Abstract

Granular activated carbon was doped with iron (Fe-AC) and was used to study the removal of Safranin O (SO) using Fe-AC/H2O2 system for reactive adsorption and Fe-AC for adsorption. Fe-AC and H2O2 doses were optimized to obtain maximum removal of SO. Maximum removal was found to be 96.1% after 5 h using 1.0 g/L Fe-AC and 5.0 mM hydrogen peroxide doses for 10 mg/L initial SO concentration. Kinetic study suggested the suitability of the pseudo first order model for reactive adsorption. The Langmuir isotherm explained well the sorption of SO onto Fe-AC. Parallel-pore-reactive-adsorption-model was applied and validated. By fitting the experimental data to the model, it is observed that the surface reaction rate coefficient, kr, was found to be 5 times that of the apparent rate constant, kapp. Parameters such as the external liquid film mass transfer coefficient, macro-pore and micro-pore diffusivities were estimated by regression analysis. Pore diffusion and surface reaction were found to be rate controlling for adsorption and reactive adsorption, respectively. An oxidative degradation of SO took place via hydroxylation and ring cleavage processes.

ABBREVIATIONS

     
  • A

    Total interfacial surface area, m2

  •  
  • a

    Interfacial surface area, m2

  •  
  • C

    Liquid phase concentration, mg/L

  •  
  • Molar liquid phase concentration, mol/m3

  •  
  • Initial liquid phase concentration, mg/L

  •  
  • Surface liquid phase concentration, mg/L

  •  
  • Surface liquid phase concentration in macropore, mg/L

  •  
  • Surface liquid phase concentration in macropore, mg/L

  •  
  • Ce

    Liquid phase concentration at equilibrium, mg/L

  •  
  • Dma

    Macropore diffusivity coefficient, m2/s

  •  
  • Dmi

    Micropore diffusivity coefficient, m2/s

  •  
  • f

    Fraction of total adsorptive capacity in macro-pores

  •  
  • External liquid film mass transfer coefficient, m/s

  •  
  • Macropore surface reaction rate constant, s−1

  •  
  • k, b

    Langmuir equation parameters

  •  
  • Freundlich equation parameters

  •  
  • N

    Number of data points

  •  
  • Solid-phase adsorbed concentration at equilibrium, mg/g

  •  
  • Solid-phase adsorbed concentration in macro-pores, mg/g

  •  
  • Solid-phase adsorbed concentration in micro-pores, mg/g

  •  
  • Solid-phase adsorbed concentration on macro-pores external surface, mg/g

  •  
  • Solid-phase adsorbed concentration on micro-pores external surface, mg/g

  •  
  • r

    Radial variable, m

  •  
  • Particle density, g/L

  •  
  • t

    Time variable, s

  •  
  • V

    Liquid phase volume, m3

  •  
  • X

    Experimental data

  •  
  • Mean of experimental data

  •  
  • Y

    Model simulated data

  •  
  • Mean of model simulated data

HIGHLIGHTS

  • 1.

    SO degradation using reactive adsorption followed pseudo-first-order kinetics.

  • 2.

    Surface – parallel-pore-reactive adsorption model has been applied for SO.

  • 3.

    SO liquid film mass transfer and surface reaction rate coefficient were predicted.

  • 4.

    SO macro-pore and micro-pore diffusivities were estimated.

  • 5.

    SO degradation occurred via hydroxylation and ring cleavage.

INTRODUCTION

Intensification process, reactive adsorption, has a unique advantage of combining reactive degradation with adsorption in a single step for organic pollutant removal (Gupta et al. 2015). Reactive adsorption is a hybrid process in which the pollutant and/or its degradative intermediate(s) formed during the reaction adsorb on to the adsorbent surface (Sharma et al. 2016). Toxic dyes are considered to be a major component of organic contaminants (Sharma et al. 2017a). Complete eradication of color via reactive degradation of dye molecules may produce relatively less harmful wastewater (Nogueira et al. 2009). Granular activated carbon (AC) is prominently used in treatment of effluents due to its high adsorption capacity (Chang et al. 2015; Garg et al. 2019; Rafatullah et al. 2010). Removal of organic moiety can be considerably improved through catalytic oxidation by the iron loading on AC (Xu et al. 2015). Reactive adsorption has been found to be a potential method for wastewater treatment as it catalyzes the oxidative degradation of a variety of pollutants using the Fe2O3/activated-carbon/H2O2 system (Zazo et al. 2006). Generated hydroxyl radical is extremely reactive (E0 = 2.80 V) (Pereira et al. 2011), oxidizing organic compounds swiftly and non-selectively (Neamu et al. 2004). Furthermore, reactive adsorption of organic pollutants gets rid of the sludge generation problem found in Fenton's process (Chen et al. 2011) and reduces the design complexity and capital investment.

Safranin O (SO) is an azine dye (El-Kemary & El-Shamy 2009) and is commonly used to dye silk, wool, tannin, cotton, fibers, silk, leather and paper. Also it has wide applications in textile, trace, and biological laboratory purposes (El-Kemary et al. 2011). The ingestion of SO-containing water causes distinct acute impact on health, such as irritation to the mouth, tongue, throat, lips and pain in the stomach that may lead to vomiting, nausea and diarrhea (Hayat et al. 2011; Kaur et al. 2015). Thus, removal of SO dye from wastewater is highly desirable.

Before implementing any technology, modeling provides a better understanding of the process behavior without performing a set of expensive and time-consuming experiments. Therefore, modeling and simulation are crucial for the design of processes like reactive adsorption. Models for adsorption are available in the literature (McKay & McConvey 1981; McKay et al. 1984) that considered porous adsorbents with mono-disperse pore structure. Peel et al. (1981) proposed the branched pore model for more accurate and realistic visualization of poly-disperse pore structure of activated carbon (Peel et al. 1981). Models are scarcely proposed for reactive adsorption of organic compounds. However, a model for removal of phenol was developed and reported using parallel pore diffusion (Gupta et al. 2016; Sharma et al. 2016). Earlier a model for reactive adsorption was proposed without considering the intricacies of the pore structure of the adsorbent and the model predictions were not authenticated by experimental results (Molga 2008). To our knowledge, experimental and modeling studies on reactive adsorption of a dye have not been reported so far.

Gupta et al. (2016) developed a model to study removal of phenol using reactive adsorption and the developed model was also used in this study for describing the removal of dye using reactive adsorption. In the model, to estimate the surface concentration of SO, mass transfer and/or surface reaction approach was used. For modeling purpose, it was assumed that the structure of AC is consisted of parallel micro- and macro-pores. Also, meso-pores have been combined together with micro-pores. Iron oxide modified AC (Fe-AC) was used as a reactive adsorbent. The impregnation of iron particles was considered in macro-pores only and micro-pores do not contain iron due to their smaller pore size (Sharma et al. 2017b). The model predictions were supported by the experimental data obtained for adsorption as well as reactive adsorption. The adsorbed SO concentration profile in Fe-AC particle was envisioned by the model. In addition, the external liquid film mass transfer coefficient, macro-pore and micro-pore diffusivities were estimated by regression analysis. Furthermore, photolytic degradation of SO is given in the literature in the presence of the Fe+3/H2O2 system (Abdullah et al. 2007). However, the authors did not study the degradation mechanism and intermediates formed. Therefore, to obtain a better understanding of the mechanistic details of the Fe-AC/H2O2 dye degradation process, the reaction intermediates were identified and a SO degradation pathway was proposed.

EXPERIMENTAL

Synthesis of Fe-AC

Granular activated carbon doped with iron (Fe-AC) was produced using FeCl3 (Mondal et al. 2007). The preparation method of Fe-AC and analytical characterization of adsorbents (TEM-EDS, XRD, BET surface area) have been reported elsewhere (Sharma et al. 2017b). Analytical techniques confirmed that the synthesized Fe-AC contained iron oxide. The presence of Fe+3 on GAC helps in producing hydroxyl radicals that participate in the reactive adsorption of dye (Pereira et al. 2011; Pinto et al. 2012).

Batch studies of SO reactive adsorption

The Fe-AC/H2O2 catalytic system was used to study the oxidative degradation of SO in aqueous solution. For batch kinetic studies of reactive adsorption, 100 mL of aqueous solution of SO (10 mg/L) was treated with 1.0 g/L of Fe-AC and 5.0 mM H2O2, respectively at 30 ± 1 °C at its natural pH of 6.5 (almost neutral condition). For a defined time interval, absorption spectra of the samples were measured using a UV–visible spectrophotometer. ESI-MS analyses were used to identify the intermediates formed during SO degradation.

Batch studies of SO adsorption

Equilibrium study was carried out using a dose of 1.0 g/L of Fe-AC in 100 mL of SO (10–50 mg/L) at 30 ± 1 °C and 150 rpm to reach the equilibrium of the solid-solution mixture. Kinetic study of SO adsorptive removal was also conducted using 10 mg/L initial concentration and 1.0 g/L of Fe-AC for 5 h to compare its performance with reactive adsorption.

RESULTS AND DISCUSSION

Removal efficiencies

SO removal efficacy was investigated using reactive adsorption for different feeds of Fe-AC and H2O2. The optimized doses were found to be 1.0 g/L and 5.0 mM of Fe-AC and hydrogen peroxide, respectively. Figure 1 shows that the amount of dye removal was increased with a greater amount of iron on the AC. An ample amount of hydroxyl radicals were generated due to the enhanced availability of iron on the adsorbent surface with increasing Fe-AC doses (Chen et al. 2011).

Figure 1

Pictorial presentation of SO degradative removal using reactive adsorption.

Figure 1

Pictorial presentation of SO degradative removal using reactive adsorption.

Figure 1 also shows that for 1.0 g/L Fe-AC dose, SO decolorization considerably increased with hydrogen peroxide dose. However, the acceleration of dye removal was found with hydrogen peroxide doses from 1.0 to 5.0 mM and when the concentration of H2O2 was raised from 5.0 to 7.0 mM, a decrease in SO removal was observed. This is attributed to hydroperoxyl radicals (HO2) produced by reaction of hydroxyl radicals and locally accessible H2O2 (Equation (1)) that results in scavenging of OH (Neamu et al. 2004; Kasiri et al. 2008). 
formula
(1)

The oxidation reaction of dye degradation diminishes considerably due to the lesser reactivity of hydroperoxyl radicals as compared to OH, with reaction rate constants lower than 2 × 104 M−1s−1 (Kasiri et al. 2008). For an Fe-AC dose of 0.8 g/L, sufficient OH radicals were formed at 1.0 mM H2O2 dose for removal of dye but at higher H2O2 doses dye removal decreased due to hydroxyl radical scavenging. Almost constant behavior for SO removal was observed at 0.6 g/L Fe-AC dose, which may be due to a balanced phenomenon of hydroxyl radical generation and its scavenging up to 1.0–5.0 mM H2O2 dose, and then at a higher H2O2 dose (7.0 mM) formation of HO2 dominated, which reduced dye degradation efficiency.

Therefore, optimization of H2O2 and Fe-AC doses was done to maximize the dye degradation. 96.1% maximum SO removal was achieved at optimized conditions (1.0 g/L Fe-AC and 5.0 mM H2O2) with 10 mg/L initial SO concentration, and further experiments were performed with these conditions.

Kinetics

Pseudo first order kinetics of dye decoloration can be given as 
formula
(2)
A correlation coefficient of 0.982 was found for the data between and t (Figure S1(a)) for SO to the above equation for the pseudo first-order-kinetic model, which validates its suitability. The kapp value 1.7 × 10−4 s−1 was obtained using pseudo first-order-kinetics. The pseudo second order kinetic system can be expressed as 
formula
(3)
1/Ct versus t were linearly plotted (Figure S1(b)). An upward concave curvature (correlation coefficient value of 0.756) was found and therefore, pseudo second-order-kinetics were not considered for dye decoloration kinetics.

Adsorption equilibrium

The nature of SO sorption on Fe-AC was examined using the Langmuir isotherm model (Figure S2). The Langmuir isotherm equation is given by 
formula
(4)

Reactive adsorption is a surface phenomenon in which hydroxyl radicals are formed on the adsorbent surface and oxidize the organic moieties. Therefore, the Langmuir isotherm, which assumes monolayer adsorption, was used with the regression coefficient 0.996. Parameters obtained for the Langmuir isotherm are k = 22.7 L/g and b = 0.909 L/mg.

Reactive adsorption modeling

A mathematical model was used for adsorption and reactive adsorption of SO (Gupta et al. 2016). According to the model, the solute molecules (SO) pass through the external liquid film and are sorbed on the sorbent surface, which consists of numerous macro-pores (f) and micro-pores (1–f). The sorbate adsorption and diffusion takes place relatively faster in macro-pores than in micro-pores due to the larger pore diameter. It was an assumption that the macro-pores and micro-pores are parallel to one another, are distributed homogeneously within the particles and give radial transport. The existence of iron particle clusters in activated carbon was confirmed by TEM analysis. These iron particles were in sizes ranging from 50 to 100 nm (Sharma et al. 2017b) which is aligned with the literature (Hristovski et al. 2009). For the modeling purposes, micro-pores and meso-pores were grouped together. However, GAC particles consist of micro-pores (<2 nm), meso-pores (2–50 nm) and macro-pores (>50 nm) (Groen et al. 2003). Therefore, it was considered that iron particles' impregnation would occur in macro-pores only. Thus, reactive sites are present on macro-pores only and micro-pores do not contain any reactive sites due to the absence of iron. However, it is believed that adsorption of SO occurs in macro-pores as well as micro-pores.

For a batch-wise removal of SO using reactive adsorption, a mathematical model was employed. It is considered that molecules of SO pass through an aqueous film that lies outside the particle surface, and then these molecules are further adsorbed onto the macro-pore and micro-pore surfaces of the particle. In reactive adsorption, dye molecules react with OH radicals generated on the surface of the macro-pores as per pseudo-first-order reaction kinetics. The mechanism of reactive adsorption can be described by three resistances in series: (i) outer resistance across the film of liquid, (ii) resistance through macropore diffusion and (iii) surface reaction resistance. It is inferred that, at any moment, there was equilibrium between the surface concentration of dye and the amount of SO adsorbed on the particle surface. The following equations presented this mathematical model:

External fluid phase mass balance: 
formula
(5)
Macro-pore mass balance: 
formula
(6)
Micro-pore mass balance: 
formula
(7)
Initial and boundary conditions are: 
formula
(8)
 
formula
(9)
 
formula
(10)
 
formula
(11)
 
formula
(12)
Relationship at the solid-liquid interface can be written as: 
formula
(13)
 
formula
(14)
Surface concentrations of dye are in equilibrium with the adsorbed amount on the surface at any instant. Therefore, 
formula
(15)
 
formula
(16)
 
formula
(17)
To evaluate the fitness of data obtained from experiment and prediction from the model, the below-mentioned statistical indices was employed. 
formula
(18)

The model equations were solved using a MATLAB program. The program uses factors like sorbent mass, size and density of the Fe-AC particle, initial SO concentration, volume of aqueous solution, kinetic data and Langmuir equilibrium constants, including time and intervals, and apparent rate constant for reactive adsorption. The fraction of macro-pores, f, was taken from literature (Peel et al. 1981). Primarily, values of kL, Dma, and Dmi were obtained for adsorption. These values do not change in the case of reactive adsorption. Hence, using these values and kapp, further kr was evaluated for reactive adsorption. The values of kL, Dma, Dmi, and kr are given in Table 1. The program gives results such as the theoretical kinetic data related to the solute concentration in the aqueous phase with time; solid-phase dye concentration at the surface and time along with the dimensionless distance across the particles for macro- and micro-pore diffusion during adsorption and reactive adsorption. By fitting the experimental data to the model, it was found that the surface reaction rate constant, kr, is 5 times the apparent rate constant, kapp. This may be attributed to the inclusion of diffusional resistance in calculation of kapp values, which considerably reduces the effective value of the rate constant to be taken into account for the surface reaction coefficient.

Table 1

Parameters of reactive adsorption model

Parameters Value 
Film mass transfer coefficient, kL 8.15 × 10−5 m/s 
Macropore diffusivity, Dma 1.1 × 10−12 m2/s 
Micropore diffusivity, Dmi 6.9 × 10−13 m2/s 
Surface reaction rate constant, kr 8.5 × 10−4 s−1 
Macropore fraction, f 0.66 
Parameters Value 
Film mass transfer coefficient, kL 8.15 × 10−5 m/s 
Macropore diffusivity, Dma 1.1 × 10−12 m2/s 
Micropore diffusivity, Dmi 6.9 × 10−13 m2/s 
Surface reaction rate constant, kr 8.5 × 10−4 s−1 
Macropore fraction, f 0.66 

It can be observed from Figure 2 that predicted SO concentrations in bulk using the model are in accordance with experimental values for adsorption. However, the simulated values deviate from experimental values, being relatively higher for the first hour with negligible deviation later. This higher deviation may be due to the particle's surface hydrophobicity, which results in initial slow film diffusion. After a certain period, the deviation is minimized as equilibrium is acquired across the film and only pore diffusion dominates. In addition, the SO concentration was reduced by only 58.4% in 5 h from the bulk solution using adsorption (Figure 2).

Figure 2

Change in SO concentration in aqueous solution during adsorption and reactive adsorption.

Figure 2

Change in SO concentration in aqueous solution during adsorption and reactive adsorption.

Radial concentration profiles of SO, obtained from the model, within the macro-pore and micro-pore regions of the Fe-AC particle during 5 h are presented in Figures 3 and 4, respectively. The concentration profiles of SO in macro-pores of adsorbent particles (Figure 3) for 5 h, indicated that the quantity of diffused SO in the sorbent particle was up to 65% of the particle radius. Furthermore, adsorbate diffused slowly towards the centre of the particle, possibly due to the lesser value of macro-pore diffusivity. Figure 4 shows that the diffusion rate of SO is comparatively slow in micro-pores of the sorbent particle. It can be observed from concentration profiles that 51% of micro-pores have adsorbed SO via diffusion. There is comparatively lower adsorption of SO towards the centre of the particle in micro-pores compared to macro-pores, which may be due to the relatively smaller value of micro-pore diffusivity. The outer part of the adsorbed concentration profile stipulates that the SO adsorbed concentration increased over a very short duration. Afterwards, SO concentration lowers as time proceeds due to the diffusion in adjacent layers in macro- and micro-pores. Comparative observation of macro-pore and micro-pore profiles indicates that the movement of concentration front towards the adsorbent center is almost similar in macro-pores and micro-pores.

Figure 3

Concentration profile of SO in macro-pores during adsorption.

Figure 3

Concentration profile of SO in macro-pores during adsorption.

Figure 4

Concentration profile of SO in micro-pores during adsorption.

Figure 4

Concentration profile of SO in micro-pores during adsorption.

The reactive adsorption experimental data of SO are found to fit well to values obtained using the parallel-pore-diffusion-model (Figure 2). During reactive adsorption, the SO concentration reduces and was almost fully vanished (<0.5 mg/L) in 5 h. Therefore, this process is found to be much more effective than pure adsorption in terms of both time required for treatment and contaminant removal efficiency. Surface reaction dominates over pore diffusion in the case of reactive adsorption. The sorbed concentration of SO in macro-pores depletes rapidly due to oxidative degradation of SO in the presence of highly reactive OH. Figures 5 and 6 show the concentration profiles of SO within the reactive sorbent particle for a duration of 5 h in the macro-pores and micro-pores respectively. The adsorbed SO concentration profiles in the external part of the macro-pores increases initially within a very short period due to the high value of the film mass transfer coefficient and then depletes rapidly due to the chemical reaction. In 5 h, only 20% of the particle is diffused with SO diffused, indicating that a considerable part of the sorbent is still vacant. However, the maximum amount of sorbed SO at the outer surface is the same as for simple adsorption. SO concentration depletes very fast due to the high rate of reaction and it does not give enough time to move towards the center of the particle as macro-pores are highly reactive (Fe/H2O2). In the case of reactive adsorption, the removal of SO dye occurred entirely on the external macro-pore surface. This represents the fast reaction of SO occurring in the exterior surface of macro-pores, and narrower pores offer strong diffusion resistances. The maximum amount of adsorbed SO in the micro-pore is almost similar to pure adsorption. However, the adsorbed concentration profiles of micro-pores increase initially due to the high film mass transfer coefficient, but it decreases with time.

Figure 5

Concentration profile of SO in macro-pores during reactive adsorption.

Figure 5

Concentration profile of SO in macro-pores during reactive adsorption.

Figure 6

Concentration profile of SO in micro-pores during reactive adsorption.

Figure 6

Concentration profile of SO in micro-pores during reactive adsorption.

Reactive adsorption study of SO using ESI-MS

Intermediates formed during oxidative degradation of SO were identified using ESI-MS analyses. The results suggested that Fe-AC catalyzed the dye degradation by forming hydroxyl radicals from hydrogen peroxide. Figure 7(a)–7(c) show the ESI-MS results obtained for different time intervals of batch reaction. Figure 7(a) illustrates the peak corresponding to SO (m/z = 315) at t = 0 before initiation of the SO degradation reaction.

Figure 7

ESI–MS spectra of SO aqueous solution and its degradation products during reactive adsorption at (a) t = 0 h, (b) t = 1.5 h, and (c) t = 5 h.

Figure 7

ESI–MS spectra of SO aqueous solution and its degradation products during reactive adsorption at (a) t = 0 h, (b) t = 1.5 h, and (c) t = 5 h.

m/z = 315 signal intensity decreased with the intermediates formation, as can be seen in respective ESI-MS analysis with a 1.5 h reaction period (Figure 7(b)). There are two possible reaction pathways for dye degradation: (1) hydroxylation (Oliveira et al. 2007) and (2) ring cleavage (Chen et al. 2011). Hydroxylated intermediates of SO were produced due to the involvement of highly reactive OH (Sharma et al. 2018). The intermediates formed can be observed in Figure 7(b) at m/z = 333, 349, and 365. Ring cleavage of SO and hydroxylation result in the formation of intermediates at m/z = 111, 126, 151, and 167 in 1.5 h. The intensity of hydroxylated intermediates of SO (m/z = 333, 349, and 365) reduced and then completely vanished as the reaction proceeded due to ring cleavage of SO chromophore and addition of the OH group (Figure 7(c)). Intermediates at m/z = 111, 126, 151, and 167 can be seen after 5 h of reaction. Furthermore, complete mineralization of SO to carbon dioxide and water can be achieved by supplying an ample amount of OH radicals and giving plenty of time for the degradation reaction. A proposed degradation pathway is shown in Figure 8.

Figure 8

Proposed oxidative degradation pathways of SO in aqueous solution during reactive adsorption using Fe-AC/H2O2.

Figure 8

Proposed oxidative degradation pathways of SO in aqueous solution during reactive adsorption using Fe-AC/H2O2.

CONCLUSIONS

The performance of granular activated carbon doped with iron (Fe-AC) was found to be effective for the removal of SO using reactive adsorption (Fe-AC/H2O2) and adsorption (Fe-AC). Maximum eradication of dye was found to be 96.1% (initial SO concentration: 10 mg/L) after 5 h of reactive adsorption using 1.0 g/L and 5.0 mM dosages of Fe-AC and hydrogen peroxide, respectively. Kinetics of dye degradation were represented by the pseudo-first-order model. The equilibrium of SO sorption onto Fe-AC was well represented by the Langmuir isotherm. The parallel-pore-reactive adsorption model was validated using adsorption and reactive adsorption experimental data for SO. The surface reaction coefficient was 5 times the apparent rate constant. The external liquid film-mass-transfer-coefficient, macro-pore and micro-pore diffusivities were estimated. In addition, the rate controlling mechanisms in pure adsorption and reactive adsorption were attributed to pore diffusion and surface reaction, respectively. Degradation of the SO molecule followed hydroxylation and ring cleavage processes and an oxidative degradation pathway has been proposed.

REFERENCES

REFERENCES
Abdullah
F. H.
,
Rauf
M. A.
&
Ashraf
S. S.
2007
Photolytic oxidation of Safranin-O with H2O2
.
Dyes and Pigments
72
(
3
),
349
352
.
Chang
J.
,
Ma
J.
,
Ma
Q.
,
Zhang
D.
,
Qiao
N.
,
Hu
M.
&
Ma
H.
2015
Adsorption of methylene blue onto Fe3O4/activated montmorillonite nanocomposite
.
Applied Clay Science
119
,
132
140
.
Chen
C. C.
,
Chen
W. C.
,
Chiou
M. R.
,
Chen
S. W.
,
Chen
Y. Y.
&
Fan
H. J.
2011
Degradation of crystal violet by an FeGAC/H2O2 process
.
Journal of Hazardous Materials
196
,
420
425
.
El-Kemary
M.
&
El-Shamy
H.
2009
Fluorescence modulation and photodegradation characteristics of safranin O dye in the presence of ZnS nanoparticles
.
Journal of Photochemistry and Photobiology A: Chemistry
205
(
2–3
),
151
155
.
El-Kemary
M.
,
Abdel-Moneam
Y.
,
Madkour
M.
&
El-Mehasseb
I.
2011
Enhanced photocatalytic degradation of Safranin-O by heterogeneous nanoparticles for environmental applications
.
Journal of Luminescence
131
(
4
),
570
576
.
Garg
D.
,
Kumar
S.
,
Sharma
K.
&
Majumder
C. B.
2019
Application of waste peanut shells to form activated carbon and its utilization for the removal of Acid Yellow 36 from wastewater
.
Groundwater for Sustainable Development
8
,
512
519
.
Groen
J. C.
,
Peffer
L. A. A.
&
Perez-Ramirez
J.
2003
Pore size determination in modified micro- and mesoporous materials. Pitfalls and limitations in gas adsorption data analysis
.
Microporous and Mesoporous Materials
60
,
1
17
.
Gupta
V. K.
,
Sharma
M.
&
Vyas
R. K.
2015
Hydrothermal modification and characterization of bentonite for reactive adsorption of methylene blue: an ESI-MS study
.
Journal of Environmental Chemical Engineering
3
(
3
),
2172
2179
.
Gupta
V. K.
,
Sharma
M.
,
Singh
K.
&
Vyas
R. K.
2016
Reactive adsorption of phenol onto Fe-GAC: parallel pore batch modeling and experimental studies
.
Journal of the Taiwan Institute of Chemical Engineers
0
,
1
9
.
Hayat
K.
,
Gondal
M. A.
,
Khaled
M. M.
,
Yamani
Z. H.
&
Ahmed
S.
2011
Laser induced photocatalytic degradation of hazardous dye (Safranin-O) using self synthesized nanocrystalline WO3
.
Journal of Hazardous Materials
186
(
2–3
),
1226
1233
.
Hristovski
K. D.
,
Westerhoff
P. K.
,
Möller
T.
&
Sylvester
P.
2009
Effect of synthesis conditions on nano-iron (hydr)oxide impregnated granulated activated carbon
.
Chemical Engineering Journal
146
(
2
),
237
243
.
Kasiri
M. B.
,
Aleboyeh
H.
&
Aleboyeh
A.
2008
Degradation of Acid Blue 74 using Fe-ZSM5 zeolite as a heterogeneous photo-Fenton catalyst
.
Applied Catalysis B: Environmental
84
(
1–2
),
9
15
.
Kaur
S.
,
Rani
S.
,
Mahajan
R. K.
,
Asif
M.
&
Gupta
V. K.
2015
Synthesis and adsorption properties of mesoporous material for the removal of dye safranin: kinetics, equilibrium, and thermodynamics
.
Journal of Industrial and Engineering Chemistry
22
,
19
27
.
McKay
G.
&
McConvey
I. F.
1981
External mass transfer of basic and acidic dyes on wood
.
Journal of Chemical Technology and Biotechnology
31
(
7
),
401
408
.
McKay
G.
,
Allen
S. J.
,
McConvey
I. F.
&
Walters
J. H. R.
1984
External mass transfer and homogeneous solid-phase diffusion effects during the adsorption of dyestuffs
.
Industrial & Engineering Chemistry Process Design and Development
23
(
2
),
221
226
.
Molga
E.
2008
Modelling of reactive adsorption processes
.
Chemical and Process Engineering
29
,
683
699
.
Nogueira
F. G. E.
,
Lopes
J. H.
,
Silva
A. C.
,
Gonçalves
M.
,
Anastácio
A. S.
,
Sapag
K.
&
Oliveira
L. C. A.
2009
Reactive adsorption of methylene blue on montmorillonite via an ESI-MS study
.
Applied Clay Science
43
(
2
),
190
195
.
Oliveira
L. C. A.
,
Ramalho
T. C.
,
Goncalves
M.
,
Cereda
F.
,
Carvalho
K. T.
,
Nazzarro
M. S.
&
Sapag
K.
2007
Pure niobia as catalyst for the oxidation of organic contaminants: mechanism study via ESI-MS and theoretical calculations
.
Chemical Physics Letters
446
(
1–3
),
133
137
.
Peel
R. G.
,
Benedek
A.
&
Crowe
C. M.
1981
A branched pore kinetic-model for activated carbon adsorption
.
AIChE Journal
27
(
1
),
26
32
.
Pereira
M. C.
,
Cavalcante
L. C. D.
,
Magalhães
F.
,
Fabris
J. D.
,
Stucki
J. W.
,
Oliveira
L. C. A.
&
Murad
E.
2011
Composites prepared from natural iron oxides and sucrose: a highly reactive system for the oxidation of organic contaminants in water
.
Chemical Engineering Journal
166
(
3
),
962
969
.
Pinto
I. S. X.
,
Pacheco
P. H. V. V.
,
Coelho
J. V.
,
Lorencon
E.
,
Ardisson
J. D.
,
Fabris
J. D.
,
Souza
P. P. d.
,
Krambrock
K. W. H.
,
Oliveira
L. C. A.
&
Pereira
M. C.
2012
Nanostructured δ-FeOOH: an efficient Fenton-like catalyst for the oxidation of organics in water
.
Applied Catalysis B: Environmental
119–120
,
175
182
.
Rafatullah
M.
,
Sulaiman
O.
,
Hashim
R.
&
Ahmad
A.
2010
Adsorption of methylene blue on low-cost adsorbents: a review
.
Journal of Hazardous Materials
177
(
1–3
),
70
80
.
Sharma
M.
,
Vyas
R. K.
&
Singh
K.
2016
Theoretical and experimental analysis of reactive adsorption in a packed bed: parallel and branched pore-diffusion model approach
.
Industrial & Engineering Chemistry Research
55
(
20
),
5945
5954
.
Sharma
K.
,
Dalai
A. K.
&
Vyas
R. K.
2017a
Removal of synthetic dyes from multicomponent industrial wastewaters
.
Reviews in Chemical Engineering
34
(
1
),
107
134
.
Sharma
K.
,
Vyas
R. K.
&
Dalai
A. K.
2017b
Thermodynamic and kinetic studies of methylene blue degradation using reactive adsorption and its comparison with adsorption
.
Journal of Chemical & Engineering Data
62
(
11
),
3651
3662
.
Sharma
K.
,
Vyas
R. K.
,
Singh
K.
&
Dalai
A. K.
2018
Degradation of a synthetic binary dye mixture using reactive adsorption: experimental and modeling studies
.
Journal of Environmental Chemical Engineering
6
(
5
),
5732
5743
.
Xu
J. h.
,
Gao
N. y.
,
Zhao
D. y.
,
Zhang
W. x.
,
Xu
Q. k.
&
Xiao
A. h.
2015
Efficient reduction of bromate in water by nano-iron hydroxide impregnated granular activated carbon (Fe-GAC)
.
Chemical Engineering Journal
275
,
189
197
.
Zazo
J. A.
,
Casas
J. A.
,
Mohedano
A. F.
&
Rodríguez
J. J.
2006
Catalytic wet peroxide oxidation of phenol with a Fe/active carbon catalyst
.
Applied Catalysis B: Environmental
65
(
3–4
),
261
268
.

Supplementary data