Combination of Biochar and Advanced Oxidation Processes for the Sustainable Elimination of Pharmaceuticals in Water

Abstract

The presence of pharmaceuticals in aquatic ecosystems is an issue of increasing concern. Regardless of the low concentration of pharmaceuticals in water, they can have a toxic effect on both humans and aquatic organisms. Advanced oxidation processes (AOPs) have been described as a promising technique for eliminating pharmaceuticals due to their high efficiency. However, the cost associated with the application of these processes and their high reagents and energy requirements have affected the implementation of AOPs at large scales. Biochar has been suggested to be used as a catalyst in AOPs to overcome these limitations. Biochar is considered as an alternative heterogeneous catalyst thanks to its physicochemical characteristics like its specific surface area, porous structure, oxygen-containing functional groups, electrical conductivity, persistent free radicals (PFRs), modifiable properties, and structure defects. This carbonaceous material presents the capacity to activate oxidizing agents leading to the formation of radical species, which are needed to degrade pharmaceuticals. Additionally, AOP/biochar systems can destroy pharmaceutical molecules following a non-radical pathway. To enhance biochar catalytic performance, modifications have been suggested such as iron (Fe) impregnation, heteroatom doping, and supporting semiconductors on the biochar surface. Although biochar has been efficiently used in combination with several AOPs for the mineralization of pharmaceuticals from water, further research must be conducted to evaluate different regeneration techniques to increase biochar’s sustainable applicability and reduce the operational cost of the combined process. Moreover, operational conditions influencing the combined system are required to be evaluated to discern their effect and find conditions that maximize the degradation of pharmaceuticals by AOP/biochar systems.

1. Introduction

Water pollution with pharmaceuticals represents a threat to living organisms. When present in the environment, pharmaceuticals can permeate biological membranes, reaching cells and tissues with high efficacy and specificity due to their properties. Pharmaceuticals and their metabolites enter water bodies through the discharge of wastewater and the inadequate post-consumption disposal of pharmaceuticals. Considering that pharmaceuticals have been found in different water bodies like surface water, groundwater, wastewater, and drinking water, conventional treatments applied to treat wastewater are inefficient in the elimination of these compounds [1,2]. Traditional treatments, including coagulation–flocculation, biological degradation, and filtration are not designed to effectively eliminate pharmaceuticals in wastewater; therefore, only a reduced fraction of the pharmaceutical concentration is eliminated [3]. Additionally, pharmaceutical properties (i.e., volatility, polarity, adsorption, persistence, and lipophilicity) affect the efficiency of conventional treatment and the removal rate of these pollutants. Consequently, alternative treatments have been evaluated to assess their efficiency in the degradation of pharmaceuticals. Over the last few years, research has been focused on pharmaceutical fate, distribution, toxicity, potential elimination techniques, and, hence, the hazard that the presence of pharmaceuticals exerts on the environment and living organisms [4].

Advanced oxidation processes (AOPs) are described as a promising alternative treatment for the elimination of these contaminants in water due to their high degradation efficiency. AOPs are based on the generation of reactive radical species that transform pharmaceutical compounds into simpler molecules or even mineralize them into inorganic substances and water molecules (H₂O) [5]. AOPs like Fenton, sonolysis, photocatalyst, and sulfate-based AOPs, among others, have been applied for the treatment of pharmaceuticals, exhibiting high degradation efficiencies, rapid degradation rates, non-selectivity, and eco-friendly characteristics. Regardless of the advantages of AOPs in the treatment of pharmaceuticals, their application is limited due to high energy costs, use of reactants, and operational costs. Therefore, to increase the applicability of AOPs and reduce their operational costs, the combination of these processes with other physicochemical techniques has been suggested [4]. One of the techniques that can be coupled with AOPs is adsorption. This hybrid process is based on the application of adsorbents as heterogeneous catalysts. Adsorption coupled with AOPs has the advantage of high efficiency, low energy consumption and other associated operating costs, attainable catalyst recovery, and an increase in process productivity [6,7]. Among the adsorbing materials that have been coupled with AOPs, biochar stands out due to its sustainability and low production costs.

Biochar is a by-product derived from the generation of energy by biomass. This material presents properties including its abundant feedstock availability, oxygen-containing surface functional groups, porous structure, large surface area, and electrical conductivity. These properties make biochar a material that can potentially be used coupled with AOPs to activate oxidizing agents and degrade pharmaceuticals in water [8]. In addition to these physicochemical properties, biochar can be used as a catalyst in AOPs due to its sustainability and environmentally friendly preparation [9,10]. It can be prepared from waste biomass, presenting several issues in its final disposal, i.e., animal manure, wood residues, sewage sludge, agricultural and industrial by-products, invasive plants, etc. [11]. Therefore, the production of biochar can be seen as a “waste to wealth” strategy to aid in the problem of waste management and disposal [12]. Biochar is produced by biomass heat decomposition using several methods such as gasification, pyrolysis, torrefaction, microwave treatment, and hydrothermal carbonization. The selection of the production technique depends on the biochar characteristics that are demanded since these generation methods influence the physicochemical properties of this material and affect its catalytic performance [10]. In this regard, the use of biochar has a positive impact on the circular economy since a by-product derived from a specific process is utilized in the remediation of water pollution with pharmaceuticals [13,14]. Consequently, biochar is considered as a material that can help in the accomplishment of sustainable development goals like water treatment, environmental remediation, climate change, and energy production [9].

Under this scenario, this work is focused on the potential of the combination of biochar with AOPs in the treatment of pharmaceuticals in water. First, the distribution, occurrence, and fate of pharmaceuticals are discussed, as well as their environmental impacts. The physicochemical properties that make biochar a promising catalyst are also explained, along with the degradation pathways pharmaceuticals can follow when treated by AOP/biochar coupled systems. Subsequently, biochar modifications that enhance its physicochemical properties and its catalytic potential are discussed. Additionally, biochar regeneration methods are described as well as the limitations and challenges associated with the implementation of the combined system. This study aims to demonstrate that sustainable materials, such as biochar, can be used in alternative water treatments to reduce the hazard ascribed to the presence of pharmaceuticals in the environment, promoting the sustainability and circularity of water treatment advanced processes.

2. Pharmaceuticals: Concept, Classification, and Physicochemical Characteristics

Pharmaceuticals are organic substances that are used to diagnose, modify, prevent, and treat diseases to improve human or animal health. This definition can also be extended to illicit drugs [15]. In recent years, there has been an increase in the production of pharmaceuticals due to the benefits pharmaceuticals confer to society [16,17]. The rate of money expended per person to buy pharmaceuticals has decreased. However, their consumption has increased due to the availability of generic drugs and the decrease in production costs [15].

Pharmaceuticals can be classified according to their chemical structure, use in treating diseases, and action on their biological target. According to the pathology in which pharmaceuticals are used, they can be classified as anti-hyperglycemic, antidepressants and antiepileptics, cytostatic, beta-blockers and hormones, analgesics and sedatives, and antivirals and antibiotics for the treatment of diabetes, mental disorders, cancer, cardiac and endocrine system diseases, pain and inflammation, and virus and bacterial infections [18,19].

Pharmaceuticals are composed of organic molecules with hydrocarbon groups, which can exhibit hydrophilic or hydrophobic character [20].

Table 1. Types of pharmaceuticals. Examples, chemical formulae, and molecular structures.

Type of Pharmaceutical	Example Name and Chemical Formula	Molecular Structure
Anti-hyperglycemic	Meglitinide C17H16ClNO4
Antidepressants	Fluoxetine C17H18F3NO
Cytostatic drugs	Docetaxel C43H53NO14
Beta-blockers	Atenolol C14H22N2O3
Analgesics	Acetaminophen C8H9NO2
Sedatives	Alprazolam C17H13ClN4
Antivirals	Zanamivir C12H20N4O7
Antibiotics	Azithromycin C38H72N2O12
Hormones	Ethinylestradiol C20H24O2
Antiepileptics	Carbamazepine C15H12N2O
Non-steroidal anti-inflammatory drugs	Ibuprofen C13H18O2

presents an example for each type of pharmaceutical, along with the chemical formula and structure. Some pharmaceuticals are polar molecules with ionizing groups in their structure, and after completing the metabolic path, can get adsorbed, distributed, and metabolized, with the subsequent change in their chemical structures [21]. Some pharmaceuticals like non-steroidal anti-inflammatory drugs have a simple molecular structure and are easy to degrade. Nevertheless, other pharmaceuticals like steroids, cytostatic drugs, and antifungals have large and complex molecular structures that limit the natural conversion of these compounds through biological, physical, and chemical processes. In this regard, their use and discharge into the environment can negatively affect human health and ecosystems [22].

3. Environmental Impacts Associated with the Presence of Pharmaceuticals in Water

3.1. Occurrence and Distribution of Pharmaceuticals in Aquatic Environments

The accelerated growth of the human population and the dependence on pharmaceuticals have contributed to the widespread presence of these compounds in the environment [3]. The discharge of pharmaceuticals into the environment is becoming a significant hazard for living organisms. Their release results from various mechanisms that follow their use either in the original form or as a metabolite. The sources of pharmaceuticals are urban wastewater, clinics, hospitals and other healthcare facilities, landfill leachate, pharmaceutical industries’ wastewater effluents, aquaculture farms, and animal husbandry runoff [3,23,24,25]. Human consumption is an important source of the release of pharmaceuticals into the environment. When these compounds are ingested, a portion of the pharmaceutical is metabolized, another portion remains unchanged, and another portion is conjugated with the inactive compound attached to the molecule. These substances are excreted and end up in the sewage system and, subsequently, in water bodies [26]. Another noteworthy source of pharmaceuticals is the disposal of unused or expired medicaments in landfills [5]. Landfill leachate releases pharmaceuticals into surrounding water bodies and soil through direct discharge or leaked leachate. Some landfills have wastewater treatment facilities to treat the leachate; however, the removal percentage of pharmaceuticals is estimated to be approximately <15% [27]. Furthermore, the production of pharmaceuticals is also a major source of pharmaceutical pollution mainly in areas where environmental legislation is lax and the discharge of these compounds is not regulated. Pharmaceutical raw materials are generally manufactured in low-income countries, whose non-restricted regulation results in the discharge of high levels of medicines in the environment [28].

Pharmaceuticals have been detected in wastewater, groundwater, drinking water, and surface water. The conventional treatments used in wastewater treatment plants cannot eliminate pharmaceutical molecules. Hence, these compounds can enter surface and drinking water, creating a risk for humans and aquatic organisms [29]. Pharmaceuticals can be bioaccumulated in fish, representing another pathway for these compounds to enter the human body [30]. In the case of groundwater, when a river that receives high volumes of wastewater effluents recharges an aquifer, the latter can be polluted with pharmaceuticals [31]. The presence of these compounds in aquifers is also related to the infiltration of wastewater used to irrigate agricultural lands and fields. Additionally, when municipal biosolids are used as fertilizers, pharmaceuticals can leachate through the soil to groundwater since medicines are not completely removed in wastewater treatment plants and can end up in sewage sludge. The application of manure and slurry to the soil also acts as a source of pharmaceuticals since animals excrete a portion of the compounds and the metabolites generated during the metabolization of the given pharmaceuticals [25]. The sources of groundwater contamination with pharmaceuticals can subsequently pollute surface water through surface runoff [31]. Figure 1 represents the distribution of pharmaceuticals in the environment and their pathways.

When pharmaceuticals are in the environment, they undergo biotic and abiotic reactions. The compound can be transformed through aerobic and anaerobic degradation, photodegradation, adsorption, hydrolysis, and sedimentation [20]. The transformation that a pharmaceutical molecule can undergo depends on its physicochemical characteristics and the environmental conditions [32]. For lipophilic substances, their concentration in water is decreased by the adsorption of the molecules in particulate matter. In turn, in the case of hydrophilic substances, their concentration decreases by dilution depending on the water flow rate. Adsorption is an important process that can determine the fate of pharmaceuticals in water [33]. Through the process of adsorption, the partitioning of the pharmaceuticals between water and sediments occurs and influences the bioavailability of the compound in aquatic organisms. In fact, sediment and suspended solids in water can become a sink of pharmaceuticals [34].

Pharmaceuticals are often found in water at trace or ultra-trace (µg/L or ng/L, respectively) levels; therefore, their detection, quantification, and removal are challenging. The detection and quantification of pharmaceuticals represent a limitation due to the need for robust and sensible analytical methods [35].

Table 2. Commonly detected pharmaceuticals in water matrixes and their concentration.

Pharmaceutical	Water Matrix	Concentration (µg/L)	Country	Ref.
Non-steroidal anti-inflammatory compounds
Ketoprofen	Pig production farm wastewater treatment plant effluent	14.7	Costa Rica	[37]
Naproxen	Pig production farm wastewater treatment plant effluent, surface water, and urban wastewater effluent	0.0544–2.8	Costa Rica, Brazil, and Colombia	[37,38,39]
Ibuprofen	Surface water	0.22–6.95	Brazil and Argentina	[38,40]
Diclofenac	Hospital wastewater	3.04	Colombia	[39]
Analgesics
Acetaminophen	Pig production farm wastewater treatment plant effluent, surface water, urban wastewater effluent, and hospital wastewater	0.12–50.9	Costa Rica, Argentina, and Colombia	[37,39,40]
Paracetamol	Surface water	3.67	Brazil	[38]
Antiepileptics
Carbamazepine	Surface water, urban wastewater effluent, and hospital wastewater	0.10–1.39	Brazil, Argentina, and Colombia	[38,39,40]
2-hydroxy-Carbamazepine	Surface water	0.079	Argentina	[40]
Epoxy-Carbamazepine		0.103
Beta-blockers
Atenolol	Surface water	0.98	Brazil	[38]
Propranolol		0.023
Hormones
17-β-estradiol	Surface water	0.0035	Brazil	[38]
Estrone		0.0099
Antihypertensives
Valsartan	Surface water	2.50	Argentina	[40]
Losartan	Urban and hospital wastewater effluents	1–1.19	Colombia	[39]
Antibiotics
Metronidazole	Surface water, urban and hospital wastewater effluents	0.108–3.54	Argentina and Colombia	[39,40]
Sulfamethoxazole	Surface water and urban wastewater effluent	0.088–0.35
N-Acetyl-Sulfamethoxazole	Surface water	0.811	Argentina	[40]
Azithromycin	Urban and hospital wastewater effluents	3.88–6.93	Colombia	[39]
Ciprofloxacin		0.62–5.56
Metronidazole		0.26–3.54
Doxycycline	Urban wastewater effluent	0.078
Clindamycin	Hospital wastewater	8.34

presents several studies reporting the presence of pharmaceuticals in different water matrixes. Despite the low pharmaceutical concentration in water, they represent a potential threat to aquatic environments [36].

3.2. Environmental Impacts of Pharmaceuticals

Pharmaceuticals are known as emerging pollutants due to their potentially harmful toxicity and risks. These compounds can generate detrimental effects on humans and aquatic species even at low concentrations. For example, pharmaceuticals in drinking water can cause health complications like fever, headaches, inflammation, faintness, chills, and vomiting. Studies on the short-term toxicity of pharmaceuticals have shown the generation of acute toxicity on aquatic organisms and long-term chronic effects owing to the continuous release and their pseudo-persistent character [32,36]. Pharmaceuticals can change the ability of microbes to metabolize different carbon sources, affecting the metabolic diversity of microorganism communities [41]. Environmental hazards related to the presence of pharmaceuticals are mainly attributed to their sub-lethal effects that can detriment the population’s health and influence the fitness of individuals. The feminization and masculinization of male and female fishes, respectively, behavioral changes in aquatic organisms, and histopathological changes to tissues are some of the effects induced by exposure to pharmaceuticals [22]. The sub-lethal effects associated with the presence of pharmaceuticals are intensified in the environment due to the combination of effects of different pharmaceuticals and other chemicals that act synergistically [42]. The presence of pharmaceuticals, namely, antibiotics, can lead to the growth of new strains of organisms like fungi, bacteria, and algae that resist antibiotic effects, leading to recurrent infections in both humans and animals [43,44]. Additionally, in general terms, pharmaceuticals are compounds with a high persistence and low biodegradability [45]. The pollution of surface water and groundwater with pharmaceuticals can substantially threaten animal and human health through the pollution of drinking water [3]. When a water body is contaminated with pharmaceuticals, the presence of such recalcitrant substances can decrease the content of oxygen, affecting the health of aquatic organisms inhabiting the water body [20].

Pharmaceuticals are designed to resist biodegradation before having a therapeutic effect. Therefore, these compounds can bioaccumulate in the tissues of living organisms and lead to histology, behavior, and metabolome changes in aquatic and terrestrial organisms by indirect exposure to pharmaceuticals when these are discharged into the environment [40,46]. Salgado-Costa et al. [40] showed that pharmaceuticals (e.g., acetaminophen, desloratadine, and phenazone) bioaccumulated in the fat tissues of Rhinella arenarum. It has been shown that the effects of global warming can regulate the impact of pharmaceuticals on aquatic organisms. For example, an elevation in water temperature can increase evaporation, resulting in a decrease in water levels and, subsequently, an increase in pharmaceutical concentration. In contrast, high temperatures can cause faster degradation of these compounds since temperature accelerates the kinetics of degradation reactions. An increase in the global temperature can raise the defense mechanism of the organisms, causing a decrease in the bioaccumulation of several pharmaceuticals due to the acceleration of metabolism and the removal of metabolites. Nevertheless, for other pharmaceuticals, the increase in the metabolic rate can lead to co-adsorption and an increase in the bioaccumulation of these compounds [46]. In

Table 3. Toxicity of pharmaceuticals.

Pharmaceutical	Sentinel Organism	Experimental Conditions	LC50, EC50, LOEC, or NOEC	Observed Effects	Ref.
Pharmaceutical wastewater (PWW)	Sinapis alba	t = 72 h T = 25 °C The soil in which the seeds were planted was hydrated with PWW	Not reported	PWW inhibited 79.42% of the plant root growth, affecting the plant root length and nutrient intake. PWW exhibited high phytotoxicity (20% of the seed germination was inhibited).	[47]
Diclofenac (DF)	Daphnia manga	t = 24, 48 h [Pharmaceutical] = 0.01–1000 mg/L	EC50 = 9.80 mg/L	The pharmaceutical caused negative impacts. The toxicity of pharmaceuticals was intensified when they were mixed.	[48]
Sulfamethoxazole (SFM)			EC50 = 43.97 mg/L
DF + SFM			EC50 = 13.59 mg/L
DF	Atyaephyra desmarestii	t = 96 h T = 20 and 25 °C [DF]20°C = 1.4–7.2 mg/L [DF]25°C = 0.5–8.9 mg/L	LC50 = 6.3 and 6.4 mg/L for 20 and 25 °C, respectively	A decreased in the respiration rate was observed when in contact with the pharmaceutical and an increase in temperature. DF and CBZ even at low concentrations produced respiratory deficiencies in shrimp.	[49]
Ibuprofen (IB)		[IB]20°C = 1.9–30.2 mg/L [IB]25°C = 0.9–35.2 mg/L	LC50 = 13.3 and 10.1 mg/L for 20 and 25 °C, respectively
Carbamazepine (CBZ)		[CBZ]20°C = 57.7–176.3 mg/L [CBZ]25°C = 9–164 mg/L	LC50 = 94.3 and 66.4 mg/L for 20 and 25 °C, respectively
Pig farm wastewater	Lactuca sativa	t = 6 d T = 20 °C	Not reported	Observed >98% germination inhibition.	[37]
Pig farm wastewater treatment plant effluent				Observed >90% germination inhibition. The use of treated wastewater for irrigation led to a potential risk due to the recirculation of antibiotic-resistant bacteria and antibiotic-resistance genes in the surrounding ecosystem.
Amitriptyline (AMI)	Danio rerio embryos	[AMI] = 0, 0.003, 0.03, 0.3, 3, and 10 mg/L	Not reported	- Overall, 100% mortality for 10 mg/L of AMI (t = 24 h). - Pericardial and yolk sac edemas and tail malformation observed at 3 mg/L of AMI. - After 96 h, 0.03 and 3 mg/L induced DNA damage.	[50]
Ketoprofen (KP)	Pseudokirchneriella subcapitata	[KP, DKP] = 1.95–2000 µg/L t = 96 h	NOEC = 7.8 µg/L LOEC = 15.6 µg/L EC50 = 240.2 µg/L	The presence of the substance inhibited the growth of the microalgae.	[51]
Dexketoprophen (DKP)			NOEC = 3.9 µg/L LOEC = 7.8 µg/L EC50 = 65.6 µg/L
Montelukast (MTL)	Daphnia magna	[MTL] = 1–200 mg/L t = 48 h	EC50 = 16.4 mg/L	Daphnids showed morphological changes. Bioaccumulation of MTL in the gut.	[52]
Losartan potassium (LOS)	Desmodesmus subspicatus	[LOS] = 0–38 mg/L t = 72 h T = 23 °C Photoperiod = 24 h	EC50 = 27.93 mg/L	LOS was classified from moderate to low aquatic toxicity.	[53]
	Daphnia magna	[LOS] = 0–380 mg/L t = 48 h T = 20 °C Comet assay [LOS] = 0.25, 2.5, 200 mg/L t = 48 h	EC50 = 303.7 mg/L	LOS induced genotoxic effects.
Albendazole (AB)	Danio rerio embryos	[FB] = 0.0022–0.22 mg/L t = 48 h	NOEC = 0.022 mg/L	Exposure to AB increased individual heart rate.	[54]
Fenbendazole (FB)		[FB] = 0.0022–0.22 mg/L	NOEC = 0.020 mg/L	FM caused the underdevelopment of the head and eyes. and the tail curve.
Oxfendazole (OD)		[OD] = 0.10–10 mg/L	NOEC = 4.6 mg/L
Flumethrin (FM)		[FM] = 0.10–1.0 mg/L	NOEC = 0.022 mg/L
Azithromycin (AZ) 3′-Decladinosyl azithromycin (AZ3)	Danio rerio larvae and GES-1 human cells	For Danio rerio [AZ, AZ3] = 0.1, 0.5, and 0.8 mmol/L; t = 6 h For GES-1 cells [AZ, AZ3] = 600 µmol/L; t = 24 h	Not reported	AZ3 (AZ metabolite) exhibits a higher gastrointestinal toxicity. AZ and AZ3 induced cytotoxic effects in GES-1 cells.	[55]
AZ	Raphidocelis subcapitata	t = 72 h T = 22 °C	EC50 = 0.051 mg/L NOEC = 0.010 mg/L LOEC = 0.033 mg/L	AZ was classified as very toxic for aquatic environments. AZ affected microalgal photosynthetic capacity and electron transfer chain. AZ induced genotoxic effects.	[56]

, the toxicity effects caused by several pharmaceuticals are reported, as well as the lethal median concentration (LC₅₀), effective median concentration (EC₅₀), lowest observed effect concentration (LOEC), and no observed effect concentration (NOEC) values.

As observed in Table 3, pharmaceuticals can generate different toxic effects on aquatic organisms. Therefore, it is imperative to evaluate a wastewater treatment technology that can efficiently eliminate these compounds in water. Advanced wastewater treatments, such as the use of AOPs, have been reported to be effective in removing pharmaceuticals in water, achieving elimination rates >99% [32].

4. Advanced Oxidation Processes Applied in the Treatment of Pharmaceuticals

AOPs are identified as an efficient, eco-friendly, and promising chemical technology to eliminate recalcitrant compounds like pharmaceuticals in water [2,5]. AOPs consist of chemical reactions that promote the degradation of organic pollutants; the compounds suffer a transformation of their chemical structure rather than a fading [57]. In AOPs, reactive species are generated to degrade or even mineralize pollutants present in the solution [5,58]. These reactive species are produced from oxidants like hydrogen peroxide (H₂O₂), ozone (O₃), peracetic acid (C₂H₄O₃), and persulfate (S₂O₈²⁻) by applying energy sources like electromagnetic and ultraviolet (UV) radiation, ultrasound, or electric current to activate them. Additionally, AOPs can use transition metals like iron (Fe) or metal oxides like titanium dioxide (TiO₂) and zinc oxide (ZnO) as catalysts to activate the oxidant. The activation of H₂O₂, O_3, C₂H₄O₃, and S₂O₈²⁻ generates radical species like hydroxyl (OH^•), sulphate (SO₄^•−), acetoxyl (CH₃C(O)O^•), and acetyl peroxyl (CH₃C(O)OO^•) radicals [59]. OH^• are strong and non-selective oxidants, which can attack a wide range of organic compounds by simple electron transfer (Equation (1)); radical electrophilic addition, resulting in their inclusion to the aromatic rings or C=C bonds (Equation (2)); and hydrogen abstraction from the N-H, C-H, and O-H bonds in the pollutant molecule (Equation (3)) [60,61]. SO₄^•− is also a radical with a high oxidation potential; however, it is much more selective and has a longer half-life (≈ 40 µs) than OH^• (≈ 20 ns). SO₄^•− can oxidate organic compounds by electron transfer (Equation (4)) [62,63]. Depending on the operational conditions, AOP applied, and aqueous matrix, other weaker reactive species can be generated like singlet oxygen (¹O₂) and superoxide radicals (O₂^•−), which are also involved in the destruction of contaminants [63].

$${\mathrm{R}}^{\mathrm{n}}+{\mathrm{O}\mathrm{H}}^{•}\to {\mathrm{R}}^{\mathrm{n}+1}+{\mathrm{O}\mathrm{H}}^{-}$$

(1)

$$\mathrm{R}+{\mathrm{O}\mathrm{H}}^{•}\to \mathrm{R}{\mathrm{O}\mathrm{H}}^{•}$$

(2)

$$\mathrm{R}\mathrm{H}+{\mathrm{O}\mathrm{H}}^{•}\to {\mathrm{R}}^{•}+{\mathrm{H}}_2\mathrm{O}$$

(3)

$${\mathrm{R}}^{\mathrm{n}}+{{\mathrm{S}\mathrm{O}}_4}^{•-}\to {\mathrm{R}}^{\mathrm{n}+1}+{{\mathrm{S}\mathrm{O}}_4}^{2-}$$

(4)

Several AOPs have been applied in the elimination of pharmaceuticals in water. Processes like UV photolysis of hydrogen peroxide (UV/H₂O₂) and ozone (UV/O₃), (photo-)Fenton and (photo-)Fenton-like processes, ultrasonication, heterogeneous photocatalysts with TiO₂ and ZnO, heterogeneous Fenton processes, electro-Fenton processes, and persulfate-based AOPs have exhibited high degradation efficiencies ranging from 79 to 98% [19,64,65,66,67,68]. In

Table 4. Characteristics, advantages, and disadvantages of AOPs applied in the treatment of pharmaceuticals.

Type of AOP	Characteristics	Advantages	Disadvantages	Ref.
UV/H2O2 system	- OH• is generated by the photolysis of H2O2. - The rate of H2O2 depends on the pH and increases with the increased in pH. - H2O2 level needs to be optimized to obtain the maximum OH• yielded. - UV < 280 nm.	- Effective degradation of different organic molecules. - No generation of sludge. - Easy scaling up. - Ease of handling and storage of reagents. - Non-selective degradation by OH•.	- Requirement of an excess of H2O2 since only 5 to 10% is photolyzed. - High levels of H2O2 can act as OH• scavengers. - High chemical costs. - The removal of the excess of H2O2 is required. - High energy demand. - Potential formation of toxic by-products. - Limited light penetration in samples with high turbidity.	[60,70,71]
Persulfate-based process	- Peroxymonosulfate (HSO5−) or peroxydisulfate (S2O82−) activation. - SO4•− are selective to electron donating groups. - SO4•− have a slower reaction towards nitro (-NO2) and carbonyl (-C=O) groups.	- High production of SO4•−. - SO4•− have a higher oxidation potential and half-lifetime than OH•. - High efficiency in the degradation of micropollutants. - No sludge production. - Low scavenging effect by the organic matter in water.	- High costs associated with oxidants. - High reagent doses needed in complex water samples. - Formation of toxic by-products in the presence of chlorine (Cl−) and bromide (Br−) ions. - The removal of residual sulfate ions (SO42−) is required.	[60,71]
UV/O3 system	- O3 is coupled with UV radiation to generate OH•. - Specially used in drinking water treatment.	- UV/O3 degrades organic pollutants without secondary pollution. - High formation rate of OH•. - OH• have a non-selective character. - Disinfection capacity. - No generation of sludge.	- Low mineralization efficiency for complex organic molecules. - Limitations associated with the production and manipulation of O3. - Formation of bromate ions (BrO3−) that have carcinogenic potential. - High reagent doses are needed in complex water samples.	[60,71,72]
(Photo-)Fenton and (photo-)Fenton-like processes	- H2O2 activation by ferrous (Fe+2) and ferric (Fe+3). - pH and H2O2 and iron levels used need to be controlled. - Redox cycling of Fe2+/Fe3+ occurs to activate H2O2.	- Rapid reaction kinetics between iron and H2O2, especially when radiation is used. - Ease of operation. - High degradation of organic substances.	- A pH of 2.8 is required. - Generation of ferric hydroxide sludge, which needs a subsequent treatment. - High costs of reagents. - Neutralization of the treated effluent before discharge is required. - When radiation is used, operating costs are augmented.	[73,74]
Ultrasonication	- Generation of cavitation bubbles by high-frequency sound waves. - Pressure waves with a frequency > 20 kHz. - Cavitation bubble collapse leading to the sonolysis of H2O and the formation of OH•. - In water treatment plants, waves are generated with a piezoelectric transducer.	- Chemical-free process. - Environmentally friendly. - Acoustic bubble formation can be controlled and optimized. - Non-reliant on optical radiation. - High degradation efficiencies.	- High energy demand.	[75,76,77,78]
Heterogeneous photocatalysis process	- Photoexcitation of the solid catalyst (semiconductors). - Generation of electron (e−)–hole (h+) pairs. - Nanostructured TiO2 is the preferred semiconductor. - Intermediate production of H2O2.	- Complete mineralization of many organic substances. - Possible use of solar irradiation. - No generation of sludge. - No need to eliminate residual reagents.	- Increase in catalyst dose for the treatment of complex matrices. - Degradation efficiency highly dependent on water sample composition. - Limited light penetration in samples with high turbidity. - Low quantum yield of radicals. - Slow kinetics. - Lower energy efficiency.	[60,78]
Heterogeneous (photo-)Fenton process	- (Photo-)Fenton and (photo-)Fenton-like processes using heterogeneous catalysts. - Iron is supported in a solid surface. - Heterogeneous catalysts must have high adsorption capacity.	- No secondary pollution. - Catalyst is easily recuperated. - Circumneutral pH can be used. - Catalysts can be recycled and reused. - Stronger oxidation than homogeneous (photo-)Fenton.	- The development of heterogeneous catalysts with high activity is required. - Heterogeneous catalysts need to be prepared to have good efficiency in complex water matrix. - Iron can leach from the catalyst surface. - Disposal of solid catalysts. - When radiation is used, operating costs are augmented.	[79,80]
Electro-(photo-)Fenton process	- Electrochemistry and (photo-)Fenton reactions are combined. - The production rate of H2O2 can be controlled by the variation in the current in the electrochemical cell. - Can be homogeneous or heterogeneous depending on the type of the cathode and the (photo-)Fenton catalyst.	- In situ generation of H2O2. - No additional input of H2O2. - Accelerated redox cycle of Fe2+/Fe3+. - Fe2+ can be (photo-)electrogenerated. - Low Fe-sludge production.	- Electro-fouling in the cell can be generated. - An acidic medium is required. - Long time to achieve complete mineralization of organic pollutants is required. - Passivation of electrodes. - When radiation is used, operating costs are augmented.	[81,82,83,84]

, the characteristics, advantages, and disadvantages of the AOPs mentioned above are listed. In spite of the high degradation efficiencies achieved by AOPs, their implementation is limited due to their high operating costs associated with the electricity and reactants demanded and the formation of by-products that can be more toxic than the parent compounds [58]. Therefore, the use of low-cost materials in AOPs represents a promising and challenging issue [2,57]. Furthermore, the development of a low-cost and easy-to-produce catalyst remains a challenge that may limit the widespread use of AOPs [69].

As an alternative to remove pharmaceuticals in water, the combination of AOPs with adsorption has been described. A hybrid AOP/adsorption process can reduce the operating costs associated with the implementation of AOPs, increasing process productivity and applicability [7]. Adsorbing materials such as clay, metal–organic frameworks, bioadsorbents, activated carbon, and biochar have been applied in combination with AOPs to remove pharmaceuticals from water [4]. Carbon-based materials like activated carbon and biochar present a high potential for the elimination of pharmaceuticals in wastewater as heterogeneous catalysts with a high content of active sites and thermal and chemical stability [85]. Among these adsorbing materials, biochar has gained attention due to its low cost, stability, sustainability, tunable physicochemical properties, and abundance in functional groups [86]. Due to these characteristics, the coupling of biochar with AOPs for the elimination of pharmaceuticals in water has been an issue of interest within the scientific community over the last few years (Figure 2). Therefore, the hybrid biochar/AOP process seems to be a potential way to increase the applicability of AOPs, reduce the negative impacts associated with the presence of pharmaceuticals in the environment, and contribute to sustainable development.

5. Coupling of AOPs and Biochar in the Treatment of Pharmaceuticals

Biochar is a porous carbon-rich material produced from the thermal decomposition of waste biomass [69]. Biochar can be produced from different biomass residues like sewage sludge, forest and agro-waste, river sediments, plant roots, food and animal waste, and invasive plants [87]. The thermochemical methods utilized to produce biochar are pyrolysis, gasification, hydrothermal carbonization, and torrefaction. Pyrolysis is the most used method to produce biochar. In this process, the operating temperature for biochar production is in the range of 300 to 900 °C, and it is conducted under an inert atmosphere achieved by applying a nitrogen (N₂) flow [88]. In the case of gasification, temperatures exceeding 700 °C are used; during the process, heat is transferred from biomass to gasification agents like oxygen, steam, and air. This thermochemical method produces lower amounts of biochar (5–10 wt%) since the majority of the feedstock is converted into syngas. Concerning hydrothermal carbonization, a biomass–water solution is heated at temperatures between 180 and 250 °C, under high-pressure conditions above the water saturation pressure [89]. Finally, torrefaction is performed at temperatures from 200 to 300 °C in an oxygen-deficient atmosphere and at a low heating rate [88]. When biomass undergoes thermochemical decomposition, volatile matter leaves the biomass surface, creating a porous structure. Additionally, the hydrogen/carbon (H/C) and oxygen/carbon (O/C) ratios decrease in biochar compared to biomass, resulting in biochar with a higher content of aromatic compounds and fewer aliphatic bonds [90].

5.1. Degradation Pathways of Pharmaceuticals by Coupling Biochar with AOPs

The application potential of biochar as a catalyst in AOPs is related to its characteristics. Biochar is mainly composed of hydrogen (H), carbon (C), and oxygen (O). Additionally, this carbonaceous material can be rich in nitrogen (N) and sulfur (S), depending on the biomass source. A decrease in the H/C rate in biochar indicates an increase in biochar aromaticity, resulting in strong electron accepting–giving effects on organic pollutants through the π electron structure in the biochar surface. On the other hand, a decrease in the O/C rate leads to a reduction in the hydrophilicity and polarity of the surface of biochar, which influences the interaction of biochar with polar organic molecules [91]. Biochar physicochemical properties, including pore volume, abundance of oxygen-containing functional groups, electrical conductivity, carbon structure defects, persistent free radicals (PFRs), and the surface area are essential when biochar is coupled with AOPs since these properties affect the catalytic potential of biochar [86].

In addition to the biomass source used, biochar characteristics are influenced by the thermochemical method utilized in its production [86]. The porous structure and surface area of biochar change its adsorption capacity and the distribution of catalytic sites on its surface. Biochar adsorption capacity is the driving force for biochar degradation of pharmaceuticals. Large surface areas and mesoporous structures in biochar are suitable for the exposure of active sites for catalysts without affecting electron transfer mechanisms. The biomass source and the temperature of the method applied for the generation of biochar affect the surface area and porous structure of this adsorbing material. When biochar is produced by a secondary feedstock such as sludge, manure, and food waste, the biochar surface area and pore volume are lower compared to biochar produced from primary feedstock (e.g., wood, fruit peels, and corn straw). The reduction in the surface area and the pore volume is attributed to the higher ash content in secondary feedstocks that block pores. Regarding the temperature and residence time, the biochar surface area and pore volume increase with the increase in temperature and residence time due to the loss of volatile compounds in biomass like aliphatic and volatile organic compounds [92]. Fast heating rates (>10 °C/min) can cause the collapse of pore structures, leading to a reduction in its pore volume and surface area [93]. Nevertheless, excessive temperatures can cause the collapse of the carbon structures, resulting in a decrease in both characteristics [94].

The oxygen functional groups on the surface of biochar are mainly carboxyl, carbonyl, hydroxyl, and ester. These functional groups affect biochar catalytic, adsorption, and redox capacity. Oxygen functional groups can activate oxidizing agents like H₂O₂, S₂O₈²⁻, and O₃ into reactive radical species since these oxidants accept electrons donated from the oxygen functional groups, leading to the generation of reactive species [92]. Additionally, the oxygen-containing functional groups can adsorb pharmaceutical molecules on the biochar surface and increase the possibility of collisions between pharmaceuticals and reactive radical species, raising the degradation efficiency of the process [69,92]. Surface functional groups in biochar such as hydroxyl (-OH), carboxyl (-COOH), and carbonyl (-C=O) can activate S₂O₈²⁻ and generate SO₄^•−, as presented in Equations (5)–(7). Furthermore, hydroxyl groups contained on the biochar surface can transfer one electron to H₂O₂ to generate OH^•, as presented in Equation (8) [95]. Subsequently, the reactive species formed can oxidate pharmaceutical molecules following the reactions presented in Equations (1)–(4). The oxygen functional groups -C=O and -COOH are more efficient in the activation of oxidants than -OH. This is attributed to the presence of a filled 2p orbital in the O atom of the -OH bond that lacks isolated electrons, hindering the transfer of electrons from -OH to the oxidizing agent [96]. The amount of oxygen-containing functional groups is highly influenced by the thermochemical decomposition method used in the production of biochar. As a matter of fact, due to the high temperatures and the presence of an oxidative atmosphere during gasification, organic compounds in biomass are oxidated and liberated, resulting in biochar with a low content of oxygen functional groups [97].

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathrm{O}\mathrm{H}+{{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-{\mathrm{O}}^{•-}+{{\mathrm{S}\mathrm{O}}_4}^{•-}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_4}^{-}$$

(5)

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathrm{O}\mathrm{O}\mathrm{H}+{{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-{\mathrm{O}\mathrm{O}}^{•-}+{{\mathrm{S}\mathrm{O}}_4}^{•-}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_4}^{-}$$

(6)

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}=\mathrm{O}+{{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-{\mathrm{C}\mathrm{O}}^{•-}+{{\mathrm{S}\mathrm{O}}_4}^{•-}+{{\mathrm{S}\mathrm{O}}_4}^{2-}$$

(7)

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathrm{O}\mathrm{H}+{\mathrm{H}}_2{\mathrm{O}}_2\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-{\mathrm{O}}^{•-}+{\mathrm{O}\mathrm{H}}^{•}+{\mathrm{H}}_2\mathrm{O}$$

(8)

When biochar is produced from biomass with a high level of lignocellulose like wood, the resulting biochar is rich in PFRs. The PFRs result from the homolytic breakage of the α-β-alkyl-aryl ether, C-O, and C-C bonds in lignocellulose-rich biomass. PFRs such as the oxygen-containing functional groups aid in activating the oxidizing agent following an electron transfer reaction to produce reactive radical species, as presented in Equations (9) and (10). Within these reactions, PFRs and the oxidant act as an electron donor and an electron acceptor, respectively. The content of PFRs in biochar depends on the temperature used in its production; an increase in temperature results in a decrease in the content of PFRs [98]. PFR formation takes place at low and medium temperatures. When the biochar production temperature surpasses a specific range, the condensation of carbon clusters occurs, and graphite-like structures are formatted. Graphite-like structures present a great number of paramagnetic unpaired electrons; therefore, its formation results in a decrease of PFRs [91]. On the other hand, the electrical conductivity of biochar represents a key factor that can affect its catalytic capacity. The activation of oxidizing agents follows an electron transfer reaction, leading to the transferring of electrons and generation of reactive radical species to degrade pharmaceuticals [92].

$${\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{e}}^{-}\to {\mathrm{O}\mathrm{H}}^{•}+{\mathrm{O}\mathrm{H}}^{-}$$

(9)

$${{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}+{\mathrm{e}}^{-}\to {{\mathrm{S}\mathrm{O}}_4}^{•-}+{\mathrm{O}\mathrm{H}}^{•}+{\mathrm{O}\mathrm{H}}^{-}\mathrm{o}\mathrm{r}{{\mathrm{S}\mathrm{O}}_4}^{2-}$$

(10)

Biochar, as a carbon material, can present carbon defects on its structure like curvatures, vacancies, and edge defects. These carbon defect structures can aid in the activation of oxidants, with the subsequent generation of radical species. This process is initiated by the adsorption of the oxidant onto the biochar surface and generates dangling sigma (σ) bonds with π-electrons that are delocalized and can donate electrons from the biochar surface to the oxidant. As consequence, radical species are generated, as presented in Equations (11)–(13). The depleted active sites can be regenerated by the electron transfer process since H₂O₂ can donate an electron to form hydroperoxyl radicals (HO₂^•−), as described in Equation (14) [95]. The degree of structure defects in biochar is analyzed with Raman spectra through the relation of the intensity of the D and G bands, which are related to the defects and atomic carbon deformations and the graphitic structures (I_D/I_G), respectively. It has been reported that the I_D/I_G increases with the rise in the temperature used in the production of biochar. This suggests that biochar with higher structure defects is obtained at high biochar production temperatures. However, too many structure defects can damage the physical strength and structural stability of biochar [91].

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathsf{\pi }+{\mathrm{H}}_2{\mathrm{O}}_2\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-{\mathsf{\pi }}^{+}+{\mathrm{O}\mathrm{H}}^{-}+{\mathrm{O}\mathrm{H}}^{•}$$

(11)

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{d}\mathrm{e}\mathrm{f}\mathrm{e}\mathrm{c}\mathrm{t}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}\to {{\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r}}_{\mathrm{d}\mathrm{e}\mathrm{f}\mathrm{e}\mathrm{c}\mathrm{t}}}^{+}+{\mathrm{O}\mathrm{H}}^{-}+{{\mathrm{S}\mathrm{O}}_4}^{•-}$$

(12)

$$\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{d}\mathrm{e}\mathrm{f}\mathrm{e}\mathrm{c}\mathrm{t}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}\to {{\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r}}_{\mathrm{d}\mathrm{e}\mathrm{f}\mathrm{e}\mathrm{c}\mathrm{t}}}^{+}+{\mathrm{O}\mathrm{H}}^{•}+{{\mathrm{S}\mathrm{O}}_4}^{2-}$$

(13)

$${\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathsf{\pi }}^{+}+{\mathrm{H}}_2{\mathrm{O}}_2\to \mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathsf{\pi }+{\mathrm{H}}^{+}+{{\mathrm{H}\mathrm{O}}_2}^{•-}$$

(14)

In addition to the radical pathways described above, in which the biochar coupled with AOPs can oxidize pharmaceutical molecules, there is a non-radical pathway that can also occur when biochar is used in sulfate-based AOPs. Two types of non-radical pathways have been reported: the electron transfer process and that associated with ₁O². In the electron transfer process, the pharmaceutical molecule adsorbed in the biochar transfers an electron to the oxidizing agent through the surface of the biochar. Therefore, the pharmaceutical acts as the electron donor, suffering oxidation; in turn, the biochar is the electron conductor, and the oxidant acts as the electron acceptor, suffering reduction [99]. Several studies have reported that the oxygen functional groups -OH and C=O on the surface of biochar can interact with O-O bonds in the S₂O₈²⁻ molecule and form a complex. The complex formed between biochar and S₂O₈²⁻ can oxidase organic molecules through the electron transfer mechanism [100]. It is reported that the electron transfer degradation mechanism is enhanced when the redox potential of the catalytic complex is higher than the redox potential of the pharmaceutical molecule [96]. Biochar prevents the recombination of the electrons and hole pairs, thus increasing the degradation efficiency of the process. The electron transfer mechanism is affected by the biochar porosity and degree of graphitization [91,99].

The ¹O₂, C=O groups on the surface of biochar play a significant role in its generation [99]. The oxygen atoms in the C=O bond present a high electron density that facilitates a catalytic reaction with S₂O₈²⁻ and leads to ¹O₂ generation [96]. As observed in Equation (15), S₂O₈²⁻ can be activated by C=O groups, producing O₂^•−. After activation, S₂O₈²⁻ suffers hydrolysis and generates reactive species like OH^•, O₂^•−, and SO₄^•−, as presented in Equations (16)–(18). O₂^•− has been reported to be involved in the generation of ¹O₂ rather than in the degradation of organic molecules (Equation (19)) [101]. Defect structures on the surface of biochar can also interfere in the generation of ¹O₂; for example, oxygen vacancies (Ov) have been described as reactive active sites for its production in the activation of peroxymonosulfate (HSO₅⁻) [102]. Ov in biochar can react with hydroxyl ions (OH⁻) and form a complex called an activated hydroxyl complex (≡Ov-OH) [96,103]. HSO₅⁻ molecules are adsorbed in the ≡Ov-OH complex, generating the elongation of the O-O bond contained in HSO₅⁻ and the production of the peroxymonosulfate anion radical (SO₅^•−), as observed in Equations (20) and (21). Afterward, SO₅^•− reacts with itself, leading to the production of ¹O₂, as presented in Equation (22). The self-reaction of SO₅^•− is enhanced in alkaline conditions [96]. ¹O₂ is a highly selective species that has a better affinity for deprotonated species; therefore, a basic pH is expected to enhance pharmaceutical degradation by ¹O₂ [95]. Additionally, ¹O₂ can react mostly with unsaturated pharmaceutical molecules by electrophilic addition [99].

$${{\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}=\mathrm{O}+\mathrm{S}}_2{\mathrm{O}}_8}^{2-}\to \mathrm{C}-\mathrm{O}+{{\mathrm{S}\mathrm{O}}_4}^{2-}+{{\mathrm{O}}_2}^{•-}$$

(15)

$${{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}+{\mathrm{H}}_2\mathrm{O}\to {\mathrm{H}{\mathrm{O}}_2}^{•}+{{\mathrm{S}\mathrm{O}}_4}^{2-}+{\mathrm{H}}^{+}$$

(16)

$${{\mathrm{S}}_2{\mathrm{O}}_8}^{2-}+{{\mathrm{H}\mathrm{O}}_2}^{•}\to {{\mathrm{S}\mathrm{O}}_4}^{2-}+{{\mathrm{S}\mathrm{O}}_4}^{•-}+{{\mathrm{O}}_2}^{•-}+{\mathrm{H}}^{+}$$

(17)

$${{\mathrm{S}\mathrm{O}}_4}^{•-}+{\mathrm{O}\mathrm{H}}^{-}\to {{\mathrm{S}\mathrm{O}}_4}^{2-}+{\mathrm{O}\mathrm{H}}^{•}$$

(18)

$${{\mathrm{O}}_2}^{•-}+{\mathrm{O}\mathrm{H}}^{•}\to {}^1\mathrm{O}_2+{\mathrm{O}\mathrm{H}}^{-}$$

(19)

$$\equiv \mathrm{O}\mathrm{v}-\mathrm{O}\mathrm{H}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}\to \equiv \mathrm{O}\mathrm{v}-\mathrm{O}\mathrm{H}\cdots {{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}$$

(20)

$$\equiv \mathrm{O}\mathrm{v}-\mathrm{O}\mathrm{H}\cdots {{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}+{{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}\to {{\mathrm{H}\mathrm{S}\mathrm{O}}_5}^{-}+{{\mathrm{S}\mathrm{O}}_5}^{•-}+\equiv \mathrm{O}\mathrm{v}-\mathrm{O}\mathrm{H}+2{\mathrm{H}}^{+}$$

(21)

$${{2\mathrm{S}\mathrm{O}}_5}^{•-}\to {}^1\mathrm{O}_2+2{{\mathrm{S}\mathrm{O}}_4}^{2-}$$

(22)

5.2. Biochar Modifications to Be Used in AOPs

5.2.1. Iron (Fe) Impregnation

Due to their high redox reactivity, metals have become attractive species to be supported on the biochar surface, improving its catalytic capacity when combined with AOPs. Among the different metals with variable valences like manganese (Mn), cobalt (Co), copper (Cu), and iron (Fe) (Mn²⁺/Mn³⁺, Cu⁺/Cu²⁺, Co²⁺/Co³⁺, and Fe²⁺/Fe³⁺, respectively), Fe is the generally used metal to modify biochar [104]. During Fe impregnation, metal oxides or metal salts are impregnated in the biochar surface and/or porous structure. Fe ions are fixed in porous biochar or onto the surface after mixing biochar with an Fe salt solution. The incorporation of Fe on the surface of biochar can increase the number of active sites on the surface of biochar [105]. Li et al. [106] modified maize straw biochar with Fe and used the Fe-modified biochar in the activation of HSO₅⁻. The authors first produced biochar by pyrolysis; afterward, 1 g of biochar was submerged into 50 mL of distilled water with 4.96 g of iron (II) sulfate heptahydrate (FeSO₄·7H₂O). The solution was shaken for 12 h, and after 10 min of setting, the formed precipitates were washed with distilled water to remove the excess Fe. The Fe-impregnated biochar was then washed with a solution of 100 mL of distilled water and 30 mL of sodium tetrahydridoborate (NaBH₄) for 40 min and mixed at 100 rpm. Finally, the Fe–biochar was washed with deoxygenated distilled water and ethanol (C₂H₆O) and dried in a vacuum for 24 h. The biochar obtained had a good dispersion of Fe on the surface and a higher catalytic capacity [106]. In turn, Wang et al. [85] reported a different approach to produce Fe-modified biochar. In this case, the authors submerged the sludge (biomass source) in a solution of iron (III) chloride (FeCl₂) before conducting pyrolysis. Sludge was added to the solution and stirred for 6 h; afterward, the mixture was placed without mixing for 2 h. Subsequently, the modified biomass was dried in an oven at 105 °C to reduce humidity and then, it was pyrolyzed. Wang and collaborators concluded that the presence of Fe on the biochar was the primary reason for the activation of oxidizing agents. As reported in the two studies mentioned above, with both modifications before and after the thermochemical decomposition process, a biochar with a high catalytic power able to activate oxidants was achieved. Fe-modified biochar has been widely used to eliminate organic compounds in wastewater by applying heterogeneous Fenton or Fenton-like processes [107,108].

Regardless of the advantages of the modification of biochar with Fe, it has been reported that the use of Fe–biochar in AOPs can exhibit low oxidation efficiencies of organic pollutants. Therefore, some modification methods have been alternatively proposed to increase the content of Fe in biochar. The introduction of transitional metals like Mn, enabling the loading of metal oxides (MnFe₂O₄) on the surface of Fe-modified biochar has been one of the approaches proposed. The introduction of Mn also increases the specific surface area of biochar and the content of functional groups like -C=O and -C=C and that of Fe. In addition, this procedure enhances the electrochemical performance of biochar [109,110]. Luo et al. [110] modified biochar with Fe and Mn by impregnating pine sawdust with a solution of manganese sulfate (MnSO₄) and ferrous sulfate heptahydrate (FeSO₄·7H₂O). The biomass with Fe and Mn was pyrolyzed and the resulting biochar was used as a catalyst to activate S₂O₈²⁻ for the elimination of metronidazole in water. The authors concluded that the addition of Mn increased the content of Fe and, therefore, the catalytic activity of the biochar. Additionally, the Mn contained on the biochar surface participated in the activation of S₂O₈²⁻, leading to the generation of radicals and the subsequent degradation of metronidazole [110].

Fe modifications can also enhance the electrical conductivity of biochar, which represents another reason why the catalytic power of biochar increases when modified with metals [92]. To reduce the cost associated with the generation of metal-modified biochar, the use of Fe-containing biomass, such as sewage sludge that is generally rich in Fe, has been suggested [111].

5.2.2. Acid–Base Modification

In the acid–base modification, biochar is doped with reactants like hydrochloric acid (HCl), phosphorous acid (H₃PO₄), nitric acid (HNO₃), sulfuric acid (H₂SO₄), potassium hydroxide (KOH), sodium hydroxide (NaOH), and potassium carbonate (K₂CO₃) [90]. To dope biochar with these reactants, biomass or biochar is washed in a solution of either an acid or base. Although the acid–base modification can be conducted before or after biomass thermochemical decomposition, post-treated biochar has shown a better enhancement of its physicochemical properties than pre-treated biochar [86]. Among the advantages of this type of modification, the rise in hydroxyl and carboxyl group levels on the biochar surface is named when acid washes are conducted. During basic modifications, the content in hydroxyl groups increases. Acid–base modifications also increase the surface area and porosity of biochar, properties that are closely related to its catalytic potential [98]. Zeng et al. [112] modified peanut shell biochar with H₂SO₄. During this procedure, peanut shell biomass was pyrolyzed at 450 °C for 15 h under a N₂ flow. Then, the calcined biomass was ground and sulfonated with a concentrated solution of H₂SO₄ at 200 °C by using a N₂ flow for 10 h. Lastly, the mixture was diluted with deionized water and dried at 120 °C for 10 h. Zen and coworkers found that the acid-modified biochar had large amounts of functional groups like hydroxyl, carboxyl, and sulfonic acid groups, which enhanced the catalytic capacity of the biochar [112]. Even though the acid–base modification can enhance the physicochemical properties of biochar, this modification method has important disadvantages, including the corrosion of the equipment used in the generation of biochar due to the use of strong bases and acids. Additionally, this type of modification can cause the collapse of the pore structure of biochar, negatively affecting its catalytic potential [90,113].

5.2.3. Biochar as a Supporting Material in the Photocatalytic Process

Photocatalysis is a process where semiconductors are used to catalyze the generation of reactive species. This type of process involves a light source, a semiconductor (catalyst), and an oxidizing agent. TiO₂ and ZnO are commonly used semiconductors in the degradation of organic matter by the application of photocatalytic oxidation processes, where TiO₂ is the most widely used due to its high stability, easy availability, high sensitivity to UV light, and inexpensive nature [90]. The exposure of TiO₂ to UV radiation causes the generation of electrons (e⁻) within the conduction band (CB) and holes (h⁺) in the valence band (VB). The accumulation of e⁻ in the CB occurs due to the generation of h⁺ in the VB. The generated vacancies react with H₂O to produce OH^•, as presented in Equations (23)–(28) [3]. Photocatalysts can suffer from rapid deactivation, charge carrier recombination, slow kinetics, low photostability, and high requirements of energy [114].

Regardless of these advantages, TiO₂ has a wide bandgap, low electron yield, difficulty to be recycled, and a narrow light absorbance in the UV–visible range [90,115]. Therefore, research has been directed towards the development of a carbonaceous material that can overcome the limitations of TiO₂, and biochar has been proposed as a stable supporting material for the semiconductor [90]. The support of TiO₂ in biochar surfaces enhances their photocatalytic potential, hence increasing the degradation efficiency of the process [116]. When biochar is used as a supporting material for TiO₂, the bandgap of the semiconductor is reduced and the UV–visible absorbance range is increased [115]. Moreover, the large electrical conductivity of biochar enhances the transferring of photogenerated e⁻ and inhibits the recombination of e⁻-h⁺ pairs [92]. Delocalized π e⁻ on the surface of biochar enables efficient charge separation, acting as a bridge for e⁻ tunnelling and reducing the recombination of e⁻-h⁺ pairs. The oxygen functional groups and the PFRs on the surface of biochar contribute to the overall degradation of pharmaceuticals since H₂O₂ is activated [114]. Additionally, the availability of free e⁻ on the surface of biochar increases the availability of e⁻ in the CB to generate charged oxygen and OH^•. Due to the adsorption capacity of biochar, pharmaceutical molecules can get adsorbed in the biochar surface, facilitating the interaction between pharmaceutical molecules and the reactive species, therefore enhancing the degradation efficiency of the process [115,117].

$${\mathrm{T}\mathrm{i}\mathrm{O}}_2+\to {\mathrm{T}\mathrm{i}\mathrm{O}}_2({\mathrm{e}}^{-}+{\mathrm{h}}^{+})$$

(23)

$${{\mathrm{h}}_{\left(\mathrm{V}\mathrm{B}\right)}}^{+}+\mathrm{B}\mathrm{i}\mathrm{o}\mathrm{c}\mathrm{h}\mathrm{a}\mathrm{r} \mathrm{s}\mathrm{u}\mathrm{r}\mathrm{f}\mathrm{a}\mathrm{c}\mathrm{e}-\mathrm{O}\mathrm{H}\to {\mathrm{O}\mathrm{H}}^{•}$$

(24)

$${{\mathrm{e}}_{\mathrm{C}\mathrm{B}}}^{+}+{\mathrm{O}}_2\to {{\mathrm{O}}_2}^{•-}$$

(25)

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

(26)

$${{2\mathrm{H}\mathrm{O}}_2}^{•-}\to {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{O}}_2$$

(27)

$${\mathrm{H}}_2{\mathrm{O}}_2+\mathrm{h}\mathrm{v}\to {2\mathrm{O}\mathrm{H}}^{•}$$

(28)

To support semiconductors on biochar, several methods like sol–gel, impregnation–calcination, and solvothermal have been used; among these, the sol–gel technique is the most widely used method [115]. Lazarotto et al. [118] prepared a composite of biochar and TiO₂ using the impregnation–calcination method, and residues from coffee grounds were used as biomass. The authors mixed biomass with TiO₂ by mechanical manual mixing until complete homogenization was achieved. Afterward, the mixture was pyrolyzed at 650 °C for 2 h in an inert atmosphere. The modification or support of TiO₂ on biochar enhanced the photocatalytic potential of the process and, therefore, the degradation efficiency was improved due to the presence of phenolic groups on the biochar surface, which contributed to e⁻ transfer [118]. Qu et al. [119] supported TiO₂ on rice straw biochar by the sol–gel method, and biomass was pyrolyzed at 400 °C for 1 h to obtain the modified biochar. The biochar was added to a solution of H₃PO₄ for 12 h; subsequently, the mixture was heated with magnetic stirring, passed through a furnace at 800 °C for 1 h, washed with distilled water, and dried. The acid-treated biochar was then added to a solution of TiO₂ and C₂H₆O for 2 h to finally be calcinated at 800 °C for 1 h. The modification of biochar with TiO₂ by the sol–gel method also enhanced the photocatalytic potential of TiO₂ [119].

The solvothermal method is a combination of hydrothermal and dry-heated treatment to facilitate the generation of TiO₂ particles on the surface of biochar [115]. Peñas-Garzón et al. [120] used this method. Biochar was submerged in a C₂H₆O solution at room temperature to create solution A. Solution B was generated by the dilution of titanium butoxide (Ti(OBu)₄) into C₂H₆O, and solution B was added by drops into solution A under continuous stirring to ensure the homogenization of the solutions. To produce the hydrolysis of Ti(OBu)₄, a solution of C₂H₆O and ultra-pure water was added by drops. Subsequently, the mixture was stirred for 5 min and transferred to an autoclave at 160 °C for 3 h. The solid particles were separated by centrifugation and washed with deionized water and C₂H₆O to finally place the modified biochar in a furnace at 60 °C. Figure 3 shows a flow chart of the solvothermal process followed by the authors [120]. With this method, the photocatalytic process was also improved, as well as the ease of recovery of the catalyst [120].

Even though TiO₂ has been widely used in photocatalytic processes, TiO₂ has been reported to produce genotoxic effects, including direct DNA damage, oxidative stress, and interference with DNA repair [121]. Furthermore, Leite et al. [122] found that TiO₂ was able to produce alterations in the digestive glands and gills of the mussel Mytilus galloprovincialis. Moreover, the exposure to TiO₂ caused cellular damage, showing that TiO₂ can generate detrimental effects on aquatic organisms. Therefore, researchers have suggested the use of ZnO instead of TiO₂ due to advantages like large excitation binding energy, high absorbance efficiency in the UV–visible spectrum, and low cost [123]. Additionally, ZnO is an environmentally friendly material that may not pose a risk to living organisms [124]. Amir et al. [123] supported ZnO on biochar derived from the pyrolysis of Calotropis gigantea leaves. To produce the ZnO–biochar composite, the authors used the impregnation–calcination method, in which biochar was submerged within a solution containing ZnO and mixed by sonification. Afterward, the solution was centrifugated to separate the solids from the liquids, and the solids were calcinated at 400 °C. The ZnO–biochar was used in the degradation of ciprofloxacin, obtaining a degradation efficiency of 98.55%, which was higher than that obtained with ZnO nanoparticles (41%). This increase in degradation efficiency was attributed to the functional groups on the biochar surface that enhanced the photocatalytic activity of ZnO and to the synergetic effect of the adsorption–photocatalytic combined process [123]. As observed, the catalytic power of the process was enhanced by the presence of biochar since a higher degradation efficiency was obtained. In this regard, biochar can be described as an efficient, low-cost, and sustainable supporting material for semiconductors like TiO₂ and ZnO.

5.2.4. Heteroatom Doping

Biochar can be modified with heteroatoms like phosphorous (P), nitrogen (N), boron (B), and sulfate (S) by adding them to the biomass before thermochemical decomposition occurs to ensure the distribution of the heteroatom through the surface of biochar [98]. This type of modification can form new active sites on the biochar surface and enhance its surface area and the content of functional groups [125,126]. N heteroatom doping is the most widely studied and has been shown to enhance the electrical conductivity of biochar, which is an important property in the activation of oxidizing agents [88]. For N-doping, the use of the self-doping technique has been reported. This technique consists of the utilization of N-rich biomass (e.g., biomass with a high content of proteins) to ensure the production of biochar with a high content and uniform distribution of N on the surface [98]. On the other hand, Shi et al. [125] prepared a P-doped biochar by immersing cyanobacteria biomass in a solution of H₃PO₄ and subjected it to pyrolysis under a N₂ flow to generate an inert atmosphere. The P-doped biochar was able to activate S₂O₈²⁻ since P improved the electron donor properties of the biochar; therefore, the degradation of acetaminophen was enhanced, and a removal efficiency of 100% with a retention time of 90 min was obtained. Zhang et al. [126] doped corncob biomass with different heteroatom precursors. To obtain N-doped biochar, corncobs were mixed with urea; for S-doped biochar, corncobs and sodium thiosulphate (Na₂S₂O₃) were mixed; and for N-S-doped biochar, biomass was mixed with thiourea. The treated biomass was placed in a high-pressure reactor to conduct hydrothermal carbonization at 240 °C for 2 h to finally be pyrolyzed at 900 °C for 2 h. The three heteroatom modifications conducted by the authors enhanced the activation potential of the biochar. The modified biochar was used in the treatment of a solution containing chlortetracycline, ofloxacin, and sulfadiazine. The degradation efficiency for the non-doped, N-doped, S-doped, and NS-doped biochar were, respectively, 82.38, 93.78, 91.11, and 81.55% [126]. As observed in the two studies mentioned above, the doping of biochar with N, S, and P enhanced the catalytic activity of the carbonaceous material since its electrical conductivity was improved by heteroatom doping. Therefore, the selection of the heteroatom to modify biochar must be guided by economical aspects to increase the applicability of the process.

5.2.5. Biochar as a Cathode in the Electro-Fenton Process

In the electro-Fenton process, the generation of H₂O₂ is an in situ process in which H₂O₂ is formed by an electron oxygen reduction reaction that takes place under the catalytic action of the cathode, as presented in Equation (29) [98]. In this process, Fe in its ferrous state (Fe²⁺) is added externally to start the Fenton reaction (Equation (30)) [12]. Transitions metals like nickel (Ni), Cu, Mn, and Co can also be used in Fenton-like reactions to generate OH^• [127]. The electrogeneration of H₂O₂ in the cathode is a limitation of the application of this type of AOP since it is imperative to develop a low-cost and reliable cathode to produce H₂O₂ and increase the applicability of the process effectively. The application of biochar as a cathode has been gaining attention due to its properties like high surface area and stability, tuneable physicochemical characteristics, and self-doped heteroatoms when it is produced from biomass that is rich in heteroatoms [128]. Indeed, oxygen functional groups on the surface of biochar can provide a higher selection for the two-electron oxygen reduction reaction [127]. Hu et al. [128] produced biochar by the pyrolysis of distillers grains at 800 °C under an inert atmosphere for 2 h; the biochar was immersed in HCl to remove metals, filtered, and washed with distilled water and C₂H₆O to finally be dried at 60 °C. The authors found that at a pH of 2, the maximum yield of H₂O₂ was obtained (15.46 mmol/L). Additionally, the removal of bisphenol A was evaluated using the biochar cathode in the electro-Fenton process. It was found that within 30 min of contact time, the removal of bisphenol A was almost 100%, showing that biochar significantly increased the performance of the electro-Fenton process. Furthermore, the authors found that the process was able to reduce the toxicity associated with bisphenol A to Daphnia magna [128].

$${\mathrm{O}}_2+2{\mathrm{H}}^{+}+{2\mathrm{e}}^{-}\to {\mathrm{H}}_2{\mathrm{O}}_2$$

(29)

$${\mathrm{F}\mathrm{e}}^{+2}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{F}\mathrm{e}}^{+3}+{\mathrm{O}\mathrm{H}}^{•}+{\mathrm{O}\mathrm{H}}^{-}$$

(30)

Biochar with and without modification has been coupled with different AOPs like the persulfate-based oxidation process, photocatalysis with TiO₂, Fenton and Fenton-like processes, the peroxyacetic acid oxidation process, electrochemical AOPs, and ultrasound-assisted photocatalysis to eliminate pharmaceuticals in water [85,86,88,115,129,130]. Table 5 presents studies where biochar has been combined with AOPs for the degradation of pharmaceuticals. As shown in Table 5, biochar used as a catalyst in AOPs is mainly produced by pyrolysis. Considering that the thermochemical method can influence the physicochemical properties of biochar, biochar from other thermochemical decomposition techniques must be evaluated to analyze the influence that thermochemical methods have on the catalytic potential of biochar. When biochar is produced from gasification, bio-gas production is higher compared with other techniques. Therefore, to make an impact on the circular economy, biochar produced from the gasification of biomass to generate energy can be used as a catalyst and recover a sub-product to eliminate pharmaceuticals from wastewater.

In most of the studies compiled in Table 5, the process was conducted under a specific pH value, and the evaluation of the effect of different pH values was not assessed. Since the degradation efficiency of biochar coupled with an AOP is strongly influenced by the adsorption capacity of the carbonaceous material, parameters such as pH must be optimized, as they not only affect the activity and stability of the oxidants but can also affect the form in which the pharmaceutical molecules are present in water and the charge in the biochar surface. Consequently, the adsorption capacity of biochar is affected. It has been found that weak acidic and neutral pH conditions are more beneficial in the degradation of organic compounds like pharmaceuticals in AOP/biochar processes. When biochar has a positive charge on its surface, the adsorption of anions such as S₂O₈²⁻ is favored, promoting the oxidative degradation of pharmaceuticals. Moreover, under alkaline conditions, SO₄^•− and OH^• can react with OH⁻ and slow down the degradation of pharmaceuticals [131]. Therefore, when evaluating the efficiency of AOP/biochar systems, different pH values must be assessed to identify the effect of the solution pH on the degradation efficiency.

Within this combined process, the elimination of pharmaceuticals has been mainly attributed to the radical pathway, as most of the processes presented described the presence of OH^•, SO₄^•−, and O₂^•− as responsible for the degradation of pharmaceuticals. However, some studies reported that the degradation resulted from the action of non-radical pathways like electron transfer and ¹O₂. This can be ascribed to the selectivity of radicals and the affinity of pharmaceutical molecules towards certain oxidant species. To evaluate the influence of radical and non-radical pathways, studies of the inhibition of degradation by various radical scavengers and the formation of complexes between the oxidant and biochar must be conducted. In addition to the affinity of pharmaceuticals to a certain oxidant species, it is worth noting that biochar biomass sources, methods used in its production, and modifications affect the physicochemical properties involved in the activation of oxidants by biochar [91] since these parameters can favor a specific degradation pathway depending on the properties that impact biochar catalytic activity.

Table 5. Pharmaceutical treatment by combining AOPs with biochar.

Pharmaceutical	AOP	Biochar Type and Generation Method	Operational Conditions	Efficiency and Radical Species Involved in the Process	Ref.
Triclosan (TCS)	Peroxyacetic acid (PAA) process	Hydrolyze rice straw Pyrolysis at 700 °C with N2 flow	[TCS] = 10 mg/L [PAA] = 1.3 mmol/L [Biochar] = 2.9 g/L t = 120 min	- Overall, 90% removal. - O2•− was the main radical involved. - 1O2 had a significant influence. - An electron transfer process was produced by the degradation of TCS with O2•−. - PFRs were the primary activation mechanism of PAA.	[129]
SFM	Sulfate-based process	Coconut shell Pyrolysis at 700 °C with N2 flow	[SFM] = 0.126 mg/L pH = 5 [HSO5−] = 0.5 mmol/L [Biochar] = 150 mg/L	- Overall, 85% removal. - SO4•− and OH• were the species involved in the degradation by the radical pathway. - 1O2 was responsible for the non-radical pathway. - SFM and HSO5− were the electron donor and acceptor, respectively. - Biochar acted as an electron transfer mediator.	[132]
Rifadin (RF)	Sonophotocatalytic process	NiCr–biochar	[RF] = 15 mg/L pH = 8 [NiCr–biochar] = 1 g/L Light = 50 W LED US = 150 W t = 90 min	- Overall, 80.3% removal. - OH• was the main radical involved in the degradation, but O2•− was also involved. - The presence of biochar significantly enhanced the degradation efficiency.	[130]
Acetaminophen (ACP)	Sulfate-based process	Cyanobacteria biomass treated with H3PO4 Pyrolys at 500 °C with N2 flow P–biochar	[ACP] = 7.56 mg/L [S2O82−] = 2 mmol/L [P–biochar] = 0.1 g/L [Thiosulfate] = 2 mmol/L t = 90 min	- Overall, ≈ 100% removal. - Electron transfer was the main degradation pathway. ACP was the electron donor, S2O82− was the electron acceptor, and P–biochar was the mediator. - SO4•− and OH• acted as secondary species.	[125]
Sulfadiazine (SZ)	Sulfate-based process	Red mud Pyrolysis at 800 °C with N2 flow	[SZ] = 20 mg/L [Biochar] = 0.2 g/L [S2O82−] = 2 mmol/L pH = 3 t = 20 min	- Overall, 99.7% removal. - SO4•− and O2•− were the main species involved. - 1O2 played a secondary role. - The concentration of zero valence iron (Fe0) in biochar enhanced its catalytic potential.	[133]
Sulfapyridine (SF)	Sulfate-based process	Maize cob modified with Fe Pyrolysis at 600 °C	[SF] = 10 mg/L T = 22 °C [HSO5−] = 1 mmol/L [Biochar] = 0.1 g/L pH = 8.2 t = 30 min	- Overall, 80% removal. - 1O2 was the main species involved in the degradation of SF.	[134]
Tetracycline (TC)	Sulfate-based process	Sewage sludge Pyrolysis under a N2 flow	[TC] = 100 mg/L [S2O82−] = 4.2 mmol/L [Biochar] = 0.2 g/L pH = 2.17 T = 25 °C	- Overall, 82.24% removal. - OH• and SO4•− were the main reactive species involved in the degradation of TC.	[135]
Ciprofloxacin (CIP)	Photocatalytic process	Calotropis gigantea leaves Pyrolysis at 520 °C Biochar modified with ZnO	[CIP] = 100 mg/L [ZnO–biochar] = 1 g/L pH = 7 Light = 3 W LED in the range of 385–750 nm t = 240 min	- Overall, 98.5% removal of CIP. - Overall, 98.1% removal of chemical oxygen demand. - O2•− and OH• were the dominant reactive species. - Supporting ZnO on biochar helped in slowing the recombination of e−-h+ pairs.	[123]
SFM	Sulfate-based process	Red mud and sewage sludge Pyrolysis at 700 °C with N2 flow	[Biochar] = 1.6 g/L [HSO5−] = 0.15 mmol/L [SFM] = 5.07 mg/L T = 25 °C t = 50 min	- Overall, 82.5% removal. - 1O2 was the main specie involved in the degradation of SFM. - Ov, C=O, and graphitic carbon had a role in 1O2 production.	[102]

Table 5. Pharmaceutical treatment by combining AOPs with biochar.

Pharmaceutical	AOP	Biochar Type and Generation Method	Operational Conditions	Efficiency and Radical Species Involved in the Process	Ref.
Triclosan (TCS)	Peroxyacetic acid (PAA) process	Hydrolyze rice straw Pyrolysis at 700 °C with N2 flow	[TCS] = 10 mg/L [PAA] = 1.3 mmol/L [Biochar] = 2.9 g/L t = 120 min	- Overall, 90% removal. - O2•− was the main radical involved. - 1O2 had a significant influence. - An electron transfer process was produced by the degradation of TCS with O2•−. - PFRs were the primary activation mechanism of PAA.	[129]
SFM	Sulfate-based process	Coconut shell Pyrolysis at 700 °C with N2 flow	[SFM] = 0.126 mg/L pH = 5 [HSO5−] = 0.5 mmol/L [Biochar] = 150 mg/L	- Overall, 85% removal. - SO4•− and OH• were the species involved in the degradation by the radical pathway. - 1O2 was responsible for the non-radical pathway. - SFM and HSO5− were the electron donor and acceptor, respectively. - Biochar acted as an electron transfer mediator.	[132]
Rifadin (RF)	Sonophotocatalytic process	NiCr–biochar	[RF] = 15 mg/L pH = 8 [NiCr–biochar] = 1 g/L Light = 50 W LED US = 150 W t = 90 min	- Overall, 80.3% removal. - OH• was the main radical involved in the degradation, but O2•− was also involved. - The presence of biochar significantly enhanced the degradation efficiency.	[130]
Acetaminophen (ACP)	Sulfate-based process	Cyanobacteria biomass treated with H3PO4 Pyrolys at 500 °C with N2 flow P–biochar	[ACP] = 7.56 mg/L [S2O82−] = 2 mmol/L [P–biochar] = 0.1 g/L [Thiosulfate] = 2 mmol/L t = 90 min	- Overall, ≈ 100% removal. - Electron transfer was the main degradation pathway. ACP was the electron donor, S2O82− was the electron acceptor, and P–biochar was the mediator. - SO4•− and OH• acted as secondary species.	[125]
Sulfadiazine (SZ)	Sulfate-based process	Red mud Pyrolysis at 800 °C with N2 flow	[SZ] = 20 mg/L [Biochar] = 0.2 g/L [S2O82−] = 2 mmol/L pH = 3 t = 20 min	- Overall, 99.7% removal. - SO4•− and O2•− were the main species involved. - 1O2 played a secondary role. - The concentration of zero valence iron (Fe0) in biochar enhanced its catalytic potential.	[133]
Sulfapyridine (SF)	Sulfate-based process	Maize cob modified with Fe Pyrolysis at 600 °C	[SF] = 10 mg/L T = 22 °C [HSO5−] = 1 mmol/L [Biochar] = 0.1 g/L pH = 8.2 t = 30 min	- Overall, 80% removal. - 1O2 was the main species involved in the degradation of SF.	[134]
Tetracycline (TC)	Sulfate-based process	Sewage sludge Pyrolysis under a N2 flow	[TC] = 100 mg/L [S2O82−] = 4.2 mmol/L [Biochar] = 0.2 g/L pH = 2.17 T = 25 °C	- Overall, 82.24% removal. - OH• and SO4•− were the main reactive species involved in the degradation of TC.	[135]
Ciprofloxacin (CIP)	Photocatalytic process	Calotropis gigantea leaves Pyrolysis at 520 °C Biochar modified with ZnO	[CIP] = 100 mg/L [ZnO–biochar] = 1 g/L pH = 7 Light = 3 W LED in the range of 385–750 nm t = 240 min	- Overall, 98.5% removal of CIP. - Overall, 98.1% removal of chemical oxygen demand. - O2•− and OH• were the dominant reactive species. - Supporting ZnO on biochar helped in slowing the recombination of e−-h+ pairs.	[123]
SFM	Sulfate-based process	Red mud and sewage sludge Pyrolysis at 700 °C with N2 flow	[Biochar] = 1.6 g/L [HSO5−] = 0.15 mmol/L [SFM] = 5.07 mg/L T = 25 °C t = 50 min	- Overall, 82.5% removal. - 1O2 was the main specie involved in the degradation of SFM. - Ov, C=O, and graphitic carbon had a role in 1O2 production.	[102]

6. Biochar Regeneration Techniques

When adsorbing materials are used coupled with AOPs, the material must present a high regeneration capacity to make a positive impact on the operational cost associated with the application of the process, making it more sustainable since it gives the possibility to use the catalyst in various degradation cycles [98]. The regeneration capacity of biochar and its efficiency in subsequent degradation cycles can be affected by (i) the reduction in the surface area, mass loss, and pore size blocking that results in the loss of active sites; (ii) degradation by-products can get adsorbed on the biochar surface and block the active sites for the activation of oxidizing agents; (iii) the catalytic active sites on the biochar surface may decrease after the degradation process; and (iv) when biochar is doped with metal ions, these can leachate from the biochar surface and decrease the catalytic potential of the material [92]. Studies on the reusability of biochar used in AOPs have shown that after degradation, the surface of biochar is oxidized and the degree of defect structures in the biochar is reduced, decreasing the reusability of this material. In addition, when C=O and O-C=O groups serve as electron-donating species to activated oxidants, lower activity groups such as -OH and C-O-C are generated, reducing the reusability of catalysts [101]. Therefore, practical and efficient methods for regenerating and reusing biochar are essential [6].

Shi et al. [125] evaluated the regeneration capacity of P-doped biochar previously used in the degradation of acetaminophen. After the first degradation cycle, the authors separated the P–biochar from the solution, washed it with deionized water, and dried it at 60 °C for 12 h to use it in another degradation cycle. After three cycles, the degradation efficiency decreased from approximately 100% to 67.4% due to the loss of active sites on the biochar surface capable of activating the oxidant. When biochar is only washed with distilled water, its properties, including the surface area, porous structures, and active sites for catalysis, are not regenerated; therefore, its catalytic potential is decreased. Tong et al. [129] also analyzed the regeneration capacity of biochar; nonetheless, instead of using distilled water, the authors used C₂H₆O to wash biochar after the first degradation cycle of triclosan. With this method, after 10 cycles, a removal efficiency of 77% was reached. The decrease in the efficiency of degradation was attributed to the obstruction of the catalytic active sites on the biochar surface by the degradation by-products. Regardless of the decrease in the degradation efficiency, the biochar exhibited high stability, allowing it to retain most of the catalytic sites during the degradation and regeneration process [129]. The difference in the efficiency of distilled water and C₂H₆O to regenerate biochar can be linked to C₂H₆O being a solvent with a higher affinity towards the by-products generated during the process, making it able to desorb the molecules from the biochar surface, unblocking the catalytic active sites.

Thermal treatment is another method that has been described for the regeneration of biochar used in AOPs. Zhu et al. [136] regenerated biochar previously used as a catalyst in the degradation of organic compounds by a sulfate-based AOP with a pyrolysis treatment. After the first cycle was conducted, the biochar was separated from the solution and pyrolyzed at 300 °C. The degradation efficiency decreased from 97.8% to 90.1% in the third degradation cycle. This led to the conclusion that thermal regeneration can remove the degradation by-products formed, unblocking the catalytic active sites of the biochar. Thermal treatment has been described as the most efficient method to regenerate biochar since it can restore its surface area, as well as its porous structure, by eliminating the by-products contained on the biochar surface. However, thermal treatment can increase the cost of the application of the process since a large amount of energy is needed [92]. In this regard, different regeneration methods must be evaluated to discern between their advantages and disadvantages. Regeneration methods should be selected considering overcoming several challenges like secondary pollution and high costs associated with regeneration methods [10].

7. Future Perspectives of Coupling AOPs with Biochar in the Elimination of Pharmaceuticals

Parameters such as the solution pH and the biochar particle size affect the adsorption of pollutants into biochar due to their effect on the surface area and active sites. Given the importance of adsorption when biochar is coupled with AOPs, it is necessary to assess the effects of these factors when biochar is used as the catalyst. Therefore, further research must be performed to optimize these parameters in a hybrid process.

It was observed that in most studies where biochar is coupled with AOPs in the elimination of pharmaceuticals, the concentrations of the studied pollutant were above the trace level. Considering that pharmaceuticals are present in trace and ultra-trace levels in water, even in pharmaceutical industry wastewater effluents, it is imperative to evaluate the efficiency of the process working at the concentration in which pharmaceuticals are found in wastewater and the environment to have a representative study of the process efficiency.

Research evaluating the coupling of AOPs with biochar in real wastewater must be conducted to study the influence of real water matrix constituents and increase the applicability of the combined process. These constituents can block biochar active sites and decrease their ability to activate oxidizing agents. When using a hybrid biochar/AOP system, the degradation of pharmaceuticals is expected to take place on the surface of biochar; therefore, other substances contained in the water matrix can compete with pharmaceuticals to get adsorbed on the biochar surface, decreasing the degradation efficiency of the process. In this regard, real scenarios must be evaluated to understand the action mechanisms of biochar/AOPs in real wastewater and increase system feasibility. Additionally, the process must be evaluated on an industrial scale since most of the studies have been conducted on a laboratory scale.

On the other hand, future studies must be focused on the regeneration of biochar to increase its applicability. The development of an efficient regeneration method that allows the use of biochar as a catalyst in AOPs in different degradation cycles not only increases the applicability of the process but also creates a higher impact on sustainable development and circularity since the amount of the catalyst that must be disposed of is decreased.

The final disposal of the expended catalyst is an aspect that must be also considered. With the possibility of degradation by-products occupying the surface of biochar, it is important to assess different disposal methods that reduce the risk of these degradation by-products entering the environment and posing a threat to living organisms. In addition, and in order to ensure the sustainable use of biochar in environmental remediation, measurements must be taken to control the release of toxic by-products from biochar [137]. Therefore, research must be directed towards the final disposal of expended catalysts.

8. Conclusions

Pharmaceuticals are toxic substances that can be released into aquatic environments through different sources. Their presence in the environment has become a risk to humans and animals due to their continuous release derived from anthropogenic activities, where wastewater is the main source for pharmaceuticals to enter the environment. Therefore, their removal from wastewater is essential to reduce the risk to living organisms.

One of the treatments that have been proposed for the treatment of pharmaceuticals presenting high degradation efficiencies is the application of AOPs. Nevertheless, their implementation is limited by the high energy and reactant consumption costs. Therefore, the coupling of AOPs with biochar is an alternative to reduce the costs associated with the application of AOPs and increase the performance of the process.

Biochar physicochemical properties are described as the mechanisms for the activation of oxidants in AOPs such as H₂O₂, O₃, S₂O₈²⁻, and HSO₅⁻. Biochar oxygen-containing functional groups, PFRs, and carbon defects are important in the degradation of pharmaceuticals through radical pathways. This means the degradation of these pollutants by the action of reactive oxygen species (e.g., OH^• and SO₄^•−). On the other hand, biochar electrical conductivity plays a crucial role in the non-radical degradation pathways through processes like electron transfer, in which biochar acts as the electron conductor to oxidize pharmaceuticals. Additionally, biochar adsorption capacity has been shown to play a synergetic role when coupled with AOPs. The efficient adsorption of pharmaceutical molecules by biochar aids in the degradation of the pollutant by increasing the possibilities of collision between the reactive species generated from the oxidant and the pharmaceutical molecule, raising the degradation efficiency of the process. In this regard, the adsorption capacity of biochar is an aspect that needs special consideration when coupling it to AOPs.

Considering the influence of the physicochemical properties of biochar on its catalytic performance, further studies need to be focused on the optimization of parameters that influence biochar properties, such as the feedstock, temperature, heating rate, and residence time. The optimization of these factors may assist in the development of biochar with high catalytic activity, thereby increasing its applicability. It is highlighted that biochar physicochemical properties can be enhanced by various modifications. Modifications with metals, acids, bases, and heteroatoms can improve the catalytic potential of biochar by increasing the activation of oxidants. Furthermore, the capacity of biochar to be used as a supporting material makes it able to be used in processes like photocatalytic AOPs, increasing the efficiency of pharmaceutical degradation.

When heterogeneous catalysts like biochar are used in AOPs, the regeneration capacity of the catalyst is a significant property that can affect the performance of the catalyst. The regeneration of biochar when used in AOPs needs further research to develop a regeneration method that not only presents a low cost of regeneration but also decreases the chance of generating secondary pollution. Therefore, the regeneration of biochar is a research gap that must be covered to increase the industrial applicability of the biochar/AOP hybrid process.