Environ. Sci. Technol. 1997, 31, 1068-1073
3,3′-Dichlorobenzidine Transformation Processes in Natural Sediments MARIANNE C. NYMAN,† ARTO K. NYMAN,‡ LINDA S. LEE,‡ LORING F. NIES,† AND E R N E S T R . B L A T C H L E Y I I I * ,† Environmental and Hydraulic Engineering Area, School of Civil Engineering, Purdue University, West Lafayette, Indiana 47907-1284, and Department of Agronomy, Purdue University, West Lafayette, Indiana 47907-1150
Release of 3,3′-dichlorobenzidine (DCB), an intermediate in dye manufacturing, and its congeners are of environmental concern due to their carcinogenic nature. To elucidate the fate of these compounds in sediment/water systems, sediment/water mixtures were spiked with DCB and incubated at 20 °C for 12 months under anaerobic conditions. The sediments used in this study were collected from Lake Macatawa (Holland, MI) and ranged in composition from silty-clay to sandy. Dehalogenation of DCB to benzidine appeared to take place through a transient intermediate, 3-monochlorobenzidine, which was observed in timecourse analyses of the sediment/water mixtures. No metabolites were observed in autoclaved samples, suggesting that dehalogenation of DCB in anaerobic sediment/ water systems was mediated by microbial activity. The product of dehalogenation (benzidine) is more toxic to humans than the parent compound, DCB. From sediment/water distribution experiments, DCB showed greater affinity for the sediment phase than its non-chlorinated derivative, benzidine. Therefore, progressive dehalogenation of DCB in anaerobic lake sediments is expected to yield a greater total concentration of aromatic amines in the solution phase, a shift to a more toxic form, and greater potential for transport in the environment.
Introduction Several million kilograms of 3,3′-dichlorobenzidine (DCB) and benzidine were produced in the United States up to 1977 for the production of dyes and pigments (1). Recognition of the carcinogenic nature of DCB and its lesser-chlorinated congeners including benzidine resulted in reduction in their use (1, 2). Benzidine has been found to be carcinogenic in the human bladder and in oral passages in animals. DCB has been likewise acknowledged to induce cancer in animals and is considered a potential carcinogen in humans (2). The carcinogenity of DCB toward humans is believed to be attributable to dehalogenation in the digestive system, resulting in benzidine formation (3). The U.S. Environmental Protection Agency (4) established water quality criteria for DCB and benzidine of 0.010 µg/L and 0.12 ng/L, respectively; ambient water containing DCB and benzidine at these * Author to whom all correspondence should be addressed; telephone: (317)-494-0316; fax: (317)-496-1107; e-mail: blatch@ecn. purdue.edu. † School of Civil Engineering. ‡ Department of Agronomy.
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concentrations was estimated to result in an incremental increase of human cancer risk of 10-6 over the lifetime of an exposed population. Several factors will govern the transport and fate of hydrophobic organic chemicals in sediment/water environments including episodic sediment resuspension, microbial transformations, sorption/desorption processes, and photochemical transformations. The vast majority of hydrophobic organic compounds are contained within the sediment phase under normal conditions. However, episodic resuspension substantially contributes to transport of these compounds in lake systems (5). Sediment resuspension behavior allows small particles to remain in the mobile water phase for long periods of time, allowing them to travel long distances. During resuspension, contaminant desorption is promoted by exposure of contaminated sediment particles to relatively “clean” water. Microbially mediated reactions and sorption are major processes affecting the fate of organic compounds in aquatic systems (6-10). Aryl halides have been shown to undergo microbially-mediated dehalogenation under anaerobic conditions (11-19). For example, chloroanilines and polychlorinated biphenyl congeners have been shown to undergo microbially-mediated reductive dehalogenation in sediment/ water systems, yielding less chlorinated congeners (15-20). Given the information that exists regarding microbiallymediated dehalogenations of these structurally similar chlorinated organic compounds (e.g., chloroanilines and PCBs), it is reasonable to hypothesize that DCB may also undergo microbially-mediated dehalogenation in a similar manner, though the process may be slow (11-20). The distribution of DCB and its congeners in sediment/ water systems will be affected by sorption processes. Several models for estimating sorptive equilibria for organic compounds have been developed based on the physical and chemical properties of the compound and sorbent (9, 10, 21). The effects of cation exchange capacity (CEC) and organic carbon (OC) content of soils and sediments, solution pH, and the solute’s dissociation constant (pKa) on sorption of aromatic amines have been widely studied (6, 7, 9, 10, 2123). Sorption of aromatic amines has been observed to be positively correlated to CEC and OC content and to be influenced by speciation (i.e., the pH-pKa relationship). Blatchley et al. (24) found the sorptive capacity of silty-clay sediments for DCB to be much greater than for sandy sediments. They attributed this to lower OC content and surface area of the sandy sediments. In contrast, Sikka et al. (6) found no obvious relationship between organic carbon content and sorption of DCB to sediment. Aromatic amines also have been shown to undergo covalent bonding with natural soil organic matter, in particular humic substances from sediments, soils, and natural water (25, 26). The objective of this study was to examine the processes most likely to govern DCB fate in contaminated sediments: anaerobic microbial transformations and sorption. Experiments were conducted to evaluate the behavior of DCB in anaerobic sediments over a period of approximately 1 year. The sediment/water distribution behavior of these compounds was examined with a range of sediment samples. Results from these investigations are presented as part of a comprehensive environmental fate assessment of DCB and its congeners in surface water systems.
Experimental Methods Materials. Acetonitrile, methanol, acetone, and glacial acetic acid were all HPLC grade as supplied by Fisher Scientific. Benzidine (95% purity) and DCB (99% purity) were purchased
S0013-936X(96)00571-8 CCC: $14.00
1997 American Chemical Society
TABLE 1. Selected Properties of DCB and Benzidine compd
formula
DCB benzidine a
Ref 27.
b
C12H10Cl2N2 C12H12N2
Ref 7. c Ref 8.
d
mol wt (g/mol) 253.13 184.23
melting pta (°C) 132-133 115-120
aq solubilitya (mg/L) 3.99 400.00
pKa,1
pKa,2
log Kow
1.6b
3.2b
3.3c
4.3c
3.5b 1.34d
Ref 28.
from Sigma Chemical and Chem Services, respectively, and used as received. Selected properties of DCB and benzidine are given in Table 1. Sediment and water samples were collected from Lake Macatawa, Holland, MI. Input of DCB to the lake is believed to be from a single point source through an industrial wastewater treatment facility’s permitted effluent diffuser. No other sources of DCB or its related compounds are known to exist in Lake Macatawa. Sediment pH was determined by rotating 5.0 g of dry sediment with 5.0 g of lake water for 1 day and measuring pH of the suspension using a combination electrode connected to a Corning Model 130 pH meter (29). Total organic carbon content of sediments was estimated by weighing oven-dried sediments (103 °C for 24 h) before and after combustion at 550 °C for 24 h. Cation exchange capacity (CEC) was estimated by equilibrating 5.0 g of air-dried sediment with 1.0 N ammonium acetate for 30 min and analyzing the supernatant for K+, Na+, Mg2+, and Ca2+ using atomic absorption spectroscopy. This method assumed that CEC sites were primarily occupied by the cations analyzed and that cation displacement was entirely attributable to the ammonium ion (30). Dry and wet sieving were used for sandy and silty-clay sediments, respectively, to determine the particle size distribution and to separate the size fractions. Particle size distributions were confirmed using a settling tube and an optical particle size analyzer (Brinkmann Instruments, Model 2010). Biotransformation Studies. Biotransformation of DCB was studied using glass serum bottles containing 50 g of wet sediment and 180 mL of lake water, flushed with N2 gas to create an anaerobic environment, and sealed using open closure screw caps with Teflon-lined septa. Lake water was used in the experiments to mimic the original environment. Stock solutions of DCB were prepared by gravimetric addition of solid DCB to methanol. Bottles were spiked with 1.5 and 4 mL of stock solution, resulting in initial concentrations of approximately 2 µmol of DCB/kg of sediment or 9.9 µmol of DCB/kg of sediment, respectively. Some of the experiments were performed in duplicate. Autoclaved (121 °C, 150 min) sediment/water samples, subsequently spiked with DCB, served as controls. The bottles were incubated in the dark at 20 °C for 12 months. At several times during the course of the investigation, subsamples (8 mL) were taken using a syringe with a 19-G needle (inside diameter of 0.686 mm). Prior to sampling, bottles were shaken to facilitate collection of a homogenized sample. The slurry collected in the syringe was subjected to centrifugation at 1350g for 75 min, and both the supernatant and sediment were analyzed for benzidine congeners. Sorbed chemical concentrations were evaluated by extracting the sediment for 1 day with 3.0 mL of methanol, followed by analysis of the methanol phase for benzidine congeners. All bottles were evaluated individually. Prior to the biodegradation experiments, sediments were extracted with methanol to estimate the background concentrations of DCB and benzidine. These concentrations ranged from nondetectable to 2.18 × 10-2 µmol/kg sediment or 0-0.22% of the total initial DCB burden spiked into the biodegradation bottles. Batch Isotherms. Sorption isotherms on whole sediments were measured independently for both DCB and benzidine using a batch-equilibration method. Isotherms were
conducted in borosilicate glass tubes (8 mL) with Teflonlined caps. Benzidine was dissolved in lake water and added to the sediment. Due to DCB’s limited aqueous solubility, it was introduced to the tube using a plating method (31) wherein DCB dissolved in acetone was added to empty tubes and the solvent was allowed to evaporate prior to adding sediment and lake water. Sediment mass to lake water volumes ranged from 0.02 to 0.32 g/mL for different sediment samples to achieve sorption of approximately 50% of the chemical added. Preliminary studies were performed to estimate the time required for equilibration. Sorptive equilibrium was assumed to be reached when successive analyses of the liquid phase showed no significant change (e4%) in concentration. Results suggested that equilibrium was achieved within 5 days for sandy sediments and within 6-8 days for silty-clay sediments. Sample tubes were wrapped in aluminum foil during equilibration to minimize photolysis. Following equilibration, samples were centrifuged at 1350g for 75 min, and supernatant was analyzed for DCB or benzidine. Each isotherm was duplicated. Sorption to glassware and volatilization of DCB or benzidine during an 8-day equilibration were found to be negligible in preliminary studies where the initial concentration of DCB or benzidine was compared to the final concentration for both compounds after 8 days of equilibration. Sorption coefficients were estimated using the Freundlich model: Se ) KfCeN where Se and Ce represent sorbed (µmol/g) and solution (µmol/mL) concentrations at equilibrium, respectively, and Kf (µmol1-N mLN g-1) and N are fitted Freundlich coefficients. Solution concentrations were measured directly, whereas sorbed concentrations (Se) were estimated by difference: Se ) (Ci - Ce)(V/m) where Ci is the initial concentration (µmol/mL) of the chemical, V is the solution volume (mL), and m is the dry soil mass (g). Ci for DCB was estimated by dividing the chemical mass (µmol) plated in the equilibration tube by the volume of lake water added to the tube. HPLC Analysis. Supernatants and solvent extracts from the biotransformation studies and isotherm measurements were analyzed using high-performance liquid chromatography (HPLC) with ultraviolet (UV) detection. The HPLC system used was a Shimadzu automated gradient system (LC10AD pump, FCV-10AL solvent delivery system, SCL-10A system controller, SIL-10A auto injector) equipped with a UV-VIS detector (Shimadzu SPD-10A) operated at 285 nm and a reverse-phase column from Supelco Inc. (LC-1, 4.6mm i.d. × 25 cm) with a guard column (10 mm × 4.6 mm) packed with Spherisorb silica C1 (Alltech Services). The mobile phase consisted of 30/70 to 50/50 v/v mixtures of acetonitrile and 0.1 M acetate buffer (pH 4.7) and was delivered to the system at approximately 1.5 mL/min. Retention times (tR) of 8.7 and 3.8 min for DCB and benzidine, respectively, resulted with the 40/60 v/v acetonitrile/buffer solution. Some samples were run under two different mobile phase conditions to confirm peak identification in the chromatographic results. The average peak area from no less than two replicate injections was used to estimate sample concentration by comparison with a calibration curve developed using standards consisting of known amounts of DCB and benzidine dissolved in methanol.
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TABLE 2. Properties of Sediments Used in This Study particle size distribution (%) sanda
silt and claya
sample
pH
total OC (%)
CEC (cmolc/kg)
>841 µm
250-841 µm
149-250 µm
74-149 µm
