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. 2024 Feb 12;9(8):8954–8972. doi: 10.1021/acsomega.3c06491

Hyperbranched-Polyethylenimine-Functionalized Coal Fly Ash as an Adsorbent for the Removal of Hexavalent Chromium and Reuse as a Dye Photocatalyst

Molahlegi Orienda Sebabi †,‡, Nonhlangabezo Mabuba †, Kriveshini Pillay †, Soraya Phumzile Malinga †,*
PMCID: PMC10905742  PMID: 38434891

Abstract

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Coal fly ash (CFA) has been extensively researched as an adsorbent for heavy metals, but its application is limited by its low adsorption capacity. The modification of CFA with hyperbranched polymers results in improved adsorption capacities. Hyperbranched polyethylenimine (HPEI) is a hyperbranched polymer containing NH2 groups that can bind with heavy metal ions through complexation or electrostatic interactions. In this study, CFA–HPEI adsorbents with various HPEI loadings (1–5%) were prepared and evaluated for the removal of Cr(VI). The successful incorporation of HPEI onto CFA was confirmed using Fourier transform infrared, elemental analysis, and X-ray photoelectron spectroscopy (XPS). The 3% CFA–HPEI loaded adsorbent resulted in optimum results when the effect of pH and adsorbent dosage was studied. The pseudo-second-order kinetics model best described the adsorption kinetics at an initial concentration of 20 mg/L. The Freundlich adsorption isotherm model best fitted the equilibrium adsorption data with a maximum adsorption capacity of 85.93 mg/g. The Cr-loaded adsorbent was reused as a photocatalyst to degrade methylene blue (MB) in the presence of visible light. The loaded adsorbent degraded 98.9% of MB (5 mg/L) within 180 min and was accompanied by compounds with m/z of 173 and 234, corresponding to the intermediate degradation of Azure A. The XPS analysis confirmed the coexistence of Cr(III) and Cr(VI) on the surface of the adsorbent. In addition, the loaded adsorbent exhibited good stability following MB degradation with no structural changes observed. Thus, CFA–HPEI adsorbents can be utilized as low-cost adsorbents for the remediation of toxic Cr(VI) from water and wastewater. The Cr-loaded CFA–HPEI adsorbent can be effectively reused as a photocatalyst, thus reducing environmental pollution.

1. Introduction

Water pollution due to heavy metals is one of the major problems currently facing the world. Effluent from a municipal wastewater treatment was found to contain Cu (0.18 mg/L), Zn (0.16 mg/L), and Pb (0.19 mg/L), which were above the permissible limits set for wastewater treatment plant effluent. Thus, the quality of the effluent that was released into the nearby river was severely impacted by this.1 The presence of Cr(VI) in water and wastewater is a great threat due to its toxic and carcinogenic nature.2 The main causes of Cr(VI) pollution are divided into two categories, namely, anthropogenic sources (industrial and agricultural activities, landfills, fuel combustion, and mineral processing) and natural sources (volcanic eruptions and weathering of rocks and minerals).3 In comparison to Cr(III), hexavalent chromium [Cr(VI)] compounds are known to be more soluble and mobile in aqueous media.4 Since Cr(VI) is a strong oxidizing agent, it can easily diffuse through cell membranes and oxidize biological molecules.5,6 Pollution of water bodies due to Cr(VI) may result in different health complications like chronic ulcers, hemorrhage, cancer, kidney damage, nervous system failure, and dermatitis.7,8 The United States Environmental Protection Agency recommends permissible limits of 0.5 mg/L in industrial effluent for Cr(VI).9

Several techniques have been applied to remove toxic heavy metals from water and wastewater. These comprise chemical coagulation, electrochemical treatment, flocculation, precipitation, reverse osmosis, biological processes, γ radiation, photo catalysis, and adsorption.10 The utilization of most of these processes is limited by either the cost of operation or the formation of unwanted byproducts.11 One of the most effective techniques for treating water is adsorption through low-cost adsorbents; this is because the adsorbents are inexpensive, efficient, simple to use, and easy to operate. High-quality activated carbon is still an expensive material despite its widespread use as an adsorbent in water and wastewater treatment plants. Low-cost adsorbents can be derived from natural materials (such as chitosan, zeolites. and clays), agricultural wastes (such as straw, bark, grout, seeds, peels, biochar, and biorefinery wastes), and industrial wastes [such as bagasse fly ash, slags, and coal fly ash (CFA)]. These adsorbents have drawn attention as they are (i) available locally, (ii) can be activated using low-cost and accessible reagents, and (iii) have high adsorption efficiencies. They possess a wide array of properties, which are significant for the removal of different pollutants from water and wastewater. For instance, chitosan, a linear polysaccharide, contains numerous hydroxyl groups, which act as adsorption sites for the removal of heavy metal ions. On the other hand, zeolites contain aluminum silicates, which induce properties for ion exchange and specific pore size which are crucial in the adsorption of heavy metals. Biochar, a carbon-based material resulting from the pyrolysis of biomass also attracted attention due to the presence of carboxyl and hydroxyl groups, which play a role in adsorption through ion exchange and electrostatic attraction.12 These low-cost adsorbents have all been studied for their capability to remove toxic chemicals from water and wastewater.13 Yogeshwaran and Priya (2021) used sugar cane bagasse for the removal of Cr(VI), Pb(II), and Zn(II) from water. High removal rates of the targeted heavy metals were observed at 95.65, 87.26, and 83.32% for Cr(VI), Pb(II), and Zn(II), respectively.14 Sugar cane bagasse mainly contains –OH groups from cellulose and hemicellulose, which facilitates adsorption through electrostatic attraction between the adsorbent and the positively charged heavy metal ions. In line with the latter study, bentonite clay was studied for the removal of Pb(II) and Cu(II) ions from wastewater. The results attained in this study indicated that bentonite could remove up to 87.7 and 89.0% of Pb2+ and Cu2+, respectively. The high removal rates of the targeted heavy metals were attributed to the presence of aluminum silicates in the bentonite clay. The adsorption process was mainly driven by the ion exchange between negatively charged aluminum and positively charged heavy metal ions.15 Studies have shown that CFA can also be utilized in different applications such as in agriculture and also as an emulsifier. Adilbekova et al. (2023) modified CFA with Tween 20 and effectively applied the modified CFA as an emulsifier for water-in-oil emulsions due to its negatively charged surface, which is capable of attracting positively charged N atoms of asphaltene molecules thereby destabilizing the thin film causing the water droplets to combine.16 The CFA modified with humic substances displayed good properties when applied in biological cultivations as soil conditioner due to the slow release of the humic substances during washing and the presence of the microbes in the material. The humic substances act as ligands for bioinorganic applications, which facilitates soil conditioning.17

In this study, CFA was applied because it is composed of Al2O3, SiO2, Fe2O3, CaO, MgO, and carbon for adsorption of Cr(VI). Its physical properties such as porosity, particle size distribution, and surface area are also crucial for the adsorption process.18−20 Although the direct application of CFA is beneficial for remediation of water from pollutants, it is limited by its low adsorption capacity and selectivity toward the analyte. Chemical modification of CFA involves the addition of functional groups on the surface of fly ash particles, thereby increasing its chemical reactivity or easy functionalization with other compounds to increase the density of surface-active sites. Different compounds have been evaluated for the activation and functionalization of CFA toward the removal of different pollutants from water and wastewater. These include metal oxides, carbon-based compounds, and amines. The effectiveness of Fe3O4-loaded fly ash composite for the remediation of Congo red dye from water was investigated. The Fe3O4-loaded fly ash composite was found to possess better magnetic characteristics, which resulted in improved adsorption capacity when compared with that of raw CFA and NaOH-treated CFA.21 Carbon-based compounds such as amine, polydopamine, and β-cyclodextrin were used to modify CFA to prepare polydopamine-β-cyclodextrin–CFA composite for the remediation of uranium from water. The maximum adsorption capacity of 536.7 mg/g was obtained with the removal efficiency of 95.6% at pH 5.0 through complexation, hydrogen bonding, and electrostatic interaction.22 On the other hand, CFA functionalized with amines such as 3-aminopropyltriethoxysilane have been applied for the adsorption of Cr(VI) from wastewater. Optimum adsorption conditions were obtained at pH 2 with a maximum adsorption capacity of Cr(VI) of 498.39 mg/g. The removal of Cr(VI) was found to be through electrostatic attraction between the protonated NH3+ of NH2–FAT and negatively charged HCrO4–.23

Hyperbranched polyethylenimine (HPEI), an imine-based hyperbranched polymer consisting of a diatomic core, interior branching units of polyethylenimine, and peripheral NH2 functional groups, has been used to functionalize different adsorbents to enhance their performance. Zeng et al. (2018) functionalized carboxymethyl chitosan with HPEI to remove Hg(II) from wastewater. The concentration of Hg(II) ions was reduced from 798.1 to 0.02 mg/L, which was well below the permissible limit (0.05 mg/L) set for industrial wastewater in China.24 This was due to the large number of –NH2, which acted as metal binding groups toward Hg(II) ions. Humin, a byproduct generated in the preparation of humic acid, was modified with HPEI for the remediation of Cd(II) from water. Humin can be applied as an adsorbent for the remediation of heavy metals, but its direct application is limited by its low adsorption capacity. To overcome this drawback, humin was functionalized with HPEI to enhance the available adsorption sites, thereby improving efficiency. The X-ray photoelectron spectroscopy (XPS) analysis show that the N atoms from HPEI played an important role in the adsorption of Cd(II), and this resulted in an adsorption capacity of 11.975 mg/g. The performance was almost 5 times than that of unfunctionalized humin.25 The removal of Ag+ ions and the recovery of Ag0 using polyethylenimine-modified potassium tungsten oxide (K2W4O13) as an adsorbent were investigated.26 Potassium tungsten oxide can act as an adsorbent as well as a reducing agent. The incorporation of PEI improved the removal efficiency of the adsorbent and its reduction properties. The significant improvement in efficiency is a result of the interaction between N atoms of the PEI and Ag+ ions, and electron transfer occurs resulting in the reduction of Ag+ ions to Ag0.26

In this study, CFA was modified with HPEI for the removal of Cr(VI) from water. The modification was made possible by first activating the CFA to introduce the −OH groups on the material, followed by adding glutaraldehyde, which is a cross-linker through the reaction of the aldehyde group on the glutaraldehyde and the −OH group on the CFA. The addition of HPEI resulted in the reaction between the aldehyde functional group and the N atom from the HPEI molecule. After the application of the CFA–HPEI adsorbent to remove Cr(VI), chromium is reduced to less toxic Cr(III) by electrons from HPEI (reducing agent causes transfer of electrons from the lone pair of electrons on the N atoms of the HPEI). The reduction step displays a novel method that alleviates Cr(VI) in wastewater. Another innovative aspect of this work is that the adsorbent is reused as the photocatalyst to degrade the methylene blue (MB) dye in wastewater due to the presence of Cr(III), which is safe for environment. The CFA–HPEI works as the multifunctional adsorbent and a photocatalyst which has a high potential for water treatment in municipalities and wastewater treatment plant.

2. Reagents and Methods

2.1. Reagents

Ammonium nitrate (NH4NO3, ≥99%), glutaraldehyde (Grade II, 25% in H2O), hydrochloric acid (HCl, 32 wt %), HPEI (MW 25,000 g/mol), magnesium sulfate heptahydrate (MgSO4·7H2O, ≥99%), and methanol [CH3–OH, high-performance liquid chromatography (LC) 99.9%] were all purchased from Sigma-Aldrich, South Africa). Potassium dichromate (K2Cr2O7, ≥99%), sodium chloride (NaCl, ≥99%) and sodium hydroxide (NaOH, ≥99%) were all purchased from Merck, South Africa, while MB (C16H18ClN3S·xH2O, AR) was purchased from Reflecta Laboratory Supplies. CFA was collected from a coal-fired power station in South Africa. All working solutions were prepared using deionized water.

2.2. Pretreatment, Activation, and Functionalization of CFA

The pretreatment of the CFA was performed according to the method reported by Chen et al. (2018) through washing with deionized water (1:6 m/v) while stirring at room temperature for 1 h.27 The resulting slurry was filtered off and dried in a 105 °C oven at for 12 h. The dried solid was ground into a fine powder and characterized. The CFA was mixed with NaOH (5:8 w/w) and calcined at a temperature of 300 °C for 3 h. The fused mass was allowed to cool to room temperature, ground, and washed with deionized water. The resulting slurry was filtered and dried for 12 h in an oven at 105 °C.28 CFA (20 g) was dispersed in methanol. The glutaraldehyde solution (1% w/v) was added dropwise to the mixture. The mixture was left to stand for 30 min under continuous stirring. The HPEI solution in methanol (1–5% w/v) was added. The mixture was stirred at 180 rpm for 24 h at 35 °C for cross-linking.29 CFA mainly contains Al2O3 and SiO2, which were converted to Al–(OH)4– and Si–(OH)4– by treatment with NaOH to introduce the −OH groups on the surface. The functionalization of the activated CFA was performed adopting the methods as outlined by Troung et al. (2020) and Xiao et al. (2010).20,30 The addition of glutaraldehyde resulted in a reaction between the aldehyde group and the –OH on the CFA forming hemiacetal, successively with another –OH resulting in a full acetal. The HPEI added reacted with the aldehyde functional group on the glutaraldehyde and the N atom on the HPEI as shown in Scheme 1.

Scheme 1. Reaction Scheme of (a) Aldehyde Group of Glutaraldehyde and the –OH on the CFA Forming Acetal, (b) Possible Cross-linking through the Reaction of the Aldehyde Group with the N Atom on the HPEI Resulting in the Formation of (c) CFA–HPEI Adsorbent.

Scheme 1

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2.3. Characterization

X-ray fluorescence (XRF) spectroscopy was used to determine the elemental composition of raw CFA. The sample was poured into an aluminum cup and pressed to form a pellet under high pressure (20 tons) for 60 s. MagiX PRO XRF spectrophotometer (Malvern Panalytical, United Kingdom) was used for analysis, and it was operated at 40 kV and 40 mA. The crystallinity measurements of the CFA were carried out using an X’Pert PRO X-ray Diffractometer (Malvern Panalytical, United Kingdom) using Ni-filtered Cu-Kα (λ = 0.1540 nm) radiation equipped with an X’Celerator detector. The generator was operated at 40 kV and 40 mA. The samples were scanned over 2θ degree angle of 4–80°. The surface chemical state of the adsorbents was determined using an ESCALAB 250Xi X-ray Photoelectron Spectrometer (Thermo Fisher Scientific, USA) equipped with a monochromatic Al kα X-ray source (1486.6 eV). The functional groups present were determined using a Fourier transform infrared (FTIR) spectrometer (PerkinElmer, Germany). Samples were combined with KBr at a dilution ratio of 1:99 to make pellets before FTIR analysis. The surface morphology of the materials was determined using a TESCAN VEGA3 scanning electron microscopy–energy-dispersive X-ray spectrometry (Czech Republic). The high voltage was operated at 20 kV. The surface area measurements were determined using Micromeritics ASAP 2460 (USA) Brunauer–Emmett–Teller (BET) surface area measurements by N2 adsorption at −196 °C. Samples were degassed at 90 °C for 9 h before analysis. The zeta potential measurements were conducted using a Malvern Panalytical Zetasizer Nano-ZS (United Kingdom). The samples were dispersed in deionized water prior to analysis. An elemental analyzer (FLASH 2000 series, Thermo Fischer Scientific, USA) was utilized to determine the elemental composition of the samples. The optical properties of the adsorbents were determined using a Shimadzu UV-2450 spectrometer in the diffuse reflectance spectroscopy (DRS) mode. The band gaps were calculated from the DRS data. The photodegradation products of MB were analyzed by LC–mass spectrometry (MS). A Waters Synapt G1 HDMS mass spectrometer with Waters Acquity UPLC (USA) was used.

2.4. Batch Experiments

2.83 g of potassium dichromate (K2Cr2O7) was dissolved in deionized water to make up a 1000 mg/L stock solution. The adsorption experiments were conducted in a thermostatic shaker set to 160 rpm. All experiments were conducted in triplicate and results reported with error bars. The effect of adsorbent dosage and pH was investigated to optimize the adsorption behavior (see the Supporting Information). Varying amounts (10–70 mg) of CFA and CFA–HPEI adsorbents (1–5%) were suspended in 20 mL of Cr(VI) solution (20 ppm) at 25 °C. A 20 mL aliquot of the 1000 ppm stock solution was accurately pipetted into a volumetric flask. The actual concentration was determined using a UV–vis spectrophotometer through the 1,5-diphenylcarbazide method.

The effect of pH was studied by varying the solution pH from 2 to 11. This was achieved by adjusting the solutions with either 0.1 M HCl or 0.1 M NaOH solution. The Cr(VI) concentrations of the solutions were measured with a UV–visible spectrophotometer (Shimadzu UV-1800) at 540 nm using the 1,5-diphenylcarbazide method.31 The removal percentage of Cr(VI) was calculated using eq 1.

2.4. 1

where C0 and Ce (mg/L) represent the initial and equilibrium concentrations of Cr(VI), respectively.32

The amount adsorbed at time t expressed as qt was calculated using eq 2.

2.4. 2

where qt (mg/g) is the amount of Cr(VI) adsorbed per unit mass of the adsorbent at time t, Ct(mg/L) is the Cr(VI) concentration at time, m(g) is the mass of the adsorbent, and V(L) is the volume of Cr(VI) solution. The adsorption capacity of the adsorbent at equilibrium, qe (mg/g), was calculated using eq 3.32

2.4. 3

2.5. Adsorption Kinetics

Based on the results obtained during optimization when studying the effect of pH and adsorbent dosage, the 3% loaded CFA–HPEI adsorbent was used to investigate the kinetics associated with the adsorption of Cr(V1). Adsorption kinetics is important in determining the rate-determining step in an adsorption process in order to provide an understanding into the design and modeling of an adsorption-based water treatment process. The adsorption kinetics data was evaluated using pseudo-first-order and pseudo-second-order kinetic models. The linear eqs 4 and 5 of the two models are displayed below

2.5. 4
2.5. 5

where t(min) is time and k1 and k2 are the rate constants (min–1).33

The rate-limiting step for the uptake of Cr(VI) was determined using the intraparticle diffusion model in eq 6.

2.5. 6

where ki (mg/g min0.5) represents intraparticle diffusion rate constant and C is the intercept related to boundary layer thickness.34

2.6. Adsorption Isotherms

The adsorption isotherms were performed at temperatures of 25, 35, and 45 °C. The concentration ranged from 20 to 500 mg/L at pH 2.0. The equilibrium adsorption data was evaluated using Langmuir and Freundlich adsorption isotherm models presented by eqs 7 and 8, respectively

2.6. 7

where qm (mg/g) represents the maximum adsorption capacity, b (L/mg) is the adsorption free energy, and Ce (mg/L) is the adsorbate concentration at equilibrium

2.6. 8

where KF (mg1–1/n L1/n g–1) is the Freundlich constants related to adsorption capacity and n (dimensionless parameter) is the adsorption intensity.35 The thermodynamics parameters for the adsorption of Cr(VI) by the CFA–HPEI (3%) adsorbent such as changes in enthalpy (ΔH0), entropy (ΔS0), and standard Gibbs free energy (ΔG0) were calculated using the following equations

2.6. 9
2.6. 10
2.6. 11

where Kc is the adsorption affinity at the lowest experimental Cr(VI) concentration, R denotes ideal gas constant (J/mol/K), and T is the absolute temperature (K).36

2.7. Interference Studies

The effect of other coexisting ions like Cl–, NO3–, SO42–, Cu2+, Fe2+, and Zn2+ on the adsorption of Cr(VI) was also investigated. The experiments were performed with 20 mL of 20 mg/L Cr(VI) solutions containing 10, 20, 30, 40, or 50 mg/L of Cl–, NO3–, and SO42– and 30 mg of the adsorbent. All solutions were adjusted to pH 2.0. The adsorption was allowed to reach equilibrium after which the residual Cr(VI) was determined. Table S1 illustrates the preparation of the solutions.

2.8. Reusability Studies

100 mg of the CFA–HPEI-Cr was transferred into a cylindrical glass reactor containing 500 mL solution of 5 mg/L MB solution. The mixture was allowed to equilibrate for 30 min in the dark before being irradiated with a visible lamp (250 HW). During irradiation, the mixture was agitated vigorously. At regular interval times, 10 mL aliquots were taken, filtered using 0.45 μm syringe filters, and analyzed using a UV–visible spectrophotometer to determine the residual MB. The maximum absorbance was obtained at a wavelength of 665 nm. The photocatalytic degradation was calculated using eq 12.

2.8. 12

The kinetics rate of degradation of MB by CFA–HPEI (3%) adsorbent and Cr-loaded CFA–HPEI (3%) adsorbent was determined using the pseudo-first-order eq 13.

2.8. 13

where C0 (mg/L) is the initial concentration at t0, Ct (mg/L) represents MB concentration during visible light irradiation at different time intervals t (min), and the rate constant (K) is the slope of the graph.37

2.9. Determination of Degradation Products Using LC–MS

The photodegradation products of MB were analyzed by LC–MS. A Waters Synapt G1 HDMS mass spectrometer with Waters Acquity UPLC (USA) was used. The LC was equipped with a Waters Acquity T3 C18 column (150 × 2.1 mm) operated at 60 °C. The MS was equipped with an electrospray ionization source operated in positive mode with a capillary voltage of 2.5 kV, cone voltage of 30 V, drying gas flow rate of 7.5 L/min, and an operating pressure of 8500 psi. The eluents used were A (water with 0.1% formic acid) and B (acetonitrile with 0.1% formic acid) set at linear gradient flow from 100% A to 2% A in 6 min at a flow rate of 0.4 mL/min with a sample injection volume of 10 μL.

2.10. Data Analysis

All experiments were reported as the mean of three replicates. The kinetics data and equilibrium isotherm models were evaluated using Origin Pro 8.5. The X-ray diffraction (XRD) data were processed with HighScore Plus 4.9, while XPS data were processed with Thermo Avantage 5.9.

3. Results and Discussion

3.1. Characterization of CFA and CFA–HPEI (1–5%) Adsorbents

3.1.1. XRF Analysis

The CFA was primarily composed of SiO2 (46%), Al2O3 (31%), CaO (5.9%), and Fe2O3 (2.4%). As depicted in Table 1, there were also small amounts of TiO2, MgO, P2O5, K2O, and Na2O present. Similar results were obtained for CFA collected from other power stations in South Africa.38−40 The elemental analysis results show that this CFA is class F fly ash as defined by the American Society for Testing Materials Standard C618.41

Table 1. Chemical Composition of CFA.
component %
SiO2 46
Al2O3 31
CaO 5.9
Fe2O3 2.4
TiO2 1.5
MgO 1.4
P2O5 1.0
K2O 0.72
Na2O 0.60
LOI 2.35
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3.1.2. XRD Analysis

Figure 1 depicts the XRD patterns for CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents. The results showed that CFA consists mainly of quartz (SiO2) and mullite (Al6Si2O13) as confirmed by ICSD 01-070-3755 and ICSD 01-083-1881, respectively. These results correspond with recent studies reported on the analysis of fly ash in South Africa.42,43 As reported, the activation of CFA with NaOH resulted in the destruction of the crystalline structure of CFA as shown by the sharp decrease in the intensity of the diffraction peaks of CFA, with no new crystalline phases formed.28,44 While the HPEI loading onto the activated CFA–NaOH was varied from 1 to 5%, the diffraction patterns obtained show that the variation in HPEI did not result in differences in the crystal structure of the activated CFA. They all showed low-intensity peaks associated with the presence of quartz and mullite.

Figure 1.

Figure 1

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XRD patterns for CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents.

3.1.3. SEM–EDS Analysis

The surface morphology of CFA, CFA–NaOH, and CFA–HPEI adsorbents was analyzed using SEM–EDS. The CFA particles were spherical in shape with varying particle sizes, with an average particle size of 4.05 μm, as shown in Figure 2. The surface morphology of CFA particles show relatively leathery surfaces, as depicted in Figure 3a.45,46 The energy-dispersive spectrometry (EDS) spectrum (Figure 3h) reveals that the CFA is primarily composed of O, Al, Si, Ca, and Fe. This data corresponds with the data obtained from the EDS analysis and is also similar to other studies reported on South African CFA.47,48 The alkali-activated CFA (Figure 3b) shows a change in the morphology of the CFA particles: the smooth spherical structure of CFA particles changed to irregular agglomerates with rough surfaces. The EDS (Figure 3i) displays the presence of Na that was attached to the surface of CFA during the activation process. After the functionalization of CFA–NaOH with HPEI, the EDS spectra for CFA–HPEI (1–5%) adsorbents (Figure 3j–n) show a significant increase in the quantity of carbon, which is attributed to the presence HPEI. An increase in HPEI loading on the CFA did not result in any significant change in the surface morphology of the CFA–HPEI (1–5%) adsorbents, as shown in Figure 3c–g.

Figure 2.

Figure 2

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Particle size distribution for CFA.

Figure 3.

Figure 3

Figure 3

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SEM micrographs for (a) CFA, (b) CFA–NaOH, (c–g) CFA–HPEI (1–5%) adsorbents, and EDS spectra for (h) CFA, (i) CFA–NaOH, and (j–n) CFA–HPEI (1–5%) adsorbents.

3.1.4. FTIR Analysis

FTIR analysis was carried out to identify the functional groups present in CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents and to monitor the change in intensity or shift in peak positions. As shown in Figure 4, the CFA shows a peak at 1096 and 551 cm–1, which depicts the presence of Si–O–Si symmetric and asymmetric stretch vibrations and Fe–O vibrations, respectively. CFA–NaOH shows a broad peak at 3440 cm–1 (Si–OH) and a peak at 1651 cm–1 due to the presence of –OH vibrations.49 The HPEI-functionalized CFA shows a change in the shape and intensity of the peaks due to the –OH vibrations (3440 cm–1) and Si–O–Si vibrations (992 cm–1) caused by the interaction of CFA–NaOH and HPEI. There are also new peaks at 2847 and 2971 cm–1 because of the presence of sp3-hybridized carbons in the HPEI chain. The additional peak at 1482 cm–1 is attributed to the presence of C–H bending vibrations, confirming the successful functionalization of CFA.49,50 In addition, the peak at 1651 cm–1, attributed to C=N stretch vibrations is enhanced, which further confirms the successful incorporation of HPEI onto CFA.51

Figure 4.

Figure 4

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FTIR spectra of CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents.

3.1.5. BET Analysis

The specific surface area of raw CFA was found to be 1.24 m2/g. The activation of CFA using NaOH resulted in a significant increase in specific surface area to 17.9 m2/g (Table 2). Similar results were reported by Meng et al. (2020) and Woolard et al. (2000) after treating CFA with NaOH. In both studies, the specific surface area was found to increase from 1.05 to 62.66 m2/g44 and from 1.58 to 23.6 m2/g.52 After functionalization of the activated CFA with HPEI, the specific surface area decreased to 15.3 m2/g for the CFA–HPEI (1%), with an increase from 15.0 to 16.6 m2/g as the CFA–HPEI loading increases (Table 2). The incorporation of HPEI in the CFA resulted in a decrease in the specific surface area, which is due to the blockage of pores by the HPEI.49 There is no significant difference in BET surface area for the 3-, 4-, and 5%-HPEI-loaded adsorbents, which suggests optimum loading. The CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents exhibited features of non porous or macroporous adsorbents with reversible BET type 2 isotherm (Figure 5).

Table 2. BET Surface Areas of CFA, CFA–NaOH, and CFA–HPEI (1–5%) Adsorbents.
  Surface area (m2/g)
CFA 1.25
CFA–NaOH 17.9
CFA–HPEI (1%) 15.3
CFA–HPEI (2%) 15.0
CFA–HPEI (3%) 16.2
CFA–HPEI (4%) 16.6
CFA–HPEI (5%) 16.1
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Figure 5.

Figure 5

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N2 adsorption–desorption isotherms of CFA, CFA–NaOH, and CFA–HPEI (1–5%) adsorbents.

3.1.6. Elemental Analyses

The elemental analyses of the raw CFA and the functionalized CFA were performed, and the results are given in Table 3. The raw CFA had an average of 3.86 and 0% of carbon and nitrogen, respectively. The functionalization of CFA with HPEI resulted in an increase in both carbon and nitrogen contents: the CFA–HPEI (1%) adsorbent had an average of 4.48 and 0.83% of carbon and nitrogen, respectively. An increase in HPEI loading to 2 and 3% resulted in an increase in the percentages of both carbon and nitrogen. Further increase in HPEI loading to 4 and 5% resulted in a decrease in carbon and nitrogen content: an average of 5.38% of carbon and 1.69% of nitrogen were obtained from the CFA–HPEI (4%) adsorbent, while the CFA–HPEI (5%) adsorbent had an average of 3.44 and 0.75% of carbon and nitrogen, respectively. The decrease in carbon and nitrogen content was caused by the hindrance effect of the –OH groups on the CFA by the bulky HPEI molecules that might have occurred as the percentage of HPEI increased. The presence of nitrogen in the functionalized CFA confirms the successful incorporation of HPEI onto CFA.29,53

Table 3. Elemental Analyses Results of CFA and CFA–HPEI (1–5%) Adsorbents.
  CFA CFA–HPEI (1%) CFA–HPEI (2%) CFA–HPEI (3%) CFA–HPEI (4%) CFA–HPEI (5%)
C 3.87 4.48 5.10 6.93 5.38 3.44
H 0.37 2.09 2.17 3.57 2.24 1.92
N   0.83 1.58 1.86 1.69 0.75
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3.2. Application of CFA–HPEI Adsorbents for the Removal of Hexavalent Chromium (Cr(VI))

3.2.1. Effect of Adsorbent Dosage on the Uptake of Cr(VI)

Adsorbent dosage is an important factor in adsorption studies as it influences the adsorption capacity of an adsorbent at a specific initial concentration.54 The adsorbent dosage ranged from 0.5 to 3.5 g/L (Figure 6). Raw CFA showed a low percentage removal of Cr(VI) with a maximum removal of 5.5% for 3.0 g/L adsorbent dosage. The maximum removal percentage was obtained from the CFA–HPEI (3%) and CFA–HPEI (4%) adsorbents with 98% removal obtained with adsorbent dosages of 1.5 g/L.

Figure 6.

Figure 6

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Effect of adsorbent dosage on the removal of Cr(VI) by CFA and CFA–HPEI (1–5%) adsorbents.

3.2.2. Effect of pH

pH is a critical factor for the adsorption of Cr(VI) because it affects the surface charge of the adsorbent and the Cr ionic speciation.55 The pH was varied from pH 2 to 11, as depicted in Figure 7. The maximum Cr(VI) removal was obtained at pH 2 for the CFA–HPEI (1–5%) adsorbent and at pH 3 for the CFA. The adsorbents showed removal percentages between 65 and 95%, while the raw CFA displayed a low removal percentage of about 10%. In acidic medium (pH 2–6), Cr(VI) exists as dichromate (Cr2O72) and bichromate (HCrO4) ions, which are major components in solution, while chromate (CrO42) ions dominate at pH above 6. Thus, the ions can be adsorbed on the positively charged surface through electrostatic attraction as shown by the zeta potential analysis in Figure 8.56,57 The point of zero charge for CFA was 3.62, indicating that the adsorbent carried a positive charge below pH 3.6 and a negative charge above pH 3.6, with the zeta potential decreasing from +20 to −10 mV as the pH was increased from 2 to 12. The functionalization of CFA with HPEI resulted in the point of zero charge slightly increasing to 4.8, 4.78, 4.6, 4.40, and 4.5 for CFA–HPEI (1–5%) adsorbents. This showed that HPEI functionalization resulted in an increase in the zero point of charge of the adsorbent because of the protonation of the amine groups on HPEI.58,59 Additionally, in acidic medium, the NH2 groups in HPEI are protonated and form NH3+ which will attract the opposite charge on the dominating negatively charged Cr(VI) species, resulting in the uptake being higher at lower pH values. As the pH was increased, the removal percentage decreased due to less protonation of the NH2 groups and also due to competition between the negatively charged Cr(VI) species and OH– groups.60

Figure 7.

Figure 7

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Effect of pH on the removal of Cr(VI) by CFA and CFA–HPEI (1–5%) adsorbents.

Figure 8.

Figure 8

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Zeta potential of CFA, CFA–NaOH, and CFA–HPEI (1%–5%) adsorbents.

3.2.3. Adsorption Kinetics

As depicted in Figure 9a, the influence of contact time on the removal of Cr(VI) by CFA–HPEI (3%) adsorbent was studied for three different initial concentrations (20, 50, and 75 mg/L). An increase in the initial concentration resulted in an increase in the amount adsorbed, with equilibrium attained after 120, 150, and 180 min for the three different initial concentrations, respectively. Initially, Cr(VI) ions migrate to the boundary layer, then diffuse through to the adsorbent surface. Lastly, they diffuse onto the surface of the adsorbent material.8 The 75 mg/L solution will require more contact time to reach equilibrium because of the higher number of Cr(VI) ions present. In addition, at the initial time of contact, there is a large number of vacant adsorption sites and also a high concentration gradient of Cr(VI) between the adsorbent and the bulk of the solution. As adsorption equilibrium was reached, the adsorption capacity remained unchanged because of the decrease in number of vacant sites available for adsorption and also a decrease in concentration gradient.61,62 The kinetics data for pseudo-first-order and pseudo-second-order kinetic models are displayed in Table 4. As shown in Figure 9b, the pseudo-second-order kinetic model better defined the adsorption of Cr(VI) with R2 values of 0.9836, 0.9868, and 0.9954 for Cr(VI) initial concentrations of 20, 50, and 75 mg/L, respectively. The qe values obtained using the pseudo-second-order kinetic model were comparable to qe experimental values. Adsorption kinetics are generally controlled by either external or intraparticle diffusion or both.63 If a plot of qt versus t0.5 gives a linear relationship and the graph passes through the zero point, intraparticle diffusion is the rate-determining step.64Table 5 presents the correlation coefficients for the graphs displayed in Figure 9c. As the concentration is increased from 20 to 70 mg/L, the correlation coefficient (R2) declines. Moreover, since the three graphs do not pass through the zero point this suggests that intraparticle diffusion is not the rate-determining step in the removal of Cr(VI) by the CFA–HPEI (3%) adsorbent.

Figure 9.

Figure 9

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(a) Effect of contact time on the adsorption of Cr(VI) by CFA–HPEI (3%) adsorbent at initial concentrations of 20, 50, and 75 mg/L, (b) linear fit for pseudo-second-order kinetic model, and (c) intraparticle diffusion model for adsorption of Cr(VI) by CFA–HPEI (3%) adsorbent.

Table 4. Kinetic Parameters for the Adsorption of Cr(VI) by CFA–HPEI (3%) Adsorbents.
  initial concentration (mg/L)205075
pseudo-first-order
linear fit      
R2 0.9448 0.9296 0.9505
k1 (min–1) 0.01199 0.004721 0.006379
qe (mg/g) 8.126 14.12 12.38
pseudo-second-order
linear fit      
R2 0.9836 0.9868 0.9954
k2 (g/mg/min) 0.002668 0.002025 0.002477
qe (mg/g) 13.28 27.67 32.45
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Table 5. Intraparticle Diffusion Model Parameters for the Adsorption of Cr(VI) by CFA–HPEI (3%) Adsorbent.
  intraparticle diffusion
initial concentration (mg/L) Ki (mg/g min) C (mg/g) R2
20 0.5120 4.562 0.9697
50 0.7478 14.89 0.9414
75 0.7951 19.66 0.9219
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3.2.4. Adsorption Isotherm Models

Adsorption isotherm experiments were performed to investigate the interaction between the adsorbent and adsorbate. The equilibrium adsorption data was studied at three different temperatures: 25, 35, and 45 °C. As shown in Figure 10a, the adsorption capacity increased with increasing temperature from 25 to 35 °C, indicating that the adsorption mechanism is endothermic. An increase in temperature to 45 °C did not result in any significant increase in adsorption capacity. The isotherm parameters obtained from the linear regression of the experimental data are summarized in Table 6. Higher correlation coefficient values (0.9804, 0.9921, and 0.9784) were obtained from the Freundlich adsorption isotherm as shown in Figure 10b, which suggests that this model fits the equilibrium adsorption data better than the Langmuir isotherm. The Freundlich isotherm model suggests that the surface of the adsorbent is heterogeneous.65 The value of n indicates the favorability of the adsorption process. It is stated that an n in the range of 2–10 shows good adsorption characteristics.66 In this study, the values of n obtained were 3.054, 2.698, and 2.962 for the three different temperatures, indicating that the adsorption process was favorable. The qm values obtained were 58.58, 85.93, and 67.03 mg/g at the three different temperatures studied. The comparison of adsorption capacity of CFA–HPEI (3%) adsorbent with that of other low-cost adsorbents is tabulated in Table 7. The data shows that the adsorbent performed much better than some of the adsorbents, while it was comparable in terms of amount adsorbed with other materials functionalized with PEI. A graph of ln Kc against Inline graphic is given in Figure 10c. ΔS0 and ΔH0 were calculated from the graph’s intercept and slope, respectively. Table 8 shows that as temperature increased, the negative value for ΔG0 increases. Together with a positive value for ΔH0, this confirms that the adsorption process was spontaneous. A positive value of ΔS0 indicates an increase in disorder at the solid–liquid interface.67

Figure 10.

Figure 10

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(a) Effect of temperature on the adsorption of Cr(VI) by CFA–HPEI (3%) adsorbent, (b) plot of log qe (qe, mg/g) against log Ce (Ce, mg/L), representing the linear Freundlich adsorption isotherm for Cr(VI) adsorption using CFA–HPEI (3%) adsorbent, and (c) graph to obtain the thermodynamic parameters for Cr(VI) adsorption using CFA–HPEI (3%) adsorbent.

Table 6. Adsorption Isotherm Parameters for the Adsorption of Cr(VI) onto the CFA–HPEI (3%) Adsorbent.
  temperature (°C)
  25 35 45
Langmuir isotherm      
linear fit      
R2 0.8512 0.9082 0.9058
b (L/mg) 15.52 12.87 14.25
qm (mg/g) 80.52 118.6 117.9
Freundlich isotherm      
linear fit      
R2 0.9804 0.9921 0.9784
N 3.054 2.698 2.962
KF (mg1–1/nL1/ng–1) 15.46 18.05 22.49
qm (mg/g) 58.58 85.93 67.03
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Table 7. Comparison of Maximum Adsorption Capacity of CFA–HPEI (3%) Adsorbent in Contrast with Other Low-Cost Adsorbent Materials Reported for the Adsorption of Cr(VI).
no adsorbent adsorption capacity (mg/g) isotherms references
1 alkali-modified CFA 5.976 Freundlich (68)
2 polypyrrole-coated secondary fly ash–iron composites 119.33 Langmuir (69)
3 natural akadama clay 4.290 Freundlich (54)
4 polyethylenimine-impregnated activated carbon 114.0 Elovich and Langmuir (70)
5 polyethylenimine–polyacrylamide–cellulose fibers 8.453 Langmuir (71)
6 polyethylenimine-functionalized magnetic montmorillonite clay 62.89 Langmuir (72)
7 polyethylenimine-modified activated-sludge-based adsorbent 86.96 Langmuir (73)
8 CFA–HPEI (3%) adsorbent 85.93 Freundlich this work
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Table 8. Thermodynamic Parameters for Cr(VI) Adsorption by CFA–HPEI (3%) Adsorbent.
temperature (°C) ΔG0 (kJ/mol) ΔS0 (kJ/mol/K) ΔH0 (kJ/mol)
25 –8.57 0.1530 37.07
35 –10.1    
45 –11.6    
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3.2.5. Interference Studies

Processes in the steel manufacturing industry produce wastewater containing Cr(VI) as well as other cations such as Cu2+, Fe2+, and Al3+ and anions such as Cl–, NO3–, and SO42–. The coexisting ions affect the binding force of the adsorbent and adsorbate during adsorption as it affects the potential and the width of the dual-layer around the adsorbent.74,75 Coexisting ions such as Cl–, NO3–, and SO42– could exhibit a significant competitive effect, while cations could contribute to a synergetic adsorption effect on the adsorption of Cr(VI).76 It is therefore critical to investigate the competitive effect of these coexisting ions on Cr(VI) remediation using CFA–HPEI (3%) adsorbent. As shown Figure 11, the concentration of the coexisting ions ranged from 10 to 50 mg/L. An increase in concentration of the coexisting ions had a slight effect on the adsorption of Cr(VI), with the 10 mg/L matrix resulting in 98% removal, while 50 mg/L resulted in 95% removal. The presence of positively charged ions is expected not to have an effect on the adsorption of Cr(VI) at pH 2 since Cr(VI) exists as a negatively charged ion (HCrO4) while the adsorbent surface is positively charged, resulting in repulsion. On the contrary, the presence of an anion like SO42– has the potential to decrease the uptake of HCrO4– because of the comparable ionic radius of SO42– and HCrO4–. The same ionic radius has comparable hydration energy,77 hence the observed decrease as the concentration of coexisting ions increased.

Figure 11.

Figure 11

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Effect of matrix on the adsorption of Cr(VI) using CFA–HPEI (3%) adsorbent.

3.2.6. Possible Mechanism for the Adsorption of Cr(VI) onto CFA–HPEI (3%) Adsorbent

The results obtained from varying the pH showed that the adsorption of Cr(VI) is pH-dependent with the maximum removal obtained at pH 2. At this pH, the NH2 groups on CFA–HPEI get protonated, forming NH3+, which attracts the negatively charged Cr(VI) species. The Cr-loaded CFA–HPEI (3%) was also analyzed using BET, SEM–EDS, XPS, FTIR, and XRD. The adsorbent ’s surface area decreased from 16.18 to 13.26 m2/g. The decrease is due to Cr(VI) adsorbed on the surface of CFA–HPEI, as shown by the EDS elemental mapping analysis depicted in Figure 12.

Figure 12.

Figure 12

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(a,c–j) EDS elemental mapping analysis of the Cr(VI)-loaded CFA–HPEI (3%) adsorbent and (b) EDS spectrum.

The XPS was used to determine the mechanism responsible for the adsorption of Cr(VI) by CFA–HPEI (3%) adsorbent. As shown in Figure 13a, the spectrum shows energy bands due to Al 2p (73.7 eV), Si 2p (101.7 eV), Ca 2p (347.2 eV), Ti 2p (358.3 eV), O 1s (531.2 eV), Fe 2p (711.2 eV), Na 1s (1071.7 eV), and Mg 1s (1303.8 eV), all attributed to the CFA, and C 1s (285.7 eV) and N 1s (400 eV) attributed to HPEI. The deconvoluted spectrum of N 1s as shown in Figure 13b shows the energy bands at 399.9 and 401.3 eV attributed to −NH– and −N+=, respectively, which are characteristic bands for HPEI. These results demonstrated that the amine groups from HPEI were successfully protonated, and Cr(VI) species which exist as negatively charged species in acidic medium could be effectively adsorbed on the protonated amino groups on surface of the CFA–HPEI (3%) adsorbent.78 The adsorption of Cr(VI) onto the CFA–HPEI (3%) adsorbent is confirmed by the appearance of peaks observed at 577.3 and 587.9 eV due to Cr 2p3/2 and Cr 2p1/2, respectively, as shown in Figure 13c. Kera et al. reported similar results for Cr(VI) adsorption using polypyrrole/2,5-diaminesulfonic acid composite.60 Furthermore, the deconvoluted spectrum of Cr 2p as shown in Figure 13d confirmed the presence of Cr(III) and Cr(VI) on the surface of the adsorbent with 77% of Cr(III), signifying reduction of the toxic Cr(VI) into less toxic Cr(III) which is less of a threat to the environment.79−82 The HPEI acted as a reducing agent by transferring electrons from the lone pair of electrons on the N atoms of the HPEI. This behavior was also observed by a study reported by Huang et al. (2019), where polyethylenimine-modified potassium tungsten oxide for the removal of Ag+ and the recovery of Ag0 was studied.26 Also, the intensity of the N 1s (400 eV) peak decreased significantly after adsorption (Figure S1), implying that the N atom plays a role in the removal of Cr(VI).83

Figure 13.

Figure 13

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(a) XPS spectrum of CFA–HPEI (3%) adsorbent before and after Cr(VI) adsorption, (b) N 1s core level spectrum before adsorption, (c,d) Cr 2p core level spectrum after adsorption.

The FTIR spectra (Figure 14) show a shift in wavenumber of the peaks associated with O–H and N–H vibrations from 3440 to 3416 cm–1. There is also a shift from 1655 to 1628 cm–1 for the peak related with O–H and N–H vibrations. The Si–O–Si peak has also shifted from 719 to 608 cm–1. This behavior was expected because of the interaction between the functional groups on the surface of CFA–HPEI (3%) adsorbent and Cr(VI).48 The XRD spectra in Figure 15 present the CFA–HPEI (3%) adsorbents before and after adsorption of Cr(VI). The peaks at 2θ degree angles of 20.93 and 29.59° associated with SiO2 and Al6Si2O13, respectively, are also observed on the diffraction pattern of the loaded adsorbent. Furthermore, no new peaks were observed in the FTIR and XRD spectra, implying that there were no changes in the structural integrity of the adsorbent after the removal process.

Figure 14.

Figure 14

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FTIR spectra of CFA–HPEI (3%) adsorbent before and after Cr(VI) adsorption.

Figure 15.

Figure 15

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XRD spectra of CFA–HPEI (3%) adsorbent before and after Cr(VI) adsorption.

3.2.7. Optical Properties of the Adsorbents

The optical properties of CFA–HPEI (3%) adsorbent and Cr-loaded CFA–HPEI (3%) adsorbent were obtained by conducting UV–vis diffuse reflectance measurements. As shown in Figure 16a,b, the clean CFA–HPEI (3%) adsorbent exhibited a band gap of 3.52 eV, while the Cr-loaded CFA–HPEI (3%) adsorbent showed a slight decrease with a band gap of 3.40 eV, respectively. The decrease in band gap of the Cr-loaded CFA–HPEI (3%) adsorbent is due to the existence of Cr adsorbed on the surface of the CFA–HPEI (3%) adsorbent. Generally, a smaller band gap enables greater absorption of visible light by the material, while photocatalytic reactions with materials which exhibit larger band gaps result in less absorption of visible light, thus lowering the photocatalytic properties.84 Due to its band gap value, the Cr-loaded CFA–HPEI (3%) adsorbent is therefore considered a better photocatalyst than the clean CFA–HPEI (3%) adsorbent.

Figure 16.

Figure 16

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Tauc plots of (a) CFA–HPEI (3%) adsorbent and (b) Cr-loaded CFA–HPEI (3%) adsorbent.

3.2.8. Reusability Study of Cr-Loaded CFA–HPEI (3%) Adsorbent in the Photocatalytic Degradation of MB

The Cr-loaded CFA–HPEI (3%) adsorbent was studied as a photocatalyst to degrade MB. MB solution (5 ppm; 500 mL) was used as working solution together with 100 mg of the adsorbent. The reaction was monitored for 180 min for both CFA–HPEI (3%) adsorbent and Cr-loaded CFA–HPEI (3%) adsorbent. The maximum absorbance of MB was detected at a wavelength of 665 nm with the intensity decreasing as the irradiation time was increased, as depicted in Figure 17a,b. Figure 17c shows 93.7 and 98.9% MB removal efficiency by CFA–HPEI (3%) adsorbent and Cr-loaded CFA–HPEI (3%) adsorbent, respectively. As depicted in Figure 17c, about 32 and 40% MB removal efficiencies are observed during equilibration time, which suggests that adsorption may be occurring between MB and CFA–HPEI (3%) adsorbent. The adsorption process may be through electrostatic interactions between the negatively charged surface of the CFA–HPEI (3%) adsorbent as confirmed by zeta potential analysis (Figure 8) and the positive charge of MB, hydrogen bonding between amine groups of HPEI and the sulfur atoms of MB, and π–π interactions between the aromatic structure of MB and the polymeric backbone of HPEI as well as the surface area of the adsorbent.85,86 The results reveal that the presence of Cr resulted in improvement in the photocatalytic activity of the material. Cr is reported to be one of the key photocatalytic materials used in the area of cleaning up environmental pollutants.37 According to the pseudo-first-order equation, the rates of MB photodegradation were found to be −0.01520 and −0.02527 min–1 for the clean and loaded adsorbent, respectively (Figure 17d). Table 9 depicts a comparison of the degradation of MB obtained in this study with other catalysts. The results show that the performance of the Cr-loaded CFA–HPEI (3%) adsorbent is comparable with that of the other catalysts studied. The adsorbent displayed good stability after degradation of MB. The FTIR and XRD illustrated no structural changes before and after application, as depicted in Figures S2 and S3. DRS analysis showed a slight decrease in band gap from 3.40 to 3.32 eV (Figures 16b and S4), confirming good photocatalytic properties of the material. Moreover, no traces of Cr(VI) could be detected in the treated MB solution as confirmed by the absence of the absorbance peak associated with Cr(VI) at 540 nm (Figure S7).

Figure 17.

Figure 17

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UV–vis spectra of photocatalytic degradation of MB by (a) CFA–HPEI (3%) adsorbent and (b) Cr-loaded CFA–HPEI (3%) adsorbent, and (c) percentage degradation of MB, and (d) rate constants.

Table 9. Comparison of Percentage Degradation of MB by Cr-Loaded CFA–HPEI (3%) Adsorbent with Other Catalysts.
no catalyst dye degradation time (min) degradation (%) refs
1 CFA heterogeneous Fenton catalyst MB 240 93.0 (86)
2 CFA-CoFe2O4 MB 60 99.0 (87)
3 3.7% anatase red mud geopolymer MB 720 94.6 (88)
4 CFA-based zeolite NaP1 MB 720 96.0 (89)
5 CFA@(r-GO/TNS)8 MB 70 55.0 (90)
6 MnFe2O4/coal fly ash  nanocomposite MB 30 100 (91)
7 coal fly ash based zinc ferrite (CFA/ZnFe2O4) composite MB 60 97.0 (92)
8 CuO/CFA based zeolite Na-X MB 180 80.2 (93)
9 Fe2O3–TiO2/fly-ash-cenosphere composite MB 60 86.81 (94)
10 CFA-HPEI–Cr adsorbent MB 180 98.9 this work
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3.2.9. Possible Mechanism for the Degradation of MB Using CFA–HPEI-Cr

As shown in Figure S5, the intensity of the MB peak appearing at a retention time of 5.1 min decreased as the irradiation time was prolonged, suggesting the successful degradation of MB in the presence of visible light. MS was used to detect the MB precursor mass and possible mass fragments resulting from the degradation process. The mass spectra of the samples collected after 30, 90, 150, and 180 min are given in Figure S6. The degradation of MB results in possible intermediate products with m/z of 270 (Azure B), 256 (Azure A), 242.2 (Azure C), and 228 (Thionin).95Figure S6 shows that after 90 min, the dominating mass is that of the precursor ion (284.1), while virtually none of the expected intermediate masses could be detected, suggesting that the MB could have disintegrated instead of following a slow and steady fragmentation process. This might be due to the presence of Cr since metals have large adsorption capability and also display strong photodegradation activity due to a decrease in electron–hole recombination. This leads to rapid degradation resulting in the intermediate masses not being detected in the solutions.96 After visible light irradiation for 150 min, the precursor ion (284.1) has decreased with the appearance of a compound with m/z of 173. After 180 min, the complete degradation of the precursor ion is accompanied by the appearance of a compound with m/z of 173 and 234, which are possible intermediate products of Azure A (256), as depicted in Figure 18.

Figure 18.

Figure 18

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Possible degradation pathway of MB by Cr-loaded CFA–HPEI (3%) adsorbent.

4. Conclusions

A CFA-based adsorbent was prepared and evaluated for the remediation of Cr(VI) from water. The CFA–HPEI adsorbents were prepared by functionalizing the CFA with different HPEI loadings (1–5%). The adsorbents were characterized using XRF, XRD, XPS, SEM–EDS, FTIR, EA, and BET. The 3%-HPEI-loaded adsorbent performed best with a maximum adsorption capacity of 85.93 mg/g at pH 2. The pseudo-second-order kinetics model best defined the adsorption kinetics, and the Freundlich adsorption isotherm model best described the equilibrium adsorption data. The XPS confirmed the existence of both Cr(VI) and Cr(III) on the surface of the adsorbent, suggesting that the remediation of Cr(VI) occurred through the electrostatic attraction of Cr(VI) onto the adsorbent followed by the reduction of Cr(VI) to Cr(III) by the CFA–HPEI (3%) adsorbent. The Cr-loaded adsorbent was reused as a photocatalyst for the degradation of MB with 98.9% degradation accompanied by adsorption within 180 min. The Cr-loaded CFA–HPEI adsorbent displayed strong photodegradation activity as it resulted in rapid degradation of MB, which was accompanied by the detection of products with m/z of 173 and 234 corresponding to intermediate degradation of Azure A. Furthermore, there was no structural change in the FTIR and XRD analyses performed before and after application, indicating that the adsorbent exhibited good stability following the degradation of MB. This study showed that the HPEI-functionalized CFA could be utilized as an adsorbent for the removal of Cr(VI) from water. In order to prevent additional environmental pollution, the Cr-loaded adsorbent can also be reused as a photocatalyst for the degradation of MB. Based on the promising results obtained from this study, other hyperbranched polymers will be explored for the functionalization of CFA targeting different emerging pollutants (i.e., pharmaceuticals and personal care products) and other heavy metal ions prevalent in water and wastewater. As part of ongoing work, the adsorbent will be supported on beads, which can be further used in columns for upscaling of the technology. This upscaled technology will then be tested against real water samples collected from pharmaceuticals, agricultural, and various manufacturing industries as HPEI polymers are effective in the removal of organic pollutants due to their hydrophobic cavities.

Acknowledgments

The South Africa/India Joint Science and Technology Research Collaboration (IND190911475982), Water Research Commission(WRC project no.: K52488//3), National Research Foundation-Competitive Support for Unrated Researchers (NRF-CSUR:SRUG210407592746), Center for Nanomaterials Science Research (CNSR: University of Johannesburg), Faculty of Science (University of Johannesburg), and Spectrum (The Analytical Facility of the Faculty of Science, University of Johannesburg) are all acknowledged by the authors for their support.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acsomega.3c06491.

  • Additional experimental details, methods, and figures (PDF)

The authors declare no competing financial interest.

Supplementary Material

ao3c06491_si_001.pdf (313.9KB, pdf)

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