819 24 Modified and Combined Systems Treatment wetlands can be designed to be “au natural” or complex. Examples of designs along a full spectrum from almost zero fossil-fuel input to highly engineered and man- aged systems can be found around the world. In this book we term these design and operation strategies as Type A and Type B wetlands. Type A systems are at the most natural, low-energy end of the scale. Type B wetlands are higher- energy and more highly engineered systems. There are good reasons to slide on this scale of design and management complexity, depending upon project goals. In this chapter we describe some of the rationale in selecting the appropriate level of management intensity. Type A treatment wetlands clearly cost less to operate than Type B systems. Type A systems may also cost less to construct, even though they typically occupy a larger foot- print area than Type B systems. They are the preferred solu- tion when a number of factors are not limiting: Relatively inexpensive land area is abundant. Land is relatively level. There is insufcient treatment potential with a Type A approach. There is insignicant potential for unauthorized human interaction with the wastewater in the wetland. Habitat and aesthetic benets outweigh the potential for nui- sance conditions related to mosquitoes or other wildlife. Type A wetlands are typically constructed free water surface (FWS) marshes and relatively low-tech constant-ow horizontal subsurface ow (HSSF) wetlands. The Type B treatment wetland-management strategy is preferred when one or more of the factors in the preced- ing list are limiting the design. Specic examples of Type B treatment wetlands include: HSSF wetlands augmented with reactive media intended for phosphorus or metals removal Addition of aerators for treatment improvement Operation of vertical ow (VF) wetlands in a ll- and-drain mode for improvement of nitrication and denitrication Combinations of VF, HSSF, or FWS wetlands to achieve specic process goals Recycling of treated efuent to improve denitri- cation and/or address inuent toxicity issues Step-feeding of inuent to reduce the impact of localized loading zones Wetlands and aquatic plant treatment systems with plant harvesting • • • • • • • • • • • High-solids treatment wetlands with frequently cleaned forebays Operating wetlands in a load-and-rest operating regime to address bed clogging or improve treat- ment efciency There are situations where these wetland modications make sense from a process design and economic standpoint, as dis- cussed in this chapter. However, the system will invariably require more operator attention, and there is often a reduc- tion in system reliability. For instance, if a treatment wet- land requires an air blower to achieve regulatory compliance, system reliability is essentially governed by the mechanical blower, not the wetland ecosystem. The same is true for dif- ferent operating regimes; if a treatment wetland requires alternating periods of loading and resting to avoid bed clog- ging, failure to follow the prescribed operating sequence will lead to system failure. Constructed wetlands have limits of performance, which have been discussed in previous chapters. It is human nature to attempt to relieve some of these constraints by modifying the basic wetland construct or its operation. It is the purpose of this chapter to explore some of the innovative ideas that have been implemented to enhance wetland performance. These fall roughly into four categories: Ecological or environmental modications Chemical additions Operational strategies Integrated sequences of natural systems, including one or more wetlands These will typically move the wetland system from a passive mode (Type A) to a mode of active management (Type B). They will also usually involve more capital and operating cost, so the potential advantages should be weighed against other options for improving performance, such as alternative technologies or increased size. 24.1 ECOLOGICAL OR ENVIRONMENTAL MODIFICATIONS Natural wetlands have been regarded as “self-designing” systems. They have also been ascribed certain functions that provide many kinds of benets: habitat, ood control, biodi- versity, and water quality improvement. Treatment wetlands target the water quality improvement functions. It is there- fore logical—to humans—to try to nudge the wetland to a condition that fosters water quality. For instance, it has been • • • • • • © 2009 by Taylor & Francis Group, LLC 820 Treatment Wetlands suggested that magnetic elds can improve wetland perfor- mance (Rasit, 2006). Ecological and environmental modi- cations that have been implemented to enhance the treatment performance of wetlands are briey discussed in the follow- ing text. MICROBIAL ENHANCEMENT Microbial processing plays a very important role in the reduction of many pollutants in treatment wetlands. In virtu- ally every case, among the many thousands of treatment wet- lands, the presence of microbial populations that will deal with the target pollutants is assumed in the design. For the major categories of contaminants, that presumption appears justied: if there is ammonia, nitriers are there; if there is nitrate, denitriers are there; if there is sulfate, sulfur bac- teria are there. However, these populations cannot establish instantaneously, and a period of some weeks of grow-in may be necessary. In general, it does not seem necessary to inocu- late a wetland with the desired microorganisms. It has been found that specialty bacteria—those that focus on particular or unusual compounds—will also require a period of time to develop. This leads to the observation of an induction period, during which little or no pollutant reduction occurs. For instance, Kadlec and Srinivasan (1995) found no degradation of naphthoic acid for about 100 hours in batch cattail mesocosms, but thereafter a steady decline (Figure 24.1). Runes et al. (2001) found that bioaugmentation of sedi- ment with small quantities of an atrazine spill-site soil (1:100 w/w) resulted in the mineralization of 25–30% of atrazine under both unsaturated and water-saturated conditions. Atra- zine and its daughter products were almost undetectable after 30 days. Unbioaugmented sediment supplemented with organic amendments (cellulose or cattail leaves) mineralized only 2–3% of atrazine. The population density of atrazine- degrading microorganisms in unbioaugmented sediment 2 3 4 5 6 7 8 9 10 11 12 0 50 100 150 200 250 300 Time (hours) Naphthoic Acid (mg/L) FIGURE 24.1 Disappearance of naphthoic acid in cattail mesocosms. A zero-order t is appropriate after an induction period. (R 2 = 0.99). (Data from Kadlec and Srinivasan (1995). Wetland treatment of oil and gas well wastewaters. Final Report DOE/MT/92010–10 (DE95000176), U.S. Department of Energy: Bartlesville, Oklahoma; graph from Kadlec and Knight (1996) Treatment Wetlands. First Edition, CRC Press, Boca Raton, Florida.) was increased from 10 2 to 10 4 /g by bioaugmentation, and increased to 6 r 10 5 /g) after incubation. A high population of atrazine degraders (approximately 10 6 g −1 ) and enhanced rates of atrazine mineralization also developed in bioaugmented sediment after incubation in ooded mesocosms planted with cattails (Typha latifolia) and with atrazine at 3.2 mg/L. Runes et al. (2001) concluded that bioaugmentation could be a way to enhance constructed wetlands. However, the persistence of degraders was not determined, and could be a constraint. WILLOW WETLANDS WITH ZERO DISCHARGE The growth habit of willows allows their use to provide zero-discharge, zero-percolation treatment for small house- holds. There are approximately 1,400 willow systems in use in Denmark, and these have been described in Chapter 3. They rely upon an excess of evapotranspiration (ET) over precipitation, which would not provide acceptable water dis- posal for regional values of ET. However, willow wetlands can be congured to enhance ET losses by using linear lay- outs in a crosswind aspect (see Figure 3.27). Willow systems are small, serving 30 PE or less, and are essentially “wind rows” that multiply regional ET by a factor of 2.5 due to the clothesline effect described in Chapter 4. The gravel beds that support the willows serve as a reservoir to accumulate wastewater during the wet seasons and periods. During dry periods, the interstitial water is drawn down. Therefore, siz- ing is determined by the water balance, which in turn is spe- cic to a particular climatic zone (Gregersen and Brix, 2001; Brix and Gregersen, 2002; Brix and Arias, 2005). Approxi- mately 120–300 m 2 are required for a single household under Danish climatic conditions. Raw sewage is pretreated in a sedimentation tank, and the supernatant is pumped onto or just under the surface of the plastic-lined bed. Treated efuent may be collected in a system of drainage pipes, and water may be recirculated back to the pumping well or the sedimentation tank. Zero-discharge © 2009 by Taylor & Francis Group, LLC Modified and Combined Systems 821 willow systems are constructed with a width of eight meters and a depth of 1.5 meters (Gregersen et al., 2003). A drain- age pipe is placed at the bottom of the bed, which is used to purge salt water from the bed after salinity rises too much. One third or one half of the willows are harvested every year to keep them in a young and healthy state with high transpi- ration rates (Brix and Arias, 2005). Harvesting at such small scale can readily be accomplished by hand labor. The willow systems will meet all regulatory require- ments because there is no outow to any receiving water bodies. Willow systems may also be used in conjunction with soil inltration. The willows will evaporate all wastewater during the growing season, but during winter there is partial discharge that is inltrated into the soil. ENGINEERED PLANTS In previous chapters it has been shown that the particular species of plants used in treatment wetlands is not critical, provided they create and sustain the desired biogeochemical cycles. However, there are differences among plants for some pollutants, most notably for trace metals and trace chemicals. Thus, in addition to the possibility of selection of hyperaccu- mulators, there is the longer-term possibility of genetic engi- neering for the desired traits. In recent years, much attention has been focused on the improvement of plants for the purpose of enhancing phytore- mediation through genetic engineering. One approach is to inuence the enzymatic pathways for assimilation and vola- tilization by overexpressing genes of rate-limiting enzymes in plants (Berken et al., 2002). Another approach is to intro- duce additional metabolic pathways for hyperaccumulators. In this way the capacity of plants to take up, accumulate, and volatilize compounds can be increased beyond that of any naturally occurring plant species. An ideal plant for environmental cleanup can be envi- sioned as one with high biomass production, combined with superior capacity for pollutant tolerance, accumulation, and degradation. With genetic engineering, it may be feasible to manipulate a plant’s capacity to tolerate, accumulate, and metabolize pollutants and thus create an improved plant for environmental cleanup. Pilon-Smits and Pilon (2002) found that mercury volatilization and tolerance was increased by genetic engineering approaches by two- to threefold more metal per plant. Nandakumar et al. (2005) found that Typha latifolia could be genetically transformed by Agrobacte- rium tumefaciens to enhance phytoremediation enzyme production. Heaton et al. (1998) proposed the use of transgenic aquatic, salt marsh, and upland plants to remove available inorganic mercury and methylmercury from contaminated soils and sediments. They noted that plants engineered with a modied bacterial mercuric reductase gene were capable of converting Hg(II) taken up by roots to the much less toxic Hg(0), which is volatilized from the plant. Plants engineered to express the bacterial organomercurial lyase gene were capable of converting methylmercury taken up by plant roots into sulfhydryl-bound Hg(II). Results from lab and greenhouse studies are beginning to show promise for transgenics, but eld studies are needed to establish their treatment potential, competitiveness in the wetland environment, and risks associated with their use. ARTIFICIAL ENCLOSURES One extreme example of environmental modication is to place the treatment wetland in an articial environment. In fact, this is commonly done for bench- and pilot-scale treat- ment units because the convenience of working indoors (especially in cold climates) outweighs added operational costs such as articial lighting (Figure 24.2). In larger-scale applications, greenhouses have been used as articial enclosures to maximize the use of available sunlight (Figure 24.3). Due to the high cost of the building envelope; greenhouse systems are designed as mechanical activated-sludge processes (U.S. EPA, 1997), although oat- ing plant racks are used on top of the aeration tanks. Plants used in the system may contribute to more stable operation due to retention of biosolids, and lower waste-activated sludge production has been reported (Austin, 2001). FIGURE 24.2 Indoor bench-scale unit treating landll leachate; White Bear Lake, Minnesota. Two 1,000-Watt lights are required to provide the necessary light intensity to support plant growth in these four mesocosms. Care must be taken to ensure that the lights do not cause the plants to dry out or burn. This setup was later modied by constructing a mini-greenhouse over the reactors using a PVC frame and a clear plastic sheet. © 2009 by Taylor & Francis Group, LLC 822 Treatment Wetlands The energy input associated with the activated-sludge process is quite high compared to treatment wetlands (see Table 1.1), and if the system is located in a cold climate, heat- ing demands can signicantly increase energy costs (Brix, 1999). Due to the high capital and operating costs associated with indoor systems, they have been used in applications where ancillary benets such as aesthetics, visitor access, and education are at a premium. 24.2 CHEMICAL ADDITIONS Wetlands do not necessarily supply sufcient reactants that may be desirable or required to foster a particular removal process. Those additional reactants may be supplied via the media for either FWS or SSF wetlands, or they may be sup- plied as additional chemical streams entering the wetland. REACTANTS VIA MEDIA The selection of media to support specic chemical reactions is a design issue. In the case of FWS units, it is common to ensure that adsorption capacity is present by the incorpora- tion of organic material in the rooting medium. Peat, com- post, or organic soils components also condition the soil and promote easier plant establishment and maintenance. In the case of HSSF systems, a much wider choice can be consid- ered, but typically materials are in three categories: Sorbents for phosphorus Reactive media for metals, including alkalinity/sul- de generators to generate precipitation reactions Materials that are designed to decompose and gen- erate a source of organic carbon for denitrication or sulfate reduction Light expanded clay aggregates, zeolites, ferrous or calcitic substrates, or organic materials can all be considered. Their phosphorus binding potential has been discussed in Chapter 10. • • • FIGURE 24.3 Greenhouse-based treatment system near Land O’ Lakes, Wisconsin. In the winter, considerable heating costs are incurred to minimize snow loads on the roof. Metals typically will bind to organics, and therefore mine drainage applications incorporate some form of compost as a layer in their substrate. This zone of wetland bed also func- tions anaerobically to reduce sulfate to sulde, which then precipitates metals (see Chapter 11). Mulch applied to the top of an SSF wetland bed can be used to supply organic carbon in excess of the plant biogeochemical cycle to fuel reactions such as denitrication if this is a process objective. All of these techniques of bed solids used as a chemi- cal addition have the implied need for periodic maintenance. Sorbents become saturated, and surface mulches become depleted in soluble materials unless the plant geochemical cycle can produce sufcient organic carbon after start-up. Therefore, these solid reactants must be replaced on some schedule, determined by the stoichiometry and loadings of the materials being treated. Because of the cost and difculty of restructuring the bed, and reestablishing the wetland, the projected lifetime of a bed solid reactant typically needs to be at least ve to ten years. Alternatively, reactant chemicals may be added as separate ows to the wetland. REACTANTS VIA ADDED STREAMS Mechanical treatment plants often utilize chemical additions to sustain particular portions of their treatment processes. Aeration or oxygen addition fosters both biochemical oxy- gen demand (BOD) reduction and nitrication. In plants that include denitrication, some means of resupplying the car- bon requirement is needed for conventional denitrication, such as utilization of an internal carbon-rich stream or the addition of an external carbon source (Metcalf and Eddy Inc., 1991). Removal of phosphorus in mechanical plants can involve the use of iron or aluminum salts to precipitate phos- phorus. These process steps all have analogs in constructed wetland technology. The processing of nitrogen in a wetland requires oxygen for nitrication and carbon for conventional denitrication (see Chapter 9). Supplies of these necessary reactants may not be sufcient to support the full requirements of required removals. Thus, HSSF wetlands are usually limited in their nitrication capabilities, and VF wetlands are usually limited in their denitrication capabilities. The capacity of a wetland to foster nitrogen reduction may be increased by inclusion of stream additions of carbon or oxygen, often at some consid- erable operational expense. The cost of such an enhancement should be weighed against other options, including the capi- tal cost of increasing the size of the system if necessary to be closer to the Type A end of the spectrum. Aer ationof FWSWetlands Although the entire water surface of a FWS wetland is exposed to air, reaeration is still subject to mass transfer resistances as discussed in Chapter 5. Oxygen supply may therefore be a constraint on BOD reduction and nitrica- tion. It has been supposed that open-water areas may con- tribute to enhanced oxygen supply (U.S. EPA, 2000a), but © 2009 by Taylor & Francis Group, LLC Modified and Combined Systems 823 that presumption is in question based on operating data (Kadlec, 2003d; 2005e). Therefore, forced-air diffusers have been considered for augmenting FWS performance, although such attempts have been few. At West Jackson County, Mis- sissippi, air injection was attempted in the deep-zone cross trenches. No conclusive data was obtained because nutria ate the plastic air diffusers. Ouray, Colorado Situated at 2,365 meters above sea level, this municipal con- structed wetland is surrounded by the San Juan Range, the youngest and steepest range of the Rocky Mountains (Camp- bell and Ogden, 1999; HDR/ERO, 2001). There is an average annual ow of about 1,500 m 3 /d, higher in summer during the tourist season. The wetlands are arranged in two parallel paths of three cells each. Each path is 21 r 143 m, for a total system area of 6,100 m 2 . Vegetation in all six cells is 80 to 90% cat- tail (Typha spp.), with the exception of Cell 3, which is 80% duckweed (Lemna minor) and 20% cattail (Typha latifolia). Other species present include common reed (Phragmites aus- tralis), curly dock (Rumex crispus), and lady’s thumb (Polygo- num persicaria). Because of concerns of sulfate reduction in the wetland inlet, the inlet deep zone of the wetlands is aer- ated with submerged perforated air distributors. The water quality from the aerated lagoon pretreatment is approximately secondary. The wetlands reduced the concen- trations of BOD to 4 mg/L, and total suspended solids (TSS) to 6 mg/L, over the period 1996–2000. These are clearly near background, and thus it is not possible to assign any advantage or disadvantage to aeration in these reductions. Ammonia in the wetland inuent was not monitored, but the wetland efuent ammonia was 9 mg/L over the same four- year period. This is a high concentration, which exceeds that expected for an aerated lagoon alone, and therefore no large advantage could have been achieved in the wetland aeration for ammonia reduction. Carbon Additions to FWS Wetlands Under many circumstances, denitrication in FWS wetlands is not carbon limited. The decomposing dead biomass is a moderately large carbon supply, although only a fraction of that material is useable (see Chapters 9 and 17). However, some waters contain large enough concentrations of nitrate that removal is constrained by the carbon supply. For instance, Stengel et al. (1987) worked with tap water at 30 mg/L nitrate nitrogen and augmented that to 110 mg/L to emulate ground- waters in northern Germany. It was concluded that, under some circumstances, insufcient carbon would be present to support full denitrication. As noted in Chapter 9, a C:N ratio of 5 or 10 to 1 has been deemed to be required for denitrication. This ratio deter- mines the potential need for an external carbon supply. For instance, for a HLR = 10 cm/d and a nitrate concentration of 50 mg/L, the NO x -N load is 5 g/m 2 ·d. Decomposition of 1,000 g/m 2 ·yr (dry weight) of biomass would yield about 500 gC/m 2 ·yr, or 1.4 gC/m 2 ·d. Thus, the C:N ratio of 1.4:5 is far less than needed to support full denitrication. Lamb–Weston, Connell, Washington The Lamb–Weston company built an integrated natural sys- tem for treating and reusing wastewater from a potato process- ing facility near Connell, Washington (Kadlec et al., 1997; Burgoon et al., 1999). Two 1-ha free water surface wetlands were constructed in series for denitrication of nitried efu- en t from VF beds (intermittent sand lters) (see Figure 24.18). The wetlands were designed to remove NO 3 -N from waste- water prior to land application. During a ve-month period of intensive study, the ow was 4,950 m 3 /d, producing a direct hydraulic loading of 25 cm/d. When coupled with an inlet nitrate nitrogen of 45 mg/L, the denitrication wetlands received a very high nitrate load of 110 kg/ha·d. The chemical oxygen demand (COD) in the nitried water was insufcient to support denitrication, even with the carbon generated by the wetland plants. Therefore, a small ow of untreated potato wastewater (COD ≈ 3,000 mg/L) was directed to the wetlands as a supplemental carbon supply. This ow raised the hydraulic loading rate (HLR) to 28 cm/d. Addition of this primary efuent resulted in a COD:NO 3 -N mass load ratio that ranged from 10 to 25. The NO 3 -N mass removal rate ranged from 50 to 99% of the inuent load. During the ve months January to May 1998, average monthly efuent NO 3 -N was 1.6 mg/L, and an average of 85% of the nitrate load was removed (Burgoon, 2001). Thus the feed-forward of untreated water provided the necessary carbon to allow nearly complete denitrication at high loading rates. Sources of carbon are not always conveniently at hand. Some wetlands have utilized methanol or molasses, but these are expensive. Multiple wastewater sources can potentially provide synergism. For example, the remediation of a high- nitrate source water could be enhanced in a wetland that also received high BOD municipal or food processing wastewa- ters. At the time of this writing, such multisource systems have been proposed but not built. Aluminum and Iron Additions to FWS Wetlands The phosphorus-binding capabilities of iron and aluminum, and their role in the wetland biogeochemistry of phospho- rus, are very well known (Reddy and DeLaune, 2007). Very early in the history of FWS technology development, it was observed that naturally occurring, blending streams of water containing iron were very effective in phosphorus removal. At Waitangi, New Zealand, phosphorus deposition occurred predominantly by reaction between treatment wetland waters (high in phosphorus, alkaline pH) with natural wetland waters (high in Fe and Al, acid pH) immediately below the conuence of their ow streams (Cooper, 1992; 1994). This study, coupled with the knowledge of mechanical plant phosphorus precipitation technology, led to investiga- tion of the direct addition of these metal salts to FWS wetlands for the purpose of enhancing phosphorus removal (Bachand et al., 1999; Bachand and Richardson, 1999). Preliminary jar © 2009 by Taylor & Francis Group, LLC 824 Treatment Wetlands test results suggested that treatment with ferric chloride and alum could effectively remove phosphorus to between 8 and 30 µg/L (Metcalf and Eddy Inc., 1997). It was hoped that chemical amendments directly within the treatment wet- lands would create a synergistic effect in which phosphorus is more effectively and efciently removed than by either of these two methods, chemical amendments and wetlands, alone. It was also hoped that addition of a polymeric coagu- lant might not be necessary because the wetland would effec- tively serve as a settling basin with very long HRT. However, small ocs formed, which did not quickly settle. Total phos- phorus concentrations in the water column remained elevated. This approach was therefore abandoned as a means of enhanc- ing wetland operational effectiveness. Aeration of SSF Wetlands Here for the rst time we encounter a signicant inuence of patents on SSF treatment wetland technology. Aeration of HSSF systems has developed both in the open literature and in the proprietary sector. Aerated subsurface ow wetlands are available as a patented technology in North America (Wallace, 1998; Dufay, 2000; Wallace, 2001b; Wallace, 2002a; Flowers, 2003; Wallace and Lambrecht, 2003). We cannot access all the proprietary data, and hence cannot independently evalu- ate it. For most patented systems, only anecdotal, qualitative information has been presented. However, we can summarize the results from the open literature and provide interpretation of it. Nothing here should be taken as supporting or refuting the claims in existing patents. The growing realization in North America during the mid-1990s that HSSF wetlands were inherently oxygen- transfer-limited systems led to considerable interest in arti- cial aeration through the application of a mechanical blower and diffuser tubing as a means to boost treatment perfor- mance (especially for ammonia) in these systems. One of the most basic applications of the technology is shown schemati- cal ly in Figure 24.4: Coarse media Impermeable liner Main bed media Water level Mulch layer Air header Air line Air FIGURE 24.4 Schematic of an aerated HSSF wetland. In the most simplistic sense, this can be thought of as “blow- ing bubbles” inside a saturated-ow HSSF wetland. However, implementation of the technology requires an understanding of the frictional loss of compressible uids, hydrodynamics, and oxygen transfer theory. Nevertheless, a variety of successful application areas have been developed, including petroleum hydrocarbons (Wallace and Kadlec, 2005), ammonia (Wallace et al., 2006a), domestic wastewater (Wallace et al., 2006b), and airport de-icing runoff (Wallace et al., 2007a). The decision to aerate a SSF wetland comes at a heavy operational cost, as operation and maintenance (O&M) of the facility is greatly increased. From an economic viewpoint, aeration is only justied when the lifecycle cost of aeration is sufciently offset by the reduction in capital cost, as the net savings of reduced wetland size less the cost of aeration equipment. Aerated wetlands typically occupy a much smaller foot- print than a nonaerated wetland, which translates into lower construction cost. Or, sufcient land may not be available for a more passive Type A wetland, or not available at an eco- nomically feasible price. At the time of this writing, performance information on aerated HSSF wetlands is a mixture of forensic analysis of full-scale systems, as well as analysis of performance data from pilot-scale systems or mesocosms. In the rst generation of aerated wetland design, assump- tions from other treatment technologies, such as lagoons, were often “borrowed” to support blower sizing and diffuser- tube placement in HSSF wetland systems. As a result, these systems were not optimally designed. In recent years, there has been an increased focus on pilot-scale systems, especially for industrial efuents. A well-designed and implemented pilot study can produce data on rate coefcients (essential for rst-order modeling), but questions about the hydrodynamic aspects of the pilot reactor versus the proposed full-scale sys- tem often remain (Noorvee et al., 2005b). Apart from reaction rates, there is the issue of oxygen transfer rates. These depend upon bed depth, interstitial © 2009 by Taylor & Francis Group, LLC Modified and Combined Systems 825 oxygen concentrations, diffuser geometry, and other factors. However, relatively gentle aeration can deliver something on the order of 50 to 100 gO/m 2 ·d to the HSSF bed. This rate of oxygen delivery is signicantly greater than rates that can be inferred from Type A nonaerated HSSF wetlands (Tables 9.26, 9.27, Chapter 9), which are usually in the range of 1 to 6 gO/m 2 ·d. The energy requirement for such gentle aeration can nev- ertheless be an important factor in overall energy usage (see Table 1.1, Chapter 1), especially compared to passive Type A wetlands, but is generally much lower than conventional mechanical treatment processes. Domestic Wastewater Treatment In a recent study, Wallace et al. (2006b) compared the treat- ment efciency of 17 full-scale aerated HSSF wetlands and 22 full-scale nonaerated wetlands, using the loading chart and statistical methods presented in Wallace and Knight, 2006. The conclusion of this study was that aeration resulted in signicant improvements in the reduction of both BOD 5 and TSS. With aeration, the equivalent wetland size needed for BOD 5 removal could be reduced by 67%; for TSS, the size reduction was approximately 36%. Ammonia Oxidation Aeration of SSF wetlands for ammonia removal has been studied in recent pilot-scale systems. For an efuent containing a mixture of cyanide/iron/ammonia from a gold mine in South America, aeration was found to increase the ammonia oxida- tion rate over tenfold, from a volumetric rate (2 TIS) of 0.52 d −1 to 5.7 d −1 (Wallace et al., 2006a). A study of ammonia oxidation at different temperatures for domestic wastewater at a pilot-scale wetland in Canada (Wallace et al., 2006a) yielded a temperature correction factor (Q) of 1.02, indicat- ing that temperature sensitivity is of some importance in aerated wetlands (compare with Table 9.17, Chapter 9). This would be consistent with a view that heavily loaded aerated HSSF wetlands are microbially dominated ecosystems. De-icing Runoff Treatment of de-icing compounds (discussed in Chapter 13) has been studied under aerated and nonaerated conditions in HSSF mesocosms. For removals of CBOD 5 associated with de-icing runoff, the volumetric rate coefcient was measured at 0.6 d −1 (2 TIS); aeration at 0.85 m 3 air per hour (per m 3 of wetland bed) improved the rate to 5.5 d −1 . Additional aeration at a rate of 2.0 m 3 air per hour (per m 3 of wetland bed) further increased the degradation rate coefcient to 15.9 d −1 . Based on the range of temperatures studied (22 to 4°C), the tem- perature correction factor was 1.03 (Wallace et al., 2007a). Petroleum Hydrocarbons Aeration has been demonstrated to improve removals of petroleum hydrocarbons through aerobic degradation and volatilization. Early work was conducted at the Gulf Stra- chan Gas Plant in Alberta, where an aeration system initially designed for freeze resistance during winter operations resulted in improved removals of total petroleum hydrocar- bons (TPH) and benzene, toluene, ethylbenzene, and xylene (BTEX) compounds (Moore et al., 2000a). A full-scale aer- ated wetland system was constructed by Williams Pipeline for a terminal facility in Watertown, South Dakota, and has been operating successfully (Wallace, 2002b). Pilot-aerated SSF wetland studies were conducted by Ferro et al. (2002) for the purpose of reducing petroleum hydrocarbons in extracted groundwater. Both aerated and nonaerated degradation rates were measured, as summarized in Table 13.2, Chapter 13 (Wallace and Kadlec, 2005). These ndings are consistent with the growing body of knowledge of biodegradation and volatilization rates for hydrocarbons under aerobic and anaerobic conditions (Suarez and Rifai, 1999). Aquaculture Waste A study at the Botanical Garden of Montreal investigated the effects of aeration in mesocosms for the treatment of aqua- culture waste (Ouellet-Plamondon et al., 2006). Aeration improved the removal of TSS, COD, and total Kjeldahl nitro- gen (TKN). For instance, the rst-order COD rate coefcient (plug ow) improved from 0.15 to 0.25 m/d for unplanted mesocosms, and from 0.19 to 0.24 m/d for Typha-planted mesocosms. Aeration was only used in the inlet region of the HSSF bed. The researchers concluded that aeration was a promising approach for improving the performance of HSSF wetlands. Oxy gen Transfer A recent study was conducted to measure the standard oxygen transfer efciency of aerated SSF beds using off-gas measure- ments (Wallace et al., 2007b). Oxygen transfer is a function of the contact time between the bubble and the water column, which depends on the water depth. It also depends on constit- uents within the wastewater that affect oxygen mass transfer at the bubble/water interface (Metcalf and Eddy Inc., 1998). The average oxygen transfer efciency in a HSSF wetland mesocosm using a standard sodium sulte solution was 4.7% per meter of water depth. As a frame of reference, coarse bubble diffusers have a transfer efciency of approximately 2.6% per meter, and ne bubble diffusers have efciencies of approximately 6.6% per meter (U.S. EPA, 1989). Because most HSSF wetlands have a bed depth of 0.3 to 0.6 m, this would put the oxygen transfer efciency in the range of 1.4 to 2.8%, depending on the bed depth. The linear diffuser tubing used in this study was a ne bubble aeration device. In an open-water tank, the expected oxygen transfer would be approximately 6.6% per meter. The lower-than-expected oxygen transfer rate could speculatively be from the inuence of the granular media, which may serve to coalesce bubbles and increase their size, reducing the oxygen transfer efciency. Air-induced mixing of the wetland was greatly reduced by the presence of the granular bed media; based on visual © 2009 by Taylor & Francis Group, LLC 826 Treatment Wetlands observations, the inuence of aeration was generally restricted to a radius of approximately 15 cm from the point of bubble introduction (Wallace et al., 2007b). Aerated wet- lands are hydraulically inefcient, as shown by tests yielding number of tanks in series (NTIS) of 1 to 3. It is apparent that a very uniform aeration grid is required to maintain oxygen- ated conditions in the wetland bed. Figure 24.5 shows the test of such an aeration system. Aeration of SSF wetlands is not a maintenance-free activ- ity. Without aeration, the system will fall back into the oper- ating envelope of nonaerated wetlands. However, because regulatory compliance is dependent on the aeration system, scheduled, proactive maintenance of blowers and other aera- tion components need to be carried out by the system opera- tor. Wetland designers should also consider the fouling of air diffusers within SSF wetlands, and provisions to replace or chemically clean diffuser assemblies should be addressed. Carbon Additions to HSSF Wetlands Early work on denitrication in HSSF wetlands was con- ducted by Gersberg et al., 1984. HSSF wetlands were fed with secondary efuent low in BOD (3.3 mg/L) and high in nitrate (18.8 mg/L). In initial trials, the removal of total nitrogen was limited by the availability of organic carbon, and ranged between 9 and 19% at hydraulic loading rates of 8.4 to 20 cm/d. Utilization of methanol as a supplemental organic carbon supply dramatically increased denitrication efciency, up to 97% at a hydraulic loading rate of 16.8 cm/d (Gersberg et al., 1984). As a lower-cost carbon supply, plant biomass was mulched and placed on top of the beds. At a hydraulic loading of 8.4–12.5 cm/d, the mean removal rate was 89% for total nitrogen, although the carbon loading (0.09 kg/m 3 ) was higher than that required for methanol (0.03 kg/ m 3 ). This nding is consistent with more recent studies on the bioavailability of plant detritus to fuel denitrication (see, for instance, Ingersoll and Baker, 1998; Baker, 1998; Hume et al., 2002a). FIGURE 24.5 Testing of HSSF wetland aeration system for BTEX degradation; Casper, Wyoming. Rows of bubbles represent the location of linear air diffuser tubes, which were placed in radial arcs within this circular wetland. The wetland basin was articially ooded to allow visual observation of the bubble pattern. Use of organic materials as a thermal insulator has the drawback of producing leachates with BOD that adds to the loading for that contaminant in domestic and municipal sys- tems (Wallace et al., 2001). However, in systems treating low BOD sources, such as many runoff and groundwater remediation projects, such an additional source of organics is desirable and may be essential to the proper functioning of the wetland. Surface mulching of HSSF wetlands is also an option for supplying carbon for denitrication of nitrate-rich inuents, as demonstrated by Gersberg et al., 1984. Mass- balance considerations to fuel denitrication processes are discussed in Chapter 9. VF wetlands are effective in oxidizing ammonia to nitrate, but typically are not efcient in denitrication pro- cesses (Cooper, 2001). One of the limitations of denitrica- tion is availability or organic carbon; and VF wetlands are often combined with HSSF wetland stages for this purpose (Seidel, 1973; Brix, 1994d; Cooper, 2001). 24.3 OPERATIONAL STRATEGIES The ow of water need not be continuous, nor need it be uni- directional from inlet to outlet of a wetland or integrated nat- ural system. There are other methods of operation, involving spatial or temporal sequencing, and the potential of recycle ows. Some of these concepts are explored in the following text. STEP FEED Step feeding refers to the introduction of water to a system at several points along the ow path. In activated-sludge mechanical plants, it refers to the introduction of settled wastewater at several points along the ow in a plug-ow aer- ation tank (Metcalf and Eddy Inc., 1991; Crites and Tchob- anoglous, 1998). In water hyacinth systems, as many as eight points of introduction have been used (Water Environment Federation, 2001). The Gustine, California, FWS system was © 2009 by Taylor & Francis Group, LLC Modified and Combined Systems 827 built with provisions for step feeding (U.S. EPA, 1988b). The reason for step feeding is, in all cases, stated to be the avoid- ance of “overloading” of the inlet zone of the system. Because the concept of loading to an interior zone of the wetland is fuzzy, a hypothetical situation is considered. The Gustine example is selected as the framework, for which a step feed at the one-third point of the length is pos- sible (Figure 24.6). The original design called for either one third or two thirds of the ow to be introduced at this interior point (U.S. EPA, 1988b). It is known that the decline of BOD through this system is exponential (see Figure 8.7, Chapter 8). Because of the high length-to-width ratio (337 m:11.6 m = 29:1), the plug-ow version of the k-C* model will be used. Calibration to the Gustine data of Walker and Walker (1990) yields k = 48 m/yr and C* = 4 mg/L. The average BOD enter- ing the system was about 200 mg/L, and the hydraulic load- ing was 3.3 cm/d. Figure 24.7 shows the effect of the step feed of one third of the inuent to the one-third point. The front end of the wetland receives a lesser ow and does a more effective job of reduction of concentration. Upon blend- ing with the step feed, the concentration goes up above that which would occur without step feed and stays above for the rest of the wetland travel distance. The outlet concentration is 29% higher for the step-feed option. It may be shown that for any step-feed operation and rst-order kinetics, step feed gives lesser performance. The overall BOD load to the wetland is not changed by step feeding because the area is xed, and the inlet ow and concentration are xed. The question then becomes: what has been accomplished? If all that is considered is BOD, the answer is, nothing. However, if it is presumed that part of BOD removal has to do with the sedimentation of solids, then TSS accumulation might factor into our thinking. The Gus- tine system reduced TSS from 75 mg/L to about 30 mg/L. Suppose that the entire accretion is to an inlet zone of 37 m length (1/9 of the length), at a bulk density of 0.5. Then the buildup would be about 1 cm per year. By step feeding, one- third of that accretion would be transferred to the step-feed zone, and the inlet buildup would be only two thirds of a cen- timeter per year. Therefore, the evaluation of step feeding as FIGURE 24.6 The step feed distribution pipe in the Gustine, California, system. FIGURE 24.7 Effect of step feed on BOD concentrations. These are model-based calculations, based on calibrations to the Gustine, California, system. 1 10 100 1,000 0 50 100 150 200 250 300 350 Distance (m) BOD (mg/L) No Step Step © 2009 by Taylor & Francis Group, LLC 828 Treatment Wetlands an alternative depends upon the necessity for spreading out the solids and the corresponding penalty in performance. RECYCLE Recycle can serve two important purposes in treatment wet- lands: (1) it brings products of reactions in the wetlands from the exit back to the inlet, and (2) it dilutes reactants entering the system. Both of these functions can be used to advantage in some wetland situations. Dilution of Influents Wetland vegetation can be sensitive to high concentrations of some pollutants in the water. Therefore, strong inuents can damage or destroy vegetation, and the problem will be most severe in the inlet of the wetland before any treatment has occurred. For example, landll leachates can be inimi- cal to wetland vegetation (Figure 24.8). Recycling treated water from the wetland system outlet can serve to dilute the incoming water, but at some price in terms of cost and level of treatment. FIGURE 24.8 A leachate seepage outbreak at the Saginaw, Michigan, closed landll. This water contained ammonia at concentrations sometimes exceeding 1,000 mg/L, and was severely toxic to all vegetation. 0 20 40 60 80 100 120 0.0 0.2 0.4 0.6 0.8 1.0 Fractional Distance Relative Concentration R = 0 R = 0.5 R = 1.0 R = 2.0 Complete Mix FIGURE 24.9 The effect of recycle on a hypothetical wetland. The fresh-feed Damköhler number is k/q = 2.0, P = 4, and C* = 0. To illustrate the concept, a hypothetical wetland is con- sidered, with a recycle of water back to the inlet. A fresh- feed Damköhler number k/q = 2.0, P = 4 TIS, and C* = 0 are chosen, to be representative of a moderate degree of removal of the contaminant (80%). Figure 24.9 shows the calculated concentration proles through the system for dif- fering amounts of recycle up to 200% of the incoming raw water. Note that as the recycle ratio goes to very high levels, the effect is to create a single well-mixed unit because of the intense recirculation. The limiting concentration reduction for this extreme is 67%. Typical recycle rates are 200% or less, for which the reduction is 72% for this example. Thus, a price in performance has been incurred by recycle. However, the concentration in the inlet region is cut in half (52%) by 200% recycle. The economic consequences of a recycle pump and pip- ing may not be large for very small systems, but for large FWS wetlands they may be a signicant contribution to cost. Recycle is being used on the large Everglades-con- structed wetlands to process seepage collected around their perimeters. © 2009 by Taylor & Francis Group, LLC [...]... FIGURE 24. 18 Layout of the Lamb–Weston integrated natural system © 2009 by Taylor & Francis Group, LLC FWS Wetlands (2) Denitrification Land application 840 Treatment Wetlands 180 TKN NH4-N ORG-N NOx-N TN Concentration (mg/L) 160 140 120 100 based on different treatment process objectives, and are considered standard process configurations These combinations have been previously discussed in detail in Chapters... wetland basin can then be immediately drained or allowed to operate as a continuous-flow, constant-water-level system for some percentage of the operational sequence The treatment performance of a full-scale HSSF wetland in Minoa, New York, significantly improved after the facility was converted from a constant-flow to a fill-and-drain mode of operation (Liebowitz et al., 2000) The presence of sulfide precipitates... such a type of wastewater Fill-and-drain wetlands operating under an accelerated cycle frequency are commonly called “reciprocating” or “tidal flow” wetlands The fill-and-drain operational sequence helps BOD removal and nitrogen removal to occur simultaneously, and some nitrate-bound oxygen may be recovered To achieve the necessary degree of oxygen transfer to drive the treatment process, cells must... Oregon Unpublished Data P-W Con Lake Nebagamon, Wisconsin Unpublished Data P-W-I Con Drummond, Wisconsin Unpublished Data P-W Nat 24 6 46 N (warm) in out 0.14 — — — — 0.13 0.17 2.36 0.94 — — Minot, North Dakota Hammer and Burckhard (2000) P-W Con 12 51 48 N (warm) in out 3 32 14 16 10 4.5 2.1 — — 381 219 Estevan, Saskatchewan Unpublished Data P-W Con 4 — 49 N (warm) in out 3 33 24 9 5 2.82 0.35 2.49 2.18... guidelines for small on-site treatment wetlands require recycle at a recycle ratio of 50% (Brix and Arias, 2005) FWS TIMED OPERATIONAL SEQUENCES The flows to FWS wetlands need not be steady and continuous Event-driven systems constitute one group of systems that receive nonsteady flows, and these have been discussed TABLE 24. 1 Comparison of Recycled and Once-Through Operation of a Two-Stage Experimental... (1999) P-W Con 11 23 30 Y in out 3.1 40 14 29 8 3.9 1.2 — — — — Fort Deposit, Alabama Bays and Knight (2002) AP-W Con 24 6 31 Y in out 0.8 92 9 34 5 3.1 1.0 — — — — Pontotoc, Mississippi U.S EPA (1999) P-W Con 2 0.04 34 Y in out 1.27 96 31 49 23 112 36 29 18 — — Duplin, North Carolina Unpublished Data P-W Con 6 0. 024 35 Y in out 4 — — — — 120 21 — — — — Kingman, Arizona Gerke et al (2001) P-W Con 6... performance of fill-and-drain SSF wetlands A recent laboratory study by Austin (2006) indicated that fill-and-drain columns filled with an expanded shale aggregate (CEC of approximately 4 meq/100 g) significantly outperformed columns filled with high-density polyethylene (HDPE) media (zero surface charge) despite the fact that the HDPE media provided more surface area for microbial biofilms Expanded-shale columns... 2,000–8,000 330–1,500 800–3,600 200–360 65 240 150–300 40–100 69–480 57 240 — 300 — — — 100–1,700 100–450 220–300 300–480 75 240 500–1,500 — 200–900 200–900 80–360 Source: Based on Sukias and Tanner (2005) In Pond Treatment Technology Shilton (Ed.), IWA Publishing, London, United Kingdom © 2009 by Taylor & Francis Group, LLC 838 Treatment Wetlands FIGURE 24. 15 The Lake Nebagamon, Wisconsin, infiltration... ecosystem units into optimal treatment systems Several examples are presented of existing integrated systems PONDS AND WETLANDS High-Sediment Waters Many urban stormwater treatment systems include sedimentation basin forebays (pond elements) followed by emergent SF wetlands (Strecker et al., 1992; Schueler, 1992) Figure 18.31, Chapter 18, provides an example of a stormwater treatment system that incorporates... example, Brix et al (2002a) performed comparison studies on a two-stage VF system treating municipal wastewater, with and without recycle (Figure 24. 10; see also Figure 16.11, Chapter 16) Table 24. 1 shows data from one pair of two-month campaigns, out of four that were conducted Several conclusions can be drawn from these results First, the wetlands are very effective in reducing BOD5 and TSS, whether . harvesting • • • • • • • • • • • High-solids treatment wetlands with frequently cleaned forebays Operating wetlands in a load-and-rest operating regime to address bed clogging or improve treat- ment efciency There. effective per- formance of ll-and-drain SSF wetlands. A recent laboratory study by Austin (2006) indicated that ll-and-drain columns lled with an expanded shale aggregate (CEC of approxi- mately. allowed to operate as a continuous-ow, constant-water-level system for some percent- age of the operational sequence. The treatment performance of a full-scale HSSF wetland in Minoa, New York,