# Synthesis and Application of Fe-Doped TiO2-Halloysite Nanotubes Composite and Their Potential Application in Water Treatment

Authors: Emmanuel Nyankson; Benjamin Agyei-Tuffour; Jonas Adjasoo; Annan Ebenezer; David Dodoo-Arhin; Abu Yaya; Bismark Mensah; Johnson Kwame Efavi
Journal: Advances in Materials Science and Engineering
 (2019)

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## Abstract

In this work, the potential application of TiO2-Fe-HNT photocatalyst-adsorbent composite in water treatment technologies was confirmed. The photocatalyst-adsorbent composite (TiO2-Fe-HNTs) was synthesized by the hydrothermal method and characterized by X-ray diffraction, thermogravimetric analysis, Fourier-transform infrared spectroscopy, scanning electron microscopy-energy dispersive X-ray spectroscopy, and diffuse reflectance spectroscopy. The adsorption and photocatalysis mechanism by the TiO2-Fe-HNT composite were examined on methylene blue dye, rhodamine blue dye, naproxen sodium (pharmaceutical drug waste), and imidacloprid (pesticide). The TiO2-Fe-HNT composite was active in UV and visible regions of the electromagnetic spectrum. The adsorption and photocatalytic efficiency increased with increasing amount of HNTs. The photocatalyst-adsorbent composite exhibited excellent removal efficiency for pharmaceutical waste (naproxen sodium) and pesticides (imidacloprid). An adsorption equilibrium data fitted well with the pseudo-second-order kinetics for both methylene blue and rhodamine blue dyes with the intraparticle model describing its rate-controlling steps. The Langmuir and Freundlich isotherm models further described the adsorption of methylene blue and rhodamine blue molecules, respectively.


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## Body

## 1. Introduction

One of the major problems faced by humanity all over the world is poor water quality. Over 2 billion people globally do not have access to potable water. 8% of this number collects drinking water directly from surface water sources such as lakes and rivers [1]. The increase in human population with its associated increase in industrialization and climate change is expected to result in an increase in demand for safe water [2]. The number of people not having access to safe water is expected to grow due to increased water contamination from industrial and human activities. The major pollutants found in surface and ground water bodies include water-soluble dyes [3], pharmaceutical waste [4], agro-chemicals [5], heavy metals [6], microorganisms [7], and hydrocarbons [8]. The presence of these pollutants may have a detrimental effect on aquatic species and humans that utilize the contaminated water in their daily activities. Contaminated water is associated with the transmission of diseases such as hepatitis A, cholera, diarrhea, and dysentery. The WHO organization has reported that approximately 845000 people die from diarrhea each year [9]. Water pollution has associated economic and social impact such as increase in expenditure on health, long risky journeys to collect portable water, and poor school attendance due to ill health. To render water bodies safe for use, these pollutants must be removed, and therefore, different water treatment technologies have been developed to reduce the worsening of clean water shortage and also meet the increased demand for potable water [10].The different water treatment technologies can be categorized into three: tertiary, secondary, and primary [11]. The primary treatment technology includes physical and chemical purification processes such as filtration, sedimentation, and coagulation, while the secondary treatment technology includes biological purification processes such as aerobic and anaerobic processes. On the contrary, oxidation and advanced oxidation processes, adsorption, distillation, and solvent extraction can be categorized under tertiary treatment technologies [3, 11]. The right method to be adopted during water treatment is dependent on the extent of pollution and the nature of the pollutant.Water bodies polluted by dyes, pharmaceutical waste, and pesticides can be treated through advanced oxidation processes (AOPs). In AOPs, reactive oxygen species (e.g., hydroxyl radicals), which aid the degradation of pollutants which are resistant to conventional treatment methods are produced. The pollutants are degraded into innocuous compounds such as H2O, CO2, and mineral acids [12]. AOPs which include sonolysis, Fenton processes, photocatalysis, and the use of UV radiation and hydrogen peroxide are nontoxic and economical and can degrade different types of organic pollutants [3]. Photocatalysis is one of the most widely studied AOPs, and its potential in the degradation of dyes has been examined on different photocatalysts such as TiO2 [13], Ag3PO4 [14], TiO2-Ag [15], and ZnO [16]. Photocatalysis involves (a) the absorption of photons (the energy of the photons must be greater than the bang gap energy), (b) excitation of photogenerated electrons from the valence band (VB) to the conduction band (CB) generating electron-hole pair, and (c) oxidation of adsorbed pollutants by the VB hole which is strongly oxidizing. The CB electrons which are strongly reducing can be used to create reactive oxygen species that can attack pollutants and degrade them [13]. TiO2 is a semiconductor photocatalyst and has been investigated by many researchers. TiO2 photocatalyst is stable, relatively cheaper, nontoxic, and readily available [17]. However, the large band gap of TiO2 (∼3.0–3.2 eV) has restricted its photoactivity to the UV region of the electromagnetic spectrum [18]. TiO2 can be made active in the UV-Vis region of the solar spectrum by doping and by coating the surface of TiO2 [15] with a metal. Among the transition metal ions, Fe3+-doped TiO2 has resulted in an enhanced photocatalytic activity. The ionic radius of Fe3+ is almost the same as that of Ti4+; hence, Fe3+ can be easily integrated into the structure of TiO2 [19]. The reduction of the electron-hole recombination after doping with Fe is responsible for the enhanced photocatalytic activity of Fe-doped TiO2 [20].The synergistic effect of adsorption and photocatalysis has been investigated by researchers. The synergistic effect that resulted from the adsorption properties of activated carbon (AC) powders and photocatalytic properties of TiO2 has been reported. The AC-TiO2 composite recorded an enhanced photocatalytic and pollutant removal efficiency when compared with that of pristine TiO2 [3, 21]. The synergistic relationship between adsorption and photocatalysis has been further examined by Sun et al. using the TiO2-sulfanyl-AC composite [22]. The enhanced pollutant removal efficiency by these composites can be linked to the excellent adsorption capacity of AC. The incorporation of adsorbent into photocatalyst therefore offers a new design for developing photocatalyst-adsorbent system with enhanced pollutant removal/degradation efficiency. In addition to AC, other adsorbent materials such as zeolite [23], graphene [24], SiO2 [25], Al2O3 [26], and multiwalled carbon nanotubes [27] have been used to develop photocatalyst-adsorbent systems. The adsorbent materials used in developing these composite adsorbent-photocatalyst systems have been reported to be expensive. In addition, there are problems associated with their treatment and regeneration [28]. More so, application of TiO2 limits the practical usage to the UV region of the solar spectrum [29].Halloysite nanotubes (HNTs) are natural 1 : 1 aluminosilicate material made up of an octahedron alumina sheet and tetrahedron silica sheet. HNTs possess rolled hollow nanotubular structure with high specific surface area of ca. 65 m2/g. The diameter and length of HNTs vary from ∼15 to 100 nm and 500 to 1100 nm, respectively [30]. Halloysite nanotubes have high adsorption capacity and are biocompatible, relatively cheap, readily available, and nontoxic. These unique properties of HNTs have enabled them to be used in different applications such as oil spill remediation, drug delivery systems, electronics, catalysis, water purification, and composite materials for construction [31–35]. Unlike the different adsorbent materials used in developing adsorbent-photocatalyst systems, naturally occurring HNTs are readily available and are low cost when compared with carbon nanotubes.The synergistic effect of photocatalysis and adsorption by the TiO2-Fe-HNT composite has not been widely investigated. The TiO2-Fe-HNT adsorbent-photocatalyst composite was developed by the hydrothermal method and characterized by XRD, TGA, FTIR, SEM-EDX, and DRS. The synergistic effect of adsorption and photocatalysis by the synthesized TiO2-Fe-HNT composite was examined on methylene blue dye, rhodamine blue dye, naproxen sodium (pharmaceutical waste), and imidacloprid (pesticide). An adsorption equilibrium studies was carried out using the water-soluble dyes (methylene blue and rhodamine blue dyes) and fitted with the Freundlich and Langmuir adsorption equilibrium isotherms, respectively. In addition, adsorption kinetics data were obtained and fitted with pseudo-first and pseudo-second-order kinetics and the intraparticle diffusion model.


## 2. Materials and Methods

### 2.1. Materials

Halloysite nanoclay (HNTs), titanium (IV) isopropoxide (TTIP), isopropyl alcohol (IPA), nitric acid (HNO3), Fe (III) chloride, methylene blue, rhodamine B, naproxen sodium, imidacloprid, benzoquinone, ethylenediamine tera-acetic acid sodium, and dimethyl sulfoxide (DMSO) were obtained from Sigma-Aldrich, UK.


### 2.2. Synthesis of TiO2-Fe-HNTs

Three different solutions were prepared and labeled A, B, and C. Solution A was prepared by adding 3.8 mL of TTIP to 19 mL of IPA under constant stirring. Acid solution with a pH of ca. 1.5 was prepared using HNO3 and labeled as Solution B. 2 mL of Solution B was added dropwise to Solution A while stirring followed by addition of a calculated amount of FeCl3 (Solution C). The resulting solution C was continuously stirred for 90 minutes. At the same time, a calculated amount of HNTs was added to 20 mL of IPA and sonicated for 90 minutes. The dispersed HNTs in IPA were added to Solution C and agitated continuously for another 90 minutes. The resulting solution was transferred into an autoclave (hydrothermal reactor) and heat treated at 180°C for 3 hours. The obtained TiO2-Fe-HNT particles were centrifuged and washed repeatedly with DI water. The particles were dried under vacuum at 70°C overnight. The amount of FeCl3 added was varied to prepare samples with 1 and 2 wt.% Fe. The amount of the HNTs added resulted in the formation of sample with 25, 50, and 60 wt.% HNTs. FeCl3 and HNTs were not included during the synthesis of TiO2.


### 2.3. Characterization

X-ray powder diffraction (XRD) analysis was conducted on the adsorbent-photocatalyst composite using the Bruker D8 diffractometer-fitted LynxEye detector. Cu-Kα radiation (λ = 0.15405 nm) and a 2theta angular range of 10–80° were used. Field Electron and Ion (FEI) Nova NanoSEM connected to the energy dispersive X-ray acquisition detector was used to study the morphology and elemental composition of the TiO2-Fe-HNT composite. Fourier-transform infrared (FTIR) spectroscopy was carried out using a Bruker Tensor 2027 FT-IR instrument. TA Instrument Q500 was used to analyze the thermal behavior (TGA). The TGA was conducted under nitrogen gas at the heating rate of 10°C/min. The diffuse reflectance spectroscopy (DRS) was conducted using the Ocean Optics USB-4000 UV-Vis spectrometer. The synthesized nanocomposites were compressed into a flat film prior to measurement.


### 2.4. Photocatalysis Experiment

The photocatalysis was conducted using the tungsten halogen UV-Vis light source. The photocatalytic activity was conducted on methylene blue, rhodamine B, naproxen sodium, and imidacloprid. In the typical photodegradation procedure, 200 mL of methylene blue solution (10 mg/L) in deionized water was used with a photocatalyst suspension concentration of 0.5 g/L. The methylene blue-photocatalyst mixture was stirred in the dark for 30 minutes and then irradiated with the tungsten halogen UV-Vis light source. Detailed adsorption studies revealed that the adsorption of dye molecules on the composite material was complete in 30 minutes. 2 mL aliquots were taken at different time intervals and centrifuged (at 6000 rpm for 10 minutes). The dye absorbance was quantified using a UV-Vis spectrometer (atλ = 665 nm). A similar procedure was used for the photocatalytic degradation of rhodamine dye (10 mg/L), naproxen sodium (8 mg/L), and imidacloprid (8 mg/L). The absorbance was taken at ∼554, ∼230, and ∼269 nm for rhodamine B, naproxen sodium, and imidacloprid, respectively.


### 2.5. Absorption Experiment

The absorption of methylene blue and rhodamine B from aqueous solution was conducted using a batch method. A calculated amount of TiO2-Fe-HNTs was added to 200 mL of different concentrations of the dye solution (2, 4, 6, 8, and 10 mg/L) and stirred continuously. 2 mL of the dye solution was sampled at different time intervals and centrifuged (6000 rpm for 10 minutes) to remove the particles. The absorbance of the supernatant was taken at a wavelength of 665 nm and 554 nm for methylene blue dye and rhodamine B dye, respectively, using the UV-Vis spectrometer.All the experiments in this paper were repeated in triplicates, and the standard mean errors were calculated (represented as error bars in the Figures1–3). The error bars were calculated from different measurements of the same sample.Figure 1
(a) Photocatalytic degradation and adsorption of methylene blue by TiO2 and TiO2-Fe-HNTs. (b) Summary of the performance of the synthesized TiO2 and TiO2-Fe-HNTs (T: TiO2; T-Fe-H: TiO2-Fe-HNT composite).
(a)
(b)Figure 2
(a) Photocatalytic degradation and adsorption of rhodamine blue by TiO2 and TiO2-Fe-HNTs. (b) Summary of the performance of the synthesized TiO2 and TiO2-Fe-HNTs (T: TiO2; T-Fe-H: TiO2-Fe-HNT composite).
(a)
(b)Figure 3
(a) Pseudo-first-order rate constants for photodegradation of rhodamine B (RhB) and methylene blue (MB). (b) Effect of Fe wt.% on the photodegradation efficiency of TiO2-Fe-HNT composite (rhodamine B).
(a)
(b)


## 2.1. Materials

Halloysite nanoclay (HNTs), titanium (IV) isopropoxide (TTIP), isopropyl alcohol (IPA), nitric acid (HNO3), Fe (III) chloride, methylene blue, rhodamine B, naproxen sodium, imidacloprid, benzoquinone, ethylenediamine tera-acetic acid sodium, and dimethyl sulfoxide (DMSO) were obtained from Sigma-Aldrich, UK.


## 2.2. Synthesis of TiO2-Fe-HNTs

Three different solutions were prepared and labeled A, B, and C. Solution A was prepared by adding 3.8 mL of TTIP to 19 mL of IPA under constant stirring. Acid solution with a pH of ca. 1.5 was prepared using HNO3 and labeled as Solution B. 2 mL of Solution B was added dropwise to Solution A while stirring followed by addition of a calculated amount of FeCl3 (Solution C). The resulting solution C was continuously stirred for 90 minutes. At the same time, a calculated amount of HNTs was added to 20 mL of IPA and sonicated for 90 minutes. The dispersed HNTs in IPA were added to Solution C and agitated continuously for another 90 minutes. The resulting solution was transferred into an autoclave (hydrothermal reactor) and heat treated at 180°C for 3 hours. The obtained TiO2-Fe-HNT particles were centrifuged and washed repeatedly with DI water. The particles were dried under vacuum at 70°C overnight. The amount of FeCl3 added was varied to prepare samples with 1 and 2 wt.% Fe. The amount of the HNTs added resulted in the formation of sample with 25, 50, and 60 wt.% HNTs. FeCl3 and HNTs were not included during the synthesis of TiO2.


## 2.3. Characterization

X-ray powder diffraction (XRD) analysis was conducted on the adsorbent-photocatalyst composite using the Bruker D8 diffractometer-fitted LynxEye detector. Cu-Kα radiation (λ = 0.15405 nm) and a 2theta angular range of 10–80° were used. Field Electron and Ion (FEI) Nova NanoSEM connected to the energy dispersive X-ray acquisition detector was used to study the morphology and elemental composition of the TiO2-Fe-HNT composite. Fourier-transform infrared (FTIR) spectroscopy was carried out using a Bruker Tensor 2027 FT-IR instrument. TA Instrument Q500 was used to analyze the thermal behavior (TGA). The TGA was conducted under nitrogen gas at the heating rate of 10°C/min. The diffuse reflectance spectroscopy (DRS) was conducted using the Ocean Optics USB-4000 UV-Vis spectrometer. The synthesized nanocomposites were compressed into a flat film prior to measurement.


## 2.4. Photocatalysis Experiment

The photocatalysis was conducted using the tungsten halogen UV-Vis light source. The photocatalytic activity was conducted on methylene blue, rhodamine B, naproxen sodium, and imidacloprid. In the typical photodegradation procedure, 200 mL of methylene blue solution (10 mg/L) in deionized water was used with a photocatalyst suspension concentration of 0.5 g/L. The methylene blue-photocatalyst mixture was stirred in the dark for 30 minutes and then irradiated with the tungsten halogen UV-Vis light source. Detailed adsorption studies revealed that the adsorption of dye molecules on the composite material was complete in 30 minutes. 2 mL aliquots were taken at different time intervals and centrifuged (at 6000 rpm for 10 minutes). The dye absorbance was quantified using a UV-Vis spectrometer (atλ = 665 nm). A similar procedure was used for the photocatalytic degradation of rhodamine dye (10 mg/L), naproxen sodium (8 mg/L), and imidacloprid (8 mg/L). The absorbance was taken at ∼554, ∼230, and ∼269 nm for rhodamine B, naproxen sodium, and imidacloprid, respectively.


## 2.5. Absorption Experiment

The absorption of methylene blue and rhodamine B from aqueous solution was conducted using a batch method. A calculated amount of TiO2-Fe-HNTs was added to 200 mL of different concentrations of the dye solution (2, 4, 6, 8, and 10 mg/L) and stirred continuously. 2 mL of the dye solution was sampled at different time intervals and centrifuged (6000 rpm for 10 minutes) to remove the particles. The absorbance of the supernatant was taken at a wavelength of 665 nm and 554 nm for methylene blue dye and rhodamine B dye, respectively, using the UV-Vis spectrometer.All the experiments in this paper were repeated in triplicates, and the standard mean errors were calculated (represented as error bars in the Figures1–3). The error bars were calculated from different measurements of the same sample.Figure 1
(a) Photocatalytic degradation and adsorption of methylene blue by TiO2 and TiO2-Fe-HNTs. (b) Summary of the performance of the synthesized TiO2 and TiO2-Fe-HNTs (T: TiO2; T-Fe-H: TiO2-Fe-HNT composite).
(a)
(b)Figure 2
(a) Photocatalytic degradation and adsorption of rhodamine blue by TiO2 and TiO2-Fe-HNTs. (b) Summary of the performance of the synthesized TiO2 and TiO2-Fe-HNTs (T: TiO2; T-Fe-H: TiO2-Fe-HNT composite).
(a)
(b)Figure 3
(a) Pseudo-first-order rate constants for photodegradation of rhodamine B (RhB) and methylene blue (MB). (b) Effect of Fe wt.% on the photodegradation efficiency of TiO2-Fe-HNT composite (rhodamine B).
(a)
(b)


## 3. Results and Discussion

### 3.1. Structural, Thermal, and Functional Groups Characterization

The XRD patterns of TiO2 and TiO2-Fe-HNT composite are shown in Figure 4. The crystallographic planes (101), (004), (200), (105), (211), (204), and (215) are characteristics peaks of anatase TiO2 (JCPDS No. 21-1272). It was observed that the crystallinity of TiO2 decreased upon the addition of Fe and HNTs. This was deduced from the broadening of the peaks. An increase in the amount of the Fe decreased the crystallinity of the synthesized composite. This is because the Fe atoms acted as defects in TiO2. This distorted the crystal structure of TiO2 and hindered the crystallinity of TiO2 [36]. The characteristic peak (110) of HNTs was observed.Figure 4
X-ray diffraction pattern of TiO2 and TiO2-Fe-HNTs (“∗” and “#” represents the crystallographic planes of TiO2 and HNTs, respectively).The XRD confirmed the synthesis of the TiO2-Fe-HNT composite. The characteristic peak of Fe was not observed probably due to the relatively lower concentration of Fe used in the synthesis of the composite.Thermogravimetric analysis of the synthesized TiO2, TiO2-Fe-HNTs, and HNTs was performed, and the results are presented in Figure 5. All the plots depict a multistage decomposition with relatively stable intermediate products. TiO2 recorded two significant weight losses. There was ca. 5.1 wt.% loss when TiO2 was heated from 60 to 250°C and ca. 1 wt.% when heated from 250 to 400°C. The weight of TiO2 remained constant at heat treatments above 400°C. The total weight loss after heat treatment was ca. 6.1%. On the contrary, HNTs recorded ca. 5.3 wt.% loss at temperatures between 60 and 400°C. Increasing the temperature from 400 to 540°C resulted in an additional 11.3 wt.% loss. A total of 16.6 wt.% loss was recorded for HNTs. The TGA curve for the composite TiO2-Fe-HNTs was a combination of the curves of HNTs and TiO2. The total weight loss recorded over the entire temperature at which the heat treatment was carried out was ca. 12.5%. The TGA curve of the composite confirmed the presence of TiO2 and HNTs.Figure 5
TGA curves for HNTs, TiO2, and TiO2-Fe-HNTs.The FTIR spectrum of the prepared TiO2-Fe-HNT composite showed three additional peaks that were not observed in the spectrum of TiO2 (Figure 6), suggesting the presence of HNTs in the composite. The broad band below 1000 cm−1 was ascribed to the stretching vibration of Ti-O [37, 38]. The strong peak at 1623 cm−1 corresponds to the symmetric vibration of absorbed water [39]. The broad peak at 3373 cm−1 was attributed to the stretching vibrations of the O-H functional group. The presence of HNTs in the composite material resulted in the appearance of three other peaks at 912, 1118, and 1032 cm−1. The band at 912 cm−1 resulted from the deformation vibration of the inner O-H groups present in HNTs. Stretching vibration of Si-O-Si was observed at 1032 cm−1, whereas apical Si-O stretching mode resulted in a peak at 1118 cm−1 [32].Figure 6
FTIR spectra of TiO2 and TiO2-Fe-HNTs.


### 3.2. Structural Morphology, Elemental, and Optical Characterization

The structural morphology and elemental characterization of the composite was investigated through SEM-EDX analysis. The SEM images and the energy dispersive X-ray spectroscopy analysis are presented in Figure7.Figure 7
Scanning electron microscopy images of (a) HNTs, (b) TiO2-1 wt.% Fe-50 wt.% HNTs at low magnification, (c) TiO2-1 wt.% Fe-50 wt.% HNTs at high magnification. (d) The secondary reference image for the area EDX data was acquired (TiO2-1 wt.% Fe-50 wt.% HNTs) and (e) EDX spectrum of TiO2-1 wt.% Fe-50 wt.% HNTs.
(a)
(b)
(c)
(d)
(e)The scanning electron microscopy images of HNTs showed tubular structure with an uneven length and diameter of the tubes (Figure7(a)). The morphology and structure of the HNTs did not change as observed in the SEM images of the composite (TiO2-1 wt.% Fe-50 wt.% HNTs). However, agglomeration of the individual tubes can be seen in Figure 7(b). This is probably due to the growth of TiO2-Fe crystals on the outer surface of the HNTs. It was observed at a relatively higher magnification (Figure 7(c)) that the TiO2-Fe crystals settled on the outer surface of the HNTs. This is expected to enhance the aqueous dispersity of the composite as HNTs have been reported to possess high dispersity in water [30]. Enhanced aqueous dispersity will increase the proximity between water-soluble pollutants and the composite and result in improved photocatalysis and adsorption. The EDX analysis as shown in Figures 7(d) and 7(e) established that Al, Si, Fe, O, and Ti are present in the composite material. The atomic ratio of Al to Si is ca. 1. From the chemical formula of HNTs Al2Si2O5OH4⋅nH2O, nSi/nAl is 1. This was confirmed from the table in Figure 7. In addition, the wt.% of Fe in the composite material used for this analysis was 1%. This was also confirmed by the EDX results which detected Fe (0.83 wt.%).Since TiO2 is a well-known photocatalyst, the effect of HNTs and Fe on the optical property of the TiO2-Fe-HNT composite was examined using diffused reflectance spectroscopy (DRS), and the result is presented in Figure 8.Figure 8
DRS analysis of TiO2 and TiO2-Fe-HNT composite.TiO2 recorded ca. 100% reflectance in the visible region of the solar spectrum as can be seen in Figure 8. This implies that TiO2 did not absorb visible light radiation. However, the reflectance by TiO2 was almost zero in the ultraviolet region, implying that TiO2 is active in the UV region of the electromagnetic spectrum. Addition of Fe and HNTs resulted in a decrease in the reflectance in the visible light region and a slight increase in reflectance in the UV region, meaning that the composite was photoactive in both ultraviolet and visible regions of the solar spectrum. The percentage of light absorption in the visible region increased with increasing wt.% of Fe. The optical band gap (BG) was also estimated using the Kubelka–Munk method. The estimated BG was 3.31, 3.24, and 3.2 eV for TiO2, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-2 wt.% Fe-50 wt.% HNTs, respectively. The Fe and halloysite nanotubes did not influence to a greater extent the optical band gap of the composite.


### 3.3. Adsorption and Photocatalysis of TiO2 and TiO2-Fe-HNT Composite

The potential usage of the synthesized nanocomposite in water purification was examined through adsorption and photocatalytic processes using model pollutants such as water-soluble dyes (rhodamine B and methylene blue), pharmaceutical drug waste (naproxen sodium), and pesticide (imidacloprid). The results for the adsorption and photocatalytic studies using methylene blue are presented in Figure1. The first thirty minutes represent the adsorption of methylene blue by TiO2 and TiO2-Fe-HNTs in the dark. The light was turned on after 30 minutes. The adsorption of methylene blue increased with increasing amount of HNTs used in the synthesis of the composite.TiO2 adsorbed ca. 4% of the methylene blue. Within the same period (30 minutes in the dark), TiO2-1 wt.% Fe-25 wt.% HNTs, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-1 wt.% Fe-60 wt.% HNTs adsorbed 28, 50, and 60% of methylene blue, respectively. Methylene blue is positively charged when in solution. On the contrary, HNTs possess a negatively charged surface. The high adsorption recorded resulted from the electrostatic interaction between the cationic methylene blue and the anionic surface of HNTs. After the adsorption process, the composite-methylene blue solution was irradiated with tungsten halogen lamp to start the photocatalysis process. The photocatalytic efficiency of TiO2-Fe-HNTs was observed to be lower than that of the pristine TiO2. As revealed in the SEM images, the TiO2 crystals were deposited onto the HNTs. Adsorption of the methylene blue dye by HNTs therefore enhanced the proximity between the pollutant and the TiO2-Fe photocatalyst. This is expected to enhance the photocatalytic activity. However, excessive adsorption of methylene blue may block the active sites of TiO2-Fe from light (solar energy) and hinder photocatalysis. The decrease in photocatalytic activity of the composite material is therefore attributed to the excessive adsorption of the cationic methylene blue by the anionic HNTs. The adsorbed methylene blue blocked the light from reaching the photoactive active sites of the TiO2-Fe. It can also be seen in Figure 1 that the photocatalytic efficiency of the composite material was fairly comparable even with increasing amount of HNTs used in the synthesis. As can be seen in the summary presented in Figure 1(b), the removal of the methylene blue is attributed to the combined effect of photocatalysis and adsorption. The total removal % of methylene blue by the TiO2-Fe-HNT composite was therefore higher than that by TiO2. However, it must be mentioned that the amount of the dye degraded through photocatalysis (as can be seen in Figure 1(b)) was relatively smaller when compared to the amount adsorbed.The photocatalytic and adsorption potential of the composite material was again examined using rhodamine B dye. This study is necessary because the adsorption studies conducted revealed that rhodamine is not “excessively” adsorbed by HNTs. The photocatalytic effect of TiO2-Fe-HNTs is expected to be more pronounced with rhodamine B than with methylene blue. The results are presented in Figure 2. TiO2 adsorbed ca. 11% of the rhodamine B dye, while TiO2-1 wt.% Fe-25 wt.% HNTs, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-1 wt.% Fe-60 wt.% HNTs adsorbed 20.7, 22.2, and 19.3%, respectively. Unlike methylene blue, the adsorption of rhodamine B by the composite was almost the same, irrespective of the amount of HNTs used in the synthesis of the composite. This suggests that the adsorption mechanism of methylene blue by the composite may be different from that of rhodamine B. The photodegradation efficiency of the composite material was higher than that of the pristine TiO2. However, it can be seen in Figure 2(b) that, unlike methylene blue (Figure 1), for all the composite materials studied, removal of rhodamine B through photocatalysis was significantly higher than removal through adsorption. This reveals that though the removal of the pollutant is through photocatalysis and adsorption, the dominant process may be dependent on the nature of the pollutants and its interaction with the HNTs.The pseudo-first-order rate constants for the photodegradation of methylene blue and rhodamine B dyes were calculated from equation (1), and the results are presented in Figure 3(a):(1)lnCtCo=kt.Generally, for the degradation of rhodamine B, the rate constants of the TiO2-Fe-HNT composites were significantly higher than that of TiO2. However, for the degradation of methylene blue, the rate constant increased in the order T-1Fe-25 H < T < T-1Fe-50 H < T-1Fe-60 H. It was also observed that the rate constants for the photodegradation by the composite were higher for rhodamine B than methylene blue. The TiO2 crystals were deposited onto the HNTs. The pronounced adsorption of methylene blue by the composite material resulted in the complete coverage of TiO2. This might have reduced the absorption of photons by TiO2 resulting in decreased photocatalytic activity and the smaller rate constant values obtained. Clay materials including HNTs have been reported to be poor adsorbents of rhodamine B dyes. This has been ascribed to the molecular size of rhodamine B and the nature of the interaction between rhodamine dyes and clays [40]. The lower adsorption of rhodamine B ensured that the TiO2 surface was not completely covered by the pollutants, and this allowed for enhanced absorption of photons. This enhanced the photocatalytic activity towards rhodamine B, and hence, the relatively higher rate constant values were observed. Therefore, to develop an efficient photocatalytic-adsorption process, a system must be designed in such a way that the amount of pollutants adsorbed does not exceed a threshold that may render the photocatalyst inactive.The potential application of the synthesized composite in adsorbing and degrading other pollutants was examined using naproxen sodium (pharmaceutical drug) and imidacloprid (pesticides). The composite-pollutant solution mixture was stirred in the dark for 24 hours. The mixture was then irradiated with light, and the photodegradation was examined over specified time intervals. The UV-Vis absorption spectrum of these pollutants after adsorption and UV-Vis light irradiation are presented in Figure9. The Fe-TiO2-HNT composite adsorbed approximately 29.5% and 23.9% of the naproxen sodium and imidacloprid (Figure 9 and Table 1). The photocatalytic degradation of the naproxen sodium was more rapid than that of imidacloprid with a recorded photocatalytic degradation of 35.9% in 4 hours. The overall removal (adsorption and photocatalysis) efficiency was 65.4 and 41% for naproxen sodium and imidacloprid, respectively. The difference in the observed photocatalytic activity resulted from the differences in the structural stabilities of imidacloprid and naproxen sodium. Naproxen sodium is negatively charged in solution. HNTs have been reported to possess a slightly negatively charged outer surface and a positively charged inner lumen (lumen is the empty cylindrical space in the HNTs) [41]. This may result in an electrostatic interaction between the naproxen sodium and HNTs. It is obvious from Figure 9 and Table 1 that Fe-TiO2-HNTs have the potential to remove (by adsorption and photodegradation) pesticides and pharmaceutical drug wastes from water bodies. It is however noted that unlike methylene blue and rhodamine B, a relatively longer time would be required for the photodegradation. Again, this may be due to the differences in the interaction between the pollutants and the Fe-TiO2-HNTs as well as the structural stabilities of the various pollutants.Figure 9
Adsorption and photodegradation of (a) naproxen sodium and (b) imidacloprid. (c) Comparing the adsorption and photodegradation of imidacloprid and naproxen sodium by the Fe-TiO2-HNT composite.
(a)
(b)
(c)Table 1
Summary of the adsorption and photodegradation efficiency.
Pollutant
Removal by adsorption (%)
Removal by photocatalysis (%)
Overall removal (%)
Photocatalysis rate constant (min−1)
Imidacloprid
23.9 ± 1.65
17.1 ± 1.87
41.0 ± 1.76
0.0012 ± 0.0001
Naproxen sodium
29.5 ± 0.91
35.9 ± 2.12
65.4 ± 1.52
0.0033 ± 0.0001The photocatalytic activity of Fe-TiO2-HNT was better than that of TiO2. This observation can be attributed to a number of reasons. Irradiation of Fe-TiO2-HNT with a light source resulted in the generation of electron-hole pair. Having in mind that the doping of TiO2 was done using FeCl3, Fe3+ can act as electron and hole trapping sites [19]. The photogenerated electrons may reduce Fe3+ to Fe2+, while the photogenerated holes may oxidize Fe3+ to Fe4+. The loss of d5 electronic configuration that arose from the reduction of Fe3+ to Fe2+ made Fe2+ unstable. Fe2+ formed may further be oxidized by the photogenerated holes back to Fe3+. As already stated, other researchers have reported that Fe3+ may act as both photogenerated electron and hole trapping sites [42]. As a result, Fe3+ may be oxidized to Fe4+ by the photogenerated holes. Fe4+ formed may be reduced back to Fe3+ and result in the formation of hydroxyl radicals that may enhance the photodegradation efficiency. The Ti3+/Ti4+ energy level lies close to the Fe2+/Fe3+ energy level. As a result, electrons trapped by Fe3+ can easily be transferred to Ti4+ at the surface of TiO2. Fe3+ therefore serves as an efficient electron trapping site [43]. The trapping of electrons by Fe 3+ may reduce the electron-hole recombination, and as a result, it may improve the photocatalytic activity. In addition, the presence of HNTs also enhanced the photocatalytic activity. The SEM images in Figure 7 revealed that the Fe-TiO2 crystals grew on the surface of HNTs. An intimate contact therefore exists between Fe-TiO2 and HNTs. The photogenerated electrons will therefore not accumulate in Fe-TiO2 and will be transferred from Fe-TiO2 to HNTs due to unique charge characteristics of HNTs (Figure 10). Due to the surface charge characteristics of HNTs, HNTs also serve as electron trapping sites which further reduced the electron-hole recombination [44, 45]. The lumen of HNTs is slightly positively charged within the pH range of 2–8; hence, the photogenerated electrons can be attracted to the positively charged lumen. More holes will therefore be available to be utilized in the photodegradation of the organic pollutants.Figure 10
Schematic representation of the photocatalytic degradation of organic pollutants.The reactive species (RS) responsible for the Fe-TiO2-HNT photocatalytic degradation of methylene blue was examined using DMSO, benzoquinone, and ethylenediamine tera-acetic acid disodium (EDTA-2Na). DMSO can be used as ⋅OH scavenger, while EDTA-2Na and benzoquinone scavenge h+ and ⋅O2−, respectively [46]. The photodegradation efficiency after the addition of DMSO, benzoquinone, and EDTA-2Na were ca. 60.45%, 69.27%, and 51.13%, respectively, when compared to 71.86% in which no scavenger was added. This suggests that the main RS responsible for the photocatalytic degradation of methylene blue by Fe-TiO2-HNTs are ⋅OH and h+. Since Fe3+ and HNTs serve as electron trapping sites, more h+ will be available for the photo-oxidation of the water-soluble dye. The photodegradation process is further enhanced by the unique adsorption properties of HNTs. Since HNTs are good adsorbents for methylene blue and rhodamine B and the Fe-TiO2 crystals grew on the surface of HNTs, the diffusion path of the pollutants to the surface active sites available for photocatalysis was reduced. This is expected to enhance the photodegradation efficiency.


### 3.4. Adsorption and Equilibrium Kinetics

The potential of the TiO2-Fe-HNT composite to adsorb MB and RhB dye molecules from solution was examined using the kinetics and isotherm models. The pseudo-first-order (PFO) (equation (2)), pseudo-second-order (PSO) (equation (3)), and Intraparticle models (equation (4)) were used to describe the adsorption kinetics of the MB and RhB dye molecules [41, 47–49]:(2)logqe−qt=logqe−k12.303t,(3)tqt=1k2qe2+tqe,(4)qt=kit0.5+Ci,where qt (mg/g) represents the quantity of dye adsorbed at time “t” in minutes, qe represent the equilibrium dye adsorption capacity (mg/g), k1,k2,andki represent the rate constant of PFO, PSO, and intraparticle models, respectively.The adsorption data were fitted unto these models with the best-fit model were selected based on the highest,R2, values as shown in Figures 11(a) and 11(b). From the study, the adsorption mechanisms were controlled by the PSO model with the highest R2 values of 0.9961 and 0.9993 as compared to the PFO model with R2 values of 0.9875 and 0.9589 for MB and RhB, respectively. The PSO model adsorption indicates that the adsorption of both MB and RhB molecules were controlled by chemisorption as a result of chemical bonds obtained through the exchange of electrons between the TiO2-Fe-HNTs and cationic dye molecules [50].Figure 11
Kinetic and sorption models describing the adsorption process. (a) Pseudo-first order. (b) Pseudo-second order. (c) Intraparticle. (d) The Langmuir and (e) Freundlich models.
(a)
(b)
(c)
(d)
(e)Moreover, the rate-controlling steps of the adsorption process were further described using the intraparticle model as shown in Figure11(c). Overall, the graph depicts a two-phase total adsorption process of the dye molecules, showing a rapid instantaneous adsorption within the first 30 minutes followed by a gradual second phase which attains saturation. The excellent adsorption process was enhanced by the aqueous dispersity of the composite provided by the TiO2-Fe crystals on the outer surface of the tubular structured HNTs as seen in Figure 7. Also the surface of the TiO2 is enriched with hydroxyl group which provides a negatively charged surface for efficient adsorption of cationic dyes [51]. In addition, the cylindrical hollow-shaped halloysite made up of several packed aluminosilicate sheets causes the inner lumen to be positively charged due to its aluminum oxide bonds, while its outer surface carries negative charges due to its silica surface. This unique bivalent surface morphology makes halloysite an excellent adsorbent to attract negative charge pollutant into its lumen and positive charge pollutant onto its outer surface [28, 52]. The presence of numerous vacant adsorption sites and dye molecules caused the establishment of a higher concentration gradient between the water-soluble dyes and TiO2-Fe-HNTs, resulting in an increase in the driving force and transfer of MB and RhB molecules from solution to readily bind with the vacant active sites present within the lumen and on the surface of the HNT. This resulted in higher instantaneous adsorption resulting from increased electrostatic attraction between the vacant adsorption sites and the dye molecules. This fast adsorption kinetics was observed in the first stage with steeper slopes during 30 minutes of adsorption. The first stage of the adsorption process obtained steeper slopes of 0.10233 and 0.14032 for MB and RhB, respectively. The adsorption of RhB attained a higher PSO rate constant as compared to MB. This indicates that the RhB molecules displayed a faster adsorption kinetic with a higher adsorption rate than MB molecules indicating that the structure of the halloysite and the nature of the molecules played a major role in the adsorption process. Though both MB and RhB are cationic dyes, MB dyes are regarded as purely cationic in nature due to the positive diethylamine group present in its molecules, while the zwitterionic nature of RhB molecules possesses both positive diethylamine and negative carboxylic group within its molecules [51]. The adsorption of MB molecules only occurred on the surface of the HNT, while the bivalent surface of the HNT allowed the adsorption of the positive diethylamine group onto the surface and the carboxylic group into the lumen of the HNT, making the RhB molecules to be preferentially adsorbed [28, 51, 52].The second stage of the dye adsorption process represented a more gradual adsorption process with rate constants of 0.05229 and 0.02903 for MB and RhB, respectively, till equilibrium was reached. The gradual diffusion process indicates that there was a lower concentration gradient and more of the vacant active sites were occupied by dye molecules leaving few active sites on the outer surface and within the lumen of the HNTs for effective adsorption. The intraparticle model also indicates that the adsorption process was controlled by film, pore, or surface diffusion or a combination during the adsorption phenomenon on the pore surface [49]. The few vacant active sites were occupied by the dye molecule until all adsorption sites got occupied implying that there was virtually no further adsorption taking place at equilibrium.Moreover, the Langmuir (equation (5)) and Freundlich (equation (6)) isotherm models were further used to analyze the interaction and adsorption capacity of the TiO2-Fe-HNTs to adsorb MB and RhB dye molecules [53, 54]:(5)Ceqe=1QmKL+CeQm,(6)lnQe=lnKF+1nflnCe,where Qm (mg/g) represents the maximum adsorption capacity at monolayer coverage and KF (mg/g) and KL (L/mg) represent the Freundlich and Langmuir adsorption constant (mg/g), respectively. Ce (mg/L) and Qe represent the concentration of the adsorbate and the adsorption capacity of the adsorbent at equilibrium, respectively. In addition, the heterogeneity factor, n, in the Freundlich model within the range of 1 to 10 confirms a favorable adsorption. A smaller value of 1/n (below unity) indicates a high adsorption intensity. Also, the separation factor, RL (equation (7)) within the range value of 0 <RL < 1 indicates the favorability of the Langmuir isotherm model during the adsorption process:(7)RL=11+KLCo,where Co (mg/L) represent the initial concentrations of adsorbate. The Langmuir equation fitted the adsorption of MB molecules with a higher R2 value of 0.9969, while the Freundlich equation fitted the adsorption of RhB molecules with a higher R2 value of 0.9757. The Langmuir isotherm model indicates that the adsorption sites were identical and energetically charged where each adsorption site was responsible for adsorbing one dye molecule. This provided a homogeneous monolayer coverage of MB molecules on the surface of the adsorbent [55]. In addition, Qm of the TiO2-Fe-HNTs was found to be 8.2 mg/g and 25.9 mg/g for MB and RhB, respectively. On the contrary, the Freundlich isotherm best fitted the RhB adsorption mechanism which assumes a nonuniform distribution and a heterogeneous multilayer adsorption coverage of the adsorbed RhB molecules as a result of the adsorption of positive diethylamine group onto the surface and the carboxylic group in the lumen of the HNT [54]. The separation factor RL lied within the range 0 < RL < 1, and heterogeneity factor, n, obtained as 1.1786 shows a higher affinity binding and a successful and favorable adsorption of MB and RhB molecules by the adsorbent [53]. The summary of parameters obtained from the kinetics and equilibrium models is presented in Table 2.Table 2
Summary of data fitted onto various kinetic models.
Models
Methylene blue (MB)
Rhodamine B (RhB)
Pseudo-first-order kinetic model
R2
0.988
0.959
K1 (min−1)
0.112
0.151
Qe (mg/g)
4.158
1.675
Pseudo-second-order kinetic model
R2
0.996
0.9993
K2 (g min/mg)
0.458
0.9523
Qe (mg/g)
4.280
10.268
Intraparticle diffusion model
K1 (min0.5 g·mg−1)
0.102
0.140
K2 (min0.5 g·mg−1)
0.052
0.029
Langmuir isotherm model
R2
0.997
0.430
Qm (mg/g)
8.169
25.867
KL
0.270
10.756
RL
0.649–0.270
0.0092–0.444
Freundlich isotherm model
R2
0.949
0.976
KF
6.085
2.184
n
6.661
1.179


## 3.1. Structural, Thermal, and Functional Groups Characterization

The XRD patterns of TiO2 and TiO2-Fe-HNT composite are shown in Figure 4. The crystallographic planes (101), (004), (200), (105), (211), (204), and (215) are characteristics peaks of anatase TiO2 (JCPDS No. 21-1272). It was observed that the crystallinity of TiO2 decreased upon the addition of Fe and HNTs. This was deduced from the broadening of the peaks. An increase in the amount of the Fe decreased the crystallinity of the synthesized composite. This is because the Fe atoms acted as defects in TiO2. This distorted the crystal structure of TiO2 and hindered the crystallinity of TiO2 [36]. The characteristic peak (110) of HNTs was observed.Figure 4
X-ray diffraction pattern of TiO2 and TiO2-Fe-HNTs (“∗” and “#” represents the crystallographic planes of TiO2 and HNTs, respectively).The XRD confirmed the synthesis of the TiO2-Fe-HNT composite. The characteristic peak of Fe was not observed probably due to the relatively lower concentration of Fe used in the synthesis of the composite.Thermogravimetric analysis of the synthesized TiO2, TiO2-Fe-HNTs, and HNTs was performed, and the results are presented in Figure 5. All the plots depict a multistage decomposition with relatively stable intermediate products. TiO2 recorded two significant weight losses. There was ca. 5.1 wt.% loss when TiO2 was heated from 60 to 250°C and ca. 1 wt.% when heated from 250 to 400°C. The weight of TiO2 remained constant at heat treatments above 400°C. The total weight loss after heat treatment was ca. 6.1%. On the contrary, HNTs recorded ca. 5.3 wt.% loss at temperatures between 60 and 400°C. Increasing the temperature from 400 to 540°C resulted in an additional 11.3 wt.% loss. A total of 16.6 wt.% loss was recorded for HNTs. The TGA curve for the composite TiO2-Fe-HNTs was a combination of the curves of HNTs and TiO2. The total weight loss recorded over the entire temperature at which the heat treatment was carried out was ca. 12.5%. The TGA curve of the composite confirmed the presence of TiO2 and HNTs.Figure 5
TGA curves for HNTs, TiO2, and TiO2-Fe-HNTs.The FTIR spectrum of the prepared TiO2-Fe-HNT composite showed three additional peaks that were not observed in the spectrum of TiO2 (Figure 6), suggesting the presence of HNTs in the composite. The broad band below 1000 cm−1 was ascribed to the stretching vibration of Ti-O [37, 38]. The strong peak at 1623 cm−1 corresponds to the symmetric vibration of absorbed water [39]. The broad peak at 3373 cm−1 was attributed to the stretching vibrations of the O-H functional group. The presence of HNTs in the composite material resulted in the appearance of three other peaks at 912, 1118, and 1032 cm−1. The band at 912 cm−1 resulted from the deformation vibration of the inner O-H groups present in HNTs. Stretching vibration of Si-O-Si was observed at 1032 cm−1, whereas apical Si-O stretching mode resulted in a peak at 1118 cm−1 [32].Figure 6
FTIR spectra of TiO2 and TiO2-Fe-HNTs.


## 3.2. Structural Morphology, Elemental, and Optical Characterization

The structural morphology and elemental characterization of the composite was investigated through SEM-EDX analysis. The SEM images and the energy dispersive X-ray spectroscopy analysis are presented in Figure7.Figure 7
Scanning electron microscopy images of (a) HNTs, (b) TiO2-1 wt.% Fe-50 wt.% HNTs at low magnification, (c) TiO2-1 wt.% Fe-50 wt.% HNTs at high magnification. (d) The secondary reference image for the area EDX data was acquired (TiO2-1 wt.% Fe-50 wt.% HNTs) and (e) EDX spectrum of TiO2-1 wt.% Fe-50 wt.% HNTs.
(a)
(b)
(c)
(d)
(e)The scanning electron microscopy images of HNTs showed tubular structure with an uneven length and diameter of the tubes (Figure7(a)). The morphology and structure of the HNTs did not change as observed in the SEM images of the composite (TiO2-1 wt.% Fe-50 wt.% HNTs). However, agglomeration of the individual tubes can be seen in Figure 7(b). This is probably due to the growth of TiO2-Fe crystals on the outer surface of the HNTs. It was observed at a relatively higher magnification (Figure 7(c)) that the TiO2-Fe crystals settled on the outer surface of the HNTs. This is expected to enhance the aqueous dispersity of the composite as HNTs have been reported to possess high dispersity in water [30]. Enhanced aqueous dispersity will increase the proximity between water-soluble pollutants and the composite and result in improved photocatalysis and adsorption. The EDX analysis as shown in Figures 7(d) and 7(e) established that Al, Si, Fe, O, and Ti are present in the composite material. The atomic ratio of Al to Si is ca. 1. From the chemical formula of HNTs Al2Si2O5OH4⋅nH2O, nSi/nAl is 1. This was confirmed from the table in Figure 7. In addition, the wt.% of Fe in the composite material used for this analysis was 1%. This was also confirmed by the EDX results which detected Fe (0.83 wt.%).Since TiO2 is a well-known photocatalyst, the effect of HNTs and Fe on the optical property of the TiO2-Fe-HNT composite was examined using diffused reflectance spectroscopy (DRS), and the result is presented in Figure 8.Figure 8
DRS analysis of TiO2 and TiO2-Fe-HNT composite.TiO2 recorded ca. 100% reflectance in the visible region of the solar spectrum as can be seen in Figure 8. This implies that TiO2 did not absorb visible light radiation. However, the reflectance by TiO2 was almost zero in the ultraviolet region, implying that TiO2 is active in the UV region of the electromagnetic spectrum. Addition of Fe and HNTs resulted in a decrease in the reflectance in the visible light region and a slight increase in reflectance in the UV region, meaning that the composite was photoactive in both ultraviolet and visible regions of the solar spectrum. The percentage of light absorption in the visible region increased with increasing wt.% of Fe. The optical band gap (BG) was also estimated using the Kubelka–Munk method. The estimated BG was 3.31, 3.24, and 3.2 eV for TiO2, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-2 wt.% Fe-50 wt.% HNTs, respectively. The Fe and halloysite nanotubes did not influence to a greater extent the optical band gap of the composite.


## 3.3. Adsorption and Photocatalysis of TiO2 and TiO2-Fe-HNT Composite

The potential usage of the synthesized nanocomposite in water purification was examined through adsorption and photocatalytic processes using model pollutants such as water-soluble dyes (rhodamine B and methylene blue), pharmaceutical drug waste (naproxen sodium), and pesticide (imidacloprid). The results for the adsorption and photocatalytic studies using methylene blue are presented in Figure1. The first thirty minutes represent the adsorption of methylene blue by TiO2 and TiO2-Fe-HNTs in the dark. The light was turned on after 30 minutes. The adsorption of methylene blue increased with increasing amount of HNTs used in the synthesis of the composite.TiO2 adsorbed ca. 4% of the methylene blue. Within the same period (30 minutes in the dark), TiO2-1 wt.% Fe-25 wt.% HNTs, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-1 wt.% Fe-60 wt.% HNTs adsorbed 28, 50, and 60% of methylene blue, respectively. Methylene blue is positively charged when in solution. On the contrary, HNTs possess a negatively charged surface. The high adsorption recorded resulted from the electrostatic interaction between the cationic methylene blue and the anionic surface of HNTs. After the adsorption process, the composite-methylene blue solution was irradiated with tungsten halogen lamp to start the photocatalysis process. The photocatalytic efficiency of TiO2-Fe-HNTs was observed to be lower than that of the pristine TiO2. As revealed in the SEM images, the TiO2 crystals were deposited onto the HNTs. Adsorption of the methylene blue dye by HNTs therefore enhanced the proximity between the pollutant and the TiO2-Fe photocatalyst. This is expected to enhance the photocatalytic activity. However, excessive adsorption of methylene blue may block the active sites of TiO2-Fe from light (solar energy) and hinder photocatalysis. The decrease in photocatalytic activity of the composite material is therefore attributed to the excessive adsorption of the cationic methylene blue by the anionic HNTs. The adsorbed methylene blue blocked the light from reaching the photoactive active sites of the TiO2-Fe. It can also be seen in Figure 1 that the photocatalytic efficiency of the composite material was fairly comparable even with increasing amount of HNTs used in the synthesis. As can be seen in the summary presented in Figure 1(b), the removal of the methylene blue is attributed to the combined effect of photocatalysis and adsorption. The total removal % of methylene blue by the TiO2-Fe-HNT composite was therefore higher than that by TiO2. However, it must be mentioned that the amount of the dye degraded through photocatalysis (as can be seen in Figure 1(b)) was relatively smaller when compared to the amount adsorbed.The photocatalytic and adsorption potential of the composite material was again examined using rhodamine B dye. This study is necessary because the adsorption studies conducted revealed that rhodamine is not “excessively” adsorbed by HNTs. The photocatalytic effect of TiO2-Fe-HNTs is expected to be more pronounced with rhodamine B than with methylene blue. The results are presented in Figure 2. TiO2 adsorbed ca. 11% of the rhodamine B dye, while TiO2-1 wt.% Fe-25 wt.% HNTs, TiO2-1 wt.% Fe-50 wt.% HNTs, and TiO2-1 wt.% Fe-60 wt.% HNTs adsorbed 20.7, 22.2, and 19.3%, respectively. Unlike methylene blue, the adsorption of rhodamine B by the composite was almost the same, irrespective of the amount of HNTs used in the synthesis of the composite. This suggests that the adsorption mechanism of methylene blue by the composite may be different from that of rhodamine B. The photodegradation efficiency of the composite material was higher than that of the pristine TiO2. However, it can be seen in Figure 2(b) that, unlike methylene blue (Figure 1), for all the composite materials studied, removal of rhodamine B through photocatalysis was significantly higher than removal through adsorption. This reveals that though the removal of the pollutant is through photocatalysis and adsorption, the dominant process may be dependent on the nature of the pollutants and its interaction with the HNTs.The pseudo-first-order rate constants for the photodegradation of methylene blue and rhodamine B dyes were calculated from equation (1), and the results are presented in Figure 3(a):(1)lnCtCo=kt.Generally, for the degradation of rhodamine B, the rate constants of the TiO2-Fe-HNT composites were significantly higher than that of TiO2. However, for the degradation of methylene blue, the rate constant increased in the order T-1Fe-25 H < T < T-1Fe-50 H < T-1Fe-60 H. It was also observed that the rate constants for the photodegradation by the composite were higher for rhodamine B than methylene blue. The TiO2 crystals were deposited onto the HNTs. The pronounced adsorption of methylene blue by the composite material resulted in the complete coverage of TiO2. This might have reduced the absorption of photons by TiO2 resulting in decreased photocatalytic activity and the smaller rate constant values obtained. Clay materials including HNTs have been reported to be poor adsorbents of rhodamine B dyes. This has been ascribed to the molecular size of rhodamine B and the nature of the interaction between rhodamine dyes and clays [40]. The lower adsorption of rhodamine B ensured that the TiO2 surface was not completely covered by the pollutants, and this allowed for enhanced absorption of photons. This enhanced the photocatalytic activity towards rhodamine B, and hence, the relatively higher rate constant values were observed. Therefore, to develop an efficient photocatalytic-adsorption process, a system must be designed in such a way that the amount of pollutants adsorbed does not exceed a threshold that may render the photocatalyst inactive.The potential application of the synthesized composite in adsorbing and degrading other pollutants was examined using naproxen sodium (pharmaceutical drug) and imidacloprid (pesticides). The composite-pollutant solution mixture was stirred in the dark for 24 hours. The mixture was then irradiated with light, and the photodegradation was examined over specified time intervals. The UV-Vis absorption spectrum of these pollutants after adsorption and UV-Vis light irradiation are presented in Figure9. The Fe-TiO2-HNT composite adsorbed approximately 29.5% and 23.9% of the naproxen sodium and imidacloprid (Figure 9 and Table 1). The photocatalytic degradation of the naproxen sodium was more rapid than that of imidacloprid with a recorded photocatalytic degradation of 35.9% in 4 hours. The overall removal (adsorption and photocatalysis) efficiency was 65.4 and 41% for naproxen sodium and imidacloprid, respectively. The difference in the observed photocatalytic activity resulted from the differences in the structural stabilities of imidacloprid and naproxen sodium. Naproxen sodium is negatively charged in solution. HNTs have been reported to possess a slightly negatively charged outer surface and a positively charged inner lumen (lumen is the empty cylindrical space in the HNTs) [41]. This may result in an electrostatic interaction between the naproxen sodium and HNTs. It is obvious from Figure 9 and Table 1 that Fe-TiO2-HNTs have the potential to remove (by adsorption and photodegradation) pesticides and pharmaceutical drug wastes from water bodies. It is however noted that unlike methylene blue and rhodamine B, a relatively longer time would be required for the photodegradation. Again, this may be due to the differences in the interaction between the pollutants and the Fe-TiO2-HNTs as well as the structural stabilities of the various pollutants.Figure 9
Adsorption and photodegradation of (a) naproxen sodium and (b) imidacloprid. (c) Comparing the adsorption and photodegradation of imidacloprid and naproxen sodium by the Fe-TiO2-HNT composite.
(a)
(b)
(c)Table 1
Summary of the adsorption and photodegradation efficiency.
Pollutant
Removal by adsorption (%)
Removal by photocatalysis (%)
Overall removal (%)
Photocatalysis rate constant (min−1)
Imidacloprid
23.9 ± 1.65
17.1 ± 1.87
41.0 ± 1.76
0.0012 ± 0.0001
Naproxen sodium
29.5 ± 0.91
35.9 ± 2.12
65.4 ± 1.52
0.0033 ± 0.0001The photocatalytic activity of Fe-TiO2-HNT was better than that of TiO2. This observation can be attributed to a number of reasons. Irradiation of Fe-TiO2-HNT with a light source resulted in the generation of electron-hole pair. Having in mind that the doping of TiO2 was done using FeCl3, Fe3+ can act as electron and hole trapping sites [19]. The photogenerated electrons may reduce Fe3+ to Fe2+, while the photogenerated holes may oxidize Fe3+ to Fe4+. The loss of d5 electronic configuration that arose from the reduction of Fe3+ to Fe2+ made Fe2+ unstable. Fe2+ formed may further be oxidized by the photogenerated holes back to Fe3+. As already stated, other researchers have reported that Fe3+ may act as both photogenerated electron and hole trapping sites [42]. As a result, Fe3+ may be oxidized to Fe4+ by the photogenerated holes. Fe4+ formed may be reduced back to Fe3+ and result in the formation of hydroxyl radicals that may enhance the photodegradation efficiency. The Ti3+/Ti4+ energy level lies close to the Fe2+/Fe3+ energy level. As a result, electrons trapped by Fe3+ can easily be transferred to Ti4+ at the surface of TiO2. Fe3+ therefore serves as an efficient electron trapping site [43]. The trapping of electrons by Fe 3+ may reduce the electron-hole recombination, and as a result, it may improve the photocatalytic activity. In addition, the presence of HNTs also enhanced the photocatalytic activity. The SEM images in Figure 7 revealed that the Fe-TiO2 crystals grew on the surface of HNTs. An intimate contact therefore exists between Fe-TiO2 and HNTs. The photogenerated electrons will therefore not accumulate in Fe-TiO2 and will be transferred from Fe-TiO2 to HNTs due to unique charge characteristics of HNTs (Figure 10). Due to the surface charge characteristics of HNTs, HNTs also serve as electron trapping sites which further reduced the electron-hole recombination [44, 45]. The lumen of HNTs is slightly positively charged within the pH range of 2–8; hence, the photogenerated electrons can be attracted to the positively charged lumen. More holes will therefore be available to be utilized in the photodegradation of the organic pollutants.Figure 10
Schematic representation of the photocatalytic degradation of organic pollutants.The reactive species (RS) responsible for the Fe-TiO2-HNT photocatalytic degradation of methylene blue was examined using DMSO, benzoquinone, and ethylenediamine tera-acetic acid disodium (EDTA-2Na). DMSO can be used as ⋅OH scavenger, while EDTA-2Na and benzoquinone scavenge h+ and ⋅O2−, respectively [46]. The photodegradation efficiency after the addition of DMSO, benzoquinone, and EDTA-2Na were ca. 60.45%, 69.27%, and 51.13%, respectively, when compared to 71.86% in which no scavenger was added. This suggests that the main RS responsible for the photocatalytic degradation of methylene blue by Fe-TiO2-HNTs are ⋅OH and h+. Since Fe3+ and HNTs serve as electron trapping sites, more h+ will be available for the photo-oxidation of the water-soluble dye. The photodegradation process is further enhanced by the unique adsorption properties of HNTs. Since HNTs are good adsorbents for methylene blue and rhodamine B and the Fe-TiO2 crystals grew on the surface of HNTs, the diffusion path of the pollutants to the surface active sites available for photocatalysis was reduced. This is expected to enhance the photodegradation efficiency.


## 3.4. Adsorption and Equilibrium Kinetics

The potential of the TiO2-Fe-HNT composite to adsorb MB and RhB dye molecules from solution was examined using the kinetics and isotherm models. The pseudo-first-order (PFO) (equation (2)), pseudo-second-order (PSO) (equation (3)), and Intraparticle models (equation (4)) were used to describe the adsorption kinetics of the MB and RhB dye molecules [41, 47–49]:(2)logqe−qt=logqe−k12.303t,(3)tqt=1k2qe2+tqe,(4)qt=kit0.5+Ci,where qt (mg/g) represents the quantity of dye adsorbed at time “t” in minutes, qe represent the equilibrium dye adsorption capacity (mg/g), k1,k2,andki represent the rate constant of PFO, PSO, and intraparticle models, respectively.The adsorption data were fitted unto these models with the best-fit model were selected based on the highest,R2, values as shown in Figures 11(a) and 11(b). From the study, the adsorption mechanisms were controlled by the PSO model with the highest R2 values of 0.9961 and 0.9993 as compared to the PFO model with R2 values of 0.9875 and 0.9589 for MB and RhB, respectively. The PSO model adsorption indicates that the adsorption of both MB and RhB molecules were controlled by chemisorption as a result of chemical bonds obtained through the exchange of electrons between the TiO2-Fe-HNTs and cationic dye molecules [50].Figure 11
Kinetic and sorption models describing the adsorption process. (a) Pseudo-first order. (b) Pseudo-second order. (c) Intraparticle. (d) The Langmuir and (e) Freundlich models.
(a)
(b)
(c)
(d)
(e)Moreover, the rate-controlling steps of the adsorption process were further described using the intraparticle model as shown in Figure11(c). Overall, the graph depicts a two-phase total adsorption process of the dye molecules, showing a rapid instantaneous adsorption within the first 30 minutes followed by a gradual second phase which attains saturation. The excellent adsorption process was enhanced by the aqueous dispersity of the composite provided by the TiO2-Fe crystals on the outer surface of the tubular structured HNTs as seen in Figure 7. Also the surface of the TiO2 is enriched with hydroxyl group which provides a negatively charged surface for efficient adsorption of cationic dyes [51]. In addition, the cylindrical hollow-shaped halloysite made up of several packed aluminosilicate sheets causes the inner lumen to be positively charged due to its aluminum oxide bonds, while its outer surface carries negative charges due to its silica surface. This unique bivalent surface morphology makes halloysite an excellent adsorbent to attract negative charge pollutant into its lumen and positive charge pollutant onto its outer surface [28, 52]. The presence of numerous vacant adsorption sites and dye molecules caused the establishment of a higher concentration gradient between the water-soluble dyes and TiO2-Fe-HNTs, resulting in an increase in the driving force and transfer of MB and RhB molecules from solution to readily bind with the vacant active sites present within the lumen and on the surface of the HNT. This resulted in higher instantaneous adsorption resulting from increased electrostatic attraction between the vacant adsorption sites and the dye molecules. This fast adsorption kinetics was observed in the first stage with steeper slopes during 30 minutes of adsorption. The first stage of the adsorption process obtained steeper slopes of 0.10233 and 0.14032 for MB and RhB, respectively. The adsorption of RhB attained a higher PSO rate constant as compared to MB. This indicates that the RhB molecules displayed a faster adsorption kinetic with a higher adsorption rate than MB molecules indicating that the structure of the halloysite and the nature of the molecules played a major role in the adsorption process. Though both MB and RhB are cationic dyes, MB dyes are regarded as purely cationic in nature due to the positive diethylamine group present in its molecules, while the zwitterionic nature of RhB molecules possesses both positive diethylamine and negative carboxylic group within its molecules [51]. The adsorption of MB molecules only occurred on the surface of the HNT, while the bivalent surface of the HNT allowed the adsorption of the positive diethylamine group onto the surface and the carboxylic group into the lumen of the HNT, making the RhB molecules to be preferentially adsorbed [28, 51, 52].The second stage of the dye adsorption process represented a more gradual adsorption process with rate constants of 0.05229 and 0.02903 for MB and RhB, respectively, till equilibrium was reached. The gradual diffusion process indicates that there was a lower concentration gradient and more of the vacant active sites were occupied by dye molecules leaving few active sites on the outer surface and within the lumen of the HNTs for effective adsorption. The intraparticle model also indicates that the adsorption process was controlled by film, pore, or surface diffusion or a combination during the adsorption phenomenon on the pore surface [49]. The few vacant active sites were occupied by the dye molecule until all adsorption sites got occupied implying that there was virtually no further adsorption taking place at equilibrium.Moreover, the Langmuir (equation (5)) and Freundlich (equation (6)) isotherm models were further used to analyze the interaction and adsorption capacity of the TiO2-Fe-HNTs to adsorb MB and RhB dye molecules [53, 54]:(5)Ceqe=1QmKL+CeQm,(6)lnQe=lnKF+1nflnCe,where Qm (mg/g) represents the maximum adsorption capacity at monolayer coverage and KF (mg/g) and KL (L/mg) represent the Freundlich and Langmuir adsorption constant (mg/g), respectively. Ce (mg/L) and Qe represent the concentration of the adsorbate and the adsorption capacity of the adsorbent at equilibrium, respectively. In addition, the heterogeneity factor, n, in the Freundlich model within the range of 1 to 10 confirms a favorable adsorption. A smaller value of 1/n (below unity) indicates a high adsorption intensity. Also, the separation factor, RL (equation (7)) within the range value of 0 <RL < 1 indicates the favorability of the Langmuir isotherm model during the adsorption process:(7)RL=11+KLCo,where Co (mg/L) represent the initial concentrations of adsorbate. The Langmuir equation fitted the adsorption of MB molecules with a higher R2 value of 0.9969, while the Freundlich equation fitted the adsorption of RhB molecules with a higher R2 value of 0.9757. The Langmuir isotherm model indicates that the adsorption sites were identical and energetically charged where each adsorption site was responsible for adsorbing one dye molecule. This provided a homogeneous monolayer coverage of MB molecules on the surface of the adsorbent [55]. In addition, Qm of the TiO2-Fe-HNTs was found to be 8.2 mg/g and 25.9 mg/g for MB and RhB, respectively. On the contrary, the Freundlich isotherm best fitted the RhB adsorption mechanism which assumes a nonuniform distribution and a heterogeneous multilayer adsorption coverage of the adsorbed RhB molecules as a result of the adsorption of positive diethylamine group onto the surface and the carboxylic group in the lumen of the HNT [54]. The separation factor RL lied within the range 0 < RL < 1, and heterogeneity factor, n, obtained as 1.1786 shows a higher affinity binding and a successful and favorable adsorption of MB and RhB molecules by the adsorbent [53]. The summary of parameters obtained from the kinetics and equilibrium models is presented in Table 2.Table 2
Summary of data fitted onto various kinetic models.
Models
Methylene blue (MB)
Rhodamine B (RhB)
Pseudo-first-order kinetic model
R2
0.988
0.959
K1 (min−1)
0.112
0.151
Qe (mg/g)
4.158
1.675
Pseudo-second-order kinetic model
R2
0.996
0.9993
K2 (g min/mg)
0.458
0.9523
Qe (mg/g)
4.280
10.268
Intraparticle diffusion model
K1 (min0.5 g·mg−1)
0.102
0.140
K2 (min0.5 g·mg−1)
0.052
0.029
Langmuir isotherm model
R2
0.997
0.430
Qm (mg/g)
8.169
25.867
KL
0.270
10.756
RL
0.649–0.270
0.0092–0.444
Freundlich isotherm model
R2
0.949
0.976
KF
6.085
2.184
n
6.661
1.179


## 4. Conclusions

The TiO2-Fe-HNT photocatalyst-adsorbent nanocomposite was successfully synthesized by the hydrothermal method and was photoactive in both UV and visible regions of the electromagnetic spectrum. SEM images revealed that the TiO2-Fe crystals were formed on the outer surface of the HNTs. The synthesized photocatalyst-adsorbent material recorded an enhanced rhodamine B, methylene blue, naproxen sodium, and imidacloprid removal (photocatalysis and adsorption) efficiency when compared to pristine TiO2. This observation was ascribed to the enhanced aqueous dispersity, adsorption, and photoactivity of the nanocomposite. An adsorption equilibrium data fitted well with the pseudo-second order for both methylene blue and rhodamine blue dyes indicating that the adsorption was controlled by chemisorption. The Langmuir and Freundlich isotherm models further described the adsorption interaction of methylene blue and rhodamine B molecules, respectively. This study shows potential application of the synthesized TiO2-Fe-HNT nanocomposite in water treatment technologies.



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*Source: 4270310-2019-12-12.xml*