Phenolic compounds are found in surface and groundwater as well as wastewater from several industries. It is necessary to eliminate phenols and phenolic compounds from contaminated water before releasing into water bodies due to their toxicity to human beings. Photocatalytic degradation seems to be a promising technology for the degradation of several phenolic compounds. Complete mineralization of phenol and phenolic compound has been achieved with TiO2-based photocatalysts under both UV and visible-light irradiation. This chapter will evaluate the conventional processes and advanced oxidation processes for the degradation of phenol and phenolic compounds. The process economics and efficiencies of different advanced oxidation processes will also be discussed. The main focus of the chapter is photocatalytic degradation processes under UV and visible light along with a detailed review of several factors affecting degradation of phenol and phenolic compounds. Photocatalytic degradation process is governed by reactions with hydroxyl radical or superoxide ion. The extent of degradation depends on light sources (UV, visible, and solar), the type of photocatalyst, and experimental conditions (pH, photocatalyst dosage, initial concentration of phenolic compounds, light intensity, electron donor concentration, etc.).Visible-light-active photocatalysts are applied by several researchers to exploit sunlight and to make the photocatalysis process sustainable. In the future, using sunlight in place of UV could make photocatalysis economically more efficient.
- dye sensitization
Phenol and phenolic compounds (chlorophenols, nitrophenols, etc.) detected in water and wastewater are toxic in nature and treated as primary water pollutants as per different countries’ regulations. Phenol is one of the first compounds included in the US EPA list of priority pollutants . Chlorophenols and nitrophenols are even more toxic than phenol itself. The exposure, health effects, and regulatory limits of phenols are mentioned in Table 1 . High concentration of phenolic compounds is present in the effluents from different industries, such as textiles, plastics, paint, paper, petroleum refining, coal processing, wood products, pharmaceuticals, and steel manufacturing . Phenols can be removed by conventional techniques such as (i) physicochemical processes and (ii) biological processes. A comparative study of different phenol degradation methods is presented in the following section. Because of several limitations of the conventional phenol degradation processes, researchers are now relying on advanced oxidation processes (AOPs) for the complete mineralization of phenols. AOPs provide much faster degradation rate with the participation of hydroxyl radical (HO•), and phenols are mineralized to CO2 and water instead of transferring the pollutants from one phase to another . Heterogeneous photocatalysis process became very popular among the AOPs, and it requires mainly three components such as (i) semiconductor photocatalyst, (ii) light energy (UV or visible or solar), and (iii) electron donor or hole acceptor. When semiconductor photocatalyst is illuminated with light energy greater than the band gap of the photocatalyst, charge carriers (i.e., electron-hole pair) are produced which ultimately generate hydroxyl radicals (HO•) in the system. Recently, photocatalytic degradation of phenol and phenolic compounds in wastewater has been extensively studied by several research groups [4–13]. Titanium dioxide (TiO2) photocatalyst is frequently used in the degradation of phenols under ultraviolet light [14–16]. TiO2 is nontoxic, photostable, insoluble under most conditions, and inexpensive and has exceptional chemical and biological inertness . There are few other photocatalysts such as ZnO, CuO, and
|Phenol and phenolic compounds||Use/exposure||Health effects||Human health for the consumption of water and organism (maximum contaminant levels) |
|Phenol||(i) Used in the production of aniline, phenolic resins, cresols, alkylphenols, dyes, pesticides, synthetic fiber, disinfectant, and antiseptic; (ii) found in industrial effluent from different industries such as pulp and paper, food|
processing, textile, pharmaceuticals, coal gasification, and petroleum refining
|Headache, skin irritation, kidney damage, liver damage||4 mg L−1|
|Chlorophenols||(i) Used in antiseptics and pesticide production,|
(ii) produced in chlorine-bleaching process
during paper making, (iii) produced via chlorination of humic matter during the chlorination of drinking water, and (iv) also produced in textile, chemical, and
in the mouth, headache, lung damage, affects the digestive
|2-Chlorophenol (0.03 mg L−1), 3-methyl-4-chlorophenol (0.5 mg L−1), 2,4-Dichlorophenol (0.01 mg L−1), pentachlorophenol (0.03 μg L−1), 2,4,6-Trichlorophenol (1.5 μg L−1)|
|Nitrophenols||(i) Formed via reaction of phenol and|
nitrite ions in water under light (UV or solar),
(ii) also produced during production and degradation of pesticides, (iii) produced from
metallurgic and electronic industry, and (iv)
used in solvents, dyes, and plastic production
muscles pain, anorexia,
|2,4-Dinitrophenol (0.01 mg L−1)|
In the first part of this chapter, conventional treatment methods for the degradation of phenol and phenolic compounds are presented, followed by the application of AOPs for such treatment. The process economics and efficiencies of different AOPs are also discussed for the degradation of phenolic compounds.
In the second part of the chapter, we focus on the photocatalytic degradation processes concerning different areas such as (i) basic principle of photocatalysis, (ii) experimental details of photocatalytic degradation of phenols, (iii) photocatalysis reaction mechanism for the degradation of phenols, and (iv) effect of different experimental parameters on degradation of phenol and phenolic compounds.
Finally, we demonstrate a dye-sensitized method to improve the photocatalytic activity and visible-light response of TiO2-based photocatalyst to perform visible-light-driven phenol degradation.
2. Degradation methodologies of phenol and phenolic compounds
It is of utmost importance to treat wastewater containing phenols before disposal to the environment in order to save the aquatic life. Physical/chemical treatment methods such as activated carbon adsorption, ion exchange, liquid-liquid extraction, chlorine oxidation, chlorine dioxide oxidation, and hydrogen peroxide oxidation are mostly applied for the removal of phenols. However, these methods are expensive and produce hazardous by-products. On the other hand, biological treatment methods for phenol degradation are superior in the above aspects, but these are applicable only for low concentration of phenols . AOPs normally produce hydroxyl radicals (HO•) as active species which have very low selectivity and can drive phenol degradation through complete mineralization . A review of the conventional and advanced degradation methods of phenol and phenolic compounds is presented in Table 2.
|1.||Chemical oxidation with|
|Phenol||Solution pH |
9, stirring for
1 h, Fe(VI)O42−
to phenol molar ratios (1:1, 10:1,
phenol initial conc. 30 ppm
|(i) Phenol oxidation follows a second-order reaction kinetics;|
(ii) at Fe(VI)/phenol of 10:1, phenol degrades 100%,
TOC decreases 57%, and COD decreases 82%;
(iii) oxidation reaction follows a radical pathway to
undergo ring opening forming intermediates such as phenoxyphenol and benzoquinone
with potassium permanganate
|Phenol and bisphenol A (BPA)||Solution pH|
(4–8.5), initial conc. of BPA
0–16.8 μM, phenol initial conc. 10 μM, permanganate conc. 181 μM
|(i) Oxidation follows a second-order reaction kinetics;|
(ii) oxidation of both phenol and BPA improve under mildly acidic conditions (pH 4–6);
(iii) oxidation of phenol delays at pH range 7.5–8.5;
(iv) manganese intermediates such as Mn(V) and
Mn(VI) form during the reactions
with ozone (O3)
|BPA||O3 concentration, solution pH, bicarbonate concentration, initial conc. of BPA 35 μM||(i) BPA oxidation with aqueous O3 follows a second|
-order rate equation at pH 7;
(ii) O3 conc. and solution pH show a significant
effect on BPA removal
|4.||Adsorption||BPA||Adsorbent: powdered activated|
carbons impregnated with iron oxide nanoparticles
|(i) Reach adsorption equilibrium in 150 min;|
(ii) adsorption follows a Freundlich isotherm;
(iii) iron oxide impregnation improves BPA removal
thin-film composite reverse
(50–1000 ppm), ionic strength (NaCl conc. 0.1–0.001 molL−1), transmembrane pressure
|(i) Phenol retention depends on transmembrane|
pressure, feed concentration, solution pH, and ionic strength;
(ii) phenol mainly diffuses through the membrane;
(iii) phenol retention exceeds 85% at alkaline pH;
(iv) electrostatic repulsion plays a more important role than size exclusion process
|6.||Nanofiltration||Phenol||Composite polyamide nanofiltration membrane, transmembrane pressure, pH, recovery rate, volumetric cross flow rate, phenol initial conc. 132 ppm||(i) 97% phenol and 96% COD removal take place by|
cross flow nanofiltration;
(ii) nanofiltration membrane successfully removes coke-oven wastewater containing phenol, oil and grease, cyanide, and ammonia
|7.||Solvent extraction||Phenol||Cumene is the extractant, pH range for extraction 1–7, phase ratio between 0.5 and 4, extraction temperature (from 25 to 55°C), phenol initial conc. 500–5000 ppm||(i) Stripping efficiency for phenol is more than 99%;|
(ii) cumene shows excellent extraction performance
on phenol in acidic solution
|8.||Biodegradation by activated sludge||BPA||Operating temperature 20°C, DO|
4 ppm, pH 7.5, sludge age 30
or 45 days,
hydraulic retention time 48 h, BPA initial conc. 40 ppm
|(i) Metabolic intermediates are 4-hydroxyacetophenone, 4-hydroxybenzaldehyde, and 4-hydroxybenzoic acid;|
(ii) biodegradation kinetics is influenced by the sludge age, BPA concentration, and the acclimation process
|9.||Enzymatic process||Phenol (in refinery effluent)||Packed bed bioreactor, biocatalyst weight 135 g, effluent pH 7, temperature|
(20–32°C), flow rate( 3–6 ml min−1), H2O2 conc.
phenol initial conc. 100 ppm
|(i) 97% phenol degradation is attained;|
(ii) H2O2 concentration, temperature, and flow rate
have a positive effect on phenol degradation;
(iii) immobilized enzyme shows better stability at
broad pH range and at high temperature
with UV, O3, and TiO2
|Phenol||A low-pressure mercury lamp|
(λ, 184.9 and 253.7 nm), circulation
flow rate for phenol 1000 ml min−1, phenol initial conc. 50–200 ppm
|(i) Phenol degradation follows a pseudo-first-order kinetics;|
(ii) O3−UV-TiO2 process achieves complete degradation
of phenol within 2 h;
(iii) formic acid, acetic acid, propionic acid, and fumaric acid are the reaction intermediate
|11.||Advanced oxidation by Fenton process||Phenol||Concentration|
of iron (II)
sulfate 0.001 mol L−1, flow rate of
0.15, and 0.3
mol per 30 min), pH 3, temperature 30°C, phenol initial conc. 0.012 mol L−1
|(i) About 94% organic degradation possible in 2 h;|
(ii) excess iron(II) is responsible for the lower efficiency
of Fenton process;
(iii) higher H2O2 flow rate provides best results
of 4-NP, light intensity, partial pressure of oxygen, photocatalyst concentration, pH, chloride ion, and temperature
|(i) Degradation rate of 4-NP follows pseudo-first-order kinetics with respect to its concentration;|
(ii) Cl− ion shows a negative effect on the degradation
|Phenol||Chelating agents (ascorbic acid|
or citric acid), solvent volumetric
ratio, electron scavenger
(H2O2), phenol initial conc.
0.1–0.4 mmol L−1
|(i) Three-dimension ordered macroporous (3D-OM) bismuth vanadates successfully remove phenol (94% removal) from wastewater under visible light;|
(ii)Bi(+V)/chelating agent optimum molar ratio is 2:1
TiO2 mass ratio, photocatalyst calcination temperature
|(i) Photocatalyst activity depends on the doping amount of S;|
(ii) maximum activity is observed when the photocatalyst is calcinated at 600°C with the mass ratio of thiourea/TiO2 of 1:1
-visible-light-multiwalled carbon nanotubes (MWNT)-TiO2 composite
|Phenol||MWNT/TiO2 ratio, MWNT surface area, photocatalyst preparation method,|
phenol initial conc. 50 ppm
|(i) MWNT-TiO2 composite photocatalysts are|
synthesized via modified sol-gel method;
(ii) the increase of MWNT/TiO2 ratio from 5 to 20%
favors the enhancement of the synergetic effect on
photocatalyst loading, triethanolamine concentration,
Pt loading on TiO2, visible
solar light intensity,
phenol initial conc. 20–100
|(i) Superoxide ion is the active species for phenol degradation;|
(ii) complete phenol degradation is achieved in 1 h
with initial phenol concentration of 20 ppm
(pH 7.0, light intensity 100 mWcm−2)
|(i) Degradation order of the different phenolic compounds: chlorophenol > trichlorophenol > dichlorophenol > phenol|||
There are two major constraints that need to be considered for industrial applications: (i) technical feasibility and (ii) economic feasibility. The overall costs of the processes are calculated by summing up the capital cost, operating cost, and maintenance cost . In the following section, we compare the costs associated with different advanced oxidation methodologies. The treatment costs of the AOPs are ranked on a 0–5 scale, 0 being the most expensive and 5 being the least. In between 0 and 5, the ranking is evaluated based on Eq. (1) :
The different process costs are compared by few authors. Saritha et al.  compare UV, UV/TiO2, UV/H2O2, UV/Fenton, Fenton, and H2O2-based AOPs for the degradation of 4-chloro-2 nitrophenol (4C-2-NP). Based on the overall costs, we find that AOP carried out using H2O2 and Fenton are least expensive having ranks of ~5, while UV is the most expensive, assigned a rank of 0. Figure 1 shows the cost ranking of the different AOPs for the degradation of 4C-2-NP. Esplugas et al.  compare UV, O3/H2O2, O3/UV, O3/UV/H2O2, UV/H2O2, and O3 processes for phenol degradation. Again, based on the overall costs, the different O3-based AOPs are least costly, while UV is the most expensive, as evident from Figure 2. We can infer from the cost comparison that incorporating a photocatalyst such as TiO2 with UV lowers the overall cost by one-third . In the future, using sunlight in place of UV could make AOPs economically more efficient.
3. Photocatalytic degradation of phenol and phenolic compounds
3.1. Basic principle of photocatalysis
The precise definition of heterogeneous photocatalysis is a tricky one; particularly as in many cases, the complete mechanism of the reactions is uncertain . In photocatalytic reactions, liquid or gas phase reactants and/or products come into contact with the light-absorbing semiconductor photocatalyst . The semiconductor material can be activated by photons with sufficient energy equal to or greater than the band gap energy (Eg) between the conduction band and valence band of the material . A photocatalytic reaction initiates with the formation of electron-hole pairs followed by oxidation and reduction reactions . In the presence of hole scavenger, the reduction reactions become predominant, whereas in the presence of electron scavenger, the oxidation reactions are the key reactions. However, there are some unwanted reactions such as recombination of electron-hole pairs which reduces the photocatalysis efficiency . Figure 3 provides a detailed mechanism of photocatalytic reactions.
3.2. Experimental details of photocatalytic degradation of phenols
Photocatalytic degradation of phenols is performed with either slurry or immobilized photoreactors . Slurry photoreactors provide a high photocatalytic surface area to reactor volume ratio but require filtration after the reaction. On the other hand, immobilized photoreactor can be used continuously without any photocatalyst separation step. However, immobilized reactors suffer from mass transfer limitations and high light scattering . Slurry photocatalyst provides much higher phenol degradation efficiency than immobilized photocatalysts . Our research group used few different types of photoreactors for the degradation of phenol and phenolic compounds. Chen and Ray  used two-phase monolithic-type photoreactor for the photodegradation of 4-nitrophenol under UV light. Sengupta et al.  used a Taylor vortex reactor (TVR) for the degradation of phenol under UV light. Chowdhury et al.  used slurry photoreactor with dye-sensitized photocatalyst, and Malekshoar et al.  used slurry photoreactor with graphene-based photocatalyst for phenol degradation under solar light. Studies by other researchers reported degradation of phenols with (i) TiO2-coated-fiber-optic cable reactor , (ii) tubular photoreactor , (iii) continuous flow photoreactor , and (iv) solar photoreactor (CPC modules and flat reactor) .
3.2.2. Light sources
Photocatalysis process efficiency largely depends on photocatalyst surface area and incident photons. Ray  combined these two factors and came up with a parameter called illuminated photocatalyst surface area (κ, m2 m−3) which mainly represents the illuminated photocatalyst inside the photoreactor to undergo the photocatalysis process. Therefore, distribution of light inside the photoreactor is a crucial factor. In majority cases of phenol degradation, photoreactors use an external light source (UV or solar) with a slurry reactor. Chen and Ray  used 125 W high-pressure Hg vapor lamp (Philips) in a swirl-flow reactor. However, such external-type photoreactors are limited by the low value of κ, and thus scale up is not possible. Sengupta et al.  used a TVR with immersion-type lamp and immobilized photocatalyst for phenol degradation and achieved a κ value of 80 m2 m−3. In such case, photoreactor, scale up is possible with larger reactor volume . Chowdhury et al.  used a solar simulator (1000 W Xe arc lamp with AM 1.5 G filter) in an external-type slurry photoreactor for dye-sensitized phenol degradation under solar-visible light. Gimenez et al.  studied the photocatalytic degradation of phenol and 2, 4-dichlorophenol under natural sunlight using compound parabolic collectors (CPCs) and the flat reactor (cylindrical tank). CPCs showed higher phenol degradation efficiency, but it is technologically more complicated than the flat reactor.
Degussa P25 TiO2 (DP25) is the most common photocatalyst used for phenol degradation under UV light. Some other commercial TiO2 photocatalysts such as Hombikat UV100, TTP, and PC500 are also used for the same. Among them, DP25 provides the highest photocatalytic activity due to slow electron-hole recombination during photocatalysis . Several visible-light-active photocatalysts such as eosin Y-sensitized TiO2/Pt , dye-sensitized TiO2, MWNT-TiO2 composite , S-doped TiO2 , and BiO4  are also used for degradation of phenol and phenolic compounds.
3.2.4. Experimental procedure
Aqueous solutions of target compounds (phenol and/or phenolic compounds) are prepared at a desired initial concentration. Solution pH is adjusted with HCl or HNO3 or NaOH solutions. In some cases, buffer solutions are used to maintain the solution pH. In the case of slurry photocatalyst, the powered photocatalysts are dispersed in the solution with ultrasonication and then mixed with a magnetic stirrer. Sometimes the photocatalyst slurry is circulated through a peristaltic pump. Appropriate electron acceptor or hole scavenger is used in the reaction mixture. At first, the dark reaction is performed to study the adsorption behavior of phenols over the photocatalyst. Then photocatalytic reactions are performed under illuminated conditions, and aqueous samples are collected at regular time interval to check the residual concentration of phenols [7, 14].
3.2.5. Analyses of phenols
Chowdhury et al.  used high-pressure liquid chromatography (HPLC) to quantify the concentration of phenol and phenolic compounds in aqueous medium The instrument is equipped with a column oven and a diode array detector. AC18 column (5 μm × 150 mm × 4.6 mm) and a mobile phase of methanol and water (67/33% v/v) at a flow rate of 0.5 ml min−1 are used. The temperature of the column oven is kept at 25°C throughout the analysis. The wavelengths of analyses for phenol and reaction intermediates catechol, hydroquinone, and 1,4-benzoquinone are done at 270, 290, 275, and 255 nm, respectively.
3.3. Photocatalysis reaction mechanism for the degradation of phenols
Here we are discussing the photocatalysis reaction mechanism for TiO2 photocatalyst only. The first step in the photocatalytic degradation is the formation of electron-hole pairs within the TiO2 photocatalyst. Most of the electron-hole pairs are recombined producing heat energy. However, hydroxyl radicals (HO•) are formed in the presence of electron acceptor (dissolved O2) while hole (h+) oxidizes water or TiO2 surface active ─OH group. Dissolved O2 reacts with the electron (e−) and generates superoxide ion (O2−•). Finally, the HO• reacts with either phenol or phenolic compounds until complete mineralization. Photodegradation mechanism of 4-nitrophenol (4-NP) under UV light is presented as follows :
Overall reaction stoichiometry shows complete mineralization of 4-NP with the involvement of HO• (Eq. (6)). Devi and Rajashekhar  described the possible degradation mechanism for phenol under natural sunlight/UV light using nitrogen-doped TiO2. Phenol mineralization went through the formation of dihydroxybenzene (catechol or resorcinol), pent 2-enedioic acid, and oxalic acid. In a parallel reaction path, benzoquinone and maleic acid were formed during the mineralization (Figure 4).
3.4. Effect of different experimental parameters on degradation of phenol and phenolic compounds
Different parameters such as solution pH, light intensity, initial concentration of target compounds, photocatalyst concentration, and electron acceptors play a significant role on photocatalytic degradation of phenol and phenolic compounds. The following section will provide a review of recent studies on the degradation of phenol and phenol derivatives.
3.4.1. Effect of solution pH
Solution pH plays a vital role in the photocatalytic degradation of phenol and phenolic compounds since it influences two surface properties of the photocatalyst: (i) band edge position and (ii) surface charge. TiO2 P25 shows a point zero charge at pH 6.8. Thus at pH < 6.8, TiO2 surface attains positive charge and can easily adsorb anionic species at the photocatalyst surface . Again, the protonation and deprotonation of phenols greatly depend on solution pH. Different phenolic compounds show different optimum pH during photodegradation. Venkatachalam et al.  studied the photocatalytic activity of Mg2+ and Ba2+-doped TiO2 nanoparticles for the degradation of 4-chlorophenol (4-CP). In the acidic pH (pH 5), 4-CP was well adsorbed on the photocatalyst surface and showed higher degradation rate than alkaline pH. Lathasree et al.  reported the photocatalytic degradation of phenol and chlorophenols with ZnO under UV light. Significant phenol degradation was achieved at neutral and mildly acidic pH. The zero point charge for ZnO was 8, and at alkaline pH, chlorophenols exist as negatively charged chlorophenolate anion. Thus the photodegradation rate was higher at acidic pH (pH < 8).
3.4.2. Effect of light intensity
Photodegradation rates of different organic compounds improve with increasing light intensity. At high light intensity when mass transfer limitation is low, the reaction rate is found to be proportional to the square root of light intensity. However, at the low-intensity level, the photodegradation rate is directly proportional to the light intensity [14, 54]. Al-Sayyed et al.  observed a similar rate shift from first order to half order in intensity while they studied photocatalytic degradation rate of 4-CP in the light intensity range of 2–50 mW cm−2. Chen and Ray  correlated 4-CP degradation rate constant (k) with light intensity (I): k ∝ I0.84 indicating that the degradation was independent of mass transfer limitation.
3.4.3. Effect of initial concentration of phenols
As the effect of phenol concentration is of importance in the process of treatment of phenolic wastewater, it is necessary to investigate its dependence. Different concentration profiles can be seen during phenol degradation at different initial concentrations. The degradation rates at same concentration with different initial concentrations are different. However, all the concentration profiles could be correlated with an exponential function as follows :
3.4.4. Effect of photocatalyst concentration
Photocatalyst concentration is a crucial parameter that has been widely studied for photocatalytic processes. The optimum photocatalyst concentrations usually vary between 0.15 g l−1 and 8 g l−1 for different photocatalyst systems and photoreactors. A large difference in optimum photocatalyst concentration (0.15–2.5 g l−1) was reported even for the same photocatalyst (DP25). Chen and Ray  expressed the photocatalytic degradation rate as follows:
3.4.5. Effect of electron acceptor
Photocatalytic degradation reaction requires the use of electron acceptor to reduce the charge carrier recombination. Oxygen is the most common electron acceptor because of its availability, higher solubility, and nontoxic nature. The partial pressure of oxygen is adjusted by mixing the oxygen stream with nitrogen stream by maintaining the total flow rate of gas at a constant value. The photocatalytic reaction of phenols will terminate if sufficient oxygen is not available in the solution . Chen and Ray  showed the improvement of 4-NP photodegradation rate with increasing oxygen partial pressure. The photodegradation rate constant reached approximately 70% of its maximum value at oxygen partial pressure of 0.2 atm. The effect of oxygen partial pressure on the photodegradation of 4-NP is described by a noncompetitive Langmuir kinetic equation as follows:
4. Dye-sensitized photocatalytic degradation of phenol and phenolic compounds
4.1. Theory of dye sensitization
The process of expanding the spectral sensitivity of semiconductor materials with a dye to the visible spectra is known as dye sensitization. Dye is typically adsorbed onto the semiconductor surface by chemical adsorption process. Chemisorbed dye molecules act as spectral sensitizer that upon excitation with visible light inject an electron into the conduction band of the semiconductor . To undergo successful electron injection, the dye molecule should include few basic properties regarding energy levels, ground-state redox potential, and surface anchoring group. Carboxylic and phosphoric acid groups form strong covalent bonds with semiconductor and provide fast electron transfer rate . Recent studies mention the use of a group of sensitizers such as poly(aniline), poly(thiophene), porphyrins, coumarin, phthalocyanines, eosin Y, alizarin red S, and carboxylate derivatives of anthracene [11, 26, 27]. Among the photosensitizers, transition metal-based sensitizers have shown best results in dye-sensitization process. Transition metals such as Fe (II), Ru (II), and Os (II) form d6 complex and undergo intense charge transfer absorption across the entire visible range . However, metal-based sensitizers are not environment friendly, and thus researchers are now focusing on the use of natural dyes as an alternative for dye-sensitization process [59–62]. Several semiconductor photocatalysts have been studied for dye sensitization including TiO2, SrTiO3, ZnO, SnO2, and Cu2O [63, 64]. Among them, TiO2 is the best photocatalyst in terms of (i) cost, (ii) availability, (iii) toxicity, (iv) stability against photocorrosion, and (v) electronic energy band structure [65, 66].
4.2. Dye-sensitized photocatalytic phenol degradation mechanism
Dye-sensitized photodegradation of phenol under visible light began through excitation of the dye molecule from its ground state to the excited state, which then assists the electron transfer to the conduction band of the semiconductor. The oxidized dye molecule (dye+) can interact with either phenol or an electron donor to return back to its ground state . Chowdhury et al.  used eosin Y (EY) as the sensitizer which provided TiO2/Pt a significant visible-light activity via dye sensitization. Eosin Y contains both hydroxyl and carboxyl end groups, which actually assists the dissociative surface adsorption of eosin Y onto the surface hydroxyl (Ti–OH2+) sites of EY-sensitized TiO2/Pt . Triethanolamine (TEOA) was used as an electron donor, which was consumed through an irreversible oxidation by extending the lifetime of eosin Y during phenol degradation (Eqs. (11) and (12)) .
(acid-base equilibrium of TEOA)
4.3. Dye-sensitized photocatalytic phenol degradation kinetics
In dye-sensitized photodegradation under the visible light, the dye molecule is first activated by visible light (λ > 420 nm) and then injects electrons into the conduction band of the semiconductor. Chowdhury et al.  described the kinetics of phenol degradation using eosin Y-sensitized TiO2/Pt with a modified Langmuir-Hinshelwood equation as follows:
Eq. (23) was used to predict the kinetic parameters of phenol photodegradation at different irradiation intensities (range, 25–100 mW cm−2). Based on a parameter estimation using the experimental data, the values of
4.4. Application of dye-sensitized photocatalyst for the degradation of phenols
Dye-sensitized photodegradation of phenol and phenolic compounds showed promising results under visible-light irradiation [7, 11, 26, 69–71]. Vinu et al.  reported degradation of phenol, 4-chlorophenol, 2,4-dichlorophenol, and 2,4,6-trichlorophenol using eosin Y and fluorescein-sensitized combustion-synthesized nano TiO2 under visible light. Eosin Y-sensitized photocatalyst showed better performance than fluorescein-sensitized photocatalyst. Phenol degradation rate was slowest among the phenolic compounds. Chowdhury et al  used eosin Y-sensitized TiO2/Pt for the degradation of phenol under the visible solar light in the presence of triethanolamine as an electron donor. About 93% phenol degradation (initial concentration 40 ppm) was achieved within 90 min using eosin Y-sensitized TiO2/Pt photocatalyst under optimum experimental conditions (pH 7.0, photocatalyst concentration of 0.8 g L−1, triethanolamine concentration of 0.2 M, 0.5% Pt loading on TiO2, visible solar light intensity of 100 mW cm2). Mele et al.  studied photocatalytic degradation of 4-nitrophenol with polycrystalline TiO2 impregnated with functionalized Cu(II)-porphyrin or Cu(II)-phthalocyanine. Cu(II)-based sensitizers provided better results for the degradation of 4-nitrophenol in comparison with metal-free sensitizers. Iliev  studied photooxidation of phenol with phthalocyanine-modified TiO2 and Al2O3 under visible light. The degree of photodegradation of phenol in the presence of phthalocyanine-modified TiO2 is much higher than phthalocyanine-modified Al2O3. Superoxide ion was considered as the active species during photodegradation. Grandos et al.  used Co(II) and Zn(II) tetracarboxyphthalocyanine (TcPcM)-sensitized TiO2 for the degradation of phenol under visible light. The photodegradation efficiencies were reported to be 4.3 and 3.3% for TcPcCo/TiO2 and TcPcZn/TiO2, respectively. Ghosh et al.  demonstrated the photocatalytic degradation of 4-chlorophenol with coumarin-sensitized TiO2 under visible LED light. The degradation rate followed a first-order kinetics and moved toward a limiting value at a photocatalyst concentration of 0.3 g L−1.
Phenol and phenolic compounds are considered as priority pollutants by US EPA because of their high toxicity. They impose severe short-term and long-term health problems to human beings. In this review, we discussed different phenol degradation methods such as physical treatments, biological treatments, and advanced oxidation processes (AOPs). AOPs provide much faster degradation rate than conventional treatment methods and undergo complete mineralization instead of transferring the pollutants from one phase to another.
Heterogeneous photocatalysis is such an AOP that limits the use of oxidizing chemicals (e.g., ozone and hydrogen peroxide) and only utilizes light (UV or solar) and photocatalyst to generate hydroxyl radicals (HO•). Several photoreactors are used with UV lamp, namely, swirl-flow reactor, Taylor vortex reactor, and two-phase monolithic-type reactor for the photocatalytic degradation of phenolic compounds. In UV-light-driven photocatalysis, hydroxyl radicals are the active species which react with either phenol or phenolic compounds until complete mineralization. Different parameters such as solution pH, light intensity, initial concentration of target compounds, photocatalyst concentration, and electron acceptors play a significant role on photocatalytic degradation of phenol and phenolic compounds. However, our economic assessment indicates that the use of UV light significantly increases the overall process cost.
Visible-light-active photocatalysts are developed to utilize the most abundant sunlight to make the photocatalysis economically feasible. Compound parabolic collectors (CPCs) are commonly used for solar photocatalytic degradation of phenol and phenolic compounds. Photocatalysts are modified via doping, dye sensitization, and coupling method to expand the photoresponse to the visible region. Among these, dye-sensitized photocatalysis is shown to be an efficient method for phenol degradation under the visible solar light. The process involves electron transfer to the conduction band of semiconductor initiated by dye sensitization under the visible solar light. Dye-sensitized photocatalysis processes are shown to be Efficient methods for the degradation of phenol and phenolic compounds under the visible solar light.