The presence of emerging contaminants in municipal wastewaters, particularly endocrine-disrupting compounds such as oestrogenic substances, has been the focus of much public concern and scientific attention in recent years. Due to the scientific uncertainty still surrounding their effects, the Precautionary Principle could be invoked for the interim management of potential risks. Therefore, precautionary prevention risk-management measures could be employed to reduce human exposure to the compounds of concern. Steroid oestrogens are generally recognized as the most significant oestrogenically active substances in domestic sewage effluent. As a result, the UK Environment Agency has championed a ‘Demonstration Programme’ to investigate the potential for removal of steroid oestrogens and alkylphenol ethoxylates during sewage treatment. Ecological and human health risks are interdependent, and ecological injuries may result in increased human exposures to contaminants or other stressors. In this context of limiting exposure to potential contaminants, examining the relative contribution of various compounds and pathways should be taken into account when identifying effective risk-management measures. In addition, the explicit use of ecological objectives within the scope of the implementation of the EU Water Framework Directive poses new challenges and necessitates the development of ecosystem-based decision tools. This paper addresses some of these issues and proposes a species sensitivity distribution approach to support the decision-making process related to the need and implications of sewage treatment work upgrade as risk-management measures to the presence of oestrogenic compounds in sewage effluent.
News reports about the potential effects of ‘chemical cocktails’ or ‘gender-bending chemicals’ on unborn male foetuses, particularly related to the presence of female hormones in tap water, are now regularly being printed in some sections of the general media. This has partly been driven by advances in analytical chemistry enabling the detection of ever lower levels of contaminants in matrices previously considered uncontaminated. Furthermore, in recent years, rapid advances of biochemical sciences and technology have resulted in the development of bioassay techniques that can contribute invaluable information regarding toxicity mechanisms at the cellular and molecular level, thereby also raising new possibilities of injury. The extrapolation of such information to predict effects in an intact individual organism for the purpose of risk assessment is however still in its infancy (Gundert-Remy et al. 2005). Moreover, as emissions of conventional priority pollutants have been substantially reduced in most developed countries through the adoption of appropriate legal measures and the elimination of the dominant pollution sources, focus has switched to compounds present in lower concentrations and that have only recently been thought of as pollutants (Birkett & Lester 2003). Certain chemicals are thought to be able to interfere with the endocrine system of a wide range of organisms and such disruption may result in adverse effects. These substances, collectively referred to as endocrine-disrupting chemicals (EDCs), have been the focus of much scientific attention in recent years. The interaction of multiple agents with the host’s endocrine system could result in a wide spectrum of potential adverse health effects and much scientific uncertainty remains regarding the risks from EDCs. Due to the scientific uncertainty surrounding them, the Precautionary Principle (PP) could be invoked for the interim management of potential risks of exposure to such substances.
The PP was enshrined into European Law in 1992 when the Maastricht Treaty modified Article 130r of the treaty establishing the European Economic Community, and in just over a decade has also been included in several international environmental agreements (Marchant 2003). The European Environment Agency undertook a review of 100 years of environmental management (European Environment Agency 2001). The ‘Late Lessons’ report provided the following working definition of the PP that was improved subsequent to further discussions and legal developments:
The Precautionary Principle provides justification for public policy actions in situations of scientific complexity, uncertainty and ignorance, where there may be a need to act in order to avoid or reduce potentially serious or irreversible threats to health or the environment, using an appropriate level of scientific evidence, and taking into account the likely pros and cons of action and inaction.
(Gee 2006, p. 154)
The report concluded that there were many lessons that should have been apparent much earlier, and that, had Precautionary Actions (PA) been taken when problems first became evident, long-term consequences and costs could have been avoided (Burger 2003). If the level of acceptable risk is ultimately a societal choice, such choice should nonetheless be informed by science, while PAs also need to be justified by a sufficiency of evidence (Gee 2006). While there are situations where potential risks clearly exceed benefits and vice versa, there is a large grey area in which science alone cannot determine policy (Kriebel et al. 2001). The question is not whether the PP itself is appropriate, but whether the proposed PA is justified and proportional to the purported risk (Rogers 2003).
Multi-criteria frameworks are particularly useful in environmental decision problems where the temporal scale adds to the uncertainty and intrinsic complexity of the issue (Balasubramaniam & Voulvoulis 2005). In the context of the application of the PP to the hazards posed by EDCs, different states of knowledge justify different types of PAs (European Environment Agency 2001). Moreover, the interest incited by the weight-of-evidence approach originally developed in the International Programme on Chemical Safety Global Assessment of the State-of-the-Science on Endocrine Disruptors (International Programme on Chemical Safety 2002) highlighted the need to develop criteria and methods for the application of such frameworks to complex toxicological issues (International Programme on Chemical Safety 2006). In this context, a simple multi-criteria framework by which to assess the current state of knowledge used to justify PAs was developed (Martin et al. 2007). Much of the controversy surrounding the association between environmental EDCs and the various health endpoints considered is probably related to the lack of convincing evidence related to the strength of such an association. While, for all the human health endpoints considered, concern arises from the analogy either with hormonally active pharmaceuticals or risk factors related to endogenous hormone levels, there is little evidence that environmental levels of weakly active compounds elicit similar effects. While 10 years of further research on the potential effects of endocrine disrupters on male reproductive health have provided some clues regarding the aetiology and mechanism of conditions such as hypospadias, cryptorchidism and testicular cancer, a meta-analysis of three of the endpoints thought to comprise the testicular dysgenesis syndrome was unable to produce quantitative evidence that such effects were associated with environmental oestrogens (Martin et al. 2008). If the benefits of PA are therefore uncertain, such measures should nonetheless be cost-effective, and balance potential risks with an evaluation of the potential social, environmental and economical costs. Such uncertainty would justify precautionary prevention risk-management measures, such as limiting human exposure to the compounds of concern.
Methodologies for human health and ecological risk assessment are developed independently. However, the need for a more integrated, holistic approach to risk assessment is increasingly being recognized (Suter et al. 2005). Ecological and human health risks are interdependent, humans depend on nature for food, water purification, hydrological regulation and ecosystem services, which can be impeded by the effects of toxic chemicals. In addition, ecological injuries may result in increased human exposures to contaminants or other stressors. Integrated exposure assessment has been put forward as a way to incorporate multiple sources, multiple exposure pathways, multiple direct effects and the possibility of indirect effects. In the context of limiting exposure to potential contaminants, examining the relative contribution of various pathways can help identify the most effective risk-management measure, whether to protect human health or sensitive biota.
One of the most comprehensive studies of endocrine disruption in wildlife is that on the impact of steroid oestrogens and oestrogenic chemicals on British fish. This research showed that sexual disruption in fish was widespread throughout rivers in England and Wales, and the severity of the condition correlated with discharge of sewage effluent (Jobling et al. 1998). Effluents are complex mixtures of organic and inorganic substances and a wide variety of structurally diverse natural and synthetic chemicals are potentially oestrogenic. Steroid oestrogens are generally recognized as the most significant oestrogenically active substances in domestic sewage effluent. This does not, however, rule out the other weakly oestrogenic compounds from contributing to the overall effect, and it is widely accepted that such compounds act additively (Gross-Sorokin et al. 2006). As a result, the UK Environment Agency has championed the inclusion of a $40 million ‘Demonstration Programme’ in the Water Industry 4th Asset Management Plan settlement, to investigate the potential for removal of steroid oestrogens and alkylphenol ethoxylates (APEOs) and their derivatives in sewage treatment works (STWs; Burke 2004).
Water pollution has traditionally been regulated using chemical objectives thought to be conservative enough to offer appropriate protection to the aquatic environment. The recognition of the need to safeguard the ‘ecological quality’ of waters led to the revision and integration of the existing fragmented water-related legislation into an integrated framework taking an ecosystem-based approach, the European Water Framework Directive (WFD) (European Community 2000; Kallis & Butler 2001). The explicit use of ecological objectives poses new challenges and necessitates the development of ecosystem-based decision tools. Initial attempts to answer some of these questions and challenges, and suggest a species sensitivity distribution (SSD) method to appraise the need and implications of STW upgrade as risk-management measures to the presence of oestrogenic compounds in sewage effluent are described in this article.
2. Levels of oestrogenic compounds in sewage effluent
A systematic review of oestrogenic compounds in European sewage effluent in the last 10 years yielded studies describing analytical methods for 193 compounds, specific isomeric forms or mixtures of such isomers (Martin et al. submitted a). This not only illustrates the complexity of the mixture of natural and anthropogenic substances that can be present in wastewaters, but also the fact that numerous chemical structures have some affinity for the oestrogen receptor. This is at least partly explained by its relatively low specificity because the binding pocket is nearly twice as large as the molecular volume of oestradiol (E2; Brzozowski et al. 1997). The most commonly analysed were also those whose levels are most likely to be of concern for the aquatic environment, namely, the natural steroid oestrogens and ethinyloestradiol (EE2), alkylphenols, diethylhexylphthate (DEHP) and to some extent bisphenol A (BPA) (Martin et al. submitted a).
Natural steroid oestrogens are female sex hormones produced predominantly by the ovaries and in small amounts by the adrenal glands. Oestrone (E1) and E2 levels generally increase by up to two orders of magnitude during pregnancy (Johnson & Williams 2004). Oestriol (E3) is only produced in significant amounts during pregnancy. All steroid oestrogens occur in sewage final effluents in a similar range of concentrations, the medians for E1 and E3 being slightly elevated when compared with E2 (figure 1). The commonest oestrogenic active compound in pharmaceutical formulation is EE2, and it is present in low nanograms per litre.
APEOs are non-ionic surfactants widely used in detergent formulations and other industrial, commercial and household products (Birkett & Lester 2003). Nonylphenol ethoxylates (NPEOs) and octylphenol ethoxylates (OPEOs) account for approximately 80 per cent and 20 per cent of the total APEO production. APEOs are partially degraded during sewage treatment, the ethoxy group becoming increasingly shorter and forming the more oestrogenic mono- and diethoxylates (NP1EO, NP2EO) derivatives as well as the alkylphenolic parent compound, nonylphenol (NP) and octylphenol (OP). Under aerobic conditions, both the final ethoxy group and the alkyl chain can be oxidized to form carboxylated derivatives (Montgomery-Brown & Reinhard 2003). Use of these chemicals in the textile industry, among others, results in regionalized inputs of NPEOs to trade waste received by STWs (around 82% of the UK textiles and woollen industry is located in the north of the country, 44% in the Yorkshire region alone) (Sheahan et al. 2002). Distributions of measured concentrations are presented graphically in figure 2, showing median levels of nonylphenol carboxylates in the 100 μg l−1 range, while levels of NPEOs were also high relative to other alkylphenolic compounds.
Phthalates are some of the most abundant and ubiquitous man-made chemicals in the environment, with DEHP being the most widely used. They are used as plasticizers, are not chemically bound to the end-product and can therefore leach into their surrounding environment. Based on the result of the systematic review mentioned above, the maximum concentration of DEHP in the final effluent from European wastewater treatment plants was 182 μg l−1, the interquartile range of reported values was 1.5–95.4 μg l−1 and the median was 5.3 μg l−1.
BPA is used in the manufacture of various plastics and epoxy resins to make a variety of common products, including baby and water bottles and medical devices. Epoxy resins are used as coatings on the inside of some food and beverage cans (Birkett & Lester 2003). The following statistics for concentrations in sewage effluent were obtained from the results on the systematic review; the maximum reported was 40.09 μg l−1, but the median and interquartile range was much lower, 0.36 and 0.08–0.36 μg l−1, respectively.
3. Significance of sewage effluent relative to other exposure pathways
(a) Steroid oestrogens
The range of concentrations of E1, E2, EE2 and E3 in various environmental compartments reported in the scientific literature are summarized in table 1. In many matrices, concentrations of E1 exceed those of E2, while, when reported, levels of E3 tend to be much lower. Natural steroid oestrogens occur in all vertebrates, and this is illustrated by the levels detected in meat, eggs and dairy produce (table 1). Although much concern has been expressed in the general media regarding the potential consequences of the presence of female hormones in tap water on male urogenital health, it should be clear that the levels detected in water are several orders of magnitude lower than those found in foods of animal origin (table 1), consumption of which is likely to be the main route of exposure for humans.
Equally, the wastes generated by husbandry practices of the same animals contain high levels of female hormones and should not be overlooked as a route of such compounds into the aquatic environment. This is well illustrated by the concentrations detected in headwater streams, impacted by agriculture and not sewage effluent (table 1). Some very high levels of steroid oestrogens, particularly EE2, have been reported for surface water and riverine sediments (table 1). While the range of reported values gives only a limited representation of actual levels, EE2 has been found to be more recalcitrant to biodegradation. Therefore, although levels discharged in sewage effluent are lower than those of its natural counterparts, EE2 is thought to persist longer and further downstream in the water column and sediments. Another interesting and often neglected source of steroid oestrogens in the aquatic environment is fisheries. Levels of E1 reported downstream of hatcheries were found to be comparable to those in a river with spawning salmon (table 1), and illustrate the need to better understand the occurrence of hormones and other oestrogenic compounds of natural origin in aquatic environments before the impact of anthropogenic sources can be fully assessed.
Natural products such as endogenous hormones cannot be controlled at source, although it could be argued that certain husbandry practices may influence the levels found in meat or other produce derived from farmed animals. Societal considerations are of particular relevance when trying to regulate releases of pharmaceuticals in to the environment, and this is particularly pertinent for the contraceptive pill. Health concerns related to the long-term use of steroid oestrogens for female contraception or hormone replacement therapy have, however, already driven research towards contraceptive formulations based on active ingredients more specifically targeting the female reproductive cycle than hormones involved in the many processes of the endocrine communication system, and clinical trials for candidate compounds are already taking place (Lakha et al. 2007).
(b) Alkylphenols, their ethoxylates and carboxylates
The ranges of some of the concentrations of these compounds reported in the scientific literature in various environmental matrices are given in table 2. A voluntary ban of APEOs in household cleaning products began in 1995 throughout Northern Europe, and restrictions on their use in industrial cleaning applications have been in place since 2000 (Petrovic & Barcelo 2004). The levels reported here may not yet reflect the effect of these controls on environmental releases. Judging from the levels in industrial wastewaters and sludges, industry appears to be an important source of such compounds in the environment (table 2). The solubility of APEOs depends on the number of polar groups forming the hydrophilic part of the molecule, and alkylphenol mono- and diethoxylates are generally described as lipophilic (Ahel & Giger 1993a). Their partition coefficients () suggest that these substances may become associated with organic matter in sediments (Ahel & Giger 1993b), as reflected by the very high levels found in sludges and sediments. Their presence in air and rain samples may again be partly due to their association with particulates (table 2). Equally, this lowers the likelihood that such compounds would leach from soils, as illustrated by the relatively low levels generally detected in groundwaters. The high levels of NPs reported by Soares et al. (2008) were detected in aquifers in the vicinity of anthropogenic sources of water pollution, such as contaminated rivers or septic systems.
Again, the levels found in drinking water compared with those found in foods imply that this would be a minor pathway for human exposure (table 2). However, these are not naturally occurring compounds, and their presence in food can only be attributed to anthropogenic sources, such as exposure of aquatic organisms to polluted waters or sediments. These compounds are also relatively persistent and have therefore the potential to bioaccumulate in environmental matrices (Ahel et al. 1993), as well as for biomagnification through the food chain (Cheng et al. 2006). Discharge of sewage effluents in surface waters and application of sewage sludge to land may therefore represent important routes of secondary human exposure.
(c) Plasticizers—bisphenol A and diethylhexylphthate
The levels of BPA and DEHP reported in the scientific literature and summarized in table 3 demonstrate the ubiquity of these compounds, as they have been detected in a wide variety of environmental matrices. They are found in all wastewaters, whether of municipal or industrial origin or resulting from solid waste disposal. Remarkably high levels have been found in wastewaters generated by paper recycling operations (table ??). BPA is relatively soluble and its partition coefficient suggest soils, sediments and other solid environmental matrices may be modest sinks for BPA (Staples et al. 1998). By contrast, DEHP has (Staples et al. 1997) and tends to accrue in solid wastes, such as compost, animal manures and sewage sludge. Despite their widespread distribution in the environment, the most important route of human exposure appears to be much more likely to result from direct exposure through the use of commercial products. This is illustrated by the levels of BPA found in canned food compared with those found in fish, as well as the levels of DEHP found in bottled water compared with those found in tap water (table ??). Both plasticizers are biodegraded fairly readily and do not persist in the environment. It is particularly interesting to note that the episodes of acute exposure to relatively high doses as a result of medical care, BPA in dental sealants and DEHP in poly-vinyl chloride (PVC) tubing used for drug delivery (table ??). This illustrates that, within the scope of the application of the PP, potential risks need to be balanced with the benefits of the application of the compounds of concern, as well as alternative products available. Further, in the context of endocrine disruption and developmental toxicity, protection of vulnerable groups, such as foetuses and infants, is of particular relevance due to the concerns related to their potential developmental toxicity. This has led Canada to ban the use of BPA in baby bottles (Chatterjee 2008), whereas in the UK, the Medicines and Healthcare products Regulatory Agency concluded that the availability of alternatives that are free of DEHP does not necessarily mean that it can be substituted without compromising the safety of the patient, pointing to the fact that the physical properties considered essential for some products used in highly demanding situations (for example, in some life-saving procedures on premature babies) cannot be achieved using the currently available alternatives to DEHP-plasticized PVC (Medicines and Healthcare Products Regulatory Agency 2007).
Risk-management options do therefore depend on the origin, application and fate of the compounds considered. Where there is sufficient evidence underlying concerns for human health, restricting or banning the use of synthetic products is an option to be considered, which will reduce the levels found in wastewaters, thereby contributing to the protection of the aquatic environment and limiting secondary human exposure via consumption of fish or shellfish. While this will depend on the balance between potential risks and benefits for the specific use of any substance, control at source is generally considered a more sustainable option.
4. Cost-effectiveness as a decision-support tool
As the analysis above demonstrated that sewage effluent is not a main pathway for human exposure, protection of the aquatic environment will be the main benefit of advanced wastewater treatment to remove oestrogenic substances. If the effects of oestrogenic compounds on endpoints affecting the reproduction, growth or survival of individuals of a particular species are well characterized, controversy remains over effects at the population level. Further, if the objective of risk management is the protection of aquatic biodiversity, ecological risks and benefits are related to the effects on communities rather than individual species.
Cost-effectiveness (CE) analysis is clearly identified in the European Water Framework Directive as the decision tool regarding measures identified to help achieve the specific objectives of the Directive (European Community 2000). Establishing linkages between specific measures and potential ecological benefits and the quantification of such benefits has been recognized as a particular challenge to the evaluation of the CE of measures (Risks and Policy Analysts Ltd 2004). The development of CE tools adapted to policies whose objectives are to conserve, protect or enhance biodiversity has been relatively neglected compared with other areas of environmental policy. This deficit is partly explained by the interdisciplinary barriers related to integrating economic and ecological models (Waetzold et al. 2006). Furthermore, research in this area has been generally restricted to conservation or land-use policies, despite the fact that, overall, inland water species appear to be declining faster than marine or terrestrial ones (Groombridge & Jenkins 2000). At present, the CE of removing pollutants from sewage has been estimated in pounds per tons or kilograms of compound removed, and neither gives a measure of effectiveness in terms of ecological benefits, nor does it allow comparison with measures aimed at other stressors.
Toxicity generally exhibits a graduated response between a stressor (the dose) and the severity of the biological effect, the number of individuals of one species or the number of species affected (the response). The principle of full recovery of costs adopted in the WFD necessitates that quantitative measures of environmental degradation be associated with specific pressures upon the aquatic environment, in turn demanding that economic methods use risk-based approaches rather than threshold dichotomies. A characterization and understanding of this dose–response relationship can be used to estimate the level of ecological improvement achieved by different levels of reduction in exposure. SSD is a statistical method that extrapolates single species test data, ranked in order of sensitivity to the substance of interest, to estimate effects at the community level as a fraction of species affected (figure 3). A method using SSD to estimate potential ecological benefits was applied to the removal of steroid oestrogens from sewage final effluent (Martin et al. submitted b) under various scenarios of dilution and pollution in the receiving water upstream of discharge. The combined oestrogenicity of the steroid oestrogens was estimated using two mixture models, namely, direct toxicity assessment (DTA) using results from oestrogenicity in vitro bioassays directly, or concentration-addition, adding the concentrations of the steroids in individual samples multiplied by their respective toxic equivalency factor (TEF) derived from the in vivo vitellogenin bioassay. These measured or calculated oestrogenicities were based on concentrations reported in the effluent of European STWs in the last 10 years (Martin et al. submitted a). Ecological benefit was expressed as the mean percentage of species spared when various level of technological performance, in terms of further removal of oestrogenicity, were applied.
While the CE of advanced wastewater treatment options will undoubtedly depend on site-specific factors, the generic results of this proof-of-concept study were combined with the capital and operational costs of two major technological options considered within the scope of the Environment Agency demonstration programme (namely, ozonation followed by granulated activated carbon (O3+GAC) and membrane filtration followed by reverse osmosis (MF+RO)) estimated by Jones et al. (2007). Total CE and annual CE per inhabitant were calculated (in million pounds per per cent of species spared or pounds per per cent of species spared per year per population equivalent, respectively) under a variety of scenarios or model assumptions for both these advanced treatment options; specifically, using ecological benefits estimated by the DTA or the TEF mixture models, assuming project lifetimes of either 15 or 30 years, assuming either 90 per cent or 99 per cent of further removal of oestrogenicity, for small, medium and large work sizes of 5000, 50 000 and 200 000 population equivalents (PEs), respectively, under various conditions of dilution in the receiving stream, whereby sewage effluent contributes from 10 to 100 per cent to the flow, and finally for background oestrogenicities from 0.1 to 10 ng l−1 E2 equivalents, respectively, in the receiving water.
Results demonstrated that CE varies widely depending on the assumptions made. In some circumstances, the cost of wastewater treatment appears prohibitively expensive. The maximum for total CE over the lifetime of the project was obtained using the DTA mixture model when upgrading large works using MF+RO technologies for 30 years, with a removal efficiency of 90 per cent where effluent only contributes 10 per cent of the river flow and the receiving water is already polluted (10 ng l−1 E2 equivalents). The estimated cost of protecting 1 per cent of species was as high as $329.5 million. Due to economies of scale, the maximum annual CE per inhabitant (PE) was obtained under similar assumptions, except for the size of the works and the lifetime of the project. The cost per year per PE of protecting 1 per cent of the species using the DTA mixture model when upgrading small works using MF+RO technologies for 15 years, with a further removal efficiency of 90 per cent where effluent only contributes 10 per cent of the river flow and the receiving water is already moderately polluted, was estimated as $104. Conversely, minima were obtained using the TEF oestrogenicity model for O3+GAC treatment assuming a removal efficiency of 99 per cent, where discharge contributed most of the river flow and water upstream of discharge was relatively pristine. The minimum annual CE per inhabitant was found when upgrading large works and operating for 30 years, whereby protecting 1 per cent of species would cost $0.17. The behaviour of this CE model is further illustrated in figures 4–⇓6. It is interesting to note that if, intuitively, CE is optimized when sewage effluent contributes most of the river flow, it is not necessarily the case that CE increases with background pollution, as illustrated by the different results obtained depending on which mixture model was used (figure 4). This also demonstrates that CE values will be sensitive to the site-specific effluent discharges of oestrogenic compounds. As both the initial capital costs and operating costs of the MF+RO option are higher than those of the O3+GAC option, the latter is expectedly always the cheaper option (figure 5). When considering the total CE over the lifetime of the project, CE increases with the size of the works. Conversely, the annual CE per PE decreases with the size of the works (figure 6). The latter measure accounts, to some extent, for the distribution of costs and benefits.
Most of the practical applications of SSDs in ecotoxicology have so far focused on the derivation of environmental quality standards, the reverse approach. As demonstrated in this work, a forward approach can also be applied to estimate the reduction in risk for different management options associated with a specific problem (Posthuma et al. 2002). The advantages of this method include: the protection of many species that have neither been nor will be tested due to experimental, ethical or financial restrictions; that the risk can be expressed as the potentially affected fraction of species, a criterion that is relatively simple to communicate to the wider public and other stakeholders; and that it can be applied to assess the combined risk associated with mixture of pollutants or even between physical (temperature, flow) and chemical pressures. A criticism of the SSD approach is that all interactions between species are ignored. It is generally assumed that at low doses, the ecological impact is overestimated by the SSD because of community resiliency and species redundancy (Posthuma et al. 2002). However, when such a significant fraction of species are predicted to be affected by the SSD, it is likely that the ecological effect is underestimated. Nonetheless, using a SSD approach to derive the ecological benefit function for the CE of surface water management options offers many advantages over existing practices. These include the ability to take the local environmental context into account, the possibility to estimate the impacts of mixtures of chemicals or even combinations of chemical and physical pressures, and the derivation of risks or benefits in a ‘common ecological currency’ that allows comparison with measures aimed at other stressors.
Whereas for synthetic compounds, management at source will generally be more sustainable than wastewater treatment, this is not applicable to some naturally occurring compounds, particularly those of faecal origin. Fish species have been found to be particularly sensitive to steroid oestrogens. The SSD method suggests that the current levels of oestrogenic substances discharged in sewage effluent can significantly impact aquatic biodiversity, yielding a mean fraction of species affected between 20 and 40 per cent (Caldwell et al. 2008; Martin et al. submitted b).
Applying the SSD as the ecological benefit function demonstrated that CE is highly dependent on the context in which the discharge is taking place. In instances where the receiving water is already polluted, either via upstream discharges of sewage effluent or diffuse sources of oestrogenic compounds, it is unlikely that ‘good ecological status’ can be achieved by tackling one point source of such substances. The question is therefore not whether or not to upgrade a specific STW, but rather which STWs to upgrade in order to optimize CE and whether to tackle diffuse pollution. Within the scope of the implementation of the WFD, a number of geographic-information-system-enabled hydrological models, such as LowFlow 2000, have been developed to predict the concentration profiles of contaminants along a river stretch or within a river basin. Combining the SSD method presented here with the outputs of hydrological models to estimate the ecological benefits in CE analysis promises to offer a versatile decision-support tool that enables options related to the location and level of STW upgrade to be explored on a river-basin basis, thereby optimizing ecological, as well as economical and social, efficiency. As the sustainability of centralized sewerage and sewage treatment is increasingly being questioned (Harremoes 1997), the ecological implications of various options for the location, size and design of STWs within a river basin could equally be explored using this method.
Although CE has clear advantages, it should not be the sole decision criterion. For example, the inevitable and environmentally undesirable increase in energy consumption and carbon footprint associated with advanced wastewater treatment technologies are not accounted for by CE analysis. Similarly, additional treatment can result in an increase in sludge production, which would have to be disposed of in a safe and environmentally sustainable manner. Sludge disposal is still at present the subject of much debate; its contamination with heavy metals and organic contaminants being one of the main concerns associated with the recycling of biosolids to arable land (Kouloumbos 2008). The decision-making process needs to address the paradox of wastewater treatment and water quality planning, i.e. the concurrent increases in energy use, carbon emissions and sludge production with increasing effluent quality (Zakkour 2002).
Moreover, the measure of effectiveness or ecological benefit in this analysis was exclusively concerned with oestrogenic compounds, more particularly steroid oestrogens. Advanced wastewater treatment will not however remove oestrogenic compounds specifically, but also a wide range of other contaminants. Ecotoxicology, aquatic toxicology in particular, and molecular biology are furthering the current understanding of the effects of combined exposure. Such research seems to support the pharmacological principles of dose/concentration addition for substances acting via the same mechanism and independent action for substances acting via different mechanisms (Silva et al. 2002; Faust et al. 2003; Zhu et al. 2006). Although such concepts ignore any interactions between substances such as synergy or antagonism, these can be integrated within the SSD approach to estimate the combined effect of mixtures. It does, however, only enable the estimation of ecological effects on species biodiversity in terms of a percentage of species affected or spared. The ecosystems approach advocated in the WFD should also encompass ecosystem structure and functions and the services they provide to humans, such as rare species of plants or animals, fish for recreational or commercial use, clean bathing waters, water purification, recharge of groundwater or flood control. The valuation of such services should also be included in the decision-making process (Water Science and Technology Board 2005).
This raises important questions regarding the environmental role of wastewater treatment. The ever higher quality standards for discharge of sewage effluent may result in some instances in dilution of the pollution in the receiving stream, while the water is often pumped back further downstream for potable water treatment and supply. As freshwater resources are under increasing stress, because of population growth, increasing pollution, poor water-management practices and climatic variations, wastewater reuse represents a sustainable option. Two major types of such reuse have been developed and practised around the world: potable uses, which can be direct, after high levels of treatment, or indirect, after passing through the natural environment (groundwater recharge is an example) and direct or indirect non-potable uses for irrigation in agriculture (Lazarova et al. 2001). The addition of treated wastewater to lake water before drinking water treatment was found to meet regulatory requirements, regardless of its influence on water quality, reducing trihalomethane but increasing nitrate levels (Comerton et al. 2006).
Moreover, the impacts of climatic change on water resources should also be taken into account when appraising measures taken to achieve ‘good ecological status’. An increase in intensity, severity and frequency of severe water events, such as droughts, storms or floods is expected. Extreme precipitation events put sewerage networks under additional pressure, and the current hydraulic capacity of the networks will be exceeded more often. The increased temperature could also lead to an increased frequency of algal blooms. Equally, in times of drought, levels of contaminants discharged by STWs or as run-off from agricultural land will increase due to higher evaporation and higher contribution of run-off and discharges to the river flow, thereby also increasing the likelihood of contaminated water reaching groundwater. Additional drinking water treatment may be necessary due to the lower quality of abstracted water, thereby incurring additional cost and energy use. Likewise, additional wastewater treatment necessary to adequately protect the aquatic environment will require infrastructure investment and further energy use (Water UK 2008).
The WFD seeks to integrate the various anthropogenic pressures on the aquatic environment, such as point and diffuse sources of pollution, hydro- and geomorphology, abstraction and flow regulation and introduction of alien species. Although it makes no specific mention of climate change, the WFD provides a valuable integrated framework for introducing climate change impacts into water resources management and river basin planning, and could ensure that adaptation measures can be implemented promptly. As SSDs can be used to assess ecological impacts associated not only with mixture of pollutants, but also other physical stressors, such as temperature or flow, investigating the application of the method in a context of adaptation to hydroclimatic changes is of particular interest.
One contribution of 12 to a Theme Issue ‘Emerging chemical contaminants in water and wastewater’.
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