordi Morato, Victor Matamoros, and JosepM. Bayona thank the Spanish Department of Education and Science for thegrants awarded in order to develop the research projects NEWWET andNEWWET2008 (CTM2005-06457-C05 and CTM2008-06676-C05).

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