Environ. Sci. Technol. 1998, 32, 3887-3892
Temporal Monitoring of Organochlorine Compounds in Seawater by Semipermeable Membranes following a Flooding Episode in Western Europe P E R - A N D E R S B E R G Q V I S T , * ,† BO STRANDBERG,† ROLF EKELUND,‡ CHRISTOFFER RAPPE,† AND A° K E G R A N M O ‡ Institute of Environmental Chemistry, Umea˚ University, S-901 87 Umea˚, Sweden, and Kristineberg Marine Research Station, Go¨teborg University, S-450 34 Fiskeba¨ckskil, Sweden
During January 1995, a severe flooding event in central and western Europe caused a significant outflow of contaminated freshwater into the Northern Sea and to Skagerrak. This water reached the Swedish west coast in mid-March. During this time, SPMD (semipermeable membrane device) sampling was performed directly in the plume. Depending on the salinity, this contaminated water was found at a depth of 22-25 m in Skagerrak. Two SPMD sampling sites were placed in this area at a depth of 24 m, and the SPMDs were changed every 1216 days during the 12-week study. In the meantime, water temperature increased from 4 to 8 °C at the 24 m depth. PCBs, DDTs, chlordane, hexachlorobenzene, R-HCH, γ-HCH (lindane), dieldrin, and other organochlorine compounds were detected in the SPMD samples. A time trend with elevated concentrations during the floodwater arrival, followed by decreasing concentrations, could be seen for most of the compounds studied. The highest levels of organochlorine compounds were found for PCBs and DDTs. The amounts found in the truly dissolved aqueous phase were respectively 4 and 2 times higher during mid-March 1995 when compared to an apparent baseline level measured during April and May. By using earlier reported relationships between water concentration and sampling rates, a truly dissolved concentration was calculated for each compound.
Introduction In January 1995, heavy rain in western Europe caused an unusually high flow of water into rivers and lakes. This resulted in a significant flood and subsequent transport of water into the North Sea. This freshwater was characterized by high turbidity, high levels of nutrients as nitrogen and phosphorus, and increased concentrations of pollutants. A portion of this freshwater was supposed to move northwards along the German and Danish coasts with the Jutland current, reaching Skagerrak and Kattegat, and finally also the Swedish west coast (Figure 1). * Corresponding author fax: +46-90-186155; phone: +46907865188; e-mail:
[email protected]. † Umea ˚ University. ‡ Go ¨ teborg University. 10.1021/es980146m CCC: $15.00 Published on Web 11/06/1998
1998 American Chemical Society
To validate that the transport of pesticides or other pollutants had taken place, an investigation was performed in the outer archipelago of Lysekil on the Swedish west coast. Because the pollutants were expected to be either dissolved in water or adsorbed to small organic carbon-containing particulates, different sampling strategies were used. In this paper, the results from the use of a relatively new technique for sampling hydrophobic pollutants from water, semipermeable membrane devices (SPMDs) (1, 2), are presented. For the calculation of total amounts of pesticides, the particulate organic carbon (POC) concentrations in seawater were measured during the SPMD sampling.
Materials and Methods Experimental. On the March 16, 1995, SPMDs were placed together with a temperature logger (Tiny-Loggers, INTAB, Stenkullen, Sweden) registering water temperature every 10 min. On the basis of oceanographic data on the floodwater distribution, the SPMD sampling devices were placed at a depth of 24 m. Site 1 was situated in the shadow of one small island, and it is possible that the current water reached site 1 later than site 2. Both sites were a few miles off the Swedish coast. Every 12 to 16 days, the SPMDs were replaced by new ones in order to study the concentration trends of polychlorinated biphenyls (PCBs), dichlorodiphenyltrichloroethanes including inpurities and metabolites (DDTs), chlordanes (CHLs), hexachlorocyclohexanes (HCHs), hexachlorobenzene (HCBz), and other organochlorine compounds. At each site, four replicate SPMDs were sampled simultanously, and they were later pooled for one composite sample per site and time period. The use of four SPMDs lowered the detection limit by a factor of 4 without significant background interferences during the cleanup of the samples. The exposed SPMDs were immediately sent to the laboratory and kept frozen until analyzed. Materials. The SPMD consisted of a thin liquid layer in a layflat tube made from a semipermeable membrane (1, 2). The device was placed in a perforated stainless steel container to protect the membranes against mechanical damage and to restrict water flow velocity at the membrane surface. Inside this container, four standard size (90 cm × 2.5 cm) polyethylene membranes (EST, St Joseph, MO, U.S. Patents 5,098,573 and 5,395,426) were mounted on fixtures to give a spread configuration. After being sampled, the membrane is placed in a clean airtight screw-capped vial or airtight steel can. The membrane surfaces were cleaned from periphyton, minerals, and particulates with a soft brusch, rinsed with clean water and a Kleenex tissue, and finally dialyzed with cyclopentane for 3 days including two solvent changes resulting in one 180-mL fraction. After dialysis, six 13C-labeled internal standards were added to the extract, and this extract was slowly evaporated to dryness. To remove some remaining codialyzed lipids and polymer waxes originating from the membrane, cleanup was performed using high-resolution gel permeation chromatography (HR-GPC). The GPC concisted of a LKB 2150 HPLC pump, a 2151 HPLC variable wavelength monitor, a 2152 HPLC controller together with an Merck AS-4000 intelligent Auto Sampler and a HP 3392 A integrator. The gel type was a Polymer Laboratories PL gel (styrene-divenylbenzene copolymer, pore size 50 Å). The column was 7.5 × 300 × 2 mm giving a total length of 600 mm and a sorbent particle size of 5 µm. The mobile phase used was a mixture of dichloromethane-n-hexane (35/65), and the flow was 0.7 mL/min. VOL. 32, NO. 24, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 1. Map of the Jutland current originating from the English Channel to the Swedish west coast. Sampling sites 1 and 2 are situated outside the city of Lysekil in Sweden. Large arrows show the main Jutland current. Smaller arrows indicate other currents. Dark color is land and the iso lines is water depth in Skagerrak and Kattegat. The second fraction (24-48 mL) from the GPC contained the contaminants of interest in this study. This fraction was reduced in volume, spiked with two 13C-labeled recovery standards, redissolved in 40 µL of tetradecane and transferred to an autosampler vial. The final analysis was performed on a HRGC/LRMS instrument (Fisons MD800) equipped with a 60-m column (Supelco PTE-5, 60 m × 0.32 mm i.d., 0.25 mm film thickness, Bellefonte, PA) and running in SIR mode. All solvents used in this study were glass distilled (Burdick and Jackson, Muskegon, MI). This study focused on organochlorine compounds such as lindane, dieldrin, penta-(PCBz) and hexa-(HCBz) chlorobenzenes, DDT components (viz. o,p- and p,p-DDT, DDD, and DDE and DDMU), some CHL substances (viz. heptachlor (hepCHL), cis-heptachlorepoxide (hepCHLep), U82, MC4, trans-chlordane (t-CHL), MC5, cis-chlordane (c-CHL), MC7, oxychlordane (oxyCHL), MC6, trans-nonachlor (t-nonaCHL) and cis-nonachlor (c-nonaCHL)), and a number of PCB (polychlorinated biphenyl) congeners from tri through decachlorinated homologues. All compounds included in this study were monitored using two main ions in the chlorine cluster, evaluated on chlorine cluster ratio, and compared with actual standards for gas chromatography retention time. PCB congener-specific identification was performed using a 1:1:1:1 PCB mixture containing four different Aroclors (3, 4), and chlordane component identification was made according to refs 5 and 6. A laboratory blank consisting of four SPMDs stored in the lab for several months was analyzed in parallel with the samples, and the analytical results were corrected for contamination in this blank sample as a worst case. 3888
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Recoveries were calculated for each of the added 13C-labeled compounds, and the final results are corrected for losses against the corresponding or nearest internal standard. No correction for influence of biofouling on the sampling rate is made since the degree of biofouling was low at the exposure temperatures and at 24 m depth. The only baseline data from this area that can be used for comparison are the late sampling intervals in this study.
Results and Discussion The hypothesis that the floodwater moved northwards was confirmed by Swedish and other oceanographic investigations along the Danish and Swedish coasts (7) as well as by satellite photos taken. From these observations, it is apparent that the water mass reached the Swedish coast in the second half of March. Simultaneously the freshwater mass was mixed with saltwater and went down to below the halocline before reaching the Swedish coastline, which has a more freshwater-influenced coastal stream. However, at the end of March and at the beginning of April, strong winds (> 10 m/s) caused a turbulent water situation resulting in additional dilution and movement of the floodwater masses away from the coast. The SPMDs were placed at 24 m depth as the water masses of interest were supposed to be found below the halocline at 20 m depth. At this depth, the linear flow will not change much, especially inside the perforated stainless steel container that is especially designed to protect the membranes and will thus reduce the water turbulence greatly, resulting in a rather constant environment.
FIGURE 2. Concentration of the detected DDTs in samples from the two sampling points, sites 1 and 2, during March-June 1995. Water temperatures at exposure sites were measured, and with some small variations the temperature increased from 4 to 8 °C during the sampling period. Vertical profiles of salinity and temperature were taken on occasions of SPMD sampling. These data show the halocline position at the different sites and were used to verify the sampling depth. For the calculation of truly dissolved water concentrations of the pollutants, uptake rate data for each compound at different temperatures is needed. In this case, data for some organochlorine pollutants at a water temperature of 26 °C were used (8). Assuming that the decrease in the uptake rate was a factor of 2 with a temperature decrease of 10 °C (J. Huckins, personal communication), it is possible to extrapolate to the uptake rate at 4-8 °C. For some pollutants, data on the relation between water concentration and uptake rate into the SPMD were not available. In these cases, a comparison to chemical compounds with similar structures has been made. For PCBs, the only available uptake rate published (2) is for 2,2′,5,5′-tetrachlorobiphenyl (PCB 52) at 18 °C. In this paper, all PCB congeners are calculated against this single isomer, assuming similar uptake of all compounds. The values calculated for PCBs are therefore considered estimates. The sampling periods were chosen in order to detect possible changes of pollutant levels resulting from floodwater reaching the Swedish coast. Since each sampling had a duration of only 12-16 days, equilibrium was not approached for the compounds of interest. Thus, the results reflect a time-weighted average of truly dissolved water concentration during this time span. The total concentration in water cannot be determinded with this method but can be estimated using partitioning coefficients and total organic carbon data (TOC) (8). The concentration of dissolved DDT and analouges detected are shown for both sites 1 and 2 in Figure 2. At site 2, the first (16-29/3) sample shows a concentration around 4 times higher than the other sampling times at this site. On the other hand, at site 1, the second sampling time gave the highest concentration, and the results from the first sampling period are about one-third less in concentration. This indicates that the polluted water mass reached site 2 first and then after some days reached site 1. The maximum sum of the water concentration of DDT and analouges is calculated to be 0.45-0.48 ng/L at both sites, but during different time intervals. The two most abundant compounds in the DDT
group are p,p′-DDT and p,p′-DDD, whereas p,p′-DDE showed a concentration of 0.01 ng/L or less. This results in a ratio of approximately 1/20 between p,p′-DDE and p,p′-DDT and a ratio of 1/2 between p,p′-DDD and p,p′-DDT throughout all sampling intervals and sites. The sum of all detected PCB congeners in Figure 3 shows a different temporal pattern. In contrast to the sum of the DDTs, the sum of the PCBs was highest at site 1 on the first sampling time, and the second sample time at site 1 and the first sampling time at site 2 are nearly equivalent in magnitude. This difference might indicate a different geographic location of the main source for PCBs as compared to DDTs. Also, in Figure 3, the profile of PCB congener groups (e.g., tri-, tetra-, penta-, hexa, hepta-, and octachlorinated biphenyls) shows a similar behavior in temporal pattern. The second sampling interval at site 1 and the first sampling interval at site 2 contain a higher chlorinated PCB mixture when compared to all other samplings in this study. When uptake rate data are avaiable for all PCB congeners, a closer comparison between the different congeners can be made. The 11 different penta-CB peaks have been compared, and it was verified that this pattern is very consistent between all samples at both sites, indicating a common source. Another large group of compounds is the chlordane mixture, and some of the constituents are presented in Figure 4. The most abundant compounds formed in technical chlordane dominate in the samples. trans- and cis-chlordane (0.10-0.26 ng/L) are detected at equal levels, whereas the minor components in technical chlordane and the metabolic product cis-heptachlorepoxide are minor components in the samples. The temporal pattern of the sum of CHLs is similar to that of sum of DDTs, where the first sampling interval at site 1 is lower than the second sampling interval. The first two sample intervals are followed by a period with lower water concentrations, consisting of one and two sample intervals, respectively. Dissimilar to the aforementioned compounds, the chlordanes increase in concentration again after the April 10-24, 1995, sampling period. Concentrations similar to the early sampling intervals were observed during the May 10-22 period at site 1, and nearly the same concentrations were detected during the April 24-May 10 time period at site 2. Lindane and dieldrin are still used, but the use is limited in the northern part of Europe. In most European countries only the γ-HCH (99% clean) is allowed to be used. Still the VOL. 32, NO. 24, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 3. Concentration of the sum of detected PCBs and the homologue pattern of tri- to octa-chlorinated PCBs in samples from the two sampling points, sites 1 and 2, during March-June 1995. Concentration is calculated by using uptake rate data from PCB 52.
FIGURE 4. Concentration of eight chlordane components in samples 1995. R- and γ-HCH water concentrations follow a similar pattern as described earlier for chlordanes (Figure 5). The ratio between R- and γ-HCH varied during the study. According to Iwata et al. (9), most of the open sea surface water samples 3890
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from the two sampling points, sites 1 and 2, during March-June showed a γ-/R-HCH ratio of 0.1-0.2 except for the Mediterranean area. In the present study, the ratio is 2.3-5.4, which indicates a very recent use of lindane (γ-HCH). However, this ratio does not show any trend in the material, and the
FIGURE 5. Concentration of r- and γ-hexachlorocyclohexane and dieldrin in samples from the two sampling points, sites 1 and 2, during March-June 1995.
FIGURE 6. Concentration of penta- and hexachlorobenzene in samples from the two sampling points, sites 1 and 2, during March-June 1995. lowest ratio is 2.3 at site 1 during sampling interval 4. In fact, the ratio between the γ and R isomers at site 2 is strikingly similar throughout the whole study with a variation between 3.7 and 4.7 while site 1 shows a larger variation. No recent use of dieldrin in northern Europe is reported. Due to the instability of this compound during the commonly used method of strong acid treatment, its occurrence in the environment will usually not be documented. With the nondestructive analytical technique used in this study, dieldrin was detected in all SPMD samples from Skagerrak. After an initial peak concentration, subsequent “baseline” levels were similar (0.21-0.30 ng/L) for both sites (Figure 5). The presence of chlorobenzenes (CBz) in the environment is a result of both technical production and unwanted formation during incineration processes. In this study pentaand hexachlorobenzenes were analyzed, and they showed no clear trend through time (Figure 6). The concentration of pentachlorobenzene (0.20-0.37 ng/L) was usually a factor
of 2 higher than HCBz (0.07-0.14 ng/L). The ratios between the two compounds in each sample show no clear relationship. During the first sampling interval, a water mass with an estimated flow rate of 10 000 m3/s in the Jutland current transported pollutants to the Swedish west coast. The time from the start of flooding in middle Europe to the arrival of the water front at the Swedish coastline was approximately 60 days. During this interval, changes of composition of chemicals can occur by evaporation, degradation, partitioning, and sedimentation processes. If our sampling points at a depth of 24 m were only affected by this current, then the total transported mass of truly dissolved γ-HCH would be 2.3 kg/day at site 1. The additional contribution by γ-HCH adsorbed to particles is estimated to be only 0.02-0.05%, based on a log KOC value of 2.96 and a particulate organic carbon (POC) concentration of 15-45 µg/L. Since the deposition rate of pollutants during the transport from the VOL. 32, NO. 24, 1998 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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source via the Jutland current to the Swedish coast is hard to calculate, the influence of other water masses at our sampling locations is unsure, and the particulate organic carbon is changing (decreasing) during the transport, only a conservative estimate can be made. The half-life (t1/2, the time it takes to degrade one-half of the present concentration) for degradation of γ-HCH in natural freshwater is estimated to be 6 weeks according to Sharom et al. (10). Using this value, the transport of γ-HCH from European rivers to the Jutland current is estimated to be larger than 6 kg/day since the transport time from the European rivers to the Swedish coast is estimated to be 8 weeks and the t1/2 for lindane was estimated to be 6 weeks multiplied with 2.3 kg/day in the arrival current. Any addition of other sources along the Jutland current will decrease the obtained value. Performing the same calculations for p,p′-DDT with a log Koc of 5.3 and a t1/2 for degradation of 1 week (10), the daily transport from the flooded area was more than 1.5 kg and the amount reaching Swedish coastal water was approximately 0.2 kg/ day. The additional amount calculated for particle-bound material contributed 4-11% in addition to the truly dissolved water concentration. In cases with compounds having a higher log Kow values and large POC levels in the water, the total substance concentration will be significantly higher than the dissolved fraction. However, this does not always mean that the bioavaiable portion for bioconcentration will be significantly different. Note that the calculations on the net flux of chemicals to the Swedish coast via floodwater were based on variables of unknown accuracy. The SPMDs sampled a multitude of organochlorine pollutants in seawater along the Swedish west coast. The concentration calculated for one sampling interval represents a time weighted average and nothing is known about the variations within the period. The time series of samples analyzed give a clear indication of increased concentrations of pollutants in the water masses during the first part of the investigation (March-April 1995). The coincidence between elevated levels of pollutants and the arrival of a water mass identified as diluted floodwater indicates that the flood event in Europe has contributed to the pollution of Skagerrak and the Swedish west coast. Three different temporal patterns can be distinguished among the analyzed compounds. The DDTs, PCBs, and dieldrin showed elevated levels in the water during the first and second sampling period, and then a stable, lower contaminant level followed. Chlordanes and HCHs shows elevated levels at the beginning followed by a temporary decrease and then a second rise in concentrations. Finally, the chlorobenzenes showed no measurable shift in concentration during the total sampling time.
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The ratios between both DDT/DDE and R-HCH/γ-HCH indicate a recent use or effluents from sources containing the original compounds. In addition, an apparent different PCB source (higher degree of chlorination) was seen during the early sampling (Figure 3). PCBs had a different temporal pattern as compared to all other compounds as they were the only compound class with the highest water concentration in the first sampling interval at site 1. The lower concentrations of DDTs and PCBs found during the later part of the present investigation are considered as background levels since no other data exist from this area, and these concentrations are also similar to earlier reported values (11). Although to ensure that this hypothesis is true, relevant comparisons could be made between our SPMD measurements and ultrafiltrated water sampling. Finally, SPMDs appear to be a promising tool to follow medium to long-term composition and concentration changes of hydrophobic micropollutants even in the open sea.
Acknowledgments The investigation was financed by Kristineberg Marine Research Centre, University of Go¨teborg, and Institute of Environmental Chemistry, Umeå University.
Literature Cited (1) Huckins, J. N.; Tubergen, M. W.; Manuweera, G. K. Chemosphere 1990, 20, 533-552. (2) Huckins, J. N.; Manuweera, G. K.; Petty, J. D.; Mackay, D.; Lebo, J. Environ. Sci. Technol. 1993, 27, 2489-2496. (3) Schwartz, T. R.; Stalling, D. L. Arch. Environ. Contam. Toxicol. 1991, 20, 183-199. (4) van Bavel, B. Analytical and environmental chemistry of polychlorinated dibenzodioxins, dibenzofurans and biphenyls. Ph.D. Thesis, Umeå University, 1995, ISBN 91-7191-123-5. (5) Miyazaki, T.; Yamagishi, T.; Matsumoto, M. Arch. Environ. Contam. Toxicol. 1985, 14, 475-483. (6) Dearth, M. A.; Hites, R. A. Environ. Sci. Technol. 1991, 25, 12791285. (7) Rydberg, L. University of Go¨teborg, Sweden, 1998, unpublished. (8) Huckins, J. N.; Petty, J. D.; Lebo, J. A.; Orazio, C. E.; Prest, H. F.; Tillitt, D. E.; Ellis, G. S.; Johnson, B. T.; Manuweera, G. K. In Techniques in Aquatic Toxicology; Ostrander, G. K., Ed.; Lewis Publishers: Boca Raton, FL, 1996; pp 625-655. (9) Iwata, H.; Tanabe, S.; Sakal, N.; Tatsukawa, R. Environ. Sci. Technol. 1993, 27, 1080-1098. (10) Sharom, M. S.; Miles, J. R. W.; Harris, C. R.; McEwen, F. L. Water Res. 1980, 14, 1089-1093. (11) Clark, R. B. Marine Pollution; Clarendon Press: Oxford, 1992; pp 1-172.
Received for review February 12, 1998. Revised manuscript received May 28, 1998. Accepted June 15, 1998. ES980146M
