activated carbon coupling to remove diethyl phthalate from water: Influence of activated carbon textural and chemical properties

Desalination 276 (2011) 359–365

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Use of ozone/activated carbon coupling to remove diethyl phthalate from water: Influence of activated carbon textural and chemical properties Tatianne Ferreira de Oliveira, Olivier Chedeville ⁎, Henri Fauduet, Benoît Cagnon ICOA - Institut de Chimie Organique et Analytique, CNRS-UMR 6005, Institut Universitaire de Technologie, Université d'Orléans, 16 Rue d'Issoudun, BP 16729, 45067 Orléans cedex 02, France

a r t i c l e

i n f o

Article history: Received 27 January 2011 Received in revised form 10 March 2011 Accepted 29 March 2011 Available online 17 April 2011 Keywords: Diethyl phthalate Ozone Activated carbon Kinetics Textural properties Surface groups

a b s t r a c t The presence of phthalates in the environment and especially in surface waters and sediments is a major environmental concern. The aim of this work was to study diethyl phthalate (DEP) removal by a water treatment process based on the coupling of ozone (O3) and activated carbon (AC). The main objective was to study the influence of AC properties on the process efficiency and on the coupling mechanism (nature and location of reactions). DEP degradation kinetics by O3/AC coupling was studied by using four commercial ACs whose chemical and textural properties had been previously determined (Boehm titration, N2 adsorption isotherm at 77 K, pHPZC determination). Degradation kinetics was correctly modelled by a pseudo-first order kinetic model based on the sum of all the effects occurring during the treatment (r2 N 0.987). Results show that degradation efficiency depends both on textural properties (microporous and external surfaces favour this treatment) and chemical functions (both acid and basic functions favour radical hydroxyl generation). Experiments performed with a radical scavenger show that in all the experimental conditions used, DEP is mainly degraded by radical reactions. Moreover, it is demonstrated that AC acts more as a radical initiator and promoter and a reaction support than as an adsorbent material. The influence of pH on the reaction efficiency and mechanism is also proved: in acidic conditions (pH b 5) radical reactions are due to O3/AC interactions, and they are due to indirect ozonation in the bulk liquid for higher pH. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Phthalate esters are intensively used in industry as additives particularly in plastics to improve the flexibility of materials and also in pharmaceuticals, lubricants, cosmetics or printing inks [1,2]. The global annual production of phthalates is about 3 million tonnes [3]. Consequently, these compounds are found in the environment, especially in surface waters and sediments [2–5]. For example, concerning watercourses in the Loire Bretagne basin (Centre region, France), about 30% of cases of chemical non-conformity are due to too high concentration of diethylhexyl phthalate (DEHP). According to Roslev et al. [2], these pollutants mainly come from industrial and municipal wastewater treatment plants, showing that existing treatment methods are not suitable to remove these compounds. Some studies show that certain phthalates (especially those having long ester hydrocarbon chain) are refractory to biological degradation and can even present toxicity for the microorganisms performing the biological treatment [1]. Their presence in the environment is a major environmental hazard, since some phthalates are suspected of being carcinogenic and/or endocrine disruptors [6–8]. Several institutions (US Environmental Protection Agency, European Union) have listed

⁎ Corresponding author. Tel.: + 33 2 38 41 72 64; fax: + 33 2 38 49 44 25. E-mail address: [email protected] (O. Chedeville). 0011-9164/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.desal.2011.03.084

phthalates as priority pollutants. Since a substitution product of these phthalates is not yet easily available, it is necessary to study wastewater treatment processes which can ensure the removal of these compounds. Ozone (O3)/Activated Carbon (AC) coupling is a recent wastewater treatment that combines different actions: direct or indirect ozonation in the bulk liquid, adsorption on AC, and direct or indirect oxidation of compounds on the AC surface [9–11]. Moreover, interaction between O3 and AC surface groups can lead to hydroxyl radical generation [12,13]. These radicals are highly oxidizing and can degrade a wide range of organic pollutants in water. Initial studies show the great potential of this treatment which can remove pollutants that are refractory to classical methods [14–16]. In a previous study, O3/AC coupling was proven to be an efficient method to remove DEP [17]. In this coupling, AC can act as an adsorbent material, a reaction support or a radical initiator or promoter. It is therefore very important to choose the appropriate AC to obtain an efficient removal of pollutants. Moreover, to control and optimize this treatment, it is necessary to know the role of AC and the nature of the reactions (molecular or radical). The aim of this work was to study the influence of the textural and chemical properties of the AC on O3/AC coupling efficiency. The degradation kinetics of diethyl phthalate (DEP), was studied with four commercial ACs whose chemical and textural properties had been previously determined. DEP was chosen as model pollutant because

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its solubility in water (1080 mg L−1 at 20 °C) permits its quantification with good accuracy and its removal by O3/AC coupling has not ever been studied [18]. A simple kinetic model, based on the sum of effects occurring during the treatment process, was used to describe the DEP degradation kinetics. To determine the nature of the reaction, the kinetic contribution of radical reactions to DEP degradation was estimated by performing experiments with tert-butyl alcohol (tBuOH) as radical scavenger. Moreover, experiments were carried out at five pH values (ranging from 2.5 to 7.2) in order to study the influence of this parameter on the O3/AC interaction. 2. Materials and method 2.1. Experimental 2.1.1. Ozonation pilot DEP degradation kinetics were performed in a 1 L batch reactor (Fig. 1) thermostated by a cryothermostat (C) and mechanically stirred (M). A dropping funnel (DF) was used to introduce solution into the reactor. The stirring mobile (impeller) and its dimensions (7 cm in diameter) were chosen to favour mass transfer between the three phases and to limit AC attrition. Ozone, produced from pure oxygen in a BMT 803N ozone generator (O3 G) supplied by the BMT company, was introduced through a porous diffuser at the bottom of the reactor (D). Ozone concentrations in the inlet and the outlet gas were measured with a BMT 964 ozone analyser (O3 A). The unconverted ozone in the outlet gas was removed by an ozone catalytic destructor (CD). The gas circuit was assembled with materials inert with respect to ozone (pipes in teflon; valves, flowmeters and non-return valves in PFA). 2.1.2. Kinetic experiments The reactor was filled with 500 mL of phosphate buffered solution, prepared with a mixture of different amounts of H3PO4 (purity up to 85%, obtained from Sigma Aldrich), KH2PO4 (purity up to 99%, obtained from Fluka) and Na2HPO4 (purity up to 99%, obtained from Fluka) at pH 2.5, 3.5, 5.6, 6.2 or 7.2 and 2 g of AC. The pH value was measured with a pH-meter at the beginning and at the end of the experiments (no pH modification was observed for all the experiments). The reactor was then thermostated at 20 °C. The dropping funnel was filled with a solution containing 0.05 g of DEP (purity up to 99.5%, obtained from Sigma Aldrich) dissolved in 250 mL of the same buffered solution. The initial DEP concentration was 0.2 g L−1. For experiments performed with tBuOH (purity up to 99.5%, purchased from Across Organics), 0.5 g of this radical scavenger was introduced into the dropping funnel in buffered solution ([tBuOH] = 30[DEP]). O3 was continuously fed into the reactor to achieve saturation of the solution, monitored by the carmin indigo method proposed by Bader and Hoigné [19]. It was assumed that saturation was obtained when the dissolved ozone concentration remained at the same equilibrium concentration during 10 min. The gas flow rate was 40 NL h−1 and O3 FI O3 A

M

D F

C D

FI PI O3 G

Fig. 1. Ozonation pilot.

2.2. Determination of AC properties The ACs used in this study (L27, X17, F22 and S21) were commercial activated carbons provided by Pica. They were washed before each experiment to eliminate any residual acidity due to their activation treatment. The characterization of the porosity was managed by conventional nitrogen adsorption–desorption isotherms at − 196 °C (using a Micromeritics ASAP 2020) on samples of approximately 0.2 g after outgassing at 250 °C for 48 h and under a residual vacuum of less than 10−4 Pa. The adsorption isotherms were carried out with relative pressures ranging from 7.9 × 10−4 Pa to 0.99 Pa and were analyzed by using the Dubinin–Radushkevich equation [20,21]. This leads essentially to the micropore volume W0 (cm3 g−1) and the characteristic energy Eo, a quantity related to the average micropore width L0 (nm) [20]. The Sing αS plot (using black carbon Vulcan 3 as a reference) was used to obtain values of the specific external surface Sext (m2 g−1) [22], assuming that for slitshaped micropores the specific microporous surface Smicro (m2 g−1) could be estimated using the specific microporous volume and the mean pore size [20]. Boehm titration was used to determine the oxygen surface groups [23]. 0.2 g of AC dried overnight at 80 °C was introduced into 25 mL of the following 0.1 mol L−1 solutions: NaOH, Na2CO3, NaHCO3 and HCl. The AC/solution mix was placed in a thermostated multi-agitation apparatus at 25 °C under mechanical stirring at 150 rpm for 48 h. The suspension was filtered through a 0.45 μm membrane filter and the excess of base or acid was titrated with 0.1 mol L−1 solutions of HCl or NaOH respectively. Following the Boehm method, the carboxylic, phenolic, and lactonic groups were quantified. The number of surface basic groups was calculated (but not identified) from the amount of HCl which reacts with the carbon. The determination of the pHPZC was obtained by the method proposed by Rivera-Utrilla and Sanchez-Polo [24]. The mesopores and macropores size distributions were determined by mercury porosimetry using a Micromeritics Autopore IV 9500 by the Laboratoire des Composites Thermostructuraux (LCTS, Bordeaux, France). 2.3. Kinetic model

CD

O2

concentration in the inlet gas was fixed at 50 g Nm−3. The dissolved ozone concentration, determined by the carmin indigo method, was 0.23 mmol L−1.When ozone saturation was obtained, the kinetic experiment begins with the introduction of the DEP solution into the reactor, always under O3 flow. Samples were withdrawn with a syringe at suitable time intervals and were filtered through a 0.45 μm membrane filter. A volume of 2 mL of filtered sample was immediately introduced into 100 μL of sulphite solution (0.1 M) to remove dissolved ozone. The concentration of dissolved ozone in the reactor was monitored by the carmin indigo method to ensure that saturation was maintained throughout the experiment. The DEP concentration in samples was monitored by HPLC (Kontron 325 system) using a Hypersil C18 column (250 mm long × 4.6 mm i.d, Thermo Scientific) with a UV detector (Spectra Physics 200) at 228 nm. Mobile phase was an acetonitrile/water mixture (70:30, v:v) introduced at a flowrate of 1 mL min−1. Some experiments were performed three times (not shown here) to determine standard deviation on kinetic constant values (this standard deviation was estimated at 0.001 min−1).

DEP degradation by O3/AC coupling results from five different effects [9–11]. Reactions can occur in the bulk liquid (direct or indirect ozonation), or on the AC surface (adsorption and direct or indirect oxidation of pollutants). Thus, DEP degradation kinetics by this treatment process cannot be easily modelled. In this study, a simplified model, based on the sum of the five effects previously described, was used in order to estimate the kinetic contribution of these effects to DEP degradation. It was assumed that no O3 diffusion limitation occurs (chemical regime, Ha b 3) [25] and that adsorption

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can be modelled with a pseudo-first order model. It was also assumed that the O3 concentration remained constant throughout the kinetic experiments (this assumption was verified by measuring the dissolved O3 concentration during the experiments). According to Von Gunten [26], it can be assumed that hydroxyl radical concentration was also maintained constant during the experiments. For ozonation reactions (direct or indirect), it is generally admitted that the kinetics is correctly described by a second-order model [27]. According to the assumptions of this model, DEP degradation kinetics by O3/AC coupling was modelled as follows: −

 • d½DEP  = k1 ½DEP ½O3  + k2 ½DEP  HO + k3 ½DEP  dt  • + k4 ½DEP ½O3  + k5 ½DEP  HO

ð1Þ

Table 1 Chemical properties of activated carbons used.

L27 S21 X17 F22

d½DEP  = kglobal ½DEP  dt

ð2Þ

with: kglobal = khomo + khetero

ð3Þ

 • khomo = k1 ½O3  + k2 HO

ð4Þ

 • khetero = k3 + k4 ½O3  + k5 HO

ð5Þ

khomo represents the kinetic constant of reactions (direct or indirect) occurring in the bulk liquid in absence of AC (s−1). khetero represents the reactions occurring on AC surface (k5 includes radical reactions occurring in the bulk liquid but these reactions are initiated and promoted by heterogeneous reactions between O3 and AC surface groups) (s−1). For experiments performed with both O3, AC and tBuOH, radical reactions are scavenged and, in accordance with Eq. 1, DEP degradation kinetics can be modelled as in Eq. 6: −

d½DEP  = kobs ½DEP  dt

ð6Þ

with: kobs = k1 ½O3  + k3 + k4 ½O3 

ð7Þ

Determination of kglobal and kobs from Eqs. 2 and 6 makes it possible to estimate the kinetic contribution of radical reactions to DEP degradation (δHO•): HO•

δ

ð%Þ =

kglobal −kobs :100 kglobal

ð8Þ

For experiments performed only with O3, the DEP degradation kinetics can be described by a pseudo-first-order model: −

d½DEP  = khomo ½DEP  dt

ð9Þ

Phenolic groups (meq g−1)

Lactone groups (meq g−1)

Total acid groups (meq g−1)

Total basic groups (meq g−1)

pHPZC

0.81 0.05 0.15 0.13

0.30 0.30 0 0.14

0.46 0.03 0.02 0.05

1.57 0.35 0.20 0.32

0.18 0.33 0.85 0.26

3.0 7.4 8.2 7.5

δ

=

kglobal −khomo :100 kglobal

ð10Þ

3. Results and discussion 3.1. Chemical and textural properties of AC Results, presented in Table 1, show that these ACs have very different chemical properties. The pHpzc varies from 3.0 to 8.2, total acid functions vary from 0.20 to 1.57 meq g−1 and basic functions vary from 0.18 to 0.85 meq g−1. There is one acid AC (L27), two neutral (F22 and S21) and one basic (X17). The results obtained by the Boehm method are in agreement with pHpzc values. The ACs also have different porous properties. The N2 adsorption–desorption isotherms for the four ACs used (Fig. 2) show that L27, F22 and X17 ACs were microporous but also contained some mesopores as indicated by the shape and the hysteresis loop of the N2 isotherms. Fig. 2 also shows that S21 is microporous only, which is confirmed by the values of its Smicro and Sext. The values of the porous properties are reported in Table 2: the mean pore size (L0) varies from 9.7 Å to 18.5 Å, the specific external surface (Sext) varies from 18 to 444 m2 g−1 and the specific microporous volume (W0) varies from 0.29 to 0.57 cm3 g−1. The mercury porosimetry results also show that these ACs are different in terms of meso- and macroporosity. Fig. 3 shows that L27 has more mesopores and macropores than F22 and X17. The mesopores and macropores size distribution of S21 confirms that this material is microporous only. 3.2. DEP degradation kinetics 3.2.1. Single ozonation The DEP ozonation kinetic constants, presented in Table 3, were obtained and discussed in a previous study [17]. It was proven that the

Adsorbed Volume (cm3 STP g-1)

−

Carboxylic groups (meq g−1)

Having determined kglobal and khomo from Eqs. 2 and 9, the kinetic contribution of heterogeneous reactions to DEP degradation (δhetero) can be estimated: hetero

where [DEP] is the concentration of DEP (mol L−1), [O3] is the concentration of dissolved ozone (mol L−1), [HO•] is the concentration of hydroxyl radicals (mol L−1), and k1 and k2 are the kinetic constants of respectively direct and indirect ozonation of DEP in the bulk liquid (L mol−1 s−1), k3 is the kinetic constant of DEP adsorption (s−1), k4 is the kinetic constants of direct oxidation of DEP on the AC surface (L mol−1 s−1), k5 is the kinetic constant of indirect ozonation of DEP due to radical reactions initiated and promoted by interaction between O3 and AC surface groups and t is the time (s). Assuming that concentration of both O3 and HO• remains constant during experiments, Eq. 1 can be simplified as follows:

361

700 600 500 400 300 200 100 0 0.0

0.2

0.4

0.6

0.8

1.0

p/p0 Fig. 2. N2 adsorption–desorption isotherms at − 196 °C of the four activated carbons used (L27 (■), S21 (●), X17 (Δ) and F22 (○)).

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Table 2 Textural properties of ACs.

L27 S21 X17 F22

Table 3 Degradation kinetics of DEP by single ozonation at different pH values.

Wo (cm3 g−1)

Lo (Å)

Sext (m2 g−1)

Smicro (m2 g−1)

Stotal (m2 g−1)

pH

khomo (min−1)

r2

0.57 0.47 0.29 0.39

18.5 9.7 15.1 12.7

444 18 130 256

616 969 384 614

1060 987 514 870

2.5 3.5 5.5 6.2 7.2

0.004 0.009 0.201 0.476 0.613

0.987 0.992 0.990 0.994 0.993

kinetic model correctly describes DEP degradation kinetics (r2 N 0.987) and that DEP can be degraded by single ozonation. Nevertheless, this degradation kinetics strongly depends on pH value and appears to be very slow in acidic conditions (Fig. 4). 3.2.2. O3/AC coupling DEP degradation was carried out with the four commercial ACs used and at the same five pH values. The results presented on Fig. 4 show that AC enhances the DEP degradation kinetics: for each AC used and at each pH value, DEP removal was faster with O3/AC coupling than with single ozonation. For example, at pH = 2.5, complete DEP removal was obtained after 6 min of treatment with ACs X17 and L27 while DEP removal was only about 20% after 100 min of treatment with single ozonation. Nevertheless, results varied markedly depending on the AC used. It also appeared that O3/AC coupling efficiency strongly depends on the AC properties. For example, at pH = 2.5, DEP removal was achieved only after 100 min of treatment with ACS21. Moreover, results show that the influence of AC on DEP degradation kinetics becomes weaker when the pH increases. At pH 7.2, the results obtained with O3/AC coupling are only slightly better than those obtained with single ozonation. It can be assumed that no influence of the AC presence on DEP degradation kinetics would be observed at pH values up to 7.2. The modelling of these results is presented on Table 4. DEP degradation kinetics was correctly described by the simple pseudofirst model used (r 2 N 0.987). These results confirm that DEP degradation kinetics depends on the AC properties. For example, at pH = 2.5, kglobal varied between 0.489 min−1 (X17) and 0.027 min−1 (S21). This phenomenon could be explained by the influence of AC chemical and textural properties on O3/AC coupling. Results show that large external and microporous surfaces both favour DEP removal. O3/ AC coupling is a gas/liquid/solid process and its efficiency is linked to mass transfer. In AC, adsorption is located in micropores, but mesopores and macropores favour intraparticle diffusion. This could explain why ACS21 presents the worst results: this AC is microporous only (Sext = 18 m2 g−1 and Smicro = 969 m2 g−1) and diffusion of O3

dV/dLog d

0.6

MESOPORES

MACROPORES

0.4

0.2

0.0 0.01

0.1

1

10

d (µm) Fig. 3. Mesopores and macropores size distributions (Hg porosimetry) of the four activated carbons used (L27 (■), S21 (●), X17 (Δ) and F22 (○)).

and DEP into AC was not favoured. Thus, a large microporous surface is not sufficient to enhance O3/AC interaction. AC L27, in contrast, which presents large external (Sext = 444 m2 g−1) and microporous surfaces (Smicro = 616 m2 g−1), is the most efficient AC in removing DEP by O3/AC coupling. L27 contains mesopores and macropores which are very important in permitting access to the microporosity. However, the efficiency of this treatment is not only linked to textural properties: ACF22 presents more suitable textural properties (Sext = 256 m2 g−1, Smicro = 614 m2 g−1) than AC X17 (Sext = 130 m2 g−1, Smicro = 384 m2 g−1) but O3/AC coupling is more efficient with AC X17 than with F22, except at pH = 7.2 (Fig. 4). This could be explained by the chemical properties of the material. AC F22 is a neutral AC having few acid (total acid = 0.32 meq g−1) and basic functions (total basic = 0.26 meq g−1), whereas AC X17 is a basic AC presenting a high amount of basic functions (total basic = 0.85 meq g−1). According to Sanchez-Polo et al. [12], basic surface groups favour O3 transformation into hydroxyl radicals: interaction between O3 and pyrrol • groups leads to the generation of O−• 2 which enhances HO generation. Moreover, it appeared that acid functions are also conducive to increasing O3/AC interaction: although AC L27 presents a microporous surface close to that of F22 (respectively 616 m2 g−1 and 614 m2 g−1), L27 yields significantly better results in DEP removal. This could be explained by the presence of a large amount of acid functions on AC L27 (total acid = 1.57 meq g−1) which enhances O3/AC interaction. According to several studies, these acid groups can enhance HO• generation [11,28]. This radical generation can result from interactions between O3 and deprotonated acid surface groups. The latter −• may lead to the formation of O−• 2 and O3 which is a radical promoter. The DEP degradation kinetic constants obtained were used to estimate the kinetic contribution of the heterogeneous reaction to DEP removal from Eq.10 (Table 5). Results show that, at acidic pH, reactions are due to heterogeneous reactions and occur in the bulk liquid when pH increases: δhetero ranged between 86.7% and 99.3% at pH = 2.5 and between 11.7% and 29.2% at pH = 7.2. This phenomenon can be explained by the evolution of ozonation kinetics with pH: in very acidic conditions (pH b 5) ozonation kinetics is slow (Table 3) and the presence of AC significantly enhances DEP removal. When the pH value increases, ozonation in the bulk liquid becomes faster and the presence of AC has less impact on DEP removal. Moreover, the kinetic contribution of heterogeneous reactions to DEP removal depends on the AC properties: results show that this contribution is strongly linked to the textural properties of ACs, the best results were being obtained with AC L27 which presents the highest microporous and external surfaces. On the contrary, the smallest values of δhetero were obtained with AC S21, confirming that an AC presenting a low mesoporous volume disfavours its interaction with AC. 3.2.3. O3/AC coupling in presence of tBuOH The above results have demonstrated that O3/AC interaction depends both on textural and chemical properties. In this coupling, AC can act as an adsorbent material, a radical promoter or initiator, or a reaction support. To determine the nature of AC action and the influence of AC properties on its action, and to further our understanding of this coupling, the same experiments were carried out in the presence of tBuOH as radical scavenger. This compound was chosen because it reacts rapidly with hydroxyl radicals (ktBuOH/HO• =

T.F. de Oliveira et al. / Desalination 276 (2011) 359–365

b

0.08

[DEP] (g L-1)

[DEP] (g L-1)

a

0.04

363

0.08

0.04

0

0 0

50

0

100

t (min)

c

d 0.08

[DEP] (g L-1)

0.08

[DEP] (g L-1)

100

50

t (min)

0.04

0.04

0

0 0

10

0

20

5

t (min)

10

t (min)

e [DEP] (g L-1)

0.08

0.04

0 0

2.5

5

t (min) Fig. 4. DEP degradation at pH 2.5 (a), 3.5 (b), 5.6 (c), 6.2 (d) and 7.2 (e) by O3/CA coupling with AC L27 (♦), S21 (□), X17 (▲), F22 (○) and by single ozonation (×).

Table 4 Modelling of DEP degradation by O3/AC coupling at different pH values. pH

L27 S21 X17 F22

2.5

3.5

5.5

6.2

7.2

kglobal (min−1)

r2

kglobal (min−1)

r2

kglobal (min−1)

r2

kglobal (min−1)

r2

kglobal (min−1)

r2

0.290 0.027 0.489 0.122

0.992 0.998 0.995 0.989

0.161 0.037 0.087 0.085

0.992 0.997 0.990 0.994

0.518 0.258 0.327 0.301

0.987 0.998 0.992 0.998

0.921 0.770 0.754 0.640

0.987 0.998 0.995 0.979

0.866 0.697 0.721 0.755

0.990 0.993 0.996 0.994

5.108 L mol−1 s−1) and slowly with O3 (ktBuOH/O3 = 3.10−2 L mol−1 s−1) [27]. Moreover, it was previously verified that, in the experimental conditions used, tBuOH does not modify DEP adsorption on AC: DEP adsorption kinetics and isotherms were carried out with and without tBuOH (not shown here), and no significant difference was observed. The comparison of DEP concentration evolution presented on Fig. 5 with those obtained without radical scavenger

Table 5 Estimation of the kinetic contribution of heterogeneous reactions to DEP removal (δhetero) at different pH values (%). pH

2.5

3.5

5.5

6.2

7.2

L27 S21 X17 F22

98.7 86.7 99.3 97.0

94.2 74.8 89.2 88.9

61.8 23.4 39.5 34.3

48.3 38.1 36.8 25.6

29.2 11.7 14.9 18.8

(Fig. 4) shows that the presence of tBuOH significantly slows down DEP removal. These results indicate that radical reactions play an important role in DEP removal by O3/AC coupling. DEP degradation kinetics was correctly modelled (r2 N 0.980) by the model previously described (Eq. 13). The kinetic constants obtained (kobs), presented on Table 6, were significantly smaller than the kinetic constants (kglobal) previously obtained. Moreover, at each pH value, the kinetic constants obtained are relatively close (for example, at pH = 2.5, kobs ranges between 0.013 min−1 (S21) and 0.038 min−1 (F22) and at pH 7.2, kobs ranges between 0.054 min−1 (S21) and 0.099 min−1 (F22)). As these experiments were carried out in the presence of a radical scavenger, the only effects occurring were adsorption and direct ozonation in the bulk liquid and on the AC surface. The results obtained (very slow degradation of DEP) indicate that these effects do not play an important role in DEP degradation by O3/AC coupling. Thus, it was proven that in this coupling, adsorption is not a preponderant effect and AC does not mainly act as an adsorbent

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b

0.08

0.08

[DEP] (g L-1)

[DEP] (g L-1)

a

0.04

0

0 0

200

100 t (min)

d

0.08

0.04

0

100 t (min)

200

0

100 t (min)

200

0.08

[DEP] (g L-1)

[DEP] (g L-1)

c

0.04

0

0.04

0 0

100 t (min)

e

200

[DEP] (g L-1)

0.08

0.04

0 0

100 t (min)

200

Fig. 5. DEP degradation in presence of tBuOH at pH 2.5 (a), 3.5 (b), 5.6 (c), 6.2 (d) and 7.2 (e) by O3/CA coupling with AC L27 (♦), S21 (□), X17 (▲), F22 (○) and by single ozonation (×).

material. It may explain why the simple kinetic model used correctly describes DEP degradation kinetics in spite of modelling adsorption by the pseudo-first order model: adsorption on AC is a minor way of DEP removal. Taking into account previous results concerning the reaction location, the main effects seem to be degradation of the pollutant by radical reactions on the AC surface or initiated and promoted by interactions between O3 and AC at acidic pH (b5). For higher pH values DEP degradation is mainly due to indirect ozonation in the bulk liquid, and AC doesn't enhance DEP removal. It indicates that in this coupling, AC is used more as a reaction support and radical initiator and promoter than as an adsorbent material. Nevertheless, AC cannot be considered as a reaction catalyst on the grounds of its modifications during the treatment process. According to Sanchez-Polo et al. [12], the ozonation of AC leads to a progressive increase in acid surface groups and a decrease in basic ones. Moreover, Alvarez et al. showed that an ozone-treated AC surface becomes more hydrophilic and

acidic, with a decrease in pHPZC [29]. A recent study has shown that in experimental conditions used, ozonation of AC leads to its chemical properties modification, without textural properties modification [30]. Estimation of the kinetic contribution of radical reactions to DEP removal presented on Table 7 confirms that DEP degradation by the O3/AC process mainly occurs by radical reactions, δHO• ranging between 52.6% and 96.1%. This demonstrates the great interest of this coupling: even in very acidic conditions, it can generate hydroxyl radicals by interaction between O3 and AC surface groups. Moreover, results show that at low pH values (pH b 5), the O3/AC coupling mechanism depends on the AC chemical properties. At this pH, it was previously shown that DEP degradation occurs on the AC surface. Therefore, the values obtained show that ACs presenting a high amount of acidic functions (L27) or basic functions (X17) favour DEP degradation by radical mechanisms, whereas smaller values of δHO•

Table 6 Modelling of DEP degradation by O3/AC coupling in presence of tBuOH at different pH values. pH

2.5 kobs (min

L27 S21 X17 F22

0.035 0.013 0.028 0.038

3.5 −1

)

r

2

0.978 0.980 0.991 0.986

kobs (min 0.035 0.011 0.024 0.031

5.5 −1

)

r

2

0.992 0.995 0.995 0.983

kobs (min 0.074 0.027 0.029 0.067

6.2 −1

)

r

2

0.993 0.991 0.994 0.987

kobs (min 0.073 0.030 0.034 0.059

7.2 −1

)

r

2

0.986 0.996 0.996 0.997

kobs (min−1)

r2

0.091 0.054 0.060 0.099

0,995 0,997 0,997 0,999

T.F. de Oliveira et al. / Desalination 276 (2011) 359–365 Table 7 Estimation of the kinetic contribution of radical reactions to DEP removal (δHO•) at different pH values (%). pH

2.5

3.5

5.6

6.2

7.2

L27 S21 X17 F22

87.9 51.8 94.3 68.9

78.3 70.3 72.4 63.5

85.7 89.5 91.1 77.7

92.1 96.1 95.5 90.1

89.5 92.2 91.7 86.9

were obtained with ACs presenting a low amount of surface functions (S21 and F22). At higher pH values, the kinetic contributions of radical reactions to DEP removal are close for each AC used. In these conditions, the presence of AC influences DEP kinetic degradation only to a minor degree, the latter mainly occurring indirectly by ozonation in the bulk liquid. 4. Conclusion The influence of AC chemical and textural properties on O3/AC coupling was determined by performing a kinetic study of DEP degradation. Compared to the single ozonation process, this coupling was found to enhance DEP degradation kinetics, and a fast and complete pollutant removal was obtained. This study has highlighted the effects of AC properties. It has been shown that high external and microporous surfaces both favour the efficiency of the treatment, by enhancing intraparticle transfer of species in AC. It has also been demonstrated that the presence of chemical functions on the AC surface (basic or acid functions) favours O3/AC interaction and the efficiency of the process. These functions initiate and promote radical reactions even in very acidic conditions. The use of tBuOH as radical scavenger furthers the understanding of the mechanism of this coupling. It appeared that DEP degradation is due to radical reactions. These reactions are mainly due to O3/AC interactions at very acidic pH values (pH b 5) to indirect ozonation at higher pH values. These radical reactions have been shown to be promoted by acid or basic AC surface groups. Thus, in this coupling, AC acts as a radical initiator and promoter and as a reaction support. The most suitable solid material to be coupled with O3 should present a large external surface with high amounts of surface groups. Moreover, as degradation mainly occurs by radical reactions, it can be assumed that a complete mineralization of the pollutant could be obtained with this treatment method. Acknowledgments The authors wish to thank Xavier Bourrain and the Agence de l'Eau Loire Bretagne for their technical and financial support, the Conseil Regional du Centre for its financial support, Pica S.A for gratuitously supplying AC and Christine Picard for her technical support. References [1] F. Alatriste-Mondragon, R. Iranpour, B.K. Ahring, Toxicity of di-(2-ethylhexyl) phthalate on the anaerobic digestion of wastewater sludge, Water Research 37 (2003) 1260–1269. [2] P. Roslev, K. Vorkamp, J. Aarup, K. Frederiksen, P.H. Nielsen, Degradation of phthalate esters in an activated sludge wastewater treatment plant, Water Research 41 (2007) 969–976. [3] S. Barnabé, I. Beauchesne, D.G. Cooper, J.A. Nicell, Plasticizers and their degradation products in the process streams of a large urban physicochemical sewage treatment plant, Water Research 42 (2008) 153–162.

365

[4] B.L. Yuan, X.-Z. Li, N. Graham, Aqueous oxidation of dimethyl phthalate in a Fe(VI)-TiO2-UV reaction system, Water Research 42 (2008) 1413–1420. [5] M. Huang, Y. Li, G. Gu, Chemical composition of organic matters in domestic wastewater, Desalination 262 (2010) 36–42. [6] A.P. Wezel, P. Van Vlaardinger, R. Posthumus, G.H. Crommentuijin, D.T.H.M. Sijim, Environmental risk limits for two phthalates, with special emphasis on endocrine disruptive properties, Ecotoxicology and Environmental Safety 46 (2000) 305–321. [7] M. Wittassek, J. Angerer, M. Kolossa-Gehring, S.D. Schäfer, W. Klockenbusch, L. Dobler, A.K. Günsel, A. Müller, G.A. Wiesmüller, Fetal exposure to phthalates – a pilot study, International Journal of Hygiene and Environmental Health 212 (2009) 492–498. [8] M. Bodzek, M. Dudziak, K. Luks-Betlej, Application of membrane techniques to water purification, Removal of phthalates, Desalination 162 (2004) 121–128. [9] M. Sanchez-Polo, R. Leyva-Ramos, J. Rivera-Utrilla, Kinetics of 1,3,6-naphthalenetrisulphonic acid ozonation in presence of activated carbon, Carbon 43 (2005) 962–969. [10] P.P.C. Faria, M.F.R. Pereira, J.J.M. Orfão, Ozone decomposition in water catalyzed by activated carbon: influence of chemical and textural properties, Industrial Engineering chemistry research 45 (2006) 2715–2721. [11] H. Valdés, A.C. Zaror, Heterogeneous and homogeneous catalytic ozonation of benzothiazole promoted by activated carbon: kinetic approach, Chemosphere 65 (2006) 1131–1136. [12] M. Sánchez-Polo, U. Von Gunten, J. Rivera-Utrilla, Efficiency of activated carbon to transform ozone into OH radicals: influence of operational parameters, Water Research 39 (2005) 3189–3198. [13] T. Merle, J.S. Pic, M.H. Manero, S. Mathé, H. Debellefontaine, Influence of activated carbons on the kinetics and mechanisms of aromatic molecules ozonation, Catalysis Today 151 (2010) 166–172. [14] M. Sánchez-Polo, J. Rivera-Utrilla, Effect of the ozone–carbon reaction on the catalytic activity of activated carbon during the degradation of 1,3,6-naphthalenetrisulphonic acid with ozone, Carbon 41 (2003) 303–307. [15] P.C.C. Faria, J.J.M. Órfão, M.F.R. Pereira, Mineralisation of coloured aqueous solutions by ozonation in the presence of activated carbon, Water Research 39 (2005) 1461–1470. [16] H. Valdés, C.A. Zaror, Ozonation of benzothiazole saturated-activated carbons: influence of carbon chemical surface properties, Journal of Hazardous Materials 137 (2006) 1042–1048. [17] T. Ferreira, O. Chedeville, B. Cagnon, H. Fauduet, Degradation kinetics of DEP in water by ozone/activated carbon process: influence of pH, Desalination 269 (2011) 271–275. [18] M. Clara, G. Windhofer, W. Hartl, K. Braun, M. Simon, O. Gans, C. Scheffknecht, A. Chovanec, Occurrence of phthalates in surface runoff, untreated and treated wastewater and fate during wastewater treatment, Chemosphere 78 (2010) 1078–1084. [19] H. Bader, J. Hoigné, Determination of ozone in water by the indigo method, Water Research 15 (1981) 449–456. [20] H.F. Stoeckli, Porosity, Carbon, in: J. Patrick (Ed.), Arnold, London, 1995, pp. 67–92. [21] H.F. Stoeckli, M.V. Lopez-Ramon, D. Hugi-Cleary, A. Guillot, Micropore sizes in activated carbons determined from the Dubinin–Radushkevich equation, Carbon 39 (2001) 1115–1116. [22] K.S.W. Sing, D.H. Everett, R.A.W. Haul, L. Moscou, R.A. Pierotti, J. Rouquerol, T. Siemieniewska, Reporting physisorption data for gas/solid systems with special reference to the determination of surface area and porosity, Pure and Applied Chemistry 57 (1985) 603–619. [23] H.P. Boehm, Surface oxides on carbon and their analysis: a critical assessment, Carbon 40 (2002) 145–149. [24] J. Rivera-Utrilla, M. Sánchez-Polo, Ozonation of 1,3,6-naphthalenetrisulphonic acid catalysed by activated carbon in aqueous phase, Applied Catalysis B: Environmental 39 (2002) 319–329. [25] O. Chedeville, M. Debacq, M. Ferrante Almanza, C. Porte, Use of an ejector for phenol containing water treatment by ozonation, Separation and Purification Technology 57 (2007) 201–208. [26] U. Von Gunten, Ozonation of drinking water: Part I, Oxidation kinetics and product formation, Water Research 37 (2003) 1443–1467. [27] J. Hoigné, H. Bader, Rate constants of reactions of ozone with organic and inorganic compounds in water—I: Non-dissociating organic compounds, Water Research 17 (1983) 173–183. [28] H. Dehouli, O. Chedeville, B. Cagnon, V. Caqueret, C. Porte, Influences of pH, temperature and activated carbon properties on the interaction ozone/activated carbon for a wastewater treatment process, Desalination 254 (2010) 12–16. [29] P.M. Álvarez, J.F. García-Araya, F.J. Beltrán, F.J. Masa, F. Medina, Ozonation of activated carbons: effect on the adsorption of selected phenolic compounds from aqueous solutions, Journal of Colloid and Interface Science 283 (2005) 503–512. [30] J.-F. Cherrier, B. Cagnon, O. Chedeville, V. Caqueret, C. Porte, Etude de la cinétique d'adsorption de l'acide gallique par un charbon actif soumis à l'ozone Récents Progrès en Génie des Procédés 98, Lavoisier Technique et Documentation, Paris, 2009.