Deciphering the role of water column redoxclines on

Confidential manuscript submitted to Global Biogeochemical Cycles

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Deciphering the role of water column redoxclines on methylmercury cycling using speciation modeling and observations from the Baltic Sea

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A.L. Soerensen1*, A.T. Schartup2, A. Skrobonja3, S. Bouchet4,5, D. Amouroux5, V. Liem-

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Nguyen3,6, E. Björn3*

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Stockholm, Sweden

Stockholm University, Department of Environmental Science and Analytical Chemistry,

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MA, USA

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Umeå University, Department of Chemistry, Umeå, Sweden

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Present address: ETH Zürich, D-USYS department, Universitätstrasse 16, CH-8092 Zürich,

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Switzerland

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pour l'Environnement et les Materiaux, IPREM UMR5254, MIRA, 64000, Pau, France

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Harvard University, John A. Paulson School of Engineering and Applied Sciences, Cambridge

CNRS/ UNIV PAU & PAYS ADOUR, Institut des Sciences Analytiques et de Physico-chimie

School of Science and Technology, Örebro University, SE-701 82 Örebro, Sweden

*corresponding author: [email protected] *corresponding author: [email protected]

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Key Points:   

Iron and Sulfur cycling in the redoxcline explain particulate and dissolved mercury concentration peaks found in the water column High in situ mercury methylation explains elevated methylmercury concentrations in anoxic water We present a conceptual model describing cycling and speciation of mercury across the redoxcline

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Abstract

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Oxygen depleted areas are spreading in coastal and offshore waters worldwide but the

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implication for production and bioaccumulation of neurotoxic methylmercury (MeHg) is

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uncertain. We combined observations from six cruises in the Baltic Sea with speciation modeling

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and incubation experiments to gain insights into mercury (Hg) dynamics in oxygen depleted

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systems. We then developed a conceptual model describing the main drivers of Hg speciation,

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fluxes and transformations in water columns with steep redox gradients. Methylmercury

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concentrations were 2-6 and 30-55 times higher in hypoxic and anoxic than in normoxic water,

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respectively, while only 1-3 and 1-2 times higher for total Hg (THg). We systematically detected

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HgII methylation in anoxic water but rarely in other waters. In anoxic water, high concentrations

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of dissolved sulfide causes formation of dissolved species of divalent inorganic Hg (HgII):

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HgS2H-(aq) and Hg(SH)20(aq). This prolongs the lifetime and increases the reservoir of HgII readily

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available for methylation, driving the high MeHg concentrations in anoxic zones. In the hypoxic

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zone and at the hypoxic-anoxic interface, Hg concentrations, partitioning and speciation are all

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highly dynamic due to processes linked to the iron and sulfur cycles. This causes a large

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variability in bioavailability of Hg, and thereby MeHg concentrations, in these zones. We find

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that zooplankton in the summertime are exposed to 2-6 times higher MeHg concentrations in

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hypoxic than in normoxic water. The current spread of hypoxic zones in coastal systems

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worldwide could thus cause an increase in the MeHg exposure of food webs.

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1. Introduction

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Methylmercury (MeHg) is a bioaccumulative neurotoxin whose production is tightly

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coupled to anaerobic microbial activity at redox transition zones in many ecosystems (Benoit et

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al., 1999; Compeau and Bartha, 1985; Schaefer al., 2014). In water columns, elevated

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concentrations of both total mercury (THg) and MeHg have been measured in hypoxic ( 100 m)(Figure 1; Myrberg and Lehmann, 2013). Similar to other

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coastal systems, it receives a high freshwater input from runoff (runoff equals 2-3% of the total

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Baltic Sea water reservoir each year; Hansson, et al., 2011a; Myrberg and Lehmann, 2013).

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Runoff typically carries excess nutrients and organic matter (OM) from agricultural impacted

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soils and sewage treatment facilities to coastal ecosystems. Phytoplankton blooms induced by

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excess OM and nutrient runoff increase microbial activity and the system’s biological oxygen

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demand (BOD), which can result in large permanently hypoxic and anoxic zones in sub-

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halocline offshore water. This is seen in the Baltic Sea and also in other systems like the Gulf of

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Mexico, the East China Sea and the Black Sea (Breitburg et al., 2018; Diaz and Rosenberg,

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2008). In the southern Baltic Sea (Baltic Proper), these zones cover 35-40% of the surface area

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(Conley et al., 2011; Diaz and Rosenberg, 2008; Hansson et al., 2011b). The Baltic Sea is

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characterized by changes in microbial communities along the salinity gradient (north-south)

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(Herlemann et al., 2011) and across the stratified layers of the redox gradient (Thureborn et al.,

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2013) as the electron acceptors transition from oxygen to nitrate, nitrite, manganese(IV,III),

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iron(III), and sulfate (Yakushev and Newton, 2013). The majority of known microbes capable of

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Hg methylation are anaerobes (including iron and sulfur reducers, and methanogens; Compeau

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and Bartha, 1985; Gilmour et al., 1992; Gilmour et al., 2013) that thrive in low and no oxygen

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conditions; such conditions are found in the hypoxic and anoxic waters of the Baltic Sea and

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represent potential hotspots for MeHg production. While the marine food web is predominantly

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located in the normoxic waters (Reissmann et al., 2009; Webster et al., 2015), zooplankton in the

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Baltic Sea have been shown to temporarily spend time in hypoxic water to avoid predators, and

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some fish (e.g. cod) forage briefly but often in hypoxic water (Neuenfeldt et al., 2009; Webster et

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al., 2015). Thus, hypoxic and anoxic zones have the potential to influence the food web MeHg

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burden through the transport of divalent inorganic Hg (HgII) or MeHg into normoxic water

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(through molecular diffusion, small scale lateral intrusions and larger scale vertical winter

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mixing; Dellwig et al., 2012; Reissmann et al., 2009) or directly through biota feeding in hypoxic

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water.

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Organic matter from runoff and decomposing plankton binds surface water Hg and

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transports it to deeper waters. There it is released during microbial OM remineralization

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providing a HgII reservoir potentially available for methylating microbes present at and below

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the redoxcline. Some previous studies of Hg at the redoxcline in the Black Sea found large

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increases in the concentration of dissolved THg (maximum concentrations >6 pM) at the

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hypoxic-anoxic interface (Cossa and Coquery, 2005; Lamborg et al., 2008). These THg peaks

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were linked to the dissolution of iron (Fe) oxides (that trap OM and Hg in their aggregates) when

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reaching low oxygen waters at the hypoxic-anoxic interface. On the other hand, data available on

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MeHg concentrations from hypoxic and anoxic waters in coastal seas present contradictory

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profiles. Malcolm et al. (2010) did not find elevated MeHg concentrations in unfiltered hypoxic

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waters of the Arabian Sea while, for the Black Sea, a dissolved MeHg peak (>1 pM) was

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reported in the hypoxic layer of the Black Sea by Lamborg et al. (2008) but in the anoxic layer

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by Rosati et al. (2018). In the Baltic Sea, elevated MeHg concentrations have been measured in

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both filtered (0.2 m filter) and unfiltered hypoxic water but the highest concentrations were

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found in anoxic water (Kuss et al., 2017; Soerensen et al., 2016). These differences in

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observations could be controlled by differences in salinity, O2, sulfide and Fe chemistry or by the

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presence/absence of anoxic zones adjacent to the hypoxic zones that influence the vertical

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distribution of both HgII methylating microbial communities and the chemical speciation, and

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thus bioavailability, of Hg. However, no systematic incubation experiments to determine rates

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and drivers for HgII methylation in hypoxic and anoxic zones have been conducted. Furthermore,

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in coastal seas, it is expected that HgII and MeHg will predominantly form complexes with

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reduced sulfur ligands, i.e. organic thiols and inorganic sulfide (Skyllberg, 2008). The

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distribution between Hg-thiol and Hg-sulfide complexes, as well as the partitioning of Hg

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between the particulate and the dissolved phases, could vary between zones with different redox

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potential and cause differences in Hg reactivity and bioavailability. While the chemical

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speciation of Hg could be of great importance for Hg cycling and exposure to biota in coastal

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seas with water column redox gradients, these processes are understudied (Fitzgerald et al., 2007;

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Krabbenhoft and Sunderland, 2013).

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The main objective of this study is to assess the impact of hypoxic and anoxic conditions

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on the production and fate of MeHg in the water column of coastal seas. To do this, we measured

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THg and concentrations of Hg species (HgII, dissolved elemental Hg (Hg0), MeHg, and

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dimethylmercury (DMHg)) and conducted incubation experiments to estimate HgII methylation

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and MeHg demethylation rate constants during six cruises in summer/fall (2014-2016) in the

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Baltic Sea. We use thermodynamic modeling to determine the chemical speciation of HgII and

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MeHg. Finally, we use these data to construct a conceptual model for Hg speciation and fate that

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is generally applicable to marine water columns with a redoxcline.

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2. Method

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2.1 Field sampling. Figure 1 shows the sampling locations in the Baltic Sea (Northern

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Europe) during our six cruises: September 2014 (SEP14-S; data published in Soerensen et al.

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(Soerensen et al., 2016)), July 2015 (JUL15-S) and July 2016 (JUL16-S) in the Southern Baltic

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(Belt Seas and Baltic Proper), and September 2014 (SEP14-N), August 2015 (AUG15-N) and

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August 2016 (AUG16-N) in the Northern Baltic Sea (Bothnian Sea and Bay; a subset of this data

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was published in Soerensen et al. (Soerensen et al., 2017)). The number of stations and samples

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collected during the cruises are given in Table S1.

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Water samples were collected using 2 L Niskin bottles (attached to a rosette) or 5 L

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Teflon-lined General Oceanic (GO-FLO) bottles following trace-metal clean protocols (Gill and

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Fitzgerald, 1987). In 2015, overlapping samples were taken with the two types of bottles and the

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THg concentrations in samples were compared (SI Figure S1). THg samples were collected in

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125 mL Teflon® (Nalgene PFA or FEP) or glass bottles (I-CHEM certifiedTM 300 series) and

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MeHg (monomethylmercury (MMHg)+dimethylmercury (DMHg)) samples were collected in

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125 or 250 mL Teflon® or HDPE bottles all of which were pre-cleaned with trace metal grade

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HCl (Merck Suprapur HCl). In 2015 and 2016, a subset of the MeHg samples were filtered

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(MeHgD) either in a portable laminar flow hood using 0.22 μm hydrophilic PTFE filters and

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Teflon filtration units or using Sterivex 0.22 μm filter units. All MeHg samples were acidified to

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0.06-0.10 M HCl (using Fisher Optima or Merck Suprapur). All samples were spiked with

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isotopically enriched MMHg for isotope dilution analysis (for information on isotopes and

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spiking levels see SI Table S2). For AUG16, THg samples were not acidified and were analyzed

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within a month of sampling (Soerensen et al., 2017); for JUL15 and JUL16 THg samples were

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preserved using 0.5% trace metal grade HCl. All samples were stored in the dark at 4°C on the

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ship and in the lab until they were processed. For the dissolved gaseous Hg species, samples

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were gently transferred from the GO-FLO bottles to 0.5 L Teflon® purging bottles using trace

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metal clean Teflon® tubing and were immediately purged on board (Bouchet et al., 2011).

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2.2 Total Hg analyses. For the JUL15-S cruise, samples were digested with bromine

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monochloride (BrCl) and prereduced with hydroxyl-amine hydrochloride (NH2OH·HCl),

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followed by addition of stannous chloride (SnCl2) to reduce HgII to Hg0 for analysis on a Tekran

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2600 following USEPA 1631 (USEPA, 2002). The method detection limit was 0.05 pM based on

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three times standard deviation of blanks. Matrix spike recoveries averaged 99% (n=3). Sample

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duplicates (RSD=9.8%, n=9) and standards for precision and recovery (103%, n=7) were

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measured every 8-10 samples. The AUG16-N cruise samples were analyzed following the same

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procedure as JUL15-S (see also Soerensen et al., 2017) but with a method detection limit of 0.2

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pM and a measurement uncertainty of 14 % (1.25 pM), respectively. For

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the JUL16-S cruise, the THg concentrations were calculated as the sum of the gaseous Hg0,

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unfiltered HgII and unfiltered MMHg concentrations.

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2.3 Dissolved Gaseous Hg species analyses. Hg0 and DMHg were purged from the

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samples onto gold-coated sand traps and carbotraps, respectively, and analyzed according to

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protocols previously described (Amouroux et al., 1998; Bouchet et al., 2011). Teflon® purging

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vessels (0.5 L) were bubbled for 30 min with a Hg-free nitrogen flow (c.a. 400 mL min−1) and

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the gas stream was dried through a moisture trap before species trapping. Gold traps were

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analyzed in the lab (within 3 weeks) by double amalgamation on gold (method detection limit ~2

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fM) while carbotraps were analyzed by cryogenic trapping–gas chromatography– ICPMS

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(method detection limit ~1 fM). RSD based on standard replicates was 7.5%.

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2.4 HgII and Methylated Hg analyses. In 2014, 2015 and AUG16 MeHg samples were

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analyzed as described in Soerensen et al. (2017). Analysis was done by isotope dilution analysis

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using thermal desorption gas chromatography inductively coupled plasma mass spectrometry

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(TDGC-ICPMS) after derivatization with sodium tetraethylborate and purge and trap on Tenax®

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adsorbent (Lambertsson and Bjorn, 2004; Munson et al., 2014). The sample concentrations were

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blank corrected by subtracting the difference (10 fM) between sample collection blanks (205

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fM, n=12) and MQ water sample levels (107 fM, n=43). The limit of detection (LOD) was 13

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fM based on three times standard deviation of replicated field blanks (n=12). The RSD of

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triplicate field samples collected from the same water depths was on average 30%. On the JUL16

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cruise, HgII and MMHg concentrations were determined by GC-ICPMS according to Bouchet et

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al. (2013) (LOD ~100 fM and ~1 fM, respectively; RSD=7% based on sample triplicates).

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Analysis was done by isotope dilution analysis after derivatization with sodium tetrapropylborate

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and liquid-liquid extraction into isooctane. After a vigorous shake, the organic phase was

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recovered and injected to the GC-ICPMS in triplicate. A laboratory intercomparison of the two

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methods shows a strong agreement (R2=0.96, n=4) with 35% difference between the two

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laboratories and methods (SI Figure S2). The DMHg fraction of the MeHg concentration is small

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in the Baltic Sea (see Results) and in this study we assume that the MeHg concentration is

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equivalent to the MMHg concentration.

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2.5 Incubation experiments. Incubation experiments to determine rate constants for HgII

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methylation (km) and MeHg demethylation (kd) in light exposed (kd-photo) and dark (kd-dark)

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samples were performed onboard (see Table S1 for number of samples on each cruise) and

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followed the experimental and data treatment procedures of Rodriguez-Gonzalez et al. (2013).

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Light incubations were performed in 2015 and 2016 on surface water samples from 5-10 m depth

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and in a few cases from 20 m, while dark incubations were performed all three years for water

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samples from up to three depths (60m). Subsamples were collected in 125 ml

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Teflon or HDPE bottles but all light incubations were done using Teflon bottles. For the photo-

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incubation experiments, all subsamples were incubated outside in an open tank with a constant

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flow of seawater with dark and control samples wrapped in aluminum foil. For the additional

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dark incubations, samples collected in water with an ambient temperature >8 °C were kept at

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room temperature (20°C) and samples collected in water with an ambient temperature LOD; 2015:

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0.03±0.0005 d-1 with 25% of samples > LOD; 2016: 0.05±0.002 d-1 with 74% of samples >

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LOD) were calculated according to Hintelmann and Evans (1997). For km the detection limit was

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calculated based on three times standard deviation of the determined Me198Hg concentration in

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all t=0 samples. For the methylation rate constants, the detection limits were 4×10-4±1×10-4 d-1

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(2014), 2×10-4±1×10-4 d-1 (2015) and 12×10-4±3×10-4 d-1 (2016; Figure S3).

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2.6 Ancillary data. QC/QA’ed data on temperature, salinity, oxygen, sulfide,

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phosphorous, nitrate, ammonium, silicate, turbidity, and chlorophyll a were collected on the

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cruises (available for download from SMHI (SMHI, 2018)). We use the term S-II for the total

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concentration of dissolved sulfide [H2S(aq)] + [SH-(aq)] + [S2-(aq)], where [S2-(aq)] is negligible at the

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relevant pH. We define the switch from normoxic to hypoxic conditions at the depth where the

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O2 level goes below 2.0 ml L-1 (Conley et al., 2009; Diaz and Rosenberg, 2008). We include one

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observation with an O2 concentration of 2.5 ml L-1 to the hypoxic pool due to other

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characteristics (BY38 50 m: concentrations of NO2- and NO3- and observed THg peaks otherwise

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seen in the hypoxic layer). In the Baltic Sea, the presence of O2 and S-II can overlap due to

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density driven small scale lateral intrusions causing water to mix across the hypoxic-anoxic

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transition zone (Berg et al., 2013; Dellwig et al., 2012). We see such overlap in our samples and

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therefore define the switch from hypoxic to anoxic conditions based on the presence of S-II above

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the detection limit (1 μM) (Soerensen et al., 2016). The shallowest depth with detectable S-II is

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also where we find the turbidity maximum (Neretin et al., 2003; Pohl and Fernández-Otero,

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2012).

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2.7 Statistics. We use Kruskal-Wallis nonparametrix test to determine statistical

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significant differences between different reservoirs. The significance level was set at 0.05 unless

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otherwise stated.

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2.8 Chemical speciation modeling. The Solgaswater (WinSGW) software (Karlsson, M

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and Lindgren) was used to model the chemical speciation of HgII and MeHg across the summer

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redoxcline. We adapted a model developed by Liem-Nguyen et al. (2017a) and a full description

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of the model used in our study is given in SI Tables S3 and S4. As input to the model we used

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information on HgII, MeHg, and S-II concentrations from the BY32 station in 2016 to represent a

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typical redox profile. For pH we used data from the close by BY31 station (SMHI, 2018) and

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dissolved organic carbon (DOC) concentrations were from Nausch et al. (2008) for the Gotland

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Sea. We derived a total thiol to DOC mass fraction of 0.15% from previous studies (Liem-

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Nguyen et al., 2017a; Skyllberg, 2008) and used this to calculate an average total dissolved thiol

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concentration of 0.25 μM.

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3. Results

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3.1 Hydrography, ancillary parameters and interannual variability. Figure S4 shows the

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spatial variability of temperature, salinity and oxygen in 2014-2016. The thermocline was found

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at 15-25 m depth across most of the Baltic Sea during the cruises except in the northern Belt Seas

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where it was located at 40 m (P2 and Anholt; SI Figure S4)(Myrberg and Lehmann, 2013). The

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halocline was at 40-80 m depth and strongest in the Baltic Proper where the deep poorly

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ventilated areas are located (Figure 1; Myrberg and Lehmann, 2013). In the Baltic Proper, a

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redoxcline was observed below the halocline at about half the stations (red and purple stations in

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Figure 1). Figure 2 (and Figure S4) shows how the redoxcline was destabilized at stations BY11

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and BY15 (purple stations) in 2015 and 2016. In December 2014, after the first cruise, a rare

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large inflow of dense, oxygen rich water with high salinity entered the Baltic Sea through the

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Belt Seas (Hansson and Andersson, 2015). The inflowing water travelled along the south-eastern

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part of the Baltic Proper and reached BY20) by July 2015 (a 6-8 months transport time from the

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Belt Sea to the eastern Baltic Proper). After this, it became indistinguishable in the water column

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at BY20, while the destabilised redoxcline slowly rebuilt at station BY11 and BY15. No effect

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was seen on BY32 and BY38 in the western part of the basin (Hansson and Andersson, 2015;

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Kuss et al., 2017).

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3.2 Total Hg concentrations and species distribution.

Figure 3 shows the spatial

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distribution of THg and MeHg averaged across cruises and Table 1 presents average

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concentrations of THg and Hg species for all cruises (a full dataset with the observations can be

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found in the SI Dataset S1). THg in individual samples ranged from 0.5 to 10.7 pM (average

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2.0±1.8; n=91). Both THg and HgII concentrations were significantly higher in hypoxic water

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compared to normoxic water in the Baltic Proper (Table 1; Table S5). Indeed, peaks between 4.9

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and 10.7 pM were often observed in the hypoxic zone of stations with a redoxcline, as

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exemplified in Figure 2 (E, F; peaks are here defined as concentrations 100% higher than the

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average of observations at the two closest depths). Among the stations presenting a steep

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redoxcline (red and purple stations on Figure 1), half show such peaks. The absence of peaks in

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some profiles is likely due to inadequate vertical sampling resolution of the hypoxic zone. In

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2016, THg concentrations (but not HgII) were significantly elevated in anoxic water compared to

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normoxic water while this was not the case in 2015. Figure 4 shows that the HgII:THg fraction

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increased significantly from normoxic to hypoxic water and then decreased again in anoxic

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water.

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Hg0 concentrations ranged from 0.02 to 0.77 pM and only showed a significant difference

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between normoxic and anoxic water (Table S5). Similarly, the Hg0:THg fraction showed a

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significant decrease from normoxic to anoxic water (Figure 4). Our observations are in the same

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range as previously reported in the Baltic Sea (Kuss et al., 2017; Kuss and Schneider, 2007;

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Wangberg et al., 2001).

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MeHg concentrations ranged from