The distribution of heavy metals and 137Cs in the central part of the Polish maritime zone (Baltic Sea) – the area selected for wind farm acquisition

Accepted Manuscript 137 The distribution of heavy metals and Cs in the central part of the Polish maritime zone – An area selected for wind farm development. Agata Zaborska, Alicja Kosakowska, Jacek Bełdowski, Magdalena Bełdowska, Marta Szubska, Jolanta Walkusz-Miotk, Adam Żak, Agnieszka Ciechanowicz, Maciej Wdowiak PII:

S0272-7714(16)30734-X

DOI:

10.1016/j.ecss.2016.12.007

Reference:

YECSS 5337

To appear in:

Estuarine, Coastal and Shelf Science

Received Date: 29 December 2015 Revised Date:

15 September 2016

Accepted Date: 17 December 2016

Please cite this article as: Zaborska, A., Kosakowska, A., Bełdowski, J., Bełdowska, M., Szubska, M., 137 Walkusz-Miotk, J., Żak, A., Ciechanowicz, A., Wdowiak, M., The distribution of heavy metals and Cs in the central part of the Polish maritime zone – An area selected for wind farm development., Estuarine, Coastal and Shelf Science (2017), doi: 10.1016/j.ecss.2016.12.007. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT 1

The distribution of heavy metals and 137Cs in the central part of the Polish maritime zone – an

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area selected for wind farm development.

3 4 Agata Zaborska1, Alicja Kosakowska1, Jacek Bełdowski1, Magdalena Bełdowska2, Marta

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Szubska1, Jolanta Walkusz-Miotk1, Adam Żak1, Agnieszka Ciechanowicz1, Maciej Wdowiak3

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Department, Powstańców Warszawy 55, 81-712, Sopot, Poland

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- Institute of Oceanology, Polish Academy of Sciences, Marine Chemistry and Biochemistry

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Environmental Protection, Av. Marszałka Piłsudskiego 46, 81-378 Gdynia, Poland

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Corresponding author email: [email protected]

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– BALTEX S.A., Śląska 53, 81-304, Gdynia, Poland

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- University of Gdańsk , Institute of Oceanography, Department of Marine Chemistry and

1. Introduction

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Heavy metals are elements that occur naturally in the environment. Their natural

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concentrations are usually very low and vary between areas of different geological origin.

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Since the beginning of the industrial era (the industrial revolution at the end of 19th century)

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large loads of metals have been emitted to the environment due to intense human activity.

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Some heavy meals are very toxic (for example, Hg, Cd, Pb and As) even at low

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concentrations, and have no known beneficial biological effects. Heavy metals enter the

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environment during production (including mining and smelting), use (batteries, pigments,

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ceramics, plastics), recycling, combustion of fossil fuels (coal, former use of leaded gasoline),

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the use of mineral fertilizers and sewage sludge application, etc. (Szefer, 2002). Metals are

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transported to the marine environment in dissolved form and/or attached to particles via air

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masses (mainly Hg, Pb), rivers (Zn, Cu, Pb, As, Hg, Cd) and water currents (Schneider et al.,

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2000), coastal erosion (Bełdowska, 2015) and groundwater discharge (Szymczycha et al.,

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2014). Artificial radionuclides (eg.

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originate from atmospheric global deposition (nuclear explosions that were conducted mainly

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in the ‘60s and ‘70s) and the Chernobyl accident in 1986. Moreover, some radionuclides are

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discharged by an inflow of Atlantic water containing effluents from reprocessing plants (for

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example, La Hague and Sellafield) and from effluents of nuclear power plants in the Baltic

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Sea catchment (Zaborska et al., 2014).

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Cs,

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Sr, plutonium isotopes) in the Baltic Sea chiefly

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ACCEPTED MANUSCRIPT The Baltic Sea is very sensitive to contamination by heavy metals and radionuclides,

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as it is a small, semi-closed reservoir with large riverine discharges from densely populated

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and industrialized catchments. Several monitoring programs control the concentrations of

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heavy metals within selected localizations of the Baltic Sea. In the case of the Polish

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Economic Area, reports concentrate on the Gdańsk Basin. Numerous scientific papers concern

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the contamination of the southern Baltic Sea sediments by heavy metals, but again most of the

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studies have been conducted in the Gdańsk Basin and Puck Bay (Szefer and Skwarzec, 1988;

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Pempkowiak et al, 1991; Pempkowiak et al., 1994; Pempkowiak et al., 2000; Szefer, 2002;

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Bełdowski and Pempkowiak, 2003; Glasby et al., 2004; Szefer et al., 2009; Hendożko et al.,

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2010; Zalewska 2011, Polak-Juszczak, 2013; Zaborska 2014, Bełdowski et al., 2014;

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Zalewska et al., 2015). In summary, the levels of heavy metals in the Polish maritime zone

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exceed the natural geochemical background in some regions. Metal concentrations are found

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to be highest in the deepest regions of the Baltic Sea (e.g. Gdańsk Basin) and are associated

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with fine grained sediments (Szefer, 2002). This study was conducted in the central part of the

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southern part of the Baltic Sea (from the latitude of the city of Kołobrzeg to the town of

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Łeba). The metal concentrations in this area that is characterized by variable bathymetry eg.

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shallow banks and deeper basins and furrows as well as variable sedimentary environments

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are not well recognized.

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Polish demand for renewable energy is increasing. The traditional ways of energy

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acquisition (fossil fuel burning) cause problems of growing carbon dioxide emission, may

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induce risk of environmental pollution and finally may deplete non-renewable resources.

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Wind energy is often described as a suitable alternative for obtaining energy in an ecological

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way (Esteban et al., 2011; Leung and Yang, 2012). In particular, offshore wind parks are

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believed to be very effective since winds are stronger at sea. Unfortunately, windmill parks

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may induce some risk to the surrounding environment. The most pronounced negative effect

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is connected to the fishery industry (OSPAR 2004, HELCOM 2010). The effects on

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biodiversity of lower trophic level organisms (plants, benthos) are often not assessed. The

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environmental risks of offshore wind energy (OWE) may also be connected to abiotic

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compounds of the marine ecosystem. Physical damage to the seabed constitutes one of the

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major pressures on the Baltic marine environment (Jennings et al., 2001). OWE may create a

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barrier to bottom sediment transport and during construction of windmills, may disturb the

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sediments deposited at the sea bottom. This disturbance will induce re-suspension of

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sediments contaminated over the last century and will thus reintroduce contaminants back to

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the water column. Therefore, it is very important to check which offshore region should be

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selected for wind farm development and what consequences of OWE construction may be

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expected. This study was part of the large project AQUILO, which was conducted to create a

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knowledge base from which investors will be able to decide on the best type of support

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structure for an offshore wind farm in a specific location in the Polish maritime area’. Several

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aspects were studied, e.g. possible destruction of benthic fauna and flora, possible changes in

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sediment transport mechanisms and possible re-contamination of the environment caused by

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bottom disturbance. The aim of this study was to determine the distribution of heavy metals in

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the sediments of the middle part of the Polish coast and to study precisely the history of

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contamination by heavy metals in selected regions. The knowledge of distribution of

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anthropogenic substances in the marine sediments will help correct decisions to be made on

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wind park localization. Finally, this study’s results will be used in a hydrodynamic model to

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assess the amount of contaminants that might re-enter seawaters during construction of OWE

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installations. .

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2. Materials and methods

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Sampling was performed during two research cruises, in May 2013 on r/v Oceania

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(cruise of the Institute of Oceanology, Polish Academy of Sciences) and in November 2013

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on r/v Santa Barbara (cruise of the Maritime Institute in Gdańsk). A sampling area of 9000

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km2 in the central part of the Polish economic zone was covered (Fig. 1). Sediments deposited

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at depths of 23 m to 84 m were collected by gravity twin corer (GEMAX) or geo-vibro corer.

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In total sediments from 46 stations were collected. 7 sediment cores of 40 cm length, intended

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for the precise study of contamination history (sediment dating), were sliced every 10 mm (till

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10 cm) and every 20 mm (till 40 cm) (stations: 30, 31, 32, 33, 34, 35, 36). Based on the

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determined age of these sediment layers, five sediment layers were selected for study of heavy

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metal concentration (0-1 cm, 1-5 cm, 5-10 cm, 10-15 cm, 15-20 cm). The 29 sediment cores

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intended for broader study of spatial heavy metal distribution were sliced every 5 cm

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(stations: 1-29). At 10 locations (stations: 37-46) only surface sediment samples were

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collected. Sediment samples dedicated for heavy metals analyses were packed in zip bags and

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frozen. Samples dedicated for determination of organic contaminants (PCBs, PAHs) were

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packed into glass jars and frozen. The study of organic contaminant distribution is the subject

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of another manuscript in preparation.

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2.2 Laboratory measurements Frozen samples were transported to the Institute of Oceanology, Polish Academy of

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Sciences (IO PAN). Sediment samples were transferred to glass Petri dishes, weighed and

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freeze-dried in a Christ Beta A Freeze Dryer. After drying, the dishes were re-weighed and

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the water content was calculated. The sediment density was measured from sub-samples

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collected in a polyethylene container of known volume. The sediment porosity was

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determined using wet and dry sediment weights and sediment density. Sediment samples were

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mixed accurately, then divided into several sub-samples according to each type of analysis.

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Grain size analyses was performed by the dry sieving method. A combination of

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several sieves (2 mm, 1 mm, 500 µm, 250 µm, 125 µm and 63 µm diameter) was used to

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separate gravel, sand and pelite (silt and clay) fractions. The Wentworth scale (1922) was used

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for sediment type determination.

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The organic matter content (Morg) was estimated by loss on ignition analyses. Weighed

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samples of sediments were transferred into porcelain dishes and combusted at 450°C for 6

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hours. The Morg was calculated as loss of material compared to the initial sample weight.

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To reveal the history of contamination by heavy metals, sediment layers were dated by 210

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Pb method. The sediment dating was performed using alpha spectrometry

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(measurement of the 210Pb daughter radionuclide 210Po) assuming secular equilibrium between

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both nuclides. Radiochemical separation of 210Po was performed by the method developed by

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Flynn (1968) and adapted by Pempkowiak (1991). Sediment samples of 0.2 g were spiked

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with 209Po tracer (AEA Technology) and digested using perchloric acid and hydrofluoric acid

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in an oven warmed to 140 °C. Later nitric acid was added and the solutions were evaporated

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in a fume hood. Dried residues were transferred to Teflon beakers with 0.5 M HCl. Polonium

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isotopes were spontaneously deposited in acid solution (ion exchange reaction) onto silver

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disks. After deposition, the disks were washed with methanol and analyzed for 210Po and 209Po

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in a multi-channel analyzer (Canberra) equipped with Si/Li detectors. The samples were

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counted for 1 day. The activity of

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recovery by comparing the measured and spiked activities of 209Po. The 210Pbsupp activity was

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calculated as the average of several

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zone of 210Pb exponential decline).

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Po in the sample was determined based on chemical

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Pb determinations in deep sediment layers (below the

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For heavy metals which are expected to be anthropogenic contaminants (Pb, Cu, Zn,

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Cd) and other metals (Cr, Ni, Fe, Mn), the method described in Vallius and Leivuori (1998)

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was used. Not fractionated sediment sub-samples of 0.5 g were digested in concentrated

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Suprapur perchloric acid and Suprapur hydrofluoric acid, in Teflon beakers, in an oven 4

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evaporated in a fume hood. The dried residues were dissolved in Suprapur 0.5M nitric acid

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and samples were transferred to polyethylene bottles. The samples were finally diluted in

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Suprapur 0.1 M nitric acid 100 or 1000 times, depending on the expected metal concentration,

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in a laminar flow cabinet. Single-element standard solutions of several concentrations (0.2-6.0

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µg·ml-1) were prepared from a stock solution (Merck). About 5-6 concentrations of each metal

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were prepared by serial dilution with dilute 0.1 M HNO3 to draw calibration curves. Prepared

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solutions were analyzed for the selected metals with a flame atomic absorption spectrometer

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(AAS Shimadzu 6800) using deuterium background correction. The Cd concentrations were

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measured in a Perkin-Elmer Sciex ELAN 9000 ICP-MS due to expected very low

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concentrations. Samples dedicated to measurement of As were treated according to the

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procedure of Loska and Wierzchuła (2006). Firstly, a dry mineralization of the sample mixed

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with Mg(NO3)2 was performed at 550 °C. Then the residues were transferred to polyethylene

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bottles using HCl and filtered by Millex filters (0.45 µm). The samples were diluted in MilliQ

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water, then 1ml of concentrated Suprapur HCl and 20% KI solution were added to reduce

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As(V) to As(III). Finally the solutions were measured in a Hydride Generation Atomic

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Absorption Spectrophotometer AAS 6800 (HVG-AAS). The total Hg concentration was

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measured via sample (0.5 g) pyrolysis in a stream of oxygen (Leco AMA 254) (Bełdowski et

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al., 2015). In total, 100 sediment samples collected at 46 stations (surface sediments and

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sediment cores) were analyzed for heavy metal concentrations.

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The

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spectrometry. Energy and efficiency calibration was performed using a multi-nuclide standard

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MBSS 2 – silicone matrix (Czech Metrology Institute). Known amounts of sediments were

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packed in counting vials with known geometry. Caesium-137 was measured in a high purity

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germanium detector (Canberra), with 40 % relative efficiency, for 3-4 days (Zaborska et al.,

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2014). Corrections for background and counting vial geometry were performed.

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Cs activity concentrations in the sediment samples were measured by gamma

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2.3 Quality assurance and control 210

Pb measurement. The efficiency of chemical separation and detection was 209

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calculated for every sample using a

Po tracer. The recovery of the whole procedure:

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chemical procedure and measurement in the alpha detector ranged from 30 % to 36 %.

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Standard reference materials (IAEA-300 and IAEA-326) were measured as quality control

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samples. The results were within the certified values given by IAEA (92-102%, n=8). One

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blank sample (not containing the sediment) was measured for every 10 sediment samples. The 5

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environmental background of the detector was checked routinely and was found to be

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negligible. Heavy metals measurement. Certified reference materials IAEA-433 and JMS-1 were

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measured simultaneously with the samples for quality control purposes. The results were

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within the certified values (93-104 %, n=16) and in agreement with the long-term internal

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laboratory reproducibility (n=±500). Results were satisfactory, as the recovery was > 90%.

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The detection limits (LOD) of every element was calculated as 3 x SDb, where SDb values

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were the standard deviations of the blank samples (measured 10 times). Blank samples were

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prepared from Suprapur 0.1M HNO3. LODs were as follows: Pb=0.03 µg·g-1, Zn=0.02 µg·g-1,

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Cr=0.03 µg·g-1, Ni=0.03 µg·g-1, Cu=0.03 µg·g-1, Fe=0.05 µg·g-1, Mn=0.03 µg·g-1, Cd=3.5

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ng·g-1, Hg=0.5 ng·g-1, As=0.2 ng·g-1. The precision of metal concentration quantification,

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assessed by triplicate analysis of the three sub-samples of the same sediment samples (whole

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procedure), was 1-3 % for AAS measurements and 1 % for ICP-MS measurements.

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materials IAEA 300, 315 and 385 and results agreed with the certified values at 91-108%

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confidence (n=15). Detector background was assessed by counting of an empty vial in the

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same detector for several days.

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Cs measurement. Quality control was assured by measurements of IAEA reference

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2.4. Calculations

Sediment accumulation rates are determined from profiles of excess (210Pbex = total

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mass sediment accumulation rates are calculated assuming an exponential decrease in

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with sediment depth (Robbins and Edgington, 1975) using a CF:CS model for most of the

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stations. The best-fit least-squares method was used. The determined rates represent

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maximum rates. In two sediment cores (stations 30 and 35) the disturbance of surface

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sediments, suggesting a variable sediment accumulation rate, was noticed, thus a CRS model

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was applied to determine sediment layer age.

Pb-supported

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Pb activity

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Pb) versus porosity-corrected sediment depth. Linear and 210

Pbex

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Anthropogenic factors (AF) were calculated to estimate the enrichment of metal

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concentrations in surface sediment layers. The measured concentrations of heavy metals were

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divided by a mean geochemical background reported in the literature (Uścinowicz, 2011).

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2.5. Statistical analyses

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The Pearson correlation analysis was performed to find correlations between particular

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metal concentrations and metal concentrations and sediment properties (e.g. grain size,

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organic matter concentration etc.). The correlations with p < 0.05 were considered to be

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significant. Correlation coefficients r≥0.8 are considered to show high correlation. All

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statistical calculations were performed in STATISTICA 10.0 licensed program. The dataset containing all results (surface samples and sediment core samples

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(n=100)) was tested by ANOVA (STATISTICA 10.0) to assess the statistical significance of

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differences in sediment properties across regions. The data did not fulfill the test assumptions

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(normality of distribution and homogeneity of variance), and therefore we applied the non-

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parametric Kruskal–Wallis test. The results were divided into 2 regions: shallow sandy areas

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(<40 m depth) and deeper silty areas (>40 m) to find differences in heavy metal

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concentrations between shallow sandy bottom regions and deeper fine-grained bottom

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regions.

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3. Results

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3.1 Sediment accumulation rates

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The study region covers a varied area composed of shallow banks, deeper basins and

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deep furrow. Due to different sedimentary conditions at the studied stations, the sediments

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vary in terms of organic matter content and sediment grain size. The organic matter content

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ranged from 0.3 % in Central Bank sediments (station 3) to 12.1 % in Bornholm Basin

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sediments (station 37). The grain size was measured in surface sediments to give an idea of

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the environmental conditions at the given location (erosion or accumulation bottom type).

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Sand content ranged from 11.2 % (station 35) to 100 % (stations 39, 41 and 42). Fine material

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(silt and clay) prevailed at stations 30, 32, 35, 36, 37 and 40 (Tab. 1). Sandy sediments

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dominated at shallower regions of Słupsk Bank, Stilo Bank and the coastal area close to

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Kołobrzeg, where water depths ranged from 20 m to 40 m. Shallow areas of the studied

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region are high energy environments influenced by bottom currents and waves. The deposited

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finer material is diluted to suspension and transported to deeper (calmer) areas.

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The total

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Pb profiles at the seven stations selected for sediment age calculation are

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shown in figure 2. The supported 210Pb ranged from 28.1 Bq·kg-1 at station 36 to 76.10 Bq·kg-1

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at station 35. This difference reflects different sources of sedimentary material in the studied

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region. The excess 210Pb (total Pb-supported Pb) ranged from 74.80 Bq·kg-1 to 271.90 Bq·kg-1.

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The Pbex profile at station 31 represents a station with mixed sediments where sediment dating 7

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by the

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suggest variable sediment accumulation rate (thus at these stations CRS modeling was

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applied).

Pb method cannot be applied. The shape of Pbex profiles at stations 30 and 35

The linear sediment accumulation rate (LAR) was estimated at 0.07±0.01 mm·yr-1 at

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stations 33 and 34, 0.10±0.01 mm·yr-1 at station 36, and 0.14±0.02 mm·yr-1 at station 32. (Fig.

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2). At station 30 the sediment accumulation rate was found to vary from 0.39 cm/yr in surface

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sediment layers (0-1 cm and 1-2 cm) to 0.15-0.20 cm/yr in deeper layers (from 3-4 cm to 22-

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24 cm). In the case of station 35 the LAR was very variable and changed from 0.43-0.70

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cm/yr in the upper 5 cm of the core while in deeper layers it was from 0.19 cm/yr (6 cm-14

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cm) to 0.06 cm/yr (14 cm-20 cm).

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3.2 Selected heavy metal and 137Cs spatial distribution

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The distribution of the most toxic metals (Cd, Hg, As, Pb, Zn, Cu) in surface

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sediments (0-5 cm) is shown in figure 3. Cadmium concentrations ranged from 0.03 µg·g-1 to

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as high as 1.32 µg·g-1. The concentrations of mercury were lower and ranged from 0.001 µg·g-

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1

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concentrations of lead ranged from 6.7 µg·g-1 to 86.1 µg·g-1. The concentrations of zinc were

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highest and ranged from 7.9 µg·g-1 to 123.1 µg·g-1. The copper concentrations ranged from

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1.35 µg·g-1 to 71.32 µg·g-1.

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to 0.019 µg·g-1. The arsenic concentrations varied from 0.28 µg·g-1 to 22.66 µg·g-1 while

The most contaminated are stations 30, 31, 32, 35, 36, 37 and 40 where most metals

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exceeded the geochemical background concentration. The concentrations of other metals that

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are not expected to be enriched due to pollution, for example nickel, chromium, iron,

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manganese and cobalt were similar to natural background (see Supplement), however in the

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case of two stations (31 and 35) manganese showed slightly higher than natural levels (see

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Supplement).

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To avoid overestimation of anthropogenic input of heavy metals in sediments, data

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normalization is often applied. Most heavy metals (e.g. Zn, Cd, Pb) in the marine sediments

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of the Baltic Sea are found to be associated with stable phases of sediments, mainly Fe-Mn

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oxide/hydroxide components (Helios-Rybicka et al., 1995; Szefer, 2002). Fe is combined with

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residual mineral components of the sediment (Szefer, 2002) and it is believed to be a very

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good indicator of the litogenic (natural) origin of sedimentary material in the Baltic Sea

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(Uścinowicz, 2011). Thus Fe is most frequently used as a normalization agent for heavy metal

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concentrations. Metal concentrations may also be normalized to organic matter content and

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the proportion of fine-grained material . The metal concentrations presented in figure 3 were 8

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real and comparable values. Fe concentrations vary in the sampling region and the highest Fe

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content was measured at the same stations as those with the highest levels of other heavy

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metals (Fig. 3). Part of the variability of heavy metal concentrations can thus be explained by

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Fe/Mn variations. The dependence of metal concentrations on pelite fraction (PF) content,

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organic matter content (Morg) and Fe concentrations was checked using regression analysis

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(Tab. 2).

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The highest correlation with the normalization factor (Fe) was noted for Zn and Cu

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concentrations (r=0.93, r=0.88 respectively). The correlations between Fe and Pb and Hg

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concentrations were moderately high (r=0.50-0.77) and lowest correlation coefficient was

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calculated for As (r=0.24) (Tab. 2). The correlations of organic matter content with Pb, Fe and

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Zn concentrations were high (r=0.80-0.91), the correlations with Hg and Cu concentrations

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were moderate (r=0.73-0.77), while very low correlation was found between Morg content and

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Cd and As (r=0.48-0.49). The pelite fraction content did correlated well with Cu

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concentrations (r=0.80), lower correlations were measured for Pb, Zn and Hg concentrations

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and lowest for Cd and As (r=0.46-0.53).

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The Kruskal-Wallis test was applied to find the differences in heavy metal

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concentrations and sediment properties (Morg and PF) between shallow (<40m) and deeper

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(>40m) regions of the studied area. All the measured parameters differed significantly

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between both regions, almost all (apart from As) with extremely low probability (< 0.001)

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(Tab. 2). The

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Cs activity concentrations in surface sediments ranged from 0.1 Bq·kg-1

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(stations: 1, 12, 17, 25) to as high as 256.9 Bq·kg-1 (station 36) (Fig. 3). As

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anthropogenic substance the natural background concentration should be within the limit of

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detection. The lowest concentrations were measured in sandy sediments, as 137Cs accumulates

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in fine grained material (e.g. Szefer, 2002). The activity concentrations of

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correlate with pelite fraction content and organic matter content (R2=0.20, R2=0.17

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respectively).

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Cs is an

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Cs did not

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3.3 Selected heavy metals and 137Cs vertical distribution

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The deepest 7 stations characterized by fine grained sediments were selected for

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sediment dating and measurement of metal concentrations at different sediment depths. The

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sediment accumulation rate was estimated at 6 stations. At one station (31) sediments were

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mixed and no net accumulation was estimated. Thus, the history of heavy metal 9

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contamination was revealed at stations: 30, 32, 33, 34, 35 and 36. The concentrations of

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metals at aged sediment layers are given in figure 4. Mercury was measured only in surface

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sediments. At station 30, the vertical concentrations of heavy metals are similar along the core

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and, for instance, Cd varied from 0.28 µg·g-1 to 0.36 µg·g-1 while As varied from 8.33 µg·g-1 to

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9.09 µg·g-1. Zn ranged from 101.6 µg·g-1 to 125.1 µg·g-1 while Pb ranged from 35.46 µg·g-1 to

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49.51 µg·g-1. At station 32, the heavy metal concentrations were slightly lower than at station

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30: Cd ranged from 0.10 µg·g-1 to 0.26 µg·g-1, As ranged from 5.64 µg·g-1 to 7.90 µg·g-1 and

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Zn ranged from 16.2 µg·g-1 to 76.4 µg·g-1. The Pb concentrations ranged from 11.41 µg·g-1 to

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34.95 µg·g-1. Generally low heavy metal concentrations were measured at station 33. The

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concentrations of Cd varied from 0.15 µg·g-1 to 0.36 µg·g-1 while As varied from 3.97 µg·g-1 to

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5.55 µg·g-1. The Pb concentration was very consistent (20.63 µg·g-1 to23.73 µg·g-1) while Zn

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varied from 52.2 µg·g-1 to 88.4 µg·g-1.

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Heavy metal concentrations showed higher variability at station 34. Cd ranged from

320

0.11 µg·g-1 to 0.48 µg·g-1, As ranged from 4.18 µg·g-1 to 20.91 µg·g-1 and Zn ranged from 29.5

321

µg·g-1 to 69.9 µg·g-1. Pb did not vary that much and ranged from 16.11 µg·g-1 to 20.01 µg·g-1.

322

Station 35, located in the Bornholm Basin, was very contaminated, and had Cd concentrations

323

reaching 1.70 µg·g-1, Pb concentrations of 19.38 µg·g-1 - 86.11 µg·g-1 and Zn concentrations of

324

134.7 µg·g-1 - 223.2 µg·g-1. Heavy metal concentrations were again lower at station 36, where

325

concentrations of Cd of 0.16 µg·g-1 - 0.40 µg·g-1 and As of 4.45 µg·g-1 - 5.98 µg·g-1 were

326

noted. Pb and Zn ranged from 20.13 µg·g-1 to 53.72 µg·g-1 and 40.7 µg·g-1 to 74.7 µg·g-1 ,

327

respectively, at this station. The

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Cs activity concentration profile in sediment cores exhibits a subsurface 137

activity peak and lower concentration at the sediment surface. The

330

concentrations range from values below the limit of detection to 151.9 Bq·kg-1 (station 30),

331

31.2 Bq·kg-1 (station 32), 121.8 Bq·kg-1 (station 33), 8.6 Bq·kg-1 (station 34), 68.5 Bq·kg-1

332

(station 35) and finally 95.1 Bq·kg-1 (station 36) (Fig. 4).

333 334

Cs activity

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Discussion

335 336

4.1. Heavy metal concentrations in surface sediments

337

The concentrations of heavy metals in surface sediments collected from the central

338

part of the Polish maritime area - a region which has been earmarked for wind farms are in the

339

range of those measured recently in Southern Baltic Sea sediments by other researchers 10

ACCEPTED MANUSCRIPT (Polak-Juszczak, 2013; Zaborska 2014; Bełdowski et al., 2014; Zalewska et al., 2015). In

341

particular, heavy metal concentrations reported by Polak-Juszczak (2013) were measured in

342

sediments deposited along the central Polish coast at Stilo Bank, very close to our study area.

343

In our study, the Pb concentrations ranged from 6.7 µg·g-1 to 86.1 µg·g-1 (Fig. 3), while the Pb

344

concentrations reported by Polak-Juszczak (2013) range from 14-15 µg·g-1. Higher Pb

345

concentrations (18.8 µg·g-1 to 146.9 µg·g-1) were reported by Zaborska (2014) for the deeper

346

region of the Gdańsk Basin. Mercury concentrations measured in this study were very low

347

and ranged from 0.001 µg·g-1 to 0.019 µg·g-1 (Fig. 3). The Hg concentrations measured by

348

Polak-Juszczak (2013) were very similar and ranged from 0.003 µg·g-1 to 0.025 µg·g-1, while

349

Bełdowski et al. (2014) report Hg levels of 0.2 µg·g-1 in the Gdańsk Basin, 0.04-0.07 µg·g-1 in

350

Bornholm Basin and only circa 0.015 µg·g-1 in the Pomeranian Bay. The As concentration in

351

this study ranged from 0.28 - 22.66 µg·g-1 (Fig. 3). The As estimations made by Szczepańska

352

and Uścinowicz (1994) showed As concentrations of 1.7 µg·g-1 - 9 µg·g-1 for the central

353

Polish coast (Słupsk and Stilo Banks) and concentrations of 10 µg·g-1 – 15 µg·g-1 in Słupsk

354

Furrow. The largest As concentrations in that study were measured in the Gdańsk Basin,

355

where they exceeded 20 µg·g-1. In our study the Cd concentration ranged from 0.03 µg·g-1 -

356

1.32 µg·g-1 (Fig. 3), while Polak-Juszczak (2013) reported values of 0.48 µg·g-1– 0.54 µg·g-1

357

for Stilo Bank. Concentrations of an order larger were measured by Zaborska (2014) in

358

Gdańsk Basin sediments (0.3 µg·g-1 – 5.0 µg·g-1). The concentration of Zn in our study varied

359

in different regions and ranged from 7.9 µg·g-1 - 123.1 µg·g-1 (Fig. 3). Polak-Juszczak (2013)

360

reported Zn concentrations of 60-66 µg·g-1 in areas of the central Polish coast while

361

concentrations amounting to 399.8 µg·g-1 have been measured in the Gdańsk Basin (Zaborska,

362

2014).

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To check the level of contamination of surface sediments the achieved values were

364

compared to Upper Continental Crust (UCC) metal concentrations (Rudnick and Gao, 2001),

365

regional geochemical background concentrations (Uścinowicz, 2011) and Polish pollution

366

threshold concentrations (Journal of Laws, 2002) (Tab. 3). It may be noticed that the metal

367

concentrations measured in this study exceed both the UCC and regional background values,

368

but not the pollution threshold values for Poland.

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369

The anthropogenic factors (AF) were calculated to show the enrichment of the surface

370

sediment layers with heavy metals (see Supplement). As the reference value the mean of the

371

geochemical background reported by Uścinowicz (2011) was applied (Cd=0.2 µg·g-1, Pb=38

372

µg·g-1, Zn=83 µg·g-1, As=13 µg·g-1, Hg=0.05 µg·g-1, Cu=40 µg·g-1). The AF for Zn was higher

373

than 1 only in sediments from stations 30 (1.3-1.5) and 35 (2.4-3.1). Half of the samples were 11

ACCEPTED MANUSCRIPT enriched in Cd. The AF for Cd was extremely high at station 35where it ranged from 6.6 to

375

9.0. Sediment collected at this station were also polluted by Pb, whose AF ranged from 1.5 to

376

2.3. In several sediment samples As slightly exceeded the geochemical background values,

377

particularly at station 34 (AF=1.3-1.5). All measured Hg concentrations did not exceed the

378

natural geochemical background concentration. The anthropogenic factor results obtained in

379

this study are in the lower range of results reported by other researchers, thus the studied

380

sediments were not highly polluted. The anthropogenic factors for heavy metals were recently

381

reported for 6 sediment cores from the Gdańsk Bay and Basin (Zaborska, 2014). The

382

maximum reported AF for Pb reached 4 and a similar AF was reported for Zn (AF=4.8),

383

while an extreme enrichment of 25 was noted for Cd. Zalewska et al. (2015) reported

384

enrichment factors reaching 8 for Cd, 6.5 for Pb and 3.5 for Hg in Gdańsk Deep sediments.

385

The enrichment values for Bornholm Deep sediments were lower and reached 3.4 for Cd and

386

Hg, 2.4 for Pb and about 2 for Zn (Zalewska et al., 2015). The enrichment factors reported

387

two decades ago (Szczepańska and Uścinowicz, 1994) are much lower than recent values. In

388

that study, the maximum AF reported for Gdańsk Basin range from 1.7 (As) to 3.0 (Pb), while

389

the maximum AF calculated for Słupsk Furrow (close to our study region) range from 1.0

390

(Cd) to 4.5 (As).

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The 137Cs activity concentrations in surface sediments ranged from 0.1 Bq·kg-1 to as

391

high as 256.9 Bq·kg-1. The largest

137

393

entersthe Baltic Sea mainly from the direct air deposition and from riverine transport. In

394

seawater 137Cs is a conservative nuclide and it is present in dissolved form (Knapińska-Skiba

395

et al., 2002). However, tt also enters the Baltic Sea attached to fine mineral particles of

396

terrestrial origin, (Zaborska et al., 2014). The contamination of the Baltic Sea by 137Cs caused

397

by the Chernobyl accident was very patchy due to changing wind directions (Zalewska and

398

Suplińska, 2013). Taking into account unpredictable riverine and ground water discharge it is

399

quite difficult to predict the distribution of 137Cs within Baltic Sea sediments. Measured 137Cs

400

activity concentrations were in range of concentrations (120 Bq·kg-1 - 220 Bq·kg-1) measured

401

recently in Gdańsk Basin (Zalewska and Suplińska, 2013; Zaborska et al., 2014; Zalewska et

402

al., 2015) and in Bornholm Basin (70 Bq·kg-1 - 114 Bq·kg-1; Zalewska and Suplińska, 2013;

403

Zalewska et al., 2015). However, surface sediments of The Gulf of Finland or the Bothnian

404

Gulf are much more polluted, with 137Cs concentrations reaching 900 Bq·kg-1 (Zaborska et al.,

405

2014).

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Cs

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Cs pollution was measured in Słupsk Furrow.

406 407

4.2. Spatial differences of heavy metal distribution 12

ACCEPTED MANUSCRIPT 408

4.2.1 Shallow coastal areas Shallow areas of the Baltic Sea (depth <40 m) are affected by waves and marine

410

currents. These areas of bottom called erosion areas do not retain fine sedimentary particles.

411

The fine material fraction is diluted from deposits and transferred by currents to deeper and

412

calmer areas (accumulation areas). Since heavy metals show affinity for fine particles and

413

organic matter, they do not accumulate in shallow sediments (Szefer, 2002). This is very

414

visible in our heavy metal concentration data. Study areas located between isobaths of 20 m

415

and 40m were characterized by the lowest concentration of heavy metals, often below the

416

natural environmental background. The Kruskal-Wallis test showed very significant

417

differences in heavy metal concentrations and sediment properties (pelite fraction content and

418

organic matter content) between shallow and deeper areas. Thus, it may be concluded that

419

shallow sandy sediments of the Southern Baltic are unpolluted.

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4.2.2 Deeper areas: Słupsk Furrow and Bornholm Basin

The largest contamination by heavy metals is visible at the outer part of Bornholm

423

Basin and in Słupsk Furrow, the deepest regions studied. Station 35, in the Bornholm Basin,

424

was particularly enriched in heavy metals. The highest accumulation of heavy metals in this

425

area is connected to the large fraction of pelite sediments and large proportion of organic

426

matter at those stations. It has often been reported that heavy metals are absorbed on fine

427

sediment particles and thus deeper areas, characterized by fine-grained sediments, are a sink

428

for heavy metals and radionuclides (Szefer, 2002; Uścinowicz, 2011; Zaborska, 2014,

429

Zalewska et al., 2015). In particular, Bornholm Basin, the deepest part of Gdańsk Basin and

430

the Gotland Basin are regarded as the most polluted regions of the Baltic Sea.

432 433

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4.3 Historical contamination of sediments by heavy metals and 137Cs

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The Baltic Sea receives heavy metals mainly by atmospheric and riverine transport

434

(Schneider et al., 2000). The history of metal emission differs depending on the metal source.

435

The largest metal emission from industrial sources started after the Industrial Revolution

436

(about 1900) and peaked after World War II. In the case of some metals, e.g. Pb, the industrial

437

source was not the main metal source; lead emission in Europe was highest in the ‘70s and

438

‘80s due to usage of leaded gasoline (Pacyna and Pacyna, 2000). Since that time both the air

439

concentration of Pb and riverine Pb flux have decreased significantly due to the introduction

440

of unleaded gasoline (Institute of Meteorology and Water Management reports). The Pb

441

concentration profile at our most polluted station (station 35) indicates a large concentration 13

ACCEPTED MANUSCRIPT peak in the ‘80s and ‘90s (Fig. 4). Thus, the results generally agree with the history of lead

443

emission. The Pb concentration, which reaches 86 µg·g-1 at that station, is twice as large as the

444

geochemical background concentration for Poland (Tab. 3). A smaller Pb concentration peak

445

is visible earlier (in the ‘30s and ‘40s) at station 36. It is earlier than expected but may happen

446

when the sampling resolution is too low (5 layers per core were studied). The 5 cm sediment

447

layer may in this case represent the mean Pb concentration for 40 years of accumulation. At

448

station 32, little sub-surface increase in Pb concentration is visible, while at station 30, Pb is

449

present at its largest concentration in the surface-most layer. In both cases the increase of Pb

450

contamination is similar to the regional geochemical background values (Tab. 3), thus these

451

sediments are little polluted. At stations 33 and 34, the concentrations of Pb are constant

452

through the whole core. Moreover, concentrations ranging from 16-24 µg·g-1 are very low and

453

similar to the UCC background value (14-19 µg·g-1) (Tab. 3). It may thus definitely be said

454

that these stations are unpolluted by Pb. The arsenic concentrations show very constant depth

455

profiles, without any enrichment after the Industrial Revolution (~1900) being visible (Fig. 4).

456

At most stations As concentration of about 5 µg·g-1 was measured and this is in the lower

457

range of the geochemical background As concentration. At one station (35), the As

458

concentration is about 17-18 µg·g-1. These values are within the regional geochemical

459

background but are higher than the mean UCC values (Tab. 3). Since no enrichment is visible

460

in the last century, this As level may be attributed to the natural As concentration. The Zn

461

concentration profiles vary for different stations (Fig. 4), with most of the measured Zn

462

concentrations being within the regional background concentration range (Tab. 3). At stations

463

33 and 34, the Zn concentrations even decrease in the top sediment layers. The largest peak of

464

concentrations was measured in sediments deposited in the ‘70s-‘80s at station 35. Moreover,

465

little increase in Zn concentration is visible in the surface-most sediment layer. The Cu

466

profiles are quite constant (20-30 µg·g-1) (Fig. 4) and similar to regional background (Tab. 3)

467

at most stations, the exception again being station 35. At station 35 the Cu reached 71 µg·g-1

468

in sediment layers dated to the ‘70s-‘80s, and a slow decrease of Cu concertation is recently

469

noticeable at this station. The Cd concentration varied little through the sediment core (Fig.

470

4). The most interesting profile, at station 33, exhibits very constant Cd concentration (0.15-

471

0.17 µg·g-1). The largest peak of Cd concentrations was noted for station 35, located in the

472

Bornholm Basin, for sediments deposited in the ‘70s-‘80s. The largest historical

473

contamination was noted at station located in Bornholm Basin. Which is probably due to the

474

greater depth of the location and the sediment geochemistry (high Morg content, fine

475

sediments).

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14

ACCEPTED MANUSCRIPT The history of heavy metal contamination of the southern Baltic Sea in dated marine

477

sediments was recently studied by Zaborska (2014) (6 stations in Gdańsk Bay and Basin) and

478

Zalewska et al. (2015) (1 station in Gdańsk Basin, 1 station in Bornholm Basin and 1 station

479

in Gotland Basin). Zaborska (2014) reported the highest metal concentrations measured in

480

layers deposited in the 1960s and 1970s, thus later than in the case of our stations. However,

481

in her study, there were also some differences in sediment age of metal peak concentrations at

482

different locations. For instance, the Cd concentration profile showed a sharp peak in layers

483

deposited in the early ‘60s at stations located closest to the Vistula river. The Cd peak at

484

other stations was present in sediments deposited in the ‘70s and ‘80s. Zalewska et al. (2015)

485

reported that the largest concentrations of heavy metals were measured in sediments from the

486

‘80s. The authors indicated a very slight decrease of Hg, Pb and Zn concentrations in

487

sediments dated to 2000 due to the reduction of emissions of these metals . In the Bornholm

488

Deep, the maximum metal concentrations (Cd = 1.21 µg·g-1 and Hg = 0.15 µg·g-1) were

489

measured in sediments deposited from the ‘80s to modern times (Zalewska et al., 2015). The profile of

490

137

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Cs activity concentrations was used to validate

210

Pb dating. The

491

137

492

concentration was noted for 1963, the period of intense nuclear weapon tests. In the Baltic Sea

493

region however, the greatest role in the environment contamination is played by the

494

Chernobyl accident in 1986. The signal from this accident usually overwhelms previous

495

contamination by this radionuclide (Knapińska-Skiba et al., 2001; Zaborska et al., 2014).

496

profile of

497

137

498

history of this radionuclide accumulation in the Baltic Sea region (Fig. 4). The well

499

pronounced peak of 137Cs activity concentration at station 30 and 33 was measured later than

500

expected – about 2005. Similarly the maximum

501

surface sediments of stations 33 and 35 and almost constant

502

station 34 may be caused by post-depositional processes or very low sediment accumulation

503

rate. Post-depositional sedimentary processes include eg. physical and biological mixing of

504

sediments, sediment erosion and re-deposition. Such a situation may be also caused by

505

ongoing flushing of heavily contaminated Baltic Sea catchment by rivers and ground water

506

discharge. Moreover,

507

concentration peaks do not always correspond with the dates of the largest radionuclide

508

deposition. The largest

509

sediments were already reported for the Baltic Sea. The recently increasing

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Cs was introduced to the northern hemisphere in 1945. The highest atmospheric

137

Cs activity concentration against sediment deposition year exhibits a peak of

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Cs concentration in ‘80s in case of stations 32 and 36, and thus corresponds well with the

137

137

Cs activity concentration measured in 137

Cs activity concentration at

Cs may migrate in sediments with pore waters and thus activity

137

Cs activity concentrations in the surface-most recently deposited 137

Cs activity 15

ACCEPTED MANUSCRIPT 510

concentrations were also reported by Mattila et al. (2006) and Zalewska et al. (2015). The

511

137

512

Bornholm Basin is observed in recent sediments (Zalewska et al., 2015).

Cs activity concentration increase till 220 Bq/kg in Gdańsk Basin and 70 Bq/kg in

513

515

5. Summary and conclusions Heavy metal concentrations and

137

Cs concentrations were measured to determine

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514

their distribution in the sediments of the middle part of the Polish coast and to study precisely

517

the history of contamination by heavy metals in selected regions. A better understanding of

518

pollutant distribution in this poorly studied region will help to inform decisions regarding

519

wind park localization. This is particularly important because during construction of OWE

520

installations some amount of contaminants may re-enter the seawaters. At most studied

521

stations the heavy metal concentrations were very low and did not exceed the natural

522

geochemical background. This study’s results will be used in a hydrodynamic model to

523

estimate the level of possible re-introduction of deposited contaminants back to the water

524

column and possible directions of contaminant transport. It may, however, be concluded that

525

shallow sandy sediments of the Southern Baltic are unpolluted and construction of OWE

526

facilities will not cause any additional contamination of the environment. Only in sediments

527

deposited at the deeper bottom areas (Słupsk Furrow and Bornholm Basin) was some

528

contamination by heavy metals visible. These regions may be regarded as moderately polluted

529

and therefore not recommended for OWE exploitation. The highest

530

concentrations were measured in sub-surface sediments, and thus any disturbance of the

531

bottom may re-introduce this recently buried isotope to the seawater.

532

534

Cs activity

Acknowledgments

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This research work was supported by The National Centre for Research and

535

Development, Poland, under the project

PBS1/A6/8/2012 'AQUILO'

(www.morceko-

536

aquilo.pl). We are grateful to the captains and crews of of r/v Oceania and r/v Santa Barbara

537

for their support and assistance at sea during sampling for the project. M.Sc. Jolanta

538

Lewandowska is thanked for help with vacuum drying of sediment samples.

539 540

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Zalewska T., Woroń J., Danowska B., Suplińska M., 2015. Temporal changes in Hg, Pb, Cd and Zn environmental concentrations in the southern Baltic Sea sediments dated with 210Pb method. Oceanologia 57 (1), 132–143.

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645 646 647 648 649 650 651 652 653 654 655 656 657 658 659 660 661 662 663 664 665 666 667 668 669 670 671 672 673 674 675 676 677 678 679 680 681

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ACCEPTED MANUSCRIPT Table 1. The description of sampling region: coordinates, water depth (m), organic matter content (Morg) (%), pelite fraction content (PF) (%) and sediment type evaluation based on mean sediment grain size. Region

Depth (m)

Longitude

Latitude

Morg [%]

PF [%]

Sediment type

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46

Słupsk Furrow Słupsk Furrow Central Bank Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Bank Słupsk Bank Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Furrow Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Bank Słupsk Furrow Słupsk Furrow Kołobrzeg Słupsk Furrow Słupsk Furrow Bornholm Basin Słupsk Furrow Bornholm Basin Słupsk Bank Kołobrzeg Kołobrzeg Kołobrzeg Słupsk Bank Słupsk Furrow Central Bank C Słupsk Furrow Stilo Bank

79.7 74.5 48.1 61.4 66.7 72.7 77.8 56.4 80.1 67.7 65.8 70.8 82.1 23.3 48.5 44.8 58.4 55.3 52.3 47.5 50.4 56.9 45.1 42.0 38.1 36.3 25.4 30.9 23.3 71.7 76.5 56.0 78.5 59.0 84.2 65.1 68.5 45.0 27.1 57.5 26.0 26.2 49.3 39.5 58.3 36.0

17°31`19 17°58`43 17°46`16 17°45`40 17°45`24 17°45`19 17°45`34 17°39`01 17°52`05 17°57`24 17°58`45 17°01`18 17°28`55 17°02`04 17°01`17 17°06`20 17°14`48 17°19`23 17°23`25 17°28`09 16°31`33 16°35`02 16°45`18 16°48`14 16°51`03 16°45`14 16°45`07 16°45`18 16°45`08 17°04`57 17°45`37 15°45`69 16°47`13 16°23`38 16°06`48 17°25`99 16°04`65 16°02`49 15°13`29 15°22`04 15°07`47 16°47`82 17°12`92 17°15`33 17°41`04 17°46`04

55°20`33 55°20`35 55°39`18 55°34`14 55°29`57 55°25`59 55°21`17 55°29`45 55°30`15 55°30`18 55°39`03 55°19`21 55°19`31 55°00`45 55°08`35 55°08`30 55°08`21 55°08`24 55°08`28 55°08`28 55°08`07 55°08`08 55°08`06 55°08`11 55°08`13 55°05`56 55°02`30 55°04`21 55°00`33 55°18`07 55°17`28 54°36`54 55°13`65 55°09`09 55°09`08 55°22`49 54°57`22 54°41`19 54°23`54 54°36`42 54°24`03 55°03`65 55°05`98 55°27`54 55°29`99 55°02`97

3.3 2.8 0.3 3.3 6.2 5.7 2.9 0.7 6.0 5.3 1.9 5.1 3.3 0.1 2.0 1.8 0.9 0.8 2.1 0.5 3.1 1.0 0.8 0.3 0.5 0.8 0.8 0.6 0.4 9.8 2.6 8.1 2.1 2.9 8.6 7.8 12.1 1.2 0.4 9.3 0.3 2.7 0.6 0.4 1.7 0.5

10 11.4 10.6 2.2 48 35.2 12.8 17.8 20.2 20.0 29.5 51.7 8.2 0.2 4.4 4.4 15.4 4.3 10.3 0.6 18.3 5.1 3.6 0.3 1.1 1.7 3.8 19.8 0.8 66.9 16.6 54.6 17.6 36.5 88.8 59.6 71.7 9.0 0.0 63.6 0.0 0.0 2.3 2.5 33.9 3.6

sand sand sand sand sandy silt silty sand sand silty sand silty sand silty sand silty sand sandy silt sand sand sand sand silty sand sand sand sand silty sand sand sand sand sand sand sand silty sand sand sandy silt silty sand sandy silt silty sand silty sand sandy silt sandy silt sandy silt sand sand sandy silt sand sand sand sand silty sand sand

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ACCEPTED MANUSCRIPT Table 2. Upper part of the table: Pearson’s correlation coefficients between measured metals concentrations and sediment descriptors: organic matter (Morg) content and pelite fraction (PF) content. Surface and sub-surface sediments are included into dataset (n=91). The correlations with p < 0.05 were significant (bolded). Correlations with “r” higher or equal to 0.8 are considered to show high correlation (marked with a star). Lower part of the table: the results (H, p values) of Kruskal-Wallis test showing differences in heavy metals concentrations and sediment properties between shallow and deeper regions of studied area. As 0.24 0.36 0.25 0.27

Hg 0.65 0.74 0.37 0.64 0.25

17.94 <0.0001

17.74 <0.0001

16.53 <0.003

6.73 <0.035

14.29 <0.0008

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Cu 0.88* 0.84* 0.52 0.86* 0.32 0.80*

Morg 0.81* 0.91* 0.49 0.80* 0.48 0.77 0.73

PF 0.58 0.72 0.46 0.65 0.53 0.74 0.80* 0.82* 17.5 <0.0002

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Pb 0.77 0.84* 0.43

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Cd 0.50 0.53

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Fe Zn Cd Pb As Hg Cu Morg H 18.02 p <0.0001

Zn 0.93*

23.0 <0.0001

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Fe

18.4 <0.0001

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32-43

200

Hg [µg·g-1]

0.01-0.05

0.02-0.08

1

As [µg·g-1]

1.5-5.1

5-20

30

Cd [µg·g-1]

0.07-0.1

0.13-0.24

Zn [µg·g-1]

52-71

70-96

Cu [µg·g-1]

25-50

6.7 - 86.1

0.001 - 0.019 0.28 - 22.66

7.5

0.03-1.32

1000

7.9 - 123.1

150

1.0 – 71.3

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Pb [µg·g-1]

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Table 3. Table contains the natural concentrations and norms used in Poland. The U.C.C natural concentration is the concentration of heavy metals in the rocks of the Upper Continental Crust measured worldwide. Regional geochemical background for Poland was estimated using deeper part of sediment cores that were deposited before industrial revolution. The threshold for polluted sediments is given the Journal of Laws (2002). Regional U.C.C. natural Polish pollution geochemical concentration Results of this study Heavy metal background threshold [µg·g-1] (Rudnick and (J.L., 2002) (Uścinowicz, Gao, 2001) 2011)

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Figure 1. The sampling stations (red dots) localized within the area preliminary selected for

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OWE acquisition. Bottom type is given on the map.

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Figure 2. The activity concentrations of the total 210Pb (Bq·kg-1) activity concentrations at 7 selected stations (30, 31, 32, 33, 34, 35, 36). Grey solid line indicates 210Pb supported. Obtained sediment accumulation rates (ω) are added to the graphs.

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Figure 3. The concentrations of selected heavy metals (Hg, As, Pb and Cd) (in µg·g-1, Hg in ng·g-1) in sediments of the southern Baltic Sea, the area selected for wind farm acquisition.

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Figure 3 (continued). The concentrations of selected heavy metals (Zn, Cu, Fe) (in µg·g-1) and the activity concentration of radioactive isotope (137Cs) (Bq·kg-1), in sediments of the southern Baltic Sea, the area selected for wind farm acquisition.

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Figure 4. History of heavy metals and 137Cs accumulation. The vertical concentrations of selected heavy metals (Hg, As, Pb, Cd, Zn, Cu) and the activity concentration of 137Cs in dated sediment cores (stations: 30, 32, 33, 34, 35, 36) collected from the southern Baltic Sea, the area selected for wind farm acquisition. The measurement uncertainty is below 5 % for heavy metals concentrations and below 15 % for 137Cs activity concentration.

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Shallow areas of the Baltic Sea were characterized by low heavy metals concentrations Largest contamination was found at deepest regions: Bornholm Basin and Słupsk Furrow The OWE construction in shallow sandy regions will not cause any additional pollution Deeper regions were moderately polluted and not recommended for OWE acquisition