Marine Pollution Bulletin 101 (2015) 860–866
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Organic polar pollutants in surface waters of inland seas Anna Orlikowska ⁎, Kathrin Fisch, Detlef E. Schulz-Bull Leibniz Institute for Baltic Sea Research, Warnemünde, Germany
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Article history: Received 30 March 2015 Received in revised form 30 October 2015 Accepted 6 November 2015 Available online 12 November 2015 Keywords: Black Sea Mediterranean Sea Baltic Sea Triazines UV-filters Polar pollutants
a b s t r a c t Available data about contamination by polar substances are mostly reported for rivers and near-shore waters and only limited studies exists about their occurrence in marine waters. We present concentrations and distribution of several polar pesticides and UV-filters in surface waters of three inland seas, the Baltic, Black and Mediterranean Sea. Many of the investigated compounds were below detection limits, however, those found in off-shore waters raise a concern about their persistence and possible adverse effect on the ecosystem. Despite a longstanding EU-wide ban we were able to detect atrazine in the Mediterranean and the Baltic Sea. Concentrations in the Black Sea were substantially higher. Runoff from agricultural and urban areas was the main transport route to marine ecosystems for investigated compounds, though irgarol in Mediterranean waters was attributed to intense maritime traffic. 2-Phenylbenzimidazole-5-sulfonic acid was the only UV-filter detected in marine waters, while benzophenone-4 was observed in the estuaries. Occurrence of UV-filters was seasonal. © 2015 Elsevier Ltd. All rights reserved.
Many organic substances used as pesticides, biocides or in personal care products and in various industrial products can be classified as polar pollutants. Their persistence depends on environmental conditions and is highly variable (Mackay et al., 2006; Santos et al., 2012). The combination of wide-spread use and hydrophilic character have resulted in their frequent detection in the natural waters (Jurado et al., 2014; Loos et al., 2009; Munaron et al., 2012). The main route of polar pollutants to enter surface water is via runoff from places of application (e.g. fields, urban areas) but also indirect input through sewerage effluents (Loos et al., 2013; Reemtsma et al., 2006; Rodil et al., 2008; Wittmer et al., 2010). Herbicides, biocides and some other polar contaminants pose a toxicological threat to aquatic ecosystems (EC, 2003, 2013; Giger et al., 2009; LeBaron et al., 2008; PAN, 2015; Tsui et al., 2014). For many, however, the risk associated with their occurrence in aquatic, especially marine, environment is still unknown. Inland seas, such as the Black Sea, the Baltic Sea and the Mediterranean Sea are particularly susceptible to pollution due to many anthropogenic activities within their large catchment areas. Coastal zones, especially areas around large cities, ports and estuaries, show high contamination levels, while open waters are less affected by hazardous substances (Bakan and Büyükgüngör, 2000; HELCOM, 2010; Kostianoy and Kosarev, 2008; UNEP/MAP, 2012). Data about contamination by polar substances is scarce and is mostly reported for rivers and nearshore waters (Carafa et al., 2007; Fent et al., 2010; Loos et al., 2009; Munaron et al., 2012; Nödler et al., 2014; Readman et al., 1993). This precludes reliable understanding of their distribution, fate and effect in ⁎ Corresponding author. E-mail address: [email protected] (A. Orlikowska).
http://dx.doi.org/10.1016/j.marpolbul.2015.11.018 0025-326X/© 2015 Elsevier Ltd. All rights reserved.
marine ecosystems. Therefore, the aim of our work was to estimate current concentration levels and distribution of various classes of polar organic pollutants in the Black Sea, the Baltic Sea and the Mediterranean Sea. Special interest was put on the fate of pollutants in the upper water layer of the open sea. Water samples from the Mediterranean and Black Seas were collected during cruise MSM33 with R/V Maria S. Merian in November 2013. Surface water samples from the Mediterranean were taken along cruise transects (Fig. S1) directly from the vessel's clean water supply system. In the northern and the central Black Sea (Fig. S2), discrete water samples from the surface and selected depths were taken via rosette Niskin water sampler. Water samples from the Baltic Sea (Fig. S3) were collected during three cruises in February (AL430), May (EMB69) and June (EMB76) 2014. Samples from AL430 were collected via the vessel's clean water supply system, while samples from EMB69 and EMB76 were taken via rosette Niskin water sampler. Additional samples were taken from German Baltic Sea estuaries in May and June 2014 (Fig. S3). Pre-cleaned 2 L amber glass bottles were used to collect water. Samples from MSM33 and from EMB76 were pre-concentrated by solid-phase extraction (SPE) on board the ship. Samples collected during AL430 and EMB69 were stored in the dark at 4 °C until SPE in the laboratory, no longer than two weeks after collection. Prior to extraction, samples were divided into two 1 L subsamples for repeated determination. Each sample was spiked with 1 mL of 5 ng mL− 1 internal standard mixture and the pH was adjusted to 2 with 5 M HCl (VWR, Germany). Sample enrichment was achieved on a Chromabond Easy extraction cartridge (3 mL, 200 mg, MarcheryNagel GmbH, Germany) conditioned with 4 mL acetone (Promochem, Germany) and 4 mL LC/MS-grade water (VWR, Germany). The sample
A. Orlikowska et al. / Marine Pollution Bulletin 101 (2015) 860–866
was loaded onto a cartridge via filtration unit with a glass-fiber filter (GF/F, 0.7 μm, Whatman) with a flow rate of 25 mL min−1. After sample enrichment, each cartridge was rinsed with 4 mL acidified (pH 2) LC/MS-grade water and gently dried. The extraction cartridges with the samples from MSM33 and EMB76 were wrapped in aluminum foil and stored at −20 °C until analysis. Polar substances were eluted with 4 mL acetone/methanol (v/v, 1/1) and 6 mL methanol/13% NH3(aq) (v/v, 97/3). The extract was evaporated to dryness with clean air at a 45 °C water bath (Turbo-Vap LV, Zymark, USA) and reconstituted in 1 mL methanol/water (v/v, 1/1) for analysis with liquid chromatograph-tandem mass spectrometer (LC-MS/MS) (Thermo Fisher Scientific, Germany). Chromatographic separation was performed using a reverse-phase Kinetex C-18 column (2.6 μm, 50 × 2.1 mm, Phenomenex, Germany) with a guard column. The mobile phase consisted of A: water (LC/MS-grade) with 0.1% acetic acid and B: methanol with 0.1% formic acid. A gradient elution program with the flow rate of 250 mL min−1 was used. Sample injection volume was 10 μL. The system operated in selected reaction monitoring (SRM) mode. Two characteristic fragments of the precursor molecular ion ([M + H]+ or [M − H]−) were monitored. The most abundant transition (based on peak area and signal to noise ratio) was used for quantification, whereas the second ion was used for confirmation. Each sample was measured in triplicate. The method detection limits (LOD) are given in Table 1. Instrument control, data acquisition and evaluation were performed with
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Xcalibur software (Thermo Fischer Scientific). For details of analytical method and study sites see supplementary materials. Surface waters of the Black Sea, the Mediterranean Sea as well as the Baltic Sea and its estuaries, were analyzed for polar pesticides (18 pesticide and 3 transformation products), UV-filters (11 compounds), disinfectant (triclocarban), bisphenol A, and nonylphenoxyacetic acid (Table 1, supplementary materials). Results showed that even though the investigated compounds are extensively used worldwide, many of them were below detection limits during our study (Table 1). Bisphenol A (BPA), used to manufacture plastics and resins, and triclocarban, a disinfectant added to many consumer products, were always below detection limits; probably due to rapid biodegradation of BPA in natural waters (Kang et al. (2006) and references therein) and effective removal (97–98%) of triclocarban in waste water treatment plants (WWTPs) (Halden (2014) and references therein). Nonylphenoxyacetic acid (NPE1C), a transformation product of nonylphenol polyethoxylate (NPnEO) surfactants, was detected in estuarine samples (b 1.0– 32.2 ng L−1) but not in marine waters. NPE1C was among the most abundant contaminants in European rivers (Giger et al., 2009; Loos et al., 2009; Loos et al., 2010). The use of NPnEO surfactants was restricted in many counties, including EU countries (EC, 2003; Giger et al., 2009), to reduce the environmental risk of their toxic metabolites. The low levels of NPE1C in this study could reflect implemented ban of use of NPnEO surfactants.
Table 1 Summary of analytical results of polar pollutants in water samples of Mediterranean Sea, Black Sea, Baltic Sea and estuaries of the German Baltic coast. (Frequency of detection, mean value and standard deviation calculated from samples above limit of detection (LOD), maximal concentration.) Compound
Mediterranean Sea (N = 34) LOD (ng L−1)
Freq (%)
Mean ± sd (ng L −1)
1.2 1.5 1.0 1.0 0.5 1.5 1.5
18 n.d. 3 n.d. 100 n.d. n.d.
1.4 ± 0.1
0.5 1.0 1.0 1.0 0.8 1.0 0.5 1.0 1.5 1.0 1.5 1.5 1.3 2.5
3 n.d. 3 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
UV-filters PBSA BP-1 BP-2 BP-3 BP-4 4-MBC EHMC OC OD-PABA Et-PABA 4-DHB
1.0 1.0 1.5 5.0 1.0 5.0 5.0 5.0 5.0 5.0 5.0
Others Bisphenol A Triclocarban Nonylphenoxyacetic acid
50 5.0 1.0
Triazine compounds Atrazine Simazine Terbuthylazine Terbutryn Irgarol Desisopropylatrazine Desethylatrazine Other pesticides Chloridazon CMD Chlorotoluron Diuron Isoproturon Bentazone 2,4-D MCPA Dichloroprop Mecoprop Metamitron Metribuzin Metazachlor Pendimethalin
Black Sea (N = 74)
Baltic Sea (N = 75)
Estuaries Baltic coast (N = 21)
Max (ng L −1)
Freq (%)
Mean ± sd (ng L−1)
Max (ng L−1)
Freq (%)
Mean ± sd (ng L−1)
1.5
100 100 100 n.d. 5 86 100
40.5 ± 3.1 9.4 ± 0.6 1.1 ± 0.2
50.7 10.9 1.7
1.9 ± 0.3 2.3 ± 0.4 1.1 ± 0.5
1.0 ± 0.6 1.6 ± 0.2 2.9 ± 0.4
1.8 1.8 4.2
100 95 83 n.d. n.d. n.d. n.d.
100 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
3.9 ± 0.3
4.9
100 100 8 9 13 3 100 n.d. n.d. n.d. n.d. n.d. 5 n.d.
4.3 ± 1.3 2.8 ± 1.8 1.8 ± 0.6 2.4 ± 0.4 0.9 ± 1.3 1.1 ± 0.0 1.8 ± 0.5
1.9 ± 0.4
2.5
n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
3 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
1.8 ± 0.7
24 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
1.5 ± 0.7
3.4
n.d. n.d. n.d.
n.d. n.d. n.d.
9.2 2.5 ± 0.6
3.6
1.6 3.8
2.3
n.d. n.d. n.d.
Max (ng L−1)
Freq (%)
Mean ± sd (ng L−1)
Max (ng L−1)
2.6 3.5 3.8
48 52 86 38 5 43 24
3.0 ± 2.1 3.4 ± 1.5 140 ± 255 3.0 ± 3.4 3.1 ± 0.8 1.9 ± 0.4
7.6 5.8 1111 10.5 1.9 4.6 2.5
95 100 62 67 100 100 76 95 n.d. 71 n.d. n.d. 57 n.d.
10.9 ± 27.6 8.2 ± 7.3 14.5 ± 29.3 13.2 ± 23.0 9.4 ± 12.8 15.5 ± 47.2 4.1 ± 4.7 8.2 ± 9.7
126 32.9 136 107 60.7 221 19.6 36.3
3.8 ± 2.6
9.7
6.6 ± 6.7
27.0
62 5 n.d. n.d. 48 n.d. n.d. n.d. n.d. n.d. n.d.
29.0 ± 38.9
170 2.5
63.2 ± 61.9
226
7.5 ± 7.9
32.2
7.4 8.9 2.7 2.9 6.6 1.1 3.2
n.d. n.d. 67
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The seven triazine compounds, i.e. atrazine (ATR), simazine (SIM), terbuthylazine (TBA), terbutryn, irgarol, desisopropylatrazine (DIA), and desethylatrazine (DEA), were measured. Terbutryn was not detected in all marine samples. The Black Sea showed considerably higher concentrations of ATR and SIM than the other two seas (Fig. 1), likely due to a ban on their use as plant protection product in the EU but not in some countries bordering the Black Sea (EC, 2004a, b; LeBaron et al., 2008). DIA and DEA, dealkylated transformation products of ATR and TBA, were detected only in the Black Sea waters. ATR (Fig. 2) and SIM were rather uniformly distributed in the surface water of the Black Sea. Small concentrations differences between near-shore and central areas could be explained by relatively high to moderate water solubility, low adsorption and high persistence (Mackay et al., 2006) in combination with the unique water circulation and mixing in the Black Sea, in which fast currents rapidly distribute dissolved pollutants and strong pycnocline and haloclines prevent deep vertical advection (Bakan and Büyükgüngör, 2000; Kostianoy and Kosarev, 2008). The shallow mixed surface layer is strongly influenced by a large influx of fresh water from the rivers, which are the main source of pollution in the Black Sea (Bakan and Büyükgüngör, 2000). Even though, nowadays Danube river cannot be consider as a significant source of ATR and SIM compounds (b5 ng L−1) (Loos et al., 2010) to the Black Sea, other rivers flowing through the countries where ATR and SIM are still in use still might. Unfortunately, to the best of our knowledge, there are no recent publications about polar pollutants levels in waters from this region. Vertical profiles of ATR and SIM in the oxic layer of the Black Sea (Fig. 3) revealed uniform distribution in the coastal zone. Slightly higher concentration in the surface could be related to the river run-off or an atmospheric input. In open waters, ATR and SIM slightly increased in the pycnocline, together with an increase of turbidity. This may be attributed to the affinity of ATR and SIM to colloidal organic matter, which was shown to have a high absorptive capacity for atrazine (Means and Wijayaratne, 1982). Low levels of ATR were detectable in the Baltic Sea and the Mediterranean Sea, despite its long ban, e.g. in Germany or Italy since 1990s. This implies a longer persistence in the environment than previously expected, or ongoing release from anthropogenic sources or nearcoast reservoirs. In the coastal Mediterranean, ATR and SIM concentrations (before the EU-wide ban up to 800 ng ATR L−1 and up to 440 ng SIM L−1) (Readman et al., 1993) have drastically decreased recently (Carafa et al., 2007; Nödler et al., 2013). In our study ATR was detected only in six samples from the Mediterranean Sea at similarly low level and its distribution pattern (Fig. 4) suggests that ATR could be transported to the Mediterranean Sea with inputs of the Black Sea
Fig. 1. Concentration of atrazine (a) and simazine (b) in Mediterranean Sea (N = 34), Black Sea (N = 72) and Baltic Sea (N = 75). Box represents 10–90% of data, whiskers show minimum and maximum values and line symbolize median. Gray field marks values below detection limits.
Fig. 2. Atrazine concentrations (ng L−1) in the Black Sea surface waters in November 2013.
waters. This is in agreement with Nödler et al. (2013), who observed an increasing gradient from West to East of the Northern Aegean Sea (1.6–15 ng ATR L−1) and an ATR concentration in Istanbul (38 ± 3.0 ng L−1) similar to our values from the Black Sea (40.5 ± 3.1 ng L−1). Therefore some countries in the Black Sea region where ATR is still legally used could be an important source of ATR to the Mediterranean Sea. The levels of ATR and SIM in the Baltic Sea also decreased over the years. In the south-western Baltic Sea concentrations in near coastal waters were up to 20 ng ATR L−1 and up to 30 ng SIM L−1 in the 1990s (Bester and Hühnerfuss, 1993; Graeve and Wodarg, 1996). In off-shore waters the concentrations were only slightly higher (1.8–5.1 ng ATR L− 1 and 2.4–6.1 ng SIM L−1) (Pempkowiak et al., 2000) than the values we observed in 2014. Along the German coastline ATR was below detection limits in 2009, but detectable a year later (Nödler et al., 2013) at the levels similar to those observed in this study. The fact that we were still able to detect ATR and SIM in the south-western Baltic Sea suggests their long persistence in the environment. Recent findings have demonstrated that ATR persists for decades in soil (Jablonowski et al., 2011) and that environmental half-life values of ATR (see data summarized in Mackay et al., 2006) based on short-term laboratory experiments are incorrect. Replacing ATR as active substance in plant protection products, TBA has become the key triazine herbicide in Europe (Carafa et al., 2007; LeBaron et al., 2008). TBA breaks down much more rapidly than ATR (LeBaron et al., 2008), and is therefore believed less likely to contaminate aquatic systems, particularly ground and marine waters. During this study the Baltic Sea concentrations of TBA were low (b1.0 to 3.8 ng L−1) (Fig. 5a), especially in spring. This is in agreement with previously reported data from this region (Bester and Hühnerfuss, 1993; Graeve and Wodarg, 1996; Nödler et al., 2013; Pempkowiak et al., 2000). TBA concentrations in the estuaries along the German Baltic Sea coast revealed high spatial and temporal variability (Fig. 5a). At the beginning of May 2014 levels were low (b 1.0–28.3 ng TBA L−1), while the concentrations at the end of June reached up to 1.1 μg L−1. This highly variable distribution in the estuarine waters is likely related to a seasonal agricultural application of TBA followed by its wash-out from the soil to the aquatic systems. Meteorological and hydrological conditions as well as large scale mixing processes, sorption and degradation also have an effect on temporal variations in herbicide runoff (Bester and Hühnerfuss, 1993; Carafa et al., 2007; Comoretto et al., 2007; Readman et al., 1993). Along with dilution, degradation and sorption on the particulate matter can cause seaward decreasing concentrations (Carafa et al., 2007). Concentrations of TBA were low in the Black Sea (max 1.7 ng L− 1). Likely due to low use of TAB in the region where ATR and SIM are still applied or due to degradation and sorption processes, which may be more pronounced in case of TBA than for the other triazines (LeBaron et al., 2008). In the Mediterranean Sea TBA was detected only in one coastal water sample (9.2 ng L−1) (Fig. 4), suggesting that the occurrence of TBA in this region was cause by inputs from the agricultural catchment areas. TBA was reported to be the most frequently detected herbicide in the rivers of southern Spain,
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Fig. 3. Vertical profiles of atrazine, simazine and chloridazon in the oxic layer at two Black Sea stations — deep station (a) and coastal station (b). The error bars represent the standard deviation of the measurements.
with the concentrations up to 534 ng L−1 (Robles-Molina et al., 2014) and river transport of TBA to the marine systems was also demonstrated in the Mediterranean coast (Carafa et al., 2007; Nödler et al., 2013; Readman et al., 1993). Irgarol is a highly effective antifouling agent, used in paints to inhibit primary colonization and the growth of algae on ship hulls and other underwater structures (Blanck et al., 2009; Tolosa et al., 1996). This triazine herbicide is believed to pose a risk to marine algae and cause a change in ecology of aquatic systems, therefore its use has been restricted in some European countries (Blanck et al., 2009; Bowman et al., 2003; Gatidou et al., 2007). Moreover, in 2013 irgarol was added to the Water Framework Directive (WFD) list of priority substances concerning the water quality of water bodies (EC, 2013). In this study irgarol was below maximum allowable concentration of 16 ng L−1 in
Fig. 4. The concentrations (ng L−1) and distribution of atrazine, irgarol, terbuthylazine, chloridazon and chlorotoluron in the Mediterranean Sea surface waters during the cruise MSM33. The error bars represent the standard deviation of the measurements.
all samples. In the Black Sea irgarol was mostly b 0.5 ng L−1 and quantified (0.6–1.8 ng L− 1) only in four surface samples. In the Baltic Sea it was below detection limits in all samples. This could be due to restriction in use of irgarol (Bowman et al., 2003; Gatidou et al., 2007) and relatively rapid photodegradation of irgarol in seawater (Lam et al., 2009). Nevertheless, we found irgarol in all Mediterranean Sea samples (2.5 ± 0.6 ng L−1) (Fig. 4). This could be attributed to intensive marine traffic in the region as leaching of this biocide from vessels coated with antifouling paints is its main way into aquatic systems. Moreover, due to its low affinity to particulate matter (logKoc = 3.0), irgarol stays mainly in the dissolved phase and is not easily removed from the surface with sedimentation of particulate matter (Tolosa et al., 1996). Additionally to triazines 13 other herbicides (2,4-Dichlorophenoxyacetic acid (2,4-D), 2-methyl-4-chlorophenoxyacetic acid (MCPA), dichloroprop, mecoprop, metamitron, metribuzin, chlorotoluron, diuron, isoproturon, chloridazon, metazachlor, bentazone, pendimethalin) were measured in the surface water of three inland seas. Most of the compounds were below detection limits in the Mediterranean and the Black Sea (Table 1). The sampling period in November is a possible reason for a low detection frequency. The spatial distribution of samples above detection limits confined to near-to-shore stations suggests that most sampling locations are not influenced by the land-based sources. Low concentrations of chloridazon (1.6 ng L− 1) and chlorotoluron (3.8 ng L−1) were detected in only one sample collected closest to the Spanish coast of the Mediterranean Sea (Fig. 4). Highest concentrations of widely used chloridazon (Buttiglieri et al., 2009; Cuevas et al., 2008) as well as of chlorotoluron were reported from coastal regions in the Mediterranean Sea (Carafa et al., 2007; Kuster et al., 2008; Munaron et al., 2012). Chloridazon is not readily biodegradable in water and its primary degradation is slow (half-life up to 105 days) (EFSA, 2007), therefore it can be transported over large distances in water. The main
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Fig. 5. The concentrations (ng L−1) and distribution of pesticide terbuthylazine (a) and UV-filter PBSA (b) in the Baltic Sea and the estuaries along the German coast. The error bars represent the standard deviation of the measurements.
removal pathways are a partitioning to the sediment and the transformation to desphenylchloridazon (EFSA, 2007). Vertical profiles of chloridazon in the Black Sea showed slightly different distribution in the water column than triazine compounds (Fig. 3) and suggest larger affinity of chloridazon to particulate matter, therefore, its possible removal from the water column by sedimentation. Significantly different adsorption coefficients of chloridazon (logKoc = 4.14) and atrazine (logKoc = 1.97) (PAN, 2015) could explain the diverse behaviors in the water body. Chloridazon in the Baltic Sea was at similar levels (4.3 ± 1.3 ng L−1) as in the Black Sea. This herbicide is widely applied in Germany and concentrations up to 890 ng L− 1 were reported in the inland surface waters (Buttiglieri et al., 2009). In the Baltic Sea we also detected a degradation product of chloridazon, i.e. methyldesphenylchloridazon (Me-DPC) (2.8 ± 1.8 ng L−1). Me-DPC is created under anaerobic condition in soil, from where it can leak to the water systems (Buttiglieri et al., 2009). We found the highest concentration of Me-DPC (8.9 ng L− 1) close to the mouth of the Oder River. It was the station with the lowest salinity and located closest to the shore. It was also the only site where metazachlor (1.6–2.5 ng L− 1, n = 4), bentazone (1.1 ng L− 1, n = 2) and chlorotoluron (1.4–2.7 ng L− 1, n = 6) were detected. Furthermore, the highest levels of isoproturon (up to 6.6 ng L−1) were measured in this region. Considerably higher amounts of the compounds in estuaries along the German coast (Table 1) and in various European rivers (Comoretto et al., 2007; Loos et al., 2009; Loos et al., 2010) suggest river run-off to be the main transport route of these herbicides to the Baltic Sea. Two other compounds detected in the Baltic Sea were 2,4-D and diuron. 2,4-D is a synthetic auxin analog, used as post-emergence herbicide to control annual broadleaf weeds in agricultural land but also in non-cropland areas, such as lawns, along roads and railroads (Mackay et al., 2006). Diuron, an inhibitor of photosynthesis, was a broad-
spectrum multi-use herbicide applied both in agricultural and urban areas, but also used as biocide in paints for facades, roofs and in antifouling paints for boats (Munaron et al., 2012; Singer et al., 2010; Wittmer et al., 2010). Nowadays diuron is banned as plant protection product in EU (EC, 2007) and is included in the list of priority substances concerning the water quality (maximum allowable concentration of 1.8 μg L−1) (EC, 2013). 2,4-D was the only phenoxyacid herbicide detected in marine waters (1.1–3.2 ng L−1). Diuron was detected in 9% of the Baltic Sea samples with average concentration of 2.4 (± 0.4) ng L− 1. In a study conducted in 2009 along the German coast of the Baltic Sea, diuron levels ranged from b 3.3 to 131 ng L−1 and owing to its use as algaecide in antifouling paints for ship hulls it was almost exclusively detected in harbor areas or marinas (Nödler et al., 2014). During this study, the levels of 2,4-D and diuron found in the estuaries were higher than in the Baltic Sea and the highest concentrations in marine waters were measured in the samples with the lowest salinity. Therefore also here rivers seem to be the main transport route to marine waters. The reported concentrations of diuron and 2,4-D in European rivers were up to 864 and 1221 ng L−1, respectively (Loos et al., 2009; Loos et al., 2010). We found the highest concentration of 2,4-D (19.6 ng L−1) in June in a stream receiving effluents from WWTP. The same stream contained highest concentrations of diuron, measured in February. Transport of compounds from urban sources to the surface waters mostly occurs via WWTPs or flushing with the rainfall from the sites, e.g. roads or farmyards, where pesticides were applied or spilled (Buttiglieri et al., 2009; Loos et al., 2013; Singer et al., 2010; Wittmer et al., 2010). UV-filters are used in large quantities in personal care products (e.g. sunscreens, shampoos, lotions) and industrial goods to prevent photodegradation (Fent et al., 2010; Jurado et al., 2014; Santos et al., 2012). Their extensive use promotes entrance of these compounds to aquatic environment, by direct inputs from aquatic recreational activities or indirectly via incomplete removal in WWTPs (Jurado et al., 2014; Rodil et al., 2008; Tsui et al., 2014). Most of investigated organic UV-filters (Table 1) were below detection limits in marine waters. None were found in the Mediterranean Sea. In the Black Sea only 2-phenylbenzimidazole-5-sulfonic acid (PBSA) was quantified in two surface water samples collected near the Crimean peninsula. In the sample collected in the coastal area (station depth = 62 m), close to the Kerch Strait, PBSA concentration was 1.3 ng L−1. Highest PBSA concentration (2.3 ng L−1) was detected in the north-west. Despite large water depth and distance from shore, salinity data (not shown) suggest this to be due to influence from river runoff introduced by typical local eddy circulation. PBSA was the single UV-filter detected in our investigation of levels and distribution of UV-filters in the Baltic Sea waters and its estuaries. It was found only in summer and at low concentrations (b1.0–3.4 ng L−1) (Fig. 5b). The concentrations of PBSA in inland waters were much higher (Fig. 5b) with the maximum (up to 170 ng L− 1) in the stream receiving the effluents from the WWTP. In the same sample we measured two other UV-filters, i.e. benzophenone-4 (BP-4) (77.7 ng L−1) and benzophenone-1 (BP-1) (2.5 ng L−1). BP-4 was observed in all estuarine samples in summer at concentrations ranging from 15.8 to 226 ng L− 1. PBSA was also detected in three locations during the spring campaign. Fate and behavior of UV-filters in the environment vary and depend on their chemical class, functionalities and octanol-water partitioning coefficient (Jurado et al., 2014; Negreira et al., 2012; Rodil et al., 2008). Thus, the higher polarity of PBSA and BP-4 than of other UV-filters (Santos et al., 2012), combined with incomplete removal in WWTPs (Fent et al., 2010; Rodil et al., 2008; Tsui et al., 2014), explains their sole presence in the investigated water samples. Additionally to dilution, photodegradation could influence PBSA levels in surface water but this process is pH dependent and inhibited by fulvic acid (Santos et al., 2012). In case of BP-4 minimal degradation was observed in river waters (Fent et al., 2010). BP-4 levels in the marine waters could be affected by electrophilic substitution of hydrogen per chlorine or bromine (Negreira et al., 2012). Seasonal changes of
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these compounds in aquatic systems were shown also in other regions. High concentrations of BP-4 (849 ± 48 ng L−1) were reported in July in the Mero River (NW Spain), while in January and March, similarly to our findings, BP-4 was below detection limits. BP-4 was also detected in March and July in seawater near the river mouth. While, PBSA was present in the river and seawater in July (34 ± 2 and 42 ± 3 ng L−1, respectively) but at lower levels than BP-4. Moreover, high concentrations of both compounds were found in the wastewater samples (Rodil et al., 2008). BP-4 concentrations of up to 36.6 ng L−1 were also reported in Besos River (Spain) (Jurado et al., 2014) and it was the most prevalent UV-filter in river Glatt in Switzerland (Fent et al., 2010). Although eco-toxicological data on UV-filters on the marine environment are still limited, it seems that the strongest influence could be expected in the coastal regions, especially in summer period. Even though the inland seas are susceptible to pollution and the investigated compounds are extensively used worldwide, many of them (e.g. bisphenol A) were below detection limits in the surface water samples during this study. Some other compounds, like diuron, isoproturon and irgarol were clearly below Environmental Quality Standards (EQS) values for the marine environment (0.2, 0.3 and 0.016 μg L−1, respectively) (EC, 2013). Nevertheless, the fact that we found several of polar organic pollutants in the off-shore waters raise a concern about their persistence and potential effects on the marine ecosystem. The low concentrations of atrazine and simazine in the Baltic Sea and the Mediterranean Sea are the consequence of the EU-wide ban on their use as plant protection products. While, significantly higher levels found in the Black Sea waters are most likely due to the fact that the use of these substances is not restricted in some countries bordering the Black Sea. It seems that persistence of these compounds in the environment could be longer than previously expected and that continuous anthropogenic release of these compounds could affect ecosystems located far from their sources. Acknowledgments We would like to thank Andrea Tschakste for her help with collecting samples during the cruises AL430 and EMB69. We thank Wael Skeff for logistic assistance with collecting the estuarine samples. We would like to acknowledge the authorities, especially Ukrainian and Turkish, for allowance to work in their territorial waters during the cruise MSM33 and thank IB-BMBF for funding (project no. 01DK12043). We would like to acknowledge the crew members and all participants of the cruises MSM33 and EMB76 and all other contributors to this work. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.marpolbul.2015.11.018. References Bakan, G., Büyükgüngör, H., 2000. The Black Sea. Mar. Pollut. Bull. 41, 24–43. Bester, K., Hühnerfuss, H., 1993. Triazines in the Baltic and North Sea. Mar. Pollut. Bull. 26, 423–427. Blanck, H., Eriksson, K.M., Grönvall, F., Dahl, B., Guijarro, K.M., Birgersson, G., Kylin, H., 2009. A retrospective analysis of contamination and periphyton PICT patterns for the antifoulant irgarol 1051, around a small marina on the Swedish West Coast. Mar. Pollut. Bull. 58, 230–237. Bowman, J.C., Readman, J.W., Zhou, J.L., 2003. Seasonal variability in the concentrations of irgarol 1051 in Brighton Marina, UK; including the impact of dredging. Mar. Pollut. Bull. 46, 444–451. Buttiglieri, G., Peschka, M., Frömel, T., Müller, J., Malpei, F., Seel, P., Knepper, T.P., 2009. Environmental occurrence and degradation of the herbicide n-chloridazon. Water Res. 43, 2865–2873. Carafa, R., Wollgast, J., Canuti, E., Ligthart, J., Dueri, S., Hanke, G., Eisenreich, S.J., Viaroli, P., Zaldívar, J.M., 2007. Seasonal variations of selected herbicides and related metabolites in water, sediment, seaweed and clams in the Sacca di Goro coastal lagoon (Northern Adriatic). Chemosphere 69, 1625–1637. Comoretto, L., Arfib, B., Chiron, S., 2007. Pesticides in the Rhône river delta (France): basic data for a field-based exposure assessment. Sci. Total Environ. 380, 124–132.
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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Baseline
Organic polar pollutants in surface waters of inland seas Anna Orlikowska ⁎, Kathrin Fisch, Detlef E. Schulz-Bull Leibniz Institute for Baltic Sea Research, Warnemünde, Germany
a r t i c l e
i n f o
Article history: Received 30 March 2015 Received in revised form 30 October 2015 Accepted 6 November 2015 Available online 12 November 2015 Keywords: Black Sea Mediterranean Sea Baltic Sea Triazines UV-filters Polar pollutants
a b s t r a c t Available data about contamination by polar substances are mostly reported for rivers and near-shore waters and only limited studies exists about their occurrence in marine waters. We present concentrations and distribution of several polar pesticides and UV-filters in surface waters of three inland seas, the Baltic, Black and Mediterranean Sea. Many of the investigated compounds were below detection limits, however, those found in off-shore waters raise a concern about their persistence and possible adverse effect on the ecosystem. Despite a longstanding EU-wide ban we were able to detect atrazine in the Mediterranean and the Baltic Sea. Concentrations in the Black Sea were substantially higher. Runoff from agricultural and urban areas was the main transport route to marine ecosystems for investigated compounds, though irgarol in Mediterranean waters was attributed to intense maritime traffic. 2-Phenylbenzimidazole-5-sulfonic acid was the only UV-filter detected in marine waters, while benzophenone-4 was observed in the estuaries. Occurrence of UV-filters was seasonal. © 2015 Elsevier Ltd. All rights reserved.
Many organic substances used as pesticides, biocides or in personal care products and in various industrial products can be classified as polar pollutants. Their persistence depends on environmental conditions and is highly variable (Mackay et al., 2006; Santos et al., 2012). The combination of wide-spread use and hydrophilic character have resulted in their frequent detection in the natural waters (Jurado et al., 2014; Loos et al., 2009; Munaron et al., 2012). The main route of polar pollutants to enter surface water is via runoff from places of application (e.g. fields, urban areas) but also indirect input through sewerage effluents (Loos et al., 2013; Reemtsma et al., 2006; Rodil et al., 2008; Wittmer et al., 2010). Herbicides, biocides and some other polar contaminants pose a toxicological threat to aquatic ecosystems (EC, 2003, 2013; Giger et al., 2009; LeBaron et al., 2008; PAN, 2015; Tsui et al., 2014). For many, however, the risk associated with their occurrence in aquatic, especially marine, environment is still unknown. Inland seas, such as the Black Sea, the Baltic Sea and the Mediterranean Sea are particularly susceptible to pollution due to many anthropogenic activities within their large catchment areas. Coastal zones, especially areas around large cities, ports and estuaries, show high contamination levels, while open waters are less affected by hazardous substances (Bakan and Büyükgüngör, 2000; HELCOM, 2010; Kostianoy and Kosarev, 2008; UNEP/MAP, 2012). Data about contamination by polar substances is scarce and is mostly reported for rivers and nearshore waters (Carafa et al., 2007; Fent et al., 2010; Loos et al., 2009; Munaron et al., 2012; Nödler et al., 2014; Readman et al., 1993). This precludes reliable understanding of their distribution, fate and effect in ⁎ Corresponding author. E-mail address: [email protected] (A. Orlikowska).
http://dx.doi.org/10.1016/j.marpolbul.2015.11.018 0025-326X/© 2015 Elsevier Ltd. All rights reserved.
marine ecosystems. Therefore, the aim of our work was to estimate current concentration levels and distribution of various classes of polar organic pollutants in the Black Sea, the Baltic Sea and the Mediterranean Sea. Special interest was put on the fate of pollutants in the upper water layer of the open sea. Water samples from the Mediterranean and Black Seas were collected during cruise MSM33 with R/V Maria S. Merian in November 2013. Surface water samples from the Mediterranean were taken along cruise transects (Fig. S1) directly from the vessel's clean water supply system. In the northern and the central Black Sea (Fig. S2), discrete water samples from the surface and selected depths were taken via rosette Niskin water sampler. Water samples from the Baltic Sea (Fig. S3) were collected during three cruises in February (AL430), May (EMB69) and June (EMB76) 2014. Samples from AL430 were collected via the vessel's clean water supply system, while samples from EMB69 and EMB76 were taken via rosette Niskin water sampler. Additional samples were taken from German Baltic Sea estuaries in May and June 2014 (Fig. S3). Pre-cleaned 2 L amber glass bottles were used to collect water. Samples from MSM33 and from EMB76 were pre-concentrated by solid-phase extraction (SPE) on board the ship. Samples collected during AL430 and EMB69 were stored in the dark at 4 °C until SPE in the laboratory, no longer than two weeks after collection. Prior to extraction, samples were divided into two 1 L subsamples for repeated determination. Each sample was spiked with 1 mL of 5 ng mL− 1 internal standard mixture and the pH was adjusted to 2 with 5 M HCl (VWR, Germany). Sample enrichment was achieved on a Chromabond Easy extraction cartridge (3 mL, 200 mg, MarcheryNagel GmbH, Germany) conditioned with 4 mL acetone (Promochem, Germany) and 4 mL LC/MS-grade water (VWR, Germany). The sample
A. Orlikowska et al. / Marine Pollution Bulletin 101 (2015) 860–866
was loaded onto a cartridge via filtration unit with a glass-fiber filter (GF/F, 0.7 μm, Whatman) with a flow rate of 25 mL min−1. After sample enrichment, each cartridge was rinsed with 4 mL acidified (pH 2) LC/MS-grade water and gently dried. The extraction cartridges with the samples from MSM33 and EMB76 were wrapped in aluminum foil and stored at −20 °C until analysis. Polar substances were eluted with 4 mL acetone/methanol (v/v, 1/1) and 6 mL methanol/13% NH3(aq) (v/v, 97/3). The extract was evaporated to dryness with clean air at a 45 °C water bath (Turbo-Vap LV, Zymark, USA) and reconstituted in 1 mL methanol/water (v/v, 1/1) for analysis with liquid chromatograph-tandem mass spectrometer (LC-MS/MS) (Thermo Fisher Scientific, Germany). Chromatographic separation was performed using a reverse-phase Kinetex C-18 column (2.6 μm, 50 × 2.1 mm, Phenomenex, Germany) with a guard column. The mobile phase consisted of A: water (LC/MS-grade) with 0.1% acetic acid and B: methanol with 0.1% formic acid. A gradient elution program with the flow rate of 250 mL min−1 was used. Sample injection volume was 10 μL. The system operated in selected reaction monitoring (SRM) mode. Two characteristic fragments of the precursor molecular ion ([M + H]+ or [M − H]−) were monitored. The most abundant transition (based on peak area and signal to noise ratio) was used for quantification, whereas the second ion was used for confirmation. Each sample was measured in triplicate. The method detection limits (LOD) are given in Table 1. Instrument control, data acquisition and evaluation were performed with
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Xcalibur software (Thermo Fischer Scientific). For details of analytical method and study sites see supplementary materials. Surface waters of the Black Sea, the Mediterranean Sea as well as the Baltic Sea and its estuaries, were analyzed for polar pesticides (18 pesticide and 3 transformation products), UV-filters (11 compounds), disinfectant (triclocarban), bisphenol A, and nonylphenoxyacetic acid (Table 1, supplementary materials). Results showed that even though the investigated compounds are extensively used worldwide, many of them were below detection limits during our study (Table 1). Bisphenol A (BPA), used to manufacture plastics and resins, and triclocarban, a disinfectant added to many consumer products, were always below detection limits; probably due to rapid biodegradation of BPA in natural waters (Kang et al. (2006) and references therein) and effective removal (97–98%) of triclocarban in waste water treatment plants (WWTPs) (Halden (2014) and references therein). Nonylphenoxyacetic acid (NPE1C), a transformation product of nonylphenol polyethoxylate (NPnEO) surfactants, was detected in estuarine samples (b 1.0– 32.2 ng L−1) but not in marine waters. NPE1C was among the most abundant contaminants in European rivers (Giger et al., 2009; Loos et al., 2009; Loos et al., 2010). The use of NPnEO surfactants was restricted in many counties, including EU countries (EC, 2003; Giger et al., 2009), to reduce the environmental risk of their toxic metabolites. The low levels of NPE1C in this study could reflect implemented ban of use of NPnEO surfactants.
Table 1 Summary of analytical results of polar pollutants in water samples of Mediterranean Sea, Black Sea, Baltic Sea and estuaries of the German Baltic coast. (Frequency of detection, mean value and standard deviation calculated from samples above limit of detection (LOD), maximal concentration.) Compound
Mediterranean Sea (N = 34) LOD (ng L−1)
Freq (%)
Mean ± sd (ng L −1)
1.2 1.5 1.0 1.0 0.5 1.5 1.5
18 n.d. 3 n.d. 100 n.d. n.d.
1.4 ± 0.1
0.5 1.0 1.0 1.0 0.8 1.0 0.5 1.0 1.5 1.0 1.5 1.5 1.3 2.5
3 n.d. 3 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
UV-filters PBSA BP-1 BP-2 BP-3 BP-4 4-MBC EHMC OC OD-PABA Et-PABA 4-DHB
1.0 1.0 1.5 5.0 1.0 5.0 5.0 5.0 5.0 5.0 5.0
Others Bisphenol A Triclocarban Nonylphenoxyacetic acid
50 5.0 1.0
Triazine compounds Atrazine Simazine Terbuthylazine Terbutryn Irgarol Desisopropylatrazine Desethylatrazine Other pesticides Chloridazon CMD Chlorotoluron Diuron Isoproturon Bentazone 2,4-D MCPA Dichloroprop Mecoprop Metamitron Metribuzin Metazachlor Pendimethalin
Black Sea (N = 74)
Baltic Sea (N = 75)
Estuaries Baltic coast (N = 21)
Max (ng L −1)
Freq (%)
Mean ± sd (ng L−1)
Max (ng L−1)
Freq (%)
Mean ± sd (ng L−1)
1.5
100 100 100 n.d. 5 86 100
40.5 ± 3.1 9.4 ± 0.6 1.1 ± 0.2
50.7 10.9 1.7
1.9 ± 0.3 2.3 ± 0.4 1.1 ± 0.5
1.0 ± 0.6 1.6 ± 0.2 2.9 ± 0.4
1.8 1.8 4.2
100 95 83 n.d. n.d. n.d. n.d.
100 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
3.9 ± 0.3
4.9
100 100 8 9 13 3 100 n.d. n.d. n.d. n.d. n.d. 5 n.d.
4.3 ± 1.3 2.8 ± 1.8 1.8 ± 0.6 2.4 ± 0.4 0.9 ± 1.3 1.1 ± 0.0 1.8 ± 0.5
1.9 ± 0.4
2.5
n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
3 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
1.8 ± 0.7
24 n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d. n.d.
1.5 ± 0.7
3.4
n.d. n.d. n.d.
n.d. n.d. n.d.
9.2 2.5 ± 0.6
3.6
1.6 3.8
2.3
n.d. n.d. n.d.
Max (ng L−1)
Freq (%)
Mean ± sd (ng L−1)
Max (ng L−1)
2.6 3.5 3.8
48 52 86 38 5 43 24
3.0 ± 2.1 3.4 ± 1.5 140 ± 255 3.0 ± 3.4 3.1 ± 0.8 1.9 ± 0.4
7.6 5.8 1111 10.5 1.9 4.6 2.5
95 100 62 67 100 100 76 95 n.d. 71 n.d. n.d. 57 n.d.
10.9 ± 27.6 8.2 ± 7.3 14.5 ± 29.3 13.2 ± 23.0 9.4 ± 12.8 15.5 ± 47.2 4.1 ± 4.7 8.2 ± 9.7
126 32.9 136 107 60.7 221 19.6 36.3
3.8 ± 2.6
9.7
6.6 ± 6.7
27.0
62 5 n.d. n.d. 48 n.d. n.d. n.d. n.d. n.d. n.d.
29.0 ± 38.9
170 2.5
63.2 ± 61.9
226
7.5 ± 7.9
32.2
7.4 8.9 2.7 2.9 6.6 1.1 3.2
n.d. n.d. 67
862
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The seven triazine compounds, i.e. atrazine (ATR), simazine (SIM), terbuthylazine (TBA), terbutryn, irgarol, desisopropylatrazine (DIA), and desethylatrazine (DEA), were measured. Terbutryn was not detected in all marine samples. The Black Sea showed considerably higher concentrations of ATR and SIM than the other two seas (Fig. 1), likely due to a ban on their use as plant protection product in the EU but not in some countries bordering the Black Sea (EC, 2004a, b; LeBaron et al., 2008). DIA and DEA, dealkylated transformation products of ATR and TBA, were detected only in the Black Sea waters. ATR (Fig. 2) and SIM were rather uniformly distributed in the surface water of the Black Sea. Small concentrations differences between near-shore and central areas could be explained by relatively high to moderate water solubility, low adsorption and high persistence (Mackay et al., 2006) in combination with the unique water circulation and mixing in the Black Sea, in which fast currents rapidly distribute dissolved pollutants and strong pycnocline and haloclines prevent deep vertical advection (Bakan and Büyükgüngör, 2000; Kostianoy and Kosarev, 2008). The shallow mixed surface layer is strongly influenced by a large influx of fresh water from the rivers, which are the main source of pollution in the Black Sea (Bakan and Büyükgüngör, 2000). Even though, nowadays Danube river cannot be consider as a significant source of ATR and SIM compounds (b5 ng L−1) (Loos et al., 2010) to the Black Sea, other rivers flowing through the countries where ATR and SIM are still in use still might. Unfortunately, to the best of our knowledge, there are no recent publications about polar pollutants levels in waters from this region. Vertical profiles of ATR and SIM in the oxic layer of the Black Sea (Fig. 3) revealed uniform distribution in the coastal zone. Slightly higher concentration in the surface could be related to the river run-off or an atmospheric input. In open waters, ATR and SIM slightly increased in the pycnocline, together with an increase of turbidity. This may be attributed to the affinity of ATR and SIM to colloidal organic matter, which was shown to have a high absorptive capacity for atrazine (Means and Wijayaratne, 1982). Low levels of ATR were detectable in the Baltic Sea and the Mediterranean Sea, despite its long ban, e.g. in Germany or Italy since 1990s. This implies a longer persistence in the environment than previously expected, or ongoing release from anthropogenic sources or nearcoast reservoirs. In the coastal Mediterranean, ATR and SIM concentrations (before the EU-wide ban up to 800 ng ATR L−1 and up to 440 ng SIM L−1) (Readman et al., 1993) have drastically decreased recently (Carafa et al., 2007; Nödler et al., 2013). In our study ATR was detected only in six samples from the Mediterranean Sea at similarly low level and its distribution pattern (Fig. 4) suggests that ATR could be transported to the Mediterranean Sea with inputs of the Black Sea
Fig. 1. Concentration of atrazine (a) and simazine (b) in Mediterranean Sea (N = 34), Black Sea (N = 72) and Baltic Sea (N = 75). Box represents 10–90% of data, whiskers show minimum and maximum values and line symbolize median. Gray field marks values below detection limits.
Fig. 2. Atrazine concentrations (ng L−1) in the Black Sea surface waters in November 2013.
waters. This is in agreement with Nödler et al. (2013), who observed an increasing gradient from West to East of the Northern Aegean Sea (1.6–15 ng ATR L−1) and an ATR concentration in Istanbul (38 ± 3.0 ng L−1) similar to our values from the Black Sea (40.5 ± 3.1 ng L−1). Therefore some countries in the Black Sea region where ATR is still legally used could be an important source of ATR to the Mediterranean Sea. The levels of ATR and SIM in the Baltic Sea also decreased over the years. In the south-western Baltic Sea concentrations in near coastal waters were up to 20 ng ATR L−1 and up to 30 ng SIM L−1 in the 1990s (Bester and Hühnerfuss, 1993; Graeve and Wodarg, 1996). In off-shore waters the concentrations were only slightly higher (1.8–5.1 ng ATR L− 1 and 2.4–6.1 ng SIM L−1) (Pempkowiak et al., 2000) than the values we observed in 2014. Along the German coastline ATR was below detection limits in 2009, but detectable a year later (Nödler et al., 2013) at the levels similar to those observed in this study. The fact that we were still able to detect ATR and SIM in the south-western Baltic Sea suggests their long persistence in the environment. Recent findings have demonstrated that ATR persists for decades in soil (Jablonowski et al., 2011) and that environmental half-life values of ATR (see data summarized in Mackay et al., 2006) based on short-term laboratory experiments are incorrect. Replacing ATR as active substance in plant protection products, TBA has become the key triazine herbicide in Europe (Carafa et al., 2007; LeBaron et al., 2008). TBA breaks down much more rapidly than ATR (LeBaron et al., 2008), and is therefore believed less likely to contaminate aquatic systems, particularly ground and marine waters. During this study the Baltic Sea concentrations of TBA were low (b1.0 to 3.8 ng L−1) (Fig. 5a), especially in spring. This is in agreement with previously reported data from this region (Bester and Hühnerfuss, 1993; Graeve and Wodarg, 1996; Nödler et al., 2013; Pempkowiak et al., 2000). TBA concentrations in the estuaries along the German Baltic Sea coast revealed high spatial and temporal variability (Fig. 5a). At the beginning of May 2014 levels were low (b 1.0–28.3 ng TBA L−1), while the concentrations at the end of June reached up to 1.1 μg L−1. This highly variable distribution in the estuarine waters is likely related to a seasonal agricultural application of TBA followed by its wash-out from the soil to the aquatic systems. Meteorological and hydrological conditions as well as large scale mixing processes, sorption and degradation also have an effect on temporal variations in herbicide runoff (Bester and Hühnerfuss, 1993; Carafa et al., 2007; Comoretto et al., 2007; Readman et al., 1993). Along with dilution, degradation and sorption on the particulate matter can cause seaward decreasing concentrations (Carafa et al., 2007). Concentrations of TBA were low in the Black Sea (max 1.7 ng L− 1). Likely due to low use of TAB in the region where ATR and SIM are still applied or due to degradation and sorption processes, which may be more pronounced in case of TBA than for the other triazines (LeBaron et al., 2008). In the Mediterranean Sea TBA was detected only in one coastal water sample (9.2 ng L−1) (Fig. 4), suggesting that the occurrence of TBA in this region was cause by inputs from the agricultural catchment areas. TBA was reported to be the most frequently detected herbicide in the rivers of southern Spain,
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Fig. 3. Vertical profiles of atrazine, simazine and chloridazon in the oxic layer at two Black Sea stations — deep station (a) and coastal station (b). The error bars represent the standard deviation of the measurements.
with the concentrations up to 534 ng L−1 (Robles-Molina et al., 2014) and river transport of TBA to the marine systems was also demonstrated in the Mediterranean coast (Carafa et al., 2007; Nödler et al., 2013; Readman et al., 1993). Irgarol is a highly effective antifouling agent, used in paints to inhibit primary colonization and the growth of algae on ship hulls and other underwater structures (Blanck et al., 2009; Tolosa et al., 1996). This triazine herbicide is believed to pose a risk to marine algae and cause a change in ecology of aquatic systems, therefore its use has been restricted in some European countries (Blanck et al., 2009; Bowman et al., 2003; Gatidou et al., 2007). Moreover, in 2013 irgarol was added to the Water Framework Directive (WFD) list of priority substances concerning the water quality of water bodies (EC, 2013). In this study irgarol was below maximum allowable concentration of 16 ng L−1 in
Fig. 4. The concentrations (ng L−1) and distribution of atrazine, irgarol, terbuthylazine, chloridazon and chlorotoluron in the Mediterranean Sea surface waters during the cruise MSM33. The error bars represent the standard deviation of the measurements.
all samples. In the Black Sea irgarol was mostly b 0.5 ng L−1 and quantified (0.6–1.8 ng L− 1) only in four surface samples. In the Baltic Sea it was below detection limits in all samples. This could be due to restriction in use of irgarol (Bowman et al., 2003; Gatidou et al., 2007) and relatively rapid photodegradation of irgarol in seawater (Lam et al., 2009). Nevertheless, we found irgarol in all Mediterranean Sea samples (2.5 ± 0.6 ng L−1) (Fig. 4). This could be attributed to intensive marine traffic in the region as leaching of this biocide from vessels coated with antifouling paints is its main way into aquatic systems. Moreover, due to its low affinity to particulate matter (logKoc = 3.0), irgarol stays mainly in the dissolved phase and is not easily removed from the surface with sedimentation of particulate matter (Tolosa et al., 1996). Additionally to triazines 13 other herbicides (2,4-Dichlorophenoxyacetic acid (2,4-D), 2-methyl-4-chlorophenoxyacetic acid (MCPA), dichloroprop, mecoprop, metamitron, metribuzin, chlorotoluron, diuron, isoproturon, chloridazon, metazachlor, bentazone, pendimethalin) were measured in the surface water of three inland seas. Most of the compounds were below detection limits in the Mediterranean and the Black Sea (Table 1). The sampling period in November is a possible reason for a low detection frequency. The spatial distribution of samples above detection limits confined to near-to-shore stations suggests that most sampling locations are not influenced by the land-based sources. Low concentrations of chloridazon (1.6 ng L− 1) and chlorotoluron (3.8 ng L−1) were detected in only one sample collected closest to the Spanish coast of the Mediterranean Sea (Fig. 4). Highest concentrations of widely used chloridazon (Buttiglieri et al., 2009; Cuevas et al., 2008) as well as of chlorotoluron were reported from coastal regions in the Mediterranean Sea (Carafa et al., 2007; Kuster et al., 2008; Munaron et al., 2012). Chloridazon is not readily biodegradable in water and its primary degradation is slow (half-life up to 105 days) (EFSA, 2007), therefore it can be transported over large distances in water. The main
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Fig. 5. The concentrations (ng L−1) and distribution of pesticide terbuthylazine (a) and UV-filter PBSA (b) in the Baltic Sea and the estuaries along the German coast. The error bars represent the standard deviation of the measurements.
removal pathways are a partitioning to the sediment and the transformation to desphenylchloridazon (EFSA, 2007). Vertical profiles of chloridazon in the Black Sea showed slightly different distribution in the water column than triazine compounds (Fig. 3) and suggest larger affinity of chloridazon to particulate matter, therefore, its possible removal from the water column by sedimentation. Significantly different adsorption coefficients of chloridazon (logKoc = 4.14) and atrazine (logKoc = 1.97) (PAN, 2015) could explain the diverse behaviors in the water body. Chloridazon in the Baltic Sea was at similar levels (4.3 ± 1.3 ng L−1) as in the Black Sea. This herbicide is widely applied in Germany and concentrations up to 890 ng L− 1 were reported in the inland surface waters (Buttiglieri et al., 2009). In the Baltic Sea we also detected a degradation product of chloridazon, i.e. methyldesphenylchloridazon (Me-DPC) (2.8 ± 1.8 ng L−1). Me-DPC is created under anaerobic condition in soil, from where it can leak to the water systems (Buttiglieri et al., 2009). We found the highest concentration of Me-DPC (8.9 ng L− 1) close to the mouth of the Oder River. It was the station with the lowest salinity and located closest to the shore. It was also the only site where metazachlor (1.6–2.5 ng L− 1, n = 4), bentazone (1.1 ng L− 1, n = 2) and chlorotoluron (1.4–2.7 ng L− 1, n = 6) were detected. Furthermore, the highest levels of isoproturon (up to 6.6 ng L−1) were measured in this region. Considerably higher amounts of the compounds in estuaries along the German coast (Table 1) and in various European rivers (Comoretto et al., 2007; Loos et al., 2009; Loos et al., 2010) suggest river run-off to be the main transport route of these herbicides to the Baltic Sea. Two other compounds detected in the Baltic Sea were 2,4-D and diuron. 2,4-D is a synthetic auxin analog, used as post-emergence herbicide to control annual broadleaf weeds in agricultural land but also in non-cropland areas, such as lawns, along roads and railroads (Mackay et al., 2006). Diuron, an inhibitor of photosynthesis, was a broad-
spectrum multi-use herbicide applied both in agricultural and urban areas, but also used as biocide in paints for facades, roofs and in antifouling paints for boats (Munaron et al., 2012; Singer et al., 2010; Wittmer et al., 2010). Nowadays diuron is banned as plant protection product in EU (EC, 2007) and is included in the list of priority substances concerning the water quality (maximum allowable concentration of 1.8 μg L−1) (EC, 2013). 2,4-D was the only phenoxyacid herbicide detected in marine waters (1.1–3.2 ng L−1). Diuron was detected in 9% of the Baltic Sea samples with average concentration of 2.4 (± 0.4) ng L− 1. In a study conducted in 2009 along the German coast of the Baltic Sea, diuron levels ranged from b 3.3 to 131 ng L−1 and owing to its use as algaecide in antifouling paints for ship hulls it was almost exclusively detected in harbor areas or marinas (Nödler et al., 2014). During this study, the levels of 2,4-D and diuron found in the estuaries were higher than in the Baltic Sea and the highest concentrations in marine waters were measured in the samples with the lowest salinity. Therefore also here rivers seem to be the main transport route to marine waters. The reported concentrations of diuron and 2,4-D in European rivers were up to 864 and 1221 ng L−1, respectively (Loos et al., 2009; Loos et al., 2010). We found the highest concentration of 2,4-D (19.6 ng L−1) in June in a stream receiving effluents from WWTP. The same stream contained highest concentrations of diuron, measured in February. Transport of compounds from urban sources to the surface waters mostly occurs via WWTPs or flushing with the rainfall from the sites, e.g. roads or farmyards, where pesticides were applied or spilled (Buttiglieri et al., 2009; Loos et al., 2013; Singer et al., 2010; Wittmer et al., 2010). UV-filters are used in large quantities in personal care products (e.g. sunscreens, shampoos, lotions) and industrial goods to prevent photodegradation (Fent et al., 2010; Jurado et al., 2014; Santos et al., 2012). Their extensive use promotes entrance of these compounds to aquatic environment, by direct inputs from aquatic recreational activities or indirectly via incomplete removal in WWTPs (Jurado et al., 2014; Rodil et al., 2008; Tsui et al., 2014). Most of investigated organic UV-filters (Table 1) were below detection limits in marine waters. None were found in the Mediterranean Sea. In the Black Sea only 2-phenylbenzimidazole-5-sulfonic acid (PBSA) was quantified in two surface water samples collected near the Crimean peninsula. In the sample collected in the coastal area (station depth = 62 m), close to the Kerch Strait, PBSA concentration was 1.3 ng L−1. Highest PBSA concentration (2.3 ng L−1) was detected in the north-west. Despite large water depth and distance from shore, salinity data (not shown) suggest this to be due to influence from river runoff introduced by typical local eddy circulation. PBSA was the single UV-filter detected in our investigation of levels and distribution of UV-filters in the Baltic Sea waters and its estuaries. It was found only in summer and at low concentrations (b1.0–3.4 ng L−1) (Fig. 5b). The concentrations of PBSA in inland waters were much higher (Fig. 5b) with the maximum (up to 170 ng L− 1) in the stream receiving the effluents from the WWTP. In the same sample we measured two other UV-filters, i.e. benzophenone-4 (BP-4) (77.7 ng L−1) and benzophenone-1 (BP-1) (2.5 ng L−1). BP-4 was observed in all estuarine samples in summer at concentrations ranging from 15.8 to 226 ng L− 1. PBSA was also detected in three locations during the spring campaign. Fate and behavior of UV-filters in the environment vary and depend on their chemical class, functionalities and octanol-water partitioning coefficient (Jurado et al., 2014; Negreira et al., 2012; Rodil et al., 2008). Thus, the higher polarity of PBSA and BP-4 than of other UV-filters (Santos et al., 2012), combined with incomplete removal in WWTPs (Fent et al., 2010; Rodil et al., 2008; Tsui et al., 2014), explains their sole presence in the investigated water samples. Additionally to dilution, photodegradation could influence PBSA levels in surface water but this process is pH dependent and inhibited by fulvic acid (Santos et al., 2012). In case of BP-4 minimal degradation was observed in river waters (Fent et al., 2010). BP-4 levels in the marine waters could be affected by electrophilic substitution of hydrogen per chlorine or bromine (Negreira et al., 2012). Seasonal changes of
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these compounds in aquatic systems were shown also in other regions. High concentrations of BP-4 (849 ± 48 ng L−1) were reported in July in the Mero River (NW Spain), while in January and March, similarly to our findings, BP-4 was below detection limits. BP-4 was also detected in March and July in seawater near the river mouth. While, PBSA was present in the river and seawater in July (34 ± 2 and 42 ± 3 ng L−1, respectively) but at lower levels than BP-4. Moreover, high concentrations of both compounds were found in the wastewater samples (Rodil et al., 2008). BP-4 concentrations of up to 36.6 ng L−1 were also reported in Besos River (Spain) (Jurado et al., 2014) and it was the most prevalent UV-filter in river Glatt in Switzerland (Fent et al., 2010). Although eco-toxicological data on UV-filters on the marine environment are still limited, it seems that the strongest influence could be expected in the coastal regions, especially in summer period. Even though the inland seas are susceptible to pollution and the investigated compounds are extensively used worldwide, many of them (e.g. bisphenol A) were below detection limits in the surface water samples during this study. Some other compounds, like diuron, isoproturon and irgarol were clearly below Environmental Quality Standards (EQS) values for the marine environment (0.2, 0.3 and 0.016 μg L−1, respectively) (EC, 2013). Nevertheless, the fact that we found several of polar organic pollutants in the off-shore waters raise a concern about their persistence and potential effects on the marine ecosystem. The low concentrations of atrazine and simazine in the Baltic Sea and the Mediterranean Sea are the consequence of the EU-wide ban on their use as plant protection products. While, significantly higher levels found in the Black Sea waters are most likely due to the fact that the use of these substances is not restricted in some countries bordering the Black Sea. It seems that persistence of these compounds in the environment could be longer than previously expected and that continuous anthropogenic release of these compounds could affect ecosystems located far from their sources. Acknowledgments We would like to thank Andrea Tschakste for her help with collecting samples during the cruises AL430 and EMB69. We thank Wael Skeff for logistic assistance with collecting the estuarine samples. We would like to acknowledge the authorities, especially Ukrainian and Turkish, for allowance to work in their territorial waters during the cruise MSM33 and thank IB-BMBF for funding (project no. 01DK12043). We would like to acknowledge the crew members and all participants of the cruises MSM33 and EMB76 and all other contributors to this work. Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.marpolbul.2015.11.018. References Bakan, G., Büyükgüngör, H., 2000. The Black Sea. Mar. Pollut. Bull. 41, 24–43. Bester, K., Hühnerfuss, H., 1993. Triazines in the Baltic and North Sea. Mar. Pollut. Bull. 26, 423–427. Blanck, H., Eriksson, K.M., Grönvall, F., Dahl, B., Guijarro, K.M., Birgersson, G., Kylin, H., 2009. A retrospective analysis of contamination and periphyton PICT patterns for the antifoulant irgarol 1051, around a small marina on the Swedish West Coast. Mar. Pollut. Bull. 58, 230–237. Bowman, J.C., Readman, J.W., Zhou, J.L., 2003. Seasonal variability in the concentrations of irgarol 1051 in Brighton Marina, UK; including the impact of dredging. Mar. Pollut. Bull. 46, 444–451. Buttiglieri, G., Peschka, M., Frömel, T., Müller, J., Malpei, F., Seel, P., Knepper, T.P., 2009. Environmental occurrence and degradation of the herbicide n-chloridazon. Water Res. 43, 2865–2873. Carafa, R., Wollgast, J., Canuti, E., Ligthart, J., Dueri, S., Hanke, G., Eisenreich, S.J., Viaroli, P., Zaldívar, J.M., 2007. Seasonal variations of selected herbicides and related metabolites in water, sediment, seaweed and clams in the Sacca di Goro coastal lagoon (Northern Adriatic). Chemosphere 69, 1625–1637. Comoretto, L., Arfib, B., Chiron, S., 2007. Pesticides in the Rhône river delta (France): basic data for a field-based exposure assessment. Sci. Total Environ. 380, 124–132.
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