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Field study of pesticide leaching in an allophanic soil in New Zealand. 1: experimental results.


This study is part of a nationwide approach for assessing pesticide behaviour in key New Zealand soils under different climatic conditions, and the suitability of simulation models to predict transport to depth, where there could be impact on groundwater quality. It follows 2 trials in the relatively dry Hawkes Bay region on sandy/gravelly soils, which ran from 1993 to 1997 (Close et al. 1999). Although the majority of groundwaters in New Zealand appear free from pesticide contamination (Ministry for the Environment 1997), the most recent national survey in 1998-99 detected pesticides in 35% of the wells sampled, albeit mostly at concentrations below 0.1 mg/[m.sup.3] (Close and Rosen 2001). Earlier national surveys (Close 1993, 1996) identified pesticides in just 12% of sampled wells, but had a greater detection limit of 1 mg/[m.sup.3]; if comparable detection limits are applied to the most recent data, then 11% of the wells would have had detectable pesticides, similar to the previous surveys. Other more localised surveys by Regional Authorities (e.g. Smith 1993; Hadfield and Smith 2000) have also found a number of pesticides in vulnerable groundwater systems, i.e. shallow groundwaters underlying free draining soils with high pesticide use.

The establishment of a site in the Waikato region offered significant contrasts to the Hawkes Bay region. Firstly, the climate is significantly wetter, with average annual rainfall around Hamilton of approximately 1200 mm per year, some 400 mm per year higher than that recorded close to the Hawkes Bay sites. Mean monthly air temperatures are marginally higher. More importantly however, allophanic soils, with higher topsoil carbon levels, are a large component of the regional soil resource. Allophane is a clay mineral with variable charge, which means that, depending on soil conditions, it can adsorb compounds with negative as well as positive charge. This may influence pesticide adsorption, which is normally considered to be controlled mainly by soil organic matter. A large variability is reported for the mobility and persistence characteristics of pesticides, although there are some 'best available' values recommended for modelling purposes (Wauchope et al. 1992). In some cases, values derived from other field studies in New Zealand have been found to be quite different to those values, particularly in the subsoil (Close et al. 1999).

The objectives of this study were to (i) monitor the movement of selected pesticides and tracers through a free-draining allophanic soil into underlying groundwater, (ii) determine their mobility and persistence characteristics, and (iii) evaluate the performance of several leaching models of different complexity. This paper describes the study and presents results of tracer and pesticide monitoring in the soil profile and groundwater. A companion paper (Close et al. 2003, this issue) describes the simulations of these results using 4 different pesticide leaching models and compares the estimates of mobility and degradation with other literature values.

Materials and methods

Soil and site

The experimental site was located about 10 km south of Hamilton, in the North Island, on a Horotiu silt loam, classified as a Typic Orthic Allophanic Soil (New Zealand Soil Classification; Hewitt 1992), or Typic Hapludand (USDA 1996). Horotiu soils are formed from silty/sandy alluvium derived from volcanic material and are extremely versatile soils, deep and well drained, and suitable for a wide variety of uses. The topsoil contains a significant amount of allophane (up to 12%). Soil profile details at the site, together with specific soil chemical and physical properties for measured horizons are given in Table 1. The depth of the Ap horizon was consistent over the site, but there was some variation in the thickness of the Bw horizons; BC horizons encountered at 0.31 m in some locations.

The site was on a large alluvial fan, with effectively zero slope, and prior to the study had been used for an experimental grape crop (1992-95), then as a grape nursery area. There had been no application of any of the chemicals applied in the study since 1992. The site had no vegetation during the period of study. This was initially because of the combined effect of the applied herbicides, and later in the study, growth was controlled by occasional applications of glyphosate. Rainfall and temperature data were obtained from Hamilton Airport (<1 km distance) and radiation data from Ruakura climate station, approximately 11 km away.

The depth to groundwater at the site varied from 3.8 to just over 6 m below ground level (bgl). The aquifer material consisted of coarse sand and bore logs indicated a layer of silt at about 6 m bgl. An array of monitoring wells, screened from 4 to 6 m bgl, was installed across the groundwater flow path to allow sampling of groundwater at different distances from the trial site (Fig. 1). Wells HW1-3 and HW6 were installed in October 1997 and the remainder in August 1998. The groundwater flow velocity and direction were estimated from 2 dye tracing experiments (described later). The first was carried out using a well array about 100 m from the site and the second was carried out using well HW2 for tracer injection at the conclusion of the study. Temporary piezometers (HT1 to 5) were installed for this tracing experiment (Fig. 1). A small-scale pump test was carried out on well HW6 and hydraulic conductivity was estimated at 131 m/day. Pump tests on a water supply well, about 275 m away, gave a hydraulic conductivity estimate of 103 m/day.



Two trials were set up at the site (Fig. 1). The main trial (Site A) was similar in design to the Hawkes Bay trials (Close et al. 1999), and measured 15 by 15 m in area. It was instrumented with tensiometers and time domain reflectometry (TDR) probes to provide information on soil moisture status at different depths, with 9 suction cups at different depths for sampling of soil water. The ceramic suction cups were obtained from Soil Moisture Equipment Corp., California, and were 1 bar, high flow, round bottom cups (56 mm long and 48 mm diam.), consisting mainly of alumina ([Al.sub.2][O.sub.3]). The ceramic cups were leached with 10% HCl then rinsed through with distilled water prior to use as recommended by Wilson (1988). The suction cups were installed from the face of a soil pit on the edge of the application area and sloped downwards at angles ranging from 10-30[degrees] so that water would collect in the bottom of the cups on application of suction and be readily sampled. Two suction cup samplers were installed at each of 0.2, 0.4, and 0.6 m depths, with individual suction cups being installed at 0.9, 1.5, and 2.5 m depths. These depths corresponded approximately to the various soil horizons (Table 1). A smaller trial (Site B), measuring 15 by 5 m with no suction cups installed, was used to examine the movement of a different suite of pesticides, and to evaluate a sampling regime consisting of only soil sampling.

Pesticides were applied in November 1997: atrazine, terbuthylazine, procymidone, and hexazinone were applied to Site A; and 2,4-D, picloram, and triclopyr to Site B. These pesticides were selected because they had been detected in groundwater in New Zealand, there was a reasonable range in leaching properties (required for model evaluation), and all the pesticides applied to each site could be analysed in a single method. Br and deuterated water were applied to both sites as tracers. A summary of properties and application rates is given in Table 2. The application rates were approximately 1-7 times normal label application rates, to ensure measurable quantities leached to depth for modelling purposes. The pesticides and tracers were applied using a pressurised hand sprayer with a fine nozzle at a rate of around 12 mm/h to minimise any preferential flow, which might result from the application. The spraying volume was 2 L/[m.sup.2] (2 mm application depth). No surface ponding was observed during application.

A small amount of irrigation (total of 90 mm) was applied to Site A during the summer of 1998 after application of pesticides using a fixed system of mini-sprinklers. Intensive irrigation was applied in August 1999 near the end of the study (156 mm over 5 days), and groundwater was sampled more frequently following this period to better define the unsaturated zone/groundwater linkage. The last sampling period was in October 1999, approximately 2 years after the pesticide application.

Sampling and analysis

The suction cups were sampled by applying a vacuum of 65 kPa for 10-20 h at intervals ranging from 1 to 4 weeks. Soil samples were taken every 3 or 4 months to a maximum depth of 1 m. Five cores were taken on each sampling occasion, separated into 10 cm depth intervals and composited for analysis. Groundwater samples were collected infrequently at the start of the study, then with increasing frequency. The soil and water samples were refrigerated after collection, and then transported to the laboratory for analysis.

The water samples were analysed for pesticides using gas chromatography with mass spectrometry (GCMS) with a detection limit of 0.1 mg/[m.sup.3] and bromide was analysed by ion chromatography. Sodium sulfate (4 g) was added to 30 mL water, together with an internal standard and extracted with 5 mL ethyl acetate, then analysed using GCMS. Two metabolites, desethyl atrazine and desisopropyl atrazine, were also analysed in the soil water samples using GCMS. The solute concentrations from the duplicate suction cup samplers were averaged for graphical presentation. The soil samples from Site A were analysed for the suite of pesticides by extracting 50 g of soil with ethyl acetate and anhydrous sodium sulfate. The extracts were cleaned up using gel permeation chromatography and analysed using high-resolution gas chromatography. The soil samples from site B were analysed for acid herbicides by extracting 35 g of soil with 100 mL of 0.1 M sodium hydroxide. The extract was filtered, acidified, and partitioned into dichloromethane before clean up by gel permeation chromatography. The purified extract was methylated using diazomethane and analysed using high-resolution gas chromatography. The detection limits were 0.01 mg/kg for the soil samples. The soil samples from both sites were analysed for bromide following extraction of 20 g soil combined with 10 g acid-washed sand with 100 mL distilled water. The soil moisture content was determined to express results as mg/kg dry weight. The detection limit for bromide in soil was 0.7 mg/kg.

Both water and soil samples were analysed isotopically by stable isotope mass spectrometry (Hulston et al. 1981). For deuterium ([sup.2]H), water was allowed to reach isotopic equilibrium with [H.sub.2] gas at 29[degrees]C using Hokko beads as a catalyst. Then the [sup.2]H/[sup.1]H ratio of the [H.sub.2] gas was measured in a double collector mass spectrometer. Oxygen-18 was measured similarly, by equilibrating the water isotopically with C[O.sub.2] gas at 29[degrees]C, and analysing the C[O.sub.2] in a mass spectrometer. Water was extracted from the soil by azeotropic distillation before analysis (e.g. Revesz and Woods 1990).


Rainfall and soil moisture

Rainfall during the study period was reasonably erratic compared with the historical monthly rainfall percentiles for Hamilton (Fig. 2). It was generally at or below the mean monthly figure prior to pesticide application, with below average rainfall continuing until May 1998. The months of June and July 1998 were extremely wet with a total rainfall of 477 mm compared with a total of 257 mm normally expected. For the remainder of the study period following these wet months, rainfall was again less than expectations, with the exception of August 1999 (Fig. 2). The total rainfall between November 1997 and October 1999 was 2230 mm, which was about 200 mm less than the expected total for this period.


Soil moisture was measured using a combination of gravimetric samples and TDR observations. Unfortunately, problems with the TDR equipment meant that measurements were not as regular or frequent as planned. Figure 3 shows the variation of soil moisture at 0.2 and 0.3 m depth with weekly totals of rainfall minus evapotranspiration. The soil moisture generally responds to excess rainfall as would be expected but the degree of correlation is limited by the sparse data.


Groundwater hydrology

Two groundwater tracing experiments were carried out at the site to determine the groundwater flow rate and direction and give an indication of dispersion characteristics. The first experiment was carried out about 100 m south of the study site. An injection well and 2 arrays of wells at 1.5 and 6 m down gradient were installed with all wells screened within and above the water table for 3 m (total depth 5.5 in bgl). Rhodamine WT dye and bromide were injected and the tracers were detected in one well in each of the monitoring arrays. The groundwater flow direction was ESE (about 106 degrees), which is similar to the direction derived from the piezometric contours. The groundwater seepage velocity was about 0.3 m/day with the hydraulic gradient at the time being 0.0017. The tracer plume was narrow (not detected at adjacent wells 0.6-0.7 in away) with the longitudinal dispersivity estimated at 0.15 and 0.2 m, and the lateral dispersivity estimated at 0.01 and 0.015 m for travel distances of 1.5 and 6 m, respectively. Monitoring of bromide and pesticides in the groundwater wells during the study indicated that the groundwater flow was much more towards the south so a second tracing experiment was carried out at the actual site at the end of the study. Tritiated water, Br and rhodamine WT were injected into well HW2 (Fig. 1), along with atrazine, hexazinone, and procymidone. Five piezometers were installed in an arc 1.5 m from the injection well (wells screened within and above the water table for 3 m) and monitored for 23 days. Further details and analysis of this experiment are given in Pang and Close (2001). The groundwater seepage velocity was calculated at 0.45 m/day with a flow direction of SSW (about 194 degrees). This direction was consistent with the results observed in the monitoring wells, particularly wells HW4, HW5, and HW10 (Fig. 1). The difference in local flow direction over a distance of 100 m is quite large (about 88 degrees). This probably reflects variable permeability of the shallow alluvial sand aquifer and changes in depth of a less permeable silt layer (about 5.5 m bgl at the site but about 9 m bgl at the site of the first groundwater tracing experiment).

Pesticides and bromide tracer

The site was monitored for just under 2 years (approx. 720 days). The soil water concentrations of bromide and hexazinone at Site A are shown in Fig. 4, with the soil concentrations for bromide and selected pesticides shown in Figs 5 and 6 for Sites A and B, respectively. It should be noted that a log scale is used for Figs 5 and 6 so that both bromide, which was applied at higher concentrations, and the pesticides could be displayed. The log scale tends to accentuate the lower concentrations. The different detection limits for bromide and the pesticides should also be noted. Atrazine and terbuthylazine are not included in Figs 4 and 5 because of the very low levels detected, but are included in the mass recoveries for the soil sampling (Table 3). Low levels (0.01 mg/kg) of atrazine and terbuthylazine were detected in soil samples down to 0.5 m on one sampling occasion (April 1998), with nothing being detected in the soil samples after November 1998. Most of the triazine mass was observed in the top 0.10 m of the profile. This was consistent with the soil water sampling, where terbuthylazine was only detected on one occasion and atrazine only on 12 occasions in various suction cups. Thus, both atrazine and terbuthylazine were fairly immobile and appear to have rapidly degraded at the site. One of the metabolites, desethyl atrazine, was detected in a number of suction cups down to 2.5 m depth and appears to be more mobile and persistent than atrazine.


Bromide, hexazinone, and picloram were highly mobile at the site. There were differences between bromide and the pesticides in the leaching patterns observed in the top 0.5 m of the profile. This probably results from the differing distributions of allophane (Table 1), which would adsorb the bromide, and organic carbon, which would adsorb the pesticides. Organic carbon is high in the top 0.2 m then the levels rapidly drop away, whereas the allophane content remains high until below 0.5 m. Procymidone was much less mobile with triclopyr showing intermediate mobility. The leaching of 2,4-D was dominated by high degradation/loss rates (Table 3), but it was also fairly immobile, not being observed below 0.2 m on any occasion. There was significant leaching in June and July 1998 (230-260 days after application) and both bromide and hexazinone moved rapidly from 0.6 m to 2.5 m (Fig. 4). This resulted from average rainfall (107 mm) in May followed by above-average rainfall in June (194 mm) and July (284 mm). The rainfall in July was above the 90 percentile for July (Fig. 2) and was 155 mm more than the average for July.

The mass recoveries (Table 3) give a quick indication of the persistence of the pesticides and when leaching below the sampling depth had occurred. There were lower mass recoveries for the persistent pesticides (bromide was not analysed due to an error in sample handling) on first sampling occasion compared with the second sampling occasion. This was probably a result of the shallow sampling depth (0.15 m) on the first occasion and there is also the possibility that some of the pesticides were discarded with root and vegetation material. The rapid disappearance of atrazine, terbuthylazine, and 2,4-D was clearly seen. There was significant leaching of most chemicals observed between Days 149 and 267, consistent with the high rainfall discussed in the previous paragraph. The higher mass recovery of bromide in April 1998 from Site B compared with Site A is probably the result of the irrigation applied to Site A over first 3 months in 1998, combined with the shallow sampling depth on that occasion.

Deuterated water

Deuterated water ([sup.2][H.sub.2]O) was added to the pesticide mix before application. Isotopic exchange with normal water (mainly [sup.1][H.sub.2]O) means that this rapidly becomes [sup.1][H.sub.2]O via the equation:

[sup.2][H.sub.2]O + [sup.1][H.sub.2]O [right arrow] [2.sup.1][H.sub.2]O

The deuterium concentration is expressed as the fractional difference between a sample and the standard, i.e.:

[[delta].sup.2]H [per thousand] = [[([sup.2]H/[sup.1]H).sub.sample]/[([sup.2]H/[sup.1]H).sub.std] - 1] x 1000

where the standard is Standard Mean Ocean Water. After addition of deuterated water, the [[delta].sup.2]H value of the pesticide mix was +4000 [per thousand].

Deuterated water is an ideal tracer in the sense that it mimics the behaviour of water. However, this means that a good fraction of it would have been lost by evaporation before reaching the soil, and from within the soil, because it is almost as volatile as normal water, unlike bromide. In addition, deuterium is naturally present in rainfall and other environmental waters (in the ratio of about 1:150) and its concentration varies because of meteorological factors. Oxygen-18 varies similarly. This leads to a relationship between the [[delta].sup.2]H and [[delta].sup.18]O values of rainfall, which is given by:

[[delta].sup.2]H = 8 x [[delta].sup.18]O + d

where d is the intercept on the y-axis; d is close to 12 [per thousand] for rainfall in the Hamilton area on average. [[delta].sup.18]O is defined by:

[[delta].sup.18]O [per thousand] = [[([sup.18]O/[sup.16]O).sub.sample]/[([sup.18]O/[sup.16]O).sub.std] - 1] x 1000

Turning the equation around gives: d = [[delta].sup.2]H - 8 x [[delta].sup.18]O. Applying deuterated water causes [[delta].sup.2]H to increase without changing [[delta].sup.18]O very much, so d increases by the same amount as [[delta].sup.2]H. d can therefore be used as a proxy for [[delta].sup.2]H, and in fact is preferred to [[delta].sup.2]H because it is less influenced by meteorological effects. (The reason that [[delta].sup.18]O does not change much with application of deuterated water is because deuterated water has [[delta].sup.18]O like normal water.) d responds mainly to applied tracer and evaporation effects, whereas [[delta].sup.18]O reflects meteorological and evaporation effects. Evaporation causes both [delta] values to increase (i.e. become less negative), and the d value to decrease because oxygen-18 is relatively more affected by evaporation than deuterium.

Figure 7 shows the d values plotted against time since application of the deuterated pesticide mix for each suction cup. The points show the d values in the soil water and the lower line shows the d values expected if deuterated water had not been applied. The difference shows the effect of deuterium tracer. It is assumed that in the absence of tracer, d would have decreased from 10 [per thousand] in early November 1997 to 3 [per thousand] in early June (210 days) because of evaporation effects in the soil. Increase in d after June was due to the input by winter rainfall. Deuterium tracer was not detected after June at any depth.


The deuterium tracer results are similar to the bromide results (Fig. 4), but appear to have shorter tails indicating less retention in the soil. Deuterium has less sensitivity than bromide. At 0.2 m depth, the initial peak is followed by a dip at 90 days like bromide, but more pronounced. Deuterium tracer can then be seen until 150-180 days. The bromide tracer lasted until 270 days. At 0.4 m depth, the peaks are flatter and more spread out than with bromide. However, samples collected closer to the date of application could have shown more deuterium. The deuterium peak lasted to about 210 days, compared with 360 days for bromide. At 0.6 m depth, deuterium transport was considerably more rapid than bromide transport, with peaks at about 120 days and peaks lasting about 240 days, compared with 240 and 720 days for bromide. Deuterium tracer was not detected in soil water at 0.9, 1.5, or 2.5 m depths.

Deuterium tracer was also not detected in the soil samples collected in 0.10 m increments down to 1 m depth in the A and B areas. The first sampling for deuterium measurement was made at 150 days after application of the tracer down to 0.5 m depth. By then, the (applied) deuterium concentrations were too low to be detected.

Appearance in groundwater

Bromide and hexazinone were detected in the groundwater in 5 of the 8 monitoring wells. Results are shown in Fig. 8 for well HW2 (centre of the plot) and 2 down-gradient wells (HW4 and HW10). Most of the wells were dry from March 1998 to late June 1998, when significant recharge occurred. There were clear peaks associated with winter recharge in 1998 and the intensive irrigation period in August 1999. The rainfall between February and June 1999 was below average (Fig. 2), and although the soil moisture levels were increasing (Fig. 3), there was little, if any, recharge in 1999 prior to the intensive irrigation in August 1999. The groundwater velocity of 0.3-0.45 m/day means that the recharge would take an average of 1-2 months to travel from the centre of the plot to the down-gradient wells.



There was evidence from the comparison of leaching patterns of bromide compared with deuterated water (Figs 4 and 7) that bromide was being slightly adsorbed by the allophane. A simple batch adsorption test was carried out for bromide on 3 soil horizons with allophane contents ranging between 5 and 12%. This showed increasing adsorption of bromide with increasing allophane content with [K.sub.d] values ranging between 0.4 and 1.3 mL/g. The pesticides in the study ranged from weak bases (atrazine, terbuthylazine and hexazinone) to weak acids (2,4-D, triclopyr, and picloram), with one non-ionic pesticide being procymidone (Weber 1994). Allophane could be expected to adsorb weak acids slightly more than would be expected from the organic matter content alone. A study of pesticide sorption by allophanic and non-allophanic New Zealand soils has been carried out by (Baskaran et al. 1996). They found that organic matter content was the dominant soil property affecting pesticide adsorption. Their data indicate that the greater adsorption by allophanic soils compared with non-allophanic soils was mainly because of the higher organic carbon content of the allophanic soils studied. [K.sub.oc] values were higher for 2,4-D and lower for atrazine for allophanic soils compared with non-allophanic soils. The effect of allophane on the mobility of bromide and the pesticides will be examined further in the companion paper that examines the estimation of leaching parameters from the model simulation results.

The use of both suction cups and soil sampling in this study allows for a discussion of the strengths and limitations of each method. There were many more suction cup samples taken in this study, which needs to be noted in the comparison. It was unfortunate that the total depths for the first 3 soil samples were not deeper as it is almost certain that some of the bromide and pesticides had leached below the lower interval sampled (Figs 5 and 6, Table 3). This was due to more rapid leaching than was anticipated and resulted in lower mass recoveries for the profile than would have been achieved if all sampling had been to a greater depth. However, from the suction cup results (Fig. 4) it is clear that bromide and hexazinone had reached 2.5 m bgl 260 days after application. As this depth is not technically feasible for soil sampling, the sampling depth will always be a limitation at some stage in the study. At this site, the practical depth for soil sampling was limited to 1 m. Depth was not such a limitation for the suction cups which were installed down to 2.5 m. The two techniques also differ in their detection limits and the fact that the soil extraction will include pesticides adsorbed to the soil as well as in solution. The difference in detection ability will be influenced by the degree of adsorption of the pesticide, the organic matter content, and the soil moisture status at the time of sampling. The US Environmental Protection Agency guidance manual for prospective groundwater monitoring studies (USEPA 1998) indicates the detection limits for soil sampling are typically an order of magnitude higher than for suction cups. This means that suction cups can better track the movement of pesticides with a greater power of detection than can soil samples. This can be seen in Fig. 5 for hexazinone, where the level at 0.9-1.0 m was below detection 388 days after application, and the deepest detection at that stage was at 0.9 m. This would imply much lower mobility than indicated by the suction cup data where significant amounts of hexazinone were observed down to 2.5 m at that time. A similar observation was made by (Patterson et al. 2000) in their comparison of coring and suction cup data. They found that atrazine migration rates were underestimated by more than 50% using core data compared with combined core and suction cup data.

A significant advantage of soil sample data is the ability to obtain a mass balance for the profile, particularly while the compounds are above the maximum sampling depth. This can give a quick estimate of persistence without the need to carry out detailed simulations that are required for the interpretation of suction cup data. Another advantage of the soil sample data is for fairly immobile pesticides where there is little transport of the pesticide. It is difficult to install suction cups in the top 0.1 m of the soil profile and, as this is often the depth with highest levels of soil organic matter, less mobile pesticides can remain in this interval for a long time. In this situation suction cup data will not provide any information, and the non-appearance of the pesticide could result from either low mobility or persistence. If a pesticide had low mobility combined with high persistence, for example DDT, it would not be detected by the suction cup but could remain a problem. An example of this type of advantage for soil sampling can be seen in the data for atrazine and terbuthylazine (Table 3), where these pesticides were detected particularly in the top 0.2 m of the profile by the soil sampling, but were only occasionally detected in the suction cups.

The combination of both soil sampling and suction cups can give a more complete description of the leaching process and the distribution of the compound through the profile. This is seen in Figs 4 and 5 where hexazinone is observed leaching down to 2.5 m in the suction cup data, but is also still present in the soil sampling data in the top 0.1-0.2 m due to the high organic matter content. We believe that a combination of both techniques is worthwhile and gives complementary information providing the strengths and limitations of both techniques are appreciated.
Table 1. Soil chemical and physical characteristics for experimental

Horizon Depth Description pH Clay Silt
 (m) (%) (%)

Ap 0-0.20 Loamy silt 5.79 20 33
Bw1 0.20-0.40 Slightly gravelly 6.43 18.5 17
 loamy silt
Bw2 0.40-0.53 Slightly gravelly 6.72 11 13
 loamy silt
Bw3 0.53-0.67 Slightly gravelly 6.80 6.5 9.5
 loamy silt
BC1 0.67-1.05 Slightly gravelly 6.37
 loamy sand
BC2 1.05-1.28 Slightly gravelly 6.41
 cemented loamy sand
BC3 1.28-1.56 Gravelly loamy sand 6.43
C >1.56 Gravelly sand

Horizon Organic C Allophane BD Total
 (%) (%) (g/[cm.sup.3] porosity
 (%, v/v)

Ap 8.00 10 0.87 63.6
Bw1 1.88 12 0.85 66.9
Bw2 0.93 12 1.02 61.6
Bw3 0.43 5 1.11 58.1
BC1 0.13 2
BC2 0.10 2
BC3 0.07 1

Table 2. Selected information on chemicals applied on each site

Selected 'best available' values from
ppdb3.html as at February 2002 with range in parentheses are given
for mobility and persistence

Chemical Primary use Mobility
 ([K.sub.oc] from

 Site A

Deuterated water Water tracer 0
Bromide Water tracer
Alrazine Arable cropping 147
 (herbicide) (38-288)
Terbuthylazine Arable cropping 220
Hexazinone Forestry 40
 (herbicide) (34-74)
Procymidone Vegetables, 1500
 (fungicide) strawberries and (1500-1945)

 Site B

Deuterated water Water tracer 0
Bromide Water tracer
Picloram General--difficult 29
 (herbicide) to kill perennial (7-48)
2,4-D General 48
 (herbicide) post-emergence (20-79)
 broadleaf weeds
Triclopyr Grassland 68
 (herbicide) (12-160)

Chemical Persistence (half-life Application
 in days from rate
 literature) (kg/ha)

 Site A

Deuterated water No degradation 13.3
Bromide No degradation 224
Alrazine 60 10
 (herbicide) (15-330)
Terbuthylazine 60 10
Hexazinone 88 10
 (herbicide) (27-218)
Procymidone 15 10
 (fungicide) (7-120)

 Site B

Deuterated water No degradation 13.3
Bromide No degradation 224
Picloram 90 (A) 10
 (herbicide) (20-277)
2,4-D 5.5 10
 (herbicide) (1-9)
Triclopyr 32 14
 (herbicide) (8-69)

(A) Picloram degradation is stated to be a function of concentration.

Table 3. Mass recoveries for tracers and pesticides for each plot

Sampling Days since Depth of Mass recovery
date application sampling (% of mass
 (m) applied)

 Site A

 Br Hexazinone

17/12/97 43 0.15 NA 38.6
2/4/98 149 0.50 64.5 72.3
29/7/98 267 0.60 48.8 19.5
26/11/98 388 1.00 10.4 16.5
22/3/99 503 1.00 16.8 14.4
29/7/99 632 1.00 4.6 0.6

Sampling Mass recovery
date (% of mass applied)

 Site A

 Procymidone Atrazine Terbuthylazine

17/12/97 44.3 23.2 21.8
2/4/98 64.5 1.6 4.7
29/7/98 36.7 1.2 2.1
26/11/98 18.8 0.2 0.7
22/3/99 21.9 0.0 0.0
29/7/99 12.6 0.0 0.0

Sampling Days since Depth of Mass recovery
date application sampling (% of mass applied)

 Site B


17/12/97 43 0.15 n.a.
2/4/98 149 0.50 84.6
29/7/98 267 0.60 18.2
26/11/98 388 1.00 18.7
22/3/99 503 1.00 30.1
29/7/99 632 1.00 8.9

Sampling Mass recovery
date (% of mass applied)

 Site B

 Picloram Triclopyr 2,4-D

17/12/97 44.2 53.1 6.0
2/4/98 54.6 70.1 4.0
29/7/98 46.2 31.9 1.9
26/11/98 28.1 10.4 0.5
22/3/99 20.7 14.9 0.7
29/7/99 20.3 6.6 0.3

n.a., Not analysed.


The authors thank Danny Thornburrow for assistance with the fieldwork. The research was funded by contracts CO3X001 (ESR) and CO9X0017 (Landcare Research) from the Foundation for Science, Research and Technology (New Zealand) and by Environment Waikato for the monitoring well installation.


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Close ME (1993) Assessment of pesticide contamination of groundwater in New Zealand. II. Results of groundwater sampling. New Zealand Journal of Marine and Freshwater Research 27, 267-273.

Close ME (1996) Survey of pesticides in New Zealand groundwaters 1994. New Zealand Journal of Marine and Freshwater Research 30, 455-461.

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Manuscript 2 July 2002, accepted 16 January 2003

M. E. Close (A,F), G. N. Magesan (B,C), R. Lee (B), M. K. Stewart (D), and J. C. Hadfield (E)

(A) Institute of Environmental Science and Research, PO Box 29-181, Christchurch, New Zealand.

(B) Landcare Research NZ Ltd, Private Bag 3127, Hamilton, New Zealand.

(C) Present address: Forest Research, Private Bag 3020, Rotorua, New Zealand

(D) Institute of Geological and Nuclear Sciences, PO Box 31-312, Lower Hutt, New Zealand.

(E) Environment Waikato, PO Box 4010, New Zealand.

(F) Corresponding author; email:
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Author:Close, M.E.; Magesan, G.N.; Lee, R.; Stewart, M.K.; Hadfield, J.C.
Publication:Australian Journal of Soil Research
Geographic Code:8NEWZ
Date:Sep 1, 2003
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