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UV treatment
Journal of Hazardous Materials 271 (2014) 275–282
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Removal of metal ions from wastewater using EB irradiation in combination with HA/TiO2 /UV treatment A.A. Zaki a,∗ , Naima A. El-Gendy b a b
Hot Laboratory and Waste Management Center (HLWMC), Atomic Energy Authority, P.O. Box 13758, Inshas, Cairo, Egypt Department of Polymer, National Center for Radiation Research and Technology(NCRRT), Atomic Energy Authority, Nasr City, P.O. Box 19, Cairo, Egypt
h i g h l i g h t s • • • •
A combined electron beam (EB) and humic acid (HA/TiO2 ) treatment process was developed. The ion metals Cu2+ , Sr2+ , and Co2+ were effectively removed from wastewater. About 50 kGy of EB was needed to remove the Cu2+ , Sr2+ , and Co2+ from wastewater. Mechanisms of interactions between HA and metal ions were suggested and discussed.
a r t i c l e
i n f o
Article history: Received 2 December 2013 Received in revised form 16 February 2014 Accepted 18 February 2014 Available online 28 February 2014 Keywords: Electron beam Photocatalysis Humic acid Removal Mechanism
a b s t r a c t The electron beam (EB) irradiation technology was applied for removal of Cu2+ , Sr2+ , and Co2+ ions from wastewater. The aim of this study is to achieve an efficient treatment process of wastewater using EB and introducing a combination of humic acid (HA) as a natural organic polymer and ultraviolet irradiation of a TiO2 (TiO2 /UV), as a suspended catalyst in the treatment of wastewater solutions (TiO2 /UV + HA). The experimental results showed that the percentage removal of Cu2+ , Sr2+ , and Co2+ ions was 41%, 87% and 75% respectively, at 125 kGy. In the presence of TiO2 photocatalyst and exposure of the investigated wastewater to ultraviolet rays before irradiation by the EB the percentage removal of Cu2+ ions became 51%, while the percentage removal of both Sr2+ and Co2+ ions was slightly improved; was 87% and 75%, respectively at the same EB dose. On the other hand, by introducing the combination of TiO2 /UV + HA, only an irradiation dose of about 50 kGy led to removal of Cu2+ , Sr2+ , and Co2+ completely from the wastewater. Mechanisms of interactions between HA and Cu2+ , Co2+ and Sr2+ metal ions were suggested and discussed. © 2014 Elsevier B.V. All rights reserved.
1. Introduction The handling of wastewater appeared to be one of the most important problems in the last few decades. Aqueous wastes may be treated using ion exchange, sorption, chemical precipitation, evaporation, reverses osmosis, filtration and solvent extraction [1]. The main features and limitation of these treatment processes are described elsewhere [2]. Occupational or environmental exposure to cobalt is associated with health effects on various organs, including the respiratory tract, the skin, the red blood cells, the thyroid gland and the myocardium [3]. Historically, effects on the thyroid gland were observed as an undesirable side-effect of cobalt therapy used for
∗ Corresponding author. Tel.: +20 2 35620445; fax: +20 2 4620796. E-mail address: [email protected] (A.A. Zaki). http://dx.doi.org/10.1016/j.jhazmat.2014.02.025 0304-3894/© 2014 Elsevier B.V. All rights reserved.
its polycythemic action, often in association with iron, in the treatment of various anemias [4]. The major industrial uses of strontium include the production of glass for color television picture tubes, ferrite magnets, and refining of zinc [5–8]. Copper is an essential micronutrient that forms part of several proteins involved in a variety of biological processes indispensable to sustain life [9]. At the same time, it can be toxic when present in excess, the most noticeable chronic effect being liver damage [10]. Radiation technologies are frequently applied for purification of wastewater and groundwater contaminated with organic pollutants [11,12]. Humic acids (HAs) are macromolecular yellow-to-black colored natural organic matter derived from the degradation of plant, algal, and microbial material [13]. Although their formation mechanism and chemical structures are not well understood, they are known to be high in carbon content (50–60%) of both aliphatic and aromatic character and rich in oxygen-containing functionalities such as carboxyl,
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phenolic, alcoholic, and quinoid groups [14,15]. The absorption of light by chromophores existing in humic acid (HA) may lead to alterations of HA structure and composition which is reflected by the changes in absorption, fluorescence, chemiluminescence, NMR, IR and EPR spectra [16,17]. Photocatalytic oxidation using TiO2 is gaining wide attentions as an advanced water treatment technology. TiO2 photocatalyst initiates upon absorbing UV photons with generating conduction band (CB) electrons and valence band (VB) holes within the particle [18,19]. In particular, the hydroxyl radicals generated through the reaction of VB holes account for the strong oxidizing power of the TiO2 photocatalytic system. Although TiO2 has been widely used as a photocatalyst, the band gap excitation of TiO2 proceeds only in the presence of UV light ( < 388 nm) [20]. Sensitized TiO2 by humic acid (HA–TiO2 ), ruthenium-based complexes or various dyes has been extensively investigated in order to extend the spectral response into the visible region [21]. The aim of this work is to achieve an efficient and energy saving process for treatment of wastewater and removal of Cu2+ , Sr2+ and Co2+ ions from waste solutions using a combination of electron beam irradiation (EB), HA sensitized UV-irradiated Ti2 O catalyst (HA–TiO2 ). 2. Materials and methods 2.1. Materials All chemicals used in this work were of analytical grade. All aqueous solutions were prepared with distilled deionized water, which was purified by a Millipore Milli-Q water (resistivity = 4.4 M cm, TOC = 0.1 mg/l) purification system. Chemical powders used to simulate wastewater were supplied from Merck; CuSO4 ·5H2 O, CoCl2 , SrCl2 ·4H2 O. Stock solutions of 100 mg/l from different metal ions were prepared using distilled deionized water. Humic acid (HA) in sodium form and TiO2 powder, Degases P25 (80% anatase and 20% rutile), were purchased from Aldrich. The TiO2 powder had a BET surface area of 55 m2 /g and an average particle diameter of 30 nm. HA was initially dissolved in alkaline solution (pH ≈ 12) at 100 mg/l and subsequently adjusted to pH 5.0 using 1 N HCl. All experiments were done at room temperature (298 ± 1 K) otherwise is mentioned. The HA filtrate was stored as stock solution and its concentration was routinely checked. 2.2. Electron-beam research facility Experiments to study the electron beam (EB) treatment process were carried out using NCRRT-Accelerator. The NCRRT-Accelerator is a direct accelerator delivered from High Voltage Company, USA; 1.5 MeV, beam current of 25 mA, power of 37.5 kW and variable scan width up to 90 cm. This accelerator is used to produce an EB with high energy which can be used for different applications in different fields.
2 h at 200 rpm. The percentage uptake of HA by Ti2 O was calculated according to the formula (C0 − Ce /C0 ) × 100, where C0 and Ce are the initial and equilibrium concentrations of HA, respectively. The dissolved humic acid was spectrophotometerically measured using the UV–vis spectrophotometer, Shimadzu-UV 2550, at = 254 nm. 2.5. Photocatalytic TiO2 /HA/metal ion process Photocatalytic TiO2 powder (Degases P25) is used in photocatalytic experiments due to its high UV-light efficiency. The TiO2 suspension was prepared by simultaneous sonication and shaking in an ultrasonic cleaning bath (Branson 3210). Solution of HA was then added to a 100 mg/l TiO2 suspension to give a desired concentration of HA and added to metal ion solutions. A removable 30 cm long UV-florescent lamp, model DESAGA Abnehmber was used to produce light for irradiating the TiO2 /HA/metal ion solution. It has two wavelengths 254 nm and 396 nm, power of 30 W and the incident light is 60 W/m2 . 2.6. Procedures of experimental setting up The liquid solution samples were divided into three groups. The first group samples were irradiated for different doses in the range 25–150 kGy. The second group samples were exposed to UV-lamp with photocatalytic TiO2 powder (Degasses P25) suspended in solution for 2 h (checked well during irradiation) then directly moved to the EB irradiation. The third group samples were mixed with HA and exposed to UV-lamp with TiO2 photocatalytic for 2 h then directly moved to EB to complete the irradiation. After EB irradiation the samples were centrifuged at 5000 rpm for 10 min before measurement on an atomic absorption spectrophotometer (AAS) as well as the metal ions in the samples of the other two groups. The AAS utilizes air acetylene flame, is a Boch Scientific model 210 VGP, USA. 2.7. Sorption mechanism Fourier transform infrared spectroscopy (FTIR) was used to investigate the sorption mechanisms of the studied metal ions onto HA. For FTIR experiments, 1 g mass of finely powdered HA placed in 250 cm3 beaker, 100 cm3 of 0.025 M metal salt solution was added and adjusted to pH 5.0. The suspension was magnetically stirred at 333 K for half an hour, and then left for 24 h to saturate with the metal ion. Resulting suspensions of HA + M complexes were filtered and rinsed with distilled water. Finally the solid complexes were dried at 353 K for 4 h. The dry complexes were mixed with 50 mg KBr and compressed to obtain a pellet. The pellets were dried at 378 K for 2 days and then, analyzed using a MIDAC 1700 model FTIR spectrometer. The base line was corrected and scanning was performed from 4000 to 400 cm−1 . 3. Results and discussion
2.3. Electron beam irradiation (EB) radiation of HA molecules 3.1. Effect of EB irradiation on HA molecules The stability of HA, 50 mg/l, was studied by applying electron beam irradiation (EB) at doses of 0–32 kGy, and its ultraviolet (UV) absorbance was checked. A Unicam UV4 double beam spectrophotometer (product of Unicam Co. Ltd., England) was used to measure absorption spectra of the EB irradiated and unirradiated HA. 2.4. Effect of pH on the uptake of HA by Ti2 O Uptake studies of HA by Ti2 O, humic acid–titanium dioxide sensitization (HA–Ti2 O), were performing by mixing 500 mg/l Ti2 O with 50 mg of HA for pH ranging from 2.0 to 13.0. The mixture was shaken in a thermostatic shaker at room temperature (298 ± 1 K) for
Humic acid (HA) substances effectively absorb non-ionizing radiation, such as ultraviolet (UV), visible or infrared radiation. Much less is known about their sensitivity to ionizing radiation, such as EB, ␥- or X-rays. The UV absorption spectra of EB irradiated HA are illustrated in Fig. 1. As shown from the figure, that the humic acid solution at pH 5.0 has a maximum absorbance in the UV region at 320 nm [22]. This absorption belongs to the aromatic rings of HA and a considerable decrease in the absorbance of peak occurred with the increase in the EB irradiation dose. The observed decrease in absorption (Fig. 1) is attributed to degradation and precipitation of HA molecules.
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277
Table 1 Dose constant values, the initial concentrations and coefficient of determination (R2 ) at 45 kGy and pH 5.0 at room temperature (298 ± 1 K) of the studied systems. System
Dose constant K1 (kGy−1 )
Initial concentration C0 (mmol/l)
Coefficient of determination (R2 )
Sr2+ + EB Co2+ + EB Cu2+ + EB Sr2+ + EB + TiO2 /UV Co2+ + EB + TiO2 /UV Cu2+ + EB + TiO2 /UV Sr2+ + EB + TiO2 /UV + HA Co2+ + EB + TiO2 /UV + HA Cu2+ + EB + TiO2 /UV + HA
0.015 0.019 0.004 0.019 0.025 0.006 0.117 0.106 0.120
0.57 0.85 0.79 0.57 0.85 0.79 0.57 0.85 0.79
0.981 0.983 0.980 0.984 0.988 0.985 0.982 0.987 0.982
where the subscription Tot, aq, ads, red, and ox are total, aqueous, adsorption, reduction, and oxidation, respectively. The experimental data for removal of contaminants by EB irradiation were analyzed by fitting into first order equation [24]: C = C0 e−K1 D Fig. 1. Ultraviolet absorbance of HA before and after EB irradiation at pH 5.0 and room temperature (298 ± 1 K).
3.2. Electron beam irradiation (EB) of water and contaminants Irradiation of water or dilute aqueous solution by EB ionizing radiations results in the formation of various species according to the following Eq. (1) [23]:
ln
C C0
= −K1 D
(6) (7)
The linearization of the data by logarithmic representation (Eq. (7)) gives the dose constant K1 (Gy−1 ). The K1 -value is useful for the comparison of the radiolysis-based experiments, D (kGy) is the EB irradiation dose, C (mol/l) and C0 (mol/l) are the concentration at the applied dose D (kGy) and the initial concentration of
(1) The values in brackets (G) are the number of micromoles of molecules or radicals produced per joule absorbed energy (mol/J) and represent the efficiency of the conversion of the high energy electron radiation to a chemical process. Among the chemical species formed during the radiolysis of water, the hydroxyl radical (• OH), hydrated electron (e− aq ) and hydrogen atom (H• ) are the most reactive species with organic compounds. In aerated solution, the dissolved oxygen concentration is approximately 3 × 10−4 M (9.4 mg/l). In this case, solvated electrons (e− aq ) and hydrogen atoms (H• ) are practically scavenged by oxygen and rapidly converted into HO2 • , which are very poor reacting species compared to the hydroxyl radical (Eqs. (2) and (3)) [23]. e− aq + O2 → O2 H•
•−
+ O2 → HO2
•
k = 1.9 × 1010 M−1 s−1 k = 2.1 × 10
10
−1 −1
M
s
(2) (3)
where k (M−1 s−1 ) is the rate constant. The metal ions were efficiently reduced to lower oxidation states by both e− aq and H• but in some cases, e− aq can lead to reduction in metal ions alone. However, re-oxidation by hydroxyl radical (• OH), and H2 O2 should be prevented. Several kinetic studies referring to reduction of H2 O2 were found elsewhere [24,25]. The experimental data were presented as the ln(C/C0 ) versus the applied absorbed dose, D (kGy) of the systems EB + metal ions, EB + metal ions + TiO2 /UV, and EB + metal ions + TiO2 /UV + HA (Figs. 2 and 3). It was found that the relative concentration of metal ions, ln(C/C0 ) decreases with the increase of the EB irradiation dose for all cases. The metal ions M and HA are divided between aqueous solution and solid materials or precipitated as follows: 2+ 2+ 2+ M2+ Tot = Maq + Mads + Mred
(4)
HATot = HAaq + HAads + HAox
(5)
the contaminant, respectively. As shown in Fig. 2 the experimental data were better fitted and linearized with minimum error (R2 ) according to (Eq. (7)); R2 > 0.98 (Table 1), for reduction of Cu2+ , Sr2+ , and Co2+ in wastewater solutions under different conditions on the entire dose range. As presented in Table 1, the values of dose constants (K1 ), at pH 5.0 and room temperature (298 ± 1 K), for Cu2+ , Sr2+ , and Co2+ in the presence of the combined TiO2 /UV + HA are about ten times larger than the K1 -values of other reactions. This may be used as a tool for comparison of the radiolysis based experiments. Fig. 3 reveals that the Cu2+ metal ions are the best to be completely removed from wastewater solutions using the combined TiO2 /UV + HA at pH 5.0 and room temperature (298 ± 1 K). 3.3. Percentage removal of metal ions from aqueous solutions The percentage removal of Cu2+ , Sr2+ , and Co2+ from aqueous solutions under different conditions is shown in Fig. 4. In the absence of HA and TiO2 , Fig. 4a–c shows that the percentage removal of Cu2+ , Sr2+ , and Co2+ ions was 41%, 83% and 60%, respectively, at 125 kGy. In the presence of TiO2 suspension and exposure of the investigated wastewater to UV before irradiation by the EB the percentage removal of Cu2+ ions became 51%, while the percentage removal of both Sr2+ and Co2+ ions was slightly improved; 87% and 75%, respectively at the same EB dose. On the other hand, by introducing the combination of TiO2 /UV + HA, only an irradiation dose of about 50 kGy led to removal of Cu2+ , Sr2+ , and Co2+ completely from the wastewater. In the presence of UV irradiated suspended TiO2 /HA, the applied EB irradiation dose decreased from about 60.88 kGy to 30 kGy to remove 75% of Sr2+ (Fig. 4a) and from about 125 kGy to 32 kGy for Co2+ (Fig. 4b), and from >150 kGy to about 38.4 kGy for Cu2+ (Fig. 4c) at the same percentage removal.
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(a) 0
(b)
-1
ln (C/Co)
ln(C/Co)
-1
-2
-3
-4
0
2+
20
40
60
80
-3 2+
-4
Sr +EB 2+ Sr +EB+TiO2/UV 2+ Sr +EB+HA+TiO2/UV 0
-2
-5
100 120 140
Co +EB 2+ Co +EB+TiO2/UV 2+ Co +EB+HA+TiO2/UV 0
20
40
Dose (kGy) (c)
0 2+
-1.5
ln (C/Co)
60 80 100 120 140 160 Dose (kGy)
Cu +EB 2+ Cu +EB+TiO2/UV 2+ Cu +EB+HA+TiO2/UV
-3.0
-4.5
-6.0
-7.5
0
20
40
60
80
100 120 140 160
Dose (kGy) Fig. 2. Comparison the effect of EB, combined TiO2 /UV and EB, and combined TiO2 /UV and EB and HA on the reduction of Sr2+ (a), Co2+ (b) and Cu2+ (c) in wastewater solutions at pH 5.0 and room temperature (298 ± 1 K).
Furthermore, it was found that the EB irradiation dose reduces from >150 kGy to about 50 kGy to completely remove the Cu2+ , Sr2+ , and Co2+ metal ions from wastewater solutions. This indicates that the energy was saved by about 66% when using TiO2 /UV + HA rather than using EB irradiation technique alone. This could be attributed to presence of HA and its role as scavenger and electron production due to UV irradiation of TiO2 [26].
h
HAaq HA∗aq h−
(8) (9)
HA upon irradiation with UV or visible light excites an electron from its highest occupied molecular orbital state to the lowest unoccupied molecular orbital state. The excited state may either undergo
ln (C/Co)
The adsorption of HA on TiO2 is a prerequisite for an efficient electron injection and subsequent oxidation of HA. Humic acid molecules adsorbed on TiO2 are excited by absorbing visible light and subsequently inject electrons to conduction band (CB) of TiO2. The elementary reaction steps of HA on the TiO2 /water interface can be represented by the following equations: Kad
-1 -2
3.4. Adsorption of HA on TiO2
HAaq + TiO2 HA–TiO2
0
-3 -4 2+
Sr 2+ Co 2+ Cu
-5 -6 0
5
10 15 20 25 30 35 40 45 50 55 Dose (kGy)
Fig. 3. Comparison the effect of HA on the reduction of Sr2+ , Co2+ and Cu2+ in wastewater solutions under the effect of EB and TiO2 /UV at pH 5.0 and room temperature (298 ± 1 K).
A.A. Zaki, N.A. El-Gendy / Journal of Hazardous Materials 271 (2014) 275–282
100
(a)
100
2+
Cu +EB+HA+TiO2 2+ Cu +EB+TiO2/UV 2+ Cu +EB
80
279
(b)
90 80
% Removal
% Removal
70 60
40
60 50 40 30
20
2+
Sr +EB+HA+TiO2/UV 2+ Sr +EB+TiO2/UV 2+ Sr +EB
20 10
0
0
20
40
60
80
0
100 120 140 160
0
20
Dose(kGy)
100
40
60
80
100 120 140 160
Dose (kGy)
(c)
% Removal
80
60
40 2+
20
0
Co +EB+HA+TiO2/UV 2+ Co +EB+TiO2/UV 2+ Co +EB 0
20
40
60
80
100 120 140 160
Dose (kGy) Fig. 4. Percentage removal of Cu2+ (a), Sr2+ (b) and Co2+ (c) from wastewater solutions under different effects of EB, combined TiO2 /UV and EB, and combined TiO2 /UV and EB and HA at pH 5.0 and room temperature (298 ± 1 K).
relaxation process by non-radiative and fluorescence paths or undergoes subsequent sensitization reactions. In the presence of TiO2 , HA after UV or visible light induced excitation subsequently sensitizes a semiconductor oxide surface by the electron transfer.
If the energy of the excited state of HA is higher than the conduction band of TiO2 , an electron is injected from the excited-state HA into the TiO2 conduction band (CB). As a result, HA is converted to a cationic radical (HA·+ ), and the injected electron in the TiO2
Fig. 5. Percentage uptake of HA onto Ti2 O versus pH (a) and, light excitation of HA and conduction electron band of TiO2 and their interactions.
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Fig. 5b illustrates the UV/vis light excitation of HA and emission of an electron to the conduction band of TiO2 . 3.5. Sorption mechanisms
Fig. 6. FTIR spectra of HA and HA + Cu2+ , HA + Co2+ , and HA + Sr2+ ion metals.
conduction band can reduce the surface-sorbed oxygen O2 produc• ing O2 − or • OOH as a reactive species that can also disproportionate to give • OH (Fig. 5a) [27]. h
HA–TiO2 −→HA∗ –TiO2
(10)
HA∗aq
(11)
+ TiO2 →
HAox + TiO2 (e− ) CB
HA∗ –TiO2 → HAox + TiO2 (e− ) CB
(12)
So, in the presence of TiO2 , the degradation/stability of HA/UV depends on whether the injected electrons transfer to dissolved O2 (Eq. (14)), or recombine with HAox (Eq. (13)). HAox –TiO2 (e− ) → HA–TiO2 CB HAox –TiO2 (e− ) + O2 CB
→
(13)
HAox –TiO2 + O− 2
(14)
HA molecules adsorbed on TiO2 (HA–TiO2 ) are excited by adsorbing UV light and subsequently oxidized due to injection of electrons to CB of TiO2 . The sorption of HA on TiO2 is a pH sensitive, the uptake of HA on TiO2 can be divided into two main pH dependent regions; before and after pH 6. At pH < 6.0, the surface of TiO2 takes positive charge and negative charge at pH > 6.0. The pKa of the HA under the experimental condition is 5.4 [26]. At pH 2.0 the sorption of HA on TiO2 surface is very high as a result of the positive-negative charge attraction between HA negatively charged groups and the TiO2 positively charged surface (Fig. 5a). HA with many acidic functional groups are deprotonated and negatively charged at alkaline solution (pH > 7.0). Therefore, the adsorption of HA at alkaline solution is strongly retarded due to the negative–negative electrostatic repulsion (Fig. 5a) [26]. The adsorption of HA on Ti2 O affects the removal of the studied ion metals from wastewater. COOH
COOH R
O
HO OH
OH
HC
COOH HO
O
O
(HC OH)4
CH N
Fig. 6 shows the FTIR spectra of the HA with and without sorbed Cu2+ , Co2+ and Sr2+ metal ions. Bands characteristic of HA are observed in the FTIR spectra in the regions of 3400 cm−1 (H-bonded OH stretching of carboxyl, phenol and alcohol), 1600–1650 cm−1 (C O stretching of COO , ketonic C O and aromatic C C conjugated with COO ), 1400 cm−1 (aliphatic C H bending and COO asymmetric stretching) [26,28]. The strong transmission peak at 1037 cm−1 represents C O of alcohol and polysaccharide and the bands below 1000 cm−1 (915 cm−1 and 687 cm−1 ) suggest the presence of substituted aromatics [26,28] (Figs. 6 and 7). The sharpening of FTIR bands at the wave number range 1624–1629 cm−1 could be attributed to the adsorption of Cu2+ , Co2+ and Sr2+ metal ions. Sharpening and broadening of bands at the wave numbers (Fig. 6) show that the metals bind to the HA at carboxylic acid sites and on OH− groups. The binding of metal ions produced the sharpening of the peaks at 1624 cm−1 –1629 cm−1 because it decreased the routes for vibrational relaxation to a lower energy level. The broadening of OH peaks at 3384, 3401, and 3417 cm−1 is caused by the binding of the metal ions or by increased hydrogen bonding (Fig. 6). As shown in Fig. 7 the EB irradiation of HA can lead to disintegrate the structure of HA such as COOH and OH resulting in more active groups. HA is able to adsorb Cu2+ , Sr2+ , and Co2+ on its surface and precipitate as a result of EB irradiation. The bond formation between metal ions and HA among others can proceed through carboxylic and phenolic groups, which leads to the liberation of H + ions and therefore to the increase of the acidity of system (Fig. 7). So, with the increasing the pH, a dramatic increase of the uptake of sorbed metal ions is expected because the principal adsorption sites COOH and COH dissociate to their anionic forms COO and CO . These dissociations cause negatively charged surfaces and cations could more easily adsorb onto the solid surface. At pH value beyond 8, it was observed a sharp decrease at sorption capacity of sorbent [29]. The major mechanisms by which metals adsorb onto HA surface have been proposed to involve anion exchange (electrostatic interaction), ligand exchange surface complexing, hydrophobic interaction, entropic effects, hydrogen bonding, and cation bridging [30]. The capacity of HA to combine with metals is usually attributed to their high contents of substituents such as carboxyl (COOH), hydroxyl (OH), and carbonyl (C O) [31,32]. The mechanisms of interaction between divalent metal ions such as Cu2+ , Sr2+ , Co2+ and oxygen functional groups on humic substances are illustrated in Fig. 8. Previous investigations have suggested a number of chelate linkages. One possibility is a linkage between a metal and two adjacent hydroxyls in a benzenediol (catechol) component, as shown
H O
HC
CH CH2 CH N
O
R
O
NH CH C O NH
O
Fig. 7. The expected structure of HA.
H
O
OH
O
O
O
O
O
COOH O
O OH
COOH
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OH
+ M
(a)
O
C
OH
2+ + M
OH
C
C
(c)
O
2+
M
+
OH
C
C
O
O
M
+
2H+
Acknowledgements
M O
+
2H
+
O
O
O
OH
O
C
C
C
C
CH
(e)
O
O
O OH
(d)
+
O
O C
2H
+
M
OH
(b)
The mechanisms of interactions between ketone carbonyls, phenol, carbonyls, and COOH groups of HA and Cu2+ , Co2+ and Sr2+ metal ions were suggested and discussed. The removal of metal ions Cu2+ , Sr2+ , Co2+ from wastewaters by HA is as a result of the presence of oxygen functional groups such as carboxyl (COOH), hydroxyl (OH), and carbonyl (C O) on HA.
O
2+
CH
OH
O
C
C
M O
+
M
O
2+
CH
C
CH
CH
M O
(f)
O
C
C CH
O
+
M
O
2+
CH
C
281
CH
Fig. 8. The possible reaction mechanisms of Sr2+ , Co2+ and Cu2+ metal ions uptake by HA.
in Fig. 7 [31,33]. Several investigations have suggested two main types of chelate linkages, one involving two carboxylic groups in close proximity to form a phthalate-like ring, seen in Fig. 8b. The other involves a carboxyl and an adjacent phenolic OH to form a salicylate-like ring, shown in Fig. 8c [31]. Infrared spectroscopic measurements of humic substances yielded results indicative of the participation of conjugated ketone structures, which are known to form complexes with metal ions [31]. The mechanisms of interactions between ketone carbonyls and divalent metals are shown in Fig. 8d–f. The ease of dissociation of a particular COOH group depends on the nature and position of the other functional groups close to it in the same molecule. Oxygen containing functional groups adjacent to the carboxylic acid group generally causes an increase in acid strength, and dissociation occurs at a lower pH. Phenols are generally much weaker acids than carboxylic acids, but their dissociation likewise depends on the nature of other functional groups nearby in the molecule. Phenol groups form strong complexes with metal ions as seen in Fig. 8a and a phenol group adjacent to COOH Fig. 8c is a very effective chelator [34,35]. 4. Conclusions A combined electron beam (EB) and TiO2 /UV + HA treatment process was developed for removal of Cu2+ , Sr2+ , and Co2+ from wastewater. Using this developed process, about 50 kGy absorbed dose of EB was enough to treat the studied wastewater and completely remove the Cu2+ , Sr2+ , and Co2+ from liquid solutions. The energy used for treatment of the studied wastewater and removal of the Cu2+ , Sr2+ , and Co2+ ions was reduced by about 66% using the proposed process. The K1 -values; reaction constant K1 (Gy−1 ) for the studied metal ions using the combined process (TiO2 /UV + HA) are about 10 times larger than that of other treatment reactions.
The authors are grateful to the HLWMC and NCRRT of Atomic Energy Authority of Egypt for the financial and academic support throughout this work. References [1] M.I. Ojovan, W.E. Lee, W.E. Lee, An Introduction to Nuclear Waste Immobilisation, Elsevier Ltd., The Netherlands, 2005. [2] R.O. Abdel Rahman, H.A. Ibrahim, Y.-T. Hung, Liquid radioactive wastes treatment: a review, Water 3 (2011) 551–565. [3] R. Lauwerys, D. Lison, Health risks associated with cobalt exposure: an overview, Sci. Total Environ. 150 (1994) 1–6. [4] S. Robey, M. Veazey, D. Crawford, Cobalt-induced myxedema: report of a case, Med. Intell. 255 (1956) 955–957. [5] L.J. Chen, L.-Y. Tang, J.-R. He, Y. Su, Y.-L. Cen, D.-D. Yu, B.-H. Wu, Y. Lin, W.-q. Chen, E.-W. Song, Z.-F. Ren, Urinary strontium and the risk of breast cancer: a case–control study in Guangzhou, China, Environ. Res. 112 (2012) 212–217. [6] B.M. Kirrane, L.S. Nelson, R.S. Hoffman, Massive strontium ferrite ingestion without acute toxicity, Basic Clin. Pharmacol. Toxicol. 99 (2006) 358–359. [7] K. Usuda, T. Dote, M. Watanabe, H. Shimizu, Y. Tanimo, E. Yamadori, An overview of boron, lithium, and strontium in human health and profiles of these elements in urine of Japanese, Environ. Health Prev. Med. 12 (2007) 231–237. [8] J. Schauer, G.C. Lough, M.M. Shafer, W.F. Christensen, M.F. Arndt, J.T. De Minter, Characterization of metals emitted from motor vehicles, Res. Rep. Health Eff. Inst. 133 (2006) 77–88. [9] S.P. Nielsen, The biological role of strontium, Bone 35 (2004) 583–588. [10] L. Daniel, O. Manuel, U. Ricardo, A. Magdalena, Risks and benefits of copper in light of new insights of copper homeostasis, J. Trace Elem. Med. Biol. 25 (2011) 3–13. [11] J.W. Cooper, D. Curry, E. O‘Shea (Eds.), Environmental Application of Ionizing Radiation, Wiley, New York, 1998, pp. 395–415. [12] P. Gehringer, H. Eschweiler, Radiation-induced cleanup of water and waste water, J.W. Cooper, D. Curry, R. O‘Shea (Eds.), Environmental Applications of Ionizing Radiation, Wiley, New York, 1998, pp. 325–351. [13] W. Tang, G. Zeng, J. Gong, Impact of humic/fulfic acid on the removal of heavy metals from aqueous solutions using nanomaterials: a review, Sci. Total Environ. 468–469 (2014) 1014–1027. [14] F.J. Stevenson, Humus Chemistry: Genesis, Composition, Reactions, John Wiley & Sons, New York, 1994, pp. 496. [15] C. Youngmin, C. Wonyong, Visible light-induced reactions of humic acids on TiO2 , J. Photochem. Photobiol. A: Chem. 148 (2002) 129–135. [16] J. Slawinski, H. Gorski, H. Manikowski, EPR spectra of humic acids and melanins exposed to UV radiation and ozone, Curr. Top. Biophys. 23 (1998) 103–112. [17] K. Polewski, D. Slawinska, J. Slawinski, The role of orto- and para-semiquinones in the chemiluminescence and antioxidizing activity of humus substances, P.E. Stanley, L.J. Kricka (Eds.), Bioluminescence and Chemiluminescence, World Scientific, New Jersey, 2002, pp. 161–164. [18] Y. Cho, W. Choi, C.H. Lee, T. Hyeon, H.I. Lee, Visible light-induced degradation of carbon tetrachloride on dye-sensitized TiO2 , Environ. Sci. Technol. 35 (2001) 966–970. [19] D. Chen, A.K. Ray, Removal of toxic metal ions from waste water by semiconductor photo catalysis, Chem. Eng. Sci. 56 (2001) 1561–1570. [20] Y. Long, W. Shutao, Y. Hong, Y. Jie, X. Kang, Photocatalytic reduction of perchlorate in aqueous solutions in UV/Cu–TiO2 /SiO2 system, Chem. Eng. J. 226 (2013) 434–443. [21] L. Wenjuan, Z. Yanrong, M. Zhang, M. Muthu, Y. Sachio, Photoelectrocatalytic degradation of microcystin-LR using Ag/AgCl/TiO2 nanotube arrays electrode under visible light irradiation, Chem. Eng. J. 231 (2013) 455–463. [22] F.J. Rodríguez, P. Schlenger, M. García-Valverde, A comprehensive structural evaluation of humic substances using several fluorescence techniques before and after ozonation, Part II: Evaluation of structural changes following ozonation, Sci. Total Environ. 476–477 (2014) 731–742. [23] G.V. Buxton, C.L. Greenstock, W.P. Helman, W.P. Ross, Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals in aqueous solution, J. Phys. Chem. Ref. Data 17 (1988) 513–886. [24] M.G. Nickelsen, W.J. Cooper, T.D. Waite, C.N. Kuruzc, Removal of benzene and selected alkyl-substituted benzenes from aqueous solution utilizing continuous high energy electron irradiation, Environ. Sci. Technol. 26 (1992) 144–152. [25] Y. Li, A. Seacat, P. Kuppusamy, J.L. Zweier, J.D. Yager, M.A. Trush, Copper redoxdependent activation of 2-tert-butyl (1,4) hydroquinone: formation of reactive
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Contents lists available at ScienceDirect
Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Removal of metal ions from wastewater using EB irradiation in combination with HA/TiO2 /UV treatment A.A. Zaki a,∗ , Naima A. El-Gendy b a b
Hot Laboratory and Waste Management Center (HLWMC), Atomic Energy Authority, P.O. Box 13758, Inshas, Cairo, Egypt Department of Polymer, National Center for Radiation Research and Technology(NCRRT), Atomic Energy Authority, Nasr City, P.O. Box 19, Cairo, Egypt
h i g h l i g h t s • • • •
A combined electron beam (EB) and humic acid (HA/TiO2 ) treatment process was developed. The ion metals Cu2+ , Sr2+ , and Co2+ were effectively removed from wastewater. About 50 kGy of EB was needed to remove the Cu2+ , Sr2+ , and Co2+ from wastewater. Mechanisms of interactions between HA and metal ions were suggested and discussed.
a r t i c l e
i n f o
Article history: Received 2 December 2013 Received in revised form 16 February 2014 Accepted 18 February 2014 Available online 28 February 2014 Keywords: Electron beam Photocatalysis Humic acid Removal Mechanism
a b s t r a c t The electron beam (EB) irradiation technology was applied for removal of Cu2+ , Sr2+ , and Co2+ ions from wastewater. The aim of this study is to achieve an efficient treatment process of wastewater using EB and introducing a combination of humic acid (HA) as a natural organic polymer and ultraviolet irradiation of a TiO2 (TiO2 /UV), as a suspended catalyst in the treatment of wastewater solutions (TiO2 /UV + HA). The experimental results showed that the percentage removal of Cu2+ , Sr2+ , and Co2+ ions was 41%, 87% and 75% respectively, at 125 kGy. In the presence of TiO2 photocatalyst and exposure of the investigated wastewater to ultraviolet rays before irradiation by the EB the percentage removal of Cu2+ ions became 51%, while the percentage removal of both Sr2+ and Co2+ ions was slightly improved; was 87% and 75%, respectively at the same EB dose. On the other hand, by introducing the combination of TiO2 /UV + HA, only an irradiation dose of about 50 kGy led to removal of Cu2+ , Sr2+ , and Co2+ completely from the wastewater. Mechanisms of interactions between HA and Cu2+ , Co2+ and Sr2+ metal ions were suggested and discussed. © 2014 Elsevier B.V. All rights reserved.
1. Introduction The handling of wastewater appeared to be one of the most important problems in the last few decades. Aqueous wastes may be treated using ion exchange, sorption, chemical precipitation, evaporation, reverses osmosis, filtration and solvent extraction [1]. The main features and limitation of these treatment processes are described elsewhere [2]. Occupational or environmental exposure to cobalt is associated with health effects on various organs, including the respiratory tract, the skin, the red blood cells, the thyroid gland and the myocardium [3]. Historically, effects on the thyroid gland were observed as an undesirable side-effect of cobalt therapy used for
∗ Corresponding author. Tel.: +20 2 35620445; fax: +20 2 4620796. E-mail address: [email protected] (A.A. Zaki). http://dx.doi.org/10.1016/j.jhazmat.2014.02.025 0304-3894/© 2014 Elsevier B.V. All rights reserved.
its polycythemic action, often in association with iron, in the treatment of various anemias [4]. The major industrial uses of strontium include the production of glass for color television picture tubes, ferrite magnets, and refining of zinc [5–8]. Copper is an essential micronutrient that forms part of several proteins involved in a variety of biological processes indispensable to sustain life [9]. At the same time, it can be toxic when present in excess, the most noticeable chronic effect being liver damage [10]. Radiation technologies are frequently applied for purification of wastewater and groundwater contaminated with organic pollutants [11,12]. Humic acids (HAs) are macromolecular yellow-to-black colored natural organic matter derived from the degradation of plant, algal, and microbial material [13]. Although their formation mechanism and chemical structures are not well understood, they are known to be high in carbon content (50–60%) of both aliphatic and aromatic character and rich in oxygen-containing functionalities such as carboxyl,
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phenolic, alcoholic, and quinoid groups [14,15]. The absorption of light by chromophores existing in humic acid (HA) may lead to alterations of HA structure and composition which is reflected by the changes in absorption, fluorescence, chemiluminescence, NMR, IR and EPR spectra [16,17]. Photocatalytic oxidation using TiO2 is gaining wide attentions as an advanced water treatment technology. TiO2 photocatalyst initiates upon absorbing UV photons with generating conduction band (CB) electrons and valence band (VB) holes within the particle [18,19]. In particular, the hydroxyl radicals generated through the reaction of VB holes account for the strong oxidizing power of the TiO2 photocatalytic system. Although TiO2 has been widely used as a photocatalyst, the band gap excitation of TiO2 proceeds only in the presence of UV light ( < 388 nm) [20]. Sensitized TiO2 by humic acid (HA–TiO2 ), ruthenium-based complexes or various dyes has been extensively investigated in order to extend the spectral response into the visible region [21]. The aim of this work is to achieve an efficient and energy saving process for treatment of wastewater and removal of Cu2+ , Sr2+ and Co2+ ions from waste solutions using a combination of electron beam irradiation (EB), HA sensitized UV-irradiated Ti2 O catalyst (HA–TiO2 ). 2. Materials and methods 2.1. Materials All chemicals used in this work were of analytical grade. All aqueous solutions were prepared with distilled deionized water, which was purified by a Millipore Milli-Q water (resistivity = 4.4 M cm, TOC = 0.1 mg/l) purification system. Chemical powders used to simulate wastewater were supplied from Merck; CuSO4 ·5H2 O, CoCl2 , SrCl2 ·4H2 O. Stock solutions of 100 mg/l from different metal ions were prepared using distilled deionized water. Humic acid (HA) in sodium form and TiO2 powder, Degases P25 (80% anatase and 20% rutile), were purchased from Aldrich. The TiO2 powder had a BET surface area of 55 m2 /g and an average particle diameter of 30 nm. HA was initially dissolved in alkaline solution (pH ≈ 12) at 100 mg/l and subsequently adjusted to pH 5.0 using 1 N HCl. All experiments were done at room temperature (298 ± 1 K) otherwise is mentioned. The HA filtrate was stored as stock solution and its concentration was routinely checked. 2.2. Electron-beam research facility Experiments to study the electron beam (EB) treatment process were carried out using NCRRT-Accelerator. The NCRRT-Accelerator is a direct accelerator delivered from High Voltage Company, USA; 1.5 MeV, beam current of 25 mA, power of 37.5 kW and variable scan width up to 90 cm. This accelerator is used to produce an EB with high energy which can be used for different applications in different fields.
2 h at 200 rpm. The percentage uptake of HA by Ti2 O was calculated according to the formula (C0 − Ce /C0 ) × 100, where C0 and Ce are the initial and equilibrium concentrations of HA, respectively. The dissolved humic acid was spectrophotometerically measured using the UV–vis spectrophotometer, Shimadzu-UV 2550, at = 254 nm. 2.5. Photocatalytic TiO2 /HA/metal ion process Photocatalytic TiO2 powder (Degases P25) is used in photocatalytic experiments due to its high UV-light efficiency. The TiO2 suspension was prepared by simultaneous sonication and shaking in an ultrasonic cleaning bath (Branson 3210). Solution of HA was then added to a 100 mg/l TiO2 suspension to give a desired concentration of HA and added to metal ion solutions. A removable 30 cm long UV-florescent lamp, model DESAGA Abnehmber was used to produce light for irradiating the TiO2 /HA/metal ion solution. It has two wavelengths 254 nm and 396 nm, power of 30 W and the incident light is 60 W/m2 . 2.6. Procedures of experimental setting up The liquid solution samples were divided into three groups. The first group samples were irradiated for different doses in the range 25–150 kGy. The second group samples were exposed to UV-lamp with photocatalytic TiO2 powder (Degasses P25) suspended in solution for 2 h (checked well during irradiation) then directly moved to the EB irradiation. The third group samples were mixed with HA and exposed to UV-lamp with TiO2 photocatalytic for 2 h then directly moved to EB to complete the irradiation. After EB irradiation the samples were centrifuged at 5000 rpm for 10 min before measurement on an atomic absorption spectrophotometer (AAS) as well as the metal ions in the samples of the other two groups. The AAS utilizes air acetylene flame, is a Boch Scientific model 210 VGP, USA. 2.7. Sorption mechanism Fourier transform infrared spectroscopy (FTIR) was used to investigate the sorption mechanisms of the studied metal ions onto HA. For FTIR experiments, 1 g mass of finely powdered HA placed in 250 cm3 beaker, 100 cm3 of 0.025 M metal salt solution was added and adjusted to pH 5.0. The suspension was magnetically stirred at 333 K for half an hour, and then left for 24 h to saturate with the metal ion. Resulting suspensions of HA + M complexes were filtered and rinsed with distilled water. Finally the solid complexes were dried at 353 K for 4 h. The dry complexes were mixed with 50 mg KBr and compressed to obtain a pellet. The pellets were dried at 378 K for 2 days and then, analyzed using a MIDAC 1700 model FTIR spectrometer. The base line was corrected and scanning was performed from 4000 to 400 cm−1 . 3. Results and discussion
2.3. Electron beam irradiation (EB) radiation of HA molecules 3.1. Effect of EB irradiation on HA molecules The stability of HA, 50 mg/l, was studied by applying electron beam irradiation (EB) at doses of 0–32 kGy, and its ultraviolet (UV) absorbance was checked. A Unicam UV4 double beam spectrophotometer (product of Unicam Co. Ltd., England) was used to measure absorption spectra of the EB irradiated and unirradiated HA. 2.4. Effect of pH on the uptake of HA by Ti2 O Uptake studies of HA by Ti2 O, humic acid–titanium dioxide sensitization (HA–Ti2 O), were performing by mixing 500 mg/l Ti2 O with 50 mg of HA for pH ranging from 2.0 to 13.0. The mixture was shaken in a thermostatic shaker at room temperature (298 ± 1 K) for
Humic acid (HA) substances effectively absorb non-ionizing radiation, such as ultraviolet (UV), visible or infrared radiation. Much less is known about their sensitivity to ionizing radiation, such as EB, ␥- or X-rays. The UV absorption spectra of EB irradiated HA are illustrated in Fig. 1. As shown from the figure, that the humic acid solution at pH 5.0 has a maximum absorbance in the UV region at 320 nm [22]. This absorption belongs to the aromatic rings of HA and a considerable decrease in the absorbance of peak occurred with the increase in the EB irradiation dose. The observed decrease in absorption (Fig. 1) is attributed to degradation and precipitation of HA molecules.
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277
Table 1 Dose constant values, the initial concentrations and coefficient of determination (R2 ) at 45 kGy and pH 5.0 at room temperature (298 ± 1 K) of the studied systems. System
Dose constant K1 (kGy−1 )
Initial concentration C0 (mmol/l)
Coefficient of determination (R2 )
Sr2+ + EB Co2+ + EB Cu2+ + EB Sr2+ + EB + TiO2 /UV Co2+ + EB + TiO2 /UV Cu2+ + EB + TiO2 /UV Sr2+ + EB + TiO2 /UV + HA Co2+ + EB + TiO2 /UV + HA Cu2+ + EB + TiO2 /UV + HA
0.015 0.019 0.004 0.019 0.025 0.006 0.117 0.106 0.120
0.57 0.85 0.79 0.57 0.85 0.79 0.57 0.85 0.79
0.981 0.983 0.980 0.984 0.988 0.985 0.982 0.987 0.982
where the subscription Tot, aq, ads, red, and ox are total, aqueous, adsorption, reduction, and oxidation, respectively. The experimental data for removal of contaminants by EB irradiation were analyzed by fitting into first order equation [24]: C = C0 e−K1 D Fig. 1. Ultraviolet absorbance of HA before and after EB irradiation at pH 5.0 and room temperature (298 ± 1 K).
3.2. Electron beam irradiation (EB) of water and contaminants Irradiation of water or dilute aqueous solution by EB ionizing radiations results in the formation of various species according to the following Eq. (1) [23]:
ln
C C0
= −K1 D
(6) (7)
The linearization of the data by logarithmic representation (Eq. (7)) gives the dose constant K1 (Gy−1 ). The K1 -value is useful for the comparison of the radiolysis-based experiments, D (kGy) is the EB irradiation dose, C (mol/l) and C0 (mol/l) are the concentration at the applied dose D (kGy) and the initial concentration of
(1) The values in brackets (G) are the number of micromoles of molecules or radicals produced per joule absorbed energy (mol/J) and represent the efficiency of the conversion of the high energy electron radiation to a chemical process. Among the chemical species formed during the radiolysis of water, the hydroxyl radical (• OH), hydrated electron (e− aq ) and hydrogen atom (H• ) are the most reactive species with organic compounds. In aerated solution, the dissolved oxygen concentration is approximately 3 × 10−4 M (9.4 mg/l). In this case, solvated electrons (e− aq ) and hydrogen atoms (H• ) are practically scavenged by oxygen and rapidly converted into HO2 • , which are very poor reacting species compared to the hydroxyl radical (Eqs. (2) and (3)) [23]. e− aq + O2 → O2 H•
•−
+ O2 → HO2
•
k = 1.9 × 1010 M−1 s−1 k = 2.1 × 10
10
−1 −1
M
s
(2) (3)
where k (M−1 s−1 ) is the rate constant. The metal ions were efficiently reduced to lower oxidation states by both e− aq and H• but in some cases, e− aq can lead to reduction in metal ions alone. However, re-oxidation by hydroxyl radical (• OH), and H2 O2 should be prevented. Several kinetic studies referring to reduction of H2 O2 were found elsewhere [24,25]. The experimental data were presented as the ln(C/C0 ) versus the applied absorbed dose, D (kGy) of the systems EB + metal ions, EB + metal ions + TiO2 /UV, and EB + metal ions + TiO2 /UV + HA (Figs. 2 and 3). It was found that the relative concentration of metal ions, ln(C/C0 ) decreases with the increase of the EB irradiation dose for all cases. The metal ions M and HA are divided between aqueous solution and solid materials or precipitated as follows: 2+ 2+ 2+ M2+ Tot = Maq + Mads + Mred
(4)
HATot = HAaq + HAads + HAox
(5)
the contaminant, respectively. As shown in Fig. 2 the experimental data were better fitted and linearized with minimum error (R2 ) according to (Eq. (7)); R2 > 0.98 (Table 1), for reduction of Cu2+ , Sr2+ , and Co2+ in wastewater solutions under different conditions on the entire dose range. As presented in Table 1, the values of dose constants (K1 ), at pH 5.0 and room temperature (298 ± 1 K), for Cu2+ , Sr2+ , and Co2+ in the presence of the combined TiO2 /UV + HA are about ten times larger than the K1 -values of other reactions. This may be used as a tool for comparison of the radiolysis based experiments. Fig. 3 reveals that the Cu2+ metal ions are the best to be completely removed from wastewater solutions using the combined TiO2 /UV + HA at pH 5.0 and room temperature (298 ± 1 K). 3.3. Percentage removal of metal ions from aqueous solutions The percentage removal of Cu2+ , Sr2+ , and Co2+ from aqueous solutions under different conditions is shown in Fig. 4. In the absence of HA and TiO2 , Fig. 4a–c shows that the percentage removal of Cu2+ , Sr2+ , and Co2+ ions was 41%, 83% and 60%, respectively, at 125 kGy. In the presence of TiO2 suspension and exposure of the investigated wastewater to UV before irradiation by the EB the percentage removal of Cu2+ ions became 51%, while the percentage removal of both Sr2+ and Co2+ ions was slightly improved; 87% and 75%, respectively at the same EB dose. On the other hand, by introducing the combination of TiO2 /UV + HA, only an irradiation dose of about 50 kGy led to removal of Cu2+ , Sr2+ , and Co2+ completely from the wastewater. In the presence of UV irradiated suspended TiO2 /HA, the applied EB irradiation dose decreased from about 60.88 kGy to 30 kGy to remove 75% of Sr2+ (Fig. 4a) and from about 125 kGy to 32 kGy for Co2+ (Fig. 4b), and from >150 kGy to about 38.4 kGy for Cu2+ (Fig. 4c) at the same percentage removal.
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(a) 0
(b)
-1
ln (C/Co)
ln(C/Co)
-1
-2
-3
-4
0
2+
20
40
60
80
-3 2+
-4
Sr +EB 2+ Sr +EB+TiO2/UV 2+ Sr +EB+HA+TiO2/UV 0
-2
-5
100 120 140
Co +EB 2+ Co +EB+TiO2/UV 2+ Co +EB+HA+TiO2/UV 0
20
40
Dose (kGy) (c)
0 2+
-1.5
ln (C/Co)
60 80 100 120 140 160 Dose (kGy)
Cu +EB 2+ Cu +EB+TiO2/UV 2+ Cu +EB+HA+TiO2/UV
-3.0
-4.5
-6.0
-7.5
0
20
40
60
80
100 120 140 160
Dose (kGy) Fig. 2. Comparison the effect of EB, combined TiO2 /UV and EB, and combined TiO2 /UV and EB and HA on the reduction of Sr2+ (a), Co2+ (b) and Cu2+ (c) in wastewater solutions at pH 5.0 and room temperature (298 ± 1 K).
Furthermore, it was found that the EB irradiation dose reduces from >150 kGy to about 50 kGy to completely remove the Cu2+ , Sr2+ , and Co2+ metal ions from wastewater solutions. This indicates that the energy was saved by about 66% when using TiO2 /UV + HA rather than using EB irradiation technique alone. This could be attributed to presence of HA and its role as scavenger and electron production due to UV irradiation of TiO2 [26].
h
HAaq HA∗aq h−
(8) (9)
HA upon irradiation with UV or visible light excites an electron from its highest occupied molecular orbital state to the lowest unoccupied molecular orbital state. The excited state may either undergo
ln (C/Co)
The adsorption of HA on TiO2 is a prerequisite for an efficient electron injection and subsequent oxidation of HA. Humic acid molecules adsorbed on TiO2 are excited by absorbing visible light and subsequently inject electrons to conduction band (CB) of TiO2. The elementary reaction steps of HA on the TiO2 /water interface can be represented by the following equations: Kad
-1 -2
3.4. Adsorption of HA on TiO2
HAaq + TiO2 HA–TiO2
0
-3 -4 2+
Sr 2+ Co 2+ Cu
-5 -6 0
5
10 15 20 25 30 35 40 45 50 55 Dose (kGy)
Fig. 3. Comparison the effect of HA on the reduction of Sr2+ , Co2+ and Cu2+ in wastewater solutions under the effect of EB and TiO2 /UV at pH 5.0 and room temperature (298 ± 1 K).
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100
(a)
100
2+
Cu +EB+HA+TiO2 2+ Cu +EB+TiO2/UV 2+ Cu +EB
80
279
(b)
90 80
% Removal
% Removal
70 60
40
60 50 40 30
20
2+
Sr +EB+HA+TiO2/UV 2+ Sr +EB+TiO2/UV 2+ Sr +EB
20 10
0
0
20
40
60
80
0
100 120 140 160
0
20
Dose(kGy)
100
40
60
80
100 120 140 160
Dose (kGy)
(c)
% Removal
80
60
40 2+
20
0
Co +EB+HA+TiO2/UV 2+ Co +EB+TiO2/UV 2+ Co +EB 0
20
40
60
80
100 120 140 160
Dose (kGy) Fig. 4. Percentage removal of Cu2+ (a), Sr2+ (b) and Co2+ (c) from wastewater solutions under different effects of EB, combined TiO2 /UV and EB, and combined TiO2 /UV and EB and HA at pH 5.0 and room temperature (298 ± 1 K).
relaxation process by non-radiative and fluorescence paths or undergoes subsequent sensitization reactions. In the presence of TiO2 , HA after UV or visible light induced excitation subsequently sensitizes a semiconductor oxide surface by the electron transfer.
If the energy of the excited state of HA is higher than the conduction band of TiO2 , an electron is injected from the excited-state HA into the TiO2 conduction band (CB). As a result, HA is converted to a cationic radical (HA·+ ), and the injected electron in the TiO2
Fig. 5. Percentage uptake of HA onto Ti2 O versus pH (a) and, light excitation of HA and conduction electron band of TiO2 and their interactions.
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Fig. 5b illustrates the UV/vis light excitation of HA and emission of an electron to the conduction band of TiO2 . 3.5. Sorption mechanisms
Fig. 6. FTIR spectra of HA and HA + Cu2+ , HA + Co2+ , and HA + Sr2+ ion metals.
conduction band can reduce the surface-sorbed oxygen O2 produc• ing O2 − or • OOH as a reactive species that can also disproportionate to give • OH (Fig. 5a) [27]. h
HA–TiO2 −→HA∗ –TiO2
(10)
HA∗aq
(11)
+ TiO2 →
HAox + TiO2 (e− ) CB
HA∗ –TiO2 → HAox + TiO2 (e− ) CB
(12)
So, in the presence of TiO2 , the degradation/stability of HA/UV depends on whether the injected electrons transfer to dissolved O2 (Eq. (14)), or recombine with HAox (Eq. (13)). HAox –TiO2 (e− ) → HA–TiO2 CB HAox –TiO2 (e− ) + O2 CB
→
(13)
HAox –TiO2 + O− 2
(14)
HA molecules adsorbed on TiO2 (HA–TiO2 ) are excited by adsorbing UV light and subsequently oxidized due to injection of electrons to CB of TiO2 . The sorption of HA on TiO2 is a pH sensitive, the uptake of HA on TiO2 can be divided into two main pH dependent regions; before and after pH 6. At pH < 6.0, the surface of TiO2 takes positive charge and negative charge at pH > 6.0. The pKa of the HA under the experimental condition is 5.4 [26]. At pH 2.0 the sorption of HA on TiO2 surface is very high as a result of the positive-negative charge attraction between HA negatively charged groups and the TiO2 positively charged surface (Fig. 5a). HA with many acidic functional groups are deprotonated and negatively charged at alkaline solution (pH > 7.0). Therefore, the adsorption of HA at alkaline solution is strongly retarded due to the negative–negative electrostatic repulsion (Fig. 5a) [26]. The adsorption of HA on Ti2 O affects the removal of the studied ion metals from wastewater. COOH
COOH R
O
HO OH
OH
HC
COOH HO
O
O
(HC OH)4
CH N
Fig. 6 shows the FTIR spectra of the HA with and without sorbed Cu2+ , Co2+ and Sr2+ metal ions. Bands characteristic of HA are observed in the FTIR spectra in the regions of 3400 cm−1 (H-bonded OH stretching of carboxyl, phenol and alcohol), 1600–1650 cm−1 (C O stretching of COO , ketonic C O and aromatic C C conjugated with COO ), 1400 cm−1 (aliphatic C H bending and COO asymmetric stretching) [26,28]. The strong transmission peak at 1037 cm−1 represents C O of alcohol and polysaccharide and the bands below 1000 cm−1 (915 cm−1 and 687 cm−1 ) suggest the presence of substituted aromatics [26,28] (Figs. 6 and 7). The sharpening of FTIR bands at the wave number range 1624–1629 cm−1 could be attributed to the adsorption of Cu2+ , Co2+ and Sr2+ metal ions. Sharpening and broadening of bands at the wave numbers (Fig. 6) show that the metals bind to the HA at carboxylic acid sites and on OH− groups. The binding of metal ions produced the sharpening of the peaks at 1624 cm−1 –1629 cm−1 because it decreased the routes for vibrational relaxation to a lower energy level. The broadening of OH peaks at 3384, 3401, and 3417 cm−1 is caused by the binding of the metal ions or by increased hydrogen bonding (Fig. 6). As shown in Fig. 7 the EB irradiation of HA can lead to disintegrate the structure of HA such as COOH and OH resulting in more active groups. HA is able to adsorb Cu2+ , Sr2+ , and Co2+ on its surface and precipitate as a result of EB irradiation. The bond formation between metal ions and HA among others can proceed through carboxylic and phenolic groups, which leads to the liberation of H + ions and therefore to the increase of the acidity of system (Fig. 7). So, with the increasing the pH, a dramatic increase of the uptake of sorbed metal ions is expected because the principal adsorption sites COOH and COH dissociate to their anionic forms COO and CO . These dissociations cause negatively charged surfaces and cations could more easily adsorb onto the solid surface. At pH value beyond 8, it was observed a sharp decrease at sorption capacity of sorbent [29]. The major mechanisms by which metals adsorb onto HA surface have been proposed to involve anion exchange (electrostatic interaction), ligand exchange surface complexing, hydrophobic interaction, entropic effects, hydrogen bonding, and cation bridging [30]. The capacity of HA to combine with metals is usually attributed to their high contents of substituents such as carboxyl (COOH), hydroxyl (OH), and carbonyl (C O) [31,32]. The mechanisms of interaction between divalent metal ions such as Cu2+ , Sr2+ , Co2+ and oxygen functional groups on humic substances are illustrated in Fig. 8. Previous investigations have suggested a number of chelate linkages. One possibility is a linkage between a metal and two adjacent hydroxyls in a benzenediol (catechol) component, as shown
H O
HC
CH CH2 CH N
O
R
O
NH CH C O NH
O
Fig. 7. The expected structure of HA.
H
O
OH
O
O
O
O
O
COOH O
O OH
COOH
A.A. Zaki, N.A. El-Gendy / Journal of Hazardous Materials 271 (2014) 275–282
OH
+ M
(a)
O
C
OH
2+ + M
OH
C
C
(c)
O
2+
M
+
OH
C
C
O
O
M
+
2H+
Acknowledgements
M O
+
2H
+
O
O
O
OH
O
C
C
C
C
CH
(e)
O
O
O OH
(d)
+
O
O C
2H
+
M
OH
(b)
The mechanisms of interactions between ketone carbonyls, phenol, carbonyls, and COOH groups of HA and Cu2+ , Co2+ and Sr2+ metal ions were suggested and discussed. The removal of metal ions Cu2+ , Sr2+ , Co2+ from wastewaters by HA is as a result of the presence of oxygen functional groups such as carboxyl (COOH), hydroxyl (OH), and carbonyl (C O) on HA.
O
2+
CH
OH
O
C
C
M O
+
M
O
2+
CH
C
CH
CH
M O
(f)
O
C
C CH
O
+
M
O
2+
CH
C
281
CH
Fig. 8. The possible reaction mechanisms of Sr2+ , Co2+ and Cu2+ metal ions uptake by HA.
in Fig. 7 [31,33]. Several investigations have suggested two main types of chelate linkages, one involving two carboxylic groups in close proximity to form a phthalate-like ring, seen in Fig. 8b. The other involves a carboxyl and an adjacent phenolic OH to form a salicylate-like ring, shown in Fig. 8c [31]. Infrared spectroscopic measurements of humic substances yielded results indicative of the participation of conjugated ketone structures, which are known to form complexes with metal ions [31]. The mechanisms of interactions between ketone carbonyls and divalent metals are shown in Fig. 8d–f. The ease of dissociation of a particular COOH group depends on the nature and position of the other functional groups close to it in the same molecule. Oxygen containing functional groups adjacent to the carboxylic acid group generally causes an increase in acid strength, and dissociation occurs at a lower pH. Phenols are generally much weaker acids than carboxylic acids, but their dissociation likewise depends on the nature of other functional groups nearby in the molecule. Phenol groups form strong complexes with metal ions as seen in Fig. 8a and a phenol group adjacent to COOH Fig. 8c is a very effective chelator [34,35]. 4. Conclusions A combined electron beam (EB) and TiO2 /UV + HA treatment process was developed for removal of Cu2+ , Sr2+ , and Co2+ from wastewater. Using this developed process, about 50 kGy absorbed dose of EB was enough to treat the studied wastewater and completely remove the Cu2+ , Sr2+ , and Co2+ from liquid solutions. The energy used for treatment of the studied wastewater and removal of the Cu2+ , Sr2+ , and Co2+ ions was reduced by about 66% using the proposed process. The K1 -values; reaction constant K1 (Gy−1 ) for the studied metal ions using the combined process (TiO2 /UV + HA) are about 10 times larger than that of other treatment reactions.
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