Application of Biochar as Functional Material for Remediation of Organic Pollutants in Water: An Overview

: In recent years, numerous studies have focused on the use of biochar as a biological material for environmental remediation due to its low-cost precursor (waste), low toxicity, and diversity of active sites, along with their facile tailoring techniques. Due to its versatility, biochar has been employed as an adsorbent, catalyst (for activating hydrogen peroxide, ozone, persulfate), and photocatalyst. This review aims to provide a comprehensive overview and compare the application of biochar in water remediation. First, the biochar active sites with their functions are presented. Secondly, an overview and summary of biochar performance in treating organic pollutants in different systems is depicted. Thereafter, an evaluation on performance, removal mechanism, active sites involvement, tolerance to different pH values, stability, and reusability, and an economic analysis of implementing biochar for organic pollutants decontamination in each application is presented. Finally, potential prospects to overcome the drawbacks of each application are provided.


Introduction
Over the past few years, the rapid growth of various industries has led to the release of toxic pollutants to the environment. In particular, various refractory pollutants such as antibiotics, phenolic compounds, dyes, and heavy metals are detected in numerous water bodies and soil [1][2][3][4][5]. In particular, pollution due to refractory pollutants in water bodies is of emerging concern due to its greater mobility compared to soil. While these pollutants pose a threat to public health, they also disrupt the microbial ecosystem [6,7]. For instance, the release of trace antibiotics will lead to the development of multidrugresistant strains, which will result in less effective clinical treatment effects of conventional antibiotics [7]. Phenolic compounds such as chlorophenols are toxic and may induce

Biochar Active Sites and Characteristics
Biochar is mainly composed of carbon, oxygen, heteroatoms (i.e., N, S), and minerals (Na, C, K). The composition distribution is highly dependent on the biochar precursor and the synthesis conditions and hence varies significantly from one biochar to another. For instance, lignin-rich (lignocellulosic) biomass sources generally yield highly carbonized and graphitized biochar materials, whereas non-lignocellulosic biomass-derived biochar will be enriched with inorganic species. For example, rice husk (contains 15-17% of silica [24]) derived biochar will encompass high Si content, whilst biomass sources such as fish bone and human hair will endow the corresponding biochar with heteroatomic species (i.e., N and S) [22,25,26]. Moreover, the surface area and active sites are influenced with biomass selection. For example, PFRs formation is related to lignin content in biomass [20], while the cellulose content in biomass is proportionally related with high surface area and porosity levels [27]. Additionally, the manipulation of biochar synthesis temperature will affect the content and composition of biochar. Higher pyrolysis temperatures will increase the graphitic carbon content via carbonization and graphitization while lowering oxygen and hydrogen contents along with enhancing the porous structure of the biochar [28][29][30].
The main active sites of biochar include oxygen-functional groups (OFGs), sp. 2 hybridized carbon, heteroatoms (i.e., N, S, P, B), and PFRs. Biochar generally has a relatively large specific surface area (SSA) with high porosity to host these active sites for envi-π-π EDA interactions, complexes adsorption, Lewis acid-base interactions, pore filling, partition uncarbonized fraction, dipole-dipole interactions, Coulombic attraction, spectrometer exchange, and acceptor interactions [22]. The adsorption process undertaken is controlled by the nature of the adsorbate, biochar properties, and operational condition (i.e., pH, pressure of water matrix, rate). The adsorption process is divided into three stages. First, there is external mass transfer of organic pollutants from the aqueous solution to the biochar surface (external diffusion), which is followed by the diffusion of organic molecules into the pores of biochar (internal diffusion) and ending with the adsorptive interactions ( Figure 1) [14,43]. Generally, OFGs can promote the adsorption of organics by various mechanisms viz, H-bonding, ion exchange, π-π interactions, and Coulombic attraction [44]. For example, OFGs on the biochar surface can promote rapid chemisorption via H-bonding with -OH groups of the organic pollutant as the -OH groups serve as a H-donor [45]. For H-bonding, whether present on the biochar surface or organic pollutant, groups such as -NH and -OH serve as H-donors, whilst groups such as benzene rings, F, N, and O act as H-acceptors for H-bonding. In addition, the presence of -OH and -COOH on the surface of biochar makes it more hydrophilic, thereby promoting hydrophobic interactions with organic pollutants with a high level of hydrophobicity. On the other hand, pollutants with strong electron withdrawing functional groups (i.e., F, N-heteroaromatic rings) acting as electron acceptors can allow the adsorption processes of encompassed π-π EDA interaction with biochar [46]. Carbon structural properties may also participate in the adsorption process. The graphitic carbon structure promotes the adsorption of organics. Particularly, the adsorption process takes place between the highly graphitized carbon layers and the aromatic structure of the pollutant through π-π interactions because the graphitized biochar will act as an excellent πelectron donor, and the pollutant will act as an acceptor. For instance, Kim et al. [36] studied the effect of pyrolysis temperature on the adsorption and catalytic activity of biochar from apple tree pruning waste for the removal of ACT. Considering the ACT pKa (9.4) and its log K ow (0.46) with the adsorption reaction taking place at a medium pH value (5.9-6.1), both electrostatic attractions and hydrophobic interactions between the biochar and ACT were not the reason for the enhanced adsorption efficiency. Consequently, it was concluded that the enhanced adsorption was ascribed to the π-π EDA interactions, especially that the enhancement in the adsorption was in good agreement with the enhancement of the graphitization degree of the synthesized biochar. Similarly, the maximum adsorption (32.9%) of sulfamethoxazole (SMX) by Spirulina residue-derived biochar (SDBC 900) was ascribed to enhanced graphitization inducing greater π-π interactions between the carbon surface and the SMX aromatic rings [29]. Heteroatoms can enhance the adsorption efficiency of biochar. For instance, biochar derived from peanut shells can only remove 20% of diethyl phthalate, while co-doping with N and S allowed 60% of the same pollutant to be removed [47]. N doping can alter the electronic structure of carbon and enhance the interactions between the N-doped biochar with the organic pollutant [14,48]. Additionally, N species serve as a Lewis basic sites, so they will promote the adsorption of pollutants through Lewis acid-base interactions and can undergo electrostatic attraction with the negatively charged OFGs of pollutants ( Figure 1) [14]. Oppositely, B species can provide Lewis acid sites [49] that can be linked to Lewis base sites on the pollutant for Lewis acid-base interactions. F doping can promote biochar adsorption efficiency by H-bonding with -NH and -OH groups and/or other H-donors of the pollutants.

H 2 O 2 Activation
Chemical oxidants such as H 2 O 2 , PS, and O 3 are capable of treating organic effluents. However, due to their relatively mild oxidative potentials, catalytic activation using biochar to generate reactive species is desirable. The activation process follows an electron transfer regime between the activator (biochar) and the oxidant, and these activation mechanisms are influenced by the properties of the peroxide (O-O) bond (i.e., bond distance and dissociation energy). Among the common oxidants, H 2 O 2 , with a redox potential of 1.76 V vs. NHE [55], is highly preferable in water treatment because its decomposition products are only oxygen and hydrogen, which is considered environmentally friendly to water treatment [56].  [57], which will consequently attack the organic molecules. The fission of the H 2 O 2 bonds by biochar activation is taken place by the surface active sites of biochar with electron-donating capacities, amongst which electron-rich OFGs are most effective in H 2 O 2 activation. For example, Sun et al. [58] found that after using biogas residue-derived biochar to treat some organic pollutants in the H 2 O 2 system, the contents of C-OH, C-O, and C=O in the biochar were reduced compared with those before the catalytic cycle, indicating the contribution of these species as the catalytic active sites. In addition, the internal electron transfer from amorphous carbon (sp. 3 [59] found that pyridinic N decreased from 24.10 at % to 15.86 at % and pyridinic N decreased from 32.24 at % to 22.25 at % after using sludge-derived biochar in H 2 O 2 activation for CIP removal, indicating the importance of these N species in sharing electrons with neighboring C atoms compared to other N species (i.e., pyrrolic N and graphitic N). Similarly, Sun et al. [58] found that the graphitic and pyridinic N content of biogas residue-derived biochar decreased after the reaction, stipulating that graphitic and pyridinic N were more operative than pyrrolic N in activating H 2 O 2 .
Other than the formation of reactive radicals, H 2 O 2 activation by biochar can promote nonradical pathways, in which the produced species experience less oxidative potential than the radical pathways. A nonradical pathway is more resistant toward being affected by conventional scavengers, inorganic ions, and pH changes [61,62]. During this pathway, the breakage of the O-O bond is not a prerequisite. According to reports, the nonradical pathways in biochar/H 2 O 2 include the production of singlet oxygen ( 1 O 2 ) and surface-bound • OH. For example, it was found that the activation of H 2 O 2 by low temperature (300 • C) derived biochar was dominated by • OH generation, whereas high pyrolysis temperature (800 • C) derived biochar catalytic reaction was dominated by 1 O 2 generation [58]. 1 [64]. Although the surface-bound radicals are more resistant toward different water matrixes, for free radicals, the organic pollutants have higher probability for productive reaction with ROS as they exist in the bulk of the solution, allowing reactions on and nearby the catalysts' surface. However, for the organic molecules to react with surface-bound radicals, the interaction between the oxidant and substrate is confined to take place on the catalyst surface.

O 3 Activation
O 3 is widely used for wastewater disinfection due to its high redox potential (E • = 2.08 V). Ozone may also directly oxidize organic pollutants with the targeting of C=C or N=N bonds [69]. However, direct ozonation is limited by selective interactions with organic molecules under acidic conditions. Additionally, O 3 direct reaction with some organics such as saturated carboxylic acids and inactivated aromatics is slow, making it difficult to achieve complete mineralization [70]. Nonetheless, enhanced ozonation can be achieved by catalytic activation for escalated ROS generation and faster O 3 decomposition. Table 3 provides an overview on the performance and mechanism of biochar in O 3 activation. Apparently, OFGs of biochar are efficacious in O 3 activation by the adsorption of dissolved O 3 and its catalytic activation. For example, Zhang et al. [71] used sludge-derived biochar for catalytic ozonation to treat phenol. The elemental analysis of biochar before and after catalytic activation illustrated the disappearance of carbonyl groups and emergence of carboxyl group after the reaction. From these results and upon studying the effects of catalytic parameters, it can be concluded that the electron-rich C=O active sites are of paramount importance in adsorbing dissolved O 3 , then initiating its rapid decomposition to produce • O 2 − that will cause the oxidation of C=O groups to O-C=O. Moreover, chromene-like and pyrone-like OFGs can reactively react with O 3 through the Criegee mechanism to form H 2 O 2 , which then reacts with the surface-active sites of carbonaceous catalyst or O 3 to yield • OH [72]. In addition, O 3 adsorption and decomposition to form • OH prefers the Lewis acid sites on the biochar surface [73,74]. Therefore, it is expected that B doping can endow Lewis acid sites on the biochar surface [75] and promote the O 3 activation performance. Meanwhile, Li et al. [76] found that the high density of C=C groups supplied efficient delocalized π-electrons. These electrons will react with H 2 O to form hydroxide (OH − ) and H 3 O + that will correspondingly yield HO 2 • , • O 2 − , and • OH after reaction with O 3 .   [76] Similar to H 2 O 2 activation, O 3 activation by biochar can bring rise to nonradical oxidative pathways, such as surface-confined • OH, 1 O 2 , surface-O 3 complex, surface adsorbed atomic oxygen (*O ad ), and direct electron transfer pathway [74,78]. *O ad species, with 2.43 V oxidation potential, are formed when O 3 molecules are adsorbed on defective edges (i.e., zigzag, vacancies, and armchair edges). Chemisorbed O 3 molecules on these sites are prone to dissociation into *O ad and free peroxide species. The *O ad can attack the adsorbed organic pollutants by the nonradical pathway or react with H 2 O and generate surface-adsorbed • OH [72]. Electron transfer can occur from the organic molecules to the O 3 with the biochar acting as an electron tunnel. Specific active sites such as graphitic N in biochar can provide an electron mobility region for the successful flow of electrons from the pollutant to O 3 to achieve the nonradical degradation of these organics ( Figure 2) [79]. However, these nonradical pathways are rarely reported in a biochar/O 3 system compared to other carbonaceous allotropes (e.g., CNTs, hollow sphere carbon) due to limited research on the mechanism of biochar as an O 3 activator.

PS Activation
Recently, sulfate radical (SO 4 •− )-based AOPs (SR-AOPs) have gained attention for organic pollutant degradation. Generally, SO 4 •− can be achieved through the activation of PS such as peroxymonosulfate (PMS) and peroxydisulfate (PDS). Table 4 provides an overview on biochar performance in PS activation to remove organic pollutants. In the radical pathway, PS can generate SO 4 •− through undergoing peroxide bond cleavage. SO 4 •− (+2.6 V vs. NHE) possesses comparable standard redox potential with • OH (+2.80V vs. NHE), which makes it a suitable substitute for the conventional use of • OH in AOPs [80]. It should be noted that • OH may be present in SR-AOPs due to the reaction of SO 4 •− with water to form • OH radicals [80,81]. On the other hand, PS can also be activated through a nonradical pathway, in which free moving radicals are not involved in organics mineralization. In the biochar/PS system, a nonradical pathway can proceed through surface-bound radicals [40,82], surface-confined complexes [83], electron mediation mechanism, and 1 O 2 generation. Similar to H 2 O 2 activation, surface-bound radicals such as SO 4 •− and • OH can oxidize organic pollutants in close proximity to the biochar. Surface-bound radicals are favored by improved PS adsorption on biochar. For instance, Ding et al. [82] reported that the PMS adsorption by biochar (derived from a phytoremediation plant) on the graphitic N site facilitates the formation of adsorbed radicals through receiving electron from the surrounding carbon of the biochar, while the adsorbed pollutant will act as an electron donor and undergo oxidation. Meanwhile, biochar can be used as an electron mediator to generate a direct electron transfer pathway through the transfer of electrons between PS and the pollutant. Typically, a higher extent of graphitization favors the electron mediator pathway due to the enhanced pollutant adsorption (by π-π interaction) and increased surface conductivity of the biochar [29,39,84]. Lastly, 1 O 2 generation can be achieved by the reaction of PS with the ketonic groups of biochar [85][86][87]. Furthermore, there are also reports suggesting the involvement of O 2 •− as an intermediate for 1 O 2 generation [88,89].
Generally, biochar with adequately graphitized C, high defective degree, better active sites (ketonic O and/or heteroatom dopant), and large SSA are favorable for PS activation. Defects in sp. 2 carbon lattice can induce charge localization, thereby facilitating PS adsorption for activation to generate various ROS for pollutant degradation [33,90,91]. The enrichment of π-π* transitions in the aromatic rings of the biochar provides it with the ability to generate free radicals. For instance, Zhu et al. [40] used sawdust-derived biochar (0.5 g L −1 ) for CA (0.1 mM) removal under PDS (10 mM) activation. The degradation rate after 60 min was 97.8%. Upon scavenging studies, it is concluded that the π-π* transitions activate PDS to form SO 4 •− and correspondingly generate • OH. The contribution of π-π* transition was reflected in the reduction of its content from 5.68% to 4.46% after catalytic usage. The conjugated π-aromatic structure can be used as electron-donating moieties [41] that enhance the catalytic performance via the electron donation process to PS for O-O bond splitting. However, this effect is less commonly reported due to the presence of other active sites on biochar that experience stronger electron donation capacity, such as electron-rich ketonic groups, which will compete with the aromatic conjugated system.
The active sites of biochar can be modulated by using different biomass and heteroatom doping. For instance, Meng et al. [13] found that biochar derived from cellulose-rich biomass favors PS activation due to its high ketonic group content and larger SSA when compared with lignin-rich biomass. Furthermore, heteroatom doping also shows promising results in PS activation. For instance, the doping of N atoms into biochar has been shown to improve PS activation by increasing the charge density (graphitic N) of surrounding carbon and also improving the electron-donating properties of the biochar (pyridinic and pyrrolic N) [15,29,92], while doping B species can introduce a Lewis acid group, which can facilitate PS adsorption and activation [49]. The multi-heteroatom doping of biochar utilizes possible synergistic interactions between different heteroatoms to alter the charge and spin density of C in biochar, which may be beneficial for PS activation. For instance, the synergistic interaction of N (graphitic N) and S (thiophenic S) in biochar has been shown to be able to redistribute the charge density of the C lattice in biochar, which will improve the electron-donating properties of biochar and facilitate PS activation [93][94][95].
Apparently, the role of PFR in activating PS is inconclusive. There are studies suggesting that PFR can effectively activate PS [96,97]. For instance, Zhang et al. [97] found that biochar prepared by thermal treatment at 500 • C has an antibiotics removal rate equivalent to that of PMS activation with biochar prepared at 900 • C (active sites: defective sites), due to the presence of carbon-centered PFRs in lower synthesis temperature that can activate PMS for radical generation. On the other hand, Li et al. [33] found that PFRs do not contribute significantly to activating PDS for ofloxacin degradation. Hence, more studies are needed to ascertain the role of PFR in PS activation. Despite its potential, the uses of PFR may be limited because PFRs are easily consumed, which will result in poor reusability in activating PS.

Photocatalysis
Photocatalytic degradation of organic pollutants has been extensity studied in the last few decades. Typically, metals oxides such as TiO 2 , ZnO, and BiVO 4 have been commonly used as a photocatalyst for environmental application. During photocatalytic reactions, the absorption of electromagnetic radiation by the photocatalyst is crucial to activate the photocatalyst and produce ROS, which can be used to degrade organic pollutants. The energy of the absorbed light irradiation must be greater or equal to the bandgap energy (Eg) of that photocatalyst in order to excite the electrons from the conduction band (CB) to the valence band (VB) [104]. When the excited electrons (e cb − ) migrate to CB, positive holes (h vb + ) will be formed in VB. Then, the electron/hole pair will migrate to the surface of the semiconductor to react with the adsorbed substrates. Generally, e cb − will react with the electron acceptors adsorbed on the catalyst surface to reduce it, whereas h vb + will react with electron donors on the catalyst surface to oxidize it [105,106] [107][108][109].
Recently, biochar has been found to have a semiconductor-like structure and serves as an active carbonaceous photocatalyst. Table 5 provides the performance and mechanism of biochar as a photocatalyst to remove organic pollutants. When used as a photocatalyst, biochar OFGs (i.e., C=O, C-OH) can behave as photosensitizers, as they allow the excited electron to be transferred to the adsorbed electron acceptor (e.g., O 2 ) for ROS generation [110]. Among OFGs, quinone moieties are capable of generating 1 O 2 . The quinone-like moieties are first excited under light irradiation to singlet states. The excited states will undergo an intersystem crossing to yield excited triplet states quinone moieties through additional photonic excitation. These species further react with dissolved oxygen to produce 1 O 2 by an energy transfer regime [111]. The light photons can be absorbed by the defective sites that serve as VB, leading to the activation of biochar and the excitation of electrons. Due to the excellent conductivity of biochar, the excited electron will migrate to OFGs that are serving as biochar CB. The e cb − of OFGs with effective redox potentials will react with adsorbed O 2 to generate • O 2 − and h vb + on defects will directly attack the adsorbed pollutants (Figure 3) [112]. Hence, the highly graphitized structure is important in photocatalysis as it allows better electron transfer from the biochar to the absorbed species to form ROS.  Regardless, there are limited studies on the potential of biochar as a photocatalyst and even less reports on the effect of heteroatomic doping on the photocatalytic activity of biochar. Since S doping can alleviate the HOMO-LUMO transitions and induce spin charges in the carbon framework [113], S doping into biochar is expected to have a positive effect on the biochar photocatalytic performance under visible light. In addition, because N is easily doped into the carbon structure, it allows the facile tuning of carbonaceous local electronic properties due to different electronegativity with carbon atoms [114]. It can have a fruitful effect on the photonic activation of biochar.

Comparison
The performance of biochar in treating organic pollutants varies significantly through different applications. Herein, the use of biochar in adsorption, Fenton reactions, O 3 activation, PS activation, and photocatalysis is evaluated with respect to their performance, organics removal mechanism, tolerance to water matrixes, reusability, and economics. It is noteworthy that a direct comparison between different applications is not feasible due to their different mechanisms in the remediation of organic pollutants.

Performance
The prime variable to be considered when comparing the different applications of biochar is the performance of each application in terms of organics removal. Organics decontamination rates via biochar through different applications is difficult to be compared and evaluated due to the diverse factors affecting the treatment speed, viz. organic pollutant nature, water matrix properties, and biochar active sites. With respect to the properties of organic pollutants, adsorption is the most advantageous in removing oxidation-resistant organic pollutants. On the other hand, AOPs can mineralize recalcitrant pollutants or decompose pollutant molecules into less recalcitrant products, which can be then treated by conventional biodegradation techniques. When using biochar as an adsorbent, the adsorption process will continue until the sorption capacity of the biochar is achieved. For the oxidant systems (O 3 , H 2 O 2 , PS), the oxidative reaction will extend until (i) the catalytic active sites of the biochar are fully deactivated, or (ii) all the oxidant/pollutant molecules are consumed. In the photocatalytic reactions, the reasons for ceasing the catalytic oxidation include the deactivation of biochar active sites, the occupation of the active sites of organic molecules, and the discontinuity of light irradiation. Generally, in all cases, controlling the operational parameters is important to ensure effective treatment. In adsorption, the insufficient sorption capacity of biochar and surface deactivation of biochar can be solved by increasing the load of biochar catalyst. This will provide extra active sites for adsorption and catalytic reactions. Ensuring that the biochar composition is free of toxic components (i.e., polyaromatic hydrocarbons, PFRs, CN − compounds [22]) is crucial to prevent secondary pollution. In photocatalysis, although increasing the biochar loading is beneficial, increasing the biochar loading beyond the optimal point may have a negative impact. An excessive increase in photocatalyst loading can cause light scattering and decrease light penetration into the active sites on the catalyst surface, leading to a less efficient catalytic reaction [115,116]. In addition, for the oxidant systems, increasing the oxidant concentrations can improve the catalytic reaction significantly. For instance, increasing the PMS dosage from 2.0 to 16.0 mM resulted in an increase in the degradation efficiency of 1,4-dioxane from 22.7% to 88.8% [91]. Nonetheless, the opposite trends have also been reported. For example, increasing the H 2 O 2 dosage to an excessive point will decrease the CIP degradation efficiency, which was ascribed to the depletion of the biochar surface active sites by the excessive H 2 O 2 content [59]. However, the low resistance of the biochar surface toward excessive H 2 O 2 dosage can be ascribed to the high amorphous structure (sp. 3 -carbon = 25.06%) of the used biochar. Since amorphous C is less stable (compared to graphitized C structure), amorphous C atoms can be considered as vulnerable sites for undesired cannibalistic reactions on the biochar surface during oxidative reactions. Other drawbacks of increasing the oxidant content include higher cost and undesirable competition between the oxidant molecules and organic pollutants over biochar surfaceactive sites [117].
All application of photocatalysis, PS, H 2 O 2 , and O 3 activation must undergo the adsorption of organics before the degradation reaction occurs. However, for a single adsorption system, the goal is to increase the adsorption capacity of the biochar. As for other applications, moderate adsorption is usually desired to avoid an excessive accumulation of organic molecules and adverse occupation of biochar surface active sites, thereby hindering the degradation reaction. During the photocatalytic reaction, excessive adsorption of pollutants on the surface of biochar will prevent the efficient absorption of light irradiation and quench the catalytic reaction. For the oxidant system, immoderately adsorbed organics molecules will compete with the oxidant molecules over the biochar surface active sites, thereby preventing the activation of these oxidants, because the adsorption of the oxidant may be a prerequisite of its activation [33]. For instance, Huang et al. [66] found that increasing the adsorption time of SMT on wheat straw-derived biochar prior to H 2 O 2 addition into the system had an inhibitory effect on H 2 O 2 activation. Similarly, PMS activation was found to be hampered by the addition of organics pollutants to the biochar systems [118]. Therefore, for oxidant systems and photocatalytic applications of biochar, optimized adsorption is best to be below 50% of the total removal to allow the degradation and catalytic activation of biochar to occur along with the synergistic effect of adsorption removal.

Pollutant-Biochar Interactions
Organic pollutants can be decontaminated by interacting with the biochar through (i) physical interactions, (ii) chemical interactions, and (iii) redox reactions. Physisorption of organic pollutants is the physical interactions between the biochar surface and the organic pollutants, whilst the chemical interactions are when the organic pollutant is chemically bounded (grafted) into the biochar. Physisorption is promoted by the electrostatic forces and intermolecular attraction between the biochar and the organic pollutant that is controlled by Van deer Waals weak forces by utilizing the biochar surface charge. Physisorption by electrostatic attraction is highly influenced by the surface change of biochar and pollutants. Chemisorption, on the contrary, occurs when stronger chemical bonds such as H-bonds and coordination bond, and covalent bonds are formed between the pollutant and the active sites of biochar [119,120]. In addition, the enthalpy of physisorption (20-40 kJ/mol) is less than that of chemisorption (200-400 kJ/mol) [120,121]. Both types of adsorption can co-exist in one system. Redox reactions in biochar/pollutant systems (i.e., AOPs) involve performing oxidation and reduction reactions that will mineralize, degrade, or alter the chemical composition of the organic pollutant rather than just adsorb it. This can be achieved through nonradical interaction in the presence of oxidant.

Biochar Active Sites Involvement and Produced ROS
The contribution of biochar active sites and the way they serve as active sites varies by applications. Likewise, the produced reactive species (during AOPs) are also different. From the overview in Section 3, Table 6 summarizes the contribution and function of biochar active sites in each application. During the adsorption processes, as no degradation reaction is taking place, PFRs and electron-donating defects are not reported as adsorption sites. To compare oxidants systems, it is predicted that O 3 , H 2 O 2 , and PS activation all share the same active sites involvement due to their similar activation process of O-O bond breakage. Nonetheless, PS activation makes full use of biochar active sites, whilst H 2 O 2 activation and O 3 activation have no aromatic structure contribution. In addition, H 2 O 2 activation does not utilize defective sites. The dissimilarity between the PS activation against H 2 O 2 activation and O 3 activation in terms of the contribution of aromatic structure and defective sites can be ascribed to the variance between O-O properties of each oxidant. Indeed, there is less reports on H 2 O 2 activation and O 3 activation compared to PS activation. Therefore, there is a high chance of all biochar active sites contributing for all oxidant systems. Hence, it is suggested to conduct more research on the biochar/oxidant systems, focusing on the activation mechanism. As for photocatalysis, all considered active sites contribute to the activation expect for the aromatic structure.
With regard to the active species generated during AOPs, all AOPS share the ability to generate ROS, such as • OH, • O 2 − , and 1 O 2 . However, due to limited reports, it is difficult to compare the efficiency of each system in terms of the formation of active species. Biochar acting as an electron tunnel between oxidant pollutant has only reported for O 3 [79] and PS activation [99]. For nonradical pathways, all oxidant systems have surfaceconfined active species. For instance, nonradical PS activation can yield a larger variety of surface-confined species, encompassing surface-bound SO 4 •− , surface-bound • OH, and surface-activated PS complexes. Secondly, O 3 activation can provide the biochar surface with surface-bound • OH, surface-bound O 3 complexes, and *O ad . Lastly, H 2 O 2 activation can only generate surface-bound • OH. Generally, free ROS has a stronger redox potential compared with the surface-bound species due to the surface-confinement effect [122]. However, the redox potential of surface-bound SO 4 •− , surface-bound • OH, and surfaceactivated PS complexes is currently unknown, whereas that of *O ad is 2.43 V [72]. As for other distinctions between the systems, the formation of SO 4 •− has a high oxidative potential (2.5-3.1 V), and selectivity toward organics with electron-rich groups, unsaturated bonds, and aromatic π-electrons [22] is constrained to PS activation. Additionally, electron tunneling in biochar/H 2 O 2 has not been reported, which may be ascribed to its lower redox potential compared to PS and O 3 . Moreover, the electron transfer process is expected to be more easily realized in PDS systems than in PMS and H 2 O 2 systems as the former has a higher redox potential ( [123]), making it more prone to serve as an electron acceptor. As for photocatalysis, it is the sole application that can produce reductive h VB + , which can directly degrade pollutants. Figure 4 summarizes the generated active species of each application of biochar.

Tolerance to pH Changes
The most important factor affecting water treatment is the pH of the reaction solution. The pH of the reaction mixture determines the surface charge of the biochar and the existing speciation of the pollutants. Generally, the point of zero charge (PZC) of biochar is between 7.1 and 10.5 [124]. As such, when the biochar surroundings' pH is equal to the PZC of that catalyst, the catalyst surface will be neutral with a zero net electric charge. At any pH below the PZC, the catalyst surface will be positively charged (protonated) and its surface will be negatively charged (deprotonated) at pH values above the PZC. Furthermore, strong alkaline conditions can accelerate the dissociation of the acidic functional groups of biochar [112]. Hence, a strong acidic condition is expected to promote the decomposition of alkaline surface sites of biochar. Certain biochar active sites are not influenced by pH. For instance, the carbon hybridization status and aromatic structure are not influenced by changes in pH. The effect of pH on the performance of PFRs in initiating redox reactions remains uncertain. Yu et al. [125] found that oxygen-centered PFRs on N-doped char were the main active site for PMS activation to remove endocrine-disrupting compounds. The removal performance over a wide pH range of 4-10 was almost unaffected. This may indirectly point out that the ability of PFR to perform redox reactions is not affected by changes in pH.
During the adsorption process, a higher pH can promote the dissociation of the -OH groups of pollutant and affect the H-bonding adsorption [126]. On the other hand, adsorption mechanism such as π-π EDA interactions are not affected by pH changes, as it only requires the successful interactions between biochar as an electron donor and pollutant as electron acceptor, which is not affected by the surrounding pH [126]. As mentioned earlier, all AOPs require adsorption prior to catalytic reactions to provide productive interactions between the biochar surface active sites and the pollutant molecules. Consequently, the effect of pH on the initial adsorption of pollutants has a similar effect on the efficiency of O 3 activation [76], PMS activation [127], PDS activation [102], and H 2 O 2 activation [59]. As for reactive species formation, • OH prefers intermediate to alkaline conditions, as it promotes a higher conversion rate of OH − to • OH. Furthermore, SO 4 •− functions well at a wide pH range of 2.0-8.0 [128]. Most efficiently, 1 O 2 is capable of adapting to pH changes [22]. In addition, under strong alkali conditions, PMS mainly exists as SO 5 2− , which prevents the sufficient SO 4 •− generation [129].

Stability and Reusability
The reusability of biochar is highly important as it improves time and cost efficiency. During the adsorption process, the biochar usually reaches its adsorption capacity, and the active adsorption sites are fully occupied. Therefore, a regeneration process is essential after saturation. Nevertheless, the regeneration of biochar via the desorption process is relatively facile, and it can be done through thermal or chemical regeneration. Meanwhile, during the application in a highly oxidizing environment (i.e., AOPs), the biochar-active sites may undergo chemical oxidation. This is because the generated ROS and radicals may not be able to distinguish between the organic pollutant and the biochar surface, as both are carbon, causing the cannibalistic attacks on the biochar surface. As a result, without careful control, biochar can only be used for several consecutive cycles without the need of regeneration processes. For instance, when using sludge-derived biochar as an O 3 activator, the activation performance is reduced from 95.4% to 59.3% in the fourth cycle. Upon studying the chemical properties of the catalyst surface, it was noticed that the active carbonyl group (C=O) completely disappeared from the biochar and was replaced by the carboxyl group (O-C=O), indicating that the surface was oxidized during the activation process [71]. Likewise, Oh et al. [30] found that in a biochar/PMS system, the graphitic N would be undergo reconstruction to form pyridinic N [30]. In addition, for photocatalytic reactions, the biochar surface active sites are endangered by both cannibalistic reaction by radicals and light-induced decomposition. During the photocatalytic oxidation, extensive light irradiation, especially UV light, can induce the decomposition of the surface functional groups of biochar [111], showing the importance of using visible light irradiation. Generally, the active sites can be restored via thermal treatment [29]. Additionally, since the major changes of the biochar surface after the AOPs cycle are the oxidation of the active sites, the chemical regeneration by the reducing agents can be efficient in restoring the deactivated sites. For instance, Hou et al. [130] found that treating the spent carbonaceous catalysts with sodium borohydride (NaBH 4 , strong reducing agent) can remove O-C=O groups and increase basic sites in restoring some activity; however, it caused the undesirable loss of N species.
For all AOPs, synthesizing biochar with high aromaticity can promote stability. An aromatic structure with a well-packed carbon structure and bond electron resonance can provide tolerance toward undesirable oxidative and cannibalistic reactions on the biochar surface (unlike alkyl moieties), which can enhance the durability of the active sites. For separation, metallic doping is often used to bear the biochar with magnetic properties to facilitate separation. Nevertheless, metal leaching can cause secondary pollution. To overcome these drawbacks, a biochar/polymer composite can be fabricated for undemanding separation. This is a win-win strategy, as the biochar can enhance the mechanical, thermal, and electrical properties of the polymer [131,132], and the polymer film will ease the reusability and catalyst stability. Amongst polymers, organics decontamination using floating (low-density) polymer-based materials (e.g., linear low-density polyethylene (LLDPE)) is highly desirable as no stirring is required, separation of catalyst is facile, and the floating of the catalyst allows excellent utilization of light irradiation (in the case of photocatalysis). In addition, a biochar/polymer composite has a longer lifespan of e/h pairs and thus allows better photocatalytic activity [133]. In addition, optimizing the reaction condition to ensure the maximized mineralization of organic pollutants will reduce the content of adsorbed by-products on the biochar surface and maintain the concentration of active sites.

Economics
Economically, all applications share the synthesis expenditure of biochar (capital cost) in which the total production cost is estimated to be approximately $0.91 per kg, with 40% yield along with the production of useful amounts of bio-oil [52]. As per application, operational cost varies significantly. The adsorption process is the cheapest compared to AOPs because it does not require oxidant or external energy sources (e.g., light, sonication). The desorption cost, on the other hand, is dependent on the desorption technique. In addition, since adsorption is only transferring the pollutants from one phase to another, secondary treatment is required to mineralize these organics. Therefore, the total cost of treatment of organics via adsorption includes and is influenced by the cost of the secondary treatment method. As for photocatalysis, the major cost factor is the irradiation source. Conventional light sources (i.e., mercury lamps, fluorescence lamps) are relatively expensive and produce noticeable heat while irradiating, causing an additional loss of electrical cost. Solar irradiation has been often reported as an efficient light source for photocatalytic reactions to overcome this drawback [134][135][136]. However, atmospheric changes need to be considered. Light-emitting diodes (LED) are getting attention as light sources for photocatalytic reactions [137][138][139] due to their low cost, high light intensity with less electrical power, and lower heat generation compared with fluorescence and mercury lamps [140,141].
As for the biochar/oxidant systems, the catalytic reaction cost is a function of the price of the oxidant. Between the oxidants, O 3 is the most expensive oxidant, as its generation requires the use of an electrical discharge over pure oxygen or air. This process consumes significant amounts of energy and capital cost; therefore, it is considered economically undesirable [69]. Optimized O 3 production over a range of only 10 wt% to 12 wt% is estimated to cost $2.00/kg [142]. In addition, the actual O 3 dosage in the catalytic reactions is experimentally measured by the difference between the inlet O 3 volume and the outflow [143], which indicates that the measurable amounts of O 3 can be lost without being completely used, causing an inefficient utilization of the oxidant, hence leading to higher cost. PS and H 2 O 2 rarely or will not encounter this problem. As for the other oxidants, PMS (Oxone), PDS (Na 2 S 2 O 8 ), and H 2 O 2 (30%) are priced at $26.4/kg, $50.8/kg, and $192/L, respectively (Sigma Aldrich).

Conclusions and Future Prospects
Thus far, biochar, as a carbonaceous material with various active sites, has been successfully utilized in the decontamination of organic pollutants from water by different applications. Among the numerous applications of biochar in wastewater treatment, adsorption, H 2 O 2 activation, O 3 activation, PS activation, and photocatalysis hold the superlative potential in the treatment of organic pollutants. These applications were compared with respect to performance, mechanism of removal with emphasis on pollutant-biochar interactions and biochar active sites' involvement, tolerance to changes in water pH, stability of biochar after consecutive cycles of pollutant treatment, and economic factors. Each application has its own advantages and limitations. Adsorption is superior in terms of simplicity of operational procedures, ability to remove oxidation-resistant pollutants, and cost-wise. AOPs utilize more active sites of biochar and are able to mineralize organics pollutants or form less resistant organic by-products, so they can be treated by conventional methods. All AOPs have their own advantages. For instance, PS activation is effective in utilizing all of the biochar active sites along with having a monopoly on the capability to produce SO 4 •− . H 2 O 2 is a safe oxidant with nonharmful decomposition products (hydrogen and oxygen). O 3 is a strong oxidant with various activation pathways. Photocatalysis does not require the use of any oxidant. Nonetheless, limitations are present in each application. (i) The low durability of biochar active sites during AOPs is repeatedly reported. In addition, (ii) the performance of pristine biochar for environmental remediation of organic pollutants is often slow with low efficiency. (iii) The cost of light source during photocatalytic reactions is a drawback. (iv) Adsorption is unable to mineralize or degrade organic pollutants.
(v) Biochar-pollutant and biochar-oxidant interactions are highly influenced by water pH changes, hindering its application in real wastewater. To enhance biochar performance in environmental remediation, several suggestions can be made: i.
The fabrication of robust biochar with high graphitization and aromaticity degree can prevent poor durability. In addition, compositing with polymers can protect the active site of biochar from cannibalistic reactions and can also simplify the separation of the catalyst. ii.
Heteroatom (i.e., N, S, B, F, P) doping is often found to have a fruitful effect on the performance of biochar in environmental remediation. However, co-doping and triple-doping are rarely reported. Hence, more studies are needed on multidoped biochar, with systematic investigations on the interactions between the multi-dopants within the biochar structure and its effect on the biochar performance. Moreover, as heteroatomic doping of biochar can be achieved by in situ and post-treatment methods, a comparison between the two techniques is needed to determine the most efficient method. iii.
For photocatalytic applications, LED lamps can be used instead of conventional light sources to avoid an additional cost of electricity and better utilization of energy. iv.
Synergistic removal by adsorption with other AOPs methods may provide constructive results. Nonetheless, the optimization of adsorption contribution is crucial to avoid undesirable competition over biochar active sites and hinder removal performance. v.
Specific tailoring of biochar active sites that are unlikely to be affected with pH changes and/or can produce species that are resistant to changing pH (i.e., nonradical pathways) can endow the biochar with better performance over different water matrixes. Data Availability Statement: Not applicable.

Conflicts of Interest:
The authors declare no conflict of interest.