agricultural pesticides and selected degradation products in five

Name /entc/26_1208 08/16/2007 03:37PM Plate # 0-Composite pg 65 # 1

Environmental Toxicology and Chemistry, Vol. 26, No. 12, pp. 000–000, 2007� 2007 SETAC

Printed in the USA0730-7268/07 $12.00 � .00

AGRICULTURAL PESTICIDES AND SELECTED DEGRADATION PRODUCTS IN FIVETIDAL REGIONS AND THE MAIN STEM OF CHESAPEAKE BAY, USA

LAURA L. MCCONNELL,*† CLIFFORD P. RICE,† CATHLEEN J. HAPEMAN,† LETICIA DRAKEFORD,†JENNIFER A. HARMAN-FETCHO,† KRYSTYNA BIALEK,† MICHAEL H. FULTON,‡ ANDREW K. LEIGHT,§ and

GREGORY ALLEN�†U.S. Department of Agriculture, Agricultural Research Service, Environmental Management and Byproduct Utilization Laboratory,

10300 Baltimore Avenue, Beltsville, Maryland 20705‡National Oceanic and Atmospheric Administration, National Ocean Service, Center for Coastal Environmental Health and

Biomolecular Research, Charleston, South Carolina 29412, USA§National Oceanic and Atmospheric Administration, National Ocean Service, Center for Coastal Environmental Health and

Biomolecular Research, Oxford, Maryland 21654, USA�U.S. Environmental Protection Agency, Chesapeake Bay Program Office, Annapolis, Maryland 21401

(Received 29 December 2006; Accepted 22 June 2007)

Abstract—Nutrients, sediment, and toxics from water sources and the surrounding airshed are major problems contributing to poorwater quality in many regions of the Chesapeake Bay, an important estuary located in the mid-Atlantic region of the United States.During the early spring of 2000, surface water samples were collected for pesticide analysis from 18 stations spanning the ChesapeakeBay. In a separate effort from July to September of 2004, 61 stations within several tidal regions were characterized with respectto 21 pesticides and 11 of their degradation products. Three regions were located on the agricultural Delmarva Peninsula: TheChester, Nanticoke, and Pocomoke Rivers. Two regions were located on the more urban Western Shore: The Rhode and SouthRivers and the Lower Mobjack Bay, including the Back and Poquoson Rivers. In both studies, herbicides and their degradationproducts were the most frequently detected chemicals. In 2000, atrazine and metolachlor were found at all 18 stations. In 2004,the highest parent herbicide concentrations were found in the upstream region of Chester River. The highest concentration for anyanalyte in these studies was for the ethane sulfonic acid of metolachlor (MESA) at 2,900 ng/L in the Nanticoke River. The degradationproduct MESA also had the greatest concentration of any analyte in the Pocomoke River (2,100 ng/L) and in the Chester River(1,200 ng/L). In the agricultural tributaries, herbicide degradation product concentrations were more strongly correlated with salinitythan the parent herbicides. In the two nonagricultural watersheds on the western shore, no gradient in herbicide concentrations wasobserved, indicating the pesticide source to these areas was water from the Bay main stem.

Keywords—Chesapeake Bay Pesticides Herbicides Metolachlor Atrazine

INTRODUCTION

The Chesapeake Bay Watershed (area � 166,000 km2), lo-cated in the mid-Atlantic region of the United States, includesportions of New York, Pennsylvania, Delaware, Maryland,Virginia, West Virginia, and the District of Columbia. Thisregion is home to approximately 15 million people. The wa-tershed contains approximately 150 major rivers and streamswith the Susquehanna River contributing 50% of the fresh-water flow to the 869-km long Chesapeake Bay main stem.Excessive levels of nutrients, sediment, and toxins from thewater sources and the surrounding airshed are major problemscontributing to poor water quality in many regions of thisecosystem. Toxins are defined as chemicals that may affect thereproduction, development, and, ultimately, the survival ofliving resources. Categories of toxins generally include in-dustrial chemicals such as polychlorinated biphenyls, metals,and solvents; combustion byproducts such as polyaromatichydrocarbons and dioxins; and pesticides from agricultural,home garden, and urban sources.

In 1993, three regions of concern with probable adverseeffects with respect to toxins were identified within the Ches-apeake Bay watershed: The Elizabeth, Baltimore Harbor, andAnacostia Rivers [1] (http://www.chesapeakebay.net/pubs/

* To whom correspondence may be addressed([email protected]).

233.pdf). Since then, 10 areas of emphasis were added alongwith eight areas with low probability for adverse effect and20 areas with insufficient or inconclusive data [2] (http://www.chesapeakebay.net/pubs/792.pdf). In 2000, the Chesa-peake Executive Council adopted the Chesapeake 2000 BayAgreement, committing to fulfill the 1994 toxics strategy goalof a ‘‘Chesapeake Bay free of toxics by reducing or eliminatingthe input of chemical contaminants from all controllable sourcesto levels that result in no toxic or bioaccumulative impact on theliving resources that inhabit the Bay or on human health’’ [3](http://www.chesapeakebay.net/pubs/chesapeake2000agreement.pdf). From the Chesapeake 2000 Bay Agreement, a new toxics2000 strategy was developed that called for completing thetoxics characterization, meaning that all areas with insufficientor inconclusive data will be characterized as either a regionof concern, an area of emphasis, or an area with low probabilityfor adverse effects [4] (http://www.chesapeakebay.net/pubs/toxicsstrategydec2000.pdf). This effort is still underway.

In the early spring of 2000, surface and deeper water sam-ples were collected from 21 stations spanning the ChesapeakeBay main stem (Fig. 1). These samples were collected in co-operation with scientists from a larger project entitled TheChesapeake Bay Land Margin Ecosystem Research, TrophicInteractions in Estuarine Systems Project led by ChesapeakeBiological Laboratory in Solomons, Maryland, USA. These

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Environ. Toxicol. Chem. 26, 2007 L.L. McConnell et al.

Fig. 1. Sample collection locations utilized in the 2000 main stem study in the Chesapeake Bay, USA, and its tributaries. Regions in the 2004toxics characterization study are labeled, and land use area (ha) within each of the five major regions included in 2004 toxics characterizationare provided [5]. Specific sample collection locations for the 2004 study are provided in Supplementary Figures S1–S5.


Pesticides in tidal regions of Chesapeake Bay Environ. Toxicol. Chem. 26, 2007

samples were characterized for a suite of agricultural and or-ganochlorine pesticides.

In a separate study, as part of the toxics characterizationeffort, an evaluation of the ecological condition and toxicscontamination in five tidal regions of Chesapeake Bay wasundertaken. In the late summer and early fall of 2004, 61stations within five tidal regions were characterized with re-spect to toxics: On the Delmarva Peninsula (the Chester, Nan-ticoke, and Pocomoke Rivers and on the Western Shore), theRhode and South Rivers, and the Lower Mobjack Bay in-cluding the Back and Poquoson Rivers (Fig. 1). Land usewithin the five regions varies from highly agricultural in theChester, Nanticoke, and Pocomoke watersheds to primarilyurban or forested in the Rhode and South Rivers or LowerMobjack Bay (Fig. 1) [5] (http://www.chesapeakebay.net/pubs/1127.pdf). At each station, samples were collected todetermine concentrations for pesticides and metals in surfacewater. Sediment samples also were collected for chemical con-taminants, sediment toxicity bioassays, and benthic commu-nity assessments.

The objective of the present study is to present the resultsof surface water analysis for pesticides and their degradationproducts from both studies. Results will be examined from aspatial perspective to discern information regarding the en-vironmental fate of these widely used agricultural chemicalswithin the Chesapeake Bay. These data also will be comparedto previous studies investigating pesticide concentrations inthe Chesapeake Bay main stem and its tributaries. The presentstudy will add to the overall body of knowledge with respectto the fate and transport of agricultural chemicals within thislarge estuary and provide the fundamental data required forfurther modeling and risk-assessment efforts ongoing in theChesapeake Bay toxics characterization program.

MATERIALS AND METHODS

Sample collection and preparation

Sampling site selection and characterization. During the2004 toxics characterization study, collection sites within eachtidal region were selected using a random probabilistic design.Each river was divided into segments (e.g., upper vs middlevs lower portions) approximately corresponding to salinityzones exhibiting similar within-segment salinity and hydro-graphic characteristics. A random point generator was used toselect the stations within each segment. Fifteen samples eachwere collected from the Chester and Nanticoke Rivers and 11samples were collected from the Pocomoke River. Ten sampleswere collected from the Rhode and South Rivers and 10 werecollected from the Poquoson and Back Rivers and Lower Mob-jack Bay region. Detailed maps of each region are providedin Supplementary Materials (Fig. S1–S5). All samples werecollected between the dates of July 14 and September 13, 2004(Supplemental Data, Table S1; http://dx.doi.org/10.1897/06-655.S1). Basic water quality parameters (salinity, pH, tem-perature, and dissolved oxygen) were measured at approxi-mately 1 m from the surface along with the total water depthusing a Hydrolab� Minisonde 4 water-quality datalogger (HachEnvironmental, Loveland, CO, USA). All water samples in thepresent study were collected at a 1-m depth unless the totaldepth was less than 1 m.

In the 2000 main stem study, water samples were collectedfrom 18 stations spanning the entire length of Chesapeake Baymain stem during the time period of April 29 to May 2, 2000(Supplemental Data, Table S2; http://dx.doi.org/10.1897/

06-655.S2). Water quality parameters (pH, salinity, and tem-perature) also were measured. Samples in this study also werecollected at 1 m and at three stations. A second, deeper watersample also was collected below the halocline.

Surface water sample collection and filtration procedures.All water samples in both studies were collected using a smallsubmersible pump that was joined by a length of Teflon� tubingto two inline, stainless steel filter holders housing a 142-mmdiameter, 1-�m pore size glass microfiber filter ([GMF], What-man, Florham Park, NJ, USA) and a 90-mm diameter, 0.7-�mpore size glass fiber filter (GF/F, Whatman), respectively. Ap-proximately 8 L of filtered water were transferred into a la-beled, precleaned, stainless steel container with an airtight seal.Containers were kept sealed during transport. Sampling equip-ment was cleaned between sites by flushing the pump, tubing,and filter heads with a solution of 1:1 deionized (DI) water:methanol through the system for 2 min. New filters were in-stalled at each station and water from each site was flushedthrough the system for 1 min prior to sample collection. Sam-ple containers were rinsed three times with filtered water fromeach site before being filled. Samples were kept on ice whileonboard the vessel and during transport to the laboratory andwere transferred to refrigerator storage at 4�C within 24 h aftercollection. Sample processing in the 2004 study was completedwithin 24 h after arrival at the laboratory. Samples collectedon the 2000 main stem study were processed immediatelyonboard the research vessel.

A daily trip blank (a clean glass bottle filled with 4 L ofDI water) was transported into the field and processed alongwith all the samples from that day to evaluate any contami-nation during transportation. Trip blanks were not required inthe 2000 main stem study because samples were not trans-ported. Daily field blanks also were conducted in both studiesby pumping 4 L of DI water through the filtration and collec-tion system into a clean stainless steel container. Blanks thenwere processed along with the samples.

Sample preparation for gas chromatography-mass spec-trometry analysis. Water samples were prepared for gas chro-matography–mass spectrometry analysis using procedurespublished previously [6]. An extraction surrogate, diazinon-diethyl-d10 (Cambridge Isotope Laboratories, Andover, MA,USA), was added to each 4-L sample just prior to extraction.Each sample was extracted using a solid phase extraction car-tridge containing 200 mg of hyper-crosslinked styrene-divinylbenzene copolymer extraction resin (ENV�, Biotage Products,Charlottesville, VA, USA). After extraction, the solid phaseextraction cartridge was dried with high-purity (99.99%) ni-trogen and was kept frozen (�20�C) until elution for a max-imum of 14 d. Cartridges were eluted with certified high-puritysolvents using 6 ml of dichloromethane followed by 9 ml of3:1 acetone:acetonitrile (all solvents were chromatographicgrade, Fisher Scientific, Pittsburgh, PA, USA). The extract wasconcentrated to a final volume of 0.5 ml under high-puritynitrogen, transferred to a 2-ml amber glass vial, and stored at�20�C until analysis. Atrazine-ethylamine-d5 (47.04 �g/ml)and 2,2�,3,4,4�,5,6,6�-octachlorobiphenyl ([PCB 204], 4.05�g/ml; both obtained from Cambridge Isotope Laboratories)were added as internal standards.

Three times during the 2004 toxics characterization study,three 4-L samples were collected, two samples were extractedto evaluate the precision of the sample extraction method, andthe third was spiked with a mixture of target analytes to eval-uate extraction efficiency in the river water matrix. Following



each day of the tributary and main stem trips, a 4-L DI watersample was spiked with a mixture of target analytes prior toextraction to evaluate extraction efficiency in the absence ofany matrix interferences.

During the 2000 main stem study, duplicate samples werecollected at every station; however, duplicate samples at threeof the 18 stations were lost during processing. A third samplealso was collected at one site on each day of the cruise toevaluate extraction efficiency in the sample matrix.

Sample preparation for high-performance liquid chro-matography/mass spectrometry analysis. Forty out of 61 tidalwater samples also were processed for high-performance liquidchromatography/mass spectrometry (HPLC/MS-MS) analysis.None of the 2000 main stem samples were analyzed by HPLC/MS-MS. A separate 4-L water subsample from each bulk sam-ple collected was extracted using an solid phase extractioncartridge (500 mg, Oasis HLB, Waters, Milford, MA, USA);each cartridge was preconditioned with 5 ml each of methanol,acetonitrile, and methanol again. Triphenylphosphate (Supel-co, Bellefonte, PA, USA) was added to each sample prior toextraction as a surrogate to evaluate extraction efficiency. An-alytes were eluted with methanol (2 � 5 ml) followed byacetonitrile (2 � 2.5 ml). The combined eluant was concen-trated to 0.5 ml under high-purity nitrogen and exchanged withan equivalent volume of methanol. Each sample extract wastransferred into a vial and brought to 1-ml volume with theaddition of 0.5 ml of DI water. Each sample was spiked with10 �l each of the internal standards monocrotophos (10 �g/ml; Supelco), [U-13C]-2,4-dichlorophenoxyacetic acid (100�g/ml; Cambridge Isotope Laboratories), and terbutylazine(35.6 �g/ml; Dr. Ehrenstorfer, Augsburg, Germany).

Analytical methods

Pesticide determination by gas chromatography–massspectrometry. Sample extracts were analyzed on an Agilent6890 gas chromatograph coupled to a 5973 inert mass spec-trometer (Agilent Technologies, Santa Clara, CA, USA) usingboth electron impact ionization and negative chemical ioni-zation modes (Table 1). Target analytes included both currentlyused agricultural pesticides and historically used organochlo-rine chemicals. In each case, selected ion-monitoring modewas utilized to maximize sensitivity. Separations wereachieved utilizing a 30-m DB-17 MS capillary column (J&WScientific, Folsom, CA, USA), 0.25-mm inner diameter, 0.25-�m film thickness, helium carrier gas, constant flow at 1 ml/min, injection carried out in pulsed splitless mode. The gaschromatography parameters under electron impact conditionswere as follows: 10.85 psi injection port pressure, 40 psi pulsepressure; temperature program of injector port temperature270�C, initial temperature 130�C, hold 1 min, 5�/min to 250�C,15�/min, hold 3 min. The detector interface, source, and quad-rupole temperatures were 300, 230, and 100�C, respectively.

Gas chromatography parameters for negative chemical ion-ization mode were: 11.85 psi injection port pressure, 50 psipulse pressure; temperature program of injector port temper-ature 230�C, initial temperature 130�C, hold 1.0 min, 6�/minto 205�C, hold 4.5 min, 6�/min to 300�C, hold 6.2 min. Thedetector interface, source, and quadrupole temperatures were300, 150, and 150, respectively. The flow controller on themass spectrometer was set to allow 40% of the ionization gas(methane) into the mass spectrometer. The pressure on theregulator was set to 5 psi.

Atrazine-d5 was used as the internal standard for the quan-

tification in electron impact mode and PCB 204 was used innegative chemical ionization. A five-point calibration curvewas established for each compound and instrument responsewas linear over the calibration standards range (r2 � 0.99).Standards were rerun after every 20 to 25 sample injections.Quantification of each compound was calculated based on thearea of the ion with the largest abundance.

Recoveries of the extraction surrogate diazinon-d10 wereacceptable during both studies and averaged 111 26% and89 8% in 2004 and 2000, respectively. Overall, matrix spikerecovery values for both studies were in an acceptable range(Table 1) with most averaging 70%. The average lab spikerecovery values for analytes in the 2004 study were all 77%.Concentration values were not adjusted for recovery results.Duplicate samples were collected throughout both projects.Differences between detected chemicals in duplicate samplesranged from 4 to 14% in 2004. Eighteen duplicate sampleswere analyzed in the 2000 main stem study (SupplementalData, Table S3; http://dx.doi.org/10.1897/06-655.S3). Of thefive major analytes observed in this study, atrazine, meto-lachlor, 6-chloro-N-(1-methylethyl)-1,3,5-triazine-2,4-diamine(CIAT), diazinon, and acetochlor, the percent difference be-tween duplicate analyte measurements was �22% for all butfour measurements. Therefore, the extraction method utilizedin these studies was effective in capturing target analytes andthe analytical method provides good precision.

Method detection limits were determined for each analyteas suggested by the U.S. Environmental Protection Agency[7]: The method detection limit values were determined fromanalysis of at least seven extracts from DI water spiked at thelowest point on the calibration curve for each analyte. Eachcompound method detection limit was calculated based on thestandard deviation of the average mass determined multipliedby the appropriate Student t-value. The limit of quantitation(LOQ) was defined as three times the method detection limitvalue.

Laboratory, trip, and field blanks were extracted along witheach batch of field samples. Overall results from the blankcontrols were very good. Analytes were detected only in a fewof these control samples at levels above our LOQ. In the 2004toxics characterization study, acetochlor was detected in twolaboratory blanks and pendamethalin was detected in one lab-oratory blank. Cyanazine residues were detected in the fieldblanks collected along the Chester and Nanticoke Rivers. Lowlevels of the three chlordane isomers (�-, �-, and trans-non-achlor) were detected along with -endosulfan and chlopyrifosin one field blank collected on September 12; however, noneof these chemicals were detected in the samples from this trip.If a chemical was detected in any type of blank, the value wasmultiplied by five, and concentrations falling below this blanklevel were removed from the final dataset. Data from eightextracts were lost due to faulty vials: One field blank, onematrix spike, and five samples (P12 in the Chester; P26 andP27 in the Nanticoke; P31 in the Pocomoke, and P47 in LowerMobjack Bay). Field blanks from the 2000 study did not showany interference at levels above the LOQ.

Pesticide degradation product determination by HPLC/MS-MS. Pesticide degradation products were analyzed usingHPLC/MS-MS. Structural information for the degradationproducts and full chemical names are provided in Table 2.Extract components were separated by HPLC conducted inreverse-phase mode as modified from Schroyer and Capel [8]utilizing a Waters 2690 Alliance module (Waters, Milford, MA,

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Table 1. Target analyte list for gas chromatography-mass spectrometry analysis, spike recovery results for both the 2004 toxics characterizationstudy and the 2000 main stem study, method detection limits (MDL), and limits of quantitation (LOQ)

CompoundMS

modea m/zb

2000 Avg.matrix spikerecovery (%)

(n � 4)

2004 Avg.matrix spikerecovery (%)

(n � 3)

2004 Avg. labpike recovery(%) (n � 3)

MDLc

(ng/L)LOQd

(ng/L)

Acetochlor EI 223, 162, 146 99 23 110 7 122 15 1.3 3.9Alachlor EI 160, 188, 237 98 23 114 10 113 8 1.3 3.9Aldrin NCI 330, 237, 332 71 10 47 12 73 8 0.41 1.2Atrazine EI 200, 215, 173 101 22 114 15 131 8 2.1 6.3Atrazine-ethylamine-de

5 EI 205, 220, 179 —f — — — —CIATg EI 172, 187, 174, 145 104 24 114 13 127 6 1.8 5.4CEATh EI 173, 158, 145 115 32 114 12 122 16 2.1 6.3Chlorpyrifos NCI 313, 315, 214 91 24 104 12 101 3 1.5 4.5�-chlordane NCI 410, 408, 412 75 19 61 9 91 2 1.1 3.3�-chlordane NCI 410, 412, 408 76 20 56 10 92 2 0.82 2.5Cyanazine EI 212, 225, 240 93 32 119 5 150 28 1.5 4.54, 4�-DDDi NCI 248, 250, 320 73 24 62 53 91 3 4.6 144, 4�-DDEj NCI 318, 320, 316, 281 53 15 20 29 78 5 0.93 2.8Diazinon NCI 169, 303 92 21 114 27 104 11 8.3 25Diazinon-diethyl-dk

10 EI 183, 314, 138, 153 89 8 111 26l — — —�-endosulfan NCI 406, 408, 412 85 20 114 5 122 6 1.7 5.0 -endosulfan NCI 406, 408, 404 81 31 103 11 115 6 2.5 7.6Endosulfan sulfate NCI 386, 388, 384 83 31 140 26 128 15 2.0 5.9Fipronil NCI 384, 331, 400 NAm 99 4 106 6 2.5 7.4�-HCHn NCI 255, 257, 71 86 19 91 5 77 19 2.2 6.5�-HCHn NCI 255, 71, 257 88 19 109 5 110 10 2.7 8.2Heptachlor NCI 266, 300, 232 83 15 97 19 94 5 0.95 2.9Heptachlor epoxide NCI 318, 237, 388, 282 87 22 74 11 97 3 1.2 3.7Metolachlor EI 162, 238 95 24 109 7 150 16 1.2 3.6Pendamethalin EI 252, 281, 191, 162 89 27 98 6 112 17 1.2 3.6PCB 204e NCI 394, 428, 430 — — — — —Simazine EI 201, 186, 173 94 24 114 12 124 10 2 6Trifluralin NCI 335, 305, 336 NA 72 9 88 3 0.74 2.2Trans-nonachlor NCI 444, 442, 412 70 17 52 10 92 1 0.82 2.5Cis-nonachlor NCI 444, 442, 446 NA 56 9 94 2 0.70 2.1

a MS � Mass spectrometer, EI � Electron impact, NCI � Negative chemical ionization.b Ions used for quantitation are in italics.c MDL for 4-L sample.d LOQ based on 4-L sample.e Atrazine-ethylamine-d5 and 2,2�,3,4,4�,5,6,6�-octachlorobiphenyl (PCB 204) were used as internal standards and no MDL values were determined.f An em dash indicates XXX.g CIAT � 6-amino-2-chloro-4-isopropylamino-s-triazine.h CEAT � 6-amino-2-chloro-4-ethylamino-s-triazine.i 4,4�-DDD � 1,1-Dichloro-2,2-bis[p-chlorophenyl]ethane.j 4,4�-DDE � 1,1-Dichloro-2,2-bis[4-chlorophenyl]ethylene.k Diazinon-diethyl-d10 was used as an extraction efficiency surrogate and no MDL values were determined.l Spike recoveries for diazinon d10 include all samples.m NA � Not analyzed.n � and � forms of 1,2,3,4,5,6-hexachlorocyclohexane.

USA). The separation unit was coupled to a triple quadrupolemass spectrometer (Micromass Quattro LC; Waters), equippedwith an electrospray interface (Z-spray) for sample detectionand quantitation.

Ten-microliter sample aliquots were separated on a C18 Ul-trasphere ODS column (4.6 mm � 25 cm; Beckman Coulter,Fullerton, CA, USA). A solvent system of 70% in water con-taining 1% formic acid with 30% methanol (solvent A), andmethanol (solvent B) was used as a mobile phase with a flowrate of 1 ml/min. The following gradient profile was used:89% solvent A and 11% solvent B (1 min); linear gradient to5% solvent A and 95% solvent B (6 min); isocratic gradient(4 min); linear gradient to 89% solvent A and 11% solvent B(6 min); isocratic gradient (12 min). The sample chamber ofthe HPLC was maintained at 10�C to minimize sample lossesand/or conversions during analysis. General operating param-eters for the electrospray interface were desolvation flow of625 L/h and nebulizer flow of 69 L/h nitrogen gas. The cap-

illary voltage of 3.0 KV was used in both negative and positiveelectrospray modes and the heated zones were 450�C for thedesolvation zone and 140�C for the source.

The optimal parameters employed for the electrospray in-terface for each analyzed compound were determined by con-ducting infusion of standards for each of the compounds intothe MS prior to setting up the analytical methods (Table 2).The concentration of target compounds was determined on thebasis of five-point calibration curves obtained for standardsprepared in 50:50 methanol/water (v/v). For most analytes,standards ranged 0.005 to 1.0 ng/ml; the exceptions were me-tolachlor ethane sulfonic acid (MESA), which ranged from0.05 to 10.6 ng/ml, and metolachlor oxanillic acid (Met-OXA)and hydroxy-metolachlor (OH-Met), which both ranged from0.06 to 11.1 ng/ml. The lowest standard of each analyte rep-resented the LOQ.

Monocrotophos was used as an internal standard for quan-titation of compounds with early retention times, N-ethyl-6-



Table 2. Target analyte list for liquid chromatography-mass spectrometry analysis, chemical name and structures of analytes, instrumentalparameters, and limits of quantitation (LOQ)

Analyte nameor abbreviation Chemical namea Structure

Ionizationmodeb

Parention

(m/z)

DaughterIon

(m/z)

Conevoltage

(V)

Collisionenergy(eV)

Recovery(%) LOQc

Alachlor-ESA 2-([2,6-Diethylphen-yl][methoxy-meth-yl]amino)-2-Oxoetha-nesulfonic acid

ES� 314 121 45 23 97 11 12.5

CAAT 6-Chloro-1,3,5-triazine-2,4-diamine

ES� 146 104 32 18 15 8.6 1.25

CEAT 6-Chloro-N-ethyl-1,3,5-triazine-2,4-diamine

ES� 174 104 32 23 87 17 1.25

CIAT 6-Chloro-N-(1-methyle-thyl)-1,3,5-triazine-2,4-diamine

ES� 188 146 30 18 78 11 1.25

OIET N-ethyl-6-hydroxy-N�-(1-methylethyl)-1,3,5-triazine-2,4-diamine

ES� 198 156 40 20 89 12 1.25

2,4-D (2,4-Dichlorophenoxy)acetic acid

ES� 219 161 18 13 83 4.9 1.25

MESA 2-([2-Ethyl-6-methyl-phenyl][2-methoxy-1-methylethyl]amino)-2-oxoethanesulfonicacid

ES� 328 121 52 28 91 13 12.5

Met-Oxa 2-([2-Ethyl-6-methyl-phenyl][2-methoxy-1-methylethyl]amino)-2-oxoacetic acid

ES� 280 248 20 18 87 15 12.5

OH-MetSurrogates

2-Hydorxy-N-(2-ethyl-6-methylphenyl)-N-(2-methoxy-1-methyle-thyl)acetamide

ES� 266 234 20 18 82 16 1.25

TPP Internalstandards

Triphenylphosphate ES� 327 152 55 25 78 25d 1.25

[U-13C]-2,4-D [U-13C]-(2,4-Dichloro-phenoxy) acetic acid

ES� 225 167 18 13

Monocrotophos Dimethyl(1E)-1-methyl-3-(methylamino)-3-oxo-1-propenyl phos-phate

ES� 224 193 17 9

Terbuthylazine 6-Chloro-N-(1,1-dime-thylethyl)-N�-ethyl-1,3,5-triazine-2,4-di-amine

ES� 230 174 30 15

a Chemical Abstracts Service nomenclature.b ES� � negative electrospray mode, ES� � positive electrospray mode.c LOQ is based on instrument response at the lowest calibration standard level.d Includes all samples.



hydroxy-N�-(1-methylethyl)-1,3,5-triazine-2,4-diamine(OIET), 6-chloro-N-ethyl-1,3,5-triazine-2,4-diamine (CEAT),and CIAT, and was analyzed in positive ionization mode. Com-pounds with later retention times and detected in positive ion-ization mode, Met-OXA and OH-Met, were quantitated usingterbuthylazine as an internal standard. The degradation productMESA was analyzed in negative ionization mode and wasquantitated using labeled [U-13C]-(2,4-dichlorophenoxy) aceticacid as internal standard. This labeled pesticide has been usedbefore as an internal standard for quantitation of MESA whenanalyzed by HPLC/MS-MS [9]. Polar compounds such as chlo-rophenoxy acid herbicides or sulfonic acid degradates of chlo-roacetanilide herbicides respond well, and in similar fashion,when analyzed in negative electrospray mode. Additionally,the presence of 13C label in 2,4-D internal standard providedassurance that the possible presence of this herbicide in thesamples would not interfere with MESA quantitation. Stan-dards were rerun during the course of each sample run afterat least every 10 samples; quantitation was based on linearcurves with r2 values 0.99.

All residue concentrations in the field and laboratory blankswere below the LOQ. Recoveries of the extraction surrogatetriphenylphosphate were acceptable, averaging 77 25%. Itshould be noted that triphenylphosphate is not necessarily themost ideal surrogate for the ionic analytes, particularly MESA;however, the adequate recovery of this compound from matrixspikes suggests that the method is adequate for determinationof our analytes. In general, spike recoveries were greater than78%, with the exception of the atrazine degradation product6-chloro-1,3,5-triazine-2,4-diamine (CAAT) at 15% (Table 2).Results were not adjusted for recovery values; therefore, con-centrations of CAAT reported here are likely lower than actualenvironmental concentrations.

Study design considerations

All results presented here represent dissolved-phase con-centrations because filters were not extracted or analyzed. Pre-vious studies of filter extracts from Susquehanna River waterat a 10-L volume sample size yielded negligible particulate-phase concentrations [6]. Parent material, CIAT, and CEATwere analyzed in all samples. More-polar parent material anddegradation products (2,4-D, MESA, Met-OXA, OH-Met,OIET, CAAT) were analyzed by HPLC/MS-MS and were add-ed to the list of analytes for 41 of the 62 samples from the2004 toxics characterization study (Supplemental Data, TableS4; http://dx.doi.org/10.1897/06-655.S4). None of the samplesfrom the Rhode/South River region were analyzed by HPLC/MS-MS. The 2000 main stem study samples were analyzedby gas chromatography–mass spectrometry only.

Agricultural pesticide application in the mid-Atlantic regionbegins in the spring and early summer months and concludesin late summer. Samples from the 2004 toxics characterizationstudy were collected during mid- to late summer and early falland do not necessarily represent the maximum annual con-centrations. Samples in the 2000 main stem study were col-lected just at the onset of agricultural activity in the region(April 30–May 2). Furthermore, results presented here arefrom one point in time at each station. With continuous changesin pollutant loads and tidal conditions, rigorous statistical anal-ysis to compare results from the various tidal regions wouldnot be valid. However, because so few studies of pesticideswithin the Chesapeake Bay watershed have been conducted,

the data presented here contribute to the overall body of knowl-edge with respect to toxics in the Chesapeake Bay.

RESULTS

Overall pesticide concentration results

In the 2000 main stem study, atrazine and metolachlor weredetected at all 18 stations (Fig. 1, Supplemental Data TableS3; http://dx.doi.org/10.1897/06-655.S3). The analyte CIATwas detected at every station except CB6, and CEAT was notdetected at all. Diazinon was detected at three stations, CB3,CB4, and CB5, in the upper main stem at concentrations rang-ing from 2.3 to 18 ng/L (levels below the LOQ). Acetochloralso was detected in the upper main stem station C2 throughC7 ranging from 9.7 to 21 ng/L. None of the organochlorinepesticides were detected in these samples.

In 2004, of the 32 analytes included in the present study,only 13 analytes were present in at least 20% of stations fromany one tidal region. All 13 of these more-frequently detectedanalytes were either herbicides or herbicide degradation prod-ucts. (Complete information on concentrations of these 13 an-alytes is provided in Supplemental Data Table S4; http://dx.doi.org/10.1897/06-655.S4.) Atrazine was found in all sam-ples from each of the five tidal regions. The chemical CIATwas found almost as frequently as atrazine; however, it wasnot detected at one station in the Nanticoke River and it wasonly detected in 40% of samples from Lower Mobjack Bay.The chemical OIET was found in all 41 samples where it wasanalyzed, and CEAT and CAAT were detected less frequently.Simazine was detected in 100% of samples from the ChesterRiver and frequently was detected (80%) in samples fromthe Nanticoke and Pocomoke Rivers. Simazine also was foundin 50% of samples from the Rhode/South River area, but itwas not detected in Mobjack Bay. The highest concentrationsof atrazine and simazine were observed in the Chester Riverat 820 and 1,500 ng/L, respectively.

Metolachlor was detected in 100% of samples from theChester River, but it was observed less often in the Nanticoke(64%) and Pocomoke (45%) rivers. Metolachlor degradationproducts, OH-Met, Met-OXA, and MESA, were found in100% of all samples where they were analyzed. Metolachlorwas not detected in any samples from the Lower Mobjack Bayor the Rhode/South River area. Alachlor was not detected inany samples, but alachlor ethane sulfonic acid (Alachlor-ESA)was detected in all samples where analyzed. Acetochlor wasonly detected at stations P18 and P29 in the Nanticoke Riverat 32 and 3.4 ng/L, respectively. The herbicide 2,4-D wasdetected in 100% of samples from the Chester River and lessfrequently detected in the Nanticoke and Pocomoke Riversand in Mobjack Bay at a median concentration of approxi-mately 20 ng/L in all four areas. (Samples from Rhode/SouthRiver area were not analyzed for 2,4-D.) Pendamethalin wasdetected infrequently in only the three agricultural watershedsat a maximum of 31 ng/L in the Chester River.

The highest maximum concentration for any target analytewas found for MESA at 2,900 ng/L in the Nanticoke River.The metabolite MESA also had the greatest concentration ofany analyte in the Pocomoke River (2,100 ng/L) and in theChester River (1,200 ng/L). The presence of the degradationproducts in the same or greater concentrations than the parentherbicide is to be expected during the late summer and earlyfall time period of the study.

Only one nonherbicide chemical was detected above theLOQ in the 2004 study. Diazinon, an organophosphate insec-



ticide, was detected at station P23 at 11 ng/L (level belowLOQ). Organophosphate insecticides will undergo degradationvia hydrolysis in surface waters; therefore, a more effectivemanner to evaluate sources of these chemicals would be tomeasure the hydrolysis products. None of the other analytes,such as organochlorine insecticides, listed in Table 1 weredetected in any samples. In previous studies on the ChesapeakeBay tributaries, researchers used much larger sample size vol-umes (10–20 L) to examine these persistent organic pollutants[6,10]. Significantly larger sample volumes or higher loadingsto the analytical equipment would be required for a thoroughevaluation of organochlorine and organophosphate insecticideconcentrations in these tributary regions. Therefore the fol-lowing discussion is limited to the fate of herbicides and theirdegradation products.

DISCUSSION

A number of organochlorine pesticides are included on theToxics of Concern lists established by the Chesapeake BayProgram Office and by other jurisdictions in the watershed[11] (http://www.chesapeakebay.net/pubs/subcommittee/tsc/toxics/pdf%20finals/toxics�2000�appendix�a.PDF). Many ofthese organochlorine pesticides were included in the analysisof samples from these two studies; but the chemicals mostfrequently observed in these studies were agricultural herbi-cides and their degradation products.

Although herbicides are generally much less toxic to fish,birds, and humans when compared to many banned organo-chlorine chemicals, they are used in significant quantities eachyear within agricultural and urban watersheds. An assessmentof ecological risk from atrazine and metolachlor in ChesapeakeBay streams and tributaries found that these chemicals did notpose a significant risk [12]. However, other studies have foundthat agricultural herbicides and insecticides alter the functionof estuarine microbial food webs [13,14]. High concentrationsof atrazine were found to inhibit the growth of an importantmarsh plant, J. roemeriannus [15]. Atrazine also has beenshown to cause reproductive toxicity and sexual differentiationdisruption in amphibians [16]. These conflicting evaluationsand observations suggest that further investigations are re-quired.

Previous observations of herbicides in the Bay and itstributaries

Agricultural field crop production in the region surroundingChesapeake Bay is predominately corn, soybean, and smallgrains. It is not surprising, therefore, that herbicides are theprimary agricultural pesticides that have been detected in pre-vious studies. Glyphosate was the most-used pesticide inMaryland in 2004 followed by chlorothalonil, atrazine, fosetylaluminum (a fungicide), S-metolachlor, mancozeb, metolach-lor, chlorpyrifos, potassium salts of phosphoric acid (also usedas a fungicide), and 2,4-D [17] (http://www.mda.state.md.us/pdf/2004�pesticide�use�survey.pdf). Atrazine was detected insurface water and precipitation throughout the year by re-searchers working in the Rhode River in the late 1970s [18].Glotfelty et al. [19] carried out a detailed examination of at-razine fate in the Wye River estuary, located just north ofChester River, and characterized conditions that create en-hanced pesticide runoff to surface waters.

In the early 1990s, Foster and Lippa [20] measured seasonalpesticide loadings to Chesapeake Bay using sites at the fallline of three major Chesapeake Bay tributaries: The James,

Potomac, and Susquehanna Rivers. They found that triazineand acetanilide herbicides were detected most frequently withpeak concentrations occurring in April and May and that thelargest pesticide loads were entering the Chesapeake Bay fromthe Susquehanna River. A U.S. Geological Survey (USGS)report on water quality in the Lower Susquehanna River basinfrom 1992 to 1995 also found that the most commonly detectedpesticides in streams and groundwater were the herbicides usedprimarily on corn: Atrazine, metolachlor, simazine, prometon,alachlor, and cyanazine [21].

The Susquehanna enters the Bay from the northern end andprovides 47% of the total freshwater to the estuary; agriculturalland use in its watershed is approximately 30%. Agriculturalpesticide loads entering the Bay from the Susquehanna werecharacterized in detail by Foster et al. [10] in 1994 and laterby Liu et al. [6] in 1997 and 1998. Foster found herbicideloads were 10 to 100 times greater than for organochlorineinsecticides, polychlorinated biphenyls, and most polyaromatichydrocarbons. Liu found the highest flux rates of atrazine andmetolachlor in early June and observed pulses of diazinon andendosulfan entering the Bay in the fall and early winter months.In addition, increasing flux rates of CIAT and CEAT wereobserved after the spring application period increasing to levelsequal to or greater than the atrazine flux in the late summerand early fall. Although both studies concluded that freshwaterinputs from the Susquehanna are the largest single source ofagricultural pesticides and their degradation products to theBay main stem, these compounds also have been observed inother tributaries such as the Choptank [22,23] and PatuxentRivers [24,25] and contribute to the overall pesticide load inthe Bay.

Pesticide degradation products

Once applied to agricultural fields, pesticides can be trans-ported to other environmental compartments and can undergotransformation reactions. Once the herbicides are transportedvia leaching into the soil or runoff to nearby surface waters,they can undergo microbial and/or abiotic transformation.These rates will vary widely depending on soil and waterphysical and chemical properties and also will be dependenton the indigenous microbial community that will be affectedby previous exposure to the chemical [26,27]. The parent her-bicide and its degradation products may leach through the soilcolumn to shallow groundwater and be transported further tostreams [28]. Thus, the mixture of degradation products ob-served in surface waters may be the result of several trans-formation and transport pathways.

Of the three primary degradation products for atrazine,CIAT and CEAT are formed via dealkylation and OIET isformed via hydrolysis. The product OIET is very polar andsorbs strongly to soils giving rise to its limited mobility insoils [29]. In other watersheds, OIET has been observed insurface waters in the dissolved phase or sorbed to soil particlesas the result of runoff from agricultural soils [30]. AlthoughOIET is formed from atrazine only, CEAT is formed via de-alkylation of simazine. The degradation product CIAT also canbe a product of cyanazine degradation, but this herbicide rarelyis used in the region.

Both alachlor and metolachlor are transformed via dechlo-rination to their ethane sulfonic acid degradation products(Alachlor-ESA and MESA) in the soil environment. Both ofthese chemicals are highly susceptible to leaching [31]. TheMet-OXA metabolite of metolachlor also has been observed

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Fig. 2. Comparison of median acetanilide (a) and triazine (b) herbicideand degradation product concentrations (ng/L) observed during theperiod July 14, 2004 to September 13, 2004 in the Chester, Nanticoke,Pocomoke Rivers, the Rhode and South Rivers, and the Lower Mob-jack Bay, USA. Acetanilide results are not available for the Rhodeand South River stations. Chemical names are abbreviated as follows:Alachlor ethane sulfonic acid (Alachlor-ESA), metolachlor ethane sul-fonic acid (MESA), metolachlor oxanillic acid (Met-OXA), hydroxy-metolachlor (OH-Met), 6-chloro-1,3,5-triazine-2,4-diamine (CAAT),6-chloro-N-ethyl-1,3,5-triazine-2,4-diamine (CEAT), 6-chloro-N-(1-methylethyl)-1,3,5-triazine-2,4-diamine (CIAT), and N-ethyl-6-hy-droxy-N�-(1-methylethyl)-1,3,5-triazine-2,4-diamine (OIET).

in tile drain effluents in corn producing regions [32]. A numberof additional metolachlor degradation products, including OH-Met, also have been identified in Chesapeake Bay surface wa-ters [9].

A comparison of median concentration values for the tri-azine and acetanilide groups and 2,4-D (Fig. 2) illustrates theimportant contribution of the degradation products to the over-all pesticide load during this July to September time period.As stated above, concentrations of MESA were higher thanany other target analyte included in the 2004 study and con-tributed at least 50% to the total herbicide degradation productconcentration at all stations. The sum of all three metolachlordegradation products (MESA, Met-OXA, and OH-Met) rep-resents approximately 75% of the total degradation productconcentration at each station where they were analyzed (Sup-plemental Data, Table S4; http://dx.doi.org/10.1897/06-655.S4). The parent herbicide, alachlor, which is no longerwidely used in this region, was not detected in any of the tidalregions, but Alachlor-ESA was found in significant concen-trations at each station where an analysis was conducted. Thesefindings support earlier studies where it was shown that Alach-lor-ESA is formed two to four times faster than MESA in soils[30].

The overall findings with respect to degradation productsin the present study are similar to those observed in a USGS

project carried out in a subwatershed of the Chester and Po-comoke Rivers in the late 1990s through 2001 [32]. In theUpper Pocomoke River and the Chesterville Branch of theChester River, corn herbicides and their degradation productswere observed year-round with concentrations of the degra-dation products often exceeding the parent compounds, es-pecially MESA. Median concentration values found in thesesubwatersheds were generally within a factor of six of thoseobserved in the 2004 toxics characterization in the Pocomokeand Chester Rivers. For example, a median atrazine concen-tration of 48 ng/L was found in the Pocomoke River in thepresent study, and a median value of 20 ng/L was found inthe Upper Pocomoke USGS study. However, the medianAlachlor-ESA concentration in the Chesterville Branch was25 times greater than in the Chester River (at 2,190 and 87ng/L, respectively). Overall, metolachlor and degradationproduct concentrations were higher in the USGS study andatrazine and CIAT concentrations were higher in the 2004study (data not shown). Other atrazine degradation productswere not included in the USGS study.

A multiyear study of metolachlor and atrazine persistencein loamy soil found that metolachlor dissipated more quicklythan atrazine (metolachlor dissipation half-life � 17–28 d;atrazine dissipation half-life � 23–46 d) [33]. The greaterpersistence of the parent atrazine may account for the smallercontribution of atrazine degradation products to the total deg-radation product concentration and also may account for thepresence of the parent herbicide at stations in all five tidalregions. Of the four atrazine degradation products included inthe 2004 toxics characterization study, OIET was present atlevels approximately 50% greater than CIAT or CEAT. A sec-ondary degradation product of atrazine, CAAT, was found onlyat a few stations in the agricultural tributaries, but it was de-tected at each station of the Lower Mobjack Bay region (Fig.2). The presence of CAAT in Lower Mobjack Bay suggeststhat the atrazine signal may be weathered during transport tothis nonagricultural region.

Spatial trends in pesticide concentrations

In the 2000 main stem study, the greatest concentrationsof both atrazine and metolachlor were found at the stationsCB3, CB4, and CB5 near the mouth of the Chester River(atrazine � 38–47 ng/L and metolachlor � 12–17 ng/L). Theoverall range of herbicide concentrations found in 2000 issimilar to those observed by Liu et al. [6] in the upper andmid-Bay region of the main stem in February 1997. Atrazineand CIAT concentrations did not exhibit a strong linear rela-tionship with salinity (Fig. 3), although a very rough decreas-ing trend in concentration with increasing salinity was ob-served. This suggests that there are multiple sources of oneor both chemicals within the Bay. Degradation of atrazine toCIAT and degradation of CIAT to CAAT also may be occurringduring transport. The trend in metolachlor concentration withsalinity is stronger where the linear regression line exhibits anr2 of 0.76. However, if a theoretical mixing line were drawnfor metolachlor connecting just the two salinity endpoints,most of the other points would fall below this line (mixingline not shown in Fig. 3). This is an indication that metolachloris not conserved with dilution within the main stem. The clear-er decreasing trend in metolachlor concentrations with salinityalso indicates a primary source in the upper Bay such as theSusquehanna River.

A similar analysis of concentration and salinity can be car-



Fig. 3. Atrazine, 6-chloro-N-(1-methylethyl)-1,3,5-triazine-2,4-di-amine (CIAT), and metolachlor concentrations (ng/L) observed onApril 29 to May 2, 2000 in surface water of the Chesapeake Bay,USA main stem plotted versus salinity (ppt). Linear regression equa-tions and correlation constants (r2) for each chemical are provided.Each data point represents the average of results from duplicate sam-ples.

Fig. 4. Concentrations (ng/L) of acetanilide (a) and triazine (b) her-bicide and degradation product concentrations observed in the ChesterRiver, USA, on July 14 and 15, 2004 (sites P05–P15) plotted versussalinity (ppt). Linear regression equations and correlation constant (r2)values for each chemical are provided. Chemical names are abbre-viated as follows: Metolachlor ethane sulfonic acid (MESA), meto-lachlor oxanillic acid (Met-OXA), hydroxy-metolachlor (OH-Met), 6-chloro-N-(1-methylethyl)-1,3,5-triazine-2,4-diamine (CIAT), N-ethyl-6-hydroxy-N�-(1-methylethyl)-1,3,5-triazine-2,4-diamine (OIET), and6-chloro-N-ethyl-1,3,5-triazine-2,4-diamine (CEAT).

ried out on results from the Chester River. The maximumherbicide concentrations in the Chester River were observedat station P05 (Supplemental Data Fig. S3, http://dx.doi.org/10.1897/06-655.S5; Table S4, http://dx.doi.org/10.1897/06-655.S4). Therefore, results from station P05 to P15 at themouth of the river were plotted versus salinity to examine thebehavior of the herbicides and their degradation products.Overall, the atrazine and metolachlor degradation products ex-hibited a stronger relationship with salinity than the parentherbicides (Fig. 4). Scatter in the data at the uppermost sitesfor the parent herbicide contributed to poor linear regressioncorrelation coefficients (r2 � 0.78 for metolachlor and 0.68for atrazine). Variability in the parent herbicide concentrationsuggests that these chemicals are less stable than their deg-radation products or that there are multiple sources of thesecompounds within the river.

The degradation products CIAT, OIET, CEAT, and Met-OXA were strongly correlated with salinity (r2 � 0.76, 0.95,0.96, and 0.96, respectively). The chemicals MESA and OH-Met exhibited the most conservative behavior with respect todilution (r2 values of 0.97 and 0.98, respectively), suggestingthat these chemicals are stable during transport in the estuary.However, the concentration of these degradation products maybe affected by degradation of the parent during transport. If atheoretical mixing line connecting only the two salinity end-points were drawn for MESA (line not shown in Fig. 4), someof the points would fall slightly above the line. This suggeststhat the degradation products are entering the Chester Riveralong with the parent herbicides in significant quantities, butthat some degradation is occurring during transport. Furtherstudy of metolachlor degradation in brackish water with a morefrequent sampling regime combined with controlled laboratorykinetics experiments could be used to verify this hypothesis.

Conditions in the Chester are similar to those seen in earlierstudies of the Patuxent River where upstream sources wereidentified as the largest inputs of herbicides to the estuary[25,26]. The major source of herbicides in the Chester appearsto be in the shallow upstream regions of its watershed as well.Earlier research in the late 1990s in the Chesterville branchindicates that groundwater is the primary source of degradation

products and that a combination of surface runoff and ground-water contributes to loadings of parent herbicides [32]. Resultsfrom the Pocomoke mimic the Chester in the overall trend ofdecreasing concentration at the downstream stations, but themaximum herbicide concentrations were a factor of 10 lowerthan the Chester (Supplemental Data Table S4; http://dx.doi.org/10.1897/06-655.S4). Although metolachlor was amajor component in the upstream station samples, it was notobserved downstream. The Pocomoke River is very narrowand shallow upstream, expanding rapidly into a wide estuary.The upper Pocomoke has very poorly drained soils when com-pared to the Chester River watershed [32]. Those stations be-low P36 actually are located in what is called Pocomoke Sound(Supplemental Data Fig. S5; http://dx.doi.org/10.1897/06-655.S5), an area that is essentially part of the Bay mainstem.

At first glance, results from the Nanticoke conflict with thespatial trends observed in the other two Eastern Shore tribu-taries. The overall range of herbicide concentrations in theNanticoke was similar to those in the Pocomoke. Atrazine andsimazine concentrations at downstream Nanticoke stations P22to P25 (Fig. S4) were approximately a factor of two higher



than the upstream freshwater stations (Table S4). This apparentcontradiction most likely is due to a fragmentation in the sam-pling schedule in the Nanticoke. Stations P22 to P25 weresampled on July 18th, although the stations above and belowthis section were sampled a few weeks later on August 5thand 6th. From this observation, we can deduce that significantdegradation of the parent herbicides occurs throughout thegrowing season.

A decidedly different picture emerges from results of theWestern Shore tributaries. The range of salinity values in theLower Mobjack Bay region was narrow, centering around 16.5parts per thousand (ppt). Atrazine was the only parent herbicidedetected here and concentrations ranged from 17 to 32 ng/L,which is similar to the levels observed in Pocomoke Sound.The major source of atrazine to this tidal region actually maybe the waters of the Chesapeake Bay main stem. A very narrowsalinity range of 7 to 8 ppt also was observed in the Rhodeand South Rivers, and there was no spatial trend in herbicideconcentration downstream in either river. Although atrazinealso was the only herbicide found in this region, concentrationswere higher than in Lower Mobjack Bay, ranging from 83 to112 ng/L. The consistent atrazine levels throughout the regionalso suggest a main stem source. The concentration rangefound in the Rhode and South Rivers was similar to that foundat the mouth of the Chester River. Therefore, atrazine con-centrations entering the mid-Bay region of the Rhode andSouth Rivers are higher than those found in the Lower MobjackBay where cleaner Atlantic Ocean water intrudes and decreasesatrazine concentrations.

CONCLUSION

Results from the present research have demonstrated thatherbicides and their degradation products are more abundantthan any other modern pesticides or legacy organochlorineinsecticides included in the present study. These herbicidesand associated products are present within both agriculturaland nonagricultural tidal areas of the Chesapeake Bay through-out the year even though their use is primarily in the spring.Agriculture remains an influence on the watershed, althoughthis may be changing as societal land use evolves from ag-ronomic to urban systems. This work also suggests that thesesmaller tributaries are important contributors to the overallpesticide load to the main stem. Although the effects of thecontinuous presence of these chemicals on the most sensitivespecies is not known, this report adds to the overall body ofknowledge with respect to the fate and transport of agriculturalchemicals within Chesapeake Bay and will provide data re-quired for further modeling and risk-assessment efforts, es-pecially within the Chesapeake Bay toxics characterizationprogram.

SUPPORTING INFORMATION

Table S1. Station locations, depth, and water-quality char-acteristics for 2004 toxics characterization study.

Found at DOI: 10.1897/06-655.S1 (45 KB PDF).Table S2. Station locations, depth, and water-quality char-

acteristics for 2000 main stem study.Found at DOI: 10.1897/06-655.S2 (28 KB PDF).Table S3. Surface Water Concentration Results (ng/L) from

2000 main stem study for Herbicides and Degradation Prod-ucts.

Found at DOI: 10.1897/06-655.S3 (29 KB PDF).Table S4. Surface Water Concentration Results from 2004

toxics characterization for Herbicides and Degradation Prod-ucts.

Found at DOI: 10.1897/06-655.S4 (43 KB PDF).Figure S1. Detailed map with sample collection locations

for the 2004 toxics characterization study: Lower MobjackBay.


for the 2004 toxics characterization study: Rhode and SouthRivers.


for the 2004 toxics characterization study: Chester River.Found at DOI: 10.1897/06-655.S5 (371 KB PDF).Figure S4. Detailed map with sample collection locations

for the 2004 toxics characterization study: Nanticoke River.Found at DOI: 10.1897/06-655.S5 (371 KB PDF).Figure S5. Detailed map with sample collection locations

for the 2004 toxics characterization study: Pocomoke River.Found at DOI: 10.1897/06-655.S5 (371 KB PDF).

Acknowledgement—Mention of specific products or supplies is foridentification and does not imply endorsement by U.S. Departmentof Agriculture to the exclusion of other suitable products or suppliers.

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agricultural pesticides and selected degradation products in five

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