Advanced photocatalytic degradation of POPs and other contaminants: a comprehensive review on nanocomposites and heterojunctions

Abstract

The environmental persistence and toxicity of pollutants such as persistent organic pollutants (POPs), synthetic dyes, and pharmaceutical residues necessitate the development of effective and sustainable remediation strategies. This review underscores the urgent need for advanced approaches to eliminate these contaminants, with a particular focus on metal oxide-based photocatalysts, such as TiO₂, ZnO, WO₃, CuO, and others. We have explored their photocatalytic mechanisms, inherent limitations, and recent advancements, such as elemental doping and heterojunction engineering, to enhance their activity under visible light. Nanocomposite systems, especially those incorporating heterojunctions, have demonstrated significant improvements in photocatalytic efficiency by facilitating charge separation and promoting the generation of reactive species. This review provides an in-depth examination of the mechanisms and practical applications of these materials in the degradation of POPs, dyes, and pharmaceutical pollutants. Furthermore, it outlines current challenges and identifies promising directions for future research aimed at developing high-performance photocatalysts for environmental remediation.

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Ruhit Kumar Paul

Ruhit Kumar Paul is a research scholar currently pursuing his PhD in Material Science at the National Institute of Technology (NIT) Silchar, Assam, India, under the guidance of Prof. Md. Ahmaruzzaman. He is in the third semester of his PhD program, focusing on the synthesis and application of semiconductor-based nanomaterials for environmental and energy-related challenges. He completed his MSc in Chemistry from Assam University in 2022 and earned his undergraduate degree in 2020 from the same university.

Md. Ahmaruzzaman

Md. Ahmaruzzaman is a distinguished full-time professor in the Department of Chemistry at National Institute of Technology (NIT), Silchar, India. He earned his MSc in Chemistry from North Bengal University, where he also completed his undergraduate studies. Additionally, he pursued a postgraduate MTech in Energy Science and Technology at Jadavpur University and later obtained his PhD from the Indian Institute of Technology (IIT) Delhi. His area of research is nanoscience and technology, energy and environment, electrochemistry, metal–organic frameworks, and development of low-cost adsorbents for water treatment, etc. With extensive experience in academic research, he has supervised numerous postgraduate projects and has an impressive publication record, contributing over 350 research articles to esteemed international journals and conferences. His scholarly work has garnered over 18 300+ citations in reputed international journals. Prof. Ahmaruzzaman serves as a reviewer for several high-impact peer-reviewed journals and holds editorial positions in multiple esteemed publications. He is the Associate Editor of the International Journal of Environmental Analytical Chemistry, Section Editor of Current Materials Science, and Editorial Board Member of Scientific Reports and Discover Materials.

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1. Introduction

In the present day, the Earth's ecosystem is experiencing ongoing contamination from various pollutants, which vary in type and location. Certain pollutants persist in the environment that are resistant to degradation (chemical, biological, and photolytic reactions) and remain in the environment for extended periods.^1–4

These pollutants are organic compounds of natural or anthropogenic origin are called persistent organic pollutants (POPs). POPs pose a significant global concern due to their persistence, ability to travel long distances, and tendency to bioaccumulate in fat-rich tissues such as adipose tissue, liver, brain, and breast tissue.^5–7 Even at low concentrations, they exhibit high toxicity. Once inside the body, POPs are transported via the lymphatic system and bloodstream, primarily accumulating in adipocyte lipid droplets within adipose tissue. The liver also serves as a major storage site, especially for compounds like dioxins, particularly at higher exposure levels.⁸ Various POPs, particularly organochlorine insecticides, polychlorinated biphenyls (PCBs), polychlorinated dibenzodioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), and polybrominated diphenyl ethers (PBDEs), among others, primarily originate from human activities and have been extensively employed in a variety of products.⁹ For instance, organochlorine pesticides have been widely used in agriculture to manage diverse plant pests such as insects, mites, nematodes, fungi, and bacteria. In addition to their agricultural application, these compounds have played a significant role in public health initiatives targeting disease vectors like mosquitoes, which transmit pathogens including bacteria, viruses, and parasites. Furthermore, certain organochlorines have also been incorporated into household disinfectants.¹⁰ PBDEs, on the other hand, have been broadly utilized as flame retardants in numerous consumer products, including polyurethane foams used in furniture, mattresses, carpet underlays, and automobile seats, as well as in styrene-based plastics found in electronic devices like televisions and computers, and in flame-retardant textiles.¹¹

The concern about POPs arises from their ability to bioaccumulate and magnify within the food chain, particularly affecting top predator species, including humans. The most extensively documented and evident effects have been observed in birds and marine mammals. Numerous studies have reviewed how organochlorines (OCs), particularly DDE-a metabolic byproduct of DDT can impact eggshell thickness in birds of prey.^12–14 For instance, Haegele and Hudson¹⁵ reported that exposure to 40 ppm of p,p′-DDE for 96 days caused significant eggshell thinning in mallard hens (Anas platyrhynchos) by 15–20% during and shortly after treatment, and by 7.4% even 11 months later, demonstrating its long lasting impact.¹⁵

POPs harm living organisms and human health, notably by disrupting immune system function and decreasing the body's ability to resist viruses.^16,17 In the current era of environmental pollution, POPs are widespread in almost every individual. Exposure to these pollutants poses significant health risks, including hormone disruption, cancer, cardiovascular diseases, obesity, reproductive and neurological disorders, learning disabilities, and diabetes. Moreover, these pollutants can also induce defects in female embryos.¹ Persistent organic pollutants are also known for their semi-volatile nature, allowing them to exist in either vapor or adsorbed forms on atmospheric particles, facilitating long-distance transport through the atmosphere. This unique persistence, coupled with characteristics like semi-volatility, has led to the global distribution of compounds such as PCBs, even in regions where they have never been utilized.¹⁸

Water, on the other hand, is thought to be the most essential natural resource for life and is, therefore, the most vital of all. Almost 80% of the world's population currently faces issues with water supply and security as a result of fast industrialization, urbanization, and a lack of awareness among people that water is a vital resource.¹⁹ Additionally, drinking water resources are getting increasingly polluted as a result of human and industrial/governmental ignorance, and freshwater resources are thus becoming increasingly inaccessible.²⁰ Consequently, there has been a significant increase in the focus on water contamination as a matter of concern in recent times. The majority of aqueous contamination is caused by waste water tainted with dyes. Dye-contaminated wastewater is the result of various industries releasing their dye-contaminated effluents into the environment, including textile dyeing, paper manufacturing, food processing, paints, and cosmetics.^21,22 Discharged dyes pose serious health risks to aquatic life, including carcinogenesis and xenobiotoxicity. The presence of dyes in wastewater has a negative impact on aquatic ecosystems because the dyes make water appear very colored, which blocks sunlight from reaching the water's core.^23,24 Pharmaceutical residues are another cause of water contamination. Almost all aquatic matrices on every continent have shown evidence of pharmaceutical residues in recent decades.²⁵

Therefore, removing POPs and other pollutants, such as dyes and pharmaceutical residues, are crucial because of their toxicity and harmful effects on human health and the environment. By removing these pollutants, we can mitigate these risks and safeguard the health of both present and future generations.

At present, a range of techniques, including physical, chemical, biological, or hybrid methods, are proposed for the decomposition and mineralization of the aforementioned pollutants. The conventional methods used for the removal of such contaminants are surface adsorption,^26,27 biological degradation,²⁸ and membrane filtration.²⁹ Alternative technologies such as adsorption and coagulation merely concentrate pollutants by transferring them between phases without achieving complete “elimination” or “degradation” of the contaminants.³⁰ For example, adsorption employing activated carbon, resins, carbon nanotubes, or similar materials enriches and separates pollutants from water or gas phases without degrading them. However, if the desorbed contaminants from the saturated adsorbent are not effectively treated, there is a significant risk of secondary pollution.^16,31 In another approach, chemical oxidation and biotechnology techniques also face many disadvantages, including high costs with large chemical consumption, incomplete destruction, and prolonged overall treatment time.³² Among all the approaches, photocatalysis is promising and essential because of its high efficiency, energy-saving properties, mild reaction conditions, and minimal secondary pollution. This has led to a growing interest from scientific researchers.^33–36 This review offers a comprehensive and integrated analysis of the photocatalytic degradation of the POPs, dyes, and pharmaceutical pollutants-three major organic pollutants, bringing together these diverse pollutants within a unified framework. While many existing reviews tends to focus on a single class of pollutants, this work covers a broader approach. The study includes degradation mechanisms, photocatalyst's performance data along with their performance enhancement strategies across different classes of pollutant. Special focus is given to advanced strategies such as doping, heterojunction formation, and Z-scheme systems to highlight how these can address present limitations. Overall, this review not only highlights recent progress but also point out gaps in current research and suggests practical directions for future work aimed at more effective remediation of pollutants.

2. Types of POPs and other pollutants

2.1. Types of POPs

Since the 1970s, the use of POPs has been restricted, and the release of these compounds has been prohibited in Europe and the USA. During the Stockholm Convention in 2001, delegates from 92 nations agreed to ratify the treaty to reduce or eliminate the release of the “dirty dozen” – the original 12 POP substances. These include aldrin, chlordane, dichlorodiphenyl trichloroethane (DDT), dieldrin, endrin, heptachlor, hexachlorobenzene, mirex, toxaphene, polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins, polychlorinated dibenzofurans.³⁷ Although additional contaminants have been identified, the primary focus remains on these original 12, which include 10 intentionally produced chemicals such as aldrin, DDT, and PCBs, as well as two unintentionally produced substances, PCDDs and PCDFs.^38,39 Polycyclic aromatic hydrocarbons (PAHs), formed unintentionally through the combustion and burning of organic compounds, are classified as persistent organic pollutants. Their prevalence is particularly pronounced in densely industrialized regions, with river sediment contamination being a significant concern. Fig. 1 shows different classes and examples of POPs.⁴⁰

Fig. 1 Classification of POPs.⁴⁰ “Reproduced from ref. 40 with permission from [Elsevier], copyright [2022].”

2.1.1. Intentional POPs. Different chemical reactions involving chlorine will generate these compounds as desired products. They belong to the class of organic molecules linked with chlorine atoms, exhibiting high lipophilicity and often high neurotoxicity, and are known as organochlorine compounds (OCs). Examples include chlorinated insecticides like dichlorodiphenyltrichloroethane (DDT) and PCBs. These compounds can be categorized into industrial chemicals and organochlorine pesticides.^38,41,42
2.1.1.1. Organochlorine pesticides (OCPs). One significant class of POPs posing significant risks to plants, animals, and humans is Organochlorine Pesticides (OCPs), characterized by many chlorine atoms. These highly stable chemicals can persist in various environmental compartments for years, leading to ongoing contamination and hazards.⁴³ However, applications of these compounds have been restricted in many countries since 1970, but residues of OCP are still found in agricultural fields,⁴⁴ food products,⁴⁵ water surfaces, and sediments.⁴⁶ Mirex, hexachlorocyclohexanes (HCHs), heptachlor, endosulfan, DDT, endrin, dieldrin, and aldrin are among the various examples of OCPs.
2.1.1.2. Industrial chemicals. Some perfluorooctanoic acid and its derivatives, PCBs, polychlorinated naphthalenes, short-chain chlorinated paraffins, and perfluorooctane sulfonic acid and its derivatives are among the substances in this group. POPs in this category are widely utilized in industrial production processes. For instance, PCBs were chemicals in industrial lubricants and coolants for manufacturing transformers, capacitors, and other electrical goods.⁹ Perfluorooctanoic acid has been used to produce numerous consumer goods known for their resistance to heat, grease, oil, stains, and water. The discharge of these POPs has been identified as environmental contamination in water, soil, and air, as documented in numerous previous studies.^47,48

2.1.2. Unintentional POPs. These chemicals were generated as undesired by-products of combustion or chemical processes involving chlorine compounds. This group includes well-known compounds such as PCBs, PCDDs, and PCDFs. Other designated compounds in this group include HCB, hexachlorobutadiene, pentachlorobenzene, and polychlorinated naphthalene.⁹ These substances are often classified into polycyclic aromatic hydrocarbons (PAHs), dioxin, and furan compounds.³⁸
2.1.2.1. Polycyclic aromatic hydrocarbons (PAHs). PAHs, characterized by two or more fused benzene rings arranged linearly, angularly, or in clusters, are composed solely of carbon and hydrogen atoms, with a stable central molecular structure bound by carbon–carbon bonds.³⁸ PAHs, acknowledged as potent carcinogens or mutagens, are designated priority pollutants by the United States Environmental Protection Agency (USEPA) and the European Commission (EC). They enter aquatic and terrestrial systems through a combination of natural phenomena such as forest fires and volcanic eruptions, as well as anthropogenic sources, including incomplete combustion of fuels, cigarette smoke, vehicular traffic emissions, industrial wastewater releases, oil spillage, and tyre wear and abrasion.^49,50 Table 1 shows the occurrence of various types of PAHs. Fig. 2 represents the structure of some PAHs recognized by European Commission.⁵¹

Table 1 Occurrence of PAHs

PAH	Occurrence	Ref.
Benzo[a]pyrene	It has been reportedly found in the air, surface water, soil, and sediments. It is present in cigarette smoke as well as in food products, especially when smoked and grilled	52
Benzo[a]anthracene	It is more likely to be found mixed with other PAHs in coal tar, bitumen and asphalt. It is primarily found in gasoline and diesel exhaust, tobacco and cigeratte smoke	53
Benzo[b]fluoranthene	Benzo(b)fluoranthene is present as a component of polycyclic aromatic hydrocarbons (PAH) content in the environment usually resulting from the incomplete combustion or pyrolysis of organic matters, especially fossil fuels and tobacco	54
Indeno[1,2,3-cd]pyrene	Typical products of motor vehicle exhausts	55
Benzo[ghi]perylene	Typical products of motor vehicle exhausts, street dust
Phenanthrene	Diesel exhausts, street dust
Chrysene	Diesel exhausts, coal combustion
Naphthalene	Gasoline combustion, vascular land plants or termite activity
Dibenzothiophenes	Abundant in coal emission condensates and industrial coal emissions
Benzo[e]pyrene	Commercial waxes, petrolatum, creosote, coal tar, crude oil from S. Louisiana, Kuwait, and Qatar	56

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Fig. 2 Structure of PAHs recognized by European Commission.⁵¹

2.1.2.2. Dioxins and dibenzofurans. For decades, chlorinated dibenzo-p-dioxins (dioxins) have raised concerns due to their toxic properties, detailed below. Similarly structured, chlorinated dibenzofurans (furans) exhibit comparable chemical properties and harmful consequences, often analyzed alongside dioxins. Recently, attention has turned to a particular subgroup of PCBs known as dioxin-like PCBs (DLPCBs). It's widely acknowledged that human activities and man-made sources significantly outweigh natural processes in contributing to the environmental presence of PCDDs and PCDFs, particularly since the 1930s, coinciding with the widespread production and usage of chlorinated chemicals.^38,57,58 Human-made sources of PCDDs and PCDFs can be categorized into three primary groups: chemical processes, combustion processes, and secondary sources. The primary sources of PCDDs and PCDFs in Hong Kong include municipal garbage incinerators, coal-burning, chemical waste incineration plants, clinical waste incinerators, landfill gas combustion, crematoria, animal carcass incinerators, and cement manufacturing.^59,60

2.2. Types of other pollutants

Water pollutants, like textile dyes, are one of the major water pollutants, possess complex chemical structures, and are often carcinogenic. Their presence in water demands immediate attention, as they pose serious threats to aquatic ecosystems. Dyes are broadly categorized into two types: natural and synthetic.⁶¹ Another common method of classification is based on the ionic nature they exhibit in aqueous solutions, dividing them into anionic, cationic, and non-ionic dyes.

Dyes can also be classified as acidic or basic, depending on the functional groups that influence their colour properties. Acid dyes are anionic in nature and typically contain acid groups such as sulphonate (–SO₃⁻) or carboxylate (–COO⁻) within their molecular structure. In contrast, basic dyes are cationic and are characterized by the presence of quaternary amine groups (–NH₄⁺).⁶² Additionally, based on their chemical structure, dyes encompass a wide range of functional groups, including azo, diazo, xanthene, indigoid, anthraquinone, nitro etc.⁶² Fig. 3 shows the classification of dyes commonly employed in the textile industry.⁶² Table 2 shows structure and types of various dyes.

Fig. 3 Classification of dyes commonly employed in the textile industry.⁶²

Table 2 Structure and types of dyes

Structure	Name of the dye	Types
	Basic yellow 37 (BY37)	Cationic dye
	Red acid 27 (AR27)	Acid dye
	Direct blue 1 (BD1)	Direct dye
	Reactive red 198 (RR 198)	Reactive dye
	Vat green 1	Vat dye
	Blue disperse 7 (BD7)	Disperse dye

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3. Sources of POPs

POPs arise from applying pesticides in the environment, industrial waste, and the byproducts of industrial processes. They are emitted into the atmosphere from various industrial sources, including power stations, heating stations, incinerating plants, household furnaces, transportation, agricultural sprays, and evaporation from water, soil, and landfills. Additionally, unintentional production contributes to POPs through incineration, chemical processes, combustion, bushfires, and waste-containing PCBs. These sources are widespread, arising from activities such as obsolete oil use, equipment maintenance, building demolition, cement manufacturing, animal carcass incineration, coal combustion, landfill leaching, recycling operations, municipal and hazardous waste incineration, sewage sludge treatment, chlor-alkali and aluminum secondary plants, organochlorine pesticide production, and waste storage facilities for pesticides.^63,64 The following section elaborates on the global and regional sources of POPs and illustrates specific case studies from countries like India, Poland, and regions in Africa to highlight the diverse pathways of POP release and contamination.

3.1. Industrial emissions and waste management

Industries like textile dyeing, petrochemicals, pharmaceuticals, iron foundries, glassworks, oil refineries, and thermal power plants release inadequately treated wastewater into rivers and streams, spreading POPs. Additionally, electronic waste, which includes nearly 60 million tons produced globally, contains harmful brominated flame retardants (BFRs) such as bromodiphenyl ethers, tetrabromobisphenol A (TBBPA), hexabromocyclododecane (HBCD), and PCBs.⁶⁵

3.2. Agricultural practices and pesticide use in India

Agricultural runoff is another significant source of POP contamination in-ground, surface, and drinking water, mainly consisting of pesticides and chemical fertilizers.³⁷ In India, the narrow genetic base of high-yielding crop varieties, monocropping, and tropical climate have made crops more vulnerable to pests and diseases. To manage these threats, farmers increasingly relied on pesticides, leading to resistance and overdependence on chemical control. As a result, pesticide use surged from 154 metric tons in 1954 to around 88 000 metric tons by 2000, a 570% increase. Although stricter regulations brought usage down to about 58 634 metric tons by 2015–16, recent data show a rise again to 62 193 metric tons in 2020–21 (Fig. 4), posing serious environmental and health concerns.⁶⁶ Despite laws such as Hudson's Pesticide Bylaw (Canada 1991) and The Insecticides Act (India 1968) aimed at restricting pesticide use, both consumption and pollution have continued to increase.⁶⁷ These pesticides might have contaminated the groundwater system as a non-point source of pollution from agricultural activities. In India, insecticides are the most commonly used pesticides.⁶⁸ West Bengal, an agriculturally dominant state in eastern India with one of the world's highest population densities, about 8% of India's population uses substantial amounts of pesticides for various livelihood activities.⁶⁹ The region's heavy agricultural practices result in extensive use of POPs, which can leach into ground and surface water used for drinking.

Fig. 4 Usage of pesticides in India inbetween 2020–2021.⁶⁶

3.3. Natural and anthropogenic sources

POPs such as dioxins and dibenzofurans can be generated by natural activities such as volcanic activity and vegetation fires, with some volcanoes located in central Africa. Specifically, attention is drawn to sources and applications distinctive to southern and South Africa, encompassing natural fires, accidental incidents, and controlled vegetation burning.^70–72

3.4. Environmental monitoring and case study

Poland lacks comprehensive investigations on dioxin-related environmental contamination, and available public data are limited and relies on faulty estimations. According to the Institute for Environmental Protection, as per information provided to the Central Statistical Office, estimates dioxin and furan emissions in 1998 at 290.353 grams TEQ in Poland. As identified by the Institute, the significant contributors to dioxin emissions in Poland are illustrated in Fig. 5.³⁸

Fig. 5 Poland's primary sources of dioxin emissions.³⁸ “Reproduced from ref. 38 with permission from [Elsevier], copyright [2010].”

The Ministry of Environment of Poland, which manages environmental protection and climate related work, has recently taken important steps to measure the actual amounts of dioxin and furan emissions. For this, 118 geological samples were collected from different parts of Poland, including both urbanized areas and areas likely free from dioxin pollution. Analyzing these samples will help identify the natural background level of dioxins in the environment. In addition, Ludwicki et al. have presented a reference source for tracking these molecules in nursing mothers' breast milk and discussing the presence of other persistent organic pollutants.⁷³ The study included 462 women from Warsaw and several other areas under investigation. Table 3 summarizes the allowable concentrations (pg-TEQ g⁻¹) of PCBs, DDT, and gamma-HCH. Higher levels of persistent organic pollutants were primarily detected in women aged 30 and older and those who had given birth to more than four children.

Table 3 Allowable amounts (pg-TEQ g⁻¹) of PCBs, DDT, and gamma-HCH

Substance	The acceptable amount of fat, pg-TEQ g^−1
PCBs	2.8
DDT	4.2
Gamma-HCH	13.3

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3.5. Other emerging contaminants of concern

Other pollutants responsible for water pollution, such as dyes and pharmaceutical compounds, are released into the environment from various sources, such as textile industries, household dyes, pharma companies, etc. Natural and anthropogenic sources are the two main sources of dyes. In contrast to synthetic colours, which are created by humans using a variety of organic compounds, natural sources include things like minerals, insects, and plants.^74,75 However, because dyes are used in so many different aspects of daily life, they are sometimes released from industrial effluents along with other hazardous organic and inorganic compounds, which has a negative impact on the environment.⁷⁵ Pollutants in the environment can originate from a variety of places, including factories, wastewater treatment plants, and even from households, as shown in Fig. 6.⁷⁶ Additionally, India's pharmaceutical, chemical, paint, and other industries are growing quickly and dumping their waste water into streams either directly or after some treatment. Pharmaceuticals have been found in drinking water sources and wastewater treatment plants. A small amount of drugs in drinking water for a long time could be very bad for people's health and aquatic life.^77,78

Fig. 6 Sources and routes by which dyes enter the environment.⁷⁶

4. Transportation and distribution of pollutants

As POPs degrade slowly, they remain in the environment for an extended period. Their transportation in fresh and marine waters occurs at low environmental concentrations.³⁸ One notable feature of persistent organic pollutants is their semi-volatility, allowing for sufficient atmospheric mobility. This facilitates the entry of considerable amounts into the atmosphere and their transportation over long distances.⁷⁹ This results in the widespread dispersion of POPs worldwide, reaching regions where they have never been utilized, contributing to their extensive global distribution.⁸⁰ The unexpectedly high POPs found in Alaska's Arctic region surprise many, especially considering the bans on some of these pollutants in the United States and Canada for many years. POPs tend to migrate towards colder areas like Alaska, where they settle due to the low temperatures, persisting for extended periods without easy degradation. This persistence allows them to transition from the air and water to the soil, plants, animals, and eventually humans. The revelation of persistent contaminants in the Arctic from distant sources emerged in the late 1970s when pesticides were discovered in polar bear fat tissue, revealing the far-reaching impact of atmospheric POPs on wildlife and human health.⁸¹ During atmospheric transport, POPs undergo various processes such as degradation, deposition onto soils, vegetation, or water bodies, revolatilization, sedimentation, and bioaccumulation. These processes, illustrated in Fig. 7,⁸² play a crucial role in determining the fate of POPs in the global environment. Specifically, the study by Fernández and Grimalt (2003) highlights that the polycyclic aromatic hydrocarbons (PAHs) are predominantly transported through the atmosphere via their association with particulate matter, with wet and dry deposition being the key mechanisms responsible for their atmospheric removal.⁸³ Organochlorine compounds (OCs) infiltrate land and water systems predominantly through atmospheric gas-phase interactions. Their low volatility, coupled with temperature-dependent behavior, leads to selective accumulation in colder, high-altitude zones-where cooler climates act as environmental traps for these persistent pollutants. Together, these mechanisms drive the global distillation effect, enabling the gradual migration of POPs from warmer latitudes to colder climates. This underscores the necessity of understanding these transport mechanisms to evaluate the broader environmental consequences of these pollutants.⁸³

Fig. 7 Environmental processes occur when POPs are transported across vast distances through the atmosphere.⁸²

Given that numerous physicochemical properties influencing the environmental fate of POPs are temperature-dependent, the theory of the Global Distillation Effect predicts the migration of gas phase contaminants from warmer regions of the planet, such as tropical or temperate source areas, to colder, higher latitude regions (as shown in Fig. 8).⁸² This decrease in temperature affects the vapor pressure and Henry's law constant of these compounds, increasing their propensity to condense and accumulate on surfaces like atmospheric particles, soils, vegetation, and aquatic ecosystems, from which they can infiltrate food chains. Depending on their volatility, different POPs will condense at varying ambient temperatures, producing a fractionation of these chemicals within the Earth.^82,84,85 The movement of compounds from low to high latitudes occurs through condensation and volatilization steps, known as the grasshopper effect. These steps involve the exchange of compounds between the air and terrestrial surfaces, driven by seasonal temperature changes at mid-latitudes. Eventually, most POPs would accumulate in polar regions, where the low ambient temperature traps them-an effect referred to as the cold finger or cold trap.⁸⁴

Fig. 8 Diagram illustrating the Global Distillation Effect via the grasshopper mechanism in temperate regions, with P_v indicating the vapor pressure and T_m representing the mean air temperature in each area.⁸²

In the last ten years, several studies have shown global POP concentration patterns that align with this distribution mechanism. For instance, there's been a noticeable increase in hexachlorobenzene (HCB) levels in tree bark as one moves from the equator toward northern regions.^86,87 Latitudinal patterns have also been documented for PCBs, particularly the more volatile congeners, HCHs, and HCB, across various environmental compartments like air,⁸⁸ sediments,^89,90 and soil.^91,92 Recently, PBDE, employed as flame retardants and not listed as POPs, has been discovered in significant amounts in Arctic ringed seals.⁹³

5. Health effects

Hydrophobic POPs possess considerable fat solubility, contributing to their advantageous bioaccumulation and extended persistence, which enhances their toxicity.^94,95 In general, POPs move from one organism to another, accumulating in high concentrations, a process referred to as biomagnification. Recently, there has been growing public concern regarding POP contamination, primarily due to the identification of several compounds as hormone disruptors. These substances can potentially interfere with the normal functioning of the endocrine and reproductive systems in both humans and wildlife. Their presence in the environment poses numerous risks without benefitting the Earth. Once released, these pollutants can persist for decades, leading to various adverse effects such as cancer, congenital disabilities, learning disabilities, and immunological, behavioral, neurological, and reproductive disorders in humans and other animal species.⁹⁶ Evidence suggests that a significant number of people globally may have accumulated sufficient levels of POPs in their body fat, leading to severe health consequences such as illness and death (Fig. 9).⁹⁷ Studies conducted in both laboratory settings and natural environments have linked POPs to a range of health issues, including endocrine disruption, reproductive and immune dysfunction, neurobehavioral disorders, and cancer.^98,99 Recently, particular POPs have been linked to decreased immunity in infants and children, leading to a simultaneous rise in infections, as well as developmental abnormalities, neurobehavioral impairment, and the initiation or promotion of cancer and tumors. Additionally, some researchers are exploring the potential role of certain POPs as significant risk factors in the development of human breast cancer.^38,100 Below are outlined several prominent health issues linked to POPs.

Fig. 9 Various health problems are due to exposure to POPs.⁹⁷

5.1. Endocrine disruption

Disruption of the endocrine system leads to various diseases and health issues. Endocrine disrupting compounds interfere with hormonal regulations, leading to detrimental effect on neurophysiological function, immune system and reproductive health across both human and animal populations. All members of the “dirty dozen” group of POPs have been identified as endocrine disruptors that impair hormone functions. Research suggests that the effects of POPs during fetal development persist throughout an individual's lifespan.¹⁰¹ Most pesticides that disrupt the endocrine system are belongs to the organochlorine class, including DDT, dieldrin, toxaphene, chlordane, mirex, endosulfan, and others. These substances mimic estrogen and cause changes to the reproductive organs.¹⁰² Children displaying heightened startle responses, decreased motor reflexes, neuromuscular functions, and diminished responsiveness showed neuro-behavioral deficits. These included hypoactive reflexes, impaired visual recognition, and short-term memory impairments observed around the age of 7 months. By 4 years of age, issues such as reduced weight progression, decreased responsiveness, and impaired visual recognition were noticeable.^103,104

Langer et al.¹⁰⁵ documented extensive industrial contamination of PCBs in the Michalovce region of East Slovakia from 1955 to 1984. Despite the passage of time, residual contamination persists, with affected residents displaying abnormal thyroid function, elevated urinary iodine levels, and increased serum concentration of 15 distinct PCB congeners-pointing to continued endocrine disruption.

5.2. Reproductive problems

Reproductive issues and diseases encompass dysfunctions in both male and female reproductive systems, leading to conditions such as congenital disabilities, preterm birth, developmental disorders, low birth weight, impotence, reduced fertility, and menstrual disorders.¹ Numerous studies have revealed the detrimental effects of POPs on human health. According to a review by Gao and Wu, different cells in the testes, including testicular germ, Leydig, and Sertoli cells, exhibited varied responses to PCB pollutants.¹⁰⁶ In 2002, Damstra elucidated the effects of POPs on the reproductive system, indicating that exposure to these pollutants led to diminished sperm quality and quantity. Furthermore, they were associated with disruptions in sex ratios and the onset of early puberty.¹⁰¹ Vizcaino et al. conducted a fascinating study exploring the transmission of POPs from mothers to their offspring. Their research revealed that Spanish mothers transferred 14 organochlorine pesticides, 7 polychlorinated biphenyls, and 14 PBDEs to their unborn fetuses. These POP concentrations were notably higher in the mothers' serum than in the samples from the umbilical cord and 50 placentas examined. The authors provided insights into the active transport of POPs across the placenta.¹⁰⁷ Giwercman et al. detailed the impact of PCBs and p,p′-DDE exposures on the alteration of sperm Y : X chromosome ratio in the Faroe Islands. Additionally, they observed a diminished Y : X ratio in men from this region.¹⁰⁸

5.3. Cancer

Cancer is a severe illness where cells grow and spread uncontrollably. It's dangerous because it's hard to predict and treat. Normal cells become cancer cells for many different reasons, both inside and outside the body.¹⁰⁹ Increased levels of POPs within low-density lipoproteins contribute to the developing of several types of cancer.¹¹⁰ Several reviews discuss the cancers induced by POPs.^111,112 The International Agency for Research on Cancer has categorized 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD) as a Group 1 carcinogen, indicating it poses a significant risk to human health.¹¹³ Different experts are also exploring the potential impact of certain POPs as substantial risk factors in the development of human breast cancer.¹⁰⁰ Yu et al. studied the relationship between seafood intake and cancer susceptibility by analyzing global concentrations of various contaminants in marine products, including OCPs, PBDEs, PCDDs, PCDFs, as well as both dioxin-like and non-dioxin-like PCBs. The researchers found that populations in Europe experienced the highest dietary exposure to PCDD/Fs and DL-PCBs, followed by those in North America and Asia.¹¹⁴ In a comprehensive study conducted in Spain in 2003, Arrebola et al. assessed the bioaccumulation of POPs in human adipose tissue from 368 participants. Utilizing Cox-regression analysis, they demonstrated a positive correlation between PCB levels and cancer risk, highlighting the potential oncogenic effects of long-term PCB exposure.^115,116 Meanwhile, Mathur et al. performed an epidemiological study in Jaipur, India, focusing on the etiological role of organochlorine pesticide exposure-namely DDT, its primary metabolite DDE, dieldrin, heptachlor, and various isomers of HCH (α, β, γ)-in the incidence of breast cancer among women. They found that high levels of these POPs were associated with elevated breast cancer risks.¹¹⁷ However, the study did not mention the specific period during which the investigation was conducted.

5.4. Cardiovascular problems

Globally, cardiovascular disorders represent the foremost contributors to death, comprising a broad spectrum of ailments including hypertension, angina pectoris, cardiac arrhythmias, and more.¹¹⁸ POPs contribute to the developing of these health issues. It's widely acknowledged that POPs, being mostly lipophilic, tend to bioaccumulate in high-density lipoproteins, thereby contributing to a range of cardiovascular problems and diseases.¹¹⁰ Lee et al. documented a significant association between cardiovascular dysfunction and the accumulation of POPs in the serum of 90 individuals. Among the culprits, organochlorine pesticides, polychlorinated biphenyls, and polybrominated biphenyl were pinpointed as key contributors to these cardiovascular abnormalities.¹¹⁹ Valera et al. conducted a comprehensive investigation into the association between hypertension and the exposure to PCBs and organochlorine (OC) pesticides among Arctic populations. The study encompassed a cohort of 1614 adults from 9 towns and 13 villages in Greenland, with concentrations of these POPs quantified and analyzed. Utilizing logistic regression models where POPs were treated as continuous variables, the authors identified a significant positive correlation between higher POP concentrations and the prevalence of hypertension.¹²⁰

5.5. Other health problems

Moreover, beyond the above-mentioned health effects and diseases, POP pollution is associated with other health issues such as chloracne, transient hepatotoxicity, porphyria, neuropsychological impairment, and peripheral atherosclerosis.¹²¹ Other sources in the literature also discuss various health issues associated with the contamination of POPs.^122–133 Apart from POPs, many diseases that affect both humans and animals have been connected to textile dyes, which are extremely hazardous and may even cause cancer.¹³⁴ Fig. 10 shows the range of illnesses that textile dyes can induce.¹³⁵ These include dermatitis, central nervous system issues, and more. One possible explanation for these issues is that enzymes become inactive due to the replacement of their cofactors.¹³⁶

Fig. 10 Health risks associated with textile dyes.¹³⁵ “Reproduced from ref. 135 with permission from [Elsevier], copyright [2022].”

6. Photocatalytic degradation of POPs and other pollutants

Photocatalysis involves catalyzing a photoreaction to accelerate its speed.¹³⁷ Photocatalysis operates straightforwardly: when a photon with energy equal to or greater than the bandgap energy strikes, it generates a pair of charge carriers. This energy transformation initiates the formation of electrons and holes.¹³⁸ These electrons and holes, induced by light, interact with O₂, H₂O, and –OH group, producing active ROS like ˙OH and O₂˙⁻ with potent oxidation capabilities.¹³⁹ After fluorine, these hydroxyl radicals hold the second-strongest oxidation potential (2.8 eV).¹⁴⁰ These radicals are highly reactive, quickly interacting with organic molecules at rate constants typically ranging from 10⁶ to 10⁹ mol L⁻¹ s⁻¹.¹⁴¹ These radicals initiate sequential reactions with contaminants on the photocatalyst's surface, leading to the degradation of organic pollutants.¹⁴²

Despite certain drawbacks, such as the inability to function in solar radiation, challenges in reactor design, limitations in catalyst reusability and recovery, lower quantum efficiency, and occasional production of toxic metabolites, it's essential to address these issues in photocatalysis research before practical application. Some catalysts like CdS and PbS may also exhibit toxicity.¹⁴³ Another limitation in photocatalysis arises from the swift recombination of electron–hole pairs, which can impede reaction progress. Doping the catalyst with metal oxide nanomaterials offers a solution by hindering this recombination. This approach extends the lifespan of electron–hole pairs, facilitating redox reactions on the catalyst's surface.¹⁴⁴

6.1. Principle of heterogeneous photocatalysis

The term “heterogeneous” denotes a system where the catalyst is in the solid phase while contaminants exist in the aqueous phase. This setup uses multiple phases alongside a light source (UV/solar radiation) with the catalyst present. Semiconductors such as TiO₂, ZnO, ZnS, etc., are commonly employed as catalysts in these heterogeneous processes. The photocatalyst becomes activated upon exposure to UV radiation, initiating a redox environment within the reaction medium.^145,146 The central aspect of the entire process centers around the production of hydroxyl radicals, which are crucial for converting a range of toxic or non-biodegradable pollutants into harmless end products like CO₂, H₂O, and other environmentally beneficial compounds.¹⁴⁷

When a semiconductor absorbs a photon with energy equal to its bandgap energy, an electron transitions from the valence band to the conduction band, leaving behind a hole (h_vb⁺) in the valence band (VB). This process is known as generating an “electron–hole (e⁻/h⁺) pair.” The holes in the valence band act as potent oxidizing agents (+1.0 to +3.5 V vs. the normal hydrogen electrode, depending on pH and material), while the electrons in the conduction band serve as powerful reducing agents (+0.5 to −1.5 V vs. the normal hydrogen electrode, depending on pH and material).¹⁴⁸ To prevent recombination with the trapped hole, it's crucial for the electron in the conduction band (CB) to be captured by an electron acceptor. On the catalyst's surface, the electron and hole unite for the redox reaction, where the photo-generated electron reacts with oxygen to form superoxide anion radicals. Meanwhile, the hole (h_vb⁺) directly oxidizes the pollutant, generating hydroxyl radicals ˙OH. Additionally, the electron (e_cb⁻) efficiently reduces the oxygen adsorbed by the catalyst.^{147,149–151} To ensure effective photocatalytic degradation, the reduction of oxygen must coincide with the oxidation of contaminants. This prevents the accumulation of electrons in the CB, which would otherwise reduce the rate of electron and hole recombination, thereby impeding the degradation reaction.^152,153

6.2. Mechanism of photocatalysis

The mechanism and theory behind semiconductor-based photocatalysis have been extensively explored in various literature.¹⁵⁴ As per Herrmann,¹⁵⁵ heterogeneous photocatalytic processes consist of five steps, namely:

(1) Surface transfer of reactants, (2) at least one of the reactants being adsorbed, (3) adsorbed phase reaction of the reactants, (4) product's desorption, and (5) product's diffusion from the surface.

In the context of utilizing MX (here M: Ti, Zn etc. and X: O, S etc.) as the photocatalyst, the way that heterogeneous photocatalysis work is outlined as follows:

Photocatalyst (MX) + hν (λ < 380 nm) → MX (e_{conduction band}⁻ + h_{valence band}⁺) (1)

Here, e_{conduction band}⁻ and h_{valence band}⁺ acting as reducing and oxidizing agents, respectively. The oxidation and reduction of POPs are as follows:

6.2.1. Oxidative reaction.

Organic pollutant + h_{valence band}⁺ → oxidation products MX (h_{valence band}⁺) + H₂O → MX + H⁻ + ˙OH (2)

This reaction follows OH⁻ ion neutralization into ˙OH by the h⁺:

MX (h_{valence band}⁺) + OH⁻ → MX + ˙OH (3)

h⁺ + organic pollutant → metabolites → CO₂ + H₂O (4)

6.2.2. Reductive reaction.

Organic pollutant + e_{conduction band}⁻ → reduction products

Here, O₂˙⁻ formed by the electron as follows:

MX (e_{conduction band}⁻) + O₂ →MX + O₂˙⁻ (5)

O₂˙⁻ + H⁺ → HO₂˙ (6)

HO₂˙ + H⁺ + MX(e_{conduction band}⁻) → H₂O₂ + MX (7)

˙OH + organic pollutant → metabolites → CO₂ + H₂O (8)

In the steps outlined above, generating radical ˙OH and superoxide anion radical (O₂˙⁻) are crucial for driving the photocatalytic oxidation reaction.^30,147 Fig. 11 represents the typical process for eliminating POPs through photocatalysis.

Fig. 11 The typical process for eliminating POPs through photocatalysis.

7. Metal oxides as a photocatalyst

To date, oxides of various metals/semiconductor-based nanomaterials, including TiO₂, SnO₂, ZnO, Fe₂O₃, ZrO₂, WO₃, CuO, Cu₂O, CdS, etc., have been researched for their ability to break down photocatalytically organic pollutants into harmless gaseous products.^156,157 This section reviews the latest developments in applying different metal oxide nanoparticles and composites of these oxides for the photocatalytic degradation of organic contaminants.

7.1. Titanium dioxide

TiO₂ was the first photocatalyst to be reported. It is regarded as a promising and viable choice because of its affordable price, extensive occurrence, inertness in both chemical and biological systems, photocatalytic stability, and minimal environmental impact.^158–160 TiO₂ has five distinct forms: anatase, rutile, brookite, monoclinic, and orthorhombic. The photocatalytic performance of these is highest in anatase and then in rutile. TiO₂ has a large band gap energy (3.2 eV for anatase and 3.03 eV for rutile), permitting photons in the near-UV range to be absorbed (λ < 387 nm).¹⁶¹ Under sunlight exposure, electrons in the valence band jump to the conduction band, forming electron–hole pairs. Under alkaline conditions, the holes (h⁺) in the valence band transform hydroxyl groups and water molecules into hydroxyl radicals. Meanwhile, photoinduced electrons in the conduction band reduce dioxygen, which interacts with H₂O to form H₂O₂, which is a strong oxidizer. These potent oxidizers then react with the target pollutants near the photocatalysts, slowly decomposing them into harmless and non-toxic compounds, such as CO₂ and H₂O.¹⁶² An extensive spectrum of organic dyes, including orange G, malachite green, amido black 10B, Safranine O'azure II, rhodamine 6G, alizarin green, Sudan III, indigo carmine, solvent red 23, and others, were degraded upon UV light exposure in the presence of TiO₂ as a photocatalyst.¹⁶³ Bandala et al. found in 2002 that aldrin could be degraded into dieldrin, chlordane, and 12-hydroxy dieldrin by employing TiO₂.¹⁶⁴ According to Bin et al., dicofol can be fully degraded in under 2 hours when exposed to UV radiation and TiO₂.¹⁶⁵ In another experiment, Thomas et al. found that endosulfan could be nearly completely broken down utilizing photocatalytically active solar TiO₂.¹⁶⁶

The anatase and rutile phases of TiO₂, defined by their wide bandgap (3.0–3.2 eV), demonstrate photocatalytic activity when exposed to near-ultraviolet radiation, which accounts for only 4% of the electromagnetic spectrum.¹⁶⁷ Hence, substantial attempts have been made to modify and introduce dopants into nanostructures based on TiO₂ to enable them to function as active photocatalysts under visible light.

7.2. Zinc oxide

Zinc oxide nanomaterials exhibit significant promise as photocatalysts due to their high quantum effectiveness, excellent stability, affordability, and eco-friendly characteristics.^168,169 Zinc oxide, with a band gap of 3.25 eV, is comparable to TiO₂ (3–3.25 eV) and is commonly used to degrade water contaminants such as organic pollutants under light exposure.¹⁷⁰ Zinc oxide has a photocatalytic degradation mechanism closely similar to TiO₂; thus, ZnO nanomaterials are also being explored as a substitute for titania in water detoxification processes.¹⁷¹ In the study of semiconductors, ZnO demonstrated superior photocatalytic performance compared to other materials, attributed to its enhanced ability to absorb a more significant portion of the electromagnetic spectrum. Furthermore, researchers investigated the fluorescence emission properties of ZnO nanoparticles for their potential application in the photocatalytic degradation of various organic hazards.¹⁷² The fluorescence intensity observed in the emission spectra of ZnO is linked to the recombination mechanism of electrons and holes. The photoinduced holes are effectively scavenged by organic compounds on ZnO's surface, which competes with the charge recombination process that produces ZnO fluorescence emission. Consequently, the intensity of fluorescence emission spectra diminishes due to the photo-reaction between ZnO nanomaterials and organic pollutants. Therefore, ZnO nanoparticles demonstrate effective detection and photocatalytic degradation of organic contaminants in aqueous systems.¹⁷² According to Reddy et al., dicofol can be degraded by 88.4% using ZnO nanoparticles in 1.6 h at 450 °C in a slightly alkaline solution.¹⁷³

Furthermore, pure ZnO nanoparticles exhibit absorption only in the UV band of the electromagnetic spectrum, which accounts for approximately 4% of the total electromagnetic spectrum. Therefore, enhancing the response to visible light is crucial, as it encompasses about 43% of the electromagnetic spectrum, significantly larger than the UV range.¹⁷⁴ Moreover, ZnO's wide bandgap allows the photoinduced electrons and holes to recombine easily, limiting the material's photocatalysis capacity.¹⁷⁵ Various methods have been developed to enhance the visible light response and mitigate the speed at which photoinduced electron–hole pairs in ZnO nanoparticles recombine. One well-known method is doping with metal ions and metals, which reduces electron–hole recombination and broadens its response to the visible range of the spectrum. For example, Lu et al. synthesized cobalt-doped ZnO (ZnO : x% Co, where x = 0.5, 1, and 1.5) via a simple hydrothermal method and evaluated its photocatalytic efficiency against alizarin red dye under visible light irradiation.¹⁷⁶ The Co-doped ZnO samples exhibited a broader light absorption range compared to pure ZnO, with the 1% Co-doped sample achieving approximately 93% degradation of the dye within 60 minutes.¹⁷⁶ This enhanced performance was attributed to strong electronic interactions between Co and ZnO, which facilitated improved charge separation and transfer while suppressing the recombination of photogenerated electron–hole pairs, thereby enhancing visible-light-driven photocatalytic activity.¹⁷⁶ Similarly, Ahmad et al. prepared nickel-doped ZnO photocatalysts (Zn_1−xNi_xO, where x = 0.0–0.5) using a wet chemical method and assessed their activity against methylene blue (MB) under visible light.¹⁷⁷ While pure ZnO exhibited relatively low degradation efficiency, the Ni-doped variants showed significantly improved photocatalytic performance. This enhancement was ascribed to the role of Ni in promoting charge carrier separation by trapping electrons, thereby reducing electron–hole recombination, which is a key factor limiting photocatalytic efficiency.¹⁷⁷

7.3. Tin dioxide

Being n-type semiconductor it finds diverse applications across catalysts, gas sensing, transistors, batteries, and transparent electrodes.^178–180 The better photocatalytic propensity of SnO₂ is predicted by the elevated surface-to-volume ratio of the nanoparticles, which increases active sites. Pure SnO₂ is seldom studied because the photocatalytic process leads to the simultaneous formation of Sn²⁺ and Sn⁰. Consequently, the catalyst may be presented as SnO₂, SnO, or Sn. SnO₂, with its large E_g of 3.6 eV, like TiO₂, is also an ultraviolet-sensitive metal oxide semiconductor.¹⁸¹ Although doped SnO₂ is commonly used in the organic contaminant's photodegradation process, there are examples where the undoped material demonstrates effectiveness in the photocatalytic process. However, due to the considerable bandgap energy of pure SnO₂, UV light is necessary for triggering the metal oxide, as visible light lacks the energy required to efficiently generate e⁻–h⁺ pairs. SnO₂ nanocrystals prepared from amino acid-chelated tin salts were used in detoxification studies of rhodamine B. This approach led to a nearly 100% reduction in rhodamine B concentration after 150 minutes of irradiation.^182,183 Comparable results were achieved with nanocrystalline SnO₂ particles synthesized from tin oxalate via a membrane of eggshell in rhodamine B degradation studies. Following one hour of exposure to UV light, the author observed an astounding 95% decrease in dye concentration.¹⁸⁴ Undoped SnO₂ nanoparticles were also effectively utilized in the oxidative breakdown of methylene blue, eosin Y,¹⁸⁵ congo red,¹⁸⁶ and violet 4 dyes.¹⁸²

7.4. Iron oxide

Iron oxide, an n-type semiconductor, is widely accessible and demonstrates photocatalytic activity for degrading organic contaminants.¹⁸⁷ The activation of Fe₂O₃ under visible light is due to its reduced energy band gap (2.2 eV).¹⁸⁸ Iron oxide powder was used to photo-degrade azo dyes, producing substituted benzene and naphthalene, which were then oxidized and mineralized into CO₂ and H₂O.¹⁸⁹ Nanoparticles of Fe₂O₃ with crystal sizes of 35 nm, 100 nm, and 150 nm were used to degrade Congo red. Notably, the 35 nm Fe₂O₃ nanoparticles completely degraded Congo red both with and without exposure to visible light, suggesting no significant impact of light on the catalytic process.¹⁹⁰ This activity in the absence of light may be attributed to a Fenton-like mechanism, where Fe³⁺/Fe²⁺ redox cycling facilitates the generation of ˙OH radicals from H₂O₂, leading to oxidative degradation of the tested pollutant.^191,192 However, the apparent rate constant for Congo red decomposition was higher in light than in dark conditions.¹⁹⁰

However, the application of pure Fe₂O₃ in photocatalysis is limited by several shortcomings, such as low movement of non-dominant charge carriers, slow surface kinetics, short diffusion lengths of photoinduced carriers, and fast recombination dynamics.¹⁹³ Doping various metals or metal oxides into iron oxide nanostructures, alongside the design of iron oxide-based nanocomposites, can effectively control charge carrier recombination events, enhancing overall activity. Among them, transition metal doping has proven particularly effective in boosting the degradation rate of organic pollutants.¹⁹⁴ Nanostructured metal additives serve as light-harvesting centers, enhancing solar energy utilization at the catalyst surface.¹⁹⁵ This not only shifts the optical response by narrowing the band gap but also extends absorption into the visible region.¹⁹⁶ In particular, doping Fe₂O₃ with transition metals improves its efficiency, durability, and recyclability, making it a compelling material for advanced wastewater treatment applications.¹⁹⁷ For example, Cheema et al. synthesized Nb-doped Fe₂O₃ with varying doping concentrations (2, 4, and 6 mmol) using a hydrothermal method.¹⁹⁸ Nb doping led to a significant reduction in particle size (from 93.74 nm to 56.01 nm) and improved stability, as indicated by the zeta potential shifting from −13.04 mV to −37.31 mV.¹⁹⁸ Optical studies revealed a red shift in the band gap energy from 3.08 eV to 2.28 eV, attributed to the introduction of Nb-induced energy levels. Among the samples, 6 mmol Nb-doped Fe₂O₃ exhibited the highest photocatalytic activity, achieving 87% degradation of 10 ppm methylene blue under sunlight within 2 hours. The enhanced performance was ascribed to suppressed electron–hole recombination and increased generation of reactive radicals. Upon light exposure, electrons in Fe₂O₃ are excited from the valence band to the conduction band, forming electron–hole pairs. These charges migrate to the surface and drive redox reactions. Holes (h⁺) directly oxidize methylene blue, while electrons react with oxygen to form superoxide (O₂˙⁻), which converts into hydroxyl radicals (˙OH), enhancing degradation. The proposed mechanism under UV and sunlight is shown in Fig. 12.¹⁹⁸

Fig. 12 Mechanism of photodegradation using Nb-doped Fe₂O₃.¹⁹⁸ “Reproduced from ref. 198 with permission from [Elsevier], copyright [2025].”

7.5. Tungsten oxide

The first documented research regarding the photocatalysis of tungsten oxide (WO₃) dates back to 1976.¹⁹⁹ WO₃ is recognized because of its remarkable light-harvesting capability due to a narrow E_g (2.5–2.8 eV), tunable composition, widespread natural availability, strong h⁺-driven oxidation potential, and excellent environmental durability.²⁰⁰ Among the studied morphologies of WO₃ are nanoparticles, nanoflakes, nanowires, and nanofibers.²⁰¹ The synthesis of tungsten oxide via the solvothermal method, having a hierarchical structure with a hollow pattern, has garnered interest as a reusable adsorbent for the degradation of organic contaminants.²⁰¹

However, due to the rapid recombination of photoinduced charge carriers and a low level of conduction band above the reduction potential of O₂/O²⁻, the photocatalytic efficiency of WO₃ has been constrained, leading to a diminished rate of O₂ molecule reduction.²⁰² Various approaches, including doping, bandgap manipulation, and structural adjustments, have been used for the better photocatalytic performance of WO₃.²⁰³ For instance, Gokila et al. synthesized Y³⁺-doped WO₃ nanoparticles (2%, 4%, and 6%) using a co-precipitation method.²⁰⁴ Pure WO₃ showed densely packed particles with rough surfaces and low porosity. With 2% and 4% Y doping, the particles became more loosely packed, slightly increasing porosity and reactive sites. At 6% Y doping, the particles formed a flower-like structure with petal shapes, greatly increasing surface area and active sites, ideal for photocatalysis. Band gap analysis showed a gradual decrease from 2.82 eV (pure WO₃) to 2.67 eV (6% Y-doped), which improves charge transfer and enhances photocatalytic activity. They further evaluated the photocatalytic efficiency of undoped and Y-doped WO₃ nanoparticles for the degradation of malachite green (MG) dye under visible light. Pure WO₃ nanoparticles degraded 68% of the dye within 180 minutes, while 6 wt% Y-doped WO₃ exhibited the highest efficiency, degrading 90% of MG dye in the same duration. The degradation rate increased with Y doping concentration up to 6 wt%, but declined at higher levels. The enhanced performance is attributed to the generation of electron–hole pairs under visible light, with Y doping introducing intermediate energy levels within the band gap of WO₃.²⁰⁴ These states extend light absorption into the visible region and promote better charge separation, thereby improving the photocatalytic activity. Under visible light irradiation, electrons are excited to the conduction band of WO₃, where Y³⁺ ions serve as electron traps, capturing the electrons and being reduced to Y²⁺. This trapping effect minimizes electron–hole recombination and prolongs the lifetime of charge carriers. The captured electrons then react with adsorbed O₂ molecules to produce superoxide radicals (O₂˙⁻), while the holes remaining in the valence band oxidize water molecules or hydroxide ions to generate hydroxyl radicals (˙OH). These reactive oxygen species (˙OH and O₂˙⁻) play a key role in attacking and decomposing MG dye molecules into less toxic products.²⁰⁴

7.6. Copper oxide

Copper oxide nanoparticles exhibit excellent optical absorption characteristics and an ideal optical bandgap. Their cost-effective production is facilitated by the abundance of precursor materials in the earth's crust.^205,206 Copper nanoparticles are applicable in catalyzing the reduction of CO₂ and its conversion into specific products. Additionally, they have been employed as plasmonic photocatalysts, which enable chemical transformations and the breakdown of toxic organic pollutants into simpler, non-toxic compounds such as water and CO₂.²⁰⁷ It comprises two semiconductor phases: cupric oxide (CuO) having a lower optical band gap (1.6 eV) than the other phase, cuprous oxide (Cu₂O), with a higher optical bandgap of approximately 2.4 eV. Cu₂O, characterized by a 2.4 eV bandgap, is a p-type-based metal oxide semiconductor.²⁰⁸ Because of its optimal bandgap for photocatalysis, visible light is preferred for such activities. Muthukumaran et al. utilized the sonochemical method to synthesize cuprous oxide (Cu₂O) nanoparticles as a photocatalyst for degrading malachite green under visible light.²⁰⁵ Following 45 minutes of exposure, the fabricated nanoparticles achieved a degradation rate of 91.89%.²⁰⁵ Kerour et al. employed a green method utilizing aloevera leaf extracts to synthesize Cu₂O nanoparticles for degrading methylene blue dye.²⁰⁹ Remarkably, they found that the dye completely degraded after 10 minutes of exposure to visible light.²⁰⁹ Despite these benefits, Cu₂O faces several challenges as a photocatalyst due to its narrow bandgap and short minority-carrier diffusion length, which lead to a high recombination rate of photoinduced electrons and holes. Additionally, stability issues caused by self-reduction from photogenerated electrons and self-oxidation from holes result in severe photocorrosion. This photocorrosion is believed to hinder its photocatalytic activity by blocking light from reaching the surface and degrading the structural integrity of the material, thereby limiting its practical applications in photocatalysis.²¹⁰ However, numerous efforts have been made to implement effective strategies for enhancing the photostability and photocatalytic performance of Cu₂O. In general, most research focuses on improving charge transfer from Cu₂O to reactants or co-catalysts to prevent charge accumulation within the particles. Common approaches include morphology control, particle size tuning, optimization of reaction conditions, and the formation of composites with other materials.

On the other hand, CuO nanoparticles have attracted substantial attention in areas such as photocatalysis, supercapacitors, solar cells, electrocatalysis, biodiesel production, and water pollution remediation.^211,212 However, the photocatalytic performance of CuO has been found to be superior to that of Cu₂O.²¹³ Similar to Cu₂O, CuO is also limited by rapid photoinduced recombination and photocorrosion. Therefore, to mitigate these problems, several innovative approaches have been developed by researchers.²¹⁴ Among them, integrating CuO with carbon-based materials has garnered considerable attention due to enhanced electron transport and surface area.²¹⁵ Additionally, coupling CuO with other metal oxide semiconductors has proven effective in boosting its photocatalytic activity. For example, Sakib et al. synthesized a CuO/ZnO composite, which demonstrated a remarkable 98% degradation of methylene blue (MB) dye within 120 minutes under sunlight.²¹⁶ The formation of p–n junction heterostructures not only broadens the light absorption range but also suppresses the rapid recombination of photogenerated electron–hole pairs.²¹⁷ Furthermore, metal doping has emerged as a promising approach to further enhance the photocatalytic performance of CuO nanoparticles.²¹⁸ Despite the improvement of photocatalytic performance of CuO by these methods, fabrication of CuO with carbon-based materials, heterojunction formation, and metal doping in a single material could further enhance light absorption, promote charge separation, and improve overall stability simultaneously.

7.7. Bismuth oxide

Bismuth oxide (Bi₂O₃) can be found in a variety of crystal forms, including triclinic, orthorhombic, face-centered cubic, body-centered cubic, monoclinic, tetragonal, and hexagonal structures.²¹⁹ Bi₂O₃ demonstrates outstanding photocatalytic performance in wastewater treatment, CO₂ reduction, and water splitting when exposed to visible light. However, its photocatalytic efficiency is significantly diminished by the high recombination rate of electron–hole pairs, which impairs its ability to degrade organic pollutants.²²⁰ To enhance the overall photocatalytic efficiency of Bi₂O₃, several approaches have been developed, including modifying its surface structure, morphology, shape, and microstructure, implementing interface engineering, combining other materials, creating oxygen vacancies, and doping with metals and nonmetals.

7.8. Other oxides

Apart from the above-mentioned metal oxides, other oxides such as CeO₂, ZrO₂, MgO, etc., and spinel-type oxides such as Co₃O₄, NiWO_4, etc., also exhibit limited efficiency in pollutant degradation. In light of these constraints, spinal ferrites (MFe₂O₄, where M is a divalent metal ion) have emerged as compelling alternatives. These materials not only exhibit excellent thermal and chemical stability,^221–223 but also features a spinel crystal structure that provides additional catalytic sites, thereby enhancing their photocatalytic performance.²²⁴ Additionally, their narrow band gap (typically ≤2 eV), enables strong absorption in the visible light region, making them promising candidates for visible-light driven photocatalytic applications. It is also noteworthy that ferrites possess smaller band gaps compared to other commonly used visible-light photocatalysts such as WO₃ (2.8 eV) or AgVO₃ (2.2 eV). This narrower band gap enhances their ability to absorb visible light more effectively, offering a potential advantage over photocatalysts with wider band gaps that cover both UV and visible regions but may exhibit lower overall efficiency under visible-light irradiation.²²⁵ The application of ferrites in the photocatalytic degradation of organic pollutants holds great promise for environmental remediation. Few materials can efficiently facilitate both photo-oxidation and photo-reduction processes essential for the complete breakdown of hazardous organic compounds, while also exhibiting strong visible light absorption. To overcome these limitations, composite photocatalysts comprising two or more components have been widely investigated, often resulting in significantly enhanced photocatalytic activity.²²⁶ Another notable advantage of ferrites lies in their inherent magnetic properties. Due to the high magnetism of iron(III) oxides, ferrite-based photocatalysts, either used alone or in composites, can be easily separated from reaction mixtures using an external magnetic field.²²⁷ Moreover, spinel ferrites and their nanocomposites, especially those integrated with carbon-based materials, exhibit excellent photocatalytic activity for contaminant degradation. These systems effectively generate reactive radicals under both visible and UV light, presenting a cost-effective and efficient approach for water purification.²²⁸ Despite their advantages, ferrites with narrow band gaps often suffer from rapid recombination of photogenerated charge carriers, which limits their overall photocatalytic efficiency. However, strategies such as metal doping and the formation of composites with other materials have proven effective in enhancing their performance.²²⁸ For example, Cao et al. synthesized Ag-doped ZnFe₂O₄ via a reflux method and evaluated its photocatalytic activity against Rhodamine B (RhB) dye.²²⁹ They observed that the degradation efficiency of the undoped ZnFe₂O₄ was significantly lower than that of its Ag-doped counterpart. The enhanced activity was attributed to improved charge separation and more efficient charge transfer to the reactant molecules.²²⁹

8. Enhancement of photocatalytic activity

The effectiveness of a photocatalyst's photocatalytic activity heavily relies on its light absorption capability, crystal quality, morphology, surface composition, and electronic attributes. Despite this, its widespread application in environmental micro-scale systems is hindered by persistent barriers such as the high expense of semiconductor materials, inefficient harnessing of solar energy, and rapid recombination of charge carriers.²³⁰ Additional challenges include stability, photocorrosion, and availability.²¹⁰ The primary challenge for photocatalytic performance is linked to the properties of metal oxide, such as insufficient overlap between their band gap with available light of the electromagnetic spectrum, structural degradation, and ineffective charge separation and mobility.²¹⁴ The bandgap of a photocatalyst determines its incident photon absorption efficiency, and most known photocatalysts have a wide bandgap, limiting their activity to the UV region of the electromagnetic spectrum. Therefore, tailoring the bandgap is a crucial challenge to overcome. Furthermore, the recombination of photoinduced electron–hole pairs is a major challenge that affecting the photocatalytic performance of catalyst.²³¹ To address these shortcomings, extensive research has been conducted to enhance solar energy utilization efficiency. Several approaches have recently been proposed to improve adsorption capacity and photocatalytic activity.²³² Fig. 13 represents the different strategies that have been applied to enhance the photocatalytic activity of a semiconductor photocatalyst under visible/solar illumination.

Fig. 13 Schematic representation showing several approaches to enhance the effectiveness of metal-oxide photocatalysts.

8.1. Doping with metals and non-metals

The use of metal-doped photocatalysts to degrade various organic contaminants under visible light irradiation has been thoroughly investigated.^233,234 Metal-oxide particles can be substitutable or interstitially doped with various cations. However, the literature contains numerous inconsistent findings because the doping process (e.g., impregnation, co-precipitation, and sol–gel methods) leads to photocatalysts with varying morphological and crystalline characteristics.^235,236 This method has several drawbacks, including the thermal instability of doped metal oxides, high costs, and the need for expensive ion-implantation equipment.³² Non-metal dopants like C, N, S, B, P, F, and I may be more effective for enhancing photocatalytic activity in the visible range because their impurity states are close to the valence band edge. However, they do not serve as charge carriers and may have a reduced ability to act as recombination centers.²³⁷ Metal-ion-doped TiO₂ is extensively investigated to expand its absorption spectrum into visible light. This is beneficial as visible light is readily available in the solar spectrum. For instance, to boost UV-driven photocatalytic degradation of ATP, Sakaguchi et al. employed the sol–gel technique to incorporate lutetium(III) ions into TiO₂.²³⁸ Their results demonstrated that the doped TiO₂ showed higher reactivity than the undoped counterparts, primarily due to the presence of labile Ti³⁺ spots on the surface, which were formed by doping Lu³⁺ ions into the TiO₂ structure. In other words, the Lu³⁺ ions played a key role in generating different types of Ti³⁺ sites on the powder's surface, which led to an increased recombination time of the electron–hole pairs compared to pure TiO₂.²³⁸ Similarly, Ndabankulu et al. explored TiO₂ doped with different lanthanides (Ce, Dy, Lu, and Sm) via the sol–gel method.²³⁹ They examined the photocatalytic activity and efficacy of these lanthanide-doped TiO₂ samples for degrading caffeine under visible light irradiation.²³⁹ Sood et al. incorporated Fe³⁺ in TiO₂ through a hydrothermal synthesis route and evaluated their photocatalytic potential under visible light using para-nitrophenol as a representative contaminant.²⁴⁰ With just 0.05 mol% Fe³⁺, the system achieved up to 92% degradation in 5 hours. The improved activity was credited to enhanced crystallinity, reduced particle dimensions, better charge carrier dynamics, suppressed electron–hole recombination, and increased sensitivity to visible light introduced by Fe³⁺ incorporation.²⁴⁰ Nonmetal components were also introduced into TiO₂ photocatalysts to enhance the light absorption of metal oxides into the visible range.^32,241 In visible light, nitrogen-doped TiO₂ exhibited a remarkable degradation of lindane into benzene of up to 100%.^242,243 Under visible-light irradiation, N- and C-doped TiO₂ nanomaterials have demonstrated superior photocatalytic activity. The dual doping of two different atoms into metal oxides has attracted significant attention.²⁴⁴ Yang and colleagues found that co-doping of monocrystalline TiO₂ with Eu³⁺ and Fe³⁺ yields a synergistic effect, significantly increasing the photocatalytic degradation of chloroform in solution.²⁴⁵ Research teams have explored nonmetal co- and triple-doped TiO₂ to enhance its photocatalytic activity. For instance, Ramandi et al. utilized the sonochemical method to synthesize anatase phase TiO₂ doped with carbon, nitrogen, and sulfur.²⁴⁶ They then investigated the photocatalytic efficiency of the catalyst against diclofenac as a representative contaminant and sunlight as the light source.²⁴⁶ The mechanism of photocatalytic degradation for the C–N–S triple-doped TiO₂ is depicted in Fig. 14.²⁴⁶ Each nonmetal doping plays a critical role in boosting the photocatalytic performance of TiO₂ by introducing additional impurity states such as N 2p, S 2p, and C 1s. Carbon doping, in particular, acts as a photosensitizer that facilitates the injection of electrons into the CB when exposed to sunlight. In contrast, N and S doping alters the electronic band structure of TiO₂ by mixing the O 2p orbitals of TiO₂ with N 2p and S 2p orbitals, leading to a narrower bandgap for TiO₂. The report suggests that both electrons in the CB and trapped electrons participate in activating O₂ adsorbed on the surface, producing reactive O₂˙⁻. Among the generated species, electrons and ˙OH radicals are identified as the dominant oxidants responsible for degrading diclofenac, whereas photoinduced holes contribute minimally. O₂˙⁻ has a significant impact on diclofenac degradation.²⁴⁶ The ˙OH radicals can be generated in two distinct ways:

O₂ + e⁻ → O₂˙⁻ (9)

O₂˙⁻ + 2H₂O → 2˙OH + 2OH⁻ + O₂ (10)

h⁺ + H₂O → ˙OH (11)

Fig. 14 Diagram showing how C–S–N-tri doped TiO₂ photocatalyst increases photoactivity.²⁴⁶ “Reproduced from ref. 246 with permission from [Elsevier], copyright [2017].”

Eqn (9) demonstrates the reaction between molecular O₂ and electrons (e⁻) in the CB, resulting in the production of O₂˙⁻. Subsequently, in eqn (10), O₂˙⁻ reacts with the molecules of H₂O to generate the species ˙OH radical. Conversely, eqn (11) illustrates that h⁺ and H₂O can react on the photocatalyst surface to produce ˙OH radicals. Table 4 summarizes the degradation performance of organic pollutants using doped TiO₂-based nanomaterials.

Table 4 Overview of TiO₂-based photocatalysts used in the photo breakdown of organic contaminants

Catalysts	Procedure for synthesis	Prototype pollutant	Light	Time (min)	Degradation efficiency (%)	Ref.
TiO2	In situ complexation hydrolysis	MB	Ultraviolet	40	∼100	247
TiO2	Hydrothermal	Ofloxacin	Visible light	60	98	248
Pd-TiO2	Photodeposition	BDE-15	Ultraviolet	20	100	249
Pd-TiO2	Photodeposition	BDE-209	Ultraviolet	<60	100	249
Ag-TiO2	UV-assisted photocatalytic reduction	BDE-47	Ultraviolet	13	100	250
Pt-TiO2	Photocatalytic reduction method	BDE-47	Ultraviolet	40	>90	251
Cu-TiO2	Photocatalytic reduction method	BDE-47	Ultraviolet	30	>90	251
Fe^3+-doped TiO2	Hydrothermal	Para-nitrophenol	Visible light	300	92	240
Se^4+-doped TiO2	Sol–gel	RhB	Visible light	30	91.3	252
Zn-doped TiO2	Sol–gel	MB	Visible light	60	99.64	253
Mo- and W-doped TiO2	Sol–gel	RhB	Visible light	60	96	254
Ce-doped TiO2	Sol–gel	Caffeine	Visible light	120	—	239
Iodine-doped TiO2	Combustion process	Direct blue 199	Visible light	—	—	255
C–S–N-tri doped TiO2	Sonochemical process	Diclofenac	Solar light	180	100	246

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Ji et al. synthesized ZnO doped individually with Fe³⁺, Sn⁴⁺, and codoped with both ions using a low-temperature solution based technique.²⁵⁶ The photocatalytic efficiency of the resulting materials was assessed through the degradation of MB. Incorporating Fe³⁺ and Sn⁴⁺ into ZnO created an impurity level adjacent to the conduction band, enabling electrons in the valence band to transition with smaller energy, as illustrated in Fig. 15.²⁵⁶ As a result, there was an increase in the concentrations of photoinduced electrons and holes, enhancing the photocatalytic performance of doped ZnO nanoparticles. Alam et al. synthesized ZnO nanoparticles doped with metals from the lanthanide series (La, Nd, Sm, and Dy) using the sol–gel technique and tested the photocatalytic degradation of MB and RhB.²⁵⁷ The improved performance of the photocatalyst was attributed to the doping with these heavy metals.²⁵⁷ Nonmetal-doped ZnO photocatalysts have also been reported enhanced photocatalytic activity under visible light irradiation. Mirzaeifard et al. utilized the hydrothermal method for the synthesis of sulfur-doped ZnO. RhB was used to evaluate the photodegradation efficiency of the synthesized photocatalyst under visible light irradiation.²⁵⁸ They determined the optimal catalyst dosage for maximizing the degradation of rhodamine B.²⁵⁸ Out of those, the photocatalyst with 0.10 g and 0.5 wt% of sulfur achieved a 100% degradation rate of RhB (5 ppm) at pH = 5 within 90 minutes. Table 5 summarizes the degradation performance of organic pollutants using doped ZnO-based nanomaterials.

Fig. 15 Diagram showing the introduction of impurities in ZnO photocatalysis.²⁵⁶ “Reproduced from ref. 256 with permission from [John Wiley and Sons], copyright [2019].”

Table 5 Overview of ZnO-based photocatalysts used in the photo breakdown of organic contaminants

Catalysts	Procedure for synthesis	Morphology	Prototype pollutant	Light	Time (min)	Degradation efficiency (%)	Ref.
ZnO	Precipitation	NP	Phenol	UV	120	100	259
Sn0.05ZnO	Low-temperature solution	NP	MB	UV	120	99.61	256
Nd-doped ZnO	Sol–gel	NP	MB	UV	25	98	257
Cu-doped ZnO	Solvothermal	NP	MB	Simulated sunlight	120	81	260
Sulfur-doped ZnO	Hydrothermal	NP	RhB	Visible light	90	100	258

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Similarly, metal doping can reduce the band gap and impede rapid charge recombination in WO₃ photocatalysts. Ramkumar and Rajarajan fabricated WO₃ thin films incorporated with Ni via the chemical bath deposition technique and examined their photocatalytic activity through the degradation of MO, MB, and phenol.²⁶¹ The addition of Ni to WO₃ greatly enhanced its photocatalytic activity. Desseigne et al. fabricated 5 wt% Au-loaded WO₃ via a room temperature impregnation–reduction method.²⁶² The introduction of Au significantly narrowed the band gap of WO₃ to 2.54 eV. The photocatalytic performance of Au/WO₃ was assessed under simulated solar irradiation against RhB, MB, and MO. Complete degradation of RhB was achieved in 90 minutes in the presence of Au/WO₃, while pristine WO₃ showed only 5% degradation. For MO, Au/WO₃ achieved 50% degradation after 180 minutes, whereas WO₃ showed negligible activity. In the case of MB, Au/WO₃ 90% degradation compared to 80% with WO₃ after the same duration.²⁶² Han et al. fabricated S-doped WO₃ nanowires through a hydrothermal synthesis followed by thermal treatment at 300 °C, effectively tuning the electronic structure to reduce the band gap to 2.52 eV.²⁶³ The presence of S not only broadened visible light absorption but also facilitated efficient charge carrier separation. As a result, S-WO₃ nanowires achieved an impressive 97% degradation of MO under 3 h of visible light exposure. This pronounced photocatalytic response was primarily driven by the synergistic effects of S doping on optical and electronic properties.²⁶³ Table 6 summarizes the degradation performance of organic pollutants using doped WO₃-based nanomaterials.

Table 6 Overview of WO₃-based photocatalysts used in the photo breakdown of organic contaminants

Catalysts	Procedure for synthesis	Morphology	Prototype pollutant	Light	Time (min)	Degradation efficiency (%)	Ref.
Au-doped WO3	Impregnation–reduction method	Platelet	RhB	Simulated solar light	180	100	262
Au-doped WO3	Impregnation–reduction method	Platelet	MB	Simulated solar light	180	90	262
Au-doped WO3	Impregnation–reduction method	Platelet	MO	Simulated solar light	180	50	262
S-doped WO3	Hydrothermal	Nanowire	MO	Visible light	180	97	263
Pt-doped WO3	Photochemical impregnation	NP	MO	230 W tungsten-halogen lamp	180	80	264
Pt-doped WO3	Spray deposition method	Pillar structure	MB	Sunlight	360	97	265
Ni-doped WO3	Chemical bath deposition	Thin film	MO	Visible light	150	88	261
Ni-doped WO3	Chemical bath deposition	Thin film	MB	Visible light	150	85	261

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Li et al. produced Fe₂O₃ nanoparticles doped with various metals, starting from fabric filter dust, to degrade methyl orange.¹⁹³ These metal-doped variants exhibited an 82.99% degradation rate under visible light, compared to only 38.94% for undoped Fe₂O₃.¹⁹³ Saikala et al. synthesized Ce-doped CuO photocatalyst using a straightforward precipitation followed by thermal decomposition approach.²⁶⁶ The photocatalytic efficiency of the material was evaluated against azo dyes like MO and Trypan Blue (TB) under UV light irradiation. Remarkably, Ce/CuO exhibited nearly complete degradation of both dyes within 300 minutes. The Ce/CuO photocatalyst exhibited superior dye degradation efficiency under UV illumination compared to both bare and commercial CuO. This enhanced performance was attributed to the presence of Ce atoms, which acted as electron traps in the CB, thereby promoting effective separation of photogenerated e⁻/h⁺ pairs and suppressing recombination.²⁶⁶ Table 7 summarizes the degradation performance of organic pollutants using doped CuO-based nanomaterials.

Table 7 Overview of CuO-based photocatalysts used in the photo breakdown of organic contaminants

Catalysts	Procedure for synthesis	Morphology	Prototype pollutant	Light	Time (min)	Degradation efficiency (%)	Ref.
Ce-doped CuO	Solution combustion method	NPs	RhB	Sunlight	40	95.39	267
Ce-doped CuO	Solution combustion method	NPs	MB	Sunlight	40	87.72	267
Tb-doped CuO	Solution combustion method	NPs	MB	Visible light	120	94	268
Co-doped CuO	One-pot green synthetic route	Spherical dopants on rod-like morphology	MB	UV-vis light irradiation	100	88	269
Mn-doped CuO	Hydrothermal method	Nanosheet	MB	UV-vis light	200	87	270
Cd, Ba-codoped CuO	Co-precipitation	Platelet-like aggregates	CR	Visible light	180	98	271

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8.2. Formation of heterojunctions

A heterojunction is the boundary where semiconductors with different band structures meet, resulting in various possible band alignments.^272,273 There are three fundamental types of semiconductor-based heterojunctions: conventional heterojunction photocatalysts, p–n, and Z-scheme heterojunction photocatalysts.

8.2.1. Conventional heterojunction photocatalysts. Heterojunction photocatalysts of these classes are mainly divided into- type-I, type-II, and type-III, as shown in Fig. 16.²⁷⁴ For a type-I heterojunction photocatalyst (see Fig. 16a), semiconductor A has a higher CB, while VB is lower compared to semiconductor B.²⁷⁵ Hence, under light exposure, electrons and holes gather at the CB and VB levels of semiconductor B, respectively. Because both charge carriers are concentrated in the same semiconductor, the separation of electron–hole pairs is effectively impeded in the type-I heterojunction photocatalyst. Additionally, a redox reaction occurs predominantly on the semiconductor having lower redox potential, drastically reducing the net redox capability of the heterojunction photocatalyst.²⁷⁶ In type-II heterojunction photocatalyst (depicted in Fig. 16b), semiconductor A has higher CB and VB levels than semiconductor B. As a result, photoinduced electrons will migrate to semiconductor B, while photoinduced holes will move to semiconductor A upon exposure to light, thereby achieving spatial separation of electron–hole pairs.^277,278 Similar to type-I heterojunction, the redox capability of the type-II heterojunction photocatalyst will also diminish, as the reduction and oxidation reactions occur on semiconductor B, having lower reduction potential, and semiconductor A, having lower oxidation potential, respectively. The configuration of the type-III heterojunction photocatalyst depicted in Fig. 16c closely resembles that of the type-II heterojunction. However, it features an intensified staggered gap, ensuring no overlap between the bandgaps.^279,280 This hindered electron–hole migration and separation between the two semiconductors in the type-III heterojunction, making it ineffective for enhancing electron–hole pair separation under visible light irradiation. Among the heterojunctions mentioned earlier, it's clear that the type-II heterojunction is the most feasible for increasing photocatalytic performance due to its structure, which facilitates the spatial separation of electron–hole pairs.

Fig. 16 Diagram showing the separation of electron–hole pair in three types of conventional heterojunction photocatalysts: (a) type-I, (b) type-II, and (c) type-III.²⁷⁴

Metal oxide-based heterojunctions have been extensively researched, with TiO₂ being the first and most thoroughly studied photocatalyst. TiO₂ possesses many benefits, but its wide bandgap energy (3.2 eV vs. NHE) limits its absorption of ultraviolet light. It is a primary semiconductor for developing type-II heterojunction photocatalytic systems.²⁸¹ In the early 2000s, Bessekhouad et al. explored Bi₂S₃/TiO₂ and CdS/TiO₂ heterojunctions for the visible light-induced degradation of Orange II.²⁸² Because the wide bandgap of TiO₂ prevents excitation by visible light, the prepared composites of TiO₂ significantly improved light utilization. Specifically, CdS/TiO₂ degrades at a pace of 30 minutes, which is significantly faster than pure CdS and TiO₂ rates alone. Mugunthan et al. reported on the photocatalytic degradation of diclofenac in wastewater in the presence of composite ZnO-WO₃ semiconductor photocatalyst hydrothermally prepared and activated by visible light.²⁸³ This composite expanded the optical response into the visible range and exhibited improved stability, consistently achieving 80% degradation efficiency.

Leghari et al. employed a hydrothermal reaction method to synthesize a WO₃/TiO₂ nanocomposite.²⁸⁴ Their study revealed that organic compounds degraded according to the Langmuir–Hinshelwood isotherm. Notably, the 5.0% composite demonstrated superior photoactivity compared to pure TiO₂, P-25, and pure WO₃ when degrading MO and 2,4-DCP under both UV and visible light conditions.²⁸⁴ Similarly, under visible light exposure, Fe₂O₃-WO₃ nano-composite heterostructure demonstrated a significant increase in the photo-degradation of RhB.²⁸⁵

8.2.2. p–n heterojunctions. While the type-II heterojunction may efficiently separate electron–hole pairs in space, the resulting enhancement in electron–hole separation is inadequate to overcome the semiconductor's rapid electron–hole recombination. Hence, a concept of p–n heterojunction photocatalyst was introduced to expedite electron–hole migration across the heterojunction, thereby boosting photocatalytic performance with the aid of an additional electric field.^286,287 To create a p–n heterojunction photocatalyst, p-type and n-type semiconductors are combined. The n-type semiconductor's electron close to the p–n interface tends to diffuse into the p-type semiconductor before being exposed to light, generating a positively charged species (Fig. 17).^288–290 At the same time, a negatively charged species is produced with the diffusion of holes from the p-type semiconductor close to the p–n interface into the n-type semiconductor. As long as the system doesn't attain equilibrium at the Fermi level, electron–hole pairs will continue to diffuse. This causes the region around the p–n interface to become charged, forming a “charged” space, or what is commonly referred to as the internal electric field.^291,292 When exposed to light having energy equal to or greater than the respective bandgap values, both the p-type and n-type semiconductors become excited, resulting in electron–hole pairs. These photoinduced electrons and holes in the p-type and n-type semiconductors move under the influence of internal electric field (IEF) to the CB of the n-type semiconductor and the VB of the p-type semiconductor, respectively, thereby spatially separating the electron–hole pairs. Importantly, this separation procedure is thermodynamically viable because, in a p–n heterojunction photocatalyst, the p-type semiconductor's CB and VB are usually positioned higher than the n-type semiconductor.^293,294 Because of the combined effect of the internal electric field and band alignment synergy, p–n heterojunction photocatalysts have an accelerated electron–hole separation efficiency compared to type-II photocatalysts.²⁹⁵

Fig. 17 Diagram showing electron–hole separation under light exposure in a p–n heterojunction photocatalyst.²⁸⁸

Chen et al. synthesized p-type NiO and n-type TiO₂ via the incipient wetness impregnation method.²⁹⁶ They assessed their photocatalytic activity by studying the breakdown of MB. The results indicated that the highest activity was observed with 0.5 wt% NiO/TiO₂ compared to TiO₂ alone. This improvement in photocatalytic performance is ascribed to the synergistic effects of the p–n junction and co-catalyst mechanisms.²⁹⁶

Under visible light exposure, the p–n type CuO-TiO₂ semiconductor supported on natural zeolite demonstrated remarkable photocatalytic efficacy in degrading azo dyes. In particular, the semiconductor with zeolite achieved a 90% degradation of MB dye within 60 minutes, exceeding the performance of zeolite alone (79.1%), TiO₂-zeolite (84.1%), and CuO-zeolite (85.2%) under identical conditions.²⁹⁷ According to Nguyen and Doong, the solvothermal preparation of a binary p–n semiconducting ZnFe₂O₄/TiO₂ nanocomposite, which has a ZnFe₂O₄ weight ratio of 1%, makes it a very effective photocatalyst for the breakdown of BPA in an aqueous environment when exposed to 465 nm visible-light.²⁹⁸ Table 8 summarizes the degradation performance of organic pollutants using heterostructured photocatalyst.

Table 8 An overview of the heterostructured photocatalyst-based photocatalytic elimination of pollutants under visible light irradiation

Photocatalysts	Heterojunction type	Synthesis method	Pollutants	Time (min)	Degradation efficiency (%)	Ref.
TiO2/SnO2	Type-II	Electrospinning/hydrothermal	RhB	60	100	299
CuO/TiO2	Type-II	Impregnation technique	BDE-209	7.5	100	300
α-Fe2O3/CdS	Type-II	Hydrothermal/wet chemical	MB	120	86.7	301
ZnTiO3/BiOI	p–n	Precipitation–deposition	Rh 6G	180	>80	302
In2S3/InVO4	Type-II	In situ anion exchange	TC	60	71.4	303
BiVO4/BiOBr	p–n	Hydrothermal	NFX	60	90	304
BiOBr/FeWO4	Type-II	Solvothermal	Doxycycline	60	90.4	305
BiOBr/Fe3O4	p–n	Solvothermal	GP	60	97	306

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8.2.3. Z-scheme heterojunction photocatalysts. Depending on whether a mediator of the electron is present, a Z-scheme photocatalytic system can be classified as direct or indirect. The components of a Z-scheme heterojunction photocatalyst are a reduction semiconductor having a high CB position (SC II), and an oxidation semiconductor with a low VB position (SC I). An electron mediator is occasionally included, resulting in an indirect Z-scheme photocatalyst. Bard et al. introduced the idea of the Z-scheme heterojunction in 1979, inspired by the mechanism of plant photosynthesis.³⁰⁷ The Z-scheme heterojunction was initially utilized in liquid-phase photocatalytic reactions. The redox mediator in the liquid phase acts as an acceptor–donor pair at the interface between the semiconductors without significant direct contact. As depicted in Fig. 18a,³⁰⁸ during light exposure, the redox mediator facilitates the transfer of photoinduced electrons from the CB of SC I to the VB of SC II by functions both as an acceptor and a donor.

Fig. 18 A schematic overview of charge carrier migration in Z-scheme heterojunction photocatalysts when exposed to visible light: (a) conventional, (b) all-solid-state, (c) direct Z-scheme heterojunction, and (d) dual direct Z-scheme.³⁰⁸

Consequently, with photoinduced holes accumulating in the VB of SC I and electrons aggregating in the CB of SC II, oxidation and reduction reactions can occur effectively within this heterojunction system.²⁷⁶ Some commonly used redox electron mediators are Fe³⁺/Fe²⁺ and IO₃⁻/I⁻ pairs.³⁰⁹ Nevertheless, this system faces several challenges. For instance, conventional Z-scheme photocatalysts typically operate in the liquid phase, requiring precise pH control of the reaction environment. Moreover, they often exhibit slow charge transfer rates, which hinder their practical application in contaminant degradation. Additionally, many liquid-phase redox mediators are unstable and prone to deactivation, leading to reduced overall photocatalytic efficiency.³⁰⁹

Considering the limitations of liquid-phase Z-scheme heterojunction photocatalysts, efforts have been directed toward finding alternatives to enhance the effectiveness of indirect Z-scheme heterojunction systems. In 2006, a novel approach involving an electron solid conductor was proposed to facilitate the transport of photoinduced electrons and holes. This innovative Z-scheme photocatalyst utilized noble metal Au as a electron mediator rather than traditional redox couples and integrated both the semiconductors, CdS and TiO₂. This advancement introduced the idea of all-solid-state (ASS) Z-scheme photocatalysts.³¹⁰ In this system, when illuminated with light energy surpassing the bandgaps of the semiconductors, electrons in SC II are excited from the VB to the CB, creating holes in the VB. These electrons are then transported to the VB of SC I via conducting materials at the semiconductor interface (Fig. 18b).³⁰⁸ This mechanism ensures the preservation of redox capability and the effective separation of photogenerated carriers.³¹¹ Numerous studies have identified noble metals as excellent mediators of electrons for constructing ASS Z-scheme photocatalysts. Alternative options, such as graphene, carbon nanotubes, etc., have also been explored.³¹² Table 9 represents the overview of notable ASS Z-scheme photocatalysts.

Table 9 An overview of notable ASS Z-scheme photocatalysts

Catalyst	Mediator	Procedure for synthesis	Application	Ref.
BiVO4/CDs/CdS	CDs	Deposition- precipitation	H2 evolution rate: 1.24 μmol h^−1	313
			O2 evolution rate: 0.61 μmol h^−1
BiVO4/N-CQDs/Ag3PO4	N-CQDs	Solvothermal-precipitation	88.9% TC photodegradation in 30 min	314
AgBr/GO/Bi2WO6	GO	In situ deposition	Elimination of tetracycline hydrochloride with 84% photodegradation efficiency	315
g-C3N4/RGO/Bi2WO6	RGO	Hydrothermal	98% degradation of TCP	316
BiVO4/MoS2/RGO	RGO	Calcination	Water splitting	317
WO3/g-C3N4/Bi2O3	g-C3N4	Cocalcination	80.2% degradation of TC	318
ZnO-Au-SnO2	Au	Precipitation	Water splitting with photoconversion efficiency of 0.55%	319
BiVO4-W-WO3	W	Calcination	PEC performance: high photocurrent density of 5 mA cm^−2	320
Au-CoFe2O4/MoS2	Au	Sol–gel and hydrothermal	∼96% MO degradation	321
Ag/Ag3PO4/WO3	Ag	Deposition–precipitation process followed by photoreduction	98.7% RhB degradation	322
CdS/Ag/Bi2MoO6	Ag	Deposition–precipitation process	RhB degradation	323

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However, ASS Z-scheme semiconductor photocatalysts utilize uncommon and expensive electron mediators, and most result in the shielding effect, which significantly limits their long-term practical application.²⁸¹

Studies on the direct Z-scheme photocatalyst were explored seven years after the ASS Z-scheme heterojunction photocatalyst was proposed.³²⁴ In contrast to traditional liquid-phase and ASS Z-scheme heterostructures, direct Z-scheme photocatalysts feature two closely contacting semiconductors (Fig. 18c),³⁰⁸ removing the requirement for an electron mediator. Notably, they combine the benefits of indirect Z-scheme heterojunctions, such as efficient electron–hole pair separation and optimum redox potential, while effectively reducing backward reactions and light shielding effects without mediators.^325–327 Moreover, eliminating mediators in direct Z-scheme systems significantly lowers construction expenditure, facilitating large-scale utilization. Additionally, developing an internal electric field at the interfaces of the direct Z-scheme system compensates for the role of mediators. This speeds up the recombination of light-induced charge carriers with weaker redox potentials, thereby preserving those with higher redox capabilities. Typically, the staggered band arrangement of two semiconductors results in an IEF only when both semiconductors have different work functions.^328,329 Specifically, free electrons of the semiconductor with a minor work function tend to move toward the other semiconductor upon contact in order to equalize their Fermi levels. Consequently, the electron-donating side becomes positively charged while the electron-accepting side becomes negatively charged, resulting in the development of IEF. Therefore, for the effective establishment of a direct Z-scheme photocatalyst, one semiconductor (SC II) should have higher CB and VB positions, as well as a reduced work function than the other semiconductor (SC I).³³⁰ Table 10 represents the overview of notable direct Z-scheme photocatalysts.

Table 10 An overview of notable direct Z-scheme photocatalysts

Catalyst	Procedure for synthesis	Application	Ref.
Ag3PO4/TiO2	Hydrothermal and ultrasonic dispersion	100% degradation of RhB	331
CM-n-TiO2	Ultrasonic sol–gel	100% degradation of aroclor	332
MoS2/TiO2	Hydrothermal or solvothermal	Photocatalytic degradation, hydrogen evolution, CO2 reduction	333
TiO2/(CdS, CdSe, CdSeS)	Chemical vapour deposition	Degradation of water dyes and antibiotics	334
RGO/FeWO4-g-C3N4	Hydrothermal and mixing	92.3% degradation of RhB	335
g-C3N4/Ag2WO4	In situ precipitation	95% degradation of MO	308
g-C3N4/MnO2	Wet-chemical	91.3% degradation of RhB	336
LaFeO3/SnS2	Hydrothermal process	28.8% degradation of tetracycline (TC)	337
ZnO/CeO2	Wet chemistry process	96% RhB degradation	338
MoS2/Bi2O3	Hydrothermal process	89.6% MBT degradation, 79.3% TC degradation, 90% RhB degradation	339

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Recently, dual/double Z-scheme heterojunction photocatalytic systems involving three semiconductors have garnered significant attention. These systems exhibit subtle variations in band alignments, pathways for charge transfer, and positions of redox reaction sites. In a regular double Z-scheme heterojunction configuration, electrons (or holes) from the intermediate semiconductor (SC III) dissipate as charge carriers migrate from nearby semiconductors. This leaves the peripheral semiconductors (SC I and SC II) with photoinduced charge carriers that participate in redox reactions. At the same time, the intermediate semiconductor also contributes to oxidation or reduction reactions (Fig. 18d).³⁰⁸ In certain scenarios, both electrons and holes on the intermediate semiconductor are depleted, allowing only the photoinduced charge carriers on the peripheral semiconductors to take part in subsequent redox reactions.³⁴⁰

Despite the extensive study on heterojunction photocatalysts, achieving high efficiency remains a persistent challenge. Traditional type I/II and p–n heterojunctions face limitations such as inadequate suppression of photoinduced charge carrier recombination and slowly declining redox capabilities, which hinder their broader application. On the other hand, Z-scheme heterostructured photocatalysts attain efficient charge transfer and retain potent redox capabilities. However, due to the redox ion mediators, conventional Z-scheme photocatalysts suffer from reduced light absorption. In the case of ASS Z-scheme photocatalysts, the metal mediator is scarce and may compete for light absorption with semiconductors. Therefore, the most promising heterojunction systems accessible today are direct Z-scheme heterostructured photocatalysts.

8.3. Support materials for photocatalysis

In photocatalysis, a broad spectrum of both opaque and transparent supporting materials are utilized.³⁴¹ The photocatalyst support needs to be durable, keep the catalyst active, increase its surface area after immobilization, and effectively adsorb pollutants.³⁴²

Glass is an ideal medium commonly used as plates and beads for photocatalytic applications. It can endure high temperatures, offers good optical properties, and is cost-effective, making it a preferred choice for research.^343,344 Recent studies reveal that support materials containing carbon, particularly CNTs and rGO, are widely used. CNTs, with their cylindrical nanostructures, effectively lower the recombination rate of photoinduced electron–hole pairs.³⁴⁵ Like glass, photocatalysts are often immobilized by clays, ceramics, and other silica-containing materials. Due to their large surface areas and chemical stability, they offer a practical and environmentally benign way to separate the catalyst from the reaction mixture and the effluents.³⁴⁶ Other support materials include zeolites and numerous polymers such as polystyrene and polyaniline.^347–350 Table 11 presents several photocatalysts and their associated immobilization supports as documented in the literature.

Table 11 Different substrates are used alongside metal oxides in photocatalytic systems

Catalyst	Substrate	Pollutant	Ref.
TiO2	Quartz tubes	Amoxicillin	351
TiO2	Cellulose	Phenol	352
TiO2	Zeolites	Diclofenac (DCF)	353
WO3	Glass	Benzoic acid	354
TiO2/SnO2	PVC, chitosan	1,2-Dichlorobenzene	355

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9. Applications

9.1. Photodegradation of POPs

Persistent organic pollutants (POPs) are toxic, long-lasting chemicals that resist degradation.^356,357 They include a wide range of substances such as pesticides (OCPs, DDTs, aldrin, etc.), industrial chemicals (PCBs, PBDEs, HBCDs, PFASs, PFOS, SCCPs, HCB, HCHs), aromatic hydrocarbons, and by-products (dioxins, furans). They are mainly chlorinated, brominated, or fluorinated, contributing to their persistence and environmental risk.^27,358,359 To address their harmful effects, a variety of techniques-physical, chemical, biological, or hybrid methods have been proposed for their decomposition and mineralization. Among these, photocatalysis stands out as a promising and essential approach because of its high efficiency, energy-saving properties, mild reaction conditions, and minimal secondary pollution.³⁶⁰

For instance, in a study by Chen et al.,³⁶¹ TiO₂ and its Fe- and Cu-doped variants were synthesized via photodeposition and tested for PFOA degradation under UV light. The doping significantly enhanced photocatalytic performance, with removal efficiencies of 14% for TiO₂, 69% for Fe/TiO₂, and 91% for Cu/TiO₂. This improvement is attributed to a red shift in absorption, indicating band gap narrowing due to metal-induced localized states. Among them, Cu/TiO₂ showed superior activity, as Cu effectively trapped photo-induced charge carriers, suppressing recombination and boosting degradation efficiency. The detailed degradation and defluorination reactions were also proposed in their work.³⁶¹

TiO₂ + hv → TiO₂(e_CB⁻ + h_VB⁺) (12)

TiO₂(h_VB⁺) + H₂O → TiO₂ + ˙OH + H⁺ (13)

C₇F₁₅COOH + ˙OH → ˙C₇F₁₅COOH⁺ + OH⁻ (14)

˙C₇F₁₅COOH⁺ → ˙C₇F₁₅ + CO₂ + H⁺ (15)

˙C₇F₁₅ + H₂O → C₇F₁₅COH + H⁺ (16)

C₇F₁₅COH → C₆F₁₃COF + F⁻ + H⁺ (17)

C₆F₁₃COF + H₂O → C₆F₁₃COOH + F⁻ + H⁺ (18)

Similarly, Xu et al. prepared Pt/In₂O₃ nanorods via photodeposition and studied their photocatalytic performance for PFOA degradation.³⁶² The removal rate increased from 62% to 98% as Pt loading rose to 3%, which was found to be optimal. However, at 5% loading, performance declined due to Pt particles acting as recombination centers and aggregating, which hindered the photocatalytic process. At 3%, Pt facilitated better electron–hole pair separation, enhancing the overall activity. Fig. 19 shows the photocatalytic mechanism, where Pt enhances activity by facilitating charge transfer and reducing electron–hole recombination.³⁶² It also extends the light absorption range, improving light utilization. However, excessive Pt can block active sites on In₂O₃ and act as a recombination center, hindering PFOA degradation.

Fig. 19 Mechanisms highlight the photodegradation of PFOA.³⁶² “Reproduced from ref. 362 with permission from [Elsevier], copyright [2017].”

In 2013, Li et al. synthesized an In₂O₃-graphene composite via sonochemical method and calcination at different temperatures to study its photocatalytic activity for PFOA degradation.³⁶³ As shown in Fig. 20a,³⁶³ In₂O₃-G400 completely decomposed PFOA within 3 hours, while In₂O₃ nanoparticles achieved 87% decomposition. In₂O₃-G350 showed similar results (87.6%), while In₂O₃-G100 was less effective, decomposing only 29% of PFOA. When In₂O₃ was calcined at 400 °C, it was partially enwrapped by graphene, which improved its photocatalytic activity by enhancing charge transfer and increasing its efficiency in breaking down PFOA.³⁶³ The process of PFOA degradation is reported to occur in a stepwise manner, with one CF₂ unit removed at a time.^364,365 As shown in Fig. 20b,³⁶³ the defluorination ratios for In₂O₃-G400, In₂O₃-G350, In₂O₃-G100, and In₂O₃ nanoparticles were 60.9%, 37.7%, 12.9%, and 29.7%, respectively, in 3 h.³⁶³

Fig. 20 (a) Photocatalytic decomposition and (b) defluorination of PFOA.³⁶³ “Reproduced from ref. 363 with permission from [Elsevier], copyright [2013].”

In 2019, wang et al. prepared g-C₃N₄/Fe₃O₄ composites through a two-step process and tested their photocatalytic performance for degrading phenanthrene in soil.³⁶⁶ With Fe₃O₄ loadings of 16.32%, 23.96%, and 53.90%, the composites were effective under visible light. As shown in Fig. 21a,³⁶⁶ up to 92.26% of phenanthrene was degraded within 120 minutes under visible light, while minimal degradation occurred in the dark, confirming that g-C₃N₄/Fe₃O₄ is activated by visible light. To evaluate the influence of Fe₃O₄ content, the photocatalytic activities of Fe₃O₄, pure g-C₃N₄, and g-C₃N₄/Fe₃O₄ with different loadings were compared (Fig. 21b).³⁶⁶ Pure g-C₃N₄ showed weak activity, while loading with Fe₃O₄ enhanced performance, reaching a maximum at 23.96 wt% loading. This improvement is due to better visible light absorption and more efficient charge separation provided by the optimal Fe₃O₄ loading. Fig. 21c presents the DRS spectra of g-C₃N₄/Fe₃O₄ composites with different Fe₃O₄ loadings.³⁶⁶ Pure g-C₃N₄ has a visible light absorption edge at 460 nm, corresponding to a band gap of 2.7 eV. Upon loading with Fe₃O₄, the visible light absorption ability of g-C₃N₄ significantly increases. This enhancement is due to the loading of Fe₃O₄, which broadens the light absorption range, and the absorption intensity increases proportionally with the Fe₃O₄ loading amount. Fig. 21d shows the PL spectra of Fe₃O₄, pure g-C₃N₄, and g-C₃N₄/Fe₃O₄ composites under 325 nm excitation.³⁶⁶ Pure g-C₃N₄ shows a prominent emission peak around 460 nm, while the emission intensity of g-C₃N₄/Fe₃O₄ composites decreases as the Fe₃O₄ loading increases. This reduction indicates that Fe₃O₄ effectively suppresses the recombination of photo-induced electron–hole pairs, attributed to its good electrical conductivity. Combining DRS and PL results, the enhanced photocatalytic activity of g-C₃N₄/Fe₃O₄ is due to improved light absorption and efficient electron–hole pair separation, which is further validated by photocurrent measurements.³⁶⁶ As is well known, higher photocurrent indicates better charge carrier separation and improved photocatalytic activity.³⁶⁷ The photocurrent–time curves based on six cycles of visible light irradiation on-off mode is shown in Fig. 22a.³⁶⁶ As shown in Fig. 22a,³⁶⁶ g-C₃N₄/Fe₃O₄ exhibits a significantly higher photocurrent response than pure g-C₃N₄ under visible light, indicating enhanced charge separation and photocatalytic performance. This enhanced separation promotes the formation of radicals such as ˙OH and O₂˙⁻. To identify the dominant species, ESR measurements were performed using DMPO under visible light. As shown in Fig. 22b,³⁶⁶ no ESR signals were seen in the dark except DMPO oxidation peaks,³⁶⁸ while visible light produced a weak DMPO-˙OH signal (1 : 2 : 2 : 1 quartet) and a strong DMPO-O₂˙⁻ signal, indicating that O₂˙⁻ is the primary reactive species involved in the photocatalytic process, along with some contribution from ˙OH radicals.^368–370 Based on these experimental results, Wang et al. illustrated a possible photocatalytic mechanism for g-C₃N₄/Fe₃O₄ in Fig. 23.³⁶⁶ The enhanced activity is attributed to Fe₃O₄ loading, which extends visible light absorption and suppresses electron–hole recombination. Under visible light, g-C₃N₄ generates electrons and holes (reaction (1)), and the electrons transfer to Fe₃O₄ due to its lower CB and good conductivity, reducing recombination (reaction (2)). The electrons in Fe₃O₄'s CB reduce surface O₂ to O₂˙⁻ (reaction (3)), which can form ˙OH on reacting with H₂O (reaction (4)). Due to the lower CB of g-C₃N₄ (+1.4 eV) than that of ˙OH (+1.99 eV), it is difficult for the holes of g-C₃N₄ to generate ˙OH, explaining the weak ESR signal. Phenanthrene is then oxidized by O₂˙⁻, ˙OH, and h⁺ (reactions (5)–(7)).³⁶⁶

Fig. 21 (a) Phenanthrene photodegradation under visible light and dark conditions using g-C₃N₄/Fe₃O₄; (b) comparison of phenanthrene degradation rates for Fe₃O₄, pure g-C₃N₄, and g-C₃N₄/Fe₃O₄ with different Fe₃O₄ loadings (16.32 wt%, 23.96 wt%, and 53.90 wt%); (c) DRS spectra of bare g-C₃N₄ and g-C₃N₄/Fe₃O₄ composites with varying Fe₃O₄ concentrations; (d) PL spectra of Fe₃O₄ nanoparticles, bare g-C₃N₄, and g-C₃N₄/Fe₃O₄ at different Fe₃O₄ loadings.³⁶⁶ “Reproduced from ref. 366 with permission from [Elsevier], copyright [2019].”

Fig. 22 (a) Photocurrent response comparison between pure g-C₃N₄ and g-C₃N₄/Fe₃O₄ under visible light, indicating improved charge separation in the composite; (b) ESR spectra of DMPO-O₂˙⁻ and DMPO-˙OH spin adducts for g-C₃N₄/Fe₃O₄ under visible light, confirming the generation of reactive oxygen species.³⁶⁶ “Reproduced from ref. 366 with permission from [Elsevier], copyright [2019].”

Fig. 23 Proposed schematic mechanism illustrating the transfer and separation of photoinduced charge carriers in the g-C₃N₄/Fe₃O₄ composite under visible light irradiation, highlighting the generation of reactive species responsible for enhanced photocatalytic activity.³⁶⁶ “Reproduced from ref. 366 with permission from [Elsevier], copyright [2019].”

Recently in 2024, Brindhadevi et al. synthesized NiO nanoparticles via co-precipitation and calcination at 900 °C for 3 h.³⁷¹ The photocatalytic activity was evaluated against PAHs, showing 79% degradation of pyrene (2 μg mL⁻¹ anthracene) within 60 min (Fig. 24a).³⁷¹ Furthermore, maximum anthracene degradation (57%) was achieved at a NiO concentration of 10 μg mL⁻¹ after 5 h of treatment (Fig. 24b).³⁷¹ Fig. 24c shows that anthracene degradation by NiO nanoparticles peaked at 72% under acidic conditions (pH 5) after 300 min.³⁷¹ Higher pH levels reduced degradation due to increased repulsive interactions between NiO nanoparticles and PAH molecules. They further investigated the influence of light sources on anthracene degradation (Fig. 24d).³⁷¹ With 2 μg mL⁻¹ PAH and 10 μg mL⁻¹ NiO nanoparticles over 300 min, they achieved 70% degradation under UV light and 64% under sunlight. The improved performance under UV was attributed to efficient generation of reactive radicals and electron–hole pairs, which enhances the photocatalytic activity of NiO.³⁷¹ Table 12 shows the photodegradation of some POPs using various metal oxide photocatalyst and their composites:

Fig. 24 Evaluation of the photocatalytic efficiency of NiO nanoparticles under different parameters: (a) variation with PAH concentration, (b) effect of NiO nanoparticle dosage, (c) influence of pH on degradation performance, and (d) comparison of photocatalytic activity under UV and sunlight exposure.³⁷¹ “Reproduced from ref. 371 with permission from [Elsevier], copyright [2024].”

Table 12 Photodegradation of POPs using metal oxides and their composites

Photocatalyst	Eg/eV	Light	Target pollutant	Decomposition	Ref.
TiO2	3.14	16 W, 254 nm, UV	PFOA	86.0%	364
			PFDA	99.0%
			PFNA	99.0%
TiO2	3.14	185 nm, VUV	PFOA	98.3%	372
TiO2	3.2	UVA	PAH	93%	373
		UVC		92%
Pt/TiO2	2.92	125 W, 365 nm UV	PFOA	100%	374
Pd/TiO2				94.2%
Ag/TiO2				57.7%
Fe/TiO2	3.32	400 W, 254 nm UV	PFOA	69.0%	375
Cu/TiO2	3.29			91.0%
TiO2-rGO	2.94	150 W, UV	PFOA	86.0%	376
Sb2O3/TiO2	3.12	4 W, UV	PFOA	81.7%	377
β-Ga2O3	4.80	15 W, 254 nm UV	PFOA	36.0%	378
Needle-like β-Ga2O3	4.68	14 W, 254 nm UV	PFOA	100%	379
In-Ga2O3	4.16	200 W, UV	PFOA	100%	380
In2O3 microspheres	2.68	15 W, 254 nm UV	PFOA	100%	381
In2O3 nanoplates	2.76			100%
In2O3 nanocubes	2.72			100%
Pt-In2O3	2.81	500 W, UV	PFOA	98.0%	362
CeO2/In2O3	2.89/2.80	500 W, 254 nm UV	PFOA	100%	382
BiPO4/BiOCl	4.20/3.38	254 nm UV	PFOA	100%	383
rGO/BiOCl	3.02	500 W, UV	PFOA	90.1%	384
ZnO	3.20	28 W, 254 nm UV	PFOA	70.5%	385
ZnO	3.37	Natural sunlight	PAH	98%	386
ZnO/rGO	3.11	Solar irradiation	PFOA	90.9%	387
ZnHCF@ZnO	2.2	Sunlight	Benz[a]anthracene (BaA)	93%	388
			Benzo[a]pyrene (BaP)	90%
CrFe2O4	1.9	LED	Phenanthrene	99%	389
			Anthracene	90%
			Naphthalene	86%

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9.2. Photodegradation of dyes

The textile industry, especially in countries like India, Pakistan, Sri Lanka, and China, generates significant wastewater during processes like dyeing, printing, and finishing. These stages consume large amounts of water (120–280 L) and synthetic dyes made from coal and petroleum. About 11–15% of the dyes are washed out as effluent, leading to increased COD and BOD in water bodies.^390,391 Reactive dyes released into water can cause severe harm to both humans and aquatic organisms, including cancer risks and reproductive issues.³⁹² To combat this, photocatalytic degradation, particularly using metal oxide nanoparticles, offers a sustainable solution. These nanoparticles stand out for their high surface area, strong catalytic efficiency, and ability to produce reactive species that effectively break down a wide range of dyes in water.³⁹³ Their excellent catalytic and adsorption properties, along with the ability to produce reactive species under certain conditions, offer an effective and versatile method for degrading various industrial dyes.³⁹⁴ For instance, Torane et al. synthesized TiO₂ using a simple sol–gel method and tested its photocatalytic ability to degrade MB under sunlight.³⁹⁵ The degradation results, shown in Fig. 25,³⁹⁵ revealed a gradual reduction in MB concentration over time. In the absence of TiO₂, there was minimal degradation even after 360 minutes. However, when TiO₂ was introduced, the degradation rate was initially slow during the first 60 minutes, likely due to the time required for the catalyst to activate. After this period, the degradation rate increased significantly, and after 360 minutes, around 92% of the MB dye was degraded. This illustrates that TiO₂ effectively accelerates the degradation of MB.³⁹⁵

Fig. 25 Percentage degradation of MB dye over time with TiO₂ material.³⁹⁵ “Reproduced from ref. 395 with permission from [Elsevier], copyright [2021].”

Recently, Faryad et al. prepared g-C₃N₄/TiO₂ composites through a straightforward hydrothermal method.³⁹⁶ To improve their photocatalytic performance, boron nanoparticles were synthesized using a green process and incorporated into the composites via a two-stage wet-chemical co-precipitation method. The photocatalytic performance of the resulting composites was then evaluated against MB dye. They observed that the addition of a co-catalyst reduced the bandgap (E_g) of the host material from 2.7783 eV to 1.5678 eV, which promoted the generation of more photo-induced charge carriers in the B-doped g-C₃N₄/TiO₂ composites under visible light, as shown in Fig. 26a.³⁹⁶ The incorporation of boron into the composite significantly enhanced its photocatalytic activity compared to the pure material, leading to a higher degradation rate of MB. This improvement was further evident as the proportion of the co-catalyst increased, as shown in Fig. 26b.³⁹⁶

Fig. 26 (a) Bandgap analysis of g-C₃N₄, CNT-2%, CNT-4%, CNT-6%, CNT-8%, and BCNT-x composites; (b) evaluation of the photocatalytic degradation efficiency of g-C₃N₄/TiO₂-X (X = 2%, 4%, 6%) compared to B-doped g-C₃N₄/TiO₂-2%.³⁹⁶ “Reproduced from ref. 396 with permission from [Elsevier], copyright [2023].”

In 2022, Selvaraj et al. synthesized ZnO doped with Sm via a co-precipitation method followed by calcination at 500 °C for 2 hours and assessed its photocatalytic efficiency against methylene blue (MB) under sunlight.³⁹⁷ The photocatalytic activity of both pure ZnO and Sm-ZnO nanoflowers is presented in Fig. 27.³⁹⁷ As shown in Fig. 27a,³⁹⁷ MB degradation was monitored at different time intervals under sunlight exposure. In the absence of any catalyst, only 10% of MB degraded, indicating its high stability. However, introducing ZnO catalysts markedly enhanced degradation, with pristine ZnO achieving 57% efficiency, as depicted in Fig. 27c.³⁹⁷ As the dopant concentration increased, the degradation efficiency improved. This was due to the presence of Sm³⁺ ions, which functioned as electron traps on the surface of ZnO nanoflowers. The captured electrons reacted with oxygen species, facilitating the degradation of organic pollutants. The highest degradation efficiency, around 96%, was achieved with 5% Sm-doped ZnO after 90 minutes of sunlight exposure. Fig. 27d displays the optical absorption spectra of the MB solution with 25 mg of the 5% Sm-ZnO catalyst under solar light exposure.³⁹⁷ As the irradiation time increased, a decrease in the absorption peak intensity indicated the breakdown of MB molecules by the catalyst. The addition of Sm³⁺ ions into the ZnO structure shifted the conduction band maximum, created new energy states, and narrowed the bandgap due to their larger ionic radius. UV-Vis analysis confirmed the reduction in the bandgap with Sm³⁺ incorporation. The calculated values of ECB and EVB for 5% Sm-doped ZnO were −0.25 eV and 2.84 eV, respectively. The bandgap energy was reduced to around 3.09 eV, and the Sm³⁺ ions acted as effective electron scavenger. The trapping of electrons slowed down the recombination of electron–hole pairs, thereby significantly enhanced the photocatalytic degradation of MB. To pinpoint the primary reactive species responsible for this degradation, a series of scavenger experiments were conducted using 5% Sm-ZnO nanoparticles under optimized conditions. Different scavengers were employed to target specific active species: benzoquinone (BQ) for superoxide ions (O₂⁻), isopropanol (IPA) for hydroxyl radicals (˙OH), and EDTA for photogenerated holes (h⁺), helping to clarify their individual roles in the reaction mechanism. Fig. 28a presents the photocatalytic performance of 5% Sm-ZnO nanoflowers for MB degradation in the presence of various scavenger agents.³⁹⁷ Without any scavengers, the catalyst achieved approximately 96% degradation efficiency. However, when benzoquinone (BQ) was introduced to quench superoxide radicals, the efficiency dropped to about 72%. The use of isopropanol (IPA), which targets hydroxyl radicals, resulted in a significant decrease to 23%, indicating their major role. Similarly, adding EDTA to suppress photogenerated holes reduced the efficiency to around 30%, confirming the involvement of multiple reactive species in the degradation pathway. The results clearly indicate that hydroxyl radicals (˙OH) and photogenerated holes (h⁺) are the primary reactive species involved in the degradation process. The lowest degradation efficiency was observed with isopropanol (IPA), highlighting the crucial role of ˙OH radicals, while the presence of EDTA showed a slightly higher degradation rate, suggesting that h⁺ also plays a significant role. Although benzoquinone (BQ) suppressed the activity of superoxide radicals (O₂⁻), its impact was less severe compared to IPA, confirming that ˙OH and h⁺ are the dominant species responsible for the effective photocatalytic breakdown of MB dye. They further carried out TOC analysis to further verify the mineralization capability of the 5% Sm-ZnO photocatalyst. As illustrated in Fig. 28b,³⁹⁷ the carbon content was reduced by 87% from the initial concentration of methylene blue, with only 7.9% of the original carbon remaining after 90 minutes. This result indicates that the photocatalyst was able to effectively mineralize the methylene blue dye, converting it into environmentally benign products like CO₂ and H₂O, rather than simply decolorizing the solution.

Fig. 27 (a) Variation in MB dye concentration with time, (b) kinetic plot of ln(C/C₀), (c) efficiency of MB degradation by the catalyst, and (d) UV-vis absorption spectra of MB after treatment with 5% Sm-ZnO.³⁹⁷ “Reproduced from ref. 397 with permission from [Elsevier], copyright [2022].”

Fig. 28 (a) Reactive species trapping analysis and (b) TOC measurement results for 5% Sm-doped ZnO.³⁹⁷ “Reproduced from ref. 397 with permission from [Elsevier], copyright [2022].”

In 2021, a study was conducted where Bi₂O₃ was doped with varying amounts of Ag (5%, 10%, and 15%) to examine its photocatalytic activity against crystal violet under visible light.³⁹⁸ The photocatalytic performance of both pure Bi₂O₃ and Ag-doped Bi₂O₃ was compared by performing the photodegradation of crystal violet dye under visible light irradiation at 40 °C. The experiments were carried out using 50 mL of a 100 mg L⁻¹ crystal violet solution and 0.05 g of each catalyst, with the degradation process monitored under visible light exposure. The data shown in Fig. 29 reveals the change in the concentration of crystal violet dye over time and the effect of varying Ag content.³⁹⁸ Without light irradiation, there was little to no alteration in the dye concentration. However, the incorporation of Ag into Bi₂O₃ significantly enhanced its photocatalytic ability to degrade crystal violet. The degradation efficiency of a 100 mg L⁻¹ crystal violet solution increased from 52% to 92%, 76%, and 57% with the addition of 5%, 10%, and 15% Ag to Bi₂O₃, respectively. The Ag-doped Bi₂O₃ samples displayed improved photocatalytic performance compared to pure Bi₂O₃, with the 5% Ag-Bi₂O₃ demonstrating the most rapid degradation of crystal violet over 120 minutes. As the Ag content increased, it covered the active sites of Bi₂O₃, limiting the interaction between the catalyst and dye molecules, which resulted in lower photocatalytic activity. Moreover, excessive Ag loading introduced oxygen vacancies and defects in Bi₂O₃, which facilitated electron–hole recombination, further decreasing the overall catalytic efficiency.³⁹⁸

Fig. 29 Photocatalytic degradation of crystal violet dye (50 mL, 100 mg L⁻¹) at 40 °C with 0.05 g of catalyst: (a) Bi₂O₃, (b) 5% Ag-Bi₂O₃, (c) 10% Ag-Bi₂O₃, and (d) 15% Ag-Bi₂O₃.³⁹⁸

Sathisha et al. fabricated CuO-NiO nanocomposite via facile combustion synthesis and investigated its photocatalytic efficiency against MB under UV light irradiation.³⁹⁹ Fig. 30 summarizes the optimized parameters influencing degradation performance.³⁹⁹ As illustrated in Fig. 30a,³⁹⁹ the highest degradation of MB was achieved at a dye concentration of 5 ppm. Fig. 30b reveals that a catalyst dosage of 15 mg yielded maximum photocatalytic activity.³⁹⁹ The impact of pH, shown in Fig. 30c,³⁹⁹ was explored across a range from pH 2 to 12. Notably, acidic conditions resulted in lower degradation rates, while an alkaline environment significantly enhanced the photocatalytic performance. Fig. 30d illustrates the impact of various scavengers on the photocatalytic degradation process.³⁹⁹ The addition of tert-butyl alcohol (TBA) significantly suppressed the degradation efficiency, indicating that ˙OH play a dominant role as the primary reactive species. In contrast, the presence of EDTA showed minimal influence on the degradation rate, suggesting minor contribution of the photogenerated h⁺ to the overall photocatalytic mechanism.³⁹⁹ Hossain et al. fabricated CuO/CdS composite using an ultrasound-assisted wet impregnation technique and assessed its photocatalytic activity under visible light irradiation against RhB, MB, and MO dyes.⁴⁰⁰ The composite demonstrated impressive degradation efficiencies of approximately 93% for RhB, 83% for MB, and 80% for MO, highlighting its broad spectrum photocatalytic capability.⁴⁰⁰ Table 13 shows the photodegradation of dyes using various metal oxide based photocatalysts along with their composites.

Fig. 30 Influence of (a) dye concentration, (b) catalyst dosage, (c) pH of the solution, and (d) role of various scavengers during MB degradation using CuO-NiO nanocomposite.³⁹⁹ “Reproduced from ref. 399 with permission from [Elsevier], copyright [2023].”

Table 13 Photodegradation of dyes using various metal oxide based photocatalysts along with their composites

Photocatalyst	Eg/eV	Light	Target pollutant	Degradation	Ref.
Zn/TiO2	3.0–3.2	Direct sunlight	MB	99%	401
Ni/TiO2	2.86	UV	MO	51%	402
Fe/TiO2	1.45	Visible light	MB	99.50%	403
CuO	1.66	Solar light	MB	75%	404
Cu2O/PA/rGO	2.14	UV lamp	MO	95%	405
CuO	1.29	UV lamp	MB	95.63%	406
			MO
ZnO/CN	2.1	Visible light	MB	98%	407
ZnO/GO	3.25	UV lamp	Basic fuchsin	95%	408
LaFeO3/rGO	1.86	300 W Xenon arc lamp	Methylene blue (MB), rhodamine B dye (RhB)	MB-98%, RhB-73%	409
Fe2O3/rGO	2.2	300 W Dy lamp with a 420 nm cutoff filter	Rhodamine B (RhB)	99%	410
BiFeO3/rGO	1.8	Mercury Xenon lamp	RhB	87%	411
MgO	2.9	UV lamp	Methylene blue	75%	412
GO-MgO	4.8	Solar simulator	Methyl blue	45%	413
Co3O4	1.29 and 3.3	Mercury lamp	Methylene blue	93.80%	414
Cr : Co3O4	2.4	Fluorescent lamp (9 W)	MB	99%	415
Co3O4/RP composite	1.77	Xe lamp (300 W) with a cutoff filter (λ > 400 nm)	Malachite green (MG)	94.50%	416
CeO2	3.34	UV light	MO	70%	417
			MB	77%
SnO2	3.32	UV light	MO	78%	417
			MB	83%
WO3	2.8	Visible light	MB	75%	418

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9.3. Photodegradation of pharmaceutical pollutants

It is estimated that roughly four-fifths of wastewater worldwide is released into the environment untreated, presenting significant threats to human well-being, aquatic ecosystems, and environmental stability.⁴¹⁹ Industrial discharges, particularly those containing synthetic dyes and pesticides, are major sources of this pollution. Additionally, the frequent use of pharmaceutical products has introduced new classes of pollutants into water bodies, which are increasingly recognized for their detrimental effects on both ecological systems and human health.⁴²⁰ Contamination from these pharmaceutical compounds in water sources has been associated with numerous health risks, such as cancer, hemorrhaging, organ impairment, hormonal disruptions, and toxic effects of varying severity in the human population worldwide.⁴²¹ With the growing concern of antibiotic resistance emerging in aquatic ecosystems, there is an urgent need to establish affordable, sustainable, and efficient treatment methods to eliminate these persistent contaminants from wastewater. Tailoring the surface and interfacial properties of photocatalyst significantly boosts their functionality by facilitating improved charge separation and active site availability. Consequently, heterogeneous photocatalysis has gained momentum as a practical, affordable, and highly efficient technique for eliminating pharmaceutical residues from contaminated water sources.^422,423 In this regard, Sood et al. fabricated Bi₂O₃/TiO₂ by the hydrothermal method and employed it for the photocatalytic breakdown of the antibiotic ofloxacin under sunlight at neutral pH.⁴²⁴ The characteristic absorbance of ofloxacin at 287 nm gradually disappeared during the reaction with 10% BT catalyst, achieving nearly 92% degradation within 120 minutes (Fig. 31a).⁴²⁴ Minimal change was observed without the catalyst, while limited degradation occurred in the absence of light, highlighting the synergistic role of the photocatalyst and solar light in the degradation process. As illustrated in Fig. 31b,⁴²⁴ photolysis alone showed negligible effectiveness in degrading the drug molecule, while adsorption led to only partial removal. To evaluate the influence of Bi₂O₃ content on photocatalytic efficiency, composites with varying Bi/Ti molar ratios-1%, 5%, and 10%-were tested alongside pure TiO₂ and Bi₂O₃ under identical conditions. The results demonstrated that increasing the Bi₂O₃ content enhanced the photocatalytic performance, with 10% Bi₂O₃/TiO₂ exhibiting the highest degradation rate for ofloxacin, as reflected in Fig. 32a.⁴²⁴ The degradation rates for the solar light-induced degradation of ofloxacin solution (25 mg L⁻¹, pH 7, 0.5 g L⁻¹ catalyst dose) were 32.5%, 75.0%, 50.9%, 60%, and 92.4% for bare TiO₂, Bi₂O₃, 1% BT, 5% BT, and 10% BT catalysts, respectively. The enhanced catalytic activity of the 10% BT composite can be attributed to its increased visible light absorption, smaller particle size, and larger surface area. Comparative studies with commercially available TiO₂ derivatives such as P25, PC 50, and PC 500 also showed that the 10% BT catalyst outperformed these products, demonstrating superior degradation efficiency, as shown in Fig. 32b.⁴²⁴

Fig. 31 (a) Absorbance spectra of ofloxacin (25 mg L⁻¹, pH 7) as a function of time using 0.5 g L⁻¹ catalyst dose; (b) comparative analysis of ofloxacin removal via photocatalysis, photolysis (solar light only), and adsorption (in dark conditions).⁴²⁴ “Reproduced from ref. 424 with permission from [Elsevier], copyright [2016].”

Fig. 32 (a) Assessment of photocatalytic degradation of ofloxacin using synthesized photocatalysts including TiO₂, Bi₂O₃, and varying Bi₂O₃-loaded TiO₂ composites (x% BT) under solar light (25 mg L⁻¹, pH 7, 0.5 g L⁻¹ catalyst, 120 min reaction time); (b) comparative evaluation of 10% Bi₂O₃-TiO₂ composite with commercially available TiO₂ photocatalysts under the same experimental setup.⁴²⁴ “Reproduced from ref. 424 with permission from [Elsevier], copyright [2016].”

Similarly, Rostami et al. synthesized ZnFe₂O₄-x wt% graphene (x = 1–4) nanohybrids via the sol–gel method and investigated their photocatalytic activity for the degradation of paracetamol under visible light irradiation.⁴²⁵ As illustrated in Fig. 33a,⁴²⁵ the nanocomposites exhibit significantly lower photoluminescence (PL) intensity compared to pure ZnFe₂O₄, suggesting a suppressed recombination rate of photogenerated electron–hole pairs. Fig. 33b further demonstrates that pure ZnFe₂O₄ shows negligible photocatalytic activity under visible light, while the incorporation of graphene markedly enhances the degradation efficiency.⁴²⁵ This enhancement is primarily attributed to the improved charge separation and transfer facilitated by the graphene sheets, which serve as electron conductors. Overall, the ZnFe₂O₄-graphene nanohybrids exhibit superior photocatalytic performance for paracetamol degradation under visible light, highlighting the synergistic effect of graphene in boosting the activity of zinc ferrite. Among the synthesized composites, the ZnFe₂O₄-4 wt% graphene nanohybrids exhibited the highest photocatalytic performance, achieving approximately 97.4% degradation of paracetamol within 180 minutes. The proposed photocatalytic mechanism is illustrated in Fig. 33c.⁴²⁵ Under visible light irradiation, electrons in ZnFe₂O₄ are excited from the valence band (h⁺) to the conduction band (e⁻). These excited electrons are rapidly transferred to the graphene nanosheets, where they are captured by adsorbed oxygen molecules to generate reactive superoxide radicals (O₂˙⁻). Meanwhile, the photogenerated holes contribute to the formation of hydroxyl radicals. These reactive species, including holes, superoxide anion radicals, and hydroxyl radicals, subsequently oxidize paracetamol molecules adsorbed on the ZnFe₂O₄-graphene surface via π–π interactions and/or electrostatic attraction. Thus, graphene plays multiple crucial roles; acting as an electron acceptor, promoting charge separation, enhancing adsorption, and serving as a photosensitizer, collectively contributing to the enhanced photocatalytic degradation of paracetamol.⁴²⁵

Fig. 33 (a) PL spectra of the pure and hybrid material, (b) photodegradation of paracetamol under visible light, and (c) mechanism of photodegradation of paracetamol using the composite.⁴²⁵

Recently, Hunge et al. synthesized MoS₂/ZnO heterostructures through a hydrothermal process and assessed their photocatalytic behavior against the antibiotic ciprofloxacin.⁴²⁶ As illustrated in Fig. 34a,⁴²⁶ pristine ZnO possessed a wide band gap of 3.17 eV, whereas MoS₂ displayed a much narrower gap of 1.51 eV. Upon increasing the MoS₂ proportion in the composite, the band gap gradually decreased from 3.06 eV in MZ-10 to 2.81 eV in MZ-50. This shift towards lower energy is linked to a rise in carrier density in the conduction and valence bands, along with the development of trap states in the conduction band, which collectively lead to a downward shift in the conduction band edge and a narrower band gap. Fig. 34 presents the photoluminescence (PL) spectra of MoS₂/ZnO composites, highlighting emission features in both the near-ultraviolet and visible regions.⁴²⁶ A gradual reduction in PL intensity was observed with increasing MoS₂ content, reaching a minimum at 30% loading (MZ-30). A distinct peak at 381 nm in the UV region corresponds to exciton–exciton recombination. The observed PL intensity followed the sequence: MZ-10 > MZ-20 > MZ-50 > MZ-40 > MZ-30. The significantly suppressed PL signal in the MZ-30 sample indicates minimized electron–hole recombination, thereby reflecting superior charge separation efficiency, an essential characteristic for improved photocatalytic performance. Fig. 35a presents the extinction spectra of ciprofloxacin (CIP) degradation under visible light using the MZ-30 photocatalyst.⁴²⁶ The extinction profile of CIP revealed a primary peak at approximately 270 nm, with two additional smaller peaks at 320 and 330 nm. Over time, the intensity of these peaks gradually decreased, confirming the progressive degradation of CIP under the combined influence of the catalyst and light. This process is attributed to the efficient charge transfer between MoS₂ and ZnO, as well as the robust redox reactions taking place on the catalyst surface. Among the different composites, the MZ-30 sample demonstrated the highest photocatalytic efficiency, which can be explained by its optimized band structure and composition. The degradation of CIP using the MZ-30 catalyst reached 89% within 120 minutes of visible light exposure. In contrast, other composites showed lower degradation efficiencies: 57% for MZ-10, 76% for MZ-20, 81% for MZ-40, and 72% for MZ-50. The superior performance of the MZ-30 composite is attributed to its favorable band alignment and enhanced light absorption capabilities, which outperform the pure ZnO catalyst. Fig. 35b presents the graph of ln(C/C₀) versus time for the different MZ composites, showing a consistent pattern of decreasing ciprofloxacin (CIP) concentration over time, signaling effective degradation.⁴²⁶ To assess the impact of catalyst dosage, experiments were conducted using the MZ-30 catalyst while keeping the concentration of CIP and solution volume constant. The catalyst dose was varied from 100 to 250 mg in 50 mg increments. The degradation efficiencies observed for these doses were 91.12%, 94.06%, 96.18%, and 92.10%, respectively (Fig. 35c).⁴²⁶ The observed increase in degradation efficiency with higher catalyst dosages is due to the greater surface area available for redox reactions, leading to more charge carrier generation. However, beyond a dosage of 200 mg, the reaction rate and degradation efficiency began to decrease. This can be attributed to increased solution turbidity, which results in particle aggregation and a light screening effect, diminishing the catalyst's ability to absorb light and reducing the overall reaction rate.⁴²⁶ Table 14 shows the photodegradation of some pharmaceutical pollutants by metal oxides and their various composites.

Fig. 34 (a) Bandgap plots of pure ZnO, MoS₂ and ZnO/MoS₂ hybrid, (b) PL spectra of ZnO/MoS₂ hybrid.⁴²⁶ “Reproduced from ref. 426 with permission from [Elsevier], copyright [2023].”

Fig. 35 (a) Excitation spectra of ciprofloxacin antibiotic using 30%-MoS₂/ZnO sample, (b) plot of ln(C/C₀) vs. time, and (c) rate constant values for varying amounts of catalyst dosage.⁴²⁶ “Reproduced from ref. 426 with permission from [Elsevier], copyright [2023].”

Table 14 Photodegradation of some pharmaceutical pollutants by metal oxides and their various composites

Photocatalyst	Band gap	Light	Target pollutant	Degradation	Ref.
TiO2	3.07	UV lamp, 30 mW cm^−2	TC	90%	427
Zn-doped TiO2	3.02	Xenon lamp, 500 W	TC	85.27	428
TiO2/g-C3N4 (TCN)	3.19	Xenon lamp, 300 W	TC	82.5%	429
Bi2O3	2.8	300 W Xe	TC	80%	430
NiFe2O4/Bi2O3	1.67	150 W Xe	TC	90.78%	431
Zn-doped Cu2O	2.024	Visible light	Ciprofloxacin	94.6%	432
CeO2@WO3	2.72	Visible light	Cephalexin	96.26%	433
WO3	2.79	Visible light	Cephalexin	57.06	433
Ag-CeO2@SnO2	2.2 eV	Sunlight	TC	96%	434
			MNZ	94%
CeO2	2.86	Visible light	SMR	24%	435
La-CeO2	2.70			81%
CdO	2.4–3.5	Solar light	Levofloxacin	80%	436
			Nizatidine	70%

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10. Conclusion

The increasing levels of synthetic dyes, pharmaceutical residues, and persistent organic pollutants (POPs) in many environmental compartments have become a major worldwide problem. They are especially dangerous to aquatic ecosystems and human health because of their chemical stability, resistance to biodegradation, potential for bioaccumulation, and long-range transport characteristics. Since these pollutants have been connected to neurotoxicity, carcinogenicity, endocrine disruption, and antibiotic resistance, immediate and efficient cleanup measures are required. Their extensive presence in water, soil, and air is a serious hazard to both ecological systems and public health. Through extensive classification, structural analysis, and source identification, supported by real-world examples; this review has showed how these pollutants originate, disperse, and accumulate across environmental matrices.

To address their complicated and resistant nature, the review looked at the principles and mechanisms of heterogeneous photocatalysis. It emphasises how light energy and semiconductor materials can be used to completely degrade these contaminants. Metal oxide-based photocatalysts, such as TiO₂, ZnO, WO₃, CuO, Cu₂O, etc. were given particular attention because of their strong oxidative capabilities, environmental compatibility, and structural stability. The photocatalytic mechanism, involving light absorption, electron–hole pair generation, and formation of reactive oxygen species, was discussed in detail to establish a foundational understanding of pollutant breakdown processes. However, despite their potential benefits, these photocatalysts exhibit several significant limitations. Many metal oxides primarily absorb UV light and possess wide band gaps, which lead to the rapid recombination of charge carriers, thereby reducing their photodegradation efficiency. To overcome these challenges, material innovations that improve photocatalytic performance were covered extensively in the review. These included metal and non-metal elemental doping, heterojunction formation (e.g., p–n junctions and Z-scheme systems), and support material integration to increase surface area and electron mobility. The effects of each modification strategy on degradation efficiency, light absorption, charge carrier separation, band gap tuning, and overall performance were considered. Of these, Z-scheme heterojunctions were noted for their ability to preserve strong redox potential while facilitating effective charge separation-an ideal balance for treating persistent and toxic pollutants. The practical applicability of these fabricated materials was confirmed by a comprehensive collection of studies describing the degradation of POPs, dyes, and pharmaceutical residues under UV and visible light.

Nevertheless, a major challenge remains: after water treatment, it is still difficult to separate the catalyst particles from the solution, which hinders the scalability of the process for real-world applications. In conclusion, this review highlights the growing problem of environmental pollution and how advanced metal oxide-based photocatalysts offer a smart and effective solution. By combining knowledge of pollutants and modern technologies, it shows a clear path toward a cleaner and healthier future.

11. Challenges and future outlook

Even though metal oxide photocatalysts have great potential for degrading POPs, dyes, and pharmaceutical pollutants, there are still a number of significant obstacle that prevent their widespread use. These contaminant are extremely resistant to traditional degrading techniques due to their complex chemical structures and stable molecular bonds, which is one of the primary causes of their difficulties in being broken down.

Although photocatalytic degradation methods are extremely effective under controlled settings, they frequently have major limits when applied to stable contaminants. Most of the research on metal oxide photocatalysts, such as TiO₂ and ZnO, focuses on material that primarily absorb in the UV region, which accounts for just a small fraction of total solar light. This drastically reduces their efficiency under real-world settings. Furthermore, the quick recombination of photogenerated electron–hole pairs lowers photocatalytic performance. Although tactics such as doping, morphological tuning, and heterojunction generation (e.g., type-II and p–n systems) have been used, they are still limited by weak redox potentials and poor charge separation. In contrast, Z-scheme photocatalysts offer improved charge separation and redox capabilities; however, they also face challenges such as limited light-harvesting ability and dependence on expensive rare-earth materials. Another practical challenge is the difficulty in recovering and reusing photocatalysts after treatment, which complicates large-scale application and raises concerns about secondary contamination.

Therefore, future studies must concentrate on designing photocatalysts with strong and stable redox characteristics, high visible light activity, and easy recovery or immobilisation for reuse in order to close these gaps. Designing multifunctional composites materials that are affordable, safe for the environment, and able to target many contaminants at once is obviously necessary. In addition, long-term operational stability studies and real-world testing in complicated wastewater matrices are necessary to advance beyond lab-scale viability. To turn these developments into workable environmental solutions, photocatalysis must be integrated with complementing treatment technologies, and scalable synthesis pathways must be developed. Additionally, the use of computational tools such as DFT and machine learning (ML) is essential for predicting the performance of photocatalysts, tuning their electronic structure, and guiding the rational fabrication of novel materials. Artificial Intelligence (AI) and machine learning (ML) are revolutionizing this process by offering insight into nanomaterial toxicity, enhancing life cycle assessments, and ensuring environmental safety. In tackling complex environmental remediation problems, the integration of principles of green chemistry with computational modeling and AI-driven frameworks will enables the fabrication of advanced photocatalysts that are not only highly efficient but also sustainable, scalable, and economically feasible.

Conflicts of interest

The authors declare that they have no competing interests.

Data availability

No primary research results, software or code have been included and no new data were generated or analysed as part of this review

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