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Polycyclic aromatic hydrocarbons in lake sediments from the High Tatras
Environmental Pollution 159 (2011) 1234e1240
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Polycyclic aromatic hydrocarbons in lake sediments from the High Tatras Barend L. van Drooge a, *, Jordi López a, Pilar Fernández a, Joan O. Grimalt a, Evzen Stuchlík b a b
Institute of Environmental Assessment and Water Research (IDÆA-CSIC), Jordi Girona 18, 08034 Barcelona, Catalonia, Spain Department of Hydrology, Charles University, Vinicná 7, 12044 Prague, Czech Republic
High sedimentary PAH loads were observed in alpine lakes in the High Tatras (Eastern Europe) which are related to high PAH atmospheric deposition fluxes.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 14 October 2010 Received in revised form 19 January 2011 Accepted 25 January 2011
European alpine lake systems are used as indicators of air quality over the continent. Preliminary data showed high polycyclic aromatic hydrocarbons (PAH) loads in the High Tatras (Eastern Europe) in comparison to other mountain regions. Here, insight on the spatial distribution of PAH is provided from analysis of top-core sediments of 27 alpine lakes distributed along the High Tatras. Top-core sediment concentrations were higher than those in deep-cores, and they were higher than those observed in other European high mountain regions. The PAH profiles were uniform and comparable to those observed in aerosols and snow, indicating that atmospheric deposition was the predominant PAH input pathway to the lakes. Good agreement between estimated atmospheric deposition and sedimentation fluxes was observed. However, in several lakes in the western range higher sediment fluxes may correspond to higher PAH depositions levels. The higher concentrations may also reflect inputs from potential emission source areas. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Polycyclic aromatic hydrocarbons High mountain lake sediments High Tatras
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) in the environment are of concern due to their mutagenic, carcinogenic and teratogenic effects on organisms (Freitag et al., 1985; Howsam and Jones, 1998). The presence of these compounds in lake sediments may be due to diagenesis of organic matter (Wakeham et al., 1980) or anthropogenic activities (Hites et al., 1977). Anthropogenic pyrolytic PAH are produced by incomplete combustion of recent or ancient biomass material, such as wood or fossil fuels (Howsam and Jones, 1998). Studies in European high mountain regions have revealed the usefulness of alpine lake ecosystems as sentinel environmental indicators of the distribution patterns of the atmospheric pollution load over the European continent (Fernández and Grimalt, 2003). Alpine lakes are different from low altitude lakes at similar latitudes since they essentially receive pollutant inputs from regional or long-range atmospheric transport. Important PAH concentration differences have been observed in lake sediment among the different European high mountain regions. Preliminary data on the High Tatras, situated between Poland and Slovakia, in Eastern Europe (49100 e49140 N; 20 00 e20100 E; Fig. 1)
* Corresponding author. E-mail address: [email protected] (B.L. van Drooge). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.01.035
show high PAH concentrations when comparing to other European mountain areas (Fernández et al., 1999). This area is the only mountain range where the lake sediment concentrations exceed the “no-effect” level of sediment quality guidelines (Quiroz et al., 2010). These high concentrations are consistent with recent measurements of ambient air PAH levels and point to contributions of these compounds from regional emission sources (van Drooge et al., 2010). The present study is aimed to describe a spatial distribution of this atmospheric PAH load in the High Tatras and to get insight into the processes responsible for differential accumulations of these compounds in these lakes (Fig. 2). 2. Materials and methods 2.1. Selection of lakes for study in the High Tatras The 27 alpine lakes selected in the area of study (18 10 km) are shown in Fig. 1. All studied lakes are located between altitudes of 1580 m (above sea level) and 2145 m. They are situated above the regional timberline (Table 1). The summits of the steep High Tatras reach altitudes of 2655 m and are about 1800e2000 m above the surrounding lowlands. The present-day 2 C isotherm (timberline) is situated at 1550 m and 1650 m on the Northern and Southern slopes, respectively. The temperature lapse rates are 0.70 C/100 m and 0.67 C/100 m in the Northern and Southern slopes, respectively (Zasadni and Klapyta, 2009). These values were used to estimate the average annual air temperature of the studied lakes (Table 1). The climatic conditions and air circulation of the Tatras are dominated by air masses from the North Atlantic that have passed over the northwestern part of the European continent (Nied zwied z, 1992). The highest precipitation totals are recorded on the
B.L. van Drooge et al. / Environmental Pollution 159 (2011) 1234e1240
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Fig. 1. Location of lakes (Table 1). The border between Poland and Slovakian Republic is indicated by a thick dashed line. The meteorological stations are indicated by stars (KW ¼ Kasprowy Wierch; LS ¼ Lomnicky Stit; SP ¼ Skalnate Pleso).
northwestern slope of the mountain ridge, while the southern slope, lee side, is located in the precipitation shadow. In contrast to temperature, the change of precipitation with altitude is more uncertain. Nevertheless, there are an important differences between the meteorological stations of Lomnicky Stit (2653 m; 1776 mm/y) and Skalnate Pleso (1778 m; 1351 mm/y), both situated in the eastern part of the Tatras (Fig. 1). The station of Kaprowy Wierch (1991 m; 1716 mm/y), situated in the western part, shows similar precipitation rates as those of Lomnicky Stit (Zasadni and Klapyta, 2009). These values indicate an increase of precipitation with altitude as well as more precipitation in the western part of the High Tatras. They were used to estimate the average annual precipitation fluxes for the studied lakes (Table 1).
2.2. Sampling Sediment samples were collected in 2001 and taken in the deepest sites using a gravity coring system (Glew, 7.5 cm diameter, 30 cm long). This device was used to obtain the upper 0.5 cm core sediment of all lakes whereas a long sedimentary column, representing a record from present back to 1850 (Appleby and Piliposian, 2006) was collected in four cases Zadni, Vysne Wahlenbergovo, Velke Spisske and Ladove. Immediately after sampling, sediment cores were divided in sections of 0.5 cm and stored in pre-cleaned aluminum foil at 20 C until analysis. Moreover, Dlugi and Starolesnianske had been sampled with the same methodology in the past (Grimalt et al., 2004). 2.3. Analytical procedures
2001
0
5000
10000
15000
20000
1971
1941
1911 Ladove
1881
Starolesnianske Dlugi
1851
1821 Fig. 2. Time scales of the concentrations (ng/g) of analyzed in the High Tatras.
P PAH from the sediment cores
2.3.1. Materials Residue analysis n-hexane, dichloromethane, isooctane, methanol and acetone were from Merck (Darmstadt, Germany). Anhydrous sodium sulfate for analysis was also from Merck. Neutral aluminum oxide type 507C was from Fluka AG (Buchs, Switzerland). Cellulose extraction cartridges were from Whatman Ltd (Maidstone, England). Aluminum foil was rinsed with acetone and let to dry at ambient temperature prior to use. The purity of the solvents was checked by gas chromatographyemass spectrometry (GCeMS). No significant peaks should be detected for acceptance. Aluminum oxide, sodium sulfate and cellulose cartridges were cleaned by Soxhlet extraction with hexane:dichloromethane (4:1, v/v) during 24 h before use. The purity of the cleaned reagents was checked by ultrasonic extraction with n-hexane:dichloromethane (4:1; 3 20 mL), concentration to 50 mL and analysis by GCeMS. No interferences were detected. Sodium sulfate and aluminum oxide were activated overnight at 400 C and 120 C, respectively. d10-Anthracene, d12-benzo[ghi]perylene and d12-perylene were used as internal and surrogate standards. The standard mixture EPA mix 9 was used as external standard. All of them were purchased from Dr. Ehrenstoffer (Augsburg, Germany). 2.3.2. Sample extraction Sample sediment dry weight was calculated by weighing and data substraction before and after freeze-drying. About 0.5 g of wet sediment was extracted by sonication with methanol (20 mL; 20 min) in order to separate most of the interstitial water from the sediment. The subsequent extractions were performed with (2:1, v/v) dichloromethane:methanol (3 20 mL; 20 min). All extracts were combined and spiked with deuterated PAH recovery standards (d10-anthracene and d12-benzo[ghi]perylene). Then, they were vacuum evaporated to almost 10 mL and hydrolyzed overnight with 20 mL of 6% (w/w) KOH in methanol. The neutral fractions were recovered with n-hexane (3 10 mL), vacuum evaporated to almost
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Table 1 Characteristics of studied lakes in the High Tatras. Lake code
Official name
Latitude N
Longitude E
Altitude (m asl)
Lake area (ha)
Air temp ( C)
Precipitation (mm/y)
Total PAH (ng/g dw)
8 9a 11 12 13 14 15 17 18 19 21 22 23 26 27 29 30 31 32 37 42 43 45 47a 49 51 54
Zelené krivánske Dlugi Staw Ni zné Terianske Zadni Staw Polski Czarny Staw Ga˛ sienicowy Vysné Terianske Zmarzly Staw Ga˛ sienicowy Vysné Wahlenbergovo Czarny Staw Polski Ni zné Temnosmre cinské Capie Vysné Temnosmre cinské Wielki Staw Polski Malé Hincovo Velké Hincovo Czarny Staw pod Rysami Velké Zabie
49.1594 49.2273 49.1698 49.2134 49.2310 49.1680 49.2244 49.1642 49.2046 49.1929 49.1683 49.1891 49.2134 49.1740 49.1797 49.1888 49.1720 49.1663 49.1942 49.1523 49.1823 49.1788 49.1841 49.1800 49.1912 49.1841 49.1932
20.0085 20.0107 20.0143 20.0143 20.0199 20.0218 20.0238 20.0271 20.0277 20.0306 20.0378 20.0395 20.0404 20.0585 20.0606 20.0778 20.0786 20.0876 20.0943 20.1315 20.1551 20.1595 20.1629 20.1678 20.1701 20.1768 20.1964
2017 1784 1941 1890 1620 2109 1787 2145 1722 1674 2072 1716 1655 1923 1946 1580 1919 1998 1699 1879 2055 1972 2057 1986 1886 2011 2014
4.32 1.60 4.91 6.46 17.79 0.55 0.28 4.96 12.65 10.48 2.43 4.95 34.14 2.22 18.19 20.54 2.26 1.71 8.08 2.78 1.20 0.66 1.72 0.78 0.83 0.88 2.43
0.46 0.36 0.05 0.38 1.51 1.08 0.34 1.32 0.80 1.84 0.83 0.84 1.27 0.17 0.02 1.79 0.20 0.33 0.96 0.47 0.71 0.16 0.73 0.25 0.35 0.42 0.44
1468 1351 1430 1667 1532 1514 1615 1532 1583 1559 1495 1580 1549 1421 1432 1512 1419 1458 1571 1399 1487 1445 1488 1452 1402 1465 1466
6200 13,000 2300 14,000 6400 30,000 7100 24,000 23,000 19,000 6300 12,000 23,000 4400 9100 1800 2000 6800 11,000 3100 5700 8000 8800 18,000 5400 17,000 9900
a
Dra cie Vysné Zabie bielovodské Batizovské Pusté Vysné zbojnícke Ladové Starolesnianske Pleso Zabie javorové Prostredné sivé Vel’ké spisské
Those lakes were sampled and analyzed in 1993 and not included in the flux calculations.
dryness, and fractionated with a column containing 2 g of alumina which was eluted with 5 mL of n-hexane:dichloromethane (19:1 v/v) and 10 mL of dichloromethane:n-hexane (2:1 v/v). These combined fractions were vacuum evaporated to 500 mL and nitrogen concentrated to almost dryness. The extracts were redissolved in isooctane prior to instrumental analysis. 2.3.3. Instrumental analysis The following PAH were determined: fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, chrysene þ triphenylene, benzo[b þ j]fluoranthenes, benzo[k]fluoranthene, benzo[e]pyrene, benzo[a]pyrene, perylene, indeno [1,2,3-cd]pyrene, benzo[ghi]perylene and dibenz[a,h]anthracene. An internal standard of d12-perylene was added to the vials prior to injection. Samples were analyzed by GCeMS (Fisons 8000 Series, Mass Selective Detector 800 Series). A fused silica capillary column, HP-5 of 50 m and 0.25 mm i.d. (0.25 mm film thickness) was used. The oven temperature program started at 90 C (1 min hold), followed by a 4 /min ramp up to 300 C (15 min hold). Injector, transfer line and ion source temperatures were 280 C, 300 C and 200 C, respectively. Helium was used as carrier gas (1.1 mL/min). PAHs were determined in the electron impact and selected ion recording modes. The following mass fragments were used for identification and quantification: m/z 166,178, 202, 228, 252, 276, and 278 (dwell time 40 ms per single ion, ion window according to retention times of standards). Diagnostic ions of the corresponding perdeuterated standards, m/z: 188, 212, and 288, were also recorded. Identification was performed by combination of the EPA Mix 9 external standard and retention index methods. Calibration curves (detector response vs amount injected) were performed for each compound. The range of linearity of the detector was evaluated from the calibration curves. All measurements were performed in the ranges of linearity found for each compound. The quantitative data were corrected for surrogate recoveries which were 65% 14 and 79% 18 for d10-anthracene and d12-benzo[ghi]perylene, respectively. 2.3.4. Quality control Procedural blanks were performed with each set of nine samples to check for the presence of interfering peaks. All samples were blank corrected when necessary. The method detection limits based on signal to noise ratio of 3 in real samples ranged from 100 pg to 400 pg for individual compounds. The analytical procedure was successfully calibrated with a standard reference material with certified PAH values (marine sediment HS-4, Institute for Marine Biosciences, Canadian National Research Council).
3. Results and discussion 3.1. PAH concentrations The total PAH concentrations in the lake sediments are summarized in Table 1. The concentrations in the top-core
sediments were more than two orders of magnitude higher than those in the corresponding bottom-core sediments. Top-core P PAH concentrations ranged between 1800 ng/g dw and 30,000 ng/g dw. These high levels are in agreement with preliminary observations in the area (Fernández et al., 1999, 2000). Comparison of the average top-core PAH concentrations in the Tatras district with average concentrations found in other European mountain lake districts (Table 2) shows that the Tatra levels are about 3 times higher than those in Scotland lakes and 8 times higher than average lake values in the Pyrenees, Alps and Norway. These top-core sediment concentrations reflect recent inputs. The sediment cores obtained in Dlugi Staw and Starolesnianske Pleso exhibit a significant PAH increase from the beginning of the XX century. In Ladove, the PAH record increases sharply since 1950. In all three records the highest PAH appear to be associated with the industrial increase after 1970 (Fernández et al., 2000; Grimalt et al., 2004). 3.2. Profiles of PAH in sediments The average PAH distributions found in the lake sediments of the High Tatras are shown in Fig. 3. The top-core and bottom-core sediment composition was remarkably similar, which is consistent with previous observations indicating that long-range atmospheric PAH transport involves a significant degree of homogenization of the PAH mixtures (Simo et al., 1997). However, the concentrations
Table 2 PAH in alpine lake sediments from European mountain regions (Emerge Project). P Lat. Long. PAH (ng/g dw) Finland Norway Scotland High Tatras Alps Pyrenees
69 60 57 49 46 42
23 8 5 20 10 0
576 1353 3689 9771 1202 1345
B.L. van Drooge et al. / Environmental Pollution 159 (2011) 1234e1240
of some compounds such as perylene, pyrene, benz[a]anthracene and benzo[ghi]perylene follow different trends. Perylene is one of the compounds with lowest top-core relative concentration while in the bottom-core sample it is one of the predominant compounds (Fig. 3). The occurrence of relatively high perylene concentrations in the bottom sections of the sediment cores has also been observed in lake sediments from other mountain ranges (Grimalt et al., 2004) and in ancient sediment layers of freshwater and marine systems (Wakeham et al., 1979, 1980) and is related to its diagenetic formation. Pyrene, benz[a]anthracene and benzo[ghi]perylene were relatively less abundant in the bottom-core than in the top-core sediments. Nevertheless, their isomeric pairs, chrysene, fluoranthene and indeno[1,2,3-cd]pyrene, respectively, had comparable contributions in both core depths. This observation suggests that pyrene, benz[a]anthrancene and benzo[ghi]perylene may be less stable upon diagenesis than their isomeric pairs. The lower relative abundance of benzo[ghi]perylene in the bottom-core sediment than its isomeric pair, indeno[1,2,3-c,d]pyrene, may also reflect changes in emissions sources through history such as lower vehicle emissions due to fossil fuel combustion (Rogge et al., 1993) in the past. All top-core sediments show similar profiles, with good correlation between individual PAH compounds (0.84 < r2 < 0.98 (p > 0.01)) pointing to similar atmospheric qualitative PAH inputs to all these alpine lakes (Fig. 4). Comparison of the average sedimentary PAH profile with the PAH composition on atmospheric deposition over these high mountain lakes (Carrera et al., 2001) and with the atmospheric PAH gas and particle phase distribution in the High Tatras (van Drooge et al., 2010) shows a good agreement between the PAH composition in lake sediments and in the atmospheric particle phase (r2 ¼ 0.94; p < 0.01) and snow deposition (r2 ¼ 0.92; p < 0.01). Gas phase PAHs show a distinct composition (Fig. 3). Accordingly, PAH are incorporated into these lake sediments by transport associated mainly to atmospheric particles. Thus, despite the semi-volatile properties of these compounds, those of larger molecular weight, between benz[a]anthracene and benzo[ghi]perylene, are almost entirely particle bound, while the PAHs of lower molecular weight, between fluorene and pyrene, tend to shift towards the atmospheric particle phase at decreasing temperatures (Fernández et al., 2000; Lei and Wania, 2004).
0.75
1237
A
0.50
0.25
0.00 0.25
B
0.20
0.15
0.10
0.05
0.00 0.25
C
0.20
0.15
0.10
0.05
0.00 0.25
D
0.20
0.15
0.20
Fig. 4. PAH profiles in environmental compartments from the High Tatras; A: Air gas phase (van Drooge et al., 2010); B: Air particulate phase (van Drooge et al., 2010); C: Snow deposition (Carrera et al., 2001); D: Sediment (present study). FL ¼ fluorene, PHE ¼ phenanthrene, ANT ¼ anthracene, FLA ¼ fluoranthene, PYR ¼ pyrene, BANT ¼ benz[a]anthracene, CRY þ TRIP ¼ chrysene þ triphenylene, BBJFLA ¼ benzo [b þ j]fluoranthene, BKFLA ¼ benzo[k]fluoranthene, BEP ¼ benzo[e]pyrene, BAP ¼ benzo[a]pyrene, PER ¼ perylene, IP ¼ indeno[1,2,3-cd]pyrene, BGP ¼ benzo[ghi] perylene, DBA ¼ dibenzo[ah]anthracene.
0.15 0.10 0.05
DB A
IP
BG P
PE R
BA P
BE P
BB JF LA BK FL A
T
RI P
Y+ T
CH
PY R
BA N
FL A
AN T
PH E
0.00 FL
IP BG P D BA
PH
0.25
T
0.00
FL A PY R BA C N R Y+ T TR IP BF LA BE P BA P
0.30
AN
0.05
FL
0.35
E
0.10
Fig. 3. Average distribution (þSD) of individual PAH in top-core sediments (black bars) and bottom-core sediments (white bars). FL ¼ fluorene, PHE ¼ phenanthrene, ANT ¼ anthracene, FLA ¼ fluoranthene, PYR ¼ pyrene, BANT ¼ benz[a]anthracene, CRY þ TRIP ¼ chrysene þ triphenylene, BBJFLA ¼ benzo[b þ j]fluoranthene, BKFLA ¼ benzo[k]fluoranthene, BEP ¼ benzo[e]pyrene, BAP ¼ benzo[a]pyrene, PER ¼ perylene, IP ¼ indeno[1,2,3-cd]pyrene, BGP ¼ benzo[ghi]perylene, DBA ¼ dibenzo[ah]anthracene.
In the High Tatras, the highest ambient air PAH concentrations have been found in winter samples (van Drooge et al., 2010), probably as a consequence of higher PAH emissions in this season caused by higher energy demands on the European continent. Moreover, PAH are related to (soot) particles (Fernández et al., 2002), which increases their environmental persistence (Esteve et al., 2006) and affords their atmospheric transport to remote areas (Fernández et al., 2002; Ribes et al., 2003).
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Besides dry deposition, wet deposition may play an important role in PAH incorporation into high mountain lake systems. Due to the low ambient temperatures at least one third of the total precipitation to the lakes is in the form of snow, which is an efficient scavenger for vapors and particles (Franz and Eisenreich, 1998; Lei and Wania, 2004).
3.3. Spatial PAH variation The spatial distribution of the PAH concentrations in the sediment top cores of all studied lakes is shown in Fig. 5a. The lowest PAH P concentrations ( PAH ¼ 1800 ng/g dw) are observed in Czarny Staw (1580 m), which is the lake with lowest altitude situated in the northern slopes of the mountain range. The highest levels P ( PAH ¼ 30,000 ng/g dw) are found in Vysné Terianske pleso (2109 m) that is situated in the southern slopes. Comparison of the PAH sediment concentrations and physical variables, such as lake altitude, annual air temperature and precipitation, shows only a statistically significant correlation for precipitation (r ¼ 0.49; p < 0.02). To get further insight into the influence of atmospheric PAH deposition to the lakes and its burial into the sediments, mean annual deposition and sedimentation fluxes for benzo[a]pyrene (BAP) as a model-compound, were calculated for each individual lake and compared. Deposition fluxes were calculated by summing the dry deposition and the wet deposition in the form of rain (from May until October; 66% of total precipitation) and snow (November until April; 34% of total precipitation). The contribution of rain and snow was based on the long term precipitation measurements from the meteorological station of Skalnate Pleso (Slovakian Meteorological Institute), while the BAP concentrations of atmospheric particles were taken from measurements in the High Tatras between 2000 and 2001 (van Drooge et al., 2010). The dry deposition BAP flux was calculated according to Equation (1):
FluxBAP; dry deposition ¼ Vd CBAP; particle
ng=cm2 =y
(1)
where Vd was the deposition velocity assumed to be 1 m/h (Lei and Wania, 2004) and CBAP, particle was the average BAP concentration (ng/m3) in the atmospheric particle phase (van Drooge et al., 2010). The wet deposition in the form of rain was calculated according to Equation (2):
FluxBAP; rain deposition ¼ Wrain rain CBAP; particlerain
ng=cm2 =y
(2)
where Wrain was the washout factor for BAP, which was assumed to be 105 (Lei and Wania, 2004), rain was the estimated amount of rain (between May and October) to the lake corresponding to its altitude (mm/y) and CBAP, particleesnow was the average BAP concentration (ng/m3) in the atmospheric particle phase sampled in the rain season (van Drooge et al., 2010). The wet deposition in the form of snow was calculated according to Equation (3):
FluxBAP; snow deposition ¼ Wsnow snow CBAP; particlesnow
ng=cm2 =y (3)
where Wsnow was the washout factor for BAP, which was assumed to be 105 (Lei and Wania, 2004), snow was the estimated liquid amount of snow (between November and April) to the lake corresponding to its altitude (mm/y) and CBAP, particleesnow was the average BAP concentration (ng/m3) in the atmospheric particle phase sampled in the period of snowfall (van Drooge et al., 2010). The resulting estimated atmospheric deposition BAP fluxes are shown in Fig. 5b and range between 1.7 and 2.1 ng/cm2/y. The higher ambient air BAP concentrations in winter than in summer, 0.3 ng/m3 and 0.03 ng/m3, respectively (van Drooge et al., 2010) involve that about 79% of the BAP load to the lakes was during
P Fig. 5. A: Measured top-core sediment PAH concentrations (ng/g dw); B: Estimated BAP deposition flux (ng/cm2/y); C. Estimated BAP sedimentation flux (ng/cm2/y), D: Ratio sedimentation flux/deposition flux. For lake n 8 the values for the corresponding bar is given in the figures.
B.L. van Drooge et al. / Environmental Pollution 159 (2011) 1234e1240
snowfall (between November and April) while 16% was deposited by rain (between May and October). The remaining 5% was incorporated by dry deposition, which was assumed to be constant during the year throughout the mountain range. The relative contributions of these pathways are similar as those found in by others (Daly and Wania, 2004), which also reported snowfall as the most important deposition mode for BAP. The estimated deposition fluxes depend on the air PAH concentrations measured in Skalnate Pleso (van Drooge et al., 2010). These ambient air concentrations were high in comparison to those observed in sites from other European alpine lakes, such as those in the Alps, Pyrenees, or Norway, especially in winter (van Drooge et al., 2010). There are several urban areas situated at about 100 km of distance from the Tatras: Ostrava in the west, Krakow in the north and Kosice in the south, which may be potential PAH sources. In addition, the “Black Triangle” is situated at around 300 km to the west. This area has shown to have a large influence on the Tatras in the past (Kopacek et al., 2000). The general air circulation from the west and these potential source areas may give rise to high ambient air PAH concentrations such as those measured in Skalnate Pleso and used in the present study. These high ambient air PAH concentrations are consistent with the EMEP-model estimations of BAP emissions, where the highest continental BAP emission fluxes, higher than 20 ng/cm2/y, were estimated for a large area in southern Poland, just west of the High Tatras (Gusev et al., 2007). Moreover, a recent study (Umlauf et al., 2010) confirms the presence of very high ambient air particulate matter BAP levels (40 ng/m3) in winter in southern Poland, which was related to the practice of wood and coal combustion for domestic heating. Higher BAP air concentration may therefore occur in the western part of the High Tatras, which on its term may cause high atmospheric deposition fluxes. The estimated BAP deposition fluxes in the present study were lower, but in the same order of magnitude, as those calculated by EMEP (personal correspondence), which were estimated to be 7.4 ng/cm2/y for the High Tatras. The higher EMEP deposition loads estimated for this mountain range are probably related to the higher precipitation rates used in the EMEP-model calculation. Also, the EMEP stations were located in the lowland areas, with possible higher BAP background levels, while the one used in the present study, Skalnate Pleso, is a remote alpine site within the High Tatras. The BAP sedimentation fluxes were calculated using the average sedimentation rates in lakes from the Tatra Mountains described in Appleby and Piliposian (2006). The BAP calculation followed Equation (4):
FluxBAP; sediment ¼ Fluxsediment CBAP; sediment
ng=cm2 =y
(4)
where Fluxsediment was the sedimentation rate and CBAP, sediment was the BAP concentration (ng/g dw) measured in sediments from each individual lake. As expected, individual lakes exhibit diverse BAP sedimentation fluxes. An average sedimentation flux of 100 g/m2/y was used in the calculations due to the lack of sedimentation rates for the individual lakes. The average value was obtained from the average values of five Tatra lakes (Appleby and Piliposian, 2006) but the observed variation interval for these lakes ranged between 55 and 420 g/cm2/y, which may involve a stronger BAP sedimentation flux variability than assumed in the calculations. The BAP sedimentation fluxes obtained from the calculations described above are shown in Fig. 5d. The average value is 3.4 ng/ cm2/y which is less than two fold higher than the estimated average total atmospheric BAP deposition flux, 1.9 ng/cm2/y (Fig. 5e). The
1239
good agreement shows that independently of individual lake variability the average estimates provide consistent figures for comparison. Furthermore, a substantial positive correlation between calculated atmospheric deposition and sedimentation flux has been observed (r ¼ 0.53, p < 0.01). In any case, lakes 14, 17, 18, 19, 22, 23 and 51 have much higher sedimentation fluxes than the rest of the lakes. These fluxes were about 3.5 times higher than their corresponding estimated deposition fluxes. These lakes, except 51, are situated in the western part of the High Tatras. In fact, 18, 19, 22 and 23 are located in the same area with a maximum distance between the lakes of 1.7 km. Also 14 and 17 are next to each other at a distance of 0.5 km. One aspect to be considered concerns the effect of snow redistribution. In the mountainous terrain, snow re-suspension and deposition by strong winds or avalanches results in re-distribution and piling in the topographical depressions, where the lakes are situated. This process may have a significant influence in the lakes situated on the lee side of the range and in catchments with steep slopes, such is the case of lakes 18 (1722 m), 19 (1674 m), 22 (1716 m) and 23 (1655 m). Despite that lake waters are isolated from direct atmospheric input during winter, this additional deposition is locked up in snow and ice on the lake surface and is released at spring thaw. Then, PAH associated with this snow and ice are transferred to the lake systems which will partly increase the PAH sedimentation rates. 4. Conclusions PAH concentrations in top-core sediments from the Tatra Mountains are higher than those observed in other European high mountain regions. Despite the relative uniform composition of these compounds, there is strong geographical concentration variability among the 27 lakes studied in the area. A statistically significant correlation was observed between precipitation and PAH concentrations in sediments. A detailed account of the different atmospheric deposition and sedimentation fluxes of PAH in these lakes shows a substantial correlation between these estimated variables. However, there are several factors of variability such as snow re-distribution which may also determine specific deviations from the general trend. The observed PAH concentrations in the top-core sediments through the whole mountain range confirms the presence of high PAH loads in the High Tatras due to high regional PAH emissions, especially in winter. These regional sources are most probably situated west of the High Tatras, since the majority of lakes with high sediment PAH concentrations are located in the western part of this mountain range. Acknowledgements Technical assistance from R. Chaler and D. Fanjul is acknowledged. Financial support for this study was provided by the EU projects EUROLIMPACS (GOCE-CT-2003-505540), EMERGE (EVK1CT-1999-00032) and the Consolider-Ingenio Project GRACCIE (CSD2007-00067). References Appleby, P.G., Piliposian, G.T., 2006. Radiometric dating of sediment records from mountain lakes in the Tatra Mountains. Biologia 61, S51eS64. Carrera, G., Fernández, P., Vilanova, R.M., Grimalt, J.O., 2001. Persistent organic pollutants in snow from European high mountain areas. Atmos. Environ. 35, 245e254. Daly, G.L., Wania, F., 2004. Simulating the influence of snow on the fate of organic compounds. Environ. Sci. Technol. 38, 4176e4186. Esteve, W., Budzinski, H., Villenave, E., 2006. Relative rate constants for the heterogeneous reactions of NO2 and OH radicals with polycyclic aromatic
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hydrocarbons adsorbed on carbonaceous particles. Part 2: PAHs adsorbed on diesel particulate exhaust SRM 1650a. Atmos. Environ. 40, 201e211. Fernández, P., Grimalt, J.O., 2003. On the global distribution of persistent organic pollutants. Chimia 57, 514e521. Fernández, P., Vilanova, R.M., Grimalt, J.O., 1999. Sediment fluxes of polycyclic aromatic hydrocarbons in European high altitude mountain lakes. Environ. Sci. Technol. 33, 3716e3722. Fernández, P., Vilanova, R.M., Martínez, C., Appleby, P., Grimalt, J.O., 2000. The historical record of atmospheric pyrolytic pollution over Europe registered in the sedimentary PAH from remote mountain lakes. Environ. Sci. Technol. 34, 1906e1913. Fernández, P., Grimalt, J.O., Vilanova, R.M., 2002. Atmospheric gas-particle partitioning of polycyclic aromatic hydrocarbons in high mountain regions of Europe. Environ. Sci. Technol. 36, 1162e1168. Franz, T.P., Eisenreich, S.J., 1998. Snow scavenging of polychlorinated biphenyls and polycyclic aromatic hydrocarbons in Minnesota. Environ. Sci. Technol. 32,1771e1778. Freitag, D., Ballhorn, L., Geyer, H., Korte, F., 1985. Environmental hazard profile of organic chemicals. Chemosphere 14, 1589e1616. Grimalt, J.O., van Drooge, B.L., Ribes, A., Fernández, P., Appleby, P., 2004. Polycyclic aromatic hydrocarbon composition in soils and sediments of high altitude lakes. Environ. Pollut. 131, 13e34. Gusev, A., ozovskaya, O., Shatalov, V., 2007. Modelling POP Long-range Transport and Contamination Levels by MSCE-POP Model Draft. EMEP/MSC-E Technical Report 1/2007. Hites, R.A., Laflamme, R.E., Farrington, J.W., 1977. Sedimentary polycyclic hydrocarbons: the historical record. Science 198, 829e831. Howsam, M., Jones, K.C., 1998. Sources of PAHs in the environment. In: Neilson, A.H. (Ed.), Handbook of Environmental Chemistry. Part I. PAHs and Related Compounds, vol. 3. Springer, Berlin, pp. 137e174. Kopacek, J., Stuchlik, E., Straskrabova, V., Psenakova, P., 2000. Factors governing nutrient status of mountain lakes in the Tatra Mountains. Freshwater Biol. 43, 369e383.
Lei, Y.D., Wania, F., 2004. Is rain or snow a more efficient scavenger of organic chemicals? Atmos. Environ. 38, 3557e3571. Nied zwied z, T., 1992. Climate of the Tatra Mountains. Mountain Res. Develop. 12, 131e146. Quiroz, R., Grimalt, J.O., Fernández, P., 2010. Toxicity assessment of polycyclic aromatic hydrocarbons in sediments from European high mountain lakes. Ecotoxicol. Environ. Saf. 73, 559e564. Ribes, S., van Drooge, B., Dachs, J., Gustafsson, Ö, Grimalt, J.O., 2003. Influence of soot carbon on the soileair partitioning of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 37, 2675e2680. Rogge, W.F., Hildemann, L.M., Mazurek, M.A., Cass, G.R., 1993. Sources in fine organic aerosol. 2. Noncatalyst and catalyst equipped automobiles and heavy duty diesel trucks. Environ. Sci. Technol. 27, 636e651. Simo, R., Grimalt, J.O., Albaiges, J., 1997. Loss of unburned-fuel hydrocarbons from combustion aerosols during atmospheric transport. Environ. Sci. Technol. 31, 2697e2700. Umlauf, G., Cristoph, E.H., Eisenrreich, S.J., Mariani, G., Paradi z, B., Vives, I., 2010. Seasonality of PCDD/Fs in the ambient air of Malopolska Region, southern Poland. Environ. Sci. Pollut. Res. 17, 462e469. van Drooge, B.L., Fernández, P., Grimalt, J.O., Stuchlik, E., Torres-García, C.J., Cuevas, E., 2010. Atmospheric polycyclic aromatic hydrocarbons in remote European and Atlantic sites located above the boundary mixing layer. Environ. Sci. Pollut. Res. 17, 1207e1216. Wakeham, S.G., Schaffner, C., Giger, W., Boon, J.J., Leeuw, J.W.D., 1979. Perylene in sediments from the Namibian shelf. Geochim. Cosmochim. Acta 43, 1141e1144. Wakeham, S.G., Schaffner, C., Giger, W., 1980. Polycyclic aromatic hydrocarbons in recent lake sediments -II. Compounds derived from biogenic precursors during early diagenesis. Geochim. Cosmochim. Acta 44, 415e429. Zasadni, J., Klapyta, P., 2009. An attempt to assess the modern and the Little Ice Age climatic snowline altitude in the Tatra Mountains. Landform Anal. 10, 124e133.
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Polycyclic aromatic hydrocarbons in lake sediments from the High Tatras Barend L. van Drooge a, *, Jordi López a, Pilar Fernández a, Joan O. Grimalt a, Evzen Stuchlík b a b
Institute of Environmental Assessment and Water Research (IDÆA-CSIC), Jordi Girona 18, 08034 Barcelona, Catalonia, Spain Department of Hydrology, Charles University, Vinicná 7, 12044 Prague, Czech Republic
High sedimentary PAH loads were observed in alpine lakes in the High Tatras (Eastern Europe) which are related to high PAH atmospheric deposition fluxes.
a r t i c l e i n f o
a b s t r a c t
Article history: Received 14 October 2010 Received in revised form 19 January 2011 Accepted 25 January 2011
European alpine lake systems are used as indicators of air quality over the continent. Preliminary data showed high polycyclic aromatic hydrocarbons (PAH) loads in the High Tatras (Eastern Europe) in comparison to other mountain regions. Here, insight on the spatial distribution of PAH is provided from analysis of top-core sediments of 27 alpine lakes distributed along the High Tatras. Top-core sediment concentrations were higher than those in deep-cores, and they were higher than those observed in other European high mountain regions. The PAH profiles were uniform and comparable to those observed in aerosols and snow, indicating that atmospheric deposition was the predominant PAH input pathway to the lakes. Good agreement between estimated atmospheric deposition and sedimentation fluxes was observed. However, in several lakes in the western range higher sediment fluxes may correspond to higher PAH depositions levels. The higher concentrations may also reflect inputs from potential emission source areas. Ó 2011 Elsevier Ltd. All rights reserved.
Keywords: Polycyclic aromatic hydrocarbons High mountain lake sediments High Tatras
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) in the environment are of concern due to their mutagenic, carcinogenic and teratogenic effects on organisms (Freitag et al., 1985; Howsam and Jones, 1998). The presence of these compounds in lake sediments may be due to diagenesis of organic matter (Wakeham et al., 1980) or anthropogenic activities (Hites et al., 1977). Anthropogenic pyrolytic PAH are produced by incomplete combustion of recent or ancient biomass material, such as wood or fossil fuels (Howsam and Jones, 1998). Studies in European high mountain regions have revealed the usefulness of alpine lake ecosystems as sentinel environmental indicators of the distribution patterns of the atmospheric pollution load over the European continent (Fernández and Grimalt, 2003). Alpine lakes are different from low altitude lakes at similar latitudes since they essentially receive pollutant inputs from regional or long-range atmospheric transport. Important PAH concentration differences have been observed in lake sediment among the different European high mountain regions. Preliminary data on the High Tatras, situated between Poland and Slovakia, in Eastern Europe (49100 e49140 N; 20 00 e20100 E; Fig. 1)
* Corresponding author. E-mail address: [email protected] (B.L. van Drooge). 0269-7491/$ e see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2011.01.035
show high PAH concentrations when comparing to other European mountain areas (Fernández et al., 1999). This area is the only mountain range where the lake sediment concentrations exceed the “no-effect” level of sediment quality guidelines (Quiroz et al., 2010). These high concentrations are consistent with recent measurements of ambient air PAH levels and point to contributions of these compounds from regional emission sources (van Drooge et al., 2010). The present study is aimed to describe a spatial distribution of this atmospheric PAH load in the High Tatras and to get insight into the processes responsible for differential accumulations of these compounds in these lakes (Fig. 2). 2. Materials and methods 2.1. Selection of lakes for study in the High Tatras The 27 alpine lakes selected in the area of study (18 10 km) are shown in Fig. 1. All studied lakes are located between altitudes of 1580 m (above sea level) and 2145 m. They are situated above the regional timberline (Table 1). The summits of the steep High Tatras reach altitudes of 2655 m and are about 1800e2000 m above the surrounding lowlands. The present-day 2 C isotherm (timberline) is situated at 1550 m and 1650 m on the Northern and Southern slopes, respectively. The temperature lapse rates are 0.70 C/100 m and 0.67 C/100 m in the Northern and Southern slopes, respectively (Zasadni and Klapyta, 2009). These values were used to estimate the average annual air temperature of the studied lakes (Table 1). The climatic conditions and air circulation of the Tatras are dominated by air masses from the North Atlantic that have passed over the northwestern part of the European continent (Nied zwied z, 1992). The highest precipitation totals are recorded on the
B.L. van Drooge et al. / Environmental Pollution 159 (2011) 1234e1240
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Fig. 1. Location of lakes (Table 1). The border between Poland and Slovakian Republic is indicated by a thick dashed line. The meteorological stations are indicated by stars (KW ¼ Kasprowy Wierch; LS ¼ Lomnicky Stit; SP ¼ Skalnate Pleso).
northwestern slope of the mountain ridge, while the southern slope, lee side, is located in the precipitation shadow. In contrast to temperature, the change of precipitation with altitude is more uncertain. Nevertheless, there are an important differences between the meteorological stations of Lomnicky Stit (2653 m; 1776 mm/y) and Skalnate Pleso (1778 m; 1351 mm/y), both situated in the eastern part of the Tatras (Fig. 1). The station of Kaprowy Wierch (1991 m; 1716 mm/y), situated in the western part, shows similar precipitation rates as those of Lomnicky Stit (Zasadni and Klapyta, 2009). These values indicate an increase of precipitation with altitude as well as more precipitation in the western part of the High Tatras. They were used to estimate the average annual precipitation fluxes for the studied lakes (Table 1).
2.2. Sampling Sediment samples were collected in 2001 and taken in the deepest sites using a gravity coring system (Glew, 7.5 cm diameter, 30 cm long). This device was used to obtain the upper 0.5 cm core sediment of all lakes whereas a long sedimentary column, representing a record from present back to 1850 (Appleby and Piliposian, 2006) was collected in four cases Zadni, Vysne Wahlenbergovo, Velke Spisske and Ladove. Immediately after sampling, sediment cores were divided in sections of 0.5 cm and stored in pre-cleaned aluminum foil at 20 C until analysis. Moreover, Dlugi and Starolesnianske had been sampled with the same methodology in the past (Grimalt et al., 2004). 2.3. Analytical procedures
2001
0
5000
10000
15000
20000
1971
1941
1911 Ladove
1881
Starolesnianske Dlugi
1851
1821 Fig. 2. Time scales of the concentrations (ng/g) of analyzed in the High Tatras.
P PAH from the sediment cores
2.3.1. Materials Residue analysis n-hexane, dichloromethane, isooctane, methanol and acetone were from Merck (Darmstadt, Germany). Anhydrous sodium sulfate for analysis was also from Merck. Neutral aluminum oxide type 507C was from Fluka AG (Buchs, Switzerland). Cellulose extraction cartridges were from Whatman Ltd (Maidstone, England). Aluminum foil was rinsed with acetone and let to dry at ambient temperature prior to use. The purity of the solvents was checked by gas chromatographyemass spectrometry (GCeMS). No significant peaks should be detected for acceptance. Aluminum oxide, sodium sulfate and cellulose cartridges were cleaned by Soxhlet extraction with hexane:dichloromethane (4:1, v/v) during 24 h before use. The purity of the cleaned reagents was checked by ultrasonic extraction with n-hexane:dichloromethane (4:1; 3 20 mL), concentration to 50 mL and analysis by GCeMS. No interferences were detected. Sodium sulfate and aluminum oxide were activated overnight at 400 C and 120 C, respectively. d10-Anthracene, d12-benzo[ghi]perylene and d12-perylene were used as internal and surrogate standards. The standard mixture EPA mix 9 was used as external standard. All of them were purchased from Dr. Ehrenstoffer (Augsburg, Germany). 2.3.2. Sample extraction Sample sediment dry weight was calculated by weighing and data substraction before and after freeze-drying. About 0.5 g of wet sediment was extracted by sonication with methanol (20 mL; 20 min) in order to separate most of the interstitial water from the sediment. The subsequent extractions were performed with (2:1, v/v) dichloromethane:methanol (3 20 mL; 20 min). All extracts were combined and spiked with deuterated PAH recovery standards (d10-anthracene and d12-benzo[ghi]perylene). Then, they were vacuum evaporated to almost 10 mL and hydrolyzed overnight with 20 mL of 6% (w/w) KOH in methanol. The neutral fractions were recovered with n-hexane (3 10 mL), vacuum evaporated to almost
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Table 1 Characteristics of studied lakes in the High Tatras. Lake code
Official name
Latitude N
Longitude E
Altitude (m asl)
Lake area (ha)
Air temp ( C)
Precipitation (mm/y)
Total PAH (ng/g dw)
8 9a 11 12 13 14 15 17 18 19 21 22 23 26 27 29 30 31 32 37 42 43 45 47a 49 51 54
Zelené krivánske Dlugi Staw Ni zné Terianske Zadni Staw Polski Czarny Staw Ga˛ sienicowy Vysné Terianske Zmarzly Staw Ga˛ sienicowy Vysné Wahlenbergovo Czarny Staw Polski Ni zné Temnosmre cinské Capie Vysné Temnosmre cinské Wielki Staw Polski Malé Hincovo Velké Hincovo Czarny Staw pod Rysami Velké Zabie
49.1594 49.2273 49.1698 49.2134 49.2310 49.1680 49.2244 49.1642 49.2046 49.1929 49.1683 49.1891 49.2134 49.1740 49.1797 49.1888 49.1720 49.1663 49.1942 49.1523 49.1823 49.1788 49.1841 49.1800 49.1912 49.1841 49.1932
20.0085 20.0107 20.0143 20.0143 20.0199 20.0218 20.0238 20.0271 20.0277 20.0306 20.0378 20.0395 20.0404 20.0585 20.0606 20.0778 20.0786 20.0876 20.0943 20.1315 20.1551 20.1595 20.1629 20.1678 20.1701 20.1768 20.1964
2017 1784 1941 1890 1620 2109 1787 2145 1722 1674 2072 1716 1655 1923 1946 1580 1919 1998 1699 1879 2055 1972 2057 1986 1886 2011 2014
4.32 1.60 4.91 6.46 17.79 0.55 0.28 4.96 12.65 10.48 2.43 4.95 34.14 2.22 18.19 20.54 2.26 1.71 8.08 2.78 1.20 0.66 1.72 0.78 0.83 0.88 2.43
0.46 0.36 0.05 0.38 1.51 1.08 0.34 1.32 0.80 1.84 0.83 0.84 1.27 0.17 0.02 1.79 0.20 0.33 0.96 0.47 0.71 0.16 0.73 0.25 0.35 0.42 0.44
1468 1351 1430 1667 1532 1514 1615 1532 1583 1559 1495 1580 1549 1421 1432 1512 1419 1458 1571 1399 1487 1445 1488 1452 1402 1465 1466
6200 13,000 2300 14,000 6400 30,000 7100 24,000 23,000 19,000 6300 12,000 23,000 4400 9100 1800 2000 6800 11,000 3100 5700 8000 8800 18,000 5400 17,000 9900
a
Dra cie Vysné Zabie bielovodské Batizovské Pusté Vysné zbojnícke Ladové Starolesnianske Pleso Zabie javorové Prostredné sivé Vel’ké spisské
Those lakes were sampled and analyzed in 1993 and not included in the flux calculations.
dryness, and fractionated with a column containing 2 g of alumina which was eluted with 5 mL of n-hexane:dichloromethane (19:1 v/v) and 10 mL of dichloromethane:n-hexane (2:1 v/v). These combined fractions were vacuum evaporated to 500 mL and nitrogen concentrated to almost dryness. The extracts were redissolved in isooctane prior to instrumental analysis. 2.3.3. Instrumental analysis The following PAH were determined: fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, chrysene þ triphenylene, benzo[b þ j]fluoranthenes, benzo[k]fluoranthene, benzo[e]pyrene, benzo[a]pyrene, perylene, indeno [1,2,3-cd]pyrene, benzo[ghi]perylene and dibenz[a,h]anthracene. An internal standard of d12-perylene was added to the vials prior to injection. Samples were analyzed by GCeMS (Fisons 8000 Series, Mass Selective Detector 800 Series). A fused silica capillary column, HP-5 of 50 m and 0.25 mm i.d. (0.25 mm film thickness) was used. The oven temperature program started at 90 C (1 min hold), followed by a 4 /min ramp up to 300 C (15 min hold). Injector, transfer line and ion source temperatures were 280 C, 300 C and 200 C, respectively. Helium was used as carrier gas (1.1 mL/min). PAHs were determined in the electron impact and selected ion recording modes. The following mass fragments were used for identification and quantification: m/z 166,178, 202, 228, 252, 276, and 278 (dwell time 40 ms per single ion, ion window according to retention times of standards). Diagnostic ions of the corresponding perdeuterated standards, m/z: 188, 212, and 288, were also recorded. Identification was performed by combination of the EPA Mix 9 external standard and retention index methods. Calibration curves (detector response vs amount injected) were performed for each compound. The range of linearity of the detector was evaluated from the calibration curves. All measurements were performed in the ranges of linearity found for each compound. The quantitative data were corrected for surrogate recoveries which were 65% 14 and 79% 18 for d10-anthracene and d12-benzo[ghi]perylene, respectively. 2.3.4. Quality control Procedural blanks were performed with each set of nine samples to check for the presence of interfering peaks. All samples were blank corrected when necessary. The method detection limits based on signal to noise ratio of 3 in real samples ranged from 100 pg to 400 pg for individual compounds. The analytical procedure was successfully calibrated with a standard reference material with certified PAH values (marine sediment HS-4, Institute for Marine Biosciences, Canadian National Research Council).
3. Results and discussion 3.1. PAH concentrations The total PAH concentrations in the lake sediments are summarized in Table 1. The concentrations in the top-core
sediments were more than two orders of magnitude higher than those in the corresponding bottom-core sediments. Top-core P PAH concentrations ranged between 1800 ng/g dw and 30,000 ng/g dw. These high levels are in agreement with preliminary observations in the area (Fernández et al., 1999, 2000). Comparison of the average top-core PAH concentrations in the Tatras district with average concentrations found in other European mountain lake districts (Table 2) shows that the Tatra levels are about 3 times higher than those in Scotland lakes and 8 times higher than average lake values in the Pyrenees, Alps and Norway. These top-core sediment concentrations reflect recent inputs. The sediment cores obtained in Dlugi Staw and Starolesnianske Pleso exhibit a significant PAH increase from the beginning of the XX century. In Ladove, the PAH record increases sharply since 1950. In all three records the highest PAH appear to be associated with the industrial increase after 1970 (Fernández et al., 2000; Grimalt et al., 2004). 3.2. Profiles of PAH in sediments The average PAH distributions found in the lake sediments of the High Tatras are shown in Fig. 3. The top-core and bottom-core sediment composition was remarkably similar, which is consistent with previous observations indicating that long-range atmospheric PAH transport involves a significant degree of homogenization of the PAH mixtures (Simo et al., 1997). However, the concentrations
Table 2 PAH in alpine lake sediments from European mountain regions (Emerge Project). P Lat. Long. PAH (ng/g dw) Finland Norway Scotland High Tatras Alps Pyrenees
69 60 57 49 46 42
23 8 5 20 10 0
576 1353 3689 9771 1202 1345
B.L. van Drooge et al. / Environmental Pollution 159 (2011) 1234e1240
of some compounds such as perylene, pyrene, benz[a]anthracene and benzo[ghi]perylene follow different trends. Perylene is one of the compounds with lowest top-core relative concentration while in the bottom-core sample it is one of the predominant compounds (Fig. 3). The occurrence of relatively high perylene concentrations in the bottom sections of the sediment cores has also been observed in lake sediments from other mountain ranges (Grimalt et al., 2004) and in ancient sediment layers of freshwater and marine systems (Wakeham et al., 1979, 1980) and is related to its diagenetic formation. Pyrene, benz[a]anthracene and benzo[ghi]perylene were relatively less abundant in the bottom-core than in the top-core sediments. Nevertheless, their isomeric pairs, chrysene, fluoranthene and indeno[1,2,3-cd]pyrene, respectively, had comparable contributions in both core depths. This observation suggests that pyrene, benz[a]anthrancene and benzo[ghi]perylene may be less stable upon diagenesis than their isomeric pairs. The lower relative abundance of benzo[ghi]perylene in the bottom-core sediment than its isomeric pair, indeno[1,2,3-c,d]pyrene, may also reflect changes in emissions sources through history such as lower vehicle emissions due to fossil fuel combustion (Rogge et al., 1993) in the past. All top-core sediments show similar profiles, with good correlation between individual PAH compounds (0.84 < r2 < 0.98 (p > 0.01)) pointing to similar atmospheric qualitative PAH inputs to all these alpine lakes (Fig. 4). Comparison of the average sedimentary PAH profile with the PAH composition on atmospheric deposition over these high mountain lakes (Carrera et al., 2001) and with the atmospheric PAH gas and particle phase distribution in the High Tatras (van Drooge et al., 2010) shows a good agreement between the PAH composition in lake sediments and in the atmospheric particle phase (r2 ¼ 0.94; p < 0.01) and snow deposition (r2 ¼ 0.92; p < 0.01). Gas phase PAHs show a distinct composition (Fig. 3). Accordingly, PAH are incorporated into these lake sediments by transport associated mainly to atmospheric particles. Thus, despite the semi-volatile properties of these compounds, those of larger molecular weight, between benz[a]anthracene and benzo[ghi]perylene, are almost entirely particle bound, while the PAHs of lower molecular weight, between fluorene and pyrene, tend to shift towards the atmospheric particle phase at decreasing temperatures (Fernández et al., 2000; Lei and Wania, 2004).
0.75
1237
A
0.50
0.25
0.00 0.25
B
0.20
0.15
0.10
0.05
0.00 0.25
C
0.20
0.15
0.10
0.05
0.00 0.25
D
0.20
0.15
0.20
Fig. 4. PAH profiles in environmental compartments from the High Tatras; A: Air gas phase (van Drooge et al., 2010); B: Air particulate phase (van Drooge et al., 2010); C: Snow deposition (Carrera et al., 2001); D: Sediment (present study). FL ¼ fluorene, PHE ¼ phenanthrene, ANT ¼ anthracene, FLA ¼ fluoranthene, PYR ¼ pyrene, BANT ¼ benz[a]anthracene, CRY þ TRIP ¼ chrysene þ triphenylene, BBJFLA ¼ benzo [b þ j]fluoranthene, BKFLA ¼ benzo[k]fluoranthene, BEP ¼ benzo[e]pyrene, BAP ¼ benzo[a]pyrene, PER ¼ perylene, IP ¼ indeno[1,2,3-cd]pyrene, BGP ¼ benzo[ghi] perylene, DBA ¼ dibenzo[ah]anthracene.
0.15 0.10 0.05
DB A
IP
BG P
PE R
BA P
BE P
BB JF LA BK FL A
T
RI P
Y+ T
CH
PY R
BA N
FL A
AN T
PH E
0.00 FL
IP BG P D BA
PH
0.25
T
0.00
FL A PY R BA C N R Y+ T TR IP BF LA BE P BA P
0.30
AN
0.05
FL
0.35
E
0.10
Fig. 3. Average distribution (þSD) of individual PAH in top-core sediments (black bars) and bottom-core sediments (white bars). FL ¼ fluorene, PHE ¼ phenanthrene, ANT ¼ anthracene, FLA ¼ fluoranthene, PYR ¼ pyrene, BANT ¼ benz[a]anthracene, CRY þ TRIP ¼ chrysene þ triphenylene, BBJFLA ¼ benzo[b þ j]fluoranthene, BKFLA ¼ benzo[k]fluoranthene, BEP ¼ benzo[e]pyrene, BAP ¼ benzo[a]pyrene, PER ¼ perylene, IP ¼ indeno[1,2,3-cd]pyrene, BGP ¼ benzo[ghi]perylene, DBA ¼ dibenzo[ah]anthracene.
In the High Tatras, the highest ambient air PAH concentrations have been found in winter samples (van Drooge et al., 2010), probably as a consequence of higher PAH emissions in this season caused by higher energy demands on the European continent. Moreover, PAH are related to (soot) particles (Fernández et al., 2002), which increases their environmental persistence (Esteve et al., 2006) and affords their atmospheric transport to remote areas (Fernández et al., 2002; Ribes et al., 2003).
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Besides dry deposition, wet deposition may play an important role in PAH incorporation into high mountain lake systems. Due to the low ambient temperatures at least one third of the total precipitation to the lakes is in the form of snow, which is an efficient scavenger for vapors and particles (Franz and Eisenreich, 1998; Lei and Wania, 2004).
3.3. Spatial PAH variation The spatial distribution of the PAH concentrations in the sediment top cores of all studied lakes is shown in Fig. 5a. The lowest PAH P concentrations ( PAH ¼ 1800 ng/g dw) are observed in Czarny Staw (1580 m), which is the lake with lowest altitude situated in the northern slopes of the mountain range. The highest levels P ( PAH ¼ 30,000 ng/g dw) are found in Vysné Terianske pleso (2109 m) that is situated in the southern slopes. Comparison of the PAH sediment concentrations and physical variables, such as lake altitude, annual air temperature and precipitation, shows only a statistically significant correlation for precipitation (r ¼ 0.49; p < 0.02). To get further insight into the influence of atmospheric PAH deposition to the lakes and its burial into the sediments, mean annual deposition and sedimentation fluxes for benzo[a]pyrene (BAP) as a model-compound, were calculated for each individual lake and compared. Deposition fluxes were calculated by summing the dry deposition and the wet deposition in the form of rain (from May until October; 66% of total precipitation) and snow (November until April; 34% of total precipitation). The contribution of rain and snow was based on the long term precipitation measurements from the meteorological station of Skalnate Pleso (Slovakian Meteorological Institute), while the BAP concentrations of atmospheric particles were taken from measurements in the High Tatras between 2000 and 2001 (van Drooge et al., 2010). The dry deposition BAP flux was calculated according to Equation (1):
FluxBAP; dry deposition ¼ Vd CBAP; particle
ng=cm2 =y
(1)
where Vd was the deposition velocity assumed to be 1 m/h (Lei and Wania, 2004) and CBAP, particle was the average BAP concentration (ng/m3) in the atmospheric particle phase (van Drooge et al., 2010). The wet deposition in the form of rain was calculated according to Equation (2):
FluxBAP; rain deposition ¼ Wrain rain CBAP; particlerain
ng=cm2 =y
(2)
where Wrain was the washout factor for BAP, which was assumed to be 105 (Lei and Wania, 2004), rain was the estimated amount of rain (between May and October) to the lake corresponding to its altitude (mm/y) and CBAP, particleesnow was the average BAP concentration (ng/m3) in the atmospheric particle phase sampled in the rain season (van Drooge et al., 2010). The wet deposition in the form of snow was calculated according to Equation (3):
FluxBAP; snow deposition ¼ Wsnow snow CBAP; particlesnow
ng=cm2 =y (3)
where Wsnow was the washout factor for BAP, which was assumed to be 105 (Lei and Wania, 2004), snow was the estimated liquid amount of snow (between November and April) to the lake corresponding to its altitude (mm/y) and CBAP, particleesnow was the average BAP concentration (ng/m3) in the atmospheric particle phase sampled in the period of snowfall (van Drooge et al., 2010). The resulting estimated atmospheric deposition BAP fluxes are shown in Fig. 5b and range between 1.7 and 2.1 ng/cm2/y. The higher ambient air BAP concentrations in winter than in summer, 0.3 ng/m3 and 0.03 ng/m3, respectively (van Drooge et al., 2010) involve that about 79% of the BAP load to the lakes was during
P Fig. 5. A: Measured top-core sediment PAH concentrations (ng/g dw); B: Estimated BAP deposition flux (ng/cm2/y); C. Estimated BAP sedimentation flux (ng/cm2/y), D: Ratio sedimentation flux/deposition flux. For lake n 8 the values for the corresponding bar is given in the figures.
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snowfall (between November and April) while 16% was deposited by rain (between May and October). The remaining 5% was incorporated by dry deposition, which was assumed to be constant during the year throughout the mountain range. The relative contributions of these pathways are similar as those found in by others (Daly and Wania, 2004), which also reported snowfall as the most important deposition mode for BAP. The estimated deposition fluxes depend on the air PAH concentrations measured in Skalnate Pleso (van Drooge et al., 2010). These ambient air concentrations were high in comparison to those observed in sites from other European alpine lakes, such as those in the Alps, Pyrenees, or Norway, especially in winter (van Drooge et al., 2010). There are several urban areas situated at about 100 km of distance from the Tatras: Ostrava in the west, Krakow in the north and Kosice in the south, which may be potential PAH sources. In addition, the “Black Triangle” is situated at around 300 km to the west. This area has shown to have a large influence on the Tatras in the past (Kopacek et al., 2000). The general air circulation from the west and these potential source areas may give rise to high ambient air PAH concentrations such as those measured in Skalnate Pleso and used in the present study. These high ambient air PAH concentrations are consistent with the EMEP-model estimations of BAP emissions, where the highest continental BAP emission fluxes, higher than 20 ng/cm2/y, were estimated for a large area in southern Poland, just west of the High Tatras (Gusev et al., 2007). Moreover, a recent study (Umlauf et al., 2010) confirms the presence of very high ambient air particulate matter BAP levels (40 ng/m3) in winter in southern Poland, which was related to the practice of wood and coal combustion for domestic heating. Higher BAP air concentration may therefore occur in the western part of the High Tatras, which on its term may cause high atmospheric deposition fluxes. The estimated BAP deposition fluxes in the present study were lower, but in the same order of magnitude, as those calculated by EMEP (personal correspondence), which were estimated to be 7.4 ng/cm2/y for the High Tatras. The higher EMEP deposition loads estimated for this mountain range are probably related to the higher precipitation rates used in the EMEP-model calculation. Also, the EMEP stations were located in the lowland areas, with possible higher BAP background levels, while the one used in the present study, Skalnate Pleso, is a remote alpine site within the High Tatras. The BAP sedimentation fluxes were calculated using the average sedimentation rates in lakes from the Tatra Mountains described in Appleby and Piliposian (2006). The BAP calculation followed Equation (4):
FluxBAP; sediment ¼ Fluxsediment CBAP; sediment
ng=cm2 =y
(4)
where Fluxsediment was the sedimentation rate and CBAP, sediment was the BAP concentration (ng/g dw) measured in sediments from each individual lake. As expected, individual lakes exhibit diverse BAP sedimentation fluxes. An average sedimentation flux of 100 g/m2/y was used in the calculations due to the lack of sedimentation rates for the individual lakes. The average value was obtained from the average values of five Tatra lakes (Appleby and Piliposian, 2006) but the observed variation interval for these lakes ranged between 55 and 420 g/cm2/y, which may involve a stronger BAP sedimentation flux variability than assumed in the calculations. The BAP sedimentation fluxes obtained from the calculations described above are shown in Fig. 5d. The average value is 3.4 ng/ cm2/y which is less than two fold higher than the estimated average total atmospheric BAP deposition flux, 1.9 ng/cm2/y (Fig. 5e). The
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good agreement shows that independently of individual lake variability the average estimates provide consistent figures for comparison. Furthermore, a substantial positive correlation between calculated atmospheric deposition and sedimentation flux has been observed (r ¼ 0.53, p < 0.01). In any case, lakes 14, 17, 18, 19, 22, 23 and 51 have much higher sedimentation fluxes than the rest of the lakes. These fluxes were about 3.5 times higher than their corresponding estimated deposition fluxes. These lakes, except 51, are situated in the western part of the High Tatras. In fact, 18, 19, 22 and 23 are located in the same area with a maximum distance between the lakes of 1.7 km. Also 14 and 17 are next to each other at a distance of 0.5 km. One aspect to be considered concerns the effect of snow redistribution. In the mountainous terrain, snow re-suspension and deposition by strong winds or avalanches results in re-distribution and piling in the topographical depressions, where the lakes are situated. This process may have a significant influence in the lakes situated on the lee side of the range and in catchments with steep slopes, such is the case of lakes 18 (1722 m), 19 (1674 m), 22 (1716 m) and 23 (1655 m). Despite that lake waters are isolated from direct atmospheric input during winter, this additional deposition is locked up in snow and ice on the lake surface and is released at spring thaw. Then, PAH associated with this snow and ice are transferred to the lake systems which will partly increase the PAH sedimentation rates. 4. Conclusions PAH concentrations in top-core sediments from the Tatra Mountains are higher than those observed in other European high mountain regions. Despite the relative uniform composition of these compounds, there is strong geographical concentration variability among the 27 lakes studied in the area. A statistically significant correlation was observed between precipitation and PAH concentrations in sediments. A detailed account of the different atmospheric deposition and sedimentation fluxes of PAH in these lakes shows a substantial correlation between these estimated variables. However, there are several factors of variability such as snow re-distribution which may also determine specific deviations from the general trend. The observed PAH concentrations in the top-core sediments through the whole mountain range confirms the presence of high PAH loads in the High Tatras due to high regional PAH emissions, especially in winter. These regional sources are most probably situated west of the High Tatras, since the majority of lakes with high sediment PAH concentrations are located in the western part of this mountain range. Acknowledgements Technical assistance from R. Chaler and D. Fanjul is acknowledged. Financial support for this study was provided by the EU projects EUROLIMPACS (GOCE-CT-2003-505540), EMERGE (EVK1CT-1999-00032) and the Consolider-Ingenio Project GRACCIE (CSD2007-00067). References Appleby, P.G., Piliposian, G.T., 2006. Radiometric dating of sediment records from mountain lakes in the Tatra Mountains. Biologia 61, S51eS64. Carrera, G., Fernández, P., Vilanova, R.M., Grimalt, J.O., 2001. Persistent organic pollutants in snow from European high mountain areas. Atmos. Environ. 35, 245e254. Daly, G.L., Wania, F., 2004. Simulating the influence of snow on the fate of organic compounds. Environ. Sci. Technol. 38, 4176e4186. Esteve, W., Budzinski, H., Villenave, E., 2006. Relative rate constants for the heterogeneous reactions of NO2 and OH radicals with polycyclic aromatic
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hydrocarbons adsorbed on carbonaceous particles. Part 2: PAHs adsorbed on diesel particulate exhaust SRM 1650a. Atmos. Environ. 40, 201e211. Fernández, P., Grimalt, J.O., 2003. On the global distribution of persistent organic pollutants. Chimia 57, 514e521. Fernández, P., Vilanova, R.M., Grimalt, J.O., 1999. Sediment fluxes of polycyclic aromatic hydrocarbons in European high altitude mountain lakes. Environ. Sci. Technol. 33, 3716e3722. Fernández, P., Vilanova, R.M., Martínez, C., Appleby, P., Grimalt, J.O., 2000. The historical record of atmospheric pyrolytic pollution over Europe registered in the sedimentary PAH from remote mountain lakes. Environ. Sci. Technol. 34, 1906e1913. Fernández, P., Grimalt, J.O., Vilanova, R.M., 2002. Atmospheric gas-particle partitioning of polycyclic aromatic hydrocarbons in high mountain regions of Europe. Environ. Sci. Technol. 36, 1162e1168. Franz, T.P., Eisenreich, S.J., 1998. Snow scavenging of polychlorinated biphenyls and polycyclic aromatic hydrocarbons in Minnesota. Environ. Sci. Technol. 32,1771e1778. Freitag, D., Ballhorn, L., Geyer, H., Korte, F., 1985. Environmental hazard profile of organic chemicals. Chemosphere 14, 1589e1616. Grimalt, J.O., van Drooge, B.L., Ribes, A., Fernández, P., Appleby, P., 2004. Polycyclic aromatic hydrocarbon composition in soils and sediments of high altitude lakes. Environ. Pollut. 131, 13e34. Gusev, A., ozovskaya, O., Shatalov, V., 2007. Modelling POP Long-range Transport and Contamination Levels by MSCE-POP Model Draft. EMEP/MSC-E Technical Report 1/2007. Hites, R.A., Laflamme, R.E., Farrington, J.W., 1977. Sedimentary polycyclic hydrocarbons: the historical record. Science 198, 829e831. Howsam, M., Jones, K.C., 1998. Sources of PAHs in the environment. In: Neilson, A.H. (Ed.), Handbook of Environmental Chemistry. Part I. PAHs and Related Compounds, vol. 3. Springer, Berlin, pp. 137e174. Kopacek, J., Stuchlik, E., Straskrabova, V., Psenakova, P., 2000. Factors governing nutrient status of mountain lakes in the Tatra Mountains. Freshwater Biol. 43, 369e383.
Lei, Y.D., Wania, F., 2004. Is rain or snow a more efficient scavenger of organic chemicals? Atmos. Environ. 38, 3557e3571. Nied zwied z, T., 1992. Climate of the Tatra Mountains. Mountain Res. Develop. 12, 131e146. Quiroz, R., Grimalt, J.O., Fernández, P., 2010. Toxicity assessment of polycyclic aromatic hydrocarbons in sediments from European high mountain lakes. Ecotoxicol. Environ. Saf. 73, 559e564. Ribes, S., van Drooge, B., Dachs, J., Gustafsson, Ö, Grimalt, J.O., 2003. Influence of soot carbon on the soileair partitioning of polycyclic aromatic hydrocarbons. Environ. Sci. Technol. 37, 2675e2680. Rogge, W.F., Hildemann, L.M., Mazurek, M.A., Cass, G.R., 1993. Sources in fine organic aerosol. 2. Noncatalyst and catalyst equipped automobiles and heavy duty diesel trucks. Environ. Sci. Technol. 27, 636e651. Simo, R., Grimalt, J.O., Albaiges, J., 1997. Loss of unburned-fuel hydrocarbons from combustion aerosols during atmospheric transport. Environ. Sci. Technol. 31, 2697e2700. Umlauf, G., Cristoph, E.H., Eisenrreich, S.J., Mariani, G., Paradi z, B., Vives, I., 2010. Seasonality of PCDD/Fs in the ambient air of Malopolska Region, southern Poland. Environ. Sci. Pollut. Res. 17, 462e469. van Drooge, B.L., Fernández, P., Grimalt, J.O., Stuchlik, E., Torres-García, C.J., Cuevas, E., 2010. Atmospheric polycyclic aromatic hydrocarbons in remote European and Atlantic sites located above the boundary mixing layer. Environ. Sci. Pollut. Res. 17, 1207e1216. Wakeham, S.G., Schaffner, C., Giger, W., Boon, J.J., Leeuw, J.W.D., 1979. Perylene in sediments from the Namibian shelf. Geochim. Cosmochim. Acta 43, 1141e1144. Wakeham, S.G., Schaffner, C., Giger, W., 1980. Polycyclic aromatic hydrocarbons in recent lake sediments -II. Compounds derived from biogenic precursors during early diagenesis. Geochim. Cosmochim. Acta 44, 415e429. Zasadni, J., Klapyta, P., 2009. An attempt to assess the modern and the Little Ice Age climatic snowline altitude in the Tatra Mountains. Landform Anal. 10, 124e133.