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Review

A Review on Application of Biochar in the Removal of Pharmaceutical Pollutants through Adsorption and Persulfate-Based AOPs

by
Ziyang Kang
1,
Xigai Jia
1,
Yuchen Zhang
1,
Xiaoxuan Kang
1,
Ming Ge
1,
Dong Liu
1,
Chongqing Wang
2,* and
Zhangxing He
1,3,*
1
College of Chemical Engineering, North China University of Science and Technology, Tangshan 063210, China
2
School of Chemical Engineering, Zhengzhou University, Zhengzhou 450001, China
3
Tangshan Sanyou Group Co., Ltd., Tangshan 063305, China
*
Authors to whom correspondence should be addressed.
Sustainability 2022, 14(16), 10128; https://doi.org/10.3390/su141610128
Submission received: 22 June 2022 / Revised: 9 July 2022 / Accepted: 12 August 2022 / Published: 16 August 2022
(This article belongs to the Special Issue Green Chemistry for Environmental Sustainability)
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Versions Notes

Abstract

:
Increasing quantities of pharmaceutical pollutants have been found in aquatic ecosystems. The treatment of pharmaceutical pollutants has been a major task that people have been committed to in recent years. The removal of pharmaceutical pollutants can be achieved by adsorption and advanced oxidation processes (AOPs). Compared with other carbon materials, biochar has a strong adsorption capacity and persulfate activation ability, and more importantly, biochar is cheap and easy to obtain; thus, it has higher economic benefits. This study firstly reviews the application of biochar in the removal of drugs (tetracycline (TC), sulfamethoxazole (SMX), acetaminophen (ACT), cephalexin (CPX), levofloxacin (LEV), etc.) through adsorption and persulfate-based AOPs. In addition, we summarize the adsorption mechanism of biochar for various pharmaceutical pollutants and the main attack sites on different pharmaceutical pollutants in persulfate-based AOPs catalyzed by biochar. Finally, the challenges and prospects of biochar with respect to the removal of pharmaceutical pollutants are put forward.

1. Introduction

Energy consumption and environmental pollution have always been a difficult problem in the world [1,2,3]. Due to the rapid development of society, the demand for energy is increasing, which also enhances the dependence on green energy, and new redox batteries can help reduce the energy burden [4,5,6]. An important aspect of environmental pollution is the pollution of the aquatic environment by drugs. Drugs are widely used around the world to protect humans and animals from various diseases. Specifically, the representative drugs in this case include tetracycline (TC), sulfamethoxazole (SMX), acetaminophen (ACT), cephalexin (CPX), levofloxacin (LEV), etc. [7,8]. The worldwide consumption of antibacterial drugs increased by as much as 65% between 2000 and 2015, from 21.1 billion doses in 2000 to 34.8 billion in 2015. It is estimated that the consumption of global antibiotics in 2030 will be 200% higher than the 42 billion defined daily doses in 2015 [9]. In fact, many drugs are not completely metabolized by humans and animals, and consequently enter the ecological environment through various routes. Figure 1 shows the sources and transfer pathways of drugs in the aquatic environment [10]. On the one hand, drug residues in water and food will accumulate through the food chain for a long time, and then have a potential impact on human health. For example, a long-term accumulation of TC can inhibit children’s development and bone growth. On the other hand, after drug residues enter the aquatic environment, drug-resistant bacteria will be induced, which will adversely affect the ecosystem and destroy the ecological balance [9,11]. Both of these effects seriously threaten environmental safety and public health [12]. Therefore, it is crucial to remove pharmaceutical pollutants from the aquatic environment as soon as possible.
Various strategies, such as physical treatment, chemical treatment, and biological treatment can be used to remove pharmaceutical pollutants [13]. Most drugs have an inhibitory effect on bacterial activity, so the removal efficiency of biological treatment is low [14]. Since the concentration of most pharmaceutical pollutants is at the µg/L level, adsorption technology has the advantages of low cost, strong reproducibility, and high efficiency, which in turn allows it to be an effective method to treat low-concentration pharmaceutical pollutants [15]. In addition to this, research on the removal of pharmaceutical pollutants by advanced oxidation methods (AOPs) has attracted extensive attention [16]. Compared with other AOPs, persulfate-based AOPs have higher oxidizing properties, better selectivity, and no secondary pollution [17]. Therefore, the application of persulfate-based AOPs is also an efficient and economical method for the removal of pharmaceutical pollutants.
Carbon-based materials, such as activated carbon [18], biochar [19], carbon nanotubes [20], and reduced graphene oxide [21], are widely used in the removal of pharmaceutical pollutants through adsorption and persulfate-based AOPs. Among them, biochar prepared from agricultural by-products, sewage sludge, animal manure, and other wastes is a feasible option [22,23,24]. Biochar has many special advantages, such as a low cost, environmental friendliness, a porous structure, and abundant functional groups (e.g., –OH, –COOH and –NH2), and it is widely considered as a next generation of functional carbon material [25]. Zhang et al. [26] used biochar activated by boric acid to remove sulfadiazine by adsorption, and the adsorption efficiency of sulfadiazine could reach more than 90%, indicating that this material was a promising porous adsorbent, which could effectively remove sulfadiazine and achieve practical applications. It further confirmed the important roles of biochar and modified biochar as adsorbents for the removal of pharmaceutical pollutants. Fang et al. [27] have proposed a possible mechanism for the activation of persulfate by biochar. Persistent free radicals and oxygen-functional groups on the biochar’s surface played an important role in activating persulfate, which could greatly enhance the degradation of pollutants [28]. Therefore, biochar can be used as an efficient adsorbent and an activator of persulfate to remove pharmaceutical pollutants.
Based on the above information, the following conclusions can be drawn: (1) Drug residues are widely found in surface water and groundwater, posing a serious threat to human health and ecological balance. (2) Adsorption and persulfate-based AOPs are effective and promising removal strategies. (3) The use of biochar for adsorbents and persulfate activators is cheap and highly efficient. As shown in Figure 2, many studies have been carried out regarding the removal of pharmaceutical pollutants. But there is still a lack of detailed summaries on the mechanisms of removing pharmaceutical pollutants using both the adsorption and activation of persulfate using biochar. Therefore, this paper focuses on the mechanism of two removal paths, namely, the adsorption mechanism of the adsorption pathway and the main attack sites of free and non-free radicals towards various pharmaceutical pollutants in persulfate-based AOPs. In addition, the effects of pH, coexisting ions, the solution temperature, the biochar dosage, and the persulfate dosage on the removal of pharmaceutical pollutants are also concluded. This review helps researchers to systematically treat pharmaceutical pollution and also provides a rational support framework for the removal of other pharmaceutical pollutants in the future.

2. Removal Mechanism

As shown in Figure 3, the removal of pharmaceutical pollutants by biochar mainly includes an adsorption pathway and a degradation pathway via persulfate-based AOPs. Each path has practical value for the removal of pharmaceutical pollutants.

2.1. Adsorption

Due to the porous structure of biochar’s surface, biochar can be used for the adsorption and removal of pharmaceutical pollutants. It has the advantages of convenient operation and a simple structure design [39], and it has more economic benefits than reverse osmosis, ion exchange, and electrolysis [40]. The adsorption process usually includes three stages: external diffusion, particle diffusion, and surface reaction [41]. Adsorption mechanisms based on biochar include physical and chemical interactions [42]. In general, physical adsorption is mainly caused by the polarity of adsorbate molecules and the adsorbent surface, including processes such as pore filling, surface complexation, electrostatic attraction, hydrophobic interaction, π–π interaction, etc. [43]. It is a non-specific process guided by relatively weak attractive forces. Chemisorption is specific and can be achieved by sharing electron pairs to form chemical bonds between the adsorbent and the adsorbate [44]. Most of the adsorption removal of pharmaceutical pollutants is not limited to either physical adsorption or chemical adsorption; they can occur simultaneously and complement each other [45].

2.2. Degradation via Persulfate-Based AOPs

AOPs provide a means to degrade pharmaceutical pollutants by inducing the production of reactive oxygen species (ROS) [46]. In general, there are many advanced oxidation processes, including ozone oxidation, Fenton oxidation, persulfate oxidation, etc. Among them, persulfate-AOPs have better pH adaptability and selective oxidation, and have become an efficient strategy for degrading pharmaceutical pollutants [47]. The degradation of pharmaceutical pollutants by persulfate-based AOPs mainly includes a free radical pathway and a nonradical pathway. The nonradical pathway can be divided into electron transfer and the application of singlet oxygen (1O2).

2.2.1. Free Radical Pathway

During the degradation process of pharmaceutical pollutants, free radical pathways are extremely important [48]. Generally, there are three kinds of free radicals involved in the degradation system of persulfate-based AOPs: SO4•−, •OH, and O2•− [49]. During the degradation process, the main active species are also different due to the different pollutants and biochar. As a strong oxidant, SO4•− has a standard redox potential of 2.5–3.1 V, which can rapidly oxidize many kinds of organic pollutants and has high selectivity. In addition, the existence time of SO4•− is relatively long and SO4•− can exist in a wide pH range (2–9), which is conducive to a full reaction with pharmaceutical pollutants. Active sites on biochar’s surface can catalyze persulfate to generate SO4•− (Equations (1)–(3)). The advanced oxidation process using •OH is a recognized clean and safe technology, and •OH also has a relatively strong oxidation effect. Pollutants and their intermediates can be easily decomposed into H2O and CO2 in an instant [50]. •OH can be produced by the reaction of SO4•− and OH− (Equation (4)). O2•− has a relatively weak oxidation capacity and a limited reactivity with organic pollutants, so its main role may be to participate in the catalytic process as an intermediate [51]. O2•− can be produced by dissolved oxygen (Equation (5)). Free radicals have three very significant characteristics: a strong reactivity, paramagnetic qualities, and a short lifespan [52]. In the process of molecular bonding, the unpaired electrons of free radicals have a strong tendency to pair. Therefore, most free radicals are relatively active, highly reactive, and easily react to form stable molecules. However, the existence time of free radicals is short, and the half-life of SO4•−, which has a relatively long lifetime, is only 30–40 μs.
Biochar’s surface − OH + S2O82− → Biochar’s surface − O•+ SO4•− + HSO4−
Biochar’s surface − OOH + S2O82− → Biochar’s surface − OO•+ SO4•− + HSO4−
Biochar’s surface − O + S2O82− → Biochar’s surface − O•+2 SO4•−
SO4•− + OH− → •OH + SO42−
O2 + e− → O2•−

2.2.2. Non-Radical Pathway

Non-radical pathways also exist in persulfate-based AOPs, including singlet oxygen (1O2) and electron transfer. The non-radical pathway can play a role in a wide range of pH conditions, including basic, neutral, and acidic conditions. In addition, non-radical pathways can reduce interference from inorganic or organic compounds. Moreover, the oxidative power of persulfate can be fully exploited in a non-radical pathway.
In the electron transfer process, the pharmaceutical pollutants are the electron donor, the biochar is the electron conductor, and the persulfate is the electron acceptor [53,54]. Electrons are transferred from the adsorbed pharmaceutical pollutants to the oxidant through the surface of biochar, which prevents the rapid recombination of electron and hole pairs, thereby increasing the degradation efficiency. The rate of electron transfer is closely related to the degree of graphitization and porosity of biochar [55]. The modification of biochar can reduce the surface resistivity and improve the efficiency of electron transfer. Biochar activated by HNO3, ZnCl2, and KOH has a broad spectrum of functional groups, a high specific surface area, a more abundant pore structure, etc., which are beneficial for improving its electron transfer ability [56,57,58]. Heteroatom doping can also increase electron transfer rates. Ye et al. [59] found that the incorporation of nitrogen with a high electronegativity and unpaired electrons could increase electron density and facilitate π electron flow.
1O2 is highly selective as an oxidant since it has an unoccupied π2p molecular orbital, and thus possesses strong electrophilic reactivity. On the one hand, 1O2 can be generated by the self-degradation of persulfate (Equations (6)–(8)). On the other hand, carbonyl/ketone groups can be considered as active sites for the generation of 1O2. 1O2 can react with most of the unsaturated pharmaceutical pollutants through electrophilic addition. In a recent study, 1O2 exhibited a strong oxidative power for electron-rich organic compounds under weakly acidic conditions (pH ≥ 5) during advanced oxidation processes [60]. It can attack the -NH2 group aligned to the benzene ring to form dimer products that could be adsorbed on the biochar’s surface to efficiently remove pollutants [61]. It can also attack the sulfonamide group to achieve the efficient degradation of sulfonamide drugs [62]. But for most pharmaceutical pollutants, the kinetic rate of 1O2 is lower than that of free radicals. Therefore, various strategies and key methods should be adopted for the rational application of 1O2 to maximize the degradation efficiency.
S2O82− → 2SO4•−
SO4•− + H2O → •OH + HSO4−
4•OH → 1O2 + 2H2O

3. Adsorption and Degradation Behaviors of Various Drugs

Taking TC, SMX, ACT, CPX, and LEV as examples and supplementing them with other pharmaceutical pollutants, their removal behaviors were studied through adsorption and degradation via persulfate-based AOPs, respectively. All five major pollutants can be detected by high performance liquid chromatography-tandem mass spectrometry. Table 1 summarizes the basic properties and applications of TC, SMX, ACT, CPX, and LEV.

3.1. Tetracycline

3.1.1. Adsorption

TC can be removed by means of adsorption. The adsorption of TC can be divided into two stages. The first stage is the fast adsorption stage, which can adsorb and remove a large amount of TC in a short time [66]. Since the adsorbent can generate many micropores, TC can diffuse into the adsorbent through most of the adsorption sites exposed on the adsorbent surface [67]. The aromatic ring structure of TC molecules can bind tightly with π electrons in biochar, acting as π electron acceptors and accelerating the occurrence of π–π interactions. There are a large number of polar groups on TC molecules, such as -OH and -NH2 [68], and there are 8 H-bond donors and 11 H-bond acceptors in each TC molecule [69]. The C-H groups of biochar can form hydrogen bonds with functional groups such as phenolic hydroxyl groups on the surface of TC to adsorb TC [70]. In the second stage, the adsorption gradually slows down and eventually reaches equilibrium, at which point the biochar has reached saturation. [71].

3.1.2. Degradation via Persulfate-Based AOPs

The use of persulfate-based AOPs has been shown to be an efficient method for TC removal, which can be realized by external energy excitation and heterogeneous catalytic processes [72]. During the degradation process, TC was first degraded into intermediates, which were further cracked into small molecules, and finally completely oxidized into carbon dioxide and water molecules. Figure 4a shows the effects of three quenching agents—EtOH, TBA, and FFA—on TC degradation [73]. It was observed that the addition of EtOH and TBA had little effect on the degradation of TC, indicating that the contribution of SO4•− and •OH was small, while with the increase of the FFA addition, the degradation efficiency of TC decreased rapidly, indicating that 1O2 contributed more in the non-radical pathway. As shown in Figure 4b, the active species were captured using TEMP as traps. It was further proved that 1O2 played an important role in the degradation of TC [73]. The two degradation pathways of TC are depicted in Figure 4c [36]. In pathway I, the C-N bond in TC was cleaved and demethylated to yield P1 and P2. The dehydration of P2 yielded P3 and P6. P3 was further converted to P4 and P5 by ring cleavage. The second pathway involved C–C, C–N, and C=C bonds being attacked by reactive oxygen species (including SO4•−, •OH, and 1O2) and transferred electrons, resulting in hydroxylation, demethylation, cleavage, and decarboxylation [74]. The formation of P7, P8, and P9 were caused by the successive hydroxylation of the rings by 1O2 attacking the –NH2 group on the benzene ring, leading to its demethylation, while the formation of P10 was caused by the –CH3 and –NH2 demethylation of P9. Oxidation-induced ring cleavage yielded the ring-opening products P11 and P12. These intermediates were further oxidized to generate CO2, H2O, and other inorganic ions. Briefly, TC was degraded through dehydration, the loss of dimethylamino groups, the cleavage of amino rings, and double bond and aromatic epoxidation.

3.2. Sulfamethoxazole

3.2.1. Adsorption

The removal of SMX by biochar can be achieved by adsorption. Many adsorption experiments have shown that SMX and biochar can form various adsorption mechanisms through π–π electron donor-acceptor, hydrogen-bond, and hydrophobic interactions. SMX is composed of an aromatic benzene ring and an aromatic heterocyclic group, which can interact with the aromatic ring on the surface of the highly carbonized biochar through π–π electron coupling, thereby achieving the effect of adsorption [75]. Another important mechanism is the formation of hydrogen bonds between dissociated H and electron-friendly atoms such as N, O, and F [76,77]. Through theoretical calculation, it was found that many nitrogen and oxygen atoms in SMX are electronegative and can form hydrogen bonds with the dissociated H on the surface of biochar. Hydrophobic interactions number among the pH-dependent adsorption mechanisms for SMX adsorption by biochar [78]. SMX is more difficult to dissolve at a high pH than at a low pH, which makes its adsorption capacity on the surface of hydrophobic biochar higher [79]. Therefore, the partial adsorption of SMX on biochar can be attributed to the hydrophobic interaction between SMX and the hydrophobic surface of biochar.

3.2.2. Degradation via Persulfate-Based AOPs

SMX can be degraded by persulfate-based AOPs. The catalytic degradation of SMX involves free-radical and non-radical pathways. As shown in Figure 5a–f, SO4•−, •OH, 1O2, and O2•− all participated in the degradation of SMX, of which 1O2 and O2•− contributed the most [80]. Density functional theory calculations have proved that the S-N bonds near the S atoms in SMX are sensitive sites for radical attack [18]. As a selective oxidizing species, 1O2 exhibits ahigh reactivity towards electron-rich groups of SMX molecules. The sulfonamide group in the SMX molecule is more vulnerable to 1O2 attack, while the isoxazole ring in the SMX molecule is not a sensitive site for 1O2 attack [81]. The biochar with a lower degree of graphitization could obtain the final product after only one hydroxylation reaction, while the biochar with a higher degree of graphitization required multiple degradation steps. The specific degradation pathway is shown in the Figure 5g. First, the parent SMX was converted to product A by hydroxylation [82]. It was subsequently converted to product B through dehydroxylation and oxidation processes. Due to the interaction of SO4•− and •OH, the S-N bond was easily broken [83,84], and then product C was obtained. Finally, product D was obtained due to deamination and dehydroxylation.

3.3. Acetaminophen

3.3.1. Adsorption

The key mechanisms for ACT removal by adsorption are π–π interactions, hydrogen bonding, and electrostatic interactions. The π–electron interaction is a key mechanism for the adsorption of aromatic compounds. The phenolic hydroxyl and amide group in the ACT are powerful electron donors that can trigger the aromatic ring to obtain a large number of electrons and interact with delocalized adsorbed electrons. ACT molecules can exhibit nitrogen-resonance electron pairs in acidic media [10]. Therefore, ACT can act as a π-donor. Electrostatic interactions occur when ionized molecules of ACT are transferred to the charged adsorbent surface. However, this interaction is lower when the adsorbent is zwitterionic or when the surface net charge is zero. Electrostatic repulsion occurs when the adsorbent surface and the adsorbate have the same charge [85]. Previous studies have shown that proton acceptor groups (such as –OH and –C=O) in ACT molecules and N with strong electronegativity can have strong hydrogen-bonding interactions with biochar [86].

3.3.2. Degradation via Persulfate-Based AOPs

ACT can be degraded by free-radical and non-radical pathways, among which the radical pathway dominated by SO4•− and •OH plays an important role. According to previous studies, the acetyl amino group of ACT was easily attacked by SO4•− [87], and •OH mainly attacked the substituents on the benzene ring of ACT, and then formed hydroquinone or 4-benzoquinone to help benzene ring-opening [88]. O2•− and 1O2 contributed to ring-opening and hydrogen abstraction reaction [89]. The specific degradation process and pathway are shown in Figure 6a,b. ACT first underwent hydroxylation reaction to form P1. In pathway I, P1 was further reacted with •OH to produce P2 and was then oxidized to P3. In pathway II, since the acetyl amino group of ACT was easily attacked by SO4•−, P1 was oxidized to P4, and P5 was generated through an attack of a hydroxyl radical on aromatic ring. There were three pathways after P5. In pathway II-1, P5 generated P6 through hydrogen abstraction dominated by O2•− and 1O2, and then P3 and P7 were obtained due to a ring-opening reaction [90]. In addition, P5 could be directly mineralized to the low-molecular compound P10 via a ring-opening reaction or transformed into P8 via hydroxylation and then P8 opened the ring to the P9 [91]. In pathway III, acylamide and phenol hydroxy on P5 were attacked through the combined action of SO4•− and •OH to form P11 and were further oxidized to small molecular compounds. In pathway IV, P1 was transformed into P12 via a ring-open reaction and further degraded to P13.

3.4. Cephalexin

3.4.1. Adsorption

The adsorption mechanism of biochar to CPX is mainly hydrogen bond interactions, electrostatic interactions, and hydrophobic interactions [93]. Biochar contains abundant oxygen-containing functional groups, such as C=O, –COOH, –OH, etc., which can form hydrogen bonds with the N-H functional groups of CPX [94]. π electrons in biochar, due to electrostatic action, can combine with the aromatic ring part of the TC molecule, which promotes π–π interaction, thus enhancing adsorption [93]. When there are fewer oxygen-containing functional groups on the surface of biochar, the surface is more hydrophobic, which in turn will generate hydrophobic interactions with CPX to promote adsorption [95,96].

3.4.2. Degradation via Persulfate-Based AOPs

The degradation mechanism of CPX is shown in Figure 7a. •OH and SO4•− were the main free radicals for the degradation of CPX [64]. The degradation pathway of CPX is shown in Figure 7b. During the degradation process, P1 was first formed by the cleavage of the β-lactam ring and the addition of a hydroxyl to replace the methyl group in the CPX molecule [94], followed by the cleavage of the C–N bond at the marker of the carboxyl loss, which divided the intermediate into two groups (A and B). The A1 universal hydroxyl group replaced the amide group to form A2 and further oxidized to A3. Subsequently, A4 was obtained due to the deamination of A2, and then A5 was produced via the deamination of A4. Acetic acid and benzene were generated resulting from the further cleavage of A5 and A3 [97]. For group B, B2 was derived from the hydroxylation of B1. In addition, B3 and B4 were formed by dehydration, while H2O and O atoms were lost, respectively. On the other hand, B5 was produced via the loss of amidogen of B1. B6 and B7 were further formed due to the loss of hydroxyl and oxygen atoms [98]. Finally, these intermediates were further decomposed into CO2, H2O, and small molecular components.

3.5. Levofloxacin

3.5.1. Adsorption

The mechanism of biochar’s adsorption of LEV can be attributed to electrostatic interactions, H-bonding, functional groups’ complexation, π–π stacking interaction, and π–π electron donor–acceptor interactions [99]. The adsorption process of LEV onto biochar was dominated by monolayer-homogeneous and chemical adsorption [100]. The maximum adsorption of LEV by biochar occurred in neutral and weakly acidic environments, which may be due to the fact that most LEV exist in the form of LEV0 at pH = 6–7, and the decrease in the molecular distance caused by the weak electrostatic repulsion may promote the formation of π–π stacking between biochar and LEV [101]. In addition, the adsorption capacities of biochar were obviously increased at higher temperatures (25–45 °C) [100]. A thermodynamic analysis showed that the adsorption process was spontaneous and endothermic [102]. The adsorption effectivity increased by increasing the site energy of biochar’s surface via a site energy distribution analysis. Abundant functional groups (e.g., –OH, –COOH, etc.) on the biochar’s surface can increase the average site energy and site energy heterogeneity, thereby enhancing the adsorption of LEV [100].

3.5.2. Degradation via Persulfate-Based AOPs

The non-radical pathway dominated by 1O2 and electron transfer played an important role in the degradation of LEV, and SO4•− in the radical pathway also contributed to the degradation of LEV (Figure 8a,b). During the degradation process, the vulnerable sites of LEV included the side with the piperazinyl and quinolone moieties and the benzene ring [65]. In order to further explore the degradation pathway of LEV, the intermediate products were detected by LC-MS (Figure 8c). The degradation of LEV included the mechanisms of demethylation, depiperazinylation, and carboxylation/decarboxylation [103]. In pathway I, P14 was formed from the piperazinyl dealkylation of LEV by electrophilic and free radical reactions [104]. P11 could also be produced from the piperazinyl oxidation and opening of LVF via attacking through radicals or electrophilic reactions [105]. P13 was obtained due to decarboxylation of P14. P9 was generated resulting from the oxidation and decarboxylation acid of the piperazine ring of P13. P9 was also produced via the oxidation and carboxylation of P11 [104]. In Pathway II, P16 was obtained due to the replacement of the F− group by the hydroxyl group. P12 formed under radical attack via two pathways: (1) the decarboxylation process of P16 and (2) the cleavage of the piperazine chain P16. P10 could be generated via the oxidation of the piperazine ring and decarboxylation and hydroxyl addition of P16 mainly under the continuous effect of the SO4•− and 1O2. P8 could be produced via the cleavage of N-C bond on P12. P6, P4, P2, and P1 could be gradually formed mainly under the attack of SO4•− and 1O2 via the oxidation and destruction of the quinolone moiety [106].

3.6. Other Drugs

In addition to the abovementioned main pharmaceutical pollutants, the pollution caused by other pharmaceutical pollutants is also becoming more and more common. Table 2 and Table 3 summarize their removal procedures using biochar from the degradation via persulfate-based AOPs and adsorption, respectively.

4. Effect of Parameters

4.1. Effect of Solution pH

The solution pH affects reactive oxygen species, the surface charge of biochar, and the dissociated form of the target antibiotic. First, the generation of active species in the free radical and non-radical pathways is pH-dependent. In a free radical pathway, under acidic conditions, the dominant species is SO4•−, and under alkaline conditions, the dominant species is •OH radicals. Since under alkaline conditions the solution contains more hydroxide ions, SO4•− can react with water or a hydroxyl to generate •OH [120]. In the non-radical pathway, 1O2 exhibits a strong oxidative ability for electron-rich organic compounds under weakly acidic conditions (pH ≥ 5) and can rapidly degrade pollutants [60]. Second, when the solution pH is lower than the zero potential of the carbon material, the carbon’s surface has strong affinity for anionic pollutants [121]. Finally, the solution pH is closely related to the dissociated form and degree of deprotonation of the antibiotic. For instance, TC can exist in three forms depending on the pH: it is cationic at pH < 3.3, dipolar (zwitterionic) at 3.3 < pH < 7.7, and negatively charged at pH < 7.7 [122]. Furthermore, it should be noted that increased doses of persulfate can also lead to a decrease in pH and an increase in inorganic ions, which will reduce the efficiency of biochar with respect to cleansing water bodies [123].
The effect of the solution pH on the adsorption path to remove pharmaceutical pollutants is due to electrostatic interactions, ion exchange, π–π EDA interactions, and so on [122]. Sun et al. [124] found that the main mechanism of pH value on the adsorption removal of SMX was based on the protonation/deprotonation of SMX at different pH values. At pH < 1.7, 1.7 < pH < 5.6, and pH > 5.6, SMX presented cationic, neutral, and anionic states, respectively. When the 4 < pH < 6, the adsorption capacity for SMX was the strongest. The H-bond is the main mechanism at pH = 6, while the ion–dipole/charge-dipole force may be the key role at pH = 4–5. Furthermore, SMX absorption depends on π–π interactions.

4.2. Effect of Coexisting Ions

The degree of the adsorption of pharmaceutical pollutants by different biochar types can be attributed to space limitations [125], electrostatic interactions [126], and the properties of biochar functional groups [127]. Different ions have different effects on the removal of pharmaceutical pollutants by biochar adsorption. Anions have little effect on biochar adsorption, while cations can promote and inhibit adsorption to varying degrees [93]. Pan et al. [128] reported that when Na+ and Ca2+ existed in the solution, the surface functional groups would react with these ions to reduce the adsorption of TC by biochar. At the same time, these ions also occupied the adsorption position. The same results were reported by Hammond et al. [129] On the contrary, the presence of some ions sometimes promotes the adsorption of pollutants. Xiang et al. [29] reported that the presence of Na+ led to an increase in the adsorption capacity of LEV. This tendency may be because the salting out of pharmaceutical pollutants reduces their solubility in water, thereby enhancing the adsorption of biochar.
Persulfate-based AOPs are also affected by different ions (Cl−, CO32−, HCO3−, NO3−, NO2−, HPO42−, and H2PO4−). This is mainly because these ions have a scavenging effect on different free radicals. In general, NO3−/NO2− and HPO42−/H2PO4− have little effect under certain concentrations. Liu et al. [130] reported that although NO3−/NO2− and HPO42−/H2PO4− would block the catalytically active sites on carbon-based materials, resulting in reduced catalytic activity, their reactions with •OH and SO4•− were too slow to be ineffective at ordinary concentrations. Therefore, the effects of NO3−/NO2− and HPO42−/H2PO4− on the catalytic process are insignificant. Yang et al. [131] reported that PO43− may also promote peroxymonosulfate’s activation under certain conditions. According to previous research, Cl− exhibited a significant inhibitory effect at low concentrations. Cl− above a certain concentration value will promote the removal of organic pollutants. Since it was discovered that SO4•− and •OH can react with Cl− to generate less active Cl•, the presence of Cl− might reduce the oxidation efficiency. On the contrary, at high concentrations, the generated Cl• was converted into active chlorine (Cl2 and HClO), which improved the catalytic performance [132]. The effects of CO32− and HCO3− on the degradation were completely opposite to that of Cl−, and low concentrations of CO32− and HCO3− could enhance the catalytic degradation process. However, high concentrations of CO32− and HCO3− inhibit the catalytic degradation process. Jiang et al. [133] and Liu et al. [134] found that the hydrolysis of CO32− and HCO3− would lead to an increase in the solution pH at low concentrations. This results in the decomposition of HSO5− into SO52− and the formation of 1O2. At the same time, CO32− and HCO3− can remove SO4•− and •OH to obtain 1O2 indirectly. However, high concentrations of CO32− and HCO3− have a strong inhibitory effect on the degradation of pharmaceutical pollutants, which may be be related to the quenching effect of ions. Previous studies have also reported that CO32− and HCO3− can quench O2•−, further suppressing the generation of 1O2 and inducing a decrease in catalytic performance.

4.3. Effect of Solution Temperature

The reaction temperature is an important factor affecting the biochar-based removal of pharmaceutical pollutants. Choosing the appropriate temperature can greatly improve the degradation efficiency [135]. Generally speaking, an increase in temperature is conducive to overcoming the activation energy barrier and increasing the degradation rate of pharmaceutical pollutants. Simultaneously, due to the synergistic effect of the temperature rise and biochar, the system performance can also be significantly improved [136]. Raising the temperature can also promote the diffusion of pollutants and persulfates to the catalyst’s surface, enhancing the contact between the pollutants, oxidants, and catalysts, which in turn degrades organic pollutants more efficiently in complex heterogeneous systems. Hu et al. [36], in an experiment involving the persulfate-activated persulfate degradation of TC in passion fruit biochar, demonstrated that as the temperature rose from 20 °C to 40 °C, the degradation rate of TC also increased from 82% to 93%. A higher temperature can activate persulfate to generate more active substances, promote the electron transfer process between the catalyst and persulfate, and is beneficial for the degradation of pharmaceutical pollutants. It is worth noting that despite the good degradation performance at high temperatures, the energy consumption should still be considered, and the reaction temperature should be reasonably controlled. For adsorption, since the reaction is endothermic, properly increasing the temperature is beneficial for the reaction to proceed [137].

4.4. Effects of Biochar Dosage

The effect of using solely persulfate to degrade pharmaceutical pollutants is extremely low, so persulfate activation with biochar is required. An increase in the catalyst dosage can provide more active sites for the activation of persulfate; then, more active species will be produced in the free radical and non-radical pathway, thereby improving the degradation rate. In addition, the increase in the dosage of biochar can also provide more adsorption sites to enhance the adsorption effect. Xiong et al. [138] used peroxodisulfate as a catalyst to degrade TC. When the dosage of the N-doped prolifera-derived magnetic biochar was increased from 0.1 g/L to 0.3 g/L, the removal rate of the pollutants had been greatly improved in the adsorption stage and catalytic stage. (The effect of adsorption was improved from 21% to 49%, and the catalytic degradation effect was improved from 75% to 88%).

4.5. Effects of Persulfate Dosage

The amount of oxidant directly affects the degradation performance of the biochar/persulfate system. In order to make the degradation rate reach the optimum value, it is necessary to select the appropriate amount of oxidant for the specific experimental situation. Regarding the biochar-activated persulfate degradation of phenolic compounds in water, Kleopatra Miserli et al. [139] mentioned that the persulfate concentration increased slightly, and the degradation rate also increased. However, when the amount of persulfate was too high, the degradation rate did not change much. This phenomenon may be due to the limited amount of biochar relative to the persulfate concentration, which cannot provide enough active sites; therefore, the excess persulfate will consume the active species, converting SO4•− or •OH to S2O8− with poor oxidizing ability and oxygen (Equations (9)–(11)), resulting in a decrease in degradation performance [27]. Wang et al. [140] also demonstrated in the degradation of SMX by biochar/Fe3+ that the degradation rate of pharmaceutical pollutants was not always proportional to the amount of persulfate.
SO4•− + S2O82− → S2O8•− + SO42−
HO• + S2O82− → S2O8•− + OH−
3/2H2O + S2O8•− → 2SO42− + 3/4O2 + 3H+

5. Conclusions and Future Prospects

5.1. Conclusions

This paper reviews the research progress on biochar in the field of treating pharmaceutical pollutants, including the removal path, the removal behavior, and the influence of operation parameters. The removal process of various pharmaceutical pollutants mainly includes the adsorption and degradation via persulfate-based AOPs. In the adsorption process, the removal methods of five kinds of pharmaceutical pollutants (TC, SMX, ACT, LEV, and CPX) are similar, mainly because the aromatic ring part of the drug is closely associated with the π electrons in the biochar, thereby accelerating the occurrence of π–π interaction to remove the pharmaceutical pollutants; then, the free hydrogen on the biochar’s surface forms hydrogen bonds with the more electronegative functional groups on the surface of the drug molecule for further adsorption. The degradation via persulfate-based AOPs includes the free radical and the non-radical pathway. During the degradation process, the C-N bond on the TC molecule is attacked and cleaved, resulting in a demethylation effect. The S-N bonds near the S atoms in SMX are sensitive sites for radical attack and the sulfonamide group is more vulnerable to 1O2 attack. The acetyl amino group of ACT is easily attacked by SO4•−, and •OH mainly attacks the substituents on the benzene ring of ACT, and then forms hydroquinone or 4-benzoquinone to aid the opening of the benzene ring. For the degradation of CPX, SO4•− and •OH mainly attack β-the cleavage of the lactam ring and -OH of CPX. SO4•− and •OH attack the nitrogen atom of piperazine in LEV, leading to demethylation effects. Table 4 and Table 5 summarize the removal conditions and reusability of five major pharmaceutical pollutants (TC, SMX, ACT, LEV, and CPX) in persulfate-based AOPs and adsorption, respectively. The pH value of the solution mainly affects the species of reactive oxygen species, the surface charge of carbon material, and the dissociation form of the target antibiotic. Coexisting anions affect the degradation efficiency of pharmaceutical pollutants in the free-radical pathway but have little effect on the non-free radical pathway.

5.2. Future Prospects

The application of biochar in the field of removing pharmaceutical pollutants still have many challenges. According to the current research status, the future research and prospects are listed as follows (Figure 9).
(I) The current research on the persulfate oxidation method is mainly concentrated on the laboratory scale, and there is a lack of research on large-scale actual water bodies. Changes in pH and the generation of secondary by-products must be comprehensively considered before practical applications. Moreover, a higher concentration of persulfate must be used in the treatment of pharmaceutical pollutants, which will result in a large amount of sulfate ions remaining in the aquatic environment. The separation of sulfates in the removal of pharmaceutical pollutants has adverse effects on the environment. Therefore, it is also worth exploring how to limit the dose of persulfate and how to restore persulfate in situ.
(II) There are relatively few studies on the catalytic mechanism of biochar/persulfate-AOPs, and the degradation pathways generated by persulfate activation are diverse. According to the material properties, such as the surface functional groups, the presence of metals, oxygen vacancies, or other defects, they may involve the simultaneous generation of free-radical and non-radical pathways. There is no precise rule of thumb for forecasting either pathway. Using advanced characterization techniques, it is expected that the active surface sites of catalysts will be determined by measuring the surface changes before and after the reaction. The sample preparation for characterization may lead to changes in the catalyst surface, and in situ measurement should be considered in future research. In addition, density functional theory can be used as a useful tool to explore the reaction mechanism at the molecular level.
(III) Biochar can also be used to remediate and improve soils, helping to maintain environmental sustainability. Composting organic waste with biochar can promote the biodegradation of organic waste. But for different organic wastes and different types of biochar, different doses of biochar are required. Therefore, an appropriate biochar application strategy should be developed according to the characteristics of organic waste, compost, and soil.
(IV) The technology of applying biochar to treat pharmaceutical pollutants can be applied to remove other types of pollutants to improve the aquatic environment. Neither pristine biochar nor even biochar composites have been shown to improve the adsorption of nutrients such as phosphate and nitrate in water, which are the main compounds that pollute freshwater ecosystems and cause eutrophication. Therefore, further research is required to develop or identify a suitable material to control these substances.

Author Contributions

Conceptualization, Z.K., X.J., Y.Z. and Z.H.; writing—original draft preparation, Z.K.; writing—review and editing, Z.K., X.J., Y.Z. and X.K.; visualization, X.K., M.G. and Z.H.; supervision, D.L., C.W. and Z.H.; project administration, C.W. and Z.H.; funding acquisition, Z.H. All authors have read and agreed to the published version of the manuscript.

Funding

This work was funded by the Hebei National Science Fund for Distinguished Young Scholars (No. E2019209433), Youth Talent Program of Department of Education of Hebei Province (No. BJ2018020) and Hebei Province High-level Talents Funded Project (No. B2020003030).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. Sources and transfer pathways of drug residues in aquatic environment.
Figure 1. Sources and transfer pathways of drug residues in aquatic environment.
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Figure 2. Timeline of outstanding works with respect to removing pharmaceutical pollutants [29,30,31,32,33,34,35,36,37,38].
Figure 2. Timeline of outstanding works with respect to removing pharmaceutical pollutants [29,30,31,32,33,34,35,36,37,38].
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Figure 3. Removal mechanisms and main interactions of pharmaceutical pollutants by biochar.
Figure 3. Removal mechanisms and main interactions of pharmaceutical pollutants by biochar.
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Figure 4. Effects of three quenchers on TC degradation (a) [73]; TEMP-1O2 at different reaction times (b) [73]; Degradation pathway of TC (c) [36].
Figure 4. Effects of three quenchers on TC degradation (a) [73]; TEMP-1O2 at different reaction times (b) [73]; Degradation pathway of TC (c) [36].
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Figure 5. Quenching tests using different scavengers (a) EtOH; (b) L-histidine; (c) TBA; (d) p-BQ [81]; EPR spectra using DMPO for (e) O2•− and TEMP for (f) 1O2 [80]; (g) Intermediates and degradation pathways [18].
Figure 5. Quenching tests using different scavengers (a) EtOH; (b) L-histidine; (c) TBA; (d) p-BQ [81]; EPR spectra using DMPO for (e) O2•− and TEMP for (f) 1O2 [80]; (g) Intermediates and degradation pathways [18].
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Figure 6. The degradation process of ACT (a) [92] and the proposed degradation pathaway of ACT (b) [92].
Figure 6. The degradation process of ACT (a) [92] and the proposed degradation pathaway of ACT (b) [92].
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Figure 7. Possible mechanism of persulfate activation by biochar for CPX degradation (a) [64]; Proposed reaction pathway for the degradation of CPX (b) [64].
Figure 7. Possible mechanism of persulfate activation by biochar for CPX degradation (a) [64]; Proposed reaction pathway for the degradation of CPX (b) [64].
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Figure 8. Quenching effect on LEV degradation by different scavengers (a) [65]; i–t curves with peroxymonosulfate and LEV addition (b) [65]; Possible degradation pathway of LEV (c) [65].
Figure 8. Quenching effect on LEV degradation by different scavengers (a) [65]; i–t curves with peroxymonosulfate and LEV addition (b) [65]; Possible degradation pathway of LEV (c) [65].
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Figure 9. Schematic diagram of improving prospects of biochar in water treatment.
Figure 9. Schematic diagram of improving prospects of biochar in water treatment.
Sustainability 14 10128 g009
Table
Table 1. Basic properties and applications of TC, SMX, ACT, CPX, and LEV.
Table 1. Basic properties and applications of TC, SMX, ACT, CPX, and LEV.
Target PollutantsFormulaMolecular Weight (g/mol)Molar VolumeSolubility (mg/mL)ApplicationRef.
TCC22H24N2O8444.43270.3 m3/mol1.7For rickettsial disease, mycoplasma infection, chlamydia infection, etc., caused by sensitive microorganisms[36]
SMXC10H11N3O3S253.28173.1 cm3/molAlmost insoluble in waterFor respiratory system infection, intestinal infection, and wound infection caused by sensitive bacteria, etc.[18]
ACTC8H9NO2151.16120.9 cm3/mol14Used to relieve both fever and pain of various causes[63]
CPXC16H17N3O4S·H2O365.4231.3 m3/mol13.5Used to treat multiple site infections caused by sensitive bacteria[64]
LEVC18H20FN3O4361.37243.9 cm3/mol50For the treatment of respiratory tract infections caused by bacteria, mycoplasma, etc.[65]
Table
Table 2. The removal of other pharmaceutical pollutants by biochar through degradation via persulfate-based AOPs.
Table 2. The removal of other pharmaceutical pollutants by biochar through degradation via persulfate-based AOPs.
Target PollutantsCatalystCatalyst Dosage
(g/L)		Pollutant Concentration
(mg/L)		Persulfate Dosage (mM)pHReaction Temperature
(°C)		Reaction Time
(min)		Vulnerable AreasDegradation Efficiency.
(%)		Ref.
Metronidazole
(MNZ)		Potassium-doped magnetic biochar
(KMBC)		0.520.01.06.525180C–N,
–NO2		98.40[107]
Norfloxacin
(NOR)		Iron and nitrogen co-doped biochar material (Fe@N co-doped biochar)0.110.01072520the nalidixic and piperazine rings95[108]
Sulfadiazine (SDZ)Biochar-based iron material (MBC)1.040.01.55.162560N–H,
S–N,
C–S		91.97[109]
PhenolDemineralized biochar
(DSS700)		0.5200.010725800benzene-OH57.80[110]
Arbidol
(ARB)		Biochar-supported red mud
(RM-BC)		0.2200.672515Sulfur atom,
4-(dimethylamino) methyle group		100[111]
Gemifloxacin (GMF)Biochar nanocomposite
(Zn-Co-LDH)		0.7515205.535130C=C, –OH, C–O, C–N92.7[38]
P-hydroxybenzoic acid (HBA)Biomass-derived N-doped porous carbon
(Y-PC)		0.05205.74.525120–OH97.6[112]
Table
Table 3. The removal of other pharmaceutical pollutants by biochar through adsorption.
Table 3. The removal of other pharmaceutical pollutants by biochar through adsorption.
Target PollutantsAdsorbentAdsorbent Dosage
(g/L)		Pollutant Concentration
(mg/L)		pHTimeIsotherms/KineticsAdsorption Capacity (mg/g)MechanismRef.
Ibuprofen (IBP)Biochar from pepper stems (PS-biochar)1.0407240 minLangmuir/
Pseudo-second-order		596.6π–π interaction, pore filling,
H-bonding		[113]
CiprofloxacinGa2S3 and sulfur co-modified biochar
(Ga/S-BC)		0.25140560 minLangmuir/
Pseudo-second-order		330.21Electrostatic interaction, hydrogen bonding,
π–π interactions.		[114]
SulfadiazineTea waste biochar0.55010.97720 minLangmuir/
Pseudo-second-order 		99.47π–π interactions.[115]
NaproxenPeanutshells-derivedbiochars5.0100071440 minLangmuir/
Pseudo-second-order 		81.6π–π interactions.[116]
FluoxetineBiochar2.3507.160 minFreundlich/Pseudo-second-order70Electrostatic attractions[117]
Antibiotic fermentation residues (AFRB)Sludge (AFSB)0.0140712 hPseudo-first-order82.6Aromatic structures, the chemisorption, active sites[118]
Amoxicillin antibiotic (AMX).Giant reed (AMX)0.52507400 minSips isotherm/ Pseudo-second-order345.4Hydrogen bonding, aromatic structures, the chemisorption, active sites[119]
Table
Table 4. The removal of five major pollutants by biochar through degradation via persulfate-based AOPs.
Table 4. The removal of five major pollutants by biochar through degradation via persulfate-based AOPs.
Target PollutantsCatalystCatalyst Dosage
(g/L)		Pollutant Concentration
(mg/L)		Persulfate
Dosage (mM)		pHReaction Time
(min)		Degradation Efficiency.
(%)		Cycle TimesLast Efficiency
(%)		Ref.
TCbiochart from tannery sludge (TSBC)0.350376099.1Four times79.8[141]
nitrogen-doped biochar (N-BCX)0.22027120100Three times55[142]
SMXmagnetic graphitized biochar (GMBC)0.061035.046099.4Four times23.1[143]
nitrogen-doped biochar from pomelo peel0.1100.573095Four times80[144]
ACTnanoscale zero-valent iron biochar (BC-Fe0) derived from waste lignocellulose rice0.5101.8720100Five times93.8[92]
cobalt-impregnated biochar produced from CO2-mediated pyrolysis of Co/lignin0.0550.373090Four times90[91]
CPXthe magnetic biochar (Fe2O3@LBC) derived from loofah0.4100.4720073.9Four times50[64]
LEVinnovative magnetic MgFe2O4/BC (MMB) derived from pomelo peel0.4104.2524087.87Three times67.9[145]
Co@RBC prepared from cobalt nanoparticles embedded in biochar0.2100.54.510100Five times100[65]
Table
Table 5. The removal of five major pollutants by biochar through adsorption.
Table 5. The removal of five major pollutants by biochar through adsorption.
Target PollutantsAdsorbentAdsorbent Dosage
(g/L)		Pollutant Concentration
(mg/L)		pHTimeAdsorption Capacity (mg/g)RegenerationCycle TimesLast Adsorption Capacity (mg/g)Ref.
TCpyrolyzed sludge biochar0.220724 h54.8NaOH, ethanol, and thermal treatmentFour times35.4 (NaOH treatment)
49.8 (thermal treatment)		[146]
tea residue-based biochar11002360 min70.8Ethanol treatmentFive times42.6[70]
SMXFeCl3-activated bermudagrass
(BG)-derived biochar (IA-BCs)		10100348 h280NaOH desorption, thermal oxidation, and Fenton oxidationFour times151.2 (NaOH treatment)
128.8 (thermal treatment)		[147]
Fe-impregnated graphited biochar0.2505240 min187.31NaOH treatmentThree times34.85[148]
ACTZnAl/biochar0.11255180 min1108.43NaOH treatmentThree times332.5[63]
CPXanthriscus sylvestris-derived activated biochar0.130424 h384.31NaOH treatmentThree times115.3[93]
LEVbiochar-supported MgFe2O4 (BMF)0.310051400 min44.9NaOH treatmentFour times33.3[102]
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Kang, Z.; Jia, X.; Zhang, Y.; Kang, X.; Ge, M.; Liu, D.; Wang, C.; He, Z. A Review on Application of Biochar in the Removal of Pharmaceutical Pollutants through Adsorption and Persulfate-Based AOPs. Sustainability 2022, 14, 10128. https://doi.org/10.3390/su141610128

AMA Style

Kang Z, Jia X, Zhang Y, Kang X, Ge M, Liu D, Wang C, He Z. A Review on Application of Biochar in the Removal of Pharmaceutical Pollutants through Adsorption and Persulfate-Based AOPs. Sustainability. 2022; 14(16):10128. https://doi.org/10.3390/su141610128

Chicago/Turabian Style

Kang, Ziyang, Xigai Jia, Yuchen Zhang, Xiaoxuan Kang, Ming Ge, Dong Liu, Chongqing Wang, and Zhangxing He. 2022. "A Review on Application of Biochar in the Removal of Pharmaceutical Pollutants through Adsorption and Persulfate-Based AOPs" Sustainability 14, no. 16: 10128. https://doi.org/10.3390/su141610128

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