Unit Process Wetlands for Removal of Trace Organic Contaminants and Pathogens from Municipal Wastewater Effluents

Trace organic contaminants

Municipal wastewater effluent typically contains relatively low levels of organic matter (i.e., most wastewater treatment plants achieve BOD<10 mg/L). In addition to the biopolymers and residual organic waste that make up the bulk of the biodegradable organic matter, wastewater effluents also contain an assortment of trace organic contaminants, such as pharmaceuticals and personal care products (Kolpin et al., 2002; Ternes et al., 2004b). Trace organic contaminants in wastewater effluents are an issue of concern due to their potential to cause adverse impacts to aquatic organisms at low concentrations (Daughton and Ternes, 1999; Suárez et al., 2008) as well as their potential to contaminate downstream drinking water supplies (Snyder et al., 2003).

The ability of constructed wetlands to remove trace organic contaminants from wastewater effluents has recently received growing attention (Matamoros and Bayona, 2008). Removal efficiencies for some pharmaceuticals and personal care products in treatment wetlands (Fig. 1) suggest that trace organic contaminants generally fall into one of three groups of removal efficiency. The first group of compounds is removed efficiently (i.e., >60% removal) regardless of wetland design and includes substances such as caffeine and naproxen. The second group, which includes the majority of the compounds in Fig. 1, exhibits partial removal with varying efficiencies depending on wetland design and hydraulic residence times. The final group of compounds, which includes carbamazepine and clofibric acid, are more recalcitrant and exhibit limited removal (i.e., typically <40% removal) irrespective of wetland design. Note that some values were determined from studies in subsurface wetlands, but they are included here because data from surface flow wetlands have not been reported. Optimization of treatment wetlands has the highest potential for enhancing the removal of the compounds in the second group.

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FIG. 1.

Averages with standard deviations of pharmaceutical and personal care product removal efficiencies in treatment wetlands reported in recent studies (n=3–16). References: ^aGray and Sedlak, 2005; ^bSong et al., 2011; ^cBreitholtz et al., 2012; ^dPark et al., 2009; ^eCamacho-Muñoz et al., 2011; ^fHijosa-Valsero et al., 2011; ^gMatamoros and Bayona, 2006; ^hMatamoros et al., 2007; ⁱMatamoros et al., 2009; ^jLlorens et al., 2009; ^kMatamoros et al., 2005; ^lMatamoros et al., 2008; ^mWaltman et al., 2006.

Waterborne pathogens

Wastewater effluents contain potentially infectious microorganisms, including viruses, bacteria, protozoan (oo)cysts, and helminth eggs. Removal or inactivation of pathogens is, therefore, necessary before treated effluents are discharged or reused. Treatment wetlands that receive wastewater which has already been disinfected may provide additional treatment of pathogens that are resistant to disinfection (e.g., Cryptosporidium oocysts for chlorine or adenovirus for UV). In this case, a wetland may be used to reduce the chemical disinfection requirements and to provide an additional treatment barrier. Alternatively, treatment wetlands that receive wastewater effluent which has not been disinfected can play a primary role in pathogen attenuation. In this context, Gersberg et al. (1989) suggested that treatment wetlands with hydraulic residence times of 3 to 6 days may be as effective as conventional water treatment systems employing disinfection for the removal of pathogenic bacteria and viruses. Reliance on polishing wetlands for disinfection has the advantage over chlorination of avoiding the production of disinfection byproducts (Buth et al., 2009, 2010).

Most studies on the removal of pathogens in treatment wetlands have measured fecal indicator bacteria rather than actual pathogens. The reported removal efficiency of fecal coliforms by surface wetlands is around 1-log removal (Vymazal, 2005; Kadlec and Wallace, 2009). The few studies that have been conducted with actual pathogens (Fig. 2) show removal efficiencies up to 2-log, with average values around 1-log. The dominant removal mechanisms vary dramatically among pathogen groups. A better understanding of pathogen removal mechanisms, including attachment and sedimentation, predation, and photoinactivation, and their effectiveness for different pathogen groups is needed to improve the ability to design unit process wetlands for disinfection.

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FIG. 2.

Averages with standard deviations of pathogen removal efficiencies in surface flow wetlands receiving nondisinfected influent. References: ^aMandi et al., 1996; ^bMandi et al., 1998; ^cReinoso et al., 2008; ^dFalabi et al., 2002; ^eGerba et al., 1999; ^fQuiñónez-Díaz et al., 2001; ^gHerskowitz, 1986; ^hHill and Sobsey, 2001; ⁱSong et al., 2010; ^jKarpiscak et al., 2001; ^kKadlec and Wallace, 2009.

Sorption

Sorption of trace organic contaminants encompasses two distinct processes. Adsorption involves the interaction of a compound with a surface, typically via ion exchange or surface complexation, while absorption entails partitioning into a particle-associated organic phase. Sediments and biofilms in treatment wetlands provide numerous surfaces that may be capable of sorbing trace organic contaminants. If the contaminants exhibit a high affinity for a surface, they will eventually be buried as decaying plant litter accumulates in the wetland. If the sorbent does not degrade, the contaminants will remain in the litter layer until it is removed as a part of wetland maintenance activities. If the contaminants are weakly associated with the sorbent, or if the sorbent (e.g., plant litter) degrades, the process may simply slow the movement of contaminants through a wetland, providing more time for other transformation processes to occur.

Although sorption is assumed to be important to trace organic contaminant fate in treatment wetlands (Imfeld et al., 2009), a few studies have investigated it specifically. In these studies, log K_ow, a measure of a contaminant's hydrophobicity, generally predicts which contaminants are most susceptible to absorption. For example, in a bulrush-dominated surface flow wetland with a 30 cm deep gravel bed, the hydrophobic phthalate esters and fragrance molecules (log K_ow>4) were absorbed to organic matter in the gravel bed as well as suspended particles (Reyes-Contreras et al., 2011). Absorption was also found to slow the movement of two steroid hormones (log K_ow≈4) relative to a conservative tracer in a densely vegetated surface flow wetland (Gray and Sedlak, 2005).

Removal of chemical contaminants by absorption may be more important in subsurface flow wetlands, where flowing water encounters higher densities of particulate organic matter. For example, partial removal of the recalcitrant contaminant carbamazepine (log K_ow≈2.5) by absorption and negligible sorptive removal of less hydrophobic contaminants was observed in a study of a subsurface flow treatment wetland (Matamoros et al., 2005).

Studies of pesticide sorption in agricultural wetlands show a similar dependence on contaminant hydrophobicity (Kruger et al., 1996; Moore et al., 2002; Reichenberger et al., 2007). For example, sorption of the herbicide atrazine (log K_ow=2.75) to soil, litter, peat, and sediments from three Midwest wetlands was well described for all sorbents by an organic carbon-normalized distribution coefficient (K_OC=760 L/kg OC) (Alvord and Kadlec, 1995). Given this distribution coefficient, a wetland with about 2 kg/m² litter, containing about 40% organic carbon (Alvord and Kadlec, 1995), and a depth of 40 cm, would be capable of absorbing more than 60% of atrazine from the aqueous phase. This suggests that absorption in wetlands could be significant for compounds with a log K_ow greater than about 2.5, provided that the system is designed properly. For comparison, absorption of trace organic contaminants in activated sludge treatment plants is usually unimportant for compounds with log K_ow values less than about 4 (Ternes et al., 2004a; Wick et al., 2009).

Less hydrophobic compounds have been found to adsorb via specific ionic interactions with activated sludge in treatment plants (Stuer-Lauridsen et al., 2000; Golet et al., 2003) and soils (Tolls, 2001). In wetlands, the relatively hydrophilic fluorescent dye, rhodamine WT, sorbs significantly to plants and sediments (Lin et al., 2003; Keefe et al., 2004). At neutral pH values, rhodamine WT contains both positively and negatively charged functional groups and is, thus, likely to be adsorbed via specific interactions with charged functional groups on the sorbents (Kasanavia et al., 1999). Therefore, adsorption of ionic trace organic contaminants in treatment wetlands may be an important loss mechanism for certain compounds. However, additional research is needed to assess the overall importance of this phenomenon and ways in which it could be enhanced through wetland design.

Certain types of wetland vegetation may increase the removal of trace organic contaminants by sorption. For example, it has been suggested that wetlands dominated by bulrush (e.g., Scirpus spp.) are conducive to sorption of trace organic contaminants due to the large amounts of spongy peat formed by decomposing plants (Horne and Fleming-Singer, 2005). Duckweed (Lemna spp.), a floating macrophyte often present in open waters in treatment wetlands, sorbs trace organic contaminants such as halogenated phenols (Tront et al., 2007) and pharmaceuticals and personal care products, including fluoxetine, ibuprofen, and triclosan (Reinhold et al., 2010). However, duckweed grows in a thin layer near the water surface and the relatively small mass of the plant in wetlands likely precludes it from removing a significant fraction of the trace organic contaminants as water passes through a wetland.

Water chemistry also affects sorption of chemical contaminants in constructed wetlands (Hussain and Prasher, 2011). In particular, the pH of wetland water will affect the sorption of contaminants by changing their speciation. This phenomenon has been observed in wastewater treatment plant sludge for the acidic pharmaceuticals diclofenac (pK_a=4.6) and ibuprofen (pK_a=3.5), which absorbed to primary sludge to a greater extent than to secondary sludge, because a greater fraction of the pharmaceuticals were in their uncharged form at the lower pH value (pH=6.6 in primary versus pH=7.5 in secondary) (Ternes et al., 2004a). Basic contaminants, such as those containing amine functional groups (e.g., the β-blockers), have pK_a values near 9 and are positively charged at neutral pH values. Consequently, their sorption is likely controlled by specific interactions, as was observed in a study by Yamamoto et al. (2009). Increasing the pH of wetland water could increase the fraction of the uncharged forms of the compounds, resulting in enhanced sorption by hydrophobic interactions. Further research is needed to determine the potential for enhancing sorption in treatment wetlands through the use of natural processes that alter pH values (i.e., photosynthesis and microbial respiration).

Biotransformation

Microorganisms play a prominent role in the attenuation of trace organic contaminants in constructed wetlands (Matamoros et al., 2008; Hijosa-Valsero et al., 2010) due to the diversity of microorganisms and enzymatic activities present (D'Angelo, 2003). In surface flow wetlands, biofilms found on roots, stalks, and detritus are more important to biotransformation than planktonic microorganisms (Gagnon et al., 2007; Truu et al., 2009). Thus, it is not surprising that properties affecting biofilm growth, such as the attachment matrix, hydraulic conditions, and composition of the wastewater effluents, can strongly influence microbial ecology and contaminant transformation rates in these systems (Truu et al., 2009).

In vegetated treatment wetlands, the density and type of plants affect microbial community dynamics (Ibekwe et al., 2006; Calheiros et al., 2009) by providing labile forms of organic carbon, surfaces for biofilm growth, and oxygen gradients (Reddy and D'Angelo, 1997). The ability of decaying plants to create anoxic zones in surface and subsurface flow constructed wetlands is important for the transformation of trace organic contaminants, because some compounds are more readily transformed under aerobic conditions (e.g., ibuprofen) while others (e.g., tonalide and galaxolide) are more readily transformed under anaerobic conditions (Hijosa-Valsero et al., 2010). Further, anoxic, nitrogen-reducing surface flow wetlands have been shown to be capable of transforming certain trace organic contaminates, including atenolol, naproxen, and triclosan, possibly through amide hydrolysis and reductive dehalogenation (Park et al., 2009).

Plant biomass and DOC from plant litter, and to a lesser degree, residual organic carbon from wastewater effluent, provide an important energy source and create selective pressure for microbial community structure and function in wetland systems (Shackle et al., 2000; Gutknecht et al., 2006). Biotransformation of pharmaceuticals in wastewater effluents can be affected by both the abundance and source of organic carbon derived from decaying aquatic plants. For example, gemfibrozil and sulfamethoxazole, two compounds that are poorly removed in wastewater treatment plants, showed better removal in the presence of labile organic carbon derived from wetland plants than in the presence of the labile dissolved organic carbon in wastewater effluents (Lim et al., 2008). Other macrophyte characteristics, such as surface area and litter properties, can also affect microbial density (Bastviken et al., 2005), which may correlate with trace organic contaminant removal rates.

The rhizosphere associated with wetland plants hosts a unique community of microorganisms within aerobic microzones, due to the release of oxygen and nutrient-rich exudates (Brix, 1997; Kyambadde et al., 2004; Münch et al., 2007; Gagnon et al., 2007). Microbially mediated iron and manganese oxides formed in this region also have the potential for indirect oxidation or enhanced sorption of trace organic contaminants (Mendelssohn et al., 1995; Emerson et al., 1999). Rhizosphere-associated transformation appears to be an important process for trace organic contaminants in subsurface flow wetlands (Zhang et al., 2012). However, the rhizosphere may be less important in surface flow wetlands due to limited contact between the rhizosphere and flowing waters.

Bacteria and fungi transform macromolecules, such as cellulose and lignin, into lower-molecular-weight compounds through the excretion of extracellular enzymes. Enzyme expression studies in constructed wetlands have provided insights into the importance of extracellular enzyme activity to the processing of these recalcitrant forms of organic carbon (Shackle et al., 2000; Wright and Reddy, 2001; Francoeur et al., 2006; Hill et al., 2006; Rier et al., 2007). Many of the extracellular enzymes used to transform cellulose and lignin (e.g., laccases, phenol oxidases, and peroxidases) can also transform recalcitrant trace organic contaminants (Gianfreda and Rao, 2004; Lu et al., 2009). By modifying the quantity and type of carbon sources in constructed wetlands, it may be possible to increase the activity of extracellular enzymes. Since most extracellular enzymes utilize oxygen or hydrogen peroxide as terminal electron acceptors, their activity is expected to be higher in aerobic environments (Sinsabaugh, 2010; Porter, 2011).

Photolysis

In natural waters, photolysis of trace organic contaminants can occur via direct and indirect mechanisms (Schwarzenbach et al., 2003). Direct photolysis occurs when a contaminant absorbs sunlight and undergoes a transformation reaction. Indirect photolysis involves other chemicals, most often NO₃⁻ and colored dissolved organic matter (CDOM), which absorb light and produce reactive intermediates that subsequently react with contaminants. With regard to the transformation of organic contaminants, important reactive intermediates include hydroxyl radical (•OH) (Zepp et al., 1987; Brezonik and Fulkerson-Brekken, 1998), singlet oxygen (¹O₂) (Zepp et al., 1977), excited triplet state DOM (³DOM*) (Canonica et al., 1995; Boreen et al., 2004), organoperoxy radicals, and carbonate radical (•CO₃⁻) (Lam et al., 2003; Canonica et al., 2005).

Treatment wetlands may be conducive to indirect photolysis due to the presence of NO₃⁻ in nitrified wastewater effluent, which produces •OH via the reaction: \documentclass{aastex}\usepackage{amsbsy}\usepackage{amsfonts}\usepackage{amssymb}\usepackage{bm}\usepackage{mathrsfs}\usepackage{pifont}\usepackage{stmaryrd}\usepackage{textcomp}\usepackage{portland, xspace}\usepackage{amsmath, amsxtra}\pagestyle{empty}\DeclareMathSizes{10}{9}{7}{6}\begin{document} \begin{align*} {\rm NO}_3^{-}+{\rm H}_2{\rm O}+{\rm h}\nu \rightarrow {\rm NO}_2+ {\bullet} {\rm OH}+{\rm OH}^- \end{align*} \end{document}

CDOM derived from wastewater effluents and decaying plants may also promote indirect photolysis through a variety of mechanisms. For example, CDOM was significant to the removal of the pesticides alachlor and carbaryl in wetland waters via indirect photolysis (Miller and Chin, 2002, 2005). For those reactions where CDOM served as a source of •OH, rates of photolysis were not strongly affected by DOM concentrations because they acted as both a source of CDOM and a •OH scavenger. In addition to generating •OH, CDOM can also serve as a source of the selective oxidants ¹O₂ and ³DOM*, which have been found to be important to the indirect photolysis of certain trace organic chemicals in the environment (Gerecke et al., 2001; Latch et al., 2003).

Most constructed wetlands are not designed to include shallow open water zones. As a result, few investigators have studied the role of photolysis in trace organic contaminant removal in treatment wetlands. Matamoros et al. (2008) attributed the nearly complete removal of ketoprofen in an engineered treatment wetland with deep (1.5 m) open water zones and a long hydraulic residence time (30 days) to photolysis. While this study demonstrated that photolysis in deep, open ponds can significantly attenuate organic compounds that are particularly susceptible to direct photolysis [ketoprofen has a half-life of 2.5 min under near-surface summer-noon conditions (Lin and Reinhard, 2005)], removal efficiencies would be significantly lower for compounds with longer direct photolysis half-lives. For example, consider sulfamethoxazole, a compound that is relatively susceptible to direct photolysis [half-life of about 2 h under near-surface summer-noon conditions (Lam and Mabury, 2005)], in a 1 m deep wetland with the EPA (2000a) recommended hydraulic residence time of 3 days for open water zones. Given a typical beam attenuation coefficient (α) of 6/m at 330 nm, and ideal plug flow conditions, sulfamethoxazole would exhibit a decrease in concentration of only about 10% due to direct photolysis under daily averaged mid-summer conditions at 40° latitude (Schwarzenbach et al., 2003).

Attachment and sedimentation

Some pathogens, such as helminth eggs, have settling velocities that are high enough to be removed by sedimentation in treatment wetlands (e.g., ∼0.1 mm/s) (Sengupta et al., 2011). The presence of rhizomes and macrophyte stalks can further enhance sedimentation (Mandi et al., 1996). However, protozoan (oo)cysts and bacteria have much lower settling velocities [e.g., <0.001 mm/s for Giardia cysts (Dai and Boll, 2006)], and viruses are stable in suspension. These organisms will only be removed by sedimentation if they are attached to larger particles. As a result, the removal of cysts such as Giardia in some cases has been correlated with particle removal (Quiñónez-Díaz et al., 2001). However, particle association is not always conducive to pathogen removal. For example, in a study conducted by Boutilier et al. (2009), the association of Escherichia coli with particles appeared to decrease their removal compared with free-floating bacteria. This was because particles with diameters less than 80 μm had settling velocities that were too low to be removed by sedimentation.

In addition to attaching to settling particles, pathogens can be removed by attachment to other surfaces in treatment wetlands. However, the contribution of this mechanism to overall removal is not known. For example, in vegetated surface flow wetlands, viruses were removed by attachment to the biofilm layer of rhizomes and submerged stalks of emergent plants (Gersberg et al., 1987). Biofilms have also been shown to increase the removal of pathogen surrogates (0.1-, 1.0-, and 4.5-μm latex microspheres) in microcosm experiments using surface flow wetland water (Stott and Tanner, 2005). A better understanding of the processes that affect the association of pathogens with particles and surfaces, as well as the size distribution of particles in treatment wetlands, may provide insights into means for enhancing pathogen removal via sedimentation and attachment.

Pathogenic helminth eggs, protozoan (oo)cysts, and viruses that accumulate on wetland surfaces and in sediments will become inactivated over time. In contrast, indicator and pathogenic bacteria have the potential to grow under certain environmental conditions (Ksoll et al., 2007; Ishii et al., 2010), although there is no evidence of growth occurring in treatment wetlands. It should be noted that sloughing or resuspension of sediments has the potential to remobilize viable organisms. In addition, wetland vegetation or sediments that are removed during maintenance activities may contain viable pathogens, which could complicate efforts to reuse the material (e.g., as a soil amendment).

Predation

Predation has the potential to be an important removal mechanism for bacteria and protozoan (oo)cysts in treatment wetlands (Stott et al., 2001; Song et al., 2008), but much less is known about viruses. Grazing of particles in the size range of individual viruses is typically inefficient (Hahn and Höfle, 2001), although viruses attached to larger particles have the potential to be removed by predation. Potential predators of pathogens in treatment wetlands include nematodes, copepods, rotifers, and protozoa (Decamp and Warren, 1998). For example, grazing rates for the ciliated protozoa Paramecium were measured to be 111 E. coli/(ciliate-hour) and 170 Cryptosporidium parvum oocysts/(ciliate-hour) in lab experiments (Decamp and Warren, 1998; Stott et al., 2001). Hence, a population of 2×10⁴Paramecium per liter (Decamp and Warren, 1998) would have the potential to achieve greater than 6-log removal of E. coli or oocysts in 1 h, if these were the only particles in the water. Grazing rates in an actual wetland are expected to be significantly lower, because pathogens would represent a small fraction of the total particles.

In addition to ciliated protozoa, rotifers may play an important role in the removal of pathogens via grazing due to their dominance in the total zooplankton population in shallow wetlands (Beaver et al., 1998; Fayer et al., 2000; Trout et al., 2002; Proakis, 2003). For example, the rotifer Brachionusplicatilist rapidly removed E. coli from water under laboratory conditions, with an average feeding rate of almost 700 E. coli/(rotifer-hour) (Proakis, 2003). Rotifers from six different genera were also found to ingest Giardia cysts and Cryptosporidium oocysts in simple laboratory experiments (Fayer et al., 2000; Trout et al., 2002). However, ingestion rates in actual wetlands have not been measured, and conditions that promote predation in treatment wetlands are not sufficiently understood.

While predation may provide an effective means for removing indicator organisms and pathogens from water, the viability of the organisms after ingestion is uncertain. In general, protozoan grazing is a major mechanism of bacterial population control (Hahn and Höfle, 2001). However, there is evidence that some pathogenic bacteria are not inactivated as a result of ingestion by protozoan grazers, such as ciliates and amoebas (Barker and Brown, 1994; Meltz Steinberg and Levin, 2007). Indeed, pathogenic bacteria such as Legionella actually colonize free-living amoeba to protect themselves from unfavorable environmental conditions (Thomas et al., 2010). More research is needed on inactivation of the wide range of pathogens of concern in wastewater by the different types of grazers in treatment wetlands.

Photoinactivation

Sunlight-mediated inactivation is one of the most important disinfection mechanisms in waste stabilization ponds (Davies-Colley et al., 2000; Davies-Colley, 2005), suggesting that open water zones are promising for the removal of pathogens in treatment wetlands. There are three main mechanisms of sunlight-mediated inactivation of microorganisms (Davies-Colley et al., 1999). Analogous to direct photolysis of chemical contaminants, the absorption of UVB radiation (280–320 nm) by DNA causes direct damage to cellular DNA, primarily by pyrimidine dimer formation (Jagger, 1985). Similar to indirect photolysis of chemical contaminants, there are also indirect disinfection mechanisms, in which sensitizers absorb light and produce reactive species that damage organisms. Indirect damage may occur due to absorption of sunlight by cell constituents (endogenous sensitizers) (Davies-Colley et al., 1999; Bosshard et al., 2010) or by sensitizers in water (exogenous sensitizers), such as CDOM. ¹O₂ has been shown to be the most important reactive species produced by exogenous sensitizers during the sunlight-mediated inactivation of MS2 coliphage, which is a model for human enteric viruses (Kohn and Nelson, 2007).

Exogenous inactivation initiated by sensitizers that absorb longer wavelengths, such as CDOM, may be more important than endogenous mechanisms, because the UV wavelengths which contribute to endogenous inactivation (direct and indirect) are readily absorbed in wetland water (i.e., λ=280 to 400 nm). However, not all organisms appear to be susceptible to this mechanism (Davies-Colley et al., 1999). Photoinactivation can also be enhanced by high dissolved oxygen and elevated pH (e.g., greater than pH 9) that result from algal photosynthesis (Curtis et al., 1992; Davies-Colley et al., 1999; Ansa et al., 2011).

The importance of sunlight-mediated disinfection in surface flow treatment wetlands was illustrated when E. coli concentrations declined significantly after a thick bed of floating duckweed (Lemna spp.) had been removed from a newly constructed surface flow wetland (MacIntyre et al., 2006). A better understanding of the sunlight-mediated inactivation mechanisms and their roles in pathogen removal is needed to optimize the design and to enable prediction of the fate of pathogens in surface flow treatment wetlands.

Shallow, open-water cells

Shallow, open-water wetland cells (Fig. 4) represent a new approach for integrating photochemical processes and aerobic microbes into a unit process wetland. While photolysis can provide a means of removing contaminants, it is difficult to design treatment wetlands for photolysis, because emergent macrophytes and floating plants (e.g., duckweed) shade the water. Furthermore, chromophores in wastewater effluents and wetland water strongly absorb sunlight, especially in the important UV region of the solar spectrum, greatly slowing photochemical reactions at depths of more than about 0.5 m. To circumvent these problems, a shallow, open-water wetland cell can be used. In these cells, concrete or geotextile liners are used on the bottom of the cell to prevent emergent macrophyte growth. Under these conditions, the bottom of the basin is rapidly colonized by a periphyton mat, which can be defined as a consortium of organisms dominated by algae, aerobic bacteria, and other eukaryotes (Wetzel, 1983).

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FIG. 4.

Pilot-scale open-water wetland cell located in Discovery Bay, CA. The cell is about 20 cm deep, 400 m², and has a hydraulic residence time of about 3 days.

A pilot-scale wetland of this design (Fig. 4) allows sunlight to penetrate throughout the water column, which is typically around 20 cm. The growth of periphyton on the bottom of the cell, rather than suspended in the water column as is the case with high rate algal ponds, prevents shading of the water by algae, and the use of relatively high water velocities prevents the accumulation of duckweed on the surface. Inexpensive materials (i.e., wooden boards) are used as baffles to minimize hydraulic short-circuiting.

Photolysis in a shallow, open-water cell with a hydraulic residence time of at least 24 h would likely result in removal of chemical contaminants that are susceptible to direct photolysis, such as NDMA, ketoprofen, and diclofenac. The shallow cell might also be conducive to the removal of certain chemical contaminants by indirect photolysis. Given NO₃⁻ concentrations of 20 mg N/L and DOC concentrations of 10 mg/L, which are typical of secondary wastewater effluents, a daily averaged •OH steady-state concentration of about 4×10⁻¹⁶ M would be expected in 20 cm of water during mid-summer at 40° latitude (Zepp et al., 1987; Schwarzenbach et al., 2003). For a typical organic contaminant, which reacts with •OH at near diffusion-controlled rates (i.e., about 7×10⁹/[M·s]), ∼50% removal would be expected after 3 days in the wetland. Both direct and indirect photolysis were found to contribute to the removal of a suite of trace organic compounds in the Discovery Bay pilot system described above (Jasper and Sedlak, 2013).

The water in periphyton-containing wetland cells exhibits diurnal cycles in which dissolved oxygen concentrations and pH values increase during the day due to photosynthesis (Fletcher and Marshall, 1982; Pollard, 2010). In the pilot-scale wetland shown in Fig. 4, the pH typically increases to values of between 9 and 10 within 50 m of the inlet. The alkaline pH conditions in the shallow, open-water cell may affect the rate of direct photolysis of chemical contaminants by changing contaminant speciation (Boreen et al., 2004). Indirect photolysis rates will also be affected by changes in water pH, because •OH is scavenged by inorganic carbon under alkaline conditions forming •CO₃⁻, which may then react with contaminants (Lam et al., 2003). This shift in radical formation could lead to selective oxidation of sulfur-containing compounds, which often exhibit elevated reaction rates with •CO₃⁻(Huang and Mabury, 2000). High concentrations of dissolved oxygen could alter other indirect photolysis pathways, either by quenching intermediate triplet states (Ryan et al., 2011) or by enhancing production of ¹O₂ (Latch et al., 2003).

Pathogens would also be inactivated in a shallow wetland cell, by both direct and indirect mechanisms. Indirect mechanisms would be especially important due to the high ¹O₂ concentrations produced by DOM present in wastewater. For example, a daily averaged steady-state ¹O₂ concentration of about 2×10⁻¹⁴ M would be expected during mid-summer at 40° latitude in 20 cm of water (Haag and Hoigné, 1986; Schwarzenbach et al., 2003). Over a 3 day residence time, this would result in almost 3-log inactivation of MS2 coliphage, which was reported to be inactivated by ¹O₂ with a second-order rate of 1.3×10⁹/[M·s] (Kohn and Nelson, 2007).

In addition to modifying overlying water chemistry, a periphyton mat could remove chemical contaminants and pathogens through sorption, biotransformation, and predation. For example, researchers have found that periphyton mats present in streams are capable of sorbing, and, in some cases, of transforming trace organic contaminants such as steroid hormones, alkylphenols, nonsteroidal anti-inflammatory drugs, and the cyanotoxin microcystin-RR (Wu et al., 2010; Writer et al., 2011; Dobor et al., 2012). Biotransformation of chemicals may be encouraged by the aerobic conditions encountered at the top of the periphyton mat. In addition, the labile carbon provided by periphyton has been shown to enhance denitrification rates in anoxic wetlands (Sirivedhin and Gray, 2006), and the increased activity of extracellular enzymes, such as the phenol oxidases, which are associated with photosynthesis (Romani et al., 2003; Francoeur et al., 2006; Rier et al., 2007), might also be important. Pathogens have been found to attach to periphyton as well, although detachment at a later time is possible (Ksoll et al., 2007).

Maintenance activities in a shallow, open-water wetland cell may include removing floating vegetation as well as detritus that accumulates in the periphyton mat. Floating vegetation, such as duckweed, is capable of quickly covering a wetland, limiting the effectiveness of photolysis and potentially altering the microbial community in the periphyton mat. The growth of floating vegetation can be limited by ensuring that hydraulic residence times are less than about 3 days (EPA, 2000a) and that the outlet structure allows floating vegetation to leave with the outflow. If floating vegetation grows, the wetland can be periodically flushed by increasing the flow rate into the wetland to wash the floating vegetation out. The slow buildup of particulate matter and detritus from decomposing periphyton will also need to be removed regularly, as over a few years, it may slowly fill in the wetland. This can be accomplished by draining the wetland and removing the dried periphyton mat with a bulldozer. After the old mat is removed, a new periphyton mat will re-grow within weeks on the wetland bottom.

Macrophyte-dominated wetland cells

To enhance contaminant attenuation, vegetated wetland zones can be designed and managed to select for microorganism communities with specific and complimentary composition and functionality. Enhanced attenuation can be achieved through a unit process approach in which different cells are optimized for specific purposes. For example, cells containing cattails provide biomass that is more readily decomposed, while bulrush cells would be expected to exhibit higher activities of extracellular enzymes that are needed to break down the lignin-rich plants (Horne and Fleming-Singer, 2005).

The linking of microbial community dynamics with factors such as plant substrates, temperature, nutrient loading, and dissolved oxygen will allow for wetlands to be actively managed based on environmental indicators. Nutrient availability can also affect microbial community structure and enzyme expression. For example, an abundance of phosphorous can reduce microbial diversity (Ahn et al., 2007) and potentially metabolic diversity. To increase microbial activity or select for reducing conditions, plant biomass harvested during routine maintenance activities can be added back to specific wetland cells. Vegetated wetlands can also be designed to include deep zones to limit plant growth. Other strategies have been suggested, such as raising the water level after plants have senesced to provide additional carbon from the previously un-submerged plant detritus in the winter (Thullen et al., 2005).

Through an increased understanding of wetland microbial communities and the ability to monitor their composition and activity, it may be possible to optimize the performance of unit process cells. Past studies have relied on culture-dependent methods to discern metabolic potential with inherent and often artificial selective pressures (Fortin et al., 2000; Truu et al., 2009) as well as fingerprinting techniques to track spatiotemporal variations in dominant microbes in these systems (Boon et al., 1996; Faulwetter et al., 2009). However, little has been done to thoroughly understand microbial ecology and enzymatic regulation in engineered wetlands. This understanding is necessary to more effectively manage microbial transformation of contaminants and to develop enhanced design and monitoring tools for future engineered wetlands. A suite of culture-independent molecular-based methods such as fluorescent in situ hybridization, quantitative PCR (DeJournett et al., 2007; Bacchetti De Gregoris et al., 2011), high throughput pyrosequencing for phylogenetic analysis, and the generation of metagenomes (Jiang et al., 2011) hold immense promise for future studies. Collectively, these molecular tools can further elucidate the microbial structure and function in constructed wetland systems. Studies that characterize enzyme expression in concert with phylogenetic characterization are important to more effectively track these complimentary but not always synonymous variables (Vilchez-Vargas et al., 2010).

Bivalve filtration wetland cells

As a compliment to contaminant attenuation in shallow, open-water wetlands and macrophyte-dominated wetlands, wetland cells can be built to provide a habitat for organisms that remove contaminants through filter feeding, as shown in Fig. 5. Studies have shown that bivalves such as mussels and clams filter large volumes of water and remove organic particulate matter from the water column (Winter, 1978; Møhlenberg and Riisgård, 1979; Kryger and Riisgård, 1988; Riisgård, 2001). While bivalves occur at low densities in habitats such as coastal estuaries, rivers, and littoral zones of lakes, large populations may be supported in systems with short hydraulic residence times and high primary productivity. Thus, it may be possible to support high densities of filter feeding organisms in a constructed wetland cell where significant concentrations of wastewater- or macrophyte-derived organic matter are available.

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FIG. 5.

Schematic of bivalve uptake and removal mechanisms of organic contaminants and pathogens.

While bivalves have not yet been applied for water quality improvement in unit process wetlands, they have been considered for a variety of applications, including drinking water treatment (McIvor, 2004), algae and suspended particulate matter removal from river water (Li et al., 2010), and clarification of secondary municipal wastewater effluent (Haines, 1979). Bivalves have also been considered as a means of removing nutrients from aquaculture wastewater (Buttner, 1986; Shpigel et al., 1997) and, more recently, to remediate surface waters polluted by excessive nutrients in New York City (Cotroneo et al., 2011; NOAA, 2011).

In the process of removing particulate matter, bivalves can also accumulate and, in some cases, transform particle-associated trace organic contaminants. For example, biotransformation of polybrominateddiphenyl ethers and polycyclic aromatic hydrocarbons by a freshwater mussel (Elliptio complanata) was observed after exposure through contaminated algae (O'Rourke et al., 2004; Drouillard et al., 2007). In addition, the transformation of crude oil was accelerated more than ten times in the presence of the mussel Mytilusedilus (Gudimov, 2002). However, recalcitrant contaminants may not be transformed after ingestion and may instead accumulate in the bivalve tissue (Verrengia Guerrero et al., 2002; Drouillard et al., 2007). These contaminants may then be released through the excretion of pseudofeces and feces (Haven and Morales-Alamo, 1966; Hull et al., 2011). Depending on the affinity of the compound for the (pseudo)feces, the contaminants may then desorb and reenter the water column, be consumed by benthic organisms, or be transformed by microbes. Further research is needed to determine whether particle-associated trace organic contaminants commonly present in wastewater effluents, such as the musk fragrances, are effectively removed and transformed by bivalves in a wetland cell.

Bivalves are also capable of ingesting a wide variety of pathogens (Silverman et al., 1995; Graczyk et al., 2003, 2006; Proakis, 2003; Nappier et al., 2008). Given their ability to efficiently remove particles greater than 0.4 μm in diameter from water, bivalves may filter individual bacteria (0.5–2 μm) and protozoan (oo)cysts (2–15 μm), although not individual viruses (20–100 nm). For example, the zebra mussel (Dreissena polymorpha) was found to remove E. coli and other bacteria from pond water at a clearance rate of about 6 L/[g dry tissue·h] (Silverman et al., 1995). Given this rate, a density of 130 mussels/L (∼16 mg dry weight/mussel) would result in 90% clearance of bacteria (1-log removal) over a 3 day residence time. Further research is needed, however, to determine whether bivalve grazing results in pathogen inactivation or just accumulation. Limited studies have shown recovery of viable Giardia spp. and infectious C. parvum oocysts in Macoma spp. and oysters, respectively (Fayer et al., 1998; Graczyk et al., 1999).

Within a unit process wetland, bivalves could be implemented in shallow, open-water cells. Periphyton-derived organic matter or organic matter from wastewater or previous wetland cells would provide sufficient particulate matter for ingestion, and bivalves would enhance contaminant and pathogen removal within the cell. Bivalves could be kept in cages to protect them from predators. Cages would also allow bivalves saturated with recalcitrant contaminants to be easily removed and depurated by exposure to clean water at regular intervals if necessary (Burns and Smith, 1981; Pruell et al., 1986; Peven et al., 1996). Native species should be employed in these wetland cells to avoid invasive species entering receiving surface waters. While further research is necessary to determine appropriate bivalve species, densities, and configurations to maximize treatment efficiency in a unit process wetland, implementation of such a wetland cell could significantly increase attenuation of certain trace organic contaminants and pathogens not effectively removed in other wetland cells, such as particle-associated trace organic contaminants and pathogens that are too small to be removed by sedimentation and which are not susceptible to photoinactivation.