Thermal chemical pretreatment of waste-activated sludge for enhanced solubilization and biogas production: a review

AD

Anaerobic digestion

BNR

Biological nutrient removal

BOD

Biochemical oxygen demand

CHP

Combined heat and power

COD

Chemical oxygen demand

CST

Capillary suction time

CSTR

Continuous stirring tank reactor

DIET

Direct interspecies electron transfer

EPS

Extracellular polymeric substances

FOG

Fats, oils, grease

HRT

Hydraulic retention time

MFCs

Microbial fuel cells

MLSS

Mixed liquor suspended solids

MT

Million ton

MW

Microwave

rbCOD

Readily biodegradable chemical oxygen demand

sCOD

Soluble chemical oxygen demand

sCSTR

Semi-continuous stirring tank reactor

SV

Settling velocity

TAN

Total ammonia nitrogen

tCOD

Total chemical oxygen demand

TN

Total nitrogen

TP

Total phosphate

TS

Total solids

VFA

Volatile fatty acid

VS

Volatile solids

VSS

Volatile suspended solids

WAS

Waste-activated sludge

WWTPs

Wastewater treatment plants

Prompt industrialization and urbanization have induced enormous wastewater volume generation (Kim et al. 2013). Therefore, improving the quality of water bodies where the treated wastewater is discharged needs a more stringent effluent quality (Yang et al. 2013). Among many treatment configurations, the biological activated sludge treatment is widely applied (Wonglertarak & Wichitsathian 2014), employed in about 90% of the municipal wastewater treatment plants (WWTPs) (Cosgun & Semerci 2019). The process produces waste-activated sludge (WAS) consisting of biomass at 0.5–1% of the influent wastewater (Li et al. 2008). This sludge requires to be stabilized and disposed of in an environmentally friendly manner (Xiao et al. 2015). However, less than 30% of produced WAS is disposed of properly now (Zhen et al. 2014). In 2012, China produced around 9.1 million tons (MT) of dry WAS (Zhen et al. 2014), and it is further growing with the expansion of sewage treatment plants (Xiao et al. 2015). The production was 10 MT dry WAS in the European Union (2005), while it was 8.2 MT in the USA (2010) (Xiao et al. 2015).

The WAS, the drawback of biological treatment, has to be dealt with prudently owing to its organic and inorganic contaminants, heavy metals, and pathogenic contents (Yang et al. 2013; Maryam et al. 2021). The literature reports that handling and disposal of WAS accounts for 50–60% of the total operational cost (Maryam et al. 2021). In comparison, its treatment can cost 30–40% of capital cost (Yang et al. 2013); hence, its disposal is a significant environmental and financial concern for WWTPs (WWTPs) (Toutian et al. 2020).

However, the concerns can be offset by its volume reduction, stabilization, and valorization (Kim et al. 2013). WAS can be used for biofuel (methane and hydrogen), bio-fertilizer, and volatile fatty acid (VFA) generation since carbohydrates, proteins, and lipids are its primary organic matter (Yu et al. 2010). Moreover, extracellular polymeric substances (EPS), produced by microorganisms during aerobic treatment, have protein and polysaccharides as significant organic compounds (Yu et al. 2010). Additionally, WAS is a source of phosphorus due to phosphorus accumulation in biomass (microorganisms) (Cosgun & Semerci 2019). WAS is used as substrate in anaerobic digestion (AD) for biogas generation to be combusted in combined heat and power (CHP) plants for electricity and thermal energy production. It is also co/incinerated to recover its energy. Consequently, the WAS treatment can alleviate energy insecurity, depletion of fossil fuels, emission of greenhouse gases, and acceleration of global warming (Geng et al. 2020).

AD is a symbiotic process involving bacteria and archaea, where hydrolyzing and acidifying bacteria degrade organic matters of high molecules to monomers and convert them to VFAs (Yang et al. 2013), which can subsequently be reduced further by methanogenic microorganisms to CH₄ and CO₂ in an anaerobic environment (Rahmani et al. 2022).

The mass transformation of WAS to biogas reduces the overall volume of the digestate after dewatering by up to 50% (Kim et al. 2013) and, hence, less volume and weight need to be handled (Valo et al. 2004). Moreover, the dewatering gets improved after AD. The energy (electrical and thermal) generation via CHPs can meet up to 80% of WWTPs' energy requirement (Toutian et al. 2020). The pathogens and odor are reduced or eliminated at thermophilic AD (Roopnarain et al. 2023). Hence, the digestate, the end product of AD, can be used as a bio-fertilizer. The organic wastes, including WAS, are also anaerobically fermented to VFAs to extract different acids and to be used as an inexpensive carbon source in the BNR (biological nutrients removal) unit of WWTPs (Yu et al. 2018).

Nevertheless, the maximum and optimal potential of WAS in bioconversion is not obtained since the degradation efficiency ranges from 20 to 50% while the hydraulic retention times (HRTs) are very high, i.e., 20–50 days (Wang & Li 2016). The microbial cells and EPS are physical and chemical barriers to accessing the organic matter of WAS (Xiao et al. 2015). Humic substances comprise 15–18%, and lignocellulose constitutes 14–44% of organic matter in WAS, which may restrict AD (Wang et al. 2019). Moreover, the heterogeneity of WAS that contains bacteria, pathogens, collides, inorganic particles, heavy metals, and refractory organic matter limits its degradation (Wang et al. 2019).

To decrease the HRT and to increase the degradation efficiency of WAS in AD, different strategies are applied: replacing mesophilic with thermophilic AD, optimization of processing parameters of AD, co-digestion, pretreatment, enhancing microorganism community, providing facilitators medium to enhance direct interspecies electron transfer (DIET), etc. The former three factors are well studied and almost matured, while the latter strategies are emerging, and in-depth knowledge still needs to be improved for WAS.

The disintegration of the sludge matrix (microbial cell and EPS) is the main focus of pretreatment to release the entrapped organic matter (Ebenezer et al. 2015) so that the bioavailability of sludge is enhanced (Xiao et al. 2015). The release and solubilization of organic matter increase the soluble chemical oxygen demand (sCOD). However, the increase does not guarantee the bioavailability for bioconversion (Xiao et al. 2015) since other inhibitory organics such as humus, lignocellulose, triclocarban, norfloxacin, phenanthrene, diclofenac, and loxosone are present in WAS (Wang et al. 2019).

Different pretreatments are applied to disintegrate the cell wall to make the intracellular and extracellular substances available for biodegradation (Chang et al. 2011). They include physical (mechanical, thermal including conventional heating, steam explosion, hydrothermal, and irradiation, ultra-sonication, sludge thickening), chemical (alkali, acid, and other chemical reagents), biological (fungi, bacteria, enzymes), and their combination as hybrid pretreatment (Chi et al. 2011).

The biological sludge disintegration is usually performed by the introduction of a suitable microorganism, i.e., bacteria and fungus or externally produced enzymes to the AD process. Biological pretreatments are energy effective and environmentally friendly. However, the treatment takes longer time and selective microorganisms and enzymes are required for a specific feedstock (Table 1) (Rahmani et al. 2022).

Chemical pretreatments break down the cross-linkages within a substrate and alter the chemical composition. They have faster reaction rates and are energy efficient. However, its drawback is formation of inhibitory compounds when extreme doses are applied. Pretreating substrate with acid can achieve a higher degree of solubilization compared to other chemical reagents but the probability of inhibitory compound formation and corrosion of equipment limits its application. In contrast, alkali pretreatment forms fewer inhibitors and is more friendly to equipment corrosion (Table 1).

Physical pretreatments (mechanical) alter the structure of the substrate and physiochemical pretreatments (thermal, ultra-sonication) have both physical and chemical effects. The pretreatments increase the surface area via size reduction and solubilize the organic matter via thermal pretreatment. However, they are more energy intensive and industrial-scale application is limited.

The drawbacks of pretreatment can be avoided by hybrid pretreatments and may cause synergic effects as well. Combining physical and chemical pretreatment is the most commonly used hybrid pretreatment (Table 1). Therefore, this review attempts a comprehensive insight into thermochemical pretreatment, especially focusing on the alkaline, microwave (MW) and their combination, which is the cutting-edge trend in AD of WAS for enhanced biogas production.

Alkaline pretreatment is a chemical treatment method that uses a strong base, such as sodium hydroxide (NaOH), Calcium Hydroxide (Ca(OH)2) or Potassium hydroxide (KOH), to increase the pH of WAS to 10–12. Alkaline pretreatment is an effective way to disintegrate WAS and improve its biodegradability. It was also used in combination with other methods, such as thermal, ultrasound, etc. to enhance the efficiency of AD. Alkaline pretreatment has several advantages such as increased bioconversion potential of WAS, improved production of VFAs from WAS which are ideal substances for electricigens, and improved biogas production and conclusively, it can inhibit methane production at extreme alkaline pH, which can lead to enhanced power density and coulombic efficiency. The various methods used in literature for the alkaline pretreatment of WAS are as follows.

The reactive oxygen of CaO₂, O₃, and UV-irradiated TiO₂ disintegrates the sludge. H₂O₂ can be decomposed to release both the radicals (Wang et al. 2019). CaO₂ reaction with water produces heat, electromagnetic, alkaline, and oxidation effects (Wang & Li 2016). WAS pretreatment with CaO₂ cannot only degrade and solubilize typical organic matter but also persistent and inhibitory compounds such as humic substances, lignocellulose, phenolic endocrine disrupting compounds, triclocarban, norfloxacin, phenanthrene, diclofenac, and roxarsone (Wang & Li 2016).

The radicals in alkaline saponify uranic acid and acetyl ester while reacting with the free carboxylic group and naturalizing the acids formed during degradation (Wonglertarak & Wichitsathian 2014).

Alkaline pretreatment for WAS has varied depending on the concentrations of chemicals, pH adjustment, temperature, time, reagents, output required after pretreatment or bioconversion, and type and phase of the WAS used (see Tables 2–4).

Alkaline pretreatment does not need complex devices. Alkaline reagents are available in abundance; they have different types and high efficiencies; and they are cheap and easy to operate (Geng et al. 2020). Moreover, MFC (microbial fuel cells) can be operated better in alkaline environments (pH 8–10) (Geng et al. 2020). Alkaline pretreatments are applied to WAS for its disintegration to enhance methane production, VFA production, nutrient recovery, and stabilization. The disintegration and swell of the WAS can be tracked by the difference in volatile solid (VS), sCOD, BOD₂₀ (biochemical oxygen demand)/sCOD, turbidity, phosphorus solubilization, protein solubilization, NH₃-N concentrations, dewaterability (capillary suction time [CST], V_{5 min} [sludge filtrate volume after 5 min]), oxygen utilization, average particle size, specific surface area, change in structure and functional group, alteration in microorganism community, microorganism abundance, mathematical models, color, moisture content of thickened cake, lag phase, and methane or biogas production in AD (Ebenezer et al. 2015).

In contrast, Wonglertarak & Wichitsathian (2014) reported increased sCOD with increased pH pretreatment and concluded that the optimum pH was 8 for enhanced biogas production when testing the alkaline pretreatment of WAS at pH 8–14. This implies that pH 12 increases the damage of bounds in the sludge structure but does not necessarily increase the biodegradable organic matter. They further report that the BOD₂₀/COD ratio decreased for pH 10 and 11 and was limited at pH 12 (Wonglertarak & Wichitsathian 2014). However, pH-dependent disintegration of WAS is influenced by the source of the sludge. For example, Tyagi et al. (2014) reported an optimized pH of 12 for WAS from pulp and paper mill WWTP. It could be because of the lignocellulosic content, which is hard to solubilize. Moreover, the freshness of WAS, dewatered WAS, the EPS structure, the reagent used, the process of pretreatment (ambient temperature, mixing, time, stabilization period, etc.), and AD processing parameters may also affect the disintegration at a particular pH and subsequently the biogas production.

Generally, a higher dose of alkaline reagent causes more disintegration (Maryam et al. 2021). The untreated WAS had sCOD of 50 mg/L; at pH 10, 11, 12, and 12.5, the concentration increased 20, 40, 46, and 66 times compared to the untreated sample (Doğan & Sanin 2009). However, some cations, such as calcium, can re-flocculate WAS, and most of the oxidizable compounds can react back (Saktaywin et al. 2005). Hence, there is a decrease in the sCOD yet again when the pretreatment dose is high enough (Wang et al. 2019). Moreover, the high dose of alkaline may need neutralization of the sludge again, usually by adding acids, which increases the sludge's salinity and may restrict microbial growth. The methane production was 48% higher when WAS was pretreated with CaO₂ at the dose of 0.14 g/g VSS, and increasing the dose restricted the methane production mainly due to free radicals rather than calcium (Wang et al. 2019). High doses of alkaline (especially strong alkali with hydroxyl group) can cause damage to DNA components of microbes during AD or anaerobic fermentation (Wang et al. 2019). However, some minerals are essential for the growth and enzyme production to hydrolyze the organic matter. Calcium is vital for the stable production of enzymes, such as neutral proteases, and the growth and abundance of Firmicutes (hydrolyzing bacteria), anaerolineaceae (acidogenic bacteria), and Synergistaceae (acetoclastic methanogens) (Wang et al. 2019). In contrast, decreased heavy metal concentration in the digestate of WAS may permit its land application (Wang et al. 2019). Therefore, an optimum dose of each reagent to pretreat various substrates is required. Alkaline pretreatment (NaOH and KOH) declines dewaterability since the EPS bounds are damaged (Yu et al. 2018), and fine particles are increased (Wang & Li 2016). Waste active sludge was pretreated with NaOH (0.05–1 mol/L) and Ca(OH)₂ (0.02–0.5 mol/L) for 24 h at varying temperatures of 0–40 °C (Li et al. 2008). The alkaline concentration of 16% (0.05 mol/L) was found to be the optimum pretreatment condition to disintegrate the sludge although sCOD was increased with the increased NaOH concentration. Therefore, the dewaterability decreased in lower (<0.2 mol/L) concentrations and increased with further dose addition. It is because the higher alkaline pretreatment makes the water that is confined by floc and cell structure be released (Li et al. 2008). When VFA production is the main objective of anaerobic fermentation, dewaterability is crucial to separate VFA from the digestate (Yu et al. 2018). Moreover, in AD, improved dewaterability would be cost-effective for digestate dewatering, transportation, and landfilling (Wang et al. 2019).

The alkaline pretreatment modifies the structural and functional group of the inhibitory compound and makes them biodegradable (Wang et al. 2019). The inhibitory compounds, such as humic acid, are also degraded and have been bio-converted to biogas or VFA (Wang et al. 2019). Otherwise, the persistent organic compounds would find their way to the ecological environment and human health (Wang et al. 2019).

During the damage to the cell membrane, nutrients are ruptured along the organic matter (Yu et al. 2018). Nutrients such as phosphorous can be recovered during alkaline pretreatment. WAS was pretreated with ozone, MW, sulfuric acid, and sodium hydroxide at varying pretreatment parameters for solubilization and nutrient release (Cosgun & Semerci 2019). The study claimed that ozonation had 90 and 20% increase in PO₄-P and COD release, respectively. The increases were 153 and 24% for alkaline pretreatment (pH 10) and 580 and 48% for acid-ozonation- MW pretreatment, respectively. This implies that the proportion of nutrient release is even more than the COD. In another study by Maryam et al. (2021) WAS was pretreated with 2, 4, 6, and 8% NaOH (TS base) at 37 °C for 3 h. Furthermore, the alkaline pretreatment was combined with photocatalytic pretreatment using 0.3–0.6 g/L TiO₂ for 3 h at 37 °C. Maximum disintegration happened at 8% NaOH concentration. It was 11.3% in terms of sCOD and 8.4% in terms of VS. The disintegration, in terms of TAN (total ammonia nitrogen), was optimum at 4% NaOH dose, which was 713 mg/kg of dewatered WAS (Maryam et al. 2021). The study denotes that a mild concentration of alkaline (4% NaOH) pretreatment is sufficient for nutrient release when compared to organic matter solubilization.

The sCOD was 2–33.3% when WAS was pretreated at ambient temperature for 24 h (240 revolutions per minute, rpm) with NaOH to adjust the pH from 8 to near 14 (Wonglertarak & Wichitsathian 2014). The thermophilic AD at optimum pH (8) had 43% VS removal and a 7.64% increase in biogas production compared to the same AD with untreated WAS. The biogas increase was 18.6% in AD of ambient temperature compared to the same AD with untreated WAS (Wonglertarak & Wichitsathian 2014). Thermophilic AD of already hydrolyzed WAS with NaOH may have increased the VFA concentration enough to put the AD process under stressed. Therefore, AD at ambient temperature had higher biogas production. However, disintegration was the highest in hybrid pretreatment with increased NaOH dose and 0.4 g/L TiO₂, which was 37% in terms of sCOD. The hybrid pretreatment (4% NaOH and 0.5 g/L TiO₂) was the optimum since the methane yield increased by 71% (462 mL/gVS) and the methane daily rate was the highest (51 mL/gVS) (Maryam et al. 2021). The results suggest that mild concentrations of hybrid pretreatments are promising for increased bio-methane production.

When the alkaline dose is increased, the WAS particles usually decrease in size until disintegration is inhibited, for instance, by re-flocculation. The pH adjustment of WAS by NaOH to 8–12 decreased the average particle size from 101.2 to 82.4 μm and increased the specific surface area from 0.114 to 0.130 m²/g (Xiao et al. 2015) due to disintegration. When the disintegration is higher, the relative turbidity is higher because of the small particles suspended in the sludge (Wang et al. 2019). However, a high dose may reduce the turbidity again. The disintegration of WAS was restored while pretreating by Ca(OH)₂ at a dose above 0.2 mol/L; hence, the fine particles decreased. It was reported that d₉₀ increased from 115.7 to 135.5 μm (Li et al. 2008).

The microbial activity is affected by alkaline pretreatment. The relative activities of acidogenesis, acetogenesis, hydrogenotrophic methanogenesis, and acetoclastic methanogenesis declined by 1.7, 16.7, 41.7, and 30.2% when sludge was pretreated with 0.14 g/gVSS CaO₂, respectively, compared to untreated sludge (Wang et al. 2019). The decline was even more (11.3, 20.4, 66.7, and 57.4%, respectively) at 0.26 g/g VSS. The microbial activity at the lower dose did not decline the methane production but the poor microbial activity in the higher dose did decline the methane production. The results infer that CaO₂ had deteriorating impacts on the microbial community responsible for methane production.

In MW technology, electromagnetic radiation interacts with dipolar molecules such as water, ammonia, and hydrogen fluoride (or samples having dipolar molecules inside) to oscillate them forth and back, making fiction to produce heat (Yu et al. 2010). The mechanism of MW irradiation has both thermal and a-thermal effects (Yu et al. 2010). Thermal effects are prompted by the cell materials' denaturation (breaking of weak linkages or bonds). In contrast, a-thermal effects include electroporation, dielectric cell membrane rupture, magnetic field coupling, and selective heating (Wang & Li 2016).

Several MW pretreatments of WAS with varying time, and temperature have been performed (see Table 3). The pretreatment temperature depends on how much MW power can be absorbed. Although an MW unit is labeled for output power, many factors are involved in limiting the absorbance of the MW power. A higher volume of sample can absorb higher energy on account of power consumption proportionally (Houšová and Hoke 2002). The type of sample and, hence, the dipolar substance in the sample decide the degree of MW power absorbance (Eskicioglu et al. 2007a). WAS of different sources and TS% may have different power absorbance and, hence, different temperatures and disintegration. For instance, MW pretreatment of WAS up to 96 °C with different TS% of 1.4 and 5.4 resulted in optimum pretreatment in the higher TS concentration (Wang & Li 2016).

MW pretreatment is preferred at an optimum time and temperature as high temperature for an extended time can deactivate the enzymes, make the nitrogenous content persistent to biodegradation, and may increase ammonia concentration (Eskicioglu et al. 2007a). Moreover, it may alter the content of individual fractions of VFA. The propionate acid yield was increased when the MW heating time was increased (Jiang et al. 2021). MW pretreatment can damage the cell membrane to release intracellular materials (Figure 1) because cell membranes are permeable lipid bilayers that can absorb microwaves (Eskicioglu et al. 2007b). Moreover, cell death in MW pretreatment occurs at a temperature lower than the cell thermal death point.

The DD and effects can be analyzed via supernatant turbidity, sCOD, VS solubilization, SS reduction, lag phase reduction (Rani et al. 2013), EPS content, dewaterability (CST, specific filtration resistance, settling velocity (SV)), BOD, nutrients (N, P) solubilization, inorganic nitrogen content, structure, and function of bacterial communities (Yu et al. 2010).

MW application has numerous advantages over conventional heating. The desired temperature can be reached more rapidly; the potential for hazardous emission is minimized (Ebenezer et al. 2015), and sanitized sludge is produced (destroying pathogens and fecal coliforms). MW-pretreated WAS at 65 °C can be used in mesophilic CSTR to have class A digestate (Doğan & Sanin 2009). Moreover, MW pretreatment improves sludge dewatering (settleability), (Eskicioglu et al. 2007a), reduces foaming, produces more stable sludge, induces low heating loss (Eskicioglu et al. 2007b), improves solubilization of sludge, increases methane yield (Ebenezer et al. 2015).

On the other hand, the MW pretreatment requires high energy consumption and high energy costs (Ebenezer et al. 2015). However, the process consumes less energy considering its prompt heat up (Yang et al. 2013), especially when compared to sonication (based on unit mass) (Rani et al. 2013). The MW pretreatment of WAS at temperatures ranging from 40 to 96 °C resulted in an optimum temperature of 85 °C (14 MJ/kg TS), which increased sCOD by 21% (Ebenezer et al. 2015). At increasing MW energy beyond 45 MJ/kg TS, the sCOD decreased because of the evaporation of organics at open MW pretreatment.

Energy, temperature, contact time (ramping, holding, and cooling times), power, and total solid concentration are essential parameters in pretreating WAS by MW (Yu et al. 2010; Rahmani et al. 2022). Different analysis of MW-pretreated WAS determined that temperature, sludge concentration, and intensity were essential for disintegration (Eskicioglu et al. 2007a). The WAS from dairy was subjected to MW pretreatment of 900 W at different intensities of 50–90% and different pretreatment times of 0–15 min. It was concluded that 12 min and an intensity of 70% are the optimum pretreatment conditions for enhanced sCOD, VS removal, and biogas production (Rani et al. 2013). Moreover, they concluded that the sCOD increased with the increase in contact time and that the intensity of MW is crucial rather than the contact time. The sCOD of raw WAS was 905 mg/L. The increase in MW intensity (50–90%) increased the sCOD to 5670–7040 mg/L (Rani et al. 2013).

Thermal pretreatment via MW at 900 W to heat WAS to 60–70 °C increased methane yield by 35% (Wang et al. 2019). A higher DD of up to 65% was obtained after alkaline (pH 8–12) and MW (9.6–48 MJ/kg TS) pretreatments of WAS (Yang et al. 2013). MW-pretreated WAS at 50–96 °C (50% intensity of 1250 W) resulted in the highest methane yield at 96 °C (Eskicioglu et al. 2007a), 15 and 20% biogas increase in 1.4 and 5.4% TS, respectively. At the same pretreatment, the sCOD increases were 3.6 and 3.2 times the control at 5.4 and 1.4% sludge concentrations, respectively (Eskicioglu et al. 2007a). The results denote that WAS of higher TS content subjected to MW pretreatment produces more biogas compared to WAS of lesser TS content due to higher solubilization. The MW pretreatment of WAS (1,841 kJ/L, 12 min) increased the sCOD by 18.6%, decreased SS by 14%, and enhanced biogas production by 35% in the batch assay. The same pretreatment resulted in 67, 64, and 57% enhancement in SS reduction, VS removal, and biogas yield, respectively, in sCSTR at HRT of 15 days and OLR of 0.6 g SS/L day (Rani et al. 2013).

MW pretreatment ruptures the microbial cell and EPS matrix and releases carbohydrates, protein, and other nutrients. The carbohydrate was increased to 464 mg/L when MW energy of 2914 kJ/L was applied during WAS pretreatment (Rani et al. 2013). At the same energy consumption, the protein increased to 1846 mg/L, which confirms the effective disintegration of WAS with MW (Rani et al. 2013). At 900 W and 140 s MW pretreatment of WAS, the protein and polysaccharides increases were 297 and 654%, respectively (Yu et al. 2010). MW pretreatment of WAS (5.4% TS) at 96 °C and 625 W has increased the solubilization of proteins, sugar, and COD by 4.2, 4.5, and 3.6 times, respectively (Eskicioglu et al. 2007b). The protein solubilization was 2.4 and 2.2 times the control sample at 50 and 75 °C, respectively, at the same power rate and sludge concentrations, which indicates that power rate or temperature is a decisive factor for disintegration (Eskicioglu et al. 2007b).

Due to colloidal matters induced by the disintegration, the turbidity of pretreated sludge depends on contact time and energy consumption (Yu et al. 2010). At 100 s and 900 W, the turbidity of sludge was at an extreme level. The d₉₀ of the untreated sample was 83.7 μm; at 900 W, the d₉₀ increased to 143.4 μm (at 60 s) and then decreased to about 60 μm (at 140 s). The midway increase in d₉₀ could be because of the nonstop friction that destroyed Zeta potential. The magnitude of ultra-colloidal particles (1–100 μm) defines the filtration characteristics. A smaller fraction of ultra-colloidal particles is beneficial for dewatering; however, extreme MW pretreatment may increase the fraction and, hence, deteriorate the dewaterability of the sludge (Yu et al. 2010). The MW irradiation of 50–96 °C had a 41% improved dewaterability compared to the control (Eskicioglu et al. 2007a). The SV of MW-pretreated sludge was high initially (up to 1 h), but decreased later (Yu et al. 2010). The SV of untreated sludge was about 39.6 mm/h; after MW pretreatment with 500, 750, and 900 W, their peak values almost increased to 44–45 mm/h. The results denote that MW pretreatment can improve the dewaterability at optimized intensity and temperature of MW irradiation but can deteriorate the dewaterability at extreme temperatures and intensities due to fine and colloidal particles.

Temperature is considered the most influential parameter for solubilization and enhancement of biogas production (Valo et al. 2004). The pretreatment of WAS at 170 °C with a pH of 10 resulted in increases in sCOD, TS degradation, and biogas production of 71, 59, and 54%, respectively (Valo et al. 2004). However, pretreatment at high temperatures (>180 °C) can form recalcitrant intermediate compounds, which may be unfavorable for the subsequent processes (Kim et al. 2013). MW pretreatment follows the disintegration of floc, cell wall, membrane rupturing, hydrolysis of organics, and Maillard reaction (Chi et al. 2011) (Figure 1). The Maillard reaction occurs when the pretreatment temperature is extreme (>210 °C), and recalcitrant compounds, which are called ‘burnt sugar,’ are formed at this stage (Chi et al. 2011).

Moreover, a high dose of alkaline reagent may inhibit a stable biodegradation process and cost too much. Combining both pretreatments at typical doses is not cost-effective since the net energy and net cost would be negative (Ebenezer et al. 2015). Hybrid pretreatments are often employed to pay off both alkali and energy (Kim et al. 2013). Thermal pretreatment (130–170 °C) demands more energy (130–190 kWh/m³) than thermal-alkaline pretreatment (40–50 kWh/m³) when 60–70 °C temperature is applied (Toutian et al. 2020). A low dose of alkali acts as an add-on in thermal pretreatment (Li et al. 2008). In addition, the low-temperature heat can be supplied from the waste heat of a CHP plant, making the pretreatment further cost-effective (Toutian et al. 2020). Reduced time in MW pretreatment can avoid enzyme deactivation, formation of refractory compounds, and high concentrations of ammonia (Alqaralleh et al. 2019).

In hybrid pretreatment, alkaline pretreatment is usually applied first so that the high pH effectively disintegrates WAS before applying thermal pretreatment for further solubilization. In case the thermal pretreatment is first, the subsequent alkaline pretreatment will require more alkali to increase the pH (Toutian et al. 2020). WAS was subjected to hybrid pretreatment of CaO₂ after and before MW pretreatment (Wang & Li 2016). CaO₂ dose of 0.1 g/g VSS before and after MW pretreatment increased methane yield by 80.2 and by 70.1%, respectively (Wang & Li 2016) indicating alkali pretreatment before MW pretreatment is more favorable (see Table 4).

The hybrid pretreatment (thermo-alkaline) can increase the stability of anaerobic digester for higher organic loading rates application (Toutian et al. 2020) and improve disintegration with synergic effects (Geng et al. 2020). During the hybrid pretreatment of WAS, the highest DD was 65.9% at 38.4 MJ/kg TS and pH 11, but the VFA production was the highest at 28.8 MJ/kg TS and pH 11 (Yang et al. 2013). Moreover, the VFA production time was shortened to 25%, and VFA production increased by 100% compared to only alkaline pretreatment (Yang et al. 2013), resulting in increased readily biodegradable COD (rbCOD). Hybrid pretreatment of WAS with CaO₂ and MW increased solubilization (46% COD solubilization) and CH₄ production (increased by 80.2%) synergistically since MW enabled more hydroxyl group development from CaO₂ (Wang & Li 2016). The additive solubilization of NaOH and MW pretreatments was 20% less than the hybrid pretreatment, confirming the combined pretreatment's synergic effect (Wang & Li 2016). Pretreatment of WAS disintegrates the sludge matrix and releases the organic matter including carbohydrates, protein and biomass, nutrients including nitrogen and phosphorous, salts and metals used when treating wastewater, and bound water (Figure 1).

The dewaterability improvement in alkaline pretreatment could be worse than in thermal and thermal-alkaline pretreatments. Improved dewaterability decreases the disposal cost of sludge significantly. Hence, dewaterability can greatly affect the techno-economic assessments of the pretreatments (Toutian et al. 2020). Alkaline pretreatment of WAS has reduced the dewaterability (CST) from 72 s in untreated WAS to 1,785s at pH 12.5; however, when alkaline pretreatment (pH 12.5) was combined with MW pretreatment, the dewaterability decreased to 440 s, showing 93% improvement (Doğan & Sanin 2009). The disintegration causes an increase in fine particles and bound water. Therefore, the CST time may increase because the floc structure is disrupted by pretreatment and AD (Jang & Ahn 2013). The filterability of the sludge and CST (dewaterability) were improved after hybrid pretreatment (Jang & Ahn 2013). However, Doğan & Sanin (2009) report that there was not a considerable affect in CST values when sludge was treated with alkaline, MW and their combination but report that the highest turbidity was achieved with hybrid pretreatment. The dewaterability of digested WAS in control reactor was 138 s for CST while the dewaterability of digested WAS in hybrid pretreated reactor was 135 s for CST. Electrolysis and alkaline (5 V + pH 9.2) pretreatment improved the dewaterability, but in high alkali pretreatment, the TS% of the dewatered cake was increased slightly (Zhen et al. 2014). The CST of MW-pretreated sludge decreased slightly after a short time, but the CST exceeded the raw CST time when the pretreatment time was extended (Chang et al. 2011). However, alkaline-assisted MW pretreatment of sludge improved settleability (Chi et al. 2011).

Hybrid pretreatment DD is higher than individual and additive MW pretreatment and alkali pretreatment (see Table 4). MW pretreatment is superior than alkaline pretreatment in carbohydrate solubilization. MW pretreatment of 600 W for 16 min solubilized more carbohydrates than in pH 12 pretreatments (Doğan & Sanin 2009). However, MW-assisted pretreatment of sludge at pH 12 resulted in the highest carbohydrate solubilization (Doğan & Sanin 2009). Hybrid pretreatment of acid + ozonation + MW resulted in 48% COD release, while the COD releases were 28% for acid, and 25.2% for MW pretreatment (Cosgun & Semerci 2019). The hybrid pretreatment solubilized 51.5% of VSS at the ozone rate of 225 mg/g MLSS (Cosgun & Semerci 2019). The thermo-alkaline pretreatment of WAS has increased the sCOD to 64% of tCOD (21.3 g/L) (Yu et al. 2018), which is among the highest values of solubilization. sCOD increased significantly by 624% (Toutian et al. 2020) when hybrid pretreatment at 170 °C with pH 10 was performed. The raw waste sCOD was 2.74 g/L and increased to 19.84 g/L (Toutian et al. 2020), comparable to sCOD reported by Yu et al. (2018), although alkalinity and pretreatment temperatures in the two studies were different. Hybrid pretreatment of WAS with STPP and MW resulted in a 28% increase in sCOD (Ebenezer et al. 2015). There was an 18 times increased sCOD compared to control when WAS was pretreated in hybrid with 20 meq NaOH/L and MW 1000 (135 °C) for 10 min (Jang & Ahn 2013). For MW pretreatment alone, the solubilization was 8–17.5%, and for alkaline pretreatment, it was 14.3%. Therefore, the solubilization values indicate that the hybrid pretreatment is inducing a synergic effect (Jang & Ahn 2013).

Conventional heating uses 5.2 times more energy to reach the same temperature; therefore, using MW is an obvious benefit in terms of both cost and time. Moreover, the COD solubilization was 7 and 8.5% (1.3 g/L) for conventional heating and MW pretreatment, respectively, while the initial sCOD was 51 mg/L. The MW combined with alkali (pH 12, 30 min) improved the COD solubilization to 46%, while only alkali pretreatment induced 18% COD solubilization (Chang et al. 2011). The minimum specific energy for the disintegration of microbial cells is reported to be 1 MJ/kg SS (Chang et al. 2011). Moreover, the sCOD increase is influenced by the TS% of WAS. The higher solid concentration can result in higher sCOD (Chang et al. 2011; Shi et al. 2015). However, different concentrations of sludge pretreated in hybrid (MW and alkali) had a minor effect on solubilization (Jiang et al. 2021). Alkaline pretreatment (pH 10) combined with MW pretreatment of WAS and FOG (fats, oils, grease) at (95–175 °C) for 1 min showed the maximum solubilization of 68.2% with 20% FOG at 175 °C (Alqaralleh et al. 2019). MW pretreatment (70–700 W, up to boiling temperature, 0–6 min), alkaline pretreatment (0–7.55 g/L NaOH, 0–48 h), and water elutriation at WAS concentration of 6–16 g/L was carried out to study the effect of solubilization (Jiang et al. 2021). MW pretreatment alone induced 12.4% solubilization, while MW with alkaline had 54.7% solubilization. The three pretreatment together had a slight increase of 55.6% solubilization. However, it had a maximum of 81.9 and 73.7% of total phosphate (TP) and total nitrogen (TN) release, respectively (Jiang et al. 2021). The alkaline pretreatment is more effective than MW pretreatment, while the hybrid pretreatment is the most effective and has a synergic effect. The COD solubilization could be up to 68%, while the average is about 50%. The sCOD improvement by the pretreatments is up to a hundredfold compared to raw WAS because the sCOD of the raw waste is usually very low, and most of them are dewatered.

Hybrid pretreatment of WAS is used not only for the disintegration in AD to produce biogas and VFA but also for nutrient release and recovery (Cosgun & Semerci 2019). Ozonation, pH 10, and hybrid (acid + ozonation + MW) pretreatments caused 89.5%, 152.6, and 579% increase in phosphorous release, respectively (Cosgun & Semerci 2019), while the initial phosphorous content was 1.9 mg/g MLSS. Hybrid pretreatment of alkali and electrolysis released the protein and polysaccharide of biopolymers. Electrolysis disintegrates the microbial cell with ohmic heating, electrophoresis, and electro-osmosis mechanisms (Zhen et al. 2014). NaOH was superior when comparing the thermo-NaOH, themro-CaO₂, and thermo-mixed alkali pretreatments for nutrient release. It released 1.6 g/L and 401 mg/L TN and total phosphorous, respectively (Yu et al. 2018). For the thermo-NaOH, thermo-CaO₂ pretreatments, the releases were, respectively, (1 g/L, 37 mg/L) and (1.1 g/L, 83 mg/L) for the TN and total phosphorous (Yu et al. 2018).

Hybrid pretreatment of MW and alkaline improves the methane yield, daily methane rate, and content, especially in sCSTR operation mode. MW pretreatment improved the methane production by 11.3%, whereas the hybrid pretreatment improved the production by 18.9% (Doğan & Sanin 2009) for the batch scale. The increase in methane yield was 55% at the steady state of 15 days HRT. Methane yield increased by 80.2% when hybrid pretreatment of MW and CaCO₂ was applied to WAS, while the CH₄ content increased by 25.4% compared to control (Wang & Li 2016).

The hybrid pretreatment of electrolysis and alkaline was tried for the disintegration to produce biogas (Zhen et al. 2014). After investigating voltages of 5–20 and pH values of 9.2–12, the optimum pretreatment values were found to be 5 V and pH 9.2. The mesophilic AD of the pretreated sludge had a 20.3% increase in methane yield compared to control (Zhen et al. 2014). Hybrid pretreated WAS used as feed for sCSTR at varying HRTs (15–5 days) indicated a biogas increase in the range from 103 to 205% range, while the methane content was between 63 and 69% (Jang and Ahn, 2013). The hybrid pretreatment at an optimum condition of 170 °C with 0.05 g/g SS NaOH for 1 min resulted in a 27% increase in methane yield (Chi et al. 2011). There was a 4.6-fold increase in methane yield of hybrid pretreatment compared to control, while it was 1.9 and 3.4 times for MW and alkali pretreated WAS (Jiang et al. 2018). The results from the above different studies confirm that the hybrid pretreatments combining alkali with voltage, conventional heating, and MW irradiation have induced increased and synergic biogas yield. The sludge was pretreated by MW and alkali, which resulted in maximum methane production, up to five times that of untreated sludge (Jiang et al. 2018). The improvement was because NaOH reduced the inhibition induced by MW and that led to improved sludge solubilization (Jiang et al. 2018). However, the biogas yields are not at the pace of the DD. Moreover, the methane content improvement was intangible in most of the studies.

Hybrid pretreatments had synergic effects on COD and VS removal in the digestate of AD. The hybrid pretreatment resulted in 48.3% VSS removal efficiency in the batch assay and 35.4% in sCSTR. Moreover, the tCOD removal in sCSTR was 30.3% more compared to the control (Doğan & Sanin 2009). The hybrid pretreatment of CaO₂ and MW resulted in 55.6% VSS reduction after mesophilic AD at batch scale (Wang & Li 2016).

The removal efficiency of VS depends on the HRT. In hybrid pretreatment (alkaline + MW), the removal efficiency decreased while HRT decreased; however, the relative improvement increased (Jang & Ahn 2013). This is because the lower HRT is insufficient for flocculated WAS but is relatively good for disintegrated sludge to bio-convert the solubilized matters. It is concluded that the maximum VS removal of MW + NaOH pretreated WAS in AD is 50–60% (Jang & Ahn 2013). After the hybrid pretreatment of WAS and thermophilic (55 °C) AD of 30 days HRT in sCSTR, 18% tCOD was removed, while the increase for VS removal and methane yield were 28 and 17%, respectively (Chi et al. 2011).

Microbial activity can be measured indirectly by enzyme activities (Wang & Li 2016) since hydrolysis of protein and carbohydrates is carried out by protease and α-amylase enzymes. Wang & Li (2016) reported increase of protease and α-amylase activity after hybrid pretreatment (MW + alkali) of WAS. The protease activity was 35 and 50 μmol/min gVSS in control and hybrid pretreated samples, respectively, on day 0.5. The α-amylase activity was 32.4 and 53 μmol/min gVSS in control and hybrid pretreated samples, respectively, on day 0.5. The increase of the enzymes at hybrid pretreatment indicates that hydrolytic enzymes of EPS are released into the pretreated sludge (Wang & Li 2016); therefore, the hydrolysis would improve further.

Hybrid pretreatment has effects on microbial interaction and microbial diversity. The pretreatment facilitates improved microbial synergy. For example, hydrogenotrophic and acetoclastic methanogen growth were improved when chemical-thermal hybrid pretreatment was applied on WAS (Wang et al. 2016). In contrast, pretreating WAS with CaO₂ alone deteriorated the methanogens abundancy. However, in hybrid pretreatment, MW pretreatment alleviated the inhibitory effect of CaO₂ on methanogens. Moreover, the activities of methanogenesis, hydrolytic, and acid-forming enzymes were improved. The relative abundance of acetate-utilizing methanogen (Methanosaeta sp.) improved from 26.8 to 30.7% while the improvement for H₂/CO₂-utilizing methanogen (Methanospirillum sp.) was from 29.9 to 44.2% (Wang et al. 2016).

Pretreatments can disrupt the trapped bacteria in the sludge matrix and facilitate the release of more organic matter from the biomass and sanitization of the sludge. In electrical-alkali hybrid pretreatment of WAS, rod-shaped bacteria were not detected, which indicates the disintegration of EPS and microbial cells. However, they were detected when only electrolysis pretreatment was performed (Zhen et al. 2014). Alkaline pretreatment of WAS at 1.5 g/L NaOH reduced the fecal coliform number less than 1,000 Most probable number (MPN)/100 mL with a 99.99% removal efficiency when compared to control (Zhen et al. 2014).

Pretreatment of feedstock that causes hydrolysis a rate-limiting step of AD is a promising approach to enhance the biodegradation. However, for readily biodegradable, soluble organics or already hydrolyzed organics where acetogenesis and methanogenesis become the rate-limiting steps, improving the interspecies electron transfer (IET) via abiotic materials is a promising strategy (Wang et al. 2021). Interspecies electron transfer is not facilitated by the conventional soluble redox medium of H₂, formate, or protein of death cell, which are called indirect IET (IIET), nor by the biological mechanism such as membrane-related cytochromes and conductive pili, which are called biological direct IET (DIET), at a rate to achieve stable and fast AD. In IIET, the electron shuttle is driven by diffusion, which makes methanogenesis a limited step in AD (Nguyen et al. 2021). The IIET and biological DIET may happen in parallel, but IIET via H₂ and formate is the primary route (Baek et al. 2018). Established and fast IET is crucial to keep balance, for example, between VFA oxidizing bacteria and hydrogenotrophic methanogens for enhanced and efficient bio-methanation (Baek et al. 2018; Gahlot et al. 2020) because they are in a syntrophic relationship. The syntrophic microorganism cannot break down the complex organic matter individually. Therefore, their syntrophic relationship helps the metabolic by-products of a microorganism serve as a substrate of the other microorganism.

Efficient valorization of wastes via AD is limited owing to the discrepancy in the syntrophic relationship of inter-microbial matrix. Although there are several reviews on the pretreatment of organic wastes but few dedicated reviews on the amendment of AD with DIET materials for enhanced bio-methanation are available. This review aims to provide insight into improving AD acetogenesis and methanogenesis steps via the supplementation of DIET materials. The significance and mechanism of DIET materials and the various materials employed as DIET accelerants are investigated. In addition, the compatibility of DIET materials for AD and their application in recovering the stability of AD with high VFA, ammonia, and organic loading concentrations are studied in detail. Eventually, the effects of DIET on enhanced biogas production and microbial community and diversity are presented.

The structure of DIET-enhancing materials affects the degree of improvement in AD. The structures of tantalum oxide and niobium oxide are hexagonal crystalline, and the structures of tungsten oxide and hafnium oxide are monoclinic crystalline (Yun et al. 2021). These transitional metal oxides were used to investigate their effect on the biodegradation of dairy manure and sewage sludge (Yun et al. 2021). The four metal oxides (tantalum oxide, niobium oxide, hafnium oxide, and tungsten oxide) of different nano-sized were added to AD at 30 mg/L. All oxides enhanced biodegradation in terms of biogas production, COD, and VS removal due to enhanced DIET effect by providing superior crystal structure, electrical conductivity, and electron exchange capacity. However, tantalum oxide was better than the other oxides. It increased biogas yield, VS reduction, and COD removal by 88.9, 60.9, and 70.5, respectively (Yun et al. 2021). Therefore, crystal structure is vital in DIET enhancement. For example, it is observed that 100 mg/L Ni (Fe₂NiO₄) could enhance biogas production while 100 mg/L Ni (Fe₄NiO₄Zn) inhibited the AD process which is due to the crystal structure and concentration of release metal (Yun et al. 2021). However, the price for the provision of metal oxide could not be compensated by increased biogas production except for niobium oxide when the capacity of the AD plant is 10,000 m³ (Yun et al. 2021).

Nano-sized graphene and AC of different doses were used to study their effect on AD of ethanol at mesophilic temperature for 12 days (Lin et al. 2017). It was revealed that AC at 20 g/L and graphene at 1 g/L are the optimum concentrations to enhance AD. With graphene addition, bio-methane production was 659 mL/g, corresponding to a 25% increase compared to control, while AC addition induced a 12.8% increase. The daily methane rate increased by 19.5 and 13.7% for graphene and AC, respectively (Lin et al. 2017). The biodegradation enhancement with low graphene dose is due to its high electrical conductivity and specific surface area (small-size particles) compared to AC. Moreover, its higher dose induced cytotoxicity, which was not appropriate for biodegradation (Lin et al. 2017).

The DIET enhancement is directly related to the surface area used for DIET. The larger specific surface area provides more considerable microbial colonization than smaller surface areas, enhancing functional microorganisms and stimulating DIET (Xu et al. 2020). Many microorganisms can transfer and receive electrons for DIET via abiotic materials; hence, they are biocompatible (Gahlot et al. 2020). The addition of Carbon-based materials (CMs) to AD has led to the enrichment of bacteria such as Geobacter sp., Thauera sp., Gordonia sp., Syntrophomonas, Clostridium, Spirochaeta, and Bacteroides (Nguyen et al. 2021) and their syntrophic methanogenic partners. Granular activated carbon was used to enhance DIET for improved biogas production in AD of WAS (Yang et al. 2017). AD of WAS was amended with the biochar dose ranging from 0 to 5 g/L of 2 mm particle size. The digestion resulted in a 17.4% increase in methane yield and 6.1% more sludge reduction than control. The increase was mainly due to the enrichment of Geobacter, which was facilitated by 5 g/L granular AC (Yang et al. 2017). The enhancement of Geobacter is limited when AD of the complex substrate is amended with CMs; however, using soluble substrates such as ethanol has proved to enrich the bacteria. Therefore, it suggests that improvement in biogas production includes but is not limited to microorganism enrichment. In contrast, there could be a positive association between the addition of CMs and the enrichment of Methanosarcina (Wang et al. 2021).

The addition of Granular activated carbon (GAC) and zeolite at the rate of 1 g/L revealed that GAC contributed to the promotion of methanogenesis, but zeolite did not contribute since it was nonconductive (Zhang et al. 2017). It is further concluded that the enhancement of methanogenesis by GAC is not associated with the increase of microbial biomass but rather with the enhancement of DIET (Zhang et al. 2017). Zhang et al. (2017) observed that microorganisms were mostly in a suspended phase instead of the surface of the GAC or zeolite carriers. In addition, GAC has played the role of c-type cytochrome in DIET promotion since the cytochrome concentration of the medium culture was decreased.

Sulfidated microscale Zero-valent iron (ZVI) at different doses of 2–10 g/L is added to the AD of WAS and food waste (Chen et al. 2020). After mesophilic AD for 15 days, 10 g/L S-mZVI was the optimum dose. It increased methane yield by 1.33 times (264.8 mL/g VS) (Chen et al. 2020). The enhancement was due to the stimulation of DIET since the DIET-related microorganisms such as Syntrophomonas, Methanosarcina, and Methanobacterium increased with the addition of S-mZVI (Chen et al. 2020).

GAC (25 g/L) addition to glucose wastewater was investigated in UASB with 2 g/L/day COD OLR and 6 h HRT at 20 °C for 170 days (Guo et al. 2020). It increased volatile suspended solids (VSS) from 20 to 26.6 g VSS/reactor. It removed COD 5.8 times (1.66 g COD/L/day) more than control. RNA-based analysis showed a high difference in microorganism community between reactors, while DNA-based analysis showed little difference in microorganism community. GAC biofilm was enriched with bacteria (Geobacter [22%], Syntrophus, Desulfovibrio, and Blvii28) and archaea (Methanospirillum, Methanosaeta [49%]) which are electroactive DIET-related microorganisms (Guo et al. 2020).

The AD of ethanol amended with graphene changed the dominant bacterial group to Geobacter, Pseudomonas, and Levilinea, while Methanosaeta, Methanobacterium, and Methanoliea were the most dominant in the archaeal communities (Lin et al. 2017). Geobacter is the most effective microorganism in oxidizing organic matter, Levilinea is a fermentative bacterium, and Pseudomonas is VFA oxidizing bacteria that produces an electric current in microbial fuel cell (MFC) (Lin et al. 2017).

Citrus peel waste (CPW) was fermented to produce H₂ and VFA to obtain value-added products. Subsequently, they were used in AD to produce CH₄. 15 g/L CPW produced 13.3 mmol H₂/L and 1.34 g/L acetic acid at pH 8.5 and 30 °C for 22 h. In AD, after 29 days at 30 °C and pH 7, the methane production was 50.2 mmol/L. Escherichia and Clostridium, cellulolytic enzyme producers, were the main bacteria in the fermentation stage, while Methanoplasma and Methanosarcina were mainly involved in methane production (Camargo et al. 2021).

AD of WAS is limited due to microbial cells and EPS in the sludge matrix. Thermal, alkaline, and hybrid pretreatment of WAS are carried out to disintegrate the sludge and solubilize the organic matters for enhanced biogas production and valorization of the waste. Higher doses of pretreatments increase solubilization but do not promise rbCOD. However, hybrid pretreatments (thermal-alkaline) have synergic effects compared to mono pretreatments. Maximum COD solubilization ranged from 2 to 37% in alkaline, 21 to 260% in MW, and 28 to 624% in hybrid pretreatments. Turbidity, CST, VFA, and soluble protein and carbohydrate increase further with extreme pretreatment. Particle size, dewaterability, and alkalinity are reduced with pretreatment. Introduction of DIET to pretreated sludge can avoid the deficiencies and enhance biogas production. In future perspectives, the pretreatment of WAS can improve sludge properties and make it possible to reuse it as class A bio-solids for land application. Pretreatment can also reduce the volume of sludge sent which can reduce the transportation cost. However, these methods have some challenges such as alkaline pretreatment, alkali salt conversion, longer reaction time, hemicellulose dissolution, and retention of residual chemicals in AD. In thermal pretreatment, heating and cooling time for the sludge before and after reaction is important and the formation of free radicals in the process also has a challenge. Furthermore, major challenges in MW are the reactor scale, MW absorbent, energy, and economics.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.