Multiphase Chemistry of Ozone on Fulvic Acids Solutions - American

Nov 13, 2008 - By means of a wetted-wall flow tube, we studied the multiphase chemistry of ozone on aqueous solutions containing fulvic acids (FA), ta...
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Environ. Sci. Technol. 2008, 42, 9165–9170

Multiphase Chemistry of Ozone on Fulvic Acids Solutions MARCELLO BRIGANTE,† BARBARA D’ANNA, PIERRE CONCHON, AND CHRISTIAN GEORGE* Universite´ de Lyon, Lyon, F-69626, France; Universite´ Lyon 1, Lyon, F-69626, France; CNRS, UMR5256, IRCELYON, Institut de Recherches sur la Catalyse et l’Environnement de Lyon, Villeurbanne, F-69626, France

Received June 4, 2008. Revised manuscript received October 14, 2008. Accepted October 15, 2008.

By means of a wetted-wall flow tube, we studied the multiphase chemistry of ozone on aqueous solutions containing fulvic acids (FA), taken as proxies for atmospheric “humic like substances”, so-called HULIS. In these experiments, the loss of gaseous O3 was monitored by UV-visible absorption spectroscopy at the reactor outlet (i.e., after contact between the gaseous and liquid phases). Measurements are reported in terms of dimensionless uptake coefficients (γ) in the range from 1.6 × 10-7 to 1.3 × 10-5 depending on ozone gas phase concentration (in the range from 6.6 to 34.4 × 1011 molecules cm-3) and fulvic acid aqueous concentration (in the range from 0.25 to 2.5 mg L-1) and pH (in the range from 2.5 to 9.2). The measured kinetics were observed to follow a Langmuir-Hinshelwood type mechanism, in which O3 first adsorbs on the liquid surface and then reacts with the Fulvic Acid molecules. The reported uptake coefficients are greatly increased over those measured on pure water, demonstrating that the presence in solution of fulvic acids does greatly enhance the uptake kinetics. Accordingly, the chemical interactions of fulvic acids (or HULIS) may be a driving force for the uptake of ozone on liquid organic aerosols and can also represent an important mechanism for the O3 deposition to the rivers and lakes.

Introduction Humic type substances comprise the major organic components of soils and sediments, and play an important role during physical and chemical processes in the natural environment (1–4). Their presence and photoexcitation can also modify or promote the degradation of many pollutants presents in aquatic environments (5–7). The reactive nature of humic substances is largely due to the presence of oxygen-containing functional groups which react with metals (8, 9) and organic compounds (10). These chemical functionalities include sCOOH, phenolic-, enolic sOH, alcoholic sOH, and CdO of quinones, hydroxyquinones, and ketones (11). A major component of the humic substances, present both in soils and natural waters, is fulvic acid (FA) (12), defined as the yellow to yellow-brown humic fraction soluble in water under all pH conditions. The size * Corresponding author e-mail: christian.george@ircelyon. univ-lyon1.fr. † Current address: Laboratoire de Photochimie Mole´culaire et Macromole´culaire, University Blaise Pascal at Clermont-Ferrand, France. 10.1021/es801539y CCC: $40.75

Published on Web 11/13/2008

 2008 American Chemical Society

of fulvic acids (FA) is smaller than humic acids (HA) with molecular weights which range from approximately 1000 to 10 000 (13). Fulvic acids have an oxygen content twice that of humic acids and many carboxyl (sCOOH) and hydroxyl (sOH) groups. In addition to being ubiquitous in soil and surface water, humic like substances (HULIS) have been recently identified in tropospheric organic aerosols. These large molecules do exhibit similar chemical properties to the fulvic acids (UV-vis absorbance, fluorescence) as obtained from field measurements (14). Therefore fulvic acids have been suggested as their representative molecules in atmospheric aerosols (15). However, atmospheric HULIS may have masses smaller than 1000 Da (16–19) and therefore be slightly lighter than fulvic acids found in surface water. As the recognition of their presence in the atmosphere is very recent, our knowledge on their sources and chemistry is spare and certainly warrants further investigations. Currently, a significant number of investigations have been focused on the formation and characteristics of these oligomeric molecules (20). HULIS seems to be formed directly in situ in the troposphere via photo-oxidation of primary biogenic and anthropogenic precursors (14) or directly emitted (21, 22) (i.e., by biomass burning or wind erosion). The chemical structure of HULIS therefore strongly depends on the nature of its precursors. In addition, it also appears that atmospheric HULIS may take part in many atmospheric processes such as light scattering and absorption (22, 23) and cloud formation (24–27). To our knowledge, there are only a limited number of studies investigating how fulvic acids may alter the uptake kinetics of atmospheric trace gases. The reaction between FA and ozone, taken as an efficient disinfectant agent (28), has been previously investigated, and some semivolatile products were identified after ether-extraction, methylation with diazomethane, and analysis by GC-MS (29, 30). Nevertheless, there is missing information on the associated uptake kinetics possibly expressed in terms of dimensionless uptake coefficients γ, considered as useful input parameters for modeling studies. Accordingly, we studied the multiphase chemistry, at the air-water interface, between gaseous ozone and aqueous solutions containing fulvic acids by means of a wetted-wall flow tube apparatus allowing the determination of the uptake coefficients under a variety of conditions.

Experimental Methods Fulvic acids solutions. Suwannee River FA were obtained from the International Humic Substance Society (IHSS) and were used without furthermore purification. A fresh solution was prepared, before each experiment, using milli-Q water (Milli Q 50 system). To determine the FA concentration and its stability in solution, UV-vis absorption spectra (190-600 nm) were recorded with a double beam Uvikon 930 spectrophotometer (Kontron instruments) before and after each experiment (31). These measurements provide a direct measure of the stability of the solution along with the measure of the dilution factor by simply comparing the UV spectra with the one of the commercial FA solutions. The pH of the solutions was adjusted using NaOH or H2SO4 before each experiment as required. All experiments were carried out using an FA concentration in the range from 0.25 to 2.5 mg L-1. The solutions were not buffered, i.e., the pH may have changed during the course of the reaction. It must be underlined that in the case of a surface reaction (as discussed below), buffering the bulk solution may have little impact on the “pH” at the surface but would introduce other concenVOL. 42, NO. 24, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 1. ln (A0/A) of ozone gas phase (absorbance at 255 nm) versus the exposure time to the fulvic acids solution in the flow tube for different ozone concentrations: 7.4 × 1011 molecules cm-3 (∆), 9.8 × 1011 molecules cm-3 (2), 1.4 × 1012 molecules cm-3 (O), 3.0 × 1012 molecules cm-3 (b). The first 10 cm (contact time ∼1 s) were not used for the analysis because the laminar flow profile may not be fully established within the first few cm of tube length and also this length is required for the relative humidity to reach 100%. trated compounds in our solutions and especially ions that may alter the interfacial chemistry. This was avoided during these experiments (32, 33). Ozone Generation. Ozone was generated by flowing 250 mL min-1 of pure air through a UV light source (UVP-ozone generator) and was used without furthermore dilution. The ozone volume mixing ratio was controlled by varying the air flow with a mass flow meter (BROOKS 5850 TR) and the intensity of UV irradiation by an adjustable aperture leading to ozone concentrations in the range from 6.6 to 34.4 × 1011 molecules cm-3. This ozone flow was then fed into the top of the flow tube, which was maintained at atmospheric pressure. Pure air (Air Liquide Alphagaz 1) was used as carrier gas. Wetted-Wall Flow Tube (WWFT). This technique has been previously used to investigate the multiphase reactions between atmospherics oxidants (e.g., O3 or NO2) and aqueous solutions containing phenolic compounds (34) or biogenically derived compounds (35). Therefore only a brief description is given below. The principle of this technique relies on measuring the loss rate of gaseous ozone flowing along a vertical aligned flow tube, which inner walls are covered by a liquid film (containing fulvic acids) flowing down by gravity. The flow tube is a 70 cm long and 1.0 cm large (internal diameter) Pyrex tube, kept at constant temperature by circulating thermostatted water through an outer jacket. The reagent solution containing fulvic acid is injected at the top of the WWFT using a peristaltic pump, and evenly distributed over the inside walls using an annular reservoir dispenser system, made of Teflon, leading to a homogeneous film flowing downward on the inner walls of the flow tube. The maximum liquid surface area established in these experiments was 188 cm2. The liquid flow rate was 2.5 mL min-1 and the resulting film thickness is 8.7 × 10-3 cm with a Reynolds number of 0.8, which indicates laminar flow condition (turbulence will occur for NRe > 2300). Ozone is introduced through a movable injector and its decay is measured at five different tube positions corresponding to 10, 30, 50, 60, and 70 cm of gas/liquid interaction lengths. These positions correspond to contact times between 2 and 13 s, given the overall flow velocity in the WWFT. The first 10 cm are not used in the data analysis because the 9166

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laminar flow profile may not be fully established within the first few cm of tube length, also this length is required for the relative humidity to reach 100%. After this “injection zone”, the gas flow is well established and the vapor pressure is in equilibrium with the liquid film, allowing kinetic measurements to be made under reproducible conditions. The gas extracted from the flow tube passed through an optical absorption cell having an optical path length of 1 m. The ozone partial pressure was monitored by optical detection at 254 nm using a spectrograph coupled to a CCD camera. Typical experimental conditions employed O3 concentrations in the range of 6.6 to 34.4 × 1011 molecules cm-3. For the conversion of the measured absorbance into concentration of ozone (molecules cm-3), an absorption cross section of 1.1 × 10-17 cm2 molecule-1 at 255 nm has been used (36). The measured loss rate can be interpreted in terms of a first order process with respect to the gas phase concentration of the reactant. As shown in Figure 1, the measured kinetics are first order with respect to the gas phase concentration of the reacting species and therefore obeys following relationship: ln

A0 ) kobst A

(I)

where A0 and A are the ozone absorbance before (at time ) 0 s) and after reaction with the liquid film, t is the average gas residence time and kobs is the observed rate coefficient for the ozone loss at the liquid film surface. The derived pseudofirst order coefficient is related to the uptake coefficient (γ) though the following equation: kobs ) γ〈c〉 ⁄ 2r

(II)

where r is the radius of the tube and 〈c〉 is the ozone mean thermal velocity(8RT/πM)0.5. Error Analysis and Diffusion Correction. The experimental errors reported below were derived at the 2σ level simply from the scattering of the experimental data. However, another systematic error may arise from gas phase diffusion limitations. In fact, during our experiments, gas phase diffusion limitations may be introduced when a significant radial gas phase concentration gradient builds up. This

FIGURE 2. Evolution of the uptake coefficient (γ) as a function of different ozone concentrations (from 6.6 to 34.4 × 1012 molecules cm-3) in the flow tube for fulvic acids solutions at 0.5 mg L-1 in milli-Q water. The line shows the fit of the data to an inverse function of [O3(g)], as implied by the Langmuir-Hinshelwood mechanism.

FIGURE 3. Evolution of the uptake coefficient of ozone (1.5 × 1012 molecules cm-3) uptake onto fulvic acids solutions (0.5 mg L-1) at different pH values (adjusted using NaOH or H2SO4) using an O3 concentration of 1.5 × 1012 molecules cm-3. happens when the surface chemical loss is faster than the mass transport in the gas phase. To avoid these errors, our data were systematically corrected for gas phase diffusion limitations using the Cooney-Kim-Davis (CKD) method (37, 38) which describes numerically the coupled diffusion and wall losses in a cylindrical geometry as those found in a flow tube. The diffusion coefficient of ozone in air were taken from the review from Massman (39) to be 0.137 cm2 s-1. At P ) 1 atm and T ) 293 K, the corrections of the uptake coefficient for slow gas phase diffusion were between 1 and 16%, with larger correction required for the uptake coefficients approaching 10-5. Chemicals. Sodium Hydroxide (98%, Across Organics), Sulfuric acid (96%, Across Organics), Air gas (ALPHAGAZ Air 1) were used as received without any further purification.

Results and Discussion Figure 1 reports the ozone decay on a FA solution (0.5 mg L-1) for different ozone concentrations as a function of the

exposure time (t). The pseudo first order rate coefficients (k) are derived from the linear regression of each exponential decay. For ozone mixing ratios of 3.4 × 1012 and 1.45 × 1011 molecules cm-3, the measured decay rates are very similar and the derived first order rate constants are close to (3.0 ( 0.5) × 10-2 s-1. When lowering the O3 concentration (down to 9.5-7.4 × 1011 molecules cm-3), the decay rates become faster with rate constants up to 1.5 × 10-1 s-1. These observations allow us to derive the kinetics of gaseous ozone uptake onto/into FA containing solutions as a function of the initial concentration of ozone. In Figure 2 we report the uptake coefficient (γ) as a function of the ozone concentrations. It can be seen that γ increases from 1.6 × 10-7 to 1.3 × 10-5 when the ozone concentration decreases from 4.7 × 1012 to 6.9 × 1011 molecules cm-3, with a particularly strong increase of γ observed around 1.0 × 1012 molecules cm-3. The nonlinear dependence of the uptake coefficient on the ozone concentration is suggestive of a surface mediated reaction, described by a Langmuir-Hinshelwood type mechVOL. 42, NO. 24, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 4. (A) Variation of the uptake coefficient with the square root of the fulvic acid concentration. Solid line is the calculated uptake using formula (III) and the dashed line is calculated taking into account surface reaction (IV). (B) Ozone uptake coefficient as a function of the fulvic acid concentrations in milli-Q water solutions using an O3 concentration of 1.5 × 1012 molecules cm-3. anism (40, 41). Such a mechanism is characterized by a rapid adsorption equilibrium between gaseous ozone and the liquid surface, followed by a reaction with fulvic acid molecules at the interface (in opposition to FA solvation followed by a bulk reaction). The same surface mechanism has been postulated in other cases, such as for gaseous ozone reacting with chlorophyll at the seawater surface (35). In this study, we also investigated the pH dependence (from 2.5 to 9.2) of the uptake rate for a FA concentration of 0.5 mg L-1. By comparing the γ values reported in Figure 3 for an ozone concentration of 1.5 × 1012 molecules cm-3, it is evident that the uptake coefficient is constant in acid to neutral conditions, but under alkaline conditions it increases with increasing pH reaching a maximum value of 1.4 × 10-6. It must be emphasized that the ozone loss in the blank experiment (i.e., without fulvic acids) at different pHs is very low with uptake coefficients at/or below our detection limit (γ e 10-7). Alvarez-Pueblaa and co-workers (42) have demonstrated that a correlation exists between the ionic state of fulvic acids, which is connected to the pH, and their aggregation in solution. They proposed different FA structures as a function 9168

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of the pH. First at pH ∼ 3.4, FA molecules are protonated and Van der Walls forces “create” a flexible structure. Then with increasing pH, carboxylic and phenolic groups become ionized and the electrostatic repulsion generated by the charge increment tends to expand the FA molecules. These chemical functions are then more readily available for reactions with ozone. Jiang et al. (43) showed a correlation between ozone and pH, in the particular case of the oxidation of nitrobenzoic acid (NBA). Where, an increase of pH induces the formation of carboxylate (sCOO-), an electron-donor, and favors the electrophilic attack by ozone. But, as shown in Figure 3, faster ozone uptake are observed for pH g 7. This result is incompatible with a pKa of 4-5 typical for the aromatic carboxyl group deprotonation. Therefore, the observed pH trend cannot simply be explained by the sCOOH deprotonation. Our results do support the studies on aqueous phenol degradation by ozone (44, 45), where faster degradation of phenols is observed at pH g 8. Furthermore, our results seems to be in good agreement with a possible deprotonation of polyphenols (pKa ∼ 9-10) considered as an important fraction of fulvic acids in soils (46), as for the standard FA used in these experiments. However, there is

currently no information available for the presence of polyphenols in atmospheric HULIS, i.e., the extrapolation of these observations to atmospheric conditions is therefore still speculative. It is well-known that ozone can react indirectly with organic and inorganic compounds in water. Xiong et coworkers (29), using a high ozone concentration (g0.5 mg L-1), proposed a faster reaction with fulvic acids introduced by a radical type chain reaction, initiated by the presence of hydroxyl anions (i.e., under alkaline conditions). Staehelin and Hoigne´ (47, 48) have observed that organic and inorganic solutes can accelerate ozone consumption in water interacting with •OH by converting •OH into HO•2. Again, these observations are in agreement with our observations. Under conditions where bulk liquid phase reaction drives uptake from the gas phase, the uptake coefficient is expected to be described by the following equation (49):

γ)

√

4HO3RT kII[FA]DO3

(III)

where HO3 is the Henry’s law constant for ozone (HO3) 1.13 × 10-2 mol L-1 atm-1), R is the gas constant, DO3 is the aqueous phase diffusion coefficient (D ) 1.176 × 10-5 cm2 s-1) (50), T is the temperature, kII is the second order rate constant for the reaction FA + O3, and [FA] represents the bulk fulvic acid concentration. Equation III assumes a rapid gas/liquid phases equilibrium (as described by the Henry’s law) and a homogeneous bulk phase reaction as unique chemical driving force for the uptake process. Figure 4a shows a plot of the uptake coefficients versus the square root of the fulvic acid concentration (dependence introduced by eq III. The observed dependence is reasonably linear. However, the intercept is largely negative which has no physical meaning. In other words, the linear relationship is accurate only for the highest FA concentrations, while the equation does not capture the features of the uptake coefficients at low FA concentrations. Considering the concentration region where some linearity is found (as shown in Figure 4b), and assuming a molecular weight between 1000 and 10 000 g mol-1 (51), we can estimate the second-order rate coefficient for reaction (IV) between O3 gas phase and fulvic acid solution to be comprised between 5.6 × 107 and 5.6 × 108 mol L-1 s-1. The existence of the large negative intercept in Figure 4a could also be indicative of FA consumption at the air/ interface, which would lower the uptake rate and lead to a misinterpretation of the relationship between the uptake coefficient and the fulvic acid concentration. However, simple calculations show that FA concentration remains reasonably constant at the low ozone mixing ratios used in these experiments. Another explanation for this negative intercept has been discussed previously in different chemical systems where such observations have been demonstrated to be suggestive of a surface reaction between the in-coming gas (i.e., ozone) and the scavenger (in this case the fulvic acid). Indeed, the fact that eq III does not provide a satisfactory description of the kinetics at low FA concentrations suggests that the bulkphase reaction is not the exclusive loss process for ozone. The behavior shown in Figure 4a has been noted previously when there is a competition between bulk and surface reaction (52–54). Indeed, Hu et al. (54) showed that in systems where there is a surface component to the reactive uptake, there is a transition between a surface controlled regime (in this case at low FA concentration) and a bulk regime (at high FA concentration). In such a case, the uptake depends linearly with the reactant (surface reaction) and then switches to bulk control, as tentatively indicated by the following:

γ)

4HO3RT√kbulkDO3

+ γsurf

(IV)

where γsurf represent the uptake coefficient associated to the surface reaction between O3 and FA solution. Note that eq IV is a highly simplified equation taking into account surface reactivity. At low FA concentrations, its surface concentration is expected to be a linear function of the bulk concentration. The linear dependence between the ozone uptake coefficient and fulvic acid concentrations is displayed in Figure 4b, the FA concentrations varied from 0.25 to 2.5 mg L-1. The experiments were performed at 293 K and 1.5 × 1012 molecules cm-3 of ozone. Given the results presented above, it is reasonable to postulate the existence of a surface reaction as well as a bulk reaction. The dashed line in Figure 4a shows a fit to the data assuming a FA concentration dependence which is both linear and square root, as suggested by eq IV. This function shows a better fit than that of the dashed line, consistent with a surface component to the uptake of ozone by solutions containing fulvic acids. The calculated lifetime associated with the dark heterogeneous reaction of O3 is a few months and is not a significant sink for ozone in the atmosphere. However, the presence of a surface reaction involving ozone and HULIS could be very important for the surface properties of the aerosol (such as hygroscopicity) and the corresponding aging of the aerosols. Finally, the chemical interaction between ozone and fulvic acid are finally relatively fast and might be the driving force for the ozone deposition on surface water.

Acknowledgments C.G. acknowledges EUCAARI for a postdoctoral fellowship provided to M.B.

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