Journal of Environmental Quality 30:121-130 (2001)
© 2001 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
TECHNICAL REPORT
ORGANIC COMPOUNDS IN THE ENVIRONMENT
Photodegradation of Selected Herbicides in Various Natural Waters and Soils under Environmental Conditions
Ioannis K. Konstantinou,
Antonios K. Zarkadis and
Triantafyllos A. Albanis
Department of Chemistry, University of Ioannina, Ioannina 45110, Greece
Corresponding author (talbanis{at}cc.uoi.gr)
Received for publication December 22, 1999.
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ABSTRACT
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The photochemical degradation of herbicides belonging to different chemical groups has been investigated in different types of natural waters (ground, river, lake, marine) and distilled water as well as in soils with different texture and composition. Studied herbicides and chemical groups included atrazine, propazine, and prometryne (s-triazines); propachlor and propanil (acetanilides); and molinate (thiocarbamate). The degradation kinetics were monitored under natural conditions of sunlight and temperature. Photodegradation experiments were performed in May through July 1998 at low concentrations in water samples (2–10 mg/L) and soil samples (5–20 mg/kg), which are close to usual field dosage. The photodegradation rates of all studied herbicides in different natural waters followed a pseudo–first order kinetics. The half-lives of the selected herbicides varied from 26 to 73 calendar days in waters and from 12 to 40 d in soil surfaces, showing that the degradation process depends on the constitution of the irradiated media. The presence of humic substances in the lake, river, and marine water samples reduces degradation rates in comparison with the distilled and ground water. On the contrary, the degradation in soil is accelerated as the percentage of organic matter increases. Generally, the photodegradation process in soil is faster than in water. The major photodegradation products identified by using gas chromatography–mass spectrometry (GC–MS) techniques were the hydroxy and dealkylated derivatives for s-triazines, the dechlorinated and hydroxy derivative for the anilides, and the keto-derivative for the thiocarbamate, indicating a similar mode of degradation for each chemical category.
Abbreviations: DOM, dissolved organic matter • ECD, electron capture detection • FTD, flame thermionic detection • GC, gas chromatography • MS, mass spectrometry • TLC, thin-layer chromatography
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INTRODUCTION
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SINCE herbicides such as triazines, anilides, and carbamates are widely used for weed control in agricultural corps, their behavior in the environment is vitally important. As a consequence of their widespread use, residue levels ranging from 0.1 to 10 µg/L have been detected in surface waters (Durand et al., 1992; Chiron et al., 1993; Pereira and Hostettler, 1993; Albanis et al., 1998; Chiron et al., 1995). Relatively high water solubilities and hydrolysis half-lives of these suggest that they may partition into ground and surface waters (Chiron et al., 1995). Therefore, information about possible degradation mechanisms in the environment is important in order to simulate the persistence of these compounds and to identify the factors that influence their behavior.
Among the different transformation processes (biotic and abiotic), photodegradation is an important factor influencing the fate of pesticides in the field (Fielding et al., 1992). Several authors have reported on pesticide photodecomposition under a variety of irradiation conditions and different solutions. Most of these works use artificial solar sources of irradiation (xenon arc lamps or mercury lamps) or ultraviolet (UV) light combined with catalyst particles such as TiO2 or Fe2O3. (Khan and Gamble, 1983; Taboada et al., 1995; Mansour and Feicht, 1994; Pelizzetti et al., 1992). Photolysis rates and phototransformation products are actually dependent on the intensity and wavelength distribution of the light used. To accelerate photolysis testing, high-pressure Hg lamps with a wavelength output of approximately 254 nm were typically used in laboratory studies (Durand et al., 1994). These studies pose difficulties for extrapolation to environmental conditions, as only wavelengths of light >290 nm would be useful for fate measurements (Wan et al., 1994).
Although modeling pesticide behavior under laboratory conditions is a very useful tool for environmental studies, there is a need to conduct photodegradation studies in natural conditions taking into consideration factors such as temperature variation the daily and seasonal variation of natural sunlight intensity (Chiron et al., 1995).
There is relatively little information in the open literature on the phototransformation of pesticides under natural environmental conditions (Takahashi et al., 1985; Kochany and Maguire, 1994; Lartiges and Carrigues, 1995; Hebert and Miller, 1990). The present work tries to approximate appropriate environmental conditions using natural sunlight and natural water and soil samples. In such media the species that can absorb light, except for herbicides, are dissolved organic matter (DOM) and inorganic compounds, which play an important role on the photochemistry of herbicides. Thus, photodegradation can occur by direct or indirect absorption of light (Zepp and Cline, 1977; Choudhry and Webster, 1985). Competitive sunlight absorption by water and soil chromophores, variable sorption of pesticides on organic water and soil colloids, and competing biotic and abiotic transformation processes complicate the photodegradation study in natural conditions.
In direct photolysis, the substance absorbs UV-visible light energy and undergoes transformation, whereas during indirect photolysis light energy is absorbed by other constituents of the media (water, soil) (Torrents et al., 1997). The excited species can then either transfer the energy to the substance (sensitization), undergo an electron transfer with the substance, or lead to the formation of reactive species, such as singlet oxygen or hydroxy radical, which enter into a series of reactions (Zepp et al., 1985; Mill et al., 1980; Haag and Hoigne, 1986).
The present study deals with the influence of the water and soil constitution on the rate of photodecomposition and with the identification of the major photoproducts that were formed for six of the most used herbicides in Europe (Fielding et al., 1992). Atrazine (6-chloro-N2-ethyl-N4-isopropyl-1,3,5-triazine-2,4-diamine), propazine [6-chloro-N,N'-bis(1-methylethyl)-1,3,5-triazine-2,4-diamine], and prometryne [N,N'-bis(1,methylethyl)-6-(methylthio)-1,3,5-triazine-2,4-diamine] were selected from the s-triazine group; propachlor (2-chloro-N-isopropylacetanilide) and propanil [N-(3,4-dichlorophenyl)propanamide] from the anilides group; and molinate (S-ethyl hexahydro-1H-azepine-1-carbothioate) from the thiocarbamates group. Water pollution by triazines and chloroacetanilides as well as molinate is the highest at estuarine areas of certain European countries, showing that many of these compounds are transported significant distances from their application sites (Readman et al., 1993; Barcelo et al., 1996; Albanis and Hela, 1998).
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MATERIALS AND METHODS
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Chemicals
Compounds tested in this study included atrazine, propazine, and prometryne (s-triazines group); propachlor and propanil (acetanilides); and molinate (thiocarbamate). All compounds were residue analysis–grade, purchased from Riedel-de Haën (Seelze, Germany) and used without further purification. The physicochemical properties of these herbicides are shown in Table 1. Pesticide-grade hexane, methanol, acetone, dichloromethane, and ethyl acetate were purchased from Pestiscan (Labscan Ltd., Dublin, Ireland). Reagent-grade sodium sulfate was purchased from Baker (Deventer, the Netherlands). The sodium sulfate used for drying organic extracts and the disposable pipettes were heated to 500°C for 24 h before use. All glassware was rinsed with pesticide-grade solvents before use. Organic free water for photochemical experiments was prepared with a Milli-Q system from Millipore–Waters (Mississauga, ON, Canada). Empore extraction disks were manufactured by 3M and distributed by Varian (Harbor City, CA). The solid-phase extraction (SPE) disks used were 47 mm in diameter and 0.5 mm thick. Each disk contained about 500 mg of C18 bonded silica (92 ± 2%) and 10 ± 2% PTFE. The particle characteristics were: 8-µm particle size, 60-Å pore size, and irregular shape (Table 1).
Water and Soil Sampling
Natural waters used in experiments were collected from the Epirus region of Greece. Water was collected from Ioannina (ground water), Lake Pamvotis, the Louros River, and the Ionian Sea. The province of Epirus is located in the northwest part of Greece and spreads out in an area of 9203 km2, the greatest part of which is mountainous and with plains covering only 15%. Agriculture and mixed farming are the major economic activities. The Louros River, with a basin of 925 km2 and an average annual flow rate of 19.2 m3/s, drives through the plains of Preveza and Arta to Amvrakikos Gulf. Lake Pamvotis is a moderately sized (22 km2), shallow, eutrophic lake, with the city of Ioannina lying along its southwestern shoreline. The natural water samples were obtained from the top meter of each water body and refrigerated at 4°C prior to use. Their physicochemical characteristics are given in Table 2.
The soils used in the experiments were collected from three different regions of Greece (Preveza, Orestiada, and Livanates) and were from fields with no previous history of persistent pesticide use. The soils selected for this study correspond to intensively cultivated areas in Greece. Field-moist soils were passed through a 2-mm sieve to remove stones and large plant fragments and homogenized. Textural classes of soil samples were characterized as sandy clay loam (SCL), clay loam (CL), and sandy loam (SL) for the Preveza, Orestiada, and Livanates regions, respectively. Their physicochemical characteristics are shown in Table 3.
Photolytic Experiments in Water and Soil-Sorbed Phase
Water
Aqueous solutions of the selected herbicides were prepared for irradiation under natural sunlight at 2 to 10 mg/L. Outdoor experiments were carried out in capped pyrex glass reservoirs in the Ioannina area (University Campus, roof terrace of the Department of Chemistry) during the period May through July 1998. Water samples, with no previous treatment of filtration or sterilization, were spiked with each pesticide by adding the standard pesticide in methanol using reaction reservoirs of 1000 mL before starting the photolytic experiments. The mixtures were homogenized by magnetic stirring and were exposed outdoors without stirring after an equilibration period of 12 h. The pesticide concentrations in water varied and were below their solubility level in water but of sufficient concentration to carry out the kinetic studies performed by GC with electron capture detection (ECD) and flame thermionic detection (FTD) (operating conditions described below). A dark control experiment was also conducted by exposing dark pyrex glass reservoirs filled with the same pesticide solutions and covered with aluminum foil in the same environmental conditions.
Incident solar radiation was measured with a radiometer (Eppley Lab., Newport, RI) that measured radiation (W/m2) in the wavelength range of 285 to 2800 nm. The mean sunlight intensity at the beginning, middle, and end of the day was estimated respectively as 187, 606, and 309 W/m2 during May; 270, 729, and 335 W/m2 during June; and 247, 743, and 409 W/m2 during July. The average total daily shortwave radiation for this period was 674 W/m2, with a 10-h mean sunshine duration from sunrise to sunset. The mean daily air temperature was 22°C, and maximum and minimum air temperatures were 37.2 and 7.8°C, respectively, during this period. The mean daily cloudiness was 3.4 (measured in octals).
Soil Sorbed Phase
Soil thin-layer chromatography (TLC) plates (20 x 20 cm, 1 mm thickness) were prepared with fine soil particles passed through a 2-mm sieve. The soils were homogenized and equilibrated with one-half of the soil-sample's weight of aqueous solutions of each herbicide previously spiked with methanol standard solutions in order to achieve initial concentrations at the range of 5 to 20 mg/kg, which are close to usual field dosages. The content of methanol in aqueous solutions was kept below 10%. Then the plates were dried and exposed to natural sunlight for 15 d during July 1998. A dark control experiment was also conducted by covering analogous soil TLC plates with aluminum foil.
Extraction and Analyses
Water samples of 5 mL were withdrawn from the glass reservoirs at different time intervals of 0, 2, 4, 8, 16, 32, 45, and 64 d. The samples were extracted twice with 2.5 mL n-hexane for 1 min using a vortex, dried with a small amount of Na2SO4, and analyzed by GC with ECD and FTD. At the end of experimental period the final remaining water volume (about 900 mL) was divided in two aliquots for the photoproduct analysis. The first part was extracted three times with 50 mL of an n-hexane and acetone mixture (1:1, v/v). The extracts were combined, dried with Na2SO4, concentrated to 1 mL, and analyzed by GC–MS for the identification of final photolytic by-products (operating conditions described below). The second part was extracted by means of solid-phase extraction as follows. The C18 extraction disks were conditioned with 10 mL of acetone for 2 h and 5 mL of methanol modifier was added to the residue to allow better extraction. The disks were placed in the conventional filtration apparatus and washed with 5 mL of dichloromethane and ethyl acetate (1:1, v/v) solvent mixture, under vacuum and with 3 mL of methanol for 3 min, with no vacuum applied. The disks were not allowed to dry (Lacorte et al., 1993) and the samples were allowed to percolate through the disks with a flow rate of 50 mL/min under vacuum. The compounds trapped in the disks were eluted twice with 5 mL of dichloromethane and ethyl acetate (1:1, v/v) solvent mixture. The fractions were evaporated to 0.5 mL in a gentle stream of nitrogen. This extract residue was dissolved into 1 mL of isooctane and evaporated to a final volume of 0.5 mL prior to GC–MS analysis.
A 2- by 2-cm zone of the soil TLC plates was scraped off and extracted twice with 10 mL acetone using a vortex. After the first extraction the samples were placed for 10 min in a sonication bath. The mixtures were centrifuged at 4000 rpm for 10 min to separate the supernatant from the soil residue. The volume of the combined supernatants was fixed at 5 mL and was analyzed by GC with ECD and FTD.
Chromatographic Conditions
The analysis of the s-triazines and molinate was performed using a Shimadzu (Kyoto, Japan) 14A gas chromatograph equipped with FTD and a Shimadzu AOC-20i autoinjector (1.5-µl injections). A DB-1 capillary column (30 m x 0.32 mm i.d.) containing dimethylpolysiloxane (J&W Scientific, Folsom, CA) was used. The temperature program was: 150°C for 2 min, from 150 to 210°C with a rate of 5°C/min, hold at 210°C for 10 min, from 210 to 270°C with a rate of 10°C/min, and hold at 270°C for 3 min. Helium and nitrogen were used as the carrier and make-up gases, respectively. The detector gases were hydrogen and air, and the ion source of the FTD was a salt of alkali metallic (Rb2SO4) bonded to a 0.2-mm spiral of platinum wire. The temperatures were set at 220°C for the injector and 250°C for the detector.
A Shimadzu 14B gas chromatograph equipped with ECD and a Shimadzu AOC-20I autoinjector (1.5-µl injections) was used for the analysis of propanil and propachlor. A DB-5 column (30 m x 0.32 mm i.d.) containing (5% phenyl) methylpolysiloxane (J&W Scientific, Folsom, CA) was used and the same temperature program as described above was followed. The temperatures were set at 250°C for the injector and 300°C for the detector. Helium was used as the carrier and nitrogen as the make-up gas.
A Shimadzu QP 5000 GC–MS instrument was used for by-product identification. A DB-1 capillary column (30 m x 0.32 mm i.d.) containing dimethylsiloxane (J&W Scientific, Folsom, CA) was used. The injector temperature was 220°C. The column temperature program was: 55°C for 2 min, from 55 to 210°C with a rate of 5°C/min, hold at 210°C for 20 min, from 210 to 270°C with a rate of 20°C/min and hold at 270°C for 4 min. Helium was used as the carrier gas at 96.5 kPa (14 psi). The interface was kept at 270°C. The MS was operated in electron impact mode with an ionization potential of 70 eV and the spectra were obtained in full scan mode.
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RESULTS AND DISCUSSION
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Photodegradation in Water
The photodegradation rates of all studied herbicides in different natural waters followed a first-order degradation curve, Ct = C0 exp 
, where Ct is the concentration of a herbicide at time t, C0 is the herbicide initial concentration, and k is the rate constant (Table 4). The half-life time (T1/2) corresponds to a period of time at which the pesticide concentration is equal to half of the initial concentration given by the equation T1/2 = ln 2/k. The photodegradation constants (k) were calculated by subtracting the exponents of the different degradation curves representing the apparent degradation and the degradation owed to hydrolysis, volatilization, and adsorption (blank experiment). In this way, the considered k constants and T1/2 refer to the real photochemical reaction excluding the contribution of other factors (Tables 4 and 5). Figures 1 and 2
show the degradation curves of the six selected herbicides in four various natural waters and distilled water. The photodegradation rates of s-triazine herbicides (atrazine, propazine, and prometryne), anilide herbicides (propachlor and propanil), and molinate decreased in the following order: marine water < lake water < river water < distilled water and ground water.
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Table 4. Equations and correlation coefficients (R2) describing the photodegradation rates of the selected herbicides in various natural waters under sunlight
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Fig. 1. Photodegradation of atrazine, propazine, and prometryne in four natural waters and distilled water under environmental conditions. (b) = blank experiment
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Fig. 2. Photodegradation of propachlor, propanil, and molinate in four natural waters and distilled water under environmental conditions. (b) = blank experiment
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Photodegradation of Triazines
The photodegradation of s-triazine herbicides in distilled water proceeded via direct photolysis. No effect was observed on direct photolysis rate relative to the alkyl substitution in the lateral chains, since the half-life times are similar and hydroxy-derivatives are the identified by-products by GC–MS (Table 6a). These derivatives have already been identified by other authors (Pape and Zabik, 1970). The mechanism of triazine direct photolysis has been reported elsewhere (Torrents et al., 1997; Barcelo et al., 1993). Thus, the photoreaction site must not involve the alkyl group but the chlorine or methyl-thio group in the second position, further supporting that dechlorination and hydroxy-derivative formation is the major pathway in direct photolysis. On the contrary, the presence of N-dealkylated products in the natural water samples shows that the indirect process of degradation that took place owed to the DOM.
The photodegradation rate of all the investigated herbicides was lower in the natural waters than in distilled and ground water. The above result could be due to the optical filter effect (quenching) of the organic matter that could act as one of the important sunlight-absorbing components of the aquatic environment (Larson and Weber, 1994, p. 359–413; Frimmel and Hessler, 1994). Particulate matter such as sediment particles and microorganisms suspended in a water column may scatter incident light, greatly reducing penetration of light beneath the surface. Photoreaction of the colored dissolved organic matter (CDOM) in sunlight caused a decrease in UV and visible absorbance that occured most rapidly in the UV-B (280–315 nm) region (Gao and Zepp, 1998). This region of the spectrum is the most important for the photodegradation of the selected herbicides because their absorbance above 320 nm is negligible.
In other cases, sorption protects substrates from photolysis, possibly by competitive light attenuation, by migration of the pollutant into regions of the particle where light does not penetrate, or by quenching of the excited states of substrates by constituents of the particles (Larson and Weber, 1994, p. 359–413). Thus, another reason for the observed filter effect is that triazines were partially bound to DOM by a reversible physical sorption (Van der Waals forces) and this fraction was never available to photolysis action.
However, a sensitization effect of humic and other substances of natural waters cannot be excluded. The sunlight absorbance by the DOM and other organic chromophores as riboflavin and flavin could provide a rich variety of photochemical reactions (Rejto et al., 1983). The resulting excited states of the DOM, and reactive transients that were produced from DOM, could participate in energy transfer, electron transfer, and free radical reaction, which affect the fate of aquatic pollutants (Larson and Weber, 1994, p. 359–413). Benzoquinone and certain substituted benzoquinones are capable of abstracting a hydrogen atom from water to generate OH· (Ononye and Bolton, 1986; Vaughan and Blough, 1998). Other biological origin solutes such as tryptophan and tyrosine could generate hydrogen peroxide in sunlight (Draper and Crosby, 1983). However, small concentrations of the produced reactives could not strongly affect pesticides belonging to the above categories. Several studies on the relationship of hydroxy reactivity to pesticide persistence (Mabury and Crosby, 1994; Haag and Yao, 1992) show that these pesticides were more stable in the attacking of OH· than other water contaminants. In addition, most of the abundant anions and cations in both fresh and marine waters (chloride, bromide, sulfate sodium, potassium) are transparent to solar irradiation. Only a few trace metal cations, nitrate, and nitrite show any absorbance at all and their contribution to the total extinction is usually negligible in comparison with the organic species (Larson and Weber, 1994, p. 359–413).
The data in this study are in contrast with other studies that provide a sensitization effect of humic substances, indicating that DOM competes with triazines for UV visible light, decreasing the photodegradation rate. In other words, the sensitization effect is hidden by the strong filter effect. The similar half-lives for triazines suggest a mechanism both sterically and electronically controlled because the isopropyl group of propazine (prometryne) is more favored electronically but more affected sterically. The results show increased half-lives of the herbicides in marine water samples. This is consistent with ·OH scavenging by chloride and bromide ions and with the increase of the binding fraction of triazines to the DOM as the water salinity increases (Chiron et al., 1995). The filter effect of the DOM and suspended particulates, living and nonliving, present in marine water contribute less to the retardation of photodegradation in comparison with lake and river water. In Table 2, DOM and total suspended particulates were one to five orders lower in marine water. In lake water pigments derived from algae or other microorganisms were responsible for the retardation of photodegradation compared with distilled and ground water. In ground water the degradation was faster than in deionized water.
Some ions of ground water, such nitrite and nitrate, absorb light [
max = 355 
and max = 303 
, respectively] and undergo homolysis to produce free radicals (Larson and Weber, 1994, p. 359–413; Penuela and Barcelo, 1998). It seems probable that in nitrite and nitrate photolysis, the production of ·OH radicals accelerates the organic reactions leading to the degradation of the pesticide (Torrents et al., 1997). Other ions accelerate the photodegradation, such as calcium and magnesium, which take part in complexation reactions enhancing the photolysis of organic compounds (Larson and Weber, 1994, p. 359–413; Penuela and Barcelo, 1998). Also, some traces of Fe+2 ions could assist the degradation following the known photo-Fenton reactions (Larson et al., 1991).
Photodegradation of Anilides and Molinate
For anilides herbicides (propachlor and propanil), the results showed a similar behavior. In direct photodegradation in distilled water the elimination of the chlorine atoms was the predominant reaction, followed by the rupture of the amide bond. In the case of river, lake, and marine water samples, indirect processes produced different by-products, which were identified as a result of a radical reaction (Fig. 3 and 4)
. The role of DOM was the same for triazines, where it had a masking behavior that provided lower photodegradation rates for natural water samples.
Finally, molinate presented the same behavior, and the predominant reaction was the formation of the 4-keto derivative both in the direct photolysis of herbicide and an indirect process through reaction with radical species (Fig. 5)
. The identified photo by-products of the selected herbicides with their mass spectra are given in Table 6a,b.
Photodegradation in Soil
In soils, photolysis will occur within a shallow surface zone, the depth of which depends on soil characteristics and the mechanism of photodegradation. Light absorption and photolysis of organic contaminants will be influenced by sorption reactions that are related to the soil organic matter content and by singlet oxygen formation. Both direct and indirect process could be occuring depending on the depth. Hebert and Miller (1990) concluded that the vertical depth of direct photolysis on the soil surface will be restricted to a region of approximately 0.2 to 0.3 mm. Mean indirect photolysis depths were reported to be greater than 0.7 mm for outdoor experiments. Several studies (Takahashi et al., 1985; Lartiges and Carrigues, 1995; Zepp et al., 1985; Zafiriou et al., 1984; Jensen-Korte et al., 1987; Vaughan and Blough, 1998) have indicated that humic acids are capable of acting as sensitizers for the production of reactive intermediates such as singlet oxygen (1O2), hydroxyl radicals (·OH), superoxide anion (·O2-), hydrogen peroxide (H2O2), and peroxy radicals (ROO·). For example, sediment-sorbed DDE photolyzes faster than when it is dissolved, and the product mixture differs in a way that suggests an environment rich in H-donors, probably organic matter (Zafiriou et al., 1984; Miller and Zepp, 1979). Such reactive species can potentially diffuse to depths approaching 1 mm depending on moisture depth, soil porosity, and thermal gradients on the sunlight-exposed soil interface. Finally, metal oxides presented in the soil, such as ZnO, Fe2O3, and MnO2, absorb radiation in the sunlight wavelength range and could accelerate degradation by reaction of OH· and ·O2- through the well-known mechanism of semiconductor photochemistry.
Moreover, in many cases, the electronic structures, absorption spectra, and excited state lifetimes of sorbed compounds are much different from their solution properties, making it very difficult to predict what effects may result from sorption. Nitrogen-containing compounds often show different spectra when sorbed on clay minerals compared with solution spectra. This is probably due to protonation–deprotonation effects (s-triazines), charge transfer phenomena, or excited state alterations. Triplet lifetimes for many aromatic compounds are greatly extented when they are adsorbed on silica or alumina, which means they may be more susceptible to photochemical reactions (Larson and Weber, 1994, p. 359–413).
The results of the present study confirmed that the sensitizing effect is bigger than the scattering one, as the photodegradation rates in soils were faster. Both direct and indirect process could occur in soil surfaces, although we did not analyze for by-products. The half-lives of herbicides ranged from 12 to 40 calendar days using the same calculation mode as in waters (Fig. 6 and 7
; Tables 7 and 8). The degradation rates increase as the organic matter content increases. Fig. 8
shows the relation between the organic matter content and the photolysis rate of herbicides on soil surfaces. The results of this study could suggest that DOM has a low aromatic or conjugated character in contrast to the humic acids, which seem to have a significant concentration of the structural components susceptible for radical formation.
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Fig. 6. Photodegradation of atrazine, propazine, and prometryne in soil surfaces under environmental conditions. (b) = blank experiment
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Fig. 7. Photodegradation of propachlor, propanil, and molinate in soil surfaces under environmental conditions. (b) = blank experiment
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Table 7. Equations and correlation coefficients (R2) describing the photodegradation rates of the selected herbicides in soil-sorbed phase under sunlight
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CONCLUSIONS
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The photochemical degradation of six common herbicides has been investigated. The degradation rates in natural waters were lower than in distilled and ground water, showing a strong dependence on the composition of the water sample and especially on the DOM, which provides an optical filter effect. An increased optical filter effect would be seen as DOM concentrations increase, resulting also in partitioning of the herbicide into the organic fractions, increasing the solubility. In marine samples, ·OH is rapidly scavenged, affecting herbicide persistence. Nitrate and nitrite can produce ·OH but in the case of nitrite can also scavenge ·OH as well. Thus, direct photolysis becomes more important while ·OH processes contribute less to the photolytic fate of herbicides as surface-layer DOM concentrations increase. Conversely, in soils the degradation rates were accelerated with increased levels of organic matter.
It could be stated that DOM concentration and presumably the type of functional groups and aromaticity present in DOM will influence the photoprocess differently. To understand the differences between direct and indirect processes, the influence of diverse structural properties of DOM should be estimated and more experiments should be carried out both in liquid and solid phases in order to elucidate the humic substance role and the mechanisms that take place in such systems. The major photodegradation products identified were the hydroxy and dealkylated derivatives for s-triazines, the dechlorinated and hydroxy derivative for the anilides, and the keto-derivative for the thiocarbamate, indicating a similar mode of degradation for each chemical category.
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ACKNOWLEDGMENTS
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The authors acknowledge the support of the General Secretariat of Research and Technology under PENED Grants, Athens, Greece (Contract, 95-E
-298).
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INTRODUCTION
