Biological Treatments for VOC-Contaminated Off-Gas: Advances, Challenges, and Energetic Valorization Opportunities

Abstract

Volatile organic compounds (VOC) are major contributors to the burgeoning air pollution issue, predominantly from industrial areas, with well-documented environmental and health risks, which demand efficient and sustainable control policies. This review analyzes the current technological challenges and investigates recent developments in biological treatment technologies for VOC-contaminated off-gases, including biofilters, biotrickling filters, and bioscrubber, as well as emerging technologies, such as bioaugmentation and microbial fuel cells (MFCs). Operational performance, economic feasibility, and adaptability to various industrial applications are assessed, alongside opportunities for integration with other technologies, including energy recovery technologies. Biological systems offer considerable advantages regarding cost savings and lower environmental impacts and enhanced operational flexibility, particularly when combined with innovative materials and microbial optimization techniques. Nevertheless, challenges persist, such as choosing the best treatment settings suited to different VOC streams and addressing biofilm control concerns and scalability. Overall, biological VOC treatments are encouraging sustainable solutions, though continued research into reactor design, microbial dynamics, and MFC-based energetic valorization is essential for broader industrial application. These insights cover advancements and highlight the continuous need for innovative prowess to forge sustainable VOC pollution control.

1. Introduction

As environmental concerns regarding gaseous pollutants continue to grow, research into technologies for controlling polluted gases is becoming increasingly advanced and comprehensive. Volatile organic compounds (VOC) are one of the major ongoing research areas due to their known significance in environmental and health impacts, necessitating effective remediation strategies. VOC are organic chemical compounds that contain carbon and other elements, such as hydrogen, oxygen, nitrogen, chlorine, fluorine, bromine, or sulfur [1]. According to the Directive 2010/75/EU of the European Parliament and of the Council of 24 November, a VOC is defined as any organic compound as well as the fraction of creosote, having at 293.15 K a vapor pressure of 0.01 kPa or more, or having corresponding volatility under the particular conditions of use [2]. The National Emission reduction Commitments Directive (NECD) gives a different definition: non-methane volatile organic compound (NMVOC) means all organic compounds other than methane that are capable of producing photochemical oxidants by reaction with nitrogen oxides in the presence of sunlight [3]. Due to their high vapor pressure, these compounds are extremely volatile and are precursors of photochemical pollutants, since they originate oxidant pollutants by reacting with nitrogen oxides (NO_X) in the presence of sunlight. Depending on their structure and life span (from minutes to years), VOC can have different influences on the chemistry of the atmosphere, human health, and fauna [1,4,5,6,7,8,9,10]. Additionally, some show a high propensity to form oxidants, while others tend to be reasonably unreactive [11].

VOC can alter the chemistry of the atmosphere, decrease air quality, and be responsible for the formation of tropospheric ozone, peroxyacetyl nitrate (CH₃COO₂NO₂ or PAN), and secondary organic aerosols (SOA), as well as photochemical smog [1,10,12,13,14,15,16]. Therefore, VOC constitute potential health risks and a potential environmental hazard. It is well established that VOC and NO_X react to form tropospheric ozone in the presence of sunlight [17,18]. Ozone formation is a secondary contaminant that occurs between ground level and about 10–12 km (lower atmosphere or troposphere) [19]. Ozone is a precursor of OH radicals, causing an indirect effect on Earth’s climate by limiting the atmospheric lifetime of greenhouse gases such as methane [20]. Furthermore, crop productivity and vegetation can be directly affected by ozone [19,20]. SOA also affect Earth’s climate system by directly scattering and absorbing sunlight or acting as condensation nuclei for cloud droplets [20,21].

Sunlight in the near-UV can promote a cycle of reactions that generate free radicals derived from the oxidation of VOC in a complicated relationship between the photochemical reactions of ozone and its precursors. These reactions typically start with the abstraction of a hydrogen atom by OH radicals, forming water and an alkyl radical $({\mathrm{R}}^{•}$) (Equation (1)). The formed alkyl radical reacts rapidly with molecular oxygen to form the peroxyl radical (${\mathrm{RO}}_2^{•}$) (Equation (2)) [22].

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

(1)

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

(2)

Under specific conditions, chlorine (Cl) radicals can effectively initiate the oxidation of VOC. These radicals are often more reactive than hydroxyl radicals for many VOC, particularly alkanes, which are less reactive with OH [23,24]. This makes Cl radicals especially effective in oxidizing less reactive alkanes during early morning hours when OH radical levels are typically low [25]. A key source of Cl radicals is the photolysis of ClNO₂, which can significantly contribute to the formation of secondary pollutants (Equation (3)) [23,25]. In high-NO_X environments, such as urban areas, Cl radicals play a crucial role in radical propagation, which tends to dominate under these conditions [24]. The reaction pathways include hydrogen abstraction (Equation (4)) or addition to double bonds (Equation (5)), both of which lead to the formation of organic peroxyl radicals. These intermediates undergo subsequent reactions that are integral to the formation of SOA [25].

$${\mathrm{ClNO}}_2+\mathrm{hv} \to {\mathrm{Cl}}^{•}+{\mathrm{NO}}_2 (\mathrm{primary} \mathrm{source} \mathrm{of} \mathrm{chlorine} \mathrm{radicals})$$

(3)

$$\mathrm{RH}+{\mathrm{Cl}}^{•} \to {\mathrm{R}}^{•}+\mathrm{HCl}$$

(4)

$$\mathrm{RCH}=\mathrm{CH}2+{\mathrm{Cl}}^{•}\to {\mathrm{RCHCl}}^{•}-{\mathrm{CH}}_2$$

(5)

Peroxyl radicals are key intermediates in atmospheric chemistry, engaging in several crucial reactions that affect the formation of NO_X and ozone (O₃), the cycling of radicals, and the formation of aerosols. When peroxyl radicals react with nitrogen oxide (NO), they form nitrogen dioxide (NO₂) and an alkoxy radical (RO), which are significant in the context of urban air pollution and ozone layer dynamics (Equations (6) and (7)). Additionally, the self-reaction of peroxyl radicals can lead to a variety of products (such as ethers, peroxides, alcohols, and carbonyl compounds), influencing the composition and characteristics of secondary organic aerosols (Equation (8)). Another important reaction occurs when peroxyl radicals react with hydroperoxyl radicals (${\mathrm{HO}}_2^{•}$), a prevalent atmospheric reaction contributing to the formation of hydroperoxides, which play a role in cloud formation processes and the oxidative capacity of the atmosphere (Equation (9)) [26,27].

$${\mathrm{RO}}_2^{•}+ \mathrm{NO} \to {\mathrm{RO}}^{•}+{ \mathrm{NO}}_2$$

(6)

$${\mathrm{RO}}_2^{•}+ \mathrm{NO} \to \mathrm{OVOC}+{\mathrm{H}}_2\mathrm{O}+{ \mathrm{NO}}_2$$

(7)

$${\mathrm{RO}}_2^{•}+{\mathrm{RO}}_2^{•} \to \left(\mathrm{Various} \mathrm{products}\right)$$

(8)

$${\mathrm{RO}}_2^{•}+{ \mathrm{HO}}_2^{•} \to \mathrm{ROH}+{\mathrm{O}}_2$$

(9)

NOx interactions will react in a photochemical loop that contributes to tropospheric ozone formation (Equations (10)–(12)). These will occur in a photo equilibrium between NO, NO₂, and O₃ with no net formation or loss of O₃. However, in the presence of VOC in the atmosphere, it increases NO₂ concentrations and breaks the photochemical balance of O₃ [28].

$${\mathrm{NO}}_2+\mathrm{hv} \to \mathrm{NO}+\mathrm{O}{{(}^3\mathrm{P})}^{•}$$

(10)

$$\mathrm{O}{{(}^3\mathrm{P})}^{•}+{ \mathrm{O}}_2 \stackrel{\mathrm{Air}}{\to }{\mathrm{O}}_3$$

(11)

$$\mathrm{NO}+{ \mathrm{O}}_3 \stackrel{\mathrm{Air}}{\to }{\mathrm{NO}}_2+{\mathrm{O}}_2$$

(12)

Furthermore, the interactions between NO, NO₂, and O₃ can lead to the formation of N₂O₅, as illustrated by the reactions in Equations (12)–(14). This compound plays a significant role in atmospheric chemistry, particularly during the nighttime. In the absence of sunlight, it acts as a reservoir for nitrogen oxides. Once N₂O₅ is present in the atmosphere, it can react with water vapor or on moist surfaces to produce nitric acid (HNO₃), contributing to the formation of aerosols and particulate matter, leading to acid rain formation [29].

$${\mathrm{NO}}_2+{ \mathrm{O}}_3 \to {\mathrm{NO}}_3^{•}+{\mathrm{O}}_2$$

(13)

$${\mathrm{NO}}_2+{ \mathrm{NO}}_3^{•} \to {\mathrm{N}}_2{\mathrm{O}}_5$$

(14)

Peroxyacetyl nitrate is a typical secondary photochemical pollutant produced reversibly by the reaction of the peroxyacetyl (CH₃C(O)OO or PA) with NO₂ (Equation (15)) [30]. The dominant sources of PA formation are VOC, specifically oxygenated volatile organic compounds such as acetaldehyde, methylvinyl ketone, methyl glyoxal, and biacetyl [31]. It is the principal tropospheric reservoir for nitrogen oxide radicals, it regulates the spatial distribution of O₃ production, and, in low temperatures, it can transport air pollutants over greater distances [31,32]. This means that it is not only responsible for photochemical smog, but also for playing a significant role in other environmental and health aspects due to its role in air quality.

$${\mathrm{CH}}_3\mathrm{C}\left(\mathrm{O}\right){\mathrm{OO}}^{•}+{ \mathrm{NO}}_2\stackrel{\mathrm{Air}}{\leftrightarrow }\mathrm{PAN}$$

(15)

Different VOC will have different affinities for forming photochemical oxidants, with volatile aromatic hydrocarbons, toluene, benzene, ethylbenzene, and xylenes ((m, p)-xylene and o-xylene) being the main contributors to tropospheric ozone formation [20]. Additionally, the chemical processes that degrade VOC in the atmosphere are influenced by factors such as emission rates and strength, available oxidants (like ozone and hydroxyl radicals), temperature, humidity, UV radiation, and wind [28,33,34]. Both natural and human sources significantly influence air quality and climate through their transformation into SOA. These transformations, mediated by interactions with atmospheric oxidants like hydroxyl radicals, ozone, and nitrogen oxides, are crucial in forming tropospheric ozone and peroxyacetyl nitrate, key pollutants with direct health impacts [35]. A representation of the key atmospheric chemical interactions and processes is presented in Figure 1.

The organic chemicals from which VOC arise have variable lipophilicity and volatility. The formed particles have small molecular sizes and no charge, making inhalation the main route of exposure and providing ready absorption across the lungs, gastrointestinal tract, and skin [1]. Symptoms associated with VOC can have short- and long-term adverse health effects. Exposure in both animals and humans led to the detection and formulation of different health concerns, such as headaches; sleepiness; skin, throat, and eye irritation; sensitization; inflammations; and several more severe problems related to mutagenicity; teratogenicity; central nervous system effects; asthma; carcinogenicity; anaphylactic and cardiovascular diseases; and liver and kidney pathologies [1,7,36,37,38].

Exposure to VOC can have significant adverse effects on various organisms. In birds, these compounds caused neurotoxicity, a mild form of ataxia, and increased blood osmolality, particularly affecting developing individuals. In aquatic environments, significant harmful impacts on aquatic organisms interfere with their physiological mechanisms, growth, reproduction, and survival [39]. In water, VOC can lead to the death of invertebrate species, algae, fish, and amphibians, affecting the reproduction of certain fish species and impacting the biomarkers of endocrine disruption [40,41]. Mortality, genotoxicity, and epigenotoxicity have been observed in various species due to exposure to specific VOC. For instance, both benzene and toluene have been found to rapidly induce cell death in blue-green and green algae, preventing the formation of any adaptive structures or responses, such as spores or dormant cells [42]. Thaysen et al. [43] found high mortality and a reduction in reproductive output in Ceriodaphnia dubia exposed to expanded polystyrene cup leachates that contain, among other VOC, styrene, ethylbenzene, toluene, and benzene. Oncorhynchus mykiss exhibited both genotoxicity and epigenotoxicity upon exposure to naphthalene [44]. Similarly, Daphnia magna was affected by Isothiazolinone biocide [45]. Pseudokirchneriella subcapitata and Pimephales promelas growth were hindered by styrene and perchloroethylene, respectively [46,47]. Plants, such as Raphanus sativa, Brassica campestris, Lilium longiflorum, Petunia hybrida, Lepidium sativum, Lactuca sativa, and Phaseolus vulgaris, exhibited symptoms like epinasty, chlorosis, wilting, and reduced dry weight, leaf area, shoot growth, and flower production. Some plants also showed increased seed weight, indicating varying physiological impacts [39]. These effects underscore the widespread ecological and biological harm caused by VOC exposure.

On the other hand, some microorganisms can be more resilient and even effective in degrading VOC compounds, highlighting the importance of studying different microorganism strains in biological VOC degradation. Pseudomonas have been highlighted as one of the most important in VOC degradation. P. putida and P. mendocina are able to degrade aromatic hydrocarbons such as toluene [48,49]. Pseudomonas veronii can degrade a variety of simple aromatic organic compounds [50]. Variovorax paradoxus, Pseudoxanthomonas spadix, Bacillus amyloliquefaciens, Comamonas sp., Ralstonia sp., Microbacterium esteraromaticum, and Bacillus pumilus can be used in the bioremediation of BTEX (benzene, toluene, ethylbenzene, and xylene) [51,52,53,54,55,56,57]. Dehalococcoides is the only known bacterium capable of degrading dichloroethane and vinyl chloride, as it is a key degrader of chlorinated ethenes [52]. Comamonadaceae was reported to be useful in the removal of toluene [58]. Rotifers such as Brachionus calyciflorus and Brachionus plicatilis have been found to be less sensitive to toluene, hexane, xylene, and benzene [59].

VOC deteriorate the water quality index of the water bodies, show biomagnification in the food chain, and impair the physiology of living organisms. However, the leading cause of VOC emissions concerns the ozone and secondary organic aerosol formed due to atmospheric photochemical reactions. Non-methane hydrocarbons (NMHC) are a specific group of VOC species that play a critical role in the formation of tropospheric ozone, while aromatic NMHC promote the formation of SOA [20,60].

VOC emissions in industries can be divided into organized and unorganized. The first one refers to emissions that are discharged through controlled points such as stacks or vents. They are typically from fixed sources. The latter are more diffuse and not discharged from a specific single point. These are often spread over a larger area and can include leaks or other releases that are not captured by organized emission control systems. This variation in the source and containment of VOC emissions significantly influences their environmental impact and the associated health risks [61,62]. The cancer risk (CR) associated with VOC emissions also varies significantly between the industrial sectors responsible for their emissions, reflecting the concentration and toxicity of the emitted compounds. The European Chemicals Agency (ECHA) refers to the thresholds for lifetime CR, applicable across different exposure routes (inhalation, oral, dermal), including the inhalation of airborne chemicals like VOC within an acceptable risk range of 1 × 10⁻⁵ to 1 × 10⁻⁴ for workers, and numerical thresholds for the general population (consumers/environment) of 1 × 10⁻⁶ or lower [63]. On the other hand, the Environmental Protection Agency (EPA) recommends a target CR of 1 × 10⁻⁶ to 1 × 10⁻⁴ for ambient air pollutants, including VOC [64,65]. The CR for VOC is usually divided into four levels by several authors: high risk (≥10⁻³), medium risk (10⁻⁴–10⁻³), low risk (10⁻⁶–10⁻⁴), and negligible risk (<10⁻⁶) [66,67,68,69,70].

According to a study from eastern China, which collected samples from 57 sites (24 organized sites and 33 unorganized sites) of 20 representative factories in the eight industries of the chemical industry, metal smelting, textile printing and dyeing, furniture manufacturing, industrial coating, paint manufacturing, the pharmaceutical industry, and packaging and printing. The order of total VOC emissions from highest to lowest among the eight industries is as follows: packaging and printing, pharmaceutical, paint manufacturing, industrial coating, the chemical industry, metal smelting, furniture manufacturing, and textile printing and dyeing. When analyzing the cancer risks associated with VOC emissions across the various industries, a qualitative assessment revealed a distinct risk level. The pharmaceutical and furniture manufacturing sectors exhibit high cancer risks from organized emissions, indicating significant potential health impacts due to substances like cyclohexane, naphthalene, and alkanes. Conversely, the chemical industry displays a medium cancer risk for organized emissions due to chemicals like chloroform and carbon disulfide, while its unorganized emissions fall into the low-risk category. Similarly, the industrial coating, packaging and printing, and paint manufacturing sectors all demonstrate low cancer risks for both organized and unorganized emissions, which are attributed to lesser concentrations of compounds such as benzene homologs, ethyl acetate, and isopropanol. The metal smelting industry, with organized emissions linked primarily to carbon disulfide, also falls into the low-risk category. This classification helps in prioritizing regulatory focus and mitigation strategies, underscoring the need for tailored interventions in industries with higher risk profiles [71]. Massolo et al. [72] studied several VOC species in indoor and outdoor samples from different types of environments (urban, industrial, semirural, and residential areas) of the region of La Plata (Argentine) and calculated the lifetime cancer risk (LCR) associated with benzene exposure for children from the different study areas. The higher LCR of both outdoors and indoors was set to be in the industrial area [72]. In fact, most studies reveal that risks are lower in urban areas but can still negatively impact human health. Hong et al. [73] indicated a CR exceeding the permissible risk level of 10⁻⁶ in the presence of 1,3-butadiene, benzene, and carbon tetrachloride in the high-density areas of Seoul, Incheon, and Gyeonggi (South Korea). Tao et al. [74] found acrolein, trichloromethane, 1,3-butadiene, 1,2-dichloroethane, and benzene in an industrial area of Nanjing (China) surpassing the acceptable lifetime CR. 1,2-dichloroethane was found to be consistently exceeding the CR threshold in Kaohsiung Harbor (Taiwan) and Map Ta Phut (Thailand) industrial areas [69,75]. In Romania, a study conducted in an oil refining complex found elevated lifetime CR values observed among workers, surpassing the internationally recognized acceptable thresholds [76]. These highlight the variability in CR across industries, underscoring the critical need for industry-specific mitigation strategies to reduce exposure to carcinogenic VOC and to enhance regulatory measures to mitigate VOC emissions. According to Malakan et al. [77], just a 10% reduction in ambient benzene levels can decrease mortality and morbidity in the population exposed to air pollution in the area of one of the largest petrochemical complexes in Thailand. This reduction in health impacts translates directly into economic benefits, both in direct medical costs saved and indirect benefits such as fewer workdays lost due to illness.

Due to the substantial contribution of VOC to environmental and health problems, effective and sustainable remediation techniques are required. Although diverse VOC control technologies have been critically evaluated in previous studies, there is still a fundamental need for systematic synthesis, specifically for recent trends in biological treatment processes, their operational difficulties, and their potential for integration with sustainable energy recovery strategies, particularly for microbial fuel cells (MFC). Consolidating current knowledge, identifying gaps, and putting forward creative sustainable strategies are crucial given growing environmental awareness and stricter regulatory frameworks. In order to close this gap, this review critically examines the state of the art and prospects for biological treatments, such as MFC technologies, biofilters (BFs), biotrickling filters (BTFs), bioscrubbers (BSs), bioaugmentation strategies, and biofilters, for VOC-contaminated off-gas streams. This review addresses major technological and economic challenges, recognizes advancements in biofilm optimization and advanced materials, and discusses potential directions to integrate energetic valorization strategies to improve process sustainability.

2. Industrial VOC Gaseous Emissions

The introduction of VOC in the atmosphere can occur via both biogenic and anthropogenic emissions [14,78]. Biogenic emissions refer to VOC biosynthesis. For example, ethanol, butanol, and acetone can be produced by microorganisms, causing high local concentrations [78]. Biogenic emissions can interact with anthropogenic emissions, influencing the climate and air quality by facilitating biogenic SOA and ozone formation [79,80].

Organic solvents are widely used in various processes, producing VOC as unsought by-products. Several examples of VOC-producing industries can be enumerated: the manufacture of cosmetics, plastics, liqueurs, resins, gums, coatings, metal production, adhesives, sealants, asphalt compounding, motor vehicle assembly, paint dilution in the furniture industry, and the burning of gasoline, wood, coal, or natural gas [1,13,78,81]. Wu et al. [13] conducted a study to assess the potential of ozone and SOA formation driven by anthropogenic VOC emissions in China. In their study, the researchers point out that industry solvent use contributed to 24.7% of ozone formation and 22.9% of SOA formation. Industry process and domestic combustion also represent a considerable slice, representing 23.0% and 17.8% of ozone and 34.6% and 19.6% of SOA formation, respectively.

VOC emissions vary significantly across industries, both in terms of composition and associated health risks. In the pharmaceutical industry, emissions are dominated by cyclohexane, naphthalene, and ethyl acetate from organized sources (systematically emitted), with tetrahydrofuran and methyl methacrylate contributing to unorganized emissions (unintended releases). The chemical industry releases chloroform, carbon disulfide, and ethyl acetate, while halohydrocarbons and alkanes are more prominent in unorganized emissions. The textile printing and dyeing sector primarily emits ethylene, carbon disulfide, alkenes, and oxygenated volatile organic compounds (OVOC). Furniture manufacturing is a major contributor, with emissions characterized by naphthalene, butane, and isobutane posing a high carcinogenic risk. Similarly, industrial coating processes release benzene homologs, xylene, and ethylbenzene. The packaging and printing industry emits VOC like ethyl acetate and isopropanol, comparable to those released during paint manufacturing. In metal smelting, VOC emissions primarily consist of carbon disulfide and ethyl acetate [71]. These underline the diversity and complexity of VOC emissions, emphasizing the need for targeted mitigation strategies tailored to each industry’s emission profile and its associated risks. In

Table 1. Industrial anthropogenic sources of VOC in the atmosphere.

Industry	Volatile Organic Compound	References
Pharmaceutical	Halogenated Alkanes: Dichloromethane, Chloroform, 1,2-Dichloropropane, 1,1-Dichloropropanone, Dichloronitromethane, Cyanogen Chloride, Chloroethane; Halogenated Aromatics: Chlorobenzene, Bromobenzene; Halogenated Alkenes: Tetrachloroethylene; Halogenated Methanes: Bromoform; Esters: Ethyl Acetate; Ketones: Acetone; Aromatic Hydrocarbons: Toluene	[82,83]
Thermal pyrolysis of high-value products from organic residues	Chickpea Aromatic Hydrocarbons: Methylbenzene; Aldehydes: 2-Methylbutanal, 3-Methylbutanal; Furans: 2-Methylfuran; Alkanes: Octane; Alkenes: 2-Octene; Sulfur Compounds: Dimethyl disulfide; Nitriles: 4-Methylpentanenitrile; Cycloalkenes: 1,3-Dimethyl-1-cyclohexene	[84,85]
	Soybeans Aromatic Hydrocarbons: Methylbenzene, Ethylbenzene; Nitrogen Compounds: 1-H-Pyrrole, 1-Methyl-1-H-pyrrole; Phenols: Phenol; Aldehydes: 2-Methylbutanal, 3-Methylbutanal; Furans: 2-Methylfuran
Cigarette	Alcohols: Ethanol, Methanol; Ketones: Acetone; Aromatic Hydrocarbons: Toluene; Phenols: Phenol; Alkanes: Butane; Aldehydes: Formaldehyde	[86]
Tire	Alkanes: n-Pentane, 3-Methylhexane, Cyclohexane; Alkenes: Isoprene; Halogenated Hydrocarbons: Methylene Chloride, Carbon Tetrachloride, 1,2-Dichloropropane, Dichlorodifluoromethane; Aromatic Hydrocarbons: m,p-Xylene, Toluene, Ethylbenzene, Styrene	[87]
Rubber and plastic products	Halogenated Hydrocarbons: Dichloromethane, 1,2-Dichloroethylene, 1,2-Dichloropropane, Carbon Tetrachloride, Chlorobenzene; Sulfur Compounds: Carbon Disulfide; Aromatic Hydrocarbons: Benzene, Styrene, Naphthalene, Ethylbenzene, Xylene, Toluene, 1,2,4-Trimethylbenzene, 1,3,5-Trimethylbenzene; Ketones: 4-Methyl-2-Pentanone, Acetone, 2-Hexanone; Esters: Methyl methacrylate, Ethyl acetate; Alcohols: Isopropanol, Ethanol; Alkanes and Isomers: Heptane and its isomers, 2-Methylpentane, 3-Methylpentane; Alkenes: 1,3-Butadiene, 1-Decene, 1-Dodecene; Ethers: Tetrahydrofuran, Methyl tert-butyl ether	[62,88,89]
Petrochemical	Alkanes and Isomers: Ethane, Propane, 2,3-Dimethylbutane, 2-Methylpentane, 2,2-Dimethylbutane, n-Undecane, Dodecane; Alkenes: Ethylene, Propylene; Aromatic Hydrocarbons: Benzene, Toluene, m,p-Xylene, o-Xylene; Halogenated Hydrocarbons: Chloroform; Alcohols: Ethanol, Isopropanol; Ketones: Acetone, 2-Hexanone	[88,90]
Glass	Aromatic Hydrocarbons: Benzene, Toluene, Ethylbenzene, m/p-Xylene, o-Xylene, 4-Ethyltoluene, 1,2,4-Trimethylbenzene, 1,3,5-Trimethylbenzene; Alkanes: n-Hexane, n-Heptane; Alkenes: 1,3-Butadiene, 1-Decene, 1-Dodecene; Halogenated Hydrocarbons: Carbon Tetrachloride, 1,2-Dichloroethane, Tetrachloroethylene, Chlorobenzene; Esters: Ethyl Acetate; Alcohols: Isopropanol; Ketones: Acetone; Ethers: Tetrahydrofuran; Sulfur Compounds: Carbon Disulfide	[91]
Fiber board	Halogenated Hydrocarbons: Dichloromethane; Epoxides: 2,3-Epoxy-2-methylbutane; Alkanes: Butane; Alcohols: Ethanol, 1-Butanol, 1,2-Propanediol; Esters: Ethyl acetate, 2-Methylacrylic acid methyl ester; Aldehydes: Acetaldehyde; Furans: 3-Methyl-2(5H)-furanone; Ethers: Tetrahydrofuran, 1,4-Dioxane; Amino Acids: N-Formylglycine; Carboxylic Acids: Acetic acid; Amides: N,N-Dimethylformamide	[92]
Furniture	Coating Aromatic Hydrocarbons: Benzene, Toluene, Xylene (m-, p-, and o-xylene), Ethylbenzene, Styrene, p-Dichlorobenzene, 1,2,4-Trimethylbenzene, Chlorobenzene; Aldehydes: Formaldehyde, Acetaldehyde, Nonanal, Decanal; Esters: Butyl acetate, Dimethyl adipate, Dimethyl glutarate, Dimethyl succinate, Ethyl acetate; Ketones: Acetone; Alcohols: Ethanol, Isopropanol, Benzyl alcohol, Methanol; Alkanes: Hexane, Heptane; Alkenes: Propylene, Butylene; Halogenated Hydrocarbons: Dichloromethane, Trichloromethane, Tetrachloroethylene	[93,94,95]
	Adhesives Alkanes and Isomers: 2-Methylhexane, 3-Methylhexane, n-Hexane, n-Heptane, 3-Methylpentane, 3,4-Dimethylheptane; Cycloalkanes and Isomers: Methylcyclopentane, Ethylcyclopentane, Methylcyclohexane, 1,2,4-Trimethylcyclopentane, 1,2,3-Trimethylcyclopentane, 1,3-Dimethylcyclopentane, 1-Methylcyclohexane; Alcohols: Propylene Glycol, 1-Butanol, 5-Methyl-2-(1-methylethyl)hexan-1-ol, 4,8-Dimethylnonan-1-ol, 6-Methyloctan-1-ol, 1-(2-Butoxy-1-methylethoxy)propan-2-ol, Dihydro-α-Terpineol; Glycols: 2-(2-Butoxyethoxy)ethanol, 2-(2-Butoxyethoxy)ethanol acetate; Esters: 2-(2-Butoxyethoxy)ethyl acetate, Butyl acetate, Ethyl acetate; Ketones: Acetone; Aldehydes: Formaldehyde, 2-Hydroxybenzaldehyde; Halogenated Hydrocarbons: 2-Chlorohexane; Ethers: 1-Methoxypropan-2-ol; Terpenes: Alpha-Pinene; Carboxylic Acids: Acetic Acid
Painting	Engine manufacturing Aromatic Hydrocarbons: Benzene, (1-Methyldodecyl)-Benzene, 1,3-Dimethyl-p-Xylene, Ethylbenzene, o-Xylene, Toluene; Phenols and Derivatives: Phenol, 4,4′-(1-Methylethylidene)bis-, Phenol derivatives (e.g., 2,4′-Isopropylidenedi-phenol); Alcohols: Benzyl alcohol, Ethanol, 2-(2-Butoxyethoxy)-; Aldehydes: Decanal, Nonanal; Alkanes: Dodecane, Hexadecane, Octadecane, Tetradecane, Undecane, Eicosane; Cycloalkanes: Cyclotetradecane; Carboxylic Acids: n-Hexadecanoic acid	[96,97]
	Car painting line Alcohols: Ethanol, 2-Propanol, 1-Propanol, Isobutanol, 1-Butanol; Aldehydes: Propanal, Butanal, Pentanal, Hexanal; Ketones: Acetone, 2-Butanone, 2-Hexanone; Carboxylic Acids: Acetic Acid, Propanoic Acid; Aromatic Hydrocarbons: Benzene
Oil and gas development	Alkanes: Ethane, Propane, n-Butane, Isobutane, n-Pentane, n-Hexane; Cycloalkanes: Cyclohexane; Alkenes: Propylene, Ethylene, Isoprene; Aromatic Hydrocarbons: Benzene, Toluene, Ethylbenzene, Xylenes (o-, m-, p-xylene), Styrene, Naphthalene, Methylnaphthalene, Dimethylnaphthalene; Halogenated Hydrocarbons: Dichloromethane, Chlorinated hydrocarbons; Phenols: Phenol	[98,99]
Ceramic	Aromatic Hydrocarbons: Benzene, Toluene, Xylene (o-xylene, m-xylene, p-xylene); Aldehydes: Formaldehyde, Acetaldehyde, Benzaldehyde, Hexanal, Heptanal; Carboxylic Acids: Propionic acid, Acetic acid, Oleic acid, Stearic acid; Alkanes: Hexadecane, Heptadecane, Octadecane; Cycloalkanes: Cyclotetradecane; Phenols: Phenol; Ketones: Acetone, 2-Pentanone	[100]
Paint	Aromatic Hydrocarbons: Benzene, Toluene, Ethylbenzene, Xylene (m/p-xylene), Styrene; Alkanes: n-Hexane, n-Heptane, n-Octane, n-Nonane, n-Decane; Halogenated Hydrocarbons: Trichloroethylene, Tetrachloroethylene, Dichlorofluoromethane; Esters: n-Butyl acetate; Ketones: Acetone	[101]
Printing	Alkanes: n-Nonane, n-Decane, n-Undecane; Aromatic Hydrocarbons: Toluene, m/p-Xylene; Alcohols: 2-Propanol; Esters: Ethyl acetate	[102,103]
Footwear	Alkanes: 3-Methylhexane, 2-Methylhexane, Heptane, n-Hexane, n-Heptane, n-Octane, Ethane; Alkenes: 1-Hexene, Propylene, Ethylene, 4-Methyl-1,3-pentadiene; Alkynes: Acetylene; Aromatic Hydrocarbons: Toluene, Benzene, Chlorobenzene; Ketones: 2-Butanone, 2-Methylethylketone, Acetone, 2-Hexanone; Esters: Ethyl acetate, Methyl Methacrylate; Ethers: Tetrahydrofuran; Halogenated Hydrocarbons: 1,2-Dichloroethylene, Chlorobenzene; Cycloalkanes: Methylcyclopentane, Methylcyclohexane; Sulfur Compounds: Carbon Disulfide; Aldehydes: Acrolein	[88,104]
Electronic equipment	Alcohols: Ethanol, Isopropanol; Alkanes: 2,3-Dimethylpentane, n-Undecane, Dodecane, 2-Methylhexane, 2-Methylpentane, 3-Methylpentane, n-Pentane; Cycloalkanes: Cyclohexane; Alkenes: Isoprene; Halogenated Hydrocarbons: Dichlorodifluoromethane, Trichloroethylene; Esters: n-Butyl acetate; Ketones: Acetone	[88,105]
Metal product	Alkanes: n-Undecane, Dodecane, n-Decane, Decane, Isopentane, Propane; Halogenated Hydrocarbons: 1,1,2-Tetrachloroethane; Alcohols: Isopropanol; Aromatic Hydrocarbons: Naphthalene; Ketones: Acetone	[88]

, different industrial anthropogenic sources of VOC are presented.

In EU27, VOC emissions have reduced by 61.6% since 1990, but only by 4.7% from 2019 to 2022. In 2022, total NMVOC emissions stood at 6291 thousand tonnes per year [106]. This slower pace of reduction in recent years can be partly attributed to diminishing opportunities for further technological advancements, as many readily available solutions have already been implemented. Continued progress will likely require innovative approaches, targeted policy measures, and sector-specific interventions to address remaining emission sources effectively. It has been estimated by the European Solvents Industry Group (ESIG) that total solvent emissions for the EU27 + UK were 2323 thousand tonnes·year⁻¹ in 2013 and 2209 thousand tonnes·year⁻¹ in 2021 [107,108]. This means that solvent emissions have been reduced, but large quantities of NMVOC are still released into the atmosphere. Even though most EU countries are below the targets for 2030 NMVOC emissions, some, like Bulgaria, Czechia, Estonia, Greece, Hungary, Ireland, Italy, Lithuania, Portugal, Romania, Slovenia, and Spain, are still above their targets [106].

According to Simayi et al. [109], VOC emissions in China were shown to be decreasing after peaking in 2016, with chemical and coking industries being responsible for most of the reduction. However, VOC emissions from petroleum industries and industrial painting continue to increase, mostly due to China’s industrial high VOC emission areas shifting from key areas to surrounding areas. Implementing emission control measures is expected to reduce VOC emissions by 15% or 30% if stricter rules are employed in 2030 relative to 2019.

From a regulatory perspective, the Industrial Emissions Directive (IED) (2010/75/EU) establishes the legislative framework for industrial VOC emission control in the EU [2]. It administers practical implementation through Best Available Techniques (BAT) Reference Documents (BREF), which offer relevant guidance for specific sectors. These BREF often recommend biological treatments as efficient and sustainable options for VOC abatement [110,111,112]. Similarly, the United States (US) Environmental Protection Agency (EPA) issues Control Techniques Guidelines (CTG) and Alternative Control Techniques (ACT) documents, serving as recommendations for state and local air pollution control agencies. Some of these documents often identify biological treatments, such as biofiltration, as viable options for controlling VOC emissions in specific industries, including wastewater treatment, food processing, and wood products manufacturing [113].

While these recommendations support the technical and environmental viability of biological methods, their large-scale adoption is still constrained by economic and institutional barriers. The EU and US regulatory instruments of the IED and its BAT conclusions, and the CTG and ACT documents produced by the EPA, provide some level of guidance, but do not explicitly prioritize biological treatments over thermal or chemical alternatives. As a result, adoption often depends on local regulatory interpretation, industry preferences, and cost-effectiveness. Nevertheless, increasing pressure to reduce greenhouse gas emissions and operational costs is leading industry toward greener technologies. Instruments such as carbon footprints, green investment and subsidies, and taxation of fossil-fuel-intensive methods could have a significant economic incentive from industries. Strengthening these mechanisms and incorporating explicit support for biological treatments in policy structures could significantly accelerate their broader application across sectors.

3. Biological Methods for VOC Emission Treatment

Various physical, chemical, and biological treatment methods are available to manage VOC in the air, either by recovering them or facilitating their destruction. Biological solutions are usually found to be simple to implement, easy to perform, require low investment, and have low operational costs and minimal secondary pollution [19,114,115]. In these systems, microorganisms are usually fixed to organic or inorganic beds capable of degrading several compounds present in a fluid stream [116]. Under aerobic conditions, the pollutant is adsorbed in an aqueous phase, suffers oxidation, and is converted to H₂O and CO₂ by active microorganisms [52,117]. In biofiltration, the removal efficiency (RE) of a single-component VOC is only related to its bioavailability, while for multicomponent VOC mixtures, the characteristics of each compound in the mixture are also relevant. This can result in microorganism competition, VOC toxicity, catabolic repression, or substrate competition [118]. Biofiltration can be enhanced by modifications of physicochemical and biochemical biofilm properties, such as adding silicone oil, which improves the transfer of oxygen and VOC into the biofilm [119,120]. A schematic representation of the interactions of a biofilm pollutant degradation in a biological setting is presented in Figure 2. Packing media [121,122,123,124], microorganism population and interactions [125,126,127,128], operation mode and configuration [129,130,131,132], VOC loads and empty bed residence time [133,134], available nutrients [135,136,137], and VOC hydrophilicity and chemical structure [138,139,140,141] all affect the efficiency of the biologic system. The degradability of VOC in aqueous environments is strongly influenced by the equilibrium between the gas phase and dissolved concentrations of a compound that can be described by Henry’s Law [142]. The octanol–water partition coefficient (Log K_OW) is a widely accepted metric of hydrophobicity, quantifying the compound’s tendency for a non-polar (lipophilic) versus polar (aqueous) environment. VOC with high Log K_OW values (>4) are typically hydrophobic, exhibiting strong affinity for organic matter, bioaccumulative potential, and low aqueous solubility. Conversely, VOC with low Log K_OW values (≤1) are more hydrophilic, showing greater solubility in water and a tendency for rapid transport in aqueous media [143]. When referring to scrubbing technologies, the affinity of hydrophobic compounds for the absorption liquid can be assessed using the equilibrium partitioning coefficient between the gas and liquid phases (gas–liquid partitioning coefficient, K_GL, dimensionless). A lower K_GL indicates a greater preference for the liquid phase, leading to more efficient VOC removal from the waste gas for the liquid phase and, therefore, the removal of VOC from the waste gas is higher [144]. Some authors categorize VOC based on their dimensionless Henry’s Law constant (H) at 25 °C: hydrophilic VOC (H < 0.1), moderately hydrophilic VOC (0.1 ≤ H ≤ 1.0), and hydrophobic VOC (H > 1.0) [138,142,145]. These classifications highlight the varying affinities of VOC for water and their subsequent environmental behavior in VOC removal from waste gas. Higher solubility and lower surface tension facilitate microbial uptake and degradation [141]. Microorganisms thriving in biofilms can more readily metabolize less hydrophobic VOC due to their higher availability in the aqueous phase. This increased solubility facilitates the transport of these compounds into the biofilm, where microbial communities can efficiently degrade them. For this reason, strategies to increase the solubility of certain compounds have been put into practice. Surfactants can reduce surface tension and form micelles, increasing the solubility of hydrophobic VOC, thus reducing the mass transfer resistance of hydrophobic compounds [146,147]. Biosurfactants like rhamnolipids and surfactin and vegetable oils offer additional benefits, including biodegradability and reduced toxicity [148,149,150]. However, it is important to find the right balance when using such strategies, as an excess of surfactants might inhibit microbial activity or reduce VOC availability to microorganisms due to a surplus of micelles [151].

Biological systems are often regarded as economically advantageous to other technologies. The study conducted by Dobslaw and Ortlinghaus [152] highly supports this claim. The authors state that biological processes generally have lower investment costs (20–50%) compared to thermal oxidation technologies. The compiled information mentioned that the operational costs are also significantly lower, primarily driven by electricity costs for compressors and circulation pumps. For example, thermal oxidation has investment costs of between EUR 12.8 and 34.83 per m³·h⁻¹ and operational costs ranging from EUR 1.5 to 7.67 per 1000 m³ of treated air. On the other hand, BFs have investment costs that range from EUR 1.53 to 12.8 per m³·h⁻¹ of treated air and operational costs between EUR 0.23 and 1.0 per 1000 m³ of treated air depending on the size and application. Additionally, BTF investment costs ranges from EUR 1.41 to 18.78 per m³·h⁻¹ of treated air and operational costs between EUR 0.04 and 1.27 per 1000 m³ of treated air. A detailed economic assessment by Prado et al. [153] evaluated a full-scale compost-based biofilter treating 20,000 m³·h⁻¹ of VOC-contaminated air with an Empty Bed Resident Time (EBRT) of 60 s. The total investment cost was approximately EUR 56,500, corresponding to a normalized investment cost of EUR 2.83 per m³·h⁻¹ of treated air. Annual operating costs were estimated at around EUR 25,750, including electricity, labor, and water consumption, and EUR 9820 per year for packing material replacement, assuming a typical two-year lifespan for compost. When expressed per unit volume of air, the operating costs ranged between EUR 0.04 and 0.34 per 1000 m³ of air. Sensitivity analysis further revealed that gas flow rate, EBRT, and packing material durability were the most influential factors in overall cost performance. The study conducted by Estrada et al. [154] supports these claims, as the authors’ comparative analysis demonstrates that biological techniques exhibit significantly lower operating costs (up to 6 times) and higher robustness compared to their physical/chemical counterparts. The packing material emerges as a critical parameter influencing operating costs by comprising 40–50% of the total operating costs.

In the comparison performed by Senatore et al. [117], the advantages of investment and operational costs become evident. The investment costs of chemical scrubbers are set to be between EUR 15 and 30 per m³·h⁻¹ and adsorption situate at EUR 5–12 per m³·h⁻¹, while the biological solutions of BF, BTF, and BS have costs in the order of EUR 6–15, 8–28, and 10–32 per m³·h⁻¹, respectively. In terms of operating costs, chemical scrubbers have similar costs to biological ones, situated in the range of EUR 5–6 per m³·h⁻¹. On the other hand, adsorption is between EUR 10 and 200 per m³·h⁻¹. Regarding the biological options, it is possible to understand that the price is between EUR 2 and 6 per m³·h⁻¹. On average, BS need less empty bed residence time in the biological methods for VOC treatment than BFs and BTFs. This means that it is possible to have a lower volume for higher inlet loading rates. Investment cost per EUR per m³·h⁻¹ decreases with increasing working flow rate and increases with empty bed residence time.

Deshusses and Cox [155] employed a computer model to assess the costs of a BTF with a 10,000 m³·h⁻¹ airstream contaminated with 1.5 g·m⁻³ of toluene as a function of the nutrient supply. The authors attained values between USD 1.90 and 3.84 per 1000·m³ of treated air. Deshusses and Webster [156] estimated actual operating costs for an 8.7 m³ pilot/full-scale biotrickling, reaching values between USD 8.71 and 14.05 per 1000 m³ for lower and upper end cost, respectively. The total construction costs of the designed reactor were USD 177,500, with about one-half of the total reactor for personnel and engineering time, whereas around 20% was for monitoring and control equipment. Delhoménie and Heitz [157] refer to values between USD 3 and 10 per m³·h⁻¹ and an investment of USD 10–70 per m³·h⁻¹ of air. The same authors performed a compilation on costs of other technologies (adsorption, incineration, catalytic oxidation, absorption, and condensation), with biofiltration operational costs representing approximately 15% to 50% of the operational costs of other technologies and about 14% to 70% of the investment costs needed for other technologies. Vikrant et al. [158] estimated the reactor unit cost for bed volumes between 5 and 1000 m³, obtaining values between USD 2400 and 9000 ·m⁻³. Nevertheless, airstreams ranging from 10,000 to 60,000 m³·h⁻¹ are considered to be 55–65% cheaper than regenerative catalytic oxidation [159]. Tomatis et al. [160] refer a cost of USD 895,418 ·year⁻¹ for a regenerative thermal oxidizer (RTO) and USD 583,757 ·year⁻¹ for a catalytic thermal oxidizer (CTO) treating 169,920 m³·h⁻¹, while Chou et al. [161] refer an initial investment of a typical RTO or CTO of between USD 18 per m³·h⁻¹ with a recuperative heat recovery system and USD 27 per m³·h⁻¹ with regenerative system.

In Gospodarek et al.’s [162] work, a strategy to compare the selected biological methods of BF, BTF, and BS technologies in terms of performance, costs, technical aspects, and sustainability on removing either hydrophilic or hydrophobic VOC, as well as odorous inorganic compounds, was proposed. The authors based their study and decisions on a preliminary method selection based on a decision tree. The method employed derives from a pairwise comparison model using numerical judgments from an absolute scale of numbers. In simple terms, the model provides a way to objectively compare the biological treatments by considering the division into various aspects of a given field. From their studies, BFs and BTFs exhibit good performance when treating hydrophobic compounds. It is suggested that high concentration streams are better treated with BFs, while a BTF offers an advantage in the removal of hydrophilic compounds and when concentrations are low. On the other hand, compared to BS and BTF, BS presents moderate or even low performance; but, according to the decision tree, BS should be preferred if Henry’s Constant is higher than 0.0075 mol·m⁻³·Pa⁻¹. In fact, from the model, Henry’s Law and inlet concentration are the most important parameters to consider in method selection. Economic and sensitivity analysis ultimately depends on investment, operating (electricity, water, chemicals, labor), and packing material replacement costs [154,163].

3.1. Biofilter

The oldest biological method for VOC treatment is biofiltration. In this technique, the contaminated air is directed to a packing medium of a fixed inorganic or organic bed bioreactor in which active microorganisms are immobilized and no liquid phase is required [116,157,158,164]. Initially, simple BFs would be employed using soil beds, with the latest technology using advanced bed materials [6]. A variety of different packings have been studied for VOC removal in order to improve microbial diversity, gas–liquid mass transfer, and VOC degradation [139,165,166,167]. Different bacteria and fungi can promote VOC biodegradation to biomass, metabolic end-products, or carbon dioxide and water [115,168]. When selecting the packing, high porosity is desirable to facilitate the homogenous distribution of gas flow, as well as a high water retention capacity to avoid media drying [158]. The moisture level in the packing materials shall be maintained under ideal conditions, since the microorganisms’ activity can decrease at low humidity levels. Anaerobic zones might be formed at high levels, reducing the oxygen available for biological activity [117].

A BF should operate close to the equilibrium phase to maintain maximum performance [169]. To this extent, the nutrient limitation should be prevented by supplying excess nutrients (especially ammonia and phosphate), thus maximizing the rate of the exponential phase. Moreover, nitrogen sources have been shown to affect the removal performance of VOC, with ammonium leading to a higher BFBF performance than nitrate nitrogen [135]. Certain byproducts formed during biological treatment need to be regularly removed from the system. If these byproducts are allowed to build up, their accumulation can negatively affect the process. To the same extent, excess biomass ought to be removed at the same rate as it is generated to prevent the clogging of the packing. By periodically removing byproducts, the system can remain in an active and productive state, avoiding this stationary or stagnation condition [169].

Liu et al. [170] studied an in situ compost BF treatment from municipal solid waste treatment plants. The system was highly efficient in removing alkylated benzenes (>80%), but poorly adequate in removing terpenes (30%). RE increases with EBRT, with 65 s providing the best results (49.8–95%). During the 140 days of operation, nitrogen, carbon, phosphorous, and organic matter in the compost media had no significant change. On the other hand, due to ammonia sorption from the odor stream, ammonium nitrogen, nitrate nitrogen, and soluble total nitrogen content gradually increased with time.

Chen et al. [171] tested a new BF for VOC removal comprising carriers with a density slightly lower than water forming a suspended carrier with a porous surface. The BF bed had a height of 30 cm, a diameter of 16 cm, and a working volume of 6.0 L. The suspended BF treated the toluene contaminated waste stream with an average RE higher than 90.2% in a toluene loading range of 11.0–58.5 g·m⁻³·h⁻¹. The authors attribute the high efficiency to toluene-degrading bacteria, i.e., Acinetobacter sp. Tol 5, Burkholderia, and Comamonadaceae. Additionally, the authors report that clogging and biomass accumulation, major drawbacks of a fixed BF, were nonexistent.

Aizpuru et al. [172] tested the performance of a BF packed with active carbon (AC) in an effluent composed of alcohol, ketones, esters, and aromatic and chlorinated compounds. The authors employed a control abiotic humidified filter operating in the same conditions as the BF. At a steady state, for a load of 110 g VOC·h⁻¹·m⁻³ AC, the AC system achieved high removal efficiencies, except for chlorinated compounds and p-xylene, which presented 25% and 64% removal efficiencies, respectively. The microorganism’s consortia reduced the adsorption site accessibility by forming a biofilm on the AC surface, improving the removal of biodegradable VOC with low affinity for AC. Additionally, pollutants with reduced affinity for AC were only removed in the upper section of the reactor. For this reason, the affinity of the packing media and, consequently, the height of the reactor play a crucial role in the VOC RE.

Han et al. [173] studied two BFs with an effective volume of 1.5 m³ to control VOC and odors from a landfill in the sealing zone and the leachate treatment. The flow rate was fixed at 60 m³·h⁻¹, translating into an empty bed residence time of 1.5 min in a packing of polyurethane foam cubes with a density and porosity of 20 kg·m⁻³ and 95%, respectively. From the collected data, it becomes perceptible that the concentrations of amines, sulfides, and TVOC in the leachate treatment zone were 1.59, 1.68, and 2.93 times higher than those in the sealing zone. The two zones’ gaseous pollutant concentration variability led to the formation of different bacteria consortia, with the bacterial populations diverse at different heights in the two zones. In total, from the same inoculum, the dominant bacteria of the BF in the sealing zone gradually evolved into Mycobacterium (67.82%) and Bacillus (27.90%), while Bacillus (72.69%) and Pseudomonas (21.59%) were the dominant bacteria of the BF in the leachate treatment zone. Even so, the average removal rate of the total VOC, sulfides, and amines in both BFs exceeded 80%.

Muszyński et al. [174] investigated bioaerosol emissions from a semi-technical-scale BF treating waste gas from a food industry plant and explored strategies to reduce bacterial and fungal emissions using membrane technology. The authors also assessed the efficiency of the membranes in reducing VOC emissions. The BF demonstrated VOC removal efficiencies ranging from 88% to 99% depending on the bed material and membrane used. The CB bed (Stumpwood Chips + Pine Bark, No Membrane) achieved 88–91% RE, while the addition of Membrane 1 (CB + M1) improved efficiency to 94%, and Membrane 2 (CB + M2) resulted in 93% removal. The CBC bed (Stumpwood Chips + Pine Bark + Compost, No Membrane) exhibited a higher RE of 95–96%, which further increased to 98–99% with Membrane 1 (CBC + M1) and 97% with Membrane 2 (CBC + M2). These results indicate that the CBC bed outperformed the CB bed, and the integration of membranes enhanced VOC removal performance, with Membrane 1 providing the highest and most stable efficiency.

Compost-based packing can be a cost-effective alternative to being employed as a pack in BFs. Guzmán-Beltrán et al. [175] studied a packing material made from chicken manure and sugarcane bagasse to treat VOC in wastewater treatment plants, achieving maximum removal efficiencies of 71.5% for toluene under controlled laboratory conditions. Toluene removal proved challenging due to its hydrophobic nature, with significant efficiency drops at concentrations above 40 mg·m⁻³, indicating toxicity to the microbial community. The BF demonstrated recovery potential after stress conditions, suggesting resilience under fluctuating pollutant loads. Moisture optimization (increased to 50%) and microbial acclimatization were crucial for improving VOC degradation, highlighting the importance of biofilm activity and pollutant transport mechanisms. This work underscores the potential for integrating compost-based BFs into WWTP to mitigate VOC emissions sustainably.

Lai et al. [176] studied the effect of the filling method on the degradation of ethyl acetate in BFs constructed with magnetite, activated carbon, perlite, and gravel added via layered filling or mixed filling. The authors discovered that, in layered filling, Burkholderia, Acinetobacter, unclassified_Rhizobiales, and unclassified_Burkholderiales were dominant, while in BF2 mixed filling, unclassified_Chloroflexi, unclassified_Bacteroidetes, and unclassified_Anaerolineaceae dominated. Therefore, the ethyl acetate RE and nutrient use of total phosphorus, total nitrogen, and ammonia nitrogen differed for each filling. Lower inlets (224 g·h⁻¹·m⁻³) provided better ethyl acetate removal, especially for layered filling. Medium and high inlets (336 and 448 g·h⁻¹·m⁻³) drastically decreased the RE, with layered filling working better for higher inlets.

The simultaneous removal of ammonia (NH₃) and VOC is achievable, as demonstrated by Shang et al. [177] in a study using a pilot-scale BF with mature compost as the filling medium. The BF was tested for three EBRT (30, 60, and 100 s). The analysis identified dimethyl sulfide (DMS), dimethyl disulfide (DMDS), dimethyl trisulfide (DMTS), and trimethylamine (TMA) as the primary odor contributors based on their odor activity values. Optimal performance was achieved with a 60 s with RE ranging from 76.7 to 100% depending on the VOC. For NH₃, the RE averaged between 85% and 89% RE. The BF also demonstrated high elimination with maximum values of 1301 mg·m⁻³·h⁻¹ for DMS, which is a compound known to be insoluble in water. Removing bioaerosols, odors, and VOC in wastewater treatment plants is also achievable from sludge dewatering rooms. This was demonstrated by Liu et al. [178], who examined a full-scale integrated reactor consisting of a two-phase BF with an activated carbon adsorption zone. The two-phase BF consists of a suspended phase and an immobilized phase with low-pH and neutral-pH zones designed for hydrophobic and hydrophilic compound removal, respectively. At an inlet flow rate of 5760 m³·h⁻¹ the RE of odors, VOC and bioaerosols were recorded as 98.5%, 94.7%, and 86.4%, respectively. The operational cost of such treatment was estimated at 0.0027 CNY·m³.

The article of Liang et al. [179] presents a comprehensive study on the effectiveness of a pilot-scale integrated system combining a spray tower and a BF for the removal of VOC emitted from a textile dyeing wastewater treatment plant. The integrated system showed high removal efficiencies for different groups of VOC, notably aliphatic hydrocarbons (67.9%), nitrogen- and oxygen-containing compounds (66.7%), halogenated hydrocarbons (52.1%), and a lower RE of aromatic hydrocarbons (11.7%). The BF’s microbial community was found to be dominated by Proteobacteria, suggesting that they are well adapted to the specific pollutants emitted by the textile dyeing processes. The spray tower efficiently removed water-soluble and larger particulate VOC, serving as a pretreatment stage that enhanced the subsequent biofiltration process. The authors highlight that, by lowering the levels of hazardous VOC, the system effectively reduced cancer and non-cancer health risks to workers and nearby communities.

Authors such as Cortés-Castillo et al. [180] explore innovative approaches to treat VOC with BFs. In their study, the authors aim to improve hexane removal by applying homogeneous magnetic fields using Helmholtz coils operating under magnetic fields of 0, 10, and 30 mT for 191 days. Applying magnetic fields resulted in improved hexane removal efficiencies of 11% and 15% for 10 mT and 30 mT, respectively. The authors found changes in biomass distribution and reduced pressure drops in magnetic-field-treated BFs, contributing to system stability and longevity. The microorganisms also exhibited an increased production of exopolysaccharide components like glucuronic acid and carbohydrates, facilitating more efficient pollutant transport and degradation.

BF remain a versatile and effective technology for VOC and odor control. Advances in packing materials, operational strategies, and novel approaches such as magnetic fields continue to enhance their efficiency. However, challenges such as biomass management, pollutant specificity, and scalability must be addressed for widespread industrial application.

3.2. Biotrickling Filter

In BTF, the gas is forced to pass through a packing bed that is continuously rinsed with an aqueous solution. The biofilm will attach, grow, and feed off the aqueous-solution-striped contaminants and externally added nutrients. Therefore, in these types of systems, ensuring the best mass transfer of the pollutant from the gas phase to the aqueous phase is essential, with the biofilm being the key element in waste gas treatment [19,158].

For the absorption phenomenon to occur, the mass of the pollutant is to be transferred from the gas to the liquid phase. Therefore, the two phases must come into contact to permit the change toward equilibrium, either through co-current flow (streams flow in the same direction) or countercurrent flow (streams flow in opposite directions) [19,181]. Countercurrent operation is usually preferable as it allows for an increased mass transfer driving force, especially at gas entrance and liquid exit. The maximum gas flow is limited by the pressure drop and the flooding limit [181]. The moisture content of BF packing materials can be controlled through adjustments in the relative humidity of the inlet gas, the occasional spraying of the packing, or immersing part or all of the packing material in a nutrient solution, with intermittent liquid spraying proving to be an effective strategy for the biofiltration of both hydrophilic and hydrophobic VOC [129]. To avoid clogging by biofilm growth and to obtain a high flooding point, the packing material should have a high bed porosity (>70%). It should also have a high external specific surface area to promote the external mass transfer of pollutants into the liquid phase due to significant liquid film trickling [19]. Packing materials with a high surface area are usually preferable for offering better performances when compared with other packings under similar conditions [114]. As expected, the liquid flow and recycling rates are documented critical parameters. This is due to the contact between the microorganisms and the pollutants occurring after the diffusion of the pollutant in the liquid media [158,182]. For this reason, EBRT is a fundamental parameter as it influences both mass transfer from the air to the water and from the water to the biofilm; thus, biodegradation is achieved by the microorganisms. For cost-effective purposes, EBRT should be as low as possible and an adequate time for the loading rate to satisfy the required RE should be ensured [114]. Other factors such as pH, temperature, nutrients, and O₂ are very case-specific as it highly depends on the microorganisms, the biofilm, and the composition of the waste gas [19,183]. A schematic representation of a BTF for off-gas treatment is presented in Figure 3.

Packing material in BFs plays a critical role in the efficiency of biological air or water treatment processes. This material provides a large surface area to support the growth of microorganisms that degrade pollutants. The choice of packing material can influence both the microbial community’s development and the BF’s overall performance. Cox et al. [184] employed seven layers of polyvinyl chloride (PVC) to form a structured packing in a pilot-scale BTF at the Hyperion Treatment Plant in Los Angeles. For 10 months, the authors obtained removal efficiencies above >98% for H₂S. However, VOC removal was limited to toluene and benzene and only when pH was neutralized during operation, with no removal of xylenes or chlorinated VOC, regardless of experimental conditions in the reactor.

The study of Sun et al. [185] aims to enhance the efficiency of BTF for removing BTEX by modifying the surface of polyurethane (PU) sponge packing material to improve biofilm formation and operational performance. The modification process involved increasing surface roughness and introducing positive charges through chemical treatments. The modified sponges significantly improved microbial growth, biofilm adhesion, and operational efficiency. Specifically, the modified PU demonstrated a BTEX RE of 61%, higher CO₂ production, and better biofilm stability due to reduced extracellular polymeric substances (EPS) and denser biofilm structures. Additionally, the modified PU exhibited a higher concentration of Actinobacteria, a bacterium capable of degrading complex organic compounds.

Wang et al. [186] employed a BTF with a simulated gas phase composed of o-xylene ethyl acetate and dichloromethane entering at the bottom of the unit with a counter-current constant liquid flow rate of 30 L·h⁻¹. The bed packing consisted of self-made porous polyurethane balls, giving a 5.9 L filtering layer with a bed porosity of 47%. The elimination capacity for low inlet loading rates increased steadily and linearly with the increased loading rate. However, when it reached 100 g·m⁻³·h⁻¹, the system deviated considerably from the 100% removal. When gradually increasing the inlet rate to 170 g·m⁻³·h⁻¹, variation in the RE was negligible. This is due to the mass transfer limitations between the gas phase and liquid phase/biofilm. At these inlet concentrations, the microorganism’s capacity to perform a further metabolic activity is hindered. Furthermore, the authors report that, after 1 month of operation, o-xylene RE went from 90% to 80% at a loading rate of 80 g·m⁻³·h⁻¹ (EBRT of 60 s), with the authors attributing this phenomenon to excess biomass accumulation, which might have lowered the mass transfer flux of o-xylene from the gas phase to the liquid phase/biofilm.

Liu et al. [187] inoculated a BTF with sludge from a chlorobenzene-containing wastewater treatment plant to investigate the degradation of gaseous chlorobenzene, as well as the mass transfer kinetics and microbial community dynamics. The authors highlight overcoming the limitation of the gas-to-liquid/solid mass transfer of chlorobenzene, achieving >85% removal efficiencies, even for an inlet concentration of 700 mg·m⁻³, with an oxygen concentration of 10% and an EBRT of 80 s.

Balasubramanian et al. [188] tested the efficiency of a BTF in removing methanol, ethanol, acetone, and toluene. The inlet loading rate varied from 52 to 419 g·m⁻³·h⁻¹, while the EBRT was varied from 25 to 69 s. Packing was made with polyvinyl chloride made in a spiral shape, with an average size of 13 mm, resulting in a surface area of 600 m²·m⁻³. The mixed VOC gases were completely degraded up until 240 g·m⁻³·h⁻¹. The authors emphasize that toluene was the most resistant to degradation, followed by acetone. Single components resulted in significantly higher efficiency than in a mixed VOC system due to the competitive interactions between the different pollutants.

Wu et al. [189] studied the performance and bacterial diversity of a BTF filled with conductive packing material to treat toluene. The authors tested two BTF, one biotrickling filter-original (BTF-O) and one biotrickling filter-electrical (BTF-E). Both had an effective bed height of 900 mm and both were packed with conductive packing material with an average diameter of 20 mm. The difference between the BTF is that, in the BTF-E, an electric current was applied to the system. Due to the microorganisms’ acclimatization, initial performance was low, with removal efficiencies around 1%. On the 20th day, the RE increased to about 50%. During this period, the inlet concentration of toluene ranged from 0.136 to 0.271 g·m⁻³ and the corresponding EBRT was 61.5 s. The authors provided different inlet concentrations to both BTF, while applying +1 V or −1 V to the BTF-E. In all the tests, the voltage applied resulted in lower removal efficiencies and microorganism elimination capacity in the BTF-E compared to the BTF-O. Thus, the developed CPM was proven to be efficient for BTF application, but BTF removal performance was lower when a voltage was applied. Nevertheless, the authors point out that a BTF can biodegrade pollutants and convert part of the chemical energy in pollutants to electrical power.

Dobslaw et al. [190] used a low-cost plasma-enhanced BTF process carried out with artificial waste air containing methane, n-butanol, or ammonia at 25 ppm, 50 ppm, 75 ppm, or 100 ppm. It entailed a three-stage system with a dielectric barrier discharge (DBD) as a non-thermal plasma stage, a low-cost zeolite adsorber with catalytic activity, and a BTF. Compared to other catalysts, the authors provide a valid and low-cost alternative capable of achieving efficiencies higher than 95%. Even so, this system requires an input of energy and two extra steps before bio removal.

The paper from Álvarez-Hornos et al. [191] sets a perfect example of the advantages of BTF over RTO. The authors studied the ability of a pilot-scale BTF to reduce VOC emissions in the plastic coating sector for one year. Different solvents affected the BTF performance differently. For instance, products containing more than 60% biodegradable compounds would require an EBRT of between 30 and 40 s to meet legal requirements. In contrast, a minimum EBRT of 80 s was required for emissions mainly composed of hydrophobic VOC. The authors’ results prove that the annual cost (TAC) of a BTF system was 1.63 TAC per (Nm³·h⁻¹) compared to the 2.45 TAC per (Nm³·h⁻¹) of thermal treatment technology. Additionally, CO₂ emissions had an estimated reduction of 66%. The authors then estimated the sizing of the industrial-scale BTF to be 750 m³ for a flow rate of 80,000 Nm³·h⁻¹ and a VOC concentration of 220 mg_C·Nm⁻³ of the waste gas.

In a more conventional approach, Khoramfar et al. [192] employed a sequential biotrickling–biofiltration unit to remove VOC from the headspace of crude oil storage tanks. Alkanes and cycloalkanes (butane, hexane, pentane, and others) were the predominant compounds, with lower concentrations of aromatics (toluene, benzene, xylene, and others) being found. Both the BF and BTF operated under the same conditions of water and air flow and same bed heights (1.21 m) consisting of crossflow plastic packing (BTF) and an engineered medium (BF), with both entailing a bed volume of 1.42 m³. The only different condition was the spraying frequency, which was continuous for the BTF and every 2 min for the BF. The waste air passed first through the BTF and then the BF, obtaining a range of 55–60% VOC removal, with the percentage of biodegradable VOC at the headspace of the storage tanks remaining almost unchanged during the operating time of the unit.

Bak et al. [193] used a pilot-scale BTF filled with polyethylene Ralu rings for the biodegradation of a VOC mixture containing styrene, ethanol, and dimethyl sulfide under dynamic pollutant loads. The results demonstrated that VOC RE was above 95% at low loads, but decreased to 80% at moderate and dropped to 55% at high loads, with ethanol degradation remaining stable throughout. Dimethyl sulfide RE dropped at higher loads (>5 g·m⁻³·h⁻¹) due to oxygen depletion and toxic intermediates, while styrene degradation efficiency declined beyond 110 g·m⁻³·h⁻¹ due to mass transfer limitations. The microbial biofilm was dominated by Pseudomonas spp. and Cedecea davisae, with sulfur-oxidizing bacteria contributing to a synergistic effect that enhanced styrene degradation. The system exhibited high robustness, recovering within 10 days after a pollutant overload and 14 days after clogging issues caused by excessive biofilm growth.

The study of Deng et al. [194] investigates the enhanced removal of mixed VOC with different hydrophobicities—n-hexane (highly hydrophobic) and dichloromethane (moderately hydrophobic)—in a BTF with the addition of the non-ionic surfactant Tween 20 to a packing medium of polyurethane foam. Tween 20 enhanced the removal of hydrophobic VOC, increasing n-hexane RE by 15.0–20.5%. DCM was completely removed at an inlet concentration of 300 mg·m⁻³, regardless of Tween 20 addition. The maximum ECs were 20.1 g·m⁻³·h⁻¹ for n-hexane and 81.1 g·m⁻³·h⁻¹ for DCM. Shorter EBRT negatively affected n-hexane removal, confirming mass transfer limitations. Tween 20 increased biofilm roughness by 20.5 times, and adhesion force increased from 5.1 nN to 8.3 nN, improving biofilm stability. EPS secretion increased, particularly for proteins, which helped bacteria adhere and form a stronger biofilm. Moreover, biofilm hydrophobicity increased from 21% to 66% (control) and from 29.8% to 82.1% (Tween 20 BTF), facilitating the uptake of hydrophobic VOC. The specific oxygen uptake rate increased significantly with Tween 20, indicating enhanced microbial metabolic activity. Confocal laser scanning microscopy showed an increase in viable cells from 55% to 90% in the Tween 20 BTF.

Silva et al. [195] conducted experiments on a lab-scale BTF to treat off-gas streams contaminated with n-butyl acetate by utilizing a packed-bed column filled with light expanded clay aggregate and comparing it to an absorption column filled with the same material. The researchers tested three distinct operational modes: open circuit using fresh-water, closed circuit by recirculating the water, and closed circuit by adding microorganisms through bioaugmentation. The authors highlight the success in achieving efficiencies of 85% to 99% RE in open circuits. However, efficiency dropped to 0% when transitioning to the closed-circuit mode, primarily due to the accumulation of VOC in the recirculated water. When using microorganisms, the VOC present in the water were biodegraded, and efficiency reached values of 25% to 45% and elimination capacities were between 12 and 55 gVOC·m⁻³·h⁻¹ for low EBRT (24 to 68 s). These values approached the prescribed discharge limits set for certain factory chimneys (75 mgC·L⁻¹), indicating significant potential for more sustainable waste gas treatment.

Rybarczyk et al. [196] tested two different BTF setups: BTF A used pre-incubated polyurethane foam discs, while BTF B used the continuous recirculation of liquid inoculum for inoculation. The study reports that, towards the end of the experiment (around day 120), BTF A reached an average BTEX elimination capacity of 28.8 ± 0.4 g·m⁻³·h⁻¹ compared to 23.1 ± 0.4 g·m⁻³·h⁻¹ for BTF B. This indicates a more effective removal process in BTF A, which utilizes pre-incubated packing elements.

Two setups of BTF filled with polyurethane foam (1 cm³) as the packing material were studied by Xue et al. [58] to explore the biofiltration of toluene in the presence of ethyl acetate (BTF A) or n-hexane (BTF B). Therefore, the study monitored the removal efficiencies of toluene when influenced by varying concentrations of these secondary VOC under a constant EBRT of 30 s. In BTF A, the presence of EA generally enhanced the RE of toluene, achieving up to 95.4% efficiency at the lowest concentration of EA tested. Higher concentrations of EA reduced this efficiency. As EA concentration increased to 1750 mg·m⁻³ and then to 2800 mg·m⁻³, toluene RE significantly dropped to 45.2% and 56.8%, respectively. On the other hand, BTF B had a more varied effect, generally inhibiting toluene removal compared to the effects observed with ethyl acetate. The introduction of n-hexane at 850 mg·m⁻³ resulted in a toluene RE drop to 48.1%. Increasing n-hexane concentration to 1750 mg·m⁻³ and 2800 mg/m³ saw slight improvements in toluene RE to 68.3% and 70.1%.

BTFs are effective air pollution control technologies that employ both physical and biological processes to remove VOC and odors from gas streams. While the technology allows for efficient pollutant removal through optimized mass transfer processes, it faces challenges such as biomass accumulation, clogging, and maintaining the right moisture and nutrient balance to support microbial activity. Overcoming these operational hurdles is essential for enhancing the scalability and overall applicability of BTFs in various industrial applications as an environmentally sustainable option for treating a wide range of pollutants.

3.3. Bioscrubber

BSs take a different approach than BFs and BTFs as they consist of a two-stage unit. In the first phase, the pollutant in a gas phase is transferred into the liquid phase. The operation can be performed with different technologies: an empty spray column, a packed column, or a venturi scrubber. After the gas–liquid contactor, biodegradation is performed in a different biologic reactor, usually in an aerobic-activated sludge reactor. The liquid phase is then transferred to a sedimentation tank where the sludge settles before the treated water is recycled back to the top of the column [19,158]. Most of the time, BSs are employed to treat readily soluble VOC, so it is common to employ polymers to trap hydrophobic compounds [138]. At high pollutant loads and when the contamination of the gas stream with oxygen is nondesirable, BSs are preferred above BFs [19]. Figure 4 gives an example of a BS configuration.

Wongbunmak et al. [197] used a fixed-film bioscrubber (FFBS) inoculated with the BTEX-degrading bacterium Microbacterium esteraromaticum SBS1–7 in an Aquaporousgel^® supporting material. The experiment duration was 172 days for a total gas flow rate of 0.012, 0.021, and 0.045 m³·h⁻¹, corresponding to an EBRT of 1.7, 3.6, and 6.4 min, contaminated with toluene or styrene. Regarding FFBS, the maximum elimination capacity of 203 g·m⁻³·h⁻¹ was achieved for a toluene ILR of 295 g·m⁻³·h⁻¹. Additionally, when ILR was controlled at ≤66 g·m⁻³·h⁻¹, the RE was close to 100%. When changing from toluene to styrene, the system only required 24 h adaptation to reach efficiencies above 95%. Efficiency decreased when the authors increased the ILR and, consequently, the EBRT. In the same study, the authors compared FFBS with BTF configuration, with EC (elimination capacity) and RE being 83 g·m⁻³·h⁻¹ and 67%, respectively, for FFBS. On the other hand, the BTF maximum EC was 104 g·m⁻³·h⁻¹ with an RE of 38%. Additionally, the authors addressed the EC per number of cells in suspension, obtaining 28.27 g·m⁻³·h⁻¹ per 1012 CFU and 22.71 g·m⁻³·h⁻¹ per 1012 CFU for FFBS and BTF, respectively. This means that factors other than the number of viable cells play a major role in the treatment efficiency of the FFBS.

Malhautier et al. [198] compared the influence of using polypropylene Media 2H TPK12 packing with an atomizing absorption column composed of seven atomizing nozzles at intervals of 0.5 m on VOC removal performance. The authors state that, regardless of the selected absorption column for an applied inlet load of around 850–870 gVOC·m⁻³ packing material·h⁻¹, the BS efficiencies were always approximately 50%.

Bravo et al. [199] tested an anaerobic BS for controlling VOC from printing press air emissions with an on-site pilot unit. The main pollutants present in the waste gas were ethanol, ethyl acetate, and 1-ethoxy-2-propanol. The unit worked under airflows ranging between 184 and 1253 m³·h⁻¹ with an average concentration of 1126 ± 470 mg_C·Nm⁻³. The scrubber unit had a total height of 3.06 m and a diameter of 0.5 m, and the packing had a height of 2.0 m. Three scrubber configurations (crossflow and vertical-flow packings and spray tower) were tested. For the biological degradation of VOC, granular sludge from a brewery wastewater treatment plant was chosen. The authors highlight that, despite the inlet load being highly variable, the granular sludge was very consistent, reaching removal efficiencies of 93%, and was capable of producing biogas with 94% methane for 3 h of hydraulic residence time. In total, the effective water volume of the whole system was 16 m³, with 8.7 m³ attributed to the anaerobic reactor, 2 m³ to the scrubber bottom tank, and the remaining volume to two intermediate tanks. Despite the high fluctuations in the waste gas emissions, with interruptions during nights and weekends and temperature oscillations, the stable conversion of alcohols, esters, and glycol ethers to enrich methane biogas was observed. One major disadvantage of this process was the necessity to continuously introduce several macronutrients and sodium carbonate for pH control, as well as Ca, Mg, trace metals, and yeast extract to maintain regular operation inside the anaerobic reactor. Additionally, biomass growth was prevented by adding chemical cleaning products.

Chien et al. [200] studied the efficiency of a parallel-plate wet scrubber (PPWS), working in counter-flow, for low concentrations of the soluble gas removal of HCl, HNO₃, and CH₃COOH with 0.02–2 ppmv at different retention times and liquid-to-gas ratios. To make the liquid film uniform and enhance the hydrophilicity, the parallel plates were made of polypropylene coated with nano-TiO₂. pH was controlled with NaOH and set at a range of 7.6–8.4. The use of the plates greatly enhances the absorption of acid gases, improving the removal efficiencies, especially for concentrations ranging from 0.02 to 1 ppmv. When the retention time was 0.5 s, the RE of the combined system allowed for the removal of 91% of HCl, 95% of HNO₃, and 98% for CH₃COOH at a concentration range of 0.02–2 ppmv. One advantage of the system is the possibility of adjusting the gaps between the plates to control the gas retention time. The gas flow rate can be increased further by increasing the size and number of the polypropylene plates.

Oliveira et al. [201] tested the effect of hexane concentration in the inlet gas stream on the treatment performance of a concentric tube internal loop airlift bioreactor with a 7 L capacity, inoculated with a strain of Pseudomonas aeruginosa isolated from soil contaminated with crude oil. In total, 100% RE was obtained for concentrations of 0.62 and 1.26 g·m⁻³, while for 2.15 and 4.5 g·m⁻³, the efficiency under continuous feed was 55% and 35%, respectively. The authors assessed biomass growth during the continuous feed experiment time frame, concluding that the concentrations of 2.15 and 4.5 g·m⁻³ hindered microorganisms’ growth, increasing the adaptation period. On the other hand, 0.62 and 1.26 g·m⁻³ attained similar biomass growth and adaptation periods. Furthermore, the authors analyzed the effect of discontinuous organic compound concentration by employing positive and/or negative step variation through time in the inlet feed operation. Total contaminant removal from the effluent was verified as 0.62, 1.26, 2.15, and 4.5 g·m⁻³ concentrations, while for a concentration of 6.2 g·m⁻³, a 70% efficiency was registered. Starvation periods of up to 3 h did not impact the bioreactor’s performance.

In summary, BSs are a two-stage unit used for VOC treatment, employing gas–liquid contact in a trickling bed and biologic reactors for biodegradation. Key aspects to consider with BSs include (i) pollutant transfer, (ii) biodegradation, (iii) VOC solubility, species, and concentration, (iv) packing material, (v) nutrient and pH control, (vi) continuous feed vs. discontinuous operation, and (vii) unit design. Understanding these important aspects will aid in designing and optimizing BSs for effective VOC treatment, offering an environmentally friendly solution for air pollution control.

Emphasizing biological VOC removal,

Table 2. Comparison of air treatment technologies for VOC removal: efficiency, costs, and industrial applicability.

Technology	Efficiency (% VOC Removal)	Investment Costs (EUR per m3·h−1 Treated Air)	Operating Costs (EUR per 1000 m3 Treated Air)	Scalability/ Industrial Suitability	References
Biofilter (BF)	High: typically 80–99%, dependent on VOC type	Low: 1.53–12.8	Low: 0.23–1.0	Suitable; highly scalable; effectiveness dependent on packing media and moisture control	[152,153,157,158,174,177]
Biotrickling filter (BTF)	High: typically 80–99%, excellent for hydrophilic VOC	Moderate: 1.41–18.78	Low: 0.04–1.27	Highly scalable; requires effective control of liquid recirculation; optimal for soluble VOC	[117,152,155,157,158,191]
Bioscrubber (BS)	Moderate to high: optimal for highly soluble VOC	Moderate to high: 10–32	Low: 2–6	Moderate scalability; best suited for VOC with high aqueous solubility	[117,158,162]
Thermal Oxidation	Very high: ~95–99%, independent of VOC type	Very high: 12.8–34.83	High: 1.5–7.67	Scalable; energy-intensive; economically suitable mainly for concentrated VOC streams	[152,160,161,191]
Chemical Scrubbing	Low to high: typically 40–99%, dependent on chemical agent	Moderate to high: 15–30	Moderate: 5–6	Moderate scalability; depends heavily on VOC solubility; chemical management and disposal issues limit broader applicability	[117,153,202]
Adsorption	High: typically 90–99%, dependent on adsorbent and regeneration	Moderate: 5–12	High: 10–200	Scalable but operational costs increase significantly with adsorbent regeneration frequency	[117,157]

contrasts the three bio-based technologies BF, BTF and BS against thermal oxidation, chemical scrubbing and adsorption in terms of removal efficiency, investment and operating costs, scalability and industrial suitability.

3.4. Bioaugmentation

According to Speight [203], bioaugmentation is the addition of pre-grown microbial cultures to enhance microbial populations at a site to improve contaminant clean-up and reduce clean-up time and cost. It has been frequently used in the remediation of soil and groundwater contaminants [204,205], including VOC [206,207]. Only some studies regarding bioaugmentation for VOC removal in BTF could be found. Nevertheless, multiple microorganisms have been found to be able to degrade VOC bacteria such as Pseudomonas putida, P. flurorescence, Rhodocous faccians, Alcaligens xyloxidans, Burkholderia cepacia, Hypomicrobium, Xanthobactor, and Acinetobactor. Examples of fungi include Paecilomyces variotii, Pacilomyces, Phanerochaete chrysoporium, Fusarium solani, and Cladosporium sphaerrospermum [208].

Chen et al. [209] explored the possibility of treating volatile organic sulfide compounds through bioaugmentation with Alcaligenes sp. SY1 and P. putida S-1 in a BTF with polyurethane foam particles (10–15 mm diameter, a bulk porosity of 87.3%) as packing. The authors used two different inoculums: a mixture of solid composite microbial inoculant (SCMI) and activated sludge at equal biomass and an inoculum of activated sludge without SCMI. The VOC used for the operation were propanethiol (PT) and dimethyl sulfide (DMS), often found in paper mills and pharmaceutical factories. A gas–liquid countercurrent operation was adopted in the experiment. The mixed gas was introduced from the bottom, and the nutrient solution was sprinkled from the top of the BTF with a spraying rate of 6 L·h⁻¹. The authors successfully inoculated a BTF with the strains becoming the dominant species among various microorganisms (Alcaligenes (21.6%) and Pseudomonas (51.3%) on day 65. The bioaugmentation achieved better pollutant removal efficiencies. Under the best EBRT (20 s), efficiencies could be maintained at 100% at an inlet concentration of 24 mg·m⁻³ and 65 mg·m⁻³ for DMS and PT, respectively.

Hu et al. [210] were faced with unsatisfactory performance for some recalcitrant volatile organic compounds in the pharmaceutical industry due to slow microbial adaptation and growth. The tests were conducted with o-xylene and dichloromethane in a BTF inoculated with specified strains of Zoogloea resiniphila HJ1 and Methylobacterium rhodesianum H13 to overcome the unsatisfactory removal efficiencies. After day 17, the removal efficiencies achieved were above 96.5% for a net influent concentration of 450–600 mg·m⁻³ with an EBRT of 30–75 s.

Cheng et al. [211] prepared a composite microbial agent, mainly composed of bacteria and fungi, and evaluated the performance in removing mixed waste gases containing α-pinene, n-butyl acetate, and o-xylene. The authors chose activated carbon, wheat bran, and sawdust as carrier components. The maximum RE after 24 h, in continuous degradation tests, was 59%, 97%, and 70% for α-pinene, n-butyl acetate, and o-xylene at a concentration of 180 mg·L⁻¹. These results were obtained for a microbial agent with more wheat bran in its composition.

Dewidar et al. [127] investigated a more unconventional approach, using a fungi BTF in the presence of rhamnolipid to remove styrene from waste gas under acidic conditions. The authors gathered experimental data and developed an empirical neural network model based on data. To collect the necessary data, the authors performed nine phases: phase I to III with an EBRT of 90 s and an ILR of 33.9–186.1 g·m⁻³·h⁻¹, phase IV to VI with an EBRT of 60 s and an IRL of 30.6–183.2 g·m⁻³·h⁻¹, and, lastly, an EBRT of 30 s with 25.5–152.7 g·m⁻³·h⁻¹. As EBRT decreased, lower styrene removal was obtained for similar ILR. Removal efficiency further declined for higher ILR with lower EBRT. The optimum observed RE was achieved at 90 s EBRT and recorded a maximum EC of 173.7 g·m⁻³·h⁻¹ corresponding to a styrene RE of 93.3%. The authors constructed an artificial neural network through multilayer perception in terms of ILR, EBRT, and pressure drop with the data collected. The idea is to predict styrene RE (%) using easily measurable inputs. The obtained model was found to have a correlation coefficient of 0.9947, validating the power that these models can have to improve scale-up construction. Additionally, once again, EBRT is found to be a critical parameter when discussing VOC RE. On the other hand, ILR and the pressure drop effect on the BTF performance were found to be negligible. The neural network confirms that a combination of low inlet LR with high EBRT increases RE. ILR between 110 and 180 g·m⁻³·h⁻¹ and an EBRT of 85–90 s are the best conditions to achieve RE > 97%.

Putmai et al. [212] focused their efforts on isolating bacteria from a local swine farm for VOC degradation. The authors identified Proteobacteria and Actinobacteria as potential VOC-degrading bacteria. Both free-cell and immobilized bacteria experiments have shown the capability of Proteobacteria to degrade phenol and Actinobacteria to degrade p-cresol in water, respectively, at concentrations of up to 500 ppm, within 12 h, and without any loss of activity. Sun et al. [213] achieved higher toluene removals with a BTF inoculated with Burkholderia spp. and under a higher ILR when compared with a BTF inoculated with activated sludge, which validates bioaugmentation potential. Furthermore, higher efficiency was also evident at broader temperature ranges and during the gradual decrease in pH.

Sun et al. [214] focused their research on the biodegradation of a ternary gas mixture containing dimethyl sulfide, propanethiol, and toluene using a BTF packed with a polyurethane pall ring and seeded with Alcaligenes sp. SY1 and Pseudomonas putida S1 and analyzed pollutant removal kinetics and microbial dynamics. Toluene RE improved in the mixture, probably due to co-metabolism, while DMS and propanethiol degradation were negatively affected by competition for microbial resources and oxygen limitation. Maximum biodegradation rates were 256.41 g·m⁻³·h⁻¹ for DMS, 204.08 g·m⁻³·h⁻¹ for propanethiol, and 90.91 g·m⁻³·h⁻¹ for toluene in single treatment, but decreased for sulfur compounds in the mixture, indicating substrate interactions. Microbial analysis revealed that Proteobacteria and Bacteroidetes dominated, with Thiobacillus, Pseudomonas, and Alcaligenes playing key roles in the degradation of the pollutants. Initially, Pseudomonas had a high abundance (20.08% at day 20). On day 75, it decreased to 4.67%. It then slightly recovered to 7.25% on day 120. Alcaligenes sp. SY1 remained more stable: It started at 13.08% on day 20. By day 75, it was still at 12.09%. It increased slightly to 15.55% by day 120. Thiobacillus became dominant, increasing from 0.37% to 40.24% in 75 days, then slightly decreasing to 32.72% by day 120.

Li et al. [215] performed the biodegradation of binary chlorinated volatile organic compounds (dichloromethane (DCM) and 1,2-dichloroethane (1,2-DCE)) using a bioaugmented BTF inoculated with Methylobacterium rhodesianum H13, Starkeya sp. T-2, and activated sludge. The BTF achieved high removal efficiencies (80% for DCM and 72% for 1,2-DCE) and superior mineralization (65.9%). The system performed well at inlet loads below 50 g·m⁻³·h⁻¹, but higher loads caused a sharp decline in RE due to microbial degradation limitations and toxicity effects. Interaction analysis revealed that 1,2-DCE enhanced DCM biodegradation, while DCM inhibited 1,2-DCE removal, influencing microbial competition. The BTF tolerated shock loads up to 400 mg m⁻³, but concentrations above 600 mg·m⁻³ led to significant efficiency losses. M. rhodesianum H13 outcompeted Starkeya sp. T-2, ensuring long-term system stability, while Hyphomicrobium and Sphingomonas also contributed to pollutant removal.

Lv et al. [137] tested the degradation of seven aromatic VOC (benzene, toluene, ethylbenzene, chlorobenzene, m-xylene, p-chlorotoluene, and p-chlorotrifluorotoluene) with a synthetic bacterial consortium. The VOC-degrading bacteria were composed of Bacillus benzoevorans, Rhodococcus qingshengii, Pseudomonas sp., Aeromonas sp., and Achromobacter sp. The consortium proved to be very effective in degrading bezene (>85%), but was much lower for the other six VOC (<45%). The addition of organic carbon and nitrogen promoted microbial growth and increased the degradation efficiency to a certain extent. The efficiency was highly dependent on the source, but substantially improved the degradation of rate of toluene (>80%), ethylbenzene (>84%), m-xylene (>72%), p-chlorotoluene (>45%), and p-chlorotrifluorotoluene (>54%). The authors address that the degradation rate of VOC with nitro and halogen atoms in their chain could hinder biodegradability.

The work of Rybarczyk et al. [196] used a different kind of bioaugmentation; by adding enzyme extracts from white rot fungi with a fresh mineral salt medium to the BTFs, it resulted in enhanced performance, particularly following episodes of acidification (around pH 4.0) within the system. The introduction of these extracts along with the fresh mineral salt medium resulted in a rapid recovery of the system’s pH to around 6.0–6.5 and a corresponding increase in BTEX RE.

Bioaugmentation may not always be successful. For example, Boada et al. [131] investigated the removal of both siloxanes (D4 and D5) and VOC (toluene, limonene, and hexane) in a BTF inoculated with anaerobic-activated sludge. The authors introduced species such as Methylibium sp., Alicycliphilus denitrificans, Rhodococcus erythropolis, and Pseudomonas aeruginosa. However, only a few bioaugmented strains (Methylibium sp. and Alicycliphilus denitrificans) were consistently detected in the reactor. The authors highlight that packing material (lava rock vs. activated carbon) influenced microbial selection, with activated carbon enhancing bacterial adhesion and increasing microbial diversity, but it did not substantially improve the integration of bioaugmented strains. However, limonene and toluene were completely removed, with an average hexane RE of 14%. The authors hypothesize that the growth of the bioaugmented microorganism in the system may have been prevented due to the competition with other bacteria already established from the inoculated anaerobic-activated sludge.

Overall, bioaugmentation holds promise as a viable approach for VOC treatment in BTFs, providing a cost-effective and efficient solution for reducing VOC contaminants and improving environmental remediation efforts. It requires the careful selection of microorganisms, thorough planning of application methods, and meticulous management of environmental conditions to achieve desired outcomes effectively and efficiently. Further research and implementation of bioaugmentation strategies can contribute to more sustainable and effective VOC remediation technologies.

3.5. Bioreactors for Bioscrubber VOC Treatment Configuration

Bacteria growing in biofilm cells have increased stability when compared to freely suspended cells concerning variations in the influent composition and abrupt changes in load, temperature, and toxicity. Therefore, biofilm increases the potential for pollutant removal, mainly due to the wide variety of microbial functional groups present in these environments [216]. Fixed film carrier technology increases treatment capacity, reduces sludge production, increases process stability and sludge settleability, and reduces carbon footprint, costs, and solids in the sedimentation tank [217,218]. These are ideal characteristics since it is undesirable to recircle a high amount of solids back to the trickling filter, and they increase liquid-phase VOC RE and, thus, the overall system efficiency.

Cheng et al. [219] investigated biological treatment feasibility for two kinds of wastewaters, collected from wet scrubbers in the semiconductor industry, contaminated with VOC. The authors performed preliminary batch tests to assess the sludge biodegradability capacity under different initial substrate/biomass ratios to assess the RE of the microorganisms under aerobic and anoxic conditions. In the aerobic experiments, dissolved oxygen was kept above 3.0 mg·L⁻¹, while anoxic batches were filled with nitrogen gas to ensure an oxygen-free environment. Due to the VOC volatilization under aerobic conditions, wastewater containing highly volatile organic matter seems to not be suitable for aerated treatment conditions. As for the anoxic environment, the reduction was only observed for sludge with nitrate addition, setting a stoichiometric ratio of chemical oxygen demand (COD) to nitrate (COD/N) of about 3.5 gCOD·gN⁻¹.

Cheng et al. [219] used MBBR tests to treat VOC-containing wastewaters from wet scrubbers in the semiconductor industry. Two lab-scale MBBR with 4.2 L of working volume were individually operated, one operating under aerobic conditions and the other under anoxic conditions. Due to the preliminary batch test, wastewater containing a high VOC concentration was fed to the anoxic reactor, while the wastewater with a low VOC concentration was fed to the aerobic reactor. Porous polyurethane carriers (BioNET) were chosen for biofilm development in a filling ratio proportion of around 20% (v/v). The experiments with the aerobic bioreactor lasted 280 days, achieving the highest COD removal rate of 98.9 mgCOD·L⁻¹·h⁻¹ with 81% COD RE at a hydraulic retention time of 1 day. On the other hand, the anoxic bioreactor reached above 80% COD RE, attaining a COD removal rate of 16.5 mgCOD·L⁻¹·h⁻¹ for an HRT of 2 days after 380 days of operation.

He et al. [220] tested the removal of toluene by anoxic denitrification in biofilm reactors. The authors added nitrate as an electron acceptor in one reactor and compared it to a traditional bioreactor. In the denitrification reactor, toluene removal was promoted by aerobic degradation, intensifying the concentration gradient inside the biofilm, with toluene being removed by denitrification inside the biofilm microenvironment. When the inlet concentration increased from 50 mg·m⁻³ to 3440 mg·m⁻³, the RE of the denitrifying reactor remained over 94.1%, while the RE of the traditional bioreactor decreased to 82.9%. The highest removal capacity of the denitrifying reactor was 127.2 g·m⁻³·h⁻¹, which was 11.8% higher than that of the traditional bioreactor.

Anoxic configuration was also employed to treat ethanethiol [221] and dimethyl sulfide [222]. In both articles, the organic compounds were successfully degraded to elemental sulfur, and other forms of sulfur were formed. All inlet ethanethiol concentrations had an average RE of 91% and EC of 24.74 g·m⁻³·h⁻¹. In similarity, dimethyl sulfide degradation in the bioreactor tank led to 89% removal.

The use of biofilm-based technologies in wastewater treatment, especially in handling VOC in varied industrial contexts, shows significant advantages over traditional methods that rely on freely suspended microbial cells. Biofilms offer enhanced stability under fluctuating operational conditions, such as shifts in influent composition, load, temperature, and toxicity, making them particularly effective in environments where such changes are common.

3.6. Combination of Bioreactors with Other Technologies

This section explores coupled treatments with physical and chemical processes to more effectively address a wider range of contaminants. We will discuss various combinations of bioreactors with other technologies, such as ozonation, UV irradiation, and nanotechnology, and provide a detailed look at their mechanisms, benefits, and applications. An Advanced Oxidation Processes (AOP)-integrated approach enhances the transformation of toxic organic compounds into non-toxic odorless byproducts, primarily through the action of hydroxyl radicals generated by AOP. These radicals are highly effective in breaking down complex contaminants, which are then further processed by biological systems, resulting in complete mineralization [223,224].

3.6.1. Ozone

Ozone injection can effectively manage biomass accumulation and maintain low bed pressure drops, which are essential for the long-term operational stability of BFs. In a study performed by Xi et al. [225], optimal ozone concentrations ranging from 180 to 220 mg·m⁻³ were determined to effectively stabilize bed pressure and preserve the toluene removal capacity (EC of 140 g·m⁻³·h⁻¹). Higher doses of ozone significantly reduced the rates of biomass accumulation, thus minimizing the risks of bed clogging, which is a common issue in long-term biofiltration systems. The study of Senatore et al. [226] demonstrates that pre-treating the VOC-laden air with ozone significantly enhances the biodegradability of the pollutants, leading to improved removal efficiencies in the BS unit. A toluene RE of up to 95% under an ozone loading rate of 920 g·m⁻³·h⁻¹ corresponded to an EC of 13.81 ± 2.0 g·m⁻³·h⁻¹. The presence of ozone-modified byproducts was found to support microbial growth and activity, suggesting a synergistic interaction between chemical and biological treatment mechanisms. Saingam et al. [227] also studied the impact of ozone injection on BFs over a prolonged period (160 days) on toluene-contaminated waste gases. The research indicates that injecting 200 mg·m⁻³ of ozone not only significantly mitigates biomass accumulation (only 46% of that in the control without ozone), but also enhances toluene mineralization rates from 83% to 91%. The continuous addition of ozone in BFs has also shown substantial benefits in controlling biomass clogging and enhancing system performance in treating ethyl acetate vapors. Ozone at concentrations around 90 ppb not only maintains higher removal efficiencies, but also significantly modifies the microbial community within the BFs due to resulting in more acidic operating conditions [228,229]. This highlights the potential of this approach to outperform conventional biological treatment methods by addressing the limitations typically associated with the biodegradation of hydrophobic compounds like toluene, as well as a lower tendency for clogging.

3.6.2. Photodegradation

Photodegradation is an advanced oxidation technology known for the generation of oxidative radicals that play a crucial role in altering the structure and physicochemical properties of various VOC. In this process, high-energy photons can be excited and activate substances under mild conditions. These activated substances and photons are highly effective in rapidly breaking down VOC into simpler, non-harmful compounds, offering an efficient solution for VOC degradation [230]. Ultraviolet light (UV) treatment plays a crucial role in mitigating common BF issues, particularly biomass overgrowth and system clogging, by inhibiting microbial proliferation, thereby reducing the pressure drop and maintaining consistent airflow through the BF [230]. Regarding the costs of this type of system, the BF system has higher initial costs but lower operational costs compared to the UV system. The costs for the combined system fall between those of the standalone systems, with operational costs increasing and capital costs decreasing as the UV system/BF ratio rises [231]. Potential drawbacks include the generation of partially oxidized VOC intermediates, some of which may be more toxic than the parent compounds, and the formation of residual ozone, which can pose secondary environmental and health risks. Proper system optimization includes the precise control of UV intensity, exposure time, and integration with downstream biofiltration [230,232]. This type of approach has been shown to be able to improve the removal of VOC that offer resistance to biodegradation, such as toluene [233,234], xylene [233,235], chlorobenzene [236], hexane [237], ethylbenzene [238], styrene [239], α-Pinene [240], and dichloromethane [241].

An example of the practical application of this technology is the study of Zeng et al. [242], which exploits the integration of a photocatalytic reactor (PCR) with UV light combined with TiO₂ and Ni catalysts with a pilot-scale BTF to enhance the treatment of VOC emitted from an automobile paint-manufacturing workshop. The study highlights the role of the PCR in pre-treating waste gases before biofiltration. The standalone BTF struggled with high inlet VOC loads, achieving removal efficiencies of only 72–77% for aromatic compounds, whereas the combined PCR-BTF system achieved nearly 99%. The study also found that the pre-treatment with UV photocatalysis mitigated the negative effects of fluctuating VOC concentrations. Another example is the pilot-scale experiment to treat VOC emitted from a paint production facility performed by He et al. [243], which integrated a BTF and photocatalytic oxidation (PCO) UV-activated TiO₂ photocatalyst stainless steel system. The combined BTF-PCO system achieved an RE of 95.6%, outperforming standalone PCO (88.7%) and BTF (73.7%). It was particularly effective in improving the removal of ethyl acetate, toluene, ethylbenzene, xylene, ethyltoluene, and trimethylbenzene.

3.6.3. Non-Thermal Plasma

Non-thermal plasma (NTP) can be defined as gas consisting of electrons, highly excited atoms and molecules, ions, radicals, photons, and neutral particles in which the electrons have a much higher energy than the neutral gas particles [244]. Plasma, often referred to as the fourth state of matter, consists of a mixture of free electrons, ions, and photons created from a gas by applying a strong electric field. This field ionizes the gas, making it electrically conductive and changing the density of its charged particles across a broad range. NTP can be produced at atmospheric pressure in air or with supporting gases, such as He, Ne, Ar, O₂, and N₂ [245]. NTP may be generated by a variety of electrical discharges such as corona, DBD, radiofrequency, microwave, spark, atmospheric pressure jets, and gliding, mostly under low pressure [246,247,248]. The NTP technique is a promising technology for the destruction of VOC owing to its unique features such as simple technology process, quick start-up, strong adaptability, capability of handling multiple pollutants simultaneously, high treatment efficiency, versatility to a wide range of temperatures, and convenient operation management [247,249]. Hybrid systems integrating NTP with biological reactors significantly enhance VOC RE compared to standalone biofiltration. Energy consumption, design, and ongoing maintenance remain challenges in optimizing the combined systems [250], but research shows that plasma–biofiltration hybrid technologies can be economically viable.

DBD has been applied by Schiavon et al. [251] to pre-treat an air stream contaminated by a mixture of VOC (toluene, n-heptane, p-xylene, ethylbenzene, and benzene). The NTP treatment significantly leveled the concentration before reaching the BF and effectively generated more soluble VOC, which helped to reduce the VOC concentration. However, benzene was more resistant and required higher energy input and the authors highlight that further optimization is needed to prevent potential toxicity effects of the technology on biofilm health. DBD was also tested by Martini et al. [252]. In their case, VOC from a waste composting facility were mainly composed of α-pinene, ethyl isovalerate, and dimethyl disulphide. Once again, NTP effectively improves VOC solubility and biodegradability, making biofiltration more efficient. However, the authors addressed high energy consumption, a key challenge in the application of NTP. The study explored the transition from a laboratory-scale reactor to an industrial-scale system, identifying optimal energy input and RE for different process setups. To battle this weakness in the process, the authors proposed a multi-stage treatment configuration that reduces energy costs by 49%.

Huang et al. [253] compared pulsed corona discharge and direct current corona discharge plasma to treat toluene gas and later combined the plasma pre-treatment with a trickling BF inoculated with Acinetobacter baumannii. Pulsed corona discharge plasma achieved a higher toluene RE (77.11%) than DC corona discharge plasma (72.62%) under optimal conditions. The energy efficiency of pulsed corona discharge was also higher, reaching 1.515 g·kW⁻¹·h⁻¹ at 4.0 kV, while DC corona discharge exhibited lower energy efficiency. Regarding the BF, it achieved 82.71% toluene removal under optimized conditions (1000 mg·m⁻³ initial toluene concentration, 40 L·h⁻¹ air flow rate). Combining the plasma pre-treatment with biofiltration significantly improved RE, reaching a maximum of 97.84%, even at increased toluene concentrations (3000 mg·m⁻³), where single BFs showed reduced performance.

3.7. Microbial Fuel Cell

In recent years, MFCs have received significant attention due to their capacity to produce green energy from wastewater. MFCs can harness the energetic potential of various waste streams that can be degraded by microbes [254,255,256,257,258,259,260]. Therefore, MFCs are bioelectrochemical systems that generate electricity by using electrochemically active microorganisms as catalysts [261]. Pollutants can be removed either at the anode as electron donors via microbial catalyzed oxidization or at the cathode as electron acceptors through reduction. Some contaminants can act as electron mediators at the anode or cathode [262].

The components of an MFC are an anode, a cathode, and the electrolyte. An ion exchange membrane may be used to separate the anodic and cathodic chambers. Microorganisms can thrive on the anodic chamber in a planktonic or biofilm state, oxidizing substrates and producing electrons, protons, and other metabolites as reaction products. The electrons released by the microbes are gathered by the anode, which conducts them into the cathode through an external circuit. To maintain the balance of the system, the produced protons percolate through an ion membrane or by diffusion to the cathode to be reduced by the electrons received. It is the established flow of electrons that generates an electric current through an external loader [262,263]. The protons reduced at the anode in the presence of oxygen will form water as a by-product; thus, oxygen is continuously consumed to maintain the potential for electricity generation according to Equation (16) [262,263].

$${\mathrm{O}}_2+4{\mathrm{H}}^{+}+4{\mathrm{e}}^{-}\to 2{\mathrm{H}}_2\mathrm{O} \left({\mathrm{E}}_0=1.23 \mathrm{V}\right)$$

(16)

The cathodic reaction can be improved by the use of catalytic-coated electrodes, such as platinum, activated carbon, metal oxides, or bio-based catalysts, which enhance electron transfer efficiency, reduce activation energy barriers, and ultimately lead to increased power density and stability in microbial fuel cell operation [264]. Alternatively, electron donors to oxygen can be used to induce higher voltage output, recalcitrant pollutant treatment, or valuable chemical recover, such as ferriciyanide, nitrogen species, persulfate, permanganate, manganese dioxide, mercury, copper, chromium, triiodide, hydrogen peroxide, carbon dioxide, perchlorate, vanadium, uranium, chloroethenes, and 2-Chlorophenol [262].

At the anode, the process is predominantly carried out by electrogenic bacteria, which can use direct electron transfer (DET) mechanisms, often through structures such as conductive pili, nanowires, or cytochromes. These structures allow electrons to move from the microbial cells directly to the electrode without mediators. This type of mechanism is supported by specialized membrane-bound proteins, notably the c-type cytochromes, which are heavily involved in shuttling electrons to the electrode surface. Alternatively, mediated electron transfer (MET) can occur, where soluble redox mediators shuttle electrons from the microbial cells to the anode. These mediators can be naturally secreted by biochemicals such as flavins or artificially added chemicals like methylene blue [265,266,267,268,269]. At the cathode, the process is less extensively explored but involves either direct or mediated electron transfer mechanisms like those at the anode. In direct cathodic electron transfer, electrons move from the cathode to specific acceptors in microbial cells through direct contact. This process may involve outer membrane cytochromes of bacteria. On the other hand, indirect methods include the use of redox mediators, which are reduced at the cathode and then interact with microbes to facilitate the reduction reactions necessary for processes like denitrification or bioremediation [270,271].

Power output and COD RE vary with the type of MFC (e.g., single versus double chamber, Figure 5), type of wastewater, MFC configuration and design, anode and cathode materials, inoculated microbes, electrode surface area, influent COD concentration, dissolved oxygen, hydraulic retention time, temperature, pH, salts and solid pollutants, type and electrolyte conductivity, and specific organic loading rate [255,263,267,272,273,274,275,276,277,278]. A simple molecular structure that supports rapid microbial metabolism enhances power generation and electron transport wastewaters with complex compositions, such as those from municipal or agricultural sources, providing long-term stability to MFC. On the other hand, high concentrations of inhibitory chemicals such as heavy metals or toxic organics found in landfill leachate or mining wastewater are generally unsuitable for MFC [278]. Even so, MFC with specific microorganism consortia have been found to be effective in degrading chlorinated organic compounds, such as trichloroethylene, dichloroethane, and chlorophenol [279,280]. Metals like chromium [281], copper [282], silver [283], and cadmium [284] can be electrochemically reduced on the cathode’s surface, changing from a more toxic state to a less toxic one. For example, Cr(VI), which is highly toxic, can be reduced to the less harmful Cr(III) [281]. In this process, the cathode interacts directly with the heavy metals, facilitating their reduction and deposition onto the cathode surface, thereby purifying the water from these toxic substances [285].

To increase the efficiency of an MFC key genera such as Geobacter and Shewanella (particulary Geobacter sulfurreducens and Shewanella oneidensis), they must be robust in their electron transfer capabilities. For instance, Geobacter species are renowned for their ability to reduce iron and other metals through extracellular electron transfer mechanisms, significantly influencing biogeochemical cycling and contaminant remediation processes. Shewanella species, on the other hand, are noted for their versatile respiratory capabilities, which allow them to utilize a variety of organic and inorganic substances as electron acceptors [269,286]. However, there are many other electroactive microorganisms such as Pseudomonas aeruginosa, Bacillus subtilis, Methanococcus maripaludis, Saccharomyces cerevisiae, and Klebsiella pneumoniae, among many others [287,288].

Non-exoelectrogens can play crucial supportive roles in bioelectrochemical systems by helping in the degradation of complex organic materials, which exoelectrogens cannot directly utilize and by producing substrates, nutrients, or other beneficial compounds that enhance the activity of exoelectrogens, or even adjusting pH levels or removing inhibitory substances [289]. This is the case of Clostridium cellulolyticum and Enterobacter aerogenes, mentioned for their role in fermentative processes, contributing indirectly to the electron transfer chain by breaking down complex organic substances into simpler forms that can be utilized by other electrogenic bacteria [290,291].

ElMekawy et al. [292] point out additional factors for variations in the power output, such as (i) activation overpotential, which is the potential difference required for the cell to produce a specific current and can be hindered by electron accumulation at the electrode surface [293]; (ii) ohmic losses associated with a voltage drop due to the resistance in the flow and transfer of electrons in the electric circuit and the movement of ions through the electrolyte and membrane [294,295]; and (iii) concentration polarization occurring when a component increases or decreases in concentration at the boundary layer of the membrane surface [296]. It is common to incur a substantial loss of electrons due to the incomplete mineralization of the substrate in the anode chamber, the conservation and growth of the anode bacteria, and resistance associated with the system operation [297]. For this reason, bacteria capable of oxidizing organic matter and transferring the electrons to an electron acceptor outside of their cells (exoelectrogenic bacteria) have to be able to maintain a flow of electrons for the complete mineralization of toxic pollutants, to obtain a short operation time, and to accelerate electron transfer [292,298]. Additionally, the performance of a co-culture-inoculated MFC is significantly impacted by the microbial community consortia, with particular emphasis on intermicrobial interactions.

Wu et al. [299] tested potassium ferricyanide (0–200 mM), an electron acceptor, in a dual-chambered MFC for benzene biodegradation. The biomass pellet to be used in the MFC was developed by the authors through a series of acclimatization and stabilization periods of bacteria collected from the oil cracking wastewater treatment plant of Nan-Ya Plastics. The MFC chambers had 0.8 L, separated by a PEM with a reactional area of 16.6 cm², and both anode and cathode were made of carbon cloth connected to a 1 kV external resistor in a loop. The power produced was in the range of 0.0276 to 2.1 mW·m⁻², achieving maximum production for 200 mM of potassium ferricyanide in the cathode chamber. To achieve the complete benzene degradation of 10.87 mg·L⁻¹, a retention time of 22–24.5 h and a potassium ferricyanide concentration between 0 and 150 mM were required. When increasing the concentration to 200 mM, the time needed for complete degradation increased to 35 h.

In another study of Wu et al. [300], electricity generation by an MFC under varying toluene doses with and without the addition of pyocyanin (as the electron mediator) was examined. The configuration of the MFC was the same as for the previously stated paper, with the toluene medium used to generate electricity being inserted in the anode chamber, while 50 mM of potassium ferricyanide solution, used in the process as an electron acceptor, was introduced into the cathode chamber. Electricity generation increased with toluene concentration, being 2.07, 2.74 and 4.69 mW·m⁻² for 11, 33, and 55 mg·L⁻¹, respectively. However, the total removal time increased from 1.2 days at an initial toluene dose of 11 mg·L⁻¹ to 5.0 days for an initial dose of 55 mg·L⁻¹. When the authors added pyocyanin, electricity generation was greatly increased from 4.69 to 21.76 mW·m⁻². The authors reveal that the use of this toxin lowered internal resistance from 500 to 100 Ω. Therefore, the system impedance was reduced, electron density increased, and the mineralization of toluene was accelerated.

Wu and Lin [301] combined a biotrickling filter with a microbial fuel cell (BTF-MFC) inoculated with acclimated microorganisms from a petroleum refinery wastewater treatment plant to treat ethyl acetate off-gas while generating electricity. The authors conducted all the tests over a period of 150 days. Conductive coke was used as the packing medium in a trickling filter composed of two layers, separated by porous acrylic plates, with dimensions of 8 × 8 × 15 cm³. The anode and the cathode were made of graphite rods 1.6 cm in diameter and 15 cm in length, and polyvinyl alcohol membrane electrode assembly with 13 × 6 cm², respectively. Each layer can operate independently or in series. The BF-MFC system achieved removal efficiencies between 91.3 and 99.2% for organic loadings of 34.3–65.1 g·m⁻³·h⁻¹ and an EBRT of 61–115 s. During these intervals, the system produced between 538 and 596 mV. When EBRT was decreased to 46 s and 20 s by increasing the organic loading rate, the removal dropped to 81.3% and 33.4%, respectively. Additionally, the closed-circuit voltage decreased from 504 mV to 450 mV. The authors attribute this decrease to the inability of the biofilm on the conductive coke surface layer not being able to withstand the increase in organic loading rate, representing a zero-order reaction of biofilm reaction control.

Wu et al. [302] complemented Wu and Lin’s work [301] by investigating the continuous power production in a BTF-MFC with waste gas contaminated with ethyl acetate. The authors used a polyvinyl alcohol membrane electrode to deliver protons from the anode to the cathode effectively. Increasing the ethyl acetate concentration from 0.18 to 1.44 g·m⁻³ increased the voltage from 49.4 mV to 658 mV. Elimination efficiency was near 100% for organic loading rates ranging from 14.41 to 29.58 mg·L⁻¹. The maximum power density (49.1 mW·m⁻²) and elimination capacity (83.8 g·m⁻¹·h⁻¹) were obtained for 400 g·m⁻³ of ethyl acetate and 113.15 g·m⁻³·h⁻¹ of inlet loading rate. However, the RE was only 73.86%.

Oveisi et al. [272] explored the ability of an MFC to degrade and produce energy from styrene in the concentration range of 27–105.7 mg·L⁻¹ with an adapted and unadapted activated sludge inoculum. Additionally, the authors tested methylene blue, methyl orange, and methyl red as electron acceptors at a concentration of 300 μM as a means to improve the system’s performance. The first conclusion drawn by the authors is that the unadopted inoculum led to higher degradation efficiency and higher power production. This could be related to the presence of an aromatic degrading enzyme in the adapted inoculum. The second conclusion is that the addition of electron intermediate materials slightly increased the time required to reach steady-state voltage, but also increased the generation of electricity. In particular, methylene blue’s high permeability across the cell membrane favored energy production, increasing the steady voltage of the system by 22%. Moreover, by adding the soluble dyes, the rate of styrene degradation improved, especially in the presence of methylene blue.

A different approach is the one performed by Dai et al. [303]. The article presents a detailed study on a novel vertically configured photocatalytic microbial fuel cell (photo-MFC) designed for the degradation of gaseous toluene and electricity generation. The photo-MFC system achieved a toluene RE of 60.65% under light conditions. This was significantly higher than the removal efficiencies achieved by MFC alone (17.81%) and photocatalysis alone using ZnO/ND (37.16%). The peak power density of the system was 120 mW m⁻² under light, which was about 1.57 times higher than that under dark conditions (76 mW m⁻²). The current density increased from 0.78 A/m² in dark to 1.07 A/m² under light. The reaction kinetics followed pseudo-first-order kinetics with the kinetic constant for toluene degradation being considerably higher under illuminated conditions (0.0117 min⁻¹) compared to dark (0.0020 min⁻¹). The system’s efficiency in toluene removal varied with the external resistance: 60.65% RE with 50 Ω and 49.37% with 1000 Ω, showing that lower resistance favored higher degradation rates. The microbial community was dominated by Proteobacteria and Firmicutes.

Liu et al. [304] investigated the RE of isopropyl alcohol vapor at concentrations between 0.36 and 4.42 g·m⁻³ by a hollow trickling bed microbial fuel cell. The design accommodates a trickling bed anode in a plexiglass reactor divided into three horizontal sections. The upper section contained a gas inlet and nozzles sprinkling the circulating liquid onto the packing material. The reaction occurred in the middle section composed of conductive coke as packing material to support microbial attachment and biofilm formation. The inoculated microorganisms were from an activated sludge from a petroleum refining industry. Coke particles helped to transport the electrons from the microbes to the main anode. This anode was composed of six graphite rods (surface area: 217 cm²) inserted vertically into the packing material and connected with titanium wires. The total volume of trickle bed including main anodes and auxiliary anodes was 4417 cm³. The carbon cloth cathode that lined the inner tube was directly exposed to the air and separated from the anode chamber by a permeable exchange layer made of polyvinyl hydrogel with a volume of 224 cm³. The liquid was collected in the lower section (2.5 L) and circulated back to the top of the system at a flow rate of 650 mL·min⁻¹. The air flow rate was of 4.417 L·min⁻¹. With this simple system, the authors achieved an RE of 95% and a mean voltage of 153.3 mV for an inlet concentration of 0.36 g·m⁻³. Maximum voltage was generated for 0.73 g·m⁻³, but RE was reduced to 91%. The authors attribute the low voltage at lower inlet concentrations to the incomplete growth of the biofilm at low inlet concentrations. Higher concentrations hindered both energy production and RE.

Liu et al. [305] propose to improve the performance of the BTF using MFC in treating exhaust gas by adjusting the oxygen content of the anode tank. Three main conclusions can be drawn from this article. The first is that the use of carbide porous ceramic rings as anodes instead of coke increased the biofilm mass, and exhibited a higher average RE. However, the power density of ceramic rings was 5.64–14.8% lower than that of coke. Therefore, it is not a suitable cathode. The second is that the microaerobic conditions could be controlled and improved by the aforementioned method, which is achieved by continuously adding sodium sulfite. The chemical decreased the average dissolved oxygen and generated a voltage of 477 mV with a maximum power density of 71.7 mW·m⁻³. It demonstrates that the competition of oxygen and anodes for electrons can be solved with this method. The third conclusion is that the combination of MFC (with carbon cloth) and a BTF can effectively increase power due to the electrochemical reaction.

Liu et al. [306] studied the application of a trickle bed MFC using the activated sludge of a wastewater treatment plant as inoculum. The designed reactor was composed of coaxial acrylic cylinders divided into three segments. The mains anode was made of a graphite rod and a cathode carbon cylinder, and encapsulated sodium sulfite was used as the deoxidizing packing material. The results showed a high efficiency in removing isopropanol with a mean inlet of 0.74 g·m⁻³ to 0.002 g·m⁻³ and a power and current density of 486.6 mW·m⁻³ and 3126 mA·m⁻³. Additionally, the authors highlight the capability of the microorganisms to convert isopropanol to CO₂, reducing the expected acetone concentration in the outflow to values between 8 and 31% of the initial isopropanol.

Wang et al. [307] constructed a BTF with a single-chamber MFC at the bottom, allowing for toluene removal and energy production in the same column. The authors achieved a stable RE of 80% and a stable current of 0.25 mA with a maximum power density output of 65.4 mW·m⁻³ for toluene gas as the sole substrate with an EBRT of 180 s. The authors also found that various intermediate products including cresol, benzyl alcohol, hydroxybenzoic acid, and dihydroxybenzoic acid were produced and could be further used as the substrate for electricity production by electrochemically active bacteria in the anode chamber.

The study of Yang et al. [308] on the integration of a microbial fuel cell with a bio-trickling filter for the removal of ethyl acetate from gaseous streams revealed significant enhancements in the biofilter’s performance. Two types of BTF were compared: one with a built-in proton exchange membrane (M-BTF) and a conventional one (O-BTF). The M-BTF demonstrated a superior biodegradation capacity, faster startup, produced more biomass, and exhibited higher removal efficiencies, peaking at 9.9%, which was better than the O-BTF. Upon initiation, both BTFs achieved an RE of around 60% for ethyl acetate with an EBRT of 91 s. However, M-BTF (with microbial fuel cell integration) demonstrated a faster startup, reaching an 85% RE in just 5 days compared to 8 days for O-BTF. During the study, M-BTF consistently produced more biomass compared to O-BTF and a maximum output voltage of 536 mV was noted for M-BTF, which correlated positively with the increase in inlet ethyl acetate loadings. M-BTF was found to be selective for electroactive bacteria such as Pseudomonas, Flavobacterium, and Geobacter.

Despite the promise shown by MFC technology, three major drawbacks have yet to be settled for real-world feasibility: insufficient power output, electrodes stability and durability, and high costs (electrode materials, membranes, and cathode catalyst) [309,310,311,312]. In fact, efficiency declines when increasing scale, often by factors of 100 to 10,000, caused by increased internal resistance, non-uniform mixing, and pressure on microbial activity and cathode efficiency [312]. Therefore, the influence of MFC components on the conversion of organic substrates into electrical charge (Coulombic efficiency) and power output are key areas underlying its development [313]. Thus, industrial confidence in establishing the technology is yet to be achieved. Some authors have explored pilot-scale MFC, although these have not been specifically applied to VOC.

Rossi et al. [314] aimed to evaluate the performance of a pilot-scale air cathode MFC (850 L of reactional volume) using domestic wastewater. The authors reported low energy efficiency covering around 5% of energy consumed for pumping, cathode fouling, and leakage that decreased performance by 60%, and several operation issues particularly during weekends and periods with lower wastewater availability. Even so, the authors were able to improve efficiency when coupling the MFC with a BF, improving chemical and biological oxygen demand and ammonia removal by 83.6%, 70.5%, and 84.0%, respectively.

Blatter et al. [315] also studied the performance and reliability of a large-scale (1000 L) MFC, designed for municipal wastewater treatment, operating continuously for 18 months. The authors point out major drawbacks: (i) sediment accumulation, (ii) voltage reversals, (iii) power fluctuations, and (iv) limited energy efficiency. They recommend optimizing sediment prevention and management strategies, as well as increasing automation and monitoring to better handle fluctuations in wastewater composition.

Heinrichmeier et al. [316] tested the reliability and performance of a 1000 L single-chamber MFC system operating under real conditions at a municipal wastewater treatment plant over a 6-month period using submerged multi-electrode modules with air-cathodes. The authors demonstrated the feasibility of integrating large MFC modules into existing wastewater infrastructure while consistently removing part of the COD without significantly affecting nitrogen levels for subsequent biological treatment. However, electricity generation decreased over time due to cathode fouling and salt precipitation. In their context, the authors propose cleaning the cathode with a 15% HCl solution, but the recovery was incomplete due to the slow dissolution of dolomite deposits. Additionally, rain events caused a drop in conductivity and power output to be decreased by around 51–53%. Future recommendations include better cathode cleaning strategies and improving energy yields and electrode longevity. Heinrichmeier et al. [317], in an earlier study using a 45 L pilot-scale MFC, pointed out that long inoculation times provoke the temporary inhibition of biofilms under high free ammonia conditions; cells dried due to pump clogging; and design inefficiencies related to a high proportion of dead space inside the MFC caused inefficiency.

Several improvement areas have been proposed by Jalili et al. [318] to enhance the cost-effectiveness and scalability of MFC, namely (i) the development of anodes using metal oxides and conductive polymers, (ii) the investigation of cathodes based on heteroatom-doped carbons and transition metal complexes, (iii) the design of proton exchange membranes with high conductivity and antifouling properties, (iv) the optimization of reactor configurations to mitigate internal resistance and enhance mass transfer, (v) the construction of compact, portable, and single-chamber MFCs with air cathodes, (vi) the development of modular and stackable MFC systems, possibly employing 3D printing technologies, (vii) the exploration of cooperative microbial communities and metabolic engineering approaches to increase electricity generation, and (viii) the elucidation of direct interspecies electron transfer mechanisms among electroactive microbes. Bird et al. [319] recommend simplifying the approach by adopting standard performance metrics for pollutants removal, loading rates, power density, coulombic efficiency, and energy recovery to use equations and heatmap tools to predict MFC design and to focus on integrating MFC into decentralized or pre-treatment systems.

To boost MFC efficiency, Ali et al. [320] propose nanotechnology-based modifications of anode materials using carbon nanostructures, such as carbon nanotubes (CNT), graphene, Fe₃O₄ nanoparticles, and carbon nanofibers (CNF), which enhance electron transfer, increase surface area for biofilm formation, and improve overall electrocatalytic performance. These nanoscale enhancements not only increase power output, but also contribute to more stable and efficient wastewater treatment. Li et al. [321] propose an alternative strategy for developing cost-effective and scalable MFC systems, suggesting biotic separators consisting of a porous skeletal material (e.g., nonwovens, sponges, filter cloths) colonized by an in situ formed microbial biofilm. These are self-regenerating and offer a low area-specific resistance and hydrodynamic continuity, thereby enhancing both operational simplicity and long-term performance. The author demonstrates that biotic separators reduce capital costs by orders of magnitude as low as USD 0.2 ·m⁻². This is a major improvement when compared to costs of non-permeable membranes, which can represent a significant part of the investment (e.g., Nafion or CEM: USD 80–1400 ·m⁻²). Compared to conventional ion exchange membranes, biotic separators also reduce capital costs and simultaneously contribute to pollutant degradation (COD removal and simultaneous nitrification–denitrification reactions), making them particularly attractive for decentralized wastewater treatment applications.

MFC represents a promising approach for both wastewater treatment and sustainable green energy generation, but capital costs per unit of treated wastewater are still 20–50% higher than those of conventional activated sludge systems [312]. Ongoing research reveals valuable insights into the optimization of energy generation, encompassing innovative components, advanced materials, ion migration mechanisms, intermicrobial interactions, efficient electron exchange pathways, the role of electron mediators, and the influence of operational variables. By integrating MFC with other complementary approaches, a potential roadmap to robust industrial implementation is conceivable. However, further investigations are imperative to consolidate these findings and pave the way for the practical realization of MFC in real-world applications.

4. Conclusions and Future Perspectives

Growing environmental and public health concerns related to VOC have driven the need to develop more sustainable treatment technologies. Biological systems, particularly BFs, BTFs, and BSs, have emerged as efficient technologies for VOC elimination with tremendous advantages on cost-efficiency, operational simplicity, and environmental impact when compared to conventional physicochemical techniques, including adsorption and thermal and catalytic oxidation.

However, the wide structural and chemical diversity of VOC presents challenging barriers to their complete biodegradation. The success of the biological treatment of VOC often depends upon compound-specific factors, such as hydrophobicity, toxicity, and bioavailability. Recent strategies, such as the use of new packing materials, biodegradable surfactants, and bioaugmentation, have shown promising improvements in system performance, suggesting that these could be important areas for future research and application.

A new and promising technology is on the energetic valorization of VOC-contaminated wastewater, generally a byproduct of biological off-gas treatment, through technologies such as MFC. These systems not only enable wastewater remediation, but also enable the production of bioenergy. Yet, limitations such as low power output, operational complexity, and material cost still constrain their scalability and broader application.

The future of biotechnologies for VOC removal is expected to rely on integrative and hybrid approaches. Potentially fruitful lines of future research and development include the following:

• Developing multifunctional packing materials that enhance pollutant degradation;
• Exploring synergistic systems that combine conventional biofiltration with other technologies (e.g., MFC, photocatalysis);
• Genetic engineering to create robust microbial consortia capable of degrading a broader range of VOC under diverse conditions;
• Machine learning to predict microbial interactions to optimize bioaugmentation and biofilm stability;
• Conducting comprehensive life cycle assessments and economic analyses to validate sustainability and cost-effectiveness at industrial scales;
• Establishing regulatory frameworks and best practices tailored to the specific VOC profiles of different industrial sectors.

This review successfully achieved its goal of mapping the current state of biological VOC treatment, emphasizing both technical viability and future valorization pathways. Environmental biotechnology will play a pivotal role in the green industrial transition, enabling the effective abatement of air pollution regarding VOC and facilitating support to circular economy strategies through energy and resource recovery. The integration of bioaugmentation and energy recovery systems marks a promising direction. Future research will focus on optimizing reactor design for complex industrial VOC mixtures and exploring the regulatory and economic incentives necessary to accelerate the industrial implementation of these technologies.