Dual-Purpose Biological Systems: Enhancing Wastewater Treatment and Biogas Generation with Duckweed and Microorganisms—A Systematic Review

Abstract

At present, treated wastewater may still contain residual nutrients and micropollutants, including heavy metals, pharmaceuticals, and dyes, which can negatively affect receiving water bodies. Increasingly stringent environmental regulations, including Directive (EU) 2024/3019, require both enhanced removal of these contaminants and greater integration of renewable energy sources in wastewater treatment plants. This paper presents a review of biomass-based wastewater polishing technologies employing biological agents such as microalgae, fungi, bacteria, co-cultures and duckweed for the removal of residual contaminants from treated effluents. The compiled data indicate that while optimal conditions can drive pollutant removal efficiencies beyond 90%, system performance varies widely depending on species selection, wastewater characteristics, and operational conditions (e.g., pH, temperature, salinity, nutrient availability, and light intensity). In addition to effluent polishing, the produced biomass can be valorized for bioenergy generation, contributing to renewable energy production and supporting circular economy principles in wastewater treatment plants. Despite these benefits, biomass harvesting remains a major technical and economic bottleneck, often representing a significant share of operational costs and limiting large-scale implementation. Overall, biomass-based treatment technologies are a promising approach for improving effluent quality and supporting renewable energy objectives; however, further advances in biomass recovery are required for broader application.

1. Introduction

Water pollution has recently become a major problem and priority for society and public authorities [1]. The significant contamination of aquatic systems by micropollutants (MPs) has attracted public attention in recent years (e.g., the removal of MPs like pharmaceuticals or personal care products is not sufficient). Modern lifestyles, urbanization, and population growth have also increased the amount of wastewater produced. Current wastewater treatment plants (WWTPs) are not fully prepared for these new challenges [2,3,4]. As a result, pollutants present in wastewater are often discharged into aquatic environments, including rivers, lakes, streams, and even groundwater, making WWTPs significant pathways for contamination [5,6]. Usually, WWTPs can remove N and P, whereas personal care products, pesticides, endocrine-disrupting chemicals, and pharmaceuticals are present in effluent. Changes in the design and operation of WWTPs are needed to ensure effluent quality [7]. Wastewater, especially from agriculture and households, contains a high amount of P and N [8,9]. The volume and quality of wastewater from households, hospitals, or industry are of concern for humans and the environment [10,11].

WWTPs have common features, but they need to be adjusted to local conditions (like wastewater type, seasonal variations, or the number of connected inhabitants). Commonly, the treatment process consists of the following stages: preliminary, primary, secondary, and tertiary. The first stage is a mechanical purification stage, which removes coarse impurities. The second one is a biological stage involving biological processes, removing organic substances. The third stage is usually chemical purification [2,11].

Numerous techniques have been considered as a potential fourth purification stage. These range from methods targeting nutrient removal (particularly N and P) to advanced techniques aimed at the removal of HMs and micropollutants (MPs) like dyes, or pharmaceuticals [2,11]. However, no single treatment technology has demonstrated universal and highly efficient removal of all contaminant groups. Instead, the performance of advanced treatment processes depends strongly on the type and characteristics of the target pollutants [12]. The European Commission defined this purification stage as the additional, advanced, and final treatment, allowing for the elimination of a broad spectrum of pollutants. Technologies proposed for this treatment stage have also been investigated for the removal of nutrients, organic matter, HMs, dyes, and other emerging contaminants [6,13,14,15,16]. Therefore, this review considers a broad range of polishing technologies and evaluates their performance with different contaminant groups. Throughout this paper, the term “polishing treatment” is used to describe the final wastewater treatment stage, aimed at reducing residual contaminant concentrations prior to discharge or reuse.

Polishing treatment is applied by numerous WWTPs and includes techniques like biological treatment, disinfection, sand filtration, or soil aquifer. Mostly, it is dedicated to removing excess P and N and final treatment; thus, it is crucial in cases when treated effluent is released to eutrophication-sensitive ecosystems. There are some drawbacks related to these treatments, like high energy costs, greenhouse gas release, or sludge formation [17,18].

The selection of the most appropriate treatment technology depends on various factors, such as specific reuse applications, the environmental impact, the desired effluent quality, and capital and operating costs. The wastewater-to-biogas energy recovery process helps to decrease the costs and greenhouse gas emissions [19,20]. Wastewater treatment based on microorganisms gained attention due to its combination of environmentally friendly treatment with enhanced biomass production [3].

In many cities, authorities are facing challenges in wastewater management due to (a) the large wastewater volume; (b) the fact that current WWTPs may not be efficient in removing contaminants of emerging concern; and (c) problems in sewage sludge management [21]. According to the sustainable development objectives, a circularly managed WWTPS should treat wastewater to meet quality standards, reuse treated wastewater, reuse biogas as an energy source, and recycle sludge as a soil conditioner [21]. Insufficiently treated wastewater has a negative impact on the surrounding environment, while wastewater quality standards have become increasingly stringent. Strict quality requirements have led to the need to look for new treatment technologies [19].

Each country has specific requirements for releasing wastewater to the receiving environment [4]. According to the Water Framework Directive (WFD) wastewater in Europe requires further reductions in nitrogen (N) and phosphorus (P) discharges (below the current 1–2 mg/L) to meet the limit set by the Urban Wastewater Treatment Directive (UWWTD)–Directive 2024/3019 of 27 November 2024 [22]. The document concerns urban wastewater treatment. Its main goal is to protect the environment against insufficiently treated wastewater effluent. There are the main principles and several deadlines are described in the Directive regarding the minimum nutrient percentage removal [22,23]. The Directive [22] requires Member States to implement mandatory treatment involving nutrient removal (specifically N and P) for all urban wastewater treatment plants where there is a 150,000 population equivalent (PE) and above by 31 December 2039 because these plants represent an important remaining source of nitrogen and phosphorus discharge. Requirements for micropollutants removal specify the levels of total phosphorus concentration at 0.5 mg/L and total nitrogen concentration at 8 mg/L in treated urban wastewater. However, discharges from WWTPs treating municipal wastewater from agglomerations with a PE of at least 10,000 but below 150,000 are supposed to achieve a biogen concentration in treated effluent of 0.7 mg/L total phosphorus and 10 mg/L total nitrogen by the end of 2045.

The Directive [22] also emphasizes that currently applied treatment processes already remove some MPs to a limited extent; however, additional purification, specifically MPs removal, is necessary to ensure the elimination of a wide range of remaining MPs from municipal wastewater. Consequently, Member States are required to ensure that discharges from wastewater treatment plants treating wastewater with a load of at least 150,000 PE achieves an 80% reduction in MPs.

The Directive [22] sets progressive targets for the use of renewable energy in wastewater treatment plants, requiring this to comprise 20% of total annual energy consumption by 31 December 2030, 40% by 2035, 70% by 2040, and full (100%) coverage by 2045.

One of the possible ways to meet the wastewater quality criteria using a sustainable approach is by applying microorganisms (e.g., microalgae, fungi or bacteria). Such a solution allows for bioenergy generation and a decrease in CO₂ emission, as well as lowering the total process costs [24,25]. Traditional wastewater treatment techniques have some limitations and high operational costs. Systems involving microorganisms have advantages over other methods, such as a compact design, ease of operation, or reduced energy consumption [24].

Recently, wastewater quality guidelines have become stricter. Current treatment techniques are supposed to allow for the removal of nutrients as well as MPs (e.g., heavy metals (HMs) and pharmaceuticals), which are becoming more common in the environment. A sustainable, circular, and eco-friendly solution for wastewater purification is needed. One of the possible ways to meet the new quality limits is the application of biological agents like microalgae, bacteria, fungi, duckweed, or their co-cultures. In comparison to other techniques, these are characterized by various advantages. Wastewater leads to intensive biomass growth and consequent harvesting of the produced biomass, which can be used for bioenergy generation. Unlike dedicated biogas plants, dual-purpose systems utilize anaerobic digestion as a waste-to-energy recovery route for the biomass generated during wastewater treatment.

The novelty of this review lies in its focused, comparative approach, which spans multiple biological agents, ranging from microalgae, duckweed, and fungi to co-cultures. While the existing literature often evaluates these bioremediation systems in isolation, this paper uniquely synthesizes their distinct pollutant removal mechanisms alongside a critical analysis of their operational advantages and disadvantages. Furthermore, this study explicitly bridges the gap between waste mitigation and resource recovery by evaluating biomass yield dynamics and subsequent biogas production potential, offering a holistic roadmap for circular bioeconomy applications.

The paper provides a comprehensive analysis of the application of biological agents in the removal of pollutants (including HMs, nutrient, pharmaceutics, and dyes) from wastewater. It aims to identify the most suitable genera for MPs’ removal, as well as to highlight the challenges and future directions associated with this approach. This article aims to: (i) identify microorganism species for wastewater treatment and characterize the key objectives of each technology, including its advantages and disadvantages, (ii) evaluate the wastewater treatment efficiency through specific microorganisms, (iii) assess the potential of biofuel production by microorganisms, and (iv) establish challenges and directions for future research.

2. Methodology

2.1. Selection Process (Inclusion and Exclusion Criteria)

A multi-stage screening process was used to ensure the relevance and quality of the selected papers. At the beginning, 301 records (in total) were identified through database searches (Web of Science: n = 80, Google Scholar: n = 221). The 20 duplicates were found and removed. The remaining 281 articles underwent an evaluation based on the following steps:

(1)

Initial Title and Abstract Screening: First, articles were identified based on their titles. At this stage, 281 records were comprehensively screened. A total of 166 papers were excluded because they did not focus on biological processes or biomass-related treatments, the full text was not available, or they lacked technical alignment with the scope of this study.

(2)

Full-Text Eligibility Assessment: The remaining 115 articles were retrieved for a full-text review to assess their suitability and compatibility with the article scope. After 29 papers were excluded due to incomplete data, the final selection of 86 studies was established. Articles were included if they provided empirical data on treatment efficiency, biomass growth, biogas yield or essential parameters describing the fundamental mechanisms and parameters of the treatment process.

As this is a structured qualitative review focused on mapping current practices, no formal risk of bias, certainty, or reporting bias assessments were conducted.

The PRISMA 2020 guidelines were followed to improve the transparency of the literature identification and study selection process. The PRISMA checklist is provided in Supplementary Table S2 [26]. The review protocol was not prospectively registered. The literature selection process is illustrated in a PRISMA-style flow diagram (Figure 1), while detailed characteristics of the publications included after the screening process are provided in Supplementary Table S1.

Criteria for exclusion included the following:

• Studies focusing solely on strictly chemical or mechanical treatment methods.
• Papers where the full text was unavailable or lacked peer-review validation, or failed to provide clear descriptive insights into the core process mechanisms.
• Outdated studies that have been superseded by more recent, comprehensive research (focus was placed on the last 10 years). Selected landmark papers or foundational methodological studies older than 10 years (e.g., Abdel-Raouf et al., 2012 [27]) were exceptionally included due to their critical relevance to the field.

2.2. Data Synthesis

The selected literature was categorized based on the type of microorganism used (algae, fungi, or bacteria) and the end-product of the treatment process (e.g., treated effluent, biomass for feed, or biogas for energy). Data from the included studies were extracted and organized according to the type of microorganism used, which served as the basis for the thematic synthesis and comparison of treatment approaches. Additionally, the operational parameters and core biological mechanisms reported in the studies were synthesized to enable a critical comparison of different biotechnological processes in modern wastewater treatment. A narrative synthesis was conducted with thematic grouping of studies. This approach was chosen due to the heterogeneous and qualitative nature of the available data, which did not allow for statistical pooling or meta-analysis.

3. Plant and Microbial Candidates for Wastewater Treatment Characterization

3.1. Microalgae General Objectives

Microalgae remove pollutants in different ways: adsorption, biodegradation, or bioaccumulation (all processes can occur singly or simultaneously) [27,28] [They eliminate organic and inorganic compounds, and their life-cycle is short. However, they can reproduce throughout the whole year. Algae work efficiently in anaerobic as well as aerobic conditions [29]. Municipal and animal manure wastewater are the two most commonly applied substrates for microalgae cultivation [28].

Aquatic treatment systems are made of one or more shallow ponds in which water-tolerant organisms are cultivated and then harvested. Microalgal cultures offer a means of enhancing nutrient removal due to their ability to use inorganic N and P for growth. In addition, they are capable of removing HMs [28,30]. Different microalgae, like Chlorella Chlorococcum sp., Pseudokirchneriella subcapitata, Coelastrum sp., and Nannochloropsis gaditana, Dunaniella, Scenedesmus sp., Phormidum, Botryococcus, Chlamydomonas, and Arthrospira, Oedogonium [25], were shown to be efficient in wastewater treatment. In order of abundance, algae found in water ponds can be listed as Chlorella, Ankistrodesmus, Scenedesmus, Euglena, Chlamydomonas, Oscillatoria, Micractinium, and Golenkinia [27]. In addition, tolerance to organic pollutants varies from species to species—Euglena, Oscillatoria, Chlamydomonas, Scenedesmus, Chlorella, Nitzschia, Navicula, and Stigeoclonium have been shown to be most resistant to organic pollutants [25,27,30,31,32].

Some efforts are being made to develop hyperconcentrated algal cultures (algal biomass concentration higher than 1.5 g/L). They were proven to be highly efficient in N and P removal in a short time (even less than 1 h). Enteromorpha, Cladophora, Sphacelaria, Ectocarpus, Ceramium, Polysiphonia, Herposiphonia, and Oscillatoria may be applied in systems made of algal mat [27]. If CO₂ availability was too low, this was shown to decrease algae development. ${\mathrm{N}\mathrm{H}}_4^{+}$ concentrations exceeding 100 mg/L or 30 mg/L (at pH 9) inhibit bacterial growth. N, P, and micronutrient presence are crucial for algae—their limitation may significantly inhibit growth. The C:N ratio in domestic wastewater may not be high enough [25,31]. For some algae, N concentration is classified as a limiting factor for the growth rate. ${\mathrm{N}\mathrm{H}}_4^{+}$ comprises about 1–10% of biomass and is a preferred energy source. A high N content led to an improvement in N assimilation. In addition, a too-low N/P ratio may hinder algal growth. Mixing is also an important factor during treatment with algae. Spirogyra sp. has shown fourfold and twofold higher N and P removal efficiencies, respectively, under low air mixing (18 mL/min) compared to high mixing (206 mL/min) [30,33].

${\mathrm{N}\mathrm{H}}_4^{+}$ concentrations exceeding 100 mg/L or 30 mg/L (at pH 9) inhibit bacterial growth. It was shown that carbon addition may lead to an increase in algal productivity [33]. Microalgae are naturally able to attain carbon from several sources (including CO₂). If CO₂ availability is too low, this decreases algae development. A CO₂ concentration at levels of 5–15% may increase nutrient removal and lipid production by microalgae [34].

An O₂ content higher than 20 mg/L may inhibit photosynthesis due to O₂ radical formation, which leads to a decrease in microalgal growth [33].

Increasing temperature, up to a point, promotes algae growth and productivity. The optimal temperature for algae cultivation was shown to be 15–28 °C. A temperature above 35 °C may decrease growth rate, whereas temperatures below 10 °C may decrease nutrient removal by 20 and 46% for TP and TN, respectively. The optimal temperature for nitrification was established between 28 and 32 °C. Temperatures exceeding 45 °C or below 5 °C cease the nitrification process [33].

Increasing light up to a certain point increases the photosynthesis and microalgae growth—after the breakpoint, it decreases. Usually, for freshwater algae, saturation amounts to 200–400 µE/m²). The easiest way to protect algal biomass from light limitations is to decrease the depth of the culture vessel and implement-mixing [33]. The presence of cyanobacteria may inhibit microalgae growth due to the release of substances that inhibit microalgae growth. Viruses and parasites (like protozoa or rotifers) influence wastewater treatment rate. It was also noted that a high amount of organic substances, or acetate, may inhibit nutrient uptake or lead to internal damage to algae cells [33].

pH also plays a critical role—it alters the ionization state of various functional groups present on the algae surface. A pH exceeding the isoelectric point enhances adsorption. A pH higher than 9 influences algae growth due to decreased photosynthetic activity [33].

Hydraulic retention time (HRT) also influences the treatment process with algae. Too short an HRT may lead to algae culture being washed out. An HRT between 4 and 7 days allows for the best results to be obtained [33].

Mixing is a factor that prevents cell sedimentation on the pond bottom. On the other hand, too intense mixing may damage the formation of biomass, which finally decreases settling efficiency [33].

The initial density is a factor that results in self-shading, the accumulation of autoinhibitors, and a reduction in photosynthetic efficiency if it is too high [27]. The high salt content in treated wastewater decreases biodegradation activity due to the inhibition of biological movement [29].

Algae-based processes are not efficient for all wastewater types; they largely rely on a single species’ performance, unlike bacteria-based processes, which use multiple bacterial cultures for N and P removal [30].

It was shown that the production costs of microalgae biomass can reach 2.3–4.5 €/kg. When wastewater is used as a nutrient source, it can be decreased to 1.4–3.6 €/kg [35].

However, specifically, Chlorella manufacturing costs reach 4.87 $/kg, with an energy consumption equal to 0.96 kWh/kg [25,31].

Growing microalgae as a by-product of wastewater treatment may decrease the cost of microalgae cultivation. However, there are some limitations linked to economic and energy assessments. The harvesting method for algae should focus on reducing operating and maintenance costs, improving biomass recovery, and minimizing environmental impact. The initial recovery stage may involve gravity-settling (which is typically conducted), dewatering, and drying. Dewatering and drying are energy-intensive and costly, and they still pose a technological challenge. Centrifugation is a fast and efficient harvesting method, but operation and investment costs are high. Sedimentation and flotation are perceived as more economical and suitable alternatives to harvesting microalgae. Spontaneous flocculation may occur in cultures cultivated under sunny conditions with limited CO₂ or by increasing pH. Fe₃O₄ was also applied in microalgae harvesting. These particles adhere to cells, and are then removed thanks to an external magnetic field. Recovered particles can be reused multiple times [17].

3.1.1. Contaminants’ Removal

Algal systems can treat human sewage, livestock wastes, agro-industrial wastes, and industrial wastes, including wastes from piggeries or food-processing factories. They remove toxic compounds like Pb, Cd, Hg, Sc, As, and Br [30]. They have the ability to eliminate coliform bacteria and decrease COD (chemical oxygen demand) and BOD (biochemical oxygen demand) Chlorella vulgaris was shown to decrease inorganic N and P in wastewater by about 50.2–86 and 78–97.8%, respectively [30]. It was noted that phosphorus was removed more efficiently from domestic wastewater than from industrial wastewater. Phormidium laminosum immobilized in polymer foam removed ${\mathrm{N}\mathrm{O}}_3^{-}$ in a continuous-flow system at a rate of about 90%. A multi-stage algal system is recommended to achieve the full potential of nitrogen removal by algal biomass incorporation [27]. It was observed that the season of the year influenced nutrient removal—nitrogen removal rates were 695.4–2383.4 mg/m²/d and phosphorus removal rates were 70.4–111.8 mg/m²/d in winter and summer [30].

Chlorella miniata, C. vulgaris, C. reinhardtii, and Sphaeroplea were successfully used for HM removal [36]. Some species, e.g., Chlorella vulgaris, Spirogyra sp., and Oscillatoria sp., can remove dyes like Methylene blue, or malachite green [36]. Nanochloris sp. was able to remove ibuprofen, trimethoprim, triclosan, ciprofloxacin, or carbamazepine. Removal efficiencies varied from 20% (for carbamazepine) up to 100% (for ciprofloxacin) [36]. Desmodesmus sp. was reported to remove Fe and Mn at rates of 40–80% when initial concentrations amounted to 10–20 mg/L [37]. Chlorella and Scenedesmus were shown to be efficient in treating olive oil and industrial wastewater [25,31]. The pollutant removal performances of microalgae species are summarized in

Table 1. Microalgae efficiency in removing contaminants from wastewater.

Microalgae	Wastewater Type	Pollutant	Removal Efficiency, %	Reference
Chlorella sp.	Municipal	COD	≤90.8	[28]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	≤93.9
		TN *	≤89.1
		TP **	≤90.6
Chlorella vulgaris	Agricultural	COD	≤96.0	[28]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	≤81.2
		TN *	≤91
		TP **	≤85.3
Scenedesmus sp.	Agricultural	COD	96.2	[38]
		TN *	88.2
		TP **	71.5
Chrococcus sp.	Sewage treatment plant, municipal, slaughterhouse	COD	45–72	[38]
		${\mathrm{N}\mathrm{O}}_3^{-}$	90–98
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$
		${\mathrm{P}\mathrm{O}}_4^{-}-\mathrm{P}$
Chlorella vulgaris	Factory wastewater Municipal wastewaters	N	81.5–92.6	[37]
		P	80.5–94.1
		N	90.2
		P	85.5
C. sorokiniana	Municipal Mixture of municipal and piggery wastewater	N	35–93	[37]
		P	65–100
		N	100
		P	40–60
Chlorella vulgaris, Scenedesmus obliquus, Consortium C.	Urban wastewater	N	84–98	[37]
		P	92–100
Nannochloropsis gaditana Chlorella Scendesmus Tetradesmus	Real wastewater (COD = 498 mg/L) (laboratory scale)	COD	54	[24]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	47
Chlorella sorokiniana	Real wastewater (laboratory scale) COD = 49 mg/L ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$ = 34.1 mg/L ${\mathrm{P}\mathrm{O}}_4^{3-}-\mathrm{P}$ = 4.9 mg/L	COD	37	[24]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	70
Chlorella sp. Scendesmus sp.	Real wastewater (laboratory scale) COD = 249 mg/L TN = 63 mg/L TP = 6.7 mg/L	COD	97	[24]
		TN *	88
		TP **	88
Cyanobacteria Diatoms Green algae	Real wastewater COD = 593 mg/L ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$ = 72 mg/L ${\mathrm{P}\mathrm{O}}_4^{3-}-\mathrm{P}$ = 16 mg/L	COD	91	[24]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	99
		TP	49
Chlorella and Diatoms	Real wastewater (laboratory scale) COD = 517 mg/L ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$ = 86 mg/L ${\mathrm{P}\mathrm{O}}_4^{3-}-\mathrm{P}$ = 43 mg/L	COD	95	[24]
		${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	99
		TP	46
Chlamydomonas sp.	Municipal wastewater (secondary effluent) (laboratory scale)	TN *	77.57	[33]
		TP **	100
Coelastrum microporum	Municipal wastewater (laboratory scale)	TN *	88	[33]
		TP **	89
Mucidosphaerium pulchellum	Domestic wastewater (laboratory scale)	TN	79	[33]
		TP **	49

* TN—total nitrogen; ** TP—total phosphorus.

.

The removal efficiency of different compounds (e.g., TN, TP, BOD, COD) shows high variability, ranging from approximately 30–40% under less favorable conditions to values exceeding 90% under optimized conditions (Table 1). Wastewater type and composition strongly influence treatment performance.

3.1.2. Biomass Production

Oedogonium cultured in situ for 12 months based on municipal wastewater achieved a biomass productivity equal to 9–15 g/m²/d. Chlorella sp. and Scenedesmus sp. have become widely applied in wastewater treatment due to their fast growth rate and high adaptability to different wastewater streams. A significant change in the microalgae productivity rate is observed in different seasons. For example, it reached 8.9 vs. 15.8 g/m² for Oedogonium in winter and summer, respectively [30]. The microalgae showed the highest growth rate under red and blue light (yellow or green light was less efficient) [27]. Monoraphidium dybowskii exhibited a productivity of 18,000 mg/m²d and a CO₂ fixation rate equal to 33,000 mg/m²d [25,31].

Table 2. Examples of biomass productivity by different microalgae species.

Wastewater Type	Microalgae	Biomass Yield/Productivity	Additional Information	Reference
Aquaculture wastewater	Parachlorella kessleri	0.015 g/(L·d)	Nutrient uptake (mg/L/d) ${\mathrm{N}\mathrm{H}}_4^{+}$-$-$3, ${\mathrm{N}\mathrm{O}}_3^{-}-$0.14, TP–0.7	[30]
Anaerobic digestate	Chlorphyceae	0.06 g/(L·d)	Nutrient uptake (mg/L/d) ${\mathrm{N}\mathrm{H}}_4^{+}-$20, TP − 4	[30]
Municipal wastewater	Scendeesmus obliquus	0.08 g/(L·d)	Nutrient uptake (mg/L/d) TN–4.4, TP–0.9	[30]
Agro-industrial wastewater	Chlamydomonas sp.	0.3 g/(L·d)	Nutrient uptake (mg/L/d) ${\mathrm{N}\mathrm{H}}_4^{+}-$19.2, ${\mathrm{N}\mathrm{O}}_3^{-}-$1.5, TP–4.5	[30]
The Cleveland Bay municipal WWTP	Oedognium	3.57 t/ha (after 49 d)	-
Domestic wastewater	Chlorella variabilis	1.72 g/L	-	[36]
Domestic wastewater	Scenedesmus abundans	3.55 g/L	-	[36]
Municipal wastewater	Scenedesmus obliquus	0.22 g/L	-	[36]
Domestic wastewater	Chlorella sp.	0.73–1.38 mg/(L·d)	-	[36]
Municipal wastewater	Scenedesmus sp.	1.81 g/L	-	[36]
Municipal wastewater	Scenedesmus sp.	1.1 g/L	-	[36]
Municipal wastewater	Chlorella sorokiniana	1 g/L	-	[36]
-	Mucidosphaerium pulchellum	0.1889 kg/m3/d	-	[33]
-	Chlamydomonas sp.	0.0552 kg/(m3·d)	-	[33]
-	Chlorella zofingiensis	0.29616 kg/(m3·d)	-	[33]
-	Coelastrum microporum	0.044 kg/(m3·d)	-	[33]
-	Oocystis sp.	0.02525 kg/(m3·d)	-	[33]

provides an overview of biomass yield/productivity across selected microalgae strains.

The growth rate was dependent on microalgae species as well as wastewater type. Depending on microalgae species, it may reach up to 0.29616 kg/m³/d or 3.55 g/L. The lowest productivity was shown by Parachlorella kessleri (Table 2).

3.1.3. Energy Production

Numerous studies [27,35,39] have demonstrated the significant role of wastewater in biofuel production. Microalgae can be applied in the production of biogas, biodiesel, bioethanol, biohydrogen, and bioelectricity [36]. As a biofuel, microalgae have a lot of advantages, like fast growth, relatively simple cultivation, no occupation of arable land, alleviation of the greenhouse effect, and high lipid content [28].

This kind of biofuel is highly biodegradable and does not contain sulfur or other toxic materials [31]. Biodiesel production from microalgae includes two different steps: lipid extraction from microalgal cells, and the transesterification of lipids using alcohol and a catalyst [25]. The biomethane potential of C. vulgaris and mixed culture of native algae species (dominated by Scenedesmus sp.) varied between 154 and 252 L CH₄/kg VS [25], whereas Scenedesmus and Chlorella may produce 0.25–0.5 L CH4/g VS input (after 11 days, at temperature 35–50 °C). When equal masses of wastewater sludge and Spirulina were co-digested, the doubled methane production was observed. Also, after adjusting the C:N ratio to around 20–25:1, the methane production was doubled from 0.6 to 1.2 L CH₄/L VS at 35 °C [27]. It was shown that biogas produced via anaerobic digestion by microalgae biomass is mostly made of methane (55–75%) and CO₂ (25–45%) [18]. Flocculation as a harvesting method was established as the least impactful and most feasible method for biofuel production [17].

Before the production of microalgal-based biofuel, some obstacles need to be overcome. The main challenges are: the optimization of microalgal harvesting/dewatering, the provision of renewable and sustainable nutrient supply, improvements in oil content and composition, an increase in efficiency, and a reduction in the cost of lipid extraction [29,40]. Biomass production demands high financial outlays, and harvesting accounts for about 50% of the total cost of biodiesel production [40].

3.1.4. Advantages/Disadvantages

Microalgae application in bio-treatment has some advantages, like photosynthetic capabilities, converting solar energy into useful biomass, and incorporating nutrients (N, P) responsible for eutrophication. Microalgae, due to photosynthesis, increase pH, which provides a disinfecting effect. Other advantages include renewability and the utilization of solar energy. In comparison to higher plants, microalgae biomass production is significantly greater. Secondly, they can be cultivated in shallow ponds on hardpan soils. Arable land or fresh water is not needed [27]. They are characterized by cost-effectiveness, easy-to-use operating systems, high biomass volume, energy efficiency, and the ability to synthesize valuable products [41]. Higher productivity and a low water footprint are some of the advantages related to microalgal biomass [42].

One of the major drawbacks is the recovery of microalgae from the treated effluent [30]. The costs of harvesting and chemical extraction are high, which may contribute to about 60% of the total operating costs [30]. The small cellular size and high water content lead to the problem of keeping processing costs acceptable [43]. Also, usually, the wastewater C:N ratio is not high enough to meet the requirements of algal biomass [25,31].

While common challenges in microalgae systems often focus on biomass–liquid separation, this issue can now be readily addressed through advanced biofilm and granular sludge technologies [30]. Consequently, the critical disadvantage shifts toward low-temperature inhibition. Vulnerability to temperature fluctuations is one of the main drawbacks related to microalgae application. Excessively high temperatures may exert an inhibitive effect [32]. Surpassing the optimal temperature range for microalgae (15–28 °C) results in a significant decrease in biomass productivity [35]. Furthermore, it has been observed that the relationship between temperature and phosphate removal is more complex than that for nitrate. Interestingly, the highest removal efficiencies were noted at 15 °C rather than at 25 °C [9]. At temperatures oscillating around 5 °C, even polar strains exhibited a significant decrease in removal efficiencies. It was also noted that at lower temperatures, algae become more sensitive to light, which inhibits their growth and general performance [9]. Conversely, at higher temperatures, more nitrogen is present in the form of free ammonia, which can freely diffuse into algal cells and inhibit their growth [44].

The algae biomass may be further used as feedstock, eco-friendly fertilizer, or biofuel [25,31,36]. Cultivated Chlorella vulgaris and Scenedesmus obliquus showed a protein content reaching 45%, which is very suitable for animal feed [28].

Microalgae can also help reduce greenhouse emissions by sequestering CO₂ because they use it as the main carbon source. This means that they can be applied to purify flue gas (especially through burning processes) [35,37,45]. Microalgae are able to fix CO₂ at rates amounting to 56.26–85.72 mg CO₂/(L·d) [32].

The wastewater environment may have high temperatures of around 30–40 °C and resistant species of microalgae are needed in this case [35].

It was assessed that a polishing treatment process aimed especially at ${\mathrm{N}\mathrm{H}}_4^{+}$, ${\mathrm{P}\mathrm{O}}_4^{3-}$, ${\mathrm{N}\mathrm{O}}_3^{-}$ removal may be four times more expensive than the primary treatment [27].

3.2. General Objectives with Duckweed

Duckweed includes 37 species and can be divided into five genera: Spirodela, Landoltia, Lemna, Wolffiella, and Wolffia [46]. To obtain satisfactory treatment results, optimization is needed, e.g., an adequate ratio of nutrients, light intensity, and fronds density [47]. Duckweed mainly requires sunlight to treat contaminated water, with the simultaneous production of protein and nutrients, and can be easily harvested [48]. Duckweed was applied to eliminate pollutants from domestic wastewater, septage, or agricultural wastewater. In comparison to other aquatic plants, duckweed has greater potential in wastewater treatment, and generally, lower costs than other treatment systems [47].

A system that includes a pond covered with a duckweed mat seems to treat wastewater jointly with bacteria. On the other hand, it should be noted that the bacterial decomposition may lead to anaerobiosis in the water. Duckweed converts the main minerals C, N, and P into protein. Duckweed also removes other organic materials thanks to its ability to directly use of simple organic substances and their assimilation as carbohydrates and amino acids [27]. Duckweed was shown to increase COD removal in shallow batch systems. Fat duckweed and great duckweed showed high potential in treating domestic wastewater [49].

A harvesting schedule is important to maintain growth and nutrient removal. Removing half of the biomass every third day is the most practical option [49].

There are some factors influencing the performance of duckweeds: pH, light intensity, temperature, salinity, and the dissolved oxygen content. The growth rate is inversely proportional to the increase in the dissolved salt content of the water [47]. Duckweed floats on the water surface and can be collected by simple mechanical skimming, making the harvesting process substantially less technically demanding and costly [50]. Harvesting costs were estimated as 3% of the capital cost (16,200 $/year) [51]. This stands in contrast to microalgae, which require energy-intensive mechanical separation methods such as centrifugation or membrane filtration due to their microscopic cell size, which largely contributes to the harvesting costs (0.5–2.0 €/kg biomass) [52].

3.2.1. Contaminants Removal

It was shown after only 3 days, the N and P content was lowered by duckweed, and may accumulate up to 9.1 t N/ha/year and 0.8 t P/ha/year [48]. Duckweed can significantly decrease COD, N, and P content. Nutrient elimination from wastewater showed a linear correlation to the initial nutrient concentrations (to a certain concentration). For TP, it was equal to <2.2 mg/L, and ≥11.6 mg/L for N [53]. Other reports showed duckweed’s capacity to remove N at a rate of 446 mg/(m²·d) [54]. The primary mechanism of nitrogen removal by duckweed is its assimilation by fronds and roots [54]. The significant efficiency of duckweed in wastewater treatment lies in its symbiosis with microorganisms. The submerged parts of the plant (roots and the underside of its fronds) provide a vast physical surface area that serves as a substrate for biofilm development. It supports dense colonies of beneficial bacteria and algae. The plant sustains these microorganisms by transporting oxygen to the root zone, which subsequently creates optimal conditions for aerobic bacteria to decompose organic matter. The cooperation between the plant and its resident microbes is responsible for removing up to 75% of the total nitrogen and phosphorus load from the wastewater [49]. Duckweed removes pharmaceuticals and antibiotics through multiple concurrent mechanisms: direct uptake into plant tissue, adsorption onto frond surfaces, and plant-associated microbial degradation by bacteria colonizing the root zone and frond surfaces. Among these, plant uptake was identified as the dominant removal pathway by Allam et al., 2015 [55], while sorption and microbial degradation act as secondary contributors. For antibiotics resistant to hydrolysis and photolysis, such as sulfamethoxazole, duckweed uptake under illumination conditions becomes particularly critical for effective removal [56].

A toxicity evaluation of 13 different pharmaceuticals showed that all compounds were toxic for duckweed at concentrations below 100 mg/L [48]. Most antibiotics are toxic to duckweed, but duckweed can tolerate and remove those compounds from the environment (paracetamol, caffeine, and triclosan were removed in amounts equal to 97.7, 98.0, and 100%). L. minor and L. gibba were able to recover after exposure to herbicide diuron (concentration between 0.4 and 208 µg/L) [48]. Flumequine was degraded in 96% of all tested concentrations. Duckweed was also shown to have the ability to remove dyes (methylene blue was eliminated in 98%). Some species (L. gibba, L. minor, S. polyrhiza, W. globosa, W. australiana, L. valdiviana) also can remove HMs, e.g., As. W. globosa accumulated 2–10 times more As in comparison to S. polyrhiza and Azolla. For S. polyrhiza, it was observed that production and growth rate decreased in the presence of boron [48]. A similar effect was observed in L. minor during exposure to cobalt stress [57]. It was shown that duckweed can remove Cd, Cr, Ni, and Pb at rates of about 44.93, 32.26, 74.49, and 79.1%, respectively. The removal efficiency showed a significant positive correlation with pH. Duckweed is a hyperaccumulator for Pb, Cr, and Ni, but a poor accumulator for Cd [58]. According to other reports [59], duckweed decreases Cr content by about 98%, and Pb, Fe, Cd, and Cu by about 30%.

The pollutant removal performance of duckweed is summarized in

Table 3. Removal efficiency of chosen parameters by duckweed.

Pollutant	Removal Efficiency, %	Scale Maturity	Reference
	59	Laboratory scale	[60]
BOD	50–95	Laboratory-scale	[49]
	80–90	Full-scale	[61]
	54	Laboratory scale	[60]
COD	50–95	Laboratory scale	[49]
	60–80	Full-scale	[61]
TSS	63	Laboratory scale	[60]
	80–90	Full-scale	[57]
TDS	9	Laboratory scale	[60]
EC	5	Laboratory scale	[60]
NH3 − N	63	Laboratory scale	[60]
	80–90	Full-scale	[61]
NO2 − N	64	Laboratory scale	[60]
NO3 − N	63	Laboratory scale	[60]
TN ****	63	Laboratory scale	[60]
	60–90	Full-scale	[61]
TP *****	65	Laboratory scale	[60]
	9–61	Laboratory scale	[49]
	70–95	Full-scale	[61]
Fecal coliform bacteria	96	Laboratory scale	[60]
Total bacteria count	98	Laboratory scale	[60]
HMs	38–80	Full-scale	[61]

TN ****—total nitrogen; TP *****—total phosphorus. The described experiments did not describe harvesting method and economic aspects.

.

It was shown that the removal efficiency of different compounds by duckweed significantly varied. This could reach about a few percent (5%) for EC and TDS or even over 90% for TP, BOD, or total bacteria count (Table 3).

There were experiments regarding duckweed growth in different solution salinities. The initial COD, phosphate, and total nitrogen Kjeldahl (TKN) were equal to 175.4, 13.60, and 15.68 mg/L, respectively [47]. The influence of wastewater conductivity on duckweed growth and the corresponding pollutant removal efficiencies is presented in

Table 4. Duckweed growth rates, TKN, COD, and phosphate removal efficiency [47].

Conductivity, µS/cm	Duckweed Growth, g/d	COD Removal, mg/m2d	${\mathbf{P}\mathbf{O}}_4^{3-}$ Removal, mg/m2d	TKN * Removal, mg/m2d
200	2.92	84.67	12.66	9.73
600	4.74	577.77	30.26	64.99
1000	4.82	642.74	31.58	77.99
1400	3.14	413.19	25.37	69.63
1800	3.44	275.46	22.38	77.99
2200	2.56	91.82	26.86	75.2
2600	2.36	137.73	23.38	77.99
300	1.50	137.73	17.90	77.99

* TKN—total nitrogen Kjeldahl.

.

Duckweed growth was significantly influenced by wastewater conductivity. The highest values were obtained for 1000 µS/cm; after that point, a decrease in growth was observed. A similar tendency was observed for COD and ${\mathrm{P}\mathrm{O}}_4^{3-}$ removal rates. NTK elimination from wastewater was less determined by conductivity (Table 4).

3.2.2. Biomass Production

In some reports [60], the observed duckweed growth rate amounted to 1.2 g/m²d. It was shown that it grows rapidly on the surface of municipal, dairy, swine, industrial, and aquaculture wastewater [53]. It grew to 8.0, 6.9–8.2, and 7.0 g/m²/d in anaerobic, aerobic, and anoxic conditions, respectively. The COD, TN, and ${\mathrm{P}\mathrm{O}}_4^{3-}$ concentration in raw wastewater reached 38–114, 22.9–32.9, and 2.9–3.3 mg/L, respectively [53]. At pH 7.8, a significant production of 55 kg dry wt./ha·d is gained [54]. According to other data [54] Lemna minor exhibits a growth rate of 29 g/(m²·d) [54]. Chen et al. [62] performed research into duckweed biomass production at a time frame varying between 4 and 16 days. In all cases, a significant decrease in biomass production over time was observed. For example, L. punctata biomass production was lowered from 68.50 to 24.17 g of dry mass/m² after 16 days. In the case of S. polyrhiza, this value decreased from 46.66 to 12.83 g of dry mass/m² after 16 days [62]. The average duckweed biomass production, tracked over the 16-day cultivation period, is summarized in

Table 5. Average duckweed biomass production [62].

	S. polyrhiza	L. minor	L. punctata	Polyculture
Average biomass production *, g of dry mass/m2	25.95	33.42	38.02	35.94

* Raw wastewater quality—TN = 6 mg/L; TP = 0.56 mg/L.

.

3.2.3. Advantages/Disadvantages

Duckweed is highly tolerant to elevated levels of ${\mathrm{N}\mathrm{H}}_4^{+}$ in water solutions. L. minor can grow well at ${\mathrm{N}\mathrm{H}}_4^{+}$ concentrations up to 84 mg/L [48]. This makes duckweed suitable for treating wastewater from domestic, agricultural, and aquaculture sources (often containing high amounts of ${\mathrm{N}\mathrm{H}}_4^{+}$) [48]. Duckweed is also characterized by high resistance to water salinity and has the ability to remove N and P at NaCl concentrations of up to 4.5 g/L NaCl [63]. However, a salinity over 2000 µS/cm is limiting factor, inhibiting duckweed’s growth rate [48]. Duckweed is relatively inexpensive to operate and maintain, and it is functionally simple. It allows for results similar to or even better than conventional wastewater treatment systems to be obtained. At present, this solution is recommended even for large-scale applications [58]. Duckweed can be harvested easily and is relatively tolerant of cold. However, a dense cover of duckweed showed a tendency to inhibit oxygen entering the water via diffusion and hinders phytoplankton production due to poor light penetration [49]. Duckweed has a relatively large size and floats on the water surface; thus, its harvesting is simpler in comparison to other microorganisms (e.g., algae) [64].

3.3. Fungi General Objectives

The use of fungi in wastewater treatment has attracted much more attention in comparison to bacteria. Fungi are simple to grow and produce a high amount of biomass. They have some specific, useful features—porosity, quick adaptation to different conditions, net-like texture, rapid growth, and efficient dewatering properties—which make them suitable for sludge dewatering, biosorption, or biomass aggregation [41]. They can survive in different environmental conditions and on various substrates. Two important factors make fungi the ideal candidate for wastewater treatment (mycoremediation): they secrete many extracellular enzymes and have a fungal hyphal mesh, which protects their sensitive internal organelles [29]. A few fungi strain was listed as capable of digesting various environmental contaminants, including dyes, pharmaceutical drugs, aromatic hydrocarbons, and HMs. Fungi can enhance pollutant removal (e.g., HMs) by limiting their bioavailability and converting them to less useful toxic forms [29,65]. Pleurotus pulmoniarus, Stachybotrys sp., Cephalosporium aphidicola, Aspergillus parasitica, Verticillum Terrestre, Candida sp., Acremonium sp., Glomus sp., Minimedua sp., Talaromyces, Hydnobolites, Peziza. Saccharomyces cerevisiae, Galactomyces geotrichum, Trichosporon beigelii and Candida krusei have the ability to remove dyes from wastewater [29,41,65]. Yeast has become a promising option in wastewater treatment, and it can be applied, e.g., for COD, as well as for mono- and polyphenol removal [29]. It also eliminates HMs from the environment. Yeast is easy to cultivate, can grow in a wide pH range, and has faster growth rates than molds [29]. Phanerochaete, Pleurotus, Penicillium, Fusarium, Geoderma, Lepiota, Trametes, Daldenia, and Aspergillus were shown to be suitable for industrial wastewater treatment [41]. Mycelial pellets or granules (spherical and scattered fungal mycelial cells) show biosorption abilities. Some filamentous fungi have a natural tendency to develop pellet morphology (self-immobilization in pellets) [66]. The occurrence of contamination or genetic instability within the fungal population influences wastewater hydrolysis, fungal production, and finally the overall product yield [65]. Some other examples of fungi used in the treatment of specific wastewater are listed as follows: Myrotherium verrucaria and Trametes hirsuta (degrade cellulose-rich waste), A. niger (apple distillery wastes), P. chrysosporium (lignin removal), Pleurotus ostreatus (lignocellulosic biomass degradation), Alternaria tenuis, A. niger, Trichoderma viride (plastic degradation), Humicola grisea (raffinose removal), and P. chrysosporium (veratryl alcohol elimination) [65]. Aspergillus niger may grow on various substrates, in wide temperature ranges (6–47 °C), and at pHs between 1.4 and 9.8 [67].

A few main factors can be listed as responsible for fungal growth and metabolism [66,68,69]:

• pH—depending on the fungi type, different pH values are desired for the most efficient performance. Generally, fungi need to have the ability to grow in a pH over 7,
• Temperature—this plays an important role in fungal growth, metabolism, and electricity generation using fungal cells. It facilitates the cell’s metabolism and enzymatic reactions. The majority of fungi are mesophiles and thrive at temperatures between 20 and 40 °C.
• Ionic strength—a higher ionic conductivity influences the fungal cell.
• Salinity—about 90% of fungi can grow at a salinity between 3 and 6; fungi species can grow better in a salty environment.
• The occurrence of contamination or genetic instability within the fungal population influences wastewater hydrolysis, fungal production, and finally the overall product yield.

3.3.1. Contaminants Removal

Trichotbecium roseum was found to be ideal for wastewater treatment (due to its protein content and ability to remove N and P). Yeast strains were able to remove COD, TN, ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$, and phosphate at rates of about 0–72, 22–93, 27–90, and 12–100%, respectively [65]. Some seven fungal species isolated from a wastewater stabilization pond were able to decrease BOD, ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$, and ${\mathrm{P}\mathrm{O}}_4^{3-}$ content by 53–72, 49–77, and 34–77%, respectively. O₂ availability was shown to be a limiting factor in biodegradation [65]. Pollutant removal efficiencies demonstrated by various fungal strains are compiled in

Table 6. Contaminant removal efficiencies achieved by fungal biomass.

Pollutant	Removal Efficiency, %	Reference
COD	0–72	[65]
BOD	53–72	[65]
TN	22–93	[65]
${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	27–90	[65]
	49–77	[65]
${\mathrm{P}\mathrm{O}}_4^{3-}$	12–100	[65]
	34–77	[65]

.

The specific types of pollutants remediated by various fungal strains across different wastewater matrices are summarized in

Table 7. Types of pollutants removed by particular fungi [29].

Fungi	Pollutant	Wastewater Type
Aspergillus niger, Rhizopus oryzae, Saccharomyces cerevisiae, Penicillum chrysogenum	Cr	Industrial
Candida sphaerica	Fe, Zn, Pb	Industrial
Candia sp.	Cu, Ni	Industrial
Candida porapsilosis	Hg	Industrial
Sphaerotilusnatans Gloephyllumsepiarium	Cr	Industrial
Aspergillus niger	Fe	Industrial
Trametes versicolor	Salicylic acid, codeine, ceflalexine, acridone, ciprofloxacine, propanolol	Urban wastewater
Penicillumcorylophilum	Suspended solids	Domestic wastewater

.

According to Table 7 various fungi species can eliminate HMs, as well as suspended solids or drugs, from wastewater.

R. oryzae was shown to be able to remove Cu and Cd from water solutions. T. harzianum, A. Niger, A. flavus, and P. expansum showed efficiency in kitchen waste biodegradation [41].

Fusarium spp., Trichoderma spp., Aspergillus spp., Rhizopus spp. allowed to for COD to be removed at rates of 92% [70].

Experiments showed that A. oryzae, after 24 h cultivation, decreased the COD value by about 50–60% [70].

3.3.2. Biomass Production

A. oryzae gained a dry biomass production equal to 49.16 g/L at pH 3.85, a TKN equal to 1.40 g/L, and COD reaching 58.7 g/L. Rhizopus oligosporus showed a production of dry biomass equal to 5 g/L, while R. arrhizhus showed a value of less than 2 g/L—they were cultivated on wastewater from a potato chips plant with a COD reaching 20–30 g/L [70]. Cephalosporium eichborniae at pH 3.75 produced 0.61 g of dry weight (per gram of carbohydrate) [65]. Ganoderma lucidum growing on wastes reached a biomass yield productivity of 7.8–20.1 g/L [71]. A. niger harvested on glucose or xylose may produce 4.6–5.8 g/L of biomass. Within this application, bagasse, as a carbon source production value, reached 1.9 g/L. Also, promising results were obtained by using glycerol as a substrate. This allowed for a productivity equal to 8.2 g/L and even 35.29 g/L to be obtained for A. niger (COD and TN content in raw wastewater reached 107 and 2.32 g/L, respectively) [67].

3.3.3. Energy Production

Hydrolytic and ligninolytic fungi are suitable candidates for the production of biofuels or bioethanol. There are some species that have the potential to produce biodiesel or electricity: Candida sp., Colletotrichum sp., Saccharomyces cerevisiae, Penicillium sp., and Rhizopus sp., Aspergillus sp. [68]. Unlike microalgae, the implementation of oleaginous fungi in biodiesel manufacturing is, so far, very limited. A simple and fast growth rate in bioreactors, unaffected by light intensity, and the ability to utilize a wide range of lignocellulosic waste biomass as a carbon source, make fungi attractive substrates for biodiesel production. Fungi pelletization makes harvesting much easier and cheaper in comparison to microalgae. On the other hand, their tough cell walls hinder wastewater treatment efficiency [40]. Fungus species have been recognized as “novel cell factories” due to their capacity for high-quality biofuel and bioelectricity generation [68,69].

3.3.4. Advantages/Disadvantages

The organic wastes are transformed by fungi into industrially important biochemicals and other valuable compounds, making them advantageous compared to the bacterial cultures applied in wastewater treatment plants [29]. In comparison to other microorganisms, fungi have more advantages. Filamentous fungi can survive better under harsh conditions, they produce more biomass, they are easier to separate in comparison to bacteria [41], and they metabolize complex carbohydrates better than bacteria [65]. Cellular enzymes produced by fungi enhance the biodegradation of recalcitrant compounds like dyes, phenolic compounds, or polyaromatic hydrocarbons. In addition, fungi are the source of important byproducts like amylase, chitin, or lactic acids. In comparison to bacteria, they have more genes, which results in greater resistance to inhibitory substances, and they can adapt better to the environment [65]. Fungal hyphal growth provides greater protection for their sensitive organelles. Fungi application in wastewater treatment offers: higher degradation rates, efficient fungi biomass separation from the mixed liquor, and the possibility to separate valuable fungi byproducts [65]. Fungal biomass is easy to harvest and can be applied as animal feed [70]. Immobilized cyanobacteria on hollow fibers can produce high amounts of hydrogen, whereas Mastigocladus and Phormidium can directly generate electricity [27].

However, when evaluating process scalability, the HRT emerges as a critical operational factor, given that fungal growth kinetics can be significantly slower than bacterial processes under certain environmental conditions. In conventional continuous systems, this difference in growth rates creates a high risk of biomass washout. To overcome this limitation, strategies such as fungal biomass immobilization are employed; by improving biomass retention and increasing cell density, immobilization fundamentally enhances system stability and allows for a substantial reduction in the operational HRT required for efficient treatment [72].

One of the main issues related to fungi application is the necessity of maintaining a pH lower than 5, which is required for enzymatic activity and significantly influences metal ion availability [65]. Another major drawback is that certain bacterial cultures actively compete with fungi for nutrients, which decreases fungal viability. Consequently, this creates a vital need to develop methods capable of preventing unwanted bacterial growth without adversely affecting the fungi [65].

3.4. Bacteria General Objectives

The uptake and N reduction by cyanobacteria is a photosynthetically driven process. Light, temperature, pH, and carbon source availability influence ${\mathrm{N}\mathrm{O}}_3^{-}$ assimilation. The bacteria culture Phormidium bohneri was applied for N removal from effluents coming from swine manure [27]. The thermophilic cyanobacterium Phormidium laminosum was used for N elimination from polluted waters. Phormidium sp. attached to chitosan particles was used for N (including ${\mathrm{N}\mathrm{H}}_4^{+}$, ${\mathrm{N}\mathrm{O}}_3^{-}{,\mathrm{N}\mathrm{O}}_2^{-}$) and orthophosphate removal from secondary urban effluents [27]. In dialysis culture, bacteria are separated by a semi-permeable dialysis barrier [27]. Bacteria have the ability to biosorb metal ions into the cell wall. This can be active (led by living bacterial cells) or passive (performed by living and dead cells) [29].

Abundance, growth under controlled conditions, size, and resistance to environmental changes led to bacteria becoming important biosorbents [29]. Under optimal conditions, cyanobacteria may reach higher growth rates than plants [27]. First, COD plays a decisive role; an organic matter concentration exceeding 300 mg COD/L, combined with a simultaneous COD:N ratio higher than 2, has been shown to inactivate Anammox bacteria. Furthermore, Anammox activity is strictly inhibited by methanol concentrations as low as 0.5 mM [33]. Temperature also serves as a fundamental regulator of gas transfer, microbial activity, and the settling properties of biological solids, with P removal efficiency significantly increasing within the 20–37 °C range. Additionally, the DO levels remain highly dependent on the specific bacterial type utilized. pH values must also be carefully controlled—nitrification demands a pH of approximately 7.5–9, whereas optimal performance occurs between 7 and 8 [33]. Finally, the co-occurrence of heterotrophic protists can further enhance the treatment process by contributing to nutrient cycling and carbon mineralization. The operational stability and efficiency of biological wastewater treatment are strictly governed by the DO content, which serves as a fundamental switch for bacterial metabolism; maintaining DO levels above 2 mg O₂/L is critical for sustaining efficient nitrification [33]. Depending on the precise regulation of anaerobic, anoxic, or aerobic conditions, distinct pathways, such as nitrification, denitrification, organic matter degradation, or enhanced biological phosphorus removal, can be selectively driven. Furthermore, temperature acts as a primary kinetic driver. Within the bacterial system, an increase of approximately 10 °C typically doubles the pollutant removal rate. This temperature dependence is closely linked with pH stabilization, which must be strictly maintained between 7.2 and 7.5. Surpassing this optimal pH range shifts the chemical equilibrium, converting, ${\mathrm{N}\mathrm{H}}_4^{+}$ into free ammonia (NH₃). This uncharged form can easily penetrate bacterial cell membranes, exerting a severe inhibitory effect on the entire treatment process [33].

Contaminants Removal

${\mathrm{P}\mathrm{O}}_4^{3-}$ uptake by cyanobacteria was shown to be a hyperbolic function of the external ${\mathrm{P}\mathrm{O}}_4^{3-}$ concentration [27]. The specific types of contaminants degraded by bacterial strains, along with their corresponding biochemical removal mechanisms, are summarized in

Table 8. Types of contaminants removed by bacterial species and their corresponding removal mechanisms [29].

Bacteria	Contaminant	Removal Mechanism
Bacillus sp.	RED HE78 (dye)	Enzymatic degradation by azoreductase and laccase
Aeromonas sp.	Methyl orange (dye)	Enzymatic degradation by laccase, azoreductase and NADH-DCIP
Escherichia coli	Methyl orange (dye)	Enzymatic degradation
Micrococcus sp.	Reactive Red–120 (dye)	Enzymatic degradation by azoreductase and laccase
Bacillus sp.	Ponceau 4 R (dye)	Enzymatic degradation by azoreductase
Anoxybacillus sp.	Direct Black G (dye)	Enzymatic Azoreductase, pyruvate kinase, quinone reductase
Acinetobacter sp., Bacillus sp. Pseudomonas aeruginosa, Cellulosimicrobium sp.	Cd	-
Bacillus sfirmus, Staphylococcus sp.	Pb	-
Enterobacter cloacae, Klebsiella pneumoniae, Bacillus licheniformis	Hg	-
Lysinibacillus sphaericus, Bacillus safensis	Co, Cu, Cr, Pb, Cd	-
Geobacter sp., Pseudomonas fluorescens, Vibrio harveyi, Pseudomonas aeurigonsa	Mn, Fe, Cu, U, Zn
Pseudomonas aeruginosa, Aeromonas sp.	Cu, Ni, Cr, U

.

Depending on the bacteria type, the pollutant removal mechanism differs. They are able to remove various pollutants, e.g., HMs or dyes (Table 8).

According to the data [73], Spirulina platensis was able to decrease the COD value. Phormidium valderianum lowered BOD and COD content. Phormidium valderianum showed the ability to remove ${\mathrm{N}\mathrm{O}}_3^{-}$ and Cyanobacteria after 7 days of incubation and decreased the total amount of bacteria present in wastewater [73].

3.5. Hybrid Systems’ (Co-Cultures) General Objectives

Co-culturing microalgae with fungi and their symbiotic role in wastewater has been investigated for some time [36]. Application of the microalgae–bacteria consortium in wastewater treatment has also gained attention. Microalgae are effective in treating effluents, eliminating inorganic nutrients from wastewater, and reducing greenhouse gas emissions. The nitrification process led by bacteria in the symbiotic system enhanced the assimilation of oxidized N forms (${\mathrm{N}\mathrm{O}}_3^{-}{,\mathrm{N}\mathrm{O}}_2^{-}$) by microalgae Chlorella vulgaris. In this system, the proliferation of nitrogen-fixing bacteria was observed [24]. In the duckweed system, 40 genera of bacteria showed high abundance (including Rhodopseudomonas, Nitrospira, and Nitrosomonas), whereas in the duckweed–bacteria system, N uptake by duckweed was the main means of N removal [62].

The configuration and environmental performance of diverse microalgal–bacterial co-culture systems are documented in

Table 9. Examples and effects of co-culture [17].

Microalgae/Bacteria	Effect
Microalgae, aerobic granular sludge enriched by Nitrospirae and Bacilliarophyceae	Improved N and C removal.
Lobomonas rostrata and Mesorhizobium loti	The secretion of vitamin B12 allows for the growth of dependent microalgae.
C. vulgaris, B. licheniformis, G. lucidum	Fast growth performance. COD, TN, and TP removal >80%.
C. vulgaris, A. beijerinckii	Microalgae biomass production increased by 71.8%. COD, ${\mathrm{P}\mathrm{O}}_4^{3-},$ and ${\mathrm{N}\mathrm{H}}_4^{+}$ removal rates increased by 20.8, 18.5 and 8.9%, respectively.

.

Various factors influence plant–fungi interaction, including plant species, water type, climate, and other microorganisms. Cooperation between those organisms has numerous important functions, including metal-chelating, siderophores’ emissions, denitrification, and detoxification [29]. The natural symbiosis between fungi, microalgae, and cyanobacteria has existed for millions of years. In this type of symbiosis, fungi consume sugars and nutrients produced by algae during photosynthesis. For instance, fungi protect the algae by retaining water; they also serve as a larger capture area for mineral nutrients, and they may also provide minerals obtained from the substrate of their living processes [24,74]. The attraction between fungi and microalgae plays an important role in the pelletization process [71]. The interaction between algae and bacteria has the character of nutrient-exchange. Bacteria mineralize various organic substrates and other valuable compounds (minerals, vitamins). On the other hand, bacteria can act as algae parasites—they may use algae as a nutrient source [42].

Within the utilization of organic substrates, bacteria produce CO₂ as a by-product, which is mineralized into inorganic compounds and is directly taken by microalgae to support their growth. Microalgae utilize generated CO₂ to produce carbohydrates and O₂ through photosynthesis. Simultaneously, the generated O₂ is needed by heterotrophic bacteria for living functions. The pH and dissolved O₂ highly determine bacterial activity [33]. It was shown that microalgae, filamentous cyanobacteria, and fungi at radiation levels of 3800 Wh/m²d form stable aggregates [24].

The bacteria present in the microalgae–bacteria consortium mainly originate from activated sludge, digestion effluent, and other specific sewage sources. The interaction between microalgae and bacteria is complex, and competition for nutrients may inhibit mutual relationships. On the other hand, bacteria may promote microalgae growth by providing CO₂, nutrients, vitamins, or volatile organic compounds [32].

3.5.1. Contaminants Removal

Co-culturing Chlorella sp. with Penicillium sp. can achieve removal efficiencies that amount to 46.13, 13, and 88.4% for COD, TN, and ${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$, respectively [36]. The obtained results showed that microbial consortia can efficiently eliminate organic matter, ${\mathrm{N}\mathrm{H}}_4^{+}$, and HMs [29].

The contaminant removal efficiencies achieved by the selected co-cultures are summarized in

Table 10. Contaminant removal efficiencies by co-cultures.

Organisms	Pollutant	Removal Efficiency, %	Reference
Fungal and bacterial biomass	COD	34 (pH = 3.5)	[65]
		80 (pH = 4.0)
		68 (pH = 4.5)
Aspergillus sp. Chlorella sp.	COD	70.68	[71]
	TN	67.09
	TP	88.39
Aspergillus fumigatus Chlorella protothecoides	${\mathrm{N}\mathrm{H}}_4^{+}-\mathrm{N}$	73.71	[71]
	${\mathrm{P}\mathrm{O}}_4^{3-}$	55.56
Ganoderma lucidum Chlorella vulgaris	COD	79.74	[71]
	TN	74.28
	TP	85.37
Chlorella pyrenoidosa and landfill leachate	TN	90	[42]
Chlorella vulgaris and Ganoderma lucidum (fungus)	COD	79	[42]
	TN	76
	TP	85

.

Various co-cultures may obtain significantly different removal efficiencies depending on pollutant type. The lowest one (55.56%) was obtained for ${\mathrm{P}\mathrm{O}}_4^{3-}$, whereas for TP and TN, efficiencies may exceed 70% (Table 8).

According to some experiments [75], microalgae decreased COD content by about 60.5%, and co-culture with microalgae and bacteria increased COD removal efficiency to 73.1%. A similar tendency was observed for ${\mathrm{P}\mathrm{O}}_4^{3-}$ and ${\mathrm{N}\mathrm{H}}_4^{+}$.

3.5.2. Biomass Production

The biomass yield/productivity achieved by various microalgae-based co-culture systems is summarized in

Table 11. Biomass production in co-cultures.

Organisms	Wastewater Type	Biomass Yield or Productivity	Reference
Aspergillus sp. Chlorella sp.	Molasses wastewater	4.215 g/L	[71]
Aspergillus fumigatus Chlorella protothecoides	Swine manure wastewater (25% dilution)	~2 g/L	[71]
Ganoderma lucidum Chlorella vulgaris	Biogas slurry (CO2 initial concentration 55%)	0.174 g/(L·d)	[71]
Chlorella pyrenoidosa and landfill leachate bacteria	Municipal wastewater and landfill leachate	1.58 g/L	[42]
C. vulgaris and Ganoderma lucidum (fungus)	Anaerobically digested swine wastewater	4.77 g/L	[42]

.

The biomass productivity significantly varied. It amounted to 1.58 g/L for Chlorella pyrenoidosa cultured with landfill leachate bacteria, and up to 4.77 g/L for C. vulgaris combined with Ganoderma lucidum (Table 11).

Some reports [74] showed the synergetic effects of the co-cultivation of A. niger, A. fumigatus, A. oryzae, and C. echinulata combined with microalgae on biomass production and lipid yields. These were caused by the fact that microalgal cell wall carbohydrates are utilized as a carbon source by fungi. In the C. echinulate and C. vulgari co-culture, cellulase activity induction was observed. This activity was not detected in the C. echinulata mono-culture [74].

3.5.3. Advantages/Disadvantages

In comparison to traditional wastewater treatment processes microalgae–bacteria symbiosis is characterized by certain benefits. This system promotes microalgae growth as well as improving wastewater treatment efficiency [75]. They have unique characteristics like reduced power consumption or biomass refundability. On the other hand, in some cases, bacteria–microalgae behavior is hard to predict. For example, when the algae–bacteria ratio is 1:3 and 5:1, the N removal efficiencies are slightly different (89 and 91%, respectively) [76]. Also, some organisms may prefer a neutral pH whereas others prefer a low one. Due to this, it may be hard to establish optimal conditions for all microorganisms present in co-culture [42].

4. Discussion

Ongoing changes in the modern world (like rapid industrialization, urbanization, and lifestyle modifications) have led to a significant deterioration and change in wastewater quality. Recently, various novel MPs, like pharmaceuticals, HMs, or microplastic, have become more common in the environment [2]. Treated effluent may still contain relatively high amounts of nutrients like P or N (which are prone to lead to eutrophication). To adjust to ongoing changes, the new regulations regarding wastewater parameters have become stricter, and at present, effluent quality is supposed to be enhanced [22]. This led to the necessity of looking for new purification methods or improving the currently used ones. The application of biological agents (like microalgae, fungi, bacteria, or duckweed) in wastewater purification is a promising method characterized by some advantages in comparison to classical techniques (e.g., compact design, ease of operation, reduced energy consumption, or limited greenhouse gas emissions) [38,42,60].

Microalgae and duckweed application, according to this review, seems to be the most popular method. Microalgae species like Parachlorella kessleri, Chlorphyceae, Scendeesmus obliquu, Scenedesmus abundans, Chlorella variabilis, Chlamydomonas sp., Chlorella vulgaris, and Scenedesmus obliquus were shown to have potential in wastewater treatment [17]. It was noted that co-cultures (e.g., microalgae–bacteria like Lobomonas rostrata and Mesorhuzobium loti) may obtain a better performance than a single culture. These species support each other’s growth through creating a more efficient system than single-species cultures [17].

As most published studies report performance exclusively in terms of relative removal efficiencies (%), evaluations of effluent quality against discharge standards are often limited. However, in selected cases where absolute data are available, the technologies demonstrate exceptional compliance. For instance, certain duckweed systems (in monoculture or co-culture) were able to reduce TN from 6.0 mg/L to below 0.5 mg/L, and TP from 0.56 mg/L to below 0.1 mg/L [62], easily satisfying the strict international regulatory discharge limits. For the remaining studies, high percentage efficiencies (>85–90%) indicate high potential for compliance with the guidelines.

While the overall removal efficiencies of the co-culture are comparable to those obtained with monocultures, the primary advantage lies in their enhanced operational robustness and process sustainability. Beyond numerical removal rates, co-culture systems offer significant techno-economic benefits, such as reduced energy consumption due to internal gas exchange (O₂/CO₂ syntrophy). Furthermore, the biological interaction promotes superior sludge sedimentation and granular stability, which directly addresses the challenges of biomass recovery and footprint reduction (compact design) that typically limit traditional and monoculture-based systems [24].

Organisms present in biomass eliminate nutrients (N, P), as well as MPs like pharmaceuticals, dyes, or HMs. They also decrease COD and BOD contents [29,60]. There are some factors that are crucial for effective wastewater treatment, e.g., pH, temperature, salinity, nutrient concentration, light intensity, mixing, the presence of viruses or parasites, the presence of compounds that are toxic to the applied organisms, or the type of treated wastewater. Biomass production is significantly linked to microorganism genera that are present in wastewater, as well as salinity and wastewater quality [24,33,37,77]. The biomass production may vary between a decimal part of a gram per liter up to a few grams per liter. Biomass can be used for the production of various biofuels, like biodiesel, gaseous fuels, or solid fuels. The biogas produced via anaerobic digestion by microalgae biomass is mostly made of methane (55–75%) and CO₂ (25–45%) [18].

There are still some challenges related to biomass application in wastewater treatment. Microalgae recovery from effluent significantly contributes to treatment costs and may account for over half of the operating costs. This is mainly related to the microalgae’s small size and high water content. In addition, the C:N ratio present in wastewater may not be high enough to meet the requirements of biomass cultivation [30].

Biomass usage is becoming more popular—it complies with the sustainable development assumptions and the circular economy. However, there are still some limitations that need to be overcome. The microorganisms must be selected based on the local economic context and wastewater characteristics. Whereas microalgae show very high nutrient removal efficiency, their practical applicability in large-scale systems is still limited by harvesting costs [17].

Anammox (anaerobic ammonium oxidation) is a biological process in which ammonium is oxidized to nitrogen gas under anoxic conditions, using nitrite as the electron acceptor [78]. Compared to conventional biological nitrogen removal, the Anammox process offers several advantages, including a significant reduction in oxygen demand (up to ~60%), elimination of the need for organic carbon for denitrification, and lower excess sludge production (up to ~80%). However, its application to domestic wastewater is limited. Municipal wastewater typically has a relatively high C:N ratio and fluctuating composition, which is not optimal for Anammox activity. In addition, the process is highly sensitive to low temperatures (<15 °C) and seasonal variations (10–25 °C), which may significantly reduce bacterial activity [79]. For these reasons, Anammox is currently considered more suitable for sidestream and high-strength nitrogenous wastewaters, such as landfill leachate and certain industrial effluents, rather than mainstream domestic wastewater treatment [80].

From a technical point of view, fungal-based systems or co-cultures (e.g., microalgae–fungi) appear to be a more viable solution for industrial applications because they form larger aggregates (pellets), which significantly simplifies biomass separation and decreases the general cost of the micropollutants’ removal stage [66].

To effectively treat pharmaceuticals and other complex drugs, biological systems are supposed to involve enzymatic pathways. Fungi have the ability to secrete extracellular enzymes such as laccases and peroxidases that allow them to degrade a wide range of less degradable MPs [66]. It is suggested that, for maximum efficiency, a two-stage process or a symbiotic algae–bacteria–fungi system should be employed, where fungi initiate the breakdown of complex molecules, making them more accessible for subsequent microbial metabolism [74].

HMs may be removed through ion exchange, membrane filtration, or chemical precipitation. Those methods are expensive and may result in secondary sludge pollution [81]. It is supposed to be noted that biological treatment is effective for treating low-concentration effluents. For example A. niger was able to remove HMs (Hg and Pb) in low concentrations (below 25 mg/L) [82].

It should be emphasized that while anaerobic digestion (AD) is a mature and well-established technology [83], its application in this study is not an isolated objective but a component of an integrated, dual-purpose biological system. Within the framework of a circular economy and in response to the stringent requirements of Directive 2024/3019 [22], biomass yield remains a significant parameter. The biomass growth rate is directly proportional to the nutrient (N, P) uptake efficiency from wastewater [39]. Furthermore, focusing on the yield of these specific microorganisms is essential to the economic viability of the entire process. High biomass productivity, especially in co-culture systems, facilitates bio-flocculation and pelletization, which significantly reduces harvesting costs, which form a significant part of the total expenses. This approach ensures that the treatment costs are partially offset by energy recovery, aligning with the circular economy principles. Taking into account the yield of microorganisms, this remains relevant from a process optimization perspective. Therefore, biogas production and biomass growth tendencies allow for a more comprehensive system understanding and can be used to develop a more sustainable biomass-energy approach.

A limitation of this review process is the restriction to English-language publications, which may have led to the exclusion of relevant regional studies. Additionally, only major electronic databases were searched, and the gray literature and patents were not included, which may have omitted additional technical insights into large-scale applications.

5. Challenges and Future Directions

The maintenance of a viable and stable microorganisms culture faces technical obstacles. Based on the review–life cycle analysis (LCA), carbon balance, particular process conditions, and economic analysis are still not well described. It is recommended to perform those analyses at an early stage of the project to ensure the best outcomes [84].

One of the challenges related to microalgae application in wastewater treatment is the nutrient composition. The nutrient concentrations should be balanced through mixing different types of wastewaters [25]. During the selection of microorganisms for a treatment method, a few aspects should be considered: fast growth rate, high photosynthetic rate, and strong environmental tolerance [30]. When developing a harvesting method for biomass, the main aim is to improve biomass recovery, focus on reducing operating and maintenance costs, and minimize environmental impact. The influence of harvesting method on the quality of treated wastewater and its eventual reuse should be considered when choosing a suitable treatment method [17].

Cell dimensions, metabolic activity, and cell density are crucial to harvesting efficiency [85]. Biomass recovery is one of the most significant factors that should be considered when choosing microorganisms for wastewater treatment. Filtration, flotation, sedimentation, or centrifugation can be used for biomass separation. Some of these techniques are linked to high costs. The integration of various methods is recommended to obtain the best results [28]. Future experiments should focus on: low-energy harvesting, e.g., via magnetic separation [85], utilizing the pellet-forming nature of fungi [69], and facilitating easier gravity-based separation without expensive chemical coagulants. Developing mats and solid carriers where the biomass grows on surfaces enables simple mechanical scraping and eliminates the need for complex suspension-thickening [43].

The commonly applied biological treatment stage is primarily optimized for the removal of organic loads (BOD, COD) and nutrients (N, P). The dominant bacterial species present in activated sludge often lack the specialized metabolic pathways needed to degrade complex MPs [86]. On the contrary, specific microorganisms possess unique mechanisms, which allow them to remove persistent MPs (e.g., fungi secrete specific enzymes) [66].

The selection of plant and microbial candidates may be optimized under laboratory conditions; several parameters remain difficult to control in large-scale wastewater treatment systems. These include fluctuations in wastewater temperature, variability in wastewater composition, a slow growth rate, limited control over CO₂ supply, and challenges in maintaining specific microorganism cultures. Consequently, it may be challenging to transfer the performance observed in controlled environments to large, full-scale applications [87]. Notably, the described experiments were performed mainly at a laboratory scale [24,33]. A limitation of this study is that most experiments were conducted at laboratory scale. Consequently, the obtained results may not fully represent the behavior of the process under industrial or real wastewater treatment conditions, where operational complexity and environmental variability may significantly affect treatment efficiency. While the laboratory-scale experiments demonstrate proof of concept, full-scale systems are influenced by additional factors, such as temperature changes and fluctuations in wastewater flow and quality. Therefore, further pilot- and industrial-scale investigations are necessary to validate the practical applicability and scalability of the proposed approach.

One of the findings of this review is a systemic lack of standardized operational data, particularly regarding energy consumption (kWh/m³) for mixing and aeration. Most studies treat economic and energetic requirements as a ‘black box’, focusing solely on removal efficiency. Future research must bridge this gap by providing detailed information regarding operational conditions (including scale maturity), which is also highlighted in the Conclusions section.

6. Conclusions

This review shows that biomass-based systems are a promising option for wastewater polishing, allowing for the removal of residual nutrients and micropollutants under various operational conditions. However, their performance is highly variable and depends strongly on biomass type and environmental factors.

Despite their potential for resource recovery, two key barriers currently limit large-scale application: process instability under fluctuating conditions and inefficient biomass harvesting. In addition, the lack of consistent techno-economic data from full-scale studies makes it difficult to assess real-world feasibility.

Overall, improving process stability and developing cost-effective biomass recovery methods are the most critical steps needed for the successful implementation of these systems.