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Release of PAHs and heavy metals in coastal environments linked to leisure boats
Marine Pollution Bulletin 127 (2018) 664–671
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Release of PAHs and heavy metals in coastal environments linked to leisure boats
T
⁎
Jenny Egardta, , Martin Mørk Larsenb, Pia Lassenc, Ingela Dahllöfa a
Department of Biological and Environmental Sciences, University of Gothenburg, Carl Skottsbergsgata 22B, 413 19 Göteborg, Sweden Department of Bioscience – Marine Diversity and Experimental Ecology, Aarhus University, Frederiksborgvej 399, 4000 Roskilde, Denmark c Department of Environmental Science – Environmental Chemistry and Toxicology, Aarhus University, Frederiksborgvej 399, 4000 Roskilde, Denmark b
A R T I C L E I N F O
A B S T R A C T
Keywords: PAH Heavy metals Leisure boat WFD EQS
Leisure boats are responsible for elevated levels of heavy metals and PAHs in sediments in- and near marinas and natural harbours. As these compounds are released directly into the water column they also pose a threat to organisms in the pelagic environment. Passive samplers were deployed during peak and post tourist season in the water column of natural harbours, leisure boat waterways and small marinas to measure the dissolved fraction of PAHs and metal ions. Differences between seasons indicative of leisure boat activities were found as PAH composition differed between peak and post season for natural harbours and waterways, where heavier PAHs increased during peak season. During peak season, metal samplers were covered by biofouling, which likely affected the uptake. Post season metal concentrations differ between locations, with concentrations exceeding quality standards at near mainland locations where boats are maintained, compared to the sites in the archipelago.
1. Introduction Coastal areas around the world are used for recreation and private leisure boats are a common feature in these areas. In Sweden, 14% of households own at least one boat and it is estimated that the total number of leisure boats is around 500,000, excluding rowing boats and kayaks (STA, 2015). Leisure boats are associated with emissions of hazardous substances, such as Polycyclic Aromatic Hydrocarbons (PAHs) from exhaust fumes and uncombusted fuel, and heavy metals from antifouling paints that are mainly emitted when boats are moving, but also leach from hulls when boats are moored, albeit at a lower rate (Valkirs et al., 2003). Previous studies have shown a correlation between leisure boats and elevated levels of antifouling paint components and PAHs at heavily boated sites and small marinas (Boyle et al., 2016; Eklund et al., 2009). We have also shown that leisure boats are a likely source of banned antifouling substances like TBT (tributyltin) found in surface sediments of natural harbours (Egardt et al., 2017). The aim of this study was to investigate whether leisure boats also contribute to elevated water concentrations of PAHs, Cu and Zn, and if so, what parts of the coastal areas that are most at risk. Concentration limits for sediment and water have been set within the EU for hazardous substances that are deemed to be a risk to the
⁎
aquatic environment, defined in the EUs Water Framework Directive (WFD) as Environmental Quality Standards (EQS). Compounds that are of extra concern are known as priority substances, (Directive 2008/ 105/EC, Annex II) which include five individual PAHs, with both an Annual Average (AA-EQS) and a maximum allowable concentration (MAC-EQS) that should not be exceeded (Directive 2000/60/EC; PAH5-6-rings EQS dossier, 2011; Directive 2013/39/EU). Legislation exist both on regional level within the WFD, and on national level in the member states. For other hazardous substances such as Cu and Zn, AAEQS and MAC-EQS are defined on national level, and in Sweden these are set by the Agency for Marine and Water Management (HVMFS 2015:4, HVMFS 2013:19). PAHs are hydrophobic compounds, made up of two or more fused benzene rings (Cerniglia, 1992). They can be pyrogenic or petrogenic in origin, where pyrogenic are formed by incomplete combustion of organic matter and fossil fuels, and petrogenic are natural components of coal and crude oil. Outboard engines release their exhaust fumes below the surface and thereby discharge PAHs directly into the water, but fuel is also release uncombusted. About one third of the boats in Sweden have engines that are of the older two-stroke model (> 25 years old), (STA, 2015) which have the highest release of uncombusted fuel. Even newer two-stroke engines wash out 20% of the fuel uncombusted. The four-stroke engines
Corresponding author. E-mail address: [email protected] (J. Egardt).
https://doi.org/10.1016/j.marpolbul.2017.12.060 Received 9 August 2017; Received in revised form 21 December 2017; Accepted 22 December 2017 0025-326X/ © 2017 Elsevier Ltd. All rights reserved.
Marine Pollution Bulletin 127 (2018) 664–671
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The concentration of the metals in the DGT was calculated according to Zhang and Davison (1995), in short, the metal accumulated in the gel layer (M) was calculated from
release ten times less of the uncombusted fuel compared to the newer two-strokes and are from that view-point, considered more environmentally friendly (Alin and Astnäs, 2001). Copper and zinc are common ingredients in antifouling paints where Cu is the active biocide and Zn is used for its polishing properties (Thomas and Brooks, 2010; KemI, 2017). Both Cu and Zn are essential metals, but can be toxic to organisms at different concentrations depending on the taxa. In more saline, oceanic waters, higher concentrations of Cu are generally needed to prevent fouling as Cu toxicity decreases with increased salinity (Kwok and Leung, 2005), although recent studies have shown there is no difference in amount of fouling organisms when using paints with 13- or 34 weight-% Cu (CHANGE, 2014). Dissolved metal ions in the water column are generally considered more hazardous compared to complex bound metals, since they have the potential to directly interact with molecular binding sites in an organism. Therefore, several models have been developed to assess toxicity and bioavailability of metals. The biotic ligand model (BLM) stresses the need to consider the presence of other metal ions competing for the same uptake mechanisms and binding sites, but also the amount of organic and inorganic complexes that bind to ions as this will affect the metals bioavailability and hence its toxic effect (Di Toro et al., 2001). Here we used passive samplers to measure dissolved concentrations of metal ions and PAHs in the water column in three different types of coastal areas: natural harbours, leisure boat waterways and small marinas in the Kosterhavet Marine National Park on the Swedish west coast. The samplers were deployed during peak and post boating season to establish, 1) how much of the compounds were present during the different seasons at the different location types, 2) if there were a difference between or within seasons that could be explained by boat presence at the sampled locations.
Ce (VHNO3 + Vgel )
M=
(1)
fe
where VHNO3 is the volume of HNO3 (10 mL), Vgel is the volume of the Chelex-100 gel (0.15) mL and fe is the elution factor (0.8 used as default for 1 M HNO3) and Ce is the measured concentration in the 10 mL HNO3 solution. From this concentration, the concentration of metals in the water phase was calculated from the temperature dependent diffusion coefficient of the individual metals tabulated by Zhang & Davison:
CDGT =
M∆g DtA
(2)
where Δ g is the thickness of the diffusive gel (0.08 cm) and the filter membrane (0.014 cm), D is the diffusion coefficient, t is deployment time and A is exposure area (3.14 cm2). Performance reference compounds (PRC, Table 1.) were added to the SR membranes prior to deployment by soaking them in a water/ methanol solution (40/60%) together with the PRCs. The amount of water was slowly increased to 85% over 48 h on a shaker to force the PRC compounds into the membrane. SR membranes were deployed together with the DGTs and transported in cold storage to the lab after retrieval. Membranes were then added labelled internal PAH standards for recovery calculations (deuterated PAHs), and then extracted in 50 mL methanol on a shaker for 12 h. The extracts were evaporated to 1 mL, spiked with injection standards (C13 and deuterated PAHs) and analysed for the labelled spiked injection standards and PRCs together with 22 unlabelled PAHs on GC–MS (Gas Chromatography-Mass Spectrometry), including the 16 EPA PAHs and dibenzothiophene, 2-mehtylphenantrene, 3,6 dimethylphenanthrene, benz(a)fluorine, 1-methylpyrene, benz(e)pyrene and perylene. Benz(b + j + k)fluoranthen and chrysene/triphenylene were measured as sums due to overlapping chromatography. PAHs were calculated based on (Booij and Smedes, 2010; Smedes and Booij, 2012). The fraction of PRCs retained in the silicone sheet was calculated according to:
2. Methods Passive samplers were deployed during the peak season (July 2015) and post season (April–May 2016) to collect metal ions and PAHs. Locations included two small marinas (Mar), four natural harbours (NH) and four leisure boat waterways (WW), where all natural harbours and two of the waterways (WW 1, 4) are located in the archipelago and two waterways (WW 2, 3) are close to the mainland (Fig. 1). Diffusive gradient in thin film (DGT) samplers for metal ions, mainly Cu and Zn, were placed at all ten locations during both seasons. Silicone rubber (SR) membranes for PAHs were deployed at five locations in the archipelago during peak season (2 membranes per location) and at nine locations during post season (1 membrane per location) (Fig. 1). The post season deployments were carried out in two steps, half of the samplers were deployed in April at the archipelago locations, and the other half were deployed in May in the marinas and in one of the near mainland waterways. One of the archipelago locations, the natural harbour NH 3 was sampled both in April and May to also catch variation at the start of the season. SR and DGT samplers were attached without a canister to the same line, secured by anchor and buoy, at between 1.0 and 1.5 m depth, together with a temperature logger (HOBO Pendant® Temp/Light, 64 K). Deployment time was set to 12 (peak season) and 14 days (post season) to ensure reasonable detection limits for the SR membranes. Transport blanks were used for both DGT and SR membranes. DGTs were purchased from DGT Lancaster, UK, ready for deployment. After collection, samplers were transported in cold storage to the laboratory, where they were dismantled. The Chelex-100 resin was extracted in 15 mL centrifuge tubes (polyethylene) with 2 mL of 1 M nitric acid that covered the resin according to Zhang and Davison (1995). Extraction was performed overnight on a shaker. After extraction, the Chelex-100 resin was removed and 8 mL of 18 MΩ MilliQ was added before analysis on ICP-MS (Inductively Coupled Plasma-Mass Spectrometry).
f=e
−Bt
Kpw mM 0.47
(3)
where t is the time of deployment (in days), m is the mass of the sampler (0,0049 ± 0,0002 kg), M is the molar mass of the PAH measured and Kpw is the distribution constant between silicone and water from literature. B is a proportional constant modelled from the dissipation data of the PRCs for each PAH. D10-Anthracene was not detected in most deployed samples and thus could not be used. Estimated Bs varied from 2,9 to 17,7 (average 10,0). The estimated B was used to calculate Rs (equivalent water sampling rate, l d− 1) (4) (Smedes and Booij, 2012):
Rs = B M 0.47
(4)
Rs for Naphthalene were 0,3 to 1,8 (average (1,0) and for Dibenzo (ah)anthracene) 0,2 to 1,3 (average (0,7). The concentration in water Cw was estimated by (5) (Smedes and Booij, 2012):
Cw =
Nt
(
−Bt
Kpw m 1 − e
Kpw mM0.47
)
(5)
Concentrations below limit of detection were treated as missing data. Comparisons were made between and within the two seasons for both PAHs and metal ions. Post season comparison was made between archipelago and near mainland locations including marinas. For peak season PAHs, the sampled locations were separated into waterways and natural harbours, whereas metal samples were separated into three 665
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Fig. 1. Map of sampled sites, adapted from Egardt et al., 2017. White circle = post season PAH, black circle = peak season PAH, white triangle = post season metal ions, black triangle = peak season metal ions.
statistically significant (ANOVA, p > 0.05). However, we decided to remove naphthalene, as the temperature difference between seasons were 6–12 degrees (see supplementary information, Appendix A, Table 3), which increases the volatility substantially for this compound during peak season compared to the other PAHs (Alaee et al., 1996). Removing naphthalene changed the pattern between seasons (Fig. 2b), as the concentrations in natural harbours NH 1–3 during peak season then exceeds post season. All locations sampled at both seasons, except NH 1 at peak season, also decrease markedly in total concentration. Of the remaining four locations only sampled at post season, NH 4 through Mar 2, Mar 2 is the only one that shows a large reduction in PAH concentration when naphthalene was removed. Mar 2 is subject to regular year-round boat traffic as the ferry stop is located here. The composition of total PAHs differed significantly between post and peak seasons (ANOSIM, p = 0.008, R = 1) (Fig. 3). This dissimilarity was mainly due to a reduced contribution of 2ringed PAHs during peak season, and to fluorene, anthracene and 1methylpyrene which were only found during peak season. Furthermore, the heavy PAHs (5-6 rings) were only found during peak season, albeit only at one location (Fig. 4a). It is worth noting that there were generally more types of PAHs detected during peak season (9–17 types) than post season (9–11 types). When grouping PAHs by number of rings, all three of the natural harbours sampled during both seasons contained proportionally more of the 2-ringed PAHs at post season (NH 1–3), and the heavier 5–6 ringed PAHs were only found at one location during peak season (Fig. 4a–b). There is also a clear difference between post season composition in PAHs between April and the start of the boating season in May for location NH 3 (Fig. 4b). There was no significant difference in PAH composition between waterways (WW) and natural harbours (NH) during peak season (Fig. 5a). However, during post season there was a significant difference between near mainland and marina locations (grey) and archipelago (black) (ANOSIM, p = 0.011, R = 0.864). The difference was due to the high relative concentration of PAHs at Mar 2 (Fig. 2a–b) but also the difference in composition (Fig. 4b.) at the near mainland and
Table 1 Performance reference compounds (PRC) added to SR membranes prior to deployment. PRC Antracene D10 Acenapthene D10 Fluoranthene D10 Benzo(ghi)perylene D12 Benzo(k)fluoranthene D12
locations types: marinas, waterways and natural harbours. Descriptive comparisons were made using non-metric multi-dimensional scaling (nMDS) based on Bray-Curtis as dissimilarity measure between locations and differences were analysed with analysis of similarities (ANOSIM) in the software, PAST (University of Oslo, Norway). Differences in total concentrations between and within seasons were made with one-way ANOVA. (Microsoft Excel, 2016). Thirty-eight interviews were conducted with tourists during the summer 2015 to find out what types of antifouling paints, and subsequently amount of active ingredient, that were most frequently used (see supplementary information, Appendix A). Moored boats were counted during 11 days over five weeks in July and August where boat types and sizes were noted and an estimated hull area was calculated using specifications for the different boat types, such as length at water level, beam, depth and keel type (reactiveresins.com/boat-bottom-calculator.html). Hull area (m2) combined with known leaching rates (μg cm−2 day− 1) from three commercial paints (KemI, 2017) generated an assessment on the potential for Cu pollution to the natural harbours. 3. Results Total PAH concentrations were higher during post season (Fig. 2a), but the difference in total PAH concentration between seasons, and within location types within post and peak seasons, were not 666
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Fig. 2. Total concentrations of sampled PAH per location for both seasons. a.) All sampled PAHs. b.) Sampled PAHs minus naphthalene. Error bars (SD) are added for locations with double membranes; all locations during peak season but only NH 3 during post season. NH 3 was pooled within post season (April–May) as PAH concentrations at this location when naphthalene was removed were very similar, and naphthalene was not detected at this location in May.
There was no significant difference between different location types during peak season (Fig. 8a), when all heavy metals were taken into account, but comparison during post season between near mainlandmarina and archipelago locations did show a significant difference (R = 0.44, p = 0.033). Near mainland and marina sites are areas where boats are maintained and launched generally display higher concentrations than the archipelago (Fig. 6b). NH 4 is located close to a commercial ferry and has a small private marina nearby, which may be why this location did not resemble the other natural harbours, but instead grouped together with near mainland waterways. Five out of 38 visitors interviewed knew the exact paint type they had used, and two of these were banned paints. Twenty-seven of the participants used legal paints, with Cu as an active ingredient, but only three knew the amount of the active substance. When asked why they chose their specific paint, 38% said it was because it's the paint they had always used, 21% claimed it was the most efficient and only 5% stated they had chosen the paint for environmental reasons. The four natural harbours were subject to very different leaching potential as the number of boats and boat sizes varied depending on location. Copper did not vary substantially during peak season between the natural harbours, but Zn did (Fig. 6a). Zn concentration during peak season ranked NH 2 < NH 3 < NH 1 < NH 4, which reflects the leach scenario for Cu (Table 2). This pattern was different during post season where NH 1 < NH 3 < NH 2 < NH 4 for Zn.
marina locations. The concentration range in the archipelago was 0.96–2.39 ng PAH/L and at the near mainland and marina location it was 1.2–4.73 ng/L, so the total concentrations in the archipelago was lower, although not significantly. Concentrations of metals were substantially higher during post season (Fig. 6). This was most likely due to presence of heavy biofouling on the DGTs at peak season that can reduce uptake markedly as sorption of metals occur on the fouling, which decreases the concentration gradient and in turn affects accumulation in the DGT (Devillers et al., 2017). Copper was uniformly distributed during peak season and well below AA-EQS (Fig. 6a), but Zn varied between all location types. During post season, WW 1, 4 and NH 1–3 had similar concentrations for Cu, with Zn being more variable, which correspond to results from post season distribution (Fig. 8b). NH 4 is located in the vicinity of the ferry stop, which may account for the elevated Cu and Zn concentrations (Fig. 6b). AA-EQS for both Cu and Zn was exceeded at this location. Both of the near mainland waterway (WW) locations had elevated concentrations of Cu as well as Zn. These locations have houses with waterfront access nearby and boats may therefore be painted and launched there. The marina locations had low Cu concentrations, which may be due to more particles that bind Cu-ions, but these locations were high in Zn and exceeded AA-EQS. There was a significant difference between seasons (p = 0.0005, R = 0.407) for the composition of all heavy metals (Fig. 7). It is mainly the lower overall concentrations during peak season due to the heavy biofouling on the DGTs (Fig. 6a–b) that account for the difference.
Fig. 3. Post and peak season PAH composition for natural harbours (NH) and waterways (WW). White = peak season, black = post season. Naphthalene not included. Post season concentrations for location NH 3 have been pooled in this analysis.
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Fig. 4. Composition of 2, 3, 4 and 5-6-ringed PAHs at the different locations. a.) peak season, b.) post season, m = near mainland location. Dashed line delimits post season samples between WW1-NH3 (April) which are the same locations sampled during peak season. NH 3 (May)Mar 2 were sampled in May and only during post season. Naphthalene was not excluded in this comparison.
4. Discussion
concentrations resemble those sampled in Danish coastal areas and off shore values for the north Atlantic (Nizzetto et al., 2008). However, benzo(g,h,i)perylene has a MAC-EQS of 0,82 ng/L (HVMFS 2015:4) and the measured value at NH 1 during peak season was 0,69 ng/L. The MAC-EQS for PAHs are set for the concentration in an entire water sample, and the SR membranes only take up the dissolved fraction. As PAHs are likely to be bound to dissolved organic matter or particles, it is reasonable to believe that when including all fractions in the water column, the EQS may very well be exceeded. The concentrations of Cu exceeded AA-EQS at one location during post season and Zn exceeded its AA-EQS at five locations (Fig. 6b). However, potential leaching load (Table 2) show that a substantial amount of Cu is likely to be released from boat hulls which affirms that DGT at peak season were uptake-inhibited due to biofouling. Effects of fouling on uptake of some metal elements, like Cu, Pb and Cd, have been shown as metal sorption occur on the fouling (Devillers et al., 2017) and as algae also utilize metals like Cu this adds to the reduced uptake and real concentrations are therefore likely to exceed post season values. This is problematic since AA-EQS are then likely to be exceeded for both Cu and Zn for several locations, which could have serious localized effects on the biota in the area. EQSs for metals are set for the fraction smaller than 45 μm, and therefore contain more than just the free ion form, which is collected by the DGTs. Macro algae have been proven to accumulate not only free Cu2 + but also organically complexed Cu (Ytreberg et al., 2011) and hence the effect of sampled concentrations of Cu and Zn in this study may be underestimated. The EQSs are set for single compounds and do not take into account interactions that may be additive, synergistic or antagonistic. Copper, for example, exhibit synergistic effects with the biocide irgarol but
There was a significant difference in composition of both PAHs and metal ions between the sampled seasons (Figs. 3, 7), indicating that leisure boat activities influence the contamination profile in the water mass. Natural Harbour 3 can also exemplify this, where PAH composition changed between April and May 2016, where the May sampling occurred after a period of good weather at the start of the boating season. However, within season, it was only during post season that differences in composition of both compounds types between the location types were detected, where near mainland and marina locations differed from archipelago locations. This suggests that during summer when boats are trafficking the entire area everything is everywhere and no distinction between locations can be made. Total concentrations of the dissolved bioavailable fraction of both PAHs and metal ions were higher during post season, which is a somewhat counterintuitive result. However, PAHs are hydrophobic and will be adsorbed onto particulates or dissolved organic matter (Smol and Włodarczyk-Makuła, 2017, references therein). During peak season, many of the shallow locations will be subject to sediment resuspension due to boating activities (propellers, anchors) and PAHs may therefore be less available for uptake in dissolved form. The relative reduction of 2-ringed PAHs due to higher water temperatures also contributes to lower bioavailable concentrations. Experimental artefacts for the metals can also be linked to adsorption to particles, but most likely, the much lower concentrations during peak season were due to biofouling on the DGT samplers that prevents uptake. None of the individual PAHs exceeded their EQS-values and
Fig. 5. PAH composition during a.) Peak season, separated into waterway (WW, white) and natural harbour (NH, black). b.) post season, separated into near mainland and marina locations (WW 3, Mar, grey) and archipelago locations (NH, WW 1,4 black), NH 3 post season concentrations have been pooled.
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Fig. 6. Cu and Zn concentrations from DGT membranes. a.) Peak season concentrations. a.) Post season concentrations. Dashed lines are AAEQS for Cu (4.3 μg/L) and Zn (3.4 μg/L), m = near mainland.
how well they cope with chemical pollution. Not exceeding EQSs for single compounds therefore does not necessarily protect the environment or assure good ecological status. In this study, we found metal ions and PAHs present close to, or above their MAC- and AA-EQSs in natural harbours of a marine national park. We therefore cannot rule out that the compounds alone or their combined effect exert pressure on the environment. Furthermore, the impact of leisure boats may not affect the pelagic community the most, but rather benthic communities and filter feeding organisms that are exposed both to the truly dissolved as well as the particulate fractions.
when mixed with another biocide, diuron, the effects are antagonistic (Gatidou et al., 2007). The single compound EQS also fails to consider the added stress in organisms of being exposed to several different stressors at the same time. While looking into experiments with the aim of studying interactions, Crain et al. (2008) found that synergy was the most common outcome with two interacting compounds, but interaction type (antagonistic, synergistic) varied depending on, for instance, trophic level and the combined stressors. The addition of a third stressor, however doubled the amount of synergistic interactions. In a study on periphyton, when combining five non-similarly acting compounds present at their individual NOEC (no observed effect concentration) levels, there was a clear effect of the mixture (Backhaus et al., 2011). Another study tested mixtures of up to 19 compounds at their individual AA-EQSs and found both acute and chronic effects in toxicity tests on different organism groups (Carvalho et al., 2014). The marine environment is exposed to several chemical compounds simultaneously. When monitoring chemicals along the Swedish west coast, 30 of the 172 analysed compounds, which included PAHs and biocides, were detected at the chosen reference site (Gustavsson et al., 2017). The fact is that organisms everywhere are constantly exposed to several different compounds, but are also stressed by other abiotic factors like changes in temperature, pH and oxygen, which will affect
Acknowledgements The study was funded by Alderman and Mrs. Ernst Colliander's Foundation, Wilhelm and Martina Lundgren Science Foundation, (2015-0706) the foundation Birgit och Birger Wåhlströms Minnesfond för den bohuslänska havs- och insjömiljön, Herbert och Karin Jacobssons Foundation (7/v16) and the Royal and Hvitfeldska Foundation which are kindly acknowledged. We would also like to acknowledge Amandine Vuylsteke for help in the field, the staff at Sven Lovén Centre for Marine Infrastructure, Tjärnö for help during fieldwork and the staff at the department of Bioscience and the department of Environmental Science, Aarhus University; Charlotte Dahl Schiødt Fig. 7. Post (black)- and peak (white) season metal ion composition. Cu, Zn, Pb, Cr and Cd were included in the analysis, but Cu and Zn were contributing most to the pattern. NH 3 concentrations for post season have been pooled.
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Fig. 8. Metal ion composition within seasons. a.) peak season; marina (grey), natural harbour (black) and waterway (white). b.) post season; near mainland waterways (WW 2–3) and marinas (grey) and archipelago, natural harbour (NH 1–4) and waterway (WW 1,4) (black) NH 3 post season concentrations have been pooled.
Biodegradation 3, 351–368. CHANGE, 2014. http://changeantifouling.com/wp-content/uploads/2014/10/Sstad.pdf. Crain, C.M., Kroeker, K., Halpern, B.S., 2008. Interactive and cumulative effects of multiple human stressors in marine systems. Ecol. Lett. 11, 1304–1315. Devillers, D., Buzier, R., Grybos, M., Charriau, A., Guibaud, G., 2017. Key role of the sorption process in alteration of metal and metalloid quantification by fouling development on DGT passive samplers. Environ. Pollut. 230, 523–529. Di Toro, D.M., Allen, H.E., Bergman, H.L., Meyer, J.S., Paquin, P.R., Santore, R.C., 2001. Biotic ligand model of the acute toxicity of metals. 1. Technical basis. Environ. Toxicol. Chem. 20, 2383–2396. Directive 2000/60/EC Of the European parliament and of the council of 23 October 2000, Establishing a Framework for Community Action in the Field of Water Policy. Directive 2008/105/EC Of the European parliament and of the council of 16 December 2016, on environmental quality standards in the field of water policy, Amending and Subsequently Repealing Council Directives 82/176/EEC, 83/513/EEC, 84/156/EEC, 84/491/EEC, 86/280/EEC and Amending Directive 2000/60/EC of the European Parliament and of the Council. Directive 2013/39/EU Of the European parliament and of the council of 12 August 2013 Amending Directives 2000/60/EC and 2008/105/EC as Regards Priority Substances in the Field of Water Policy. Egardt, J., Dahllöf, I., Nilsson, P., 2017. Sediment indicate the continued use of banned antifouling compounds. Mar. Pollut. Bull. 125, 282–288. Eklund, B., Elfström, M., Gallego, I., Bengtsson, B.E., Breitholtz, M., 2009. Biological and chemical characterization of harbour sediments from the Stockholm area. J. Soils Sediments 10, 127–141. Gatidou, G., Thomaidis, N.S., Zhou, J.L., 2007. Fate of Irgarol 1051, diuron and their main metabolites in two UK marine systems after restrictions in antifouling paints. Environ. Int. 33, 70–77. Gustavsson, B.M., Magnér, J., Carney Almroth, B., Eriksson, M.K., Sturve, J., Backhaus, T., 2017. Chemical monitoring of Swedish coastal waters indicates common exceedances of environmental thresholds, both for individual substances as well as their mixtures. Mar. Pollut. Bull. 122, 409–419. HVMFS, 2013. Havs- och vattenmyndighetens författningssamling. Havs och vattenmyndighetens föreskrifter om ändring i Havs- och vattenmyndighetens föreskrifter (HVMFS 2013:19) om klassificering och miljökvalitetsnormer avseende ytvatten. Swedish Agency for Marine and Water Management. HVMFS, 2015. Havs- och vattenmyndighetens föreskrifter om ändring i Havs- och vattenmyndighetens föreskrifter (HVMFS 2013:19) om klassificering och miljökvalitetsnormer avseende ytvatten. Swedish Agency for Marine and Water Management. KemI, 2017. Swedish Chemicals Agency. Kwok, K.W.H., Leung, K.M.Y., 2005. Toxicity of antifouling biocides to the intertidal harpacticoid copepod Tigriopus japonicus (Crustacea, Copepoda): effects of temperature and salinity. Mar. Pollut. Bull. 51, 830–837. Nizzetto, L., Lohmann, R., Gioia, R., Jahnke, A., Temme, C., Dachs, J., Herckes, P., Di Guardo, A., Jones, K.C., 2008. PAHs in air and seawater along a North-South Atlantic transect: trends, processes and possible sources. Environ. Sci. Technol. 42, 1580–1585. Smedes, F., Booij, K., 2012. Guidelines for passive sampling of hydrophobic contaminants in water using silicone rubber samplers. In: ICES Techniques in Marine Environmental Sciences No. 52, (20 pp). Smol, M., Włodarczyk-Makuła, M., 2017. The effectiveness in the removal of PAHs from aqueous solutions in physical and chemical processes: a review. Polycycl. Aromat. Compd. 37, 292–313. STA, 2015. Båtlivsundersökningen 2015. Swedish Transport Agency, Norrköping. Thomas, K.V., Brooks, S., 2010. The environmental fate and effects of antifouling paint biocides. Biofouling 26, 73–88. Valkirs, A.O., Seligman, P.F., Haslbeck, E., Caso, J.S., 2003. Measurement of copper
Table 2 Potential leaching pressure (g Cu location− 1 day− 1) for three different commercial paints to the natural harbours (NH) based on average boat counts and boat sizes at sampled locations during peak season. Leaching rates for paints used were: Low (Hempel light), 0.97 μg cm−2 day− 1, Medium (Biltema west coast =), 2.37 μg cm−2 day− 1, High (Hempel Mille Extra), 3.5 μg cm−2 day− 1. (Leaching rates with permission from KemI). Medium and High leaching paints are known to be used in the area.
NH NH NH NH
1 2 3 4
Low
Medium
High
0.6 0.3 0.8 3.0
1.4 0.8 1.9 7.2
2.1 1.2 2.8 10.6
and Ellen Christiansen for performing the PAH analysis and Gitte Jacobsen for assistance with the metal analysis. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.marpolbul.2017.12.060. References PAH-5-6-rings EQS dossier, 2011. Prepared by the sub-group on review of the priority substances list (Under Working Group E of the Common Implementation Strategy for the Water Framework Directive). In: 5-6 RINGS POLYAROMATIC HYDROCARBONS (PAH), PAH-5-6-rings EQS dossier. Alaee, M., Whittal, R.M., Strachan, W.M.J., 1996. The effect of water temperature and composition on Henry's Law Constant for various PAH's. Chemosphere 32, 1153–1164. Alin, J., Astnäs, T., 2001. Jämförande studie av utombordsmotorers emissioner till vatten. Avdelningen för kemisk miljövetenskap, Chalmers Tekniska Högskola, Göteborg. Backhaus, T., Porsbring, T., Arrhenius, A., Brosche, S., Johansson, P., Blanck, H., 2011. Single-substance and mixture toxicity of five pharmaceuticals and personal care products to marine periphyton communities. Environ. Toxicol. Chem. 30, 2030–2040. Booij, K., Smedes, F., 2010. An improved method for estimating in situ sampling rates of nonpolar passive samplers. Environ. Sci. Technol. 44, 6789–6794. Boyle, J.F., Sayer, C.D., Hoare, D., Bennion, H., Heppel, K., Lambert, S.J., Appleby, P.G., Rose, N.L., Davy, A.J., 2016. Toxic metal enrichment and boating intensity: sediment records of antifoulant copper in shallow lakes of eastern England. J. Paleolimnol. 55, 195–208. Carvalho, R.N., Arukwe, A., Ait-Aissa, S., Bado-Nilles, A., Balzamo, S., Baun, A., Belkin, S., Blaha, L., Brion, F., Conti, D., Creusot, N., Essig, Y., Ferrero, V.E.V., Flander-Putrle, V.V., Fürhacker, M., Grillari-Voglauer, R., Hogstrand, C., Jonáš, A., Kharlyngdoh, J.B., Loos, R., Lundebye, A.K., Modig, C., Olsson, P.E., Pillai, S., Polak, N., Potalivo, M., Sanchez, W., Schifferli, A., Schirmer, K., Sforzini, S., Stürzenbaum, S.R., Søfteland, L., Turk, V., Viarengo, A., Werner, I., Yagur-Kroll, S.S., Zounkova, R., Lettieri, T., 2014. Mixtures of chemical pollutants at European legislation safety concentrations: how safe are they? Toxicol. Sci. 141, 218–233. Cerniglia, C.E., 1992. Biodegradation of polycyclic aromatic hydrocarbons.
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ceramium tenuicorne: a test of the free ion activity model. Environ. Sci. Technol. 45, 3145–3153. Zhang, H., Davison, W., 1995. Performance characteristics of diffusion gradients in thin films for the in situ measurement of trace metals in aqueous solution. Anal. Chem. 67, 3391–3400.
release rates from antifouling paint under laboratory and in situ conditions: implications for loading estimation to marine water bodies. Mar. Pollut. Bull. 46, 763–779. Ytreberg, E., Karlsson, J., Hoppe, S., Eklund, B., Ndungu, K., 2011. Effect of organic complexation on copper accumulation and toxicity to the estuarine red macroalga
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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul
Release of PAHs and heavy metals in coastal environments linked to leisure boats
T
⁎
Jenny Egardta, , Martin Mørk Larsenb, Pia Lassenc, Ingela Dahllöfa a
Department of Biological and Environmental Sciences, University of Gothenburg, Carl Skottsbergsgata 22B, 413 19 Göteborg, Sweden Department of Bioscience – Marine Diversity and Experimental Ecology, Aarhus University, Frederiksborgvej 399, 4000 Roskilde, Denmark c Department of Environmental Science – Environmental Chemistry and Toxicology, Aarhus University, Frederiksborgvej 399, 4000 Roskilde, Denmark b
A R T I C L E I N F O
A B S T R A C T
Keywords: PAH Heavy metals Leisure boat WFD EQS
Leisure boats are responsible for elevated levels of heavy metals and PAHs in sediments in- and near marinas and natural harbours. As these compounds are released directly into the water column they also pose a threat to organisms in the pelagic environment. Passive samplers were deployed during peak and post tourist season in the water column of natural harbours, leisure boat waterways and small marinas to measure the dissolved fraction of PAHs and metal ions. Differences between seasons indicative of leisure boat activities were found as PAH composition differed between peak and post season for natural harbours and waterways, where heavier PAHs increased during peak season. During peak season, metal samplers were covered by biofouling, which likely affected the uptake. Post season metal concentrations differ between locations, with concentrations exceeding quality standards at near mainland locations where boats are maintained, compared to the sites in the archipelago.
1. Introduction Coastal areas around the world are used for recreation and private leisure boats are a common feature in these areas. In Sweden, 14% of households own at least one boat and it is estimated that the total number of leisure boats is around 500,000, excluding rowing boats and kayaks (STA, 2015). Leisure boats are associated with emissions of hazardous substances, such as Polycyclic Aromatic Hydrocarbons (PAHs) from exhaust fumes and uncombusted fuel, and heavy metals from antifouling paints that are mainly emitted when boats are moving, but also leach from hulls when boats are moored, albeit at a lower rate (Valkirs et al., 2003). Previous studies have shown a correlation between leisure boats and elevated levels of antifouling paint components and PAHs at heavily boated sites and small marinas (Boyle et al., 2016; Eklund et al., 2009). We have also shown that leisure boats are a likely source of banned antifouling substances like TBT (tributyltin) found in surface sediments of natural harbours (Egardt et al., 2017). The aim of this study was to investigate whether leisure boats also contribute to elevated water concentrations of PAHs, Cu and Zn, and if so, what parts of the coastal areas that are most at risk. Concentration limits for sediment and water have been set within the EU for hazardous substances that are deemed to be a risk to the
⁎
aquatic environment, defined in the EUs Water Framework Directive (WFD) as Environmental Quality Standards (EQS). Compounds that are of extra concern are known as priority substances, (Directive 2008/ 105/EC, Annex II) which include five individual PAHs, with both an Annual Average (AA-EQS) and a maximum allowable concentration (MAC-EQS) that should not be exceeded (Directive 2000/60/EC; PAH5-6-rings EQS dossier, 2011; Directive 2013/39/EU). Legislation exist both on regional level within the WFD, and on national level in the member states. For other hazardous substances such as Cu and Zn, AAEQS and MAC-EQS are defined on national level, and in Sweden these are set by the Agency for Marine and Water Management (HVMFS 2015:4, HVMFS 2013:19). PAHs are hydrophobic compounds, made up of two or more fused benzene rings (Cerniglia, 1992). They can be pyrogenic or petrogenic in origin, where pyrogenic are formed by incomplete combustion of organic matter and fossil fuels, and petrogenic are natural components of coal and crude oil. Outboard engines release their exhaust fumes below the surface and thereby discharge PAHs directly into the water, but fuel is also release uncombusted. About one third of the boats in Sweden have engines that are of the older two-stroke model (> 25 years old), (STA, 2015) which have the highest release of uncombusted fuel. Even newer two-stroke engines wash out 20% of the fuel uncombusted. The four-stroke engines
Corresponding author. E-mail address: [email protected] (J. Egardt).
https://doi.org/10.1016/j.marpolbul.2017.12.060 Received 9 August 2017; Received in revised form 21 December 2017; Accepted 22 December 2017 0025-326X/ © 2017 Elsevier Ltd. All rights reserved.
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The concentration of the metals in the DGT was calculated according to Zhang and Davison (1995), in short, the metal accumulated in the gel layer (M) was calculated from
release ten times less of the uncombusted fuel compared to the newer two-strokes and are from that view-point, considered more environmentally friendly (Alin and Astnäs, 2001). Copper and zinc are common ingredients in antifouling paints where Cu is the active biocide and Zn is used for its polishing properties (Thomas and Brooks, 2010; KemI, 2017). Both Cu and Zn are essential metals, but can be toxic to organisms at different concentrations depending on the taxa. In more saline, oceanic waters, higher concentrations of Cu are generally needed to prevent fouling as Cu toxicity decreases with increased salinity (Kwok and Leung, 2005), although recent studies have shown there is no difference in amount of fouling organisms when using paints with 13- or 34 weight-% Cu (CHANGE, 2014). Dissolved metal ions in the water column are generally considered more hazardous compared to complex bound metals, since they have the potential to directly interact with molecular binding sites in an organism. Therefore, several models have been developed to assess toxicity and bioavailability of metals. The biotic ligand model (BLM) stresses the need to consider the presence of other metal ions competing for the same uptake mechanisms and binding sites, but also the amount of organic and inorganic complexes that bind to ions as this will affect the metals bioavailability and hence its toxic effect (Di Toro et al., 2001). Here we used passive samplers to measure dissolved concentrations of metal ions and PAHs in the water column in three different types of coastal areas: natural harbours, leisure boat waterways and small marinas in the Kosterhavet Marine National Park on the Swedish west coast. The samplers were deployed during peak and post boating season to establish, 1) how much of the compounds were present during the different seasons at the different location types, 2) if there were a difference between or within seasons that could be explained by boat presence at the sampled locations.
Ce (VHNO3 + Vgel )
M=
(1)
fe
where VHNO3 is the volume of HNO3 (10 mL), Vgel is the volume of the Chelex-100 gel (0.15) mL and fe is the elution factor (0.8 used as default for 1 M HNO3) and Ce is the measured concentration in the 10 mL HNO3 solution. From this concentration, the concentration of metals in the water phase was calculated from the temperature dependent diffusion coefficient of the individual metals tabulated by Zhang & Davison:
CDGT =
M∆g DtA
(2)
where Δ g is the thickness of the diffusive gel (0.08 cm) and the filter membrane (0.014 cm), D is the diffusion coefficient, t is deployment time and A is exposure area (3.14 cm2). Performance reference compounds (PRC, Table 1.) were added to the SR membranes prior to deployment by soaking them in a water/ methanol solution (40/60%) together with the PRCs. The amount of water was slowly increased to 85% over 48 h on a shaker to force the PRC compounds into the membrane. SR membranes were deployed together with the DGTs and transported in cold storage to the lab after retrieval. Membranes were then added labelled internal PAH standards for recovery calculations (deuterated PAHs), and then extracted in 50 mL methanol on a shaker for 12 h. The extracts were evaporated to 1 mL, spiked with injection standards (C13 and deuterated PAHs) and analysed for the labelled spiked injection standards and PRCs together with 22 unlabelled PAHs on GC–MS (Gas Chromatography-Mass Spectrometry), including the 16 EPA PAHs and dibenzothiophene, 2-mehtylphenantrene, 3,6 dimethylphenanthrene, benz(a)fluorine, 1-methylpyrene, benz(e)pyrene and perylene. Benz(b + j + k)fluoranthen and chrysene/triphenylene were measured as sums due to overlapping chromatography. PAHs were calculated based on (Booij and Smedes, 2010; Smedes and Booij, 2012). The fraction of PRCs retained in the silicone sheet was calculated according to:
2. Methods Passive samplers were deployed during the peak season (July 2015) and post season (April–May 2016) to collect metal ions and PAHs. Locations included two small marinas (Mar), four natural harbours (NH) and four leisure boat waterways (WW), where all natural harbours and two of the waterways (WW 1, 4) are located in the archipelago and two waterways (WW 2, 3) are close to the mainland (Fig. 1). Diffusive gradient in thin film (DGT) samplers for metal ions, mainly Cu and Zn, were placed at all ten locations during both seasons. Silicone rubber (SR) membranes for PAHs were deployed at five locations in the archipelago during peak season (2 membranes per location) and at nine locations during post season (1 membrane per location) (Fig. 1). The post season deployments were carried out in two steps, half of the samplers were deployed in April at the archipelago locations, and the other half were deployed in May in the marinas and in one of the near mainland waterways. One of the archipelago locations, the natural harbour NH 3 was sampled both in April and May to also catch variation at the start of the season. SR and DGT samplers were attached without a canister to the same line, secured by anchor and buoy, at between 1.0 and 1.5 m depth, together with a temperature logger (HOBO Pendant® Temp/Light, 64 K). Deployment time was set to 12 (peak season) and 14 days (post season) to ensure reasonable detection limits for the SR membranes. Transport blanks were used for both DGT and SR membranes. DGTs were purchased from DGT Lancaster, UK, ready for deployment. After collection, samplers were transported in cold storage to the laboratory, where they were dismantled. The Chelex-100 resin was extracted in 15 mL centrifuge tubes (polyethylene) with 2 mL of 1 M nitric acid that covered the resin according to Zhang and Davison (1995). Extraction was performed overnight on a shaker. After extraction, the Chelex-100 resin was removed and 8 mL of 18 MΩ MilliQ was added before analysis on ICP-MS (Inductively Coupled Plasma-Mass Spectrometry).
f=e
−Bt
Kpw mM 0.47
(3)
where t is the time of deployment (in days), m is the mass of the sampler (0,0049 ± 0,0002 kg), M is the molar mass of the PAH measured and Kpw is the distribution constant between silicone and water from literature. B is a proportional constant modelled from the dissipation data of the PRCs for each PAH. D10-Anthracene was not detected in most deployed samples and thus could not be used. Estimated Bs varied from 2,9 to 17,7 (average 10,0). The estimated B was used to calculate Rs (equivalent water sampling rate, l d− 1) (4) (Smedes and Booij, 2012):
Rs = B M 0.47
(4)
Rs for Naphthalene were 0,3 to 1,8 (average (1,0) and for Dibenzo (ah)anthracene) 0,2 to 1,3 (average (0,7). The concentration in water Cw was estimated by (5) (Smedes and Booij, 2012):
Cw =
Nt
(
−Bt
Kpw m 1 − e
Kpw mM0.47
)
(5)
Concentrations below limit of detection were treated as missing data. Comparisons were made between and within the two seasons for both PAHs and metal ions. Post season comparison was made between archipelago and near mainland locations including marinas. For peak season PAHs, the sampled locations were separated into waterways and natural harbours, whereas metal samples were separated into three 665
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Fig. 1. Map of sampled sites, adapted from Egardt et al., 2017. White circle = post season PAH, black circle = peak season PAH, white triangle = post season metal ions, black triangle = peak season metal ions.
statistically significant (ANOVA, p > 0.05). However, we decided to remove naphthalene, as the temperature difference between seasons were 6–12 degrees (see supplementary information, Appendix A, Table 3), which increases the volatility substantially for this compound during peak season compared to the other PAHs (Alaee et al., 1996). Removing naphthalene changed the pattern between seasons (Fig. 2b), as the concentrations in natural harbours NH 1–3 during peak season then exceeds post season. All locations sampled at both seasons, except NH 1 at peak season, also decrease markedly in total concentration. Of the remaining four locations only sampled at post season, NH 4 through Mar 2, Mar 2 is the only one that shows a large reduction in PAH concentration when naphthalene was removed. Mar 2 is subject to regular year-round boat traffic as the ferry stop is located here. The composition of total PAHs differed significantly between post and peak seasons (ANOSIM, p = 0.008, R = 1) (Fig. 3). This dissimilarity was mainly due to a reduced contribution of 2ringed PAHs during peak season, and to fluorene, anthracene and 1methylpyrene which were only found during peak season. Furthermore, the heavy PAHs (5-6 rings) were only found during peak season, albeit only at one location (Fig. 4a). It is worth noting that there were generally more types of PAHs detected during peak season (9–17 types) than post season (9–11 types). When grouping PAHs by number of rings, all three of the natural harbours sampled during both seasons contained proportionally more of the 2-ringed PAHs at post season (NH 1–3), and the heavier 5–6 ringed PAHs were only found at one location during peak season (Fig. 4a–b). There is also a clear difference between post season composition in PAHs between April and the start of the boating season in May for location NH 3 (Fig. 4b). There was no significant difference in PAH composition between waterways (WW) and natural harbours (NH) during peak season (Fig. 5a). However, during post season there was a significant difference between near mainland and marina locations (grey) and archipelago (black) (ANOSIM, p = 0.011, R = 0.864). The difference was due to the high relative concentration of PAHs at Mar 2 (Fig. 2a–b) but also the difference in composition (Fig. 4b.) at the near mainland and
Table 1 Performance reference compounds (PRC) added to SR membranes prior to deployment. PRC Antracene D10 Acenapthene D10 Fluoranthene D10 Benzo(ghi)perylene D12 Benzo(k)fluoranthene D12
locations types: marinas, waterways and natural harbours. Descriptive comparisons were made using non-metric multi-dimensional scaling (nMDS) based on Bray-Curtis as dissimilarity measure between locations and differences were analysed with analysis of similarities (ANOSIM) in the software, PAST (University of Oslo, Norway). Differences in total concentrations between and within seasons were made with one-way ANOVA. (Microsoft Excel, 2016). Thirty-eight interviews were conducted with tourists during the summer 2015 to find out what types of antifouling paints, and subsequently amount of active ingredient, that were most frequently used (see supplementary information, Appendix A). Moored boats were counted during 11 days over five weeks in July and August where boat types and sizes were noted and an estimated hull area was calculated using specifications for the different boat types, such as length at water level, beam, depth and keel type (reactiveresins.com/boat-bottom-calculator.html). Hull area (m2) combined with known leaching rates (μg cm−2 day− 1) from three commercial paints (KemI, 2017) generated an assessment on the potential for Cu pollution to the natural harbours. 3. Results Total PAH concentrations were higher during post season (Fig. 2a), but the difference in total PAH concentration between seasons, and within location types within post and peak seasons, were not 666
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Fig. 2. Total concentrations of sampled PAH per location for both seasons. a.) All sampled PAHs. b.) Sampled PAHs minus naphthalene. Error bars (SD) are added for locations with double membranes; all locations during peak season but only NH 3 during post season. NH 3 was pooled within post season (April–May) as PAH concentrations at this location when naphthalene was removed were very similar, and naphthalene was not detected at this location in May.
There was no significant difference between different location types during peak season (Fig. 8a), when all heavy metals were taken into account, but comparison during post season between near mainlandmarina and archipelago locations did show a significant difference (R = 0.44, p = 0.033). Near mainland and marina sites are areas where boats are maintained and launched generally display higher concentrations than the archipelago (Fig. 6b). NH 4 is located close to a commercial ferry and has a small private marina nearby, which may be why this location did not resemble the other natural harbours, but instead grouped together with near mainland waterways. Five out of 38 visitors interviewed knew the exact paint type they had used, and two of these were banned paints. Twenty-seven of the participants used legal paints, with Cu as an active ingredient, but only three knew the amount of the active substance. When asked why they chose their specific paint, 38% said it was because it's the paint they had always used, 21% claimed it was the most efficient and only 5% stated they had chosen the paint for environmental reasons. The four natural harbours were subject to very different leaching potential as the number of boats and boat sizes varied depending on location. Copper did not vary substantially during peak season between the natural harbours, but Zn did (Fig. 6a). Zn concentration during peak season ranked NH 2 < NH 3 < NH 1 < NH 4, which reflects the leach scenario for Cu (Table 2). This pattern was different during post season where NH 1 < NH 3 < NH 2 < NH 4 for Zn.
marina locations. The concentration range in the archipelago was 0.96–2.39 ng PAH/L and at the near mainland and marina location it was 1.2–4.73 ng/L, so the total concentrations in the archipelago was lower, although not significantly. Concentrations of metals were substantially higher during post season (Fig. 6). This was most likely due to presence of heavy biofouling on the DGTs at peak season that can reduce uptake markedly as sorption of metals occur on the fouling, which decreases the concentration gradient and in turn affects accumulation in the DGT (Devillers et al., 2017). Copper was uniformly distributed during peak season and well below AA-EQS (Fig. 6a), but Zn varied between all location types. During post season, WW 1, 4 and NH 1–3 had similar concentrations for Cu, with Zn being more variable, which correspond to results from post season distribution (Fig. 8b). NH 4 is located in the vicinity of the ferry stop, which may account for the elevated Cu and Zn concentrations (Fig. 6b). AA-EQS for both Cu and Zn was exceeded at this location. Both of the near mainland waterway (WW) locations had elevated concentrations of Cu as well as Zn. These locations have houses with waterfront access nearby and boats may therefore be painted and launched there. The marina locations had low Cu concentrations, which may be due to more particles that bind Cu-ions, but these locations were high in Zn and exceeded AA-EQS. There was a significant difference between seasons (p = 0.0005, R = 0.407) for the composition of all heavy metals (Fig. 7). It is mainly the lower overall concentrations during peak season due to the heavy biofouling on the DGTs (Fig. 6a–b) that account for the difference.
Fig. 3. Post and peak season PAH composition for natural harbours (NH) and waterways (WW). White = peak season, black = post season. Naphthalene not included. Post season concentrations for location NH 3 have been pooled in this analysis.
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Fig. 4. Composition of 2, 3, 4 and 5-6-ringed PAHs at the different locations. a.) peak season, b.) post season, m = near mainland location. Dashed line delimits post season samples between WW1-NH3 (April) which are the same locations sampled during peak season. NH 3 (May)Mar 2 were sampled in May and only during post season. Naphthalene was not excluded in this comparison.
4. Discussion
concentrations resemble those sampled in Danish coastal areas and off shore values for the north Atlantic (Nizzetto et al., 2008). However, benzo(g,h,i)perylene has a MAC-EQS of 0,82 ng/L (HVMFS 2015:4) and the measured value at NH 1 during peak season was 0,69 ng/L. The MAC-EQS for PAHs are set for the concentration in an entire water sample, and the SR membranes only take up the dissolved fraction. As PAHs are likely to be bound to dissolved organic matter or particles, it is reasonable to believe that when including all fractions in the water column, the EQS may very well be exceeded. The concentrations of Cu exceeded AA-EQS at one location during post season and Zn exceeded its AA-EQS at five locations (Fig. 6b). However, potential leaching load (Table 2) show that a substantial amount of Cu is likely to be released from boat hulls which affirms that DGT at peak season were uptake-inhibited due to biofouling. Effects of fouling on uptake of some metal elements, like Cu, Pb and Cd, have been shown as metal sorption occur on the fouling (Devillers et al., 2017) and as algae also utilize metals like Cu this adds to the reduced uptake and real concentrations are therefore likely to exceed post season values. This is problematic since AA-EQS are then likely to be exceeded for both Cu and Zn for several locations, which could have serious localized effects on the biota in the area. EQSs for metals are set for the fraction smaller than 45 μm, and therefore contain more than just the free ion form, which is collected by the DGTs. Macro algae have been proven to accumulate not only free Cu2 + but also organically complexed Cu (Ytreberg et al., 2011) and hence the effect of sampled concentrations of Cu and Zn in this study may be underestimated. The EQSs are set for single compounds and do not take into account interactions that may be additive, synergistic or antagonistic. Copper, for example, exhibit synergistic effects with the biocide irgarol but
There was a significant difference in composition of both PAHs and metal ions between the sampled seasons (Figs. 3, 7), indicating that leisure boat activities influence the contamination profile in the water mass. Natural Harbour 3 can also exemplify this, where PAH composition changed between April and May 2016, where the May sampling occurred after a period of good weather at the start of the boating season. However, within season, it was only during post season that differences in composition of both compounds types between the location types were detected, where near mainland and marina locations differed from archipelago locations. This suggests that during summer when boats are trafficking the entire area everything is everywhere and no distinction between locations can be made. Total concentrations of the dissolved bioavailable fraction of both PAHs and metal ions were higher during post season, which is a somewhat counterintuitive result. However, PAHs are hydrophobic and will be adsorbed onto particulates or dissolved organic matter (Smol and Włodarczyk-Makuła, 2017, references therein). During peak season, many of the shallow locations will be subject to sediment resuspension due to boating activities (propellers, anchors) and PAHs may therefore be less available for uptake in dissolved form. The relative reduction of 2-ringed PAHs due to higher water temperatures also contributes to lower bioavailable concentrations. Experimental artefacts for the metals can also be linked to adsorption to particles, but most likely, the much lower concentrations during peak season were due to biofouling on the DGT samplers that prevents uptake. None of the individual PAHs exceeded their EQS-values and
Fig. 5. PAH composition during a.) Peak season, separated into waterway (WW, white) and natural harbour (NH, black). b.) post season, separated into near mainland and marina locations (WW 3, Mar, grey) and archipelago locations (NH, WW 1,4 black), NH 3 post season concentrations have been pooled.
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Fig. 6. Cu and Zn concentrations from DGT membranes. a.) Peak season concentrations. a.) Post season concentrations. Dashed lines are AAEQS for Cu (4.3 μg/L) and Zn (3.4 μg/L), m = near mainland.
how well they cope with chemical pollution. Not exceeding EQSs for single compounds therefore does not necessarily protect the environment or assure good ecological status. In this study, we found metal ions and PAHs present close to, or above their MAC- and AA-EQSs in natural harbours of a marine national park. We therefore cannot rule out that the compounds alone or their combined effect exert pressure on the environment. Furthermore, the impact of leisure boats may not affect the pelagic community the most, but rather benthic communities and filter feeding organisms that are exposed both to the truly dissolved as well as the particulate fractions.
when mixed with another biocide, diuron, the effects are antagonistic (Gatidou et al., 2007). The single compound EQS also fails to consider the added stress in organisms of being exposed to several different stressors at the same time. While looking into experiments with the aim of studying interactions, Crain et al. (2008) found that synergy was the most common outcome with two interacting compounds, but interaction type (antagonistic, synergistic) varied depending on, for instance, trophic level and the combined stressors. The addition of a third stressor, however doubled the amount of synergistic interactions. In a study on periphyton, when combining five non-similarly acting compounds present at their individual NOEC (no observed effect concentration) levels, there was a clear effect of the mixture (Backhaus et al., 2011). Another study tested mixtures of up to 19 compounds at their individual AA-EQSs and found both acute and chronic effects in toxicity tests on different organism groups (Carvalho et al., 2014). The marine environment is exposed to several chemical compounds simultaneously. When monitoring chemicals along the Swedish west coast, 30 of the 172 analysed compounds, which included PAHs and biocides, were detected at the chosen reference site (Gustavsson et al., 2017). The fact is that organisms everywhere are constantly exposed to several different compounds, but are also stressed by other abiotic factors like changes in temperature, pH and oxygen, which will affect
Acknowledgements The study was funded by Alderman and Mrs. Ernst Colliander's Foundation, Wilhelm and Martina Lundgren Science Foundation, (2015-0706) the foundation Birgit och Birger Wåhlströms Minnesfond för den bohuslänska havs- och insjömiljön, Herbert och Karin Jacobssons Foundation (7/v16) and the Royal and Hvitfeldska Foundation which are kindly acknowledged. We would also like to acknowledge Amandine Vuylsteke for help in the field, the staff at Sven Lovén Centre for Marine Infrastructure, Tjärnö for help during fieldwork and the staff at the department of Bioscience and the department of Environmental Science, Aarhus University; Charlotte Dahl Schiødt Fig. 7. Post (black)- and peak (white) season metal ion composition. Cu, Zn, Pb, Cr and Cd were included in the analysis, but Cu and Zn were contributing most to the pattern. NH 3 concentrations for post season have been pooled.
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Fig. 8. Metal ion composition within seasons. a.) peak season; marina (grey), natural harbour (black) and waterway (white). b.) post season; near mainland waterways (WW 2–3) and marinas (grey) and archipelago, natural harbour (NH 1–4) and waterway (WW 1,4) (black) NH 3 post season concentrations have been pooled.
Biodegradation 3, 351–368. CHANGE, 2014. http://changeantifouling.com/wp-content/uploads/2014/10/Sstad.pdf. Crain, C.M., Kroeker, K., Halpern, B.S., 2008. Interactive and cumulative effects of multiple human stressors in marine systems. Ecol. Lett. 11, 1304–1315. Devillers, D., Buzier, R., Grybos, M., Charriau, A., Guibaud, G., 2017. Key role of the sorption process in alteration of metal and metalloid quantification by fouling development on DGT passive samplers. Environ. Pollut. 230, 523–529. Di Toro, D.M., Allen, H.E., Bergman, H.L., Meyer, J.S., Paquin, P.R., Santore, R.C., 2001. Biotic ligand model of the acute toxicity of metals. 1. Technical basis. Environ. Toxicol. Chem. 20, 2383–2396. Directive 2000/60/EC Of the European parliament and of the council of 23 October 2000, Establishing a Framework for Community Action in the Field of Water Policy. Directive 2008/105/EC Of the European parliament and of the council of 16 December 2016, on environmental quality standards in the field of water policy, Amending and Subsequently Repealing Council Directives 82/176/EEC, 83/513/EEC, 84/156/EEC, 84/491/EEC, 86/280/EEC and Amending Directive 2000/60/EC of the European Parliament and of the Council. Directive 2013/39/EU Of the European parliament and of the council of 12 August 2013 Amending Directives 2000/60/EC and 2008/105/EC as Regards Priority Substances in the Field of Water Policy. Egardt, J., Dahllöf, I., Nilsson, P., 2017. Sediment indicate the continued use of banned antifouling compounds. Mar. Pollut. Bull. 125, 282–288. Eklund, B., Elfström, M., Gallego, I., Bengtsson, B.E., Breitholtz, M., 2009. Biological and chemical characterization of harbour sediments from the Stockholm area. J. Soils Sediments 10, 127–141. Gatidou, G., Thomaidis, N.S., Zhou, J.L., 2007. Fate of Irgarol 1051, diuron and their main metabolites in two UK marine systems after restrictions in antifouling paints. Environ. Int. 33, 70–77. Gustavsson, B.M., Magnér, J., Carney Almroth, B., Eriksson, M.K., Sturve, J., Backhaus, T., 2017. Chemical monitoring of Swedish coastal waters indicates common exceedances of environmental thresholds, both for individual substances as well as their mixtures. Mar. Pollut. Bull. 122, 409–419. HVMFS, 2013. Havs- och vattenmyndighetens författningssamling. Havs och vattenmyndighetens föreskrifter om ändring i Havs- och vattenmyndighetens föreskrifter (HVMFS 2013:19) om klassificering och miljökvalitetsnormer avseende ytvatten. Swedish Agency for Marine and Water Management. HVMFS, 2015. Havs- och vattenmyndighetens föreskrifter om ändring i Havs- och vattenmyndighetens föreskrifter (HVMFS 2013:19) om klassificering och miljökvalitetsnormer avseende ytvatten. Swedish Agency for Marine and Water Management. KemI, 2017. Swedish Chemicals Agency. Kwok, K.W.H., Leung, K.M.Y., 2005. Toxicity of antifouling biocides to the intertidal harpacticoid copepod Tigriopus japonicus (Crustacea, Copepoda): effects of temperature and salinity. Mar. Pollut. Bull. 51, 830–837. Nizzetto, L., Lohmann, R., Gioia, R., Jahnke, A., Temme, C., Dachs, J., Herckes, P., Di Guardo, A., Jones, K.C., 2008. PAHs in air and seawater along a North-South Atlantic transect: trends, processes and possible sources. Environ. Sci. Technol. 42, 1580–1585. Smedes, F., Booij, K., 2012. Guidelines for passive sampling of hydrophobic contaminants in water using silicone rubber samplers. In: ICES Techniques in Marine Environmental Sciences No. 52, (20 pp). Smol, M., Włodarczyk-Makuła, M., 2017. The effectiveness in the removal of PAHs from aqueous solutions in physical and chemical processes: a review. Polycycl. Aromat. Compd. 37, 292–313. STA, 2015. Båtlivsundersökningen 2015. Swedish Transport Agency, Norrköping. Thomas, K.V., Brooks, S., 2010. The environmental fate and effects of antifouling paint biocides. Biofouling 26, 73–88. Valkirs, A.O., Seligman, P.F., Haslbeck, E., Caso, J.S., 2003. Measurement of copper
Table 2 Potential leaching pressure (g Cu location− 1 day− 1) for three different commercial paints to the natural harbours (NH) based on average boat counts and boat sizes at sampled locations during peak season. Leaching rates for paints used were: Low (Hempel light), 0.97 μg cm−2 day− 1, Medium (Biltema west coast =), 2.37 μg cm−2 day− 1, High (Hempel Mille Extra), 3.5 μg cm−2 day− 1. (Leaching rates with permission from KemI). Medium and High leaching paints are known to be used in the area.
NH NH NH NH
1 2 3 4
Low
Medium
High
0.6 0.3 0.8 3.0
1.4 0.8 1.9 7.2
2.1 1.2 2.8 10.6
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