Even though conservation methods like tillage may unavoidably perturb the soil surface, conservation tillage strategies, including reduced tillage (RT) and no-tillage (NT), can significantly decrease soil erosion rates (Daryanto et al., 2017). Other advantages of the conservation tillage (CT) include enhancing soil aggregation and increasing organic matter, thereby enhancing the water infiltration rates and water retention while encouraging biological activity (Plaza-Bonilla et al., 2013; Schmidt et al., 2019). Both RT and NT can be efficient methods of CT. For instance, a meta-data study of particulate P (PP) losses because of CT revealed 45% mean (concentration) and 55% (load) decreases relative to CT (Daryanto et al., 2017). However, these values did not reflect variations in climate, topography, cropping system, or length of tillage. Disaggregating the results showed that the efficacy of CT was reduced on more sloping lands (>4%) or with wetter antecedent states. Dodd and Sharpley (2016) reported a range for PP reduction from −33%–96% because of CT. Uusitalo et al. (2018) indicated that PP loads and levels were 27% and 55% lower, respectively, with NT compared to conventional till in Finland.

Clausen et al. (1996) investigated tillage influences on runoff for corn (Zea mays L.), demonstrating that RT decreased runoff by 64% and sediment losses by 99%. Adopting NT from a rice-planted watershed decreased the runoff volume and the TN and TP exports by 25.9, 8.5, and 7.8%, respectively (Liang et al., 2016). RT and NT decreased the intensity of tillage strategies and the effect of rainfall by conserving the soil surface using plant residues. Recently, it has been shown that ground covers and soil amendments like biochar, which improve soil aggregation and porosity, can be utilized to maintain the soil (Awad et al., 2012; Meier et al., 2017). Won et al. (2016) utilized rice straw mat cover, polyacrylamide, and gypsum to amend Chinese cabbage farmlands, which led to a decrease in suspended matter TN, and TP by 86.6%, 34.7% and 39.1%, respectively. Lee et al. (2015) assessed the effects of soil quality enhancement on soil loss. They found that the field soils treated with biochar and polyacrylamide decreased soil loss by 70.4% in a 33 mm d^-1 natural rainfall, whereas there was no difference in the runoff. Lee et al. (2018) conducted an integrated assessment of the impacts of biochar application on runoff quality, soil losses, and agricultural productivity. They found that farm field soils amended with 4% biochar reduced runoff by 16.8%, soil loss by 25% and N loss (via runoff) by 41.8% (2018). Biochar has generally been applied in soil remediation and possesses enormous potential in agricultural runoff control. Further research on the impacts of biochar on soil aggregation and pollutant fixation is needed in the future (Prenarathna et al., 2019; Sun et al., 2020).

CT efficiently minimizes DN in overland flow (Liu et al., 2014). These practices, however, unavoidably cause soil compaction over the long term, resulting in phosphorus buildup on the soil surface and, therefore, an increase in phosphorus concentrations in the runoff. In the northern Great Plains of Canada, it was reported that CT decreased total nitrogen levels by 41%, whereas total phosphorus levels increased by 42% (Tiessen et al., 2010).

Another option to manage pollutant loss during agricultural runoff is reduced tillage (RT). Conversion of CT to RT to minimize phosphorus (both in surface soil and leachate from plant residues) losses during runoff was reported to be a good practice (Liu et al., 2014). Consequently, it decreased the levels of TDP by 46% and TP by 38% and reduced loads of TDP by 56% and TP by 42%. This is because tillage measures could reduce soil compaction and reduce phosphorus buildup in surface soil. In principle, plant residues in CT could increase the water holding capacity, resulting in higher runoff duration. Consequently, RT may diminish the exposure duration between plant residues and surface runoff which decreases phosphorus from plant residues. Daverede et al. (2003) assessed the impacts of two tillage methods (NT and chisel plough) and a range of soil P levels on the concentration and loads of dissolved reactive phosphorus (DRP). They showed that the latter practice could decrease DRP by 60%. Consequently, choosing tillage measures must be based on regional environmental circumstances, soil status, crops, and predominant eutrophication pollutants.

To achieve these goals, the fundamental approach of 4R fertilizer management is to attain desired socioeconomic outcomes while minimizing negative ecological damage (Grant and Flaten, 2019). The principles of the 4Rs are relatively the same all over the world and are considered by many farmers as a precious tool to aid in the productive use of nutrients. N and P fertilizers are widely utilized in farming. Fertilizer-based NUE differs from crop to crop. For instance, the average nitrogen fertilizer utilization efficiencies of crops including wheat, rice, and maize are 18, 31% and 37%, respectively (Cassman et al., 2002). Consequently, it is essential to intentionally boost crop yield and agronomic efficiency while reducing adverse ecological consequences and maintaining N and P resources. Yao et al. (2018) studied the impacts of deep urea placement on N loss in the paddy land and found that deep urea placement decreased N loss by 50% while maximizing rice yield. Fertilization management plots, including band placement and hole placement, decreased TN by 63.6%–77% and TP by 42.8%–53.8% (Ye et al., 2016; Liao et al., 2017). Band placement could decrease contact with soil microbes and slow microbial nitrification. Zeng et al. (2008) measured N and P runoff losses from orchard soils as influenced by fertilization depths (surface, 10 cm and 20 cm) and rates and revealed that fertilization with chemical compound fertilizer at a soil depth of 20 cm decreased TN by 36.2% and TP by 31.4% in the runoff compared to both surface and 10-cm depth fertilization. Controlled-release fertilizer (CRF) can be another way to extend the duration of nutrient release and exploit the release rate, ideally to be compatible with the metabolic requirements of crops (Irfan et al., 2018). Tan et al. (2013) assessed the impact of fertilizer strategies on N loss via runoff on a wheat–maize rotation system surrounding Nansi Lake. They concluded that to decrease N loss, in parallel to securing continued good agricultural output levels, controlled-release N fertilizer combined with wheat straw integration must be promoted during corn production. Other research on the influences of controlled-release fertilizer reported a reduction in phosphorus loss in corn and paddy fields by 33% and 62%, respectively (Kun, 2012). Attention to the 4R stewardship would also manage the risk of loss of diffuse nutrients to the environment since the risk of pollutant loss from farmlands exhibits seasonal characteristics, with maximum pollutant loading in the rainy and fall seasons.

Intense rainfall and field drainage structures could drive overland flow. The rice‒wheat system is an essential potential N source for water pollution, accounting for up to 85% of cumulative nitrogen losses (Zhao et al., 2012). This is due to traditional flooding irrigation (TFI) continuing an elevated floodwater level in the farmlands. Water-saving irrigation methods can considerably decrease floodwater levels, enhancing the buffering capacity of farmland to aid in decreasing runoff and pollutant losses. In addition, water-saving irrigation improved root development by increasing grain production more than TFI (Aziz et al., 2018). Another technology that can be applied to decrease water use in irrigated lands is alternate wetting and drying (Hao et al., 2009; Matsuo and Mochizuki, 2009; Yang et al., 2009). This technology has been reported to decrease surface runoff by 30.2%–36.7% in compared to traditional methods (Liang et al., 2013). However, the levels of pollutants did not diminish with the reduction in surface runoff when alternate wetting and drying were utilized alone because the exposure period between water and soils could not decline. Therefore, integrating or combining irrigation management with farming practices and fertilization management should be the best approach.

All the aforementioned source reduction strategies can adequately address one or more nutrient loss pathways by decreasing the precipitation effect, enhancing infiltration, managing erosion, or decreasing leaching, agricultural runoff, and pollutant levels. However, they may not stop field runoff from reaching surface water. While source reduction strategies have considerably reduced the levels of agrochemicals in farm field runoff, it is still challenging to attain acceptable release levels. The long-term buildup of agrochemicals in waterbodies could also exacerbate the threat of eutrophication. Consequently, full control of farmland runoff requires extra process retention and end-treatment approaches.

Agricultural drainage ditches are normally planned and built as a part of farm irrigation systems and can be an essential portion of agricultural watersheds. Their primary purposes encompass irrigation and draining extra water from agricultural lands (Kröger et al., 2007; Kumwimba et al., 2018). Ecological ditches are being converted from conventional farm drainage ditches (Wu et al., 2013) to minimize water flow, and intercept and mitigate pollutants and sediment from runoff (Otto et al., 2016; Kumwimba et al., 201b; Kröger et al., 2008; Cooper et al., 2004). Ignored in the past for their mitigation abilities, ecological ditches have great potential to minimize the movement of pollutants derived from cultivated lands in a similar way as constructed wetlands (CWs). Their acceptance could be ascribed to their small area requirements and possessing the characteristics of both free-water surface wetlands and streams (Kumwimba et al., 2023b).

Ecological ditches are referred to as transition zones between terrestrial and aquatic ecosystems. Depending on the conditions, ecological ditches could behave as a sink to retain pollutants and promote denitrification or as a source to release nutrients into the water column. Because of the substrate/sediment and macrophytes within these systems (Figure 6A), a unique substrate-macrophyte-microbial system is constructed, and complex physicochemical, and biological processes take place for contaminant removal in the system. Table 3 illustrates the typical ecological ditches with different purification efficiencies. Reduction rates of 38.3%–68.35% have been demonstrated for N and P by ecological ditches through hydrophyte assimilation, adsorption, and denitrification (Kröger et al., 2007; Yang et al., 2016; Vymazal and Březinová, 2018; Wang et al., 2019; Kumwimba et al., 2020b). Wang et l (2019) demonstrated NH₄-N purification efficiencies of 26.7% (winter) and 51.9% (summer) for ecological ditches purifying rural wastewater. This is because macrophytes senesce and decay in water over the low winter temperatures, and the pollutants released from dead vegetation could further pollute the ecological ditch water (Kumwimba et al., 2023a).

When well-managed, ecological ditches possess great potential to minimize nutrients like CWs while providing various ecosystem services (Dollinger et al., 2015; Kumwimba et al., 2017d). However, it should be noted that the agricultural runoff nutrient removal efficiency of ecological ditches and their sustainable long-term operation are increasingly perceived as crucial future challenges. The nutrient load of ecological ditches is determined by a complicated process referring to ditch characteristics (structural or functional features); and available maintenance (Kumwimba et al., 2018). Several management practices, including using low-grade sediment traps, various plant species, organic carbon barriers, etc., have been employed in agricultural ditches to enhance nutrient removal efficiency. Low-grade weirs can maintain water levels and improve HRT in ecological ditches, subsequently, offering reasonable nutrient purification rates in systems. Ecological ditches designed with low-grade weirs decreased 51%–83% of the influent TP loads (Flora and Kröger, 2014). On the one hand, ditch features (size, length, slope, hydrophyte cover, types, etc.) and ditch bed properties (substrates/soil media types) are key factors impacting the removal performance (Wu et al., 2013; Zhang et al., 2016a; Zhang et al., 2016b; Kumwimba et al., 2017c; Kumwimba et al., 2017e). On the other hand, the efficacy of ecological ditches in purifying agricultural runoff nutrients from farmlands is a complicated process that heavily relies on biofilms, connectivity between fields and ditches, hydraulic loading, depth, temperature, and ditch network topologies (Figure 6B), which in turn could lead to variations in the retention capacity of nutrients. In short, long-term research on the application of ecological ditches for both rural wastewater and farmland runoff treatment under various ditch management methods must be further studied.

Planted buffer areas (PBAs) within and/or between farmlands and/or along water bodies can be seen as the last line/barrier to keep farm runoff and pollutants from reaching surface water bodies. A buffer zone is composed of three components: a) aboveground plants, b) belowground plant parts (roots and rhizomes), and c) subsoil. When agricultural runoff reaches buffer zones, water infiltration occurs. PBAs could diminish nutrient concentrations via filtration, infiltration, sorption, deposition of nutrient-bounded sediments, microbial transformation, etc. (Figure 7B). The removal efficiency of nutrients in PBAs is listed in Table 3. Various factors may impact the performance of PBAs, such as field parameters (e.g., type of nutrient, application rate, timing and frequency of nutrient application, rainfall intensity, runoff characteristics, slope of land, crop type), buffer vegetation type, soil characteristics, PBA dimension, drainage area to PBA ratio, season, management, PBA structure (presence of erosion rills), etc. (Lacas et al., 2005; Prosser et al., 2020; Kumwimba et al., 2023b) (Figure 7C). Most of these factors could enhance water infiltration, sedimentation, deposition, and biochemical reactions within PBAs. However, there is a current discussion about how wide the PBAs must be to prevent environmental consequences. Nevertheless, there is increasing uncertainty about the suitable PBA width for corridors or habitats, or what the minimum PBA width should be. Another issue is that their effectiveness in mitigating dissolved pollutants has been questioned (Figure 7B) (Kieta et al., 2018; Kumwimba et al., 2023a) since these practices can become a source of dissolved pollutants to sensitive waterbodies because of the saturation of PBA soil with nutrients. PBAs are not always effective in diminishing pollutant concentrations in cold climate regions. Excessive surface runoff events occur during spring snowmelt when PBA soils and PBA plants are frozen, and infiltration is highly restricted or absent and bacterial acclivities are low or nonexistent. Frozen/flattened vegetation can potentially contribute pollutants to runoff. Approximately 70%–80% of the surface runoff and pollutant losses take place during spring snowmelt in cold climate areas. Investigation of the removal of nutrients in buffer zones during snowmelt runoff is less frequent in comparison to rainfall-driven runoff. However, nominal absorption rates, incapacity for infiltration to take place or for sediments to be trapped within buffers, and the release of pollutants with snowmelt have been ascribed to frozen buffer zone soils and senesced plants. These circumstances hamper the capability of buffers to function as thought.

Satchithanantham et al. (2019) studied the pollutant retention performance of buffer zones by estimating inlet and outlet runoff pollutant levels during rainfall and snowmelt-driven runoff. The findings showed that pollutant levels of the soils in buffer zones could be the best indicator of whether the release or removal of pollutants could take place. The findings further indicated that VBZs act as a sink to sorb nutrients in runoff at low soil nutrient levels but as a source of nutrients at elevated soil nutrient levels. More variable responses were observed for NO₃-N, but overall, the buffer zones trapped an important amount of NO₃-N in the rainy period than in winter. This finding shows that the activities of vegetation and microorganisms in buffer zones are especially significant for nitrogen elimination. Vanrobaeys et al. (2019) assessed pollutant concentrations above and below the PBA. The results showed that PBAs are effective compared with cultivated land at minimizing sediment and sediment-bound phosphorus levels but do little to decrease dissolved forms of phosphorus. The efficacy of buffer zones in trapping phosphorus exhibits strong seasonality, as they possess certain abilities in the warm period but fail to perform during the spring snowmelt period. Therefore, it was suggested that the effectiveness of buffer zones must be improved in the summer when plants are actively developing. However, mowing plants at the end of autumn to minimize plant nutrient loss over these periods should be considered (Vanrobaeys et al., 2019). In cold climate areas, buffer design and management need careful assessment. Land managers and local governments must not greatly depend on buffer zones, as they may not function well for most times of the year to minimize pollutant transport. Further investigation must be carried out at a larger scale in natural systems and concentrate on optimizing buffer design and management methods.

To make PBAs more efficient, they must be “shaped” and smoothed out to encourage uniformly dispersed flow via more of the PBA’s surface, delay surface runoff, and trap chemicals in both dissolved form or attached to sediments. In addition, PBA management should be planned so that plants can be harvested and sequestrated pollutants can be eliminated. In low-lying zones where surface runoff takes place, it is advisable to put in a broad vegetated canal and remove/harvest the herbaceous plants. This implies that plants are inserted where the overland flow takes place and then mow the plants afterwards. Failing to do so, plants will senesce, die, and release back pollutants into the overland flow and streams. The PBA will not be as efficient without adequate management.

Artificial or constructed wetlands (CWs) are engineered systems that imitate (or optimize) the biogeochemical and physical processes that take place in natural wetlands for cleaning sewage under controlled conditions (Vymazal, 2007; Kumwimba et al., 2021; Ajibade et al., 2023), while offering mutual gains such as ecosystem services and recreation. They can be categorized by application, like habitat formation, flood regulation and sewage purification. Nevertheless, from a practical or technical point of view, different configurations of constructed wetlands (e.g., free water surface flow, subsurface flow and floating wetlands) have been utilized to purify different types of water (Figure 8I).

Sewage can be decontaminated in CWs via a sequence of processes (sedimentation, absorption, substrate/filler adsorption, precipitation, and microbial degradation, etc.) based on the synergistic actions of microorganisms, substrates and macrophytes (Brix, 1995; Vymazal, 2007; Stefanakis et al., 2014; Kujala et al., 2019; Kumwimba et al., 2020; Kumwimba et al., 2021). Fig. 8II shows the effectiveness of various constructed wetlands in tropical and cold climate regions. One of the great benefits of the CWs is the minimal operational and maintenance costs, in addition to many environmental advantages such as the minimum carbon footprint. This method may be used at small and large scales; nevertheless, CW requires relatively large land areas, depending on the design. Several factors could impact the performance of constructed wetlands, including feeding mode, HRT, HLR, pH, CW type, DO, inlet and outlet configuration, temperature, vegetation characteristics, etc. Hydrophytes, microbes, and substrates are responsible for the different treatment mechanisms (e.g., plant uptake, sedimentation, adsorption, filtration, microbial degradation, phytodegradation, phytoextraction, etc.) taking place in constructed wetlands. Undoubtedly, these mechanisms are impacted by climate conditions (Fig. 8II). These inherent weaknesses have usually limited their application and long-term stability. Valkama et al. (2017) assessed the daily and seasonal variations in pollutant retention performance of a boreal wetland. They showed that TP purification efficiency (28%) was most significant in summer and minimum (5.5%) in winter. Similarly, the TN purification efficiency (82%) was most remarkable in summer and minimal (3.5%) in winter. Certainly, the retention of N in constructed wetlands is a crucial question, particularly at low winter temperatures. The key nitrogen removal pathways are biological processes, which greatly rely on microbial nitrification-denitrification processes (Ji M. et al., 2020; Kujala et al., 2019; Varma et al., 2021; Ajibade et al., 2021a; b). An evaluation of most peer-reviewed papers on the performance of full-scale CWs globally showed that N was the most susceptible parameter to be easily impacted by low cold temperatures. In contrast, P, TSS and organic matter remain relatively less affected (Ji B. et al., 2020). Aiming to enhance the inadequate O2, low influent C/N ratio and low purification rate of constructed wetlands in cold temperatures, scholars have developed many modifications and different designs based on the system’s internal components (bacterial activities, macrophytes, substrate addition, and probable bioaugmentation) or external operational factors (heat loss, external aeration, operating conditions, and DO, etc.) (Figure 9; Figure 10). These modifications and technologies can promote water purification efficiency while keeping system ecological features and economic benefits.

Scaling has been one of the greatest challenges in CW research over the last few years. Extensive works have been conducted at the microcosm and mesocosm scales, but only a few research studies have been performed on full-scale CWs. Findings of these small-scale research or lab-based experiments are usually rarely applied to full-scale CWs, although few large-scale systems have been reported in the literature (Stefanakis, 2020; Alwahaibi et al., 2021). Furthermore, numerous findings are uncertain for full-scale CW systems, primarily from the operation, maintenance, and labor perspectives. The other key challenge is the investigation of the long-term performance of CWs. Existing studies are often brief and short, less than a year, which may not provide any valuable data concerning long-term performance. To improve the application of CWs, it is essential to perform more studies on full-scale CWs purifying actual sewage, specifically concentrating on long-term performance.

Growing interest in process retention strategies brings attention to vegetation-based water treatment and its role in diffuse pollution management. While understanding the suitable design, operation and management of vegetation-based water treatment systems is still depends on empirical parameters based on long-term assessment, there are complete global information gaps in understanding the system design, operation and maintenance. Knowledge of environmental conditions at the site (weather), sewage compositions, hydrology and hydraulic characteristics, substrate/filter medium types, vegetation type, system kinds and ages, full design aspects and water-quality parameters should be involved from both sources, e.g., scholars, and water treatment companies and city managers. There is still much room for developing and enhancing the predictive ability of models (related to both agricultural and urban conditions) for a deep understanding of the extra features, including transport and purification processes of diffuse nutrients and other contaminants on a watershed scale, and thereby helping to determine and target pollutant control strategies and to assess their relative efficacy. A thorough examination of ecological and health risks related to organic contaminants must be carried out in the plant-based treatment systems to remediate and manage the risks of water reuse through the use of a risk quotient approach and embracing measurable indices like the disability-adjusted life year. The application of plant-based treatment systems for goals other than eutrophic water purification could be another major field that the authors suggest for future studies. These systems could be essential in decreasing the carbon footprint of WWTP’s and developing a circular economy. For instance, the integration of plant-based treatment systems with bioelectrochemical systems to synchronously treat eutrophic water and produce power, which can reduce carbon transmissions. There are challenges in the approval and efficient operation of plant-based treatment systems, particularly in the least-developed nations. Developing practical and adaptable frameworks for design, building, operation and maintenance is required to orient and facilitate their use, particularly in least-developed nations. Future plant-based treatment system implementation must attempt to equalize the long-term sustainability of water purification and socioeconomic benefits, beginning from the design stage. In this context, plant-based treatment systems could and must be incorporated into city ecosystem growth by inserting characteristics including green roofs and green walls, further strengthening attempts to attain the global goals.

Ecological floating beds (EFBs), also known as floating treatment wetlands, artificial or planted floating islands, are another nature-based purification method that uses buoyant structures to grow aquatic or terrestrial plants on the water’s surface. Pollutant removal in EFBs is accomplished using many mechanisms (Figure 11A). Constructed wetlands with floating hydrophytes have been extensively applied worldwide to purify stormwater, wastewater, landscape water, and rural river water and prevent harmful algal blooms with low operational and maintenance costs (Wang et al., 2020a; Kumwimba et al., 2020; Kumwimba et al., 2022). For instance, field-scale EFBs purified 60 million m³ per year of sewage at a cost of only CNY 0.0018 per m³ (Afzal et al., 2019). Similarly, the operation and maintenance cost of the hybrid EFBs (0.0138 CNY/m³) was much lower than that of traditional sewage treatment facilities (Kumwimba et al., 2023b). In contrast to other ecological purification measures, EFBs benefit from not requiring large land areas while directly purifying eutrophic river water in situ at the field scale. Furthermore, EFBs could provide amenity value by greening and beautifying the rivers and the surrounding environment. These ecological advantages with numerous environmental benefits have resulted in their broad application worldwide, becoming compulsory in certain regions to obtain improved water quality (Olguín et al., 2017; Kumwimba et al., 2020b; 2022; Colares et al., 2020).

Like CWs, the purification performance of EFBs mainly relies on their hydrophyte roots (Figure 11). Nevertheless, several problems cast doubt on the use and long-term efficacy of EFBs (Figure 11B), especially their poor purification in cold temperature conditions due to the alteration of biological activities and hydrophyte absorption processes (Tharp et al., 2019; Wang et al., 2020a; Kumwimba et al., 2020; Kumwimba et al., 2022). EFBs are ineffective in deep and static water bodies as the macrophyte root lengths limit them. For instance, the mean length of hydrophyte roots in EFBs varies between 20 and 45 cm, making it unfit for use in deep waters, resulting in poor purification performance (Chen et al., 2016). The unsatisfactory EFB performance for carbon-limited wastewater treatment, especially in cold climates, is an inevitable limitation of the application of EFBs (Wu et al., 2021). Besides, insufficient carbon sources and/or electron donors could often limit the microbial denitrification process in EFBs, leading to poor pollutant removal performance. Nutrient removal in EFBs can be unsteady and insufficient during river remediation because of the restricted development of many floating macrophytes and the attached biomass (Sun et al., 2009).

In recent years, several technological innovations aimed at enhancing the purification performance of EFBs have been achieved (Figure 12). These encompass the incorporation of artificial biocarriers, artificial aeration, bioaugmentation, and the addition of various substrates (e.g., zeolite, alum sludge, biochar, woodchip, and stereo-elastic packing) and their combinations beneath the floating mats of EFB for the promotion of microbial attachments. Water depths of 0.6–1.1 m within EFBs and a rational bed coverage of 5%–38% were suggested for EFB designs, and the hydrophyte density should be inversely proportional to the hydrophyte size (Wang et al., 2020). Other studies also recommend up to 70% of needed bed coverage using Pontederia crassipes for a 0.25 m water depth and a contact time of 3–5 days to purify a eutrophic river (Gaballah et al., 2021). These results indicate that coverage must be expanded as much as possible for hypereutrophic rivers. For waterbodies with moderate eutrophication, the bed coverage could be minimized accordingly. Harvesting of plant biomass can be one of the essential operation parameters. Colares et al. (2020) reported that macrophyte harvesting could raise operation and maintenance expenses. However, the management of macrophyte biomass by harvesting could eliminate pollutants, and it can also be used in animal feed, green matter and compost, cosmetics, therapeutics, biosorbents, biofuels, and human food, thereby decreasing the overall system expenses. Another issue to address is the suitability of different hydrophytes for EFBs. Rigotti et al., 2021 reported that Typha domingensis was better at pollutant removal than Scirpus californicus in EFB treating urban runoff.

The addition of particular bacterial cultures to a contaminated environment (also known as bioaugmentation) has become an effective strategy for controlling algal proliferation and eutrophication (Xu et al., 2017; Simon and Joshi, 2021) because of its efficient, robust, reliable, and cost-effective nature. Bioaugmentation greatly depends on the metabolic activity of microbes by breaking down and converting hazardous substances while improving environmental integrity. In bioaugmentation, bacteria are the most prevalent organisms. Biofilms, aggregates of microorganisms on biotic and abiotic surfaces, have also been demonstrated as a rather powerful constituent of environmental sewage purification technology.

Yang et al. (2019) effectively used a new plastic carbon fiber filler to develop a biofilm technology to treat eutrophic water. They found an excellent purification performance for eutrophic water pollutants (e.g., TN, TP, and COD). Common biofilm methods used to restore eutrophic water systems encompass filler contact oxidation, contact oxidation, fluidized bed aerated biological filters, thin-layer flow technology, suspended biofilm reactors (frequently also known as moving bed biofilm reactor), and steam injection approaches for enhancing geothermal steam containing pollutants (Qian et al., 2021). Many published benefits for the utilization of moving bed biofilm reactors include space savings, robust and simple reduction of contaminants, less temperature dependence, inexpensive and relatively easy processes, minimal operational and maintenance costs, low biomass loss, steady biofilm thickness, and minimal likelihood of clogging (Zhang et al., 2016). Hypereutrophic waterbodies and sewage can be treated by biological contact oxidation techniques, which have many benefits, such as promoting sludge thickening and the lack of bed clogging. Nevertheless, their effectiveness varies with seasonal temperatures. This method is eco-friendly and cost-effective in improving contaminant degradation (Simon and Joshi, 2021).

With the goal of optimizing purification performance while limiting ecological risks, many scholars have evaluated the potential of exploiting the advantages of the combining or integrating several purification processes (Addo-Bankas et al., 2022). Table 1 illustrates different purification strategies, their corresponding purification mechanisms, and the advantages and disadvantages of each strategy. These integrations could enhance the purification performance for eutrophic water treatment compared to the individual approaches. Šereš et al. (2017) combined constructed wetlands and stabilization ponds to enhance secondary sewage purification, particularly dispersed village sewage or concentrated sewage in littoral areas. It is desirable to reduce building costs in regions with several ecological drainage ditches and conduits (Ham et al., 2004). Similarly, integrating an algae pond via the O₂ enhancement of algae photosynthesis and a constructed wetland considerably improves nitrogen elimination (Zhao et al., 2016). Meanwhile, microorganisms could use the algae debris in the constructed wetland as a carbon source in treating sewage with a low carbon-to-nitrogen (C: N) ratio. Moreover, Lin et al. (2003) integrated aquaponics and greenhouse structures into constructed wetlands. Other scholars (Hu et al., 2012; Wu et al., 2014) established an integrated system of eco-wetland ponds and the extensive use of resources, which offered economic advantages to the regional freshwater aquaculture industries. Combining TiO₂-photocatalysis and CWs to purify many specific contaminants in sewage effectively has also been demonstrated in recent research (Antoniadis et al., 2007; Arana et al., 2008; Herrera-Melian et al., 2012; Wu et al., 2016). The combination of CWs with ozonation was also found effective in the treatment of complex agro-industrial effluents such as the cork boiling wastewater (Gomes et al., 2020). Another fascinating research was performed by Huang et al. (2022), in which a bypass CW system integration method, including surface flow CWs, eco-gravel beds, eco-wetland multipond systems, and underwater forest ecosystems, was implemented to enhance river water quality. The results show that the integrated technology operates stably and could enhance contaminant removal in river water. For example, NH₃-N, TP, and COD concentrations were less than 1.0, 0.2 and 20 mg L^-1, respectively, while DO exceeded 5 mg L^-1, and transparency exceeded 60 cm. Additional studies were performed on the potential of combining aeration and algae-eating fish for eutrophication and algal growth control in rivers. Integrating EFB and aluminum-based drinking water treatment residual enhanced its purification performance considerably (He et al., 2022). The findings exhibited improved N and P removal efficiencies, with mean retention of 53.3% for N and 45.3% for P, compared to the control with 27.3% for N and 12.8% for P. It was also revealed that combining fillers with EFB might enhance nutrient retention while boosting macrophyte development.