Recent progress in photocatalytic degradation of chlorinated phenols and reduction of heavy metal ions in water by TiO 2 -based catalysts

Abstract

Among the various semiconductor photocatalysts reported so far, TiO₂ is still the most promising material for real applications because of its excellent chemical and thermal stability, non-toxicity, low cost and highly oxidising photogenerated holes. This review summarises the recent progress (mainly over the last five years) in photocatalytic oxidation of non-biodegradable organic pollutants (chlorophenols) and reduction of toxic heavy metal ions in aqueous solution. The review details the recently developed strategies for improving the performance of TiO₂-based photocatalysts, with particular respect to the visible light activity, charge separation efficiency, stability, separability and adsorption capacity for the remediation of the aforementioned categories of water contaminants, as these factors heavily affect the practical application of this technology. Next, the underlying semiconductor photocatalytic mechanisms have been thoroughly addressed experimentally and theoretically, together with the proposed defect engineering to improve the photocatalytic performance. Finally, the prospect of TiO₂ photocatalysis was discussed.

Introduction

Access to potable water is one of the biggest challenges globally, especially in developing countries. Several persistent organic pollutants (pharmaceuticals, pesticides, personal care products, endocrine disrupters) are frequently detected in wastewater effluents [1–3]. Chlorophenols (CPs) and their metabolites are common and recalcitrant environmental pollutants, believed to have high bioaccumulation capability and carcinogenic effect. Figure 1 shows four major CPs (2-CP, 2,4-DCP, 2,4,6-TCP and PCP) that have been classified as first-degree toxic pollutants by the US Environmental Protection Agency (EPA 2003) [4,5].

Figure 1.

Chemical structure of major common chlorinated phenols.

Adverse effects of phenolic compounds on the human nervous system have been reported and linked to several health disorders, e.g. kidney failure, liver and lung damage [4–6]. CPs find extensive application in the chemical, forestry and wood-working industries. They are used as herbicides, insecticides, fungicides, wood preservatives and chemical intermediates [4,7–10]. Generally, these organic pollutants are released into the environment because of several man-made activities including water disinfection, waste incineration, uncontrolled use of pesticides and herbicides and as byproducts in the bleaching of paper pulp with chlorine [11].

It is interesting to note that 2-CP, 4-CP, 2,4-DCP and 2,4,6-TCP are the most significant CPs formed as byproducts of water chlorination [12,13]. Effluents from a pulp and paper mill were analysed before treatment and subsequent discharge into the Lake Baikal in Russia, 2,4-DCP concentration of (630 ± 50) µg L⁻¹, 2,4,5-TCP concentration of (2660 ± 210) µg L⁻¹ and 2,4,6-TCP concentration of (320 ± 30) µg L⁻¹ were reported [14]. After treatment, the individual chlorinated phenols were present in trace quantities, except for 2,4,5-TCP with the concentration of (560 ± 50) µg L⁻¹ [14]. Environmental pollution agencies and health organisations recommended a maximum allowable concentration for chlorinated phenols as: 0.1 µg L⁻¹ in drinking water and 200 µg L⁻¹ in wastewater [12].

Various strategies have been employed to remove CPs from the environment, as shown in Table 1. Conventional methods include coagulation-flocculation, reverse osmosis, active carbon adsorption, biodegradation, air stripping and incineration. However, these techniques have some drawbacks and limitations, e.g. toxic byproduct generation, incomplete mineralisation, low efficiency, high energy and capital cost [12–14,16]. CPs absorb light of wavelength below 300 nm; thus, they do not undergo direct sunlight photolysis. Also, they are difficult to decompose completely to CO₂ and HCl by conventional methods due to high stability of the benzene ring [17]. Therefore, it is important to find innovative and cost-effective techniques for the safe and complete degradation of chlorinated organic pollutants such as CPs.

Table 1.

Conventional approaches for the removal of chlorinated phenols [15].

Treatment approach	Drawbacks
Biological treatment	Very slow, incomplete degradation and byproducts may be more toxic than the contaminants
Adsorption technology	Expensive, due to regeneration of the adsorbent materials and post-treatment of solid wastes
Chemical precipitation	Expensive, due to high dosage of chemicals and it produces large quantity of sludge which need post-treatment
Air stripping	High capital cost due to off-gas treatment, aesthetic constraints and technically not reliable

The current trend in wastewater treatment has moved from simple phase transfer to the destruction of pollutants such as by advanced oxidation processes (AOPs). These techniques involve reactive free radical species for non-selective mineralisation of organic compounds to harmless end products. The AOPs are generally based on the generation of the hydroxyl radicals (^.OH), which have a great oxidation power, thus can almost oxidise all organic compounds to carbon dioxide and water [15]. There are several methods for generating hydroxyl radicals, e.g. Fenton-based processes, UV-based processes, Ozone-based processes and Photocatalytic processes [16,18–20].

In spite of the advantages of AOPs, there are several limitations in their use: (a) costs may be higher than competing technologies because of energy requirements, (b) harmful intermediates may be formed, (c) pre-treatment of the wastewater may be required to minimise cleaning and maintenance of UV reactor and quartz sleeves, (d) handling and storage of ozone and hydrogen peroxide require special safety precautions and (e) major challenges for the photocatalytic process are catalyst deactivation, slow kinetics, low photo efficiency and unpredictable mechanism [21]. However, AOPs are still more effective than the other techniques for wastewater treatment containing toxic and persistent pollutants [15].

Recently, application of titanium dioxide photocatalytic oxidation technology in environmental remediation has gained considerable attention as a cheap and clean alternative, though there are still various challenges to be resolved for its commercialisation. A high number of organic compounds (dyes, drug residues, pesticides and herbicides) have been eliminated from aquatic environment via TiO₂ photocatalysis [22–25]. Complementary to the reviews published before, this review only concentrates on the progress of TiO₂ photocatalysis over the past five years, thus providing a comprehensive view of the very recent progress in this field.

Basic principles of TiO₂ photocatalysis and applications

This review addressed the background in photocatalysis, and the removal of wastewater contaminants, via irradiation of semiconductors (e.g. nano-TiO₂) with a light of energy higher than its bandgap energy (Figure 2). This is the first step in photocatalysis. Subsequently, the excitation of photogenerated electrons into the conduction band takes place, leaving behind energetic photogenerated holes in the valence band. Oxidation reaction is initiated in the valence band, while reduction reaction is initiated in the conduction band of the semiconductor. This requires the semiconductor to have proper band alignment required for organic pollutant decomposition (Figure 3), e.g. the conduction band minimum (CBM) should be higher (more negative in potential) than the redox potential for oxygen reduction to superoxide radicals, and its valence band maximum (VBM) should be lower (more positive in potential) than the redox potential for water or hydroxyl ion oxidation to hydroxyl radicals [13,27,28].

Figure 2.

Diagram illustrating the principle of semiconductor photocatalysis for wastewater treatment.

Figure 3.

Diagram illustrating the bandgap energy, VB and CB positions of reported efficient semiconductors. Adapted from Wu et al. [26]. Copyright 2015, Royal Society of Chemistry. Reprinted with permission.

TiO₂ as a photocatalyst

One of the most important aspects of environmental photocatalysis is the availability of an efficient photocatalyst, which should have several respects. For example, it is relatively inexpensive, physically and chemically stable, and non-toxic. Also, the band alignment in such a photocatalyst (Figure 3) makes the photogenerated holes to be highly oxidising. In addition, photogenerated electrons have reduction potential enough to produce superoxide radical from oxygen [26,29,30]. TiO₂ is the typical material meeting these criteria. Zhang et al. [31] and Hirakawa and Nosaka [32], based on experimental evidence, proved that the conduction band electrons in TiO₂ have the potential to generate superoxide radicals (O₂ ^.− ) from dissolved oxygen in water.

Challenges with TiO₂ application in water treatment

The most studied phases of TiO₂ are anatase and rutile. The anatase phase is known to display higher photocatalytic activity than the rutile phase. The low activity observed in rutile (E _g = 3.0 eV, λ = 413 nm), despite its lower bandgap energy than anatase (E _g = 3.2 eV, λ = 388 nm), is attributed to the rapid rate of electron–hole recombination. Also, the CBM of anatase is about 0.2 eV higher than that of rutile (Figure 3). This implies that the conduction band electrons in anatase are more reductive than those in rutile [33,34]. However, the inefficient utilisation of visible light and high recombination rate of photoelectrons and holes, that drastically reduces quantum efficiency, are the major drawbacks of photocatalytic application [28,35]. The challenges with TiO₂ as a potential photocatalyst for the removal of organic pollutants in wastewater are summarised in Figure 4.

Figure 4

Limitations in TiO₂-based photocatalytic processes for the degradation of organic pollutants. Adapted from Dong et al. [34]. Copyright 2015, Elsevier. Reprinted with permission.

Visible light activity

The ultraviolet (UV) region only accounts for <5% of the entire solar spectrum, which TiO₂ is responsive to. The strategies used in extending the optical absorption of TiO₂ into the visible region include, metal-doping, non-metal doping, dyes sensitisation and direct reduction of TiO₂ (e.g. hydrogenation, chemical reduction and electrochemical reduction). Three of the methods will be addressed here.

Metal-doped TiO₂ (e.g. Ag, Pt, Au, Fe, Cu, Co, Ni, Cr, Mn and Ru). This method has been extensively investigated over the past decades for improving the photocatalytic degradation efficiency of TiO₂ for several organic pollutants under visible light (λ > 420 nm) [34,36]. Ag-doped TiO₂ and Cu-doped TiO₂ are the most reported in photocatalytic water treatment. The lowest bandgap energy reported in the literature for metal-doped TiO₂ is 1.6 eV with 7.5% Cu-doped TiO₂, prepared using a low-temperature hydrolysis reaction [37]. Khan et al. reported the synthesis of Ag-doped TiO₂ using sol–gel method and the application in photocatalytic degradation of oxytetracycline under UV–visible light irradiation. The experimental results indicated 70% and 100% photocatalytic degradation of oxytetracycline in 60 min with pure TiO₂ and 1.9 wt.% Ag/TiO₂, respectively, under UV–visible light irradiation. With only visible light irradiation (λ > 420 nm), the degradation efficiency of the 1.9 wt.% Ag/TiO₂ sample reduced drastically to 60% [38]. Generally, for metal-doped TiO₂, the metal ion doping can either be substitutional or interstitial in the TiO₂ crystal lattice and can also form a mixture of oxides [38–43]. The limiting factors with this strategy include the nature of dopant species, dopant concentration, photochemical stability and thermal treatment. Several methods employed in metal-doped TiO₂ (e.g. impregnation, co-precipitation, sol–gel, hydrothermal, spray pyrolysis, ion-assisted sputtering and chemical vapour deposition) were reported [36,44–47]. However, sol–gel method is commonly used in the preparation of nano-size-doped TiO₂ materials. The procedure is simple to carry out and does not require complicated facility [36]. The dopant dosage and calcination temperature play a significant role in photocatalytic activity of prepared TiO₂ samples [48]. The visible light activity of metal-doped TiO₂ is primarily due to the introduction of new energy levels in the bandgap of TiO₂, and photogenerated electrons can be excited from the defect state to the conduction band in TiO₂. Dopant ions above the optimum dosage could act as electron–hole recombination centres and will automatically reduce photocatalytic and quantum efficiencies [34].

Non-metal-doped TiO₂ (e.g. S, N, C, P, F and B). This method extends the visible light absorption of TiO₂ more than metal-doped counterparts. This is because their impurity states are close to the valence band edge of TiO₂ [34]. Compared to other non-metals, the N-doped and C-doped TiO₂ powders have been reported to show exceptional photocatalytic activity under visible light irradiation. Chen et al. [49] prepared C–TiO₂, N–TiO₂ and C–N–TiO₂ using sol–gel method and tested their photocatalytic activities for the degradation of methylene under visible light irradiation. In the absence of catalysts, about 5% photodegradation of methylene blue was observed in 7 h. The C–N co-doped sample exhibited the highest efficiency with about 95% methylene blue removal in 7 h, while 26, 60 and 70% degradation efficiencies were observed for unmodified TiO₂, N-TiO₂ and C-TiO₂ samples, respectively. Co-doped TiO₂ generally exhibit higher visible light absorption than single-doped TiO₂ due to a synergistic effect between the two dopants, which increases charge carrier lifetime (quantum efficiency) [48]. The major drawbacks associated with non-metal doping include the formation of oxygen vacancies in the bulk in which the created defects act as electron–hole recombination sites, thus reducing the visible light photocatalytic efficiency. More so, instability is a major concern due to liberation of doped non-metals atoms in solution during photocatalysis [50,51].

Reduced TiO₂ (TiO_2−x ). Hydrogenation of TiO₂ was reported in 2011 by Chen et al. as an effective method in shifting the bandgap of TiO₂ to around 1.5 eV (∼820 nm) with an optical onset of around 1.0 eV (black TiO₂) [52]. The photocatalytic activity of the coloured TiO₂ was compared with unmodified TiO₂ for methylene blue decomposition under solar light irradiation. Complete bleaching of methylene blue was achieved after 8 min for the black TiO₂, while the unmodified white TiO₂ under the same experimental conditions, took nearly 1 h. Under similar experimental conditions, complete degradation of phenol was achieved after 40 min for the black TiO₂, while the unmodified TiO₂ took 80 min to achieve the same. Total organic carbon (TOC) removal efficiency of the black TiO₂ was not reported by the author. Furthermore, without any form of treatment, the black TiO₂ was highly stable after eight successive photocatalytic degradation cycles. This discovery led to an extensive research into coloured TiO₂ synthesis globally. Other methods used in fabricating reduced TiO₂ include Al reduction [53], Zn reduction [54], imidazole reduction [55], NaBH₄ reduction [56], CaH₂ reduction [57] and electrochemical reduction [58]. It has been widely accepted that heating TiO₂ under vacuum or in a reducing atmosphere also leads to colour changes [59]. The coloured TiO₂ (red, yellow, blue or black) resulted from the creation of Ti³⁺ or/and oxygen vacancies in the white TiO₂. However, their activities in the visible region are still far from satisfactory [59,60].

Dispersibility

TiO₂ particles suffer from severe aggregation during photocatalytic application, which will reduce their active sites and light-harvesting capability. Since photocatalytic degradation occurs at the surface of TiO₂, mass transfer limitations need to be reduced for its effective application in water treatment [61,62]. Two possible approaches employed are:

Stabilisation by support structures (glass beads, fibre glass, glass pellets, glass sheets, silica, organo-clays, stainless steel, Al₂O₃ fibre, quartz beads, polymers, activated carbon and zeolites) [61]. However, there is still no ideal support that is suitable in terms of mechanical stability and selectivity [62]. Immobilisation on supports reduces TiO₂ surface area-to-volume ratio and the photocatalytic efficiency compared to unsupported TiO₂ particles [63].

Stabilisation by surface modification (humic acid, carboxylic acids, amino acids and polylactic acid). This method tends to generate high negatively charged density (zeta potential) on the TiO₂ surface and promotes repulsive forces among particles, thus reduces aggregation rate. The main drawback is the stability of these organic modifiers during photocatalysis [64–67].

Separability

Recovery of fine particles of TiO₂ from solution is a huge challenge in suspension systems. Thus, the recovery and reusability of TiO₂ is required before the technology can proceed beyond lab scale to industrial application. Two potential approaches applied are:

Immobilisation on solid supports (glass fibres, glass, quartz and stainless steel). This method has been discussed under TiO₂ particles stabilisation using solid supports [63]. Although this approach overcomes the solid–liquid separation problem, slurry-type reactors are still more efficient over immobilised catalyst-type reactors, due to the availability of catalyst’s surface and superior mass transfer properties [68].

Magnetic separation (Fe₃O₄, γ-Fe₂O₃, NiFe₂O₄, CoFe₂O₄, FeCo and Co₃O₄). Compared to other magnetic materials, magnetite (Fe₃O₄) has received considerable attention due to its remarkable magnetic properties, low toxicity and biocompatibility [69–71]. This approach is very convenient for separating, recovering and reusing the fine magnetic TiO₂ particles by applying an external magnetic field [72,73]. It was reported that a direct deposition of TiO₂ onto the surface of magnetic oxide particles led to severe photo-dissolution of the magnetic core. In order to overcome the challenge, the addition of silica (SiO₂) layer between the magnetic core and the TiO₂ shell provided magnetic nanoparticles with a chemically inert surface, which inhibited the oxidation of the magnetic core (Fe₃O₄) by the photogenerated VB holes in the outer layer (TiO₂) [74]. The synthesis of a multipurpose catalyst (graphene/TiO₂/SiO₂/Fe₃O₄) was reported for the degradation of 2,4-dichlorophenoxyacetic acid (a herbicide). It took the advantages of TiO₂ photocatalysis, graphene (excellent electron transfer ability and high adsorption capability), Fe₃O₄ (magnetic separation) and SiO₂ (suppressing photo-dissolution of the magnetic core). The TiO₂ composite achieved 100% degradation of 2,4-D under simulated solar irradiation, while commercial P25 recorded only 33% under the same conditions. More so, the composite was highly stable after eight successive cycles. The photocatalytic activity was dependent on the calcination temperature, magnetic core size and the silica layer content [75,76].

Adsorption capacity

The photocatalytic activity of TiO₂ is also dependent on its adsorption capacity for the pollutants in aqueous solution. Several approaches have been used to overcome the low adsorption capacity of TiO₂. Modification of TiO₂ with some chelating ligands (arginine and salicylic acid) and carbon nanomaterials (graphene, carbon nanotubes and activated carbon) have been reported to enhance the adsorption of organic water contaminants [77–79]. With respect to doping with carbon-based nanoparticles, the higher adsorption capability and better photocatalytic activity were reported to be related to their high surface area, high conductivity and high visible light absorption intensity [54,55].

Photocatalytic degradation of chlorophenols

Various studies show that the rate of degradation of CPs is dependent on, solution pH [80–84] and molecular structure – especially the number of substituted Cl groups and their positions on the benzene ring relative to the hydroxyl group [85–87]. It is well known that the initial degradation rate of most organic pollutants with TiO₂ and other photocatalysts follows the pseudo-first-order reaction model [85,88–91]. Table 2 summarises the recent findings (mainly over the past five years) on the application of TiO₂-based photocatalysts for the photocatalytic degradation of CPs in aqueous solution.

Yu et al. prepared a magnetically separable TiO₂/FeO _x microstructure decorated with poly-oxo-tungstate (POM) using hydrothermal method. The photocatalytic degradation of an endocrine disruptor 2,4-dichlorophenol (2,4-DCP) was evaluated under low-intensity solar light [92]. Non-magnetic TiO₂/POM(1%) led to the fastest 2,4-DCP degradation kinetics with 87.6%, compared to magnetically separable TiO₂/FeO _x (25%)/POM(1%) and bare TiO₂ with 76% and 50% 2,4-DCP degradation in 180 min, respectively, under simulated solar light at pH 5.0. POM was reported to scavenge electrons from the conduction band of TiO₂ to decrease the charge recombination, thus enhancing the photocatalytic (quantum) efficiency of TiO₂. Peng et al. synthesised a multipurpose photocatalyst, La-doped magnetic TiO₂ (Fe₃O₄/SiO₂/La-TiO₂) using sol–gel method and investigated its photocatalytic activity for 2,4,6-TCP degradation [93]. The magnetic La-TiO₂ exhibited a higher photocatalytic degradation activity (90% in 60 min) compared to magnetic TiO₂ (50%). However, the La-doped magnetic TiO₂ was fairly stable after seven successive cycles. The drawback was attributed to the reduction in surface hydroxyl groups and reaction sites on TiO₂.

Table 2.

Progress on TiO₂ photocatalysis for the degradation of chlorophenols over the past five years.

TiO2-based photocatalysts	Bandgap energy (eV)	Experimental condition	Performance	Ref.
0.21 mol % Cu–TiO2 (sol–gel)	Nil	(5 × 7 W) Vis LED (λ = 440–490 nm), catalyst dose (3.0 g L−1), 2-CP concentration (20 ppm, pH = 5.5)	100% degradation in 6 h	[83]
1 wt.% M–TiO2 (M = Fe, Cu, Ni)	2.5–3.4	(4 × 15 W) UV lamp (λ = 365 nm), cat dose (2.0 g L−1), 4-CP (10 ppm)	Ni (90%, 80%), Cu (90%, 73%), Fe (37%, 31%) degradation and TOC removal, respectively, in 6 h	[88]
Magnetic TiO2/FeO x /POM	2.65	Simulated solar light, cat dose (0.5 g L−1), 2,4-DCP conc (10 mg L−1, pH = 5)	76% degradation and 55.9% TOC removal in 180 min	[92]
Fe3O4/SiO2/La-TiO2 (sol–gel)	nil	UV–visible light (λ > 340 nm), catalyst dose (2.0 g L−1), TCP conc (10 µM)	90% degradation in 60 min	[93]
V2O5/TiO2/CTAB (solid-state dispersion)	2.21	64-W UV-B lamp, catalyst dose (2 g L−1), DCP (25 mg L−1, pH = 5)	100% degradation in 30 min, 97% TOC removal	[94]
C/Cl–TiO2@C3N4 NTs	nil	300 W Xe lamp (λ > 420 nm), catalyst dose (1.0 g L−1), PCP conc (20 mg L−1)	100% degradation in 4 h, 75% TOC removal in 18 h	[95]
TiO2 (P25)	3.1	500 W Xe lamp (λ > 365 nm), catalyst dose (0.02 g L−1), PCP conc (10 µM)	∼90% degradation in 15 min	[96]
Ag/TiO2 (photodeposition)	Nil	40 W UV LED (λ = 365 nm), catalyst dose (1.0 g L−1), 4-CP conc (10 ppm)	90% degradation in 60 min, 25% TOC removal in 60 min	[97]
TiO2–RGO–CoO (sol–gel)	2.83	200 W Xe lamp (visible light), catalyst dose (0.5 g L−1), 2-CP conc (10 mg L−1, pH = 6)	98.2% degradation in 8 h	[98]
N–In–Sn–TiO2 (sol–gel)	3.10	UV and vis light (λ > 400 nm), cat dose (0.125 and 0.25 g L−1), 4-CP conc (50 µM, pH = 5.38)	96% degradation in 1 h (UV), 90.7% degradation in 8 h (Vis)	[99]
8.0 wt.% B-TiO2 (sol–gel)	2.89	(18 × 8 W) UV-A lamp and (2 × 150 W) visible light (λ > 400 nm), catalyst dose (1.0 g L−1), 2,4-DCP (20 mg L−1, pH = 6.5)	90% degradation in 5 h (UV), 75% degradation in 5 h (Vis)	[100]
I/TiO2 –T (5/2) n(Ti)/n(I) ratio of 5/2 (lignin as template, T)	3.16	350 W Xe lamp (λ > 400 nm), catalyst dose (0.5 g L−1), 4-CP concentration (50 mg L−1)	100% degradation in 60 min, ∼90% TOC removal in 100 min	[101]
CdS@TiO2	2.25	Mic-LED-455 (visible light), cat dose (1.0 g L−1), CPs (10 mg L−1)	73.8% (2,4-DCP) and 77% (2,4,6-TCP) degradation in 360 min	[102]
Fe3O4@TiO2@Au	nil	150 W Xe lamp (UV–vis light), cat dose (0.5 g L−1), TCP conc (7.5 × 10−5 M, pH = 10)	97.7% degradation in 40 min	[103]
2 wt.% CuS/TiO2	2.73	125 W Hg lamp (λ = 365 nm), 4-CP (20 mg L−1, 60 mL)	87% degradation in 150 min	[104]
M/TiO2 (M = CoPc or ZnPc)	2.67–2.78	128 W Lightex LT50 lamp (visible light), cat dose (1.0 g L−1), 4-CP (0.013 M, pH = 7)	99% degradation in 30 min (with both samples)	[105]
Y2O3/TiO2 nanosheets	2.92–3.20	Natural sunlight, cat dose (1.0 g L−1), 4-CP (20 mg L−1)	98% degradation and 67.6% TOC removal in 120 min	[106]
7 wt.% (Fe2O3, Co3O4 and CuO) loaded ZnTiO3–TiO2	2.82–3.12	Natural sunlight, 0.3 mL H2O2, cat dose (2.0 g L−1), 4-CP conc (25 mg L−1, pH = 5)	100%, 76% and 85% degradation in 45 min with Fe2O3, Co3O4 and CuO, respectively	[107]
Fe–TiO2–Ag nano-sphere (0.3 wt.% Fe and 2.0 wt.% Ag)	nil	(2 × 8 W) UV-C lamp, cat dose (1.5 g L−1), 4-CP conc (40.4 mg L−1, pH = 4.86)	97% degradation in 165 min	[108]
10 wt.% GO–TiO2 (titanate nanotubes)	2.8	UV lamp (λ = 254 nm), cat dose (1.0 g L−1), 4-CP conc (50 ppm)	65% degradation in 175 min	[109]
FeOOH/TiO2 (0.14 wt.% Fe)	3.1	300 W Xe lamp (λ > 320 nm), cat dose (0.5 g L−1), 2,4,6-TCP conc (50 mg L−1)	100% degradation and 85% TOC removal in 240 min	[110]

CTAB, cetyl trimethylammonium bromide; NTs, nanotubes; RGO, reduced graphene oxide; Pc, phthalocyanine; GO, graphene oxide; POM, poly-oxo-tungstate.

Sinirtas et al. [94] prepared a binary oxide catalyst (V₂O₅/TiO₂) using solid-state dispersion method and evaluated the photocatalytic activity for 2,4-dichlorophenol degradation under UV–visible light irradiation. Photocatalytic degradation efficiencies of 55%, 60%, 70% and 85% were achieved in 30 min with TiO₂ (synthesised), commercial TiO₂ (P25), V₂O₅ and V₂O₅/TIO₂, respectively. The effect of surfactant additives (CTAB, HTAB and PVA) on the photocatalytic activity of the binary oxide catalyst was also investigated. Under similar operating conditions, the (V₂O₅/TiO₂/CTAB) sample displayed the highest activity (complete degradation in 30 min), while the samples prepared with HTAB and PVA recorded degradation efficiencies of ∼92% and ∼55%, respectively. The enhanced photocatalytic performance of V₂O₅/TiO₂/CTAB was attributed to an appropriate pore distribution and high separation efficiency of photoinduced charge carriers. However, the photochemical stability of V₂O₅/TiO₂/CTAB was not reported.

Photocatalytic reduction of heavy metals

Hexavalent chromium

The discharge of hexavalent chromium (Cr(VI)) from industrial processes, such as electroplating, leather tanning, metal fabrication, paint and pigment production into the aquatic environment has raised serious concerns over the potential water pollution issues [111,112]. In this regard, the reduction of toxic and carcinogenic Cr(VI) to the less harmful Cr(III) is imperative to mitigate their adverse effects towards biological and environmental systems. Over the years, photocatalytic reduction of Cr(VI) by TiO₂ photocatalysts has been regarded as one of the more promising treatment alternatives owing to its direct one-step photocatalytic reduction of Cr(VI) under UV light irradiation [113–115]. Table 3 summarises the recent findings (mainly over the past five years) on the application of TiO₂-based photocatalysts for the photocatalytic reduction of aqueous-based Cr(VI).

The simplified mechanism for the photocatalytic reduction process: Cr₂O₇ ²⁻ + 14H⁺ + 6e⁻ → 2Cr³⁺ + 7H₂O, indicating that the photogenerated electrons separated from the electron–hole pairs are solely responsible for the photocatalytic reduction of Cr(VI) to Cr(III). Lin et al. [131] revealed that the initial reduction rate of Cr(VI) using TiO₂ photocatalyst is following the Langmuir reaction model. They also discussed on the pH-related thermodynamic aspects of the photocatalytic reduction of Cr(VI), as well as found that the photocorrosion of TiO₂ to be relatively unfavourable in basic media. Ku and Jung [132] evaluated the photocatalytic reduction rates of Cr(VI) under various operating conditions (i.e. solution pH, initial Cr(VI) concentration, TiO₂ photocatalyst loads, light intensities and the presence of scavengers). They found a higher photocatalytic reduction rate of Cr(VI) in acidic medium than in basic medium, which is contradicting to the previous study by Lin et al. [131]. The higher photocatalytic reduction rate of Cr(VI) in acidic medium is attributed to the highly positively charged TiO₂ surfaces, as well as the presence of less competing molecules for redox reaction at the lower pH. Yoon et al. [133] immobilised TiO₂ photocatalysts onto inert substrates to overcome the photocatalysts’ recovery and post-separation issues via anodising various nanotubular TiO₂ arrays under oxygen stream in the range 450–850°C for reduction of Cr(VI). Their findings demonstrated that anodised nanotubular TiO₂ arrays annealed at 450°C exhibited the highest reduction efficiency of Cr(VI) at 98% along with a hydrogen evolution rate of 38 μmol h⁻¹ cm⁻². The presence of higher-activity anatase TiO₂ phase in the anodised nanotubular TiO₂ arrays annealed at lower temperatures (<450°C) could explain their excellent reduction efficiency of Cr(VI) [134].

Table 3.

Recent studies on TiO₂-based photocatalysts for photocatalytic reduction of Cr (VI).

TiO2-based photocatalysts	Dopants/cocatalysts	Bandgap energy (eV)	Initial concentration and pH	Light sources	Performance	Ref.
g-C3N4/Ti3+-TiO	g-C3N4 (6 wt-%)	2.7–3.1	10 mg L−1; pH 2	300 W Dy lamp with a 400 nm cutoff filter	∼90% Cr(VI) removal in 120 min	[116]
(C, N and S)-TH–TiO2	Molar ratio of TH to Ti = 1.5:1 C: 10.24 wt-% N: 0.92 wt-% S: 1.15 wt-%	2.95	10 mg L−1; pH 2	500 W Xe lamp	>99% Cr(VI) removal in 50 min	[117]
nZVI–TiO2	FeCl3·6H2O (50 wt-%)	Nil	6 mg L−1; pH 3	18 W UV-C lamp	100% Cr(VI) removal in 2 h	[118]
AgI–TiO2	AgI (5%)	2.67–2.83	8.3 mg L−1; pH 1	300 W halogen lamp	99% Cr(VI) removal in 10 min	[119]
TiO2–graphene hydrogel	TiO2 (20%)	Nil	5 mg L−1; pH 5.5	250 W Hg lamp (λ: 365 nm)	100% Cr(VI) removal within 30 min	[120]
PAN-CNT/TiO2-NH2	PAN (10 wt-%)	Nil	20 mg L−1; pH 2	120 W Xe lamp (λ: 420 nm; Ee : 100 mW cm−2)	>90% Cr(VI) removal in 30 min	[121]
Cu/Cu2O/SA/TiO2 beads	Weight ratio of TiO2 to SA (1:1); C: 12.83% w/w; H: 2.45% w/w	Cu2O/TiO2 = 1.9–2.2	5 mg L−1; pH 4	9 W UV-A lamps (λ: 365 nm); 9 W visible lamps (Ee : 7.2 mW cm−2)	100% Cr(VI) removal in 15 min under UV irradiation; >90% Cr(VI) removal in 100 min under visible light irradiation	[122]
rGO-TiO2	rGO (5 wt-%)	2.82	10 mg L−1; pH 3	Xe lamp (Ee : 300 mW cm−2)	98% Cr(VI) removal in 180 min	[123]
Algae-Ag/TiO2	AgNO3: 0.5%; TiO2: 1%; weight ratio of chitosan to PVA (1:1)	Nil	5 mg L−1; pH 2	500 W halogen tungsten lamp	∼90% Cr(VI) removal in 180 min	[124]
Fe–N–C–TiO2	Fe/Ti weight ratio = 0.30%	2.75	20 mg L−1; pH 2	500 W Xe lamp	100% Cr(VI) removal in 240 min	[125]
CeO2–TiO2	Molar ratio of Ce to Ti = 1:3	2.49	2.78 mg L−1	Visible light	>99.6% Cr(VI) removal in 60 min	[126]
Cu–TiO2	Cu (1.5 wt-%)	Nil	5 mg L−1; pH 5	6 W black-light blue lamps	>90% Cr(VI) removal in 60 min	[127]
CD–TiO2	CD (0.75 wt-%)	Nil	10 mg L−1; pH 3	LED light (λ: 365 nm)	100% Cr(VI) removal within 30 min	[128]
Bi–TiO2	BiCl3: 0.10 M	2.82	40 mg L−1; pH 4.5	32 W visible light fluorescent lamp	53% Cr(VI) removal within 120 min	[129]
Au/Pt–TiO2	Au (10 wt-%); Pt (10 wt-%)	Nil	5 mg L−1; pH 2.5	UV LED and green LED	100% Cr(VI) removal in 25 min	[130]

TH, thiourea; PAN, polyacrylonitrile; CNT, carbon nanotubes; SA, sodium ascorbate; PVA, polyvinyl alcohol; CD, carbon dots.

In order to develop highly efficient TiO₂-based photocatalysts with extended absorption into the visible light spectrum, the introduction of dopants into morphological-modified TiO₂-based photocatalysts has been proven as an effective strategy. The introduction of dopants could hinder the rapid recombination rate of electron–hole pairs and narrowing the wide bandgap of TiO₂ photocatalysts and hence, improving its visible light photocatalytic activity. Lu et al. [135] optimised the amount of vanadium (V) dopant on TiO₂ nanosheets since this sheet-like structure possesses higher exposed surface areas for Cr(VI) adsorption. It is noteworthy that the introduction of V dopant resulted in the reduction of TiO₂ bandgap leading to a high visible light photoactivity, and the co-existence of RhB and Cr(VI) greatly improved photocatalytic degradation kinetics as Cr(VI) is an efficient scavenger of photogenerated electrons. Besides, the V-doped TiO₂ nanosheets could effectively trap photogenerated electrons and introduce a higher density defect (Ti³⁺), which are essential to enhance the effective separation of photogenerated charge carriers and high quantum efficiency. They extended their works in another study by synthesising TiO₂ nanosheets loaded with Mn _x O _y and CoO _x nanoparticles for simultaneous photocatalytic reduction of Cr(VI) and decontamination of organic compounds [136,137]. Similarly, the introduction of Mn _x O _y and CoO _x nanoparticles resulted in a considerable enhancement of visible-light photoactivity, formation of active electron centres and the generation of Ti³⁺ density defects.

Recently, Xue et al. [138] reported the rational design of an interfacial heterojunction photocatalyst between 2D BiOBr nanoplates and 1D rutile TiO₂ nanorods. Figure 5(a) shows the synthesis route of the heterojunction photocatalyst in which the 1D TiO₂ nanorods are first fabricated using the molten salt flux method followed by the bonding between positively charged alkoxide complexes, Bi-(OCH₂CH₂OH)²⁺ and negatively charged TiO₂ nanorods via solvothermal reaction. This eventually leads to the nucleation of BiOBr nanoplates, which are later being vertically embedded with TiO₂ nanorods to synthesise the heterojunction photocatalyst (i.e. Figure 5(b)). The HRTEM image (Figure 5(c)) showed the interplanar spacings of 0.230 and 0.248 nm are well corresponded to the {200} and {101} facets of the rutile phase of TiO₂, respectively. While the measured 0.277 and 0.353 nm fringes of BiOBr nanoplates can be assigned to the {110} and {101} planes of tetragonal BiOBr, respectively. Additionally, the rectangular spotty electron diffraction pattern can be clearly observed in Figure 5(d), indicating the single-crystalline rutile phase TiO₂ in this heterojunction photocatalyst. More importantly, the photocatalytic results (Figure 5(e)) indicated that the TB-2 (i.e. the sample with a molar ratio of Ti to Bi, n(Ti/Bi) = 2:1) exhibited the highest photoreduction rate of Cr(VI) with an apparent rate constant of 0.025 min⁻¹ under visible light illumination. This is ascribed to the formation of an interfacial junction between BiOBr and TiO₂, which provides directional transfer nanochannels for the migration of photoinduced charge carriers and thus, enhancing quantum efficiency and reducing the charge recombination as evidenced by the PL spectra in Figure 5(f) leading to highphotoactivity as per Figure 5(g). Interestingly, similar theory on enhanced physisorption and charge transfer of PANI/MnO₂/TiO₂ nanocomposite shows the importance of TiO₂ heterojunctions to improve the visible light-driven photoreduction of Cr(IV) [139].

Figure 5.

(a) Schematic illustration for the synthesis route of BiOBr nanoplates/TiO₂ nanorods heterojunctions; (b) TEM image; (c) HRTEM image; (d) SAED image of an individual BiOBr/TiO₂ sample; (e) reaction kinetics for photocatalytic reduction of aqueous Cr(VI) over different samples under visible light illumination; (f) PL spectra with an excitation wavelength of 325 nm and (g) transient photocurrent responses under visible light irradiation for all samples. Adapted from Xue et al. [138]. Copyright 2017, American Chemical Society. Reprinted with permission.

Pentavalent arsenic

Owing to their carcinogenicity and mutagenicity nature, arsenic (As) compounds contamination in natural water can present a major threat to the environmental system and public health even in a very low concentration [140–142]. Previous studies are mainly focusing on the oxidation of acutely toxic arsenite (As(III)) to arsenate (As(V)); however, the high adsorbability of As(V) on soil could adversely influence bacteria and microscopic fungi [143,144]. Particularly, this section presents the recent findings on the photocatalytic reduction of aqueous As(V) by using TiO₂-based photocatalysts in detail. Previously, Levy et al. [145] have applied pristine TiO₂ photocatalyst to investigate its photoreduction performance over As(V) under UV light in deoxygenated aqueous suspension. They found that the direct reduction of As(V) to As(IV) by accepting photogenerated electrons at conduction band was not thermodynamically possible. Nevertheless, the addition of sacrificial electron donors (i.e. alcohols or carboxylic acids) in the photocatalytic medium could lead to the production of strongly reductive radicals and thus, promoting the indirect reductive mechanism of As(V).

Similarly, the deposition of transition and precious metals onto TiO₂-based photocatalysts has been widely attempted in order to extend their light absorption into the visible light spectrum as well as serving as active centres to inhibit the recombination of electron–hole pairs [34,146]. For instance, Andjelkovic et al. [147] prepared a Fe-doped TiO₂ photocatalyst via microwave-assisted hydrothermal method in order to provide large capacity for the adsorption and removal of As(V). They found that the adsorption efficiency of As(V) was proportional to the amount of Fe(III) dopants because it tends to provide a higher specific surface area which directly responsible for faster As(V) adsorption rate. Isotherm studies showed that the Langmuir model provides the best fitting for the adsorption data of As(V) better and the Fe-doped TiO₂ photocatalyst attained the maximum adsorption capacity of 17.35 mg g⁻¹ As(V). In another study, Liu et al. [148] deposited Fe onto morphological-modified titanate nanotubes (TNTs) for sequential As(III) oxidation and As(V) adsorption. The resultant Fe-TNTs exhibited a good adsorption performance for As(V) at acidic medium (i.e. adsorption capacity of ∼49.1 mg g⁻¹ at pH 3; removal efficiency of 98.2%) due to the opposite surface charges between Fe-TNTs and As(V) anions. It was found that there are two different forms of deposited Fe that could improve the photocatalytic activity through: (1) Fe³⁺ ions that locate in the interlayers to act as temporary electron- or hole-trapping sites, and (2) attached α-Fe₂O₃ that facilitates the charge separation from TNTs to reduce recombination of electron–hole pairs.

A more recent study by Gao et al. investigated the performance of a 3D cake-like COOH functionalised anatase (TiO₂@C) (Figure 6(a,b)) for simultaneous Cr(VI) reduction and As(III) oxidisation followed by the adsorption of As(V) [149]. The simultaneous conversion of Cr(VI) and As(III) was significantly enhanced from 14.5% to 61.8% and 38.5% to 92%, respectively, in the exposure system of TiO₂@C/As(III)/Cr(VI). It is worth mentioning that As(V), which was oxidised from As (III), has also been effectively removed under acidic condition, as shown in Figure 6(c). Moreover, they also reported that the removal rate of As(V) was essentially relied upon the pH, TiO₂@C dosage (Figure 6(d)) and initial concentration of Cr(VI)/As(III) solution (Figure 6(e)). Before the reaction, Figure 6(f) shows the full XPS spectrum of TiO₂@C containing carbon groups, oxygen and titanium elements. The characteristic peaks of As(III) (45.1 eV) and As(V) (46.1 eV) were detected on TiO₂@C with the proportion of 48.5% and 51.5%, respectively, after exposure to Cr(VI)/As(III) solution (Figure 6(g)) [150]. As a result, the removal performance of As(III)/Cr(VI) using TiO₂@C was improved through co-adsorption of As(III) and Cr(VI) ions, accelerated electrons transfer from As(III) to Cr(VI) initiated by carboxyl groups on TiO₂, and larger binding sites provided by TiO₂ surface.

Figure 6.

(a) TEM and (b) HRTEM images of TiO₂@C; (c) As(V) production in various solution (As(III) concentration: 6.5 mg L⁻¹; Cr(VI) concentration: 5.5 mg L⁻¹; TiO₂@C concentration: 0.3 g L⁻¹; pH: 3.0; time: 90 min); (d) effect of TiO₂@C dosage on removal of total As (As(III) concentration: 5 mg L⁻¹; Cr(VI) concentration: 5 mg L⁻¹; TiO₂@C concentration: 0.1, 0.2, 0.3, 0.5, 0.7 g L⁻¹; time: 90 min; T: 20 ± 2°C pH: 3.0); (e) effect of Cr(VI) initial concentration on adsorption of As(V) (As(III) concentration: 6.5 mg L⁻¹; Cr(VI) concentration: 0–7 mg L⁻¹; TiO₂@C concentration: 0.3 g L⁻¹; time: 90 min; T: 20 ± 2°C pH: 3.0); XPS spectra of (f) pure TiO₂@C full spectrum before reaction; and (g) As 3d spectra of TiO₂@C after exposed in As(III)/Cr(VI) solution. Adapted from Gao et al. [149]. Copyright 2018, Elsevier. Reprinted with permission.

Outlook

After decade’s development on different photocatalysts, TiO₂ is still widely recognized as the best semiconductor material for real application in photocatalytic water purification technology due to its low toxicity, high stability and availability. However, TiO₂ is confronted with some limiting factors such as poor visible light absorption, low photonic efficiency and low adsorption capacity. In this review, we have highlighted various strategies for TiO₂ modification (such as ion doping, cocatalyst decoration and heterojunction formation) not only to improve its light absorption but also to enhance charge separation. Furthermore, this review summarises the recent application of TiO₂-based photocatalysts for the catalytic oxidation of CPs and reduction of heavy metal ions such as chromium and arsenic in aqueous solution, which are two typical while the most challenging topics.

The co-doping of TiO₂ with a metal and non-metal is receiving considerable attention and proved to be more effective than single ion doping towards efficient photocatalytic degradation of organic contaminants in water. This improvement was attributed to an extension in the visible light absorption capacity. Apart from the suppression of photogenerated electron–hole recombination rate, the bandgap energy is also largely decreased due to the introduction of impurity energy levels above the valence band and below the conduction band. The solution pH is also an important operating variable in photocatalytic degradation of organic contaminants. In most TiO₂-based studies, it was shown that acidic pH range (4–6) is the optimum for the efficient degradation of organic substrates in water. This was attributed to the positively charged surface of TiO₂, favouring increased adsorption capacity for organic compounds. It is also worth noting that the fabrication of different bandgap alignments (i.e. Type I, II and III) and Z-scheme heterojunctions was proved to efficiently enhance charge separation of photogenerated electron–hole pairs and reduce the charge carriers recombination rate, beneficial to photocatalytic water treatment.

Reduction of a reactive, high oxidation state of metal ion to a lower oxidation state and less toxic species by TiO₂-based photocatalysts has been regarded as an effective approach for heavy metal removal from aqueous environment. Upon UV light irradiation, the anodic and cathodic redox reactions are initiated, whereby the water oxidation by photoholes is involved and thus transfer of photoexcited electrons from the conduction band of TiO₂ photocatalysts to the adsorbed heavy metal ions takes place. Acceptance of electrons into the metal atom (reduction) could suppress its reactivity and thus, the toxicity effect. Additionally, it is also showed that the equilibrium adsorption capacity of TiO₂ photocatalysts is higher at lower pH medium, signifying a higher density of adsorption sites for heavy metals. In order to extend the functionality and applicability of TiO₂ photocatalysts in visible light range, several material modification strategies such as morphological alteration, dopants introduction and heterojunction are proved to be effective to separate the photogenerated charge carriers and therefore, leveraging the photon responsiveness of pristine TiO₂ photocatalysts.

For both photocatalytic organic contaminants decomposition and heavy metal ions reduction, cocatalysts play a proved key role, which not only help charge separation but also lower the reaction barrier. The so far best cocatalyst is still noble metal, e.g. Pt. To develop a low-cost cocatalsyt is also important for TiO₂ photocatalyst application. Based on the above discussion, we proposed that the future commercial and large-scale application of photocatalysis in water treatment is feasible if the following could be solved:

The preparation of low cost, reproducible and environmental-friendly cocatalyst/TiO₂-based photocatalysts that can be photo-activated by solar light (not artificial UV light).

The efficient recovery and reuse of the photocatalysts during water purification via immobilisation on flexible and highly stable support materials.

The design of photocatalyst materials capable of selective photocatalytic oxidation of organic pollutants or reduction of heavy metal ions in water.

The design of efficient photocatalytic reactor with satisfactory efficiency in acidic, neutral and basic media, reflecting the real environment of the wastewater.

The fundamental understanding underlying the photochemical process is still far behind discovery of the photocatalysts.

The photocatalytic mechanism of a single photocatalyst is relatively simple and to some extent straightforward. The photochemical reaction involves the excitation of electrons to the conduction band of a photocatalyst while leaving positive holes at the valence band upon light illumination whereby the photon energy (hv) is greater than or equal to the bandgap energy of photocatalyst. Studies on charge carrier dynamics have been successfully investigated over the years using time-resolved analytical techniques (time-resolved absorption spectroscopies) to observe the generation, trapping, and transfer of electrons and holes during photocatalytic processes [151,152]. The generation of hydroxyl radicals and superoxide radicals in aqueous systems have been widely established in the literature through spin trapping ESR experiments and fluorescence spectroscopy. The next step is still unclear. The discovery of hydrogen peroxide in oxygen-free systems justifies the generation of hydroxyl radicals by valence band holes, though hydrogen peroxide can also be generated through oxygen reduction. It is still difficult to distinguish the hydroxyl radicals generated from both different pathways, that is, hole oxidation of adsorbed water/hydroxyl ions or through oxygen reduction. Different hydroxyl radical production pathway will provide a different route for improvement of the photocatalytic activity. More importantly, although the time-resolved IR has been successfully employed to other fields, e.g. homogeneous photocatalysis, it has not been used in water treatment. By using this cutting edge technology, the detailed reaction pathway and the rate-determining step could be elucidated, which can also indicate the factor to mitigate the toxic byproducts production during photocatalytic contaminants’ treatment and heavy metal cation reduction. In addition, the development of highly efficient photocatalysts is growing intensively, the photocatalytic mechanism within the complex photocatalyst systems and compositing structures is also important while more challenging. These time-resolved spectroscopies should be applied to these complicate systems to understand the reaction pathway. Overall, a practical photocatalytic system for water treatment still requires substantial efforts.