Skip to main content
NCBI home page
ACS Omega logoLink to ACS Omega
. 2020 Nov 10;5(46):29642–29656. doi: 10.1021/acsomega.0c00628

Polyaniline-Coated TiO2 Nanorods for Photocatalytic Degradation of Bisphenol A in Water

Shepherd Sundayi Sambaza 1, Arjun Maity 1, Kriveshini Pillay 1,*
PMCID: PMC7689664  PMID: 33251400

Abstract

graphic file with name ao0c00628_0020.jpg

Polyaniline (PANI)-wrapped TiO2 nanorods (PANI/TiO2), obtained through the oxidative polymerization of aniline at the surface of hydrothermally presynthesized TiO2 nanorods, were evaluated as photocatalysts for the degradation of Bisphenol A (BPA). Fourier-transform infrared spectroscopy analysis revealed the successful incorporation of PANI into TiO2 by the appearance of peaks at 1577 and 1502 cm–1 that are due to the C=C and C–N stretch of the benzenoid or quinoid ring in PANI. Brunauer–Emmett–Teller analysis revealed that PANI/TiO2 had almost double the surface area of TiO2 (44.8999 m2/g vs 28.2179 m2/g). Transmission electron microscopy (TEM) analysis showed that TiO2 nanorods with different diameters were synthesized. The TEM analysis showed that a thin layer of PANI wrapped the TiO2 nanorods. X-ray photon spectroscopy survey scan of the PANI/TiO2 nanocomposite revealed the presence of C, O, Ti, and N. Photocatalytic activity evaluation under UV radiation through the effect of key parameters, including pH, contact time, dosage, and initial concentration of BPA, was carried out in batch studies. Within 80 min, 99.7% of 5 ppm BPA was attained using the 0.2 g/L PANI/TiO2 photocatalyst at pH 10. The quantum yield (QY) of these photocatalysts was evaluated to be 9.86 × 10–5 molecules/photon and 2.82 × 10–5 molecules/photon for PANI/TiO2 and TiO2, respectively. PANI/TiO2 showed better performance than as-synthesized TiO2 with a rate constant of 4.46 × 10–2 min–1 compared to 2.18 × 10–2 min–1. The rate of degradation of PANI/TiO2 was also superior to that of TiO2 (150 mmol/g/h vs 74.89 mmol/g/h). Nitrate ions increased the rate of degradation of BPA, while humic acid consistently inhibited the degradation of BPA. LC–MS analysis identified degradation products with m/z 213.1, 135.1, and 93.1. The PANI/TiO2 nanocomposite was reused up to five cycles with a removal of at least 80% in the fifth cycle. LC–MS results revealed three possible BPA degradation intermediates. LC–MS analysis identified degradation products which included protonated BPA, [C14H13O2+], and [C9H11O+]. The PANI/TiO2 nanocomposite demonstrated superior photocatalytic activity with respect to improved QY and figure of merit and lower energy consumption.

1. Introduction

Water is regarded as the most vital of natural resources worldwide for the sustainability of life, yet freshwater systems are directly threatened by human activities. Currently, the world is facing a serious challenge in saving the few available sources of “clean” water.1 Rapid industrialization has led to high discharge of organic pollutants into the environment, causing serious damage to aquatic life and human well-being.2,3 Among these pollutants that are well-reported for their negative impact on water quality are a unique class of compounds referred to as emerging organic pollutants. Bisphenol A (BPA) is one of these emerging organic pollutants and is a synthetic organic compound, widely used in food packaging and production of pesticides, flame retardants, polycarbonate, and epoxy resins.4−8 BPA has been reported to act as an endocrine disruptor.4−8 It enters the aquatic system during the manufacturing process or through leaching from the products.9−11 BPA has a high water solubility (120 mg/L at 25 °C) and KOW ∼ 3.4, which is considered as a moderate potential for bioaccumulation.12 It acts by binding to the estrogen receptor or membrane-bound estrogen receptor in the cellular nucleus and influences cellular activities.9,10 In addition, BPA has been shown to affect metabolism through influencing enzyme activity in animal tissues.7

To date, various methods have been reported for the removal of BPA from wastewater. These include adsorption, biological treatment, chemical oxidation, and advanced oxidation processes.13−16 However, some of these methods have proven to be inefficient or are prone to generating a large amount of sludge.17,18 Advanced oxidation processes, on the other hand, have been consistently identified as effective methodologies for the removal of persistent organic pollutants from an aqueous environment. Among the advanced oxidation processes, the photodegradation technique, which is based on the ability of semiconductor materials to absorb a photon of energy equal to or greater than its band gap energy and generate hydroxyl (•OH) and superoxide (•O2–) radicals, has been consistently used to degrade organic contaminants.19 Photodegradation offers the advantages of high efficiency, low toxicity, simplicity in design, and ease of operation.20,21

Several catalysts have been reported in the literature for the photodegradation of BPA.22−25 Taking into consideration the TiO2 properties, including strong oxidizing power, large surface area, corrosion resistance, nontoxicity, and cost effectiveness, this semiconductor has been widely used as a photocatalyst for the degradation of organic pollutants.26,27 However, its application has been hindered by the high recombination rate of the electron–hole pairs.19 Surface modification of TiO2 with a polyaniline (PANI)-conducting polymer has been reported to alleviate this challenge by effecting charge separation. For example, commercially available irregularly shaped polyhedra P25 deposited by PANI has been found to exhibit an improved photocatalytic performance toward methylene blue because of electrons being gathered in the conduction band of TiO2, while holes are trapped in the highest occupied molecular orbital (HOMO) of PANI.28 On the other hand, spherical black anatase TiO2 nanoparticles, synthesized via a hydrothermal procedure, were deposited onto a PANI matrix, resulting in improved light absorption for photocatalytic degradation of methyl orange.29

The conducting polymer PANI has been utilized as a polymer of choice because of its easy synthesis, relatively low cost, nontoxicity, and modifiable surface chemistry.30 In addition to the stated desirable properties of PANI, we chose it for our study for its added ability to adsorb organic molecules via π–π interactions and hydrogen bonding.30 Furthermore, the incorporation of the TiO2 photocatalyst into a PANI polymer alleviates the problem of agglomeration that may occur if powders are used.31 In addition, we envision an improved photodegradation performance of the nanocomposite because of the large surface area offered by nanorods compared to other forms of TiO2.32

Although there have been numerous reports on the structural properties of TiO2/PANI such as magnetic properties, conductivity, and crystallinity which can impact their application in photocatalysis and wastewater treatment,33−35 there are still no studies reported on PANI-wrapped TiO2 nanorods for the degradation of BPA under UV radiation as per the authors’ knowledge.28,29 It is important to investigate the synergistic effect of enhanced surface area offered by PANI-wrapped nanorods in which PANI acts as the adsorbent for BPA and the TiO2 nanorods degrade BPA, leading to enhanced adsorptive photocatalytic degradation. The advantages of adsorptive photocatalytic degradation cannot be overemphasized. Indeed, the adsorption of the pollutant molecules onto the surface of the catalyst plays an important role in enhancing degradation, and composite materials, which are good adsorbents, are crucial in terms of promoting the synergy between simultaneous adsorption and degradation of pollutant molecules.36,37

When engaging in research on photocatalytic studies pertaining to industrial wastewater, it is also imperative to examine the effects of other interfering species which can coexist in industrial wastewater.36 Till date, the authors are unaware of any such studies which have been conducted for PANI/TiO2-based photocatalysts. Moreover, other important parameters such as quantum yields (QYs), figures of merit (FOMs), and energy consumption, which are important in assessing the photocatalytic efficiency for industrial applications, have not been reported for PANI/TiO2 photocatalysts. In addition, there are no reports with experimental data to explain the mechanism of degradation by PANI-supported TiO2 which was conducted in this study.

Based on the abovementioned consideration, in the present work, PANI-wrapped TiO2 nanorods were prepared for the degradation of BPA under UV radiation. The effect of nitrate and humic acid, which are likely to coexist with BPA in real industrial wastewater samples, is also reported. The QYs, energy consumption, and FOMs have been evaluated and compared with those of other nanocomposites which have been reported for BPA in order to assess the efficiency of this photocatalyst for BPA degradation in industrial applications. A detailed mechanism of degradation is further provided.

2. Materials and Methods

2.1. Reagents

All chemicals used in present research were of AR grade except the ones highlighted. Titanium isopropoxide (TTIP), isopropyl alcohol, acetic acid, sodium hydroxide (NaOH), nitric acid (HNO3), iron(III) chloride (FeCl3), aniline, BPA [2,2-bis(4-hydroxyphenyl) propane, BPA], methanol (HPLC Grade), acetonitrile (HPLC Grade), benzoquinone, potassium iodide, potassium nitrate, ammonium acetate, humic acid, and acetone were acquired from Sigma-Aldrich and used without further purification. Ultrapure water was obtained from a Merck water system (Milli-Q).

2.2. Synthesis of PANI

PANI was synthesized following a method adopted from the literature with slight modification.38 In a typical synthesis, FeCl3 (6 g) was dissolved in 40 mL of distilled water and 0.8 mL of aniline monomer was rapidly added to the oxidant solution under stirring. Distilled water (40 mL) was slowly added to the solution while stirring for about 5 min and then left to stand for 12 h. The green product was washed with acetone to stop further polymerization and then further washed with distilled water at ∼pH 7. The green product was filtered under vacuum and dried in an oven at 60 °C.

2.3. Synthesis of TiO2 Nanoparticles

TiO2 nanoparticles were synthesized using a hydrothermal method. Typically, 5 mL of TTIP was dissolved in a mixture of isopropyl alcohol and water (5:3). The solution was poured into a 100 mL beaker and stirred. Concentrated acetic acid (6 mL) was added to the solution while stirring. The solution was stirred for 3 h and then heated to 180 °C in an autoclave for 24 h. The solution was then cooled at room temperature and washed with excess ethanol to yield a white precipitate. The precipitate was washed with distilled water until ∼pH 7. The precipitate was then dried in an oven at 60 °C and calcined at 500 °C in a muffle furnace in air, producing TiO2.

2.4. Synthesis of TiO2 Nanorods

TiO2 nanoparticles synthesized following the method described earlier in Section 2.4 were then dispersed in 10 M NaOH solution and stirred until homogeneity was achieved. The mixture was placed in a Teflon-lined autoclave and then heated to 180 °C for 48 h. The solution was then cooled at room temperature and then washed with dilute HNO3 and then with distilled water until ∼pH 7. The precipitate was then dried in an oven at 60 °C and then calcined at 500 °C in a muffle furnace in air, producing TiO2 nanorods.

2.5. Synthesis of PANI/TiO2

The PANI/TiO2 nanocomposite was synthesized as follows: varying amounts of TiO2 nanorods were dispersed in 20 mL of distilled water and stirred continuously. FeCl3 (6 g) was added to the mixture and 0.2 mL of aniline monomer was rapidly added to the oxidant solution under stirring. Distilled water (60 mL) was slowly added to the solution under stirring for about 5 min and then left to stand for 12 h. The green product was washed with acetone to stop further polymerization and then further washed with distilled water at ∼pH 7. This product was filtered under vacuum and dried in an oven at 60 °C.

2.6. Characterization of the Photocatalysts

The morphology of TiO2 nanorods and the PANI/TiO2 nanocomposite was analyzed using a VEGA3 TESCAN (Bruno, Czech Republic), scanning electron microscopy (SEM), and a JEOL JEM-2100 transmission electron microscope (Peabody, Massachusetts, USA) The structural properties were analyzed by X-ray diffraction (XRD) (Philips, X’Pert PRO MPD, mineral powder diffraction) analysis, which was carried to investigate the crystallinity of the nanocomposite. The instrument uses an X-ray source of Cu Kα-radiation beam with an excitation wavelength of 0.15406 nm for the analysis. The functional groups of the materials were analyzed by Fourier transform infrared (FTIR) spectra that were recorded from 400 to 4000 cm–1 using KBr pellets.

Surface area and porosity of the TiO2 and PANI/TiO2 nanocomposite were analyzed using a Brunauer–Emmett–Teller (BET) surface area analyzer (ASAP 2020, Micromeritics Instruments, USA) using N2 adsorption and desorption isotherms. Elemental composition, states, and surface characterization were analyzed using a Physical Electronics (PHI) Quantum 200 XPS spectrophotometer with Al Kα as the excitation source. Photoluminescence measurements were performed using a PerkinElmer LS 45 fluorescence spectrometer.

2.7. Photocatalytic Degradation Studies

The photodegradation studies were carried out in a photocatalytic reactor with a capacity of 1000 mL. A glass column housing a UV lamp was fixed inside the reactor with tap water circulating around the glass column, cooling the solution. The wall of the glass column has a thickness of 3 mm. The reactor was made of Perspex, and during reactions, it was placed on a magnetic stirrer to keep the solution under agitation. The mixture of the catalyst and BPA solution was left in the dark under magnetic stirring for an hour until adsorption–desorption equilibrium was established.39 The mixture was irradiated with a UV lamp light at a predominant wavelength of 365 nm. The photodegradation performance was investigated as a function of time. In the experiments, the initial solution pH was adjusted to pH 10 from the solution pH of 6.5 with 0.1 M HCl or 0.1 M NaOH for experiments investigating the influence of pH. The suspension was continuously stirred in the dark for 30 min using a magnetic stirrer to ensure the adsorption equilibrium of BPA on the catalyst. Aliquots (10 mL) were removed from the sample at varied time intervals and filtered through 0.45 μm membrane filter discs for analysis by an Agilent high-performance liquid chromatograph () using acetonitrile/water (40/60, v/v) as the mobile phase, an elution time of 15 min, and a detector wavelength of 275 nm.

A C18 reversed-phase column (4.6 nm × 150 nm) was used for chromatographic analysis. Reproducibility of the measurements was ensured by performing them three times. The effect of pH on the photodegradation was examined by keeping 0.1 g of the photocatalyst in contact with 500 mL of BPA solutions of 5 ppm concentrations, set at pH 4.5, 7, and 10 and at a temperature of 25 °C. Similarly, the effect of photocatalyst dosage was investigated for the photodegradation of BPA by varying the mass of the photocatalyst from 10 to 250 mg in 500 mL of 5 ppm BPA solution. The effect of initial BPA concentration was investigated at different initial BPA concentrations from 1 to 30 ppm. The presented data points are mean values. The removal efficiency of BPA was calculated using eq 1

2.7. 1

where Ci (mg/L) is the initial concentration of the adsorbate in solution and Cf (mg/L) is the final concentration of the adsorbate in the filtrate.

In order to test the validity and reproducibility of the analytical data, the photodegradation experiments were conducted in triplicate and an average of three results is reported. The relative standard deviation was found to be within ±0.05%.

2.7.1. LC–MS Conditions

The analysis was conducted using a Shimadzu UHPLC-MS/MS 8030 instrument (Tokyo, Japan). The chromatographic separation was performed using a RAPTOR AR C18, 2.7 μm 100 × 2.1 mm column. The column was previously conditioned with MeCN: (5:95 v/v) at a flow rate of 0.1 mL/min. The column temperature was set at 40 °C. BPA solutions were separated following a gradient elution program shown in Table 1. To prevent sample carry over, 100% MeOH was used as blank and was injected after every standard and sample injection. A flow rate of 0.2 mL/min was used for the elution of BPA in UHPLC. For tandem mass spectrometry, the deprotonated molecule [M + H]+ was used as the precursor ion. Tandem mass spectra were acquired between m/z 30 and m/z 300 in the enhanced mass resolution mode Q3. A nebulizing gas flow rate of 3 L/min was used. A drying gas flow rate of 15 L/min was used for the mass spectrometry.

Table 1. LC Gradient Elution Program for BPAa.
  time module command value
1 0.01 pumps solvent B conc. 25
2 0.50 pumps solvent B conc. 25
3 3.50 pumps solvent B conc. 95
4 4.00 pumps solvent B conc. 95
5 4.10 pumps solvent B conc. 25
6 10.00 pumps solvent B conc. 25
7 10.10 controller stop  
Open in a new tab
a

Solvent A: water. Solvent B: acetonitrile.

3. Results and Discussion

3.1. FTIR Analysis of TiO2, PANI, and PANI/TiO2

FTIR spectroscopy analysis revealed the incorporation of PANI into a nanocomposite of PANI/TiO2 as shown by the functional groups of PANI that are present in the PANI/TiO2 nanocomposite. The main peaks of PANI were all confirmed in the spectra (b) with peaks appearing at 3662, 2955, 1577, 1502, 1301, 1250—, and 827 cm–1.40−42 The band at 3662 cm–1 corresponds to the N–H stretching mode and the peak around 2955 cm–1 corresponds to the aromatic C–H vibration.42 The peaks appearing at 1577 and 1502 cm–1 are due to the C=C and C=N stretch of the benzoid and quinoid, respectively (Figure 1). The peak appearing at 1301 cm–1 is due to the C–N vibrations.42 The peaks due to PANI were clearly observed in the PANI/TiO2 sample. The spectrum displays the major peaks of PANI such as the peak at 3662 cm–1 which is due to the N–H stretching and other peaks due to the benzoid and quinoid rings which appear at 1577 and 1502 cm–1.40 The new weak peak appearing at 917 cm–1 is probably due to the C–H out-of-plane bending vibration of the PANI composite.39 These bands reveal the existence of PANI in the synthesized PANI/TiO2 sample. At 700 cm–1, a peak assigned to Ti–O–Ti bridging stretch modes was observed on the TiO2 spectra.43

Figure 1.

Figure 1

Open in a new tab

FTIR spectra of pristine TiO2, pristine PANI, and PANI/TiO2.

3.2. XRD Analysis of Pristine PANI, TiO2, and PANI/TiO2

The results for XRD analysis are shown in Figure 2. XRD analysis of as-synthesized PANI showed two broad peaks at 2θ = 20.5 and 25.3°. This suggested an amorphous structure of the PANI and the results are in agreement with results that were obtained by Haldorai et al.44 The XRD pattern of as-synthesized TiO2 nanoparticles showed peaks at 2θ values of 25.30, 37.90, 47.99, 54.2, 62.72, 68.84, 70.28, 75.12, and 82.79°.42,45 These peaks correspond to anatase TiO2 with hkl crystal planes [101], [004], [200], [105], [211], [204], [116], [220], [215], and [303], respectively.46 This pattern confirms the formation of a pure anatase phase without any other phases with a JCPDS score of 97%. The observed 2θ values were consistent with the standard values and showed the anatase phase of TiO2 as per Joint Committee on Powder Diffraction Standard (JCPDS) card no. 01-086-1157. The peaks in as-synthesized TiO2 also appeared in the PANI/TiO2 nanocomposite showing that the TiO2 did not lose its crystallinity even after incorporation of PANI.

Figure 2.

Figure 2

Open in a new tab

XRD plot of (A) pristine PANI, (B) pristine TiO2, and (C) PANI/TiO2.

XRD analysis can be used to determine the crystallite size of the powdered material by treating the XRD data using the Scherrer equation.47−49 The particle size of the material has been reported to affect the overall performance of the photocatalyst.50 Small particle sizes have been reported to provide a high surface area to volume ratio, which improves the photocatalytic efficiency by improving the contact between the photocatalysts and the pollutant.50,51 In this study, the particle sizes were determined using the Scherrer equation48

3.2. 2

where D is the average size in nanometers, k = 0.89, λ = 0.1540 nm (Cu Kα), β = fwhm × 0.01745 in radians, and θ is the peak position.

The results show an average particle size of 16.62 nm. The same peak positions were also observed in a nanocomposite of PANI/TiO2 with a calculated average particle size of 16.6 nm.

3.3. BET Analysis of Pristine PANI, TiO2, and PANI/TiO2

The pore size distribution and specific surface area were determined via nitrogen adsorption–desorption isotherms. Figure 3 illustrates that all the as-synthesized photocatalysts can be identified as a Type V isotherm with the H4-type hysteresis loop according to the IUPAC–BET classification.52 This signifies the presence of the mesoporous structure in the photocatalysts.52Table 2 summarizes the BET specific surface area, pore volume, and pore size of all as-prepared PANI, TiO2, and PANI/TiO2 photocatalysts. The BET data showed that the addition of PANI to the calcined TiO2 resulted in an increase in the surface area from 28.2179 m2/g of TiO2 to 44.8999 m2/g of PANI/TiO2. A similar trend was observed in the pore volume and pore size because of the incorporation of PANI into the PANI/TiO2 nanocomposite. The increase in these important parameters may result in high degradation performance of PANI/TiO2.

Figure 3.

Figure 3

Open in a new tab

BET surface area of [A] pristine TiO2 nanorods, [B] PANI/TiO2, and [C] TiO2–/PANI.

Table 2. BET Specific Surface Area, Pore Volume, and Pore Size of as-prepared PANI, TiO2, and PANI/TiO2.

photocatalyst SBET (m2/g) pole volume (cm3/g) pore size (nm)
PANI 4.3093 0.033007 30.64
TiO2 28.2179 0.097520 13.82
PANI/TiO2 44.8999 0.145248 35.62
Open in a new tab

3.4. SEM Analysis of (A) TiO2 Nanorods and (B) PANI/TiO2

Figure 4 shows the rod-like TiO2 photocatalyst (image A). Image B in Figure 4 shows denser rod-shaped materials of different sizes.

Figure 4.

Figure 4

Open in a new tab

SEM images of pristine TiO2 (A) and PANI/TiO2 (B).

3.5. TEM Analysis of (A) TiO2 Nanorods and (B) SAED Analysis TiO2 Nanorods

Transmission electron microscopy (TEM) analysis confirmed that nanorods of TiO2 were synthesized, as shown in Figure 5a,b. These uniformly dispersed TiO2 nanorods had a diameter size range of 40–100 nm. TEM analysis revealed that the TiO2 nanorods were wrapped by a thin layer of PANI forming in the nanocomposite of PANI/TiO2 (Figure 5c,d). EDX elemental analysis in Figure 5e showed that the nanocomposite was mainly composed of carbon (C), oxygen (O), and titanium (Ti) elements. The chloride ion (Cl–) species and Fe came from the preparation of the PANI/TiO2 nanocomposite. Copper (Cu) atoms shown were due to copper grids used during TEM analysis.

Figure 5.

Figure 5

Open in a new tab

TEM images of (a,b) pristine TiO2 (c,d) PANI/TiO2, (e) EDS of PANI/TiO2, and (f) diameter of the TiO2 nanorods.

3.6. TGA of Pristine PANI, TiO2 Nanorods, and PANI/TiO2

Thermogravimetric analysis (TGA) was employed to investigate the thermal stability of the TiO2 and PANI/TiO2 nanocomposite. TGA profiles of the TiO2 and PANI/TiO2 nanocomposite are shown in Figure 6. TGA results for PANI show four major weight losses. The weight loss in the first stage, from 50 to 150 °C for the peak of the pristine TiO2 and PANI/TiO2 nanocomposite, is due to the loss of adsorbed and bound water on the nanocomposite.53 The weight loss at 150–250 °C was due to loss of water of crystallization.54 The weight loss that occurred after 380–500 °C was due to the removal of functional groups and breaking of bonds forming the PANI polymers and also due to loss of gaseous emissions, for example, carbon monoxide, which might have occurred.54 The final weight loss at 800 °C was due to the thermal degradation of the carbonaceous material.55 TGA for the PANI/TiO2 nanocomposite is shown in Figure 7. The results indicate three major weight losses. The first weight loss at 25–150 °C was due to loss of moisture.56 Another weight loss observed from 180 to 700 °C was due to loss of low molecular weight oligomers of PANI.26 After 700 °C, there was no weight loss observed for the nanocomposite because of the stability of TiO2, which was incorporated into PANI to form a nanocomposite. This is in agreement with observations made on pristine TiO2.57

Figure 6.

Figure 6

Open in a new tab

TGA profile of pristine PANI, TiO2, and PANI/TiO2.

Figure 7.

Figure 7

Open in a new tab

(a) Luminescence spectra of pristine TiO2, pristine PANI, and PANI/TiO2. (b) UV absorption spectra of pristine TiO2, pristine PANI, and PANI/TiO2 and Tauc’s plot of pristine TiO2 and PANI/TiO2.

3.7. Photoluminescence and UV Absorption Analysis

Photoluminescence (PL) and UV spectroscopy provide information on the charge separation of materials.55,58,59 PL reveals the migration of charges (electrons and holes) and the formation of defects in the photocatalyst.55 This helps in understanding the recombination rate of charges at the surface of the photocatalyst. Figure 7a shows results of emission of the materials at 530 nm for all sample materials. The intensity of pristine TiO2 spectra is higher than that of pristine PANI and that of PANI/TiO2. The results show that the incorporation of PANI greatly improved the performance of the nanocomposite by reducing the recombination rate of the nanocomposite. This is attributed to the ability of PANI to affect charge separation by transferring electrons to the conduction band of TiO2 and also trapped holes, leading to charge separation.39Figure 7b shows results of UV and DRS analysis of the photocatalysts. The results showed that PANI has a stronger absorption than either the PANI/TiO2 nanocomposite or pristine TiO2. The wrapping of TiO2 nanorods with PANI improved the absorption properties of the PANI/TiO2 nanocomposite. This was also reflected in Tauc’s plot with the band gap estimation of PANI/TiO2 at 3.1 eV compared to that of pristine TiO2, which was at 3.2 eV.

3.8. X-ray Photoelectron Spectroscopy

The surface characterization and chemical states of the PANI/TiO2 nanocomposite were examined by X-ray photoelectron spectroscopy (XPS). Figure 8a represents the XPS survey scan of the PANI/TiO2 nanocomposite, which revealed the presence of C, O, Ti, and N. The XPS results shown in Figure 8b reveal a peak at 284.75 eV, which was assigned to carbon from the PANI polymer.60 A small peak at 288.1 eV corresponding to C–O–Ti was observed.61 The peak in Figure 8c at 399.3 eV due to the N 1s peak corroborates positively with quinoid amine and benzenoid amine, which is consistent with the structure of PANI as reported by Golczak et al.(62)Figure 8d presents the core level O 1s spectra (530 eV) due to O bonds such as Ti–O–Ti and Ti–O–H.60 Ti 2p core level spectra are presented in Figure 8e at 459.2 eV and 464.9 eV due to Ti 2p1/2 and Ti 2p3/2, respectively.61,63

Figure 8.

Figure 8

Open in a new tab

(a) XPS survey scan for TiO2/PANI. (b) Core level XPS spectra for (b) C 1s, (c) N 1s (d) O 1s, and (e) Ti 2p of TiO2/PANI.

3.9. Application Results

3.9.1. Selection of Photocatalysts

The photocatalytic action of pristine TiO2 and PANI/TiO2 was evaluated using a photocatalyst dosage of 0.2 g/L and 5 mg/L BPA solution at solution pH 6.5. Figure 9 shows the photocatalytic degradation of BPA under UV radiation. It is clearly observed that the photocatalytic degradation removal of PANI/TiO2 was higher than that of pristine TiO2. After 80 min irradiation, 79.8% degradation of BPA was achieved by PANI/TiO2 compared to a removal efficiency of 60.7% that was observed for pristine TiO2. The enhanced photoresponse of PANI/TiO2 can be attributed to the modification of TiO2 with PANI. This is in agreement with the observed results from BET analysis that showed an increase in the surface area of the nanocomposite after modification of TiO2 with PANI (Figure 3). The contribution of PANI to the overall degradation efficiency is further highlighted by the increased removal efficiency of BPA in graph (b) compared to graph (a) under darkness. The removal in graph (b) can be attributed to adsorption via π–π stacking and hydrogen bonding of BPA and PANI.43 The slight removal that was observed on graph (a) can be attributed to photolysis. PANI/TiO2 was chosen for the rest of the experiments.

Figure 9.

Figure 9

Open in a new tab

Degradation of BPA on (a) pristine TiO2 in the dark and (b) PANI/TiO2 in the dark. Photocatalytic degradation on (c) pristine TiO2 and (d) PANI/TiO2.

3.9.2. Effect of pH of BPA Solution on Its Photodegradation

Solution pH plays an important role in determining the efficiency of the photodegradation process. This is because it affects the adsorption–desorption of the organic molecules on the surface of the photocatalyst. In addition, pH influences the generation of hydroxyl radicals and active oxygen species that occur on the surface of the photocatalyst.38Figure 10 shows a comparison of the degradation of BPA (5 ppm) with a dosage of 0.2 g/L at different initial BPA solution pH’s ranging from 4 to 10. The results showed that the degradation of BPA depended strongly on the initial solution pH. The reported point of zero charge of TiO2 from the literature is 6.25 while that of PANI is 5.8.64,65 This implies that the surface of the PANI/TiO2 was negatively charged in an alkaline solution and positively charged in an acidic solution. On the other hand, BPA had a negative charge because of the hydroxyl molecules in its structure.66 These surface properties influenced the initial adsorption of BPA on the catalysts. The results in Figure 10 show that the BPA removal was 99.8% at pH 10 compared to 79.6% at pH 7 and the lowest removal was at pH 4 (50%). The high removal at pH 10 was due to an increase in diffused hydroxyl radicals in the bulk solution that degraded BPA.41,42,67 Therefore, a pH of 10 was used as the optimum pH for the rest of the experiments.

Figure 10.

Figure 10

Open in a new tab

Effects of initial pH on the degradation of BPA using PANI/TiO2 under UV light.

3.9.3. Effect of Photocatalyst Dosage on the Photodegradation of BPA

The effect of dosage on the photodegradation of BPA was determined in order to optimize the efficiency of the photodegradation process. Figure 11 shows a comparison of the degradation of BPA at different initial photocatalyst dosages. The results show that photodegradation efficiency improved with an increase in the dosage. The increase in dosage increased the number of •OH radicals, which participated in the degradation of BPA.41 The results shown in Figure 11 show that the highest degradation of 99.9% was attained within 80 min for dosages of 250 mg. However, despite achieving 94.5% degradation efficiency after the same irradiation time, 100 mg dosage in 500 mL of BPA was chosen as the optimum dosage for the rest of the study. A high dosage may have reduced the degradation efficiency because of aggregation of the photocatalyst which may have reduced the surface area of the photocatalyst for light absorption.14 The BPA solution was degraded up to 40, 50, 59.6, and 80.5% by 10, 20, 50, and 75 mg photocatalyst dosages after 80 min of irradiation, respectively.

Figure 11.

Figure 11

Open in a new tab

Effect of photocatalyst dosage on the photodegradation of BPA by PANI/TiO2.

3.9.4. Effect of Initial Concentration of BPA on Photodegradation

The effect of initial BPA concentration on its photodegradation was investigated over a concentration range of 1–30 ppm under an optimal pH of 10 and a photocatalyst dosage of 0.2 g/L. The photodegradation efficiency was influenced by the initial BPA concentration, as shown in Figure 12. Complete degradation of 1 ppm BPA solution was achieved within 80 min. 99.8% degradation was attained for 5 ppm BPA solution within the same time and only 45.4% was achieved when degrading 30 ppm BPA. After 80 min of irradiation, 68%, 75.8%, and 85% degradation was achieved for 10 ppm, 15 ppm, and 20 ppm BPA solutions, respectively. The degradation efficiency decreases with increase in the concentration of BPA. This can be attributed to the increase in the BPA molecules that were adsorbed on the surface of the photocatalyst.39 This resulted in a reduction in the photons that were able to reach the surface of the photocatalyst surface, and therefore, less hydroxyl radicals were generated resulting in reduced photodegradation efficiency.14

Figure 12.

Figure 12

Open in a new tab

Effect of initial concentration on the photodegradation of BPA by PANI/TiO2.

3.9.5. Kinetic Studies

A kinetic study was conducted on the degradation of BPA by as-synthesized PANI, TiO2, and PANI/TiO2. Table 3 shows the rate constants for the photodegradation of BPA corresponding to the Langmuir–Hinshelwood kinetics, which follows the relationship given below. The rate constant was calculated using the following expression68

3.9.5. 3

where Co and C are the initial and final concentration, respectively, k is the rate constant, and t is the irradiation time. The rate constant for the degradation of BPA is shown in Table 3. The results show that PANI had the lowest rate constant of 7.04 10–4 min–1. Pristine TiO2 had a rate constant of 2.18 × 10–2 min–1. PANI/TiO2 shows a better performance than as-synthesized PANI and TiO2 with a rate constant of 4.46 × 10–2 min–1. The results show a linear agreement with pseudo-first-order kinetics with high R2 values. However, the R2 value for the degradation of BPA by PANI was 0.6895, which was the lowest, suggesting that the removal of BPA did not fit well with the pseudo-first-order kinetic model.

Table 3. Rate Constants for the Degradation of BPA by PANI, TiO2, and PANI/TiO2.
photocatalyst R2 rate constant (K min–1)
PANI 0.6895 7.04 × 10–4
TiO2 0.9688 2.18 × 10–2
PANI/TiO2 0.9589 4.46 × 10–2
Open in a new tab

The rate of degradation was further evaluated as 150 mmol/g/h for PANI/TiO2 and 74.89 mmol/g/h for TiO2. Thus, PANI/TiO2 is almost 50% more efficient at degrading BPA than TiO2 and even shows enhanced rates of degradation when compared to TiO2 nanotubes developed for the degradation of CO2 into CO and CH4.68

3.9.6. Effect of the Nitrate Ion on the Degradation of BPA

Nitrates have been reported to influence the degradation of organic matter in natural water as it is a source of hydroxide radicals.69 However, some reports indicate that NO3– ions also negatively affect the process of degradation of organics in water as they can scavenge reactive oxygen species.70 The effects of the NO3– ion on the degradation of BPA was investigated using environmentally relevant concentrations (0.05, 0.1, 0.5, and 1.0 mM) using 0.2 g/L photocatalyst dosage and 5 mg/L BPA solution with a pH of 6.5. The results in Figure 13 showed that the presence of nitrate ions in the BPA solution greatly influenced the degradation efficiency. The removal efficiency increased with an increase in the concentration of nitrates. The increase in the degradation rate was due to the ability of NO3– ions to induce the production of OH• radicals according to eqs 5–7.69,71

3.9.6. 4
3.9.6. 5
3.9.6. 6
Figure 13.

Figure 13

Open in a new tab

Effect of NO3– ion concentration on the photodegradation of BPA by PANI/TiO2.

3.9.7. Effect of Humic Acid on the Degradation of BPA

Natural organic matter has been reported to have an influence in the fate of environmental organic contaminants. Guetzloff and Rice reported that the presence of humic acid resulted in an increase in the solubility of polycyclic aromatic hydrocarbon (PAH).72 Some reports also indicated that humic acids had no effect on degradation of organic pollutants.73 In this study, the effect of humic acids in solution during degradation of BPA (5 mg/L) was investigated by adding 0, 1, 3, and 5 mg/L humic acid. The results shown in Figure 14 show that the presence of humic acids in solution at different concentrations consistently inhibited the degradation of BPA. The highest inhibition was observed with the addition of 5 mg/L. The inhibition was attributed to the blockage of the active sites by adsorbed humic acid on the PANI/TiO2 photocatalyst.74 Similarly, Li and Hu reported an inhibition of degradation of tetracycline in the presence of humic acid.45 They concluded that humic acid quenched OH• radicals that are critical for the degradation process.

Figure 14.

Figure 14

Open in a new tab

Effect of humic acid concentration on the photodegradation of BPA by PANI/TiO2.

3.9.8. Byproduct Identification

BPA was degraded using PANI/TiO2 and the intermediate products were identified using LC–MS. Samples were collected from the reactor at 0; 20; 60; and 120 min of irradiation. They were then filtered using 0.45 μm filter discs. The high-performance liquid chromatograms for BPA standard and the degradation products are shown in Figure 15a. The retention time (tR) of BPA was shown to be at 4.4 min, which was evidenced by the single chromatogram in the samples for BPA standard and that at 0 min irradiation time. However, after irradiation, the peak due to the parent BPA reduced in size and completely disappeared with irradiation. Subsequently, new peaks appeared in chromatograms with tR at 1.0; 1.7; and 3.0 that were attributed to degradation products. A peak with retention time ∼0.7 was attributed to the degradation of the PANI polymer. However, there is need for other techniques to assist with its identification.

Figure 15.

Figure 15

Open in a new tab

(a) Chromatograms of BPA standard and samples after 0, 20, 60, and 120 min of irradiation. (b) Mass spectra of BPA standard and samples after 0, 20, 60, and 120 min of irradiation.

Figure 15b shows the full mass spectrum scans of the samples. The results show that BPA could be ionized in the positive mode producing a molecular ion abundance at m/z 229.1, which corresponded to the protonated BPA, [M + H]+. The results show that a fragment of BPA with a m/z of 213.1 was produced after 20 min of irradiation.75 This fragment resulted from a loss of a methyl group, producing a carbocation, [C14H13O2+]. Another peak with m/z 135.1 was produced and this was attributed to [C9H11O+] which was produced as a result of the cleavage of the alkyl-phenyl bond and possible degradation of BPA.76 A peak observed at m/z 93.1 was attributed to the cleavage of two base phenolic molecules of BPA producing a fragment with the formula C6H5O. These degradation products were consistent with those reported in the literature.75−77

From the observed degradation products, a degradation pathway is proposed in Figure 18.

Figure 18.

Figure 18

Open in a new tab

Proposed mechanism for the degradation of BPA using PANI/TiO2.

3.9.9. Proposed Mechanism for BPA Degradation

Reactive species such as hydroxyl (OH•), superoxide (•O2–), and holes (h+) are responsible for the degradation process (Figure 16).78 In a study, Brooms et al. suggested a mechanism for the degradation of p-cresol by TiO2/ZnO/PANI.39 The authors suggested that PANI played an active role in the degradation of p-cresol by TiO2/ZnO/PANI by transferring electrons to the conduction band of TiO2/ZnO and also trapped holes, leading to charge separation.39 In another study, Masid et al. suggested the same mechanism when they used PANI/TiO2 to degrade a dye, reactive blue 4.79 However, there is no experimental data to support the suggested degradation mechanisms. In this study, we provide experimental data showing which reactive oxygen species were responsible for the degradation. Different scavengers of reactive species were added to the degradation solution, and the degradation performance was evaluated after irradiation under optimized conditions. In this study, potassium iodide (KI) was used as a scavenger for positive holes.80Figure 17 shows that only 30% degradation was achieved after the addition of potassium iodide (KI). This suggests that the positively charged holes were actively involved in the degradation of BPA. In the presence of benzoquinone (BQ) which was used as scavenger for superoxide radicals,81 48% degradation was achieved, which is lower than the 99.8% achieved in the absence of quenchers. This suggests that the superoxide radicals were also involved in the degradation of BPA, although their involvement was not as dominant as that of the positive holes. In the presence of 10% volume isopropyl alcohol (IPA) which was used as a scavenger for hydroxyl radicals,78 a slight decrease in the degradation efficiency with an overall degradation of 70% after 80 min of irradiation under UV light was observed. This suggests that the hydroxyl radicals were not the dominant reactive species in the degradation of BPA by PANI/TiO2.

Figure 16.

Figure 16

Open in a new tab

Proposed degradation pathway for BPA using PANI/TiO2 under UV light.

Figure 17.

Figure 17

Open in a new tab

Photodegradation of 5 ppm BPA by PANI/TiO2 after addition of different reactive species scavengers under UV irradiation.

The mechanism for the degradation of BPA was therefore proposed to occur, as shown in Figure 18. After absorbing UV light irradiation, PANI can release electrons from the HOMO into the lowest energy unoccupied molecular orbital. PANI can transfer the photogenerated electron (e–) into the conduction band of the TiO2. The electrons were then scavenged by oxygen molecules, forming superoxides that are highly oxidizing, and they degraded BPA. Simultaneously, electrons might migrate from the VB of TiO2 to the HOMO of PANI, leaving behind positively charged holes (h+). The positive holes might react with moisture, forming hydroxyl radicals that are also highly oxidizing, and they degraded BPA.

3.9.10. Performance Efficiency of PANI/TiO2

In order to assess the efficiency of PANI/TiO2 as a photocatalyst for the degradation of BPA, the following parameters were evaluated

3.9.10. 7
3.9.10. 8

These parameters are tabulated for comparison with those of other photocatalysts developed for BPA degradation.

Table 4 clearly shows that PANI/TiO2 supersedes most photocatalysts developed for BPA with respect to the QY, energy consumption, and FOM. Moreover, this photocatalyst also surpasses the degradation performance of the rGO/ZrO2/Ag3PO4 developed by Anwar and Park82 for the degradation of 4-nitrophenol and also gave the same FOM. It further shows a higher QY than that developed by Cortes et al.(83) for the degradation of CO2. Hence, PANI/TiO2 can be considered as a highly efficient photocatalyst for industrial applications because of its relatively higher QY and FOM and lower energy consumption.

Table 4. Comparison of Efficiency Parameters for the Degradation of BPA (Anwer and Park, 2018).
photocatalyst light source energy consumption (W h/μmol) photocatalyst dosage (g/L) QY (molecules per photon) FOM
Ag3PO4–graphene vis 14.6 0.5 4.69 × 10–6 69
GO/Bi2Fe4O9 UV 135.0 0.5 7.24 × 10–7 50
GO/Bi2Fe4O9 vis 165.5 0.5 3.65 × 10–7 50
PANI/TiO2 (this study) UV 1.46 0.5 9.86×10–5 100
Open in a new tab

4. Conclusions

PANI-wrapped TiO2 nanorods were successfully synthesized by oxidative polymerization of aniline on hydrothermally presynthesized TiO2 nanorods. The photodegradation of BPA under UV radiation was investigated. The use of nanostructured TiO2 wrapped in PANI improved the photodegradation efficiency of the photocatalyst because of the synergistic effect of adsorption by PANI and photodegradation by TiO2. PANI affected charge separation and reduced the rate of recombination. LC–MS analysis identified degradation products with m/z 213.1, 135.1, and 93.1. The performance metrics showed that PANI/TiO2 had a relatively higher QY and FOM than most developed photocatalysts and lower energy consumption and enhanced rates of degradation and hence is a highly favorable photocatalyst for industrial applications. The degradation was however enhanced in the presence of nitrate ions and inhibited in the presence of humic acid. In summary, we report that the degradation of BPA using the PANI/TiO2 nanocomposite has the potential to help alleviate the threat of water contamination by BPA and it is a highly efficient photocatalyst. However, humic acid must be removed from the wastewater sample matrix, as this inhibits the photocatalytic activity.

Acknowledgments

The authors wish to thank Prof. Philiswa Nomngongo for allowing the use of the HPLC instrument from her laboratory. The authors would also like to thank NRF-Thutuka and the Centre for Nanomaterials Research for funding this study.

The authors declare no competing financial interest.

References

  1. Karaouzas I.; Smeti E.; Vourka A.; Vardakas L.; Mentzafou A.; Tornés E.; Sabater S.; Muñoz I.; Skoulikidis N. T.; Kalogianni E. Assessing the ecological effects of water stress and pollution in a temporary river - Implications for water management. Sci. Total Environ. 2018, 618, 1591–1604. 10.1016/j.scitotenv.2017.09.323. [DOI] [PubMed] [Google Scholar]
  2. Kolpin D. W.; Furlong E. T.; Meyer M. T.; Thurman E. M.; Zaugg S. D.; Barber L. B.; Buxton H. T. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. Streams. 1999–2000:  A national reconnaissance. Environ. Sci. Technol. 2002, 36, 1202–1211. 10.1021/es011055j. [DOI] [PubMed] [Google Scholar]
  3. Paz-Alberto A. M.; Sigua G. C. Phytoremediation: A green technology to remove environmental pollutants. Am. J. Clim. Change 2013, 02, 71. 10.4236/ajcc.2013.21008. [DOI] [Google Scholar]
  4. Erler C.; Novak J. Bisphenol A exposure: human risk and health policy. J. Pediatr. Nurs. 2010, 25, 400–407. 10.1016/j.pedn.2009.05.006. [DOI] [PubMed] [Google Scholar]
  5. Milić N.; Četojević-Simin D.; Milanovic M.; Sudji J.; Milosevic N.; Curic N.; Abenavoli L.; Medic-Stojanoska M. Estimation of in vivo and in vitro exposure to bisphenol A as food contaminant. Food Chem. Toxicol. 2015, 83, 268–274. 10.1016/j.fct.2015.07.003. [DOI] [PubMed] [Google Scholar]
  6. Rajasärkkä J.; Pernica M.; Kuta J.; Lašňák J.; Šimek Z.; Bláha L. Drinking water contaminants from epoxy resin-coated pipes: A field study. Water Res. 2016, 103, 133–140. 10.1016/j.watres.2016.07.027. [DOI] [PubMed] [Google Scholar]
  7. Richter C. A.; Birnbaum L. S.; Farabollini F.; Newbold R. R.; Rubin B. S.; Talsness C. E.; Vandenbergh J. G.; Walser-Kuntz D. R.; vom Saal F. S. In vivo effects of bisphenol A in laboratory rodent studies. Reprod. Toxicol. 2007, 24, 199–224. 10.1016/j.reprotox.2007.06.004. [DOI] [PMC free article] [PubMed] [Google Scholar]
  8. Vandenberg L. N.; Hauser R.; Marcus M.; Olea N.; Welshons W. V. Human exposure to bisphenol A (BPA). Reprod. Toxicol. 2007, 24, 139–177. 10.1016/j.reprotox.2007.07.010. [DOI] [PubMed] [Google Scholar]
  9. Acconcia F.; Pallottini V.; Marino M. Molecular mechanisms of action of BPA. Dose-Response 2015, 13, 1559325815610582. 10.1177/1559325815610582. [DOI] [PMC free article] [PubMed] [Google Scholar]
  10. Bhandari R. K.; Deem S. L.; Holliday D. K.; Jandegian C. M.; Kassotis C. D.; Nagel S. C.; Tillitt D. E.; vom Saal F. S.; Rosenfeld C. S. Effects of the environmental estrogenic contaminants bisphenol A and 17α-ethinyl estradiol on sexual development and adult behaviors in aquatic wildlife species. Gen. Comp. Endocrinol. 2015, 214, 195–219. 10.1016/j.ygcen.2014.09.014. [DOI] [PubMed] [Google Scholar]
  11. Le Fol V.; Aït-Aïssa S.; Sonavane M.; Porcher J.-M.; Balaguer P.; Cravedi J.-P.; Zalko D.; Brion F. In vitro and in vivo estrogenic activity of BPA, BPF and BPS in zebrafish-specific assays. Ecotoxicol. Environ. Saf. 2017, 142, 150–156. 10.1016/j.ecoenv.2017.04.009. [DOI] [PubMed] [Google Scholar]
  12. Repossi A.; Farabegoli F.; Gazzotti T.; Zironi E.; Pagliuca G. Bisphenol A in edible part of seafood. Ital. J. Food Saf. 2016, 5, 5666. 10.4081/ijfs.2016.5666. [DOI] [PMC free article] [PubMed] [Google Scholar]
  13. Koduru J. R.; Lingamdinne L. P.; Singh J.; Choo K.-H. Effective removal of bisphenol A (BPA) from water using a goethite/activated carbon composite. Process Saf. Environ. Prot. 2016, 103, 87–96. 10.1016/j.psep.2016.06.038. [DOI] [Google Scholar]
  14. Malato S.; Fernández-Ibáñez P.; Maldonado M. I.; Blanco J.; Gernjak W. Decontamination and disinfection of water by solar photocatalysis: Recent overview and trends. Catal. Today 2009, 147, 1–59. 10.1016/j.cattod.2009.06.018. [DOI] [Google Scholar]
  15. Yüksel S.; Kabay N.; Yüksel M. Removal of bisphenol A (BPA) from water by various nanofiltration (NF) and reverse osmosis (RO) membranes. J. Hazard. Mater. 2013, 263, 307–310. 10.1016/j.jhazmat.2013.05.020. [DOI] [PubMed] [Google Scholar]
  16. Zielińska M.; Bułkowska K.; Cydzik-Kwiatkowska A.; Bernat K.; Wojnowska-Baryła I. Removal of bisphenol A (BPA) from biologically treated wastewater by microfiltration and nanofiltration. Int. J. Environ. Sci. Technol. 2016, 13, 2239–2248. 10.1007/s13762-016-1056-6. [DOI] [Google Scholar]
  17. Reif R.; Suárez S.; Omil F.; Lema J. M. Fate of pharmaceuticals and cosmetic ingredients during the operation of a MBR treating sewage. Desalination 2008, 221, 511–517. 10.1016/j.desal.2007.01.111. [DOI] [Google Scholar]
  18. Schröder P.; Navarro-Aviñó J.; Azaizeh H.; Goldhirsh A. G.; DiGregorio S.; Komives T.; Langergraber G.; Lenz A.; Maestri E.; Memon A. R.; Ranalli A.; Sebastiani L.; Smrcek S.; Vanek T.; Vuilleumier S.; Wissing F. Using phytoremediation technologies to upgrade waste water treatment in Europe. Environ. Sci. Pollut. Res. Int. 2007, 14, 490–497. 10.1065/espr2006.12.373. [DOI] [PubMed] [Google Scholar]
  19. Wang M.; Ioccozia J.; Sun L.; Lin C.; Lin Z. Inorganic-modified semiconductor TiO2 nanotube arrays for photocatalysis. Energy Environ. Sci. 2014, 7, 2182–2202. 10.1039/c4ee00147h. [DOI] [Google Scholar]
  20. Kalyanasundaram K.Photochemistry; The Royal Society of Chemistry, 2013; Vol. 41, pp 182–265. [Google Scholar]
  21. Zapata P. A.; Rabagliati F. M.; Lieberwirth I.; Catalina F.; Corrales T. Study of the photodegradation of nanocomposites containing TiO2 nanoparticles dispersed in polyethylene and in poly(ethylene-co-octadecene). Polym. Degrad. Stab. 2014, 109, 106–114. 10.1016/j.polymdegradstab.2014.06.020. [DOI] [Google Scholar]
  22. Bechambi O.; Sayadi S.; Najjar W. Photocatalytic degradation of bisphenol A in the presence of C-doped ZnO: Effect of operational parameters and photodegradation mechanism. J. Ind. Eng. Chem. 2015, 32, 201–210. 10.1016/j.jiec.2015.08.017. [DOI] [Google Scholar]
  23. Bilgin Simsek E.; Kilic B.; Asgin M.; Akan A. Graphene oxide based heterojunction TiO2 −ZnO catalysts with outstanding photocatalytic performance for bisphenol-A, ibuprofen and flurbiprofen. J. Ind. Eng. Chem. 2018, 59, 115–126. 10.1016/j.jiec.2017.10.014. [DOI] [Google Scholar]
  24. Reddy P. V. L.; Kim K.-H.; Kavitha B.; Kumar V.; Raza N.; Kalagara S. Photocatalytic degradation of bisphenol A in aqueous media: A review. J. Environ. Manage. 2018, 213, 189–205. 10.1016/j.jenvman.2018.02.059. [DOI] [PubMed] [Google Scholar]
  25. Wu W.; Shan G.; Wang S.; Zhu L.; Yue L.; Xiang Q.; Zhang Y.; Li Z. Environmentally relevant impacts of nano-TiO2 on abiotic degradation of bisphenol A under sunlight irradiation. Environ. Pollut. 2016, 216, 166–172. 10.1016/j.envpol.2016.05.079. [DOI] [PubMed] [Google Scholar]
  26. Li Y.; Liu L.; Guo M.; Zhang M. Synthesis of TiO2 visible light catalysts with controllable crystalline phase and morphology from Ti-bearing electric arc furnace molten slag. J. Environ. Sci. 2016, 47, 14–22. 10.1016/j.jes.2015.11.032. [DOI] [PubMed] [Google Scholar]
  27. Ong W.-J.; Tan L.-L.; Chai S.-P.; Yong S.-T.; Mohamed A. R. Highly reactive {001} facets of TiO2-based composites: synthesis, formation mechanism and characterization. Nanoscale 2014, 6, 1946–2008. 10.1039/c3nr04655a. [DOI] [PubMed] [Google Scholar]
  28. Yang C.; Dong W.; Cui G.; Zhao Y.; Shi X.; Xia X.; Tang B.; Wang W. Enhanced photocatalytic activity of PANI/TiO2 due to their photosensitization-synergetic effect. Electrochim. Acta 2017, 247, 486–495. 10.1016/j.electacta.2017.07.037. [DOI] [Google Scholar]
  29. Kavil J.; Ullattil S. G.; Alshahrie A.; Periyat P. Polyaniline as photocatalytic promoter in black anatase TiO2. Sol. Energy 2017, 158, 792–796. 10.1016/j.solener.2017.10.049. [DOI] [Google Scholar]
  30. Tan X.; Wang J.; Pang X.; Liu L.; Sun Q.; You Q.; Tan F.; Li N. Indocyanine green-loaded silver nanoparticle@polyaniline core/shell theranostic nanocomposites for photoacoustic/near-infrared fluorescence imaging-guided and single-light-triggered photothermal and photodynamic therapy. ACS Appl. Mater. Interfaces 2016, 8, 34991–35003. 10.1021/acsami.6b11262. [DOI] [PubMed] [Google Scholar]
  31. Chen X.; Li H.; Wu H.; Wu Y.; Shang Y.; Pan J.; Xiong X. Fabrication of TiO2@PANI nanobelts with the enhanced absorption and photocatalytic performance under visible light. Mater. Lett. 2016, 172, 52–55. 10.1016/j.matlet.2016.02.134. [DOI] [Google Scholar]
  32. Zheng Y.; Chen C.; Zhan Y.; Lin X.; Zheng Q.; Wei K.; Zhu J.; Zhu Y. Luminescence and photocatalytic activity of ZnO nanocrystals:  Correlation between structure and property. Inorg. Chem. 2007, 46, 6675–6682. 10.1021/ic062394m. [DOI] [PubMed] [Google Scholar]
  33. Radoičić M.; Saponjic Z.; Ciric-Marjonovic G.; Konstantinovic Z.; Mitric M. S.; Nedeljkovic J. Ferromagnetic Polyaniline-TiO2 Nanocomposites. Polym. Compos. 2012, 33, 1484–1493. 10.1002/pc.22278. [DOI] [Google Scholar]
  34. Radoičić M.; Saponjic Z.; Jankovic I. A.; Ciric-Marjonovic C.; Ahrenkiel S. P.; Comor M. I. Improvements for Photocatalytic Efficiency on polyaniline modified TiO2 nanoparticles. Appl. Catal., B 2013, 136–137, 133–139. 10.1016/j.apcatb.2013.01.007. [DOI] [Google Scholar]
  35. Radoičić M.; Ciric-Marjonovic G.; Saponjic Z. V.; Mitric M.; Konstantinovic Z.; Stoiljkovic M.; Nedeljkovic J. M. Structural and magnetic properties of nanocomposites based on nanostructured polyaniline and titania nanotubes. J. Mater. Sci. 2013, 48, 5776–5787. 10.1007/s10853-013-7370-1. [DOI] [Google Scholar]
  36. Anwer H.; Mahmood A.; Lee J.; Kim K.-H.; Park J.-W.; Yip A. C. K. Photocatalysts for the degradtion of dyes in Industrial Effluents: Opportunities and Challenges. Nano Res. 2019, 12, 955–972. 10.1007/s12274-019-2287-0. [DOI] [Google Scholar]
  37. Raza N.; Raza W.; Gul H.; Azam M.; Lee J.; Vikrant K.; Kim K.-H. Solar light active silver /titanium oxide/silica heterostructures for photocatalytic removal of organic dye. J. Cleaner Prod. 2020, 254, 120031. 10.1016/j.jclepro.2020.120031. [DOI] [Google Scholar]
  38. Bhaumik M.; Maity A.; Srinivasu V. V.; Onyango M. S. Removal of hexavalent chromium from aqueous solution using polypyrrole-polyaniline nanofibers. Chem. Eng. J. 2012, 181–182, 323–333. 10.1016/j.cej.2011.11.088. [DOI] [Google Scholar]
  39. Brooms T. J.; Otieno B.; Onyango M. S.; Ochieng A. Photocatalytic degradation of P-cresol using TiO2/ZnO hybrid surface capped with polyaniline. J. Environ. Sci. Health, Part A: Toxic/Hazard. Subst. Environ. Eng. 2018, 53, 99–107. 10.1080/10934529.2017.1377583. [DOI] [PubMed] [Google Scholar]
  40. Deshmukh P. R.; Patil S. V.; Bulakhe R. N.; Pusawale S. N.; Shim J.-J.; Lokhande C. D. Chemical synthesis of PANI–TiO2 composite thin film for supercapacitor application. RSC Adv. 2015, 5, 68939–68946. 10.1039/c5ra09233g. [DOI] [Google Scholar]
  41. Sandhya K. P.; Haridas S.; Sugunan S. Visible light induced photocatalytic activity of polyaniline modified TiO2 and Clay-TiO2 composites. Bull. Chem. React. Eng. Catal. 2013, 8, 145. 10.9767/bcrec.8.2.4949.145-153. [DOI] [Google Scholar]
  42. Sboui M.; Nsib M. F.; Rayes A.; Swaminathan M.; Houas A. TiO2–PANI/Cork composite: A new floating photocatalyst for the treatment of organic pollutants under sunlight irradiation. J. Environ. Sci. 2017, 60, 3–13. 10.1016/j.jes.2016.11.024. [DOI] [PubMed] [Google Scholar]
  43. Fleaca C. T.; Scarisoreanu M.; Morjan I.; Luculescu C.; Niculescu A.-M.; Badoi A.; Vasile E.; Kovacs G. Laser oxidative pyrolysis synthesis and annealing of TiO2 nanoparticles embedded in carbon–silica shells/matrix. Appl. Surf. Sci. 2015, 336, 226–233. 10.1016/j.apsusc.2014.11.106. [DOI] [Google Scholar]
  44. Haldorai Y.; Nguyen V. H.; Shim J.-J. Synthesis of polyaniline/Q-CdSe composite via ultrasonically assisted dynamic inverse emulsion polymerization. Colloid Polym. Sci. 2011, 289, 849. 10.1007/s00396-011-2400-5. [DOI] [Google Scholar]
  45. Li S.; Hu J. Photolytic and photocatalytic degradation of tetracycline: Effect of humic acid on degradation kinetics and mechanisms. J. Hazard. Mater. 2016, 318, 134–144. 10.1016/j.jhazmat.2016.05.100. [DOI] [PubMed] [Google Scholar]
  46. Erdogan D. A.; Ozensoy E. Hierarchical synthesis of corrugated photocatalytic TiO2 microsphere architectures on natural pollen surfaces. Appl. Surf. Sci. 2017, 403, 159–167. 10.1016/j.apsusc.2017.01.107. [DOI] [Google Scholar]
  47. Chen Y.; Wang K.; Lou L. Photodegradation of dye pollutants on silica gel supported TiO2 particles under visible light irradiation. J. Photochem. Photobiol., A 2004, 163, 281–287. 10.1016/j.jphotochem.2003.12.012. [DOI] [Google Scholar]
  48. Speakman S. A.Estimating Crystallite Size Using XRD; MIT Center for Materials Science and Engineering, 2014. [Google Scholar]
  49. Tamura N.; Gilbert P. U. P. A.. Methods in Enzymology; De Yoreo J. J., Ed.; Academic Press, 2013; pp 501–531. [DOI] [PubMed] [Google Scholar]
  50. Amano F.; Ishinaga E.; Yamakata A. Effect of particle size on the photocatalytic activity of WO3 particles for water oxidation. J. Phys. Chem. C 2013, 117, 22584–22590. 10.1021/jp408446u. [DOI] [Google Scholar]
  51. Tabaei H. S. M.; Kazemeini M.; Fattahi M. Preparation and characterization of visible light sensitive nano titanium dioxide photocatalyst. Sci. Iran. 2012, 19, 1626–1631. 10.1016/j.scient.2012.07.005. [DOI] [Google Scholar]
  52. Ruiz H.; Zambtrano M.; Giraldo L.; Sierra R.; Moreno-Pirajan J. Production and characterization of activated carbon from oil-palm shell for carboxylic acid adsorption. Orient. J. Chem. 2015, 31, 753–762. 10.13005/ojc/310217. [DOI] [Google Scholar]
  53. Sambaza S. S.; Masheane M. L.; Malinga S. P.; Nxumalo E. N.; Mhlanga S. D. Polyethyleneimine-carbon nanotube polymeric nanocomposite adsorbents for the removal of Cr6+ from water. Phys. Chem. Earth, Parts A/B/C 2017, 100, 236–246. 10.1016/j.pce.2016.08.002. [DOI] [Google Scholar]
  54. Mi F.-L.; Wu S.-J.; Lin F.-M. Adsorption of copper(II) ions by a chitosan–oxalate complex biosorbent. Int. J. Biol. Macromol. 2015, 72, 136–144. 10.1016/j.ijbiomac.2014.08.006. [DOI] [PubMed] [Google Scholar]
  55. Ravidhas C.; Juliat Josephine A.; Sudhagar P.; Devadoss A.; Terashima C.; Nakata K.; Fujishima A.; Moses Ezhil Raj A.; Sanjeeviraja C. Facile synthesis of nanostructured monoclinic bismuth vanadate by a co-precipitation method: structural, optical and photocatalytic properties. Mater. Sci. Semicond. Process. 2015, 30, 343–351. 10.1016/j.mssp.2014.10.026. [DOI] [Google Scholar]
  56. Sun X.; Yang L.; Li Q.; Liu Z.; Dong T.; Liu H. Polyethylenimine-functionalized poly(vinyl alcohol) magnetic microspheres as a novel adsorbent for rapid removal of Cr(VI) from aqueous solution. Chem. Eng. J. 2015, 262, 101–108. 10.1016/j.cej.2014.09.045. [DOI] [Google Scholar]
  57. Ba-Abbad M. M.; Kadhum A. A. H.; Mohamad A. B.; Takriff M. S.; Sopian K. Synthesis and catalytic activity of TiO2 nanoparticles for photochemical oxidation of concentrated chlorophenols under direct solar radiation. Int. J. Electrochem. Sci. 2012, 7, 4871–4888. [Google Scholar]
  58. Gurushantha K.; L. Renuka L.; Anantharaju K. S.; Vidya Y. S.; Nagaswarupa H. P.; Prashantha S. C.; Nagabhushana H. Photocatalytic and Photoluminescence studies of ZrO 2 /ZnO nanocomposite for LED and Waste water treatment applications. Mater. Today: Proc. 2017, 4, 11747–11755. 10.1016/j.matpr.2017.09.091. [DOI] [Google Scholar]
  59. Ye R.; Barron A.. Photoluminescence Spectroscopy and Its Applications. Retrieved from the OpenStax-CNX Web site: http://cnx.org/content/m38357/1.2, 2011.
  60. Omaish Ansari M.; Khan M. M.; Ansari S.; Raju K.; Lee J.; Cho M. H.. Enhanced Thermal Stability under DC Electrical Conductivity Retention and Visible Light Activity of Ag/TiO2@polyaniline Nanocomposite Film; American Chemistry Society, 2014. [DOI] [PubMed] [Google Scholar]
  61. Parveen N.; Ansari M. O.; Han T. H.; Cho M. H. Simple and rapid synthesis of ternary polyaniline/titanium oxide/graphene by simultaneous TiO2 generation and aniline oxidation as hybrid materials for supercapacitor applications. J. Solid State Electrochem. 2017, 21, 57–68. 10.1007/s10008-016-3310-8. [DOI] [Google Scholar]
  62. Golczak S.; Kanciurzewska A.; Fahlman M.; Langer K.; Langer J. Comparative XPS surface study of polyaniline thin films. Solid State Ionics 2008, 179, 2234–2239. 10.1016/j.ssi.2008.08.004. [DOI] [Google Scholar]
  63. Linh N. T. T.; Tuan P. D.; Dzung N. V. The shifts of band gap and binding energies of titania/hydroxyapatite material. J. Compos. 2014, 2014, 283034. 10.1155/2014/283034. [DOI] [Google Scholar]
  64. Chou J.-C.; Liao L. P. Study on pH at the point of zero charge of TiO2 pH ion-sensitive field effect transistor made by the sputtering method. Thin Solid Films 2005, 476, 157–161. 10.1016/j.tsf.2004.09.061. [DOI] [Google Scholar]
  65. Qureshi U. A.; Gubbuk I. H.; Ersoz M.; Solangi A. R.; Taqvi S. I. H.; Memon S. Q. Preparation of polyaniline montmorillonite clay composites for the removal of diethyl hexyl phthalate from aqueous solutions. Sep. Sci. Technol. 2016, 51, 214–228. 10.1080/01496395.2015.1088029. [DOI] [Google Scholar]
  66. Park Y.; Sun Z.; Ayoko G. A.; Frost R. L. Bisphenol A sorption by organo-montmorillonite: implications for the removal of organic contaminants from water. Chemosphere 2014, 107, 249–256. 10.1016/j.chemosphere.2013.12.050. [DOI] [PubMed] [Google Scholar]
  67. Zhang J.; Nosaka Y. Mechanism of the OH radical generation in photocatalysis with TiO2 of different crystalline types. J. Phys. Chem. C 2014, 118, 10824–10832. 10.1021/jp501214m. [DOI] [Google Scholar]
  68. Porkodi K.; Kumar K. V. Comments on“ photocatalytic properties of TiO2 modified with platinum and silver nanoparticles in the degradation of oxalic acid in aqueous solution” Langmuir Hinshelwood kinetics-A theoretical study. Appl. Catal., B 2008, 79, 108–109. 10.1016/j.apcatb.2007.09.041. [DOI] [Google Scholar]
  69. Blaszczak-Boxe C. S.; Saiz-Lopez A. Nitrate photolysis in ice and snow: A critical review of its multiphase chemistry. Atmos. Environ. 2018, 193, 224–241. 10.1016/j.atmosenv.2018.09.002. [DOI] [Google Scholar]
  70. Zhan M.; Yang X.; Xian Q.; Kong L. Photosensitized degradation of bisphenol A involving reactive oxygen species in the presence of humic substances. Chemosphere 2006, 63, 378–386. 10.1016/j.chemosphere.2005.08.046. [DOI] [PubMed] [Google Scholar]
  71. Mack J.; Bolton J. R. Photochemistry of nitrite and nitrate in aqueous solution: a review. J. Photochem. Photobiol., A 1999, 128, 1–13. 10.1016/s1010-6030(99)00155-0. [DOI] [Google Scholar]
  72. Guetzloff T. F.; Rice J. A. Does humic acid form a micelle?. Sci. Total Environ. 1994, 152, 31–35. 10.1016/0048-9697(94)90548-7. [DOI] [Google Scholar]
  73. Liang Y.; Britt D. W.; McLean J. E.; Sorensen D. L.; Sims R. C. Humic acid effect on pyrene degradation: finding an optimal range for pyrene solubility and mineralization enhancement. Appl. Microbiol. Biotechnol. 2007, 74, 1368–1375. 10.1007/s00253-006-0769-8. [DOI] [PubMed] [Google Scholar]
  74. Lim B.-R.; Do S.-H.; Hong S.-H. The impact of humic acid on the removal of bisphenol A by adsorption and ozonation. Desalin. Water Treat. 2015, 54, 1226–1232. 10.1080/19443994.2014.896751. [DOI] [Google Scholar]
  75. Voutsa D.Plastics in Dentistry and Estrogenicity; Springer, 2014; pp 51–77. [Google Scholar]
  76. Majedi S.; Lai E. Mass spectrometric analysis of bisphenol A desorption from titania nanoparticles: Ammonium acetate, fluoride, formate, and hydroxide as chemical desorption agents. Methods Protoc. 2018, 1, 26. 10.3390/mps1030026. [DOI] [PMC free article] [PubMed] [Google Scholar]
  77. Zhao H.; Xiang L.; Li J.; Yang Z.; Fang J.; Zhao C.; Xu S.; Cai Z. Investigation on fragmentation pathways of bisphenols by using electrospray ionization Orbitrap mass spectrometry. Rapid Commun. Mass Spectrom. 2016, 30, 1901–1913. 10.1002/rcm.7666. [DOI] [PubMed] [Google Scholar]
  78. Torki F.; Faghihian H. Visible light degradation of naproxen by enhanced photocatalytic activity of NiO and NiS, scavenger study and focus on catalyst support and magnetization. Photochem. Photobiol. 2018, 94, 491–502. 10.1111/php.12906. [DOI] [PubMed] [Google Scholar]
  79. Masid S.; Tayade R.; Rao N. Efficient visible light active polyaniline/TiO2 nanocomposite photocatalyst for degradation of Reactive Blue 4. Int. J. Photocatal. 2014, 119, 190–203. [Google Scholar]
  80. Naraginti S.; Li Y.; Wu Y.; Zhang C.; Upreti A. R. Mechanistic study of visible light driven photocatalytic degradation of EDC 17α-ethinyl estradiol and azo dye Acid Black-52: phytotoxicity assessment of intermediates. RSC Adv. 2016, 6, 87246–87257. 10.1039/c6ra20702b. [DOI] [Google Scholar]
  81. Pelaez M.; Falaras P.; Likodimos V.; O’Shea K.; de la Cruz A. A.; Dunlop P. S. M.; Byrne J. A.; Dionysiou D. D. Use of selected scavengers for the determination of NF-TiO2 reactive oxygen species during the degradation of microcystin-LR under visible light irradiation. J. Mol. Catal. A: Chem. 2016, 425, 183–189. 10.1016/j.molcata.2016.09.035. [DOI] [PMC free article] [PubMed] [Google Scholar]
  82. Anwer H.; Park J.-W. Synthesis and characterization of a heterojunction rGO/ZrO2/Ag3PO4 nanocomposite for degradation of organic contaminants. J. Hazard. Mater. 2018, 358, 416–426. 10.1016/j.jhazmat.2018.07.019. [DOI] [PubMed] [Google Scholar]
  83. Cortes M. A. L. R. M.; Hamilton J. W. J.; Sharma P. K.; Brown A.; Nolan M.; Gray K. A.; Byrne J. A. Formal quantum efficiencies for the photocatalytic reduction of CO2 in a gas phase batch reactor. Catal. Today 2019, 326, 75–81. 10.1016/j.cattod.2018.10.047. [DOI] [Google Scholar]

Articles from ACS Omega are provided here courtesy of American Chemical Society

RESOURCES