Pollutant removal within constructed treatment wetlands

Wetlands receive the majority of their resources from terrestrial systems and offer a useful combination of conditions for supporting the biogeochemical cycles locally (Keddy, 2010). For instance, due to their sink function, wetlands accumulate relatively high amounts of carbon (Dodds and Whiles, 2010), which is used as a food source by the prevailing microorganism community (see Box 2.1). Aside from creating new biomass, gaseous carbon-based by-products are excreted by these microorganisms, including CO2 (under aerobic conditions) and CH4 (under anaerobic conditions). Yet, due to the high amount of carbon within the wetland, oxygen is often depleted throughout the majority of the water column, causing mostly anaerobic conditions to occur. Consequently, wetlands tend to contribute to climate change by exhausting CH4

and N2O, which are created in anaerobic conditions and have a higher global warming potential than CO2 (i.e. around 28 and 265 times at the century scale (IPCC, 2014), respectively). At the water surface, however, oxygen diffuses into the water column and allows for the presence of an aerobic boundary layer. Due to this layer, a heterogeneous environment exists, with complementary processes occurring in the top (aerobic) and bottom (anoxic) layers.

These biochemical processes have been the basis for applying a wetland configuration within the framework of water treatment, with wastewater originating from domestic, agricultural, industrial and storm water sources (Kadlec and Wallace, 2008; Vymazal, 2010). Similar to conventional treatment systems, these natural counterparts rely on the activity of microorganisms to mineralise or transform waste products into new resources (i.e. nutrients, see further) without extensively applying chemicals, electricity or artificial aeration, although research on how these factors impact treatment performance is ongoing (Donoso et al., 2019; Gao et al., 2017).

The microbial conversion of organic material into new resources supports the survival and reproduction of primary producers (phytoplankton, macrophytes). Moreover, due to their sink function and associated biogeochemical processes, freshwater wetlands can produce up to 10 times more biomass than lakes and streams (i.e. around 1100 g∙m^-2∙y^-1 versus 110 g∙m^-2∙y^-1, respectively) (Dodds and Whiles, 2010). This biomass, in turn, acts as a food source for heterotrophic organisms (zooplankton, macroinvertebrates, fish), including herbivores and detritivores. Hence, the presence of these biogeochemical processes provides the basis for complex food web development.

Within the remainder of this section, the attention is focused on (1) the most frequently occurring and reported pollutants within CTWs and (2) additional key aspects that require study to improve and evaluate the applicability of CTWs. The different biotic groups that benefit from the provided resources will be discussed in the next section (see Section 2.3).

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Box 2.1: Microorganisms in constructed treatment wetlands

The terminology ‘microorganism’ as used here overarches several biotic groups, including Archaea, bacteria, fungi and microscopic algae. The identification and classification of microorganisms is increasing, though researchers acknowledge the idea that the majority remains undiscovered (Cavicchioli et al., 2019; Saccá et al., 2017). The composition of such a microorganism community is highly case-dependent and often hard to control due to the complex interplay and dependency among species. For instance, He et al. (2017) indicated that the increased usage of saltwater to replace freshwater during activated sludge treatment potentially affects the performance of the system and illustrated that increased salinity decreased bacterial activity and sludge floc size. Nevertheless, microorganism presence remains essential in developing and maintaining the biogeochemical nutrient cycles that underlie the high-valued attenuation capacity of natural systems (Cavicchioli et al., 2019; Saccá et al., 2017).

Within the considered FWS CTWs, microorganisms can occur in the sediment, in the sludge layer, suspended in the water column and attached to alternative substrates (including stones, vegetation, liners and pipes). The latter often combines with non-motile algae and the resulting micro-community is generally referred to as periphyton, which is described in more detail in Section 2.3.1.2, along with its importance for supporting the development of aquatic food webs. Transformation of pollutants throughout CTWs is highly dependent on the activity of these microorganisms and therefore benefits from the creation of additional surface area. Consequently, higher removal efficiencies are theoretically obtained for treatment systems characterised by a flow through a substrate (i.e. subsurface flow) rather than on top of a substrate (i.e.

free water surface), although this has been contradicted by field observations (Kadlec, 2009).

Presence of microorganisms within CTWs is crucial for developing aquatic food webs, while a variety of factors (e.g. temperature, wastewater type, macrophyte presence) dynamically influences the prevailing community composition. For instance, Wang et al. (2016) observed that reduced temperatures negatively affected the performance of the microorganisms and, consequently, the removal efficiency of the system.

Moreover, they found that plant presence has a positive effect on microbial abundance, being further extended by Hernández-Crespo et al. (2016) who stated that combining multiple plant species supports a more diverse microbial community.

KEY ISSUES MULTIFUNCTIONAL WETLANDS

27 2.2.1 Wastewater pollutants and removal within CTWs

Within this subsection, specific attention is given to total suspended solids (SS), biochemical oxygen demand (BOD), chemical oxygen demand (COD), total nitrogen (tN) and total phosphorus (tP), as they are the main focus of both legislation and research. Complementary topics dealt with in literature include metals, pharmaceuticals and personal care products (PPCP), pesticides, faecal contamination and endocrine disruptors (ED) (Vymazal, 2009).

SS represent all the particulate matter being suspended in the water column, covering a fraction of the overall BOD, COD, tN and tP content. Due to the low flow conditions within FWS CTWs, SS is mostly reduced via settling and complemented with decantation and filtering (e.g. due to macrophyte presence) supporting removal efficiencies up to 80 % (Kadlec and Wallace, 2008; Rousseau et al., 2004b; Verhoeven and Meuleman, 1999). Consequential to the settling of SS, a sludge layer is formed at the bottom, being a mix of non-degradable (e.g. sand, silt) and degradable solids, allowing the latter to dissociate and, ultimately, dissolve within the water column or dissipate into the atmosphere. The mineralisation underlying this dissociation is a complex concert of pollutant-specific processes (see further), often resulting in reduced sludge volumes.

Within both the settled solids and water column, organic pollutants (BOD and COD) are subjected to biochemical processes conducted by microbial activity. Both aerobic respiration (conversion of organic-C into CO2) and anaerobic fermentation (conversion of organic-C into CH4) support the (partial) removal of BOD and, hence, COD (Kadlec and Wallace, 2008). The openness of FWS allows for the diffusion of oxygen into the water column, yet this rate tends to be lower than the overall oxygen demand and causes oxygen depletion near the bottom. The resulting gradient separates the aerobic layers with CO2-production at the surface from the anaerobic layers with CH4-production near the bottom. Still, removal efficiencies up to 90 % for BOD and 80 % for COD have been reported, though these can be as low as 50 % for BOD and 60 % for COD (Galanopoulos et al., 2013; Healy et al., 2007; Kivaisi, 2001; Wang et al., 2017).

At nutrient level, both sedimentation and microbial activity play a role, though the importance of each process depends on the type of nutrient under consideration. For instance, the nitrogen cycle is highly diverse, including anions (NO₂⁻ and NO₃⁻), cations (NH₄⁺) and gaseous forms (NH₃, N₂O and N₂). With N₂ being the main component of the atmosphere, the transformation of organically bound nitrogen via ammonification (production of NH₄⁺ and NH₃ following a pH-based equilibrium), nitrification (conversion of NH₄⁺ into NO₃⁻ via NO₂⁻ in aerobic conditions) and denitrification (conversion of NO₃⁻ into N₂ in anaerobic conditions with a C-source) into nitrogen gas (see Figure 2.2) does not pose any significant environmental impact.

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However, the gaseous nitrous oxide (N₂O) produced during incomplete denitrification potentially leaks into the atmosphere where it contributes to global warming (IPCC, 2014; Song et al., 2012). Moreover, with nitrification occurring in the aerobic top layer and denitrification taking place in the anaerobic (sludge) layer (Figure 2.2) (Vymazal, 2010), the overall nitrogen removal efficiency is highly dependent on the diffusion process, resulting in lower values compared to carbon removal efficiencies (up to 86 %, but going as low as 14 % (Wang et al., 2017)). Nevertheless, when all conditions are present to support ammonification, nitrification and denitrification, FWS CTW can remove nitrogen indefinitely (Zedler, 2003).

Figure 2.2: Illustration of the nitrogen cycle within wetlands. Dissolved ammonium (𝑁𝐻₄⁺) equilibrates with ammonia (𝑁𝐻₃), which is oxidised to nitrite (𝑁𝑂₂⁻) and nitrate (𝑁𝑂₃⁻) in the aerobic zone. In the anaerobic zone, nitrite is reduced to nitrous oxide (𝑁₂𝑂) and nitrogen gas (𝑁₂). The latter two can escape into the atmosphere as a gas, as well as ammonia (𝑁𝐻₃).

In contrast to carbon and nitrogen removal, limits occur with respect to phosphorus removal due to the absence of a gaseous form. Phosphate-ions (PO₄³⁻) adsorb on the substrate surface, which, over time, results in lower removal efficiencies due to saturation effects (Bolton et al., 2019; Vohla et al., 2011). For instance, Wang et al. (2017) reported a decrease in phosphorus removal efficiency from 91 % down to only 35 % due to saturation effects within the substrate. Studies on characteristic substrate saturation curves indicate that phosphorus breakthrough in operational CTWs can be delayed by using different substrates (Bolton et al., 2019; Park and Polprasert, 2008).

Unfortunately, no stand-alone solutions to this substrate saturation are currently available, which implies that CTWs cannot act as completely independent treatment systems.

Each of these removal processes is steered by a plethora of abiotic variables, ranging from manageable (e.g. retention time, substrate depth) to unmanageable (e.g.

temperature, precipitation) variables. Research related to these variables indicates an improved performance within a warmer climate, broad open spaces and higher retention times (Garfí et al., 2012; Kadlec, 2009; Kotti et al., 2010).

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29 For instance, Wang et al. (2017) observed that cold climates had an upper limit of 80 % removal for BOD, while Kadlec and Wallace (2008) reported BOD removal efficiencies over 80 % in warm climates, thereby illustrating the positive effect of increased metabolic rates of the prevailing microorganism assemblage due to elevated temperatures (Kadlec, 2009). In contrast, the removal of SS is negatively correlated with temperature, as elevated microorganism productivity increases their suspension potential, hence causing higher SS levels to occur within the wetland effluent compared to the influent (Kadlec and Wallace, 2008).

Similarly, steering biotic variables range from manageable (e.g. presence of vegetation) to limitedly or completely unmanageable (e.g. microorganism assemblage). The effect of macrophyte presence on pollutant removal has been studied for decades and has been reported as one of the factors controlling temperature and nitrogen removal in wetlands (García-Lledó et al., 2011; Vymazal, 2007; Vymazal, 2013; Wang et al., 2017).

Macrophytes influence pollutant removal at several levels. For instance, plant structures within the water column provide a surface for microorganisms to grow and interact, thereby supporting improved contact between the organic pollutants in the water phase and the heterotrophic bacteria (Brix, 1997; Fan et al., 2016). Moreover, within the root zone, oxygen is released and results in micro-aeration of the substrate, thereby locally supporting the presence of aerobic bacteria active in the oxidation of both organic matter and nitrogen compounds (Brix, 1997; Vymazal, 2013). Contrasting these indirect effects, a direct effect on nutrients is exerted via uptake and assimilation into biomass (Beutel et al., 2014; Dierberg et al., 2002). However, this type of nutrient removal has been observed to account for maximally 10 % of the total incoming load and is potentially returned to the water phase when biomass is not harvested (Hernández-Crespo et al., 2016; Merlin et al., 2002; Park and Polprasert, 2008).

The extensive range of variables identified to exert an influence on wetland performance in combination with the reported case studies to be found throughout literature, illustrates that many research gaps still exist, especially due to the limited comparability of different systems (Thomaz and Cunha, 2010). Moreover, the effect of these variables is not restricted to altering pollutant removal, but extends to the water body receiving the effluent of the treatment system. Similar to the effect of conventional wastewater treatment plants (Ort and Siegrist, 2009; Zhou et al., 2009), both quantity and quality of the effluent have the potential to cause changes in the abiotic conditions downstream of the CTW discharge point. However, the intensity of these changes remains highly dependent on the actual flow of the discharge, which is often several magnitudes smaller than conventional systems due to being applied at a smaller scale. Despite the importance of discharge flow, attention in the following section is mostly directed at the treatment performance of artificial wetlands.

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2.2.2 Improving treatment to accommodate clean water and sanitation

Constructed treatment wetlands provide the potential to reduce the amount of incoming suspended solids and biodegradable organic compounds up to 85 % and 80

%, respectively, although this highly depends on the type of water being treated (Hijosa-Valsero et al., 2010; Verhoeven and Meuleman, 1999). However, environmental conditions greatly affect microbial processes, causing difficulty in reaching stable effluent concentrations, while the absence of strong oxidative compounds within the treatment system impedes the removal of highly recalcitrant organic compounds (Donoso et al., 2018). This provides two main areas for further research: (i) improve the understanding of how prevailing conditions affect the treatment efficiency and (ii) determine the potential impact of recalcitrant compounds on freshwater conditions.

Improved understanding of the treatment performance implies the combination of experiments, analyses and simulations. A multitude of experimental studies discussing separate case-studies can be found in literature, applying a range of wastewater compositions (Garfí et al., 2012; Wang et al., 2017), different kinds of vegetation (Maine et al., 2007; Vymazal, 2013) and a variety of substrate types (Sakadevan and Bavor, 1998;

Vohla et al., 2011), yet provide a limited basis to support an overall, holistic comparison.

For instance, Donoso et al. (2017) assessed the operating conditions (i.e. temperature, water flow) of FWS CTWs treating diffuse nutrient pollution and concluded that FWS CTWs provide an alternative measure to fight the eutrophication of waterways. Despite the fact that this result supports the applicability of FWS CTW as a mitigation measure, only superficial information related to the influence of prevailing conditions on treatment performance can be extracted from this type of studies. This highlights the need of more in-depth research to obtain a better process-based understanding of CTW performance and the inherent influence of environmental conditions.

Secondly, despite providing relatively high removal efficiencies for specific pollutants, trace concentrations do occur within effluents that are discharged into the environment, especially in the case of recalcitrant compounds. Effects caused by their discharge are highly case-specific and depend on the prevailing freshwater conditions on the one hand and on the pollutant load and discharge frequency on the other hand. For instance, exceeding the official effluent standards causes an unequivocal drop in absolute water quality, while the relative change can be higher for high-quality compared to low-quality surface waters. To illustrate this, Donoso et al. (2018) studied the relevance of COD discharge limits for CTWs treating animal manure by assessing the occurrence of macroinvertebrates in the receiving river. They observed the presence of pollution-sensitive taxa downstream of the discharge point, despite the standard-exceeding COD concentrations in the effluent, suggesting that the existing COD-standards might be too stringent.

KEY ISSUES MULTIFUNCTIONAL WETLANDS

31 Aside from indicating the limited environmental effect, Donoso et al. (2018) did not specify the COD compound composition, making this kind of conclusion inference overly simplistic and inappropriate towards other types of wastewater. For instance, high concentrations of insecticides can result in high COD concentrations in the effluent, simultaneously causing drastic effects on the downstream macroinvertebrate assemblage. Hence, a more in-depth characterisation of COD compounds and how they behave within the CTW is required prior to adapting the standards.

Progress within these fields is crucial to optimise the treatment process and limit the environmental impact. This requires the collective consideration of societal, environmental and operational aspects (Becerra Jurado et al., 2009; Mereta et al., 2012;

Truu et al., 2009), as illustrated in Figure 2.3. However, most studies only focus on a subset of these aspects, with limited research applying a holistic approach. For instance, De Troyer et al. (2016), assessed the water quality of the rivers and wetlands around Jimma (Ethiopia), considering both chemical and biological indicators. They acknowledged the potential of wetlands as a promising technique for wastewater treatment, though concluded that further societal awareness and stakeholder participation were needed to implement CTWs in regions affected by water pollution, limited sanitation and overall poverty. Similarly, other reports highlighted the capacity of natural and CTWs for wastewater treatment, while concluding that implementation is impeded due to stakeholders lacking insight into the integrated functioning of CTW ecosystems (Donoso et al., 2017; Hefting et al., 2013; Vymazal, 2010). These observations highlight the need for (i) including societal aspects into CTW research and (ii) assigning a budget for educating and involving local communities, confirming that restoration success is determined by merging science, society and politics (Catalano et al., 2019;

Jähnig et al., 2011).

Figure 2.3: Illustration of the required input to improve implementation of constructed treatment wetlands. By combining only two aspects, successful long term implementation is impeded due to the lack of societal, environmental or operational input.

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