Process Intensification of Anaerobic Digestion of Biowastes for Improved Biomethane Production: A Review

Abstract

Anaerobic digestion is a widely adopted technique for biologically converting organic biomass to biogas under oxygen-limited conditions. However, several factors, including the properties of biomass and its complex structure, make it challenging to degrade biomass effectively, thereby reducing the overall efficiency of anaerobic digestion. This review examines the recent advancements in commonly used pretreatment techniques, including physical, chemical, and biological methods, and their impact on the biodegradability of organic waste for anaerobic digestion. Furthermore, this review explores integrated approaches that utilize two or more pretreatments to achieve synergistic effects on biomass degradation. This article highlights various additives and their physicochemical characteristics, which play a vital role in stimulating direct interspecies electron transfer to enhance biomethanation reaction rates. Direct electron interspecies transfer is a crucial aspect that accelerates electron transfer among syntrophic microbial communities during anaerobic digestion, thereby enhancing biomethane formation. Finally, this article reviews potential approaches, identifies research gaps, and outlines future directions to strengthen and develop advanced pretreatment strategies and novel additives to improve anaerobic digestion processes for generating high-value biogas.

1. Introduction

Since the Industrial Revolution, most of the world’s energy has been derived from fossil fuels. CO₂ and other greenhouse gases released from the combustion of fossil fuels play a significant role in global warming and climate change. Key greenhouse gases, such as CO₂, CH₄, N₂O, fluorinated gases, and water vapor, which are produced by anthropogenic activities, harm the ozone layer and contribute to the greenhouse effect. The atmospheric methane levels have increased from 1712 ppb in 1990 to 1775 ppb in 2000, 1797 ppb in 2010, 1872 ppb in 2020, and 1934 ppb in 2025 [1].

Figure 1 illustrates the current trend in global methane emissions by various sectors. Agriculture contributes to 3.6 billion tons of worldwide methane emissions, followed by fugitive emissions (3 billion tons), and unmanaged waste disposal (1.5 billion tons) [2]. Agriculture is identified as the primary source of methane emissions globally. In the agricultural sector, livestock produce methane through their digestive processes, known as enteric fermentation [3]. Additionally, rice cultivation significantly contributes to global methane emissions [4]. Furthermore, fugitive emissions account for a significant share of methane. Fugitive emissions refer to unintended gas leaks that occur during fracking, conventional oil and gas extraction processes, and the transportation of these resources. Methane can also escape from landfills, composting sites, anaerobic digesters, and industrial activities as fugitive emissions. A significant factor in methane emissions is the accumulation of unmanaged and organic waste in landfills and dumping sites. The global warming potential of methane is estimated to range from 27 to 30 over a century and 84 to 87 over two decades [5].

A fundamental aspect of tackling global warming is shifting from fossil fuels to adopting biofuels and effectively managing waste to prevent the release of greenhouse gases into the atmosphere [6]. Importantly, biofuels stand out as a viable and eco-friendly substitute for traditional fossil fuels. The production of biofuels is crucial for addressing global energy challenges and expanding energy options. Biofuels provide several notable advantages. Firstly, they represent a renewable energy option that can diminish carbon footprints and dependence on conventional fossil fuels. Secondly, biofuel production can promote rural growth by generating economic opportunities and enhancing living conditions. Lastly, by reducing greenhouse gas levels in the atmosphere, biofuels can help mitigate the effects of climate change.

Organic waste materials from agriculture, forests, municipalities, sewage treatment plants, and livestock farming can be transformed into biofuels through various thermochemical and biological biorefinery methods [7]. Common thermochemical biorefinery techniques include torrefaction, pyrolysis, liquefaction, gasification, and transesterification [8]. These thermochemical methods yield biofuels, including torrefied biomass, bio-oil, synthesis gas, and biodiesel [9]. In contrast, bioconversion methods involve anaerobic digestion and fermentation, generating biofuels (e.g., biogas, biohydrogen, biopropane, biopropanol, bioethanol, and biobutanol) as well as high-value biochemicals and organic acids (e.g., lactic, acetic, butyric, succinic, levulinic, and hyaluronic acid) [10]. In recent times, thermochemical and biological biorefineries have been integrated to benefit from waste minimization, effective byproduct utilization, reduced resource and energy input, and a circular economy [11].

Anaerobic digestion has been demonstrated as a scalable and practical technology for converting garden waste, food waste, wastepaper, and other organic components of municipal solid waste into biogas [12]. Gathering all accessible organic waste materials and processing them through anaerobic digestion for biogas production can lead to a reduction in greenhouse gas emissions by approximately 3.3–4.4 billion tons of CO₂ equivalent, which is about 10–13% of the global greenhouse gas emissions [11]. On a worldwide scale, the potential energy generation from the currently available and recovered feedstocks is estimated to be 10,100–14,000 terawatt-hours. This energy could represent roughly 6–9% of the primary energy consumed or about 23–32% of the coal used worldwide.

Biogas can generate electricity through combustion or be processed into clean natural gas suitable for use in vehicles. With a typical heating value of 20–25 MJ/m³, biogas comprises a mixture of CH₄, CO₂, H₂, H₂S, N₂, and small quantities of NH₃. However, raw biogas undergoes several purification and conditioning processes to eliminate impurities, such as moisture, volatile matter, CO₂, H₂S, N₂, and NH₃, enhancing its heating value [13]. Biogas can also be upgraded to produce renewable natural gas (RNG), clean natural gas (CNG), or liquefied natural gas (LNG), which can be used as gaseous biofuels for transportation, residential heating and cooking, and industrial processes. The upgrading of biogas involves the removal of impurities (e.g., H₂S, NH₃, moisture, and siloxane), CO₂ separation, compression, and liquefication. Adsorption using molecular sieves, silica gel, activated carbon, and biological scrubbers, along with cryogenic separations, can remove impurities and CO₂ from biogas [14]. Additionally, pressure swing adsorption, water scrubbing, membrane separation, and chemical adsorption are also used to remove CO₂ from biogas.

The effective use of anaerobic digestion can lead to improved sanitation, decreased waste volume, and increased biomethane-rich biogas production [15]. Anaerobic digestion also yields a digested slurry, similar to compost, which can be utilized as a fertilizer in agriculture. The semisolid digestate is generally repurposed for composting or as a soil enhancer. The anaerobic digestion process involves a diverse microbial community, where hydrolysis facilitates the breakdown of complex molecules, making them accessible for further degradation by microorganisms [16]. Given these social, environmental, and economic benefits, anaerobic digestion plays a crucial role in promoting various United Nations Sustainable Development Goals (UN SDGs) by transforming organic waste into bioenergy, thereby enhancing energy security and producing biofertilizer for sustainable agriculture. Anaerobic digestion addresses several UN SDGs, particularly #6 (Clean Water and Sanitation), #7 (Affordable and Clean Energy), #11 (Sustainable Cities and Communities), #12 (Responsible Consumption and Production), and #13 (Climate Action).

Lignocellulosic biomass consists of agricultural crop residues and forestry refuse, primarily cellulose, hemicellulose, and lignin. Cellulose is made of hexose sugars, whereas hemicellulose is made of pentose and hexose sugars and sugar acids, such as glucuronic acid and galacturonic acid. On the other hand, lignin, a phenylpropane polymer that provides rigidity to plant cell walls, is highly resistant to degradation by acids, solvents, and most hydrolytic enzymes [17]. Considering the cost and toxicity challenges associated with physical and chemical pretreatments, studies have utilized diverse microorganisms for biomass pretreatment, thereby enhancing its biodegradability.

Various studies have employed physical pretreatments to break down the complex compositions of organic biomass. Physical pretreatments, such as applying shear stress, ultrasonication, and microwaves, have been proven effective in enhancing the concentration of readily available fermentable sugars for microbial degradation during anaerobic digestion and increasing biogas production [18]. Additionally, chemical pretreatments, such as dilute acids, bases, organic solvents, and oxidizing agents, can significantly affect biogas production during anaerobic digestion. Lomwongsopon and Aramrueang [19] reported that alkaline (i.e., KOH) pretreatment significantly improved biomass digestion and biomethane yields by 18–33%.

The biological pretreatment method is recognized as the most eco-friendly and cost-effective, although the process is time-intensive. Lignocellulolytic fungal strains, like Trichoderma viride, Ceriporiopsis subvermispora, Phanerochaete chrysosporium, and Trametes versicolor, are reported to decrease lignin content from 21% to 70%, thereby enhancing biomethane yield from 23–154% from anaerobic digestion [20,21,22]. Many recent studies have focused on novel microbial consortia that can degrade complex lignocellulosic biomass into sugar monomers [23,24]. Consequently, suitable pretreatment and its effect on other process parameters involved in anaerobic digestion are equally crucial for the complete utilization of biomass and efficient generation of biomethane. Moreover, combined pretreatment approaches, such as alkali hydrothermal pretreatment, can potentially increase the efficiency of the anaerobic digestion system by reducing the consumption of alkali agents and thereby enhancing the biodegradability of the sludge [25]. Hence, combining physical, chemical, and biological pretreatment approaches offers valuable insights for developing an effective pretreatment technique for biomass hydrolysis before anaerobic digestion.

Recent studies have focused on individual pretreatment approaches and their impact on various aspects of the anaerobic digestion system. The novel aspect of this review lies in exploring integrated pretreatment approaches that combine different pretreatment modes to induce a synergistic effect on the efficiency of the anaerobic digestion process. This review also discusses recent advances in biomass pretreatment technologies and their potential implications and effectiveness on the process intensification of anaerobic digestion. Moreover, the role of beneficial additives in the anaerobic digestion system and their impact on efficiency are explored and discussed. Furthermore, the role of such catalysts in stimulating direct interspecies electron transport and altering microbial community dynamics is also highlighted.

2. Anaerobic Digestion

Anaerobic digestion, also known as biomethanation, is a widely employed biological technology used to treat a diverse range of organic waste, including agricultural residues, municipal solid waste, sewage sludge, and industrial effluents, to produce biogas. This technique efficiently decreases chemical oxygen demand and pathogenic bacterial load, generating biogas. Anaerobic digestion can be categorized based on the input and management of organic matter in wet anaerobic digestion (solids < 15%) and dry anaerobic digestion (solids > 15%) [26]. Anaerobic digestion involves four stages, namely, hydrolysis, acidogenesis, acetogenesis, and methanogenesis, as illustrated in Figure 2 [27,28]. Upon analyzing the microbial community structure using 16S rRNA amplicon sequencing, Bacillus, Lactobacillus, and Ruminococcus were identified as the dominant acidogenic genera [29]. Kurade et al. [30] observed the dominance and significant role of Propionispira and Sporanaerobacter in acetogenesis. However, the dominance of microbial species depends on various process parameters, such as pH, temperature, type of feedstock, and hydraulic retention time [31].

Hydrolysis involves converting complex and high-molecular-weight compounds (e.g., lipids, carbohydrates, and proteins) into simpler molecules (e.g., short-chain fatty acids, amino acids, and sugars). The hydrolysis rate is influenced by the type of macronutrients, with the biodegradation of amorphous cellulose occurring more rapidly than that of crystalline cellulose, as well as by substrate level, pH, with an optimal range of 5–7, and temperature, with an optimal range of 30–50 °C [32]. Equation (1) represents the hydrolysis reaction that converts lipids into glycerol and long-chain fatty acids.

$${\mathrm{C}}_{57}{\mathrm{H}}_{104}{\mathrm{O}}_6+3{\mathrm{H}}_2\mathrm{O}\to {\mathrm{C}}_3{\mathrm{H}}_8{\mathrm{O}}_3+3{\mathrm{C}}_{18}{\mathrm{H}}_{34}{\mathrm{O}}_2$$

(1)

Hydrolysis is followed by the acidogenic phase, during which low-molecular-weight compounds are incorporated into a metabolic cycle [33]. In this phase, gaseous products, like H₂, H₂O, and CO₂ are formed, whereas intermediate products, like butyric acid, valeric acid, acetic acid, and propionic acid, collectively known as volatile fatty acids, are generated. Acidogenic microorganisms are the most active and abundant in anaerobic digestion [34]. Acetogenic bacteria favor the breakdown of organic matter into acetic acid during the acidogenic phase. The acidogenic and acetogenic phases provide maximal energy, enhance microbial growth, and withstand an acidic pH range of 5–6. Such rapid development might hinder the anerobic digestion process due to a sudden drop in pH, causing pH shock to microbial species, particularly when acids are not metabolized rapidly [35]. Additionally, varying concentrations of CO₂ and H₂O can be observed in this phase, depending on the characteristics of the feedstock. Specifically, feedstocks with higher carbohydrate content generate higher concentrations of H₂ and CO₂. Equations (2) and (3) show the general reaction for long-chain fatty acid and glycerol acidogenesis, respectively.

$${\mathrm{C}}_{18}{\mathrm{H}}_{34}{\mathrm{O}}_2+16{\mathrm{H}}_2\mathrm{O}\to 9{\mathrm{C}}_2{\mathrm{H}}_4{\mathrm{O}}_2+15{\mathrm{H}}_2$$

(2)

$${\mathrm{C}}_3{\mathrm{H}}_8{\mathrm{O}}_3\to {\mathrm{C}}_3{\mathrm{H}}_6{\mathrm{O}}_2+{\mathrm{H}}_2\mathrm{O}$$

(3)

In the next stage, acetogenic microorganisms transform glucose into volatile fatty acids, predominantly acetic acid. The hydrogen generated is metabolized quickly to prevent the inhibition of the next stage, i.e., methanogenesis. In this phase, the microbial growth rate is slower, with a doubling time of 2–4 days. Equations (4) and (5) illustrate the transformation of butyric acid and acetic acid into butyrate and acetate, with H₂ and CO₂ representing the acetogenesis reaction, respectively. Methanogenesis is the last stage of anaerobic digestion, where acetic acid, CO₂, and H₂ are converted into CH₄, the primary component of biogas [36].

$${\mathrm{C}}_3{\mathrm{H}}_6{\mathrm{O}}_2+2{\mathrm{H}}_2\mathrm{O}\to {\mathrm{C}}_2{\mathrm{H}}_4{\mathrm{O}}_2+3{\mathrm{H}}_2+{\mathrm{C}\mathrm{O}}_2$$

(4)

$${\mathrm{C}}_4{\mathrm{H}}_8{\mathrm{O}}_2+2{\mathrm{H}}_2\mathrm{O}\to 2{\mathrm{C}}_2{\mathrm{H}}_4{\mathrm{O}}_2+2{\mathrm{H}}_2$$

(5)

$${\mathrm{C}}_2{\mathrm{H}}_4{\mathrm{O}}_2\to \mathrm{C}{\mathrm{H}}_4+{\mathrm{C}\mathrm{O}}_2$$

(6)

$$\mathrm{C}{\mathrm{O}}_2+4{\mathrm{H}}_2\to \mathrm{C}{\mathrm{H}}_4+2{\mathrm{H}}_2\mathrm{O}$$

(7)

Methanogenesis is carried out by two groups of microorganisms, especially acetoclastic and hydrogenotrophic methanogens. Acetoclastic methanogenic bacteria utilize acetate, whereas hydrogenotrophic methanogens utilize hydrogen, with doubling times of 2–3 days and 5–6 h, respectively [37]. Other methanogens consume methylated products, formates, and alcohols, and are commonly referred to as methylotrophic, formotrophic, and alcoholotrophic methanogens. Acetoclastic methanogens, which account for approximately 70% of biomethane production, are sensitive to pH fluctuations and the concentrations of trace elements in the anaerobic digestion media. Despite being acknowledged as slow-growing microorganisms in anaerobic digestion, the action of hydrogenotrophic methanogens contributes to maintaining low hydrogen partial pressure during biomethanation [38].

The concentration of organic acids and pH in the bioreactor adversely affect biomethane generation due to the sensitivity of methanogenic bacteria to fluctuating pH in single-stage anaerobic digestion. Conversely, two-phase anaerobic digestion has emerged as a promising technology, which involves hydrolysis, or the breakdown of complex polymers, such as carbohydrates, proteins, and lipids, into simpler ones in the first stage, followed by a methanogenic stage that provides optimal growing conditions for each microbial group [39]. Zhao et al. [40] explored temperature staging based on a two-stage anaerobic digestion process, which resulted in a 47% removal of volatile solids. Furthermore, this approach enabled on-demand biogas production with desired characteristics, as the hydrolysis product could be dynamically added to the bioreactor. However, the process could be significantly stabilized, and inhibitions due to pH and organic loading rate could be prevented by two-stage separation, providing a simpler and reliable digestion process.

The use of anaerobic membrane bioreactors is another promising approach to intensify the biomethanation process. This bioreactor configuration overcomes various limitations of conventional reactors, thereby increasing the efficiency of anaerobic digestion. It combines anaerobic digestion with membrane filtration, serving as a barrier for solid–liquid separation [41]. The biodegradability of organic material can be increased by phase separation and pretreatment techniques. Anaerobic membrane bioreactors offer high density and retention of beneficial microbial communities for effective biodegradation and biomethane production [42]. These bioreactors can operate at high organic loading rates due to high retention of microorganisms [43]. Recent studies have investigated the potential of these bioreactors to enhance biomethane production at hydraulic retention times of less than 11 days [44,45]. The efficient valorization of sludge and the highest biomethane production of 0.69 L/day were observed in a single-stage anaerobic digestion utilizing a membrane bioreactor at a hydraulic retention time of 12 days [46].

Bioaugmentation is another promising approach that involves the introduction of a specific microbial strain or consortium of microorganisms to enhance anaerobic digestion efficiency, stability, and biomethane production. For instance, hydrolysis is regarded as a rate-limiting step, and the introduction of species, such as Clostridium and Bacteroides, or other hydrolytic microorganisms, can enhance the hydrolysis phase or accelerate the breakdown of complex substrates into simpler molecules. On the augmentation of lignin-degrading microbial consortia, Ozsefil et al. [47] found an overall increase of 141% in biomethane generation. Hydrogenotrophic methanogenic bacteria are known to utilize hydrogen for biomethane production [48]. Shang et al. [49] introduced a psychrophilic methanogenic bacterial strain to the anaerobic digestion of corn straw, reporting a 26% improvement in biomethane production.

Perman et al. [50] simulated an industrial-scale anaerobic digestion facility with a capacity of 2100 m³ using a plug flow bioreactor, which utilized garden, agricultural, and food waste under thermophilic conditions. They observed that lab-scale and industrial-scale anaerobic digesters had similar process efficiencies with volatile solids reduction in the range of 41–43%. Bona et al. [51] used wood-derived biochar as an additive in pilot-scale anaerobic digestion, utilizing the organic fraction of municipal solid waste, which resulted in significant changes in microbial community structure. Azizi et al. [52] added carbon cloth as an additive to a pilot-scale reactor with a capacity of 114 L. They observed a 22% increase in biomethane production, reporting the abundance of Methanosarcina, Methanobacterium, and Pseudomonas that facilitated direct interspecies electron transfer.

Zhang et al. [53] utilized a 30 L anaerobic digester to valorize food waste by incorporating biochar as an additive, resulting in a 33% overall increase in biogas production. The addition of biochar improved stability at a larger scale. Biochar’s particle size reduction enhanced the uniformity of its spatial distribution. However, a significant biomethane enhancement was not observed due to excessive microbial colonization. A continuous stirred reactor of capacity 50 L, fed with swine manure, was amended with coconut shell biochar [54]. A pilot-scale digester, measuring 700 L, was fed with 15 g/L biochar and had a retention time of 30 days. A total of 345 g of biochar was dosed after each discharge to replace the biochar that was lost on discharging. The bioreactor dosed with biochar substantially increased biomethane production, thereby demonstrating the feasibility and potential of intensification through biochar amendment at a larger scale [55].

A full-scale plug flow reactor with a volume of 800 m³, located in Sweden, was modified to incorporate biochar made from food waste and cow manure. The addition of biochar intensified biomethane generation by stimulating direct interspecies electron transfer, thereby enhancing process stability at high organic loading rates and improving the syntrophic interactions among microbial communities [56]. These potential studies represent the intensification and energy improvement brought at a larger scale using conducive materials. Some large-scale anaerobic digesters operating worldwide, with varying capacities and biomethane potential, are listed in

Table 1. Select case studies on small, large, and demonstration-scale anaerobic digestion facilities operating internationally.

Name and Location	Feedstock	Digester Type	Biogas Production Potential	Project Start Year	Capacity	Primary Use	Reference
ARC Bio Fuel Private Limited (India)	Cow and poultry manure	Continuous stirred tank reactor	5000 m3/day	2016	-	Bio-CNG	Global Methane Initiatives [57]
Frantoio Oleario Domenico Cassese (Italy)	Olive oil byproduct	Two-stage bioreactor	Biogas to generate 100 kW of electricity	-	400 m3	Combined heat and power	Tamborrino et al. [58]
Govind Godham Gaushala (India)	Cattle manure	Floating drum	150 m3/tank	2014	-	Cooking and electricity generation	Global Methane Initiatives [57]
Jinhua Kitchen Waste (China)	Kitchen waste	Continuous stirred tank reactor	3944 m3/day	2016	1200 m3	Electricity generation	IEA [59]
Kern Cluster (USA)	Manure and food waste	Covered lagoons	5 million diesel gallons equivalents	2013	-	Electricity generation, injection into natural gas pipelines, and Bio-CNG	Global Methane Initiatives [57]
Lily Group (China)	Pigment wastewater	Up-flow anaerobic sludge bed	2.7 million m3/year	2021	1500 m3 per reactor	Injected into natural gas pipelines	IEA [60]
Noblehurst Farms (USA)	Manure, food waste, whey, and process water	Mixed bioreactor	432,000 ft3/day	2015	1,336,710 gallons	Combined heat and power	USEPA [61]
W. Hamburger Facility (Austria)	Wastewater	Up-flow anaerobic sludge blanket	17,600 Nm3/day	2016	3500 m3	Combined heat and power	IEA [62]

.

3. Process Parameters Influencing Anaerobic Digestion

3.1. Temperature

The selection of the optimum temperature is crucial in the anaerobic digestion system as it influences the rates of hydrolysis and methanogenesis. Moreover, it is an essential aspect for microbial communities to thrive and metabolize the feedstock compositions at a faster rate. Additionally, it impacts the gas transfer rate and settling properties of biosolids. The temperature profile is categorized into psychrophilic (<25 °C), mesophilic (30–40 °C), and thermophilic (50–60 °C), where most anaerobic digestion processes are carried out at mesophilic and thermophilic temperatures. Although anaerobic digestion at psychrophilic temperatures has low energy input and high process stability, it is typically characterized by a slower biomass digestion rate, poor pathogen removal efficiency, and low biomethane yields [63]. Conversely, anaerobic digestion at thermophilic temperatures has relatively higher energy requirements, a faster biomass digestion rate, and excellent pathogen removal efficiency and biomethane yields, but it is sensitive to fluctuations in process stability.

Rapid temperature changes can disrupt microbial populations, resulting in a decrease in methane production. Mesophilic bacteria exhibit greater resilience and adaptability to environmental changes. Although thermophilic bacteria operate more effectively, they are more susceptible to variations in pH, ammonia levels, and organic loading rates. Some examples of psychrophilic bacteria involved in anaerobic digestion in colder climates or unheated digesters are Acetobacterium psychrophilum, Clostridium psychrophilum, Psychrobacter, and Methanogenium frigidum [63,64]. Due to their stability and efficiency, mesophilic bacteria commonly used in anaerobic digestion include Bacteroides, Clostridium butyricum, Methanobacterium formicicum, and Methanosaeta concilii [31]. Characterized by faster bioconversion and better pathogen removal, some thermophilic bacteria used in anaerobic digestion include Caldicellulosiruptor bescii, Methanosarcina thermophila, Methanothermobacter thermautotrophicus, and Thermoanaerobacterium thermosaccharolyticum [65].

The primary drivers of effective methanogenesis at 45 °C are Methanosarcina, Anaerolinea, Caldicoprobacter, and Bathyarchaeia [66]. In the anaerobic digestion of food waste, the biomethane content decreased at approximately 15 °C, resulting in no biogas production. Moreover, the concentration and activity of methanogenic bacteria in an anaerobic digester affect the temperature and pH, elevating them to a certain level [67]. A temperature range of 45–55 °C has been reported to increase the pH value, likely due to a reduction in the production of volatile fatty acids [68].

3.2. Hydraulic Retention Time

Hydraulic retention time refers to the average residence time of the substrate inside the bioreactor during anaerobic digestion. It plays a vital role in the growth of microbial species. It can also be one of the reasons for the collapse of anaerobic digestion, due to the washing away of microorganisms resulting from an inadequate hydraulic retention time. However, a longer hydraulic retention time could increase operating costs due to higher heating costs, increased energy input for substrate mixing and pumping, and a larger bioreactor volume [69].

In a study on the anaerobic digestion of sludge, a sudden decrease in microbial diversity was observed following a reduction in hydraulic retention time from 30 to 10 days [70]. In another study, biomethane yield decreased from 0.28 to 0.12 L/g VS (volatile solids) following a decrease in hydraulic retention time from 20 days to 3 days [71]. It also affected the acetate content, increasing it from 38 mg/L to 376 mg/L.

Wu et al. [72] utilized oily food waste as a substrate for biomethane production to optimize operational parameters, such as temperature and hydraulic retention time. The optimized temperatures were 35 °C (mesophilic) and 55 °C (thermophilic). For stable operation, a 4.5% volatile solids content was chosen as the optimum feeding rate. For the first stage, the hydraulic retention time exceeded 10 days. This was because β-oxidation of long-chain fatty acids is slow at shorter hydraulic retention times. Additionally, the accumulation of these acids can lead to the inhibition of anaerobic microflora, resulting in the collapse of the anaerobic digestion system. Sillero and Solera [73] evaluated the impact of varying hydraulic retention times on the anaerobic digestion of poultry manure, wine vinasse, and sewage sludge. Their study assessed the mono-digestion, bi-digestion, and tri-digestion of these organic wastes. The retention times varied from 6 days to 20 days. It was observed that maximum biomethane yield was generated by tri-digestion at a hydraulic retention time of 13 days.

3.3. pH and Volatile Fatty Acids

The pH level of an anaerobic digestion system plays a crucial role in determining the stability and overall efficiency of the biochemical process, including adjustments in reaction rates. An efficient anaerobic digestion process corresponds to the continuity of biochemical reactions in each phase, where metabolites are transformed quickly to enter the next phase, resulting in no accumulation of intermediates. For instance, the accumulation of volatile fatty acids will decrease the pH of the anaerobic digestion system, resulting in the process’s collapse and inhibition of methanogenesis. Therefore, alkaline sources enhance biomethane generation [74]. In this regard, biochar plays an active role in accelerating the consumption of volatile fatty acids, mitigating the collapse of the anaerobic digestion system, and maintaining the required pH [75]. Biochar is a solid carbon-rich porous material obtained from carbonization, pyrolysis, or gasification of organic waste biomass [76]. Depending on the feedstock type and other process conditions, a volatile fatty acid concentration above 6 g/L can inhibit the system in solid-state anaerobic digestion. Significant biogas generation occurs at concentrations below 6 g/L of volatile fatty acids, which pose no stress to hydrolytic or acidogenic bacteria, resulting in greater reactor stability [77,78].

In a study on the co-digestion of kitchen waste and cattle manure, the highest biomethane and volatile solids degradation was observed at an initial pH of 7.5 [79]. In a study where granular activated carbon and algal biochar were used to enhance the anaerobic digestion process of sewage sludge, no significant difference was observed between the buffering capacities of these materials and those of deionized water [80]. It was observed that FeCl₃-modified biochar produced from coconut shells exhibited 81% removal of volatile fatty acids [81]. Additionally, it was reported that the addition of biochar to an anaerobic digestion system enhanced volatile fatty acid production by 11–81%, promoting the hydrolysis and acidification rates of organic matter and increasing biomethane yield [81,82]. At neutral pH, a steady generation of acetic acid, butyric acid, lactic acid, and biomethane was observed, indicating the tendency towards acidogenesis and methanogenesis.

4. Pretreatment of Biomass

The recalcitrant nature of organic biomass necessitates the development of suitable pretreatment technologies, particularly for agricultural residues with a high lignin content, which hinders their direct use. Moreover, suitable pretreatment is necessary to utilize the biomass entirely and increase the activity of microorganisms in the anaerobic digestion process.

Table 2. Pros and cons of different physical, chemical, and biological biomass pretreatment technologies.

Method	Advantages	Disadvantages	Reference
Physical pretreatment methods
Size reduction	Well-established technology. Full-scale employment. Low cost. Reduces biomass particle size. Increases biomass surface area.	Requires subsequent pretreatments. Creates fine biomass particles that can cause fire hazards upon accumulation.	Kim et al. [83]
Drying	Easy operation. Increases the heating value of biomass.	Cost-extensive. Energy-intensive.	Lytras et al. [84]
Microwave	Fast process. Cost-effective. Less heat loss. Uniform heating.	Energy-intensive. High operational costs. Relatively less popular.	Arman et al. [85]
Ultrasonication	High efficiency. Lower operating temperature. Causes preliminary degradation of lignocellulosic biomass.	Cost-intensive. Limited to wet digestion technology.	Zhao et al. [86]
Chemical pretreatment methods
Acid and alkali	Easy handling. Easy operation. Widely used. Removes hemicellulose and amorphous cellulose.	Requires a high concentration of acids and bases for degrading lignin. Generation of byproducts (phenols and furfural) inhibits fermentation. Remediation of waste products is required. Expensive equipment due to corrosive acids and bases.	Nanda et al. [87]
Oxidizing agents	Easy operation. Less toxicity.	Cost-intensive. Explosive nature.	Gomes and Lucas [88]
Ionic liquids	Selective to degrade lignin. Reusable.	High cost of ionic liquids. Relatively less popular.	Pérez-Pimienta et al. [89]
Biological pretreatment technologies
Bacterial or fungal pretreatment	Less resource-intensive. Requires ambient conditions. Environmentally friendly. Wide spectrum of products.	Longer reaction time. Relatively low product yield. Prone to contamination. Requires prior biomass pretreatment.	Rani and Dhoble [90]
Enzymatic saccharification	Faster reaction. Requires ambient conditions. Environmentally friendly. Product-selective.	High cost of enzymes. Recycling enzymes is tedious. Changes in operating conditions impede saccharification rates.	Xiu et al. [91]
Integrated pretreatment methods
Physical, chemical, and biological pretreatment	Co-processing. Combined effect. Effective resource management. Circular economy approach. Recycling waste is possible. Less waste and byproduct formation.	High cost due to sequential processes. Requires more time. Product quality could be compromised by additional processes.	Du et al. [92]

presents the various pretreatment approaches employed, along with their advantages and disadvantages. The physical pretreatment process involves pulverizing or size-reducing biomass before anaerobic digestion to provide a more accessible surface area for hydrolysis. It helps in the solubilization of various constituents of biomass. Some physical pretreatment techniques involve extrusion, microwave heating, and ultrasonication. By removing hemicellulose and lignin, the chemical pretreatment approach provides a viable alternative for reducing the crystallinity and degree of polymerization of cellulose, thereby enhancing its biodegradability.

Table 3. Different physical pretreatment methods used for anaerobic digestion.

Feedstock	Reaction Conditions	Main Observations	Reference
Pretreatment: Extrusion
Cassava starch and sugarcane bagasse	Temperature: 129 °C Screw rotation: 132 rpm Barrel diameter: 16 mm	A 42% increase in CH4 yields. An 86% reduction in chemical oxygen demand.	Fasheun et al. [93]
Food waste	Pressure: 35 MPa	Biomethane yield of 580–605 m/g VS. Biomass particle size reduced to 16 mm.	Kong et al. [94]
Organic fraction of municipal solid waste	Grate size: 6 mm	Biomethane yield of 445 mL/g VS. Balanced carbon/nitrogen ratio.	Mu et al. [95]
Pretreatment: Microwave
Food and lipid waste	Power: 0–1000 W Time: 0–50 s	Higher soluble chemical oxygen demand. Biomethane yield of 739 mL/g VS.	Yue et al. [96]
Food waste	Power: 600 W Time: 20 min	Increment in soluble chemical oxygen demand from 16 to 65%. Biomethane yield of 76 kJ/g VS. Nearly 20% increase in biomethane yield.	Liu et al. [97]
Fruit and vegetable waste	Power: 200–500 W Temperature: 98 °C Time: 10 min	Increase in the pH value. Increased volatile fatty acid production. Biogas yield of 542 NmL/g VS.	Agrawal et al. [98]
Slaughterhouse sludge	Power: 302–500 W Temperature: 45–66 °C Time: 3–7 min	Enhanced hydrolysis of complex organic molecules. Biogas yield of 490 NmL/g VS. Nearly 5-fold increase in biogas yields.	Arman et al. [85]
Pretreatment: Ultrasonication
Cow dung	Frequency: 20 kHz Power input: 650 W	Soluble chemical oxygen demand increased by 94%.	Xu et al. [99]
Sewage sludge	Frequency: 24 kHz	Biomass particle size reduced by 57%. Biomethane yield of 177 NmL/g VS. Nearly 78% increase in biomethane yield.	Zhao et al. [100]
Waste active sludge	Frequency: 26 kHz Power range: 0–2000 W	Soluble chemical oxygen demand increased by 0.6–3.6 times. Biomass particle size decreased from 158 μm to 67 μm. Nearly 26% increase in biogas yield.	Zhao et al. [86]

Abbreviations: volatile solids (VS).

summarizes several studies on the effects of various physical pretreatment methods on anaerobic digestion.

The biodegradability of organic biomass can be enhanced through pretreatment with specific chemicals, including acids, bases, deep eutectic solids, and ionic liquids. These chemicals facilitate the hydrolysis of complex sugars into simple sugars, making them more accessible to anaerobic microorganisms. Alkaline pretreatment breaks down the bonds that bind lignin to other polymers, increasing the internal surface area of organic biomass through a saponification reaction. By eliminating acetate groups from hemicellulose and partially solubilizing lignin, alkaline pretreatment also improves access to cellulose and hemicellulose [101]. Chemicals have a better influence on fermentation and are easier to handle and operate. However, the main drawback of employing chemicals for pretreating biomass is that they may produce secondary pollutants and inhibitory compounds that need further recovery and could impede fermentation.

Table 4. Different chemical pretreatment methods used for anaerobic digestion.

Feedstock	Reaction Conditions	Main Observations	Reference
Pretreatment: Acids and bases
Cassava pulp	Temperature: 30 °C Time: 0–240 h KOH concentration: 0.5–3%	Soluble chemical oxygen demand increased by 10–18 times. Hemicellulose content decreased by 48–58%. Biomethane yield increased by 18–33%.	Lomwongsopon and Aramrueang [19]
Cotton stalk	Temperature: 25–30 °C Time: 24 h KOH and NaOH concentration: 1.5–6% w/w	Increase in chemical oxygen demand. Biomethane yield of 177 mLg/VS. Nearly 226% increase in biomethane yield.	Zhang et al. [102]
Sludge	Temperature: 25 °C Time: 30 min Peracetic acid concentration: 1–4 mM/g VS	Increased volatile fatty acids concentration. Biomethane yield of 298 mL/g VS. Nearly 40% increase in biogas yield.	Ren et al. [103]
Water hyacinth	Temperature: 121 °C Time: 0–90 min H2SO4 concentration: 1–5%	Improved hydrolysis of cellulose and partial removal of lignin. Nearly 131% increase in biogas yield.	Sarto et al. [104]
Pretreatment: Oxidizing agents
Agricultural residues	Time: 40 min Ozonation: 5 mg/L	Significant reductions in total solids and chemical oxygen demand. Biomethane yield of 287 L/kg/VS.	Almomani et al. [105]
Digestate	Ozonation: 0.018 g O3/g TS	Improvement in chemical oxygen demand and reduced latency period. Nearly 22% increase in biomethane yield.	Domínguez et al. [106]
Waste-activated sludge	Time: 48 h Calcium peroxide concentration: 0.02–0.26 g/g VS	Degradation of humus and lignocellulosic components. Nearly 1.3 times the increase in biomethane yield.	Wang et al. [107]
Waste-activated sludge	Periodate concentration: 10–100 mg/g TS	Soluble chemical oxygen demand increased by 2.3 times. Biomethane yield increased from 100 to 146 L/kg VS.	Guo et al. [108]
Pretreatment: Ionic liquids
Agave bagasse	Cholinium lysinate, Ethanolamine acetate, and 1-Ethyl-3-methylimidazolium acetate Ionic liquid concentration: 50–90%	Significant increase in sugar yields and lignin reduction. Near complete consumption of volatile fatty acids. Biomethane yield of 0.28 L/g COD.	Pérez-Pimienta et al. [109]
Grass (Axonopus compressus)	Temperature: 120 °C Time: 2 h Butylmethylimidazolium chloride, 1-Butyl-3-methylimidazolium acetate, 1-butyl-3-methylimidazolium tetrafluoroborate, and 1-Butyl-3-methylimidazolium hexafluorophosphate	Partial removal of lignin and hemicellulose. Biomethane yield of 221 mL/g VS.	Li and Xu [110]
Wheat straw, barley straw, and grape stem	1-Ethyl-3-methylimidazolium acetate Ionic liquid concentration: 5–25%	High conversion of sugars of 158 mg/g. Biomethane yield of 368 mL/g VS.	Marin-batista et al. [111]

Abbreviations: chemical oxygen demand (COD); total solids (TS); volatile solids (VS).

summarizes several studies on the effects of various chemical pretreatment methods on anaerobic digestion.

By breaking down complex organic chemicals into simpler and more easily biodegradable ones, biological pretreatment alters waste composition through enzymatic or microbiological processes. Biological pretreatments can increase the hydrolysis rate, pH stabilization, substrate conversion rate, sugar production, and adaptability in mild environmental conditions without generating inhibitory molecules [112].

Table 5. Different biological pretreatment methods used for anaerobic digestion.

Feedstock	Reaction Conditions	Main Observations	Reference
Barley straw and natural meadow hay straw	Temperature: 25–27 °C Time: 40 days Consortium: Barley straw-adapted microbial communities	Reduced the carbon/nitrogen ratio to an optimal level. Biogas yield of 27 L/kg TS.	Raut et al. [24]
Corrugated board and tissue paper	Temperature: 37 °C Composting inoculum (5 mL) and straw-decomposing inoculum (2 g)	Higher production of volatile fatty acids. Biomethane yield of 343 mL/g VS.	Song et al. [113]
Food waste	Temperature: 60 °C Time: 24 h Fungus: Aspergillus oryzae CICC 40214 (5% w/w)	Increased soluble chemical oxygen demand. Nearly 17% increase in biomethane yield.	Cui et al. [114]
Organic fraction of municipal solid waste	Time: 0–180 min Enzyme cocktail	Enhanced organic matter solubilization. Nearly 10–13% v/v increase in biomethane yield.	Demichelis et al. [115]
Pearl millet, wheat, and rice straw	Temperature: 18–21 °C Time: 28 days Fungus: Pleurotus ostreatus (20% mycelium)	Nearly 40% removal of lignin. Biomethane yield of 187 mL/g VS. Nearly 83% increase in biomethane yield.	Talwar et al. [116]
Rice straw	Temperature: 35 °C Time: 48 h Co-culture: Sporocytophaga CG-1 and Bacillus clausii HP-1	Improved hydrolysis of cellulose and hemicellulose and removal of lignin. Biomethane yield of 93 mL/g VS.	Wang et al. [117]
Rice straw	Temperature: 30 °C Time: 4 weeks Fungi: Pycnoporus Sanguineus MTCC 137 and Trichoderma longibrachiatum MTCC 2478	Nearly 71% removal of soluble chemical oxygen demand. Nearly 21% increase in biogas yield.	Rani and Dhoble [90]
Rice straw	Temperature: 25–30 °C Time: 10 days Consortium: Punjab Agriculture University decomposer: 20% v/w	Increase in reducing sugars yield. Biogas yield of 187 L/kg. Nearly 46% increase in biogas yield.	Sahil et al. [118]
Tall wheat grass (Agropyron elongatum)	Temperature: 28 °C Time: 4 weeks Fungus: Flammulina velutipes	Significant degradation of hemicellulose and lignin. Biomethane yield of 398 Ndm3/kg VS. Nearly 120% increase in biomethane yield.	Lalak et al. [119]

Abbreviations: total solids (TS); volatile solids (VS).

summarizes several studies on the effects of various biological pretreatment methods on anaerobic digestion. Biological pretreatment utilizes diverse microorganisms and enzymes to break down and degrade various biodegradable compositions. Several methods for administering pretreatment are available, including bacterial, fungal, enzymatic, and ensiling methods. Similarly, biological pretreatment enhances biogas generation by reducing pollutants that are toxic to anaerobic microorganisms, such as sulfide and ammonia.

4.1. Extrusion

Extrusion involves applying mechanical and thermal stresses to the substrate by combining two pretreatment techniques. It has many temperature-regulating screw components. The substrate is continually fed into the machine and moved through a small area where the barrel walls and screws provide high shearing stress, which alters the fiber structures [120]. Fasheun et al. [93] employed the extrusion pretreatment technique on cassava starch and observed an increase in its degradability by lowering the adaptive phase from 41 days to 23 days. This increase in biomethane generation was achieved by 42% in a single-stage anaerobic digestion process. It was also observed that the extrusion did not significantly affect the two-stage anaerobic digestion system. Extrusion not only enhances biomethane generation but also affects the overall efficiency of the anaerobic digestion process by reducing the chemical oxygen demand.

In another study, extrusion was employed to separate substances that are difficult to degrade from food waste, and it was stated that extrusion pretreatment could increase the process efficiency and management of organic waste [94]. While extrusion aids the digestibility and accessibility of feedstock to anaerobic microorganisms, it also negatively affects them because of the accumulation of volatile fatty acids [120].

4.2. Microwave

Microwave energy is utilized in the pretreatment of biomass to facilitate the degradation of lignocellulosic biomass into cellulose, hemicellulose, and lignin monomers. Microwave heating can effectively break down lignin and transform or break down hemicelluloses into oligosaccharides and other complex organic waste constituents. Microwave pretreatment offers the benefits of higher yields of fermentable sugars, minimal reliance on chemicals, and shortened reaction times. It operates by applying microwave irradiation to biomass, which can modify the structure of cellulose, eliminate lignin and hemicellulose, and improve the accessibility of cellulose for enzymatic action. It offers other advantages, including uniform heating and energy efficiency [121].

Anaerobic bacteria can readily be used to produce biohydrogen. Microwave pretreatment can readily create hydrothermal conditions, eliminate chlorine, and stabilize heavy metals. However, this procedure requires less time and temperature, suggesting that microwave pretreatment is generally more effective than conventional pretreatment methods [122]. It is observed that microwave pretreatment of sludge increases the maximum and initial degradation rates compared to untreated sludge. Additionally, pretreated sludge exhibits a lower lag phase during anaerobic digestion. The desired temperature is reached very quickly in this case because of significantly less heat loss. Microwave technology heats the interior, reducing heat loss through convection and conduction, which explains the rapid temperature rise. Conventional heating, on the other hand, heats from the outside. Therefore, heat loss in these pathways is unavoidable. Microwave’s non-contact heating method has a distinct advantage over traditional bulk heating in furnaces, as it does not require direct contact between the material and the heating source.

Other significant benefits of microwave irradiation include penetrating radiation, improved heating efficiency, adjustable electric field distributions, instantaneous power and temperature control, and control over response and reaction [85]. Agrawal et al. [98] used microwave technology to pretreat vegetable and fruit waste for co-digestion with anaerobic sludge. There was a 43% increment in the biogas generated from the co-digestion. The increment was attributed to microwave pretreatment, which enhanced the solubilization of the feedstock, resulting in 10% more biomethane production than the untreated feedstock.

It should also be noted that microwave pretreatment has some drawbacks, such as excessive degradation of cellulose, hemicellulose, lignin, and lipids in biomass. This leads to the formation of phenolics and furanic compounds (e.g., furfural and 5-hydroxymethylfurfural), which act as inhibitors for microorganisms during anaerobic digestion and/or fermentation [123]. The furanic compounds cause harm to the nucleic acids of microorganisms and hinder the function of various enzymes involved in glycolysis. Conversely, phenolic compounds specifically alter the membrane permeability of microorganisms, leading to the leakage of intracellular substances and the inactivation of crucial enzymatic systems [124].

4.3. Ultrasonication

High-frequency sound waves are utilized in the ultrasonication process to further pretreat the lignocellulosic structure for subsequent processes. It involves subjecting the substrate to ultrasonic waves, typically those with frequencies exceeding 20 kHz. This technique has been demonstrated to increase biomethane generation on a large scale, producing 3–10 kW of energy for each kW of ultrasonic power applied. It was observed that upon ultrasonication pretreatment of sludge, the soluble chemical oxygen demand increases, and the floc size of the sludge decreases, which is responsible for the enhancement in biomethane production [100].

The effective frequency range for sludge dissociation is 20–40 kHz. A longer ultrasonication time leads to variations in the amount of soluble organic matter, and more complex reactions are observed. Since the released organics and polymers may flocculate during ultrasonic pretreatment, the stabilization of soluble chemical oxygen demand and carbohydrates during the 10–20 min of the ultrasonic period is most likely the result of organics being trapped in a floc structure. In a study, Xu et al. [99] utilized cow dung for biomethane production after subjecting the substrate to ultrasonication pretreatment with a frequency of 20 kHz and a power input of 650 W. It was observed that there was a significant increase in soluble chemical oxygen demand and biomethane production by 945% and 1.4 times, respectively.

Ultrasonication also has several drawbacks, including high energy consumption by ultrasonic generators and transducers, the need for regular maintenance to prevent cavitation and erosion of reactor materials, and the formation of inhibitory products [125]. Similar to microwave pretreatment, the inhibitory products generated from the over-treatment of biomass by ultrasonication include furanic compounds, phenolics, volatile fatty acids, and free radicals, which pose significant harm to microorganisms during anaerobic digestion and fermentation. Hence, preventing the formation and/or removal of these inhibitory compounds is crucial for microbial growth during the bioconversion of biomass.

4.4. Acids and Bases

Hydrochloric acid, phosphoric acid, sulfuric acid, nitric acid, maleic acid, and organic acids are all used in acid pretreatment. Using diluted acids as a pretreatment for lignocellulosic biomass has been extensively studied [126]. The raw material’s surface is pretreated by spraying or soaking it in an acid solution, followed by a specific amount of time spent heating it to a temperature range of 140–200 °C. By breaking down the glycosidic linkages, the acid makes it easier for the hemicellulose fraction to be hydrolyzed by enzymes into oligomers and monomers. This makes the hemicellulose fraction more accessible to microorganisms, improving biogas production. The output of different inhibitory compounds, such as aldehydes, phenolic acid, ketones, acetic acid, furfural, and 5-hydroxymethyl furfural, and the loss of dry matter, along with corrosiveness, hinder the process of methanogenesis in addition to increasing the capital costs because the reactor configuration requires expensive non-metallic containers, even though the use of concentrated acid is highly suited for cellulose hydrolysis [127].

Wang et al. [128] studied the pretreatment behavior of corn stalk with 1–7% H₂SO₄. They reported that maximum levels of inhibitors were found in a bioreactor with 5% H₂SO₄ pretreated corn stalk. An optimum concentration of H₂SO₄, i.e., 5%, can disrupt hydrogen and covalent bonds along with van der Waals forces, and more cellulose can be decomposed and transformed into soluble components, like glucose [104]. Zhang et al. [102] compared the effects of acid and alkali pretreatments. It was observed that more lignin and carbohydrates could be reconstructed by alkali than by acid pretreatment. The biomethane content decreased when the optimum range of the alkali concentration was reached because of the dissolution of much degradable organic matter in the biomass.

Pretreatment of organic biomass with peracetic acid enhances biogas production. In one study, peracetic acid was used in concentrations ranging from 1 to 4 mM/g VS, with a dosage of 2 mM/g VS for pretreatment, resulting in maximum biogas production. Using peracetic acid for pretreating anaerobic sludge is an economical and feasible method [103]. Additionally, the concentration of volatile fatty acids increases with the addition of peracetic acid, resulting in enhanced biogas production. The prevention of propionic acid accumulation resulted in a biogas output at peracetic acid concentrations of 3 and 4 mM/g VS that was inferior to that at a peracetic acid concentration of 2 mM/g [103]. A high quantity of peracetic acid exhibits potent bactericidal properties, whereas excessive levels of volatile fatty acids adversely impact anaerobic microorganisms, disrupting microbial metabolic activities within the system [129].

4.5. Oxidizing Agents

Oxidizing agents help disrupt complex polysaccharides through oxidation reactions, including the oxidative cleavage of aromatic nuclei, side-chain displacement, the formation of alkyl–aryl ether linkages, and electrophilic substitution. Despite their various advantages, they have multiple drawbacks, such as high cost and an explosive nature when these agents are used at high concentrations. Calcium peroxide pretreatment increases the surface area of biomass-dissolving lignin, making cellulose more accessible to anaerobic microorganisms. The hydrolysis of complex organic molecules, such as cellulose and hemicellulose, is facilitated by calcium peroxide pretreatment, resulting in the formation of simple monomers, including sugars, that microorganisms can easily digest [130].

Calcium peroxide pretreatment requires more chemical addition, resulting in increased input compared to other pretreatment methods, such as free nitrous acid and thermal pretreatment. However, calcium peroxide pretreatment can encourage the dissolution of sludge flocs and aid in the degradation of recalcitrant organic or hazardous contaminants in sludge because of its strong oxidizing and alkaline properties [131]. A commonly used solid oxidant, periodate, has been considered an effective and potent oxidant for destroying water contaminants and inactivating microorganisms [131]. Because of periodate’s low reactivity, it must be activated before use for contaminant degradation. Various agents, including reducing agents, alkalis, and transition metals, were employed for activation as they generate reactive species such as iodate oxygen, singlet oxygen, –OH, and O₂⁻. However, a recent study used waste-activated sludge to activate periodate, which substantially pretreated the sludge for efficient biomethane production, as the activation generated reactive agents important for anaerobic digestion efficiency [103].

Additionally, strong oxidants like ozone can degrade and sterilize organic contaminants. Ozone can undergo a slow, selective reaction directly with an organic substrate. Under acidic conditions, this reaction is significant for unsaturated molecules and those with amine or acid groups [88]. However, because of this route, the organic molecules only partially mineralize. Additionally, ozone (O₃) can react with hydroxide ion ($\mathrm{O}{\mathrm{H}}^{-})$ to produce hydroperoxyl radical ($\mathrm{H}{\mathrm{O}}_2^{.}$) and superoxide radical (${\mathrm{O}}_2^{.-}$) (Equation (8)). Superoxide radical and ozone can react to produce ozonide radical (${\mathrm{O}}_3^{.-}$) and oxygen (Equation (9)). Finally, ozone reacts with hydroxyl radical ($\mathrm{O}{\mathrm{H}}^{.}$) to generate hydrotetraoxide radical ($\mathrm{H}{\mathrm{O}}_4^{.}$) (Equation (10)).

$${\mathrm{O}}_3+\mathrm{O}{\mathrm{H}}^{-}\to \mathrm{H}{\mathrm{O}}_2^{.}+{\mathrm{O}}_2^{.-}$$

(8)

$${\mathrm{O}}_2^{.-}+{\mathrm{O}}_3\to {\mathrm{O}}_3^{.-}+{\mathrm{O}}_2$$

(9)

$${\mathrm{O}}_3+\mathrm{O}{\mathrm{H}}^{.}\to \mathrm{H}{\mathrm{O}}_4^{.}$$

(10)

In a study, ozone was used for the pretreatment of digestate from an anaerobic digester, and it was found that the lag phase increased to 10 days in ozonated substrates [86]. However, a substantial increase in biomethane production occurred after the lag phase. Moreover, integrating substrate ozonation with nanoparticles increased the total solids and volatile solids removal by 27% and 24%, respectively.

4.6. Ionic Liquids

Ionic liquids have been regarded as potential agents for handling, processing, and pretreating biomass, as these low-melting-point salts have a tunable nature and low volatility. In a study, mixtures of 1-ethyl-3-methylimidazolium acetate and water (i.e., 70% and 50%) generated sugars at the same rate as untreated Agave bagasse, indicating that ionic liquids lose effectiveness in the presence of water. The study reported that there was no change in the structure and composition following biomass pretreatment [109].

Padrino et al. [132] investigated the use of 1-ethyl-3-methylimidazolium acetate for pretreatment to enhance feedstock recovery under both mesophilic and thermophilic conditions. For this purpose, 3 g of barley straw was immersed in 57 g of 1-ethyl-3-methylimidazolium acetate, accompanied by mechanical stirring in a reactor at 95 °C. To separate the particles and precipitate the sample, 250 mL of deionized water was added, and the mixture was vacuum filtered. It was observed that the overall biomethane recovery increased by 28% and 80% in thermophilic and mesophilic conditions, respectively. This suggested that the biomass was effectively pretreated and that the anaerobic digestion process was more suitable for ionic liquids under thermophilic conditions. Zhang et al. [133] reported that the cost of ionic liquids was very high and that recycling and recovering them could be a viable option. They also noted that when ionic liquids were added to water or ethanol, which are regarded as anti-solvents, recovery from the heated mixture became possible.

4.7. Biological Pretreatment

In a study by Talwar et al. [116], pearl millet, wheat, and rice straw were subjected to fungal pretreatment using Pleurotus ostreatus. Nearly 40% of the lignin was removed in the case of pretreated mixed biomass. There was an overall 83% increase in biomethane generation for the pretreated biomass compared to the untreated biomass. Their study suggests that fungal cultivation before anaerobic digestion effectively disintegrates recalcitrant structures conjugated with lignin, thereby enhancing biomethane production.

In another study, Wang et al. [117] pretreated rice straw with a co-culture of Sporocytophaga CG-1 and Bacillus clausii HP-1 and reported a significant increase in biomethane production. The spore-forming bacteria hydrolyzed the cellulose in filter paper to create sugar monomers. A small portion was utilized for metabolism, and excessive sugar prevented cellulose from functioning when co-culturing with B. clausii. It was possible to quickly adjust the sugar concentration in the culture media, which helped decrease the sugar’s inhibitory effect. Additionally, it was found that these cultures complemented each other in the transformation process of cellulose degradation intermediates, and this substantially accelerated substrate conversion.

Sahil et al. [118] conducted a study on biomass pretreatment with a consortium containing bacteria (e.g., Bacillus, Delftia, Pseudomonas, Lysinibacillus, Arthrobacter, and Paenibacillus) and fungi (e.g., Aspergillus and Trichoderma). They reported a significant increase in the reducing sugar content. The reducing sugars in the consortium-pretreated rice straw increased by 79% compared to the untreated rice straw. Additionally, the protein content increased in the pretreated samples, which was attributed to the disruption of the lignocellulosic complex by hydrolytic microorganisms, thereby increasing the concentration of accessible fermentable sugars.

A study on the enzymatic pretreatment of food waste using Aspergillus oryzae CICC 40214 demonstrated that, in addition to enhancing biomethane production, it can effectively decrease the abundance of antibiotic resistance genes and the risk of their dissemination [114]. The study also reported a significant increase in hydrolysis and anaerobic digestion efficiency. The majority of antibiotic resistance gene variation was driven by horizontal gene transfer and the emergence of changes in microbial community composition.

Compared to physical and chemical pretreatments, biological pretreatments encounter several drawbacks, such as slower reaction rates, longer reaction times, inefficiency in degrading lignin and non-biodegradable materials, expensive hydrolytic enzymes, issues with recycling enzymes, inactivation of enzymes with fluctuations in process temperature and pH, regular maintenance of microbial cultures, presence of inhibitors and impurities in biomass hydrolysate, and undesired contamination [134,135].

4.8. Integrated Pretreatment Methods

Mixed pretreatments provide a collaborative effect on the substrate and are more effective than individual pretreatments.

Table 6. Different integrated pretreatment methods used for anaerobic digestion.

Pretreatment	Reaction Conditions	Main Observations	Reference
Acid-hydrothermal and deep eutectic solvents	Feedstock: Poplar residues Temperature: 170 °C Acetic acid (0–10%) Three-component deep eutectic solvent: 600 g	Nearly 83% removal of xylan, 95% recovery of cellulose, and 14% removal of lignin. Biomethane yield of 208 L/kg VS. Nearly 98–148% increase in biomethane yield.	Xie et al. [137]
Alkali and thermal pretreatment	Feedstock: Waste-activated sludge Temperature: 80 °C NaOH and Ca(OH)2 Dose: 1:4, 2:3, 1:1, 3:2, and 4:1	Increase in soluble chemical oxygen demand. Biomethane yield of 430 mL/g VS. Nearly 172% increase in biomethane yield.	Zou et al. [138]
Alkali and thermal pretreatment	Feedstock: Swine manure Temperature: 121 °C NaOH (3% w/v)	Increase in soluble chemical oxygen demand. Biomethane yield increased from 30 to 205 mL/g VS.	Sousa et al. [139]
Alkaline and photocatalytic pretreatment	Feedstock: Waste-activated sludge NaOH Temperature: 37 °C TiO2 photocatalyst (0.3–0.6 g/L)	Increase in soluble chemical oxygen demand. Biomethane yield of 462 NmL/g VS. Nearly 71% increase in biomethane yield.	Maryam et al. [140]
Hydrothermal pretreatment and co-hydrothermal pretreatment	Feedstock: Garden waste, chicken manure, duck manure, and pig manure Temperature: 140 °C and 180 °C Time: 30 min	Increased hydrolysis of lipids into fatty acids and glycerol. Nearly 37% increase in biomethane yield.	Zou et al. [141]
Microwave-assisted ammonization	Feedstock: Rice straw Microwave power: 660 W Ammonia water (25% w/w) Temperature: 30 °C	Lignin removal and saponification denatured ester bonds in rice straw. Biomethane yield of 5080 mL/d. Nearly 57% increase in biomethane yield.	Liu et al. [136]
Microwave-assisted chemical thermohydrolysis	Feedstock: Straw, cattle manure, and milking station wastewater Temperature: 150 °C Microwave power: 12 kW HCl (0.05–0.20 g/g TS) NaOH (0.05–0.20 g/g TS)	Increased dissolved organic compounds and chemical oxygen demand. Biogas yield of 703 L/kg VS.	Debowski et al. [142]
Thermal hydrolysis and photocatalysis	Feedstock: Waste-activated sludge Temperature: 170 °C TiO2 photocatalyst (0.1–1.0 g/L)	Soluble chemical oxygen demand increased from 292 mg/L to 4066 mg/L. Nearly 66% increase in biomethane yield.	Chen et al. [143]
Ultrasound-assisted thermal pretreatment	Feedstock: Microalgae (Synechocystis) Temperature: 70–110 °C Ultrasound power: 400 W Frequency: 25 kHz	Increased concentration of volatile fatty acids. Biogas yield of 706 NmL. Nearly 1.4–5.6-fold increase in biogas production.	Abedi et al. [144]

summarizes several studies on the impact of various integrated pretreatment methods on anaerobic digestion. In a study, rice straw was pretreated at a microwave power of 660 W, followed by 24 h pretreatment with ammonia water (i.e., 25% w/w) [136]. It was found that enhancing process efficiency by subjecting mixed microwave and ammoniation pretreatment to rice straw resulted in 24%, 27%, and 7% degradation of cellulose, hemicellulose, and lignin, respectively. Additionally, microwave-assisted ammoniation pretreatment resulted in a 57% increase in biomethane production. Microwave heating and ammonia break the ester bonds between carbohydrates and lignin, a process known as saponification.

Debowski et al. [142] performed thermohydrolysis of lignocellulosic biomass, exposing electromagnetic microwave radiation followed by acid and alkaline hydrolysis. Alterations in the amount of organic molecules in the liquid or dissolved phase frequently validate the efficacy of technological approaches [145]. The effectiveness of the thermohydrolysis of biomass is directly demonstrated by the magnitude of the release and the rise in the concentration of dissolved organic molecules. Since low-molecular-weight organics are the substances that microorganisms can most readily metabolize in the first stage of fermentation, an increase in their content in the liquid phase controls the rate of reactions in the anaerobic system [146].

Microwave and alkaline pretreatment of wheat straw significantly affected biomethane generation and lignin removal, where the optimal conditions were found to be 126 °C and 15 min at a 1.5% NaOH concentration [146]. In another study, microwave and hydrothermal-assisted ionic liquid pretreatment was applied to wheat straw, revealing that the optimal conditions were 360 W power and a temperature of 200 °C [147]. This integrated pretreatment technique removed lignin by 35% and 25% in the case of microwave-assisted ionic liquid and hydrothermal-assisted ionic liquid pretreatments, respectively.

Xie et al. [137] utilized poplar residues for biomethane generation with acid hydrothermal pretreatment and deep eutectic solvent pretreatment. This integrated pretreatment significantly decreased lignin and xylan contents and increased biomethane generation by 98–148%. The disaggregation of lignin and hemicellulose occurs upon pretreatment with deep eutectic solvents. This is because lower viscosity increases the dissolution of lignocellulosic biomass at high temperatures. In another study, thermal hydrolysis followed by photocatalysis pretreatment was applied to waste-activated sludge, which increased biomethane content by 66% [143]. Photocatalysis following thermal hydrolysis enhances the availability of organic substrates for producing biomethane and promotes further sludge cleavage [148]. However, the addition of photocatalysis in energy anaerobic digestion encourages the breakdown of refractory materials, enhancing biodegradability and increasing biomethane yield [140].

5. Direct Interspecies Electron Transfer in Anaerobic Digestion

Out of the four phases of anaerobic digestion (i.e., hydrolysis, acidogenesis, acetogenesis, and methanogenesis), the rate-limiting step varies depending on the type and treatment of the feedstock. However, syntrophic acetogenesis is often regarded as an obstacle that governs the process rate [149]. In this phase, C₂–C₆ organic acids or alcohols are converted to electron donors with a low molecular weight. It mainly consists of acetic acid and hydrogen, which are consumed in the next step, known as methanogenesis. This is known as indirect interspecies electron transport. In past studies, various attempts have been made to enhance this route by stabilizing the hydrogen partial pressure and increasing biomethane generation. However, a new route involves the direct transfer of electrons from acetogenic or acidogenic microorganisms to methanogenic microorganisms, a mechanism known as direct interspecies electron transfer [150]. Figure 3 illustrates the various types of electron transfer mechanisms involved in anaerobic digestion.

In a typical anaerobic digestion process, electrons are transferred by mediated and direct interspecies electron transfer. Mediated interspecies electron transfer involves the role of shuttle molecules or the formation of metabolized intermediates, resulting in the transfer of electrons between species. The mediated interspecies electron transfer mechanism requires favorable conditions related to partial pressure to provide ideal thermodynamics, producing more metabolites while maintaining existing metabolites at low concentrations. A similar phenomenon can be observed in the anaerobic digestion system, where hydrogenotrophic methanogenic bacteria consume hydrogen to lower partial pressures, i.e., <10⁻⁴ atm, followed by the action of syntrophic bacteria that carry out the secondary fermentation of fatty acids and alcohols to produce hydrogen [151].

The thermodynamics and mechanism of direct interspecies electron transfer differ from those of mediated interspecies electron transfer, as it does not require metabolite generation or exchange. In the case of direct interspecies electron transfer, electron transfer occurs directly from one organism to another, either through cell-to-cell contact mediated by pili or via electrically conductive materials. Various carbon-based and noncarbon-based agents can activate or stimulate direct interspecies electron transfer among certain exoelectrogenic bacteria and methanogenic archaea. The most prevalent conducive material is biochar, which is utilized at a larger level to stimulate the electron transfer mechanism.

Direct interspecies electron transfer is the direct transfer of electrons from electroactive microorganisms to methanogenic bacteria, facilitating the reduction of CO₂ to CH₄. In comparison to traditional syntrophic relationships, fermentative bacteria degrade organic compounds, producing H₂, which is then consumed by methanogens or other electron acceptor species. In contrast, direct interspecies electron transfer employs direct electron transfer. Among species, it can be carried out via two routes, such as using (i) bioelectric connections like cytochrome c, pili, or electron transport proteins and (ii) connections created by conducive materials. For instance, conducive pili and c-type cytochrome-associated genes in Geobacter species are crucial in direct interspecies electron transfer [152].

The role of stimulated direct interspecies electron transfer in maximizing biomethane production has been explored in various studies using conducive materials, such as magnetite, stainless steel, biochar, graphene, and carbon cloth [153]. Conductive materials act as electron shuttles, enabling them to transfer electrons over long distances and eliminating the need for intermediate products, such as hydrogen.

An abundance of electric species, such as Geobacter and Pseudomonas, was observed upon the addition of conducive materials, and these species play a crucial role in stimulation through direct interspecies electron transfer [154]. Zhao et al. [155] concluded that the abundance of electrical species increased in a bioreactor amended with conducive materials. Electrogenic bacteria can attach themselves to conducive materials to exchange electrons with high conductance, resulting in reduced energy consumption, without needing e-pili or c-cytochrome [156]. However, in the case of methanogens, the abundance of acetate-utilizing methanogens, such as Methanosaeta and Methanosarcina, increases with the addition of conducive materials [157].

The upscaling of anaerobic digestion based on direct interspecies electron transfer mechanisms can be achieved by adding conducive materials, such as biochar, hydrochar, carbon cloth, and magnetite. However, strategies to prevent washout and ensure homogeneous distribution of conducive materials are crucial in the anaerobic digestion process at the pilot scale. This can be achieved using an advanced bioreactor configuration, such as a fixed-bed reactor and optimal continuous dose selection. Biochar, as a conducive material, is usually preferred over other conducive materials because of its high conductivity, porous structure, high specific surface area, non-toxicity, low cost, and sustainable production method [158]. Some studies found that biochar can have negative impacts on the anaerobic digestion process, and modifications can release toxic elements that can inhibit microbial growth [159].

Magnetite and biochar are promising conducive materials for stimulating the direct interspecies electron transfer mechanism. However, continuous anaerobic digestion faces significant challenges relating to the loss of these materials in the effluent due to poor retention and fine size. This necessitates repeated dosing. To overcome these limitations, studies have explored methods such as filtration and immobilization of biochar on a carrier, which can be polymer beads, natural fibers, or biofilms. This approach can effectively increase the conductive material’s retention and colonization of microorganisms, provided by stable attachment sites. For instance, Zhuravleva et al. [160] utilized a non-conductive diatomite layer to spatially separate conducive materials, such as granular activated carbon and magnetite, from the organic fraction of municipal solid waste, obtaining promising results, including the alleviation of volatile fatty acids accumulation. Moreover, packed bed systems and up-flow anaerobic sludge blanket anaerobic digesters can also be utilized, as they provide better retention compared to conventional stirred bioreactors.

6. Novel Additives for Anaerobic Digestion

The biological conversion for biogas production is limited because of its low efficiency, which results from poor electron transfer among bacteria, electroactive bacteria, and methanogenic archaea for biohydrogen and biomethane production, respectively [161]. Recently, numerous studies have investigated the impact of various carbon-based additives, including granular activated carbon, carbon cloth, graphite, and biochar, on enhancing the efficiency of anaerobic digestion systems by stimulating electron transfer to overcome limitations [162,163].

Table 7. Different additives used in anaerobic digestion.

Feedstock	Additive Properties	Gas Yield	Reference
Activated sludge	Additive: Biochar Preparation: Sludge pyrolyzed at 500 °C for 1 h Surface area: 110 m2/g Pore diameter: 0.024 Å	Nearly 70% increase in biomethane yield	Zeynali et al. [164]
Activated sludge	Additive: Biochar Preparation: Pyrolyzed at 500 °C for 2 h Modification: Acylated homoserine lactone	Nearly 52% increase in biomethane yield	Li et al. [165]
Activated sludge	Additive: Biochar Preparation: Fruitwood pyrolyzed at 500 °C for 2 h Modification: Magnetic biochar with FeSO4·7H2O and FeCl3·6H2O Specific surface area: 34 m2/g	Nearly 22% increase in biomethane yield	Jin et al. [166]
Activated sludge	Additive: Biochar Preparation: Fruitwood pyrolyzed at 500 °C for 2 h Specific surface area: 3.5 m2/g	Nearly 42% increase in biomethane yield (135 mL/g)	Li et al. [167]
Activated sludge	Additive: Glycerol	Nearly 1.3-fold increase in biomethane yield	Wang et al. [168]
Brewing wastewater and sludge	Additive: Biochar Preparation: Sludge pyrolyzed at 300 °C for 2 h Modification: Doping of nano zero-valent iron	Nearly 1.4-fold increase in biomethane yield (212 mL)	Li et al. [169]
Cattle manure and corn straw	Additive: Hydrochar Preparation: 220 °C for 4 h Modification: FeCl3	Nearly 34% increase in biomethane yield	Yang et al. [170]
Cheese whey wastewater and wine sludge	Additive: Hydrochar Preparation: Hydrothermal carbonization of agrowaste at 250 °C	Biomethane yield of 130–140 mL/g VS	Liakos et al. [171]
Cow manure and corn straw	Additive: Hydrochar Preparation: Hydrothermal carbonization of bio-residue at 220°C for 4 h Modification: Fe-modification Specific surface area: 192 m2/g	Nearly 36% increase in biomethane yield	Ren et al. [172]
Propionate	Additive: Hydrochar Preparation: Hydrothermal carbonization at 260 °C for 1 h Surface area: 42 m2/g	Nearly 57% increase in biomethane yield	Shi et al. [173]
Rice straw	Additive: Biochar Preparation: Corn stover pyrolyzed at 600 °C for 1 h Surface area: 464 m2/g Pore diameter: 1.80 nm	Nearly 37% increase in biomethane yield of 230 L/kg VS	Bhujbal et al. [174]
Sewage sludge	Additive: Biochar Preparation: Oil sludge and wheat straw pyrolyzed at 600 °C for 2 h Surface area: 9.4 m2/g	Biomethane yield of 144 mL/g VS	Feng et al. [175]
Sheep manure	Additive: Carbon nanotubes Surface area: 78 m2/g Pore size: 14 nm	Nearly 34% increase in biomethane yield	Hao et al. [176]
Sludge	Additive: Co-immobilized hydrogel Modification: Mixed with immobilized microorganisms involved in anaerobic digestion	Nearly 1.3 times increase in biomethane yield	Chan et al. [177]
Swine manure	Additive: Hydrochar Preparation: Hydrothermal carbonization of swine manure digestate of 180–260 °C for 45 min Modification: Fe-modified	Biogas yield of 265 mL/g TS	Shen et al. [178]

Note: The reaction temperature in these studies ranged from 35 to 38 °C. Abbreviations: total solids (TS); volatile solids (VS).

summarizes the effects of different additives on anaerobic digestion.

The thermochemical conversion of biomass waste produces biochar, a carbon-rich material [179]. Adding biochar enhances anaerobic digestion with several advantages, including effective hydrolysis, stimulation of direct interspecies electron transfer, immobilization capability, and buffering capacity [166]. Moreover, pyrolysis process parameters, such as residence time, heating rate, and temperature, can influence the physicochemical properties of biochar. Variations in surface functional groups, porosity, specific surface area, and ash content can be observed. Therefore, the parameters set for generating biochar are crucial for enhancing the performance of anaerobic digestion. Several studies have been conducted on the addition of biochar to improve anaerobic digestion, and current research also focuses on modifying biochar. Biochar can potentially stimulate electron transfer in direct interspecies electron transfer among certain exoelectrogenic bacteria and methanogenic archaea, resulting in enhanced biomethane generation [180].

To support the closed-loop and circular economy approach and minimize waste residues and byproducts, the solid digestate of biomass pretreatment and/or anaerobic digestion can be used as feedstocks for carbonization to produce biochar. Biochar is regarded as a low-cost carbonaceous material that significantly enhances the production of biomethane from anaerobic digestion by stimulating direct interspecies electron transfer. Biochar does not produce biomethane but increases the biomethane production rate and reduces the lag phase for microbial growth during anaerobic digestion [173]. The temperature used for biomass carbonization affects the properties of biochar, particularly its proximate composition (i.e., moisture, ash, volatile matter, fixed carbon), ultimate composition (i.e., carbon, hydrogen, oxygen, nitrogen, and sulfur), elemental composition, surface area, porosity, pH, and conductivity. It is recommended to explore the inhibition of biomethane production due to biochar amendment at the initial stages and further study its adsorption potential on volatile fatty acids and various toxic compounds.

The abundant functional groups, porous structure, excellent electrical conductivity, and thermal stability of biochar make it a suitable carbon-based additive for biomethane production [181,182,183]. This mechanism can significantly enhance the efficiency of converting biomass into biomethane. Combining specific pretreatment methods and incorporating suitable additives offers a promising insight and potential approach to improve and optimize biomethane generation from organic biomass. This approach addresses efficiency and significant effect-related limitations associated with single pretreatment, and by integrating additives, it offers promising advancements in enhancing biogas yields, reducing retention times, and increasing overall process efficiency.

Biochar and activated carbon, products derived from thermochemical conversion, especially carbonization of waste biomass, have been investigated as novel carbon-based additives to enhance anaerobic digestion rates and biomethane productivity [184,185,186]. Several studies have investigated the separation and reuse of biocarbon materials, which have improved biomethane yield and anaerobic digestion efficiency [187,188]. Furthermore, in continuous operations, additives such as magnetic biochar and magnetite can be separated from the effluent because of their strong magnetic properties. They can be recycled back into the continuous bioreactors. Upon adding recycled magnetite, an 80% increase in biomethane was observed [189]. The retrieved or recycled biochar contains numerous microbial attachments, and the material can be effectively managed to maintain these microbial species, which can be active participants in direct interspecies electron transfer. However, a consistent direct interspecies electron transfer demands timely dosing of conducive materials. Therefore, future research studies can focus on exploring reliable, reusable, and conducive materials, as well as optimized reactor configurations.

Li et al. [167] supplemented biochar and acylated homoserine lactone in the anaerobic digestion of activated sludge and reported a 52% increment in biomethane yield. It was observed that biochar modified with acylated homoserine lactone produced less biomethane than when the individual components of biochar and acylated homoserine lactone were added separately. Wu et al. [190] investigated biomethane generation from a sole acylated homoserine lactone and found that the methane generation was negligible. Therefore, the readily biodegradable acylated homoserine lactone can enhance biomethane production. However, if fixed on biochar, it is ineffective. The pores for microbial accumulation are utilized by fixing acylated homoserine lactone, which decreases biomethane yield. Also, free acylated homoserine lactone promotes intracellular communication. Moreover, the co-addition of biochar and acylated homoserine lactone enhances the overall efficiency of anaerobic digestion.

In another study, oil sludge and wheat straw were utilized for biochar production, and it was found that increasing the ratio of wheat straw to oil sludge to an optimal level in co-pyrolysis resulted in biochar generation with a high surface area, which was beneficial for microbial accumulation [175]. Moreover, co-pyrolysis of oil sludge and wheat straw contained a certain level of oxygen and iron elements, which was observed to be crucial and conducive for stimulating direct interspecies electron transfer and was of great significance for increasing biomethane yield. Furthermore, adding biochar obtained by co-pyrolysis increased the rate of hydrolysis of polysaccharides and proteins compared to the normal methanogenesis process during the initial phase of the process.

Several studies have reported the activation of direct interspecies electron transport in Clostridium with electroactive methanogenic bacteria [130,145]. Bhujbal et al. [174] reported the enrichment of Methanobacterium and Methanosarcina in biochar-supplemented bioreactors. Additionally, Methanobrevibacter was found to dominate in the control group, where no biochar was added. The study concluded that there is a clear relationship between biochar supplementation and the enrichment of acetoclastic and hydrogenotrophic methanogens. Methanobacterium, Methanobrevibacter, and Methanoculleus are three major hydrogenotrophic methanogens that are observed to utilize formate and hydrogen with carbon dioxide to generate biomethane [191].

Methanosarcina is regarded as a multi-type of methanogen as it can reduce CO₂ via direct interspecies electron transport and utilize H₂ and CO₂ to generate CH₄. Methanosarcina is also known to play a crucial role in the methanogenesis phase of anaerobic digestion, as it establishes direct interspecies electron transfer with electroactive bacteria [130]. Additionally, Methanothrix can function as an acetoclastic methanogen, accepting extracellular electrons with acetate as the sole energy source through direct interspecies electron transfer during methanogenesis [192].

The more specific the area of biochar and porosity, the more attachments there are to such microorganisms and their enrichment. Additionally, in biochar-supplemented anaerobic digestion systems, surface oxygen functional groups, such as C–O, C–OH, and C=O, can potentially enhance biomethane production. The higher dose of zerovalent iron-doped biochar results in increased production of hydrogen due to corrosion, which disrupts the anaerobic digestion process. The enhancement in the anaerobic digestion system resulting from the addition of hydrochar is attributed to the pores present on its surface, which facilitate microbial attachment and improve electron transfer.

It has been observed that the presence of electrotrophic methanogenic and electrogenic bacteria eliminates the intermediate metabolism step for electron transfer, thereby increasing the rate of methanogenesis [170]. When hydrochar generated from swine manure digestate was used as an additive in the anaerobic digestion of swine manure and digestate, hydrogenotrophic methanogens, including Methanomassiliicoccus, Methanoculleus, Methanofollis, Methanosphaera, and Methanobrevibacter, were found to be dominant in only hydrochar-containing bioreactors. In contrast, in bioreactors containing Fe-modified hydrochar, Methanosarcina was observed as the dominant hydrogenotrophic methanogen. Methanosarcina utilizes a variety of substrates, including H₂, CO₂, acetate, methylamines, and formate. Iron-modified hydrochar can induce Methanosarcina to be a dominant methanogen in anaerobic digestion.

7. Conclusions and Perspectives

Although anaerobic digestion is scalable, adaptable, and commercially viable, various factors and research gaps still require investigation to enhance the efficiency and effectiveness of this process. To improve biomass diversification, appropriate pretreatment is necessary to efficiently break down the complex structures of varied organic biomass types, such as lignocellulosic biomass, ensuring that anaerobic digestion is practical and versatile. Advanced pretreatment methods, particularly integrated approaches that combine physical, chemical, and biological pretreatments, can effectively treat resistant biomass. It is crucial to understand that every pretreatment method has its pros and cons. For instance, size reduction is a commonly utilized and economical approach for biomass pretreatment that enhances the surface area of biomass for hydrolysis. Acidic and alkaline treatments are also frequently employed, in a cost-effective manner, following the size reduction in biomass to eliminate lignin and extract cellulose and hemicellulose from lignocellulosic biomass for anaerobic digestion. Saccharification refers to a biological pretreatment method that facilitates the hydrolysis of biomass to generate fermentable sugars through the action of enzymes and hydrolytic microorganisms. The intensity of pretreatment techniques affects the digestion rate of biomass, the yield of products, and the generation of inhibitory compounds, such as furanic and phenolic compounds, making the optimization of pretreatment a vital step before anaerobic digestion.

Anaerobic digestion represents a viable technology for managing organic biomass and producing sustainable and efficient bioenergy, alleviating the burden on waste management organizations and addressing environmental issues. A comprehensive analysis and evaluation can shed light on the influences of additives, such as biochar, organic compounds, and quorum-sensing molecules, on promoting electron transfer and biomethane production. The mechanism of direct interspecies electron transfer facilitates electron movement between electroactive bacteria and methanogens, which produce biomethane. This process is characterized by energy-efficient electron transfer occurring without the creation of intermediates. Carbon-based substances, including biochar, activated carbon, and hydrochar, facilitate this mechanism, thereby boosting biomethane production. Their impact on changes in microbial community composition can reveal energy-efficient pathways for electron transfer and enhance biomethane production. To advance anaerobic digestion technology, challenges, such as maintaining process stability, ensuring cost-effectiveness, integrating advanced pretreatment techniques with additives, and assessing the effects of these conditions, must be addressed. Bioaugmentation can also facilitate the introduction of beneficial species that are scarce or required for efficient waste valorization in biogas production. Moreover, bioaugmentation using methanogenic bacteria is also a potential approach to maximize biomethane production.

Anaerobic digestion has traditionally proven to be a scalable and adaptable technology for transforming carbonaceous waste, whether rural or urban, into valuable products, such as biogas and organic fertilizer. With continual technological advancements, the future direction of anaerobic digestion is increasingly concentrating on enhancing process intensification to improve efficiency, resource recovery, commercialization, and profitability. Innovations are leading to the incorporation of advanced physical, chemical, and biological pretreatment methods to enhance the degradability of biomass and accelerate digestion rates for biomethanation. The combination of anaerobic digestion with resource recovery techniques, such as nutrient recovery and upgrading biogas to biomethane, is also gaining popularity, in line with circular economy objectives. Furthermore, the integration of real-time monitoring and machine learning-based control systems is anticipated to optimize operational parameters, maintain stability, and maximize output. Additionally, technoeconomic analysis could provide crucial insights into anaerobic digestion by evaluating both technical performance and economic viability. It helps stakeholders gauge the cost-effectiveness of different feedstock supply chains, bioreactor configurations, and product streams to identify cost drivers, estimate return on investment, and assess payback periods, thereby facilitating informed decision-making for the scale-up and commercialization of anaerobic digestion. These advancements collectively strive to enhance anaerobic digestion as a more practical and competitive alternative for waste management and renewable energy generation in both urban and rural environments.