Not to be cited without prior reference to the author ICES CM 2007/J:10
Passive sampling devices to assess the diffusive transfer of persistent hydrophobic organic contaminants at the sediment-water interface in the coastal marine environment C. Tixier*, F. Léauté, K. Héas-Moisan, P. Le Gall, D. Munaron, C. Munschy, J. Tronczyński
Abstract : Persistent hydrophobic organic contaminants accumulate in the coastal marine environment and can enter trophic webs. In marine ecosystems, the sediment-compartment may act as a burial sink and also as a diffusive source of contaminants to water and organisms. For a better understanding of the fate and bioavailability of contaminants at the sediment-water interface, we studied the bio-geo-chemical behaviour of various model hydrophobic organic contaminants (polycyclic aromatic hydrocarbons PAHs and polychlorinated biphenyls PCBs). The site studied is a Mediterranean lagoon located south of France. This lagoon presents an intensive shellfish farming activity and receives input from different human activities, i.e. urban activities, industries, port activities and agriculture. Two types of passive sampling experiment with low-density polyethylene (LDPE) strips were carried out to assess the diffusive mobility of PAHs and PCBs. Firstly, LDPE strips were exposed in laboratory conditions to sediments collected in the lagoon. The analysis of these strips enables us to determine the freely dissolved concentration of contaminants in pore water. Another type of experiment was conducted by exposing LDPE strips to the overlying water column, in situ at the sediment-water interface. Using this combined approach based on passive sampling techniques, concentrations of freely dissolved persistent organic contaminants were determined to assess their diffusive exchanges at the sediment-water interface. These experiments provide field data for the fate and ecological modelling of organic contaminants in the Thau lagoon. These data may also serve to better assess environmental quality standards for marine coastal sediments. Keywords: passive sampling, LDPE, sediment, organic contaminants, diffusive transfer * Contact author: Céline Tixier, IFREMER (French Research Institute for Exploitation of the Sea)- DCN-BE, Laboratory of Biogeochemistry of Organic Contaminants, BP 21105, 44311 Nantes, France [tel: +33 (0)240 374134, fax: +33 (0)240 374075, e-mail:
[email protected]].
1. Introduction In semi-enclosed marine ecosystems, such as lagoons, exchanges at the sediment water interface play a key role in the fate of persistent hydrophobic organic contaminants (HOCs) accumulated in sediments. Among the various processes governing the behaviour of contaminants, diffusive exchanges of freely dissolved contaminants between the water column and surface mixed sediment layer can determine the role of the sediment bed as a source and/or sink of contaminants. Recent work revealed quite high levels of contamination by PAHs and PCBs in the sediment of Thau lagoon, one of the largest Mediterranean lagoons, located in the South of France1. This lagoon is subject to different human activities: tourism, urban development, industries and agriculture2,3. Fishing, fish farming and shellfish farming, are other intensive activities taking place in the lagoon. With standing stocks of oysters (Crassostera gigas) at 40 000 tons and annual production of about 15 000 tons of oysters and 4 000 tons of mussels (Mytilus galloprovincialis), the shellfish farming may exercise a major influence on the functioning of Thau lagoon ecosystem4,5,6. On the other hand, the quantitative assessments of contaminant exchanges between sediments and water column in such closed and shallow marine ecosystems is necessary for the understanding and prediction of the future trends of marine habitats contamination by persistent HOCs. Our study aims at assessing the diffusive exchanges of HOCs at the sediment-water interface- i.e. the determination of the equilibrium state of these chemicals between sediment and overlying waters in Thau lagoon. Freely dissolved concentrations of various hydrophobic organic contaminants (polycyclic aromatic hydrocarbons -PAHs and polychlorinated biphenyls -PCBs) were therefore determined by passive sampling experiments with low-density polyethylene (LDPE) strips exposed to the sediments in the laboratory and at the sediment-water interface, in situ, in the lagoon. LDPE strips have been previously successfully used to assess freely dissolved concentrations of various HOCs in various water bodies 7,8,9,10 and also in heavily contaminated harbor sediments10,11,12. Our study also provides some more scientific information on the suitability of using such sampling technique in sites exposed to diffuse chronic contamination.
2. Materials and method 2.1. Sampling site 2 Thau lagoon covers a surface area of 75 km , with an average depth of 4-5 m and a maximum depth of 11 m. Its 2 catchment area covers 250 km . The lagoon is isolated from the Mediterranean Sea by an offshore bar and is connected to the sea by two channels. The experiments were carried out at one station in the eastern part of the lagoon, where most of the anthropic activities are located. 2.2. Passive sampler preparation Single layered strips were prepared from low-density polyethylene (LDPE) lay-flat tubing (2.5 cm wide) containing no additives (Brentwood plastics, MO, USA). Strips length was 30 cm or 50 cm for the exposure in sediments or in water respectively. The strips were pre-extracted twice with hexane and then spiked with seven Performance Reference Compounds (PRCs) covering a logKOW range from 4.5 (Anthracene-d10) to 7.3 (CB204)13. Strips were kept in a freezer (-20oC) until further treatment. 2.3. Passive sampling experiment in sediment (laboratory conditions) Sediment cores were collected manually by divers using Teflon® tubes (9 cm i.d.). A homogenized composite sample made up of the four top centimetres of the core was used for sediment exposure. Twelve glass jars were then filled with 280 g of wet sediment and one LDPE strip of 30 cm. These jars were capped with aluminiumlined lids and tumbled on a rotating system over a period of two months. Exposure was performed in a UVprotected climate laboratory at 20°C. Strips were regularly sampled for kinetics study. Unexposed strips were also used as control samples over the exposure time period. 2.4. Passive sampling experiment in water (field exposure) A special device was designed for this experiment. It consisted of a rectangular stainless steal frame of 1 m length and 60 cm wide. Sixteen strips of 50 cm were mounted horizontally on the frame. The frame was deployed by divers horizontally at the sediment-water interface, the strips being exposed at about 2.5 cm above the sediment bed and at around 7.5 m depth. Strips were regularly sampled after 16, 43, 64, 101 and 140 days of exposure. Composite sample were made from two strips exposed during the same period of time. Exposed strips were transported to the laboratory in glass jars put into a cooler and were kept frozen until processing. Field blanks and unexposed strips kept in the laboratory were analysed to check contamination from transportation and handling. The water temperature in the lagoon ranged from 9 to 14 °C and the salinity was about 37‰ during the exposure time period. 2.5. Extraction, clean-up and analyses Aliquots of sediment were taken to determine dry weight and organic carbon content (Thermoquest CHNS-O elemental analyzer). About 10 g of dry weight sediment sample were extracted twice with dichloromethane using an accelerated solvent extractor (ASE, Dionex). After exposure, strips were washed with milli-ro water and extracted twice by soaking overnight in hexane. Hexane extracts were combined and internal standards added. Elemental sulfur was removed from all extracts by treatment with copper powder activated with hydrochloric acid. The clean-up and fractionation of all extracts were made by adsorption chromatography on a two layer silica/alumina column. The first two fractions, eluted with hexane and hexane/dichloromethane, were analysed after concentration, for PAHs and alkylated homologues by GC/MS in selective-ion monitoring14. The first fraction was also analysed for PCBs by GC/ECD according to the procedure described earlier15. Internal QA/QC procedures including laboratory and field blanks, analyses of replicate samples for precision determination and use of recovery standards added to each sample before extraction made it possible to verify loss of analytes during the entire sample work-out. The laboratory is also routinely participating in the QUASIMEME (Quality Assurance of Information for Marine Environment Monitoring in Europe) intercomparison exercises for PAHs and PCBs. 2.6. Non linear and linear Regression: Non-linear and linear regressions were performed in the Origin System 7.5 SR2 version 7 (OriginLab Corporation, MA, USA).
3. Results and discussion: 3.1. Passive sampling experiments in sediments
Uptake kinetics: Accumulation process in the membrane obeys first order kinetics for all the compounds studied and, as described by Huckins et al.16, the concentration in the membrane after a time t is:
CLDPE (t ) = CLDPE [1 − exp(− ket )] eq
(eq.1) -1
eq
where C LDPE (ng/g) is the concentration in the LDPE strip at equilibrium and ke (d ) is the exchange rate constant. Figure 1 presents the uptake curves obtained for two PAHs and three PCBs over the exposure time of two months. 250000 250
600
R2 = 0.99
500 2
R = 0.99 400
300
(a) 200
100
0
CB153 CB180 CB66
Concentration in LDPE strip (ng/g)
Indeno[1,2,3-cd]pyrene Benzo[e]pyrene
Concentration in LDPE strip (pg/g)
Concentration LDPE strip(ng/g) (ng/g) Concentration in in LDPE strip
700
200000 200
R2 = 0.98
150000 150
100000 100
(b) R2 = 0.99
50000 50
R2 = 0.96 0
0
10
20
30
40
50
0
10
20
Time (day)
30
40
50
Time (day)
Figure 1: Uptake of two PAHs (a) and three PCBs (b) by LDPE strips exposed to sediment.
The uptake curves fit very well with equation 1 with a significant coefficient of determination for all the contaminants. Replicate analyses at 30 and 49 days show low relative standard deviation and good reproducibility (RSD between 2 and 8%). PAHs up to four aromatic rings and PCBs up to four chlorine atoms reach sorption equilibrium within ten days of incubation. Most of the studied compounds have reached equilibrium after 30 days. At the end of the experiment (49 days), all compounds have reached more than 98% of sorption equilibrium. The exchange rates between LDPE strip and water were determined for each contaminants by a regression fit of the uptake. As shown in figure 2, the estimated exchange rate constant ke is well correlated to the hydrophobicity of the compounds represented here by the logarithm of the octanol-water partition coefficient. The elimination rate constant for the added PRCs are in agreement with the value of the exchange rate constant observed for the native contaminants. Similar results were obtained by Booij et al.10 with LDPE strips exposed to contaminated sediment slurries from harbors. 0,5
PAHs PCBs PAH PRCs PCB PRCs
Log ke
0,0
yPCB = -0,5782x + 3,0672 2
R = 0,9294
-0,5
-1,0 yHAP = -0,6195x + 3,0801 2
R = 0,7985 -1,5 4
4,5
5
5,5
6
6,5
7
7,5
8
Log Kow Figure 2 : Exchange rate coefficients (ke) as a function of log Kow17,18 for passive sampling experiments with LDPE strips exposed to sediments
Determination of dissolved porewater concentrations: At equilibrium, the concentration of the contaminants in the membrane is related through the LDPE_water partition coefficient (KLDPE_water) to the aqueous dissolved eq
concentration ( C water ) and may be expressed as: eq C LDPE (t ) = K LDPE − water C water [1 − exp(− k e t )]
(eq.2)
In our study, the water concentration is determined from the concentration at equilibrium in the LDPE strips fitted to equation 1 and from the LDPE_water partition coefficients calculated according to Booij et al.19. These authors have reported KLDPE_water values obtained at different temperatures. KLDPE_water used in our calculations were adjusted to the experimental temperature of 20°C. This estimation of aqueous concentrations in pore water assumes no depletion during the experiment. However, if the sediment presents a low sorption capacity, the
presence of the LDPE strip in the system can lead to a depletion of contaminants in both the pore water and the solid phase of the sediment. In a recent work, Huckins et al.16 summarized some conditions regarding experimental design to avoid the depletive conditions. Firstly, at equilibrium, the amount of contaminant accumulated in the membrane should be much smaller than the initial amount of this contaminant in the sediment. This condition is fulfilled in our experiments because the amounts accumulated in the membrane range from 0.1 to 1% and 2 to 10% of the total amount of PAHs and PCBs respectively. Secondly, the authors recommended a sampler volume to carbon organic mass ratio of 0.05 ml.g-1. In our experiment, this ratio is of 0.12 ml.g-1. Hence, the depletion cannot be excluded for the most hydrophobic contaminants and the pore water concentration determined in our experiment may be underestimated. To control the level of depletion, strips were spiked with PRCs before exposure to the sediment20. At equilibrium, the amount of PRCs dissipated from the strips may be considered as equal to the amount of PRCs accumulated in the sediment (the amount of PRCs in the water phase being negligible). Therefore, it is possible to calculate a distribution factor (DF) corresponding to the ratio of the amount of PRCs remaining in the membrane and the amount of PRCs in the sediment at equilibrium. This ratio also represents the sorption capacity of the membrane relative to the sorption capacity of the sediment. A high value of DF means that the sorption capacity of the LDPE strip is very important compared to the one of the sediment and that the depletion of contaminants from the particulate phase of the sediment will be quite high. Figure 3 presents the value of DF obtained for the different PRCs used in the present study. 2,5 HAP PRCs PCB PRCs
2,0
DF
1,5 1,0 0,5 0,0 4
4,5
5
5,5
6
6,5
7
7,5
Log Kow
Figure 3: Distribution factor for the different PRCs as a function of log Kow17,18
DF values range from less than 0.1 for fluoranthene-d10 to more than 2 for CB204. PRC dissipation curves clearly show that all added PRCs reach equilibrium within the exposure time. Therefore, in our experimental conditions, using LDPE strips, the exponential increase of DF with the logKow cannot be related to a nonequilibrium state, unlike suggested by Smedes et al.20. It appears that for the more hydrophobic compounds (log Kow > 6.5), the sorption capacity of the LDPE strips seems to be much higher than their sorption in the sediment. Therefore, for these compounds, we estimated that an underestimation of the pore water concentration by a factor higher than 1.5 could be expected. However, for the majority of the less hydrophobic contaminants studied, the pore water concentration may be underestimated only by a factor lower than 1.3. Hence, the results presented in this paper are not corrected for this depletion phenomenon. The determined PAH dissolved concentration levels in pore water range from 60 pg/L for dibenz[a,h]anthracene to 2660 pg/L for pyrene. The dissolved concentration levels of PCBs are much lower, from 0.3 pg /L for CB 194 to 66 pg/L for CB 153. 3.2. Passive sampling experiments in overlying water Time-averaged water concentrations at the sediment-water interface were calculated according to equation 2, from the exchange rate coefficient determined by regression fit of the uptake and from the LDPE_water partition coefficients calculated according to Booij et al.19. KLDPE_water values were adjusted to the in situ temperature. The water concentration levels range from 8 pg /L for indeno[1,2,3-cd]pyrene to 2 ng/L for C2-pyrenes/anthracenes. The concentration levels of PCBs are much lower, from 0.1 pg /L for CB 194 to 9 pg/L for CB 153. 3.3. Water-sediment equilibrium: Figure 4 shows the distribution patterns of PAHs in sediment, pore water and overlying water. The high molecular weight PAHs are present in higher proportion in pore water than in overlying water, where these compounds are close to detection limits. However, enrichment of lower molecular weight PAHs is observed in both water phases. PAHs of three/four aromatics rings account for 18% of the total amount of parent compounds present in water and only for 9% in the sediment. Enrichment of three to four aromatic rings alkylated PAHs is also observed in overlying water. These compounds account for more than 60% of lower weight PAHs. For PCBs (data not shown), enrichment of the low chlorinated PCBs is observed in both water phases: three to four chlorinated PCBs represent around 40% of the total PCB amount in the overlying water and only 18% in the sediment. However, the level of contamination is around seven times higher in pore water than in overlying water.
4000
pg/L
3000
Overlying water
2000 1000
*
0 4000 pg/L
*
**
3000 2000
Pore water
1000
*
0 1200 ng/g dry weight
*
**
900
Sediment
600
C1-BNTs
C3-DBT
Benzonaphthothiophenes
C2-DBT
C1-DBT
Dibenzothiophene
Benzo[g,h,i]perylene
Dibenz[a,h]anthracene
Perylene
Indeno[1,2,3-cd]pyrene
Benzo[a]pyrene
C1-BFLs
Benzo[e]pyrene
Benzo[k]fluoranthene (BFL)
C2-CHR
Benzo[b+j]fluoranthene (BFL)
C1-CHR
Chrysene/Triphenylene (CHR)
C2-PY/Fl
Benz[a]anthracene
C1-PY/Fl
Pyrene( PY)
C3-P/A
Fluoranthene (Fl)
C2-P/A
C1-P/A
Anthracene (A)
0
Phenanthrene (P)
300
Figure 4: Distribution patterns of PAHs in sediment, pore water and overlying water. *No octanol-water partition coefficient available for the calculations
100
6 aromatic rings 5 aromatic rings 3/4 aromatic rings
10
CB 194
CB 180
CB 170
CB 1 83
CB 156
CB 174
CB 153
CB 151
CB 138
CB 118
CB 101
CB 87
CB 49
CB 52
CB 44
CB 31
CB 28
C2-CHR
C1-CHR
C3-P/A
C1-PY/Fl
C2-P/A
C1-P/A
C3-DBT
C2-DBT
C1-DBT
Indeno[1,2,3-cd]pyrene
Benzo[e]pyrene
Benzo[g,h,i]perylene
Benzo[k]fluoranthene (BFL)
Benzo[b+j]fluoranthene (BFL)
Benz[a]anthracene
Chrysene/Triphenylene (CHR)
Pyrene( PY)
Fluoranthene (Fl)
Anthracene (A)
Phenanthrene (P)
0.1
Dibenzothiophene
1
Benzo[a]pyrene
Conc. Pore water / Conc. Overlying water
Figure 5 presents the ratio of the pore water/water concentrations for PAHs and PCBs. This ratio also corresponds to the ratio of the chemical activity of the contaminants in these environmental compartments and thus can be cautiously used to indicate the direction and extent of the diffusion between pore water and overlying water21. A ratio close to 1 means pore water-water equilibrium and a ratio higher than 1 means net diffusion from pore water to water. For PAHs, the ratio seems to increase with increasing hydrophobicity of the contaminants, this ratio ranging from 2 to 12 for the three to six aromatic rings PAHs respectively. Similar results are obtained for the alkylated derivatives with a ratio increasing with the degree of substitution of the aromatic ring. The ratio is more or less constant for all PCB congeners, independently of the number of chlorine atoms. According to these results, the system appears to be close to equilibrium for low molecular weight PAHs. On the other hand, for PCBs and high molecular weight PAHs, diffusion from the sediment to the water is expected.
Figure 5: Ratio of dissolved pore water to overlying water concentrations for PAHs and PCBs
4. Conclusion: Passive sampling experiments with LDPE strips enabled the determination of truly dissolved water concentration of persistent hydrophobic organic contaminants in both water column and pore water of sediments. The truly dissolved concentrations of contaminants, which are very difficult to determine with classical methods, can be related to the availability of contaminants for mass transfer and uptake by organisms. Furthermore, by combining the results obtained for pore water and overlying water, the direction and the extent of the diffusive exchange at the sediment-water interface was determined in Thau lagoon. It appears that low molecular PAHs are close to or at equilibrium with regards to the sediment-water interface, while PCBs and higher molecular weight PAHs are not at equilibrium. For these contaminants, the sediment in Thau lagoon constitutes a source of contamination to the water column. Further investigation on present day fluxes and temporal trends of contaminants in the sediment will enable better prediction of long-term contamination by HOCs from the sediment of Thau lagoon. These results will also be used for ecological modelling of organic contaminants in the lagoon. Finally, passive sampling technique appears as a useful tool to assess HOCs exchanges at the sediment-water interface in marine habitats. However, good practice of passive sampling technique requires comprehensive understanding of the uptake process of contaminants as well as thorough calibration of the device used. .
References: 1.
F. Léauté, Tixier C., Munschy C., Tronczyński J. (in preparation). Present-day distributions and past background concentrations of persistent organic contaminants (PAHs and PCBs) in the sediments of the Mediterranean lagoon (Thau) (to be submitted in 2007 to Marine Pollution Bulletin).
2.
Labonne M., Othman D.B., Luck J.-M., (2001). Pb isotopes in mussels as tracers of metal sources and water movements on a lagoon (Thau Basin, S. France). Chem. Geol., 181: 181-191.
3.
Monna F., Lancelot J., Bernat M., Mercadier H., (1997). Taux de sédimentation dans l'étang de Thau à partir des données géochronologiques, géochimiques et des repères stratigraphiques. Oceanol. Acta., 20: 627-638.
4.
Gangnery A., Bacher C., Buestel D., (2001). Assessing the production and the impact of cultivated oysters in the Thau lagoon (Mediterranee, France) with a population dynamics model. Can. J. Fish Aquatic Sci., 58: 1012-1020.
5.
Duport E., Gilbert F., Poggiale J.-C., Stora G.(2007). Macrofauna and sediment reworking quantification at the Thau Lagoon. Estuar. Coastal Shelf Sci. 72: 522-533
6.
Plus M., Chapelle A., Lazure P., Auby I., Levavasseur G., Verlaque M., Belsher T., Deslous-Paoli J.-M., Zaldívar J.-M., Murray C.N., (2003). Modelling of oxygen and nitrogen cycling as a function of macrophyte community in the Thau lagoon. Cont. Shelf Res., 23: 1877-1898.
7.
Müller J. F., Manomanii K., Mortimer M.R., McLachlan M.S., (2001). Partitioning of polycyclic aromatic hydrocarbons in the polyethylene/water system. Fresenius Journal of Analytical Chemistry, 371: 816-822.
8.
Carls M. G., Holland L.G., Short J.W., Heintz R.A., Rice S.D., (2004). Monitoring polynuclear aromatic hydrocarbons in aqueous environments with passive low-density polyethylene membrane devices. Environ. Toxicol. Chem., 23(6): 1416-1424.
9.
Adams R. G., Lohmann R., Fernandez L.A., MacFarlane J.K., Gschwend P.M., (2007). Polyethylene devices : passive samplers for measuring dissolved hydrophobic organic compounds in aquatic environments. Environ. Sci. Technol., 41(4): 1317-1323.
10. Booij K., Hoedemaker J. R., Bakker J.F., (2003). Dissolved PCBs, PAHs, and HCB in pore waters and overlaying waters of contaminated harbor sediments. Environ. Sci. Technol., 37(18): 4213-4220. 11. Vinturella A. E., Burgess R. M., Coull B.A., Thompson K.M., Shine J.P., (2004). Use of passive samplers to mimic uptake of polycyclic aromatic hydrocarbons by benthic polychaetes. Environ. Sci. Technol., 38(4): 1154-1160. 12. Lohman R., Burgess R.M., Cantwell M.G., Ryba S.A., MacFarlane J.K., Gschwend P.M., (2004). Dependency of polychlorinated biphenyl and polycyclic aromatic hydrocarbon bioaccumulation in Mya Arenaria on both water column and sediment bed chemical activities. Environ. Toxicol. Chem. 23(11): 2551-2562.
13. Booij K., Smedes F., van Weerlee E. M., (2002). Spiking of performance reference compounds in low density polyethylene and silicone passive water samplers. Chemosphere, 46: 1157-1161. 14. Tronczyński J., Munschy C., Héas-Moisan K., Guiot N., Truquet I., Olivier N., Men S., Furaut A., (2004). Contamination of the Bay of Biscay by polycyclic aromatic hydrocarbons (PAHs) following the T/V “Erika” oil spill. Aquat. Living Res., 17 : 243-259. 15. Johansson I., Héas-Moisan K., Guiot N., Munschy C., Tronczyński J., (2006). Polybrominated diphenyl ethers (PBDEs) in mussels from selected French coastal sites : 1981-2003. Chemosphere, 64: 296-305. 16. Huckins J N., Petty J. D., Booij K., (2006). Monitors of Organic Chemicals in the Environment. In: Semipermeable Membrane Devices. Springer, New-York. 17 Hawker D. W., Connell D. W., (1988). Octanol-water partition coefficients of polychlorinated biphenyl congeners. Environ. Sci. Technol., 22 : 382-387 18 Mackay D., Shiu W. Y., Ma K. C., (2000). Physical-Chemical Properties and Environnemental Fate. Handbook, Chapman & Hall/CRCnetBase. 19 Booij K., Hofmans H. E., Fisher C.V., Van Weerlee E.M., (2003). Temperature-dependant uptake rates of nonpolar organic compounds by semipermeable membrane devices and low-density polyethylene membranes. Environ. Sci. Technol., 37(2): 361-366. 20 Smedes F., van der Zande T., Tixier C., Davies I.M., (2007). ICES Passive sampling trial survey for water and sediment (PSTS) 2006-2007. Part 2: Laboratory intercomparaison, analytical issues and lessons learned. International Council for the exploitation of the Sea, Annual Science Conference, 2007, Helsinki, Finland. ICES CM 2007/J/03. 21 Reichenberg F., Mayer P., (2006). Two complementary sides of bioavailability/accessibility and chemical activity of organic contaminants in sediments and soils. Environ. Sci. Technol., 25(5): 1239-1245.
