Grantee Research Project Results
Final Report: Role of Reduced Sulfur Species in Promoting the Transformation of Triazines in Estuaries and Salt Marshes
EPA Grant Number: R826269Title: Role of Reduced Sulfur Species in Promoting the Transformation of Triazines in Estuaries and Salt Marshes
Investigators: Roberts, A. Lynn , Salmun, H.
Institution: The Johns Hopkins University
EPA Project Officer: Chung, Serena
Project Period: February 1, 1998 through January 31, 2001 (Extended to January 31, 2002)
Project Amount: $304,163
RFA: Exploratory Research - Environmental Chemistry (1997) RFA Text | Recipients Lists
Research Category: Water , Land and Waste Management , Air , Safer Chemicals
Objective:
The objectives of this research were to: (1) determine the rates of abiotic reactions of triazines (and related species) with inorganic reduced sulfur nucleophiles; (2) examine the ability of azines to bind covalently to natural organic matter (NOM); and (3) examine the potential impact of reduced sulfur species on the fate of azines in anoxic coastal marine environments. We hypothesized that reactions in sulfidic hydrologic environments between well-defined reduced sulfur species (particularly polysulfide ions and aromatic thiolate moieties present in NOM) and chloro-s-triazine herbicides, fungicides, and reactive dye compounds, could provide a significant sink for such organic contaminants.
Summary/Accomplishments (Outputs/Outcomes):
Our accomplishments and results can be divided into seven major topics: (1) method development for determination of polysulfides; (2) determination of hydrogen sulfide and polysulfide abundance in anoxic marine waters; (3) determination of second-order rate constants for reactions of haloazines with reduced sulfur nucleophiles in laboratory settings; (4) comparison of predicted rates of reaction to measured rates in laboratory-spiked samples of natural sulfidic water; (5) attempts to measure concentrations of azine transformation products in field samples; (6) mass balance calculations for atrazine within the Chesapeake Bay; and (7) incorporation of information pertaining to atrazine input and potential reactions with reduced sulfur nucleophiles into a numerical model of atrazine fate and transport within the Chesapeake Bay.
Method Development for Determination of Polysulfides. The first portion of our efforts were related to the development of methods suitable for the laboratory and field analysis of polysulfides. Not only were such analyses important for verifying concentrations of polysulfides in laboratory solutions used to determine rate constants for reactions with haloazines, but also an understanding of concentrations of polysulfides encountered in natural settings, as well as factors controlling their abundance, which are key to understanding their potential impact in anoxic marine waters. To analyze polysulfides, we refined a method (Borchardt and Easty, 1984) based on the conversion of polysulfides to sulfur (0) and subsequent derivatization with triphenylphosphine to produce triphenylphosphine sulfide. This derivative readily can be measured via gas chromatography/mass spectrometry (GC/MS), GC/flame ionization detector (FID), or high performance liquid chromatography (HPLC). Agreement between polysulfide concentrations measured in laboratory solutions (prepared by equilibrating HS- and S8 [s]) with values predicted on the basis of published stability constants (Giggenbach, 1974) was excellent, and revealed no systematic bias as a function of pH. A tendency for measured values to exceed predicted values was noted at very low polysulfide concentrations (< 10-5 M S(0)poly); this was attributed to incomplete retention of colloidal sulfur by the filter employed (0.02 µm). Such filter breakthrough imposes a practical detection limit for determination of polysulfides on the order of 3 µM. This is comparable to levels studied by researchers using electrochemical techniques (Rozan, et al., 2000; Kariuki, et al., 2001). An advantage of the triphenylphosphine method is that it allows for derivatization (to relatively stable products) in the field, minimizing the likelihood of experimental artifacts; moreover, it avoids the problem of fouling electrodes by NOM frequently encountered in electrochemical measurements in natural waters.
Determination of HS- and Polysulfide Abundance in Anoxic Marine Waters. Method development efforts were followed by field analyses of an anoxic, sulfidic marine water. The field site selected consisted of the anoxic hypolimnion of Lower Mystic Lake (Arlington, MA). This meromictic lake previously received tidal inputs of sea water that were halted when dams were constructed in 1909 and 1966 (Duval and Ludlam, 2001). The denser saline water became anoxic below the chemocline, and concentrations of total hydrogen sulfide were found to increase with depth, attaining values as high as 16 mM. The polysulfide profiles we measured (with concentrations up to 60 µM) for this natural water closely paralleled profiles for total hydrogen sulfide. Furthermore, measured values, generally were in excellent agreement with values anticipated if HS- and Sn2- in this water were in equilibrium with S8 (s). In addition, analyses of field samples revealed the presence of particulate S8 in the water, with concentrations attaining a maximum about 2 m below the chemocline. The existence of excess S8 (s), and the close correspondence measured with predicted polysulfide concentrations, suggests that equilibrium considerations govern the distribution of polysulfides within the hypolimnion of this lake. Measured concentrations of hydrogen sulfide and polysulfides are of the same order of magnitude as values previously reported for salt marsh porewaters, suggesting that water from the hypolimnion of this lake represents a readily accessible model of anoxic marine porewaters.
Determination of Second-Order Rate Constants for Reactions of Haloazines With Inorganic and Organic Sulfur Nucleophiles. After analytical methods for analyses of polysulfides had been established, our efforts turned to experimental determination of rates of reaction of haloazines with reduced sulfur nucleophiles. Rates of reaction of three commonly used chloro-s-triazine agrochemicals (atrazine, anilazine, and cyanazine) with polysulfide dianions and HS- species were determined in aqueous solution at 25°C. Experiments with atrazine were conducted by varying the identity and concentration of the nucleophile, as well as solution pH, to verify that reactions were first-order in total polysulfide dianion (Sn2-) concentration.
To provide information pertinent to the fate of reactive dye compounds, reactions of HS- and Sn2- with seven halogenated pyridine, pyrimidine, and quinoxaline compounds also were investigated. A fluorinated analog of atrazine was synthesized, and its reactions were studied. For the reactions of chloroazines with HS- and Sn2-, a general trend was observed in which an increasing number of chlorine substituents resulted in increased reactivity. Moreover, ring aza nitrogens, as well as the presence of a fused ring system (bicyclic heterocycles), were found to confer increased reactivity.
Large differences in the reactivity of haloazines toward polysulfide dianions versus HS- were observed, with ratios of kSn2-/kHS- ranging from 75 to more than 3,000. For example, atrazine is recalcitrant to reaction with HS-, yet reacts readily with Sn2-. Cyanazine reacts with HS- at a slow but measurable rate over a 24-day period. Rate constants for reaction with polysulfides vary by seven orders of magnitude, and kHS- by at least five orders of magnitude. Products were identified by derivatization with methyl iodide or pentafluorobenzyl bromide, followed by GC/MS/electron impact ionization (EI) analysis (and, in some cases, by GC/MS analysis with positive and negative chemical ionization). Results indicate substitution of halogen by sulfur. The products, the first-order dependence on the concentration of the nucleophile, and the qualitative structure-reactivity trends observed, are all consistent with a nucleophilic aromatic substitution (SNAr) reaction mechanism.
A quantitative structure-activity relationship (QSAR) comparing chloroazine reactivity to calculated lowest unoccupied molecular orbital (LUMO) energies was developed in the hope that it might provide a useful tool for predicting the environmental fate of untested azines. This QSAR indicated a weak correlation between LUMO energy and reactivity, no doubt reflecting its neglect of steric effects, as well as the subtle effects of transition state structure on the rate of reaction. Better correlations were observed within individual classes of chloroazines than for chloroazines as a whole.
Somewhat mixed results were obtained when we attempted to investigate the role that organic sulfur nucleophiles might exert on the fate of halogenated azines. Although atrazine was found to react readily with the model compound thiophenolate, attempts to generate aromatic thiolates through a 1,4-nucleophilic (Michael) addition of HS- to quinones (Weber, et al., 1996) present in NOM (obtained from the Great Dismal Swamp, VA, with a total organic carbon content of 40 mg/L) yielded solutions in which no reaction of added atrazine could be discerned after 2 weeks. It is possible that the quinone content of the Great Dismal Swamp NOM was too low to yield appreciable concentrations of aromatic thiolates through such an addition process.
Laboratory-derived rate constants for reactions of haloazines with inorganic sulfur nucleophiles were extrapolated to an environmentally relevant concentration of hydrogen sulfide and polysulfides (reported for salt marsh sediment porewaters at pH 7.1 [Boulègue, et al., 1982]). The resulting half-lives range from less than 1 minute (for anilazine) to approximately 4 days (for atrazine). Even shorter half-lives would be anticipated for the highest hydrogen sulfide and polysulfide concentrations measured in Lower Mystic Lake. Polysulfides, by virtue of their very high reactivity, are predicted to play a much more important role than HS- in abiotic reactions of haloazines. The half-life for atrazine is much shorter than the reported uncatalyzed hydrolysis half-life of 1,800 years at a pH of 6.97 (Plust, et al., 1981). For purposes of comparison, exchange coefficients for solutes between the water column and the sediment porewaters underlying the Chesapeake Bay are on the order of 0.1-1 m/d; characteristic times (for a 5-m water column) are thus on the order of 5-50 days. The mean residence time of water in the Chesapeake Bay is approximately 8 months (Nixon, et al., 1996). These results suggest that abiotic reactions with reduced sulfur nucleophiles-particularly polysulfides-present in sediment porewaters potentially could exert a substantial effect on the fate of azine agrochemicals or reactive textile dyes in this estuary.
Comparison of Predicted Rates to Measured Rates of Reaction in Laboratory-Spiked Samples of Natural Sulfidic Water. To test the ability of laboratory-derived rate constants to predict haloazine fate in the much more complicated milieu of natural waters, several samples from a Lower Mystic Lake were fortified with sulfur, and the pH was adjusted to produce varying concentrations of polysulfides. These modified waters then were allowed to react with atrazine. Some experiments also were conducted in unaltered lake water samples. A sharp increase in reactivity was observed on increasing polysulfide concentration. The observed atrazine transformation products in each case were consistent with chlorine displacement. The observed transformation rates very favorably compared to those predicted by the laboratory-derived second-order rate constants and the measured concentrations of [Sn2-] in each solution. Similar results were obtained for cyanazine, except that a mixture of transformation products (indicative of hydrolysis of the nitrile moiety in addition to displacement of chlorine) was observed.
Preliminary Field Analyses of Atrazine and Aatrazine Degradation Products in the Chesapeake Bay. "Back of the envelope" calculations suggested that reactions with reduced sulfur nucleophiles within sediment porewaters might represent an important loss mechanism for atrazine within the Chesapeake Bay. Attempts to determine a mass balance for atrazine in the Chesapeake Bay also identified the lack of information concerning the vertical distribution of atrazine within the water column of the Chesapeake Bay as a significant data gap. Therefore, preliminary attempts to determine vertical profiles of atrazine, as well as concentrations of its environmental transformation products, were carried out for a single site within the Upper Chesapeake Bay. The site selected (near the Bay Bridge between Kent Island and Annapolis) was known to undergo seasonal hypoxia each summer. In addition to the vertical profile, samples also were obtained from the Susquehanna River immediately upgradient of the Conowingo Dam 10 days prior to the Chesapeake Bay sampling event. This timing was selected to sample the water parcel that represented the input to our sampling point, based on the assumption of a 10-day mean hydraulic residence time at the point sampled for our profile. Metolachlor (and several of its transformation products) also were included as analytes, as this herbicide is known to react more readily with polysulfides than atrazine (Loch, et al., 2002). It also is an important herbicide introduced to the Bay. The environmental transformation products expected of these two herbicides were: desethyl atrazine and desisopropyl atrazine (for which reference materials were commercially available); the sulfur-substituted products mercaptoatrazine and the mercaptometolachlor (analyzed as the pentafluorobenzyl derivatives, which were synthesized in our laboratory); and two additional metolachlor degradates: deschlorometolachlor (N-[2-ethyl, 6-methylphenyl]-N-[2-methoxy-1-methylethyl] acetamide) and deschloroacetylmetolachlor (N-[2-ethyl-6-methylphenyl]-N-[2-methoxy-1-methylethyl] aniline), both of which were synthesized in our laboratory.
Analyses were conducted using solid phase extraction (300 mL samples of water extracted on Oasis HLB cartridges, subsequently eluted with 1 mL of solvent), derivatization (where appropriate), and GC/MS analysis using selective ion monitoring (with 100 µL injections using a programmed temperature vaporizing injector). For all samples from the Chesapeake Bay and the Susquehanna River, 13C3-atrazine was added as a surrogate. The mean recovery of the surrogate was 105 percent. Recoveries for other analytes were determined by extracting 100 mL of deionized water spiked with target analytes. Further recovery studies were conducted by spiking herbicides into a natural water containing a high NOM content. Because of the preliminary nature of this study, only single samples were taken at each point along the depth profile, at a single "snapshot" obtained on July 13, 2001.
The vertical profile of the Chesapeake Bay reveals atrazine concentrations ranging from 280-340 ng/L, and metolachlor concentrations between 36-63 ng/L. No pronounced spatial trend in concentration was observed for either of these herbicides. Concentrations of desethyl atrazine and desisopropyl atrazine ranged from 20-50 ng/L, and did not vary substantially with depth. The two metolachlor degradates, deschlorometolachlor and deschloroacetylmetolachlor, were identified in some of the samples obtained. Deschlorometolachlor concentrations were relatively constant with depth at around 4 ng/L (note that the reference material had not been fully purified at the time of sampling; concentrations, therefore, were determined assuming a molar response factor similar to metolachlor). Concentrations of deschloroacetylmetolachlor increased from nondetectable at shallow depths, to 35 ng/L at a depth of 23 m. The mercaptoatrazine product was not detected. Although small amounts of mercaptometolachlor were detected, concentrations were not quantified, owing to difficulties encountered in purification of the reference material.
If we assume that the principal route of atrazine and metolachlor introduction to the upper Chesapeake Bay is via fluvial input from the Susquehanna River, and that the concentration of atrazine in sea water is negligible, we can account for dilution with sea water at the site sampled using the measurements of salinity we conducted. When we do so, we find that measured concentrations of atrazine in our profile are greater than those predicted from mixing of the Susquehanna River water with sea water. This may reflect additional sources of atrazine within the Chesapeake Bay, limitations in our assumption of the mean hydraulic residence time of our sampling point, or more simply that our "point" measurement of the input from the Susquehanna River does not accurately capture spatial or temporal variations in fluvial input. More detailed field investigations would be required to fully resolve this issue. For metolachlor, the measured concentrations in the Chesapeake Bay were less than predicted for dilution with sea water. Moreover, the ratio of metolachlor to atrazine in the Susquehanna River sample was greater than we measured in our profile. These observations suggest that natural degradation processes within the Chesapeake Bay are attenuating metolachlor more rapidly than atrazine.
Mass Balance Calculations for Atrazine Within Chesapeake Bay. Available data were compiled to identify important inputs of atrazine and to estimate the resident mass of atrazine in the Chesapeake Bay. The objectives of this portion of the project were to: (1) quantify the mass loading of this herbicide to the Chesapeake Bay; (2) use the data to estimate a resident atrazine mass; and (3) possibly identify from the data the potential transformation processes that may contribute to its degradation. A simple mass balance for the northern section of the Chesapeake Bay based on these data, summarized here, illustrated that to assess and predict the long-term trends of atrazine for different loading scenarios, more comprehensive field data and more sophisticated models are needed that better capture the relevant physical and chemical processes.
The total amount of atrazine entering the whole Chesapeake Bay via the various pathways is 4,700 kg/yr, and the amount entering the Upper Chesapeake Bay is 2,150 kg/yr. Atrazine inputs derive mainly from: (1) surface runoff, estimated to contribute about 60 percent of the total atrazine load; (2) groundwater flux, estimated to contribute about 23 percent of the total atrazine load; and (3) atmospheric input: wet and dry deposition, estimated to contribute about 17 percent of the total atrazine load to the Chesapeake Bay. Based on some necessary and reasonable assumptions given the quality of available data, we estimated that the average resident atrazine mass in the Chesapeake Bay is 32,760 kg (0-125,890 kg, 95 percent confidence interval), of which 6,224 kg (0-23, 919 kg, 95 percent confidence interval) is assumed to be in the northern 110 km of the Chesapeake Bay. To obtain this estimate, we assumed that the bulk of this herbicide resides in the water column and in the sediment layer. Samples for atrazine concentration in surface waters typically are taken from the top 0.5 m in the Chesapeake Bay, and it is implicitly assumed that this concentration is valid throughout the well-mixed upper layer. Reported concentrations for the data collected during 1976-1993 range between 0.03-4.3 µg/L. To compute the average mass of atrazine in the surface layer, we used the mean concentration over a 6-m thick, well-mixed surface layer, and multiplied it by the Chesapeake Bay's surface area, 7,800 km2 (total) and 1,480 km2 (Upper Chesapeake Bay only). For reference, we note that these values are of the same order as those reported by Schottler and Eisenreich (1997) for the atrazine inventory in, for example, Lake Michigan of the Great Lakes.
A first step toward understanding atrazine behavior in surface waters and sediment porewaters of the entire Chesapeake Bay is to understand its behavior in the 110 km long northern section. We focused our study on this region. Total input and output can be computed for the Upper Chesapeake Bay to provide a simple budget analysis, and the results compared to resident mass that was calculated from measured concentrations. The fluvial contributions from the Susquehanna, Chester, and Choptank Rivers yield a total surface flux of 1,800 kg/yr into this region. The input in groundwater to the upper region is 200 kg/yr, and the atmospheric deposition is 150 kg/yr. In summary, a total of 2,150 kg/yr is loaded to the Upper Chesapeake Bay with surface runoff estimated to contribute about 84 percent of the total load, groundwater flow contributing about 9 percent of total atrazine load, and atmospheric input constituting 7 percent of the atrazine load.
Using the range of concentrations reported in the literature, the total mass of atrazine resident in this part of the Chesapeake Bay ranges from 270 kg for the lowest reported concentrations to the unrealistic value of 38,700 kg for the upper limit. However, if we use the mean concentration value of 0.7 µg/L, we obtain a total mass of about 6,000 kg. The outflow from the Upper Chesapeake Bay over a depth of 6 m is approximately 11,000 m3/s (U.S. EPA, 1994). The rate of mass of atrazine removed by this flow is obtained by multiplying the outflow value by the concentration value. The lowest concentration value (0.03 µg/L), if assumed constant throughout the year, results in about 10,400 kg/yr removal of atrazine from this portion of the Chesapeake Bay, or 5,200 kg/yr if that concentration is present only half of the year. Yet, we can only account for 2,150 kg of atrazine as input; hence, this simple mass balance argument indicates that either a substantial source of atrazine is unaccounted for or, more likely, that concentrations are lower and are not uniform in time and space. Furthermore, atrazine must be leaving this region in the top layer by some diffusion-like mechanism. To achieve steady state, atrazine concentrations in the outflow surface water need to be 0.006 µg/L, but there is no evidence that supports this estimate.
Numerical Modeling of Atrazine Fate and Transport Within the Chesapeake Bay. Our laboratory and field investigations were complemented by a comprehensive modeling effort designed to investigate the behavior and distribution of atrazine in the Upper Chesapeake Bay. The main objectives of this portion of the project were to:
· Incorporate the compiled data into a comprehensive hydrodynamical model that can capture the most important physical and chemical processes occurring in an estuarine environment.
· Predict the horizontal and vertical distribution of atrazine in the Upper Chesapeake Bay.
· Investigate the relative importance of photolysis and a potential nucleophilic aromatic substitution reaction with polysulfides in anoxic sediment porewaters with respect to the degradation of atrazine in the water column of the Bay.
· Determine the minimum rate constant required for: (1) photolysis to be considered a significant removal mechanism, and (2) for the reaction between atrazine and polysulfides in the sediment porewaters to be considered a significant removal mechanism.
Initially, two models were adapted and unsuccessfully used to predict the behavior of this herbicide. These models were Modeling of Anthropogenic Substances in Aquatic Systems (MASAS) and a three-dimensional hydrodynamic model, INTROGLLVHT (the latter developed by J. E. Edinger Associates, Inc.). The multi-box model known as MASAS (developed by Ulrich, 1991), although successfully used to describe atrazine behavior in Swiss Lakes, cannot account for important dynamics occurring in large estuaries. In particular, the model assumes that the rates of vertical mixing are slower than reaction rates, and that both are much slower than horizontal mixing rates. Such assumptions are not valid for estuaries. INTROGLLVHT was studied at length as a potential model for our studies. Although correct flow patterns and salinity spatial distributions were generated using this model, there were several limitations. These included the inability to allow input data that varied with time which, in particular, limited the impact of the time-varying flow rate of the Susquehanna River and the simulated time of the numerical experiments, and the limitation of a uniformly sized rectilinear grid, which does not allow for a detailed shoreline or the use of one "column" of cells (or grid boxes) to represent the main channel. However, experiments conducted with this model served as excellent tools to become acquainted with the physical and chemical properties of the Chesapeake Bay, as well as to prepare the modeler for the use of a more complex model.
From the compilation of available data, current concentrations of atrazine and major inputs to the Chesapeake Bay were identified. This information, combined with the relevant physical and chemical processes occurring in the Upper Chesapeake Bay and the results from the analysis of field sampling, was incorporated into a three-dimensional numerical model. The model used for the final study, Generalized Environmental Modeling System for Surface Waters (GEMSS), also was developed by J.E. Edinger Associates, Inc. GEMSS is a comprehensive software package used for three-dimensional, time-varying simulations of rivers, lakes, estuaries, and coastal water bodies. GEMSS includes time-varying boundary conditions, as well as other time-varying input data. A curvilinear grid gives the user the ability to initialize the system in more than one region. The model captures horizontal advection and horizontal and vertical diffusion of atrazine as well as transformation of atrazine via photolysis and the potential reaction between atrazine and polysulfides present in the sediment porewaters. This finite difference model is designed in a modular fashion to allow coupling of existing modules (hydrodynamic module, water quality module, etc.) with user defined modules (J.E. Edinger Associates, Inc., Wayne, PA, unpublished work, 2001). The hydrodynamic and water quality modules were adapted to simulate the Chesapeake Bay environment and are coupled with two user-defined modules that simulate photolysis and reaction with polysulfides.
We have applied this model to the Upper Chesapeake Bay. A significant fluvial input of atrazine enters this region. The detailed information needed to set up the model includes bathymetry, latitude, average wind speed, initial temperature and salinity, and the inflow rate, location, and concentration of atrazine within tributary rivers, groundwater, and runoff. Location of the tidal boundary, mean tidal height, tidal amplitude, tidal period, and salinity, temperature and atrazine concentration at the tidal boundary are needed. In addition, it also is necessary to estimate the Chezy coefficient (which amounts to parameterizing the vertical mixing due to turbulent processes in the estuary), surface heat exchange coefficient, equilibrium temperature of heat exchange, and pseudo-first-order rate constant for the reaction between atrazine and polysulfides. Extensive model simulations were conducted for the time April 1, 1994 to October 1, 1994, as the most comprehensive data set for atrazine loading to the Upper Chesapeake Bay is available for this time. These months also capture the period of anoxia in the Chesapeake Bay when polysulfides are most likely to be present; hence, the reaction between atrazine and polysulfides most likely is to occur.
The model results show that the distribution of atrazine mass is highly variable in time and non-uniform in space. For all model runs, atrazine is present at the bottom of the Chesapeake Bay, where it may diffuse into sediment porewaters and react with naturally occurring polysulfides. Our results showed that atrazine always is transported down the water column, across the pycnocline into the bottom layers of the Upper Chesapeake Bay. They also show that atrazine is always present in the surface layers where it may be degraded via photolysis, and that the horizontal distribution of atrazine is variable. Predicted concentrations of atrazine in the Upper Chesapeake Bay range from 50 ng/L to 400 ng/L, which is of the same order of magnitude as the measurements reported in the literature. In addition, the distribution of atrazine mass is both temporally and spatially inhomogeneous. For all modeling scenarios, atrazine mass was present at the sediment-water interface where it may diffuse into the sediment porewaters and react with naturally occurring reduced sulfur species. Using the rate constants reported in the literature (for photolysis) or (for polysulfides) determined as part of this study, our results also show that neither photolysis nor the reaction with polysulfides is a significant sink for atrazine in the Upper Chesapeake Bay. Sensitivity analysis of the rate constants for photolysis and the reaction with polysulfides showed that the minimum values of these parameters required for these transformation processes to be considered a significant removal mechanism are 103 d-1 and 102 s-1, respectively. These values are several orders of magnitude greater than the highest reported values for haloazines (for which our laboratory studies reveal likely are to be encountered at environmentally relevant concentrations of polysulfides), and we therefore conclude that these processes likely are not important sinks for haloazines.
Our modeling efforts produced useful information concerning the magnitude of the rate constants that are necessary for photolysis and the nucleophilic aromatic substitution reaction between agrochemicals and reduced sulfur species to significantly decrease the atrazine concentration in the Upper Chesapeake Bay. This work can serve as a basis for future investigations of the behavior of herbicides in aquatic environments. For instance, the model results can be used to guide future field sampling to validate the simulated atrazine concentrations, and longer time scales can be modeled to determine if these transformation processes could represent removal mechanisms over long periods of time. Biodegradation and input of atrazine via groundwater and atmospheric deposition also may be incorporated into the model for a more accurate representation of the fate and transport of this herbicide in the Upper Chesapeake Bay.
References:
Borchardt LG, Easty DW. Gas chromatographic determination of elemental and polysulfide sulfur in kraft pulping liquors. Journal of Chromatography 1984;299:471-476.
Boulègue J, Lord C, Church TM. Sulfur speciation and associated trace metals (Fe, Cu) in the pore waters of Great Marsh, Delaware. Geochimica et Cosmochimica Acta 1982;46:453-464.
Duval B, Ludlam SD. The black water chemocline of meromictic Lower Mystic Lake, Massachusetts, USA. International Review of Hydrobiology 2001;86:165-181.
Giggenbach W. Equilibria involving polysulfide ions in aqueous sulfide solutions up to 240 degrees. Inorganic Chemistry 1974;13:1724-1730.
Kariuki S, Morra MJ, Umiker KJ, Cheng IF. Determination of total ionic polysulfides by differential pulse polarography. Analytica Chimica Acta 2001;442:277-285.
Loch AR, Lippa KA, Carlson DL, Chin YP, Traina SJ, Roberts AL. Nucleophilic aliphatic substitution reactions of propachlor, alachlor, and metolachlor with bisulfide (HS-) and polysulfides (Sn2-). Environmental Science and Technology 2002;36:4065-4073.
Nixon SW, Ammerman J, Atkinson LP, Berounsky VM, Billen G, Boicourt WC, Boynton WR, Church TC, DiToro DM, Elmgren R, Garber JH, Giblin AE, Jahnke RA, Owens NJP, Pilson MEQ, Seitzinger S, et al. The fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 1996;35:141-180.
Plust SJ, Loehe JR, Feher FJ, Benedict JH, Herbrandson HF. Kinetics and mechanism of hydrolysis of chloro-1,3,5-triazines. Atrazine. Journal of Organic Chemistry 1981;46:3661-3665.
Rozan TF, Theberge SM, Luther GW III. Quantifying elemental sulfur, S(0), bisulfide (HS-) and polysulfides (Sx2-) using a voltammetric method. Analytica Chimica Acta 2000;415:175-184.
Schottler SP, Eisenreich SJ. Mass balance model to quantify atrazine sources, transformation rates, and trends in the Great Lakes. Environmental Science and Technology 1997;31:2,616-2,625.
Ulrich M. Modeling of chemicals in lakes - development and application of user-friendly simulation software (MASAS and CHEMSEE) on personal computers. Ph.D. Dissertation. Federal Institute of Technology, Zurich (ETH), Switzerland, 1991, No. 9,632.
U. S. Environmental Protection Agency. Response of the Chesapeake Bay water quality model to loading scenarios. Chesapeake Bay Program, Technical Report Series 101/94, 1994.
Weber EJ, Spidle DL, Thorn KA. Covalent binding of aniline to humic substances. 1. Kinetic studies. Environmental Science and Technology 1996;30(9):2,755-2,763
Journal Articles on this Report : 5 Displayed | Download in RIS Format
Other project views: | All 23 publications | 5 publications in selected types | All 5 journal articles |
---|
Type | Citation | ||
---|---|---|---|
|
Hladik ML, Jans U, Lippa KA, Roberts AL. Measurement and interpretation of polysulfides in a natural water. Aquatic Sciences. |
R826269 (Final) |
not available |
|
Lippa KA, Roberts AL. Reactions of atrazine with reduced sulfur species and model NOM compounds. American Chemical Society Division of Environmental Chemistry Preprints 1998;38(2):128-130. |
R826269 (Final) |
not available |
|
Lippa KA, Roberts AL. Nucleophilic aromatic substitution reactions of chloroazines with bisulfide (HS-) and polysulfides (Sn2-). Environmental Science and Technology 2002;36(9):2008-2018. |
R826269 (Final) |
not available |
|
Lippa KA, Klotz JC, Roberts AL. Transformations of chloro-s-triazine and chloroacetanilide agrochemicals in natural and polysulfide-fortified sulfidic waters. Environmental Toxicology and Chemistry. |
R826269 (Final) |
not available |
|
Salmun H, Goetchius K, Kolloru VS, Edinger JE. Modeling the fate and transport of atrazine in the upper Chesapeake Bay. Journal of Hydrology. |
R826269 (Final) |
not available |
Supplemental Keywords:
chemical transport, environmental chemistry, hydrology, modeling, Chesapeake Bay, agriculture, textile dyes., Scientific Discipline, Air, Waste, Ecosystem Protection/Environmental Exposure & Risk, Ecology, Chemistry, Fate & Transport, Engineering, Chemistry, & Physics, Biology, fate and transport, waste treatment, toxicology, estuaries, salt marshes, sulfur, kinetic models, Triazines, agriculture, halogenated hydrocarbons, water quality, herbicidesProgress and Final Reports:
Original AbstractThe perspectives, information and conclusions conveyed in research project abstracts, progress reports, final reports, journal abstracts and journal publications convey the viewpoints of the principal investigator and may not represent the views and policies of ORD and EPA. Conclusions drawn by the principal investigators have not been reviewed by the Agency.