Effect of fluctuating soil humidity on in situ bioavailability and degradation of atrazine

Chemosphere 84 (2011) 369–375

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Effect of fluctuating soil humidity on in situ bioavailability and degradation of atrazine Anastasiah Ngigi a,b, Ulrike Dörfler b, Hagen Scherb c, Zachary Getenga a, Hamadi Boga d, Reiner Schroll b,⇑ a

Department of Physical Sciences, Masinde Muliro University of Science and Technology, Kakamega, Kenya Helmholtz Zentrum München, German Research Center for Environmental Health (GmbH), Institute of Soil Ecology, 85764 Neuherberg, Germany c Helmholtz Zentrum München, German Research Center for Environmental Health (GmbH), Institute of Biomathematics and Biometry, 85764 Neuherberg, Germany d Department of Botany, Jomo Kenyatta University of Agriculture and Technology, Nairobi, Kenya b

a r t i c l e

i n f o

Article history: Received 10 January 2011 Received in revised form 30 March 2011 Accepted 31 March 2011 Available online 4 May 2011 Keywords: Drying–rewetting cycles In situ bioavailability Metabolism Mineralization

a b s t r a c t This study elucidates the effect of fluctuating soil moisture on the co-metabolic degradation of atrazine (6-chloro-N2-ethyl-N4-isopropyl-1,3,5-triazine-2,4-diamine) in soil. Degradation experiments with 14Cring-labelled atrazine were carried out at (i) constant (CH) and (ii) fluctuating soil humidity (FH). Temperature was kept constant in all experiments. Experiments under constant soil moisture conditions were conducted at a water potential of 15 kPa and the sets which were run under fluctuating soil moisture conditions were subjected to eight drying–rewetting cycles where they were dried to a water potential of around 200 kPa and rewetted to 15 kPa. Mineralization was monitored continuously over a period of 56 d. Every two weeks the pesticide residues in soil pore water (PW), the methanol-extractable pesticide residues, the non-extractable residues (NER), and the total cell counts were determined. In the soil with FH conditions, mineralization of atrazine as well as the formation of the intermediate product deisopropyl-2-hydroxyatrazine was increased compared to the soil with constant humidity. In general, we found a significant correlation between the formation of this metabolite and atrazine mineralization. The cell counts were not different in the two experimental variants. These results indicate that the microbial activity was not a limiting factor but the mineralization of atrazine was essentially controlled by the bioavailability of the parent compound and the degradation product deisopropyl-2-hydroxyatrazine. Ó 2011 Elsevier Ltd. All rights reserved.

1. Introduction Pesticides are favourably degraded by microorganisms in soil (Rüdel et al., 1993). But it has to be considered that for most chemicals just a certain amount of these compounds is bioavailable and only this portion can be degraded by microorganisms (Katayama et al., 2010). The microbial breakdown is not only dominated by the activity of the microbes but also by the mass transfer of pesticides to microorganisms (Bosma et al., 1997) and the prevalent water regime in soils (Han and New, 1994). On the microscopic scale, pesticides and pesticide degrading microorganisms are differently distributed in soils, and pesticides must mostly diffuse to the more or less immobile microbes to be metabolized by them. Hampered mass transfer of chemicals to the degrading microbes could therefore be a limiting factor in biodegradation (Bosma et al., 1997). Thus, soil moisture is one of the most important parameters regulating pesticide bioavailability and degradation.

⇑ Corresponding author. Tel.: +49 89 3187 3319; fax: +49 89 3187 3376. E-mail address: [email protected] (R. Schroll). 0045-6535/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2011.03.068

Differences in soil moisture seem to have a more intensive effect on the degradation of chemicals than differences in soil temperature. Degradation half-life of isoproturon increased by a factor of 10–15 when soil water potential was reduced from 660 MPa to 56 kPa in soil material of a clay-enriched illuvial horizon, whereas a change in soil temperature from 10 °C to 22 °C was of minor importance (Alletto et al., 2006). In laboratory experiments, a soil water potential of 15 kPa was identified for optimal pesticide mineralization (Ilstedt et al., 2000; Schroll et al., 2006). In contrast to laboratory studies, where soil humidity in most cases is constant, under outdoor conditions the soil is exposed to varying water regimes. Depending on the soil properties, precipitation events, air temperature, and uptake by plants, soil moisture will vary to a great extent during the vegetation period. These variations in soil water will have a considerable effect on sorption behaviour of pesticides, their bioavailability, and finally their transformation and degradation. Diverse and even contrary results of drying–rewetting are reported in literature, varying from favouring, inhibiting, or no effects of soil drying on the sorption and degradation behaviour of pesticides in soils. García-Valcárcel and Tadeo (1999) found lower concentrations of hexazinone and simazine in soil solution after several

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drying–rewetting cycles (from 90% to 20% field capacity) in a sandy loam soil. In five soils ranging from sandy loam to clay, the adsorption of the herbicide imazaquin was increased when the soils were dried to 50%, 25% or 0% of field capacity with subsequent returning to field capacity (Goetz et al., 1986). The increased imazaquin adsorption was explained by reduction in water film thickness coating the soil minerals and thus favouring the access of imazaquin to active surfaces. The effect of drying and rewetting cycles on the desorption behaviour of diuron and terbuthylazine was studied in detail by Lennartz and Louchart (2007). They performed classical batch experiments and adsorption was followed by one to three drying cycles (<3% water content) before conducting the desorption experiments. Drying resulted in decreased desorption of the pesticides. The authors suggest that drying cause structural modifications of the soil organic matter due to shrinking-like processes, which impede diffusion and the pesticide molecule is trapped in micropore spaces. Summarizing these findings, the authors found a higher sorption of the selected pesticides after soil drying. White et al. (1998) found inhibiting as well as favouring as well as no effects of drying–wetting cycles (‘‘dried to constant weight’’ and ‘‘remoistened to 1.1 bar’’) on the mineralization of phenanthrene and di(2-ethylhexyl)phthalate, depending on the incubation and aging periods for both compounds. In contrast to the upper cited literature, the following findings would suggest a potentially higher bioavailability and degradation of pesticides under FH conditions. Belliveau et al. (2000) postulated that the intraparticle diffusion of a dissolved pesticide must be preceded by the intraparticle diffusion of water. In laboratory experiments, Belliveau et al. (2000) and Gamble et al. (2000) could demonstrate that after drying the mobility of water in an organic rich soil was reduced and only very slow. The mobility of the pesticides 2,4-D and atrazine was comparable to the slow water penetration (Belliveau et al., 2000). The aim of the present study was to further elucidate the impact of FH conditions on pesticide bioavailability, pesticide degradation and pesticide mineralization in an agricultural soil. The pesticide selected for this study was atrazine (6-chloro-N2-ethyl-N4-isopropyl-1,3,5-triazine-2,4-diamine), which was used globally in the past to control pre- and post-emergence broadleaf and grassy weeds in major crops. Atrazine can be very persistent in soil (Jablonowski et al., 2009) and it is by far the most frequently found xenobiotica in ground water and increasing restrictions of its use have been introduced (Premazzi and Stecchi, 1990).

2. Materials and methods 2.1. Soils The soil material was a humic cambisol from an agricultural field (Kelheim; latitude 48.917°, longitude 11.867°, altitude 348 m) in Germany without atrazine history for the past 20 years. Therefore, we could expect to select a soil with co-metabolic and thus non-growth-linked degradation dynamic; using soil material with a growth-linked degradation dynamic might have complicated the interpretation of the results. Moreover, co-metabolic degradation dynamic is still the rule but not the exception in agricultural soils (Krutz et al., 2010). The relevant soil characteristics are: clay (<2 lm) 11%, silt (2–63 lm) 19%, sand (63 lm–2 mm) 70%, org. C 1.3%, total N 0.1%, pH (CaCl2) 6.9, and water content of 18.1% at a water potential of 15 kPa and a soil density of 1.3 g cm 3. Before the experiments were started, the soil samples (depth 0–10 cm) were sieved (<2 mm) and kept at room temperature (20 ± 1 °C) for 5 d after moistening to a water potential close to but below 15 kPa.

2.2. Chemicals 14

C-ring-labelled atrazine, with a specific radioactivity of 351.5 MBq mmol 1 and a radiochemical purity of >98% was purchased from Sigma–Aldrich (St. Louis, MO, USA). Non-labelled atrazine and the metabolite standards deethylatrazine (DEA), deisopropylatrazine (DIA), deethyl-deisopropylatrazine (DEDIA), 2-hydroxyatrazine (OH-ATR), deethyl-2-hydroxyatrazine (OHDEA), deisopropyl-2-hydroxyatrazine (OH-DIA), and deethyldeisopropyl-2-hydroxyatrazine (OH-DEDIA) were obtained from Ehrenstorfer (Augsburg, Germany). All chemicals had a purity of >99%. Scintillation cocktails were obtained from Packard (Dreieich, Germany). All other chemicals and solvents were of analytical grade and were purchased from Merck (Darmstadt, Germany). 2.3. Degradation experiments 2.3.1. Pesticide application Atrazine degradation experiments were carried out in an aerated closed laboratory system as described previously (Schroll and Kühn, 2004) with 50 g soil (dry weight; d.w.) in 100 mL double walled glass incubation vessels. 14C-labelled and non-labelled atrazine were mixed and dissolved in methanol to give a final specific radioactivity of 66.3 Bq lg 1. A volume of 0.5 mL of this application standard with a radioactivity of 33.1 kBq was applied dropwise with a Hamilton syringe to an aliquot of 3.5 g of oven-dry, grinded soil. The aliquot was stirred with a spatula for 2 min until a homogenous distribution of the herbicide was achieved. After evaporation of methanol the aliquot was mixed for another 2 min with fresh soil (46.5 g dry weight) yielding a total sample amount of 50 g dry soil per experiment with a pesticide concentration of 10 mg kg 1. This concentration corresponds to a realistic agronomic application rate of 500 g ha 1 when assuming that the distribution of the herbicide might be at 3–4 mm depth shortly after the application in a field with a soil density of approximately 1.3 g cm 3. The soils were then transferred to the incubation flasks, compacted to a density of 1.3 g cm 3 and water content was adjusted to a water potential of 15 kPa to obtain maximum pesticide mineralization at CH conditions (Schroll et al., 2006). 2.3.2. CH and FH experimental set up For CH conditions the incubation vessels were placed in the dark at 20 ± 1 °C and connected to a trapping system. Three times per week humidified air (1.0 L h 1) was drawn via a pump through the system for 1 h. After passing through the flasks, the air was trapped in a series of four wash bottles filled with 10 mL 0.1 M NaOH to trap 14CO2 from mineralization process. After each aeration the trapping solution was collected and the traps were filled with fresh 0.1 M NaOH solution. The weight of the incubation vessels was monitored gravimetrically every week to control for any water loss. For experiments with FH conditions, the incubation vessels were placed in the dark at 20 ± 1 °C and connected to a trapping system. For the drying cycle non-humidified air (3.0 L h 1) was continuously drawn via a pump through the system for one week. The air, having passed through the flasks, was sampled and trapped as it was described above. During the drying cycles the incubation flasks were weighed daily to monitor the water loss gravimetrically. After one week of drying, the rewetting of the soil was done immediately after weighing the incubators to adjust the water content to the original value corresponding to a water potential of 15 kPa. After rewetting, the next drying cycle was started. In total, eight drying and rewetting cycles were conducted. From the collected 0.1 M NaOH solutions of both variants, 2 mL aliquots were taken and mixed with 3 mL of scintillation cocktail Ultima Flo AF (Packard, Dreieich, Germany). Radioactivity

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was detected by scintillation counting using a TriCarb 2800 TR (Packard, Dreieich, Germany). The detection limit for 14CO2 was 0.05 lg kg 1 soil (d.w.). 2.3.3. Soil sampling After every two rewetting cycles, exactly 14 d, four replicates from the experiments with CH and FH were sampled. Thirty grams of each soil sample (d.w.) were used for PW extraction to determine the pesticide bioavailability (Folberth et al., 2009). Subsequently, these soil samples were extracted with methanol to determine the quality and quantity of extractable residues as well as NER. Another aliquot (1 g of each soil sample) was taken directly from the incubators and was used for counting the cultivable bacterial cells. 2.4. Pesticide in situ bioavailability In situ bioavailability of atrazine was determined using the PW extraction method as described by Folberth et al. (2009). In brief, 30 g (d.w.) of soil was transferred to a custom-built centrifuge container. This container consists of an upper and lower part. The upper cup contained the soil sample to be extracted. In the lower part, PW was collected. Both pieces were connected by a canal. A glass frit with 120 lm average pore size was placed underneath the soil sample to prevent the canal from being clogged by soil particles. The top of the upper cup was sealed with aluminum foil and closed with an aluminum cap. Centrifugation was carried out using a Beckman J2-21 centrifuge and a Beckman JA-14 rotor (Beckman, Krefeld, Germany) at 21 °C for 90 min with a relative centrifugal force of 9000g. Radioactivity in the extracted water was measured by liquid scintillation counting using 4 mL of the scintillation cocktail Ultima Gold XR (Packard, Dreieich, Germany) and 0.1 mL PW aliquot to quantify the dissolved pesticide concentration in soil solution. The total dissolved pesticide amount was calculated by multiplication of the concentration of the aliquot with the total amount of water in the soil (Folberth et al., 2009). The PW samples were stored at 20 °C before HPLC analysis. 2.5. Solvent extraction of soil, clean up and analysis After centrifugation the same soil aliquots were extracted with methanol in an accelerated solvent extractor (ASE 200, Dionex, Idstein, Germany) at 90 °C, with a pressure of 10 MPa (Gan et al., 1999). Aliquots of 0.5 mL of each extract were mixed with 4.5 mL Ultima Gold XR and measured by liquid scintillation counting. Subsequently, extracts were concentrated with a rotary evaporator to a volume of 2–3 mL. The concentrated methanolic soil extracts were dissolved in 250 mL distilled water. These solutions as well as samples of PW were cleaned up with Isolate Triazine columns (500 mg, Separtis, Grenzach-Wyhlen, Germany). After extraction, the SPE columns were dried under a gentle nitrogen-stream and eluted with 10 mL methanol. The eluate was concentrated to a volume of 1 mL with a rotary evaporator and further concentrated to a volume of 0.2 mL under a gentle nitrogen-stream. The samples were immediately analyzed by HPLC or stored at 20 °C before analysis. For residue analysis 20 lL of each soil extract or PW sample were injected to a HPLC system that was equipped with a L-6200 Intelligent Pump (Merck-Hitachi, Darmstadt, Germany), a UV/VIS detector (220 nm, Merck-Hitachi, Darmstadt, Germany), and a radioactivity detector LB 506 C1 (Berthold, Wildbad, Germany); column: LiChrospher 100 RP-18, 2 lm, 4  250 mm (MerckHitachi, Darmstadt, Germany). The mobile phase consisted of 0.003 M KH2PO4, pH 3 (A) and acetonitrile (B) at a flow rate of 0.8 mL min 1. The gradient program was: T0 min 20% (A); T10 min 38% (A); T24 min 75% (A); T29 min 75% (A); T33 min

20% (A); T40 min 20% (A). Parent compound and metabolites were identified by comparison of their retention times with reference substances. The method detection limits – based on radioactivity detection – were as follows: atrazine: 23.5 lg kg 1 soil (d.w.); DEA: 20.5 lg kg 1; DIA: 18.9 lg kg 1; OH-DIA: 16.9 lg kg 1. 2.6. Quantification of

14

C-labelled NER

After ASE, soil material was dried and homogenized intensively. Three aliquots of each soil sample were filled into combustion cups and mixed with 3–4 drops of saturated aqueous sugar solution to guarantee a complete oxidation of the 14C. The combustion was conducted with an automatic sample-oxidizer 306 (Packard, Dreieich, Germany). 14CO2 was trapped in Carbo- Sorb E (Packard, Dreieich, Germany) and mixed with Permafluor E (Packard, Dreieich, Germany) prior to scintillation counting. The detection limit for NER was: 9.0 lg kg 1 soil (d.w.). 2.7. Bacterial cell counting For extraction of bacterial cells from soil, the following solution was used: 0.1 g NaCl, 0.02 g CaCl
2H2O, 0.2 g MgSO4
7H2O, 5.0 g Tween 80. This solution was adjusted to 1 L with Milli Q water and autoclaved for 20 min at 121 °C. Soil bacteria were extracted from the soil by mixing 1 g fresh soil with 99 mL of the extraction solution in a 200 mL jar. The mixture was shaken vigorously at 150 rpm for 1 h. The soil particles were then allowed to sediment for 10 min before several dilution steps were conducted. Dilutions of 10 2 were spread in duplicates on plates with LB medium. The LB was prepared by mixing 10 g trypton enzymatic digest from casein, 5 g yeast extract, 5 g NaCl, 15 g Agar and 0.1 mg cycloheximide (per 1 L Milli Q water). The mixture was sterilized for 20 min at 121 °C. The number of colony forming units (CFU) was determined after 3 d of incubation at 25 °C. 2.8. Data analysis For statistical data analyses, the program package SAS 9.1 was used to set up regression models for trend analyses with procedure REG. Time varying treatment effects were modelled as main effects, eventually adjusted for interaction of independent treatment variables with time (days). For ordinary or inverse-variance weighted mean value comparisons, procedure TTEST was used. (SAS Institute Inc: SAS/STAT User’s Guide, Version 9.1. Cary NC: SAS Institute Inc.; 2003). 3. Results and discussion 3.1. Soil water potential In the CH experimental set up, the soil water potential was kept constant at 15 kPa, whereas in the FH experimental set up, the water potential varied between 15 kPa and around 200 kPa at each drying and rewetting cycle (Fig. 1). Thus, the water regime in the various experimental set ups was quite different and the selected drying method was appropriate in achieving an efficient and reproducible reduction in soil humidity under laboratory conditions. 3.2. Mineralization of atrazine The cumulative mineralization at the end of the investigation period of 56 d was relatively low for both, CH and FH (Fig. 2). Under constant water regime 3.4% and under fluctuating water regime 5.1% of the applied pesticide was mineralized. Since the soil had

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soil water potential (kPa)

300 250 200 150 100 50 0 0

10

20

30

40

50

60

Time (d)

14

cum. mineralization (% of appl. C)

Fig. 1. Water tension of the soil during the drying–rewetting cycles. Bars indicate standard deviation of 16 (beginning) to four (end) samples.

6 constant soil humidity (CH)

5

fluctuating soil humidity (FH)

4 3 2 1 0 0

10

20

30

40

50

60

Time (d) Fig. 2. Cumulative mineralization of 14C-atrazine in a humic cambisol under constant (CH) and fluctuating soil humidity (FH). Bars indicate standard deviation of 16 (beginning) to four (end) samples, respectively.

no history of atrazine use for the past 20 years, the low mineralization is a characteristic of a co-metabolic biodegradation. However, the variant FH showed significantly (p < 0.0001) higher mineralization compared to the control (CH). For a further explanation of the mineralization behaviour of atrazine, every two weeks four replicates were sampled and analyzed for atrazine and its transformation products (see Section 2.3.3). At each sampling time, mass balances were established. The recoveries were quite good and ranged between 96.1% and 96.6% of the applied 14C in the variants CH and FH, respectively. 3.3. Identification of the factors governing mineralization of atrazine To elucidate the degradation pathway and to identify the factors that govern the mineralization of atrazine in soil humic cambisol, the amounts of extractable atrazine, atrazine metabolites, NER, and the mineralized atrazine were compared at the four sampling points (days 14, 28, 42 and 56). Several significant correlations could be identified. 3.3.1. Relationship between total atrazine and NER As can be seen from Table 1, NER levels are increasing for both CH and FH variants from sampling to sampling, while the respective atrazine levels are deceasing in both cases. To test whether there is any relationship between total atrazine and the formation of NER, we calculated respective correlations between both values. There exist significant correlations between the totally extractable atrazine (atrazine in methanol extract plus atrazine in PW) and the formation of NER for both variants, CH (p = 0.0124) and FH (p = 0.0114) (Fig. 3). Both regressions were not significantly different from each other as neither the main effect of CH vs. FH

(p = 0.5762) nor the interaction of atrazine with FH (p = 0.6615) were significant, and therefore data from both variants could be calculated in a common regression (y = 0.70x + 5.35; R2 = 0.98), meaning that the processes leading to NER formation are identical for both variants. Nevertheless, a remarkable difference between both variants was identified: in the variant FH atrazine was higher and NER were lower than in the control (CH) at each respective sampling point (Table 1 and Fig. 3). This is reflected by significantly different inverse-variance weighted mean values of atrazine for CH (3.1, SE = 0.2) and for FH (4.1, SE = 0.3) (p = 0.0362). Belliveau et al. (2000) have shown that after drying soil samples the penetration rate of both water and pesticide into micropores and strong binding sites is reduced, thereby impeding the formation of NER. The lower NER formation under FH conditions could be explained by these mechanisms. Since the NER formation was apparently retarded in variant FH, more atrazine was potentially available for other processes, like mineralization. In fact, the FH variant showed higher mineralization than the control: in Fig. 2, the increase of cumulative mineralization is significantly (p < 0.0001) accelerated under FH conditions compared to CH conditions. In general, the formation of NER in both variants increased considerably with time, which is in accordance with former results (Clay and Koskinen, 1990). 3.3.2. Relationship between atrazine and its metabolites Atrazine is stepwise degraded until it is finally mineralized. In methanol extracts and in PWs we found the degradation products DIA, DEA, and OH-DIA (Table 1). Depending on the extraction procedure we found significant correlations (i) for total atrazine and total DIA in CH (Fig. 4a, p = 0.0148) and (ii) for methanol extractable atrazine and DIA in FH (Fig. 4b, p = 0.0287). The other two correlations (see Fig. 4a and b) were not significant, and this was most likely caused by the relative small amount of data pairs. As can be seen from Fig. 4a and b, with decreasing atrazine concentration, the concentration of DIA increased. Since the slope of the regression lines were not statistically different (p > 0.3), it can be concluded that the degradation processes and transformation rates – forming DIA from atrazine – are identical for both variants. Under CH conditions, DIA was formed to a lower extent than in the soil under FH conditions, which was caused by the lower atrazine concentration in the soil with CH regime. Thus, the DIA formation rate was limited by the availability of atrazine in the control (CH). N-dealkylation of s-triazines is an important degradation pathway in many microorganisms, and Behki and Kahn (1986) found that the formation of deisopropylatrazine is favoured over the formation of deethylatrazine. In our study, we also found much higher total DIA residues than total DEA residues (Table 1). 3.3.3. Relationship between DIA, OH-DIA and mineralized atrazine The intermediate product DIA was compared with the amount of mineralized atrazine. For each 14-d sampling interval, the individual 14-d-mineralization-rates were calculated for both variants; for this purpose the measured 14CO2 radioactivity was transformed via the respective molecular weight into atrazine equivalents and correlated with the total DIA residues at each sampling point. Since the results above showed in both variants identical processes by which atrazine is degraded to DIA, the following correlations were calculated with the common data pool from both variants. Fig. 5a shows a significant positive correlation (p = 0.0045) between DIA and pesticide mineralization. Most of the above calculated correlations were conducted with total residues of atrazine and DIA in methanol extract plus PW: these total residues represent the potentially available atrazine and DIA residues. Since we found a significant correlation between DIA and pesticide mineralization, and since microbial degradation

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Table 1 Concentrations of atrazine and degradation products in methanol (MeOH) extract and pore water (PW), cumulative mineralization and non extractable residues (NER) in soils with constant (CH) and fluctuating moisture (FH) conditions at 4 different sampling times: after 14 days (T14), 28 days (T28), 42 days (T42), 56 days (T56). n = 4 ± standard deviation. Pore water: HPLC analysis was conducted with pooled samples from 4 replicates to exceed detection limit, therefore no standard deviation can be given. Moisture condition

T14

CH

T42

T56

MeOH extract

PW

MeOH extract

PW

MeOH extract

PW

MeOH extract

PW

Atrazine (lg g 1) Deisopropyl-2-hydroxyatrazine (lg g Deisopropylatrazine (lg g 1) Deethylatrazine (lg g 1)

4.53 ± 0.30 0.05 ± 0.04 0.60 ± 0.14 0.07 ± 0.06

0.96 0.17 0.11 0.37

3.80 ± 0.18 0.09 ± 0.05 0.80 ± 0.12 0.01 ± 0.01

0.04 0.14 0.06 0.52

3.05 ± 0.04 0.03 ± 0.01 0.89 ± 0.02 n.d.

0.06 0.23 0.06 0.75

2.61 ± 0.09 0.01 ± 0.03 0.84 ± 0.06 n.d.

0.11 0.21 0.13 0.55

5.09 ± 0.15 0.09 ± 0.05 0.69 ± 0.15 0.04 ± 0.03

0.86 0.15 0.10 0.33

4.30 ± 0.04 0.07 ± 0.04 0.77 ± 0.10 n.d.

0.08 0.20 0.09 0.70

3.60 ± 0.04 0.05 ± 0.03 0.92 ± 0.04 n.d.

0.12 0.27 0.10 0.92

3.31 ± 0.21 0.05 ± 0.01 1.01 ± 0.05 n.d.

0.05 0.26 0.08 0.81

Atrazine (lg g 1) Deisopropyl-2-hydroxyatrazine (lg g Deisopropylatrazine (lg g 1) Deethylatrazine (lg g 1)

FH

T28

Atrazin and degradation products 1

)

1

)

T14

T28

T42

T56

CH

Mineralization and NER Cum. mineralization (lg g NER (lg g 1)

1

0.03 ± 0.00 1.6 ± 0.2

0.08 ± 0.01 2.5 ± 0.1

0.19 ± 0.01 3.3 ± 0.1

0.34 ± 0.02 3.6 ± 0.1

FH

Cum. mineralization (lg g NER (lg g 1)

1

0.05 ± 0.00 1.3 ± 0.1

0.11 ± 0.01 2.1 ± 0.1

0.26 ± 0.03 2.8 ± 0.0

0.51 ± 0.05 3.0 ± 0.0

)

)

(a) 1.2

4.0 3.5 3.0

-1

y = -0.66x + 5.18 2 R = 0.98 CH FH

2.5 2.0

y = -0.72x + 5.46 2 R = 0.98

1.5 1.0 0.5 0.0 0.0

1.0

2.0

3.0

4.0

5.0

deisopropylatrazine (µg g )

-1

non extractable residues (µg g )

n.d. = not detectable.

6.0

7.0

y = -0.11x + 1.42 2 R = 0.88

1.1 1.0

CH FH

0.9 0.8

y = -0.10x + 1.24 2 R = 1.0

0.7 0.6 0.5 0.0

-1

atrazine (µg g )

1.0

2.0

3.0

4.0

5.0

6.0

7.0

-1

atrazine (µg g )

3.3.4. Relationship between bacterial cell counts and mineralization There were no significant differences in the cell counts for the constant and fluctuating water regimes (p > 0.10). No direct rela-

-1

occurs preferably in the water phase, we analyzed the relationship between the metabolites in PW and mineralized atrazine. Here we found a significant correlation (p = 0.0173) between OH-DIA and mineralization (Fig. 5b). Dechlorination of DIA to the respective hydroxylated compound OH-DIA was demonstrated by Behki and Kahn (1986). OH-DIA is an important intermediate product in atrazine mineralization since ring cleavage apparently occurs only after hydroxylation (Kaufman and Kearney, 1970). This metabolite is a polar compound and was detected almost exclusively in the PW fraction (Table 1). Therefore, the method of in situ PW extraction gave a valuable insight into the degradation mechanisms of atrazine, and we could show that higher atrazine mineralization at FH was caused by higher levels of OH-DIA in comparison to the control (CH). From the detected correlations we suggest that the pathway ATRAZINE – DIA – OH-DIA is an important route in the breakdown of atrazine in soil humic cambisol. Although we could not elucidate the complete pathway of atrazine degradation in this study, we could identify the main intermediate products, which play an important role in the mineralization of atrazine in this soil.

(b) 1.1 deisopropylatrazine (µg g )

Fig. 3. Correlations between total atrazine residues and non-extractable residues (calculated in atrazine-equivalents) in a humic cambisol under constant (CH) and fluctuating soil humidity (FH). Bars indicate standard deviation of four samples.

y = -0.18x + 1.56 2 R = 0.96

1.0 CH FH

0.9 0.8

y = -0.14x + 1.25 2 R = 0.77

0.7 0.6 0.5 0.0

1.0

2.0

3.0

4.0

5.0

6.0

-1

atrazine (µg g ) Fig. 4. Correlations between atrazine and deisopropylatrazine residues in a humic cambisol under constant (CH) and fluctuating soil humidity (FH). Bars indicate standard deviation of four samples: (a) for total atrazine and total deisopropylartazine, and (b) for MeOH extractable atrazine and MeOH extractable deisopropylatrazine.

tionship could be established between the cell counts and the mineralization of atrazine. This shows that the mineralization of atrazine is limited by the availability of atrazine and its degradation products in soil solution rather than by the bacterial activity.

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0.4

-1

mineralized atrazine (µg g )

(a)

0.3 y = 0.51x - 0.31 2 R = 0.76

0.2

0.1

0.0 0.4

0.5

0.6

0.7

0.8

0.9

1.0

1.1

-1

deisopropylatrazine (µg g )

0.3

-1

mineralized atrazine (µg g )

(b)

0.2

y = 1.24x - 0.15 2 R = 0.64

0.1

0.0 0.1

0.2

0.3 -1

deisopropyl-2-hydroxyatrazine (µg g ) Fig. 5. Correlations between deisopropylatrazine (DIA) residues, deisopropyl-2hydroxyatrazine (OH-DIA) residues and mineralized atrazine in a humic cambisol under constant and fluctuating soil humidity: (a) correlation between total DIA and mineralized atrazine. Bars indicate standard deviation of four samples respectively, and (b) Correlation between OH-DIA in soil pore water and mineralized atrazine.

3.4. Influence of FH conditions on the bioavailability of atrazine and metabolites In the soil with FH conditions, higher residues of atrazine were detected, indicating a higher bioavailability of atrazine under such conditions. At the sampling day 14 (T14) a 10–20-fold higher atrazine concentration was detected in the PW fraction as compared with the later sampling times (Table 1). This could be a hint that the in situ adsorption process was not at equilibrium after 14 d and that atrazine adsorption proceeds very slowly under such conditions. Gamble et al. (2000) suggest that the sorption of chemicals can be described by a two-step model: a rapid equilibration at the surface of soil particles followed by a slow diffusion into the soil particles where they are more strongly bound. Applying this theory to our results, this would mean that at least the second step was not accomplished after 14 d. In literature it is reported that drying–rewetting cycles lead to higher sorption, lower bioavailability, and lower biodegradation of organic chemicals caused by soil structure modifications, which favours pesticide adsorption and/or impedes desorption (Goetz et al., 1986; White et al., 1998; García-Valcárcel and Tadeo, 1999; Lennartz and Louchart, 2007). These findings and explanations could not be confirmed by our results. Most likely one of the reasons for these apparent discrepancies between our results and results from the literature are the varying humidity conditions in the soils of the various experiments. Some effects (e.g. shrinking and swelling) mentioned in the literature were found when soils were intensively dried, which was not the case in our study. Moreover, the direct comparability of the various experiments and re-

sults is hampered because some authors expressed water content in percentage and others in soil water potential. For a better comparison we would suggest to use water potential data for describing water availability in soils. This would enable a direct comparability of pesticide in situ bioavailability in different soils. Using an identical soil density will also help to harmonize experimental conditions e.g. when in situ sorption, desorption or turnover of chemicals in soils is studied (Schroll et al., 2006). In our study soil water potential and thus soil water content was varied during the drying–rewetting processes (Fig. 1). Diffusion of solutes (e.g. nutrients, pesticides) in soils is directly related to the cross section for flow (Papendick and Campbell, 1981) and thus the lower water contents during the drying period most likely reduce the diffusion of atrazine to sorption sites. Thus, more atrazine is available for microbial processes leading to a higher degradation to intermediate products and finally to a higher mineralization in the soil with FH regime (Fig. 2). This hypothesis is supported by the studies of Gamble et al. (2000) and Belliveau et al. (2000). In laboratory experiments Gamble et al. (2000) demonstrated that water considerably influences the interactions of organic chemicals with both the surfaces and the interiors of soil particles. By using magnetic resonance imaging Belliveau et al. (2000) could affirm the connection between water uptake and atrazine uptake in an organic rich soil. After wetting of an air-dried soil, some of the water penetrated only very slowly into the soil to the strong binding sites. Parallel studies on atrazine adsorption revealed a little bit slower mobility of atrazine to the micropores as it was observed for water. This shows that in general drying and rewetting can impede the diffusion of water and pesticide to the adsorption sites resulting in a higher bioavailability of the organic chemicals.

4. Conclusion The increased co-metabolic mineralization of atrazine in soil humic cambisol under FH conditions was due to a higher bioavailability of atrazine and its metabolites under these conditions. The method of in situ sampling of soil PW turned out to be a valuable tool for elucidation of the degradation mechanisms of atrazine. With help of this tool deisopropyl-2-hydroxyatrazine was identified as an important intermediate product in atrazine degradation and mineralization in this humic cambisol. Since slow atrazine adsorption was observed in this soil and since the increased atrazine mineralization under FH regime was most likely due to the sorption behaviour, in situ long term sorption studies should be conducted for a better understanding of such complex interactions in soils.

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