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Published online 5 January 2006
Published in J Environ Qual 35:312-323 (2006)
DOI: 10.2134/jeq2004.0025
© 2006 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

Vadose Zone Processes and Chemical Transport

Fate of Diuron and Linuron in a Field Lysimeter Experiment

L. Guzzellaa,*, E. Caprib, A. Di Corciac, A. Barra Caracciolod and G. Giulianod

a Istituto di Ricerca sulle Acque - CNR, via Della Mornera 25, 20047 Brugherio (MI), Italy
b Istituto di Chimica Agraria ed Ambientale, Università Cattolica del Sacro Cuore, Via Emilia Parmense 84, 29100 Piacenza, Italy
c Dipartimento di Chimica, Università La Sapienza, Piazza Aldo Moro 5, 00100 Roma, Italy
d Italy Istituto di Ricerca Sulle Acque - CNR, Via Reno 1, 00198 Roma, Italy

* Corresponding author (guzzella{at}irsa.cnr.it)

Received for publication January 20, 2004.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
The environmental fate of herbicides can be studied at different levels: in the lab with disturbed or undisturbed soil columns or in the field with suction cup lysimeters or soil enclosure lysimeters. A field lysimeter experiment with 10 soil enclosures was performed to evaluate the mass balance in different environmental compartments of the phenylurea herbicides diuron [3-(3,4-diclorophenyl)-1,1-dimethyl-urea] and linuron [3-(3,4-dichlorophenyl)-1-methoxy-1-methylurea]. After application on the agricultural soil, the herbicides were searched for in soil, pore water, and air samples. Soil and water samples were collected at different depths of the soil profile and analyzed to determine residual concentrations of both the parent compounds and of their main transformation products, to verify their persistence and their leaching capacity. Air volatilization was calculated using the theoretical profile shape method. The herbicides were detected only in the surface layer (0–10 cm) of soil. In this layer, diuron was reduced to 50% of its initial concentration at the end of the experiment, while linuron was still 70% present after 245 d. The main metabolites detected were DCPMU [3-(3,4-dichlorophenyl)-1-methylurea] and DCA (3,4-dichloroaniline). In soil pore water, diuron and linuron were detected at depths of 20 and 40 cm, although in very low concentrations. Therefore the leaching of these herbicides was quite low in this experiment. Moreover, volatilization losses were inconsequential. The calculated total mass balance showed a high persistence of linuron and diuron in the soil, a low mobility in soil pore water (less than 0.5% in leachate water), and a negligible volatilization effect. The application of the Pesticide Leaching Model (PELMO) showed similar low mobility of the chemicals in soil and water, but overestimated their volatilization and their degradation to the metabolite DCPMU. In conclusion, the use of soil enclosure lysimeters proved to be a good experimental design for studying mobility and transport processes of herbicides in field conditions.

Abbreviations: DCA, 3,4-dichloroaniline • DCPMU, 3-(3,4-dichlorophenyl)-1-methylurea • DCPU, 3-(3,4-dichlorophenyl)urea • Koc, organic carbon partition coefficient • PELMO, Pesticide Leaching Model • PUF, polyurethane foam cartridge • TP, transformation product • TPS, theoretical profile shape


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
PESTICIDES are widely used in Italian food production. Italy ranks second in Europe (after France) and third worldwide in pesticide use, with over 50910 Mg of active ingredients sold every year. The phenylurea herbicides diuron and linuron are among the most used for weed control; they have a high level of phytotoxicity and act in plants as photosynthetic inhibitors. The effects of phenylurea herbicides on nontarget organisms are currently being studied and arouse concern (Simon et al., 1998). Diuron has been found to be highly toxic for some nontarget organisms (Nebeker and Schuytema, 1998; Teisseire et al., 1999), and its potential toxicity at cellular and subcellular levels has also been demonstrated (Chauhan et al., 1998). Linuron has been classified as possibly carcinogenic in humans by the USEPA (Wagner, 1983). However, above all, the metabolites have been reported to be more toxic than their parent compounds (World Health Organization, 1993; Simon et al., 1998; Tixier et al., 2000).

Linuron and diuron are relatively persistent in soil, with a quite variable DT50 value (degradation time for 50% of the substance) in the field in a range from 30 to 180 d for diuron and from 30 to 150 d for linuron (Rao and Davidson, 1980; Cullington and Walker, 1999). Phenylurea herbicides together with their metabolites have been detected as contaminants of ground waters (Field et al., 1997; Spliid and Køppen, 1998).

After introduction into agricultural soils, several factors may affect the environmental fate of these herbicides. The environmental characteristics, such as agricultural practices, soil profile, and hydrological conditions, will determine their vertical and horizontal distribution in the field. Depending on these factors, herbicides can evaporate into the atmosphere, adsorb onto the soil, run off into rivers, be leached to ground water, accumulate in biota, and/or degrade via abiotic and biotic processes. The latter are involved in their complete degradation. Microbial metabolism is, in fact, recognized as the primary force in transformation and mineralization of phenylurea herbicides (Geißbuhler et al., 1973; Vroumsia et al., 1996; Gillian et al., 2001). Diuron and linuron are reported to be a substrate for bacterial growth (Engelhardt et al., 1971; Cullington and Walker, 1999; Widehem et al., 2002) and to be transformed to their metabolites (Haggblom, 1992). Nevertheless, some transformation products (TPs), such as DCPMU and DCA, cause concern because they show toxicological effects greater than their parent compounds (Lo et al., 1990; World Health Organization, 1993; Valentovic et al., 1997; Simon et al., 1998; Bauer et al., 1998; Tixier et al., 2000).

The environmental fate of herbicides can be studied at different levels: in the lab with disturbed or undisturbed soil columns (experiments particularly useful for calculation of the organic carbon partition coefficient [Koc] and the DT50 of herbicides) (Mallawantantri et al., 1996), or in the field with suction cup lysimeters (Byers et al., 1995) or soil enclosure lysimeters (Bergstrom, 1990; Lennartz et al., 2001). The first type of lysimeter is used particularly for studying the transport processes of herbicides from surface soil to ground water (Biocca, 2001; Guzzella et al., 2000), while the second type for measuring the mass balance of the herbicides in different environmental compartments (Bergstrom, 1990). In the latter type, the quantification of the amount of leaching herbicides is very accurate and precise, because the pore water is collected at the bottom of the different lysimeters and it can be used to quantify water volume collected and herbicide amount leached. The amount of herbicide leaching from the bottom of lysimeters takes into account both matrix transport and preferential flow transport in soil. The latter is one of the most important factors in significantly increasing the risk of ground water contamination. In fine textured soils, large and discontinuous macropores consisting of shrinkage cracks, earthworm channels, or root holes operate as preferential flow pathways and can cause rapid movement of chemicals through the unsaturated zone (Klavidko and Timmenga, 1990).

In this study, the environmental fate of two widely used phenylurea herbicides, diuron and linuron, was studied in a field lysimeter experiment. Ten soil enclosure lysimeters were installed and the herbicide mass balance was calculated using data from pore water obtained by lysimeters, from soil by soil coring, and from air by polyurethane foam cartridge (PUF) sampling. The idea that pushed the authors to use soil enclosure lysimeters was that of testing the goodness and effectiveness of this experimental model in respect to the use of suction cups lysimeters (Guzzella et al., 2000, 2001). The role plaid by bacteria in the herbicide degradation was also investigated by studying bacterial abundance in soil samples measured with an epifluorescence direct count method that can estimate living organisms.

The field data were also used to simulate the herbicide transport with PELMO (Klein, 1995). PELMO is a well-known European registration model used for the calculation of the predicted environmental concentrations (PECs) in soil, runoff and drainage water, air, and leaching water. The comparison between predicted and experimental concentrations in the different matrices allows us to point out some possible discrepancies between the real situation and modeled events.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 
Experimental Field and Soil Profile Description
The experimental site (Monza, Italy) is on the central Lombardy plane. The plane was formed from the depositions of the Seveso, Lambro, and Adda Rivers. The geological substrate consists of silty gravel with noncalcareous sand and of silty-gravelly sands, sometimes clayey, both calcareous and noncalcareous. The soil is a Typic Hapludalf (USDA Soil Taxonomy) with Typic and Fluventic Dystrudepts and Oxyaquic Hapludolls (Ente Regionale Sviluppo Agricolo Lombardo, 2001). A 50-cm deep soil core was sampled before the herbicide application (April 2001). The main geopedological and chemical characteristics of the soil profile are shown in Table 1. The texture was silt loam with a clay content that increased from the topsoil to the Bt horizon. The organic carbon content in the topsoil horizon was quite high (2.6%) and decreased with depth, especially below 37 cm. The same decreasing trend was detected for cation exchange capacity (CEC) values and Ca, Mg, and K contents. The experimental area was a 4.6- x 6.4-m rectangle (29.44 m2) not treated with herbicides before the experiment.


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  		Table 1. Geopedological and chemical characteristics of the soil profile.

 
Lysimeter Setup
Ten lysimeters (28-cm i.d.) were installed in a circle (Fig. 1 ) and identified by numbers from 1 to 10: five had a 40-cm depth (even numbers) and five were at a 20-cm depth (odd numbers). The soil enclosure lysimeters were prepared by pushing a 28-cm-i.d. PVC tube into the soil using a heavy hammer and reaching the depths of 20 or 40 cm. The soil enclosures were transported to the laboratory, where a Teflon draining bottom and a glass funnel for collecting leaching waters was connected under the lysimeters (Fig. 2 ) and then the soil enclosure was reinstalled in the soil in the same position the next day. We used the same soil horizon to fill the hole around the soil enclosure. A Teflon tube was connected to the glass funnel bottom of the lysimeters and to a dark glass 500-mL bottle to collect the water sample. The soil enclosure was allowed to settle for 15 d before starting the treatments. The whole experiment was undertaken for 245 d for all the lysimeters except for Lysimeters 2 and 3 that were transported to the laboratory after 52 d because they were used for studying herbicide bioaccumulation in anellids (Basker et al., 1994). The results of the bioaccumulation experiment are not shown in the present paper.



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  		Fig. 1. Position of lysimeters (empty circle) and of paper filters (filled circle) in the experimental plot. The position in the x–y scale is given in meters.

 


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  		Fig. 2. Scheme of a soil enclosure lysimeter; the plate is located between the lysimeter and its base.

 
Meteorological Station and Air Sampling System
A meteorological station with a data acquisition system (Model AD2; Silidata, Modena, Italy) was placed inside the experimental field to make hourly measurements of wind speed and direction at different heights (20, 75, and 210 cm), air temperature and humidity, and soil temperature and humidity at the 5-cm depth. An air sampling system (polyurethane foam cartridge [PUF] connected with a tygon tubing system to a pump) was installed at the meteorological station.

Chemicals and Herbicide Application
Analytical reference standards of linuron [3-(3,4-dichlorophenyl)-1-methoxy-1-methyl-urea], diuron [3-(3,4-diclorophenyl)-1,1-dimethyl-urea], DCPMU [3-(3,4-dichlorophenyl)-1-methylurea], DCPU [3-(3,4-dichlorophenyl)urea], and DCA (3,4-dichloroaniline) were purchased from Dr. Ehrenstorfer (Germany) with a certified purity greater than 95%.

Linuron and diuron were applied on 5 June 2001. Commercial herbicide solutions of Siolcid (linuron 37.6%; Siapa, Italy) and Toterbane (diuron 50%; Chimiberg, Albano S. Alessandro, Italy) were used; 1.51 L of this solution, containing a total amount of 1.7 g of linuron and 6.8 g of diuron, were applied uniformly on the plot with a metallic bar connected to a system of 10 nozzles at 30-cm intervals. The bar was kept 70 cm above the soil, as is usual in agricultural practices. The theoretical applied doses (0.0564 g m–2 of linuron and 0.229 g m–2 of diuron) corresponded to the agronomic advised doses. Six paper filters (9 cm Ø) were randomly distributed over the experimental plot before the herbicide was applied, collected at the end of the treatment, and analyzed to evaluate the real applied dose.

During application of the herbicides, the surface soil water content was 4% (w/w), the wind speed 0.1 m s–1, and the temperature 34°C (at a 5-cm depth), while the air temperature was 28°C. Rainfall was recorded every day during the experiment; in addition, a rain event (100 mm d–1) was simulated on the 16th day, because the period was very dry.

The main chemio-dynamic characteristics of diuron and linuron (Tomlin, 1997) are: solubility in water at 25°C = 42 mg L–1 for diuron and 75 mg L–1 for linuron; Koc = 0.49 m3 kg–1 for diuron and 0.5 to 0.6 m3 kg–1 for linuron; and vapor tension = 3.1 x 10–6 mmHg for diuron and 1.5 x 10–5 mmHg for linuron.

Volatilization Study
The volatilization of pesticides from the soil surface was calculated with the theoretical profile shape (TPS) method described by Wilson et al. (1982), used to determine the gaseous mass transfer from field experiments conducted in a circular plot. The TPS method has the advantage over the aerodynamic method (Yates, 1996) in that: (i) the large fetch requirement is not necessary, (ii) measurement of the air concentration and wind speed are needed at only one height, and (iii) the sensor is placed at a height that is relatively insensitive to the atmospheric stability so temperature, wind gradients, and stability correction are unnecessary. This approach is based on the Trajectory Simulation Model (TSIM) (Wilson et al., 1982) and was used (Yates, 1993) to simulate movement of pesticides away from a treated field. Volatilization at the soil surface V (mg m–2 s–1) was calculated according to:

[1]
where u(t) is the wind velocity (m s–1) and c(t) is the herbicide concentration in the air (mg m–3) at sampling height Zinst. The term {Omega} (dimensionless), the horizontal to vertical flux ratio, was calculated from the TSIM on the basis of the soil roughness (Z0) and the circular surface radius. With the TSIM, the hypothetical movement of the particles in the atmosphere is traced from the source to the point of measurement. Since the height profiles of the theoretical position of the volatilized pesticides cross each other for stable and unstable atmospheric conditions, it is possible to identify a single height, Zinst, were measurements can be performed for all atmospheric conditions. Surface roughness Z0 was calculated at 0.8 cm through the wind logarithmic profile, obtained by measuring wind speed at three different heights (20, 75, and 210 cm) for 3 d continuously. The term Z0 was the constant of the exponential function obtained by interpolation. The value of Zinst was calculated to be 23 cm, with a corresponding value of 5.0 for {Omega}. The PUF sampling plugs were positioned at the center of the plot. Air samples were collected daily for the first 13 d after the herbicide application.

Water and Soil Sample Collection
Water
Twelve soil pore water samples (about one for each week or after rain events) from each lysimeter were collected during the whole experiment in dark glass bottles (500 mL) using a vacuum pump (Millipore, Billerica, MA). Samples were kept at 4°C until analysis. Sampling was performed after rain events or rain simulation (irrigation).

Soil
Soil was sampled by driving a 40-cm PVC tube (5.6-cm i.d.) with a hammer into the ground inside the experimental area but outside the soil lysimeters. Core samples (in triplicate) were taken on the 1st, 8th, 16th, 28th, 90th, 147th, and 245th day after herbicide applications.

Cores were subdivided for the analysis into four portions: 0 to 10, 10 to 20, 20 to 30, and 30 to 40 cm. At each sampling the organic content, the water content, and the surface soil temperature were also measured. For each sampling and for each layer considered, three soil subsample replicates were collected for chemical and three replicates for microbial analysis. Subsampling in the different horizons was performed taking care that no contamination by other soils occurred.

Chemical Analysis
Water
The herbicide analyses were performed by filtering water samples through a 0.45-µm cellulose nitrate filter (Whatman, Maidstone, UK). Herbicide extractions were performed by concentrating from 100 to 1000 mL of water depending on the recovered water volume through 200-mg Lichrolut EN solid-phase extraction (SPE) cartridges (Merck, Darmstadt, Germany) (Pozzoni and Guzzella, 2000). The extracts were concentrated to <1 mL by a gentle N2 flux (TurboVap II; Zymark, Hopkinton, MA) and recovered to 1 mL in 25 mM KH2PO4.

Extracts were analyzed by high performance liquid chromatography (HPLC) Diode Array 1050 (Hewlett-Packard, Palo Alto, CA) with an automatic sampler, a LiChrospher 100-RP18 5-µm column (250 mm x 4.6 mm) (HPLC Technology, Welwyn Garden City, UK). The mobile phase was H2O/CH3CN: t = 0 to 15 min 60:40 v/v, up to 20 min 25:75 v/v, t = 20 to 30 min 25:75 v/v, up to 40 min 60:40 v/v, 0.7 mL min–1 flux, 35°C. The herbicide recoveries were greater than 80% (Guzzella et al., 2001).

Quantitative analysis was performed at a 250-nm wavelength using standard solutions for HPLC calibration in KH2PO4 25 mM. The detection limits for 1-L water samples concentrated to 1 mL were 0.05 µg L–1 for diuron, linuron, DCPMU, and DCPU, and 0.1 µg L–1 for DCA.

Soil
The residue concentrations of diuron, linuron, DCPMU, DCPU, and DCA were determined using a liquid chromatography–mass spectrometry system (Di Corcia et al., 1999; Crescenzi et al., 2000). After sampling, soil samples were frozen in sealed dark glass bottles and stored at –20°C. Then for the determination of the herbicide content, soil samples were thawed and air-dried at ambient temperature (20–22°C) for 96 h.

To extract the herbicides from the soil samples, a home-made apparatus was used, applying a hot phosphate-buffered water extraction system coupled on-line with liquid chromatography–mass spectrometry and using a Finnigan Navigator AQA benchtop mass spectrometer (Thermoquest, Milford, MA) (Crescenzi et al., 2000). During the chromatographic run, the electrospray (ES) liquid chromatography–mass spectrometry (LC–MS) system was operated in the positive-ion mode and signals were acquired by time-scheduled selected ion monitoring (SIM).

The detection limits for 25-g soil samples extracted and concentrated to 0.5 mL were 2 µg kg–1 for DCA and 1 µg kg–1 for diuron, linuron, DCPU, and DCPMU.

Soil Microbial Analysis
The bacterial abundance (number of bacteria per gram of soil) was measured by the epifluorescence direct count method, using as a fluorescent dye DAPI (4',6'-diamino-2-phenylindole) (Barra Caracciolo et al., 1999, 2001), since it allows distinction of bacteria (that appear with a luminescent blue color) from nonliving bacterium sized particles (that show up as yellow).

Air
Paper filters and PUF were extracted thrice with acetone (60 mL) in an ultrasonic bath for 15 min. The acetone extracts were concentrated by Rotavapor (30°C) and N2 flux and made up to 1 mL with CH3CN/KH2PO4.

Analysis of the chemicals extracted from paper filter and PUF was performed using an HPLC DAD 1100 (Agilent Technologies, Palo Alto, CA) with a Luna C18 column (250 x 4.60 mm), 5 µm (Phenomenex, Torrance, CA). The mobile phase was KH2PO4 2 mM/CH3CN (70:30 v/v, 30:70 v/v from 2 to 30 min). Quantitative analysis was performed at 212- and 254-nm wavelengths using external standard. The detection limits for a 1-m3 air sample concentrated to 1 mL were 0.1 ng m–3 for diuron, linuron, DCPMU, DCPU, and DCA.

Model Simulation
PELMO (Klein, 1995) is a model that simulates chemical movement in the unsaturated soil system within and below the plant root zone. PELMO is based on PRZM-1, but processes have been added to it, to overcome limitations of PRZM-1 (Klein et al., 2000).

PELMO is a well-known European registration model used for the calculation of the predicted environmental concentrations (PECs) in soil, runoff and drainage water, air, and leaching water, as required by European legislation. It also represents one of the most assessed models in recent years. It has been evaluated many times comparing the prediction to the experimental data for leaching (Vanclooster et al., 2000), runoff (FOCUS, 2001), and air (Ferrari et al., 2003; Wolters et al., 2003).

FOCUSPELMO 1.1.1, the version used in this study, is the application running on a Windows platform based on PELMO 3.2. The inputs required by the model are divided into three categories: Scenario, Pesticide, and Weather. Some of those were measured (i.e., soil bulk density, initial soil water, sand fraction, clay fraction, organic carbon, pH value, rainfall, evaporation, maximum and mean temperature, relative humidity), some others were default values due to the lack of defined information (i.e., volatilization layer of air, diffusion coefficient in air, Freundlich exponent, dispersion in soil, biodegradation factor) or obtained by Tomlin (1997) (i.e., molar mass, vapor pressure, aqueous solubility, Koc); the DT50 (degradation time for 50% of the substance) values used as input in the model were 135 d for diuron, 52.5 d for linuron, 26.4 d for DCPMU, 1.6 d for DCPU, and 0.4 d for DCA to CO2, following the degradation pathway described above.


   RESULTS AND DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
CONCLUSIONS
 REFERENCES
 
Herbicide Application
The theoretical herbicide dose applied was 229 mg m–2 for diuron and 56 mg m–2 for linuron (Table 2), corresponding to 6.75 g of diuron and 1.66 g of linuron. The analysis of the paper filters located on the experimental field together with those of the soil samples, collected just after the herbicide application, showed that the distribution of diuron and linuron was not actually uniform. It was presumably due to the dispersion phenomenon caused by the wind speed or to the distribution into neighboring areas and/or to a lack of homogeneity in the manual herbicide application. Besides, the use of six filters might have not been sufficient to study the herbicide application variability. The herbicide mass loss during application is well-known phenomenon in agricultural practices and it was described by Vischetti et al. (1997) and Capri et al. (1995).


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  		Table 2. Theoretical and real herbicide doses applied on the field experiment.

 
The average doses applied of diuron and linuron were 35 and 67% of the theoretical ones (Table 2). The mean applied dose was estimated to be 80 mg m–2 of diuron and 38 mg m–2 of linuron, corresponding to 4.92 mg lysimeter–1 of diuron and 2.33 mg lysimeter–1 of linuron. The total applied amount was 2.352 g of diuron and 1.118 g of linuron with a loss of 65% for diuron and of 33% for linuron.

Volatilization
The analysis of the PUF filters, by which air was sampled during the 13 d after the herbicide application, showed that diuron was never detected, while linuron was found in only two filters (Table 3). The sampling flux was constant (2 L min–1) during the experiment and the volatilized active principle amount was determined through the TPS model: total volatilized amount for linuron was calculated to be 2.63 x 10–5 µg m–2. The volatilization loss was therefore considered negligible.


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  		Table 3. Linuron and diuron concentrations in air samples and volatilization evaluation by the theoretical profile shape (TPS) model.

 
Herbicides and Transformation Products in Pore Water and Soil Samples
The volume of water collected during the 161 d for the 10 lysimeters sampled varied from 2 to 7 L. The lowest volumes were collected under Lysimeters 2 and 3, because sampling under these two lysimeters continued for only 52 d; Lysimeters 2 and 3 were transported to the laboratory on Day 52 for the analysis of earthworms and soil. The total mass of herbicides recovered from each lysimeter was not related to the total volume of water collected under the lysimeter, as shown in Fig. 3 .



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  		Fig. 3. Total volume (L) of soil pore water collected from each lysimeter and diuron and linuron recovered mass (µg).

 
The daily rainfalls are shown in Fig. 4 . Plotting the herbicide concentrations in the leachates vs. time, it was possible to distinguish two different lysimeter behaviors: in the first group (Lysimeters 1, 2, 7, and 10), the peak of herbicide concentrations was evident by Day 16, in concomitance with the first rain event (100 mm of artificial rain). Then the herbicide concentrations decreased to become undetectable after 97 d. In the second group (Lysimeters 3, 4, 5, 6, 8, and 9) the maximum concentration peaks were detected for Day 20. Then the concentrations of the herbicides decreased quickly, showing some secondary peaks between 40 and 90 d in concordance with other precipitation events (Fig. 4).



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  		Fig. 4. Amount of precipitation (mm), before leachate collection, throughout the experiment.

 
As representative of the first group of lysimeters, the trend for Lysimeter 1 is shown in Fig. 5 , while the trend for Lysimeter 5 is shown in Fig. 6 as representative of the second group.



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  		Fig. 5. Lysimeter 1 herbicide concentrations (µg L–1) in soil pore water samples throughout the experiment.

 


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  		Fig. 6. Lysimeter 5 herbicide concentrations (µg L–1) in soil pore water samples throughout the experiment.

 
The herbicide concentrations seemed to follow the rain events (Fig. 4, 5, and 6), but with different behavior for the two groups of lysimeters. This difference might be ascribed to several factors such as local variations in soil permeability, soil structure, and herbicide biodegradation. Heterogeneity in field conditions was observed also by other authors (Gooddy et al., 2002; Sorensen et al., 2003). Diuron and linuron showed similar trends, but linuron concentrations were always five times lower than diuron concentrations, in accordance with the lower solubility of linuron (Tomlin, 1997) and with the lower applied amount (the linuron applied amount was 50% lower than the diuron one).

In Fig. 7 the average herbicide concentrations (all 10 lysimeters) are shown at left and the average TP concentrations at right. Beyond the herbicide peaks of diuron and linuron observed after the first precipitation events (at the 16th and 20th day), it is possible to see some secondary concentration peaks between Days 38 and 97.



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  		Fig. 7. Mean concentrations (µg L–1) of diuron, linuron, and transformation products (TPs) in soil pore water samples from all 10 lysimeters.

 
The TPs mainly detected were DCPMU and DCA and their concentrations were always lower than 3 µg L–1. No significant differences were observed between lysimeters located at the two different depths (20 and 40 cm), both for parent compounds and for TPs. The presence of TPs was in accordance with the general pathway of degradation of these herbicides. In fact, it involves successive N-demethylation in the case of diuron and a N-demethylation and a demethoxylation in that of linuron. Following these stepwise N-dealchilations (Field et al., 1997; Berger, 1999), the metabolites are hydrolyzed to DCA (Dalton et al., 1996) that subsequently may be further degraded to CO2 (El-Fantroussi, 2000; Sorensen et al., 2001). DCPMU was detected during the first part of the experiment, while DCA was detected at a significant concentration only at the end of the experiment. Therefore DCA derived presumably from DCPMU degradation.

For the herbicide concentrations in soil (Fig. 8 ), the applied herbicides and their TPs were measured at four depths (0–10, 10–20, 20–30, and 30–40 cm), but they were detected only in the surface layer (0–10 cm). Moreover the herbicide concentrations in the soil samples were very variable during the first 28 d and they started to decrease substantially only after 90 d. The concentrations of diuron were approximately twice those of linuron, according to the different applied amount; but diuron was reduced to 50% of its initial concentration after about 150 d, while linuron was still 70% present at Day 245. Therefore, linuron was more persistent than diuron, in accordance with other experiments (Rao and Davidson, 1980; Cullington and Walker, 1999).



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  		Fig. 8. Mean concentrations (µg L–1) of diuron, linuron, and transformation products (TPs) in soil samples (0–10 cm).

 
Regarding the metabolites, DCPMU was the main transformation product detected, DCPU was detected only in the last part of the experiment and in trace amounts, while DCA was detected even at the beginning of the experiment, presumably because trace amounts of DCA were present in the commercial herbicide formulations and/or that some bacterial populations could have hydrolyzed the parent compounds directly to this metabolite, as recently found possible in soil (Sorensen et al., 2003). A greater amount of DCPMU was detected in soil than in the pore water, presumably because DCPMU may be readily adsorbed to the organic matter of soil (Gooddy et al., 2002), which is in accordance with other studies (Funari et al., 2000) that found the Koc of DCPMU (284) is greater than that of DCPU (149).

The bacterial abundance (bacteria per gram of soil) in soil samples was assessed during the experimental period (from 0 to 245 d) and at Day 300 for each soil layer considered. As shown in Table 4, there was a negative correlation (p < 0.01) between bacterial abundance and depth until Day 28, then no significant decrease was detected between Days 90 and 147. At Day 245 there was a relative increase in bacterial abundance in the surface layers and at Day 300, 2 mo after the end of the experiment, it was possible to observe the initial trend in bacterial abundance again, decreasing with depth as before.


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  		Table 4. Bacterial abundance in soil samples at different depths assessed at 1, 8, 16, 28 90, 147, 245, and 300 d from the start of the experiment. Each value is the mean (±standard error) of nine replicates.

 
The organic carbon, water content, and soil temperature in the surface soil samples were plotted against the microbial abundance (Fig. 9 ). No significant correlation was observed between these parameters (the regression coefficients were respectively R = 0.13, 0.18, and 0.23). The falling of bacterial abundance measured at Day 90 might be ascribed to the formation of herbicide metabolites in soil (see TPs in Fig. 8) and consequently this phenomenon might have determined a toxic effect on microorganisms; in fact, DCPMU and DCA are reported to be more toxic to microorganisms than the parent compounds themselves (World Health Organization, 1993; Tixier et al., 2000; Widehem et al., 2002). However, this part of experiment was undertaken without a negative control with no herbicide application, therefore a one way direction interpretation of the observed relation between soil depth and bacterial abundance cannot be argued.



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  		Fig. 9. Bacterial abundance (number of bacteria per gram of soil) vs. soil abiotic factors (percent organic C; percent H2O; temperature) in surface soil samples. The vertical bars represent the standard errors.

 
After 300 d, the metabolite presence can be presumed to have been negligible, and the bacterial abundance showed the initial trend with depth again.

Comparing soil concentrations with pore water ones, it is evident that diuron and linuron can be transported to a greater depth than 10 cm; they were detected in the soil pore water at depths of 20 and 40 cm, even if they were not measured in the soil samples collected at a depth greater than 10 cm. A possible explanation of this observed anomaly may be in the different limits of detection of soil samples; they were greater than those for water, owing to difficulties in the extraction of these herbicides from the clay-enriched soil horizons. Large fractions of residues not extractable by organic solvents have been observed following treatment of soil with phenylurea herbicides in various laboratory studies (Scheunert and Reuter, 2000). The herbicide-soil binding significantly decreases the mineralization of the phenyl structure of herbicides in soil systems because of the lower availability of the compounds to soil biodegradation (Sorensen et al., 2003). The residual concentration in soil and water samples detected in this experiment showed that the leaching of diuron and linuron was in any case low and that their mobilization to deeper soil horizons was not considerable.

Mass Balance in Lysimeters 2 and 3
A mass balance of the applied herbicides, calculated on the basis of molarity, was estimated at the end of the experiment (at Day 52) for Lysimeters 2 and 3 (Table 5). In Lysimeter 2, 34% of the initial amount of diuron and 41% of that of linuron was still present in the soil at Day 52 (Table 5), while the total amount of TPs in soil was 37% of the herbicide applied doses. The leaching amount represented less than 0.43%, the volatilized one less than 0.001%, while 27% of the total applied mass was missed.


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  		Table 5. Mass balance of diuron and linuron in Lysimeters 2 and 3.

 
In Lysimeter 3, 13% of the initial amount of diuron and 21% of that of linuron was still present in soil at the end of the experiment (Table 5), while the total TPs amount was 38% of the applied doses. The leaching amount represented less than 0.9% and the volatilized one less than 0.001%, while 41% of the total applied mass was missed.

The mass balance for diuron and linuron showed a missing fraction between 27 and 41%. It may be explained by herbicide mineralization to CO2, by the formation of metabolites not analyzed in the present investigation (i.e., aniline derivates), and partially by the low extraction efficiency of some very hydrophobic compounds, such as chloroanilines, that may be strongly bounded to the soil components.

Mass Balance in the Experimental Plot
The mass balance of the studied herbicides (Table 6) was estimated considering a mean applied dose of 2360 mg for diuron and of 1120 mg for linuron on the field area (mean applied amount x plot area, see Table 2). Considering the mean of the 10 lysimeters (Table 7), the recovery of the compounds in leachate water was 0.29% for parental compounds and 0.005% for metabolites. The total mass of the leachate compounds was calculated as 8.48 mg of diuron, 1.68 mg of linuron, 0.28 mg of DCPMU, 0.03 mg of DCPU, and 0.07 mg of DCA.


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  		Table 6. Mass balance of diuron and linuron in the experimental plot.

 

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  		Table 7. Amounts of herbicides and transformation products (TPs) collected from water samples and percent of herbicide leaching.

 
The total amount of volatilized compounds was estimated as 77.4 x 10–5 µg of linuron. The residual concentrations of diuron, linuron, and their metabolites in soil were calculated from the last sampled soil; 35% of the applied amount was missed. This estimation is similar to that obtained for Lysimeters 2 and 3.

PELMO Application
The PELMO model slightly overpredicted the mobility of linuron, diuron, and DCPMU as found in the experiment (Table 8). The model predicted that these compounds would reach the 10- to 20-cm soil layer by Day 90. In the experiment, DCPMU was detected in soil samples below 10 cm of depth only once (at Day 8) and linuron and diuron were never detected in soil samples below 10 cm. The persistence of linuron, diuron, and DCPMU was well predicted by the model: most predicted concentrations are within a factor of 2 of the measured concentrations. These data confirm the efficiency of the model in predicting the fate of pesticides in soil enclosures such as these small lysimeters, as reported by other authors (Francaviglia et al., 2000; Vanclooster et al., 2000). Although the original purpose of the PELMO model was to help the registration of the pesticide in Europe, our study indicates that it can be easily used to interpret the results from field experimental studies.


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  		Table 8. Comparison of estimated (Pesticide Leaching Model, PELMO) and measured concentrations of diuron and linuron and transformation products (TPs) in soil samples.

 
The model can be used in the calibrated or noncalibrated form to obtain a better understanding, in time and mass, between modeling results and a specific set of real data (Table 8). However, in spite of the fact that PELMO is a capacity model, it did not predict the exact hydraulic dynamic of the different TPs in the soil profile. In fact, although DCPU was detected just once in the experimental field (at Day 8), PELMO predicted measurable concentrations of DCPU throughout the experiment. Moreover, predicted DCA concentrations were generally within an order of magnitude greater than the experimental values.

On the other hand, the experimental and modeling approaches were in agreement if we looked to the total amount of TP compounds: the experimental data showed that the leaching out of the soil columns was less than 1% and according to PELMO, the amount leaching to below 1 m of depth was effectively negligible. Thus linuron, diuron, and their TPs did not appear to be particularly mobile.

Volatilization of diuron and linuron was estimated by PELMO to be 0.6 and 5.8 µg m–2, respectively, much greater than the measured values (Table 3); consequently, in regards to volatilization, there is a need for the development of an advanced model to remove these limitations (Wolters et al., 2002, 2003; Ferrari et al., 2003).


   CONCLUSIONS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
REFERENCES
 
From the results we conclude that the use of soil enclosure lysimeters is a good experimental design for studying mobility and transport processes of herbicides in field conditions. Particularly, they are useful for mass balance calculations. The total mass balance of the applied herbicides in the experimental plot showed that linuron and diuron were very persistent herbicides in soil (50% for diuron and 70% for linuron remaining at the end of the experiment). Fortunately, these compounds were not particularly mobile (<0.5% leaching out of the lysimeters) and they did not volatilize significantly. But, they could generate environmental problems owing to their high persistence, with successive agricultural applications producing an accumulation of residual concentrations in soil and in some situations, such as preferential flow phenomenon, they could be leached to the subsurface soil and to ground water.

The uncalibrated PELMO model slightly overpredicted the mobility, biodegradation, and volatilization of diuron and linuron; and the experimental data showed a greater soil persistence of diuron and linuron compared to the PELMO simulation. On the other hand, the experiment with soil enclosures showed considerable spatial variability of the pesticide distribution during herbicide application and this variability can greatly affect the results of the mass balance calculations. The temporal-spatial variability of the measured mass in the small plot of this experiment can be expected as well in large plots and catchments, increasing the risk assessment uncertainty.


   ACKNOWLEDGMENTS
 
The present study was conducted with the financial support of the Strategic Project Environment and Territory (2000) promoted by the National Research Council of Italy. The authors thank Simona Rullo, Adolfo De Paolis, Luciano Previtali, Federico Ferrari, and Fiorenzo Pozzoni for their technical help in the setting up, sampling, and analyzing operations.


   REFERENCES
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS AND DISCUSSION
 CONCLUSIONS
 REFERENCES
 




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 Appl. Environ. Microbiol.Home page
S. R. Sorensen, C. N. Albers, and J. Aamand
Rapid Mineralization of the Phenylurea Herbicide Diuron by Variovorax sp. Strain SRS16 in Pure Culture and within a Two-Member Consortium
Appl. Envir. Microbiol., April 15, 2008; 74(8): 2332 - 2340.
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