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Environmental implications of soil remediation using the Fenton process
Available online at www.sciencedirect.com
Chemosphere 71 (2008) 43–50 www.elsevier.com/locate/chemosphere
Environmental implications of soil remediation using the Fenton process Ricardo D. Villa, Alam G. Trovo´, Raquel F. Pupo Nogueira
*
UNESP – Sa˜o Paulo State University, Institute of Chemistry of Araraquara, R. Prof. Francisco Degni s/n, P.O. Box 355, 14801-970 Araraquara, SP, Brazil Received 17 September 2007; received in revised form 22 October 2007; accepted 23 October 2007 Available online 18 December 2007
Abstract This work evaluates some collateral effects caused by the application of the Fenton process to 1,1-bis(4-chlorophenyl)-2,2,2-trichloroethane (DDT) and diesel degradation in soil. While about 80% of the diesel and 75% of the DDT present in the soil were degraded in a slurry system, the dissolved organic carbon (DOC) in the slurry filtrate increased from 80 to 880 mg l1 after 64 h of reaction and the DDT concentration increased from 12 to 50 lg l1. Experiments of diesel degradation conducted on silica evidenced that soluble compounds were also formed during diesel oxidation. Furthermore, significant increase in metal concentrations was also observed in the slurry filtrate after the Fenton treatment when compared to the control experiment leading to excessive concentrations of Cr, Ni, Cu and Mn according to the limits imposed for water. Moreover, 80% of the organic matter naturally present in the soil was degraded and a drastic volatilization of DDT and 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene was observed. Despite the high percentages of diesel and DDT degradation in soil, the potential overall benefits of its application must be evaluated beforehand taking into account the metal and target compounds dissolution and the volatilization of contaminants when the process is applied. 2007 Elsevier Ltd. All rights reserved. Keywords: Air contamination; DDT; Diesel; In situ chemical oxidation; Metals; Organic matter
1. Introduction The increasing contamination of soil and water by organic compounds has motivated the development and application of several remediation processes (Khan et al., 2004) such as the advanced oxidation processes (AOPs). These processes are receiving great emphasis because of their ability to rapidly oxidize refractory organic contaminants (Pignatello et al., 2006). The Fenton process is one of the most common AOPs and it has been the subject of considerable interest for the remediation of contaminated soils. The Fenton reaction consists in the catalytic decomposition of hydrogen
*
Corresponding author. Tel.: +55 16 3301 6606; fax: +55 16 3301 6692. E-mail addresses: [email protected] (R.D. Villa), alamtrovo@ hotmail.com (A.G. Trovo´), [email protected] (R.F.P. Nogueira). 0045-6535/$ - see front matter 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.10.043
peroxide (H2O2) due to soluble iron(II) in acidic medium, generating hydroxyl radicals (OH). The hydroxyl radical (OH) is a strong oxidizing agent o (E = 2.73 V versus NHE, Wardmann, 1989) capable of non-selectively oxidizing a variety of organic contaminants. One of the main advantages of Fenton processes is the possibility of in situ remediation, eliminating the need for excavating contaminated soils and thus allowing the rapid treatment of refractory contaminants when compared to other remediation processes (Monahan et al., 2005). The use of Fenton reactions for the treatment of contaminated soils was initially investigated by Watts et al. (1990), who was the first to observe pentachlorophenol mineralization. Subsequently, several other works appeared which emphasized the efficiency of Fenton processes for the remediation of soils contaminated with – among other organic compounds – pesticides (Watts et al., 1994; Villa and Nogueira, 2006), fuel (Kong et al., 1998) and explosives
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
(Liou et al., 2003). However, few works have evaluated the collateral effects of soil remediation using Fenton processes. Therefore, there is little information about other consequences of the oxidation process such as soil organic matter (OM) oxidation, metal and target compounds dissolution, the effect on microorganisms and the volatilization of contaminants caused by the strong exothermic characteristic of the Fenton reaction. The aim of the present work was to evaluate some of those collateral effects – namely OM degradation, the dissolution of metals and target compounds, and the volatilization of contaminants in the remediation of soils contaminated with 1,1-bis(4-chlorophenyl)-2,2,2-trichloroethane (DDT), 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene (DDE) and diesel using the Fenton reaction. 2. Materials and methods 2.1. Chemicals Standard DDT and DDE, at 99% purity were purchased from Supelco. The diesel used was a commercial product. Isooctane (Mallinckrodt) and n-hexane (Baker) were used to prepare the DDT and diesel standard solutions, respectively. A neutral and deactivated 70–290 alumina mesh (Vetec) was used in the DDT extraction process. Pesticide-grade solvents n-hexane, acetone and dichloromethane (Tedia) were used for the extraction of both DDT and diesel. H2O2 30% (w/v) from Synth was used in the degradation experiments, while Na2SO4 (Baker) was used as a desiccation agent. HNO3 (Synth) was used in the decomposition of the samples for metal determination. The solutions of Fe2+ were prepared by dissolving FeSO4 Æ 7H2O (Carlo Erba) in 0.1 M H2SO4 (Synth). Analytical grade metals solutions (Zn, Pb, Cr, Ni, Cu, Cd, Mn and Co) (MV Laboratories, Inc.) were used to construct analytical curves for the determination of these metals. All the glassware was cleaned with Extran MA-01 10% (v/v) from Merck. 2.2. Sampling and soil contamination The soil used in this work was collected from a former pesticide warehouse in the state of Mato Grosso, Brazil. That soil was seriously contaminated with DDT as a consequence of inadequate storage of this substance after its use was restricted by law (Villa et al., 2006). As a consequence of the natural degradation of DDT, the soil has also contamination by DDE since this compound is one of the main degradation products of DDT. After sampling, the soils were passed through a 3.0 mm sieve and dried in air at 25–30 C for 48 h. The soil used to study diesel degradation was collected near the pesticide warehouse (about 30 m). This soil was not contaminated and presented the same characteristics of the soil used in the DDT degradation experiments. The diesel spiked soil was prepared by adding 200 ml of 50 g l1 of n-hexane diesel solution to
2.0 kg of soil. The spiked soil was then vigorously homogenized and left standing for 24 h to eliminate the solvent, which completeness was confirmed by weight loss. 2.3. Experimental conditions of degradation For the diesel degradation experiments, a slurry was prepared by directly mixing 500 ml of 12 mM FeSO4 aqueous solution and 150 g of soil or silica contaminated with diesel at initial concentration of 5.0 mg g1. For DDT degradation, 500 ml of 6.0 mM of FeSO4 aqueous solution and 150 g of contaminated soil (1.6 mg g1) were used. A total volume of 160 ml of 7.0 and 10 M H2O2 was pumped into the slurry for the DDT and diesel degradation, respectively. In both cases, the H2O2 was added using a peristaltic pump (Ismatec model 78017-10) programmed to deliver 1.1 ml of H2O2 solution every 20 min for 12 h. After this period of time, the additions were interrupted for 12 h to allow for the consumption of the H2O2 and then restarted. All the experiments were carried out at ambient temperature and with mechanical stirring at 110 rpm. The initial pH of the slurries was adjusted to between 2.5 and 3.0 by the addition of 3 M H2SO4 solution before starting the experiments, i.e. the optimum pH value for the Fenton reaction in soil (Watts et al., 1990). No further pH adjustment was made during the experiments. Metal dissolution was evaluated by comparing the metal concentrations in the filtrate of the diesel-degradationexperiment slurry (under the same oxidation conditions as described previously) to those of the filtrate of a slurry composed of only 150 g of the diesel-contaminated soil and 500 ml of distilled water. The volatilization of the DDT and DDE was evaluated in an experiment where the reagents were directly pumped to the dry soil. In this experiment, 650 ml of 20 mM Fe and 650 ml of 4.0 M H2O2 solution were merged immediately before its addition to 2.0 kg of soil contaminated with DDT (1.6 mg g1). In the first 30 min of experiment, the reagent flow rate injection was 15 ml min1 with the aim of humidifying the soil. After this time, the pump was programmed to start injecting again every 3 h at a flow rate of 10 ml min1 for a period of 5 min in order to keep the soil humidified. This experiment was carried out for 48 h. The air was collected at a height of 25 cm above the surface of the contaminated soil using a polyurethane cartridge as the absorber (5 cm diameter, 4 cm long, having a mass of 2.30 g) through which the air was pumped at a flow rate of 4 ml min1. A schematic representation of the reactor and the air sampling is shown in Fig. 1. In order to evaluate the extent of the DDT and DDE volatilization, one sample of air was taken during the soil remediation. The air was firstly pumped through the polyurethane cartridge for 48 h after the injection of the Fenton reagents at a flow rate of 4.0 l min1. The concentration of DDT and DDE adsorbed on the polyurethane were determined after soxlet extraction and compared to the concentration obtained in a separate experiment without the
R.D. Villa et al. / Chemosphere 71 (2008) 43–50
45
Polyurethane cartidge
Air pumping (4.0 L min-1). 18 cm 15 cm H2O2
Fe 2 + pH = 2.5
1.0 cm Contaminated soil 23 cm
Net for soil retention
Fig. 1. Schematic representation of the non-slurry system and air sampling device.
addition of the Fenton reagents for the same time and flow rate. 2.4. Chemical analysis DDT and DDE extraction from the soil was carried out using the method developed by Villa et al. (2006). In summary, after drying, grinding and homogenizing the soil, triplicate samples of 0.50 g of soil were mixed with 1.0 g of deactivated alumina. This mixture was transferred to glass columns (1.0 cm internal diameter · 44 cm length), containing 2.0 g of neutral alumina, where the pesticides were eluted with 210 ml of n-hexane:dichloromethane (7:3) solution. The extracts were concentrated and then dissolved in 6.0 ml n-hexane. The DDT and DDE were quantified using gas chromatography with an electron capture detector using a Varian 3300 chromatograph and a DB-5 capillary column (0.32 mm · 30 m). The analyses were carried out under the following conditions: injector temperature 280 C; initial furnace temperature 110 C; final furnace temperature 250 C; a heating rate of 10 C min1; detector temperature 330 C; N2 as carrier gas at a flow rate of 5.0 ml min1. The DDT present in the aqueous filtrates was extracted using liquid–liquid extraction and 50 ml of the sample (filtrate) with 10 ml of n-hexane/dichloromethane solution (85:15) (APHA, 1989). All extractions were carried out in triplicate and the DDT and DDE concentrations were determined under the same conditions as those for the soil extracts. The extraction of DDT and DDE from the polyurethane cartridges was carried out in a Soxlet apparatus for 24 h using 200 ml of hexane:acetone (1:1) as extracting solution. After extraction, the extracts were concentrated and dissolved in 5.0 ml hexane. DDT and DDE were then determined using gas chromatography under the same conditions as for extracts from soil and water. Diesel extraction from the soil was carried out by adding 7.0 ml of n-hexane–dichloromethane solution (1:1) and 1.0 g of sodium sulfate to 3.0 g of spiked soil in an airtight
sealed glass tube (2.0 cm internal diameter and 15 cm long). The mixture was then mechanically stirred at 240 rpm for 2 h. After this time the mixture was centrifuged and the extract collected. This extraction procedure was repeated two more times. The soil extracts were analyzed by gas chromatography using a Shimadzu 14B chromatograph, a DB-5 capillary column (0.32 mm · 30 m) and a flame ionization detector. The following chromatographic conditions were used: injector temperature 280 C; initial furnace temperature 45 C; final furnace temperature 250 C; heating rate 12 C min1; detector temperature 330 C; H2 as carrier gas at a flow rate of 10 ml min1. The diesel concentration was determined in terms of total petroleum hydrocarbons – TPH (EPA 8015B, 1996). Prior to determination of metal content, the soil samples were prepared according to the standard digestion method 6010 (Keith, 1996). The determination of metals in the slurry filtrate was done directly after filtration through a 0.45 lm membrane. In all cases, the quantification was done by flame atomic absorption spectrometry using an Analyst 300 spectrometer (Perkin Elmer). All determinations were carried out in triplicate followed by a blank check. The dissolved organic carbon (DOC) in the slurry filtrates was determined using a Total Organic Carbon analyzer (TOC-5000A Shimadzu) after filtration of the slurry through filter paper and then through a 0.45 lm membrane. The soil texture was determined using the pipette method (Suguio, 1973). The density of the soil was determined using a calibrated tube and the soil pH was determined in water (EMBRAPA, 1997). The soil OM content was determined by thermogravimetric analysis in a Simultaneous Thermal Analysis Module, model number SDT2960 (TA Instruments). 3. Results and discussion The soil used in this work is slightly acidic and has an OM content of 7.5% as shown in Table 1. The metal
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
A
800
-1
0.8 600 0.4 400 0.2
-1
DDT and DDE (mg g )
1.2
200
3.1. Organic matter degradation and contaminant dissolution
0.0
0
12
24
36
48
60
0 72
Time (h)
1000
B 5
800
-1
600 3 400
2
-1
Diesel (mg g )
4
200
1
0
DOC (mg L )
After 64 h of degradation process, the DDT concentration in the soil decreased of 75%, while the DOC in the slurry filtrate increased from 80 to 880 mg l1 (Fig. 2A). The concentration of the DDE decreased of only 40%. However, considering that DDE is the main degradation product of DDT, it can be produced and degraded simultaneously during the experiments (Villa and Nogueira, 2006). The OM present in the soil consists of a complex mixture at several stages of decomposition, resulting from chemical and/or biological degradation of vegetal and animal residues (vanLoon and Duffy, 2000). The greater the OM oxidation stage, the greater the tendency for its solubilization (Rocha and Rosa, 2003). The oxidative conditions of the reaction medium contributed to further oxidation of the OM and consequently to an increase of the DOC concentration in the slurry filtrate. It was also observed that the DDT concentration in the slurry filtrate increased from 12 to 50 lg l1 after 64 h of reaction, probably as a consequence of soil OM dissolution. There is strong evidence that the soluble OM acts as a surfactant, increasing the solubility of hydrophobic compounds, such as DDT (Ding and Wu, 1995; Worrall et al., 2001; Spark and Swift, 2002). The initial amount of DDT in 150 g of soil was 240 mg. Considering that the final concentration in 500 ml of the slurry filtrate achieved 50 lg l1 after the Fenton treatment, it can be calculated that less than 0.01% of the DDT present in the soil was dissolved. However, the concentration of 50 lg l1 in water is 50 times higher than the maximum concentration allowed by the Brazilian Ministry of the Environment (CONAMA, 2007) for surface water destined for irrigation or recreation, which is 1.0 lg l1. The solubilization of contaminant as seen here
1000 1.6
DOC (mg L )
concentrations in the studied soil are within the range reported in literature for all metals analyzed (vanLoon and Duffy, 2000). Interpolating the sand, silt and clay percentages (Table 1) into a textural triangle cited by Brady and Buckman (1983), it was observed that this soil can be classified as loam, where silt and sand are predominant sized fractions. Loam soils present a small surface area when compared to clay soils, and therefore they retain less water, are easily drained and generally have a low plasticity (Brady and Buckman, 1983). These characteristics work together to facilitate the mobility of many contaminants through the soil, mainly the water soluble ones.
0
12 24 36 48 60 72 84
0
Time (h) Fig. 2. (A) DDT and DDE degradation in soil and increase of DOC concentration in the slurry filtrate. Experimental conditions: 6.0 mM FeSO4; additions of 1.1 ml of 7.0 M H2O2 every 20 min (145 additions, total amount: 38.1 g). –j– DDT; –m– DDE; –d– DOC; (B) Diesel degradation in soil and increase of DOC in the slurry filtrate. Experimental conditions: 12 mM FeSO4; additions of 1.1 ml of 10 M H2O2 solution every 20 min (145 additions, total amount: 54.4 g). –– Diesel; –d– DOC; –$– DOC from non-contaminated soil; –D– DOC without Fenton reagents.
can be particularly critical in cases of high toxicity, mainly when in situ remediation is undertaken, since it facilitates the mobility of contaminants through the soil, offering contamination risks to surface water and underground
Table 1 Characteristics of the soil used in this work pH (H2O) 5.7
qb
OM
(g cm3)
(%)
1.5
7.5
fOC
Sand
Silt
Clay
Textural classification
Zn
Cr
Ni
Cu
Cd
Mn
16
27
66.2 · 103
492
(mg kg1) 4.35
41
47
12
Loam
12
64
R.D. Villa et al. / Chemosphere 71 (2008) 43–50
water reservoirs. Furthermore, the amount of contaminants dissolved in the filtrate may be higher for a higher volume of water used for the slurry. A similar behavior was observed for diesel oxidation when compared to DDT degradation. When 80% of the diesel was degraded, the DOC concentration in the slurry filtrate reached 870 mg l1 (Fig. 2B). In the experiment without any addition of H2O2, the DOC remained very low over an 84 h period, evidencing that the OM dissolution was a result of the strong oxidizing conditions of the medium. In the absence of diesel, under the same conditions of iron and H2O2 concentration, the DOC concentration in the slurry filtrate reached 715 mg l1. The difference of 155 mg l1 of DOC in the filtrate of diesel-contaminated and non-contaminated soil can be related to the components of the diesel and its degradation products. Furthermore, in a separate experiment where soil was replaced by silica (the only source of OM was the diesel), the DOC reached 160 mg l1 after 84 h of experiment, reinforcing the hypotheses of the solubilization of the diesel compounds. It can be stated that although most of the DOC concentration in the slurry filtrate originates from the natural OM of the soil, diesel is also a source of DOC due to the formation of soluble degradation products. Considering that diesel is a complex mixture formed by aromatic compounds and other unsaturated hydrocarbons that can be easily oxidized in the presence of H2O2, generation of several soluble compounds, such as carboxylic acid, aldehydes and ketones is possible (Solomons and Fryhle, 2000). A detailed knowledge of the toxicity of these compounds is of great importance to the safe application of the Fenton processes in soil remediation. During diesel degradation, 80% of the OM initially present in the soil was degraded, evidencing the aggressiveness of the Fenton remediation processes on the soil. Besides contributing to soil deterioration, the Fenton process probably contributes to its sterilization. Ferguson et al. (2004) observed the elimination of great part of the microorganisms initially presents in soil after the application of the Fenton process in the remediation of soil contaminated with diesel. They had attributed the death of the microorganisms to the strong oxidizing conditions of the medium, the change in pH and the temperature increase. 3.2. Dissolution of metals during oxidation processes In soil, metals can: (i) participate in dissolution and in precipitation reactions; (ii) be electrostatically absorbed in the exchange sites; (iii) be incorporated into inorganicphase surfaces (specific adsorption); and/or (iv) be associated to the organic fraction (Sodre´ et al., 2001). In general terms, the most important factors that influence ion sorption in soils are the pH of the medium, the valence and the ionic radius of the metal (Gonc¸alves et al., 2000). The low pH (2.0–3.0) required for the Fenton reaction in soil
47
and the strong oxidizing conditions of the medium can modify those interactions and facilitate the dissolution of metals during the remediation process. The metals analyzed in this study were chosen based on their environmental risks. Lead, cadmium and nickel, besides being highly toxic, are also extensively used metals, and therefore are commonly found in the environment (Baird, 1995). Zinc, copper, chromium, manganese and cobalt are common in many types of soil (Selim and Spark, 2001). It was observed that the concentration of some of the metals analyzed increased considerably in the slurry filtrate (Table 2) after 84 h of Fenton reaction (addition of 1.1 mol of H2O2 and 3.0 mmol of Fe2+). Zinc was the most dissolved metal. Even in the control experiment, using only soil and water, 25% of the original amount of Zinc in the soil was dissolved. After the Fenton reaction, this percentage reached 61%, an increase of 36% of dissolution in relation to the control experiment. Copper was the second most dissolved metal in the slurry filtrate. About 12% of the amount of this metal originally present in the soil was dissolved exclusively after the Fenton reaction since the concentration in the control experiment was below the detection limit of the method. In relation to lead and cadmium, no significant dissolution was observed. Although zinc and copper were the most dissolved, chromium, nickel, copper and manganese offer the greatest risks, since their concentrations in the slurry filtrate exceed the maximum allowed value in water (CONAMA, 2007). In a similar experiment however using kaolinite, the displacement of metals was also observed by Monahan et al. (2005) during Fenton treatment, where zinc, copper and cadmium were released to aqueous phase, while lead and nickel were not. The authors attributed the release of metals to the low pH of the Fenton reaction and to the formation of the superoxide radical anion ðO 2 Þ and the hydroperoxide anion ðHO2 ) at high H2O2 concentrations, which may reduce metals. However, in the complex soil matrix, the solubilization of metals may be a consequence of different mechanisms. Clay minerals as well as organic compounds contain adsorption sites such as –OH, –COO and –NH2 where several metals can be adsorbed. The pH reduction to 2.5– 3.0, required for the Fenton process, permits more effective competition of H+ with metals for these adsorption sites, thus contributing to its respective desorption and solubilization. The low pH also increases the solubility of many oxides and hydroxides of metals present in soil (Matos et al., 1996). In this work, although no effort was made to keep the low pH of the slurry, no significant changes were observed reaching a final value of 2.3. Furthermore, the strong oxidizing conditions can change the metal’s oxidation state and increase their solubility, or degrade the OM of the soil displacing them into the solution. Rock et al. (2001) evaluated the effect of applying H2O2, without soluble iron addition, to different types of soils on
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
Table 2 Metal concentrations present in the slurry filtrates and limits imposed by legislation Metals
[Metal] in soil (mg kg1)
[Metal] in SF Controla
[Metal] in SFb
MAVc
Increase with respect to the control (%)
2.2 60.080 0.20 0.43 1.0 66.2 · 103 4.9 0.20
5.0 0.033 0.050 0.025 0.013 0.01 0.5 0.2
36 0 0.9 8.9 12 0 3.1 0.4
1
(mg l ) Zn Pb Cr Ni Cu Cd Mn Co a b c
12 23 64 16 27 66.2 · 103 492 153
0.9 60.080 0.036 60.04 66.4 · 103 66.2 · 103 2.7 60.078
Metal concentration in slurry filtrate (SF) in the control experiment (water and soil). Metal concentration in slurry filtrate in the experiment with the addition of Fenton reagents. Maximum allowed value of metals in water by Brazilian Legislation.
the solubilization of Cr and observed that in all cases the addition of H2O2 increased the solubility of this metal. Some hypotheses were considered for the chromium solubilization due to the addition of H2O2, such as: (i) the capacity of the H2O2 to oxidize species of Cr(III) of low solubility present in soil such as hydroxides, for example, generating a more soluble species such as the HCrO 4 (Eq. (1)) þ 3H2 O2aq þ 2CrðOHÞ3ðsÞ ! 2HCrO 4ðaq:Þ þ 2H þ 4H2 O
ð1Þ
(ii) the formation of peroxide soluble species of Cr(VI) (Eq. (2)) from minerals of low solubility such as chromatite (CaCrO4) 2H2 O2ðaq:Þ þ CaCrO4ðsÞ þ Hþ ! CrðOÞðO2 Þ2 ð OHÞ þ Ca2þ þ 2H2 O
ð2Þ
(iii) the solubilization of the chromium associated with the degradation of OM present in soil, since part of this metal was associated to this matrix. In general, the solubilization of metals present in soil is a complex process that involves reactions of adsorption/ desorption, precipitation/dissolution, complexation and oxidation/reduction of diverse ionic species present in the medium (Sobrinho et al., 1998). Studies on the solubilization of metals during soil remediation using oxidative processes are scarce and mechanisms involved are still not known. However, the increase in metal concentrations to values beyond the maximum allowed by Legislation, limits the application of the Fenton process for the remediation of soils having similar characteristics to those studied in this work. 3.3. Volatilization of DDT and DDE from the soil during Fenton reagent application In this study an experiment was also done where the Fenton reagents were applied directly to the soil (Fig. 1), simulating an in situ remediation process. In this case, the application of Fe2+ and H2O2 to the soil is localized and the concentrated solutions of these reagents enter into direct contact with the soil. This provokes an abrupt
increase in temperature, since the Fenton reaction is highly exothermic. It also provokes gas emissions, mainly in the neighborhood of the application point of the oxidizing solution. In this experiment, the temperature increased from 25 to 55 C in the first 30 min of reaction and started to decrease slowly after 70 min from the start of the experiment. After 48 h of reaction, about 50% of the DDT and 54% of the DDE was degraded. In a control experiment carried out by injecting water into the soil, the DDT and DDE concentrations in the air collected by the polyurethane cartridges above the reactor (Fig. 1) during 48 h were 78 and 27 ng m3, respectively. In the experiment where the Fenton reagents were injected, the concentration of the DDT and the DDE increased to 2472 and 5620 ng m3, respectively. The volatilization of a contaminant present in the soil depends on its properties, on the characteristics of the contaminated soil and on different processes which contribute to the volatilization, such as soil movement, mass transfer due to water evaporation and diffusion of contaminant due to gradient concentration on the surface (Jury et al., 1987; Cousins et al., 1999). The soil–air partition coefficient of a pesticide (KSA) allows us to evaluate the extent of its volatilization by considering the pesticide and soil properties simultaneously. The KSA value can be determined (Eq. (3)), from the values of the soil’s organic carbon fraction (fOC), the bulk density 0 of the soil (qb), dimensionless Henry‘s law constant (H ) and from the octanol–water partition coefficient of the pesticide (KOW) K OW K SA ¼ 0:411f OC qb ð3Þ H0 Using Eq. (3) and the data presented in Tables 1 and 3, the KSA values determined for DDT and DDE in the soil studied were 6.6 · 1010 and 1.0 · 1010 kg m3, respectively. The lower value of KSA for DDE indicates its higher tendency to volatilize from the soil when compared to DDT, as observed by the results of this work. Spencer et al. (1996) have evaluated the volatilization of DDT and its degradation
R.D. Villa et al. / Chemosphere 71 (2008) 43–50 Table 3 Physical–chemical properties of DDT and DDE DDT KOW H0 DHvap (kJ mol1) Vapor pressure 25 C (Pa)
DDE 6
8.13 · 10 3.32 · 104 106 2.50 · 105
3.24 · 106 8.61 · 104 87.2 8.66 · 104
products from contaminated site after 23 years and also observed a higher volatilization of DDE when compared to DDT. Hippelein and McLachlan (2000) reported that KSA is strongly dependent on the temperature and observed that the KSA value for polychlorinated biphenyls decreased almost 3800 times with an increase of temperature from 5 to 60 C. This result suggests that the observed drastic increase in the DDT and DDE concentration in the sampled air after the Fenton reaction, obtained in the present work, could be a consequence of the sharp increase of the temperature.When considering only one property of DDT and DDE such as vapor pressure, the effect of temperature on the volatilization process becomes more evident. The Clausius–Clapeyron equation (Eq. (4)) correlates the change of vapor pressure with the temperature: p DH Vap 1 1 ln 2 ¼ ð4Þ T1 T2 p1 R where p1 and p2 are vapor pressure for DDT at temperatures T1 and T2, respectively, DHvap is the vaporization enthalpy and R is the gas constant (8.314 J K1 mol1). The substitution of the values of DHvap and p1 (Table 3), and those of temperatures T1 (25 C) and T2 (55 C) measured before and after the experiment, respectively, and that of R in Eq. (4), results in p2. The value of p2 was 1.24 mPa for DDT and 21.7 mPa for DDE. This corresponds to a vapor pressure 50 times higher than the value of p1 for DDT and 25 times higher for DDE, which explains the increase in volatilization. It is also observed that p2 for DDE is about 20 times higher than p2 for DDT. These results demonstrate the effect of the temperature on the volatilization of contaminants present in the soil. Furthermore, the intense emission of gases, such as CO, CO2 and O2, generated from the oxidation of the organic matter and the H2O2 decomposition, probably contributed to the volatilization of DDT and DDE. 4. Conclusions The Fenton process efficiently remediated soil contaminated with DDT and diesel, achieving 75% and 80% degradation, respectively, in slurry systems. However, the oxidizing conditions of the Fenton process contributes to the dissolution of the OM originally present in the soil and to the increase of water solubility of hydrophobic compounds such as DDT achieving 50 lg l1, far above the
49
limits imposed for water. The oxidizing medium and the low pH value (2.0–3.0) also favor the dissolution of metals originally present in the soil. Copper was the most dissolved metal by the Fenton process, but also chromium, nickel and manganese may cause environmental risk since their final concentrations in the slurry filtrate exceeded the legal limits for water. The application of the Fenton process, directly in the soil, resulted in an intense volatilization of DDT and DDE, which can contribute to the dissipation of the contaminant into the environment and offer great occupational risks to the personal involved in the remediation process, especially in the case of DDT contamination, whose main contamination route in mammals is via inhalation (WHO, 1989). The application of the Fenton process to soils must be carefully evaluated beforehand, taking into account not only the degradation of target compounds, but also the potential collateral effects of its application. Acknowledgements The authors thank CAPES and FAPESP (05/00172-0) for the scholarship awarded to R.D. Villa and A.G. Trovo´, respectively. The authors also thank Prof. Dr. M.R. de Marchi for making the CG-FID equipment available, Prof. Dr. M.S. Crespi for the TG analysis, Prof. Dr. M.L. Ribeiro for making the CG-ECD equipment available, Prof. Dr. J.C. Rocha for the metal determination and Mr. David Nicholson for revising the manuscript. References APHA/AWWA/WEF, 1989. Standard methods for the examination of water and wastewater, 17th ed., American Publish Health Association, Washington, DC. Baird, C., 1995. Environmental Chemistry. W.H. Freeman and Company, New York. Brady, N.C., Buckman, H.O., 1983. Nature and Properties of Soils, 7th ed., Macmillan, London. CONAMA – Conselho Nacional do Meio Ambiente, 2007. (accessed 15.05.07). Cousins, I.T., Beck, A.J., Jones, K.C., 1999. A review of the processes involved in the exchange of semi-volatile organic compounds (SVOC) across the air–soil interface. Sci. Total Environ. 228, 5–24. Ding, J.Y., Wu, S.C., 1995. Partition-coefficients of organochlorine pesticides on soil and on the dissolved organic-matter in water. Chemosphere 30, 2259–2266. EMBRAPA – Empresa Brasileira de Pesquisa Agropecua´ria, 1997. Manual de Me´todos de Ana´lise de Solo, Rio de Janeiro. EPA 8015B, 1996. Nonhalogenated Organics Using GC/FID. EPA, Revision 2, 28p. Ferguson, S.H., Woinarski, A.Z., Snape, I., Morris, C.E., Revill, A.T., 2004. A field trial of in situ chemical oxidation to remediate long-term diesel contaminated Antarctic soil. Cold Reg. Sci. Technol. 40, 47–60. Gonc¸alves, A.C., Luchese, E.B., Lenzi, E., 2000. Evaluation of phytoavailability of the cadmium, lead and chromium in soybean cultivated in the latossolo vermelho escuro, treated with commercial fertilizers. Quim. Nova 23, 173–177. Hippelein, M., Mclachlan, M.S., 2000. Soil/air partitioning of semivolatile organic compounds. Part II: Influence of temperature and relative humidity. Environ. Sci. Technol. 34, 3521–3526.
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Jury, W.A., Winer, A.M., Spencer, W.F., Focht, D.D., 1987. Transport and transformation of organic chemicals in a soil/air/water ecosystem. Rev. Environ. Contam. Toxicol. 99, 119–164. Keith, L.H., 1996. Compilation of EPA’s Sampling and Analysis Methods. CRC Press LLC, Boca Raton. Khan, F.L., Husain, T., Hejazi, R., 2004. An overview and analysis of site remediation technologies. J. Environ. Manage. 71, 95–122. Kong, S., Watts, R.J., Choi, J., 1998. Treatment of petroleum contaminated soils using iron mineral catalyzed hydrogen peroxide. Chemosphere 37, 1473–1482. Liou, M.J., Lu, M.C., Chen, J.N., 2003. Oxidation of explosives by Fenton and photo-Fenton processes. Water Res. 37, 3172–3179. Matos, A.T., Fontes, M.P.F., Jorda˜o, C.P., Costa, L.M., 1996. Heavy metals mobility and retention forms in a Brazilian oxisol. Rev. Bras. Cieˆncia Solo 20, 379–386. Monahan, M.J., Teel, A.L., Watts, R.J., 2005. Displacement of five metals sorbed on kaolinite during treatment with modified Fenton’s reagent. Water Res. 39, 2955–2963. Pignatello, J.J., Oliveros, E., MacKay, A., 2006. Advanced oxidation processes for organic contaminant destruction based on the Fenton reaction and related chemistry. Environ. Sci. Technol. 36, 1–84. Rocha, J.C., Rosa, A.H., 2003. Substaˆncias Hu´micas Aqua´ticas: Interac¸a˜o Com Espe´cies Meta´licas. Editora UNESP, Sa˜o Paulo. Rock, M.L., James, B.R., Helz, G.R., 2001. Hydrogen peroxide effects on chromium state and solubility in four diverse, chromium-enriched soils. Environ. Sci. Technol. 35, 4054–4059. Selim, H.M., Spark, D.L., 2001. Heavy Metals Release in Soil. CRC Press LLC Corporate, Boca Raton. Sobrinho, N.M.B.A., Velloso, A.C.X., Costa, L.M., Oliveira, C., 1998. Heavy metal mobility in soil columns treated with an acid iron smelter residue. Rev. Bras. Cieˆncia Solo 22, 345–353. Sodre´, F.F., Lenzi, E., Costa, A.C.S., 2001. Applicability of adsorption models to the study of copper behaviour in clayey soils. Quim. Nova 24, 324–330.
Solomons, G., Fryhle, G., 2000. Organic Chemistry. John Wiley & Sons, Inc., New York, USA, pp. 820. Spark, K.M., Swift, R.S., 2002. Effect of soil composition and dissolved organic mater on pesticide sorption. Sci. Total Environ. 298, 147–161. Spencer, W.F., Singh, G., Taylor, C.D., Lemert, R.A., Cliath, M.M., Farmer, W.J., 1996. DDT persistence and volatility as affected by management practices after 23 years. J. Environ. Qual. 25, 815–821. Suguio, K., 1973. Introduc¸a˜o a` Sedimentologia. Ed. da Universidade de Sa˜o Paulo, Sa˜o Paulo. VanLoon, G.W., Duffy, S.J., 2000. Environmental Chemistry. Oxford University Press, New York. Villa, R.D., Dores, E.F.G.C., Carbo, L., Cunha, M.L.F., 2006. Dissipation of DDT in a heavily contaminated soil in Mato Grosso, Brazil. Chemosphere 64, 549–554. Villa, R.D., Nogueira, R.F.P., 2006. Oxidation of p, p 0 DDT and p, p 0 DDE in highly and long-term contaminated soil using Fenton reaction in a slurry system. Sci. Total Environ. 371, 11–18. Wardmann, P., 1989. Reduction potentials of one-electron couples involving free radicals in aqueous solution. J. Phys. Chem. Ref. Data 18, 1637–1723. Watts, R.J., Udell, M.D., Rauch, P.A., 1990. Treatment of pentachlorophenol-contaminated soils using Fenton’s reagent. Hazard. Waste Hazard. Mater. 7, 335–345. Watts, R.J., Kong, S., Dippre, M., Barnes, W.T., 1994. Oxidation of sorbed hexachlorobenzene in soils using catalyzed hydrogen peroxide. J. Hazard. Mat. 39, 33–47. WHO (World Health Organization), 1989. DDT and its derivatives, Environmental Health Criteria, Geneva. Worrall, F., Fernandez-Perez, M., Johnson, A.C., Flores-Cesperedes, F., Gonsalez-Pradas, E., 2001. Limitations on the role of incorporated organic matter in reducing pesticide leaching. J. Contam. Hydrol. 49, 241–262.
Chemosphere 71 (2008) 43–50 www.elsevier.com/locate/chemosphere
Environmental implications of soil remediation using the Fenton process Ricardo D. Villa, Alam G. Trovo´, Raquel F. Pupo Nogueira
*
UNESP – Sa˜o Paulo State University, Institute of Chemistry of Araraquara, R. Prof. Francisco Degni s/n, P.O. Box 355, 14801-970 Araraquara, SP, Brazil Received 17 September 2007; received in revised form 22 October 2007; accepted 23 October 2007 Available online 18 December 2007
Abstract This work evaluates some collateral effects caused by the application of the Fenton process to 1,1-bis(4-chlorophenyl)-2,2,2-trichloroethane (DDT) and diesel degradation in soil. While about 80% of the diesel and 75% of the DDT present in the soil were degraded in a slurry system, the dissolved organic carbon (DOC) in the slurry filtrate increased from 80 to 880 mg l1 after 64 h of reaction and the DDT concentration increased from 12 to 50 lg l1. Experiments of diesel degradation conducted on silica evidenced that soluble compounds were also formed during diesel oxidation. Furthermore, significant increase in metal concentrations was also observed in the slurry filtrate after the Fenton treatment when compared to the control experiment leading to excessive concentrations of Cr, Ni, Cu and Mn according to the limits imposed for water. Moreover, 80% of the organic matter naturally present in the soil was degraded and a drastic volatilization of DDT and 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene was observed. Despite the high percentages of diesel and DDT degradation in soil, the potential overall benefits of its application must be evaluated beforehand taking into account the metal and target compounds dissolution and the volatilization of contaminants when the process is applied. 2007 Elsevier Ltd. All rights reserved. Keywords: Air contamination; DDT; Diesel; In situ chemical oxidation; Metals; Organic matter
1. Introduction The increasing contamination of soil and water by organic compounds has motivated the development and application of several remediation processes (Khan et al., 2004) such as the advanced oxidation processes (AOPs). These processes are receiving great emphasis because of their ability to rapidly oxidize refractory organic contaminants (Pignatello et al., 2006). The Fenton process is one of the most common AOPs and it has been the subject of considerable interest for the remediation of contaminated soils. The Fenton reaction consists in the catalytic decomposition of hydrogen
*
Corresponding author. Tel.: +55 16 3301 6606; fax: +55 16 3301 6692. E-mail addresses: [email protected] (R.D. Villa), alamtrovo@ hotmail.com (A.G. Trovo´), [email protected] (R.F.P. Nogueira). 0045-6535/$ - see front matter 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.10.043
peroxide (H2O2) due to soluble iron(II) in acidic medium, generating hydroxyl radicals (OH). The hydroxyl radical (OH) is a strong oxidizing agent o (E = 2.73 V versus NHE, Wardmann, 1989) capable of non-selectively oxidizing a variety of organic contaminants. One of the main advantages of Fenton processes is the possibility of in situ remediation, eliminating the need for excavating contaminated soils and thus allowing the rapid treatment of refractory contaminants when compared to other remediation processes (Monahan et al., 2005). The use of Fenton reactions for the treatment of contaminated soils was initially investigated by Watts et al. (1990), who was the first to observe pentachlorophenol mineralization. Subsequently, several other works appeared which emphasized the efficiency of Fenton processes for the remediation of soils contaminated with – among other organic compounds – pesticides (Watts et al., 1994; Villa and Nogueira, 2006), fuel (Kong et al., 1998) and explosives
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
(Liou et al., 2003). However, few works have evaluated the collateral effects of soil remediation using Fenton processes. Therefore, there is little information about other consequences of the oxidation process such as soil organic matter (OM) oxidation, metal and target compounds dissolution, the effect on microorganisms and the volatilization of contaminants caused by the strong exothermic characteristic of the Fenton reaction. The aim of the present work was to evaluate some of those collateral effects – namely OM degradation, the dissolution of metals and target compounds, and the volatilization of contaminants in the remediation of soils contaminated with 1,1-bis(4-chlorophenyl)-2,2,2-trichloroethane (DDT), 2,2-bis(4-chlorophenyl)-1,1-dichloroethylene (DDE) and diesel using the Fenton reaction. 2. Materials and methods 2.1. Chemicals Standard DDT and DDE, at 99% purity were purchased from Supelco. The diesel used was a commercial product. Isooctane (Mallinckrodt) and n-hexane (Baker) were used to prepare the DDT and diesel standard solutions, respectively. A neutral and deactivated 70–290 alumina mesh (Vetec) was used in the DDT extraction process. Pesticide-grade solvents n-hexane, acetone and dichloromethane (Tedia) were used for the extraction of both DDT and diesel. H2O2 30% (w/v) from Synth was used in the degradation experiments, while Na2SO4 (Baker) was used as a desiccation agent. HNO3 (Synth) was used in the decomposition of the samples for metal determination. The solutions of Fe2+ were prepared by dissolving FeSO4 Æ 7H2O (Carlo Erba) in 0.1 M H2SO4 (Synth). Analytical grade metals solutions (Zn, Pb, Cr, Ni, Cu, Cd, Mn and Co) (MV Laboratories, Inc.) were used to construct analytical curves for the determination of these metals. All the glassware was cleaned with Extran MA-01 10% (v/v) from Merck. 2.2. Sampling and soil contamination The soil used in this work was collected from a former pesticide warehouse in the state of Mato Grosso, Brazil. That soil was seriously contaminated with DDT as a consequence of inadequate storage of this substance after its use was restricted by law (Villa et al., 2006). As a consequence of the natural degradation of DDT, the soil has also contamination by DDE since this compound is one of the main degradation products of DDT. After sampling, the soils were passed through a 3.0 mm sieve and dried in air at 25–30 C for 48 h. The soil used to study diesel degradation was collected near the pesticide warehouse (about 30 m). This soil was not contaminated and presented the same characteristics of the soil used in the DDT degradation experiments. The diesel spiked soil was prepared by adding 200 ml of 50 g l1 of n-hexane diesel solution to
2.0 kg of soil. The spiked soil was then vigorously homogenized and left standing for 24 h to eliminate the solvent, which completeness was confirmed by weight loss. 2.3. Experimental conditions of degradation For the diesel degradation experiments, a slurry was prepared by directly mixing 500 ml of 12 mM FeSO4 aqueous solution and 150 g of soil or silica contaminated with diesel at initial concentration of 5.0 mg g1. For DDT degradation, 500 ml of 6.0 mM of FeSO4 aqueous solution and 150 g of contaminated soil (1.6 mg g1) were used. A total volume of 160 ml of 7.0 and 10 M H2O2 was pumped into the slurry for the DDT and diesel degradation, respectively. In both cases, the H2O2 was added using a peristaltic pump (Ismatec model 78017-10) programmed to deliver 1.1 ml of H2O2 solution every 20 min for 12 h. After this period of time, the additions were interrupted for 12 h to allow for the consumption of the H2O2 and then restarted. All the experiments were carried out at ambient temperature and with mechanical stirring at 110 rpm. The initial pH of the slurries was adjusted to between 2.5 and 3.0 by the addition of 3 M H2SO4 solution before starting the experiments, i.e. the optimum pH value for the Fenton reaction in soil (Watts et al., 1990). No further pH adjustment was made during the experiments. Metal dissolution was evaluated by comparing the metal concentrations in the filtrate of the diesel-degradationexperiment slurry (under the same oxidation conditions as described previously) to those of the filtrate of a slurry composed of only 150 g of the diesel-contaminated soil and 500 ml of distilled water. The volatilization of the DDT and DDE was evaluated in an experiment where the reagents were directly pumped to the dry soil. In this experiment, 650 ml of 20 mM Fe and 650 ml of 4.0 M H2O2 solution were merged immediately before its addition to 2.0 kg of soil contaminated with DDT (1.6 mg g1). In the first 30 min of experiment, the reagent flow rate injection was 15 ml min1 with the aim of humidifying the soil. After this time, the pump was programmed to start injecting again every 3 h at a flow rate of 10 ml min1 for a period of 5 min in order to keep the soil humidified. This experiment was carried out for 48 h. The air was collected at a height of 25 cm above the surface of the contaminated soil using a polyurethane cartridge as the absorber (5 cm diameter, 4 cm long, having a mass of 2.30 g) through which the air was pumped at a flow rate of 4 ml min1. A schematic representation of the reactor and the air sampling is shown in Fig. 1. In order to evaluate the extent of the DDT and DDE volatilization, one sample of air was taken during the soil remediation. The air was firstly pumped through the polyurethane cartridge for 48 h after the injection of the Fenton reagents at a flow rate of 4.0 l min1. The concentration of DDT and DDE adsorbed on the polyurethane were determined after soxlet extraction and compared to the concentration obtained in a separate experiment without the
R.D. Villa et al. / Chemosphere 71 (2008) 43–50
45
Polyurethane cartidge
Air pumping (4.0 L min-1). 18 cm 15 cm H2O2
Fe 2 + pH = 2.5
1.0 cm Contaminated soil 23 cm
Net for soil retention
Fig. 1. Schematic representation of the non-slurry system and air sampling device.
addition of the Fenton reagents for the same time and flow rate. 2.4. Chemical analysis DDT and DDE extraction from the soil was carried out using the method developed by Villa et al. (2006). In summary, after drying, grinding and homogenizing the soil, triplicate samples of 0.50 g of soil were mixed with 1.0 g of deactivated alumina. This mixture was transferred to glass columns (1.0 cm internal diameter · 44 cm length), containing 2.0 g of neutral alumina, where the pesticides were eluted with 210 ml of n-hexane:dichloromethane (7:3) solution. The extracts were concentrated and then dissolved in 6.0 ml n-hexane. The DDT and DDE were quantified using gas chromatography with an electron capture detector using a Varian 3300 chromatograph and a DB-5 capillary column (0.32 mm · 30 m). The analyses were carried out under the following conditions: injector temperature 280 C; initial furnace temperature 110 C; final furnace temperature 250 C; a heating rate of 10 C min1; detector temperature 330 C; N2 as carrier gas at a flow rate of 5.0 ml min1. The DDT present in the aqueous filtrates was extracted using liquid–liquid extraction and 50 ml of the sample (filtrate) with 10 ml of n-hexane/dichloromethane solution (85:15) (APHA, 1989). All extractions were carried out in triplicate and the DDT and DDE concentrations were determined under the same conditions as those for the soil extracts. The extraction of DDT and DDE from the polyurethane cartridges was carried out in a Soxlet apparatus for 24 h using 200 ml of hexane:acetone (1:1) as extracting solution. After extraction, the extracts were concentrated and dissolved in 5.0 ml hexane. DDT and DDE were then determined using gas chromatography under the same conditions as for extracts from soil and water. Diesel extraction from the soil was carried out by adding 7.0 ml of n-hexane–dichloromethane solution (1:1) and 1.0 g of sodium sulfate to 3.0 g of spiked soil in an airtight
sealed glass tube (2.0 cm internal diameter and 15 cm long). The mixture was then mechanically stirred at 240 rpm for 2 h. After this time the mixture was centrifuged and the extract collected. This extraction procedure was repeated two more times. The soil extracts were analyzed by gas chromatography using a Shimadzu 14B chromatograph, a DB-5 capillary column (0.32 mm · 30 m) and a flame ionization detector. The following chromatographic conditions were used: injector temperature 280 C; initial furnace temperature 45 C; final furnace temperature 250 C; heating rate 12 C min1; detector temperature 330 C; H2 as carrier gas at a flow rate of 10 ml min1. The diesel concentration was determined in terms of total petroleum hydrocarbons – TPH (EPA 8015B, 1996). Prior to determination of metal content, the soil samples were prepared according to the standard digestion method 6010 (Keith, 1996). The determination of metals in the slurry filtrate was done directly after filtration through a 0.45 lm membrane. In all cases, the quantification was done by flame atomic absorption spectrometry using an Analyst 300 spectrometer (Perkin Elmer). All determinations were carried out in triplicate followed by a blank check. The dissolved organic carbon (DOC) in the slurry filtrates was determined using a Total Organic Carbon analyzer (TOC-5000A Shimadzu) after filtration of the slurry through filter paper and then through a 0.45 lm membrane. The soil texture was determined using the pipette method (Suguio, 1973). The density of the soil was determined using a calibrated tube and the soil pH was determined in water (EMBRAPA, 1997). The soil OM content was determined by thermogravimetric analysis in a Simultaneous Thermal Analysis Module, model number SDT2960 (TA Instruments). 3. Results and discussion The soil used in this work is slightly acidic and has an OM content of 7.5% as shown in Table 1. The metal
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
A
800
-1
0.8 600 0.4 400 0.2
-1
DDT and DDE (mg g )
1.2
200
3.1. Organic matter degradation and contaminant dissolution
0.0
0
12
24
36
48
60
0 72
Time (h)
1000
B 5
800
-1
600 3 400
2
-1
Diesel (mg g )
4
200
1
0
DOC (mg L )
After 64 h of degradation process, the DDT concentration in the soil decreased of 75%, while the DOC in the slurry filtrate increased from 80 to 880 mg l1 (Fig. 2A). The concentration of the DDE decreased of only 40%. However, considering that DDE is the main degradation product of DDT, it can be produced and degraded simultaneously during the experiments (Villa and Nogueira, 2006). The OM present in the soil consists of a complex mixture at several stages of decomposition, resulting from chemical and/or biological degradation of vegetal and animal residues (vanLoon and Duffy, 2000). The greater the OM oxidation stage, the greater the tendency for its solubilization (Rocha and Rosa, 2003). The oxidative conditions of the reaction medium contributed to further oxidation of the OM and consequently to an increase of the DOC concentration in the slurry filtrate. It was also observed that the DDT concentration in the slurry filtrate increased from 12 to 50 lg l1 after 64 h of reaction, probably as a consequence of soil OM dissolution. There is strong evidence that the soluble OM acts as a surfactant, increasing the solubility of hydrophobic compounds, such as DDT (Ding and Wu, 1995; Worrall et al., 2001; Spark and Swift, 2002). The initial amount of DDT in 150 g of soil was 240 mg. Considering that the final concentration in 500 ml of the slurry filtrate achieved 50 lg l1 after the Fenton treatment, it can be calculated that less than 0.01% of the DDT present in the soil was dissolved. However, the concentration of 50 lg l1 in water is 50 times higher than the maximum concentration allowed by the Brazilian Ministry of the Environment (CONAMA, 2007) for surface water destined for irrigation or recreation, which is 1.0 lg l1. The solubilization of contaminant as seen here
1000 1.6
DOC (mg L )
concentrations in the studied soil are within the range reported in literature for all metals analyzed (vanLoon and Duffy, 2000). Interpolating the sand, silt and clay percentages (Table 1) into a textural triangle cited by Brady and Buckman (1983), it was observed that this soil can be classified as loam, where silt and sand are predominant sized fractions. Loam soils present a small surface area when compared to clay soils, and therefore they retain less water, are easily drained and generally have a low plasticity (Brady and Buckman, 1983). These characteristics work together to facilitate the mobility of many contaminants through the soil, mainly the water soluble ones.
0
12 24 36 48 60 72 84
0
Time (h) Fig. 2. (A) DDT and DDE degradation in soil and increase of DOC concentration in the slurry filtrate. Experimental conditions: 6.0 mM FeSO4; additions of 1.1 ml of 7.0 M H2O2 every 20 min (145 additions, total amount: 38.1 g). –j– DDT; –m– DDE; –d– DOC; (B) Diesel degradation in soil and increase of DOC in the slurry filtrate. Experimental conditions: 12 mM FeSO4; additions of 1.1 ml of 10 M H2O2 solution every 20 min (145 additions, total amount: 54.4 g). –– Diesel; –d– DOC; –$– DOC from non-contaminated soil; –D– DOC without Fenton reagents.
can be particularly critical in cases of high toxicity, mainly when in situ remediation is undertaken, since it facilitates the mobility of contaminants through the soil, offering contamination risks to surface water and underground
Table 1 Characteristics of the soil used in this work pH (H2O) 5.7
qb
OM
(g cm3)
(%)
1.5
7.5
fOC
Sand
Silt
Clay
Textural classification
Zn
Cr
Ni
Cu
Cd
Mn
16
27
66.2 · 103
492
(mg kg1) 4.35
41
47
12
Loam
12
64
R.D. Villa et al. / Chemosphere 71 (2008) 43–50
water reservoirs. Furthermore, the amount of contaminants dissolved in the filtrate may be higher for a higher volume of water used for the slurry. A similar behavior was observed for diesel oxidation when compared to DDT degradation. When 80% of the diesel was degraded, the DOC concentration in the slurry filtrate reached 870 mg l1 (Fig. 2B). In the experiment without any addition of H2O2, the DOC remained very low over an 84 h period, evidencing that the OM dissolution was a result of the strong oxidizing conditions of the medium. In the absence of diesel, under the same conditions of iron and H2O2 concentration, the DOC concentration in the slurry filtrate reached 715 mg l1. The difference of 155 mg l1 of DOC in the filtrate of diesel-contaminated and non-contaminated soil can be related to the components of the diesel and its degradation products. Furthermore, in a separate experiment where soil was replaced by silica (the only source of OM was the diesel), the DOC reached 160 mg l1 after 84 h of experiment, reinforcing the hypotheses of the solubilization of the diesel compounds. It can be stated that although most of the DOC concentration in the slurry filtrate originates from the natural OM of the soil, diesel is also a source of DOC due to the formation of soluble degradation products. Considering that diesel is a complex mixture formed by aromatic compounds and other unsaturated hydrocarbons that can be easily oxidized in the presence of H2O2, generation of several soluble compounds, such as carboxylic acid, aldehydes and ketones is possible (Solomons and Fryhle, 2000). A detailed knowledge of the toxicity of these compounds is of great importance to the safe application of the Fenton processes in soil remediation. During diesel degradation, 80% of the OM initially present in the soil was degraded, evidencing the aggressiveness of the Fenton remediation processes on the soil. Besides contributing to soil deterioration, the Fenton process probably contributes to its sterilization. Ferguson et al. (2004) observed the elimination of great part of the microorganisms initially presents in soil after the application of the Fenton process in the remediation of soil contaminated with diesel. They had attributed the death of the microorganisms to the strong oxidizing conditions of the medium, the change in pH and the temperature increase. 3.2. Dissolution of metals during oxidation processes In soil, metals can: (i) participate in dissolution and in precipitation reactions; (ii) be electrostatically absorbed in the exchange sites; (iii) be incorporated into inorganicphase surfaces (specific adsorption); and/or (iv) be associated to the organic fraction (Sodre´ et al., 2001). In general terms, the most important factors that influence ion sorption in soils are the pH of the medium, the valence and the ionic radius of the metal (Gonc¸alves et al., 2000). The low pH (2.0–3.0) required for the Fenton reaction in soil
47
and the strong oxidizing conditions of the medium can modify those interactions and facilitate the dissolution of metals during the remediation process. The metals analyzed in this study were chosen based on their environmental risks. Lead, cadmium and nickel, besides being highly toxic, are also extensively used metals, and therefore are commonly found in the environment (Baird, 1995). Zinc, copper, chromium, manganese and cobalt are common in many types of soil (Selim and Spark, 2001). It was observed that the concentration of some of the metals analyzed increased considerably in the slurry filtrate (Table 2) after 84 h of Fenton reaction (addition of 1.1 mol of H2O2 and 3.0 mmol of Fe2+). Zinc was the most dissolved metal. Even in the control experiment, using only soil and water, 25% of the original amount of Zinc in the soil was dissolved. After the Fenton reaction, this percentage reached 61%, an increase of 36% of dissolution in relation to the control experiment. Copper was the second most dissolved metal in the slurry filtrate. About 12% of the amount of this metal originally present in the soil was dissolved exclusively after the Fenton reaction since the concentration in the control experiment was below the detection limit of the method. In relation to lead and cadmium, no significant dissolution was observed. Although zinc and copper were the most dissolved, chromium, nickel, copper and manganese offer the greatest risks, since their concentrations in the slurry filtrate exceed the maximum allowed value in water (CONAMA, 2007). In a similar experiment however using kaolinite, the displacement of metals was also observed by Monahan et al. (2005) during Fenton treatment, where zinc, copper and cadmium were released to aqueous phase, while lead and nickel were not. The authors attributed the release of metals to the low pH of the Fenton reaction and to the formation of the superoxide radical anion ðO 2 Þ and the hydroperoxide anion ðHO2 ) at high H2O2 concentrations, which may reduce metals. However, in the complex soil matrix, the solubilization of metals may be a consequence of different mechanisms. Clay minerals as well as organic compounds contain adsorption sites such as –OH, –COO and –NH2 where several metals can be adsorbed. The pH reduction to 2.5– 3.0, required for the Fenton process, permits more effective competition of H+ with metals for these adsorption sites, thus contributing to its respective desorption and solubilization. The low pH also increases the solubility of many oxides and hydroxides of metals present in soil (Matos et al., 1996). In this work, although no effort was made to keep the low pH of the slurry, no significant changes were observed reaching a final value of 2.3. Furthermore, the strong oxidizing conditions can change the metal’s oxidation state and increase their solubility, or degrade the OM of the soil displacing them into the solution. Rock et al. (2001) evaluated the effect of applying H2O2, without soluble iron addition, to different types of soils on
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R.D. Villa et al. / Chemosphere 71 (2008) 43–50
Table 2 Metal concentrations present in the slurry filtrates and limits imposed by legislation Metals
[Metal] in soil (mg kg1)
[Metal] in SF Controla
[Metal] in SFb
MAVc
Increase with respect to the control (%)
2.2 60.080 0.20 0.43 1.0 66.2 · 103 4.9 0.20
5.0 0.033 0.050 0.025 0.013 0.01 0.5 0.2
36 0 0.9 8.9 12 0 3.1 0.4
1
(mg l ) Zn Pb Cr Ni Cu Cd Mn Co a b c
12 23 64 16 27 66.2 · 103 492 153
0.9 60.080 0.036 60.04 66.4 · 103 66.2 · 103 2.7 60.078
Metal concentration in slurry filtrate (SF) in the control experiment (water and soil). Metal concentration in slurry filtrate in the experiment with the addition of Fenton reagents. Maximum allowed value of metals in water by Brazilian Legislation.
the solubilization of Cr and observed that in all cases the addition of H2O2 increased the solubility of this metal. Some hypotheses were considered for the chromium solubilization due to the addition of H2O2, such as: (i) the capacity of the H2O2 to oxidize species of Cr(III) of low solubility present in soil such as hydroxides, for example, generating a more soluble species such as the HCrO 4 (Eq. (1)) þ 3H2 O2aq þ 2CrðOHÞ3ðsÞ ! 2HCrO 4ðaq:Þ þ 2H þ 4H2 O
ð1Þ
(ii) the formation of peroxide soluble species of Cr(VI) (Eq. (2)) from minerals of low solubility such as chromatite (CaCrO4) 2H2 O2ðaq:Þ þ CaCrO4ðsÞ þ Hþ ! CrðOÞðO2 Þ2 ð OHÞ þ Ca2þ þ 2H2 O
ð2Þ
(iii) the solubilization of the chromium associated with the degradation of OM present in soil, since part of this metal was associated to this matrix. In general, the solubilization of metals present in soil is a complex process that involves reactions of adsorption/ desorption, precipitation/dissolution, complexation and oxidation/reduction of diverse ionic species present in the medium (Sobrinho et al., 1998). Studies on the solubilization of metals during soil remediation using oxidative processes are scarce and mechanisms involved are still not known. However, the increase in metal concentrations to values beyond the maximum allowed by Legislation, limits the application of the Fenton process for the remediation of soils having similar characteristics to those studied in this work. 3.3. Volatilization of DDT and DDE from the soil during Fenton reagent application In this study an experiment was also done where the Fenton reagents were applied directly to the soil (Fig. 1), simulating an in situ remediation process. In this case, the application of Fe2+ and H2O2 to the soil is localized and the concentrated solutions of these reagents enter into direct contact with the soil. This provokes an abrupt
increase in temperature, since the Fenton reaction is highly exothermic. It also provokes gas emissions, mainly in the neighborhood of the application point of the oxidizing solution. In this experiment, the temperature increased from 25 to 55 C in the first 30 min of reaction and started to decrease slowly after 70 min from the start of the experiment. After 48 h of reaction, about 50% of the DDT and 54% of the DDE was degraded. In a control experiment carried out by injecting water into the soil, the DDT and DDE concentrations in the air collected by the polyurethane cartridges above the reactor (Fig. 1) during 48 h were 78 and 27 ng m3, respectively. In the experiment where the Fenton reagents were injected, the concentration of the DDT and the DDE increased to 2472 and 5620 ng m3, respectively. The volatilization of a contaminant present in the soil depends on its properties, on the characteristics of the contaminated soil and on different processes which contribute to the volatilization, such as soil movement, mass transfer due to water evaporation and diffusion of contaminant due to gradient concentration on the surface (Jury et al., 1987; Cousins et al., 1999). The soil–air partition coefficient of a pesticide (KSA) allows us to evaluate the extent of its volatilization by considering the pesticide and soil properties simultaneously. The KSA value can be determined (Eq. (3)), from the values of the soil’s organic carbon fraction (fOC), the bulk density 0 of the soil (qb), dimensionless Henry‘s law constant (H ) and from the octanol–water partition coefficient of the pesticide (KOW) K OW K SA ¼ 0:411f OC qb ð3Þ H0 Using Eq. (3) and the data presented in Tables 1 and 3, the KSA values determined for DDT and DDE in the soil studied were 6.6 · 1010 and 1.0 · 1010 kg m3, respectively. The lower value of KSA for DDE indicates its higher tendency to volatilize from the soil when compared to DDT, as observed by the results of this work. Spencer et al. (1996) have evaluated the volatilization of DDT and its degradation
R.D. Villa et al. / Chemosphere 71 (2008) 43–50 Table 3 Physical–chemical properties of DDT and DDE DDT KOW H0 DHvap (kJ mol1) Vapor pressure 25 C (Pa)
DDE 6
8.13 · 10 3.32 · 104 106 2.50 · 105
3.24 · 106 8.61 · 104 87.2 8.66 · 104
products from contaminated site after 23 years and also observed a higher volatilization of DDE when compared to DDT. Hippelein and McLachlan (2000) reported that KSA is strongly dependent on the temperature and observed that the KSA value for polychlorinated biphenyls decreased almost 3800 times with an increase of temperature from 5 to 60 C. This result suggests that the observed drastic increase in the DDT and DDE concentration in the sampled air after the Fenton reaction, obtained in the present work, could be a consequence of the sharp increase of the temperature.When considering only one property of DDT and DDE such as vapor pressure, the effect of temperature on the volatilization process becomes more evident. The Clausius–Clapeyron equation (Eq. (4)) correlates the change of vapor pressure with the temperature: p DH Vap 1 1 ln 2 ¼ ð4Þ T1 T2 p1 R where p1 and p2 are vapor pressure for DDT at temperatures T1 and T2, respectively, DHvap is the vaporization enthalpy and R is the gas constant (8.314 J K1 mol1). The substitution of the values of DHvap and p1 (Table 3), and those of temperatures T1 (25 C) and T2 (55 C) measured before and after the experiment, respectively, and that of R in Eq. (4), results in p2. The value of p2 was 1.24 mPa for DDT and 21.7 mPa for DDE. This corresponds to a vapor pressure 50 times higher than the value of p1 for DDT and 25 times higher for DDE, which explains the increase in volatilization. It is also observed that p2 for DDE is about 20 times higher than p2 for DDT. These results demonstrate the effect of the temperature on the volatilization of contaminants present in the soil. Furthermore, the intense emission of gases, such as CO, CO2 and O2, generated from the oxidation of the organic matter and the H2O2 decomposition, probably contributed to the volatilization of DDT and DDE. 4. Conclusions The Fenton process efficiently remediated soil contaminated with DDT and diesel, achieving 75% and 80% degradation, respectively, in slurry systems. However, the oxidizing conditions of the Fenton process contributes to the dissolution of the OM originally present in the soil and to the increase of water solubility of hydrophobic compounds such as DDT achieving 50 lg l1, far above the
49
limits imposed for water. The oxidizing medium and the low pH value (2.0–3.0) also favor the dissolution of metals originally present in the soil. Copper was the most dissolved metal by the Fenton process, but also chromium, nickel and manganese may cause environmental risk since their final concentrations in the slurry filtrate exceeded the legal limits for water. The application of the Fenton process, directly in the soil, resulted in an intense volatilization of DDT and DDE, which can contribute to the dissipation of the contaminant into the environment and offer great occupational risks to the personal involved in the remediation process, especially in the case of DDT contamination, whose main contamination route in mammals is via inhalation (WHO, 1989). The application of the Fenton process to soils must be carefully evaluated beforehand, taking into account not only the degradation of target compounds, but also the potential collateral effects of its application. Acknowledgements The authors thank CAPES and FAPESP (05/00172-0) for the scholarship awarded to R.D. Villa and A.G. Trovo´, respectively. The authors also thank Prof. Dr. M.R. de Marchi for making the CG-FID equipment available, Prof. Dr. M.S. Crespi for the TG analysis, Prof. Dr. M.L. Ribeiro for making the CG-ECD equipment available, Prof. Dr. J.C. Rocha for the metal determination and Mr. David Nicholson for revising the manuscript. References APHA/AWWA/WEF, 1989. Standard methods for the examination of water and wastewater, 17th ed., American Publish Health Association, Washington, DC. Baird, C., 1995. Environmental Chemistry. W.H. Freeman and Company, New York. Brady, N.C., Buckman, H.O., 1983. Nature and Properties of Soils, 7th ed., Macmillan, London. CONAMA – Conselho Nacional do Meio Ambiente, 2007.
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