A review on enhanced nutrient removal in Constructed Wetlands: Insights into operating conditions and plant variation

doi:10.21203/rs.3.rs-1564173/v1

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A review on enhanced nutrient removal in Constructed Wetlands: Insights into operating conditions and plant variation

https://doi.org/10.21203/rs.3.rs-1564173/v1

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In the last few decades, nutrient-rich wastewater has become a worldwide problem due to exponentially increased anthropogenic activities causing eutrophication or algal blooms in waters and affect the water bodies with their biological and chemical activities. Physico-chemical techniques have shown promising results, while operational cost, hefty sludge generation, and energy requirement are the major drawbacks. On the other hand, biological processes are cost-effective and efficient, although they are associated with design complexity and extensive monitoring. Among biological techniques, constructed wetland (CW) is an economical, efficient, and green technique for nutrient removal comprising of inherent mechanisms, such as microbial degradation, rhizofiltration, phytodegradation, sorption, etc. This major highlight of the present study is to apprise and critically analyze the ideal conditions and system insights to enhance the nutrient removal efficiency of CW. Macrophytes play a vital role in CW and its planting practices, namely monoculture and polyculture, also influence the removal efficiency of the CW. In the present study, an in-depth analysis reflected that polyculture exhibits better treatment for TN and TP compared to monoculture practices. The observed optimum conditions for nutrient removal in CW were polyculture in combination with higher temperature (23ºC-29^ºC), artificial aeration, advanced substrate, and bioaugmentation. The review emphasized the process and design criticality of CWs in removing nutrients from various types of wastewater, which may be instrumental for water research and field applications.

Removal mechanisms

Polyculture

Monoculture

Influencing parameters

sustainability

The most prominent cause of nutrient pollution in the water resources is the exponential growth of industries and major agricultural activities to fulfil the rapidly growing human demands (Iamchaturapatr et al., 2007). Since the middle of the 20th century, surface water pollution caused by an excessive nutrient supplement from human activities has become a significant threat to water resource sustainability and environmental safety worldwide (Liu et al., 2017). The sources of nutrients are broadly classified into two categories, point sources and non-point sources. The non-point sources of nutrients are less controlled and generate nutrients in large quantities than point sources (Liu et al., 2017). It was reported that a point source (Industrial effluent, WWTP effluent, etc.) discharge has a significant effect on the water body and may carry the nutrient load for several kilometers downstream of the input (Haggard et al., 2005). Discharge of industrial wastewater into water bodies without treatment disrupts ecological sustainability, leads to algal bloom, and creates risk for human health (excess nitrate contaminant may cause the blue baby syndrome in infants) (Chiranjeevi et al., 2019). To control the impacts of agriculture on the quality and ecological biodiversity of water resources, agricultural engineers and environmental agencies should find a cost-effective and efficient technique to manage nutrient pollution (Tanner et al., 2005).

Constructed wetland (CW) is a new way or technology designed to remove contaminants by natural processes in a controlled environment with better removal efficiency. CW works on natural contaminant removal processes along with saving costs because this requires less or no power consumption, human labour and has lower construction or maintenance costs (Bois et al., 2021; Rousseau et al., 2008). Studies suggest that CW's application with other existing nutrient removal technology might be an effective solution for wastewater treatment (Liu et al., 2015). Various types of wetlands, such as Free water surface flow CW (FWSCW), Horizontal subsurface flow CW (HSSFCW), Vertical subsurface flow CW (VSSFCW), are used for nutrient removal; moreover, combined systems, i.e., hybrid systems, are used to exploit the benefits of various specifications of different CW (Vymazal, 2007; Wang et al., 2021). Nutrient removal in CW occurs by many processes; nitrogen is primarily removed by microbial activities and plant uptake, while the majority of phosphorus is removed by adsorption and plant uptake (Vymazal, 2007). The CW’s performance depends on many factors, i.e., temperature, dissolved oxygen (DO), microbial activity, pH, substrates, and macrophyte species (Varma et al., 2021). The macrophytes play an important role in the CW’s performance with their associated microbial biofilms which perform nutrient removal by transformation and storage (Dipu et al., 2011). Several types of plant species have been used in CW, such as Typha spp., Phragmites australis, Glyceria spp., Eleocharis spp., Iris spp. etc. (Shelef et al., 2013; Tanner, 1996). Furthermore, from the planting point of view, the CWs are classified into two groups: monoculture and polyculture (Tondera et al., 2020). The effect of plant species variation was studied in many articles, but the dependency of performance on the monoculture or polyculture is not described well (Liang et al., 2011; Licata et al., 2019; Picard et al., 2005; Shelef et al., 2013; Vymazal and Kröpfelová, 2011). Moreover, the studies have been documented to show inconsistent results for different nutrients (Liang et al., 2011). Out of a large number of studies considering total nitrogen (TN) and total phosphorus (TP) removal employing CW, few studies focused on different culture practices (Fraser et al., 2004; Leiva et al., 2018; Zhang et al., 2007a). Although macrophyte culture practices and their effects on the CW performance have been reported in only a few publications (Fig. 1), this review tried to focus on the possible impact assessment of monoculture and polyculture practices on CW.

This article presents a qualitative and quantitative analysis of the major nutrient sources, including point and non-point sources. It also highlights the various environmental and other impacts concerning nutrients rich wastewater. This paper provides an in-depth understanding of CW, its fundamental classification, and the different processes involved in nutrient removal. Other existing nutrient removal techniques are also described with their treatment efficiencies, advantages, and disadvantages. The role of macrophytes in CW has been addressed along with the role of other parameters (Temperature, pH, microorganisms, substrate, and DO). Additionally, the effect of monoculture and polyculture practices on the nutrient removal efficiency of CW is also determined.

Nutrients are conserved in the plant organs and other biomass while their forms keep on changing according to the environmental conditions in the nutrient cycling by various mechanisms. Nutrients move in a cycle and change their forms at distinct sources, such as the nitrification process results in the conversion of ammonia nitrogen (NH₄⁺-N) into nitrite nitrogen (NO₂⁻-N) and further due to oxidation into nitrate nitrogen (NO₃⁻-N) (Lam et al., 2020). Similarly, the phosphorus cycle plays a major role in nutrient availability in the environment. Phosphate-bearing igneous and sedimentary rocks are the most important pre-anthropogenic sources of phosphate (PO₄-P), which is released by continental weathering. The weathering profiles of the soil composes temporary reservoirs of PO₄-P that store and transform the bioavailable- organic phosphate. It has been reported that in rivers, approximately 95% of phosphate-containing compounds are particulate, of which 40% are organic (Föllmi, 1996). Human interventions also produce a high amount of TP and released it into the water which further increases water loads. The minerals generally contain TP in form of PO₄-P which is immobilized and mineralized by microbes in soil and made available for plant uptake. Concerning the extent of generation, the sources of nutrients have been classified into two types, i.e., point and non-point sources. The non-point sources are generally agricultural farms, forests, and open barren lands runoff, while significant point sources are industries and municipal wastewater. In most developed countries, point sources of nutrient pollution are well-controlled, whereas non-point sources need proper management (Han et al., 2021).

Animal products processing industries, dairy industries, and livestock wastewater have a high nutrient concentration in their effluent, also shown in Table 1. Unregulated agricultural pollution is a significant non-point source of nutrients in developed and developing countries due to intensive farming (Gottschall et al., 2007). In an ideal condition, plant use, only 50% of applied fertilizers, 2 to 20% is lost in evaporation process and, 15 to 25% react with organic compounds present in clay soil, while residual fertilizer (2 to 10%) pollute the soil and water (Savci, 2012). However, it was reported that 35 to 93% of NPK fertilizer applications could be reduced by implementing a balanced crop nutrient requirement (Chen et al., 2021). Due to the high demands and processing needs of dairy products, it results in huge wastewater generation (Schierano et al., 2020). Dairy parlours wastewater contains urine, cattle dung, detergent, and other organic and inorganic contaminants, which may severely affect the environment (Healy et al., 2007). This wastewater causes eutrophication by leaching and runoff and requires an extra financial burden for management (Allen et al., 2022). In dairy wastewater, the contaminant’s concentration is higher than in municipal wastewater which attributes to its lesser treatability (Healy et al., 2007; Kato et al., 2013). However, the volume of wastewater generated depends on the type and number of cattle handled and the farming practices (Schierano et al., 2020). It was found that a cow excretes 88.2% of daily feed phosphorous, and 50 litres of water is required per day to wash a single cow (Healy et al., 2007; Morse et al., 1992). Uncontrolled Nutrient losses from animal production need management for animals' wellness and prevent environmental sustainability (Tamminga, 2003).

Table 1
Source and occurrence of nutrient
Sl. No.	Source	Origin	Type of Nutrient	Wastewater Characteristics	References
1.	Fertilizers residual loads	USA	NH₄⁺-N 88 ± 18 mg/l, NO₃⁻-N 5.3 ± 2.3 mg/l, TP 39 ± 7 mg/l	Agriculture wastewater with high N, P, K	(Rozema et al., 2016)
2.	Agriculture wastewater	China	TN 258 kg/ha, PO₄³⁻-P 101 kg/ha	High COD	(Wang, 2006)
3.	River water, outlet from a hydropower plant	China	TN 106 ± 62 µM, TP 3.84 ± 2.4 µM	Warm with nutrient and sediment accumulation	(Wang, 2020)
4.	Livestock feed Waste	Netherland	NH₄⁺-N 5.1 mg/l	High ammonia	(Heal, 2014)
5.	Livestock’s waste	Gulf of Mexico	TN 650 mg/l, TP 380 mg/l	More nutrient	(Othman et al., 2013)
6.	Fish feed in aquaculture	Saudi Arabia	TN 22.1 ± 2.1 mg/l, NO₃⁻-N 17.4 ± 0.6 mg/l, TP 0.85 ± 0.3 mg/l	The nutrient concentration depends on fish stocking density	(Siddiqui and Al-Harbi, 1999)
7.	Dairy wastewater	USA	NH₄⁺-N 260 mg/l	Nutrient-rich wastewater with metal loading	(Lee et al., 2010)
8.	Tannery wastewater	Bangladesh	NH₄⁺-N 111 mg/l, NO₂⁻-N 4.3 mg/l, NO₃⁻-N 66 mg/l, PO₄³⁻-P 30 mg/l	Low DO (0.02 mg/l) and high NO₃ and PO₄ pollutant	(Saeed et al., 2012)
9.	Slaughterhouse	New Zealand	NH₄⁺-N 30 mg/l, TP 15 mg/l	Effluent consisted of a highly variable range of COD and TSS	(Russell et al., 1993)
10.	Metallurgical industry	Argentina	NH₄⁺-N 2.1 mg/l NO₃⁻-N 16 mg/l, NO₂⁻-N 0.745 mg/l, TP 0.189 mg/l	Wastewater with Mg, Ca, Na hardness	(Maine et al., 2007)
11.	Coffee processing wastewater	Brazil	TN 89.6 ± 11.9 g/m³, NH₄⁺-N 54.6 ± 9.8 g/m³, NO₃⁻-N 0.6 ± 0.3 g/m³	Very high COD wastewater with phenolic compounds and nutrients	(Rossmann et al., 2012)
		India	NO₃⁻-N 147 mg/l, PO₄³⁻-P 84 mg/l	The high content of TS and TDS was observed	(Sahana et al., 2018)
12.	Olive mill wastewater	Morocco	TKN 2.05 g/L, TP 0.30 g/L	Olive mill wastewater with phenolic compounds, and nutrients	(Achak et al., 2009)
		Greece	NH₄⁺-N 123 mg/l, PO₄³⁻-P 95 mg/l	Wastewater had COD, phenols, OP, ammonia and TKN	(Herouvim et al., 2011)
13.	Industrial wastewater (Mixture of metal, paper and textile effluents)	Bangladesh	NH₄⁺-N 10.5 mg/L, NO₂⁻-N 0.01 mg/, NO₃⁻-N 0.1 mg/l, TN 31.3 mg/l, TP 2.3 mg/l	Colorful wastewater with very high COD, BOD loading with nutrient and Sulphur	(Saeed and Khan, 2019)
14.	Sugar industries wastewater	Ethiopia	PO₄³⁻-P 5.9 mg/l	Highly polluted with nutrients and many heavy metals	(Sahu, 2017)
15.	Winery industry	Spain	NH₄⁺-N 28 mg/l, PO₄³⁻-P 2.3 mg/l	High BOD, COD, and TSS concentration	(Serrano et al., 2011)
16.	Sugarcane processing	Brazil	PO₄³⁻P 67.7 mg/l	Very high BOD and COD with PO₄-P	(Grismer and Shepherd, 2011)
17.	Pulp industry	India	TN 9.932 ± 1.87 mg/l, TP 30.25 ± 3.28 mg/l	Effluent had an acidic pH and high solids content	(Usha et al., 2016)
18.	Pig farm	Japan	NH₄⁺-N 1199 mg/l, TN 1371mg/l, TP 132 mg/l	The wastewater contained high COD and BOD ranges	(Kato et al., 2013)

Sugarcane processing and the winery industry generates wastewater with a high concentration of nutrients (Table 1). It was reported that for 1 litre of wine, 1.6 to 2 litres of wastewater is generated (Fernández et al., 2007). Complex wastewater generated from the meat processing industry contains a high level of nitrogen and other contaminants, wherein the wastewater of the slaughter house industry was found to contain nutrient concentration (NH₄⁺-N 30 mg/l, TP 15 mg/l) (Russell et al., 1993). Other agro-processing industrial effluents also have high nutrient contaminants, i.e., coffee processing, olive mill, pulp industry, and pig farm (Table 1). Some other industries, such as the metallurgical industry, also create an environmental problem by releasing heavy metals (Zn, Cr, Ni) and nutrients (Maine et al., 2006). Oil-producing industries discharge huge wastewater because of heavy oil production, which contains nutrients with a high organic load that amplifies treatment vulnerability (Ji et al., 2002). The tannery industry also releases complex wastewater due to the use of heavy metals, such as chromium, cadmium, arsenic, etc., along with high nutrient concentrations (NH₄⁺-N-100 mg/l, and TP- 44 mg/l) (Calheiros et al., 2007; Saeed et al., 2012).

Many researchers have found that even rivers and streams are not immune to nutrient pollution whereby the nutrient input affects the aquatic environment quality, such as the development of excess phytoplankton in the water. Excessive enrichment of aquatic and terrestrial ecosystems due to nutrient augmentation is called eutrophication and it is considered the most challenging environmental problem worldwide (Schoop et al., 2004). According to some studies, nutrient-rich wastewater may result in detrimental impacts on environmental and human health, and create eutrophication or algal bloom problems (Le Moal et al., 2019; Meena et al., 2019; Sharma and Singhvi, 2017; Ye et al., 2020). There are several ways by which nutrients can stimulate or enhance impacts of toxic species, such as due to nutrient enrichment there is an increase in the amount of harmful phytoplankton biomass, especially there may be a dominance of one species or a group of species creating ecological imbalance. Benthic flora also starts excessive growth causing the anaerobic condition in the system (Anderson et al., 2002). While the mechanism of eutrophication is not entirely determined but it was reported that the nutrient loading in water bodies is the most compelling reason (Yang et al., 2008). Eutrophication directly impacts the water quality with time, in the hypertrophic stage of the water body. Along with agricultural runoff, urban nutrients i.e. industrial and domestic sewage from human settlement shares a high load of 50% of TP inflow in lakes triggering an algal bloom situation (Schoop et al., 2004). Algal blooms resist light penetration and reduce the growth of the plants in littoral zones, and also affect the consumption of DO (Lehtiniemi et al., 2005). Anderson et al. (2002) reported numerous examples in geographic regions ranging from the largest and second-largest U.S. mainland estuaries (the Chesapeake Bay and tile Albemarle-Panllico Estuarine System) to the Inland Sea of Japan, the Black Sea, and Chinese coastal waters, where increases in nutrient loading have been linked with the development of large biomass which create the anoxia due to algal bloom and may have a toxic effect on the ecosystem. Eutrophication not only affects the water quality of a water body but its primary functions, such as ecological support to water animals, are also influenced. Due to toxic cyanobacterial blooms, the taste and odour problems in lakes, rivers, etc., are created and results in drinking water problems (Dodds et al., 2009). Other than drinking water, it also affects recreational angling, boating activities, swimming, and fishing in the water bodies, which affects the sustainable development of the economy and society (Pretty et al., 2003; Yang et al., 2008). Thus, the treatment of this nutrient-rich wastewater is a prerequisite to prevent degradation and destruction of the downstream surface water in the aquatic system. A large number of technologies are available for the treatment of nutrient loaded wastewater, but suitable technology is required which is more resilient, cost-efficient and effectively remove contaminants.

Various techniques available for nutrients removal comprises conventional technologies, such as ASP, MBR, MBBR, Extended aeration etc., while advanced techniques of nutrient removal include membrane distillation, forward osmosis, electrodialysis, membrane photobioreactor, etc. (Ye et al., 2020). Detailed analysis of different types of nutrient removal techniques and their removal efficiencies with wastewater characteristics and hydraulic loading rate has been shown in Table 2. Although these physicochemical and biological nutrient removal techniques are efficient, they have a complex design, energy-consuming mechanical equipment, and are also responsible for environmental pollution with extensive use of fossil fuels and other energy sources (Lee et al., 2009). These technologies can be broadly classified as physicochemical and biological processes.

Table 2
Nutrient removal by various biological and physicochemical processes
Treatment Method	Wastewater characteristics	HRT (H)	Initial Concentration (mg/l)	Removal / Recovery Efficiency %						References
				NH₄⁺-N	NO₂⁻-N	NO₃⁻-N	TN	PO₄³⁻-P	TP
Forward osmosis (FO)	Treated municipal wastewater from MBR technology	-	NH₄⁺-N: 15.8 NO₃⁻-N: 1.9, NO₂⁻-N: 2.6, PO₄³⁻-P: 3.4	66.7	76	58	-	92	-	(Xue et al., 2015)
Membrane distillation (MD)	Synthetic wastewater	1.75	NH₄⁺-N: 3500	99.5	-	-	-	-	-	(Qu et al., 2013)
Electrodialysis (ED)	Urine	-	NH₄⁺-N: 4850, PO₄³⁻-P: 230	92.57	-	-	-	73.91	-	(Pronk et al., 2006)
Bioelectrochemical system	mixed activated sludge and digested sludge from Wastewater treatment plant	10.4	NH₄⁺-N: 90	96	-	-	63.8	-	-	(Zhang and He, 2012)
Membrane Photobioreactor	secondary effluent of WWTP	96	TN: 24.7, TP:3.5	-	-	-	95.95	-	70	(Sheng et al., 2017)
Osmotic Membrane bioreactor (OMBR)	Activated sludge from a bioreactor plant	15.4 -22.6	NH₄⁺-N: 60.0 PO₄³⁻-P: 8.0	90	-	-	-	95	-	(Qiu and Ting, 2014)
Struvite Magnesite 85% H₃PO₄	Anaerobic digestate of piggery wastewater	6	PO₄³⁻-P: 161, NH₄⁺-N: 985	80	-	-	-	96	-	(Huang et al., 2011)
Vacuum thermal stripping acid absorption	Anaerobic digester effluent	3	NH₄⁺-N: 2915	99.9	-	-	-	-	-	(Ukwuani and Tao, 2016)
Struvite recycling	Swine wastewater	-	NH₄⁺-N: 378, PO₄³⁻-P: 105	91	-	-	-	97	-	(Huang et al., 2015)
Stripping & acid absorption	Urine	5	PO₄³⁻-P: 311 TN: 4515	-	-	-	91	98	-	(Antonini et al., 2011)
Struvite crystallization in SBR and continuous-flow reactor	Anaerobically digested swine wastewater	6–15	NH₄⁺-N: 706, PO₄³⁻-P: 40.3	90	-	-	-	-	85	(Song et al., 2011)
Microbial Nutrient Recovery Cell (MNRC)	Synthetic wastewater	24	NH₄⁺-N: 23.8, PO₄³⁻-P: 6.4	96	-	-	-	64	-	(Chen et al., 2015)
MBR	Sewage wastewater	26.47	TN: 147.76, TP: 13.83	-	-	-	65.17	-	50.65	(Leyva-Díaz et al., 2013)
MBBR	Sewage Wastewater	26.47	TN: 147.76, TP: 13.83	-	-	-	67.34	-	48.31	(Leyva-Díaz et al., 2013)
MPC & IE	Raw sewage	24	NH₄⁺-N: 27.4, TN: 34.5	74.4	-	-	87.4	-	-	(Gong et al., 2017)
RRMFC	Synthetic Wastewater	72	TN: 20.20, PO₄³⁻-P: 0.90	-	-	-	98	99	-	(Lu et al., 2019)
Hybrid MBBR	Sewage wastewater	5	NH₄⁺-N: 37, PO₄³⁻-P: 8,	93	-	-	-	60	-	(Hasan et al., 2019)
UASB + Aeration + PP	Sewage wastewater	9.24	NH₄⁺-N: 53, PO₄³⁻-P: 6	85	-	-	-	60	-	(Hasan et al., 2019)
UAFB–EGSB Biosystems	Sewage wastewater	6	NH₄⁺-N: 43, TN: 45,	92.5	-	-	73.2	-	-	(Gao et al., 2015)
SBR	Digested piggery wastewater	1	NH₄⁺-N: 900, PO₄³⁻-P: 90	99.8	-	-	-	97.8	-	(Obaja et al., 2005)
FBBR	Synthetic dairy effluent	28.7	NH₄⁺-N: 160.6, NO₃⁻-N: 306.6,	55.8	-	99.2	-	-	-	(Hamdani et al., 2020)
BAF	Domestic wastewater	1	NH₄⁺-N: 49.7, TN: 51.4, TP: 5.57	95	62.75	-	-	-	30.9	(Wang et al., 2015)
BAF& Chemical precipitation	Domestic wastewater	1.25	NH₄⁺-N: 49.7, TN: 51.4, TP: 5.57	93.83	93.31	-	-	-	30.9	(Wang et al., 2015)
PBR	Anaerobically treated black water	72	TN: 111.6, TP: 15.5	-	-	-	66	-	74	(Slompo et al., 2020)
MBBR	Secondary municipal wastewater	5.3	NH₄⁺-N: 63, TN: 68, TP: 9.7	97.2			93.6		92.5	(Almomani and Bohsale, 2020)
MBR	Sewer wastewater		NH₄⁺-N: 35.3, NO₃⁻-N: 0.18 TP: 4				89		88	(Monclús et al., 2010)
(FO: Forward osmosis, MD: Membrane distillation, ED: Electrodialysis, MPBR: Membrane Photo-bioreactor, OMBR: Osmotic Membrane bioreactor, SBR: Sequencing batch reactor, MNRC: Microbial Nutrient Recovery Cell, MBR: Membrane bioreactor, MBBR: Moving bed biofilm reactor, MPC: Membrane-based pre-concentration, IE: ion exchange, RRMFC: Resource recovery microbial fuel cell, UASB: Up-flow anaerobic sludge blanket, PP: Polishing Ponds, UAFB: Up-flow anaerobic fixed bed, EGSB: Expanded granular sludge bed, FBBR: Fixed bed bioreactor, BAF: Biological aerated filter, PBR: Photo-bioreactor

3.1. Physico-chemical techniques

Physico-chemical techniques have shown good potential to treat nutrient-rich wastewater in recent years, wherein forward osmosis (FO), reverse osmosis (RO), nanofiltration (NF), electro-dialysis (ED), and membrane distillation (MD) processes etc., are the most prevalent technologies (Xie et al., 2016). Due to the same ionic properties in the FO process, the electrostatic repulsions may generate between PO₄³⁻-P ions and membrane, which results in the retention of phosphorus on the feed side (Ye et al., 2020). The difference between FO and RO is that in the RO process, the feed solution is kept at a higher pressure than osmotic pressure (Patel et al., 2020). Due to pressure differences, nutrients are retained on the feed side, and clean water is obtained. (Van Voorthuizen et al., 2005)Van Voorthuizen et al. (2005) obtained a very high nutrient removal efficiency of 80% and 90% for NH₄⁺-N and PO₄³⁻-P respectively, from RO technology. NF is a process capable of efficient removal of dissolved solutes, including multivalent ions with lower pressure at higher flows than RO (Andrade et al., 2014). In this process, a semipermeable membrane is used to separate the nutrients from wastewater by applying pressure on the feed side. The MD process is based on the volatile property of NH₄⁺-N, where the feed solution is heated moderately and then passed through a membrane, where NH₄⁺-N is separated due to its more volatile nature compared to water (Qu et al., 2013). The solution of NH₄⁺-N recovered from this process can be utilized as fertilizer to reduce the economic loads in the agriculture sector (Qu et al., 2013). In the ED processes, the ion exchange membranes are alternatively arranged into a direct electric current field to separate the nutrients according to their electrostatic properties (Tran et al., 2014). This process can be utilized to remove phosphorus selectively from the solution with multiple anions (Zhang et al., 2013). Although the ED process has some advantages, membrane fouling is one of the biggest drawbacks as it results in a decrease in efficiency, current flow, and overall performance (Mondor et al., 2009).

Some other physicochemical processes for nutrient removal include flocculation, precipitation, adsorption, ion exchange, air stripping, and supercritical water oxidation (Adam et al., 2019; Hasan et al., 2021; Yuan et al., 2016). The flocculation method uses different kinds of flocculants or coagulants, in which contaminants aggregate in the culture medium and settle down by forming flocs (Aguilar et al., 2002). Magnesium and calcium-based materials are generally used in the chemical precipitation process, which reacts with nitrogen and phosphorus and forms struvite (MgNH₄PO₄·6H₂O) and hydroxyapatite (Ca₅(OH)(PO₄)₃) precipitate, respectively (Banu et al., 2019; Liu et al., 2019; Ye et al., 2017). These precipitation reactions are presented in Eqs. (1) and (2) (Ye et al., 2020):

$${Mg}^{+2}+{{PO}_{4}}^{-3}+{{NH}_{4}}^{+}+{6H}_{2}O\underrightarrow{ }Mg{NH}_{4}{PO}_{4}.6{H}_{2}O\downarrow$$

$${5Ca}^{+2}+{{3PO}_{4}}^{-2}+{OH}^{-}\underrightarrow{ }{{Ca}_{5}\left(OH\right)\left({PO}_{4}\right)}_{3}\downarrow$$

Struvite precipitation is the most popular and acceptable method for nutrient removal and this precipitate can be used as a fertilizer in agriculture fields (Xie et al., 2016). Struvite and hydroxyapatite can also be utilized as the raw material in the phosphate industry (Stolzenburg et al., 2015). The pH value and the ratios of N:Mg:P and Ca:P in the struvite precipitation process depends on the composition of wastewater (Ye et al., 2020). It was reported that for best nutrient removal efficiency, the ratio of N:Mg:P was 1:1.2:1.2 (Siciliano et al., 2020). However, It was reported that for the wastewater with low phosphorus concentration, recovery of struvite is less, or reaction time is longer (Jaffer et al., 2002).

Adsorption is another simple technique for nutrients removal in which a nutrient-rich solution is generated as a byproduct that could be used as a bio-compost (Sengupta et al., 2015). Metals, such as iron, magnesium, and calcium-based adsorbent are efficient in phosphorus removal (Ma et al., 2020; Weng et al., 2008). Ion exchange is another conventional nutrient removal technology, where the contaminants are adsorbed using ion-exchange resins (Liberti et al., 1981). Ion exchange is reported to have many advantages i.e. handling shock loadings, operating over a wider range of temperatures, selective contaminant removal, cost-effective and very high nutrient removal efficiency (Johir et al., 2011). In the air stripping process for removal of NH₄⁺-N, it is converted into NO₃⁻-N or NO₂⁻-N before the removal process. In the process, NH₄⁺-N, which is a weak base, reacts with water to form ammonium hydroxide (Adam et al., 2019; Yuan et al., 2016). Some alkali substances are added to wastewater during treatment, and the pH of the solution is maintained above 8.5 for easier ammonia stripping (Folino et al., 2020). This process involves the advantages of being relatively simple, no backwash or regeneration process is needed; thus, ammonia stripping is used for NH₄⁺-N recovery in the industries (Adam et al., 2019; Yuan et al., 2016). Supercritical water oxidation is a nitrogen removal process in which NH₄⁺-N is oxidized into NO₂⁻ -N at critical temperature (≈ 600 ^οC) and pressure (23 MPa) (Killilea et al., 1992). This process produces molecular nitrogen and water as a product, but due to the high cost, temperature, pressure and energy requirements, the use of supercritical water oxidation is limited (Adam et al., 2019). Another group of technologies for the treatment of nutrient-rich wastewater is the biological system, which has been discussed in the next section.

3.2. Biological processes

The biological nutrient removal technique is one of the popular options for effective nutrient removal (Hu et al., 2012). Moving bed biofilm reactor (MBBR), Membrane bioreactor (MBR), Sequential batch reactor (SBR), Up-flow anaerobic sludge blanket (UASB), Anaerobic ammonium oxidation (ANAMMOX), Completely autotrophic nitrogen-removal over nitrite (CANON) are the main biological techniques. MBBR is a biofilm-based technique in which biomass is developed on small carriers and move with water (Osmani et al., 2021). Several advantages have been observed in the MBBR system, such as simple operation, low maintenance costs, higher removal efficiency; but this technology needs aeration which involves higher energy requirements (Han et al., 2021). It was also reported that in MBBR, NH₄⁺-N, TN, and TP removal efficiency was 97.2%, 93.6% and 92.5%, respectively (Almomani and Bohsale, 2020). MBR is a membrane-based biological nutrient removal technique that is a combination of both biological treatment and membrane separation. This technique is efficient in the treatment of both nitrogen and phosphorus (Abegglen et al., 2008). In an MBR system, TN and TP removal efficiency were 89% and 88%, respectively (Monclús et al., 2010). SBR is another type of continuous biological nutrient removal technique based on the activated sludge process. This technique is popular due to design and operational flexibility, and better process control (Dutta and Sarkar, 2015). It was found that in an SBR, NH₄⁺-N, and PO₄³⁻-P removal efficiencies were 99.8%, and 97.8%, respectively. UASB is a popular anaerobic wastewater treatment method designed to handle high loading rates with low HRT (Tian et al., 2015). However, some studies found the UASB reactor to be less efficient for nutrient removal (Lettinga and Hulshoff Pol, 1991). Elmitwalli et al. (2007) achieved only 4% and 24% average removal efficiency for NH₄⁺-N and PO₄³⁻-P respectively while treating grey water in a UASB reactor at an HRT of 20 hr and temperature 18^oC. The activated sludge process is a biological PO₄³⁻-P recovery process in which polyphosphoric accumulating organisms (PAOs) are responsible for the deposition of phosphorus as a sludge product (Hu et al., 2012; Yu et al., 2013). For enrichment of these PAOs, alternative aerobic and anaerobic conditions are required, and these conditions are obtained by temporal separation in sequencing batch reactors or spatial separation in continuous reactors (Nancharaiah et al., 2016). In anaerobic conditions, the stored poly-P and glycogen are co-degraded and get detached from the cells which result in the accumulation of PO₄³⁻-P ions in wastewater. Moreover, the organic carbon from wastewater gets removed and stored as intracellular polyhydroxyalakanoates (PHAs) (Ye et al., 2020). While in aerobic conditions, the detached intracellular poly-P and intracellular glycogen reserves are restored due to oxidation of PHAs (Zuthi et al., 2013). As a result, net PO₄-P is removed from wastewater as the amount of PO₄³⁻-P ions released in wastewater during the anaerobic phase is less than the PO₄³⁻-P ions are taken up during the aerobic phase. The energy required for the biotransformation of organic compounds during the anaerobic phase is obtained by glycolysis of intracellular glycogen and the breakdown of poly-P, releasing PO₄³⁻-P ions into wastewater (Zuthi et al., 2013).

In this process, the Mg²⁺ and K⁺ could be released into the wastewater and can store energy in the form of poly-β-hydroxy alkanoates (PHAs). In excess biological phosphorus removal technology, two anaerobic and aerobic chambers are used. In the process, PO₄³⁻-P ions are released in the anaerobic chamber while their uptake is done in the aerobic tank and then removed as sludge (Pastor et al., 2008). ANAMMOX is a nitrogen removal process, where the NH₄⁺-N is oxidized by autotrophic Anammox bacteria (brocadiales) to dinitrogen gas under anoxic conditions (Jin et al., 2012). Advantages of this technique are higher nitrogen removal rate, lower operational costs, and smaller space requirements, but due to the slow-growth rate of the anammox bacteria and toxicity issues, instead of using a standalone system, this process is mainly incorporated with other methods (Jin et al., 2012; Joss et al., 2009). CANON is similar to the ANAMMOX, in which NH₄⁺-N is oxidized by anaerobic NH₄⁺-N oxidizing bacteria (Third et al., 2001). This process is dependent on the harmonious and balanced interaction between the aerobic and the anaerobic ammonium oxidizing bacteria (Third et al., 2001). Although this process is a cost-effective technique, it is very sensitive to DO and nitrogen loading rates (Hasan et al., 2021). Ecological treatment systems, such as CW are natural ways for nutrient removal that are being widely utilized for wastewater treatment (Méndez et al., 2009). These systems do not require skilled labour, high initial, operation cost, and regular monitoring, which might be advantageous for field-level applications (Jain et al., 2020). Therefore, CW is a popular alternative to conventional nutrient removal techniques (García et al., 2020; Vymazal, 2011).

CWs might be defined as engineered systems copied from natural wetlands that have been designed and constructed to utilize the natural activities involving wetland vegetation, substrate, and their associated microbial accumulation to treat nutrient-rich wastewater (Vymazal, 2007). These are artificial wastewater treatment units with shallow ponds or channels having organized plantation, which utilizes the natural biological, physical, and chemical processes to treat wastewater simulating natural wetlands within a more controlled environment (Dipu et al., 2011; Valipour and Ahn, 2016). CWs systems are fabricated and designed to get better water quality with relatively low external energy requirements, and easy operation and maintenance. This natural means of treating wastewater also offers the potential of multiple benefits, such as a source of a recreational system providing aesthetic qualities, wildlife habitats, and the superior quality effluents that can be recycled for land-scape irrigations (Iamchaturapatr et al., 2007).

The primary classification is based on the growth type of macrophytes (free-floating, floating leaves, emergent plants, submerged plants), and further classification is usually based on the water flow regime which is also depicted in Fig. 2 (Ingrao et al., 2020). According to the flow of wastewater, the CW is classified into two types, i.e., FWSCW and SSFCW (Fig. 2). The SSFCW is further classified into HSSFCW and VSSFCW. These are the most prevalently used system configurations, highlighted in (Fig. 2). The hybrid CW is the two-stage or multistage combination of HSSFCW and VSSFCW (Vymazal, 2013). The hybrid CW shows a higher nitrogen removal efficiency than the single standalone systems because it accommodates better nitrification and denitrification processes (Saeed et al., 2012; Serrano et al., 2011; Vymazal, 2005).

4.1. Parameters affecting nutrient removal efficiency of CW

The effectiveness and performance of CW technology depend on some associated factors, such as macrophytes species, microbial activity, substrate, type of wetland, DO, temperature, pH, and hydraulic loading rate, etc. (Fig. 3). The magnitude and condition of these factors at various treatment stages influence the output efficiency of technology. The factors could be divided into three categories, biotic, abiotic, and environmental factors. Biotic factors are plants and microorganisms, while abiotic factors are substrate, precipitation, evaporation, etc., used in CW. DO, temperature, and pH are the main environmental components in a CW (Jingyu et al., 2020; Sindilariu et al., 2009). Many biological processes of CW are affected by DO and temperature (Oliver et al., 2017).

4.1.1. Dissolved oxygen and artificial aeration

Nitrification or denitrification processes can be controlled by the oxygen level in the CW (Nuamah et al., 2020; Wu et al., 2014). The presence of oxygen is affected by plant respiration and microbial activities (Shelef et al., 2013). NH₄⁺-N to NO₃⁻-N conversion rate increase significantly by continuous oxygen supply or aeration. Sometimes an electromechanical device is used for a continuous supply of air (Foladori et al., 2013). Fan et al., (2013) found that TN and NH₄⁺-N removal efficiency was 63% and 58%, respectively without aeration, and 82% and 96% respectively, with artificial aeration and step feeding strategy in VSSFCW due to the formation of alternative aerobic and anaerobic zones and increase in the heterotrophic bacterial activity. Different types of aeration systems were used in the CW, such as step feeding, continuous aeration, intermittent aeration (Fan et al., 2013). It was reported that CW operating with intermittent aeration followed by step feeding resulted in removal efficiencies of NH₄⁺-N and TN to be 96% and 82%, respectively, at 14.7°C (Ji et al., 2020). Recently, the use of artificial aeration is becoming popular in CW technology (Jizheng et al., 2019). In the HSSFCW, oxygen depletion is more common, and denitrification is the dominating process (Xinshan et al., 2010). Artificially aerated CW achieved significantly higher NH₄⁺-N, and TN removal efficiency, with elevated microbial abundance than non-aerated ones (Yang et al., 2018). However, the removal efficiency of TP was not very much affected by the artificial aeration. Rossmann et al., (2012) reported that TN removal is enhanced by 12% while, TP increased only by 6% with artificial aeration. It was also reported that for 0.6 L/min aeration rate for 2 h/day, the removal rate of TN, NH₄⁺ N, and COD were 78%, 97%, and 96%, respectively (Zhou et al., 2018).

4.1.2. Temperature

The treatment performance of CW is affected by temperature variations. It affects both the biological reactions and physical processes in the wetland systems. Many biological processes regulate the removal of nutrients, which are controlled by temperature, affecting overall efficiency (El-Refaie, 2010). Evaporation water loss affects nutrient removal because the chemical budget is associated with the water budget in a CW (Lott and Hunt, 2001). El-Refaie (2010) found a direct relationship between temperature and phosphorus removal. It was reported that TP removal efficiency in winter was 10%, while in summer, it was 66.7%. A similar trend was also obtained for nitrogen removal (Mesquita et al., 2017). Nitrogen removal is more influenced by temperature variations than phosphorus; as most phosphorus removal mechanisms depend upon the physicochemical and sorption processes. In contrast, nitrogen removal depends on microbial activities as the microbes are temperature sensitive (Picard et al., 2005; Spieles and Mitsch, 1999). Temperature also affects oxygen availability or indirectly decreases the redox process (Bachand and Horne, 1999; Kadlec, 1999). Redox processes are responsible processes for phosphorus adsorption on ferrous and ferric oxide, which results in a decrease in the removal efficiency of phosphorus with decreasing temperature (Wittgren and Maehlum, 1997). This study observed that in the summer season (temperature range 23^oC − 29^oC) the removal efficiency of NH₄⁺-N, TN, NO₃-N, and TP increased by 25.4%, 28.16%, 28.5% and 28.35% respectively when compared to winter season summer season (temperature range − 11.1^oC to 14^oC) (Supplementary Table A.2).

4.1.3. pH Level

The performance of CW is affected by pH as it directly or indirectly influences the biotic and abiotic factors of the CW. The organic nitrogen conversion processes i.e. ammonification, nitrification and denitrification are pH-dependent reactions (Parde et al., 2021; Vymazal, 2007). Intracellular metabolic activity, cell growth, and biomass may be affected by pH (Çelen et al., 2007; Ranieri et al., 2013). Even the increased pH level can reduce the oxygen concentration in sediments (Yin et al., 2016). It was reported that the TN removal was decreased from 76.3 to 51.8% at stress pH level (from 7.5 to 10.5) and the NH₄⁺-N and NO₃⁻-N removal were also adversely affected. The stress in pH was due to plant assimilation and decay in the summer season (Yin et al., 2016). In summer, the pH level of submerged planted CW increased rapidly, creating stress for emergent plants (Mjelde et al., 2012). The pH level is high in the summer season because of the intensive photosynthesis activity of submerged plants and high oxygen transfer from macrophyte to CW (Yin et al., 2016). It was reported that the heterotrophic bacterial activity was the highest at neutral pH and reduces with pH variations (Meng et al., 2014). The optimal pH range for nitrifying bacteria was 7 to 8 (Antoniou et al., 1990; Painter and Loveless, 1983; Paredes et al., 2007). The pH range of 6 to 8 is ideal denitrification activity; however, the reaction rate also decreases when the pH reduces below 5, and even it becomes negligible below pH 4 (Vymazal, 2007). It was reported that at a higher pH level of approximately 10 to 11 plant growth rate started reducing (Hadad et al., 2018). The protoplasm of many vascular plants' root cells was seriously damaged at pH above 9 and in low pH (4.3 to 6.2), the free-floating species were about to die (Akçin et al., 1993; Dyhr-Jensen and Brix, 1996). T. latifolia showed a relatively low growth rate at low pH of 3.5 as compared to 5, because at pH of 3.5, the H⁺ ions concentration increased, and the electrochemical gradient decreased across the plasma membranes of the root cell (Dyhr-Jensen and Brix, 1996). It was reported that the soil, plant, microorganism, and phosphorus release was also influenced by changes in pH (Kim et al., 2016).

4.1.4. Substrates

Soil, rock, gravel, and organic materials are common substrates used in CW (Ballantine and Tanner, 2010). These substrates can be divided into three types concerning their origin, viz. naturally occurring materials (soil, sand, gravel, etc.), processed materials (alum, amended zeolite, ceramics etc.), and waste materials (fly ashes, slag, etc.) (Ballantine and Tanner, 2010; Dong et al., 2021). Nitrogen removal in CW can be obtained satisfactorily with the help of the microorganism and substrates, but phosphorus removal is obtained mainly through the adsorption process on substrates (Cui et al., 2008). Nguyen et al. (2020) depicted major phosphorus removal by adsorption (77.5%), followed by microbial assimilation (14.5%), plant uptake (5.4%), and other processes (2.6%). (Xu et al., 2006) Xu et al. (2006) reported that the sorption capacity varies with the types of substrate materials. Fine sized substrates have more surface area, which attributes to more phosphorus adsorption capacity; thus, the phosphorus removal efficiency can be increased by choosing a better substrates material and particle size (Cui et al., 2008; Xu et al., 2006). (Lima et al., 2018) Lima et al. (2018), stated that on treating low-strength sewage with E. crassipes, the phosphorus removal efficiency in a CW with gravel substrate was 25%, while with bricks, it was 87%. In another study by Vymazal (2004), it was reported that in the long term, more than 95% of phosphorus is stored in soil or litter only. Other than phosphorus adsorption, the substrates also support living organisms, microbial transportation and regulates the water flow through the CW.

4.1.5. Microorganisms

Bacterial diversity living inside plant tissues is responsible for several physicochemical and biological processes occurring in the system (Rajan et al., 2019). Thus, microbial activity is the critical factor influencing the proper functioning and maintenance of the CW (Ibekwe et al., 2003). It was reported that microbes play an essential role in organic and inorganic pollutant removal and plant growth (Shahid et al., 2020). Microorganisms are responsible for the transformation and mineralization of nutrients (Dinakar et al., 2020; Stottmeister et al., 2003). Various types of processes, such as ammonification, nitrification, denitrification, and nitrogen fixation are depended upon the activity and diversity of the microorganisms (Bañeras et al., 2012; Bell et al., 2005). It was reported that the presence of various types of bacteria, such as Nitrosopumilus, Vibrio, Pseudoalteromonas, Nitrospina, and Planctomyces contribute to effective nutrient removal and plant growth in CW (Ma et al., 2018). Various bacteria are responsible for specific functions, such as Syntrophusand, Syntrophobacter bacteria responsible for nitrification, whereas Pseudomonas, Dechloromonas, Desulfomicrobium, and Desulfobacca are denitrifying bacteria (Li et al., 2020). Lee et al. (2009) reported that a major portion of the nitrogen is removed by the denitrification process, which decreased about 60 to 70% of the nitrogen while around 20 to 30% was removed by plant uptake.

Phosphorus is degraded by microbes in the form of orthophosphate along with many other organophosphorus compounds, while its uptake is limited due to less storage capacity (Singh and Walker, 2006; Vymazal, 2007). Microorganisms attain nutrients from the root exudates of macrophytes which is one of the key factors affecting the removal efficiency of nutrients in CW.

4.2. Assessment of the role of macrophytes in CW

Macrophytes or plants are responsible for providing the necessary components for nutrient removal, such as the development of microbial communities near plant roots and fulfilling the oxygen requirement of wetland (Rossmann et al., 2012). Macrophytes play a vital role in CW processes, especially for nutrient removal via direct plant uptake and provide the substrate for microbial activities (Meng et al., 2014). All microbial activities, density, and diversity are supported by the plant rhizosphere (Carvalho et al., 2014; Truu et al., 2009). The enzyme activities play an important role in microbial production, which are also affected by root and shoot. Planted CWs have more soil microbes and redox activities than unplanted wetlands (Huang et al., 2012; Salvato et al., 2012; Tanner, 2001). Some principal features, such as root morphology, structure, and eco-physiology, are influenced by plants, and these features have an important role in nutrient retention and enzyme activities (Zhang et al., 2010). Even dead plants can produce useful organic compounds, i.e., amino acids, sugars, and volatile fatty acids, that will help in the growth of plants and microbes which increases contaminants removal efficiency (Vymazal, 2007).

It was reported that with the vegetation growth and adequate carbon availability, the denitrification process increased, and phosphorus removal processes also increased (Craft, 1996; Kadlec, 1999). Wetland macrophytes supply oxygen to the root zone which influences the aerobic activities in CW and reduce up to 90% of BOD (Bodelier et al., 1996; Moorhead and Reddy, 1988). Individual plant species have different capacities of oxygen supply to the root zone according to their ability, vascular tissues, root distribution, and associated metabolic activities (Shimamura et al., 2003). The root zone is also known as the rhizosphere, and it is divided into two parts endorhizosphere (the root interior), and the ectorhizosphere (the root exterior). The interaction area of these two zones is known as the rhizoplane, and this is the main place of interaction between the plant and microorganisms (Stottmeister et al., 2003).

Figure 4 summarizes salient features and various roles of macrophytes in CW with nitrogen and phosphorus removal mechanisms. The major role of plants in CW are related to its hydraulic function (control flow), physical functions (fixing substrate, providing shelter to microbes), environmental functions (maintain DO level), and most important treatment of nutrients by plant uptake, degradation, transport, transform, store, and phytovolatilization of nutrients. Sometimes plants also contribute to nutrient input in CW through litter fall, which may also affect the total nutrient input in the CW system (Shelef et al., 2013).

A critical function of plant roots is to maintain sediments' hydraulic conductivity (Brix, 1994; Wittgren and Maehlum, 1997). Other than treatment facilities, it also provides green space and provides shelter to many aquatic animals with a good aesthetic appearance. A study described that temperature and oxygen release in root zones are affected by plant species which alters overall wastewater treatment efficiency (Allen et al., 2002). Vascular plants or non-vascular plants (algae) are an essential part of a CW. Algae enhance the dissolved oxygen of water which indirectly increases the nutrient removal reactions. Plants properly control the water flow and spread it throughout the substrates of the CW. Moreover, they uptake nutrients, carbon, and other contaminants and fix them in their tissues. The roots of the macrophytes provide a surface for microorganism activities. Several factors, such as types of CW, quality, and quantity of wastewater, affect system’s removal efficiencies (Sklarz et al., 2009). Zhou et al. (2017) assessed the effect of vegetation on the treatment of domestic sewage in HSSFCW and found that the NO₃⁻-N removal rate is 99.9% in a vegetative CW, much higher than an unplanted wetland (82.9%).

In a substrate-free CW, Kyambadde et al. (2005) obtained the TN and PO₄³⁻-P removal efficiency in planted CW to be 73.9% and 73.5%, while in an unplanted CW, they were found to be 54.3% and 45.2%, respectively. Evapotranspiration of emergent macrophytes is an essential activity in a CW (Shelef et al., 2013). It was reported that the water losses by evapotranspiration reduce the flow velocities, increase retention times and enhance pollutant concentration in water (Headley et al., 2012). Microclimatic conditions may affect different components of CWs, i.e., the shades of plants can obstruct algal growth, protect from direct radiation, and slow down wind velocity that helps in stabilizing the upper substrate (Shelef et al., 2013). Although plant growth is high in the summer season due to the high level of solar radiation and high relative humidity. In autumn, plant leaves dry and therefore, the overall nutrient removal rate is higher in summer and lower in autumn (Licata et al., 2019). The biomass produced by the plants have various commercial uses and its production is mainly obtained by pruning and harvesting operations of plants and used as fodder for livestock, fertilizer, and sometimes as biofuel (Licata et al., 2019; Liu et al., 2012). The planted wetland may be an effective carbon sink and help in environmental production (Mander et al., 2008; Mitsch et al., 2013). Plants enhance nitrogen removal by providing stem and root surface area for microorganisms and lowering the algal growth (Oliver et al., 2017).

4.2.1. Difference between monoculture and polyculture practices in CW

Effluent quality is affected by the presence of different plant species in the wetland (Iamchaturapatr et al., 2007; Maine et al., 2007). Every plant has a specific capacity that influences nitrogen removal in CW, which is also affected by species richness, i.e. number and type of species (Zhu et al., 2012). It was reported that suitable macrophyte selection could have an important role in improving removal efficiency for different types of contaminants (Brisson and Chazarenc, 2008). Monoculture and polyculture (mixed culture) are the two types of planting practices in CW. Monoculture is the practice of growing a single type of plant species (such as Typha latifolia plantation), while polyculture refers to planting multiple types of plant species (such as Typha latifolia plantation with augustifolia or other plant species) in CW (Fig. 5. a & b).

In most of the CW studies, monoculture is practised, but it was reported that the plant diversity could increase the plant’s effectiveness for better contaminant removal (Liang et al., 2011). Zheng et al. (2016), reported that NO₃⁻-N, NH₄⁺-N, and TP removal efficiency is higher in vegetated CW than unplanted ones. Plant diversity also helps in better system functioning because competition between plant species results in more biomass production, less dominance of a single species, and reduced sensitivity to seasonal changes (Qiu et al., 2011). Better root growth and more microbial diversity were observed in the polyculture (Amon et al., 2007). Due to more release of root exudates in polyculture, the nutrient removal via plant uptake might be enhanced (Marín-Muñiz et al., 2020; Wu et al., 2012). Polyculture has shown higher nutrient removal efficiencies (Supplementary Table A.1), this might be a result of the fact that different macrophyte species have specific nutrient selectivity for plant uptake, growth rates, and microbial processes (Liang et al., 2011; Zhang et al., 2009). However, due to the similar characteristics with comparable growth and feed demand, the competition between different plants in polyculture may be more (Engelhardt and Ritchie, 2001; Zhang et al., 2007b). Liang et al. (2011) found that the plant density in polyculture (123.8 individuals/m²) was higher than in monoculture (71.3 individuals/m²). The same study reported that in the first year of planting, the biomass production in monoculture (2041.8 g/m²) was more than in polyculture (1703.7 g/m²). Moreover, after three years, the result was changed, and biomass production in polyculture (1661.5 g/m²) was higher than in monoculture (1306 g/m²) (Liang et al., 2011). The growth of polyculture is higher because it reduces the negative effect of some species and becomes more resilient to diseases (Liang et al., 2011). Geng et al. (2019) reported that the nutrient selection and the nutrient removal process, i.e., nitrification, denitrification, and plant uptake, also changed with plant species. The study also reported that all types of nitrogen removal efficiencies were better in polyculture than monoculture (Geng et al., 2019).

Figure 5c & d are showing the mean nutrient removal efficiencies for NH₄⁺-N, TN, NO⁻ ₃-N, PO₄³⁻-P, TP in CW in monoculture and polyculture practices. Average nutrients removal rates in HSSFCW for TN, NO₃⁻-N, TP were found to be 88.55%, 84.06%, 74.34%, respectively in monoculture (Supplementary Table A.1). While for VSSFCW in monoculture CW system, nutrients removal rates for TN, NO₃⁻-N, TP were 60.11%, 87.21%, 58%, respectively (Supplementary Table A.1). In the case of polyculture, the average removal efficiencies of HSSFCW were found to be TN, NO₃⁻-N, TP were 95.26%, 89.5%, 79.17% respectively (Supplementary Table A.1). While VSSFCW achieved a removal efficacy of around 81%, 71.6%, and 92.8% for TN, NO₃⁻-N, TP respectively (Supplementary Table A.1).

Nutrients are possible contaminants in most wastewater compositions, such as agro-industrial wastewater, agriculture runoff, municipal wastewater, which have detrimental impacts on biological activities and chemical composition of the waterbody, such as eutrophication. Out of many available nutrient removal techniques, CW technology is found to be one the most promising nutrient removal technology with low cost and less external energy requirement. The performance of CW varies with many factors, such as the design and type of wetland, inflow wastewater characteristics, and other parameters of CW. It is found that the macrophytes are one of the dominating factors that may change the performance of CW, especially for nitrogen removal. Various literature suggests that phosphorus removal mainly depends on substrates and some other removal mechanisms, such as plant uptake and microbial degradation may also play a minor role. Monoculture and polyculture practices are found to have a significant impact on the performance of CW.

The number of studies conducted for comparison between monoculture and polyculture practices is very limited and more research is required in this direction. From this study, it has been found that the average TN removal efficiency in the HSSFCW is higher than the VSSFCW in monoculture CW systems. Additionally, for TN and TP, removal efficiency noticeably increased in the VSSFCW with polyculture. While comparison of monoculture and polyculture shows that the average NH₄⁺-N removal efficiency is almost similar (≈ 81%) but in the case of TN, NO₃-N, and TP efficiency increased in the polyculture system. The average removal efficiency of NH₄⁺-N, TN, NO₃-N, and TP in summer seasons (temperature range 23^oC-29^oC) increased by more than 25%, this may be due to higher microbial activities and chemical reaction rates. It was observed that aeration or enhanced DO level in the system increased the removal efficiency of NH₄⁺-N and TN. However, the TP removal rates were not much affected by the aeration. Bio-augmentation was found to enhance the removal efficiency for all the nutrients in wastewater. The substrate layers also influenced the nutrient removal rates in the CW systems where the CW with activated alumina were found to remove TP by an average of 97.7%, whereas in the normal gravel-based CW systems an average removal of 22.9% was achieved.

The effect of different plant species on CW removal efficiency is determined and well-studied by many researchers and but the role of plants in biological and nutrient removal mechanisms need to be determined. Although it was found that different types of CW are efficient for specific nutrient removal, but further studies are required to find out suitable design depending on various nutrient removal selectivity. This review helps in determining the role of plants for nutrient removal in the CW with the impact of monoculture and polyculture practices. Moreover, the advantages of CW technology over other nutrient removal technology are highlighted. All parameters and components have their significance, while the effectiveness of each component may also be influenced by others.

Ethics approval and consent to participate

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Consent for publication

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Availability of data and materials

The datasets used and analyzed during the current study are available from the corresponding author on reasonable request.

Competing interests

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Funding

This research did not receive any specific grant from funding agencies in the public, commercial, or not-for-profit sectors.

Author’s contributions

Balram Sharma: Conceptualization, Investigation, Data Collection, Writing - original draft, reviewing, and editing. Tuhin Kamilya: Writing- reviewing and editing, Mahak Jain: Writing- reviewing and editing, Supervision. Ashok Kumar Gupta: Conceptualization, Writing- reviewing and editing, Supervision. Partha Sarathi Ghosal: Writing- reviewing and editing, Supervision.

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A review on enhanced nutrient removal in Constructed Wetlands: Insights into operating conditions and plant variation

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Abstract

Figures

1. Introduction

2. Source, Occurrence, And Effects Of Nutrients In Wastewater

3. Different Types Of Nutrient Removal Techniques

3.1. Physico-chemical techniques

3.2. Biological processes

4. Application Of Cw For Nutrient Removal

4.1. Parameters affecting nutrient removal efficiency of CW

4.1.1. Dissolved oxygen and artificial aeration

4.1.2. Temperature

4.1.3. pH Level

4.1.4. Substrates

4.1.5. Microorganisms

4.2. Assessment of the role of macrophytes in CW

4.2.1. Difference between monoculture and polyculture practices in CW

5. Summary Of Findings And Future Perspective

Declarations

References

Supplementary Files

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