Pure and modified TiO₂ photocatalysts and their environmental applications

Abstract

Semiconductor photocatalysis is a process that harnesses light energy in chemical conversions. In particular, its applications to environmental remediation have been intensively investigated. The characteristics of TiO₂, the most popular photocatalyst, is briefly described and selected studies on the degradation/conversion of various recalcitrant pollutants using pure and modified TiO₂ photocatalysts, which were carried out in this group, are reviewed. Photocatalytic reactions are multi-phasic and take place at interfaces of not only water/TiO₂ and air/TiO₂ but also solid/TiO₂. Examples of photocatalytic reactions of various organic and inorganic substrates that are converted through the photocatalytic oxidation or reduction are introduced. TiO₂ has been modified in various ways to improve its photocatalytic activity. Surface modifications of TiO₂ that include surface platinization, surface fluorination, and surface charge alteration are discussed and their applications to pollutants degradation are also described in detail.

Introduction

Semiconductor photocatalysis has been extensively studied as a vital device that utilizes solar photoenergy [1]. The intrinsic bandgap in semiconductors serves as a light harvesting center. Photocatalytic reactions occurring on semiconductor surface have many applications including chemical fuel synthesis (e.g., H₂ production through water splitting) [2], selective oxidation [3,4], degradation of organic compounds [5,6], disinfection [7,8], metal corrosion prevention [9–11], lithography [12,13], etc. In particular, their application to the remediation of polluted water and air has been demonstrated to be a technically viable process and TiO₂ has been the most popular and successful photocatalyst for this purpose [5, 6,14–16]. This technology has a unique versatility in that it can be applied to various environmental media of water, air, and even solid phases. Photocatalytic conversion or degradation with illuminated TiO₂ has been demonstrated for a huge number of substances, which is largely ascribed to the strong oxidation potential of its valence band (VB) holes and surface OH radicals [5,17]. TiO₂ is ideally suited as a practical environmental photocatalyst because it is extremely stable, non-toxic, safe to handle, inexpensive, and photoactive under solar light. Although the TiO₂ photocatalysis seems to be a narrowly defined subject of research, to our surprise, it has been continuously producing a large number of research papers, new findings, patents, and commercial products for more than 30 years.

Research activities on TiO₂ photocatalysis can be classified into many different subjects such as reaction kinetics and mechanisms, material synthesis and modification, thin film and coating fabrication, surface and colloid chemistry, photoelectrochemistry, charge recombination and transfer dynamics, and reactor design and engineering. The level of current TiO₂ research activities widely ranges from the very fundamental such as the flash laser spectroscopic studies of charge recombination dynamics [18–21] to the development of commercial products such as air purifiers and self-cleaning glasses [22–24]. As for the application, TiO₂ is probably second to none in its diversity, which is summarized in Table 1. Thanks to the interdisciplinary nature of TiO₂-related research and the diversity of its applications, TiO₂ is certainly one of the most frequently and thoroughly studied materials in the world. On the other hand, a tremendous number of works on TiO₂ modification to improve its efficiency have been carried out in various ways, which include impurity doping [25–27], sensitization [28–31], surface modification [32–37], and fabrication of composites with other materials [38,39]. Modified TiO₂ can be very different from pure TiO₂ in many aspects such as light absorption, charge recombination dynamics, interfacial charge transfer kinetics, surface structure and charge, and adsorption of substrates. As a result, the photocatalytic reactions of modified TiO₂ markedly differ from those of pure TiO₂. This article is intended to review selected topics related with “TiO₂ as an environmental photocatalyst” and most works discussed in this paper are limited to those performed in the author’s laboratory in the last 5 years.

Table 1 Photocatalysis research examples classified into application areas

TiO₂ photocatalysis as an advanced oxidation process

General characteristics of TiO₂ photoactivation

Titanium dioxide generates a pair of a conduction band (CB) electron and a VB hole in the solid oxide lattice upon absorbing a photon with energy greater than 3.2 eV (or λ < 388 nm) and the subsequent charge transfers at the interface initiate various kind of redox reactions under the ambient condition (in both air and water). In principle, any semiconductor with an appropriate magnitude of the bandgap and the position of band edges is able to initiate photoinduced redox reactions on its surface. Figure 1 compares the bandgap and the band edge position of various semiconductors at pH 0. Other wide bandgap semiconductors with high positive values of E _vb such as ZnO, WO₃, and SnO₂ often show some oxidative photocatalytic reactivities [17]. However, in most cases their photoactivities are lower than those of TiO₂. The strong remedial power of TiO₂ photocatalysts is mainly ascribed to the strong oxidation potential of VB holes (E _vb =  +2.7V _NHE at pH 7) or OH radicals that are produced on TiO₂ surface through the reaction of VB holes with the surface hydroxyl groups or adsorbed water molecules (reaction 1). The reaction of VB hole should accompany the CB electron transfer to maintain the electroneutrality of the catalyst particle and the typical scavenger of CB electrons is O₂ (reaction 2).

$$ {\text{h}}_{{\text{vb}}} ^ + + {\text{ > OH}}_{{\text{surf}}}\ {\text{or}}\ ({\text{H}}_{\text{2}} {\text{O}}_{{\text{ad}}} {\text{)}} \to \bullet {\text{OH}}_{{\text{surf}}} $$

(1)

$$ {\text{e}}_{{\text{cb}}} ^{\text{ - }} + {\text{ O}}_{\text{2}} \to {\text{O}}_{\text{2}} ^ - $$

(2)

Such a sensitized photooxidation in the TiO₂-mediated remediation is closely compared with the nature’s photochemical cleanup mechanisms [40].

Figure 1.

Energy-level diagram showing the CB and VB edge positions of various semiconductors at pH 0 along with selected redox potentials. The energy scales are referenced against both the vacuum level and the normal hydrogen electrode (NHE).

Photocatalytic conversion of organic and inorganic pollutants

Environmental remediation technologies that are based on the in situ generation of oxidant radical species such as OH radicals in the polluted media are named “advanced oxidation processes (AOPs)” and many physicochemical methods have been employed as AOPs [40]. TiO₂ photocatalysis is an excellent example of AOPs that utilizes artificial or solar light in environmental cleanup processes. TiO₂ photocatalysts have shown excellent performances in oxidizing and subsequently mineralizing various organic pollutants [5]. As a result of the photocatalytic oxidation, all elements present in a molecule are mineralized to inorganic species: carbon to CO₂, hydrogen to H₂O, halogens to halide ions, sulfurs to sulfates, and phosphorus to phosphates, respectively. Here we describe several examples of TiO₂ photocatalytic oxidation/reduction reactions of organic and inorganic pollutants.

Tetramethylammonium hydroxide (TMAH) is being consumed in quantity in semiconductor industry where it is used as a silicon etchant [41]. Since this chemical is very stable and recalcitrant against degradation, the conventional wastewater treatment process is not efficient at all in removing it. Figure 2 compares the time-dependent decays of (CH₃)₄N⁺ and the formation of intermediates and products in UV-illuminated TiO₂ suspension [15]. The photocatalytic degradation of (CH₃)₄N⁺ proceeded with sequential demethylation to generate (CH₃)₃NH⁺, (CH₃)₂NH ⁺₂ , CH₃NH ⁺₃ , and NH ⁺₄ . Each demethylation step consists of a series of interfacial reactions and is initiated by an H-atom abstraction from the methyl group (reaction scheme in figure 2). This provides an example how the photocatalytic oxidation can be successfully applied to degrading a recalcitrant water pollutant. However, the photocatalytic oxidation process applied to the wastewater treatment is often very costly and not easily scaled up, which hinders its commercial exploitation. The significant gap between the scientific feasibility and the practical engineering needs to be bridged by the development of more efficient photocatalytic materials and ingenious engineering.

Figure 2.

Photocatalytic degradation of (CH₃)₄N⁺ and the formation of intermediates and products as a function of the irradiation time in aqueous TiO₂ suspension at pH 3.4. The bottom reaction scheme proposes a photocatalytic demethylation mechanism.

Polychlorinated dibenzo-p-dioxins (PCDDs) are one of the most problematic and toxic pollutants and classified into persistent organic pollutants (POPs) because of their refractory character. PCDDs can be also degraded by TiO₂. Figure 3 compares the direct photolytic and photocatalytic degradation of four PCDDs (loaded on a glass plate or a TiO₂-coated glass plate) under UV irradiation and ambient air [14]. The direct photolysis with λ > 300 nm did not induce any noticeable degradation of PCDDs but their photocatalytic degradation on TiO₂ was greatly enhanced, and resulted in 85% conversion within 15-h irradiation for OCDD (octachlorodibenzo-p-dioxin). Neither stable intermediates nor less chlorinated dioxin congeners from OCDD degradation were detected under the present analytical conditions where authentic less chlorinated dioxin was easily detected. Since PCDDs and their degradation products reside on TiO₂ surface in the air throughout the photoirradiation, stable degradation intermediates, if any, seemed to be volatilized into the air or not to be extracted by toluene from the TiO₂ surface. Similar photolytic and photocatalytic behaviors were observed for HpCDD (hepta-), TCDD (tetra-), and MCDD (mono-): insignificant photolytic degradation and greatly enhanced photocatalytic degradation on TiO₂ for all of them. In situ diffuse reflectance FTIR spectra of OCDD adsorbed on TiO₂ showed that the aromatic vibrational peaks of OCDD was gradually reduced with irradiation, which indicates that the benzene ring of the dioxin molecule is destroyed as a result of photocatalytic reaction [14].

$$ {\text{C}}_{\text{6}} {\text{Cl}}_{\text{4}} {\text{O}}_{\text{2}} {\text{C}}_{\text{6}} {\text{Cl}}_{\text{4}} + \bullet {\text{OH}} \to {\text{ [C}}_{\text{6}} {\text{Cl}}_{\text{4}} {\text{O}}_{\text{2}} {\text{C}}_{\text{6}} {\text{Cl}}_{\text{4}} {\text{OH]}} \bullet $$

(3)

The photocatalytic oxidation begins with the addition of an OH radical to the dioxin benzene ring, which leads to the formation of a hydroxycyclohexadienyl radical (reaction 3). The preferred position of OH radical attack on dioxin seems to be dependent on the number and position of chlorines. The hydroxycyclohexadienyl radical immediately reacts with O₂ in the air, then the cleavage of the aromatic ring follows, which was suggested from the FTIR analysis of OCDD degradation. In general, the photocatalytic oxidation process is well suited for the destruction of POPs and other recalcitrant organic pollutants that are present at very low concentrations in environmental media.

Figure 3.

Comparison of the photocatalytic degradation of four PCDD congeners as a function of irradiation time under UV irradiation (λ > 300 nm). Direct photolysis was negligible for all congeners.

The photocatalytic reactions initiating on the TiO₂ surface have the multi-phasic character. The photocatalytic degradation reactions of organic substances take place not only at the TiO₂/water and TiO₂/air interfaces but also at the TiO₂/solid interfaces. Figures 4 and 5 show the evidences (SEM images) that the photocatalytic reactions are able to proceed at the TiO₂/(organic polymer) and TiO₂/(carbon soot) interfaces, respectively. We prepared the TiO₂ particle-embedded PVC films and irradiated them with UV light. The SEM images (figure 4) show that the degradation of the PVC matrix started from the PVC–TiO₂ interface and led to the formation of cavities around TiO₂ particle aggregates [42]. The PVC–TiO₂ film that was irradiated under nitrogen atmosphere showed little sign of degradation. The cross-sectional SEM images of soot layer on TiO₂ film (figure 5) exhibit the progressive degradation of soot as the irradiation time increases [43]. The image shows that the soot layer of ∼2 μm thickness completely disappeared after 32 h irradiation, which corresponds to a soot oxidation rate of ∼65 nm/h. The production of CO₂ from the photocatalytic degradation of both PVC and soot was confirmed by gas chromatographic measurements. Therefore, this is essentially a combustion process of organic solid in which oxygen molecules photoactivated in the ambient air condition combine with organic carbon to generate CO₂.

Figure 4.

SEM images of the PVC–TiO₂ (1.5 wt%) composite film surface. (a) before irradiation; (b) 25 h irradiated; (c) 50 h irradiated; (d) 100 h irradiated.

Figure 5.

Cross-sectional SEM images of soot-coated TiO₂ films on a glass plate. The UV light was illuminated from the TiO₂ side. (a) before illumination, (b) 18 h illuminated, (c) 32 h illuminated.

Although the oxidants should be produced on the surface of TiO₂ film, a strict two-dimensional surface reaction at the PVC/TiO₂ or soot/TiO₂ interface cannot account for the degradation of the organic solid bulk. This implies that the active oxidants generated on TiO₂ surface desorbed and migrated into the bulk of the organic solid. The phenomenon that TiO₂ photocatalyst is able to oxidize a substrate that is remote from the active surface site has been repeatedly observed in other studies [44,45]. The mobility of photooxidants should play a critical role in the photocatalytic degradation of solid substrates because the substrates are immobile in this case. The SEM images in figure 6 clearly verify that the photooxidants generated on TiO₂ are migratory. The soot film deposited alongside the TiO₂ film was degraded with developing a gap between the edges of soot and TiO₂ domains: the gap distance continuously increased with UV illumination up to 80 μm [43]. The active oxidants formed on irradiated TiO₂ surface desorb and migrate across the glass surface to reach the soot domain.

Figure 6.

Remote photocatalytic degradation of soot near the edge of TiO₂ domain. (a) the schematic illustration of the photocatalytic degradation of soot layer near the borderline of soot and TiO₂ domains where a gap distance, d develops between edges of TiO₂ and soot domain with illumination time. (b) The SEM image shows the developing gap after 6 h irradiation.

Photooxidants other than OH radicals may contribute to TiO₂ photocatalytic oxidation. We investigated the photocatalytic oxidation of arsenite (As(III)) to arsenate (As(V)) in aqueous TiO₂ suspension [16,46]. The preoxidation of As(III) to As(V) is recommended in the treatment of arsenic-contaminated waters since As(V) is less toxic and more easily removed by adsorbents. As(III) could be rapidly converted into As(V) in illuminated TiO₂ suspension but the main photooxidants do not seem to be OH radicals since the addition of excess OH radical scavenger, tert-butylalcohol, did not reduce the oxidation rate at all. It has been proposed that superoxides play an important role as oxidants (reaction 4).

$$ {\text{As(III)}} + {\text{O}}_{{\text{2}}^ \bullet } ^{\text{ - }} + {\text{2H}}^ + \to {\text{As(IV)}} + {\text{H}}_{\text{2}} {\text{O}}_{\text{2}} $$

(4)

$$ {\text{As(IV)}} + {\text{O}}_{\text{2}} \to {\text{As(V)}} + {\text{O}}_{{\text{2}}^ \bullet } ^{\text{ - }} $$

(5)

The photocatalytic oxidation rate was significantly reduced in the presence of superoxide dismutase. Superoxides are generally considered as a weak oxidant. However, in this specific case of arsenite photooxidation, they seem to be efficient oxidants.

Most photocatalytic degradation reactions occurring on TiO₂ are initiated by an oxidation step such as an OH radical attack or VB hole transfer. However, not all substances can be degraded in this way. Perchlorinated compounds such as CCl₄ and CCl₃CO ⁻₂ are good examples. CCl₄ and CCl₃CO ⁻₂ without any C–H bond react with neither OH radicals nor VB holes. However, their photocatalytic degradation in TiO₂ suspensions has been successfully demonstrated, which was ascribed to the role of CB electrons (reactions 6, 7) [47,48].

$$ {\text{CCl}}_{\text{4}} + {\text{e}}_{{\text{cb}}} ^{\text{ - }} \to \bullet {\text{CCl}}_{\text{3}} + {\text{Cl}}^{\text{ - }} $$

(6)

$$ {\text{CCl}}_{\text{3}} {\text{CO}}_{\text{2}} ^{\text{ − }} + {\text{e}}_{{\text{cb}}} ^{\text{ - }} \to \bullet {\text{CCl}}_{\text{2}} {\text{CO}}_{\text{2}} ^{\text{ − }} + {\text{Cl}}^{\text{ − }} $$

(7)

The reductive dechlorination step is followed by a series of thermal radical reactions, which lead to full degradation. Such reductive degradation is usually enhanced in the presence of electron donors (e.g., alcohols and organic acids). The reducing power of CB electrons in TiO₂ is generally not strong enough to dechlorinate chlorohydrocarbons but the reduction potentials of perchlorocompounds are positive enough to initiate reactions 6 and 7 upon reacting with CB electrons. CB electrons in TiO₂ may initiate the degradation of organic compounds in a reductive way for limited cases only.

The reactivity of CB electrons in TiO₂ can be also utilized in the reductive conversion of heavy metal ions. Photocatalytic reduction of metal ions leads to conversion to lower oxidation states (e.g., Cr(VI) → Cr(III)) [49,50] or deposition onto TiO₂ surface as a zero-valent metal (e.g., Ag⁺ → Ag⁰) [50]. For example, the platinum metal deposition on the surface of TiO₂ particles, which is frequently performed to enhance the photocatalytic activity of TiO₂, is typically done by the photocatalytic reduction of platinum ions. The UV-illumination of aqueous TiO₂ suspension in the presence of PtCl ²⁻₆ and electron donors results in the deposition of Pt⁰ on the TiO₂ surface (reaction 8) [32].

$$ {\text{PtCl}}_{\text{6}} ^{{\text{2 - }}} + {\text{4e}}_{{\text{cb}}} ^{\text{ - }} \to {\text{Pt}}^{\text{0}} + {\text{6Cl}}^{\text{ - }} $$

(8)

Figure 7 shows a TEM image of Pt nanoparticles that were photocatalytically deposited on TiO₂ via reaction 8. Such reductive conversion of heavy metal ions is enabled not only under UV irradiation but also under visible light when dyes that serve as a sensitizer are co-present. Figure 8 illustrates a case in which excited dyes transfer electrons to metal ions through TiO₂ CB [51]. As a result, dyes are oxidized and metal ions are reduced simultaneously under visible light. The ternary system (TiO₂/dye/metal ion) exhibits highly enhanced conversion efficiencies for both dye (Acid Orange 7: AO7) and heavy metal ion (Cr(VI)) under visible light, compared with the binary systems (TiO₂/dyes or TiO₂/metal ions).

Figure 7.

Pt nanoparticles deposited on the surface of TiO₂ particles (Degussa P25) through a photoreductive conversion of PtCl ²⁻₆ .

Figure 8.

Visible light-induced simultaneous oxidation of dyes and reduction of metal ions on TiO₂ particles. Conversion of (a) AO7 (C₀ = 100 μM) and (b) Cr(VI) (C₀ = 100 μM as Cr₂O ²⁻₇ ) in the binary or ternary systems under visible light illumination (λ > 420 nm). The experimental conditions were air-equilibrated; pH = 3.0; [TiO₂] = 0.5 g/L.

As discussed in the above, the photocatalytic reactions are very versatile owing to their multi-phasic nature and the kind of chemical substances that can be destructed or transformed photocatalytically is almost unlimited. The target of photocatalytic conversion is not limited to chemical pollutants. Microorganisms, pathogens, and algae can be killed photocatalytically using TiO₂ on the basis of the similar mechanisms involving photooxidants, which is being intensively studied for disinfection applications [7,8].

On the other hand, the development of photocatalytic reactors is vitally important from the practical aspect. However, compared to a large number of research works devoted to the mechanistic studies of photocatalytic reactions and the synthesis and development of high efficiency photocatalytic materials, much less effort has been made in the area of photocatalytic engineering and reactor development for commercial exploitation [52]. Most studies of photocatalytic purification of water have been carried out using TiO₂ slurry that is not very useful for practical applications. TiO₂ particles in water should be removed after treatment or immobilized on supports. Immobilized TiO₂ photocatalytic reactors are generally less efficient than the slurry-type reactors because of the reduced catalyst surface area, the reduced exposure of the catalyst surface area to light, and the mass-transfer limitation and suffer from the surface deactivation and the lack of long-term durability of immobilized TiO₂ coatings. In this respect, the slurry-type reactor coupled with membrane filtration is attractive. We recently constructed a pilot-scale photocatalyst-membrane hybrid reactor (500 L volume) and characterized its performance in terms of the degradation efficiency of organic pollutants and the degree of membrane fouling under various operational conditions [53]. Figure 9 shows the schematic diagram of the reactor system that has a submerged membrane module and an air blower that plays the multiple roles of mixing the suspension, supplying oxygen, and inhibiting the membrane fouling. 4-Chlorophenol of 100 ppb could be completely removed within 2-h batch operation. In continuous runs, no fouling of the membrane (or no suction pressure build-up) took place with an intermittent operation with the 9-min suction and 3-min pause period.

Figure 9.

Schematic of a pilot-scale photocatalyst-membrane hybrid reactor. During the reactor operation, the concentration of suspended TiO₂ particles and the flow rate of effluent were kept at a constant level of 0.5 ± 0.1 g/L and 2 L/min, respectively.

Modification of TiO₂ photocatalysts

Properties and reactivities of surface platinized TiO₂

The surface platinization of TiO₂ has been a popular photocatalyst modification technique, since Kraeutler and Bard [32] first introduced it, because the platinized TiO₂ (Pt/TiO₂) exhibits enhanced activity for many photocatalytic reactions [54–56]. The presence of Pt deposits on TiO₂ is believed to retard fast charge-pair recombination by serving as an electron sink (Schottky-barrier electron trapping) and to facilitate the interfacial electron transfer to dioxygen or other electron acceptors, which has been supported by electrochemical and time-resolved spectroscopic investigations [21,57].

To demonstrate the role of Pt deposits on TiO₂ as an electron sink, the Fe³⁺-mediated photocurrents that are collected onto an inert electrode are compared between pure TiO₂ and Pt/TiO₂ suspensions (figure 10) [58]. A higher photocurrent is obtained in the suspension of Pt/TiO₂. This should be ascribed to the fact that CB electrons are trapped in the surface Pt phase and consequently more electrons are transferred onto the collector electrode via the electron shuttle (Fe³⁺/Fe²⁺). To take another example, the electron trapping ability of the Pt phase was successfully utilized in the development of dye-sensitized TiO₂ photocatalysts. Figure 11 describes the working principle of ruthenium bipyridyl-complex (Ru^IIL₃) sensitized TiO₂ photocatalyst under visible light [30]. Figure 12 clearly shows that loading Pt nanoparticles on TiO₂ surface dramatically enhances the visible light-sensitized dechlorination of CCl₄ [30]. In the absence of Pt on TiO₂, most electrons injected from the excited sensitizer into TiO₂ CB recombine with the oxidized dye (path 2 in figure 11), which limits the overall photonic efficiency smaller than 10⁻³. With Pt loaded on TiO₂, both the electron trapping into the Pt phase (path 3) and the back electron transfer (path 2) typically occur within 1 μs [21,29], and therefore the electron trapping on Pt can compete with the back electron transfer. As a result, the interfacial CB electron transfer to the electron acceptor (CCl₄) can be highly enhanced when Pt is deposited on the sensitized TiO₂.

Figure 10.

Comparison of photocurrent generation in UV-illuminated deoxygenated suspensions of TiO₂ + Fe³⁺ and Pt/TiO₂ + Fe³⁺ with acetate used as an electron donor (D). Experimental conditions were [TiO₂] = [Pt/TiO₂] = 0.5 g/L; [Acetate]₀ = 0.2 M; pH_i = 1.95 ± 0.05; platinum collector electrode held at + 0.6 V (vs. SCE); illuminated with 30 W-black lamp; N₂ purged continuously during the test.

Figure 11.

Visible light-induced reductive degradation of perchlorinated compounds (Cl₃CX: X = Cl for CCl₄, X = –CO ⁻₂ for CCl₃CO ⁻₂ ) on Pt/TiO₂/Ru^IIL₃ particles. The number represents the major electron pathways: 1, electron injection from the excited sensitizer to CB; 2, back electron transfer to the oxidized sensitizer (Ru^IIIL₃); 3, electron migration and trapping in Pt deposits; 4, interfacial electron transfer to a perchlorinated molecule on Pt; 5, sensitizer regeneration by electron donors.

Figure 12.

Time-dependent chloride production from CCl₄ degradation on TiO₂/Ru^IIL₃ and Pt/TiO₂/Ru^IIL₃ under visible light. The effects of adding 0.1 M isopropyl alcohol (IPA) on the chloride production are compared as well. The experimental conditions were: [TiO₂] = 0.5 g/L, pH_i = 3, [Ru^IIL₃]_i = 10 μM, [CCl₄] = 1 mM, λ > 420 nm, and initially N₂-saturated.

Pt deposits on TiO₂ not only enhance the photocatalytic activity by serving as an electron sink and consequently retarding fast charge-pair recombination but also change reaction pathways by providing catalytic sites. The catalytic activity of Pt is well known but its catalytic role in photocatalytic reactions using Pt/TiO₂ has been little investigated. Here we describe a few examples.

Photocatalytic degradation of ammonia in aqueous TiO₂ suspension was slow and resulted in the quantitative conversion to NO ⁻₂ and NO ⁻₃ . On the other hand, the photocatalytic degradation of ammonia with Pt/TiO₂ was much faster and accompanied with a significant reduction in the total N-mass, which implies the presence of missing products [34]. In order to account for the deficit in the N-mass balance, GC/MS analysis was performed to detect the peak at m/e = 30 (¹⁵N₂) from the photocatalytic oxidation of isotope-labeled ¹⁵NH₃ on Pt/TiO₂. A similar experiment using naked TiO₂ did not produce even a trace of m/e = 30 signal. This confirmed that NH₃ on Pt/TiO₂ photocatalytically transformed into N₂. Pt deposits on TiO₂ stabilize the intermediate atomic nitrogen species and thus facilitate their recombination into dinitrogen. Figure 13 compares the anoxic photocatalytic conversion of dimethylamine ((CH₃)₂NH) between the pure and platinized TiO₂ suspensions as another example [59]. The photodegradation of dimethylamine on Pt/TiO₂ was much faster and yielded products different from those obtained with pure TiO₂. In particular, not only demethylated amines but also N-methylated amines were produced as a byproduct in the deaerated Pt/TiO₂ suspension. About 30% of dimethylamine was converted into trimethylamine in the deaerated Pt/TiO₂ suspension within 1 h of irradiation, whereas no conversion into trimethylamine was observed in the presence of dissolved O₂. This case also demonstrates the catalytic role of Pt in the photocatalytic reactions. The degradation of TCA on Pt/TiO₂ is discussed as the third example. The photocatalytic degradation of TCA should be initiated by CB electron transfer as mentioned earlier (see reaction 7) [48]. The resulting dichloroacetate radical rapidly reacts with dissolved O₂, leading to complete destruction with no stable intermediates produced.

$$ \bullet {\text{CCl}}_{\text{2}} {\text{CO}}_{\text{2}} ^{\text{ - }} + {\text{O}}_{\text{2}} \to \bullet {\text{OOCCl}}_{\text{2}} {\text{CO}}_{\text{2}} ^{\text{ - }} \to \to {\text{2Cl}}^{\text{ - }} $$

(9)

Therefore, the photocatalytic degradation of TCA in pure TiO₂ suspension needs O₂ as a reagent and is markedly retarded in deaerated suspension [33]. However, when Pt/TiO₂ was used as a photocatalyst, the degradation rate was enhanced in the absence of O₂, which implies that the Pt catalyst provides an alternative degradation path. It is proposed that the reaction of dichloroacetate radical with VB holes (reaction 10) is enabled on Pt/TiO₂ and is responsible for this anoxic pathway [33].

$$ \bullet {\text{CCl}}_{\text{2}} {\text{CO}}_{\text{2}} ^{\text{ - }} + {\text{h}}_{{\text{vb}}} ^ + \to {\text{CCl}}_{\text{2}} + {\text{CO}}_{\text{2}} $$

(10)

The resulting dichlorocarbene (CCl₂) could be hydrolyzed to yield chloride ions in an anoxic solution [60]. When the anoxic path (reaction 7 + reaction 10) is dominant, the presence of O₂ reduces the reactivity by scavenging CB electrons. This anoxic mechanism seems to be effective only in the presence of Pt deposits on TiO₂. As the last example of the Pt catalytic effect, figure 14 shows the photocatalytic oxidation of gaseous CO on Pt/TiO₂ photocatalysts prepared with three commercial TiO₂ samples (Degussa P25, ISK STS-01, Hombikat UV-100). The photocatalytic conversion of CO to CO₂ was quantitative in the presence of O₂ and proceeded on Pt/TiO₂ at much faster rate than on bare TiO₂ [54].

Figure 13.

Photocatalytic conversion of dimethylamine [(CH₃)₂NH] on (a) naked and (b) Pt/TiO₂ in the anoxic suspension. (C₀ = 500 μM; [TiO₂] = 0.5 g/L; pH_i = 10.4) .

Figure 14.

Time-dependent profiles of photocatalytic removal of CO on bare TiO₂ and Pt/TiO₂. Three commercial TiO₂ samples (TiO₂(D) [Degussa P25], TiO₂(H) [Hombikat UV-100] and TiO₂(I) [ISK STS-01]) are compared for their photoactivities. The experimental conditions were [O₂] = 20 vol.%, Pt loading of 3 wt%, UV intensity of 3 mW/cm², and no water vapor added.

On the other hand, the reported Pt effects in the photocatalytic degradation of substrates have not been always positive and even contradictory for the same substrate. For example, Chen et al. [61] reported that the platinization of TiO₂ drastically reduced the photocatalytic degradation rate of trichloroethylene (TCE) and Driessen et al. [62] observed a similar phenomenon. On the contrary, Crittenden et al. [63] reported that such a significant retardation in TCE degradation was not observed with Pt/TiO₂. Our recent study explains why there are discrepancies in the reported observations [64]. The photocatalytic degradation rates of TCE, PCE (perchloroethylene), and DCA (dichloroacetate) in air-equilibrated aqueous suspensions are compared in figure 15 when a series of Pt/TiO₂ samples prepared with varying Pt photo-deposition time (1, 2, 5, 30, and 60 min) were used as a photocatalyst. As for DCA degradation, a volcano-shape activity curve, which has been frequently observed in other studies [30,56] was obtained. However, the platinization effect on the degradation of TCE and PCE is drastically different from the general case. TiO₂ that was photo-platinized for 1, 2, and 5 min had detrimental effects on the degradation of TCE and PCE while the 30-min photodeposition of Pt recovered the photoactivity to a level similar to that of bare TiO₂. This result clearly demonstrates that the Pt effects in TiO₂ photocatalytic reactions depend on not only the Pt loading but also the kind of substrates, hence cannot be generalized. Our study revealed that the oxidation state of Pt deposits is very important in determining the initial degradation rates of chlorinated organic compounds. TiO₂ with oxidized Pt species (Pt_ox/TiO₂) was less reactive than TiO₂ with metallic Pt (Pt⁰/TiO₂) for all substrates tested. In particular, Pt_ox/TiO₂ strongly inhibited the degradation of TCE and PCE whereas it was more reactive than bare TiO₂ for other compounds. The main effect of photodeposition time (in figure 15) was related to the change in the oxidation state of deposited Pt species. X-ray photoelectron spectroscopic analysis showed that Pt^II and Pt^IV were the major species in the early photodeposition period (1–2 min) but Pt⁰ was dominant after longer photodeposition (30–60 min). It should be realized that the Pt effects in photocatalysis are substrate-specific and highly dependent on how the platinization is done.

Figure 15.

Effects of Pt photodeposition time on the photocatalytic degradation rates of TCE, PCE, and DCA. Experimental conditions were [Pt/TiO₂] = 0.5 g/L; [TCE]₀ = 0.4 mM; [PCE]₀ = 0.25 mM; [DCA]₀ = 1 mM; pH = 4.0; air-equilibrated [Note that the abscissa scale (photodeposition time) is not linear].

Properties and reactivities of surface fluorinated TiO₂

The surface fluorination of TiO₂ (F–TiO₂) is a simple ligand exchange between fluoride anions (F⁻) and surface hydroxyl groups on TiO₂ in water (reaction 11) [35,36].

$$ \equiv {\text{Ti-OH}} + {\text{F}}^{\text{ - }} \leftrightarrow \quad \equiv {\text{Ti-F}} + {\text{OH}}^{\text{ - }} \quad {\text{p}}K_{\text{F}} = {\text{6}}{\text{.2}} $$

(11)

It was recently reported that the surface fluorination of TiO₂ improves the photocatalytic oxidation rate of phenol [35] and tetramethylammonium (TMA) [36] in a specific pH range. Since the surface fluorides themselves should not be reactive with VB holes [E ⁰(F^•/F^-) = 3.6 V vs. NHE] [65] the higher photocatalytic oxidation rate in the F–TiO₂ suspension has been ascribed to the enhanced generation of mobile free OH radicals (reaction 12) whereas most OH radicals generated on naked TiO₂ surface prefer to remain adsorbed (reaction 13) [35].

$$ \equiv {\text{Ti - F}} + {\text{H}}_{\text{2}} {\text{O}}\,({\text{or OH}}^{\text{ - }} {\text{)}} + {\text{h}}_{{\text{vb}}} ^ + \to {\text{ }} \equiv {\text{Ti - F}} + \bullet{\text{OH}}_{{\text{free}}} + {\text{H}}^ + $$

$$ \equiv {\text{Ti-OH}} + {\text{h}}_{{\text{vb}}} ^ + \to \, \equiv {\text{Ti-OH}} \bullet ^ + $$

(13)

Effects of surface fluorination on the photocatalytic reactivities are very different depending on the kind of substrates to be degraded. F–TiO₂ is more effective than pure TiO₂ for the photocatalytic oxidation of Acid Orange 7 and phenol, but less effective for the degradation of dichloroacetate [37]. It is proposed that the OH radical-mediated oxidation pathways are enhanced on F–TiO₂, whereas the hole transfer-mediated oxidations are largely inhibited due to the hindered adsorption (or complexation) of substrates on F–TiO₂. As for the photocatalytic reduction, the dechlorination of TCA is much retarded on F–TiO₂. The photocurrents collected in TiO₂ suspensions, which are mediated by electron shuttles (methyl viologen or ferric ions) as in figure 10, are also markedly reduced in the presence of F^–. The surface ≡Ti–F group seems to act as an electron-trapping site and to reduce interfacial electron transfer rates by tightly holding trapped electrons because of the strong electronegativity of the fluorine.

The surface fluorination effect is observed not only at TiO₂/water interface but also at TiO₂/air interface. The enhanced desorption of mobile OH radicals from F–TiO₂ surface into the air could be clearly verified through a study of remote photocatalytic oxidation [66]. A stearic acid-coated glass plate and a TiO₂-coated plate were faced to each other and held together, but separated by a small intervening gap (30 μm) as illustrated in figure 16a. A black-light UV lamp (10 W: Sankyo Denki) irradiated the sample from the TiO₂ side. A control experiment was carried out with reversing the TiO₂-coated glass plate upside down. As shown in figure 16b, the remote photocatalytic degradation of stearic acids over F–TiO₂ was much faster than over bare TiO₂. It indicates that the generation of free OH radical as air-borne oxidant is enhanced over F–TiO₂.

$$ \equiv {\text{Ti-F}} + {\text{H}}_{\text{2}} {\text{O}}_{{\text{ad}}} + {\text{h}}_{{\text{vb}}} ^ + \to \, \equiv {\text{Ti-F}} + \bullet {\text{OH (air-borne)}} + {\text{H}}^ + $$

(14)

Figure 16.

(a) Illustration of the experimental setup for the remote photocatalytic oxidation of stearic acid (SA). (b) Remote photocatalytic degradation of SA over TiO₂ vs. F–TiO₂ film as a function of UV irradiation time. The degradation of SA was monitored by the IR absorption of C–H stretching band.

Surface charge modifications

In TiO₂ photocatalytic systems where the substrate adsorption is essential as a prerequisite step, the surface charge modification of TiO₂ influences the photocatalytic reactivity for ionic substrates by altering the electrostatic interaction between the catalyst surface and the substrate. The surface charge of bare TiO₂ is positive at acidic condition (pH ≤ 5) due to the presence of ≡Ti–OH ⁺₂ groups, near neutral at pH 5–7, and negative at basic condition (pH ≥ 7) owing to ≡Ti–O^– groups [5,67]. This pH-dependent surface charge of TiO₂ can be controlled by loading other metal oxides or adsorbing charged substrates. We introduced anionic surface charge by loading silica [68] or nafion polymer (cation-exchange resin) [69] onto TiO₂ and demonstrated that the surface charge-modified TiO₂ highly accelerated the photocatalytic degradation of cationic substrates. Figure 17a shows that the surface charges of both silica-loaded TiO₂ (SiO₂/TiO₂) and nafion-coated TiO₂ (Nf/TiO₂) are significantly shifted to the negative values compared with that of bare TiO₂. As a result, SiO₂/TiO₂ exhibited a highly enhanced activity for the degradation of TMA (cationic substrate) (figure 17b) [68]. Nf/TiO₂ showed an enhanced reactivity for the degradation of TMA as well [69]. The visible light-sensitized degradation of cationic dyes (e.g., methylene blue) was also enhanced with Nf/TiO₂ [69]. In particular, the sensitized degradation of rhodamine B (RhB) followed a different path when the surface of TiO₂ was coated with nafion. The N-de-ethylation of RhB that leads to the generation of rhodamine-110 was a prevailing path with Nf/TiO₂ whereas the cleavage of the chromophoric ring structure was dominant with bare TiO₂ [69]. The photocatalytic degradation of N-nitrosodimethylamine (NDMA: an emerging water pollutant) was also enhanced with Nf/TiO₂ [70]. Although NDMA is a neutral molecule, the highly concentrated protons within the nafion layer facilitate the formation of Lewis acid complex with NDMA and enhance its photocatalytic degradation.

Figure 17.

(a) Zeta potentials of bare TiO₂, SiO₂/TiO₂, and Nf/TiO₂ particles suspended in water ([catalyst]₀ = 2 mg/L) as a function of pH. (b) Photocatalytic degradation of TMA in aqueous suspensions of bare TiO₂ vs. SiO₂/TiO₂.

Conclusions

TiO₂ photocatalysis is being investigated from diverse points of view that are related with materials synthesis and modification, reaction kinetics and mechanisms, reactor engineering, and surface chemistry to take a few examples. The application areas of TiO₂ photocatalysis are very diverse, which seems to be largely responsible for the longevity of this field. Photocatalytic reactions taking place on the surface of TiO₂ can be applied to the degradation of pollutants in water, air, and even solid phases, which makes this technology very versatile. Mechanistic understanding about heterogeneous photocatalytic reactions is far from complete due to the complex nature despite intensive research efforts. The current status of knowledge suggests that the mechanisms in photocatalytic reactions are hard to generalize and should be understood on a case-by-case basis. Some examples discussed in this article also reveal such aspects. Although pure TiO₂ is a reasonably good photocatalyst, a variety of methods have been employed to improve the efficiency and overcome the inactivity in the visible light region. Only a few examples are introduced in this article. Very efficient photocatalysts that are modified from TiO₂ or new materials are yet to be developed for successful commercialization. In addition, more efforts in photocatalytic engineering and reactor development are required to commercialize the photocatalytic detoxification technology.