|Udskriftsvenlig version af siden:|
"Pesticide leaching at JYNDEVAD"
Udskriftsdato: mandag 19. marts 2018 kl. 4:26
Sidst ændret: 4. juli 2003
© Pesticidvarsling pesticidvarsling.dk
Chapter 3, PLAP - Monitoring results May 1999 - June 2002
3.1 Materials and methods
3.1.1 Site description and monitoring design
Jyndevad is located in southern Jutland (Figure 1). The test site covers a cultivated area of 2.4 ha (135 x 184 m) and is practically flat. A windbreak borders the eastern side of the test site. The area has a shallow groundwater table ranging from 1 to 2 m b.g.s. The overall direction of groundwater flow is towards northwest (Figure 13). The soil can be classified as Arenic Eutrudept and Humic Psammentic Dystrudept (Soil Survey Staff, 1999) with coarse sand as the dominant texture class and topsoil containing 5% clay and 1.8% total organic carbon. The geological description points to a rather homogeneous aquifer of meltwater sand, with local occurrence of thin clay and silt beds (Figure 14). A brief description of the sampling procedure is provided in Appendix 2. The monitoring design and test site are described in detail in Lindhardt et al. (2001) and the analysis methods in Kjær et al. (2002).
3.1.2 Agricultural management
The field was sprayed with glyphosate on 22 September 1999 prior to the sowing of winter rye (cv. Dominator) on 13 October. Weeds were sprayed on 12 November using tribenuron methyl. At the same time, potassium bromide tracer was applied. Fungicide spraying was carried out twice on 5 May and 7 June, each time using propiconazole and fenpropimorph. On 6-7 May the site was irrigated with 29 mm/ha. The winter rye was harvested on 9 August, yielding 56.2 hkg/ha of grain (water content 15%), approximately 5 hkg/ha less than average for the location.
On 24 April 2001, 49 tonnes/ha of cattle slurry was spread and incorporated. The field was ploughed two days later and sown with maize (cv. Loft) on 30 April. Herbicide spraying with terbuthylazine + pyridate was carried out on 14 May and on 30 May. The site was irrigated twice with 31 mm/ha on 4-5 July and 30 mm/ha on 23-24 July. The maize was harvested on 1 October yielding 151.4 hkg/ha (100% dry matter) cobs and stalks.
The 2002 crop was potato (cv. Oleva) for starch production sown on 22 April. Before the potatoes emerged the field was treated with metribuzin to combat weeds on 13 May. Weeds were sprayed with rimsulfuron on 23 May, at which time the potatoes had just emerged. The field was irrigated with 20 mm/ha on 13 June and 25 mm on 12 August. Fungicide spraying was carried out 10 times between 18 June and 20 August, each time using fluazinam at a rate of 0.2 l Shirlan/ha. The potatoes were harvested on 24 September with a tuber yield of 515.8 hkg/ha, equivalent to 118.8 hkg/ha (100% dry matter) and slightly less than the average for that year. It should be noted that neither metribuzin nor fluazinam are included in the monitoring programme. Management practice at the site is detailed in Appendix 3, Table A3.2.
Figure 13. Overview of the Jyndevad test site. The innermost white area indicates the cultivated land, while the grey area indicates the surrounding buffer zone. The positions of the various installations are indicated, as is the direction of groundwater flow (by an arrow).
Figure 14. Geological description of the Jyndevad site (Lindhardt et al ., 2001).
3.1.3 Model set-up and calibration
The MACRO model was applied to the Jyndevad site covering the soil profile to a depth of 5 m b.g.s., always including the groundwater table. The model was used to simulate water flow and bromide transport in the unsaturated zone during the full monitoring period July 1999-June 2002.
The model was calibrated to the observed groundwater table measured in the piezometers located in the buffer zone, as well as to measured time series of soil water content at three different depths (25, 60 and 110 cm b.g.s.) from the two profiles S1 and S2 (see Figure 13) and to the bromide concentration measured in the suction cups located 1 m b.g.s. The calibration procedure is briefly described in Section 2.1.3. For a detailed description of data acquisition, model set-up and calibration procedures, see Kjær et al. (2002).
3.2 Results and discussion
3.2.1 Soil water dynamics and water balances
The model simulations were generally consistent with the observed data, thus indicating a good model description of the overall soil water dynamics in the unsaturated zone. The model provides a good simulation of the fluctuations in the measured groundwater table well. The dynamics of the measured soil water saturation was improved during calibration, especially 0.6 m b.g.s. (Figure 15D). The model still has some difficulty in capturing the degree of the soil water saturation 1.1 m b.g.s., however (Figure 15E). One explanation could be the large inter-probe variation in this horizon (data not shown), probably due to local variations in the texture of the sand in which the probes are installed.
Table 5. Annual water balance for Jyndevad (mm/yr). Precipitation is corrected to the soil surface according to the method of Allerup and Madsen (1979).
At Jyndevad the three monitoring years differed as regards annual water balance, the first year being normal, the second dry and the third wet. The simulated actual evapotranspiration varies only slightly compared to the precipitation input, thus resulting in large variation in the modelled groundwater recharge. During the summer months the actual evapotranspiration usually counterbalances the precipitation, but periods with heavy precipitation events might result in percolation to 1 m b.g.s., thus minimizing the periods without percolation.
Figure 15. Soil water dynamics at Jyndevad: Locally measured precipitation and simulated percolation 1 m b.g.s. (A), simulated and measured groundwater level (B), and simulated and measured soil water saturation (SW sat.) at three different soil depths (C, D and E). The measured data in B derive from piezometers located in the buffer zone. The measured data in C, D and E derive from TDR probes installed at S1 and S2 (see Figure 13).
3.2.2 Bromide leaching
The autumn application of bromide was followed by high autumn precipitation with a resultant high level of infiltration and rapid leaching of bromide. The bromide concentration 1 m b.g.s. thus increased rapidly just one month after application. All of the bromide had leached from the uppermost metre of the soil about four months after application (Figure 16). The model was able to satisfactorily simulate the bromide transport at 1 m b.g.s. Both the timing and the concentration level, as indicated by the measurements, were well captured by the model (Figure 16). The model predicts a high peak concentration between the measurements in the breakthrough curve, but no monitoring data are available to confirm this.
As the suction cups located 2 m b.g.s. are mostly in the saturated zone during the winter, no attempt has been made to calibrate the model towards the measured bromide concentrations at this depth. Nevertheless, the breakthrough of bromide at 2 m b.g.s. was detected in both suction cups two months after application. The bromide concentration remained elevated until January 2002, indicating an overall transport time of 27 months from field application until the majority of the bromide had passed the suction cups (2 m b.g.s.).
Marked breakthrough of bromide was also detected in all downstream monitoring wells, with the results indicating rather homogeneous groundwater flow. Elevated bromide concentrations were thus detected in all downstream monitoring wells around July, with the temporal evolution being somewhat similar (Figure 17). The area around M3 was characterized by a more heterogeneous flow pattern, however. The bromide concentration in the upper screen of M3, located 2-3 m b.g.s., was thus only slightly elevated, while transport of the majority of the bromide took place at lower depths. Silt and clay lenses located in the upper three meters of M3 (Figure 14) may determine the flow pattern. At the end of the monitoring period the bromide concentration in the deepest filter at M5 was slightly elevated, possibly due to the groundwater flow turning slightly northwards during the last part of the monitoring period.
Figure 16. Simulated (red line) and measured (blue and green crosses) bromide concentration at Jyndevad. The data derive from suction cups installed 1 m b.g.s. and 2 m b.g.s. at locations S1 and S2 (see Figure 13). The green vertical line indicates the date of bromide application.
Figure 17. Bromide concentration in the groundwater at Jyndevad. The data derive from monitoring wells M1-M7. Screen depth is indicated in m b.g.s. The green vertical line indicates the date of bromide application.
3.2.3 Pesticide leaching
At Jyndevad, the monitoring encompassed seven different pesticides and several degradation products applied during three growing seasons as indicated in Figure 18 and Table 6. It should be noted that precipitation in Table 6 is corrected to the soil surface according to Allerup and Madsen (1979), whereas percolation (1 m b.g.s.) refers to accumulated percolation as simulated with the MACRO model It should also be noted that as tribenuron methyl (applied here as Express) and pyridate (applied here as Lido) degrade rapidly, the leaching risk is more associated with their respective degradation products, triazinaminmethyl and PHPC. For the same reasons it is the degradation products and not the parent compounds that are monitored in the PLAP (Table 6).
Table 6. Pesticides analysed at Jyndevad with the product used shown in parentheses. Degradation products are in italics Precipitation and percolation are accumulated from the date of first application (App. date) until 1 July 2002. 1 st month percolation refers to accumulated percolation within the first month after application. Cmean refers to average leachate concentration at 1 m b.g.s. The number of pesticide-positive samples is indicated in parentheses.
With propiconazole, fenpropimorph, triazinamin-methyl (degradation product of tribenuron methyl) and glyphosate, the leaching risk was found to be negligible at the Jyndevad site. Apart from two samples containing 0.03-0.04 µg/l of fenpropimorph, and three containing 0.01-0.02 µg/l of AMPA, none of these compounds or the degradation products listed in Table 6 have yet been detected.
Figure 18. Pesticide application, precipitation and irrigation (primary axis) together with simulated percolation 1 m b.g.s. (secondary axis) at Jyndevad. Pesticides applied later than April 2002 are not included.
The leaching risk of glyphosate and triazinamin-methyl should be viewed in relation to the rather wet monitoring period, with percolation occurring shortly after application of both pesticides (Figure 18). Percolation within the first month after application was 149 mm for glyphosate and 95 mm for triazinamin-methyl (Table 6.). For further information, see Kjær et al . (2002). In contrast, fenpropimorph and propiconazole were applied during summer 2000, when precipitation input was nearly normal (Appendix 4). During this period precipitation input was almost counterbalanced by actual evapotranspiration such that no percolation occurred during the first month after application (Table 6 and Figure 18).
The leaching risk of terbuthylazine, PHPC (degradation product of pyridate) and rimsulfuron will not be evaluated until the 2003 monitoring results become available, i.e. when two years of monitoring data have been collated. It should be noted, though, that apart from desethylterbuthylazine, none of these compounds or their degradation products listed in Table 6 have yet been detected in any of the water samples analysed.
Desethylterbuthylazine (degradation product of terbuthylazine) did leach during the current monitoring period. In mid October 2001, about five months after application, desethylterbuthylazine was detected in elevated concentrations in the S1 suction cups located 1 m b.g.s. The average concentration has not yet been calculated, though, as the monitoring period does not fully cover the leaching period. The highest concentration (0.06 µg/l) was detected at the very end of the current monitoring period, thus indicating that leaching of the compound from the uppermost metre of the soil has not yet ceased.
The observed leaching should be viewed in relation to the spring application, when hydrological conditions allow the applied compound a relatively long residence time in the root zone. Terbuthylazine was applied in May 2001, when precipitation input was close to normal and counterbalanced by actual evapotranspiration (Appendix 4 and Figure 18). Hence, percolation did not occur until mid July, about 1.5 months after the last application (Figure 19). Desethylterbuthylazine was first detected after 190 mm of percolation, equivalent to 1.5 pore volumes.
Figure 19. Measured concentration of desethylterbuthylazine (primary axis) together with accumulated percolation 1 m b.g.s. (secondary axis) at Jyndevad. The measured data derive from suction cups installed 1 m b.g.s. at location S1 (see Figure 13). Percolation is simulated (see Section 3.2.1.). The red vertical lines indicate date of pesticide application. Concentrations below detection limits are indicated with open squares.
Desethylterbuthylazine has not been detected in either the S2 suction cups or the downstream monitoring wells. In the monitoring well M7 located upstream of the test site, however, desethylterbuthylazine was detected in 8 of 9 analysed samples in concentrations ranging from 0.01-0.02 µg/l due to prior application of terbuthylazine on the neighbouring field located upstream of the PLAP site. This was detected in the initial screening analysis, indicating that desethylterbuthylazine was present in M7 before the monitoring started in September 1999 (Kjær et al., 2001).
It should also be noted that pesticide application prior to the PLAP has caused marked groundwater contamination with the degradation products of metribuzin. Metribuzin-diketo was detected in concentrations as high as 1.37 µg/l and exceeded 0.1µg/l in 73% of the 26 water samples analysed. Metribuzin-desamino-diketo was detected in concentrations as high as 1.83 µg/l and exceeded 0.1µg/l in 50% of the 26 water samples analysed. For further information, see Kjær et al. (2002).
The risk of pesticide leaching at Jyndevad can be summarized as follows: