Pilot-plant evaluation of TiO2 and TiO2-based hybrid photocatalysts for solar treatment of polluted water

Journal of Hazardous Materials 320 (2016) 469–478

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Pilot-plant evaluation of TiO2 and TiO2 -based hybrid photocatalysts for solar treatment of polluted water Luminita Andronic a,∗ , Luminita Isac a , Sara Miralles-Cuevas d , Maria Visa a , Isabel Oller b,c , Anca Duta a,∗ , Sixto Malato b,c a

Transilvania University of Brasov, Centre of Renewable Energy Systems and Recycling, Romania, Eroilor, 29, 500036, Brasov, Romania Plataforma Solar de Almería, CIEMAT, P.O. Box 22, 04200 Tabernas, Almería, Spain c CIESOL, Joint Centre of the University of Almería-CIEMAT, 04120 Almería, Spain d Laboratory for Environmental Research in Arid Zones, LIMZA, School of Mechanical Engineering, University of Tarapaca, Arica, Chile b

h i g h l i g h t s • • • • •

VIS-active TiO2 –based hybrid powders were tested for the mineralization of three organic pollutants. The photocatalytic removal mechanisms is not influenced by the radiation intensity. In the CPC outdoor pilot, all three pollutants reach maximum mineralization during 40–90 min. The most photocatalytic resistant was phenol (66% removal efficiency) in the CPC using TiO2 . Dichloroacetic acid and imidacloprid removal efficiencies were over 95% when using TiO2 in the CPC.

a r t i c l e

i n f o

Article history: Received 23 May 2016 Received in revised form 3 August 2016 Accepted 4 August 2016 Available online 27 August 2016 Keywords: Titanium dioxide Copper sulfide Fly ash Photocatalysis Solar radiation Phenol Imidacloprid Dichloroacetic acid

a b s t r a c t Materials with photocatalytic and adsorption properties for advanced wastewater treatment targeting reuse were studied. Making use of TiO2 as a well-known photocatalyst, Cu2 S as a Vis-active semiconductor, and fly ash as a good adsorbent, dispersed mixtures/composites were prepared to remove pollutants from wastewater. X-ray diffraction, scanning electron microscopy, energy-dispersive X-Ray spectroscopy, atomic force microscopy, band gap energy, point of zero charge (pHpzc ) and BET porosity were used to characterize the substrates. Phenol, imidacloprid and dichloroacetic acid were used as pollutants for photocatalytic activity of the new photocatalysts. Experiments using the new dispersed powders were carried out at laboratory scale in two solar simulators and under natural solar irradiation at the Plataforma Solar de Almería, in a Compound Parabolic Collector (CPC) for a comparative analysis of pollutants removal and mineralization efficiencies, and to identify features that could facilitate photocatalyst separation and reuse. The results show that radiation intensity significantly affects the phenol degradation rate. The composite mixture of TiO2 and fly ash is 2–3 times less active than sol-gel TiO2 . Photodegradation kinetic data on the highly active TiO2 are compared for pollutants elimination. Photodegradation of dichloroacetic acid was fast and complete after 90 min in the CPC, while after 150 min imidacloprid and phenol removal was 90% and 56% respectively. © 2016 Elsevier B.V. All rights reserved.

1. Introduction Natural organic compounds and synthetic organic microcontaminants such as phenols, surfactants, pharmaceuticals, polychlorinated biphenyls, fertilizers and pesticides, are constantly being released into the environment [1]. Due to their toxicity, such

∗ Corresponding authors. E-mail addresses: [email protected] (L. Andronic), [email protected] (A. Duta). http://dx.doi.org/10.1016/j.jhazmat.2016.08.013 0304-3894/© 2016 Elsevier B.V. All rights reserved.

pollutants have strict discharge limits, and phenolic compounds are on the worldwide priority pollutants list [2]. Conventional wastewater treatment removes most pollutants by cost-effective sedimentation, filtration, adsorption, or biological processes, all of which are considered relatively effective [3]. However, biologically toxic and non-degradable organics can often remain in water at concentrations higher than the ppb discharge or reuse limits. Heterogeneous photocatalysis is an eco-friendly process based on semiconductor materials [4–6]. Although it has been widely studied, scaling up photocatalysis must be made affordable by using

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Vis-active, low/moderate-cost photocatalysts (in easily filterable suspensions) that support rather simple pollutant decomposition by mineralization [7]. Titanium dioxide has been the most widely investigated photocatalyst material since it was discovered by Honda and Fujishima [8] due to its high chemical stability and photosensitivity, biological inertness, low cost and unique optical properties [9–13]. However, photocatalysis on titanium dioxide is still limited by its UV band wavelengths (220–400 nm), and potentially limited photocatalytic activity can be expected in the visible spectral region or under solar irradiation on the Earth’s surface (>290 nm). Therefore, extending the response of TiO2 to the Vis spectral region is a subject frequently addressed (but not yet solved). Many publications reported on doping (anion and cation impurities), on developing TiO2 -based thin film composites combining n–n semiconductors (TiO2 /SnO2 , TiO2 /ZnO, TiO2 /WO3 ) or n-p junctions (TiO2 /NiOx , TiO2 /Cux S, TiO2 /CuOx ), or by activating TiO2 with noble metals (Pt, Pd, Au/TiO2 ) [14], etc. A wide variety of coupled semiconductor oxide systems, such as TiO2 /Cux S [15], ZnO/TiO2 [16], WO3 /TiO2 [17], WS2 /WO3 [18], TiO2 /Fe2 O3 [19], ZnO/Fe2 O3 [20], TiO2 /SnO2 [21], are reported for wastewater pollutants photodegradation [22]. The use of narrow-band-gap semiconductors (e.g., copper sulfide, [23,24]) would solve this problem, but most of these materials are not stable in water, are themselves or produce pollutants (e.g., heavy metals), and recombination of the electrons and holes photocatalytically produced is very fast. Powder photocatalyst with similar composition immobilized on glass by doctor blade [23] or by spray pyrolysis deposition [24] show lower photodegradation efficiency comparing with powders. Obviously, the best pollutant removal efficiencies are expected with suspended nano-powders. However, their separation is difficult, adding new steps to the treatment process (e.g., microfiltration or nanofiltration) if water reuse is targeted. Therefore, the TiO2 photocatalyst is usually deposited on the surface of more easily filtered mezzo or micro-sized nanoparticles. In this study, materials with photocatalytic and adsorption properties were used in wastewater treatment, taking advantages of TiO2 as a well-known photocatalyst, Cux S as Vis-activated materials and fly-ash as a very good adsorbent for the removal of organic pollutants from wastewater. In situ formation of composite heterostructures involving these components is expected, with positive consequences on the Vis-activation and on the Cux S stability. Sol-gel TiO2 and Cux S obtained by photochemical precipitation were analyzed as single or combined photocatalysts and compared with fly ash-TiO2 composites prepared by hydrothermal synthesis, an important pathway for combining the different physical and chemical properties of individual components into one system, yielding hybrid materials. Their photocatalytic activity was studied in the degradation of phenol, imidacloprid and dichloroacetic acid in aqueous suspensions under artificial UV–vis and solar irradiation. Experiments were run under different radiation intensities in two solar- simulators and outdoors, at pilot-plant scale, at the Plataforma Solar de Almería in a non-concentrating Compound Parabolic Collector (CPC) solar radiation system. 2. Materials and methods 2.1. Materials preparation For the targeted upscaling, the single and mixed photocatalysts were prepared chemically which consumes significantly less energy than most physical processes. 2.1.1. Titanium dioxide TiO2 powder was made by the sol-gel technique, using 30 mL titanium tetraisopropoxide (TTIP) added to an HNO3 -ethanol solu-

tion under vigorous stirring for 2 h at room temperature. The gel was obtained after 24 h and then dried at 110 ◦ C for 3 h. The dried TiO2 powder was calcined in air at 450 ◦ C in a furnace for 3 h. 2.1.2. Copper sulfide Photochemical precipitation of copper sulfide was done in a 0.25 M aqueous solution of copper sulfate mixed with a 0.25 M sodium thiosulfate solution at a volumetric ratio of 1:1 at room temperature under direct UV irradiation (108 W). When the pH was adjusted to 3 by adding HNO3 65%, green precipitates developed. After 24 h of UV irradiation, the precipitates turned black, and were then washed with double deionized water and dried at 110 ◦ C for 24 h. The dry Cux S powder was calcinated at 350 ◦ C for 3 h. Reaction conditions, such as precursor concentration, weight ratio, pH and reaction time were used as previously optimized [23,24]. 2.1.3. Fly-ash titanium oxide composites The raw fly ash used in hydrothermal synthesis was collected from the electrofilters of the Combined Heat and Power plant in Brasov-Romania. Its major oxide components are SiO2 (53.7%), Al2 O3 (21.60%) and Fe2 O3 (9.56%), and their total percentage of over 70% makes this fly ash of ASTM Type F [25]. Minority oxide components with potential photocatalytic properties were found: TiO2 (1.07%) and MnO (0.08%). The fly ash was washed in double deionized water (fly ash: water = 1:10 g/mL) under mechanical stirring at room temperature for 48 h, followed by filtration, washing with double deionized water and drying at 105–120 ◦ C for 24 h. The 20–40 ␮m fraction was selected by mechanical sieving (Analysette 3 Spartan). This fraction was dispersed under mechanical stirring in NaOH 2N solution for 24 h (fly ash: NaOH 2N = 1:10 g/mL), followed by filtration, washing with double deionized water and drying at 115 ◦ C for 24 h. During these previously optimized pre-treatment processes, leachable compounds were removed, leaving the pretreated fly ash as a suitable material for wastewater treatment [26,27]. Two fly-ash-TiO2 composites were obtained from the pretreated fly ash (FA) and TiO2 (Degussa P25, cca. 80% anatase, 20% rutile) from the slurry of the powder mixture: Fly 1, using a mass ratio of FA: TiO2 = 3:1, and Fly 2 starting from an FA:TiO2 ratio of 1:1. The components were subjected to hydrothermal synthesis under mild autoclave conditions, using NaOH 2N under stirring ◦ (300 rpm) for 24 h, at 5 atm and 150 C. When hydrothermal synthesis was complete, the composite was separated by filtering and further washed and dried at 105–120 ◦ C until constant mass. Three composite materials were prepared by mixing sol-gel TiO2 with the correspondent powder (1:1 wt ratio) to form: TiO2 –Cux S, TiO2 -Fly1, TiO2 -Fly2. 2.2. Material characterizations The X-ray diffraction (XRD) measurements were used to analyze the crystalline phase(s), composition and crystallinity. Measurements were carried out at room temperature with a Brucker D8 Discover X-ray diffractometer at 40 kV and 40 mA (equipped with a Cu tube for generating CuK␣ radiation, ␭ = 0.15406 nm), incident beam in the 2-Theta mode over the 20–70◦ range, q = 0.025◦ scan step, and step time of t = 1.5s. The photocatalyst surface morphologies were characterized by (1) Atomic force microscopy (AFM, NT-MDT model BL222RNTE). The AFM images were taken in semicontact mode with Si-tips (NSG10, force constant 0.15 N/m, tip radius 10 nm) and the results were processed with the dedicated software (NT-MDT, Nova Soft), and (2) Scanning electron microscopy (SEM, Hitachi model S-3400N Type II) in BSE mode at low vacuum (50 Pa, 15 kV). Energy dispersive X-ray (EDX) was used to assess the elemental composition (Thermo Scientific).

L. Andronic et al. / Journal of Hazardous Materials 320 (2016) 469–478

UV–vis spectra were acquired using a UV-VIS-NIR Perkin Elmer Lambda 950 spectrometer equipped with an integrating sphere, in the range 250–825 nm. The band gap energy of materials were calculated using the Kubelka-Munk equation. The diffuse reflectance was converted into the absorption coefficient ␣. The overall band gap values were evaluated using the Tauc plots, by ploting (␣h)2 vs. h [28]. The specific photocatalyst surface was examined by nitrogen adsorption-desorption equilibrium at 77 K using the Autosorb IQMP (Quantachrome) analyzer. The point of zero charge of powders (pHpzc ) was estimated by potentiometric titration [29]. In each experiment the powder weight/volume ratio of the testing solution used in photocatalysis was respected, thus pHpzc measurements were run by adding 0.02 g powder to 100 mL of NaOH 0.1 M followed by titration with HNO3 0.1 M. The pH was measured after adding portions of the acidic solution and equilibrium was reached, using a digital pH meter (Schott Instruments, Prolab 2000). The pHpzc was calculated from the pH/VHNO3 experimental curve derivative.

2.3. Photocatalysis experiments The photocatalytic experiments were performed in three different photoreactors: (1) Solar Simulator 1 at Transilvania University of Brasov, which combines UV and Vis radiation by using two F18W/T8 (black light tube, UVA, typically 340–400 nm, with max = 365 nm, Philips, nominal power was 36W) and five TL-D Super 80 18W/865 lamps (white cold light tubes, Vis, typically 400–700 nm, with max = 565 nm, nominal power was 90W), placed annularly around the photoreactor. This combination was chosen because it is close to the solar spectrum (excluding IR), but with a higher UV share (28%). The total irradiance was 25 W/m2 measured with a Class A+ pyranometer (LSI LASTEM, BSR 153). The photocatalytic experiments were run in 500-mL quartz beakers under stirring. 0.2 g/L of photocatalyst were dispersed in a pollutant solution of 20 mg/L which is the optimum concentration described for CPC collectors in the pilot plant [30,31]. The solution was irradiated for 2 h at room temperature (after 15 min in the dark to reach adsorption-desorption equilibrium). During the experiments free air circulation in the photoreactor allowed to run the experiments at 25. . .28 ◦ C. At pre-set time intervals, 5–10 mL of the phenol solution was withdrawn and filtered using Milex GN 0.2 ␮m filter. The pollutant (phenol) concentrations in the aqueous solution were analyzed by UV–vis spectrometry (PerkinElmer Lambda 950) on the calibration curve recorded at the maximum phenol absorption peaks ( = 270 nm). Then the photodegradation efficiency was calculated. (2) Solar Simulator 2: The Suntest XLS+ solar light simulator (Heraeus, Germany) at the Plataforma Solar de Almería, has a 2.2 kW xenon lamp and a correction filter for the solar spectrum in the 290–750 nm range (9% is UV radiation in the 290–400 nm range). The air was permanently cooled to prevent evaporation, thus the experiments were carried out at 35 ◦ C (as the radiation in the solar simulator also includes about 45% IR). The radiation intensity in the UV region was 40 W/m2 measured with a radiant power meter, corresponding to a total radiation intensity of 443 W/m2 . The photocatalytic degradation of each pollutant (starting concentration 20 mg/L) was carried out in a borosilicate vessel containing 500 mL of synthetic wastewater and suspended photocatalyst (0.2 g/L) at unadjusted pH. The solution was irradiated for 2 h (after 15 min in the dark). At pre-set times, 5–10 mL of the pollutant solution was taken and filtered using a Milex GN 0.2 ␮m filter and each pollutant was analyzed to calculate the photodegradation efficiency.

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(3) Solar pilot plant at the Plataforma Solar de Almería (PSA, latitude 37◦ N, longitude 2.4◦ W) with direct solar irradiation in a non-concentrating Compound Parabolic Collector (CPC) radiation system. Global ultraviolet solar radiation (UVG ) was measured using a radiometer (KIPP&ZONEN, model CUV 3) mounted on a platform tilted 37◦ . To standardize the solar irradiation, Malato et al. defined [32] the duration of normalized irradiation considering the average solar UV flux on a sunny day of about 30 WUV /m2 [33]. t30W,n = t30W,n−1 + tn ·

UVG,n Vi · 30 VT

(1)

where t30W,n is the time normalized to 30 W/m2 , t30W,n−1 is the real duration of irradiation, UVG,n is the average global solar ultraviolet radiation measured during tn , Vi is the irradiated volume on the horizontal plane (22 L) and VT is the total volume (35 L) in the reactor. The use of Eq. (1) allows photocatalytic experimental data collected on different days to be compared. The photocatalytic experiments were done using two different types of water matrices: distilled water (DW, conductivity < 10 microS cm−1 , organic carbon < 0.5 mg·L−1 ) and tap water (TW, 300–450 mg/L Na+ , 5–10 mg/L K+ , 30–50 Mg2+ , 80–120 mg/L Ca2+ , − − 250–300 mg/L SO2− 4 , 250–350 mg/L Cl , 800–900 mg/L HCO3 ). The pollutant tested was dissolved in DW or in TW to form solutions with the desired concentration, the catalyst (0.2 g/L concentration) was suspended in 35 L synthetic wastewater and the suspension was recirculated for 30 min in the dark (solar collectors covered) to ensure the adsorption-desorption equilibrium in the reactor. Then, the cover was removed and the suspension was recirculated under solar irradiation to run photodegradation. Samples were taken periodically and photodegradation efficiency was calculated. In all the experiments, the photodegradation efficiency was calculated using Eq. (2):  = [(c0 − c)/c0 ] · 100

(2)

where c0 is the initial pollutant concentration and c is the pollutant concentration at a specific time. For the experiments run in Solar Simulator 2 and in the CPC plant, the concentrations were evaluated for each pollutant, as follows: 2.3.1. Phenol (PHE) analysis The concentrations of phenol or aromatic intermediates were monitored by High Performance Liquid Chromatography with UV detection (HPLC-UV), employing a C-18 column (Supelcosil LC18, 5 ␮m particle size, length 15 cm, i.d. 3 mm) at 30 ◦ C and water/methanol at 65/35 ratio as mobile phase. All samples were diluted by a factor of 2 in acetonitrile to desorb any phenol adsorbed on the catalyst and filtered through a 0.2-␮m syringe-driven filter prior to analysis. The injected volume for all samples was 20 ␮L. 2.3.2. Imidacloprid (IMD) analysis The concentration of imidacloprid was measured by UPLC (Agilent 1200 series with DAD-UV detector) using a linear gradient. The mobile phase used was a mixture of solvents A (acetonitrile) and B (ultrapure water acidified with 25 mmol L−1 formic acid). The flow rate was 1 mL min−1 , and the injection volume was 20 ␮L. The UV signal for imidacloprid was recorded at the maximum absorption wavelength (270 nm). For UPLC analyses, 9 mL aliquots were filtered through a 0.22-␮m syringe filter, then 1 mL of UPLC-grade acetonitrile was also filtered to extract any compound adsorbed. Analytical HPLC-grade organic solvents, acetonitrile, formic acid and water were used.

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Table 1 The composition of the crystalline phases, the elemental surface analysis and the band gap values of the photocatalytic materials. Sample

Crystallinity (%)

Composition of the crystalline phases from XRD

Elemental surface composition (at%)

Eg (eV)

TiO2 SG

77.1

Ti: 24.23 O: 75.77

3.00

Cux S

75.2

Cu: 22.72 S: 8.90 O: 68.38

2.49

Fly1

48

Na: 4.20 Al: 6.31 Si: 8.69 Ti: 6.19 Fe: 1.55 O: 73.06

3.88

Fly2

65.7

TiO2 anatase, tetragonal (75%) TiO2 rutile, tetragonal (20%) TiO2 brookite, tetragonal (5%) Cu1.8 S, syn, cubic (30.44%) Cu1.96 S, tetragonal (55.81%) Cu1.97 S, orthorhombic (5.61%) Chalcocyanite, syn, CuSO4 , orthorhombic (8.13%) TiO2 anatase, tetragonal (73.37%) TiO2 rutile, tetragonal (5.77%) SiO2 (4%) Al2 O3 , orthorhombic (1.62%) Hematite, syn, Fe2 O3 , rhombohedral (4.29%) Mullite, syn, Al2.35 Si0.64 O4.82 , orthorhombic (4.43%) Na8 (AlSiO4 )(OH)2 ·4H2 O, cubic (6.52%) TiO2 anatase, tetragonal (86.85%) TiO2 rutile, tetragonal (6.72%) SiO2 (0.2%) Hematite, syn, Fe2 O3 , rhombohedral (0.11%) Mullite, syn, Al2.35 Si0.64 O4.82 , orthorhombic (1.45%) Na8 (AlSiO4 )(OH)2 4H2 O, cubic (4.67%)

Na: 4.21 Al: 3.61 Si: 6.66 Ti: 13.70 Fe: 1.35 O: 70.47

3.94

2.3.3. Dichloroacetic acid (DCA) analysis DCA and the chloride anions formed during its decomposition were assessed with a Dionex DX-600 ion chromatograph system equipped with an autosampler (Dionex AS40), a quaternary gradient pump (Dionex GP50), a thermostatic column oven (Dionex LC30) and a conductivity detector (Dionex ED50) using a Dionex Ionpac AS11-HC 4 × 250-mm column. All samples were measured by direct injection of samples filtered with 0.2-␮m syringe-driven filters. In addition, mineralization was assessed by total organic carbon (TOC) measurements using a TOC analyzer (Shimadzu TOC-VCSN). The samples were filtered before analysis by 0.20-␮m PVDF filters (Millipore). 3. Results and discussion 3.1. Materials characterization The photocatalytic materials show polycrystalline structures with crystallinity varying from 48% (Fly1) to 75.2% (Cux S) and 77.1% (TiO2 ). The composition of the crystalline phases and overall crystallinity (estimated using the device software EVA 1.4) are presented in Table 1. Diffraction analysis of the TiO2 sol-gel powder shows that the main crystalline phase is anatase (tetragonal structure) (75%), along with rutile (20%) and brookite TiO2 polymorphs. This anatase:rutile ratio is close to Degussa P25 composition (although the amount of amorphous phase is higher) and should support efficient photocatalysis. The average size of the anatase crystallites calculated using the Debye-Scherrer equation is about 17 nm. The annealed Cux S powder formed by photochemical synthesis contains “copper-rich” sulfides (x = 1.8 − 1.97, 91.87%) with tetragonal Cu1.96 S as the predominant crystalline phase (55.81%) along with orthorhombic chalcocyanite. This phase composition might be suitable for a good Vis photocatalyst, even if the stoichiometric chalcocite Cu2 S polymorph was not found in the sample. The band gap of copper rich chalcocite (Cu2 S) is located in the Vis spectral region (Eg = 1.2 to 2.35 eV, [34]), enabling Vis activation. The predominant crystalline phase in the Fly1 and Fly 2 composites is TiO2 anatase, which is higher in the Fly2 sample (18%), as expected from the FA:TiO2 mass ratios used in composite preparation. As the data in Table 1 show, rutile could also be identified, but in a significantly lower percentage than

the original Degussa P25, confirming that hydrothermal synthesis lead to solubilization-reprecipitation supporting anatase development. Typical crystalline FA components were also found: SiO2 , hematite syn. (Fe2 O3 ), nonstoichiometric aluminate mullite (Al2.35 Si0.64 O4.82 ) and sodium aluminum silicate hydroxide hydrate (Na8 (AlSiO4 )(OH)2 ·4H2 O) with a cubic structure. The content of the last two compounds decreased from 10.95% (Fly1) to 6.12% (Fly2), while anatase content increased. This difference could be important in the Fly1 and Fly2 materials and in the Fly ash-TiO2 systems and may influence their photocatalytic properties. Table 1 The SEM images of the photocatalytic powders in Fig. 1 show that the powders have a relatively uniform and homogeneous porous structure. While TiO2 and Cux S have a rather random arrangement of the aggregates, the Fly1 and Fly2 morphology show small (100–400 nm) spherical particles aggregated in regular structures. The surface morphology and complexity of the irregular microstructures was further examined using atomic force microscopy. The 2D AFM images (5 ␮m × 5 ␮m scan area) and the pore size distributions are shown in Fig. 2. The results show that the TiO2 aggregates consist of very small particles (15–20 nm) in the crystallite size range. Their random distribution in aggregates forms a surface with an average roughness of 212 nm. This random aggregation is confirmed by the macropore distribution curve, which exhibits several maxima. The AFM image of the Cux S sample shows compact aggregates formed by small and spherical particles. The average roughness was 95 nm outlining a smoother and more compact macrostructure. The macropore distribution confirms increased surface regularity with most of the pores in the 200–400 nm range. This compact structure is likely to increase water stability of the photocatalyst [35]. The Fly1 and Fly2 samples show similar homogenous structures of round, aggregated particles interconnected by pores with an average size of 200 nm, and average roughness of around 38 nm for Fly1 and 78 nm for Fly2, respectively. The narrow macropore distribution curves confirm the SEM findings and the regular morphology. Photocatalysis is, by definition, a surface process, therefore the surface composition was estimated based on EDX; the results in Table 1 show a large oxygen excess in the TiO2 SG powder as a possible result of trapped water molecules in the surface micro-pores. Considering the application (as suspensions in aqueous media), this trapped water may enhance the wetting properties and support

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Fig. 1. SEM images of TiO2 , Cux S, Fly1 and Fly2.

Fig. 2. AFM images and pore histogram of the materials.

the adsorption and photocatalysis process. A similar large oxygen excess is observed also in the copper sulfide powders; this may also originate from trapped water molecules but can also be the result of partial Cux S oxidation to Cux O. As XRD data did not show any copper oxide peaks, one may conclude that, if existent, the oxide is amorphous. The EDX composition shows a Cu:S ratio close to 2, confirming that Cu+ is the most likely specie, in good agreement with the XRD data. The specific BET surface area, average pore size, total pore volume and mean area diameter of the samples are listed in Table 2.

The specific surface area of the powders was calculated based on the nitrogen adsorption/desorption isotherms. The highest specific surface corresponds to sol-gel TiO2 , confirming the AFM findings of individual crystallites loosely assembled in aggregates. Additionally, the rather large micro-pores (10 nm) can well accommodate adsorbed water molecules, confirming the EDX results. Significantly larger micro-pores are exhibited by the copper-based surface that will eventually allow a lower amount of adsorbed water, thus partial (surface) oxidation of Cux S is likely, according to the EDX data. The fly ash-TiO2 dispersion has very similar specific surfaces,

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Table 2 Textural parameters of the samples. Samples

BET (m2 /g)

Pore size (nm)

Pore volume (cm3 /g)

Micropore volume (cm3 /g)

Smicropores (m2 /g)

Sexternal [m2 /g]

TiO2 Cux S Fly1 Fly2

126.53 10.52 36.14 36.59

10.43 23.27 24.46 37.38

0.3300 0.0612 0.2210 0.3419

0.002 0.001 0.003 0.003

10.62 2.01 4.86 5.87

115.91 8.52 31.28 30.72

confirming that hydrothermal synthesis mostly affected the FA backbone. Table 2 The first step in efficient photocatalysis is pollutant adsorption on the catalyst surface. This is well supported by suitably charged surfaces and fractured morphologies. The photocatalysts analyzed in this paper are amphoteric, thus their surface is positively charged at pH below their point of zero charge (pzc) and at values over pHpzc , the surface charge is negative [36]. TiO2 and SiO2 (generically described as ≡ AOH) are considered two-protonic acids having the following dissociation steps [37]: K

a1 ≡ AOH2+ + H2 O ↔ ≡ AOH + H3 O+

Ka2

−

≡ AOH + H2 O ↔ ≡ AO + H3 O

+

(3) (4)

The copper sulfide surface tends to adsorb water or hydrolyze reducing the surface energy. The dissociation steps proposed for Cux S could be therefore described as follows: K

a1 ≡ Cux S + H + ↔ ≡ Cux SH +

Ka2

+

≡ Cux S + H2 O ↔ ≡ Cux SH + HO

(5) −

(6)

The pH values vs. the cumulative acid volume added in acid-base titration and their derivatives as plotted in Fig. 3 were used to calculate the dissociation constants (Ka1 and Ka2 ) for the equilibrium

reactions described by Eqs. (3) and (4), and the corresponding pKa1 and pKa2 . Then the pH corresponding to the point of zero charge (pHpzc ) was calculated according to Eq. (7): pHPZC =

pKa1 + pKa2 2

(7)

Fig. 3 The pHpzc was calculated as 6.15 for TiO2 , 6.32 for Cux S, 6.69 for Fly1 and 6.28 for Fly2, respectively. These points of zero charge close to neutral show that very low variations in acidity or alkalinity can actually charge the surface and promote electrostatic attraction with anions or compounds containing ␲ electrons (at pH < pHpzc ) or with cations at working pH above pHpzc . These results also show that despite the significant morphological differences among the samples, surface polarizability is similar. Additionally, these close pHpzc values support the in situ formation of the TiO2 /CuxS, TiO2 /Fly1 and TiO2 /Fly2 composites, tagreting extended Vis activity. Besides the affinity for the pollutants, the band gap values are important, as being directly linked to the activation energy type. The values in Table 1 show slightly lower band gap values of the TiO2 SG powders, as compared to the literature typical data for the titania polymorphs. However, these values are the result of the mixed polymorph phases, [38]. Fly ash materials involve Vis

Fig. 3. pH changes in the photocatalyst suspension (a. TiO2 , b. Cux S, c. Fly1, d. Fly2) vs. total HNO3 0.1 M volume (full squares) and the derivative curve (dpH/dV, open circles).

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Kinetic calculations were developed and the Langmuir Hinshelwood (L-H) model was tested using the equation proposed by Turchi and Ollis [40]. Although the L–H isotherm was rather useful in modelling the process, it is generally agreed that both rate constants and orders are only “apparent” [41–43]. They are good for describing the degradation rate, and may be used for reactor optimization, but they have no physical meaning, and may not be used to identify surface processes. Thus, while not a useful tool for describing the active species involved in oxidation, engineers and solar designers seem to have a common understanding of the usefulness of the unmodified L–H model: r =−

Fig. 4. Photodegradation efficiency of phenol in two different solar simulators (Experimental conditions: Vs=500 mL, Pollutant concentration=20 mg/L, Irradiation duration=2 h, Catalyst=0.2 g/L).

active components (as Fe2 O3 ), inactive insulators (Al2.35 Si0.64 O4.82 , Na8 (AlSiO4 )(OH)2 ·4H2 O, SiO2 , Al2 O3 ) but also TiO2 . The overall band gap energy values of Fly 1 and Fly 2 composite indicate UV activation as the wide band gap components and the insulators are prevailing (as the XRD data show); however, the Vis-active components can form heterostructures and induce VIS activation even for the Fly 1 and Fly 2 composites. When adding the Cux S powder (Eg = 2.49 eV) in the system, the dispersed mixtures/composites are even more likely to get Vis-active.

3.2. Kinetic parameters of phenol photodegradation at different simulated solar irradiation The first photocatalytic experiments used phenol as the target pollutant, as it is recognized for its stability and is proposed by the IUPAC as a reference in photocatalysis [39]. Additionally, the EU directive 2013/39/EU extended the list of priority substances along with new/updated environmental quality standards to be met by 2021 (with updated values, e.g. for phenols). The comparative experiments were done at two intensities of simulated solar radiation: 25 W/m2 (in Solar Simulator 1) and 443 W/m2 (in Solar Simulator 2). The UV share in these irradiances is 7 W/m2 in Solar Simulator 1) and 40 W/m2 in Solar simulator 2, respectively. The results are presented in Fig. 4. The results show the high phenol removal efficiencies with sol-gel TiO2 , as expected from its large specific surface and anatase:rutile ratio. Results are also promising for the fly ash dispersions, especially for Fly2.

kr · KS · C dC = 1 + KS · C dt

(8)

where r is the photocatalytic degradation rate (mmol L−1 min−1 ), C is the phenol concentration (mmol L−1 ), kr is the apparent reaction rate constant (mmol L−1 min−1 ) and Ks is the apparent adsorption constant (L mmol−1 ). Kinetic parameters during the first stages of the process can be calculated using the initial reaction rate (r0 ) by plotting 1/r0 vs. 1/C0 , and plot intercept and slope 1/kr and 1/kr KS , respectively. The kinetic constants are shown in Table 3. The results show that L-H supports good fit to the experimental data for all materials. Regardless of the radiation intensity, the apparent adsorption constant does not change, thus the mechanism is not fundamentally affected by the irradiance. However, the apparent rate constants are affected differently: the highly heterogeneous Cux S, Fly1 and Fly2 show a stronger increase in the rate constants, which triple when the total radiation intensity is increased 17.7 times (and UV intensity 5.7 times). This could be because of parallel adsorption/photodegradation reactions with rate constants affected differently by the irradiance. On the contrary, the FA-TiO2 dispersion are less reactive to the radiation intensity as a possible result of faster electron-hole recombination promoted on the inactive SiO2 . 3.3. Phenol, imidacloprid and dichloroacetic acid photodegradation on TiO2 photocatalysts under simulated solar irradiation As the phenol removal efficiencies under higher simulated solar radiation intensity (443 W/m2 ) were significantly better, further experiments were done in the Solar Simulator 2 photoreactor involving two additional pollutants highly soluble in water, imidacloprid (IMD, C9 H10 ClN2 O5 , an N-containing insecticide) and dichloroacetic acid (DCA, an aliphatic organochloride compound), and the results were compared with those for phenol (PHE). In 2015 the EU watch list (EC Decision 495/2015) was extended with 10 new substances/groups of substances, including Neonicotinoids pesticides (as imidacloprid), thus IMD is of real interest as emergent organic contaminant, while DCA represents a stable chlorine-containing pollutant.

Table 3 Photocatalytic parameters of phenol degradation using the Langmuir–Hinshelwood equation at different radiation intensities. Materials

Fly2

TiO2 -Cux S

TiO2 -Fly1

TiO2 -Fly2

Simulator 1—total radiation intensity 25 W/m2 ; UV radiation intensity: 7 W/m2 kr ·105 (mmol L−1 min−1 ) 31.6 8.03 9.83 4.72 4.71 4.69 Ks (mmol−1 L) 2 0.9932 0.9551 0.9822 R

TiO2

Cux S

Fly1

8.63 4.70 0.979

9.90 4.71 0.9827

13.3 4.70 0.9850

16.3 4.71 0.9745

Simulator 2—total radiation intensity 443 W/m2 ; UV radiation intensity: 40 W/m2 kr ·105 (mmol L−1 min−1 ) 66.9 29.7 32.2 4.73 4.62 4.35 Ks (mmol−1 L) 0.9837 0.9323 0.9804 R2

24.1 4.38 0.9624

9.22 4.95 0.9580

20.1 5.09 0.9850

21.5 5.06 0.9718

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3.4. Phenol, imidacloprid and dichloroacetic acid photodegradation on TiO2 catalyst under solar irradiation in the solar pilot plant The most active photocatalyst under simulated solar radiation, TiO2 , was selected for outdoor experiments in the CPC reactor pilot facility. TiO2 -Fly2 was also tested in the CPC under solar irradiation, as it showed good photoactivation, is significantly more economical

Fig. 5. Pollutant photodegradation efficiency in Solar Simulator 2 (Experimental conditions: Vs=500 mL, Pollutant concentration=20 mg/L, Irradiation time=2 h Catalyst=0.2 g/L, catalyst ratio=1:1 wt.).

It was intended to test and compare the activities of the catalysts in the photocatalytic degradation of an ensemble of various pollutants having different structures to enable: (i) observation of possible changes in catalyst activity patterns induced by changing the nature of the pollutant, and (ii) the detection of possible variation in classification of their activity. This strategy differs from that of most studies in which catalyst comparison is mainly based on the degradation reaction of a single pollutant type. It is well known that in photocatalysis, the nature of the substrate affects the behavior of the catalyst [44]. Both dichloroacetic acid and phenol have often been used in photocatalysis experiments. Imidacloprid has both pyridyl and imidazolidine rings, which makes it particularly interesting for the study of special behavior of N-containing compounds during photocatalytic treatment, and its photocatalytic degradation is already well known [45]. A preliminary run was made to find out the pollutant adsorption on the catalysts in the dark. After 2 h of adsorption under identical experimental conditions, no distinguishable changes in pollutant concentration were observed when catalysts were Cux S, Fly1 or Fly2. The adsorption tests on titanium dioxide showed 5–7% removal for all three pollutants. The photodegradation efficiencies for the three pollutants with different photocatalysts are shown in Fig. 5. The large-surface area TiO2 showed the best photocatalytic performance, especially for the degradation of DCA. However, it is interesting to note that the FA-TiO2 dispersion showed promising efficiencies for IMD removal while Vis-active Cux S efficiency was lower, confirming the importance of a large photocatalytic-specific surface. As expected from their composition, the Fly 1 and Fly2 substrates had the lowest response due to the rather large number of components without any photocatalytic activity. The pH of the pollutant solution significantly affected the interaction between the catalyst surface and the pollutant molecules. The starting pH of the solutions was 5.9 for PHE, 6.2 for IMD and 5.75 for DCA, respectively, and remained unchanged. According to the catalyst, pHpzc , at working pH of 5.9 and 5.75, the photocatalyst surfaces are positively charged, thus promoting at least physisorption involving the ␲ electrons in the PHE aromatic ring or DCA anion. Imidacloprid degradation at pH 6.2 on the TiO2 substrate was less favored (thus slower when compared to the other pollutants) due to the unfavorable interaction between the slightly negatively charged TiO2 surface and the imidacloprid molecules.

Fig. 6. Photodegradation and mineralization on TiO2 under solar irradiation in the CPC pilot plant of: (a) PHE; (b) IMD; (c) DCA.

L. Andronic et al. / Journal of Hazardous Materials 320 (2016) 469–478

Fig. 7. Photodegradation efficiency of DCA, IMD and PHE solutions in distilled water (DW) and tap water (TW) on TiO2 and TiO2 -Fly2 in the CPC.

and is easier to filter than TiO2 . The results in Fig. 6 show the evolution of the pollutant concentration (PHE, IMD and DCA) in distilled water over normalized irradiation time calculated with Eq. (2). Phenol is uncharged in acidic and neutral media (pKa = 9.99) and weakly adsorbs onto hydrophilic oxides surface [46]. At higher pH the phenol dissociates forming the phenolate (C6 H5 -O− ) anion and its adsorption on the negatively charged surface of the photocatalyst is nearly impossible. However, as the experimental pH was 5.9, the photocatalytic substrates are positively charged, thus as Fig. 6a shows, phenol is removed with mineralization in the first 40 min (with about 27% efficiency). Later, the hydroquinone and catechol by-products slowly accumulate and clog the surface, reducing phenol ability to (weakly) adsorb, and the mineralization rate slowly decreases for the next 100 min. After 150 min of solar irradiation the difference between the apparent phenol removal (55%) and mineralization (33%) indicates the presence of other photo-products at the end of the reaction. Similarly, the apparent removal and mineralization of imidacloprid (pKa=11.12 [47]) compared in Fig. 6b shows the clogging effect of by-products after 90 min of irradiation. However, the insecticide is easier to degrade than PHE (as a possible result of its structure with more possible breaking points in the molecule) and mineralization efficiency is 80% after 150 min. The fastest apparent removal efficiency was recorded for DCA (pKa=1.246) as a result of the strong affinity of its anion for the positively charged TiO2 surface (at pH below 6.153). However, mineralization is slower than overall removal after 60 min of irradiation (Fig. 6c) and confirms the adsorption of DCA organic intermediate oxidation products on the substrate (the proposed mechanism is included in the Supplementary material). When the synthetic wastewater was prepared using tap water (Fig. 7), slightly higher overall removal efficiencies were recorded for all three pollutants. One reason could be that the anions in natural water have more affinity for holes (h+ ) thus preventing the electron-hole recombination and favoring pollutant degradation. PHE in the tap water solution is degraded with significantly lower efficiency with the TiO2 -Fly2 dispersed catalyst than with sol-gel TiO2 , but the final removal is slightly higher (32%) than when the PHE solution is prepared using distilled water (Fig. 6). This confirms that the tap water composition supports increased efficiency, regardless the photocatalyst or the pollutant. 4. Conclusions Titanium dioxide is recognized as the most efficient photocatalyst for wastewater treatment. However, up-scalable photocataly-

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sis requires Vis-active, low-cost, filterable photocatalytic materials. Therefore, three types of materials were compared for phenol, dichloroacetic acid and imidacloprid removal: sol-gel TiO2 (as the reference), Vis-active Cux S prepared by photochemical precipitation and highly filterable TiO2 -fly ash mixture/composites. The results show that photocatalytic phenol removal efficiency is influenced by (simulated) solar radiation intensity. The reference TiO2 showed the best results, but TiO2 -fly ash dispersions (despite their much lower specific surface) may be considered promising candidates. Under solar radiation in the CPC pilot plant, all three pollutants are partially mineralized during the first 40–90 min, but by-products clog the surface and removal continues without fully oxidizing the organics. The use of tap water for preparing the pollutant solutions lead to slightly higher removal efficiencies than the synthetic wastewater prepared using double deionized water, confirming the positive effect of anions in trapping holes, thus reducing recombination. Therefore, after 150 min of solar irradiation, phenol removal efficiency reached 66% while DCA and IMD removal efficiencies were over 95%. Acknowledgment Funding under the EU 7th Framework Programme Access to Research Infrastructures Activity (SFERA 2 Grant Agreement no. 312643) is gratefully acknowledged, and the use of the facilities and researchers/technology experts; the Romanian Agency for Research (UEFISCDI, ANCS) is acknowledged for funding research on composite photocatalysts under PNII-PCCA, Sim-Photo-Ad, Grant Contract no. 217/2014. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2016.08. 013. References [1] V. Geissen, H. Mol, E. Klumpp, G. Umlauf, M. Nadal, M. van der Ploeg, S. van de Zee, C.J. Ritsema, Emerging pollutants in the environment: a challenge for water resource management, Int. Soil Water Conserv. Res. 3 (2015) 57–65. [2] A.R. Ribeiro, O.C. Nunes, M.F.R. Pereira, A.M.T. Silva, An overview on the advanced oxidation processes applied for the treatment of water pollutants defined in the recently launched Directive 2013/39/EU, Environ. Int. 75 (2015) 33–51. [3] C. Prasse, D. Stalter, U. Schulte-Oehlmann, J. Oehlmann, T.A. Ternes, Spoilt for choice: a critical review on the chemical and biological assessment of current wastewater treatment technologies, Water Res. 87 (2015) 237–270. [4] A. Mills, C. O’Rourke, K. Moore, Powder semiconductor photocatalysis in aqueous solution: An overview of kinetics-based reaction mechanisms, J. Photochem. Photobiol. A 310 (2015) 66–105. [5] M. Curti, D.W. Bahnemann, C.B. Mendive, Mechanisms in heterogeneous photocatalysis: titania under UV and visible light illumination, Reference Module in Materials Science and Materials Engineering (2016). [6] M.E. Simonsen (Ed.) Chemistry of Advanced Environmental Purification Processes of Water, Chapter 4 − Heterogeneous Photocatalysis, 2014, pp. 135-170. ˜ [7] D. Spasiano, R. Marotta, S. Malato, P. Fernandez-Ibanez, I. Di Somma, Solar photocatalysis: Materials, reactors, some commercial, and pre-industrialized applications. A comprehensive approach, Appl. Catal. B 170–171 (2015) 90–123. [8] A. Fujishima, K. Honda, Electrochemical photolysis of water at a semiconductor electrode, Nature 238 (1972) 37–38. [9] Y. Lan, Y. Lu, Z. Ren, Mini review on photocatalysis of titanium dioxide nanoparticles and their solar applications, Nano Energy 2 (2013) 1031–1045. [10] Y. Taga, Titanium oxide based visible light photocatalysts: Materials design and applications, Thin Solid Films 517 (2009) 3167–3172. [11] M. Kitano, M. Matsuoka, M. Ueshima, M. Anpo, Recent developments in titanium oxide-based photocatalysts, Appl. Catal. A 325 (2007) 1–14. [12] H. Zangeneh, A.A.L. Zinatizadeh, M. Habibi, M. Akia, M. Hasnain Isa, Photocatalytic oxidation of organic dyes and pollutants in wastewater using

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