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a U.S. Geological Survey, 1730 East Parham Road, Richmond, VA 23228
b U.S. Geological Survey, P.O. Box 25046, MS 407, Lakewood, CO 80225-0046
c U.S. Geological Survey, 5231 South 19 Street, Lincoln, NE 68512
d U.S. Geological Survey, 5957 Lakeside Boulevard, Indianapolis, IN 46278
e U.S. Geological Survey, 934 Broadway, Suite 300, Tacoma, WA 98402
* Corresponding author (thancock{at}usgs.gov)
Received for publication January 12, 2007.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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Abbreviations: BSFOD, branched, serial, first-order decay • DEA, deethylatrazine • ESA, ethanesulfonic acid • foc, mass fraction of organic carbon in the soil • Kd, soil–water distribution coefficient • Koc, organic carbon-normalized distribution coefficient • LEACHM, Leaching Estimation And CHemistry Model • OXA, oxanilic acid • RZWQM, Root Zone Water Quality Model
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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Field studies of pesticides in the unsaturated zone can identify which compounds are transported from the land surface to the water table, especially in relation to their respective sorption properties and degradation rates. For example, pesticide compounds within the same chemical class may show similar patterns of persistence in a given environmental setting, but variations in structure among compounds within the same chemical class may also result in substantial variations in reactivity in the same medium (Mackay et al., 1997; Barbash, 2007). For most pesticide degradates, data for many parameters related to sorption and persistence are not available in the published literature, making it difficult to predict their persistence and rates of transport to the water table.
The intensity of pesticide use is a significant predictor of pesticide detections in ground water (e.g., Barbash et al., 1999), but soil properties (e.g., permeability, available water capacity, and organic carbon content) and land management practices (e.g., the use of subsurface drains, irrigation, and conservation tillage) also seem to be important controlling factors (Barbash and Resek, 1996). Improvements in the accuracy with which the presence of pesticides and their degradates in ground water can be predicted, therefore, are dependent on the acquisition of a more thorough understanding of the transport and fate of these compounds in the unsaturated zone—topics that have been summarized by a number of reviews (e.g., Barbash and Resek, 1996; Flury, 1996; Bergström and Stenström, 1998).
The movement of agricultural chemicals and their degradates through the subsurface is governed primarily by their transport in water. Typically, the movement of water and solutes through the subsurface has been assumed to occur by uniform transport (i.e., by migration within a single, homogeneous flow regime analogous to piston flow). However, the transport pathways along which pesticides and their degradates move from the land surface through the unsaturated zone to ground water are highly variable in space and time. Evidence from more than three decades of research has established that nonuniform or preferential transport in the subsurface is likely to be encountered—and may even be common—within a wide variety of hydrogeologic settings (Barbash and Resek, 1996; Flury, 1996; Kjaer et al., 2001; McMahon et al., 2006).
One consequence of this phenomenon is that substantial amounts of pesticides and their degradates may move rapidly through more mobile regions in the unsaturated zone in response to individual recharge events, thereby bypassing most of the rest of the soil matrix (Barbash and Resek, 1996). The downward movement of water and solutes may also be more rapid beneath localized, topographically low areas, a phenomenon known as "focused recharge" (Delin and Landon, 2002).
Although 42 pesticides and 40 of their selected degradates were targeted in this study (Capel et al., 2008), the compounds that were the primary focus included several chloroacetanilide and triazine herbicides and organophosphorous insecticides, which are among the most abundant and widely used pesticides in US agriculture (Kiely et al., 2004). The most commonly used chloroacetanilide herbicides include acetochlor (2-chloro-N-(ethoxymethyl)-N-(2-ethyl-6-methylphenyl)acetamide), alachlor (2-chloro-N-(2,6-diethylphenyl)-N-(methoxymethyl)acetamide), and metolachlor (2-chloro-N-(2-ethyl-6-methylphenyl)-N-(2-methoxy-1-methylethyl)acetamide). Among the triazine herbicides, the most extensively used compounds include atrazine (6-chloro-N-ethyl-N'- (1-methylethyl)-1,3,5 triazine-2,4-diamine) and simazine (6-chloro-N,N'-diethyl-1,3,5-triazine-2,4-diamine). The most commonly used organophosphorous insecticides include chlorpyrifos (O,O-diethyl O-(3,5,6-trichloro-2-pyridinyl) phosphorothioate), diazinon (O,O-diethyl O-[6-methyl-2-(1-methylethyl)-4-pyrimidinyl] phosphorothioate), and malathion (diethyl[(dimethoxyphosphinothioyl)thio]butanedioate).
The two most common pesticide classes discussed in this article are the chloroacetanilide and triazine herbicides. In general, compounds in these classes undergo transformation relatively rapidly in aerobic soil and more slowly in water (Capel et al., 2008). The chloroacetanilide herbicides may be transformed by a relatively large number of different pathways (most of them microbially mediated), resulting in the production of at least 12 different degradates for acetochlor, 22 for alachlor, and 21 for metolachlor (Stamper and Tuovinen, 1998; Lee and Strahan, 2003; Hladik et al., 2005). The degradates that have been examined most extensively for alachlor, acetochlor, and metolachlor have been their ethanesulfonic acid (2-oxoethanesulfonic acid [ESA]) and oxanilic acid (2-oxoacetic acid [OXA]) metabolites. Using the examples of metolachlor and atrazine, Capel et al. (2008) illustrate the pathways by which these two classes of herbicides are transformed.
The purpose of this study was to measure the concentrations and spatial distributions of agricultural pesticides in the unsaturated zone beneath five different agricultural settings to examine the relative importance of different factors in controlling the movement and storage of the compounds above the water table. The relative importance of different chemical properties, time elapsed since application, timing and magnitude of significant hydrologic events (rain or irrigation), and recharge rate were examined by comparing the spatial distributions and concentrations of a wide range of current-use pesticides among several sites with substantial differences in their soil properties, water table depths, and climatic conditions. In some agricultural settings, lysimeters located in areas with different rates of recharge (i.e., upslope vs. downslope) were used to examine the effects of spatial variations in the rate of recharge and the importance of focused recharge (from rain or irrigation water) on pesticide transport.
This investigation represents one of the first large-scale studies of the occurrence of current-use pesticides in the unsaturated zone that have included a large number of their degradate compounds. Comparisons between the amounts of pesticides entering the unsaturated zone (including the targeted pesticide degradates) and the amount exported to the water table provide an indication of the influence of the unsaturated zone on the transport and fate of pesticides through the subsurface. Computer models were used at the Maryland site to examine the relative importance of different physical, hydrologic, and chemical processes (including preferential flow) in controlling the transport and fate of these compounds within the unsaturated zone. Such information, coupled with an understanding of local hydrogeology and agricultural practices at each of the study sites, will help to enhance current understanding of the various processes and factors that govern the transport and fate of pesticides and their degradates in the unsaturated zone.
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Materials and Methods |
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TOPNOTESABSTRACTINTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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Installation, Sampling, and Analysis
A detailed description of the overall study design, instrumentation, and chemical analyses can be found in the introductory article for this series by Capel et al. (2008). At each study location, multiple suction lysimeters were installed at various depths in the unsaturated zone for the collection of water samples (Table 2
). Lysimeters were installed 3 to 15 mo before the first sampling for pesticides, which took place at the beginning of the 2004 agricultural growing season. Each lysimeter was 30 cm long and 5 cm in diameter and consisted of a PVC tube with a porous ceramic cup at the bottom and two nylon access tubes at the top that extended to the land surface. Water samples were obtained by applying a vacuum to the lysimeter for at least 3 h before sampling. The sample was extracted from the lysimeter by applying a positive pressure to one of the access tubes and collecting the sample that flowed out from the other tube. Lysimeters were sampled at least six times throughout the 2004 agricultural growing season at each site. Unsaturated zone water samples were analyzed for 42 pesticides and 40 of their selected degradates according to the analytical protocols described by Capel et al. (2008).
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In another companion article in this issue, Fisher and Healy (2008) infer the rates of water movement and storage in the subsurface on the basis of meteorological observations, measurements of soil moisture and tension, and water table fluctuations. Nolan et al. (2006) evaluated the factors influencing ground water recharge at these sites using chloride concentrations in the unsaturated zone and ground water samples.
Also in this issue, Vogel et al. (2008) present information on the collection and analysis of pesticides and degradates examined in wet deposition (rain) at these sites. Pesticide concentrations were quantified in rain at all locations except WA, where rainfall is minimal and generally occurs during the nongrowing season (Payne et al., 2007). For this study, information on the use of irrigation water from ground water pumping and surface water diversions from nearby canals was gathered from local farmers and water control boards.
Farm management practices used during 2 yr of the study (water years 2003 and 2004) were documented by gathering information from local farmers, commercial applicators of farm chemicals, and Cooperative Extension Agents from the U.S. Department of Agiculture. This information included crop type and planting and harvest dates; pesticide types, amounts, and application dates; tillage practices; and use of tile drains and irrigation (Table 1). The information was supplemented by the results from reconnaissance of the sites by study personnel.
Data Analysis
Degradate Fraction
To examine spatial and temporal variations in the proportion of a given pesticide that was present in the form of the parent compound or any of its degradates, the following quantity, referred to as the "fraction" of the compound in question, was computed for each degradate of interest:
 | [1] |
i n, and the brackets are used to denote molar concentrations (Steele et al., 2008). In the case where only a single degradate is examined (n = 1) for a particular parent compound, a fraction value of unity indicates that only the degradate was present and no parent compound was detected, whereas a fraction of zero indicates that only the parent compound was detected. If multiple degradates are examined for a given parent compound and the fraction is equal to one, only the degradate of interest (degradate(i)) was detected; if the fraction equals zero, then degradate(i) was not detected, but there may have been other degradates present in addition to the parent compound.
Partition Coefficients
Soil–water distribution coefficients (Kd) and organic carbon–normalized distribution coefficients (Koc) were calculated for atrazine and deethylatrazine (DEA) (6-chloro-N-(1-methylethyl)-1,3,5 triazine-2,4-diamine) from the unsaturated zone field data collected at one of the MD sites (lysimeters M22a, b, c, and d). Table 3
provides depth information for these lysimeters; other site information was provided by Capel et al. (2008). Data on analyte concentrations for the aqueous and solid phase and for the water content and porosity of the solid phase were available for most of the MD M22 lysimeter sites. However, M22c was the only lysimeter for which data on the concentrations of atrazine in water were available; consequently, values for Kd and Koc for atrazine could only be calculated directly for that one sampling point. By contrast, the detection of DEA in the aqueous phase at all of the MD22 sites made it possible to compute values of Kd and Koc at all of the MD22 sites for this degradate. These calculations used data on soil properties and analyte concentrations in the solid phase from samples collected on 23 Oct. 2003 and results from the chemical analysis of water collected from the lysimeters on 9 Sept. 2004. Values of Kd (in mL g–1 dry weight) for each analyte of interest were computed by dividing its concentration in the solid phase (µg g–1, dry weight) by its concentration in water (µg mL–1) (Hamaker and Thompson, 1972). For the observed unsaturated zone data, the following equation was used for this calculation because the soil solids that were analyzed—referred to herein as the "bulk soil"—also contained the interstitial soil water:
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b is the soil bulk density (in g mL–1), computed as follows:
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s is the particle density (assumed to be 2.65 g mL–1), and 
is the porosity (pore volume in mL per total volume in mL). Koc (mL g–1 OC) was computed as follows (Hamaker and Thompson, 1972):
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Modeling
At the MD site, two unsaturated zone simulation models—LEACHM and the Root Zone Water Quality Model (RZWQM) (Ahuja et al., 2000)—were used to simulate the movement of water and pesticides through the unsaturated zone. A detailed description of the overall model design, input parameters, and data analysis can be found in the companion articles in this series by Webb et al. (2008) for LEACHM and Bayless et al. (2008) for RZWQM. Simulations using LEACHM and RZWQM were not conducted at other sites in this study because of time constraints or a lack of available data for one or more input variables.
Detailed LEACHM simulations for the MD site provided independent estimates of recharge and solute flux that could be compared with the results obtained using the methods used by Fisher and Healy (2008) and Nolan et al. (2006). In LEACHM, water movement was simulated using Richard's equation (via the WATer-FLOw, or WATFLO routine) (Hutson, 2005). Input data for the model simulations included meteorological observations for water years 1995 through 2004, soil texture, bulk density, organic matter, and information on the timing and loading of pesticides and bromide (applied as a conservative tracer)(Webb et al., 2008). Values for parameters related to the hydraulic properties of the soil and solute dispersivity were adjusted during calibration runs to match observed variations in soil moisture content and bromide concentrations. Climate, geology, and agricultural management information from the study site was used to build the model, and simulations of water and chemical transport were compared with measured values for soil–water tension, soil moisture, and the concentrations of pesticides and degradates.
Values for most of the parameters used during model runs to characterize the physical and chemical properties of pesticides are listed in Capel et al. (2008). The only exceptions involved estimates of Koc and soil aerobic half-life for atrazine, DEA, metolachlor, metolachlor ESA (2-([2-ethyl-6-methylphenyl][2-methoxy-1-methylethyl]amino)-2-oxoethanesulfonic acid), and metolachlor OXA (2-([2-ethyl-6-methylphenyl][2-methoxy-1-methylethyl]amino)-2-oxoacetic acid). Values for these parameters were obtained from other literature sources or through model calibration using observations made at the NE East site, where the largest number of these compounds was detected simultaneously. Calibrated transformation rates were obtained using a branched, serial, first-order decay (BSFOD) model in LEACHM (Webb et al., 2008). Use of the BSFOD model for this purpose was based on the assumption that the sum of the rates of production of all of the degradates of interest for a given parent compound could not exceed the rate of disappearance of the parent compound. These reaction rates were estimated from temporal variations in the compound fractions (Eq. [1]) observed in samples collected from the N22b lysimeter at the NE East site (Webb et al., 2008). Because of their inherently site-specific nature, these calibrated half-life values may be substantially different from those reported by controlled laboratory studies (e.g., Gilliom et al., 2006). Pesticide transformation rates were adjusted for variations in temperature and water content but not for pH.
For some aspects of the analysis of the results from this study, statistical tests were used to examine the significance of relations of selected variables with time or depth below the ground surface. In such cases, a significance level (
) of 0.05 was used.
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Results |
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TOPNOTESABSTRACTINTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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Unsaturated Zone Solids
Few pesticides and degradates were detected in the solid samples obtained from the unsaturated zone beneath the study sites. Because of operational difficulties, solid-phase samples were not collected from the unsaturated zone of the IN or WA sites. No pesticides or degradates were detected in solid-phase samples collected from the NE site. At the CA site, the only pesticides detected in solid samples were simazine and trifluralin (2,6-dinitro-N,N-dipropyl-4-(trifluoromethyl)benzenamine). At the MD site, metolachlor and simazine were detected in 40% (n = 10, where n is the number of samples analyzed) of the unsaturated zone solid samples, and dieldrin was detected in 20% (n = 10). At one lysimeter in MD (M21a), diazinon and pendimethalin (N-(1-ethylpropyl)-3,4-dimethyl-2,6-dinitrobenzenamine) were detected.
At the MD M22 site, located in an upland recharge area of a cropped field, atrazine and DEA were detected in 75 and 100% of the samples (n = 4), respectively. However, atrazine and DEA were detected less frequently under the middle and lower field sites in this location than at the upland recharge site. Atrazine and DEA concentrations in the soil were highest closest to the surface (Fig. 1 ) and decreased with depth in the unsaturated zone at the M22 site.
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Herbicides were the type of pesticide detected most frequently in the unsaturated zone water samples. At CA, no parent compounds and only a few acetanilide degradates (metolachlor ESA, metolachlor OXA, and alachlor ESA) were detected in unsaturated zone samples (Table 2). At both IN locations, metolachlor and atrazine were the only parent pesticides detected and were detected only in one sample; however, among the transformation products of interest, the acetanilide degradates were detected frequently, whereas DEA was detected less frequently. At the MD site, neither acetochlor nor alachlor were detected. Metolachlor and atrazine and their degradates were detected frequently at the MD site, whereas degradates of acetochlor and alachlor were detected less frequently. At both NE locations, several parent compounds were detected, albeit infrequently, and typically in the same sample (from the same lysimeter N22b). Results at the NE sites were similar to those at the MD site, with metolachlor and the atrazine degradates being detected most frequently and with few detections of acetochlor or alachlor degradates. At NE East Site N22b, the concentrations of metolachlor and atrazine and their degradates were the highest observed for this study. At the WA site, few parent compounds were detected, but degradates of acetochlor, alachlor, and atrazine were frequently detected.
At the CA, IN, WA, and NE West sites, none of the targeted insecticides or fungicides were detected, whereas at the MD site, none of the organophosphate insecticides or fungicides were detected. Only two degradates of the insecticide fipronil (5-amino-1-[2,6-dichloro-4-(trifluoromethyl)phenyl]-4-[(trifluoromethyl)sulfinyl]-1H-pyrazole-3-carbonitrile) were detected, fipronil sulfone (5-amino-1-[2,6-dichloro-4-(trifluoromethyl)phenyl]-4-[(trifluoromethyl)sulfonyl]-1H-pyrazole-3-carbonitrile) and desulfinyl fipronilamide (5-amino-1-[2,6-dichloro-4-(trifluoromethyl)phenyl]-4-[trifluoromethyl]-1H-pyrazole-3-carboxamide), which were detected in the same sample from Site M20b. At the NE East site, the organophosphate insecticide chlorpyrifos was detected in several samples from N20b and in one sample from N20c. Metalaxyl (methyl N-(2,6-dimethylphenyl)-N-(methoxyacetyl)-DL-alaninate), the only fungicide detected in this study, was found at NE East Site N22b.
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Discussion |
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TOPNOTESABSTRACTINTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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At the MD site, 2 yr had passed since the last application of metolachlor, atrazine, or simazine on corn in 2002; soybeans were planted in 2003 and 2004. Few, if any, of the parent compounds analyzed were detected in unsaturated zone water samples. This was attributed to the potential influence of a variety of factors, including the time elapsed between application and sampling, the physical and chemical characteristics of the chemicals (water solubility, sorptive properties, and degradation rates), the characteristics of the soil, and water flux at the site. The effects of these factors are discussed in later sections.
At the NE East site, atrazine and metolachlor were applied during the sampling period in 2004 and in previous years. At both of the NE sites, acetochlor was applied 1 yr before sample collection. These parent compounds and their degradates were detected frequently in unsaturated zone water samples from these sites (Table 2; data for acetochlor and alachlor not shown). At both IN sites, 1 yr had passed since the last application of atrazine and acetochlor, and at least 2 yr had passed since metolachlor had been used. Few parent compounds were detected at the IN sites. Among the transformation products, acetanilide degradates were frequently detected, but DEA was less frequently detected. The degradate acetochlor sulfinylacetic acid (SAA) (2-({[N-(ethoxymethyl)-N-(6-ethyl-2-methylphenyl)carbamoyl] methyl}sulfinyl) acetic acid) was detected more frequently and at higher concentrations at the IN and NE sites than at the other sites in this study. Because its degradate fraction (Eq. [1]) was observed to decrease with increasing subsurface residence time (data not shown), acetochlor SAA may be a short-lived intermediate transformation product; its presence in the unsaturated zone may therefore be an indication of recent application of its parent compound (acetochlor) to the land surface. Studies conducted under controlled laboratory conditions are needed to characterize the persistence of this degradate.
Rain and Irrigation
Another input mechanism by which pesticides may reach the unsaturated zone is rainfall. The concentrations of pesticides and degradates were measured in wet deposition (rain) by Vogel et al. (2008). In general, the amounts of herbicides deposited to these fields from rain (and thus potentially entering the unsaturated zone) were minor in comparison with the amounts commonly applied to the land surface for agricultural purposes, usually representing less than 1% of the applied mass (Vogel et al., 2008).
Irrigation water can also add pesticides to the land surface and, hence, the unsaturated zone. At the NE East location, ground water is pumped for irrigation, whereas at the CA and WA locations, surface water transported in canals is used for irrigation. Samples of irrigation-well and canal water taken from these sites were therefore analyzed for the same suite of pesticides and degradates examined in the lysimeter samples.
Canal water used for irrigation at the CA location did not contain detectable levels of pesticides. At the WA and NE East locations, several pesticides and degradates were detected in irrigation water. At the NE East site, several triazine and acetanilide pesticides and degradates were detected in water samples from wells completed at 28 and 32 m below land surface and screened in the same hydrologic unit as the irrigation well. In particular, metolachlor ESA and alachlor ESA were detected at concentrations of up to 8 x 10–4 µmol L–1 (0.26 µg L–1). Several triazine and acetanilide pesticides and degradates and organophosphorus insecticides (including diazinon and chlorpyrifos) were detected in irrigation water from the canal in WA. Acetanilide and triazine herbicides were detected at concentrations of up to 3 x 10–4 µmol L–1 (0.1 µg L–1). These observations suggest that irrigation water contributed negligible amounts of all applied pesticide compounds relative to the amounts applied directly to the crops.
Focused Recharge and Nonmatrix Flow
Major hydrologic events (irrigation or rain) can cause surface ponding and thus the focused recharge of water. This phenomenon can cause much of the infiltrating water to bypass the more biologically and chemically reactive root zone, allowing relatively high concentrations of unreacted parent compounds to move deeper in the unsaturated zone.This results in an increased flux of pesticides to the unsaturated zone relative to other areas nearby (Flury, 1996; Tasli et al., 1996; Iqbal, 1999).
On 29 Apr. 2004, more than 10 cm of rain—the largest daily total in more than 2 yr—fell on fields at the NE East site. This large amount of rain led to intense runoff, resulting in the ponding of water in a topographically low area near the N22 lysimeter site, located in the down-slope end rows of a corn field. Pesticides, which had recently been applied to land, were transported to this ponded area and, under conditions of focused recharge, transported rapidly downward out of the root zone, decreasing their near-surface residence time and opportunities for their transformation. This resulted in relatively high concentrations of the parent compounds in the unsaturated zone (Table 2). Figure 2 shows the elevated concentrations of pesticides and degradates from May though July 2004 for the N22b site. A similar trend in pesticide and degradate concentrations was also observed at the shallow NE East Site N22a.
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Water table fluctuations driven by high-flow events in streams may also be a source of pesticides in the unsaturated zone near the streams in question. Through a process known as bank filtration, pesticide parent compounds, degradates, and other solutes that are present in the stream water may be transported downward through the streambed into the unsaturated zone beneath or immediately adjacent to the stream. After a large storm in May 2004 caused the water table to rise above the lysimeter located at the 2.7-m depth at NE East Site N20c, increases in the concentrations of acetochlor and three of its degradates were observed in the lysimeter. The concentrations measured in samples taken from the lysimeters at this site provided evidence of degradation and downward transport of the compounds over the rest of the summer. During this period, the formation and subsequent disappearance of acetochlor SAA provided further indication of its transient nature.
Storage, Transport, and Fate of Pesticides in the Unsaturated Zone
Storage
The majority of pesticides and degradates examined for this study were not detected in water- or solid-phase samples taken from the unsaturated zone at any of the sites. Thus, most of these compounds were either taken up by the crops, retained in the root zone, carried rapidly through the unsaturated zone without being detected, or transformed in situ to other compounds for which analyses were not conducted. By contrast, atrazine and DEA were detected in many of the solid-phase samples from the unsaturated zone at MD. This section focuses primarily on the transport and persistence of these compounds at this site.
Atrazine and DEA were detected frequently in solid-phase samples from an upland recharge area of a cropped field (Site M22) and less frequently under the middle and lower field sites (M21 and M20, respectively). At Site M22, while atrazine was more prevalent than DEA in the solid phase near the land surface (Fig. 1), the concentrations of both compounds decreased with depth through the unsaturated zone. No atrazine was detected in water from the 0.6-m lysimeter (Table 2), although both atrazine and DEA were detected at greater depths. Because atrazine was not applied to this field in 2004 or 2003, but was applied in 2002 when the field was planted in corn, its presence in the near-surface solids was likely to have been a consequence of previous applications. At the M22 site, fine-grained sediments containing woody debris overlie coarse, angular gravel. In a manner similar to the capillary barrier effect discussed previously with respect to the sand/silt contact at NE East Site N22b, gravel underlying finer-grained surface material can result in enhanced moisture retention in the finer-grained material because unsaturated gravel conducts moisture poorly. The woody debris and other refractory carbon in the overlying layer might therefore represent a low-conductivity repository in which atrazine becomes sequestered (Lesan and Bhandari, 2003), resulting in reduced contact with microorganisms and more limited exchange with mobile water (Gevao et al., 2001).
Slow leaching of atrazine from the surface was indicated by other data from the MD site. Atrazine concentrations in water samples from the lysimeter at 0.6 m, beneath the organic-rich surface layers, increased during the summer of 2004. According to Steele et al. (2008), atrazine was detected at low concentrations in ground water near the water table beneath M22 and the other two MD sites farther down the ground water flow gradient (M20 and M21, located down-slope from the M22 site).
Values for Kd and Koc were calculated for atrazine and DEA from the unsaturated-zone field data collected at MD lysimeters M22a, b, c, and d using Eq. [2] and [3], respectively (Table 3). The calculated Kd value for atrazine at M22c (4.4 m below land surface) was 2.07 mL g–1, whereas those for DEA ranged from 17.8 mL g–1 at the near-surface MD site (M22a) to 1.83 mL g–1 at the deepest site (M22d). These Kd values are similar to those reported in the literature for laboratory experiments; Seybold and Mersie (1996) reported Kd values for atrazine and DEA of 1.8 and 0.99 mL g–1, respectively.
The calculated Koc value for atrazine was 296 mL g–1 OC (at M22c), whereas for DEA the values ranged from 1980 mL g–1 OC at the near-surface MD site (M22a) to 203 mL g–1 OC at the deepest site (M22d) (Table 3). Such observations are not without precedent; at a single site in Champaign County, Illinois, Roy and Krapac (1994) observed variations in Koc spanning an order of magnitude for DEA and atrazine within the soil column and observed a general decrease in Koc with decreasing Kd. Given the fact that Kd values commonly show a direct relation with foc and that foc decreased with depth at the MD site (as is typically the case), the decrease in Koc values observed with depth at the MD site is therefore not unexpected. At the MD site, the nature of the organic carbon in the soil column likely varies with depth such that carbon in deeper layers has undergone more transformation over time. It is also possible that sorption equilibrium was not reached at this site, particularly in the upper soil column where water likely moves through the unsaturated zone rapidly.
Although the near-surface MD sites for the current study exhibited higher calculated Koc values than those commonly reported by previous studies, the values obtained for the deeper locations were similar to those in the literature. Mackay et al. (1997) reported a Koc value for atrazine of 100 mL g–1 OC, whereas Seybold and Mersie (1996) reported a Koc value for DEA of 80 mL g–1 OC. Differences between the partition coefficients computed from this study and those reported by these other authors may be attributable to differences in the soil and sediment types, moisture conditions, and the amount and type of organic material in the laboratory experiments relative to the materials at this particular field site.
The downward transport of atrazine and DEA through the unsaturated zone at MD Site M22 was simulated for the period 1 Oct. 1994 through 30 Oct. 2004 using LEACHM. The simulations used Koc values from Bayless et al. (2008) for atrazine and DEA (234 and 110 mL g–1 OC, respectively) and soil aerobic half-lives from Webb and Sandstrom (2008) of 11 and 241 d, respectively, obtained through LEACHM model calibration (R.M.T. Webb and M.W. Sandstrom, personal communication, 2007). Using soil property values measured from core samples obtained during well installation at the site, a 10-m-thick unsaturated zone model was constructed. Drawing on information from local farmers, the simulations involved the application of atrazine to the land surface at a rate of 0.96 mmol m–2 (208 mg m–2) before planting corn every other year from 1994 through 2004. This amounted to four applications totaling 3.86 mmol m–2 (832 mg m–2) for this period; no atrazine was applied in 2004. Two weeks after the last atrazine application on 27 Apr. 2002, the simulations indicated that 0.39 mmol m–2 (84 mg m–2) of atrazine remained in the 10-m soil column, with 0.06 mmol m–2 (13 mg m–2) in the top 10 cm, decreasing mass in each 10-cm layer to a depth of 1.8 m, and concentrations below the reporting level of 4.6 x 10–7 µmol dm–3 (µmol L–1) at greater depths. Accounting for the combined influence of the Koc of the two compounds and the amount of organic carbon in the soils, roughly 75% of the atrazine mass and 70% of the DEA mass at any depth was sorbed to the solid phase of the soil, with the remaining fraction being present in solution. The degradation of atrazine over the 2-wk period after the 27 Apr. 2002 application produced 0.019 mmol (4.2 mg) of DEA per m2 in the top 10 cm. Atrazine applications before 2002 and continued leaching resulted in the presence of 6 x 10–6 mmol (0.013 mg) of DEA per m2 in each 10-cm layer deeper than 8 m.
A snapshot of the simulated atrazine distribution on 31 Oct. 2004, more than 2 yr after the last application on 27 Apr. 2002, inferred the following for the 10-m-deep unsaturated zone column. Of the 3.86 mmol m–2 (832 mg m–2) of atrazine applied, none would have remained as the parent compound, and none would have exited from the bottom of the column; 0.87 mmol m–2 of atrazine would have undergone transformation to DEA, with a total of 0.18 mmol m–2 remaining in the 10-m unsaturated zone column, exhibiting concentrations decreasing from approximately 4.6 x 10–3 µmol dm–3 (µmol L–1) near the surface to 4.6 x 10–5 µmol dm–3 (µmol L–1) below 8 m. The mass of sorbed DEA predicted to have been in the unsaturated zone column was similar to that observed in the field data, but clearly atrazine is more persistent than the model predicts.
Potential for Leaching to Ground Water
The potential for a pesticide to leach through the unsaturated zone to ground water depends largely on five factors: (i) rate and timing of water flux, (ii) characteristics of the soil, (iii) rate and timing of pesticide application, (iv) sorption properties of the compound, and (v) degradation rates of the compound. For a given location, the first two factors are constant for all pesticides, whereas the last three are compound specific.
Figure 3 compares the presence or absence of parent pesticide compounds in the unsaturated zone with their sorptivity (Koc) and persistence (soil half-life). Data are included only for those pesticides that are reported to have been used in the counties where these study sites are located (Capel et al., 2008) (these diagrams are similar to that used by Barbash and Resek (1996) to examine national data on pesticide occurrence in ground water). In general, the parent pesticides that were detected in the unsaturated zone water samples during this study are clustered in a group characterized by relatively low Koc values (<200 mL g–1 OC) and long half-lives in soil (>20 d). Conversely, pesticides with shorter half-lives and/or higher Koc values were generally not detected. However, another pesticide-specific factor determining potential transport of pesticides to the subsurface is use intensity, which was not considered in this simple analysis.
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The molar concentrations of metolachlor and its ESA and OXA degradates (Table 2) were used to calculate the degradate fraction (Eq. [1]) for metolachlor ESA. Because metolachlor was detected in only a few locations—most notably the shallow NE East sites (N22a, N23a, and N23b)—spatial variations in the metolachlor ESA fraction resulted primarily from variations in the relative amounts of metolachlor ESA and OXA, rather than those of the parent compound.
The metolachlor ESA fractions increased with depth at the MD sites (Fig. 4 ). At the M20 site, for example, the metolachlor ESA fraction increased from a mean of 0.3 at the 0.6-m depth to a mean of 0.9 at the 4.3-m depth (Fig. 4a). This relation was found to follow a simple linear regression of metolachlor ESA fraction vs. depth that was statistically significant (R2 = 0.934; P < 0.001; n = 12). Similarly, at Site M22, the metolachlor ESA fraction exhibited a significant, positive relation with depth (Fig. 4b), ranging from a mean of 0.5 at 0.6 m to a mean of 0.9 at 9.1 m (R2 = 0.553; P = 0.004; n = 18). No significant changes in the metolachlor ESA fraction were observed with depth at the IN sites I30 or I32 (P > 0.05). In Nebraska, while a significant increase in the metolachlor ESA fraction was observed at NE Site N20 from 0.4 to 0.9 (R2 = 0.6667 P = 0.0250), values were generally lower at the NE Sites N22 (0.1–0.4) and N23 (0.27–0.34) and did not change with depth (data not shown). The NE East sites had more recent applications of metolachlor than the other sites (MD or IN).
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By fitting the BSFOD model to temporal variations in the observed fractions for metolachlor and its ESA and OXA degradates in lysimeter N22b at the focused-recharge site lysimeter at the NE East site, Webb et al. (2008) calculated soil aerobic half-lives for metolachlor ESA and OXA of 70 and 50 d, respectively; these values are consistent with the observation by Krutz et al. (2006) that metolachlor OXA undergoes mineralization in soil more rapidly than metolachlor ESA. Use of this procedure also yielded estimated half-lives of 11 d for atrazine, 241 d for DEA, and 23 d for metolachlor.
Taken together, these results suggest that the observed increases in the metolachlor ESA fractions with time are primarily a consequence of a faster transformation rate for metolachlor OXA than for ESA, rather than any substantial difference in the mobilities (Koc) of the two compounds. In addition, the observed changes in the metolachlor ESA fraction with time and depth are consistent with the loss of the (apparently) more labile metolachlor OXA at shallower depths in the subsurface, where the microbial community is likely to be more active. Over time, this would lead to increases in the fraction of metolachlor ESA that is present in the soils and, by inference, transported to greater depths in the unsaturated zone. The concentrations of metolachlor ESA and OXA at shallower depths were found to increase throughout the year (Fig. 5), although the increase for metolachlor ESA was greater than that for metolachlor OXA (data not shown). At greater depths, the rates of production of these degradates were lower, perhaps because of diminished microbial activity.
Comparisons among the different study sites indicated that the relative concentrations of metolachlor ESA and OXA also varied in ways that were consistent with faster degradation of metolachlor OXA relative to metolachlor ESA (Fig. 6 ). At the NE East site, which experienced the most recent applications of metolachlor (in 2004, during the year of sampling), the molar concentrations of metolachlor OXA were about twice those of metolachlor ESA in the shallow lysimeters (N22a, N22b, and N23a) (Fig. 6a and 6b), suggesting a roughly twofold higher rate of formation of metolachlor OXA relative to metolachlor ESA. For the deeper lysimeter (N23b), the ratio of metolachlor OXA to ESA concentrations is about unity. At the MD sites, metolachlor had been applied more than 2 yr before sampling, and metolachlor ESA was nearly always present at concentrations higher than those of metolachlor OXA, especially in the deepest lysimeters (M22c and M22d) (Fig. 6c). At shallower depths, the ratio of metolachlor OXA to ESA was usually closer to unity. Although other factors, such as different microbial communities and different loading of metolachlor in the different areas, could also be important, these differences in metolachlor ESA fractions seem to reflect the greater persistence of metolachlor ESA relative to the OXA degradate.
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The LEACHM simulations were used to estimate recharge rates at the MD M22 site (Webb et al., 2008). At this site, a calibrated LEACHM model predicted a recharge rate of 0.11 m yr–1, which was in close agreement with the recharge rate of 0.32 m yr–1 calculated by Fisher and Healy (2008). Using the LEACHM-derived recharge rate and observed mean concentrations for metolachlor ESA, metolachlor OXA, and DEA at the deepest lysimeter, flux estimates of 1.0, 0.11, and 0.024 µmol m–2 yr–1, respectively, were obtained. These values were in relatively close agreement with those calculated directly from the lysimeter data (3.0, 0.32, and 0.07 µmol m–2 yr–1, respectively) (Table 2).
The flux estimates calculated for this study for some degradates changed with depth in the unsaturated zone column for some sites (Table 2). The flux of DEA increased with depth at MD Sites M21 and M22 and decreased with depth at IN Sites I30 and I32. At the MD sites, atrazine seemed to be forming DEA throughout the unsaturated zone, whereas at IN, a shallow static water table and flooding of the unsaturated zone may flush chemicals over time, leading to lower concentrations with depth.
The fluxes of acetochlor ESA and OXA increased with depth at the WA site to 2.8 m (W22j); however, there was more variability in the flux of these compounds at the IN sites. Alachlor ESA and OXA concentrations did not change appreciably with depth at the IN and WA sites. Acetochlor and alachlor may have been used more recently at the IN and WA sites than at the other sites; however, no information was available from local growers to evaluate this hypothesis. At MD Site M20, the fluxes of acetochlor ESA and OXA and of alachlor ESA and OXA were found to decrease with depth. The M20 site may be affected by a different hydrogeologic matrix and water flow pattern more than the other MD sites. The flux of metolachlor ESA increased with depth for many sites, including both IN sites, MD Sites M21 and M22, and NE East Site N22, whereas the flux of metolachlor OXA increased only at MD Site M20. Consistent with the earlier discussion, metolachlor ESA seems to be longer-lived than metolachlor OXA, and these results are consistent with the general loss of metolachlor OXA observed at shallower depths in the subsurface. Over time and at most sites, more metolachlor ESA was found to remain in the soils and to be transported through the unsaturated zone to the ground water (Steele et al., 2008) than was the case for metolachlor OXA.
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ACKNOWLEDGMENTS |
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TOPNOTESABSTRACTINTRODUCTIONMaterials and MethodsResultsDiscussionREFERENCES |
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