Improved charge transfer and enhanced visible light photocatalytic activity of Bi₂O₃@Fe-MOF for degradation of Rhodamine B and Triclopyr

2.1 Chemicals used

Bismuth nitrate $[Bi{\left(N{O}_3\right)}_3.5H_{2}O,99.9\%]$ , Ferric nitrate $[Fe{\left(N{O}_3\right)}_3.9H_{2}O,99.9\%]$ , Terephthalic acid (1, 4 benzenedicarboxylic acid, 98%), N, N dimethylformamide (DMF, 99.5%), and ${\mathit{HNO}}_3$ were taken from CDH chemicals. Urea ${[NH}_{2}CO{NH}_2,99.5\%]$ was utilized as the fuel. Sigma Aldrich, India, was the supplier of the model contaminants Rhodamine B and Triclopyr.

2.2 Fabrication of Bi₂O₃ and Bi₂O₃@Fe-MOF nanocomposites photocatalysts

2.3 Characterizations

To analyze the phase and structural attributes of the produced materials, SmartLab 3 kW-Rigaku instrument was used to obtain their high-quality XRD patterns. To access morphology, the field emission scanning electron microscopy technique (FESEM) (7610F Plus/JEOL) was employed. Energy dispersive x-ray spectroscopy (EDX) coupled with FESEM was used to confirm the elemental composition. Optical properties and band gap characteristics were investigated using Shimadzu-3600 plus double beam UV–vis spectrophotometer. The separation of photogenerated charge carriers was studied by Photoluminescence (PL) spectroscopy. The spectra were acquired using a Hitachi F-7000 fluorescence spectrophotometer. Furthermore, the zeta potential (ζ) of the synthesized materials was measured using a (Zetasizer Ver. 7.13) analyzer.

2.4 Photocatalytic activity procedure

To evaluate the photocatalytic degradation ability of the synthesized compounds, Rhodamine B dye and Triclopyr pesticide were employed as the representative pollutants. The degradation experiment involved the addition of 10 $\mathrm{m}\mathrm{g}$ of the photocatalyst into a 100 $\mathrm{m}\mathrm{l}$ aqueous solution of the pollutant (10 ppm). Subsequently the mixture was stirred under dark conditions until adsorption–desorption equilibrium was reached. The stirring process was continued for 60 min at a controlled temperature of 25 °C. After that, a visible light lamp (250 W) was turned on and oxygen was supplied continuously. At regular intervals, approximately 5 $\mathrm{m}\mathrm{l}$ of the experimental solution was taken out for centrifugation, facilitating the separation of photocatalyst from the solution. The remaining concentration of the dye and pesticide in the collected solution was determined using a double beam UV–visible spectrophotometer by measuring the absorbance at their respective wavelengths of maximum absorption (553 and 295 nm respectively). The percentage degradation efficiency (η) was calculated utilizing the following equation: $$\eta =\frac{C_0-C_t}{C_0}\ast 100$$ where C₀ and C_t represent the concentration of organic pollutant at the initial time and after time t, respectively.

Furthermore, the impact of pH on the degradation process was investigated across the pH range of 3–11. To understand the involvement of reactive oxygen species, scavenger’s studies were conducted, aimed at evaluating the various reactive species.

3.1 XRD analysis

The crystalline structure of the synthesized materials was studied by recording the high quality XRD patterns (Fig. 1). XRD patterns were provided the evidence of the crystalline nature of the synthesized materials as depicted in the Fig. 1. The diffraction peaks obtained at 2θ values 5.06°, 8.90°, 10.20°, 16.40°, 18.84°, 24.22°, and 35.74° agree with the previously reported works (Hao et al., 2022; Yan et al., 2019) and confirm the successful synthesis of Fe-MOF. Sharp peaks below 10° indicated the formation of well crystalline MOF material. For Bi₂O₃, the obtained diffraction positions were in agreement with the β-tetragonal Bi₂O₃ phase (PDF NO. 27-50). The distinct peaks characteristic of Fe-MOF are also obtained in the patterns of the composite materials with Bi₂O₃. This observation indicates that the structural integrity of the Fe-MOF framework remains unaffected after the incorporation of Bi₂O₃ nanoparticles. To determine the crystallite sizes of these materials, Scherrer’s equation was employed (Kumari et al., 2020). The calculated sizes were found to be 10.24, 8.85, 9.10, 8.85, 9.10 nm for Bi₂O₃, Fe-MOF, Bi₂O₃@Fe-MOF-1, Bi₂O₃@Fe-MOF-2, and Bi₂O₃@Fe-MOF-3, respectively.

Close

Fig. 1

XRD patterns of Fe-MOF, Bi₂O₃, and Bi₂O₃@Fe-MOF composites.

Export to PPT

3.2 FESEM and EDX

Surface micrographs of the synthesized samples were obtained using FESEM to study their morphology. Fig. 2A depicts FESEM images of Bi₂O₃, Fe-MOF and Bi₂O₃@Fe-MOF-3 composite. The agglomeration of pure Bi₂O₃ nanoparticles was possibly caused by electrostatic interactions, resulting in an irregular morphology (Koli et al., 2021). Fe-MOF material exhibited the rough surface with pores which is favorable for binding adsorbate molecules. The material was composed of a numerous accumulated nanoparticle, resulting in void spaces/sites i.e. pores among them. Within the composite material, similar micrographs depicted the attachment of Bi₂O₃ nanoparticles to the surface of Fe-MOF. The EDX spectrum in Fig. 2B shows all the relevant peaks for the constituents of the synthesized materials, with no additional peaks detected. Further, investigation into the elemental distribution through elemental mapping of the synthesized compounds further supported the uniform distribution of the elements.

Close

Fig. 2

(A) FESEM images of (a) Bi₂O₃, (b) Fe-MOF, and (c) Bi₂O₃@Fe-MOF-3 composite and (B) EDX constitution and elemental mapping of the synthesized compounds.

Export to PPT

3.3 UV–Vis DRS

UV visible spectra were recorded to study the optical response and band gap energies of the prepared materials. As shown in the Fig. 3A, Bi₂O₃ shows an absorption edge at a shorter wavelength (414 nm) than Fe-MOF (855 nm). Absorption bands of the composites were detected in the region 800–850 nm. The band gap energies were evaluated according to the relation: $$E_g=\frac{1240}{{\lambda }_g}$$ where ${\lambda }_g$ represents the main absorption edge wavelength in $\mathrm{n}\mathrm{m}$ . The calculated values of band gaps were 2.99, 1.45, 1.48, 1.47, and 1.46 eV for Bi₂O₃, Fe-MOF, Bi₂O₃@Fe-MOF-1, Bi₂O₃@Fe-MOF-2, and Bi₂O₃@Fe-MOF-3, respectively. It is observed that with increase in Bi content in MOF, band gap values were gradually increased.

Close

Fig. 3

(A) UV–Vis DRS spectra of all the synthesized nanomaterials and (B) FTIR spectra of all the synthesized materials.

Export to PPT

3.4 FTIR

Structures of the fabricated materials were further confirmed with FTIR spectra recorded in the range 500–4000 cm⁻¹ (Fig. 3B). The Fe-MOF has characteristic peaks at 543, 753, 1020, 1384, 1574, and 1660 cm⁻¹. The peak appeared at 543 cm⁻¹ is due to the Fe—O bond vibration (Gecgel et al., 2019). Peak 753 cm⁻¹ corresponds to the C—H bond bending vibration of the benzene ring and at 1020 cm⁻¹ is assigned to the C—O—C bond. Peaks obtained at 1384 and 1574 cm⁻¹ confirmed the carboxyl group while at 1660 cm⁻¹ is due to the C⚌O bond stretching vibration. For Bi₂O₃, vibrational band at 543 cm⁻¹ is originated from Bi—O bond while at 818 cm⁻¹ belongs to the vibration of Bi—O—Bi bonds (Motakef-Kazemi and Yaqoubi, 2020; Yang et al., 2018). The peak at 1384 cm⁻¹ corresponds to the stretching vibration of Bi—O bond (Astuti et al., 2020a). In addition, peak at 3437 cm⁻¹ indicates the stretching vibration of O—H bond of the adsorbed water molecules (Yang et al., 2018). All the characteristic peaks of Fe-MOF were retained in the Bi₂O₃@Fe-MOF composites indicating the existence of stable crystal structure of Fe-MOF in the composites.

3.5 PL spectra

It is widely recognized that the PL signals arise due to the recombination of electron-hole pair’s within semiconductor materials (Sharma et al., 2021). Intensity of a peak in a photoluminescence spectrum is directly related to the rate of electron-hole recombination in the sample being studied. In the sample with high recombination rates, the PL peak has a high emission intensity, indicating that excited electrons and holes are recombined rapidly and vice versa (Fan et al., 2014). PL emission spectra of the different synthesized materials were recorded at room temperature (Fig. 4). Among all the synthesized samples, Bi₂O₃ exhibited the highest emission intensity i.e. maximum recombination of charge carriers while Bi₂O₃@Fe-MOF-3 had lowest emission intensity in the spectrum as shown in the inset Fig. 4. Minimum peak intensity of the composite Bi₂O₃@Fe-MOF-3 indicated the highest separation of the photogenerated electron holes pairs and hence the maximum photocatalytic efficiency of this material towards degradation of organic pollutants.

Close

Fig. 4

Photoluminescence spectra of Bi₂O₃, Fe-MOF, and Bi₂O₃@Fe-MOF composites.

Export to PPT

4.1 Catalytic efficiency and degradation kinetics

The ability of the synthesized materials to degrade RhB and TC under visible light irradiation was studied. The results obtained are depicted in the Fig. 5(Aa and Ba) for both the pollutants, respectively. It can be seen that Bi₂O₃ exhibited lower adsorption and photodegradation activity for both the pollutants (50.87% for RhB and 68.89% for TC in 120 min). Fe-MOF showed strong adsorption capacity under dark conditions and moderate photodegradation activity for both the pollutants, respectively. With Fe-MOF, removal efficiencies obtained were 78.99 and 77.04% for RhB and TC, respectively. Composites Bi₂O₃@ Fe-MOF showed lower adsorption capacity for both the pollutants which may be ascribed to the decreased surface areas of the composite materials as compared to the Fe-MOF (Tang et al., 2020). However, the photocatalytic activity of the Bi₂O₃@ Fe-MOF composites was found significantly increased than the both pristine Bi₂O₃ and Fe-MOF. Out of all the composites, Bi₂O₃@ Fe-MOF-3 exhibited the highest degradation efficiency for both the pollutants (91.79 and 88.42% respectively in 120 min). The enhanced photoactivity of the composite materials is attributed to the Bi₂O₃/Fe-MOF heterojunctions owing to which effective separation and movement of the charge carriers become successful (Wu et al., 2022). Further, Fig. 5(Ab and Bb) showed the photodegradation kinetics of both the pollutants. Langmuir-Hinshelwood model was employed to study the kinetics of degradation where pseudo-first-order reaction kinetics was followed (Danish and Muneer, 2021). Curve fitting of this model was utilized to evaluate rate constants and regression constants that are tabulated in Tables S1 & S2.

Close

Fig. 5

(A) (a) Photocatalytic removal and (b) kinetic study of RhB degradation and (B) (a) Photocatalytic removal and (b) Degradation kinetics of TC by prepared samples.

Export to PPT

4.2 pH of the medium and scavengers study

The pH of the initial solution can affect the number of available adsorption sites, surface charge, and hence binding and reactivity of the photocatalyst towards organic pollutants. To examine the effect of pH conditions for the synthesized compounds, initial solution pH was selected 3, 5 (acidic), 7 (neutral), and 9, 11 (basic). Fig. 6(Aa and Ba) depict the degradation in different pH mediums of both the pollutants (RhB and TC), respectively. For RhB, when pH increased from 3 to 5, degradation also increased up to ∼99% in 120 min. Further, rise of pH from 7 to 11 lead gradual decrease in photocatalytic activity (Fig. 6Aa). This can be explained by the electrical charges of RhB molecules and the electrostatic interactions with the photocatalyst surface that dominate in the adsorption process. In highly acidic environment (pH = 3), catalyst surface becomes protonated (zeta potential = + 2.48 mV) and repels away the cationic molecules of the dye. Thus, degradation efficiency decreases at pH = 3 than pH = 5. Further, under acidic conditions, the existence of H⁺ ions result in a huge production of ^•OH radicals, hence resulting in a high photodegradation rate of organic pollutants as compared to the neutral media (Nguyen Thi et al., 2021). Moderate degradation was obtained at pH = 7. At higher pH from 9 to 11, cationic dye is adsorbed more strongly on the negatively charged catalyst surface (zeta potential = − 13.4 mV) and more adsorption of the dye molecules effectively shields the surface of the catalyst and prevented the contact of the catalytic surface with light, resulting in a decreased photocatalytic efficiency (Nguyen et al., 2018).

Close

Fig. 6

Photocatalytic decontamination of (A) RhB and (B) TC involving (a) effect of different pH mediums and (b) effect of addition of different scavengers on Bi₂O₃@Fe-MOF-3.

Export to PPT

On the other hand, Triclopyr showed maximum removal under neutral conditions and decreased activity in both acidic and alkaline conditions (Fig. 6Ba). This can be attributed to the charge of the photocatalyst surface and acid dissociation constant of Triclopyr (pK_a = 2.93) under different pH values of the solution. At acidic pH, pesticide molecules are more likely to exist in their undissociated form which result in lesser adsorption and photocatalytic activity (Poulios et al., 1998). In basic pH conditions, the removal efficiency decreased due to the repulsive forces between the negatively charged catalyst and the anions. In addition, hindrance caused by the excess OH⁻ ions also prevents the absorption of light and generation of the reactive species that are needed for an efficient degradation of the pollutants.

Trapping experiments for the different reactive species were performed to examine the role of these species in the photodegradation reaction. The scavengers like ethylene diamine tetra-acetic acid (EDTA), p-benzoquinone (BQ) and isopropanol (IPA) were used to examine the involved reactive species. BQ, IPA and EDTA are highly reactive to superoxide anion radicals (^•O₂), hydroxyl radicals (^•OH), and holes respectively. Fig. 6(Ab and Bb) clearly shows that the presence of all scavengers hinders the decomposition of the organic pollutants. EDTA exhibits the highest scavenging performance with maximum decline in efficiency of the photocatalyst against both RhB and TC, suggesting major role of holes species in the reaction mechanism for oxidation of pollutants under visible light irradiation. Whereas, lesser effect on degradation rate by IPA and BQ for RhB, suggested lower participation of hydroxyl radicals and superoxide anion radicals in the degradation process. On the other side, the reaction rate of photocatalytic oxidation of TC was significantly inhibited by both IPA and BQ addition, indicating that hydroxyl radicals (^•OH) and superoxide anion radicals (^•O₂) also played their role in the photodegradation process of the pollutant.

4.3 Mechanism of degradation

The efficiency of photocatalysis in composite structures depends on the production, separation, and migration of electron-hole pairs (Hu et al., 2023). On the basis of energy level diagram, possible photocatalytic mechanism for degradation of organic pollutants with Bi₂O₃@Fe-MOF composites is illustrated in the Fig. 7. The valence band and conduction band of Bi₂O₃ lies at higher positive potentials than the LUMO and HOMO potentials of Fe-MOF (Hao et al., 2022; Lan et al., 2022). Therefore, the Z-scheme charge transfer mechanism is a possible explanation for the observed results (Raizada et al., 2019). Under the visible light illumination, both the materials i.e. Bi₂O₃ and Fe-MOF get excited and generate electrons and holes respectively. The photogenerated electrons in the higher energy conduction band of Bi₂O₃ migrate to the low energy valence band of Fe-MOF owing to the internal field at the heterojunction interface. The quick combination of these electrons with the holes species of Fe-MOF, boosts up the interfacial separation of charged carriers. Ultimately, electrons in the conduction band of Fe-MOF and holes in the valence band of Bi₂O₃ remain reserved. According to the literature reported values of the conduction band of Fe-MOF being more negative than the O₂/^•O₂ reduction potential (−0.33 eV), the electrons in LUMO of Fe-MOF can transform the O₂ molecules into ^•O₂ radicals. In the similar manner, holes in the valence band of Bi₂O₃ produce ^•OH radicals from H₂O molecules oxidation [H₂O/^•OH (2.4 eV)] as indicated by their respective redox potential values. The reactive species thus produced (^•O₂ and ^•OH radicals) then decompose the targeted organic pollutant molecules. Therefore, the improvement in the photocatalytic performance of the composite Bi₂O₃@Fe-MOF is ascribed to the effective generation and separation of photogenerated charge carriers at the Z-scheme heterojunctions.

Close

Fig. 7

Proposed Z- scheme photocatalytic charge transfer mechanism of Bi₂O₃@Fe-MOF composites for degradation of organic pollutants.

Export to PPT