The Role of the Microalgae–Bacteria Consortium in Biomass Formation and Its Application in Wastewater Treatment Systems: A Comprehensive Review

Abstract

The optimization of wastewater treatment technologies using biological processes is no longer limited to improving the removal of organic matter and nutrients, as it is possible to reduce area and energy consumption, and recover value-added by-products. In this context, the microalgae–bacteria consortium is an alternative for reducing costs, as microalgae produce the oxygen required by bacteria to oxidize organic matter through photosynthesis. Additionally, it is possible to extract different by-products such as lipids, biofertilizers, biogas, alginate-type exopolymers, and others. Furthermore, bioflocculation occurs naturally through the adhesion of microalgae to the surface of bacterial flocs, without the addition of chemical products. This review discusses the main systems that utilize the microalgae–bacteria consortium, the metabolism of the microalgae–bacteria consortium, and its performance in removing organic matter and nutrients, as well as the effect of operating conditions on the physical properties of the biomass. Among the highlighted systems are sequencing batch and single-batch reactors, high-rate ponds, and continuous flow reactors. Among the systems discussed in this work, the sequential batch reactor configurations found better biomass formation and production of extracellular polymeric substances and the continuous flow reactors showed lower installation and operating costs. From this perspective, the potential for full-scale application of each system can be evaluated once the optimum operating conditions have been defined and the limitations of each system have been understood.

1. Introduction

The recent water crisis underscores the urgency of investigating and studying water usage to ensure access to a clean water supply and efficient wastewater management compliance with current standards and legislation [1]. Biological aerobic treatment (BAT) has been extensively employed globally and emerged as a prominent method to convert organic matter from effluents into biomass and carbon dioxide [2,3,4,5]. However, considering that traditional wastewater treatment plants are encountering technical limitations and high operational costs in removing nutrients, organic loads, and recovering biomass [6,7], systems with microalgae–bacteria consortium offer several advantages, including lower maintenance costs, compact designs, ease of operation, and reduced energy consumption [5,8]. According to Hamawand et al. (2023) [9], the water and sewerage sector is the largest electricity consumer in cities, consuming around 40% of the total energy consumption in urban areas.

In addition, instead of releasing CO₂ formed during the organic matter removal process, microalgae–bacteria consortium cultivation is gaining attention, since microalgae cultivation is effective in treating effluents and removing inorganic nutrients from wastewater [10], while also contributing to the reduction of greenhouse gas emissions [11]. Contrary to what happens in bacteria-dominant reactors via the nitrification–denitrification process, the consortium facilitates the incorporation of nitrogen into biomass through microalgae assimilation [12], especially in granular biofilm, where the microbiota is well defined. Leong et al. (2018) [13] observed that bacteria’s nitrification process in the symbiotic system facilitated the assimilation of oxidized nitrogen forms (NO₂⁻ NO₃⁻) by the microalgae Chlorella vulgaris. Similarly, Wang et al. (2016) [14] noted the proliferation of nitrogen-fixing bacteria in this symbiotic setup.

In addition, microalgae offer promising prospects for the generation of value-added by-products, including combustible options like bioethanol, biogas, and biodiesel, as well as non-combustible alternatives such as feed, vitamins, ultraviolet protection, bioplastics, and nanoparticles [15,16]. Moreover, this interaction has been associated with increased microbial community diversity [17], lipid production, sedimentation capacity [18], and biomass production for methane generation [19].

Granulation occurs in four stages: initiation, maturation, maintenance, and disintegration [20], involving algae adherence to flocculent sludge surfaces, subsequent bacterial attachment to algae, and continued consortia growth until biomass attachment and detachment achieve equilibrium [21]. Nowadays, some strategies are being utilized to monitor the algal growth process, including single-molecule techniques [22] and nanomechanical characterization [23]. Also, the consortium benefits from the degradation of organic matter by heterotrophs to bacteria while producing extracellular polymeric substances (EPSs), which are responsible for bacteria, microalgae, and cyanobacteria aggregation. At present, there is a wide array of research investigating the microalgae–bacteria consortium across various reactor configurations. These encompass continuous flow reactors, photobioreactors in single-batch and sequential-batch mode, and high-rate ponds [24,25,26,27].

In this context, the microalgae–bacteria consortium is a low-cost alternative for reducing energy consumption and gas emissions in aerobic treatment systems. This review thus outlines the sustainable potential for biomass formation and by-product recovery, along with an assessment of current systems and operational parameters utilizing this consortium for wastewater treatment. The focus lies on examining physical and biological characteristics and applied engineering, aiding in informed decision-making for future applications.

2. The Evolution of Aerobic Biological Systems towards the Microalgae–Bacteria Consortium

In aerobic systems, pollutant degradation occurs mainly through microorganisms’ respiration; biosynthesis and endogenous respiration can also occur [28]. It is well known that most of the BAT is implemented during domestic and industrial wastewater [29] on both small and large scales [30], and the most common aerobic technologies are aerated lagoons, biological filters, and activated sludge [28]. Nowadays, more studies are focused on process optimization, such as reducing the footprint and energy consumption [31].

For an efficient organic matter and nitrogen removal process, a well-defined microbiota is essential, such as the presence of heterotrophic and autotrophic microorganisms, the first group of which is responsible for organic matter degradation, and the second for the initial stages of complete nitrogen removal. Ammonia-oxidizing bacteria (AOB) or ammonia-oxidizing archaea (AOA) are responsible for oxidizing ammonia to nitrite (Figure 1), which is oxidized to nitrite by nitrite-oxidizing bacteria (NOB). Subsequently, the generated nitrate is removed from the system through denitrification, where NO₃⁻ is converted to atmospheric N₂ [32]. Additionally, there are two known NH₄⁺ oxidation processes: anammox, where nitrite is used as an electron acceptor under anoxic conditions and a low C/N ratio (0.2–2.5) [33,34], and comammox, a process where there is complete oxidation of ammonia to nitrate [35].

In systems where the biomass consists of flocculent sludge, a compartment is needed to separate the treated effluent and formed sludge [36]; however, more area for construction is necessary, which leads to the technology becoming more expensive. An excellent solution is cultivating a fast-settling aggregate, such as aerobic granular sludge (AGS) [37]. Compared to activated sludge systems, aerobic granular systems can reduce construction area requirements by up to 70% and energy consumption by 40% [38,39,40]. Therefore, the potential for symbiosis between bacteria and algae to optimize energy utilization in aerobic systems holds promise for wastewater treatment. The formation of aggregates primarily occurs through inoculation with activated sludge [19] or is obtained from a biomass-sourced upflow anaerobic sludge blanket (UASB) [41] or from a UASB reactor biomass combined with samples from eutrophied environments [38].

In recent years, an improvement on AGS has been studied. Since the 1970s, integrating microalgae into treatment systems has aimed to curb greenhouse gas emissions and supplant artificial aeration [11]. This approach has yielded commendable outcomes in nutrient and organic matter removal. Moreover, adoption of this approach yields biomass rich in intracellular lipids, fostering the production of value-added products [42,43,44]. The cultivation of microalgae and promoting the bacteria–algae consortium shows significant advantages, mostly in terms of improving the sludge sedimentation, the final effluent quality, the gas exchange, and the granular stability. Although previous studies have explored aggregates of microalgae and bacteria [45], a patent for the algal granulation process was issued based on the work of Park and Dolan (2015) [46], who observed the formation of these structures through inoculation with activated sludge and microalgae. The operational strategies and characteristics of microalgae–bacteria is discussed is the following chapters.

3. Insights into Algal–Bacterial Biomass

The microorganisms within wastewater treatment systems involving microalgae and bacteria can be categorized based on their disposition in the system as either suspended or aggregated. The latter group, known for its enhanced pollutant removal and biomass-settling capabilities, is referred to by various names, such as photogranules, oxygen photogranules, flocs of microalgae and bacteria, algal–bacterial aerobic granular sludge, algal granular sludge, or aggregates of activated sludge and microalgae [13,46,47,48,49,50,51]. Typically, these structures comprise a consortium of photoautotrophs, chemoautotrophs, and heterotrophs, stabilized and bound together within an extracellular polymeric substance matrix, thus falling within a spectrum of biogranules applicable to environmental engineering [49,52].

3.1. Formation Processes of the Microalgae–Bacteria Consortium (MABA)

The aggregates have been successfully cultivated in various configurations across different systems. He et al. (2018) [53] operated a sequential batch reactor and noted an initial aggregation of filamentous microalgae on the granules’ surface, followed by settling inside, resulting in a dark green coloration. Ahmad et al. (2017) [51] developed photogranular biofilm in a continuous regime and reported that filamentous microalgae can act as nuclei for the aggregates. In general, the presence of cyanobacteria and/or other filamentous microalgae is considered essential to the formation of MABAs [39].

Generally, a photogranular biofilm has either the appearance of filamentous aggregate (Figure 2a), due to the prevalence of filamentous microalgae, or more a compact structure (Figure 2b), which exhibits better sedimentability properties [24]. The compact form represents a mature state of filamentous aggregates, and the presence of microalgae improves the granulation process, decreasing the formation time and improving the biomass structure [48,49]. The microalgae’s species play an essential role in biofilm formation, which was reported by Arcila and Buitron (2016) [19]. The authors documented the physical characteristics of a bacteria–algae aggregate cultivated in a high-rate pond as a central aggregate with a predominance of diatoms with filamentous microalgae attached to them. In addition, later it was observed that an increase in this diatom nucleus led to fewer filamentous bacteria cells and enhanced sedimentation [54].

While most investigations into microalgae–bacteria consortia (MABAs) have been conducted in closed reactors operating in sequential batches [25,55,56], studies have also been undertaken in single-batch reactors [5,13], continuous flow systems [57,58], and high-rate ponds [24,59]. Furthermore, aggregate formation has been observed in high-rate ponds operating in a sequential batch regime [50]. Moreover, a crucial factor in MABA formation is the secretion of extracellular polymeric substances by microorganisms, primarily bacteria [47]. EPS, characterized by a permeable hydrogel primarily composed of carbohydrates and proteins, facilitates microalgae and bacteria aggregation while also acting as a carbon source and providing protection for microorganisms [52]. EPS plays a role in the formation of MABA nuclei [19].

3.2. MABA General Characteristics

One factor that contributes to the limited sedimentation capacity of microalgae, alongside factors such as cell size and low concentrations in sewage treatment, is the negative charge of the cells, hindering their aggregation. Furthermore, Daudt et al. (2019) [60] indicate that biological aggregates are mostly granular when at least 50% of the particles in the systems are larger than 0.20 mm. This helps to clarify the concept of particle size. For example, MABAs exhibit sedimentation properties comparable to other aerobic wastewater treatment systems (

Table 1. Size and sedimentation properties of aerobic biomass used in wastewater treatment.

Biomass	Size (mm)	SV (m.h−1)	SVI (mL.g−1)	Sedimentation (%)	References
AGS	0.2–16	18–130	<80	-	[49]
AS	0.0005–1	0.6–15	100–150	84.44	[13,14,49]
			76.8
MABAs	0.1–5	36–360	42.55	42.37
		8.3	57	92	[13,14,19,48,49]
				>90
MA	0.005–0.05	0.001–0.026	-	13	[13,54]

AGS: aerobic granular sludge; AS: activated sludge; MABAs: microalgae–bacteria consortia; MA: microalgae; SV: settling velocity; SVI: sludge volumetric index.

), such as activated sludge and aerobic granular sludge, offering a significant enhancement in harvesting capacity compared to microalgae alone.

The protein fraction of EPS serves to stabilize the negative charges of microalgae biomass [55]. Consequently, MABAs typically exhibit a high EPS-PN/EPS-PS carbohydrate ratio, ranging from 1.0 to 6.6, and can constitute 8% to 34% of the algal biomass composition, a critical determinant of sedimentation properties [48,59]. Also, according to Brockmann et al. (2020) [61], cyanobacteria are highly significant in MABA aggregates because they create EPS fractions in their layer that contribute to the aggregates’ stability. Furthermore, according to Satiro et al. (2022) [5], the assessment of EPS and its fractions in MABAs is significant since EPS might operate as a type of “glue” for cell adhesion in aggregates and because the viscosity and sedimentability of MABAs can be affected by these fractions.

3.3. The Feasibility and Application of Systems Utilizing MABAs

The effluent utilized for microalgae cultivation harbors organic pollutants, which microalgae absorb as cellular constituents like lipids and carbohydrates [61]. Among the genera of microalgae employed in MABA systems, Chlorella sp. and Scenedesmus sp. exhibit superior performance, attributed to their heightened capacity to absorb nitrogen, phosphorus, and carbonaceous organic matter, alongside achieving elevated growth rates [62,63]. Despite numerous studies aimed at enhancing microalgae-based technologies, extracting value-added products such as biofuels from these organisms remains a challenge.

Effective microalgae harvesting is crucial to obtain an effluent with low turbidity. However, studies indicate that harvesting costs range from 20% to 30% of the total production value due to the low sedimentability, small size, and low density of the algal biomass [64,65]. Dry algal biomass typically constitutes only 0.1% to 1.0% of the culture weight [66]. Consequently, commonly used techniques such as electrical, mechanical, and chemical methods are often limited and significantly escalate operational expenses [64,67]. A more economical alternative is bioflocculation, which involves promoting algae adherence to the surface of flocculent or granular sludge, forming aggregate without the need for chemical additives [68,69,70].

3.4. Obtaining Value-Added By-Products from MABAs

The recovery of by-products from cultivating pure cultures of microalgae is a widely discussed topic, as it promotes the substitution of fossil fuels with sustainable biodiesel, contributing to the circular economy. Microalgae store lipids intracellularly for their metabolic functions, and their content can vary depending on stress conditions imposed on the microorganisms [71], such as nutrient availability, lighting conditions, and cultivation medium [72,73]. Feng et al. (2011) [74] reported that nitrogen deficiency increases lipid production in the pure species Chlorella zofingiensis. The lipid content under the conditions of 0.04 and 1 g NaNO₃/L was 54.5% and 27.3%, respectively. The following year, the same authors confirmed this finding and inferred that phosphate deficiency also affects lipid productivity, but to a lesser extent than nitrogen deficiency.

In addition, algae–bacteria interaction in wastewater treatment systems hold promise in the field of biotechnology. Obtaining by-products from microalgal–bacterial aggregates can significantly contribute to the development of sustainable products. There are various possibilities for products to be recovered, including biodiesel [19], biopolymers [75], phosphorus [76], and biochar [77], among others, with potential applications across various economic sectors (Figure 3). Chen et al. (2021) [75] successfully recovered phosphorus and alginate-like exopolysacchrides (ALEs) in bacterial granular sludge and algal–bacterial granular sludge. The authors observed a higher ALE content in the granularity of the algae–bacteria consortium compared to the bacteria-only one. However, the recovered phosphorus content was similar in both biomasses. These findings are consistent with the work of Karakas et al. (2021) [78], who evaluated the phosphorus content recovered in conventional AGS and algal–bacterial AGS, reporting similar results (more than 83% of phosphorus recovered) for both systems.

The sub-product content and recovery is influenced by the operating conditions of the treatment system, and several studies are underway to assess which stress conditions are best suited for greater by-product production, such as light intensities and hydraulic retention time (HRT). Meng et al. (2019) [79] evaluated the effect of different light intensities on lipid production in sequential batch reactors (SBRs) and reported that the ideal light intensity range for lipid production is between 45 and 135 µmol/m².s. Katam and Bhatacharyya (2019) [80] compared different hydraulic detention times (2, 4, 6, 8, and 10 h) in reactors treating domestic sewage and observed that this parameter is directly related to the potential for lipid production: The higher the HRT, the higher the system’s lipid production. In addition, Arcila and Buitrón (2016) [19], working with MABAs developed in high-rate ponds, showed variations in methane production potential based on the hydraulic retention time. For an HRT of 10 h, the biomass formed produced 347.9 ± 2.0 mL CH₄/g.VS, while for an HRT of 6 h, the production was 290.4 ± 12.1 mL CH₄/g.VS.

4. Insights into Bacterial–Algal Biomass in Different Operational Configurations

Due to the diverse microbial community present in aerobic sludge aggregates, various metabolic pathways contribute to energy acquisition and carbon uptake. The success of process like SND (simultaneous nitrification and denitrification) and phosphorus removal hinges on the operational conditions applied during the development of aerobic granular sludge, including factors such as the composition of the feed, the duration of cycle phases, and the characteristics of the treated effluent [55,81]. Photogranulation alongside these factors is influenced by the species of microalgae attached to the biomass. Excessive growth of Stigeoclonium sp. accelerates the disintegration of photogranules, whereas the presence of cyanobacteria promotes the formation of more stable granules that are less prone to disintegration [82]. Meng et al. (2019) [79] associate luminosity with the presence of filamentous microalgae on the granular surface, which contributes to destabilization, resulting in less stable granules prone to disintegration. Conversely, He et al. (2018) [53] suggest that filamentous bacteria act as bridges, providing ample space for microalgae to adhere to aerobic granules.

In contrast with bacterial granular sludge systems, the microalgae–bacterial consortium demonstrates the ability to maintain granule stability even at low COD/N ratios [53,79]. Extracellular polymeric substances (EPSs) play a crucial role in understanding granule stability, as variations in protein (PN) and polysaccharide (PS) concentrations alter the physicochemical characteristics of the cell surface, including charge and hydrophobicity [83]. During the dark phase of the system, the increase in biomass within the granular consortium is attributed to the assimilation of CO₂ in the form of HCO₃, which is deposited as a reserve in the form of EPSs [51]. Consequently, there is a greater production of EPSs compared to aerobic granule systems, which may contribute to reducing the negative charge of the algal cell surface [44,84]. Arcila and Buitrón (2016) [58] suggest that the presence of nitrifying bacteria influences granule formation and stability, leading to the production of more sedimentable granules.

With the aim of developing systems with excellent performance in wastewater treatment, as well as promoting more efficient and less expensive microalgae-harvesting techniques, granular microalgae–bacteria were developed. Consortia have been extensively studied under different configurations. The main systems studied include sequential batch reactors (SBRs), single-batch photobioreactors (SBPs), high-rate ponds (HRPs), and continuous flow reactors (CFRs), which are shown in

Table 2. Summary table of the articles included in the study: different configurations of SBR, SBP, HRP, and CFR systems.

Reactor Model	Effluent Concentration	Operational Conditions	Treatment Performance	Biomass Characteristics	Microalgae Species	Remarks	References
SBR	Synthetic wastewater COD = 600 mg/L NH4+-N = 100 mg/L PO43–P = 23 mg/L	Volume = 1.4 L Natural lighting HRT = 8 h	96% COD removal 99% NH4+-N removal 46% TN removal	MLVSS = 5.4–5.5 g/L SVI30 = 30–40 mL/g	-	The algal–bacterial granules showed an excellent nutrient removal rate. Good EPS production contributed to granule stability.	[85]
SBR	Synthetic wastewater COD = 600 mg/L NH4+-N = 100 mg/L PO43–P = 10 mg/L	Volume = 1.4 L Natural lighting HRT = 8 h	95% COD removal 98% NH4+-N removal	MLVSS = 7.7 g/L SVI30 = 38 mL/g	-	The algae–bacteria consortium directly affected the biodiversity of the microbial community. The system showed low phosphorus removal efficiency.	[86]
SBR	Synthetic wastewater COD = 1200 mg/L NH4+-N = 200 mg/L PO43–P = 16 mg/L	Volume = 1.4 L Illumination = 121 µmol/m2.s HRT = 12 h	97% COD removal 99% NH4+-N removal 90% TP removal	MLVSS = 8.6 g/L SVI30 = 78 mL/g Granule diameter = 3.25 mm	Trebouxiophyceae sp. Bacillariophyceae sp. Chlorophyceae sp.	There was good formation of the algae–bacteria consortium and good nutrient removal. High protein content was found in the biomass.	[21]
SBR	Synthetic wastewater COD = 200 mg/L NH4+-N = 200 mg/L TP = 5 mg/L	Volume = 3.6 L HRT = 12 h	-	Dark granules were observed from day 7.	Diatomea Chlorophyceae Chrysophyceae Trebouxioplyceae	The eukaryotic algae were replaced by green algae during the granulation process.	[53]
SBR	Synthetic wastewater COD = 300 mg/L NH4+-N = 30 mg/L TP = 10 mg/L	Volume = 0.92 L Illumination = 6000 LUX	98% COD removal 50% TP removal	MLVSS = 28.9 mg/L SVI5 = 24 mL/g	Chlorella sp. Scenedesmus sp.	Microalgae predominated over granules. The granules offered good nutrient removal performance.	[87]
SBR	Real wastewater COD = 498 mg/L	Volume = 1.5 L Illumination = 200 µmol/m2s HRT = 4, 6, and 8 h	54% COD removal 47% TP removal	Stabilized granules were observed from day 8.	Nannochloropsis gaditana Chlorella, Scenedesmus, and Tetradesmus	An inconsistency was observed in the process of homogenization of the medium, which may be related to the chemical precipitation of phosphorus.	[88]
SBR	Real wastewater COD = 498 mg/L NH4+-N = 34.1 mg/L PO43–P = 4.9 mg/L	Volume = 1 L Illumination = 122 µmol/m2s	37% COD removal 70% NH4+-N removal	-	Chlorella sorokiniana	The respiration of endogenous microalgae had an impact on oxygen absorption during the dark phase of the system.	[89]
SBR	Synthetic wastewater COD = 492 mg/L TN = 101.3 mg/L TP = 5.2 mg/L	Volume = 1.72 L HRT = 12 h	85% COD removal 15% TN removal 49% TP removal	-	Green algae and diatoms	Photobioreactors with the presence of an algal–bacterial consortium are a promising technology for removing macro- and micropollutants.	[80]
SBR	Synthetic wastewater NH4+-N = 180 mg/L	Volume = 3.2 L	66% NH4+-N removal	MLVSS = 950 mg/L (day 200)	Chlorella vulgaris Nitrosomonas europaea	RNA analysis of the sludge showed an increase in nitrosomonas, which may indicate good performance in the system’s partial nitrification process.	[90]
SBR	Synthetic wastewater COD = 150 mg/L TN = 5 mg/L TP = 40 mg/L	Volume = 12.5 L Illumination = 4000 LUX	39% COD removal 74% TN removal 94% TP removal	Increase in biomass of 90%	Chlorella sp. and Scenedesmus sp. Anabaena sp. and Oscillatoria sp.	The consortium with the 5:1 algae/sludge ratio achieved greater granular structuring and biomass growth.	[91]
SBR	Real wastewater COD = 249 mg/L TN = 63 mg/L TP = 6.7 mg/L	Volume = 12.5 L Illumination = 200 µmol/m2s HRT = 2.5 days	97% COD removal 88% TN removal 88% TP removal	-	Chlorella sp. Scenedesmus sp.	The use of a moving support material in photobioreactors can be an essential factor in nutrient removal in systems with lower HRT.	[92]
SBR	Synthetic wastewater COD = 179 mg/L NH4+-N = 15.1 mg/L PO43–P = 11.5 mg/L	Volume = 500 L Illumination = 120 µmol/m2s	53% COD removal 68% NH4+-N removal	-	Chlorella sp.	Mixing wavelength photoperiods (blue:green) is a suitable strategy for increasing biomass production and removing organic matter and nutrients.	[93]
SBR	Synthetic wastewater COD = 300 mg/L TP = 5 mg/L	Volume = 0.92 L HRT = 6 h	99% COD removal 58% TP removal	MLVSS = 112 mg/L (day 40)	-	The strategy of fractionating the lighting period contributed to an increase in chlorophyll.	[94]
SBR	Synthetic wastewater COD = 300 mg/L TP = 5 mg/L	Volume = 0.92 L HRT = 12 h	99% COD removal 55% TP removal	-	Chlorella vulgaris	In the treatment containing the microalgae–bacteria aggregate, there was competition between these organisms, which reduced the performance of the bioreactor’s nitrate removal efficiency.	[95]
SBR	Synthetic wastewater COD = 400 mg/L NH4+-N = 100 mg/L PO43–P = 10 mg/L	Volume = 3.5 L Illumination = 140 µmol/m2s	74% COD removal 69% NH4+-N removal	MLVSS = 6.3 g/L	Chlorella sp., and diatoms, Navicula sp.	Increasing the C:N ratio from 4:1 to 8:1 resulted in greater biomass accumulation.	[25]
SBR	Real wastewater COD = 294 mg/L	Volume = 1.5 L Illumination = 113 µmol/m2s HRT = 12 h	88% COD removal	SVI30 = 100 mL/g	Tetradesmus sp.	Microalgae inoculation may not be necessary to develop algal–bacterial AGS when treating real municipal wastewater.	[96]
SBR	Synthetic wastewater COD = 500 mg/L NH4+-N = 50 mg/L PO43–P = 10 mg/L	Illumination = 255 µmol/m2s HRT = 8 h	99% COD removal 93% TP removal	MLVSS = 4.1 g/L	Chlorococcum sp. and Chlorella sorokiniana	The algal–bacterial granules were observed to break into filamentous flocs and relatively compact fragments due to the hydrolysis of anaerobic cores.	[97]
SBP	Real wastewater COD = 380 mg/L NH4+-N = 39.4 mg/L PO43–P = 8.8 mg/L	Volume = 14 L Illumination = 700 LUX	91% COD removal 100% TN removal 94% TP removal	-	Green and blue filamentous algae	The best inoculum ratio was 1090 mg/L of microalgae:200 mg/L of sludge, after which there was a significant increase in the sedimentability of the biomass.	[45]
SBP	Synthetic wastewater COD = 1200 mg/L NH4+-N = 45 mg/L PO43–P = 12 mg/L	Volume = 0.5 L Illumination = 200 µmol/m2s	87% COD removal 99% TN removal 84% TP removal	SVI30 = 42.5 mL/g	Chlorella sp.	The best results were achieved in systems with continuous lighting. There was growth of rhizobacteria.	[98]
SBP	Synthetic wastewater COD = 60 mg/L NH4+-N = 28.7 mg/L PO43–P = 1.3 mg/L	Volume = 2 L Illumination = 100 µmol/m2s	95% TN removal 92% TP removal	MLVSS = 2 g/L	Chlorella vulgaris	Various inoculum concentrations were studied. The best ratio was 400 mg/L algae:200 mg/L sludge.	[99]
SBP	Synthetic wastewater COD = 440 mg/L NH4+-N = 50 mg/L PO43–P = 12.8 mg/L	Volume = 2 L Illumination = 4600 LUX	91% COD removal 98% TN removal 84% TP removal	-	-	For an initial concentration of 300 mg/L of activated sludge, various initial concentrations were tested, with the best being 700 mg/L.	[100]
SBP	Real wastewater COD = 440 mg/L NH4+-N = 50 mg/L	Volume = 1 L Illumination = 70 µmol/m2s	97% TN removal	-	Chlorella vulgaris	The best inoculum ratio was 100 mg/L sludge:75 mg/L algae.	[13]
SBP	Synthetic wastewater COD = 1130 mg/L NH4+-N = 260 mg/L PO43–P = 28.5 mg/L	Volume = 1 L Illumination = 200 µmol/m2s	83% COD removal 75% TN removal 100% TP removal	-	Chlorella vulgaris	By studying the composition of the biomass, it was possible to use it for biofuels and animal feed.	[101]
SBP	Real wastewater COD = 1000 mg/L	Volume = 1 L Illumination = 100 µmol/m2s	85% COD removal 87% TN removal 86% TP removal	MLVSS = 1100 mg/L	Scenedesmus sp.	The reactors showed a greater possibility of harvesting the biomass to generate bioproducts with added value.	[5]
HRP	Real wastewater COD = 593 mg/L NH4+-N = 72 mg/L PO43–P = 16 mg/L	Volume = 50 L Illumination = 200 µmol/m2s HRT = 10, 6, and 2 days	91% COD removal 99% NH4+-N removal 49% TP removal	-	Cyanobacteria Diatoms Green algae	At an HRT of 2 days, the system collapsed. Methane production = 55.7 mL/gVSd was observed.	[19]
HRP	Real wastewater COD = 591 mg/L NH4+-N = 64 mg/L PO43–P = 15 mg/L	Volume = 50 L HRT = 10 days	84% COD removal 98% NH4+-N removal 92% TP removal	SVI30 = 40–740 mL/g	Cyanobacteria Diatoms Green algae	The production of EPSs was fundamental in the formation and structuring of the granule.	[58]
HRP	Real wastewater COD = 332 mg/L NH4+-N = 39 mg/L PO43–P = 361 mg/L	Volume = 22,000 L HRT = 10 days	98% COD removal 86% NH4+-N removal 98% TP removal	SVI30 = 45–109 mL/g	Cyanobacteria Diatoms Green algae	After 26 days of operation, the HRPs were fully functional. The bacterial community was established after 10 days of operation.	[102]
HRP	Real wastewater COD = 296.5–858.3 mg/L NH4+-N = 58–136 mg/L PO43–P = 7.8–27.7 mg/L	Volume = 5–17 m3	70% COD removal	-	Scenedesmus almeriensis, diatoms, and green algae	Operating depths needed to be optimized. Nitrogen and phosphorus consumption was influenced by operating conditions.	[103]
HRP	Real wastewater COD = 338 mg/L NH4+-N = 62 mg/L	Volume = 60 L	79% COD removal 100% NH4+-N removal	-	-	Carrying out initial batches before the continuous system period was a fundamental strategy for the biomass.	[24]
HRP	Real wastewater COD = 517 mg/L NH4+-N = 86 mg/L PO43–P = 43 mg/L	Volume = 80 L Illumination = 200 µmol/m2s HRT = 6 days	89% COD removal 89% NH4+-N removal 23% TP removal	-	Chlorella and Diatoms	Increasing the retention time of solids in reactors favored granular formation and increased the sedimentation properties of the biomass.	[104]
CFR	Synthetic wastewater COD = 300 mg/L NH4+-N = 100 mg/L PO43–P = 10 mg/L	Illumination = 1100 LUX HRT = 6 h	95% COD removal 99% NH4+-N removal 46% TP removal	MLVSS = 4.3 g/L	Phormidium sp.	The internal separator facilitated hydraulic selection and selective sludge discharge.	[51]
CFR	Synthetic wastewater COD = 300 mg/L NH4+-N = 152 mg/L PO43–P = 47 mg/L	Volume = 2 L Illumination = 200 µmol/m2s HRT = 24 h	90% COD removal 94% NH4+-N removal 9% TP removal	MLVSS = 4 g/L	Scenedesmus sp.; Closterium sp.; Chlorella sp.; Diatoms; Oscillatoria sp.	Autotrophic bacteria, heterotrophic bacteria, algae, and PAOs coexisted in the reactor.	[57]

. The table shows each article’s type of reactor; the effluent concentration in terms of the main parameters of monitoring, such as chemical oxygen demand (COD), total nitrogen (TN), NH4⁺, NO₂⁻, NO₃⁻, PO₄³, and total phosphorus (TP); operational conditions; treatment performance; biomass characteristics; microalgae species; remarks; and reference numbers. Figure 4 shows the main cycles and operation of each configuration discussed in this work.

4.1. Sequential Batch Reactor (SBR) Systems

Aerobic photogranules are preferably cultivated in photobioreactors operated in sequencing batches, where the operation can be divided into time cycles. These, in turn, are characterized by the phases of filling, reaction (aerobic, anoxic, or anaerobic), sedimentation, and discharge of treated effluent (supernatant) [105]. The filling phase consists of the raw sewage entering the system by pumping or gravity. This stage can be carried out in an ascending or descending direction. The inclusion of the anaerobic/anoxic phase during the feeding of the effluent to the reactor has shown satisfactory results in denitrification and phosphorus removal [36]. In this stage, the denitrification process takes place due to the remaining nitrite/nitrate from the previous cycle.

The aeration phase in AGS cultivation takes place solely through the insertion of artificial oxygen so that aerobic biological processes can take place. However, many systems with algal–bacterial consortia use the oxygen produced by the microalgae during photosynthesis to carry out aerobic processes [82]. Due to the significant concentration of photosynthetic microorganisms, the algae–bacteria consortium can photosynthetically produce O₂ necessary for the oxidation of organic matter and nitrification. The occurrence of these processes in systems composed solely of granular sludge is exclusively due to the provision of aeration, resulting in greater energy use [49]. Meanwhile, studies using photosynthetic aeration have shown excellent results, with ammonia nitrogen and phosphorus removals of over 90% [82]. The sedimentation phase consists of separating the biomass from the previously treated effluent while the lighter and less sedimentable biomass remains in suspension.

Like the granular sludge system, the biomass in the granular consortium has excellent sedimentation capacity, producing an effluent free of suspended solids [51,87]. In the discharge stage, the treated effluent is discharged, carrying the lower-weight biomass that remained in suspension in that cycle [105]. The operating principle of SBR facilitates the retention of high concentrations of photogranular sludge and eliminates the need for settling tanks to separate the biomass from the treated effluent, as well as the return of the sludge to the reactor [83], just like sequential batch reactors, but they need to be exposed to lighting.

4.2. Single-Batch Photobioreactor (SBP) Systems

Studies conducted in single-batch reactors typically aim to investigate the interactions between microalgae and bacteria [17,21,101]. These works, as depicted in Table 2, have been conducted using both pure cultures [13,106] and mixed cultures [17,45].

Leong et al. (2018) [13] introduced Chlorella vulgaris into the treatment system and noted that the presence of oxidized forms of nitrogen facilitated the growth of microalgae. They hypothesized that during nitrogen assimilation, microalgae excrete H⁺ to maintain intracellular pH, thus conserving energy for cell growth by preserving intracellular neutrality. Additionally, an increase in microbial diversity was observed, including the proliferation of nitrogen-fixing bacteria such as rhizobacteria [106]. Furthermore, studies exploring the application of photosynthetic oxygen have yielded promising results, with batch studies demonstrating over 90% removal of ammoniacal nitrogen and phosphate [17,45,99]. Another critical aspect addressed is biomass harvesting, a significant challenge for commercialization [17,18,44]. Wang et al. (2016) [14] also observed a positive impact on the sedimentation of bacterial biomass.

Su et al. (2012) [45] assessed sedimentability by measuring the total suspended solids (TSS) of non-sedimented biomass. They observed a reduction from 1.64 to 0.05 g/L in the algae–bacteria consortium, while the biomass with microalgae alone exhibited a TSS above 0.8 g/L. Wang et al. (2016) [44] evaluated the sludge volume index (SVI), obtaining values of 76.8 g/L for systems with activated sludge only and 42.55 g/L for co-cultivation with Chlorella sp. Leong et al. (2018) [13] assessed sedimentability through flocculation efficiency, obtaining indices of 88%, 42%, and 1.23% for the biomass of bacteria only, algae–bacteria co-cultivation, and systems with microalgae only, respectively.

4.3. High-Rate Ponds (HRPs)

The treatment of wastewater combined with the controlled production of microalgae in high-rate ponds is an adaptation of stabilization ponds, a model developed by Oswald in 1988 [107]. HRPs consist of an open, shallow lagoon with a system of rotors to move the liquid mass. They can be considered open photobioreactors that combine wastewater treatment and microalgae growth [108]. However, finding an economically viable alternative for harvesting the biomass remains challenging, and little is known about the behavior of the aggregate [19,54].

This capacity depends on the microorganisms’ sedimentation rate; however, the construction of a decanter can become economically and technically unfeasible—for example, for the production of microalgae targeting high-value-added compounds, harvesting efficiencies must be above 99% [109]. Sutherland et al. (2018) [110] treated anaerobic lagoon effluent in HRPs with an HRT of 8 days. The biomass was then sent to decanters with a solids retention time (SRT) of 9 h, and a harvesting efficiency of 4–12% was observed due to the low flock formation. Thus, the authors concluded that the 8-day HRT was considered insufficient. Park et al. (2011) [111] indicated the retention time for the separation of the HRT from the solids. In other words, it is necessary to recirculate the biomass to promote better sedimentation with shorter hydraulic retention times; in this study, the harvesting technique consisted of coagulation followed by flocculation and sedimentation [109].

In addition to efficient harvesting, HRPs must also produce biomass in large quantities, and area productivity is a parameter used to determine the most productive system [112]. Furthermore, operational and environmental factors can also influence production. Buchanan et al. (2018) [113] varied the water depth (0.32–0.55 m) and observed higher productivity in the 0.32 m pond (28.3 gSS/m²d). Similarly, Arbib et al. (2017) [112] varied the height between 0.15 m and 0.30 m and found the highest productivity (26 gSS/m²d) in the 0.30 m deep pond.

Other factors that can influence productivity are solar radiation and temperature. Couto et al. (2015) [114] varied solar radiation (387.1 to 1087.3 µE/m²d) and found no significant difference in area productivity. However, when area productivity was determined in terms of chlorophyll-a, the lowest values were obtained in the systems with the highest radiation, possibly due to the photoinhibition than can occur with higher solar radiation. As for the temperature factor, studies have shown that biomass production increases in warmer seasons [113,115]. Radiation and temperature are also parameters that can influence the performance of HRPs. Couto et al. (2015) [114] worked with various levels of radiation and observed that, although it had an effect on biomass productivity, it had no effect on nitrogen assimilation. Regarding temperature, great efficiencies have been observed in warmer seasons [114,115]. As for HRT, unlike the behavior of microalgae biomass productivity, higher values favor sewage treatment performance [116].

Hydraulic retention time (HRT) is also a parameter related to biomass production. It is known that as the HRT increases, the ratio of microalgae:bacteria decreases. It has been reported that reducing the HRT from 8 to 4 days can lead to a 24.5% increase in microalgae production [111,116]. Johnson et al. (2020) [117] recommend an HRT of 4 to 7 days for HRPs. Similarly, Park et al. (2011) [111] reported an 85% increase in microalgae biomass when the HRT was reduced from 8 days to 4 days.

The injection of CO₂ into HRPs is reported to be one of the factors responsible for increasing biomass production [8]. In addition to controlling pH, it serves as a carbon source for microalgae. It has been reported that CO₂ assimilation occurs more efficiently in shallower HRPs [118]. Another aspect identified as relevant to productivity is the pre-disinfection of sewage in the lagoons. Santiago et al. (2013) [119] observed that HRPs fed with disinfected sewage increased the productivity of microalgae but showed lower productivity of total biomass. In general, the groups identified in HRPs are Chlorophyceae, diatoms, Cyanophyceae, and Euglenaceae [110,120]. However, changes in species abundance do not influence effluent treatment performance.

Nutrient loading also affects performance in sewage treatment. Sutherland et al. (2017) [120] varied the concentration of NH₄⁺ nitrogen from 19.9 to 39.7 g/m³ and observed that the higher load provided greater physiological performance of the microalgae and better seawater treatment performance. Another important aspect is the size of the treatment systems. For example, Sutherland et al. (2020) [115] operated three HRPs of different volumes (1.5 m³, 90 m³, and 2900 m³) and observed that the lowest N-ammonia removal occurred in the pond with the largest volume.

Recent studies on algae–bacteria consortium in high-rate ponds have investigated the influence of detention time and hydraulics [19] and solar radiation [54] on aggregate structuring and sewage treatment. From the works cited in Table 2, it can be seen that the aggregates withstood 10-day and 6-day HRTs but collapsed with an HRT of 2 days. The evaluation of radiation levels showed the formation of stable aggregates with green microalgae, diatoms, filamentous cyanobacteria, and fungi at radiation levels of 3800 Wh m⁻²d⁻¹ and 2700 Wh m⁻²d⁻¹. The SRT also influences the formation of these structures. It has been observed that the formation was favored when the SRT exceeded 12 days [104]. However, is has also been noted that an SRT of 18 days can negatively affect the stability of the aggregates [24].

Lastly, Arcila and Buitrón (2017) [19] assessed sedimentation through sedimentation velocity, total suspended solids of the effluent, and sedimentation efficiency, observing that increasing the HRT also improves the sedimentation of the biomass. In addition, the methane production of the aggregates formed was also assessed, with a yield of 348 mL CH₄ gVS⁻¹. As a proposal for the formation of aggregates in high-rate ponds, Arcila and Buitrón (2017) [58] suggest that the formation of the aggregate may be associated with the appearance of a central aggregate of diatoms attached to filamentous microalgae, which, at a later stage, are no longer present, thus improving their sedimentability.

4.4. Continuous Flow Reactors (CFRs)

Currently, CFRs are the most widely used for the treatment of domestic effluents using microalgae–bacteria aggregate [50,81]. Continuous flow reactors have lower installation and operating costs, as well as simpler operation [56]. When applied to systems with microalgae–bacteria aggregation, they require greater control of the feed to ensure the structural integrity of the biomass in its stratified form [121].

Furthermore, some studies have achieved success in promoting a consortium between algae and bacteria in continuous flow reactors (Table 2). Ahmad et al. (2017) [51] compared bacterial AGS and algal–bacterial AGS in CFR with recirculation. The results showed better performance in the reactor with algae, achieving COD, TN, and TP removals of 100%, 98%, and 64%, respectively. Additionally, the CRF system exhibited greater energy efficiency compared to the SBR with algal–bacterial consortium used in the study by Huang et al. (2015) [86]. Another important operational parameter is the HRT. High HRTs result in higher costs, possibly due to the lower amount of biomass in suspension inside the reactors. In an algal–bacterial CFR, it was possible to establish an HRT of 6 h, treating low-concentration effluent (COD: 300 mg/L; N-NH₄⁺: 100 mg/L; P-PO₄: 10 mg/L). In cases where CFRs require an internal sedimentation zone, the selection of flow must be in line with the sedimentation speed of the particles, i.e., only biomass with little sedimentation will be washed out of the reactor into the final effluent [122].

Another critical aspect in CFR is the amount of dissolved oxygen (DO) injected into the system, which must facilitate the growth of heterotrophic biomass and the expected aerobic process, such as organic matter degradation and nitrification. Conversely, the production of photosynthetic oxygen is linked to the applied light intensity, with higher light intensity favoring the formation of microalgae flocs and bacteria with good sedimentability [123,124]. Therefore, parameters such as SRT, DO, and light intensity need to be well defined when planning CFR systems.

5. Conclusions

This review aimed to address the characteristics and formation process of the microalgae–bacteria consortium, commonly known as a MABA, as well as the research conducted using this consortium with several reactors’ configurations. Studies demonstrate that MABAs effectively remove COD, TN, and TP, with continuous flow reactors exhibiting lower installation and operating costs. In sequential batch reactor configurations, better extracellular polymeric substance formation is observed, resulting in more stable and sedimentable granules, whereas higher concentrations of microalgae do not favor good aggregate formation. However, high-rate ponds are the most widely used configurations on a full scale, allowing for good sedimentation of the biomass and the production of value-added products from the microalgae.

Light intensity plays a crucial role in system operation, as aggregates assimilate CO₂ during the light period. The dark phase, in the form of HCO₃⁻, facilitates the generation of biomass associated with extracellular polymeric substances. However, there is a pressing need to thoroughly investigate the performance and stability of microalgae–bacteria biomass resulting from varying reactor designs and applied operating conditions. This is essential for identifying systems with optimal financial and operational viability for real-scale implementation. It is worth nothing that achieving a more stable microalgae–bacteria consortium results in aggregates less prone to disintegration. This not only enhances nutrient removal capabilities but also enables the extraction of value-added products. Lastly, the main challenge in operating these systems is maintaining the microalgae and making operational adjustments to the reactors. These adjustments aim to increase the sedimentability of the biomass generated from effluent treatment, facilitating biomass recovery. Once separated and harvested, this biomass can serve as an alternative source of energy, animal feed, fertilizer production, and biofuel generation.