Effectiveness of Constructed Water Quality Treatment Systems for

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Effectiveness of Constructed Water Quality Treatment Systems for Mitigating Pesticide Runoff and Aquatic Organism Toxicity Marie E. Stillway,* Bruce G. Hammock, and Swee J. Teh Department of Anatomy, Physiology, and Cell Biology, School of Veterinary Medicine, Aquatic Health Program Laboratory, University of California, Davis, California 95616, United States *E-mail: [email protected].

Water quality treatment systems, such as vegetated agricultural ditches, retention ponds, and water quality treatment wetlands, are used to reduce waterborne contaminant loads. These systems have been gaining prominence for their successful mitigation of pesticides, agricultural effluence, and pharmaceutical and personal care products. Aquatic organisms exposed to the outflow runoff are the most vulnerable to toxicant exposure. While there is clear evidence of successful contaminant remediation within these systems, the link between pesticide concentration reduction and biological responses of aquatic organisms exposed to treated runoff is less clear. Generally, toxicity is reduced as water passes through the treatment systems with accelerated amelioration in those systems where vegetation is present. Contaminant concentrations are often reduced to levels below analytical detection limits, and, in most cases, organism responses (e.g., mortality, growth) are negatively correlated to pesticide concentrations. However, a complete removal of toxicity is rarely observed. Hydrophobic pesticides elicit toxicity in sediment tests, whereas hydrophilic pesticides cause more adverse effects in the water column tests. Pesticides bound to sediment particles were shown to be less bioavailable than those in the water phase. Although there is concern about source/sink dynamics of these treatment ponds in the long

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term, these treatment processes are an effective way to mitigate pesticide loads in water sources. Studies evaluating treatment system efficacy over time, and whether wetland sediment toxicity increases during that time, are needed to determine the long-term viability of these systems.

Introduction Continued worldwide population growth has increased the number of structures and landscapes needing pest control. The application of pesticides to permeable areas such as lawns, golf courses, and parks can contribute to the pesticide load in urban landscapes. Likewise, due to urbanization, there are now more non-permeable surfaces such as sidewalks, driveways, gutters, sewers, and storm drains. These impermeable surfaces drastically change runoff hydrologic characteristics by reducing transport time and increasing contaminant load into receiving waters. In addition, an increase in food demand, caused by concomitant population increases, will intensify farming and, as a result, pesticide use. Climate change impacts such as rising temperatures are expected to affect agricultural practices in terms of extending growing seasons as well as an increased prevalence of crop diseases and pests (1). Thus, pesticide use will likely increase under these scenarios. Non-point source pollution through runoff, drainage, and pesticide spray drift accounts for the majority of all surface water pollution (2, 3). In California, reported pesticide use in 2016 totaled 209 million lbs of applied active ingredients over 101 million treated acres (4). California was ranked number one in the nation in 2012 for total value of agriculture products sold including crops, nursery, and greenhouse items (5), so the need for best management practices in pesticide remediation is essential for protecting the water bodies that receive contaminated runoff in both agricultural and urban settings. Water quality treatment systems are a best management practice used to reduce waterborne contaminant loads and have been gaining prominence with their success in mitigating pesticides, agricultural effluent, and pharmaceutical and personal care products (6–8). The efficacy of these types of water quality treatment processes have been discussed in the literature regarding reductions of pesticide concentrations; reviews from Budd (9), Moore et al. (10, 11), and Vymazal and Brezinova (12) are excellent resources for those interested in the specific mechanisms under which these processes take place. However, there are concerns that over time these treatment ponds and wetlands can become sinks for some contaminants, which could have an effect on their efficacy in reducing toxicant loads as well as putting ecosystem services associated with these wetlands at risk. While there is clear evidence of successful contaminant remediation within these systems, the link between pesticide concentration reduction and biological responses of aquatic organisms exposed to the treated outflow waters is less clear. Therefore, the intent of this chapter is to synthesize the existing literature examining the efficacy of these treatment processes and their relationship to organism toxicity reductions and biota responses to treated outflow water. 436

Components of an Effective Water Quality Treatment System Constructed wetlands are land-based water treatment systems, such as vegetated agricultural ditches, bioswales, filter strips, retention ponds, and water quality treatment wetlands, that consist of trenches or shallow ponds that contain floating, emergent, or rooted wetland vegetation (13). A bioswale is a type of vegetated drainage ditch with sloped sides (14), whereas a filter strip creates a vegetated barrier between receiving water bodies and polluted inputs (12). These types of treatment systems are typically found in conjunction with agricultural areas. In comparison, constructed wetlands (either surface flow- or subsurface flow-based) have several components, can be constructed in a variety of ways, and can be implemented in urban, industrial, and agricultural land use settings. When designed in consideration of the mechanisms at work, these wetlands can be an effective tool in reducing off-site contaminant movement. These systems have several advantages, such as low operational costs and minimal maintenance, and can provide usable habitat and ecosystem services once established (8, 15). Constructed wetlands and ditch systems mitigate pesticide pollution through several mechanisms including microbial processes, sedimentation, phytoremediation, hydrolysis, and plant uptake (16). The presence of vegetation provides a substrate to which contaminants can bind and decreases flow velocity allowing for longer residence times (9, 17). Increased residence time has a direct influence on sedimentation processes that contribute to contaminant removal (9); this is because water retention enables the settling of suspended particles, sediment adsorption, and plant uptake (15). Moreover, increased residence time is a key component in the occurrence of many microbiological processes (16). However, vegetation is only one part of a combination of mechanisms that work together to reduce contaminant loads; consideration of the anticipated contaminant class or type is important when designing treatment wetland systems due to the varying physicochemical properties of the toxicant(s) in question. Physicochemical Processes Bed characteristics are essential mechanisms at work within these natural treatment processes. Organic matter, soil clay content, pH, and temperature all affect adsorption rates, and the physicochemical properties of the pesticide in question will have a large impact on its partitioning tendencies and subsequent removal from the water column. Pesticides that are highly soluble tend to have low adsorption coefficients (e.g., octanol-water [Kow] partition coefficients (18)). The octanol-water partition coefficient characterizes the hydrophobicity of a molecule and is the ratio of the affinity a molecule has for the water (polar) phase and the non-water (octanol or non-polar) phase (19); because of this, hydrophobic compounds will have high Kow values, and hydrophilic compounds will have low Kow values. Additionally, the organic carbon-water partition coefficient (Koc) is the tendency for a molecule to transfer from soil-water to soil-solids and is the ratio of the concentration of a compound absorbed by the soil organic carbon to the concentration of the compound in water (19). A molecule with a high Koc will have a greater affinity for adsorption to soils and sediments. 437

These adsorption processes are critical in predicting removal efficiencies as these mechanisms affect the amount of pesticide that will bind to the sediment. The amount of pesticides partitioned to the solid phase will have a direct effect on the amount of pesticides that are available to be adsorbed by the vegetation in a water quality treatment system. Hydrophobic compounds will have higher absorption rates (to sediments, plant material, etc.) than hydrophilic compounds (19–21) due to their affinity for the solid phase. Hydrophillic compounds remaining in the water phase will have less of an affinity to bind to solids. The partition relationships are predictive of adsorption rates and, therefore, treatment system efficacy in the removal of contaminants. For instance, the hydrophobicity of pyrethroid pesticides has been well documented due to their low water solubility, high Kow, and high Koc values. As a result, one can expect that pyrethroids will have a higher affinity for the solid phase than for water, with log Kow values ranging from 4 (esfenvalerate) to 7.6 (tralomethrin (22),) and Koc values ranging from 16,400 (permethrin) to 180,000 (lambda cyhalothrin (23),). In comparison, organophosphate pesticides with lower Kow and Koc values will have less affinity for the solid phase, such as methyl parathion which has a log Kow of 2.86 and a Koc of 5,100 (24). Water quality treatment systems can have varying success depending on the types of contaminants present. For instance, Hunt et al. (25) evaluated the effectiveness of an on-farm vegetated treatment system with regard to the removal of a number of pesticides including pyrethroids, organophosphates, and organochlorines. While diazinon was reduced as it travelled through the system, the authors attributed the reductions to dilution not adsorption, whereas the pyrethroid reduction was attributed to sedimentation. Moreover, Bouldin et al. (6), observed continued diazinon transport in their evaluation of a constructed wetland in a simulated rainfall event. Although diazinon concentrations were reduced throughout the system, detectable concentrations of diazinon, albeit low (0.06 µg/L), were still present in the outflow water after 26 days. Differences in adsorption rates were also observed by Budd et al. (26) based on pesticide class. Pyrethroid removal was very efficient, with reduction levels above 95%. The authors determined that the majority of pyrethroids in the water column were associated with suspended solids, indicating the affinity of pyrethroids for lighterweight particles as documented in other studies (26–28). Conversely, the removal efficiency for the organophosphate diazinon was a little more than half (68%). The continued presence of diazinon in the system was, in part, due to its affinity for the water phase; as with a Koc value of 1000, diazinon remains in the water phase rather than adsorbing to the sediment or vegetation.

Additional Abiotic Factors Abiotic factors are important to consider in terms of treatment efficacy. Temperature often increases the rate of degradation of chemicals and, in conjunction with pH, can result in notable differences in breakdown. For instance, research conducted by Brogan and Reyala (39) revealed that the breakdown of malathion in the presence of plants was due to alkaline hydrolysis rather than 438

sorption to the plants themselves. Brogan and Reyala (40) also demonstrated high breakdown efficiencies of less hydrophobic insecticides, such as carbaryl and carbofuran, via alkaline hydrolysis. They refer to this mechanism as the “hydrolysis-based model,” where plants increase water pH via photosynthesis, which then initiates the rapid breakdown of susceptible compounds (40). This mechanism of chemical breakdown may not necessarily apply to all chemicals, especially to those that are resistant to hydrolysis such as chlorpyrifos and diazinon (39). This demonstrates that there are several mechanisms at work in reducing pesticide loads; therefore, a water quality treatment system designed to allow a variety of processes will have the broadest success in remediation. The Role of Vegetation Within the Treatment System Vegetation is the key to removing contaminants from the water column, either directly or indirectly. Highly soluble pesticides are displaced to greater depths within the water column compared to those with less solubility (19). With pyrethroids’ low solubility, and their affinity for absorpting lightweight particles, there is the potential for increased transport during high-flow conditions. Dense vegetation acts as a buffer in high-flow conditions, reducing water flow and increasing residence time in the wetland, which in turn enhances the efficacy of the treatment system. Increased retention time enables pesticide-bound lightweight particles to settle out of the water column for further absorption to sediment and vegetation, while also allowing more soluble pesticides to remain in the system for degradation, adsorption, or hydrolysis. Researchers have studied the uptake rates in vegetated treatment systems and found accelerated removal in vegetated systems versus non-vegetated. Lizoette et al. (29) evaluated the removal efficiencies of a nutrient/pesticide mixture in vegetated and non-vegetated sections in a constructed wetland during a simulated rain event and found that the vegetated sections of the wetland were more effective in pesticide removal at 5h after dosing than the non-vegetated sections and suggest a 21-day retention time for maximum remediation capacity. Matamoros and Salvado (8) evaluated the removal efficiencies of 27 emerging contaminants in wastewater treatment plant effluent in Spain. Although the results were compound-dependent, overall concentration reduction was more efficient in the vegetated surface flow-constructed wetland compared to the non-vegetated polishing pond. Milam et al. (30) compared the reduction of methyl parathion between vegetated and non-vegetated wetland mesocosms and found that concentration reductions were due to the partitioning of the pesticide to the plants present in the vegetated wetland, suggesting a 10-day residence time for maximum methyl parathion removal. Moore et al. (31) also demonstrated higher methyl parathion removal efficiencies through plant uptake in vegetated constructed wetlands compared to non-vegetated counterparts. Methyl parathion was below detection limits in the vegetated wetland outflow water, whereas 8.83 µg/g of methyl parathion was detected in the non-vegetated wetland outflow (and at varying concentrations at all other sampling points). In this study, mass balance calculations indicated that the majority of methyl parathion was partitioned to 439

the plants in the vegetated wetland and into the sediment compartment in the non-vegetated wetland; concentrations were an order of magnitude higher in the non-vegetated wetland compared to the vegetated counterpart (31). These results demonstrate the importance of the combined mechanisms that work together in the remediation process. Researchers have demonstrated the successes of various vegetated water quality treatment processes. For example, Moore et al. (31–35), and Lizoette et al. (29, 36, 37) have conducted extensive research to evaluate the efficacy of vegetation within these treatment systems and have demonstrated the acceleration of contaminant uptake in the presence of vegetation. Mahabali and Spanoghe (38) evaluated two types of South African native plants (Nymphaea amazonum and Eburia mutata) in mesocosms dosed with ‘low’ and ‘high’ concentrations of lambda-cyhalothrin and imidacloprid and found that removal of lambda-cyhalothrin and imidacloprid were both independent of the plant type. Although a sediment-only mesocosm wasn’t evaluated, lambda-cyhalothrin was completely eliminated within 72 h and imidacloprid removal efficiency ranged from 72–100% within 216 h in both planted mesocosms (38). Moore et al. (34) found no differences in the efficacy of permethrin removal across four types of common Mississippi Delta macrophytes: cutgrass (Leersia oryzoides), cattails (Typha latifolia), burrweed (Solanum americanum), and powdery alligator-flag (Thalia dealbata). However, compared to a non-vegetated control, overall mass reductions of 71% for cis-permethrin and 76% for trans-permethrin were reported (34). Budd et al. (26) evaluated two constructed wetlands adjacent to the San Joaquin River in California, which had combinations of knotgrass, pale smartweed, and barnyardgrass as the dominant plant species. Tailwater inflow was comprised of a mixture of pyrethroids, chlorpyrifos, and diazinon. Reductions in pyrethroid concentrations ranged from 52–94% and reductions in chlorpyrifos ranged from 52–61%, while the systems were less successful in the removal of diazinon. The authors note that the high removal efficiency of the pyrethroids was due to sedimentation of pesticide-bound particles, which was enhanced by an extended retention time and vegetation density (26). These studies demonstrate that a variety of plants are suitable for contaminant reduction when cultivated in a wetland treatment system. Often times, common macrophytes such as the common reed (Phragmites australis), bulrushes (Scirpus spp.), and cattails (Typha spp.) are selected for these systems as they can withstand varied water depths, reduce slope erosion, and are resistant to weed invasion (16). This practicality of use allows for easy implementation, especially at the early planning stages of remediation, as developers can utilize native and/or resident species as part of the larger habitat, increasing ecosystem services and preserving the surrounding communities. The inclusion of vegetation in a water quality treatment system in combination with bed sediments selected specifically for pesticides requiring mitigation, have been demonstrably more effective in reducing contamination levels of outflow water than non-vegetated systems.

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The Effect of Wetland Design on Biotic Factors Constructed wetlands are designed in a variety of ways, most notably with surface or subsurface flows. These mechanisms have similar processes yet work in different ways depending on the construction of the wetland. For instance, constructed wetlands with surface flows are successful due to the aerobic and anaerobic conditions that occur in the bottom layer of detritus where chemical breakdown occurs through hydrolysis and adsorption to sediments (12). In constructed wetlands utilizing subsurface flows, where conditions are typically anoxic and/or anaerobic due to the water’s continuing percolation below the surface, chemical degradation occurs mainly through plant uptake (12). In fact, microbial processes of pesticide degradation tend to be the main processes at work within these systems (16). This makes residence time, as noted earlier, essential as it allows time for these microbiological processes to occur. Vegetation also plays a role with these microbial processes as its presence in treatment systems aerates sediments and increases microbial activity; the decomposition of plant material provides organic carbon to microorganisms during the chemical breakdown process (16). The combination of vegetation, residence time, physicochemical properties of the chemicals themselves, and bed characteristics are all important traits of constructed treatment systems.

Biological Responses of Aquatic Organisms Although the success of pesticide load reductions in water quality treatment systems have been well demonstrated, there is less research focused on the biological response of aquatic organisms exposed to the treated outflow waters of these systems. A reduction in pesticide loads may not produce a commensurate reduction in toxicity in exposed biota, depending on the organism and the bioavailability of the contaminants in question. As discussed previously, the physicochemical properties of a pesticide can determine where in the environment it will partition, be it in the water, sediment, or plant tissue. As the partitioned phase will determine the bioavailable fraction of the toxicant, an appropriate selection of test species is necessary to accurately assess the effectiveness of the treatment system. The evaluation of aquatic organism responses to the treated outflow water in these systems is an important step to determine the efficacy of constructed wetlands, ponds, and vegetated ditches. Several researchers have used aquatic organisms in their studies with varied responses. Generally, toxicity is reduced as water passes through the treatment systems with accelerated amelioration in those systems where vegetation is present. However, a complete removal of toxicity is not always observed. Water Toxicity Pelagic organisms such as zooplankton and certain fish species are useful for determining the toxicity of pesticides partitioned into the water phase, such as organophosphates. These compounds, with their higher solubility and lower 441

adsorption rates, tend to favor the polar phase. Test species such as water fleas (Ceriodaphnia dubia), fathead minnows (Pimephales promelas), and amphipods (Hyalella azteca) are frequently used in toxicity tests to evaluate water quality in ambient monitoring and regulatory programs (41). As a result, these test species can be considered an effective tool for determining the biological efficacy of toxicity reduction in treatment systems. Researchers evaluating the effectiveness of three urban bioswales used C. dubia, H. azteca, and P. promelas toxicity tests across three storms and found that pesticide loading was consistently reduced across most contaminants, bioswales, and storms, whereas water toxicity varied by species (14). Toxicity was not detected with P. promelas or C. dubia species, whereas H. azteca toxicity was significantly reduced by passage through the bioswales. In-situ toxicity tests with P. promelas and H. azteca and in-lab toxicity tests with C. dubia were used to evaluate the efficacy of two vegetated agricultural ditches in Yolo County, California (42). These vegetated ditches drained water from an alfalfa field treated with chlorpyrifos and a tomato field treated with permethrin. While the passage of water through these ditches reduced chlorpyrifos runoff by 23% and permethrin runoff by 50%, there was only a modest reduction in organism toxicity with an approximate 15% increase in survival of H. azteca tested in-situ across both ditches (P. promelas were not negatively affected (42)). Treated outflow runoff with chlorpyrifos (alfalfa) was still toxic to both invertebrate species, and treated outflow runoff with permethrin (tomato) was still toxic to H. azteca. Thus, passage through a 389–402 m section of vegetated ditches was beneficial in reducing pesticide loads, but their remedial effects with respect to organism toxicity was less so (42). C. dubia and P. promelas were used in a study examining diazinon transport through a constructed wetland in the Mississippi Delta (6). Diazinon concentrations were reduced by 99% after 48 h; however, toxicity was still observed with C. dubia exposed to water collected from sites throughout the wetland. Survival of P. promelas was not affected, and the authors noted that the diazinon concentrations were below those expected to cause acute mortality and attributed the C. dubia toxicity to other contaminants or transformation products in the wetland (6). C. dubia were also used to determine the toxicity of a two-celled vegetated on-farm treatment system on the central coast of California (25). Significant mortality of C. dubia was observed. Results from associated toxicity identification evaluations (TIEs) and analytical chemistry indicated that the toxicity observed was due to a combination of the organophosphates diazinon (9.62 µg/L), dimethoate (8.40 µg/L), and chlorpyrifos (0.762 µg/L). These concentrations were high enough to cause the observed C. dubia mortality (25). P. promelas, C. dubia, and H. azteca were used in experiments evaluating outdoor wetland mesocosms in Mississippi in terms of their reduction of methyl parathion during a simulated storm event (30). In this experiment, toxicity was both spatially and temporally reduced. Survival for P. promelas was robust in the vegetated mesocosms at 3 h and 24 h; however significant mortality was observed with P. promelas at 3 h in the non-vegetated mesocosm. C. dubia survival was positively correlated with time, with survival rates ranging from 70 –100% in the vegetated mesocosm by 10 days. Significant C. dubia mortality was observed in 442

the non-vegetated mesocosm through 96 h with 100% mortality observed at all sites. By 10 days, survival ranged from 40–100% in the non-vegetated mesocosm with the lowest survival observed at the site closest to the inlet point. Samples taken for H. azteca were collected at 10 h. The survival rate in the vegetated mesocosm was significantly reduced at the 5 m site (closest to the inlet), but exceeded 92% in all other sites; in the non-vegetated mesocosm, survival rates ranged from 62.5–100%. It can be argued that toxicity was reduced as the water passed through the treatment system; however, residual toxicity was still observed in some instances (30). Conversely, Brogan and Relyea (42) examined the effect of Elodea canadensis density on malathion toxicity to Daphnia magna and observed that increases in D. magna survival were positively correlated with macrophyte densities. In addition, significant increases in estimated 48-h LC50 values corresponded to each increase in E. canadensis density, and in some instances water purification rates increased with macrophyte density (43). Later studies by these authors attributed this reduction to alkaline hydrolysis (39, 40), which was successful at reducing the malathion toxicity to D. magna. Sediment Toxicity Toxicity tests utilizing benthic organisms such as H. azteca and Chironomus spp. are commonly applied if sediment toxicity is suspected. Like C. dubia and P. promelas, these organisms are test species applied in standardized testing (44) in ambient monitoring programs, with increasing application in regulatory permitting. H. azteca and Chironomus dilutus (formally tentans) are used in sediment toxicity tests that evaluate organism survival and growth, as the life histories of these species bring them in contact with sediment resulting in higher sensitivities to hydrophobic compounds with affinities for sediment particles. In a two-cell vegetated treatment system used along the central California coast (25), observed sediment toxicity was extensive and of high magnitude to H. azteca in both vegetated cells. Chemical analyses detected cypermethrin and lambda-cyhalothrin in the first cell and chlorpyrifos and permethrin in the second cell in concentrations above those expected to cause acute effects, with permethrin measured at 10 times the level of H. azteca LC50 (25). Chironomus tentans were used to evaluate the toxicity of methyl parathion in sediment in vegetated and non-vegetated mesocosms (43) and exhibited relatively high survivability at all sites and time points, with the exception of the 10 m site location in the nonvegetated mesocosm, which had a significant reductions in survival rates. Growth was unaffected at all sites and time points in both vegetated and non-vegetated mesocosms which indicates the efficacy of the treatment system (30). Ten-day C. dilutus sediment tests were used to determine the efficacy of three urban bioswales in Salinas, California, in conjunction with water toxicity tests during two storm events. While two of the three bioswale inflow waters were significantly toxic to this species, the toxicity was significantly reduced after passing through the bioswales (14). Chironomus spp. were used in-situ in a constructed wetland in Cape Town, South Africa (32). The midge larvae were exposed for 24 h at the inlet and outlet of the wetland during a storm event. Prior to the start of runoff, the mortality rate 443

at both locations was less than 3%, but increased to 46% at the inlet and to 6% at the outlet during the runoff event, which demonstrates the wetland’s efficacy in reducing toxicity during storms. H. azteca were used in a toxicity study evaluating the remediation efficiencies of a Mississippi wetland system (33). In a simulated pyrethroid runoff event, toxicity was evaluated through the use of sediment, detrital, and 48-h water column tests. Toxicity was of a high magnitude in both water column and leaf-litter tests, at all sites at all time points, with the survival rate never exceeding 20%. In the sediment tests, toxicity was variable but reduced in comparison to the water and detritus phases indicating that pyrethroids partitioned to the solid phase are less available to biota (33). Similar results were observed in another study evaluating vegetated and non-vegetated wetlands in Mississippi with methyl parathion (45). In-situ tests with C. tentans indicated reductions in toxicity in both vegetated and non-vegetated mesocosms. Accelerated spatial reductions in toxicity were observed in the vegetated wetland compared to the non-vegetated wetland, indicating a positive effect of vegetation in toxicity remediation (45). In a different study, sediment toxicity tests with H. azteca were used to evaluate the efficacy of diazinon removal in a constructed wetland (46). Significant reductions in survival were observed at 8 and 48 h. However, survival increased significantly by 7 days and remained constant until the end of the study period. In this particular study, sediment bioassays were conducted in conjunction with water column tests, which exhibited significant mortality at all time points after 0 h in all locations, indicating the potential for diazinon to be more toxic in the water phase rather than the solid phase (46). Taken together, these studies illustrate how the partitioning capacity of pesticides affect their bioavailability to resident organisms present in water quality treatment systems. As expected, hydrophobic pesticides elicited toxicity in sediment tests, whereas hydrophilic pesticides had more adverse effects in the water column tests. However, pesticides bound to sediment particles were less bioavailable than those in the water phase although the potential for resuspension does exist. Further studies evaluating pesticide partition tendencies in conjunction with organism responses would help elucidate this further.

Potential Concerns for Future Research: Source-Sink Dynamics One remaining question regarding constructed water quality treatment systems is to what extent their efficacy is maintained through time. As described previously, treatment systems are designed to allow contaminants, many of which are hydrophobic, to adsorb to surfaces during storms. At longer time scales, natural processes degrade adsorbed contaminants. However, if the rate of contaminant degradation is slower than the accrual rate of contaminants, toxicity within the constructed waterbody may increase. Eventually, the efficacy of these systems for mitigating storm runoff may decline, and the substrate of the waterbody may become toxic. For instance, Jeppe et al. (47) demonstrated the toxicity of urban wetland sediments and their effects on the freshwater amphipod Austrochiltonia subtenuis. These researchers linked the toxicity observed in lab 444

experiments and the absence of field populations of A. subtenuis to the retention of bifenthrin in the urban wetland sediments at concentrations above the calculated LC50 for that species (47). These urban wetlands were successful at removing chemicals from storm runoff, but Jeppe et al. noted that their research raises concerns that the associated habitat and environmental services provided by the wetlands may be at risk. Sharley et al. (48) evaluated the sediment quality of 98 urban constructed wetlands using boosted regression trees analyses and demonstrated that land use has a significant effect on the likelihood of toxicity of the wetlands. Those wetlands that are >10% industrialized have significantly higher sediment-bound concentrations of trace metals than wetlands with little industrialization, and wetlands constructed in primarily residential or urban areas have significantly less risk of becoming polluted long-term (48). Of the limited research available, there are mixed results on the efficacy of treatment systems over time (49). There is some evidence of declining efficacy as treatment systems age (50, 52–54), but overall the topic appears under-studied (49), particularly regarding pesticides. Thus, two potentially fruitful areas of research would be to examine whether treatment systems maintain efficacy for pesticide removal over time and whether sediment toxicity increases over time.

Conclusion Water quality treatment systems, such as ponds, wetlands, and vegetated agricultural ditches, have been shown to be a cost-effective, successful mitigation strategy to reduce pesticide loads to receiving waters. Based on the available research, plant composition, mass, retention time, and sedimentation type are all important factors to consider when developing the best management practices for these types of systems. As demonstrated in this review chapter, contaminant concentrations are often reduced to levels below analytical detection limits, and in most cases organism responses are negatively correlated with pesticide concentrations. Although there is concern about the source/sink tendencies of these treatment ponds in the long term, these treatment processes are an effective way to mitigate pesticide loads in source waters. Moving forward, the most efficacy can be generated by these treatment processes if land uses are taken into account during the initial planning stages, so that the optimum remediation strategies can be used to focus on specific classes of pesticides and their physicochemical properties.

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