Rivers, lakes, and streams provide many recreational activities and benefits, as well as important ecosystem services such as nutrient cycling, wildlife habitat, and flood mitigation. With the increasing demand in urban and agricultural areas for freshwater, few options are available to ensure that aquatic systems maintain their respective ecohydrological requirements (Neubauer et al., 2008). Environmental applications of water reuse include river and wetland habitat creation and augmentation of existing water sources for the express purpose of improving conditions for aquatic biota. The Florida Everglades, for example, are at risk due to a decrease of incoming freshwater (see Box 8-1). For areas such as the Everglades and others, water reuse for ecological enhancement may be a beneficial option because reclaimed water could be used to augment streamflow, restore wetlands, and/or enhance water quality (Wintgens et al., 2008; Carey and Migliaccio, 2009). In addition to ecological benefits, there may also be economic benefits (e.g., increased tourism, hurricane protection) from such projects (Carvalho, 2007; Costanza et al., 2008; see also Chapter 9).
Reclaimed water may have potential for augmenting existing surface water systems and creating new habitats. In most instances, reclaimed water used for the purpose of ecological enhancement will meet or exceed local wastewater discharge standards. Nevertheless, the ecological risk of such planned applications needs to be considered to ensure that the level of risk to the environment is acceptable relative to the benefits. The level of acceptable ecological risk in these projects will likely vary between reuse scenarios; for example, the acceptable level of risk in a newly constructed wetland may be different than in a pristine system such as the Everglades. The level and cost of the assessment will also vary depending on the scenario.
Based on these considerations, the purpose of this chapter is to (1) present what is known about risks associated with the purposeful reuse of treated wastewater for habitat restoration and creation, (2) describe methods for assessing ecological risks from a historical and state-of-the-science perspective, and (3) recommend future research needs in the area of water reuse and ecological risk assessment.
As presented in Chapter 2, treated wastewater is routinely discharged to the nation’s rivers as part of the wastewater disposal process, with nearly 99 percent of wastewater discharges receiving secondary or greater treatment (see Table 2-1). The quality of reclaimed water used for ecological applications would be no lower than that of traditional wastewater discharge, and may be treated to higher levels. Therefore, available data on the ecological effects from the chemical, physical, and biological stressors in treated wastewater effluent discharged to rivers and lakes provide a worst case scenario of effects that could occur in ecological enhancement water reuse projects.
Typical wastewater discharges contain a mixture of microbes, inorganic chemicals, and organic chemicals, some of which may cause adverse ecological effects in
Proposed Reuse Projects to Expand Environmental Water Supply in the Everglades
The Comprehensive Everglades Restoration Plan (CERP) was envisioned as a multidecadal effort to achieve ecological restoration by reestablishing the hydrological characteristics of the historic Everglades ecosystem, where feasible, and to create a water system that simultaneously serves the needs of both the natural and human systems of South Florida (NRC, 2010). The conceptual plan (USACE/SFWMD, 1999) included 68 different project components focused on restoring the quantity, quality, timing, and distribution of water in the ecosystem. The largest component of the budget for this $13 billion project is devoted to water storage, including conventional surface water storage reservoirs, in-ground reservoirs, aquifer storage and recovery, and seepage management. To provide sufficient water supply to meet anticipated future environmental, urban, and agricultural water demands in South Florida, the comprehensive plan included two water reuse projects in Miami-Dade County, which together would treat more than 200 million gallons per day (MGD; 760 million m3/d). In the preliminary project concept, the reclaimed water would be used for aquifer recharge to enhance urban water supplies and reduce seepage out of the Everglades. Additionally, reclaimed water could be provided to Biscayne Bay National Park to help meet freshwater flows to support ecosystem needs. However, the plan acknowledged the high costs of such treatment to support ecological needs and noted that other potential sources of water would be investigated before water reuse was pursued.
Pilot projects were planned to assess the “cost effectiveness and environmental feasibility of applying reclaimed water to sensitive natural areas” and to “identify treatment targets consistent with preventing degradation to natural area,” among other objectives. A pilot plant was constructed by Miami-Dade County that included several different wastewater reclamation treatment trains (e.g., with and without reverse osmosis; ozone vs. ultraviolet/advanced oxidation processes), and trace organic chemical data were collected for several months. However, the pilot project was halted in 2011 before the planned toxicity testing was initiated because of general concern about the economic feasibility of the larger ecological restoration project (Jim Ferguson, Miami Dade County Water and Sewer Department, personal communication, 2011).
receiving water bodies. However, the level of toxicant exposure and dilution within the receiving systems are key considerations when assessing toxicity. The individual constituents may arise from industrial, household, or wastewater treatment plant applications. For instance, chlorine is often used as a disinfection chemical to reduce pathogen load and disease risk in wastewater. Low levels of chlorine may cause toxicity in the receiving stream or form chlorinated byproducts capable of causing ecotoxicity. Organic chemicals in wastewater have the potential to deplete the receiving aquatic system of oxygen, thus impacting aquatic life. Suspended solids from wastewater can block sunlight, thus reducing the photosynthetic capability of aquatic plants. Reduction in sunlight penetration may reduce plant life, as well as vertebrate and invertebrate populations. All of these stressors singularly or in combination may affect aquatic life, which includes macroinvertebrates, fish, plants, and amphibians (Sowers et al., 2009; Brix et al., 2010; Slye et al., 2011). Ecological assessments of wastewater effluent-dominated surface waters have shown that aquatic life can be sustained in these types of waters; however, site-specific factors may influence the aquatic life in various locations (Brooks et al., 2006; Slye et al., 2011).
Many studies associated with municipal effluents have been focused on standard measures of water quality, such as pH, temperature, total nitrogen and phosphorus, dissolved oxygen, and the impact of the effluent on the receiving system (Howard et al., 2004; Kumar and Reddy, 2009; Odjadjare and Okoh, 2010). Regulatory agencies, such as the U.S. Environmental Protection Agency (EPA), have developed guidance documents and criteria for many of these water quality parameters on a site-specific or ecoregion basis. Further, the EPA created the National Pollutant Discharge Elimination System to prevent aquatic life impacts associated with these traditional forms of wastewater pollution. As information on new classes of environmental contaminants arise, standard methods for assessing risk (e.g., whole effluent toxicity [WET] testing) may be unable to detect the subtle changes associated with these compounds. For instance, there have been recent reports of treated wastewater causing severe lesions and developmental alterations in amphibians, which are not common sentinel testing organisms in the WET testing paradigm (Sowers et al., 2009; Keel et al., 2010; Ruiz et al., 2010).
Because stressors may be different between each reuse scenario, basic information on the effects of potential ecological stressors in treated wastewater are described in this chapter.
Nitrogen and Phosphorus
Nutrients represent one of the historical problems with direct discharge of wastewater effluent, although the nutrient discharge concentrations are highly dependent on the type of wastewater treatment provided (Carey and Migliaccio, 2009; see Box 8-2). EPA has recently focused increasing attention to the impacts of nutrients on surface water ecosystems and has encouraged states to develop and adopt numeric nutrient criteria for nitrogen and phosphorus (EPA, 2011).1 Excess nutrients to an aquatic ecosystem can be problematic, because they cause an increase in the primary productivity of the ecosystem, known as eutrophication. Eutrophication can lead to changes in dissolved oxygen concentrations, algal blooms, decreases in submerged aquatic vegetation, and fish-kills. Increases in the limiting nutrient (i.e., the nutrient needed for plant growth but which typically occurs in small quantities) will accelerate eutrophication. Typical levels of nitrate in effluents receiving secondary treatment with disinfection are between 5 and 20 mg nitrogen (N)/L. Typical levels of phosphorus in effluents receiving conventional activated sludge (i.e., secondary) treatment are 4–10 mg/L, and these concentrations can be lowered to 1–2 mg/L with biological nutrient removal (BNR) (see Table 3-2).
Ammonia is particularly toxic to aquatic organisms, with the toxicity dependent on pH and temperature. The roles of pH and temperature relate to the amount of un-ionized ammonia (NH3) in the water body. The acute and chronic criteria for ammonia (pH 8 at 25°C) are 2.9 and 0.26 mg/L, respectively (EPA, 2009a). Typical levels of ammonia in secondary effluents with disinfection are 1–10 mg/L and 1–3 mg/L with BNR (Asano et al., 2007).
Trace metals (cadmium, copper, etc) are common regulated contaminants in wastewater discharges. The toxicity of metals in aquatic systems is complex and is often related to the amount of dissolved or free metal in the water. Water quality parameters, such as hardness, pH, and organic matter, can greatly affect toxicity. When considering copper, for instance, a low pH increases the most toxic form (i.e., Cu2+) of copper. Hardness and copper toxicity are inversely proportional, whereby elevated water hardness leads to decreased copper toxicity (Erickson et al., 1996). Organic matter forms complexes with copper and reduces toxicity (Hollis et al., 1997). EPA has national water quality guidelines to protect aquatic life for most metals, but site-specific parameters may need be considered for ecological applications of reclaimed water in sensitive ecosystems, particularly in areas with little dilution of the wastewater discharge in the ecosystem (EPA, 2009a).
Santa Clara Valley Water District (SCVWD) Stream Augmentation Project
SCVWD proposed a pilot project to augment the flows of Coyote Creek with advanced-treated reclaimed water from the San Jose/SCVWD treatment plant for the purpose of ecological enhancement. Reclaimed water would be discharged into Upper Silver Creek 2 km upstream from its confluence with Coyote Creek in San Jose and released from May to October at a flow rate of 1 to 2 cubic feet per second (cfs) (2,400 to 4,900 m3/d). Baseline studies were conducted prior to the project to monitor water quality parameters (e.g. nutrients, oxygen, temperature) and algal biomass (Hopkins et al., 2002). Hopkins et al. (2002) concluded that augmentation to Coyote Creek could result in increased nutrient and ammonia levels, as well as algal biomass. Analysis of advanced-treated wastewater (from treatment plants using dual-media filtration followed by disinfection by either chlorination or chloramination) indicated that it contained measurable levels of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) at total concentrations ≤ 470 ng/L (Plumlee et al., 2008). The bioaccumulation and biomagnifacation factors for PFOS and PFOA that were used in the ecological risk assessment of Coyote Creek were based on data obtained from the Great Lakes. Because the Great Lakes and Coyote Creek are disparate water bodies, there were higher levels of uncertainty in the analysis of the risks of PFOS and PFOA in Coyote Creek. Nonetheless, the detection of these chemicals placed this project on hold in an attempt to understand the meaning of these findings.
The impact of silver- and titanium-based nanoparticles in the aquatic environment is an emerging topic of research interest. Fabrega et al. (2011) reported that concentrations of silver nanoparticles as low as a few nanograms per liter can affect fish and invertebrates, although mechanisms of toxicity, nanoparticle fate in wastewater treatment and the environment, and
ecological risk in the environment remain poorly understood.
Changes in salinity may occur with the use of reclaimed water. Typical levels of salt (measured as total dissolved solids [TDS]) in effluents receiving secondary treatment with disinfection are 270–860 mg/L (Asano et al., 2007). Although the TDS of treated wastewater is not expected to be significantly greater than that of many surface waters, ecological applications should consider the TDS of the native water before introducing reclaimed water into existing ecosystems. Currently, no federal TDS aquatic life criterion exists (Soucek et al., 2011). However, site-specific criteria have been advocated. For example, in certain regions of Alaska, a TDS criterion of 500 mg/L has been suggested for periods of salmon spawning, while a TDS criterion of 1,500 mg/L has been suggested for nonspawning periods (Brix et al., 2010).
Temperature and Dissolved Oxygen
Changes in water temperature may be associated with the use of reclaimed water for environmental purposes. Temperature can influence aquatic community structure and productivity of microbes to fish. For instance, water temperature has been shown to influence factors that affect growth in aquatic organisms (e.g., metabolic rate, respiration), which may alter community structure and trophic interactions (i.e., predator-prey dynamics) within a water body (Sobral and Widdows, 1997; Abrahams et al., 2007; Hoekman, 2010). Further, temperature can alter aquatic habitat influencing species composition and biodiversity (Jones et al., 2004). Typically, the temperature of treated wastewater discharge is in the normal range of the receiving environment.
Dissolved oxygen is an important parameter for aquatic life and is related to various water quality parameters including temperature. As temperature increases in a water body, dissolved oxygen decreases. Dissolved oxygen can also be reduced by algal blooms spurred by high nutrient concentrations. National and site-specific dissolved oxygen criteria have been developed to protect aquatic life (EPA, 1986, 2000). For instance, dissolved oxygen acute mortality criteria for non-embryo/early-life-stage freshwater fish is 3 mg/L (EPA, 1986). An increase in organism mortality and/or growth, in addition to changes in species composition, may be observed if dissolved oxygen levels fall below the developed criterion.
Boron, in the form of borates, is released into the environment from anthropogenic sources (i.e., wastewater treatment plant discharge), as well as from weathering of sedimentary rocks (Frick, 1985; Howe, 1998; Dethloff et al., 2009; see also Chapter 3). Boron in reclaimed water is generally less than 0.5 mg/L, while concentrations in surface waters are generally ≤1.0 mg/L (Butterwick et al., 1989, Asano et al., 2007). Fish, amphibian, invertebrate, and plant effects associated with boron exposure generally occur in the low to mid milligram-per-liter range (Powell et al., 1997; Howe, 1998; Laposata and Dunson, 1998; Davis et al., 2002; Dethloff et al., 2009). The concentration of boron that affects fish, amphibians, invertebrates, and plants, including landscape plants, are typically above the concentrations observed in reclaimed water (Wu and Dodge, 2005).
Trace Organic Chemicals
As discussed in Chapter 3, trace organic contaminants (e.g., pharmaceuticals and personal care products, and flame retardants) have been detected in municipal wastewater effluent and in the nation’s surface waters, creating concerns for both human and aquatic systems (Daughton and Ternes, 1999; Kolpin et al., 2002; see also Appendix A). The presence of these chemicals (e.g., carbamazepine, triclosan, brominated diphenyl ethers) is associated with normal human use of trace organic compounds. When considering the sensitivity of human and aquatic organisms to trace organic compounds detected in reclaimed water, it is important to note that aquatic organisms are generally as sensitive or more sensitive than humans to these chemicals (Table 8-1). Further, the potential toxicity for many of these compounds may be heavily influenced by water quality parameters (e.g., pH), thus complicating the risk assessment process described below (Valenti et al.,
TABLE 8-1 Comparison of Human Monitoring Trigger Levels for Potable Reuse and Aquatic Predicted No Effect Concentrations for Selected Chemicals in Reclaimed Water
|Chemical||Example Occurrence in Secondary/Advanced-Treated Water (ng/L)a,b||Human Monitoring Trigger Levels (ng/L)b||Predicted No Effect Concentrations (PNECs) for Aquatic Ecosystems (ng/L)c|
aDefined as the 90th percentile average occurrence in secondary or advanced-treated wastewater, representative of water quality required by California’s Title 22 regulations for urban irrigation (Drewes et al., 2010).
bCalculated from risk-based acceptable daily intakes (ADIs; see Chapter 6 and Box 6-5) in the California State Water Resources Control Board (SWRCB) Science Advisory Panel Report (Drewes et al., 2010).
cDerived by methods outlined under Single Chemical Risk Assessment in this chapter using Brooks et al. (2003); Cleuvers (2003); EPA (2005b); Beach et al. (2006); Costanzo et al. (2007); Caldwell et al. (2008); Capdevielle et al. (2008); Küster et al. (2010).
Natural and synthetic chemicals have the ability to mimic endogenous hormones and alter the endocrine system in aquatic organisms. Chemicals that alter the endocrine system may ultimately cause reproductive dysfunction and population-level decline of organisms. While there are a myriad of chemicals that may interact and disrupt the endocrine system (e.g., bisphenol-A, cadmium), one of the best-studied endocrine disruptors is the birth control contraceptive 17-α ethinyl estradiol (EE2; see Box 8-3; Lange et al., 2001; Maunder et al., 2007). The science is still developing with respect to biological assay for rapid detection of endocrine disruptors (discussed later in this chapter).
One of the current limitations in evaluating the ecological risk of trace organics relates to the amount of ecotoxicity data available. For many trace organics, few data are available to make a reliable assessment of risk. With respect to pharmaceuticals, for instance, the recent improvement in the European Medicines Agency guidelines for the environmental risk assessment of pharmaceuticals should reduce these data gaps. To date, few field studies have evaluated the impact that water reuse and associated trace organics may have on the environment. In addition, few studies are available linking the relationship of laboratory endocrine and reproductive responses to effects in natural systems. Although endocrine disruption is a major scientific research thrust, the detection and risk of endocrine disruptors may be different depending on the reuse scenario. Atkinson et al. (2009) and Slye et al. (2011) investigated surfactants along a 100-mile gradient on the Trinity River spanning the Dallas-Fort Worth Metroplex to Palestine, Texas, where in some areas in the summer months >95 percent of the flow comes from municipal wastewater effluent from multiple inputs. No risk to aquatic organisms could be attributed to surfactants associated with this effluent-dominated river. These two studies represent examples for how geographic information systems (GIS) and chemical and biological monitoring can be incorporated to evaluate an ecosystem dominated with effluent.
17-α Ethinyl Estradiol: A Case in Ecological Endocrine Disruption
Natural and synthetic chemicals have the ability to mimic endogenous hormones, alter the endocrine system, and lead to reproductive dysfunction in aquatic organisms. In particular, numerous studies have focused on the toxicity associated with the birth control contraceptive 17-α ethinyl estradiol (EE2) in fish (Lange et al., 2001). Fish reproduction is the most sensitive end point associated with EE2, with a laboratory predicted no effect concentration of 0.35 ng/L (Caldwell et al., 2008). A whole Canadian lake study was conducted with EE2, where lakes treated with 5–6 ng/L EE2 caused population declines in fathead minnows and other organisms (Kidd et al., 2007). Although these data supported the laboratory findings of EE2, the levels are higher than those normally expected in the environment (Hannah et al., 2009).
Many questions remain about the risk of trace organic chemicals to the environment because of the lack of associated environmental fate and effects informa-
tion. Historically, most chemicals have been tested one chemical at a time. However, mixtures of bioactive trace organic chemicals are often present in water, for which new techniques need to be developed and refined to better understand their risk to the environment. As described in Chapter 6, a mixture of chemicals may result in toxicity that is equal to, less than, or greater than the sum total of the toxicity of the individual components. Using chemicals with the same mode of action (e.g., environmental estrogens), it has been demonstrated that the combined toxicity could be predicted based on the toxicity of the individual chemicals (Thorpe et al., 2003). However, it is much more difficult to model mixture responses when the modes of action of the individual chemicals within the mixture are different. This section discusses historical as well as newer techniques that can be used to assess ecological risk even in the absence of chemical-specific data.
The ecological risk assessment (ERA) process is adapted from and is not dissimilar to the human health risk assessment process described in Chapter 6. An ERA consists of four phases: problem formulation, characterization of exposure, characterization of effects, and risk characterization (EPA, 1998a). Following the risk characterization phase, the information can be used by risk managers to determine the course of action for the particular action or question. Furthermore, the data can be used to prioritize which chemicals are of greatest concern and deserve further research.
An ERA is typically conducted to evaluate the likelihood of adverse effects in the environment associated with exposure to chemical, biological, or physical stressors (EPA, 1998a). In addition, the ERA is designed to accommodate mixtures of stressors on aquatic life and habitat. In this respect, it can be used as the foundation for determining potential adverse effects of using reclaimed water for ecological purposes. Key factors in the ERA are the end point to be evaluated (e.g., habitat, endangered species) and the sensitivity of the ecosystem, which may be different for each reuse scenario.
Once the end points of concern have been identified, an understanding of the magnitude of exposure and response to the stressors will ultimately determine the level of risk. One of the first and fastest approaches will be to conduct a literature evaluation based on the stressors of interest to determine if aquatic toxicity or water quality criterion data are available. If data are available, the assessment may be done without further testing. However, if no data are available for the contaminants or end points of interest, then testing may be necessary (described in the following sections).
Once risk exposure and effects analysis are completed, a predicted environmental concentration (PEC) and a predicted no-effect concentration (PNEC) for the stressors will be available. The PEC/PNEC ratio will determine the risk associated with the stressors. If the PEC/PNEC ratio is ≥1, then a risk exists to the environment. A ratio that is < 1 suggests that the potential risk to the environment is low. If adequate data are available to calculate a species sensitivity distribution, a more extensive probabilistic environmental risk assessment approach may be used to estimate the likelihood and the extent of adverse effects occurring (Verdonck et al., 2003).
Single Chemical Risk Assessment
The environmental safety of chemicals is most often assessed on an individual basis, irrespective of the fact that in the aqueous environment there is a mixture of chemicals. In a single chemical assessment scheme, a no-observed effect concentration (NOEC), the lowest-observed effect concentration (LOEC), and/or an effective concentration (EC)2 will be derived in a series of laboratory studies with fish, algae, and invertebrates. Typically, these studies focus on higher level end points such as survival, growth, and reproduction of the test organisms. Both the EPA and the Organisation for Economic Co-operation and Development have defined methodological protocols to conduct these studies (EPA, 2010d; OECD, 2011). In the case of fish and invertebrates, the first studies that are conducted are acute or short-term assays, which are ≤96 hours and focus on the concentration that causes 50 percent mortality in the test organisms (lethal concentration 50 percent; LC50). Following these initial mortality studies, chronic reproduction and growth studies (≥21 days) are often conducted with fish and invertebrates. Once
2 Effective concentration (ECx) is the concentration of a toxicant that produces X percent of the maximum physiological response. For example, an EC50 reflects the concentration that produces half of the maximum physiological response after a specified exposure time. NOEC and LOEC are the ecological risk parallels of the lowest observed adverse effect level (LOAELs) and no observed adverse effect level (NOAEL) for human health risk assessment as defined in Box 6-5.
a NOEC is obtained, a safety/uncertainty factor is applied, accounting for species and exposure differences, to derive the PNEC. These data can be useful in the assessment of reclaimed water because one can compare the PNEC values to the concentrations measured in the water (see Table 8-1).
Note that in this single-chemical assessment scheme, the data are usually obtained in controlled laboratory settings and do not focus on community and ecosystem attributes (e.g., nutrient cycling). Once the chemical is released into the environment, there may be interactions with other substances, such as dissolved organic carbon, that may modulate its toxicity, in addition to potential interaction with other chemical contaminants. Laboratory studies do not account for mixture interactions, where these interactions may lead to additive or greater-than-additive toxicity. Although laboratory studies can be conducted to evaluate mixtures, it is unreasonable to assume that every realistic mixture component can be studied. These sources of uncertainty with respect to potential toxicity need to be recognized. Safety or uncertainty factors can be applied with the risk assessment process to account for mixture scenarios.
A single chemical risk assessment approach is used for most trace organic chemicals, including pharmaceuticals and personal care products (see Box 8-4 for an example application of this method). In this single-chemical approach, molecular, biochemical, and physiological end points are not utilized because they are often difficult to link to higher level effects (e.g., survival, growth, and reproduction).
In the case where a PEC/PNEC ratio is >1 and the body of information suggests that a chemical may adversely affect the environment, controlled outdoor pond or stream mesocosm experiments will help to better predict its impact on populations, communities, and ecosystems. This approach was used by Kidd et al. (2007) to demonstrate that 17α-ethinyl estradiol can cause population- and community-level impacts at environmentally relevant concentrations. The benefit of these studies is that one can measure end points (e.g., species density, species richness, nutrient cycling) in a controlled exposure scenario. In addition, mesocosm experiments have been conducted where a single contaminant has been introduced into a complex effluent to evaluate potential mixture interactions (Brooks et al., 2004).
Assessing the Ecological Risk of Carbamazepine: Two Approaches
Carbamazepine is an antiepileptic drug marketed in North America and Europe. Approximately 17 percent of an ingested dose is eliminated from humans as nonmetabolized carbamazepine. Using the traditional ecological risk assessment approach, the potential ecological risk can be estimated. The predicted environmental concentration (PEC) for carbamazepine based on modeling approaches is estimated to be ≤0.658 μg/L (Cunningham et al., 2010), while the 90th percentile occurrence in reclaimed water meeting California’s urban irrigation requirements (Title 22) is <0.400 μg/L (CSWRCB, 2010). To determine the PNEC, a wide array of available ecotoxicity data are assessed. The 96-hr LC50 values for Daphnia magna (invertebrate) and Japanese medaka (fish) are 76 and >100 mg/L, respectively (Kim et al., 2007). The 72-hr algal effective concentration 50 percent (EC50; the concentration that produces a response halfway between the baseline and the maximum response) is 74 mg/L, while the 7-day duckweed growth EC50 was 25 mg/L (Cleuvers, 2003). Duckweed growth appears to be the most sensitive end point and because only acute data are available, a safety factor of 1,000 is applied to the EC50 value. The resultant PNEC is 25 μg/L. Performing the risk quotient calculation (PEC/PNEC), the risk is <0.03 indicating that adverse environmental effects are not expected in surface waters augmented with reclaimed water.
The potential environmental risk can also be estimated using the mammalian model screening approach (Huggett et al., 2003). This approach represents a rapid screening method to estimate ecological risks based on the large quantity of mammalian effects data available for pharmaceuticals. Considering the predicted environmental concentration of carbamazepine of ≤0.658 μg/L and an octanol water coefficient (log Kow) of 1.68 (Cunningham et al., 2010), the resultant fish plasma concentration is calculated as 2.6 μg/L. The human therapeutic plasma concentration for carbamazepine is 2,170 μg/L after a single administration (Revankar et al., 1999). The calculated fish plasma concentration at estimated environmental concentrations of carbemazepine is much less than the plasma concentration known to exhibit effects in humans. Therefore, the environmental risk to fish is estimated to be low. The mammalian model screening approach yielded the same conclusion as the traditional risk assessment approach, suggesting its utility in rapid screening of environmental risk associated with pharmaceuticals.
An immense amount of mammalian pharmaceutical data (e.g., toxicological impacts, pharmacokinetics, and metabolism, often in multiple organisms) may be helpful in screening potential environmental risks asso-
ciated with pharmaceuticals (Lange and Dietrich, 2002; Huggett et al., 2003, 2004). A recent analysis indicated that many human pharmaceutical therapeutic targets are present in fish (Gunnarson et al., 2008). If the therapeutic targets are similar across species, then the internal concentrations that elicit effects across species may be similar. Knowing the PEC of a pharmaceutical and its relative hydrophobicity (or aversion to water, as measured by the octanol-water coefficient, Kow), a fish plasma concentration of that pharmaceutical may be calculated. This value can then be compared to the human therapeutic plasma concentration (HTPC), which is the concentration of that drug in plasma known to cause an effect. If the fish plasma concentration exceeds the plasma concentration known to cause biological effects in humans, then the concentration of the drug in the water should be suspected of causing an ecological effect. This model can quickly help prioritize ecological risk associated with pharmaceuticals and identify specific drugs that should undergo further testing prior to ecological reuse applications (Box 8-4) (Huggett et al., 2003, 2004; Schreiber et al., 2011).
Bioconcentration and Bioaccumulation
Since the publication of Silent Spring by Rachel Carson in 1962, the bioaccumulation of chemicals in the environment has received growing attention. Bioconcentration has traditionally been defined as the accumulation of chemical substances from aquatic environments through nondietary routes, whereas bioaccumulation is the accumulation from nondietary and dietary routes (Barron, 1990). EPA has established criteria where a bioconcentration factor (BCF) or a bioaccumulation factor (BAF) > 1,000 (i.e., concentration in organisms 1,000× greater than water or food) must undergo further testing. Substances with a BCF or BAF >5,000 may be banned from commerce (Moss and Boethling, 1999) (Box 8-5). Several studies have shown a relationship between BCF and KOW (Barron, 1990), where a log KOW > 3 requires additional consideration. Both laboratory and field studies at multiple trophic levels (e.g., fish, birds) can indicate if a chemical is potentially bioaccumulated or bioconcentrated, although field measurements may be needed to confirm laboratory findings (OECD, 1996; Weisbrod et al., 2009).
Bioconcentration and Bioaccumulation of Perfluorooctane sulfonate (PFOS)
PFOS has multiple commercial uses (e.g., stain repellant) and has been detected in wastewater and reclaimed water (Plumlee et al., 2008). The log Kow for PFOS is 4.4, and laboratory fish BCF values range from 210 to 5,400, which indicate that this substance is potentially bioaccumulative (Martin et al. 2003; EA, 2004; Ankley et al., 2005). Multiple field studies have measured concentrations of PFOS in invertebrates and fish at concentrations greater than that in the surrounding environment (Kannan et al., 2005; Li et al., 2008). Concentrations of PFOS have also been measured in eagles and mink from the Great Lakes region at concentrations 5–10 times greater than in their respective prey items (Kannan et al., 2005). Given that PFOS has been measured in reclaimed water, these data indicate that PFOS has the potential to move through the food chain in areas where reclaimed water is being used for environmental enhancement. The major U.S. manufacturer of PFOS has announced a voluntary phase-out of PFOS from commerce (EA, 2004).
Effluent Toxicity Testing and Monitoring
A number of toxicity testing and biomonitoring methods are available to assess the ecological effects of reclaimed water for ecological applications. These can be divided into conventional, state-of-the-science, and blended approaches.
Conventional Approaches: Whole Effluent Toxicity Tests
The WET testing program in the United States was implemented to protect water bodies from point-source municipal and industrial discharges (Heber et al., 1996). WET programs for wastewater facilities typically consist of whole-effluent bioassays to determine whether the discharges are affecting the receiving waters (Heber et al., 1996). Typical WET laboratory bioassays include acute invertebrate and fish survival studies, subchronic fish growth studies, and chronic invertebrate reproduction studies. These tests can also be conducted to determine ecological responses to a single contaminant or specific mixtures. The typical duration for most of these studies is <7 days. Field assessments
of invertebrate and/or fish population and community structure can also be part of WET programs, but these assessments are not as frequent as laboratory testing.
State of the Science
While traditional ecotoxicology has focused on survival, growth, and reproduction as the main determinants of risk (e.g., WET testing), knowledge regarding the toxic modes of action (i.e., how the chemicals manifest their toxicity) has expanded available toxicity testing alternatives, including in vivo biomarkers or in vitro bioassays. These in vivo and in vitro markers may be specific or nonspecific for a class of chemicals.
In the past several decades, researchers have discovered that chemicals in the environment may interact with the normal estrogen, androgen, and thyroid signaling pathways in aquatic organisms (i.e., endocrine disruption) (Desbrow et al., 1998; Rodgers-Gray et al., 2001; Sumpter and Johnson, 2008). Through in vitro and in vivo screening of wastewater effluents (primary, secondary, and advanced secondary), researchers discovered that chemicals can interact directly with hormone receptors (e.g., estrogen receptors) and that these chemicals can induce changes in the fish egg yolk precursor vitellogenin (Desbrow et al., 1998). Desbrow et al. (1998) were unable to identify a relationship between the various wastewater treatment effluents studied (including primary, activated sludge, percolating filters, and sand filters) and vitellogenin production. From this knowledge, the yeast estrogen (YES) and yeast androgen receptor assays were developed for screening purposes (Arnold et al., 1996). These assays investigate the binding of aqueous chemicals to the estrogen or androgen receptors in yeast cells via colorimetric measurements. Ultimately, researchers can determine the extent to which estrogenic or androgenic chemicals are present in water. For instance, Holmes et al. (2010) utilized the YES assays to demonstrate a 97 percent reduction in total estrogenic activity in a reclaimed water treatment system that utilizes stabilization lagoons followed by coagulation, dissolved air flotation/filtration, and chlorination.
Vitellogenin production is directly linked to stimulation of the estrogen receptor. Circulating 17β-estradiol in female fish stimulates the production of vitellogenin in the liver, where it is released into the blood for incorporation in eggs. Chemicals that act as estrogen mimics (e.g., nonylphenol) increase vitellogenin production in fish, especially in male fish which only produce small quantities under normal conditions. Vitellogenin production, in either whole fish or liver cells, can therefore be used to evaluate estrogen content in municipal effluent, surface waters, and reclaimed waters. Filby et al. (2010) utilized vitellogenin as the primary method to determine the extent of estrogen content reduced by various wastewater treatment technologies.
Another promising nonspecific approach is through the use of gene expression profiling. Fish or other aquatic organisms are exposed to the water of interest, and the differential regulation of genes in the liver or gonad is determined (Garcia-Reyero et al., 2008). The analysis can help determine which biological pathways and processes, if any, are being altered by the water sample. Efforts are currently under way to bridge changes in biological pathways to adverse outcomes (termed adverse outcome pathways) at higher levels of biological organization, as well as develop genomic fingerprints for individual and chemical-specific classes (Kramer et al., 2011). An understanding of pathway data may be useful in developing new in vitro screening methodologies for chemicals of interest.
Conventional testing methodologies (e.g., WET) focus on higher level biological end points (i.e., growth, survival, reproduction). Research with endocrine-disrupting chemicals demonstrates that some of these methodologies (e.g., invertebrate reproduction) may not be sensitive enough to detect subtle biological changes that may take months or years to generate, while other responses (e.g., fish reproduction) offer more sensitive end points (Länge et al. 2001). The yeast screening, vitellogenin, and gene profiling assays offer the ability to generate screening-level biological data quickly to determine the presence and/or the relative levels of biologically active compounds in the matrix of interest. However, there is a need for assay standardization and training in order to achieve reliable results. There is also potential with some of these
assays (e.g., YES) for false-positive or false-negative results. Further, it should be recognized that at this time there is no direct link to higher level measurements (e.g., reproduction). Neither binding of a chemical to a receptor, induction of vitellogenin, nor changes in gene expression are conclusive of a population effect. They do, however, strongly suggest that more research is needed.
Because of the advantages and shortcomings with each conventional and state-of-the-science methodology, researchers are utilizing a blended approach incorporating both methodologies (Steinberg et al., 2008). Deng et al. (2008) utilized an online, flow-through fish exposure system with reproductive, endocrine (vitellogenin), and other end points to assess the ecological effects of shallow groundwater recharged by reclaimed water in the Santa Ana River Basin, California. The advantage of using a blended monitoring system is that one can achieve the rapid screening-level data associated with the newer assays, as well examine higher level end points.
The difference in ecological risk analysis using conventional vs. more state-of-the-science techniques is evident when one considers nonylphenol (Box 8-6). For nonylphenol, EPA developed ambient water criteria using conventional toxicity testing methods, while the European Union utilized new scientific methods (Box 8-6).
Currently, few studies have documented the environmental risks associated with the purposeful use of reclaimed water for ecological enhancement. Water reuse for the purpose of ecological enhancement is a relatively new and promising area of investigation, but few projects have been completed and the committee was unable to find any published research in the peer-reviewed literature investigating potential ecological effects at these sites. As environmental enhancement projects with reclaimed water increase in number and scope, the amount of research conducted with respect to ecological risk should also increase, so that the potential benefits and any issues associated with the reuse application can be identified.
The ecological risk issues and stressors in ecological enhancement projects are not expected to exceed those encountered with the normal surface water discharge of municipal wastewater. The most probable ecological stressors include nutrients and trace organic chemicals, although stressors could also include temperature and salinity under some circumstances. For some of these potential stressors (e.g., nutrients) there is quite a bit known about potential ecological impacts associated with exposure. Based on the available science, there is no reason to believe that the use of reclaimed water for environmental enhancement purposes would pose greater impacts than those already occurring in many of the nation’s surface waters impacted by wastewater discharge. Further, the presence of contaminants and potential ecological impacts may be lower if additional levels of treatment (e.g., nutrient removal, ozone) are applied.
Water Quality Criterion for Nonylphenol: United States vs. European Union
Nonylphenol is a frequently detected wastewater contaminant, most commonly used to produce nonionic surfactants. In 1998, 104 million kg of nonylphenol was produced in the United States (Harvilicz 1999). EPA has established ambient water quality criteria for nonylphenol in both saline and freshwater systems. The acute and chronic freshwater quality criteria are 28 and 6.6 μg/L, respectively, while the acute and chronic saltwater criteria are 7 and 1.7 μg/L, respectively Aquatic organism survival, growth, and reproduction end points were used to establish these criteria. Although nonylphenol has been demonstrated to cause estrogenicity in aquatic organisms (e.g., causes fish to produce vitellogenin), these data do not meet the acceptability requirements for water quality criteria by the EPA (EPA, 2005b). Therefore, these data were not utilized in establishing the criteria. In contrast, the European Union has restricted marketing and use of nonylphenol based in part on the potential for nonylphenol to be an estrogenic substance. The European Union risk assessment for nonylphenol cited a PNEC of 0.33 μg/L, based on a long-term algal study. Further, the resultant nonylphenol PEC/PNEC ratio was determined to be 1.8 (European Union, 2002).
Trace organic chemicals have raised some concerns with ecological enhancement projects, because aquatic organisms can be more sensitive to trace organic chemicals than humans. Although other stressors are well understood and treatment systems can
be developed to reduce their concentrations to acceptable levels, less is known about the ecological effects of trace organic chemicals, including pharmaceuticals and personal care products. Endocrine disruption has been, and will likely continue to be, a scientific research area and concern. More data are needed to link population level effects in natural aquatic systems to laboratory observations.
Sensitive ecosystems may necessitate more rigorous analysis of ecological risks before proceeding with ecological enhancement projects with reclaimed water. Although conventional methods (e.g., WET) of monitoring can be used, newer, more rapid and sensitive methods of biological screening (e.g., YES) are available. However, the limitations of these assays should be recognized, and as the science develops, these limitations will likely be reduced. Site-specific considerations (e.g., species present, habitat, geology) and a priori knowledge regarding specific contaminants of concern (e.g., endocrine disruptors) may suggest a more sophisticated testing program, involving field-based testing combined with lab-based bioassays.