Pesticide Decontamination and Detoxification - American Chemical

Chapter 15

Comparison of Atrazine and Alachlor Sorption, Mineralization, and Degradation Potential in Surface and Aquifer Sediments 1,*

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Downloaded by CORNELL UNIV on May 22, 2017 | http://pubs.acs.org Publication Date: October 10, 2003 | doi: 10.1021/bk-2004-0863.ch015

Sharon A. Clay , David E. Clay, and Thomas B. Moorman 1

Plant Science Department, South Dakota State Northern University, Plains Biostress Laboratory, Brookings, SD 57007 National Soil Tilth Laboratory, Agricultural Research Service, U.S. Department of Agriculture, Ames, IA 50011 *Corresponding author: Fax: 605-688-4452 2

Atrazine and alachlor have been used for controlling weeds in crops. These herbicides and their degradation products also have been detected in surface and ground waters. The objective of this study was to determine atrazine and alachlor sorption and degradation potentials at four soil depths and in aquifer matrix materials. Sorption potentials decreased with sample depth but were similar between herbicides at each depth. Mineralization potential differed between herbicides with twice as much alachlor mineralized at each depth compared with atrazine. After 112-d incubation at 10 C, atrazine degradation products were not detected in sample extract, whereas alachlor degradation products were present in both soil and aquifer matrix samples. Alachlor degradation products detected in A-horizon soil included the ethanesulfonic acid (ESA) and oxanillic acid (OAA) forms of alachlor. In aquifer extracts, two degradation products were formed, one cochromatographing with OAA-alachlor and an unidentified product. The ESA degradation product of alachlor has been detected frequently in aquifers at higher concentrations than alachlor. Its presence may be due to transport through the soil profile since this product was not detected in incubated samples.

© 2004 American Chemical Society Gan et al.; Pesticide Decontamination and Detoxification ACS Symposium Series; American Chemical Society: Washington, DC, 2003.

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Introduction Atrazine [6-chloro-iV-ethyl-N '-( 1 -methylethyl)-1,3,5-triazine-2,4-diamine] has been and continues to be one of the most used soil and foliar applied herbicides for control of broadleaf and grass weeds in com (Zea mays) in the United States. About 33 million kg of atrazine were applied in 1997 (1). Alachlor [2-chloro-A^-(2,6-diethylphenyl)-iV-(methoxymethyl)acetamide] (Figure 1) also has been widely used in mid-western United States com and soybean (Glycine max) production as a soil treatment for control of problem grass weeds. Prior to 1999, about 23 million kg of alachlor was applied annually in the United States (1). In 1994, other chloracetamide compounds, including metolachlor [2chloro-#-(2-ethyl-6-methyiphenyl)^ million kg used in 1997 and ranked number 2 behind atrazine usage) and acetochlor[2-chloro-AHetooxymethyl)-^ (15 million kg used in 1997 and ranked number 3 in soil herbicide usage), began to replace alachlor in the marketplace (1) (Figure 1). These four herbicides have been detected in aquifers throughout the United States (2). The amount of each herbicide detected in lakes, streams, resevoirs, and aquifers has been estimated to be not more than 1.5% of the total amount applied annually (3,4). The percentage of ground water samples containing these herbicides range from 0.09 for acetochlor to 30% for atrazine (2, 5). The maximum concentrations for individual detections rangefrom0.02 ug/L for acetochlor to 5.4 ug/L for metolachlor (2). Each of these herbicides degrades in the environment. Each degradate formed differs in it's degradation rate and soil retention properties (5). Atrazine is degraded microbiaily to deethylatrazine (DEA; 2-amino-4-chloro-6-isopropyl~ j-triazine) and deisopropylatrazine (DIA; 2-amino-4-chloro-6-ethylamino-.striazine) and by chemical hydrolysis to hydroxyatrazine (HA; 6-hydroxy-4ethylamino-2-isopropylamino-s-triazine) (6). The chloracetanilide herbicides degradation in soil is the result of either microbiaily mediated glutathione conjugation (7, 8) or by the hydrolysis of amine group (9). The glutathione conjugation reaction with subsequent cleavage of the tripeptide glutathione molecule yields many polar compounds including ESA (ethanesulfonic acid) and OAA (oxanillic acid) metabolites of each respective chloractetanilide herbicide (10) (Figure 2). In aquifers of the midwestern U.S., some of these herbicide metabolites have been detected morefrequentlyand at higher concentrations than the parent molecule (3). Researchers in Wisconsin reported that alachlor, metolachlor, and acetochlor were detected in 20 and 40% of monitoring and private wells tested, respectively, while only 2% of die municipal samples contained these herbicides (11) . At the same time, the ESA metabolites of each compound were detected in over 90% of monitoring and private wells and 50% of the municipal wells. In addition, OAA metabolites were detected in about 80% of the monitoring/private

Gan et al.; Pesticide Decontamination and Detoxification ACS Symposium Series; American Chemical Society: Washington, DC, 2003.

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(a)

CH2-CH3


CH2-O-CH3

CH2-CH3

W

^ 3

C

H -CH3 2

CH2-O

CH

3

^H-CHz-O-Chb C-CHz-CI Figure 1. Chemical structures of (a) alachlor, (b) acetochlor, and (c) metolachlor.


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CH2-CH

3

^2-0-^3 Ν X

C-CH -S03H 2

CH -CH Downloaded by CORNELL UNIV on May 22, 2017 | http://pubs.acs.org Publication Date: October 10, 2003 | doi: 10.1021/bk-2004-0863.ch015

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(b)

°

3

Ch^-CHjJ , CH2-O-CH3 -N

COCOOH CH2"CH3

CH2"CH3

CH -CH 2

0 3

Figure 2. Chemical structures of selected alachlor metabolites, (a) ESA metabolite of alachlor, (b) OAA metabolite of alachlor, (c) 2,6-diethylaniline and (d) 2,6-diethylacetanilide.



203 wells and 40% of the municipal wells. In studies reported by Kolpin et al. (3), the ESA-alachlor metabolite was detected 10 times more frequently and at concentrations two times greater than alachlor. The ESA metabolites have higher water solubility than the original compounds (12) and it is thought that the relatively higher concentrations and greater detection frequencies of these metabolites may be the result of faster leaching through the soil profile. In contrast, the DEA metabolite of atrazine was found at the same frequency and concentration as atrazine while DIA was detected half as often and at half the maximum concentration of atrazine (3). The stability of DEA and atrazine in the environment may play a role in thefrequencyand maximum detection. The toxicological properties of each herbicide are evaluated in stringent testing procedures. The toxicological properties of the herbicide dégradâtes often are unknown due to the numerous and varied products that can be formed under the wide range of environmental conditions. There are several questions that are relevant. Are these major metabolites deactivated? Are these metabolites detoxified? Only a few of the metabolites have been assessed for toxicity and the methods vary widely due to the type of toxicity in question. For example, toxicity can be discussed in a number of different ways including acute toxicity, chronic toxicity, life-time exposure, oncogenicity, and genotoxicity. The toxicity rating can be evaluated on an array of species including plants, invertebrates, vertebrates, or human exposure. While these dégradâtes may be less phytotoxic to both crop and weed species, their toxicity to other species may increase, decrease, or remain similar to the original herbicide. Using the Microtox method of toxicity assesment that measures acute toxicity, DEA, DIA, and HA were reported to be less toxic to photobacterium than atrazine whereas the hydoxyalachlor metabolite had a toxicity rating similar to alachlor (13). The ESA-alachlor was found to pose little or no risk of producing adverse effects when assessing subchronic, genotoxic, and developmental toxicity to mice and rats (14). In further studies, ESA-alachlor was poorly adsorbed in the rat and those injesting ESA-alachlor showed no oncogenic responses, unlike the response to alachlor (15). Aquifers of eastern South Dakota, western Minnesota, and northern Iowa are major sources of potable water for local communities. A high percentage of these aquifers are vulnerable to herbicide contamination due to: 1) their proximity to the soil surface (2 to 20 m below the surface); 2) intensive farming practices immediately above the aquifers; 3) soil types above the aquifers that are high in sand and/or low in organic matter; and 4) intense storm events where rainfall amounts can exceed 10 cm in a single storm. In studies conducted above the Big Sioux Aquifer in eastern South Dakota, atrazine was detected more often and at higher concentrations than alachlor although these two herbicides were applied at a similar application rates (16). However, alachlor was found much more frequently than atrazine in screening studies using monitoring wells across several eastern SD sites placed in several different aquifers (17).


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Although alachlor is no longer used in the U.S., the three chemical compounds have very similar structural (Figure 1) and chemical properties. Alachlor degradataion data may be useful as a model for this chemical class. Caution must be used in interpolating these data however since the ESA metabolite of metolachlor is formed more slowly and at lower concentrations in soil (18). The objective of this study was to compare atrazine and alachlor sorption, mineralization, and degradation potential, processes that are major contributors to the environmental fate of pesticides, from surface soil to aquifer sediments in laboratory studies. In addition, degradation of alachlor was compared under aerobic and anaerobic conditions.

Materials and Methods Surface and subsurface soil and aquifer matrix material (to an 8-m depth) were collected in an aseptic manner in presterilized 6.35 cm (width) PVC tubes using a hollow-stem auger from six sites over the Big Sioux Aquifer near Aurora, SD (19). The water table at the time was 7 m below the soil surface. The soil or matrix material was removed from the tubes in a sterilized laminar flow hood. The profile was divided into 6 zones that were designated as A (surface soil,from0- to 0.3-m), Β (soilfromthe Β horizon,from0.31 to 1.2-m), CI (soilfromthe C horizon,from1.4- to 3.5-m), C2 (soilfromthe C horizon, from 3.6- to 5.0-m), F (aquifer matrix material recovered just above the saturated zone,from5.2- to 6.6-m) and S (aquifer matrix materialfromthe saturated zone, from 6.8- to 8.8-m). All equipment used in this study was autoclaved or surface sterilized prior to use to prevent contamination. Bacterial counts were similar throughout the soil profile rangingfrom8.47 to 7.98 log bacteria/g for the A and F zones, respectively (19). The number of fungal organisms, however, differed by orders of magnitude in the profile,from6.81 to 1.28 log fungi/g for A and S zones, respectively. Adsorption of atrazine and alachlor to soil and aquifer material was determined for each zone using batch equilibration techniques (19). To investigate alachlor degradation and mineralization, samples that consisted of 30 g of material were transferred into sterilized glass serum bottles. In order to achieve anaerobic conditions in selected treatments, the noncapped serum bottles containing sample were exposed to an anaerobic atmosphere (N +C0 gas mixture) in an anaerobic glove box for at least 4 h or until equilibration occurred. The treatments included: 1) alachlor alone (aerobic and anaerobic conditions, all zones); 2) alachlor + carbon (aerobic and anaerobic conditions, all zones); 3) alachlor + nitrogen (aerobic and anaerobic conditions for F and S zones only); 4) alachlor +carbon + nitrogen (aerobic and anaerobic conditions, 2

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205 A, F, and S zones). Alachlor was added at 16.6 pg/kg as a 1-ml aqueous aliquot labeled with 5 kBq of C-uniformly-ring labeled herbicide. Carbon was added to the herbicide aliquot as lyophilized algae biomass at a rate of 3.3 mg/kg. This mimicked the flush of dissolved organic carbon that has been detected during spring recharge of the aquifer (20). Nitrogen was added as the N0 -N form at 0.8 mg N/kg. After treatment, sterilized water was added to ail samples. Samplesfromthe A to F zone were brought up to 30% water whereas samplesfromthe S zone were brought up to a 1:1 soil/water ratio. Samples were capped and incubated at 10 C (the ambient temperature of the Big Sioux Aquifer). Each treatment for a given zone was replicated six times. Aerobic samples were purged and aerated every 14 d up to 112 d, whereas anaerobic samples were purged only on 112 d. Compressed air was passed through a Drierite/Ascarite column to remove water and C 0 and a silver filter to remove airborne microorganisms. The purged airfromthe sample was run through KOH traps to trap C 0 and the amount of C was quantified using liquid scintillation techniques. After thefinalaeration, the remaining herbicide and dégradâtes were extracted from the sample using 30 ml of 4:1 methanol/water solution. Methanol was removed by evaporation. The total amount of C extracted was quantified in an aliquot by liquid scintillation counting. Another aliquot was spotted on thin layer chromatography plates and developed using butanol/acetic acid/water (6:2:3 v/v/v) to determine the amount of C that remained as alachlor. Alachlor had a relative mobility (Rf value) of 0.75 and 88% of the C-alachlor standard was detected in this band. Therefore, this band was scrapedfromeach sample track to quantify the amount of alachlor remaining. Other Rf bands from selected treatments were scraped to provide information on other dégradâtes. The Rf value of three dégradâtes, 2'6-diethlanaline, 2-chloro-2'6'diethlacetanilide, and ethane sulfonic acid {[2-[(2,6-diethyl-phenyl) (methoxymethyl) amino]-2-oxoethanesulfonic acid}, were similar with an Rf value of 0.65. The oxoacetic acid metabolite (OAA) {[(2,6-diethylphenyl) (methoxymethyl) amino] oxoacetic acid} had an Rf value of 0.45 (Figure 2). Other bands of more polar, unidentified C compounds were found in some treatments and had Rf values rangingfrom0.05 to 0.35. Atrazine mineralization was investigated in a very limited number of treatments. Aerobic samplesfromall zones were treated with atrazine at 16.6 pg/kg as a 1-ml aqueous aliquot labeled with 5 kBq of C-uniformly-ring labeled herbicide. The samples were aerated as described above, and the amount of C trapped was quantified. l4

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Results and Discussion Physical and chemical characteristics of the A horizon soil and aquifer matrix material have been reported in by Clay et al. (19). The soil in the A


206 horizon had a silty clay loam texture (15% sand, 56% silt, and 27% clay). The C horizon was glacial outwash, classified as a sandy loam soil, with 90% of the silt particles made up of nonreactive quartz. Over 60% of the material collected from the F and S zones was gravel and rock materials >4 mm, the remaining smaller particles were similar in makeup to the silt particles in the C horizon. Organic carbon rangedfrom2688 to 44 mg/kg in the A horizon soil and aquifer matrix material, respectively (Table 1). Atrazine sorption (Kd value) ranged from 5.4 L/kg in the A horizon to about 0.4 in the F and S zones (Table 1). Alachlor had Kd values that rangedfrom5.8 (A horizon) to 0.4 (F and S zone material. Sorption in the F and S zones is most likely overestimated for both compounds since sorption studies were conducted on particles

Pesticide Decontamination and Detoxification - American Chemical

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