Fate and movement of atrazine, cyanazine, metolachlor and selected degradation products in water resources of the deep Loess Hills of southwestern Iowa, USA

Abstract

The environmental fate and movement of herbicides widely used for weed control in corn are assessed for a deep loess soil in southwestern Iowa. Beginning in the early 1980s, the herbicide-based weed control program emphasized the application of atrazine (ATR) or cyanazine (CYN) and metolachlor (MET) for both broadleaf and grass control. Between 1992 and 1995, concentrations of ATR, desethylatrazine (DEA), desisopropylatrazine (DIA), CYN and MET were measured in rainwater, both shallow and deep vadose zone water, and well water. Results show that the frequency of herbicide detections and the range and distribution of occurrences are dependent upon both landscape position and temporal inputs of recharge water from rainfall. Generally, DIA was observed more frequently and in higher mean concentration in well water than DEA, while DEA was observed more frequently than DIA in vadose zone groundwater. A chromatographic analogy is suggested to explain the occurrence patterns observed for both parent herbicide and degradation products within the unsaturated zone water. Analysis of rainwater samples collected during this time also revealed low concentrations of ATR, CYN and MET, with the timing of the detections indicative of non-local transport. Results show that the deep loess soil conducts both water and agricultural chemicals relatively rapidly and as such represents a production system which is vulnerable to contamination of shallow groundwater by herbicide-derived chemicals. Results also illustrate the importance of including major herbicide degradation products in water resource impact assessment studies.

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Aim of investigation

Continuation of efficient grain production practices across the Midwest depends upon the development of agricultural management systems that protect and conserve the soil as a natural resource, provide profitable income to the farmer, and are environmentally benign. Management practices must be adopted in which agricultural chemicals, including both fertilizers and pesticides, are used in a manner that minimizes impacts upon surface and groundwater quality. Of particular concern in the midwestern US are the issues related to the offsite movement of herbicides applied to corn, soybean, wheat, and sorghum into non-target areas with adjacent or associated water resources.

The extensive history of research on pesticide impact assessments and contamination of surface water and groundwater resources has been summarized in several recent reports.^1–4 In the most comprehensive compilation to date, the National Water Quality Assessment Program (NAWQA) of the US Geological Survey (USGS) has collated results of studies on the occurrence, distribution, and environmental fate of pesticides in the atmosphere,⁵ in surface water,⁶ and in groundwater.⁷ Data are summarized across many scales, on different landscapes with varying land uses, and under different weather patterns. In the Midwest, where pesticide use is heavy and water tables are commonly shallow, the public awareness and concern over contamination by agricultural herbicides and insecticides are substantial.^2,8 Moreover, recent studies have confirmed the presence of several kinetically stable degradation products of common herbicides produced in the soil reaching groundwater.^9–11 Today in Iowa, studies monitoring groundwater for residues of herbicides continue a legacy which began nearly 20 years ago.¹² Because of land use patterns and hydrogeologic characteristics, the shallow groundwater beneath the deep loess hills of western and southwestern Iowa is vulnerable to contamination. Elevated nitrate levels in groundwater, often exceeding by several-fold the US Environmental Protection Agency (USEPA) standard of 10 mg l⁻¹ for nitrate-N in drinking water, provide evidence of contamination from agricultural activities on this landscape.¹³

Since the advent of the drainage basin-scale studies and intensive river quality assessments of the 1980s, several efforts have focused on water quality within the Mississippi River basin. Draining some 40% of the conterminous US, much of it devoted to production agriculture, this area extends from the headwaters of the Missouri River in eastern Montana through the grain belt and into the Ohio River in western Pennsylvania. These major rivers together with numerous tributaries all hold the potential for the discharge of agricultural chemicals. Across much of Iowa, both natural and manmade terrain features result in hydrologic conduits for the transport of chemicals to tributaries of larger drainage areas. Intensive monitoring of the lower Mississippi River has revealed salient loads of corn and soybean herbicides moving through the river system. A “spring flush” phenomenon has recently been observed.¹⁴ Since the application of these chemicals usually proceeds on a field-by-field basis, it is therefore appropriate to conduct research at the same scale of use with the objective of reducing or eliminating both the onsite movement and offsite displacement of herbicides and their degradation products into non-target areas with associated water resources. Under the prevailing weather patterns of southwestern Iowa, the loess topography conducts both water and soluble agricultural chemicals relatively rapidly. This kind of landscape response has also been observed in the “Palouse” region of Washington State, USA, wherein the mid-slope and toe-slope position recharge rates are 5–10 times greater in magnitude than those of the hilltop positions.¹⁵ As such the Loess Hills represent an ideal production-scale field laboratory in which to evaluate alternative farming systems which are sustainable with respect to environmental quality endpoints. Both surface and subsurface drainage from fields in our study area typically meanders through perennial creeks into larger rivers, which eventually discharge into the Missouri River. In addition, it is now imperative that such studies include the pesticide degradation products formed in the soil, as they assume increasing importance in assessments of the total changes occurring in groundwater quality beneath agricultural landscapes.

The importance of analyzing for major pesticide degradation products along with the parent compounds is now recognized as a necessary component of environmental impact studies.^11,16 In a 1991 study of water analyses from 303 wells across 12 midwestern states, the mean of the atrazine (ATR) plus metabolite concentrations was 53% greater than the mean of the ATR concentrations alone.¹⁷ Moreover, a censored regression analysis using the total ATR plus desethylatrazine (DEA) plus desisopropylatrazine (DIA) residue revealed other environmental occurrence factors not in evidence when using ATR concentration alone. In addition, the types of organic reactions necessary to convert cyanazine (CYN) to ATR, via the isobutyrylnitrilo substituent, are hydration, hydrolysis, and decarboxylation.¹⁸ These are the same types of reactions carried out by a consortium of microorganisms indigenous to many types of tilled soils.¹⁹ This biodegradation chemistry has been corroborated in field dissipation studies wherein soil microbes produced ATR, the precursor to DIA, from CYN in a clay-loam soil.²⁰ In a laboratory study of the aerobic degradation of CYN, including two acidic sandy-loam soils, more than 50% of the original CYN was recovered as cyanazine carboxylic acid in 40 days.²¹ Another more recent plot-scale field study reported the formation of DIA from CYN in both sandy- and silt-loam soils.²² DIA formation from CYN was faster in the silt-loam than in the sandy-loam, and in one study was observed to terminate with exhaustion of available CYN.²⁰

A review of the literature reveals numerous studies of corn herbicide movement through agricultural systems. Many have been reported on different soil types, under differing tillage practices, at many scales, and under a variety of hydrologic regimes. This study was undertaken to evaluate the impact of regionally common agronomic practices on groundwater quality by investigating the fate and movement of herbicide-derived chemicals within a field. Several hydrogeologic compartments were sampled and analyzed for DIA, DEA, ATR, CYN, metolachlor (MET), and morpholino-metolachlor (METMOR) over a 4 year period. It constitutes the first comprehensive study of corn herbicide distribution carried out on the deep loess deposits unique to the landscape immediately east of the Missouri River alluvial valley of western and southwestern Iowa. We report the results of the assessment for wet precipitation, vadose zone water, and well water.

Experimental

Sampling and analysis

The study was conducted on watershed 3 (W3), a 40 ha field at the Deep Loess Research Station in southwestern Iowa. Total relief across the study area is less than 25 m over a horizontal distance of 500 m with hydrogeologic boundaries effectively defining a small watershed. Detailed characterization of the site, soils, and geohydrology, history of agronomic practices, equipment and procedures for sampling both surface runoff and groundwater, and frequency of sample collections have been published elsewhere.¹³ A more recent series of reports of watershed-scale evaluations on southwestern Iowa’s deep loess soils addresses topography and agronomic practice,²³ hydrology,²⁴ and nitrogen use efficiency.²⁵ Wet precipitation samples were collected in 19 l buckets positioned on a wet–dry Model 301 Precipitation Collector (AeroChem Metrics, Inc., Bushnell, FL, USA) located at the southwestern corner of W3. Vadose zone water was collected from porous cup suction lysimeters at four different landscape positions with varying depths to the water table. At each position, lysimeters are nested with cups vertically separated to sample vadose zone compartments at different depths. Well water was collected from piezometers that had been purged 24 h earlier. Slug tests indicated that the saturated conductivity of the loess was of the order of 10⁻⁴ m s⁻¹, confirming recharge times of 8–16 h. Two wells screened well below the loess/till interface showed longer recharge times and saturated conductivities of the order of 10⁻⁷ m s⁻¹. Wells are installed at seven locations across the field, representing depths to the water table of 2–20 m. Both vadose zone and saturated zone samples were collected monthly, except during the time from November to March. All groundwater samples were collected by US Department of Agriculture (USDA) field personnel in accordance with USGS protocols for purging, sampling, storage, and transport of samples for groundwater quality studies.²⁶ Cleaned glass, Teflon^®, or stainless steel equipment was used in order to minimize pesticide loss by sorption or degradation. Cross-contamination between wells and samples was minimized by thorough rinsing and cleaning of all equipment. All samples were stored in clean oven-baked amber-glass bottles and held at 4 °C in the dark without additional chemical preservation until analysis.

A high-performance liquid chromatographic (HPLC) method was used for quantitative determination of DIA, DEA, ATR, CYN, MET, and METMOR in water samples. ATR, CYN, and MET were determined in rainwater; all were determined in vadose zone water; and DIA, DEA, ATR, and MET were determined in well water. Hydroxy derivatives of ATR or dealkylated ATR are not included in the analytical protocol, nor was fully dealkylated ATR itself. The method employs solid-phase extraction on cyclohexyl cartridges, liquid chromatographic (LC) separation on a reversed-phase column, and ultraviolet (UV) detection with a photodiode array (PDA) detector.²⁷ Based upon a 0.5 l sample volume, the minimum detectable concentration was 0.04 µg l⁻¹ for all analytes. Mean concentrations are computed from all values equal to or greater than 0.04 µg l⁻¹.

Quality control

All samples for HPLC determination were fortified with terbutylazine at 1.0 µg l⁻¹ immediately prior to analysis. The calculated recovery of this surrogate analyte provided a quality control check on both the extraction and instrumental determination. Surrogate recoveries consistently fell within ±15% of the fortified concentration.

Data analysis

Data obtained from the analysis of samples are presented in violin plots, which synergistically combine the features of a box-and-whisker plot (boxplot) and a density trace diagram.²⁸ There are many variations of boxplots, but each is a diagram that displays spread, center, asymmetry, and outliers for a variable.²⁹ In the boxplot, the spread, usually a rectangular box, is shown as a dark thick line connecting the upper (75th percentile) and lower (25th percentile) quartiles (the interquartile range), and thus represents the middle 50% of the data. The center is depicted as an open circle lying within the spread box; it represents the median value (50th percentile). If the data distribution is not symmetric, the median will not be in the middle of the box. The whiskers are the thin lines that connect the lower quartile to the 10th percentile and the upper quartile to the 90th percentile of the variable's empirical data distribution. Outliers are represented as points above or below the end of the whiskers. The density trace is a variation on a histogram that displays a smoothed empirical frequency distribution for the data. As with the histogram, the appropriate choice of class intervals for the number of observations is crucial for detecting and distinguishing unimodal from multimodal distributions. In the violin plot, the density trace is drawn symmetrically (mirrored) to the left and right of the vertical boxplot with the space in between shaded for emphasis.

Results and discussion

Agronomic system and chemical use

The agronomic system on W3 is representative of typical farming systems that were in place on the Loess Hills of southwestern Iowa in the early 1970s. Continuous corn was the predominant cropping system and ridge tillage along the sloping landscape contours was common. This tillage method was believed to be the best in-field approach for minimizing soil loss from wind and water erosion. Weed control relied mostly on chemicals and two ridge tillage passes, with the dominant herbicides for both broadleaf weed and annual grass control being the triazines and the chloroacetanilides.

Herbicide use history

Fig. 1 shows the complete 22 year record of ATR application on W3, beginning in 1972, as well as the MET record for the period between 1980 and 1995. Between 1973 and 1979, ATR was often co-applied as a minor ingredient with alachlor (ALA). By the early 1980s, the ATR/MET combination controlled the weeds of interest so well that it had become the mixture of choice for farmers. ATR was applied to this field every year from 1972 to 1994, with the exception of 1985 through 1989, when it was replaced by CYN, an alternative triazine that controlled approximately the same range of weed species. Generally, ATR and CYN application rates were at their current label recommendations. The total herbicide loads applied to W3 between 1972 and 1996 were 36.5, 9.8, and 34.2 kg ha⁻¹ for ATR, CYN, and MET, respectively.

Fig. 1 Annual application history on W3 for ATR (1972–1994), CYN (1985–1989), and MET (1980–1995).

Detections in wet precipitation

As semivolatile organic compounds, field-applied pesticides are known to be present in the atmosphere and to be redeposited on the land surface by washout in wet precipitation.^5,30Fig. 2 shows the concentration distribution for ATR, CYN, and MET in rainwater samples collected on W3. In the 225 rainwater samples collected from W3 between January 1993 and December 1995, ATR was detected in 31% at a median concentration of 4.7 µg l⁻¹ and MET was detected in 36% at a median concentration of 2.5 µg l⁻¹. The violin plot reveals an observed distribution of values above the minimum detection limit (MDL), indicating that the MDL was sufficient to observe the majority of herbicide occurrences in rainfall. The flask-like distributions indicate that there is a probability that some low concentrations are below the MDL. Therefore, a redeposition load analysis would likely underestimate the actual deposition. The maximum concentrations measured were 68.5 and 11.3 µg l⁻¹ for ATR and MET, respectively. DIA was not detected. During the summer 1993 floods, measured rainfall on W3 was 18.7, 19.9, 25.5, and 31.5 cm during the months of May, June, July, and August, respectively. These values correspond to the upper 90th percentile of the monthly rainfall distribution as determined from the 120 year record for the area (National Weather Service, Omaha, NE, USA). The detection frequency for each year was greatest from April through July and least from November through March. For 1993, ATR washout in rainwater was limited to the months of May through July, with an estimated maximum daily deposition rate of 40 µg ha⁻¹. In 1994, ATR was detected in wet precipitation from January through December, and in 1995, from May through December. The estimated maximum daily deposition rate for 1994 was the same as for 1993, but increased in 1995 to 80 µg ha⁻¹.

Fig. 2 Concentration distribution for ATR, CYN and MET in 225 wet precipitation samples collected on W3 between January 1993 and December 1995.

The less polar of the two principal soil metabolites of ATR produced by dealkylation, DEA, was detected in two precipitation samples. The much lower detection frequency relative to ATR is consistent with the presumption that this metabolite does not volatilize substantially more rapidly from plant or soil surfaces than ATR itself, and probably much less so. This is further supported by the observation that ATR and DEA elute within 3–4 min of each other when separated by gas chromatography (GC) on polysiloxane capillary columns, a widely used GC material similar in composition to the clay minerals of the western Iowa loess. However, the absence of published vapor pressure data or Henry's law constant estimates for DEA preclude more quantitative comparisons. Alternatively, the lower frequency of detections may result from increased sorption to the surface soil due to its greater organic matter content.

CYN, reflecting its use in only 5 years of the past 22 year record, was also detected very infrequently in precipitation. The first detection for CYN 4 years after its most recent documented application to W3 suggests non-local transport as a mechanism for this occurrence. While the estimated Henry's law constant, k_H (calculated from water solubility and vapor pressure data at 293 K) for CYN is six to seven orders of magnitude greater than that for ATR,³¹ the lower frequency of detections and lower mean concentration may merely be a consequence of much lower corn belt usage. ATR was applied on three times the acreage of that for CYN between 1992 and 1994.³²

Other recent studies have confirmed the presence of these parent herbicides in Iowa rainwater: three sites across the state, including one in southwestern Iowa, sampled between 1987 and 1990,³³ and two sites in central Iowa sampled between 1991 and 1994.³⁴ Both reported a seasonal variability of detections, with most occurring between April and July. The timing of chemical applications, together with the warmer daily temperatures, contribute to greater volatilization. No data for any metabolites or other degradation products were reported in either study.

In a regional assessment of herbicides in midwestern rainfall conducted from 1990 to 1991, ATR and both DEA and DIA were detected, with DEA occurring nearly 10 times more frequently than DIA.³⁵ In this study, formation of DEA from ATR by photochemical transformation in the atmosphere is suggested to explain the greater frequency of these detections. MET was found more frequently and in higher concentration than ATR. This can be related to its greater use in the central US³² and its higher vapor pressure.³⁶ In our study, the maximum daily deposition rates for MET are estimated to be 120 µg ha⁻¹ in 1993, 14 µg ha⁻¹ in 1994, and only 5 µg ha⁻¹ in 1995. Variation in these rates for all the compounds were presumed to have been related to the timing of application, the physicochemical properties of the compound, and environmental conditions. In addition, 1993 rainfall across the midwest resulted in 200 year flooding events in this area. METMOR was not detected in rainfall during this period.

Detections in vadose zone water from lysimeters

Herbicides were detected in the vadose zone water collected with suction lysimeters installed at four multiple depth nest locations and at five single depth locations on W3.¹³ The results of 3 years of monitoring (1993–1995), during which 340 water samples were collected, are summarized in Fig. 3. Across all depths, ATR was detected in 50% of the samples at a median concentration of 0.8 µg l⁻¹, MET was detected in 70% at a median concentration of 8.3 µg l⁻¹, and CYN was detected in 10% at a median concentration of 0.6 µg l⁻¹. Both DEA and DIA were detected in 12% and 1% of the samples, and at median concentrations of 0.3 and 0.6 µg l⁻¹, respectively. METMOR was detected in 24% of the samples at a median concentration of 7.5 µg l⁻¹. Maximum concentrations measured were 97.5 and 14.8 µg l⁻¹ for MET and METMOR, respectively. Our results are consistent with other more recent field studies conducted on silt-loam soils in Indiana where both DEA and DIA were detected in shallow vadose zone water following a single ATR application in 1994.³⁷ In this study, DEA was detected more frequently during the first year than DIA.

Fig. 3 Concentration distribution for: (a) ATR, CYN, several triazine metabolites, MET, and one degradation product in 340 vadose zone water samples collected on W3 between January 1993 and December 1995; and (b) ATR, CYN, MET, and one degradation product in vadose zone water samples collected beneath the grassed waterway on W3 between January 1993 and December 1995.

Vadose zone water content and matric potential at 2 m depth are mostly influenced by precipitation, evaporation, transpiration, and infiltration. Within this depth and in the waterways, suction lysimeters captured percolating recharge. Between 1991 and 1995, CYN was not applied directly to the area of study on W3. However, in 1994, it was applied to an area of approximately 4 ha directly upgradient from and draining directly onto W3. As a result, CYN was detected at 0.5–1.0 m depth in lysimeters installed beneath grass-covered waterways in the basin valley. This is an area just above the water table which intercepts overland runoff which has been diverted and slowed by the grassed waterway, resulting in added infiltration into the soil. The distribution of detections beneath the grassed waterways is shown in Fig. 3(b), a subset of the total observation presented in Fig. 3(a). MET movement was detected in the vadose zone at higher concentration than ATR. The presence of METMOR suggests an abiotic cyclization process, perhaps catalyzed by weakly acidic sites on mineral soil components or within the organic matter. As shown in Fig. 4, this largely geochemical process would proceed through the demethylated methoxymethyl side chain of MET. This chemistry is corroborated by laboratory hydrolysis experiments using dilute mineral acid at pH 5 which yield primarily METMOR.³⁶ Furthermore, photodegradation of MET in the presence of soil mineral and organic constituents yields METMOR as the major degradation product.³⁸ Below 2 m, the chemicals are carried by percolating water destined to become groundwater recharge. The violin plots in Figs. 3(a) and 3(b) illustrate the distribution of detections for each compound and at each depth. Parent herbicides were predominant and showed higher concentrations in the upper horizons of the unsaturated soil profile.

Fig. 4 Acid-catalyzed conversion of MET to METMOR in loess materials by soil mineral/organic matter surfaces.

Two interpretations are suggested to account for the vadose zone results. First, differences in the physicochemical properties for DEA, DIA, ATR, and MET are consistent with the “chromatographic analogy” conceptual model. For the separation of a mixture of solutes dissolved in a mobile phase moving through a uniformly porous material at constant velocity, the resolution equation defining a chromatographic separation incorporates capacity factor, selectivity factor, and efficiency factor terms. In the loess soil, as each compound moves through the vadose zone, both capacity and selectivity are influenced to varying degrees by the water solubility of the sorbate and its adsorption/desorption coefficients for that sorbent. Generally, leaching of non-ionic organic compounds through the soil is determined by the water solubility, sorption coefficient, ionicity of the sorbate and/or sorbent at soil pH (estimated from pK_a for specific functional groups or ion exchange capacity), and organic matter content.³⁹ Hence, rainfall percolating through the vadose zone gives rise to a fractionation of solutes developed over time which is analogous to a chromatographic separation. In the isocratic reversed-phase liquid chromatographic separation of these compounds using aqueous acetonitrile–water as mobile phase, the observed order of elution is DIA, DEA, ATR, and MET.²⁷ The relationship between soil sorption coefficient and HPLC capacity factor has been estimated using both isocratic⁴⁰ and gradient⁴¹ separations on a variety of packing materials. The chromatographic analogy requires several assumptions: first, that the mobile phase composition remains constant; second, that soil organic carbon content diminishes from ∼2–3% at the surface to ∼0.2–0.3% at 30 cm depth, and remains constant below that depth; third, that DIA, DEA, ATR, and MET do not differ greatly in their relative retention behavior; and, fourth, that the soil texture remains relatively unchanged throughout the profile. The analogy is illustrated in Table 1 where water solubility, soil sorption constant normalized to organic carbon (K_OC), Freundlich soil sorption constant (K_f), and hysteresis soil coefficient (ω) for DEA, DIA, ATR, and MET are compared. Both the selectivity and efficiency factors are functions of the sorbent material. The properties of each herbicide sorbate determine the magnitude of the hysteresis coefficient. Within the triazine series, greater water solubility is associated with a reduced sorption coefficient. Furthermore, studies of DIA, DEA, ATR, and MET sorption to soil confirm that desorption hysteresis occurs and that it is not significantly influenced by soil type or initial concentration of herbicide.⁴² Therefore, the residence time within a soil volume is a function of the adsorption/desorption relationship defining hysteresis, and it is this effect which accounts for the variability in frequency of detections and in mean concentration distributions in the vadose zone for all of the chemicals. In the only published report of herbicide retention on loess material, thin-layer chromatography of ATR and its metabolites on plates coated with soil taken from our study area reported mobilities generally consistent with our chromatographic analogy.⁴³

Table 1 Comparison of water solubility and the soil sorption parameters of K_OC, K_f, and ω for ATR, DIA, DEA, and MET

	Elution order^a	Retention^b[(t1/tMET) × 100] (%)	STLC^c (RF)	Water solubility^d/ mg l^−1	K OC ^d Emporia soil/l − (kg C)^−1	K f ^d (Freundlich) Emporia soil/µmol^1 − 1/n l^1/n kg^−1	Hysteresis Emporia soil^d (ω)
a Reversed-phase HPLC on C18 packing.^18 b % of reversed-phase elution, relative to MET. c STLC, soil thin-layer chromatography on plates coated with Treynor soil.^43 d ω = [(nads)^−1/(ndes)^−1 − 1] × 100.^42 e Not reported.
DIA	1	42	0.35	980	130	1.24	150
DEA	2	53	0.48	2700	110	0.89	196
METMOR	3	78	^e	^e	^e	^e	^e
ATR	4	84	0.35	33	234	1.51	223
MET	5	100	0.35	530	162	1.28	264

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Second, the indigenous consortium of microorganisms present in the loess soil may be capable of degrading ATR and its degradation products more rapidly than they are capable of degrading MET and its degradation products. Estimates of the half-life for ATR in various soils range between 40 and 70 days, whereas estimates for MET range between 30 and 100 days.⁴⁴ Annual applications of ATR began in 1972 and continued for 19 of the succeeding 22 years, while MET applications only began in 1980 and continue to date (see Fig. 1). The adaptability of a population of soil microorganisms following repeated exposure to a pesticide, and resulting in an enhanced biodegradation of that pesticide, is known to occur in soils amended with organophosphate and carbamate insecticides, as well as for chlorophenoxy acid herbicides.⁴⁵ Moreover, within the triazine series of herbicides, the further degradation of DEA to didealkylatrazine is reported to occur twice as fast in a surface soil with an ATR use history than in a soil with no ATR use history.⁴⁶ As a result, the soil biodegradation half-life for MET may be greater than that for ATR, thus making more material available for movement through the soil profile. Generally, the biodegradation of pesticides results in degradation products which are more hydrophilic than the parent compounds because they are more polar. Under the chromatographic analogy, DIA exhibits less hysteretic effect than DEA; therefore mobility through the unsaturated loess is proportionately greater. This relatively rapid transmissivity of both water and solutes has been observed for nitrate movement through this landscape, wherein monthly sampling has been shown to be insufficient to capture all of the variability in the flux of chemicals moving rapidly through this soil profile.¹³

Because DEA exhibits greater hysteresis than DIA (see Table 1), it is retained in the soil for a longer period of time, thus allowing more time for it to undergo further breakdown. This results in an increased probability that it will undergo further biodegradation. The thickness of the vadose zone on W3 varies from nearly 30 m on the ridge tops to less than 5 m in the valleys. Landscape positions at higher elevation provide the multiple month residence times necessary for enabling these processes to occur.¹²

Our data show that the deep loess material, under typical rainfall patterns, conducts both water and soluble herbicide-derived chemicals through the vadose zone at rates probably greater than those observed for glacial till landscapes in the corn belt of the upper midwestern US.

Detections in well water

Herbicides and their degradation products were also detected in the groundwater collected from wells installed at seven locations on W3.¹³ Five wells are 5 cm id galvanized steel open piezometers installed into the loess in 1989; two are screened PVC installed into the glacial till in 1991. The results of 4 years of monitoring are summarized in Fig. 5. In the 86 groundwater samples collected from W3 between January 1992 and December 1995, ATR was detected in 30% at a median concentration of 0.5 µg l⁻¹ and MET was detected in 17% at a median concentration of 0.3 µg l⁻¹. DEA was detected in 5% at a median concentration of 0.4 µg l⁻¹ and DIA was detected in 16% of the samples at a median concentration of 0.6 µg l⁻¹. ATR was detected nearly twice as frequently as MET, and no MET degradation products were detected. DIA reaches the water table more rapidly than DEA because it is adsorbed/desorbed from the vadose zone loess more rapidly. In addition, the less mobile DEA may undergo subsequent degradation resulting in fewer detections in the well water. As a consequence of its more rapid adsorption/desorption interaction with the soil (see Table 1), DIA moves more rapidly to the water table. MET moves more slowly and is retained longer in the unsaturated zone where it may undergo degradation. Most published studies of ATR leaching to shallow groundwater have not attempted to assess a nearly 20 year record of use. While the movements of ATR, DIA, and DEA into the saturated loess are documented in our studies, their occurrences are at very low concentrations and seem to be seasonally dependent. Generally, in studies focusing on ATR and MET together, the field dissipation in cultivated soils for MET is approximately 50% greater than that observed for ATR.⁴⁷ However, persistence variability for both herbicides may simply be a reflection of the natural field variability of microbial biomass. The two deepest wells on W3, both screened at more than 15 m depth below the loess/till interface, showed no detections for any of the herbicide-derived chemicals at any time during the monitoring effort. This suggests that the impacts on groundwater quality of agricultural practices using herbicides begun 28 years ago on W3 are limited to the shallow resource within the saturated loess and well above the interface.

Fig. 5 Concentration distribution for ATR, DIA, DEA, and MET in 86 well water samples collected on W3 between January 1992 and December 1995.

Acknowledgements

The authors wish to thank several anonymous reviewers for their comments. Special thanks go to Dr. Franceska D. Wilde and Dr. Ellen L. Arthur for their careful reading of the manuscript, insightful comments, and clarifying suggestions. We also thank Dr. David Meek for his helpful advice regarding the violin plots. Mention of specific products, suppliers, or vendors is for identification purposes only and does not constitute an endorsement by the US Department of Agriculture to the exclusion of others.

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Footnote

† This is the work of United States government employees engaged in their official duties. As such it is in the public domain and exempt from copyright. © US government.

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