The aim of this review paper is to identify the strengths and weaknesses of digestion reactors, characterize the technological parameters that influence the course of anaerobic wastewater treatment, present primary reactor design types and their technological performance, and indicate the potential avenues of development of anaerobic methods for contaminant removal. A novel aspect of this review paper is the presentation of innovative technological and design solutions that minimize or eliminate the weaknesses of anaerobic wastewater treatment systems. The focus is on installations with a high technology readiness level (TRL), the effectiveness of which was confirmed in a pilot and fractional–technical scale. The improvements made in these installations allow to bolster wastewater treatment effectiveness, increase the efficiency of biogas and methane production, promote the removal of biogenic compounds, and save energy.
The multifunctionality of anaerobic wastewater treatment is a great strength of these technological solutions. Sustainable biogas systems include processes for treating wastewater and other organic waste to protect the natural environment; converting low-value material into higher-value material; and producing electricity, heat, and/or advanced gaseous biofuels, including hydrogen [ 51 ]. The feasibility of biogas production can contribute to the decentralization of energy security by balancing local energy supplies [ 52 ]. The methane fermentation of wastewater and the production of anaerobic digestate also contributes to improving the efficiency of nutrient uptake in agriculture by replacing synthetic fertilizers based on fossil fuels with biofertilizers, as well as to economical water management in agriculture and industry [ 53 ]. Following the circular economy assumptions, anaerobic digestion contributes to the treatment of wastewater streams throughout the food supply chain. Waste previously considered difficult to neutralize is being reintroduced into the production cycle where water, organic material, and nutrients are returned to the soil to replace chemical fertilizers [ 54 ].
The use of anaerobic wastewater treatment systems is in line with the assumptions of a circular economy [ 42 ]. Biomethane as a source of renewable energy, digested sludge used as a biofertilizer, and treated wastewater are the main pillars of the resource recovery process [ 43 ]. Biogas is a universal energy carrier that can be used on site or stored [ 44 ]. In addition, its production and implementation technologies have a significant impact on the reduction of gas emissions from conventional energy carriers [ 45 ]. It should also be noted that anaerobic digestion processes play a key role in wastewater treatment processes in terms of ensuring their energetic neutrality [ 46 ]. Anaerobic reactors merge the processes of wastewater treatment and water recovery [ 47 ], the production of renewable energy [ 48 ], and the production of biofertilizers and soil improvers that are widely used on marginal, degraded, and low-quality soils [ 49 ]. Their use reduces greenhouse gas emissions and directly affects the energy-efficient protection of surface waters, groundwaters, and groundwater reservoirs [ 50 ].
Anaerobic reactors are quite often used to treat food-industry effluent [ 19 20 ]. They have been successfully used to degrade waste from the production of dairy goods [ 21 22 ], fruit and vegetables [ 23 ], meat [ 24 ], beer [ 25 ], spirits [ 26 ], paper/paper-pulp [ 27 ], chemicals [ 28 ], and pharmaceuticals [ 29 ]. What makes such waste particularly well-suited to anaerobic processing is its high content of labile organics, which are converted to biogas under anaerobic conditions [ 30 ]. Pre-treatment can be used for waste containing hard-to-degrade, biodegradation-resistant or digestion-inhibiting substances [ 31 ]. Pre-treatment options include microwave radiation [ 32 33 ], ultrasound [ 34 ], high-voltage disintegration [ 35 ], thermal depolymerization [ 36 ], and treatment with acids, bases [ 37 ], or strong oxidizing agents [ 38 ]. Industrial effluent has a consistent composition due to the uniform production cycle and the use of retention/equalization tanks [ 39 ]. This ensures that the ratios of carbohydrates, proteins, and lipids remain fairly stable in the long term [ 40 ]. Industrial effluent tends to have a high and uniform temperature, which further facilitates its processing in various types of anaerobic reactors [ 41 ]. The commercial applications and market share of anaerobic technologies are presented in Figure 2
Digesters offer a wide range of applications for industrial and municipal sewage treatment systems [ 5 ]. Anaerobic sewage treatment methods can efficiently biodegrade organic matter while keeping investment and operating costs relatively low [ 6 ]. Anaerobic plants and accompanying facilities do not take up much space and produce little anaerobic surplus sludge, further bolstering the attractiveness of the technology [ 7 ]. Finally, digestion technologies can also be used to produce and capture methane-rich biogas—a considerable advantage over the aerobic-activated sludge process [ 8 ]. Anaerobic methods have been shown to be particularly suited to treating food-industry effluent [ 9 ], which is characterized by high pollutant loads that can only be removed with aerobic methods in large-scale plants, with considerable energy expenditures (for the intensive aeration of the wastewater) and the production of hard-to-manage sludge [ 10 ]. The clear advantages of anaerobic wastewater treatment have been recognized by manufacturing operators, as evidenced by the growing number of new full-scale plants. Furthermore, existing energy-consuming aerobic systems have been upgraded and retrofitted with new technologies [ 11 12 ]. Table 1 presents a comparison between aerobic and anaerobic wastewater treatment methods.
The development of wastewater treatment methods entails searching for novel, efficient, and technologically feasible solutions that could replace those currently in use [ 1 ]. The aim is to achieve not only the best possible performance in terms of pollutant removal, but also expected technological effects driving down investment and operating costs [ 2 ]. These criteria have greatly increased interest in pursuing anaerobic technologies, rapidly expanding the range of available technologies and deploying anaerobic reactors [ 3 4 ]. As shown in Figure 1 , in recent years, there has been growing interest in the studies focusing on both anaerobic wastewater treatment and anaerobic biological reactors.
The most energy-efficient of these is the COreduction by H. Conversely, the acetotroph-mediated acetoclastic reaction is the least efficient. Some methanogens do not belong to any of these categories. For example, some are hydrogen-methylotrophs which use hydrogen to reduce methanol and synthesize methane [ 78 ]. Methanogenesis is the bottleneck for the entire process, since methanogenic microorganisms grow much slower than acidogenic bacteria [ 79 ]. Therefore, the efficiency of the entire process is contingent on ensuring optimal conditions for methanogenesis [ 80 81 ]. All methane-producing microorganisms are classified into a separate domain.are grouped into two fundamentally different domains of organisms: the domain(most of the extant strains of bacteria, no methanogens, few extremophilic species) and the domain(all methanogens, many extremophilic species) [ 82 ].are some of the oldest known microbes, vastly different from the otherThese differences relate not only to their metabolism, but also to some of their morphological features [ 83 ]. For example,lack the typical peptidoglycan (murein) mesh.only has an outer protein layer, whereas the cell wall ofis built of a polysaccharide constituting uronic acids, neutral sugars, and amino acids [ 84 ]. Ribosomes of methanogens are similar in size to those of, but the sequences of ribosomal RNA (especially 16S rRNA) are completely different. It is believed that approximately 70% of the methane produced in anaerobic reactors is synthesized from acetic acid by, including, and. Most of the remaining methane is generated by CO-reducing hydrogenotrophs, such as, andspp. [ 85 ]. An overview of the conversions occurring during the anaerobic digestion of organic compounds is presented in Table 3
Methanogenesis is the last step of the digestion process, and the one where biogas is produced [ 73 ]. The currently known species of methane-producing microbes are divided into three main groups according to their nutritional requirements [ 74 ]. Hydrogenotrophs oxidize Hand reduce COinto CH; methylotrophs use methanol, methylamines, or dimethyl sulfides to synthesize CH; and acetotrophs use acetate to produce CH Table 2 ).
The same groups of microbes carry out acidogenesis, by which the hydrolysis products are degraded into low-molecular-weight organic compounds, including mainly fatty acids—primarily acetic acid, followed by propionic acid and butyric acid, as well as smaller quantities of isobutyric acid, valeric acid, isovaleric acid, caproic acid, and lactic acid [ 66 ]. Acidogenesis can also produce alcohols, such as ethanol, methanol, and trace amounts of propanol [ 67 ]. Other major end products include hydrogen and carbon dioxide [ 68 ]. The preferred outcome of the anaerobic wastewater treatment technology is the direct production of acetate, carbon dioxide, and hydrogen. This fermentation pathway provides more energy for the microbes, and its products can be directly taken up by methanogens [ 69 ]. Other products must first be converted by obligatory hydrogen-producing bacteria via acetogenesis [ 70 ]. Acetates are generated through one of two pathways. The first, which is usually critical, converts organic substances generated during the preceding steps (in particular, butyric acid, propionic acid, and long-chain fatty acids) to produce hydrogen [ 71 ]. The other variant occurs when homoacetate bacteria use hydrogen to reduce carbon dioxide [ 72 ].
Hydrolysis entails the biochemical and accompanying physicochemical depolymerization of complex organic compounds into monomers or dimers that can be directly consumed by bacteria [ 62 ]. This stage is crucial for the successful biodegradation of organic compounds in subsequent steps [ 63 ]. Proteins are hydrolyzed into amino acids, and polysaccharides into simple sugars, whereas lipids are hydrolyzed into polyols and fatty acids [ 64 ]. Hydrolysis is carried out by a wide spectrum of microbes, including facultative and obligatory anaerobes of the, andgenera [ 65 ].
Anaerobic wastewater treatment is predominantly based on conversions that occur during methanogenic fermentation [ 55 ]. Anaerobic digestion (AD) is a four-step process, with each successive step conducted by different groups of microbes and yielding different end products [ 56 ]. The assorted steps of the process are mediated by facultative and obligatory anaerobes [ 57 ]. The first step involves hydrolysis—the biological decomposition of macromolecular organic matter into simpler compounds more readily taken up by anaerobic microbes [ 58 ]. During acidogenesis, the pre-hydrolyzed compounds are further degraded into organic acids, which are, in turn, converted into acetic acid during acetogenesis. This acid serves as the primary food source for anaerobic, methanogenic bacteria [ 59 ], which conduct the last step—methanogenesis—and produce gaseous metabolites [ 60 ]. A step-by-step conversion flowchart for the anaerobic digestion (AD) of wastewater is given in Figure 3 61 ].
The literature data indicates that the demand for energy in anaerobic wastewater treatment systems is within a very broad range, from 0.20 kWh/mto 5.7 kWh/mof wastewater [ 172 ]. Such a significant variability of energy consumption is due to the construction and technological diversity of anaerobic reactors; climatic conditions; the temperature of the methane fermentation process; and the characteristics, quantity, and properties of the wastewater [ 173 ]. The analyses conducted so far have also demonstrated significant differences in the unit energy demand for the removal of 1.0 kg of COD from wastewater, with values ranging from 0.27 kWh/kg COD to 10 kWh/kg COD [ 21 ]. Apart from the factors indicated above, these differences are affected by the susceptibility of wastewater to biodegradation under anaerobic conditions and by the technological parameters of the treatment process, including mainly HRT and OLR. Ample studies have proven a correlation between the initial concentration of organic compounds in wastewater, its susceptibility to anaerobic degradation, as well as the applied organic load rate and retention time in the digester [ 174 175 ]. In the case of anaerobic wastewater treatment, it is possible to partially or fully cover the energy demand by recovering it from the methane produced. The literature data indicate that the potential energy contained in biogas may range from 0.3 kWh/mto 40 kWh/mof treated wastewater [ 176 ], whereas the energy contained in methane reaches 9.17 kWh/m, and the maximum theoretical unit amount of this biogas component is 0.35 m/kg of the COD removed. Thus, removing 1.0 kg of COD can yield approximately 3.2 kWh of potential energy [ 177 ]. As in the case of energy demand, the efficiency of methane production is affected by a huge number of variables. Therefore, there is a need to optimize the anaerobic process of contaminant biodegradation [ 178 180 ]. It has been proven that in the case of highly concentrated biodegradable wastewater, the production of methane not only covers the demand for energy, but also induces a positive balance of heat and electricity. In terms of potential energy, it can yield from 5 to 20 kWh/mof treated wastewater for export [ 181 ]. There are also known examples of energetically passive or positive anaerobic technological systems [ 182 ]. It should be emphasized that the energy savings achieved during anaerobic wastewater treatment result mainly from the lack of a need to aerate the wastewater and, in certain cases, from the lack of a need to mix the digester’s content [ 183 ]. Profits are primarily related to the production of methane and its conversion into heat and electricity [ 184 ]. Undoubtedly, a thorough economic balance should also take into account the costs associated with the construction and design of digesters and the fact that they are smaller in volume and require less space in terms of the volume of sewage and the load of organic compounds [ 185 ]. It is also essential to take account of the environmental costs related to the reduction of greenhouse gas emissions and the production of renewable energy and biofertilizers. This has been confirmed by LCA and LCC analyses [ 186 187 ].
The higher tolerance of anaerobic biomass to worsening environmental conditions is attributable to differences in the population structure [ 168 ]. Boonapatcharoen et al. (2007) [ 169 ] used inoculum sludge sourced from a low-load anaerobic reactor, which was mainly colonized by-like cells. The authors did not find this surprising, asspp. thrive in low-acetate conditions, unlike other acetoclastic methanogens, e.g.,spp. Stabilized mesophilic reactors of wastewater treatment plants have low levels of acetate, creating prime conditions formicrobes to grow in the inoculum. Slow-growingspecies have a higher affinity for acetate and lower uptake threshold (5–15 μM); thus, high levels of acetate can inhibit their growth., which are faster to grow, have a minimum acetate uptake threshold of 1–2 mM. This is corroborated by Calli et al. (2010) [ 170 ], who operated mesophilic UASB reactors inoculated with-heavy sludge. Thewere replaced byin series with higher pollutant loadings. The resistance ofspecies to environmental changes may stem from their chondroitin structures (pseudo-chondroitin, methano-chondroitin). In very specific growth conditions, chondroitin forms a thick and rigid outer layer [ 171 ]. It can, therefore, be surmised that the constant exposure of anaerobic biomass to a microwave electromagnetic field creates optimal conditions for the growth ofstrains with chondroitin walls.
When the microbial community in the anaerobic sludge is highly diverse taxonomically, multiple species with the same or similar metabolic profile can co-exist [ 82 ]. This leads to a more stable and productive operation of the digester [ 164 ]. If an uncontrolled change in the medium conditions leads to the loss of some species, others will be able to sustain and continue the given reaction in the methane production cascade [ 165 ]. The dominance of a small number of species indicates that anaerobic bacteria in the reactor, especially the selective and sensitive methane-generating, have been exposed to strong environmental selection pressures (e.g., high temperature or high organic loadings) [ 166 ]. This results in the intensive growth of a few select, most adapted species [ 167 ].
One interesting effect of shortened HRT and elevated OLR, as reported in the literature, is the sharp decrease in the count and length of filamentous microbes [ 159 160 ]. This is attributed to flushing or disaggregation of these bacteria during shock loads [ 161 ]. There have been investigations into the morphology of anaerobic biofilm as a function of organic load, VFA levels, and biogas production in fluidized bed reactors. Filamentous bacteria have been shown to be more populous at lower OLR, whereas high OLR and VFA led to elevated counts of coliforms and cocci [ 162 163 ]. The effect of OLR and HRT on anaerobic wastewater treatment performance is presented in Table 6
Another typical outcome of stress brought on by excessive/unstable HRT and OLR is a drastic increase in VFAs, which directly affects the biogas composition and production rate [ 151 ]. Researchers agree that excessive OLR tends to cause VFA buildup in the medium and has an inhibitory or even toxic effect on methanogenic bacteria [ 138 153 ]. However, some authors downplay the role of high VFA levels, and instead point to skewed acid/base balance as the major cause of reactor destabilization [ 154 155 ]. This is supported by studies that did not observe any inhibition and toxicity of VFA after OLR increases and VFA buildup under careful pH control [ 156 157 ]. Under these conditions, it is common to observe significant accumulation of propionate and extensive hydrogen saturation in the medium, which facilitates the alternative electron removal pathway [ 158 ].
It has been noted across multiple studies that the accumulation of volatile fatty acids (VFAs) is a common outcome of overloading and of sudden, uncontrolled changes in hydraulic and organic loads [ 146 ]. Also considered important are variations in hydrogen partial pressure, which play an important role in maintaining the target proportions between different anaerobic reaction products [ 147 ]. The stress of fluctuating OLR and HRT may cause a shift in metabolism to a less favorable profile, upsetting the population balance between VFA producers (acids and acetogens) and consumers (methanogens, sulfate- and nitrogen-reducing bacteria). This can lead to excessive carbon dioxide and hydrogen generation. High hydrogen partial pressures (above 104 atm) can halt methanogenesis and permanently alter metabolic pathways [ 148 ]. Once this happens, sensitive and slow-growing methanogens are unable to quickly and fully eliminate H, generated by acidogenic bacteria [ 149 ]. This has been shown to inhibit the degradation of propionate, malonate, and lactate [ 150 ].
One important aspect of anaerobic wastewater treatment is maintaining the delicate balance between the general stages of the process: hydrolysis, acidogenesis, and methanogenesis [ 143 ]. Changes in flow rates and wastewater composition have been shown to be critical in affecting anaerobic wastewater treatment performance [ 144 ]. This has a direct influence on the organic load rate (OLR) for the digester and the sludge, the hydraulic retention time (HRT), and the biomass retention time (SRT) in the reactor. Research is still in progress on how these parameters relate to and impact the maintenance of wastewater AD efficiency; it is still not fully known how anaerobic reactors perform under variable conditions [ 145 ].
Depending on the reactor design, stirring may be provided by agitators (reactors with full stirring, e.g., CSTRs) or by the hydraulic pumping of wastewater (e.g., UASB reactors), or wastewater with biogas (e.g., fluidized bed reactors). The stirring can be continuous or intermittent [ 142 ].
The anaerobic digestion of waste requires effective reactor stirring or flow control to eliminate dead zones across the digester’s active volume [ 134 ]. Stirring increases the contact surface between the wastewater and the bacterial biomass, while also agitating the substrate and biochemical products [ 135 ]. It also serves to reduce temperature gradients across the reactor [ 136 ], and prevents fouling, encrustation, and dead-zone formation [ 137 ]. The method and intensity of mixing anaerobically treated wastewater must also be chosen with the specific characteristics of anaerobic sludge in mind. Anaerobic bacteria are small and do not form dense conglomerates that can be separated from treated wastewater by simple sedimentation [ 138 ]. Issues with anaerobic sludge also extend to thickening and flocculation [ 139 ]. Design and process solutions aim to reduce the removal of biomass from the reactors (which can translate directly to reduced performance) [ 140 ]. This is achieved by extending solid retention time (SRT) (extending sludge age) and reducing hydraulic retention time (HRT). This is particularly important for methanogens due to their slow generation rate [ 141 ].
The effect of pH on the dissociation of certain metabolites and their toxicity is also determined by temperature. The dissociation degree will fall at higher temperatures, meaning the metabolites can become more toxic. However, this is only the case for stepwise changes. If the temperature is raised gradually, the microbes will progressively adapt to the new equilibrium [ 128 ]. The effect of pH on anaerobic wastewater treatment performance is presented in Table 5
The pH determines synthesis and toxicity of intermediate and final metabolites during AD (VFAs, ammonia, hydrogen sulfide). It has been noted that ionized volatile fatty acids are toxic to methanogens at concentrations above 2000 mg CHCOOH/dm, whereas the non-ionized forms are harmful at levels as low as 10 mg CHCOOH/dm, or even lower [ 125 ]. For an optimal acid/base balance, it is generally accepted that total VFAs should not exceed 500 mg CHCOOH/dm, and the VFA-to-alkalinity ratio (expressed in mg CaCO/dm) should be around 0.3 [ 126 ]. The pH of the solution can also affect the toxicity of ammoniacal nitrogen. The dissolved free ammonia and ammonium ions are at an equilibrium point. Any change in pH results in an equilibrium shift with the non-ionized ammonia being toxic to microbes at concentrations as little as 50–150 mg NH– N/dm. Conversely, dissociated ammonia is far less toxic (1500–3000 mg NH/dm). Methanogens can acquire an adaptive tolerance to ammoniacal nitrogen at levels as high as 5000 mg NH/dm. This also holds true for sulfur, as its non-dissociated form (HS) is approximately twice as toxic as the dissociated S 127 ].
Potentially, methanogenesis can also proceed at low pH levels. Peat environments are host toandable to produce methane at a pH of around 4.5 [ 122 ]. Acidophilic methane-producingmay also be present in anaerobic sludge in closed digesters. After a long, gradual adaptation process, methane can be produced at 4.5 or 6.5 pH [ 123 ]. However, it should be noted that acidophilic methanogenesis has only been observed in the laboratory, with no reports to date on its effective incorporation into full-scale installations [ 124 ].
A multi-step process design can be used to avoid the constraints related to the different pH requirements of microbes across the various stages of waste degradation [ 119 ]. To that end, separate hydrolyzers are operated at high organic load rates (OLRs), short hydraulic retention times (HRTs), and acidic pH (4.5–6.3). The hydrolyzer discharge is fed into a large digester with a low OLR conductive to high methanogen activity, in which the pH is maintained at around 7 [ 120 ]. This design is typically used in agricultural biogas plants [ 121 ].
The pH level plays a major role in AD. As the pH changes, so does the solubility and form of chemical compounds [ 111 ]. The bacteria responsible for different steps of fermentation have different pH preferences for optimal growth [ 112 ]. For hydrolyzing and acidogenic bacteria, the optimal pH ranges from 4.5 to 6.3 [ 113 ], whereas methane-producingprefer neutral conditions between 6.8 and 7.5 pH [ 114 ]. Single-step reactors generally require a neutral pH for nominal operation. The pH level itself is regulated by acid or alkaline metabolites produced in the course of anaerobic degradation [ 115 ]. The pH is therefore the net result of the feedstock pH and the acid-/base-forming activity of microbes [ 116 ]. Reduced pH is an indicator of methanogenesis disruption and acidic metabolite accumulation [ 117 ]. As such, digesters are fed with alkaline substances (e.g., sodium hydroxide) to ensure high methane production. This is particularly crucial for high-load anaerobic wastewater treatment reactors. At loads approximating 20 kg COD/m/d, continuous pH adjustment is usually required to maintain methane production [ 118 ].
In practice, mesophilic processes are preferred for systems designed to ensure the high efficiency of anaerobic organic matter degradation, though thermophilic conditions are used as well in special cases [ 95 ]. This is primarily to offset the power expended for digester heating with improved digestion performance. Thermophilic AD offers faster rates, more complete organic matter degradation, and better elimination of pathogens [ 96 ]. However, when compared with mesophilic systems, these benefits do not compensate for the considerably higher energy input [ 97 ]. There are also certain drawbacks to thermophilic digestion, such as diminished stability and higher process sensitivity [ 98 ]. Thermophilic conditions lead to higher levels of volatile fatty acids compared to mesophilic processes. This holds particularly true for propionic acid, since acetogenesis is inhibited above 55 °C [ 99 ]. Anaerobic wastewater treatment performance as a function of temperature is presented in Table 4
There have been numerous reports of psychrophilic and psychrotolerant methane-producingin cold natural environments [ 89 90 ], though their metabolic activity and rate are relatively low [ 91 ]. Research on adapting these bacteria also does not bode well for their practical applicability in contaminant degradation [ 92 ]. Temperature can change the methanogenesis rate by affecting microbial activity and counts. Changes have also been observed in the efficiency of available methanogenic pathways. At low temperatures, most of the methane is produced from reduced acetates [ 93 ]. Importantly, methanogenesis, which is the bottleneck for the entire anaerobic conversion chain, is also the step most sensitive to sudden changes and spikes in temperature [ 94 ].
All biochemical processes of AD are subject to the laws of thermodynamics [ 86 ]. On the one hand, elevated temperatures will accelerate the rate of biochemical reactions catalyzed by enzymes (in accordance with the Arrhenius theory), and on the other, the constituent proteins of enzymes can denature if the temperature is too high, drawing the process to a halt [ 87 ]. Fermenting microbes can be divided into three groups according to their preferred temperature range: psychrophilic (under 25 °C), mesophilic (30–40 °C), and thermophilic (50–60 °C). Numerous studies have examined the relationship between AD performance and process temperature. According to Dhaked et al. (2010), the organic methanogenesis cascade has been observed in nature within the very broad temperature range of 0 to 98 °C. Out of the entire complex of microorganisms involved in the AD process, methanogens are the most sensitive to temperature changes, especially to excessive or severely unstable temperatures [ 88 ].
One of the frequently indicated weaknesses of anaerobic technologies is the low N and P removal rate [ 206 ]. In anaerobic processes, nutrients are only removed when incorporated into the microbial biomass [ 207 ]. This issue limits the versatility of the technology, which is why further advancement of anaerobic methods focuses on eliminating this hurdle [ 208 ]. For this reason, anaerobic technologies are very often treated as the first step in a larger processing chain [ 209 ]. In order to achieve final effluent quality sufficient to discharge to a receiving water body, digesters are usually integrated and coupled with other facilities and equipment [ 210 ]. Elaborate processing systems must control many antagonistic operation parameters, such as oxygen concentration, VFA levels, loadings, and concentrations of digestion microbes [ 211 ]. A typical treatment system alternates between anaerobic and aerobic conditions to provide efficient organics removal, orthophosphate fixation, ammonification, nitrification, and denitrification. In many cases, the process needs to be modified or supplemented with other components to improve treatment performance [ 212 ]. Therefore, there is real need to seek versatile methods that could serve as a competitive alternative to current solutions, both in terms of technological and investment attractiveness [ 213 ]. Various innovative solutions and methods can be used to overcome this deficiency of anaerobic reactors and improve their technological and commercial performance. These include active filling [ 214 ], dissolution of metals [ 215 ], zeolites [ 216 ], adsorption [ 217 ], absorption [ 218 ], inorganic coagulants [ 219 ], alkaline substances [ 220 ], integration with microalgae cultivation systems [ 221 222 ], and many other physical and chemical treatments [ 223 ]. A breakdown of the strengths and weaknesses most commonly cited in the literature is given in Table 7
Such close interdependence of the assorted digestion steps often leads to underperformance when products necessary to smoothly conduct the successive stages are produced in too large or small quantities; for example, the overaccumulation of acidogenesis products reduces pH and inhibits the activity of methanogenic bacteria [ 204 ]. The efficiency of the entire process is ultimately determined by the most sensitive link in the conversion chain, which is why methanogenic activity is used as the indicator for reactor efficiency [ 205 ].
Of course, the anaerobic processes of wastewater treatment have their limitations and shortcomings [ 200 ]. Therefore, they need to be continuously improved and supplemented with novel technologies for better process stability and performance with regard to methane yields and final effluent quality [ 201 ]. Some of the limitations of anaerobic reactors are inherent to biological methods. AD entails a cascade of various biochemical conversions to remove pollutants and synthesize biogas. The process is based on syntrophy (cross-feeding), where the metabolites from one part of the digestion process provide essential nutrients for the organisms responsible for the subsequent step [ 202 ]. This is because the microorganisms responsible for the successive digestion steps are highly specialized: from hydrolysis, which can be carried out by various bacteria and, to methanogenesis, which can only be conducted by specializedunder very specific conditions [ 203 ].
Anaerobic wastewater treatment reactors also run at much higher biomass and pollutant loads [ 196 ]. This allows them to remove the same organic loads as aerobic ones at a much smaller reactor size. The performance of anaerobic wastewater treatment is a function of several factors, including reactor design, process temperature, and wastewater type. For example, the COD removal from textile industry wastewater can range from 9% to 51% in a UASB reactor [ 197 ], whereas the range for starch-containing wastewater is 77–93% COD removed [ 198 ]. A well-designed, well-built, and well-operated anaerobic wastewater treatment reactor does not disperse aerosols into the air, nor does it produce odors. A vital advantage, especially for treating wastewater from seasonal/intermittent production, is that the reactor can be quickly started, even after extended downtime. Anaerobic sludge can be stored for months without nutrient media and still largely retains its biochemical properties [ 199 ].
The major advantage of anaerobic over aerobic wastewater treatment is the lower generation of surplus sludge. Anaerobic treatment generates 0.02 to 0.15 g/g biomass per 1 g of COD removed, whereas aerobic systems can produce approx. 0.5–0.7 g/g biomass for the same pollutant load. Seghezzo et al. (1998) [ 195 ] determined that anaerobic systems produce three to twenty times less surplus sludge than aerobic systems.
The anaerobic process has found wide use in treating highly concentrated wastewater, especially that from the food industry [ 188 ]. It has been noted that anaerobic methods are well suited for wastewater with COD levels (which primarily measure readily available organics) of more than 4000 mgO/dm. In practice, the anaerobic technology is preferred mainly for its cost efficiency, as it is considerably cheaper as a means of anaerobic wastewater treatment than traditional aerobic processes. The reduced cost of anaerobic treatment stems mainly from the obvious lack of the need for wastewater aeration. It is estimated that this can produce energy savings of up to 75% [ 183 ]. The sole energy input required for anaerobic technologies is to power wastewater stirring/pumping mechanisms and temperature control systems [ 189 ]. According to the operators of anaerobic wastewater treatment systems, heating accounts for approximately two-thirds of the energy consumption of an anaerobic plant [ 190 ]. This portion of energy intake can be reduced by using reactors that support the capture and combustion of biogas [ 191 ]. The methane produced after the full conversion cycle can be used for fuel, e.g., for heat generation or for co-generation of heat with electricity. In fact, 1 mof methane has a calorific value of 9.17 kWh, which means that biogas containing 60–70% of methane (the average range) would have a calorific value of 6–7 kWh [ 192 ]. It is estimated that 30% of the heat produced by a co-generation system will be used to heat the reactor, whereas approximately 10–15% of the production will be allocated for on-site power generation [ 193 ]. Biogas production from the AD of food industry wastewater usually falls between 0.2 and 0.5 m/kg COD removed. The biogas yield depends on a number of factors, the most important of which include the composition and quality of the wastewater, as well as the type and design of the reactor [ 194 ].
Anaerobic membrane bioreactors (AnMBRs) have been used as an alternative to traditional anaerobic digestion ( Figure 9 ) [ 286 ]. Research and operational data show that AnMBRs offer reduced sludge production and better biogas yields. It is believed that incorporating membranes into an anaerobic bioreactor design reduces the total energy input and facilitates microbial community retention for high-biomass, high-OLR processing [ 287 ]. The problem with this reactor type is membrane fouling, caused by foreign materials in wastewater. Major membrane fouling factors include high concentrations of suspended solids, protein, lipids, oils, and grease. Another limitation of AnMBRs is the high level of extracellular polymeric substances (EPS) generated by anaerobic sludge (especially granular sludge) microbes [ 288 ]. Depending on the target parameters of the treated effluent, potential treatments include microfiltration, ultrafiltration, nanofiltration, and reverse osmosis [ 289 ]. Buntner et al. (2013) [ 258 ] tested a combination of a UASB plus ultrafiltration membrane reactor on dairy wastewater. The biogas production was 150 dm/kg COD with a methane fraction of 73%, the OLR was 4.85 kg COD/m/d, whereas the COD removal reached 95–99%. Slightly worse performance was reported by Deowan et al. (2013) [ 290 ] at 90% organic removal. Both were pilot-scale installations, but the former implementation used an external membrane, whereas the latter was immersed in the bioreactor.
Hybrid reactors are designs that merge aspects of other, ordinarily discrete types of anaerobic reactors ( Figure 8 ). The designers of such solutions typically set out to harness the strengths of the given reactor type while minimizing or eliminating its limitations [ 255 ]. The most common hybrid process combines anaerobic filters with UASB reactors. This enables large organic loadings while increasing biomass retention in the system. Such hybrid reactors can also be quickly restarted after downtime [ 256 ]. Ramasamy et al. (2004) [ 257 ] used this design to degrade synthetic effluent with a COD of 10 g/dm. The organic load rate ranged from 0.82 to 6.11 kg COD/m/d and the hydraulic retention time from 4.1 to 1.7 d. Under these process parameters and mesophilic conditions, the COD removal efficiencies varied between 90% and 97%. Another type of hybrid design calls for coupling a CSTR with a membrane module for efficient anaerobic sludge separation [ 285 ], which dramatically changes its capabilities and applications.
In anaerobic fluidized-bed reactors ( Figure 7 ), the microbial biomass grows on the surface of fine media particles [ 284 ]. The rapid upward flow generated by the extensive recirculation keeps the biomass suspended. The particles that make up the fluidized bed can be made of quartz, pumice granules, or activated carbon (the particles are usually 0.2 to 3.0 mm in diameter) [ 233 ]. This structure allows fluidized-bed reactors to maintain a very high concentration of biomass of over 40 kg/m(even 100 kg/min some cases), enabling very high OLRs. The values reported in the literature range from 1 to 100 kg COD/m/d, at HRTs of 0.5 to 7 d [ 232 ]. Unlike typical biofilters, fluidized-bed reactors usually do not suffer from blockage. However, they do require relatively high energy input for the extensive recirculation of effluent. This, in addition to the operational difficulty of fluidization (keeping the fluid suspended), is the primary flaw of this design [ 234 ].
A characteristic feature of UASB reactors, much like the structurally similar EGSB reactors, is that they use anaerobic biomass in the form of granular sludge [ 279 ]. The granules range from 0.14 to 5.00 mm in size [ 242 ]. The granulation of sludge makes it easier to increase its levels in the digester, allowing installations to be operated at OLRs as high as 10.0–40.0 kg COD/m/d [ 241 ]. The granulated microbial biomass in a UASB is kept suspended by the upward-flowing wastewater and biogas [ 247 ]. The top of the device is fitted with a three-phase separator, which separates the treated effluent from the sludge and biogas bubbles [ 280 ]. UASB reactors are highly versatile, making them a popular choice for treating a variety of industrial effluents [ 281 ]. On the other hand, they are beset by the relatively long start-up times—3 to 5 months—needed to granulate the anaerobic sludge and pre-treat the effluent if it contains large amounts of lipids and suspended solids [ 282 ]. In practice, this means that UASB usually have to be accompanied by a flotation tank, further complicating their exploitation. A diagram of a UASB reactor is presented in Figure 6 . UASB reactors have been successfully used to treat dairy effluent for over 50 years [ 283 ]. OLRs and HRTs vary greatly across the literature. For example, Najafpour et al. (2008) [ 243 ] tested OLRs ranging from 7.9 to 45.42 kg COD/m/d, achieving 98% organic compounds removal from whey at 35 °C.
The performance of full-scale anaerobic beds in treating dairy wastewater was tested by Omil et al. (2003) [ 253 ]. After almost two years of measurements, the authors found a reduction of about 90% in organic matter (COD) at loads of 5–6 kg COD/m/d and at an industrial scale. The HRT was changed throughout the experiment, from 20 d at the onset to 3.33 d at the end, without any significant changes in removal performance, indicating that the reactor achieved a steady state [ 253 ]. An extremely high COD removal efficiency (98%) was achieved by Mendez et al. (1989) [ 254 ] when using an anaerobic filter for neutralizing whey, though at the expense of very long HRT (142 d). This very high removal efficiency was achieved under mesophilic conditions, despite the relatively high OLR of 9.8 kg COD/m/d [ 254 ].
In anaerobic filters, the problem of biomass washout is addressed by capturing the biomass on top of a bed, forming a biofilm ( Figure 5 ). Filtration media usually consist of ceramic and plastic structures [ 275 ]. Packed-bed anaerobic reactors have a very long start-up time, as the biofilm forms very slowly under anaerobic conditions [ 276 ]. Wastewater can be passed through anaerobic filters in an upflow or downflow configuration. According to Rajeshwari et al. (2000) [ 232 ], the organic load rate in the reactor can be as high as 40 kg COD/m/d. Filtration reactors have multiple advantages: they are simple in design, easy to operate, do not require stirring (thus enable saving energy), run stably even at high OLRs, have a high ability to withstand toxic loadings, and can accommodate fluctuations in the pollutant concentration [ 277 ]. However, one problematic aspect of such designs is the relatively large size of the digesters, since the packing takes up a huge portion of their volume. Filtration-pore blockage can also be an issue, especially at high levels of lipids and/or suspended solids [ 278 ].
In contactors, the flocculated anaerobic sludge is suspended in the wastewater. At their most basic, anaerobic contactors consist of a reaction vessel (digester) colonized by anaerobic microflora, together with a stirring system and a settling tank [ 272 ]. The sludge is recirculated from the settlement tank to the reaction chamber in order to age it and maintain the target concentration ( Figure 4 ). The advantage of this design is that it can treat wastewater with high levels of suspensions and solids, e.g., proteins and lipids [ 273 ]. Therefore, contactors are cited as a potential option for treating meat-industry wastewater, which other designs struggle with [ 24 ]. Compared to other reactor types, contactors are cheap to build and easy to operate [ 274 ]. Their main drawback lies in the processing limitations, particularly those relating to floating sludge in the settling tank, which can be washed out in the treated effluent [ 229 ]. Due to the low concentrations of sludge in the reaction vessel, these reactors can only handle smaller organic loads of 0.25–4.00 kg COD/m/d [ 230 ]. CSTRs can be considered the progenitor of contactors. In this type of reactor, the biomass retention time is the same as the hydraulic retention time. They are rarely used for wastewater treatment, but are a common feature of agricultural biogas plants [ 231 ].
The basic input data for the design of anaerobic bioreactors include the computational flow of wastewater and the physical and chemical characteristics of raw wastewater, mainly in terms of the content of organic compounds (COD, BOD, TOC), the content and forms of nutrients (P, N), and the presence of suspended lipid and protein fractions, sparingly degradable substances, or substances toxic to anaerobic bacteria [ 271 ].
Digesters and biogas tanks should be designed and constructed in a way that protects against fire or explosion, freezing of gas supply and discharge pipes, gas condensation, and corrosion caused by substances contained in the gas, including in particular ammonia and hydrogen sulfide [ 267 ]. Shell tanks made of flexible materials should be secured with a fence at least 1.8 m high. Metal and reinforced concrete biogas tanks may not be enclosed. Underground biogas tanks can only be loaded with the soil above them [ 268 ]. Digesters should be made of fire-proof materials [ 269 ]. The thermal insulation of digesters and biogas lines should be made in a way that prevents fire from spreading. In addition, digesters and biogas tanks should be equipped with lightning protection systems and protected against static electricity [ 270 ]. Electrical installations should be carried out in pipes with a degree of protection not lower than IP-54 or IP-68 if diving mixers are used. Flexible plastics intended for the construction of biogas tanks should have a tearing strength of at least 15 N/cm, a methane permeability not higher than 10 cm/mh bar, resistance to temperatures ranging from −30 °C to +50 °C, as well as the ability to discharge electrostatic charges.
Cubature constructions used in anaerobic wastewater treatment systems must be resistant to aggressive environments; their wall surfaces must not corrode in contact with wastewater fed to the installation [ 261 ]. An important feature of anaerobic digesters is their sufficient mechanical resistance. Their construction should ensure the safety of their operation in terms of resistance to pressures related to the retention and transport of wastewater, as well as resistance to loads related to the assembly of the necessary equipment and accompanying elements [ 262 ]. Gas-tightness is a prerequisite for anaerobic installations where the biogas produced is stored, transported, and ultimately used for energetic purposes [ 263 ]. Leaks reduce the energy and economic efficiency of the technological system, pose a threat to the occupational health and safety of personnel, and exert an adverse effect on the environment through the emission of CH, CO, and HS into the atmosphere [ 264 ]. Heat losses should be minimized as well, which is particularly important in facilities operated in cold climate zones [ 265 ]. The methane fermentation process is usually carried out under mesophilic (30–40 °C) or thermophilic (50–60 °C) conditions. Therefore, it is necessary to ensure the effective thermal insulation of the digester to reduce convective heat energy losses [ 266 ].
Anaerobic wastewater treatment reactors are designed to minimize the inefficiencies resulting from slow-growing biomass, which is a problem intrinsic to anaerobic digestion [ 247 ]. The main challenge is keeping the SRTs high and the HRTs low [ 248 ]. There are many different designs of anaerobic wastewater treatment reactors currently in use. Anaerobic-only designs include UASB (upflow anaerobic sludge blanket) and EGSB (expanded granular sludge bed) with high-density granulated anaerobic biomass [ 249 ]. In reactors with full stirring (CSTR, continuous-flow stirred tank reactor, contactors), the biomass takes the form of suspended sludge and is recirculated into the system by auxiliary installations (settlement tanks, separators) [ 250 ]. In bed reactors, the anaerobic microbial community forms a biofilm that adheres to the bed. However, in the case of anaerobic beds, a large portion of the biomass may be suspended above the bed and not firmly attached [ 251 ]. The listed reactor types do not represent the totality of the available anaerobic wastewater treatment solutions. New designs and improvements to existing ones continue to be introduced. For example, a traditional CSTR can be expanded with a membrane module [ 252 ], which dramatically changes its capabilities and applications. Wastewater treatment parameters across different types of digestion bioreactors are given in Table 8
Biohythane production from wastewater via AD is currently based on two-step, physically separated processes that ultimately produce biohydrogen and biomethane [ 293 ]. Such set-ups usually integrate two separate anaerobic reactors, which requires advanced monitoring and costly, complex operation [ 293 ]. A multistage anaerobic hythane reactor (MAHR) is an innovative solution that integrates the two steps of the digestion in a single digester [ 293 ]. An experimental MAHR has been constructed, consisting of an internal downflow reactor with an anaerobic filter and an external upflow sludge blanket to boost microbial enrichment and thermodynamic efficiency of the associated bioreactions. The MAHR was run for 160 d, with performance assessed on the basis of biogas production, metabolic flux, and microbial population composition against a typical anaerobic high-rate UASB. Biohythane production with an optimized hydrogen volume ratio (10–20%) and high methane content (75–80%) was achieved in the hythane zone (MH) and methane zone (MM) of the MAHR, respectively. In addition, the MAHR showed better adaptability at high organic loadings (120 g COD/dm/d) and provided better organic compounds conversion, with a methane production rate 66% higher than the UASB. A thermodynamic analysis indicated that the hydrogen extraction in MH significantly decreased the hydrogen partial pressure (<0.1% vol.), which drove acetogenesis in MM. An analysis of the metabolic flux and microbial function further demonstrated that the MAHR showed superior performance to the UASB, which was primarily attributed to improved acetogenesis and acetoclastic methanogenesis [ 293 ].
The ABFSB has been tested on a laboratory scale. The experimental reactor was constructed of PMMA (poly(methyl methacrylate)) to a total working volume of 4.77 L. The medium had a total weight of 32 g, which translated to an area of 43.8 m/g (98% porosity). The feed pump output was 0.4 dm/h (6.6 cm/min) so that the hydraulic retention time would be 12 h. The process temperature was 30 ± 1 °C [ 292 ]. The reactor was inoculated with granular sludge sourced from a UASB reactor for treating poultry slaughterhouse wastewater. The matrix inoculation procedure was carried out by grinding the granules and immersing the medium in the biomass for 2 h at room temperature. Synthetic sewage was used for the experiment [ 292 ]. The ABFSB was operated at four different COD/SOratios, namely, 6.1, 3.5, 1.7, and 0.72. The best COD and sulfate removal performance (82 and 89%, respectively) was achieved at a COD/[SO] of 1.7. The variants also boasted a higher overall kinetic apparent parameter (0.96/h). The kinetic parameters indicate that sulfate-reducing bacteria played a major part in removing organic matter over a COD/[SO] range of 6.1 to 1.7 [ 292 ].
In this design, the digestion process is coupled with phase separation in an upflow anaerobic bioreactor with fixed-structured bed (ASTBR) as the methanogenic phase [ 236 ]. The technology was tested on sugarcane vinasse, a type of effluent from ethanol production. The organic loadings were within a range of 15–30 kg COD/m/d [ 236 ]. The ASTBR was successful in removing over 80% of the COD from the vinasse, and demonstrated stable long-term operation (240 days), even for OLR values as high as 30 kg COD/m/d. A UASB performed much worse under similar process conditions, suffering from acid accumulation and impaired performance [ 236 ].
A labyrinth-flow vertical anaerobic reactor (LFAR) consists of three parts, each having a different function: hydrolysis–acidogenesis, methanogenesis, and sedimentation–clarification ( Figure 13 ) [ 291 ]. The reactor is equipped with a two-pump system: the first pump delivers raw wastewater to the reactor; the second, recirculating pump provides complete mixing and circulation of sludge and wastewater in the hydrolysis tank. The flow is directed downward in the methanogenesis tank, and upward in the clarifier. A diagram of the LFAR is presented in Figure 13 . The hydrolysis tank is packed with magneto-active filling (MAF) comprising PE-HD beads. The beads contain neodymium magnets and are coated with Cu-Fe powder. The metals (90 Fe/10 Cu) account for 5.0% of the weight of a single bead [ 291 ]. The LFAR-MAF has been operated at the following parameters: temperature = 35 °C; HRT = 2 d; and organic load (in COD) = 6.0 g COD/dm/d. The best AD performance was noted for 60% and 80% MAF in the filling [ 291 ]. The MAF had a significant impact on methane fractions when the filling contained 60% to 100% magnets, with over 70% CH 291 ]. The experiments showed that the MAF had a positive effect on COD removal when the beads contained 20%, 40%, and 60% FM. The organic compounds removal in these series reached 75–76%, but subsequent increases in the number of SMF beads did not improve COD removal any further [ 291 ]. MAF was also shown to ensure low levels of HS in the biogas.
This digester is partitioned with perforated plates into four sections, each for one of the processing steps (hydrolysis, acidogenesis, acetogenesis, and methanogenesis) ( Figure 12 ). Stirring in the chamber is provided by a shaft rotating at 1 rpm and fitted with stirring rods. The reactor is heated using microwave generators with magnetrons. Each section is equipped with a set of four 20 kHz, 95 W ultrasonic transducers [ 21 ]. In the upper part of the HFAR, each partition is equipped with ports for sludge and biogas collection, as well as monitoring temperature and pH. The COD and TOC removal in the HFAR has been measured at approximately 85% at OLR of 2.0 g COD/dm/d. The biogas yield was 0.33 ± 0.03 dm/g COD removed, with a methane content of 68.1 ± 5.8%. Peak performance was achieved with ultrasonic generators working in a 2 min on/28 min off scheme. The nutrient removal was low across all series of the experiment, ranging from 2.04 ± 0.38 to 4.59 ± 0.68% for phosphorus and from 9.67 ± 3.36 to 20.36 ± 0.32% for nitrogen [ 21 ].
The general FAF-R process design consists of a retention (hydrolysis) tank equipped with a vertical agitator. A fluidized active filling (FAF) is introduced into the exact digestion reactor at a ratio of 1/5 of the digester’s active volume of the chamber [ 214 ]. Much like the hydrolysis tank, the reactor is fitted with a vertical agitator rotating at 60 rpm and working on a 30 min on/30 min off operating scheme. Due to the density of the FAF, the tank acts as a contactor with fluidized filling during the mixing-on periods. During the mixing-off period, the FAF floats to the top, where a filtration layer forms [ 214 ]. COD removal rates in the reactor have been reported at over 74%. The biogas yields ranged from 356 ± 25 to 427 ± 14 dm/kg COD removed and the biogas contained approximately 70% methane. Increasing the organic load to 8.0 kg COD/m/d led to a significant decrease in AD and effluent treatment efficiency, which correlated with decreased pH and a rise in FOS/TAC to 0.44 ± 0.2. Phosphorus removal rates with active filling ranged from 64.4 ± 2.4 to 81.2 ± 8.2%, depending on the stages. The concentration of total suspended solids in the treated effluent was low [ 214 ]. A diagram of the M-SHAR reactor is provided in Figure 11
The structure of the M-SHAR consists of two separate sections, each with their own function. The lower section, filled with granular sludge, is made up of two centrally arranged chambers (inner and middle) [ 32 ]. The design provides for separate hydrolysis/acidogenesis/methanogenesis zones and a meandering (labyrinthine) flow of the wastewater being treated. The top part of the M-SHAR houses a microwave-heated biological filter. Raw sewage is transported to the lower section, filled with granular sludge, and then separated from it in the course of treatment. Afterwards, the liquid is pumped onto the microwave-heated anaerobic filter, where the dissolved organic matter is further biodegraded. Studies have shown that microwave heating improves wastewater treatment performance. At OLRs of 15 to 25 kg COD/m/d, the microwave-heated bioreactor boasted 5% higher COD removal and 12–20% higher biogas production [ 32 ]. The M-SHAR reactor is illustrated in Figure 10
Anaerobic wastewater treatment methods can efficiently biodegrade organic matter and are still being actively developed. Such technologies are predominantly used in industries that generate high volumes of readily biodegradable waste. The popularity of digesters stems from their multiple major advantages over aerobic-activated sludge or biofilter processes. The most frequently cited benefits are the effective treatment of high-OLR wastewater, the ability to accommodate high OLRs, relatively small size and space requirements, markedly lower investment and operation costs compared with aerobic technologies, and the ability to halt and resume operation (which is a priority for intermittent/seasonal producers). Anaerobic systems generate less surplus sludge, provide better sludge stabilization, and lead to improved sanitary indicators, which directly contribute to easier neutralization and management. Another crucial aspect is their ability to produce methane-rich biogas that can be harnessed for energy purposes. The aforementioned benefits have been demonstrated not only in experimental research or pilot-scale units, but predominantly in large-scale plants. There is currently a wide variety of solutions available on the market, which greatly enhances the versatility of anaerobic wastewater treatment—an important consideration for deployment. The high technical readiness level (TRL) and thorough understanding of the technology makes it easier to select the appropriate process for the type and volume of wastewater, pollutant load, biodegradability parameters, or digestion inhibitors in the system.
The nature of AD reactors and their advantages over aerobic technologies make them an excellent choice from the standpoint of sustainable development, as well as for programs intended to improve waste management, protect the environment, reduce energy waste, and promote renewable energy. Since AD produces biogas and reduces carbon dioxide emissions, its deployment is also in line with the principles of a circular economy, energy and material recycling, bio-economy, and the European Union climate and energy package.
One of the limitations of anaerobic wastewater treatment is that the processing is biologically mediated, making it susceptible to external physical and chemical factors. Optimal and stable functioning of digestion reactors is a balancing act between environmental conditions and the rate/products of biochemical reactions. Achieving this balance requires precise adherence to key process parameters, including temperature, organic load rate, hydraulic retention time, anaerobic sludge levels in reactors, method and frequency of wastewater feeding, homogenization of waste in retention/equalization tanks, stirring method, and wastewater amendment with microelements and essential nutrients as necessary.
The low nitrogen and phosphorus removal rates of anaerobic wastewater treatment are considered its biggest drawback. These biogenic substances are mainly responsible for accelerating the eutrophication and degradation of natural water bodies. The fixation of nitrogen and phosphorus in digesters is driven solely by the growth of bacterial biomass, of which these substances are an important component. The N and P removal percentage is usually within single digits. The resultant anaerobically treated effluent cannot be discharged directly to a receiving water body, as it exceeds permissible nitrogen and phosphorus concentrations.
Anaerobic digestion has been thoroughly explored in terms of its biochemical pathways, taxonomy of fermentation microorganisms, optimal medium conditions, inhibitors of pollutant degradation, and final performance. The current understanding of the technology and its shortcomings enables the identification of areas of growth and improvement. Granular sludge reactors are currently a popular choice, as they largely prevent sludge washout and reduce the risk of effluent recontamination. There have also been bioreactors built and tested with structurally and functionally separated zones for specific steps of the AD cascade. Various packings have been explored to increase the active surface for anaerobic microbe growth, boost nutrient removal, and stimulate biochemical conversion. Finally, physical factors have been utilized as well, including microwave radiation, static magnetic field, electromagnetic field, and ultrasound. Such treatments ensure optimal medium conditions and promote biochemical reactions, as exemplified by the demonstrated non-thermal effects of electromagnetic microwave radiation.
Integrating AD with microalgal biomass production systems is an important next step in the development of anaerobic wastewater treatment. This is borne from the need to remove N and P, as well as to reduce the costs of microalgal culture via novel, cost-effective technologies. Numerous studies have shown that nitrogen and phosphorus in pre-treated wastewater can serve as valuable sources of nutrients. After all, pre-treated sewage has low concentrations of organic matter and contains readily available mineralized nitrogen and phosphorus. Incorporating such methods in microalgae cultivation systems becomes even more justified when considering the fact that anaerobically treated effluent cannot be discharged directly into the environment, and requires complex, costly post-processing (including nitrification, denitrification, and biological/chemical phosphorus removal).
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