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Marine Pollution Bulletin 186 (2023) 114426
Available online 5 December 2022
0025-326X/© 2022 Elsevier Ltd. All rights reserved.
Review
Mercury in the Polish part of the Baltic Sea: A response to decreased
atmospheric deposition and changing environment
Agnieszka Jędruch
a
,
b
,
*
, Lucyna Falkowska
a
,
1
, Dominika Saniewska
a
, Agnieszka Grajewska
c
,
Magdalena Bełdowska
a
, Włodzimierz Meissner
d
, El˙
zbieta Kalisi´
nska
e
, Kazimierz Duzinkiewicz
f
,
J´
ozef M. Pacyna
g
a
University of Gda´
nsk, Faculty of Oceanography and Geography, Institute of Oceanography, Marszałka J´
ozefa Piłsudskiego 46, 81-378 Gdynia, Poland
b
Polish Academy of Sciences, Institute of Oceanology, Powsta´
nc´
ow Warszawy 55, 81-712 Sopot, Poland
c
Institute of Meteorology and Water Management – National Research Institute, Jerzego Waszyngtona 42, 81-342 Gdynia, Poland
d
University of Gda´
nsk, Faculty of Biology, Wita Stwosza 59, 80-308 Gda´
nsk, Poland
e
Pomeranian Medical University, Faculty of Pharmacy, Medical Biotechnology and Laboratory Medicine, Powsta´
nc´
ow Wielkopolskich 72, 70-111 Szczecin, Poland
f
Gda´
nsk University of Technology, Faculty of Electrical and Control Engineering, Gabriela Narutowicza 11/12, 80-233 Gda´
nsk, Poland
g
AGH University of Science and Technology, Faculty of Energy and Fuels, Adama Mickiewicza 30, 30-059 Krak´
ow, Poland
ARTICLE INFO
Keywords:
Baltic Sea
Climate change
Hg
Minamata convention
Risk assessment
Temporal trends
ABSTRACT
Our review of the literature showed that since the beginning of the socio-economic transformation in Poland in
the 1990s, the downward trend in Hg emissions and its deposition in the southern Baltic Sea was followed by a
simultaneous decrease in Hg levels in water and marine plants and animals. Hg concentrations in the biota
lowered to values that pose no or low risk to wildlife and seafood consumers. However, in the rst decade of the
current century, a divergence between these two trends became apparent and Hg concentrations in sh, herring
and cod, began to rise. Therefore, increasing emission-independent anthropogenic pressures, which affect Hg
uptake and trophodynamics, remobilization of land-based and marine legacy Hg deposits, as well as the structure
of the food web, can undermine the chances of reducing both the Hg pool in the marine environment and human
Hg exposure from sh.
1. Introduction
Mercury (Hg) is an element that naturally occurs in the Earth’s crust.
However, anthropogenic activities, such as mining, fossil fuel combus-
tion, and the use of Hg in products and industry, have signicantly
accelerated its emission and release into the environment, leading to
widespread global pollution (Outridge et al., 2018; Wang et al., 2019).
Hg is one of the most potent neurotoxins known. It accumulates in or-
ganisms and biomagnies through the food chain, resulting in exposures
that affect the health of the top predators, including humans. However,
Hg also negatively affects organisms at lower trophic levels, altering
their physiology and behavior (Weis, 2014; Eagles-Smith et al., 2018).
To address the global problem of Hg pollution, the Minamata
Convention on Mercury was established in 2013 by the United Nations
(UNEP, 2013). Actions to reduce Hg emissions, such as the phase-out of
Hg and its compounds from commercial products and air pollution
control technologies, were already taken much earlier, starting around
1990. As a result, Hg concentration in the air and its atmospheric
deposition decreased signicantly (e.g., by 30–40 % between 1990 and
2010 in the northern hemisphere) (Zhang et al., 2016). In general, Hg
trends in wildlife followed those in the atmosphere; however, declines in
biotic Hg were often not proportional. In some areas, despite the
reduced atmospheric inow of Hg, its level in the organisms remained
the same or even increased (Obrist et al., 2018; Jędruch et al., 2021).
This pattern became more apparent over the past two decades, which
was a consequence of climate-driven changes in the biogeochemical Hg
cycle, as well as the release of historical Hg from marine and terrestrial
deposits (Hsu-Kim et al., 2018; Wang et al., 2019). In the Hg budget for
2018, direct anthropogenic emissions (about 2200–2500 Mg year
−1
) are
not the dominant source in the recent global Hg cycle. A more important
* Corresponding author at: University of Gda´
nsk, Faculty of Oceanography and Geography, Institute of Oceanography, Marszałka J´
ozefa Piłsudskiego 46, 81-378
Gdynia, Poland.
E-mail address: ajedruch@iopan.pl (A. Jędruch).
1
Deceased
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
https://doi.org/10.1016/j.marpolbul.2022.114426
Received 29 August 2022; Received in revised form 21 November 2022; Accepted 23 November 2022
Marine Pollution Bulletin 186 (2023) 114426
2
role is played by the reemission and remobilization of Hg accumulated
over the years on land and in the ocean (about 3900–5000 Mg year
−1
)
(Outridge et al., 2018; Zhou et al., 2021). The growing contribution of
recycled Hg, including the legacy of previous anthropogenic emissions,
is driven by climate change and increasing anthropogenic pressures
affecting all ecosystems (Reusch et al., 2018; Reckermann et al., 2022).
The consequences of the perturbed global Hg cycle are expected to be
more severe for the marine environment, compared to terrestrial and
freshwater ecosystems (Amos et al., 2013). The marine ecosystem is
particularly vulnerable to Hg contamination, as it originates from
different sources, including not only atmospheric deposition but also
waterborne input and direct point sources located on land and sea
(Outridge et al., 2018; Liu et al., 2021). Once introduced into the marine
environment, Hg remains there for a long period by cycling between
matrices and changing chemical forms. Marine organisms are more
exposed to Hg and have a higher tissue concentration than terrestrial
biota since bioaccumulation and biomagnication of Hg occur primarily
in aquatic systems (Lavoie et al., 2013). Hg uptake and trophic transfer
depend on physicochemical parameters (e.g., nutrients, dissolved
organic carbon, temperature, O
2
, pH, redox) and biological factors (e.g.,
bacterial biodiversity and activity, productivity, structure of the food
web) (Eagles-Smith et al., 2018; Jędruch et al., 2019a, 2019b; Wang
et al., 2019; C´
ordoba-Tovar et al., 2022). Therefore, the impacts of
climate-driven disturbances (e.g., water warming, acidication, uctu-
ations in hydrological cycles, increasing frequency of storms) and other
changes occurring in aquatic ecosystems (e.g., shifts in ecosystem
structure and species diversity, habitat destruction) on Hg bio-
magnication in the marine food chain are crucial (Dijkstra et al., 2013;
Bełdowska et al., 2013, 2016a, 2016b; Schartup et al., 2019; Fernandez
et al., 2022).
As shown in previous studies, the cascading effects of environmental
perturbations will lead to enhanced remobilization of the legacy pool of
Hg, as well as its accelerated accumulation in aquatic organisms in the
coming decades (Dietz et al., 2021a, 2021c; Reckermann et al., 2022). It
requires action by both the scientic community and policymakers to
determine the current status of Hg distribution and dispersion, predict
their future changes, and establish recommendations to protect human
health and ecosystems, supporting the implementation of the Minamata
Convention. Nevertheless, stabilization of Hg concentration at the
present-day level may require more aggressive reductions in primary
anthropogenic emissions than initially assumed (Amos et al., 2013;
Schartup et al., 2019). Given that efforts of the Minamata Convention
may be countered by climate change, evaluation of its effectiveness re-
quires biomonitoring of multiple species that represent different trophic
levels (Wang et al., 2019). Projecting changes in the future ocean in-
volves determining how ecological perturbations and changes in Hg
emission will impact the current Hg cycle, which, in turn, is not possible
without understanding the oceans of the past (Amos et al., 2013). Due to
its unique morphological and hydrological characteristics, exposure to
multiple stressors, and a simplied ecosystem, the Baltic Sea (BS) can
serve as a ‘time machine’ to study the consequences and mitigation of
future coastal disturbances, as well as a unique model area for other
coastal regions (Reusch et al., 2018; Dietz et al., 2021c; Reckermann
et al., 2022). The BS is also referred to as the most polluted ecosystem in
the world, although its environmental quality has improved consider-
ably over the past few decades (Dietz et al., 2021b). Given that Poland
leads the list of Hg emitters in Europe (EMEP, 2020), our literature re-
view focuses on the Polish part of the BS and aims to assess the current
state of Hg pollution and indicate the pace and direction of future
changes.
2. Material and methods
2.1. Study area
The BS (Fig. S1) is a semi-enclosed inland water body surrounded by
nine developed and industrialized countries and ve more belonging to
the catchment area, representing 85 million inhabitants. For decades,
the BS was exposed to various anthropogenic stressors, such as pollu-
tion, eutrophication, and overshing (Reusch et al., 2018; Reckermann
et al., 2022). Also, the effects of climate change, including warming of
water, rising sea level, increase in strength and frequency of extreme
wheatear phenomena, and a drastic reduction of sea ice cover became
more pronounced in recent years (R´
o˙
zy´
nski and Lin, 2021; Meier et al.,
2022). Their combined interaction will increase the Hg pool in the BS
environment, intensifying the outow of Hg from land and its remobi-
lization from marine sediments (Saniewska et al., 2014b, 2018; Beł-
dowska et al., 2016a; Jędruch et al., 2021; Kwasigroch et al., 2021), but
also increase the amount of naturally generated Hg, enhance its
methylation, and stimulate absorption (Bełdowska et al., 2016b;
Jędruch et al., 2018a; Schartup et al., 2019; C´
ordoba-Tovar et al., 2022).
The southern part of the BS, with a highly populated and predomi-
nantly agricultural catchment, is facing stronger human pressure,
compared to the mostly pristine north (Reckermann et al., 2022). The
largest rivers owing into the southern BS are the Vistula and Odra
(Fig. S1), which are the second and third largest rivers that ow into the
entire BS in terms of catchment area (HELCOM, 2015). Poland is located
centrally in the southern strip of the BS, and the Polish coastline (length
of 529 km, excluding internal lagoons) and maritime areas (area of
38,347 km
2
, excluding internal lagoons) cover a large part of the
southern BS. The largest share of the catchment of the southern BS is
located within the boundaries of Poland (88 % of the Vistula and 85 % of
the Odra river basins). Poland is considered a major emitter of Hg in the
southern BS region, both in terms of atmospheric deposition and
waterborne load (HELCOM, 2011, 2015; UNEP, 2006; EMEP, 2020).
2.2. Data collection
The aim of the literature search was to identify data relevant to Hg
concentration in abiotic and biotic compartments of the Polish part of
the BS ecosystem. The rst studies on Hg concentration in the southern
BS date back to the 1970s (Ku´
zma, 1971; Chodyniecki et al., 1975;
Zimak and ´
Cwiertniewska, 1976; Gajewska and Nabrzyski, 1977; Pro-
tasowicki and Chodyniecki, 1980). However, due to the uncertainty of
these results, this study considers only data published after 1990.
The search was restricted to English and Polish language. The sci-
entic literature search was based on international databases, i.e., Sci-
enceDirect (https://www.sciencedirect.com), Springer Link
(https://www.link.springer.com), Wiley Online Library (https://onli
nelibrary.wiley.com), and Google Scholar (https://scholar.google.
com). A grey literature searching strategy involved grey literature da-
tabases, e.g., DiVA (http://www.diva-portal.org), government agencies
and academic institutions websites, Google search engine, as well as
consultations with experts. The terms ‘mercury’, ‘Hg’, and ‘Baltic Sea’
were key descriptors used in the search, however, ‘heavy metals’, ‘trace
metals’, ‘pollution’, ‘Polish maritime areas’, ‘Poland’, ‘Gulf of Gda´
nsk’,
‘Puck Bay’ and ‘Pomeranian Bay’ (Fig. S1) were also considered. In this
review, works concerning the Hg concentration in water, bottom sedi-
ments, vegetation, invertebrates, sh, coastal and waterbirds, and ma-
rine mammals, were included. Since the available works and data sets
mostly refer to total mercury concentrations, our study focuses on that
form. Additional information on this section can be found in the Elec-
tronic Supplementary Material.
3. Emissions and releases of Hg
3.1. Atmospheric deposition
Annual Hg emission in the BS region, based on inventory data of the
HELCOM contracting parties from 2018, is estimated at 31.0 Mg. Poland
is responsible for the emission of 8.7–9.6 Mg of Hg per year (KOBIZE,
2019; EMEP, 2020) (Fig. 1), which is about one-third of the emitted Hg
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
3
pool, making Poland a major Hg supplier among the HELCOM countries.
In Poland, 71 % of Hg is emitted from the energy sector (with 5–7 %
coming from the households), while industrial processes are responsible
for 26 % of total national Hg emissions (Pyka and Wierzchowski, 2016;
KOBIZE, 2019).
Due to the reforms of the social, political and economic system
initiated in the 1990s, the emission of Hg from the Polish territory
decreased by more than 40 % in the past three decades (KOBIZE, 2019)
(Fig. 1), while the average decline among HELCOM countries was esti-
mated at 62 % (EMEP, 2020). Following the downward trend in Hg
emission, atmospheric deposition of Hg in the BS decreased from 1990
by about 35 % (Fig. 1). In 2018, the annual atmospheric deposition of Hg
in the BS was estimated at 2.8 Mg, representing almost 70 % of the total
Hg input (HELCOM, 2021). According to the results of the GLEMOS
model (https://en.msceast.org/index.php/j-stuff/glemos) used by
EMEP, the atmospheric input of Hg introduced to the BS in 2018 by
Poland was 0.17 Mg, which is 6.2 % of the total Hg deposition of all
countries. As in the case of emissions, it places Poland in the lead of the
HELCOM contracting parties (EMEP, 2020). However, this is not re-
ected in Hg concentrations measured in the air over Poland (Siudek
et al., 2011; Bełdowska et al., 2012; Saniewska et al., 2014a; Lew-
andowska et al., 2018; Korejwo et al., 2020), as well as Polish water-
courses (Saniewska et al., 2014b, 2018, 2022; Bełdowska et al., 2016a;
Jędruch et al., 2017) and terrestrial biota (Nawrocka et al., 2020;
Jędruch et al., 2021), which were not elevated and typical of low or
moderately polluted regions of Europe.
3.2. Riverine input
In 2018, the annual waterborne input of Hg into the BS was esti-
mated at 1.6 Mg, which was about two times less than atmospheric
deposition (HELCOM, 2021). Approximately 60 % of this load (0.9 Mg
year
−1
) was introduced from the territory of Latvia. Polish rivers
introduced to the sea about 0.09 Mg of Hg year
−1
, which constitutes 6 %
of the total riverine Hg entering the BS and is one of the smallest con-
tributions among the HELCOM contracting parties. Given that Poland
covers 18 % (311,900 km
2
) of the BS catchment area, the area-specic
riverine load estimated for Poland (0.3 g km
−2
) was the lowest among
other HELCOM countries (HELCOM, 2021). Although HELCOM has
been collecting data on the waterborne load of Hg to the BS since 1994,
the assessment of inputs over time is difcult due to the uncertainty and
scarcity of the results. These issues are particularly related to older
datasets; consequently, only results from 2012 to 2018 are included in
the latest HELCOM (2021) report (Fig. S2). Although the interannual
variability of riverine Hg inputs to the BS (from 0.8 to 2.8 Mg year
−1
),
this period is too short to determine any tendency. Similarly, no such
trend was noted for Poland, whose share in the waterborne load of Hg in
years 2012–2018 varied from 0.02 to 0.35 Mg year
−1
(mean: 0.17 Mg
year
−1
). However, these values allow demonstrating how incorrect the
previous estimations were. Data presented by HELCOM until 2008
(Knuuttila, 2008), indicated that the riverine input of Hg from Polish
territory in years 1994–2006 ranged from 1.12 to as much as 103.58 Mg
year
−1
(mean: 17.94 Mg year
−1
), which exceeds recent estimations by
up to several orders of magnitude (Fig. S2). These loads were continu-
ously corrected in further HELCOM publications (2011, 2015, 2018a, b);
however, outdated estimations ingrained in the scientic literature and
media reports. Problems with data availability and reliability concern
not only Poland but also eastern European countries, while the complete
time series of Hg riverine inputs to the BS are only available for Ger-
many, Sweden, and Finland (HELCOM, 2021).
3.3. Re-emission of legacy Hg
The contribution of recycled Hg in the current Hg budget in the BS is
not known (Reckermann et al., 2022). In Poland, the topsoil is the main
terrestrial Hg storage (Jędruch et al., 2021). Given that retention of Hg
in soils is affected by environmental variables and hydrological condi-
tions (Saniewska et al., 2018; Zhou et al., 2021), the predicted climate
evolution will enhance Hg runoff to receiving water bodies. The increase
in annual precipitation and the frequency of heavy rains in the southern
BS region is particularly important (R¨
ais¨
anen, 2017). During these pe-
riods, the daily outow of Hg from the soil was estimated at 90–150 % of
the daily wet deposition ux (Saniewska et al., 2018). In 2010, during
the ood on the Vistula river, the load of Hg introduced into the BS was
Fig. 1. Temporal changes of annual total Hg emission (Mg) from Polish territory and estimated Hg deposition (Mg) to the southern Baltic Sea (reproduced from a
study by Jędruch et al. (2021) without changes under the terms of the Creative Commons Attribution 4.0 International License https://creativecommons.org/lice
nses/by/4.0/).
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
4
ve times greater than in previous years, more than 75 % of which was
contributed by oodwaters (Saniewska et al., 2014b).
Climate change models indicate intensied coastal erosion, favored
by storm surges, rising sea level, and decreased sea ice cover (R´
o˙
zy´
nski
and Lin, 2021). Soft cliffs, located not only in Poland but also in Ger-
many or Lithuania, are particularly vulnerable to erosion. Given the
large mass of sediments introduced into the marine environment, the
inuence of abrasion on the Hg input cannot be ignored. The Hg load to
the BS as a result of the abrasion of the Polish coast in 2011–2014 was
estimated to be 0.06 Mg year
−1
, which is approximately one-third of the
annual input of the Vistula river in that period (Bełdowska et al., 2016a).
Importantly, most of the Hg pool in eroded coastal sediments and soils
that crumble to the sea is labile (64 % on average, with the domination
of Hg adsorbed on sediment or organic particles) (Kwasigroch et al.,
2018), which means that Hg stored in coastal deposits can enter the
marine trophic chain.
Although marine sediments are seen as the Hg sink, over half of Hg
deposited in the southern BS sediments is labile (67 % on average, with
the domination of Hg bound to organic matter) and may be remobilized
to the water column (Bełdowski et al., 2009; Kwasigroch et al., 2021).
Projected changes in water dynamics, induced by greater wave energy
and more frequent storms, will enhance the resuspension of sediments
and the release of Hg. One of the most important consequences of
climate change is the decrease in oxygen conditions in the bottom zone,
indirectly affected by temperature, salinity stratication, precipitation,
and runoff patterns in the southern BS region (Reckermann et al., 2022).
It is unclear whether the frequency of inows of more oxygenated water
from the North Sea to deep basins of the BS will change; however, their
number in recent years was lower compared to earlier decades. An in-
crease in the hypoxia area will most probably increase the benthic ux of
Hg and its release from sediments (Emili et al., 2011), which is especially
dangerous in regions with elevated Hg concentrations (Siedlewicz et al.,
2020; Kwasigroch et al., 2021).
Due to military activity in the BS region during World Wars I and II,
thousands of tons of chemical and conventional munitions were sunk in
its waters. The presence of munitions was also conrmed in the Polish
part of the BS, in the Gulf of Gda´
nsk (Bełdowski et al., 2019; htt
p://www.chemsea.eu/). In most cases, the dumped material contained
Hg, since back then mercury fulminate was used as a primary explosive.
Sunken munitions are estimated to contain more than 300 Mg of Hg. Due
to the progressive corrosion of the munition shells, Hg-containing ma-
terials are leaking out and spreading on the marine bottom (Scharsack
et al., 2021). Studies conducted at dumpsites located in the southern BS
indicated that sediments collected close to the dumped ammunition
contained a few times more Hg than samples distant from the object
(Bełdowski et al., 2019; Siedlewicz et al., 2020). In addition to muni-
tions, there are also hundreds of ships that have been sunk in the BS
(Sokołowski et al., 2021). Fuel and ammo tanks make wrecks another
potential source of Hg. Studies conducted in the area of two ships sunken
in the Gulf of Gda´
nsk, the s/s Stuttgart and the t/s Franken, showed that
Hg concentrations in sediments near the wrecks contained up to two
orders of magnitude more Hg than sediments from the reference areas
(Rogowska et al., 2015; Siedlewicz et al., 2020). Climate change is ex-
pected to accelerate the corrosion of munitions shells and steel bodies of
wrecks, promoting the distribution of their content in sediments (Beł-
dowski et al., 2019; Scharsack et al., 2021). Hypoxia can alter the
degradation process of chemical warfare agents, leading to greater
persistence of degradation products in sediments, and promote Hg
methylation by anaerobic bacteria. While rising water temperature is
likely to increase the solubility of stable Hg compounds and the
desorption of Hg from sediment particles (Reckermann et al., 2022;
Scharsack et al., 2021).
4. Hg level in abiotic environment
4.1. Seawater
The total Hg concentration in the surface water of the open part of
the southern BS ranged from 0.1 to 0.8 ng L
−1
and averaged approxi-
mately 0.3 ng L
−1
(Table S1). These values are slightly higher compared
to the other offshore waters of the BS (Pohl and Hennings, 2009; Kuss
et al., 2017; Soerensen et al., 2018), as well as typical level in surface
waters of the open ocean (Cossa et al., 2020). Concentrations measured
in the Polish waters are comparable to other inland seas, the Mediter-
ranean Sea and the Black Sea, but signicantly lower than the level
determined in areas considered polluted, e.g. Gulf of Trieste, the East
China Sea or Minamata Bay (Table S1). Hg level in water changed in the
water column. In deeper waters, the Hg concentration was on average
two-three times higher than in the surface layer. In coastal areas, Hg
level in the Polish part of the BS reaches several ng L
−1
, a range typical of
land-sea transition zones (Stein et al., 1996). These values increased
more than tenfold during and after the occurrence of extreme hydro-
logical phenomena, such as storms or heavy rainfall (Saniewska et al.,
2014b, 2018). As Hg which enters the BS via rivers is mainly associated
with suspended particulate matter (SPM), the particulate-bond Hg
concentration in coastal waters, reaching 250 ng g
−1
dry weight (dw), is
several times higher than in open sea, where the average range is from
more than 10 to less than 100 ng g
−1
dw (Murawiec et al., 2007; Jędruch
et al., 2017). The level of particulate Hg is also shaped by the intensity of
primary production, as Hg is adsorbed onto microalgae, mainly as labile
halides (Bełdowska et al., 2018b). This is particularly important during
phytoplankton blooms, especially of ciliates, when the concentration of
suspended Hg increases approximately ve times (Jędruch et al., 2017;
Bełdowska and Kobos, 2018).
The concentrations of Hg in the Polish part of the BS are currently
lower than those recorded in previous years, which is a consequence of
reduced Hg emission in Poland and in other EU countries, described in
detail by Jędruch et al. (2021). The most signicant decrease in Hg
concentration in surface waters occurred in 1993–2001 when it has been
reduced by seven times from over 17 to 1 ng L
−1
(Fig. 2, Table S2). In the
rst years of the 21st century, Hg level in the Polish waters of the BS
uctuated from 1 to almost 5 ng L
−1
. Due to a lack of regular monitoring
after 2009, further investigation of temporal changes in concentrations
is not possible. As the 2010–2014 data concern the Gda´
nsk Deep only,
they are too local to be representative of the studied area. However, the
results of the study conducted in 2015–2016 by Soerensen et al. (2018)
indicate that the concentration of Hg in the waters of the southern BS
remains at a low level both in the surface and bottom layers.
4.2. Bottom sediments
Given, that the Hg concentration in sediments is related to the grain
size composition and the content of organic matter, sandy sediments,
where the content of organic matter was low, contained less Hg than soft
sediments in a range from 2 to 50 ng g
−1
dw (Jędruch et al., 2015;
Kwasigroch et al., 2021) (Fig. 3). In many cases the Hg level was lower
than 30 ng g
−1
dw, which is a geochemical Hg background for BS sed-
iments (Leipe et al., 2013). Elevated Hg concentrations, reaching 264 ng
g
−1
dw were detected in the muddy sediments of the Gda´
nsk Deep,
which is a typical depositional area inuenced by direct discharge of the
Vistula river. However, even in the deepest regions of the study area, the
proportion of the most dangerous form of Hg, methylmercury (MeHg),
was relatively low and did not exceed 2 % (Siedlewicz et al., 2020).
Nevertheless, considering that the level of 150 ng g
−1
dw was set as a
safe threshold for Hg in sediments (NOAA, 1999), in some areas an
adverse effect on the marine organisms lining near the bottom may be
observed. Concentrations of Hg in sediments of the Polish part of the BS
were lower compared to its other BS regions considered to be polluted,
such as the Gulf of Finland (industrial waste, including wood- and paper
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
5
mills) or Kiel Bay, Mecklenburg Bay and western Arkona Basin (histor-
ical dumping sites of military waste) (Leipe et al., 2013; Bełdowski et al.,
2019). Sediments of the southern BS contained far less Hg than heavily
contaminated areas in Europe, such as the northern Adriatic near the Hg
mining district in Slovenia (Table S3).
Based on data from the accumulation region of the Gdansk Deep,
concentration of Hg in surface sediments of the Polish part of the BS
uctuated since the 1990s without showing any trend (Fig. 4a,
Table S4). This variability was modied by changes in the Hg net input
to the sediments shaped by river inows, especially oods (Saniewska
et al., 2014b; Jędruch et al., 2015). In the Gdansk Basin, two oods have
occurred in the last three decades - one in 1997 and one in 2010. Hg
concentration in the sediments increased by about two times after the
ood, however, returned to the level preceding the event relatively
quickly, within a few years. That was the result of the remobilization of
labile Hg forms to the water column or their uptake by benthic
Fig. 2. Temporal changes of total Hg concentration (ng L
−1
) in surface and sub-halocline water (unltered) of the Polish part of the southern Baltic Sea (raw data and
references are presented in Table S2).
Fig. 3. Total Hg concentration (ng g
−1
dw) in surface sediments (0–5 cm) of the Polish part of the southern Baltic Sea in years 2016–2017 (based on raw data from a
study by Kwasigroch et al. (2021) kindly shared with the authors under the terms of the Creative Commons Attribution 4.0 International License https://creativeco
mmons.org/licenses/by/4.0/).
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
6
organisms, as well as sediment resuspension and redeposition in deeper
areas of the bottom (Bełdowski et al., 2009; Jędruch et al., 2019a,
2019b). Considering the results obtained in sediment cores, however,
the decline in Hg concentrations after 1990 is apparent (Fig. 4b,
Table S5). According to the latest ndings (Kwasigroch et al., 2023), the
decrease in Hg level was close to 40 %, similar to the trend in atmo-
spheric Hg emissions and deposition (Fig. 1) (Jędruch et al., 2021).
5. Hg level in biotic environment
5.1. Current status of Hg in marine biota
5.1.1. Vegetation
The Hg level in the primary producers of the southern BS varied in a
wide range. The lowest values were measured in macroalgae and
vascular plants in which the Hg concentration averaged 14 and 9 ng g
−1
dw, respectively, and rarely exceeded 20 ng g
−1
dw (Bełdowska et al.,
2016b; Jędruch et al., 2019b; Jędruch and Bełdowska, 2020). The
elevated Hg concentration, reaching 51 ng g
−1
dw, was measured in
vascular plants from the Pomeranian Bay (Bełdowska et al., 2015)
(Table S6), which may be attributed to the increased Hg concentration in
surface sediments compared to other areas of the Polish coastal zone
(Fig. 3). Hg level measured in the beach-cast plant biomass was gener-
ally higher than in fresh macrophytes and averaged 21 ng g
−1
dw, with a
maximum reaching 47 ng g
−1
dw (Graca et al., 2022). A higher Hg
concentration compared to those observed in macrophytes was also
measured in benthic microora, epiphytes (biolm growing on plants),
and epilithon (biolm growing on rocks) (Bełdowska et al., 2018a). The
Hg concentration in samples collected in Puck Bay averaged from 31 to
108 ng g
−1
dw in the epiphyton and about 65 ng g
−1
dw in the epilithon
(Jędruch et al., 2019b). It was also noted that the presence of epi-
psammon (e.g., sediment-associated diatoms) in sediments has a posi-
tive effect on Hg accumulation, increasing its concentration by up to two
times. Among the primary microproducers, the highest Hg level, with an
average Hg level of approximately 100 ng g
−1
dw, and a maximum of
over 600 ng g
−1
dw, was measured in phytoplankton (Bełdowska and
Kobos, 2018). Among the investigated species, an autotrophic cosmo-
politan ciliate Mesodinium rubrum showed the highest ability to accu-
mulate Hg. The dominant form of Hg, accounting for about 60 %, was
halides loosely absorbed on the surface of phytoplankton (Bełdowska
et al., 2018b). Part of it crosses the cell walls and is stored in the cyto-
plasm, becoming readily assimilable to zooplankton, making phyto-
plankton a critical element shaping the tropic transfer of Hg within the
marine food web (Harding et al., 2018; Schartup et al., 2018).
5.1.2. Pelagic and benthic invertebrates
Concentrations of Hg in zooplankton collected in the coastal zone of
the Polish part of the SB varied from 16 to 204 ng g
−1
dw with the
average level around 106 ng g
−1
dw (Bełdowska and Mudrak-Cegiołka,
2017; Jędruch et al., 2019b) as in the case of phytoplankton. In areas
more distant from the shore, concentrations of Hg in zooplankton were
about half as high (Table S6). High Hg concentrations were observed in
zooplankton when their biomass was dominated by rotifers Synchaeta
spp. and copepods Acartia spp. Elevated Hg concentrations were also
measured when a high proportion of the biomass consisted of planktonic
larvae of benthic invertebrates, mainly barnacle Amphibalanus improvi-
sus nauplii. Preliminary results show that in the case of zooplankton, the
contribution of Hg adsorbed on the surface is lower than in phyto-
plankton, and the dominant form with a proportion of 70 % is absorbed,
cytoplasmic Hg (Jędruch, unpubl.). This means that the transition be-
tween the prevailing mechanism of Hg entry into the organism occurs
already at low lower trophic levels of the pelagic ecosystem (Harding
et al., 2018; Graca et al., 2022).
The level of Hg concentration in the zoobenthos of the coastal lagoon
of the southern BS, Puck Bay (Fig. S1), ranged widely from 7 to 250 ng
g
−1
dw and depended on the food preferences of the species. Hg con-
centration increased with trophic level, and the highest Hg level was
measured in the soft tissues of the Harris mud crab Rhithropanopeus
harisii and the woolly crab Eriocheir sinensis, averaging 71 and 111 ng
g
−1
dw, respectively (Jędruch et al., 2018b, 2019b). Both species are
omnivorous; however, live and dead animal tissues account for a sig-
nicant proportion of their diet. In comparison, the Hg level measured in
the soft body of the herbivorous grazing snail Perinigia sp. averaged 17
ng g
−1
dw, and in the suspension-feeding mussel Mytilus trossulus of the
same region was 23 ng g
−1
dw. These values were too low to pose a
threat to the sh or birds (Evers et al., 2011). However, organisms that
inhabit deeper areas of the bottom, where Hg tends to accumulate in
sediments, had more Hg in soft tissues. As demonstrated in the study
Fig. 4. Temporal changes of total Hg concentration (ng g
−1
dw) sediments of the Polish part of the southern Baltic Sea based on: a. data from surface sediments and
b. sediment cores collected at Gda´
nsk Deep (ca. 100 m depth) (Fig. S1) (raw data and references are presented in Tables S4 and S5).
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
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conducted in Puck Bay, in the area with locally elevated Hg concen-
trations in sediments (Fig. 3), the mean concentration of Hg in
M. trossulus mussel (145 ng g
−1
dw) was more than six times higher
compared to organisms sampled in the coastal zone (Jędruch et al.,
2019a) (Table S6). However, even values at this level remain a low risk
to potential consumers (Evers et al., 2011). The total share of bioavail-
able Hg forms in benthic invertebrates was greater than 90 % and about
60 % of Hg occurred as organic form (Jędruch et al., 2018b, 2019a;
Jędruch and Bełdowska, 2020), which means that most of the accumu-
lated Hg can be transferred to higher trophic levels.
5.1.3. Fish
In the case of sh from the southern BS, a clear effect of the trophic
level on the Hg concentration in muscle tissue was visible (Fig. 5a).
Planktivorous sh, European sprat Sprattus sprattus and Baltic herring
Clupea harengus, were characterized by the lowest average Hg concen-
tration of 16 and 24 ng g
−1
wet weight (ww), respectively. In predatory
sh, the Hg concentration increased, averaging from 42 ng g
−1
ww in
sea trout Salmo trutta to 151 ng g
−1
ww in the European perch Perca
uviatilis. However, this ‘traditional’ relationship between Hg concen-
tration and trophic level does not always occur (Buck et al., 2019). For
some benthivorous sh (e.g., European ounder Platichthys esus, turbot
Scophthalmus maximus, European eel Anguilla anguilla), the concentra-
tion of Hg were higher compared to some piscivorous species (e.g.,
Baltic cod Gadus morhua, Atlantic salmon Salmo salar). It could be
related to the fact that in areas where Hg accumulates in sediments, sh
of lower trophic levels, but feeding on benthic organisms, may have a
higher Hg concentration compared to pelagic sh (Chˆ
ateauvert et al.,
2015). The effect of trophic position on Hg concentration in the livers of
BS sh was not unequivocal and the average Hg concentration did not
exceed the level of accumulation in the muscles (Fig. 5a). The distri-
bution pattern in the organs of BS sh, with a predominance of accu-
mulation in the muscles compared to the liver, shows that sh of the
southern BS are slightly polluted with Hg (Kwa´
sniak and Falkowska,
2012; Kwa´
sniak et al., 2012), as other researchers have concluded on the
example of sh from freshwater ecosystems (Svobodov´
a et al., 1999;
Havelkov´
a et al., 2008).
Hg concentration in the liver and muscles of the BS sh was similar to
the Hg level in sh caught in various regions of the North Sea and North
Atlantic (Table S7). The average Hg concentrations in sh muscles
(except sprat) exceeded the threshold value of 20 ng g
−1
ww set for sh
by HELCOM (2018a) as a criterion of good environmental status.
Nevertheless, the thresholds of allowed Hg concentrations in consumed
sh set by the US EPA (300 ng g
−1
ww) were exceeded in a limited
number of cases. However, none of the measured Hg concentration were
higher than the WHO human health criterion for Hg in sh (500 ng g
−1
ww). The obtained results showed that in most of the investigated spe-
cies mean Hg level in muscles was low enough (below 100 ng g
−1
ww) to
pose ‘no risk’ for wildlife (Dietz et al., 2021b) (Fig. 5a). Mean Hg con-
centrations in muscles of turbot, European eel, and European perch were
within the range of values corresponding to a ‘low risk’ category
(100–300 ng g
−1
ww). Only in a few muscles samples of European eel
and European perch, and livers of the Baltic cod, the Hg concentration
posed a ‘moderate risk’ for potential health effects (300–500 ng g
−1
ww), e.g., reduced reproduction, changes in biochemical processes and
damage to cells and tissues (Evers et al., 2011). There was no case of Hg
concentrations exceeding 500 ng g
−1
ww, the lower limit of the ‘high
risk’ category (500–2000 ng g
−1
ww) (Fig. 5a), which means that, when
it comes to Hg contamination, sh from the southern BS can be
consumed safely both by humans and wildlife. Moreover, according to
the calculations by Kwa´
sniak et al. (2012), an adult consumer of sh
caught in the Polish economic zone of the BS would have to eat about 30
kg of herring or 26 kg of cod per month to exceed the allowable refer-
ence doses.
5.1.4. Coastal and waterbirds
As in the case of sh, the trophic position had an impact on Hg
concentration, both in muscles and livers of the southern BS avifauna.
The average Hg level in omnivorous species that consume food both
from natural and anthropogenic sources, i.e., mallard Anas platyrhyn-
chos, black-headed gull Larus ridibundus, common gull Larus canus,
herring gull Larus argentatus, varied from 37 to 103 ng g
−1
ww in
muscles and from 73 to 317 ng g
−1
ww in livers (Fig. 5b). Birds foraging
in the coastal zone having a high proportion of sh in their diet, i.e.,
white-tailed eagle Haliaeetus albicilla, great black-backed gull Larus
marinus, or typical piscivorous species, i.e., razorbill Alca torda, black-
throated loon Gavia arctica, great cormorant Phalacrocorax carbo, com-
mon merganser Mergus merganser, had higher Hg tissue concentrations.
The mean level of Hg in sh-eating birds ranged from 144 to 731 ng g
−1
ww in muscles and from 347 to 4292 ng g
−1
ww in the livers.
Birds associated with aquatic ecosystems are often treated as an in-
dicator of environmental pollution (Binkowski et al., 2016). The Hg
concentrations in tissues of birds inhabiting coastal areas of the southern
BS were similar or lower compared to the other regions of the world
(Table S7). In most cases, Hg levels in studied birds did not exceed a
value of 1000 ng g
−1
ww (Fig. 5b), indicating a ‘low risk’ of adverse
effects of Hg on avian health (Kalisi´
nska et al., 2014b; Scheuhammer
et al., 2015). In a few livers samples of piscivorous species, i.e., white-
tailed eagle, great black-backed gull, Hg concentrations were higher
than 2000 ng g
−1
ww, a threshold of ‘moderate risk’ of negative impacts
of Hg. However, Hg concentration in the tissues of these birds did not
exceed a level of 5000 ng g
−1
ww, treated as a ‘high risk’ threshold for
waterbirds, beyond which the effects of Hg poisoning become apparent
(Zillioux et al., 1993; Badzinski et al., 2009). Higher Hg concentrations
were measured in tissues of great cormorants, in which about a quarter
exceeded the hepatic Hg concentration of 5000 ng g
−1
ww (Misztal-
Szkudli´
nska et al., 2018). This means that the growth, individual
development, reproduction, metabolism, and behavior of this species
may be affected (Ackerman et al., 2016; Scheuhammer et al., 2015).
Concentration exceeding the ‘severe risk’, lethal level of 20,000 ng g
−1
ww (Kalisi´
nska et al., 2014a, 2014b) was measured in the liver of only
one cormorant. Such high values were not recorded in other cormorant
populations in Europe and Asia (Table S7). In the muscles of cormorants
from the southern BS the Hg level of 1000 ng g
−1
ww was exceeded in
11 % of 55 subjects, while the maximum concentration measured was
1544 ng g
−1
ww (Misztal-Szkudli´
nska et al., 2018). The authors explain
these high values by the fact that cormorants were sampled at the
beginning of the breeding season, when they forage intensively and
accumulate the largest amounts of Hg, and before the molting during
which Hg is eliminated from the body (Misztal-Szkudli´
nska et al., 2018).
Despite the fact that birds from coastal populations, unlike those from
inland populations, can tolerate higher Hg levels without adverse effects
(Scheuhammer et al., 2015), elevated Hg levels may reduce the number
of cormorants in the coast of the southern BS. Decreased reproductive
success of the birds may be also caused by higher mortality of eggs and
nestlings (Zillioux et al., 1993).
5.1.5. Marine mammals
Seals and cetaceans occupy the top of the BS food pyramid and are
typical endpoints in the biomagnication of Hg in the marine trophic
chain. Consequently, the Hg concentrations measured in tissues of ma-
rine mammals were the highest among the investigated groups of or-
ganisms, exceeding by more than ve orders of magnitude the
concentrations in organisms from the initial links of the trophic chain
(Fig. 5c). As a result of biomagnication, the share of organic Hg in their
tissues was also the highest reaching as much as 95 % of total Hg
(Wiener et al., 2003; Jędruch et al., 2018b). Although the diet of the BS
mammals largely overlaps, each species has its own preferences. The
main prey of the grey seal is herring, followed by sprat and cod, the
harbor seal feeds mainly on benthic sh (e.g., eel, goby, cod, ounder,
whitesh), the ringed seal feeds on small sh (e.g., herring, smelt,
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Marine Pollution Bulletin 186 (2023) 114426
8
Fig. 5. Total Hg concentration (ng g
−1
ww) in the muscles and livers of: a. sh, b. birds, and c. mammals from the southern Baltic Sea (values and references are
given in Table S8) and estimated risk categories for Hg-associated health effects in wildlife: NRC – no risk, LRC – low risk, MRC – moderate risk, HRC – high risk, SRC
– severe risk category (detailed information regarding the assumptions are provided in Electronic Supplementary Material).
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
9
sculpin, perch, stickleback) and benthic fauna (e.g., crustaceans, bi-
valves), while the harbor porpoise prefers fatty pelagic shes like her-
ring and sprat (Teilmann et al., 2017). These diet differences, together
with daily food intake, affect Hg uptake and therefore Hg concentration
in individual species (Fig. 5c). Consequently, the largest BS seal,
H. grypus, had a higher Hg concentration than other seals and porpoises.
Many marine mammals possess key characteristics of sentinel spe-
cies, including showing clear responses to environmental variability or
change and indicating anthropogenic impacts on the ecosystems (Hazen
et al., 2019). Despite this, they rarely become indicator organisms of a
toxic threat to the marine environment (Bełdowska and Falkowska,
2016; Grajewska et al., 2020). However, given that marine mammals are
top predators, the state of their populations indicates the state of the
entire BS ecosystem. Hg level in seals from the Polish coast was higher
compared to other regions of the BS (Table S7). A possible cause is that
seals from the southern BS feed higher up in the food chain and more in
the benthic food web in comparison to other areas (Helsingen, 2011). In
contrast, compared to other European and North Atlantic regions, grey
seals from the southern BS population have similar Hg levels, especially
in livers, and lower in the case of muscles. The risk analysis of possible
Hg-associated health effects for marine mammals carried out by Dietz
et al. (2021b) showed that the estimated risk for BS populations was not
higher than in the bordering waters of the North Sea and North Atlantic.
Taking into account the data presented in this study (Fig. 5c), Hg
concentrations in the livers of the ringed seal and harbor porpoise pose
‘no risk’ for negative health effects, while in the case of the harbor and
grey seals this risk was ‘low’.
5.2. Time trends of hg at different trophic levels
5.2.1. Low-trophic level organisms
The Hg level in plants has shown a downward trend in the last de-
cades, both for macroalgae (Fig. 6a) and vascular plants (Fig. 6b). In
macroalgae, there was a clear decrease in Hg concentration from 62 ng
g
−1
dw in 1988 to 11 ng g
−1
dw in 2006, when the level of Hg con-
centration was reduced by about 3 ng g
−1
year
−1
. Since 2006, the Hg
concentration in macroalgae has remained relatively constant, not
exceeding 14 ng g
−1
dw. In the case of vascular plants, the downward
trend in the Hg level persisted longer, until 2011 (Fig. 6b). Between
1998 and 2011, Hg concentration decreased from 48 ng g
−1
dw to 9 ng
g
−1
dw, averaging less than 2 ng g
−1
dw year
−1
. This indicates a slower
response of vascular plants to the reduction of Hg emission and conse-
quently its concentration in the abiotic environment, compared to
macroalgae that also uptake Hg from sediments via roots. On a long-
term scale, the decrease in Hg concentration in sediments (Fig. 4b) is
a slower process than in water (Fig. 2). Considering the dependence of
Hg concentration in marine plants on the amount of Hg emission, it is
explicit for macroalgae and vascular plants (Table 1).
Fig. 6. Linear regression model of temporal changes in Hg concentration in low-trophic level organisms: a. macroalgae, b. vascular plants (ng g
−1
dw), c.
zooplankton and d. soft tissues of benthic invertebrates (ng g
−1
dw) from the southern Baltic Sea (circles represent means, whiskers represent ranges, dashed line
indicates a statistically signicant trend, solid line indicates changes in anthropogenic emission of Hg from the Polish territory (Fig. 1), raw data and references are
presented Table S9).
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Marine Pollution Bulletin 186 (2023) 114426
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The number of studies for planktonic organisms is much smaller than
for phytobenthos. Determining the temporal variability of Hg was only
possible for zooplankton. The Hg level in these organisms dropped from
240 ng g
−1
dw in 1995–1998 to 62 ng g
−1
dw in 2013 and the decline
averaged 10 ng g
−1
dw year
−1
(Fig. 6c). The concentration of Hg in
zooplankton correlated well with the emission of Hg into the environ-
ment (Table 1). A decrease in Hg concentration was also observed in
benthic invertebrates, in which the average metal level decreased from
125 ng g
−1
dw in 1998 to 26 ng g
−1
dw in 2017 and dropped by more
than 4 ng g
−1
dw year
−1
(Fig. 6d). Unlike zooplankton, the Hg con-
centration in zoobenthos was not strongly dependent on the Hg emission
(Table 1), due to the time shift of the effects of Hg deposition in sedi-
ments. The use of plant organisms and invertebrates as bioindicators of
Hg pollution in the southern BS environment is hindered by several
factors, such as relatively low abundance and large population uctua-
tions, high seasonality, difcult collection and analysis, and feeding
habits that are not always sufciently well known (Soto et al., 2011).
5.2.2. Intermediate consumers
During the last three decades, the Hg level in pelagic (sprat, herring)
and benthic sh (cod) in the southern BS showed statistically signicant
downward trends (Fig. 7), indicating a strong relationship with Hg
emission (Table 1). Statistically signicant trends in Hg level variability
in sh correspond well with other literature reports. A similar estimation
of Hg downward trends in the Polish part of the BS sh was described by
Polak-Juszczak (2009, 2013). However, taking into account HELCOM
(2018a) monitoring data from 1998 to 2016 on Hg level in herring
caught in the Gda´
nsk Basin and the Słupsk Bank (Fig. S1), despite
uctuations, there was no apparent temporal trend. A decrease in Hg
concentration relative to the 1990s was, in turn, noted in the case of
herring muscles from the northern BS, Bothnian Bay and Northern Baltic
Proper, as well as its western parts, Arkona Basin and Kattegat (HEL-
COM, 2018a). A relationship between changes in Hg emissions and wet
deposition in the last decades and the Hg level in sh has also been
demonstrated in other aquatic ecosystems in Europe and North America
(Drenner et al., 2013; Wang et al., 2019; Grieb et al., 2020).
The slowdown of the Hg emissions trend at the beginning of the 21st
century at a global and national level, resulted in the stabilization of Hg
levels in the sprat from the Polish BS (Fig. 7a). However, around 2006,
the downward trend of Hg concentration in herring and cod changed
direction (Fig. 7b, c). After this year, a nearly twofold increase in Hg
concentration in the herring muscles was recorded, while in cod mus-
cles, this increase was more than threefold. As a result of this change, the
relationship between Hg concentration in cod and Hg emission was poor
and not statistically signicant (Table 1). As demonstrated in the study
by Haase et al. (2020), a shift in the diet of cod from benthic to sh prey
was also observed at the same time. Although a gradual decrease in the
proportion of benthic invertebrates (e.g., polychaetes or isopod crusta-
cean Saduria entomon) in the diet of cod was already observed in earlier
years, in 2006 sh started to dominate the diet of smaller cod, while the
largest cod fed almost exclusively on sh. The increased importance of
sh prey is related to the decline in benthos populations resulting from
the worsening oxygen conditions and the expansion of hypoxic and
anoxic areas. It was also observed that the contribution of herring to the
diet of cod increased with a simultaneous drop in the share of sprat. This
is probably due to the decline in sprat abundance in the southern BS and
the increase in the herring stock. However, in parallel, the size of the
herring began to decrease, making it a more suitable food even for
smaller cod (Casini et al., 2016). The decrease in the size of the herring
may, in turn, have increased the Hg concentration in their tissues,
increasing the Hg load in the cods that consumed them. Generalization
of the cod feeding strategy, which could be described as ‘choosing what
is available’, resulted in lower nutritional conditions and reproductive
success. Therefore, it is one of the explanations for the decline in the
condition and abundance of cod that occurred during the past two de-
cades, in addition to overshing (Haase et al., 2020).
A U-turn for Hg concentration was noted around 2005 for European
perch caught in Mecklenburg Bay in the German part of BS (HELCOM,
2018a). A similar change in Hg trend in the early 2000s was also
recorded for Arctic char Salvelinus alpinus, rainbow smelt Osmerus mor-
dax, and lake trout Salvelinus namaycush from Canadian lakes (Blukacz-
Richards et al., 2017; Wang et al., 2019; Morris et al., 2022). This
pattern was already observed more than ten years earlier in freshwater
and marine sh from North America (Grieb et al., 2020), attributing it
both to climate change and to the increase in anthropogenic Hg emis-
sions from Asia (Monson, 2009; Gandhi et al., 2014). Schartup et al.
(2019) pointed out the climate change and dietary shifts initiated by
overshing as key factors increasing MeHg concentration in predatory
sh. The study indicated that due to the restoration of the herring
population in the Gulf of Maine, the diet of cod changed from small
herring to larger herring and lobster, resulting in a higher concentration
of MeHg by 6–20 %. At the same time, another species, the spiny dogsh
Squalus acanthias, switched from squid and other cephalopods, which
exhibit higher MeHg concentrations than other prey sh, to herring,
leading to a decrease in the MeHg level by 33–61 %. In addition, the
model developed by Schartup et al. (2019) predicts that an increase of
1 ◦C in seawater temperature would lead to a 32 % increase in MeHg
concentration in cod.
5.2.3. Top predators
Despite that many studies indicated that marine top predators are
sensitive to changes in food supply (e.g., Yurkowsky et al., 2018; Dietz
et al., 2021c; Reckermann et al., 2022), the available data are insuf-
cient to determine the impact of changes in sh structure and the
resulting changes in Hg concentration on coastal birds or marine
mammals from the southern BS. In the case of birds, the only results that
allow to compare the temporal variability of Hg concentration concern
feathers. The Hg level in the rectrices (ight feathers) of adult Herring
gulls from the Polish BS coast collected in 1992–1993 and 2010–2013
allowed to conclude that over 20 years Hg contamination decreased on
average by 1.5–1.9 % annually. Given that the trophic status of the
herring gull determined at that time using stable isotopes was close to
the current one (Szumiło-Pilarska et al., 2017), the diet of this species
has not changed. The observed decrease in Hg concentration can
Table 1
Relationship between total Hg concentration in organisms from the Polish part of the southern Baltic Sea and the anthropogenic emission of Hg from Polish territory
with the comparison (Mann–Whitney test) between period of stabilized Hg emission and before (Fig. 1).
Mean Hg level
Period Equation r p Until 2002 After 2002 p
Vegetation Macroalgae 1989–2018 Hg
biota
=4.0025Hg
emission
-27.5346 0.9233 0.0000 52 12 0.0317
Vascular plants 1989–2018 Hg
biota
=2.9256 Hg
emission
-16.6952 0.9189 0.0002 41 11 0.0319
Invertebrates Zooplankton 1998–2014 Hg
biota
=57.7334 Hg
emission
-494.7910 0.9991 0.0009 240 94 1.0000
Zoobenthos 1989–2018 Hg
biota
=6.6377 Hg
emission
-5.4399 0.5840 0.1686 152 41 0.0455
Fish Sprat 1995–2016 Hg
biota
=1.5560 Hg
emission
-0.3872 0.7342 0.0101 23 15 0.1025
Herring 1994–2018 Hg
biota
=6.2703 Hg
emission
-44.2084 0.7754 0.0011 42 26 0.1655
Cod 1997–2016 Hg
biota
= − 7.9107 Hg
emission
+132.2246 −0.4474 0.1087 37 50 0.3329
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Marine Pollution Bulletin 186 (2023) 114426
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Fig. 7. Linear and quadratic polynomial regression
model of temporal changes in total Hg concentra-
tion in the muscles of sh: a. sprat Sprattus sprattus,
b. herring Clupea harengus, and c. cod Gadus morhua
(ng g
−1
ww) from the southern Baltic Sea (circles
represent means, whiskers represent ranges, dashed
line indicates a statistically signicant trend, solid
line indicates changes in anthropogenic emission of
Hg from the Polish territory (Fig. 1), raw data and
references are presented Table S9).
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
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therefore be explained by the reduction in Hg emission and the resulting
decrease of Hg concentration in air inhaled and ingested food, probably
other than sh.
In the case of mammals exposed to Hg through eating sh, changes in
feeding patterns combined with the warming of water may be more
noticeable. Impact of feeding ecology on Hg exposure revealed in the
case of narwhal Monodon monoceros from Northwest Greenland. As a
result of the decline in sea ice cover over almost 50 years, the diet of
narwhal changed from ice-associated and benthic prey, such as halibut
or cod, to pelagic pray, such as capelin, which resulted in a steep in-
crease in Hg accumulation (Dietz et al., 2021c). Changes in feeding
ecology may also be a possible reason for the change in the trajectory of
the Hg concentration trend to an upward one, observed around 2004 for
common dolphins Delphinus delphis from the French North Atlantic coast
(Fernandez et al., 2022). As for the Hg concentration in the prey itself,
an increased intake of pelagic sh by mammals inhabiting the southern
BS, along with decreased consumption of demersal sh, could even be
benecial for them. It is related to the fact that pelagic species, such as
sprat or herring, contain less Hg than benthic sh, including ounder or
eel (Fig. 5a). However, the change in the diet of southern BS seals has not
been conrmed, e.g., due to insufcient results of isotopic investigations
or analysis of gastric contents. Secondly, taking into account the
occurring and forecasted changes in the BS ecosystem, and their effects
on Hg concentration in sh or Hg metabolism by top predators, the di-
rection of possible change is difcult to predict.
6. Conclusions
The southern BS has improved with respect to Hg contamination
over the last three decades, which resulted from a signicant reduction
in Hg emissions and a simultaneous decrease in its atmospheric depo-
sition by about one-third. Due to the lack of reliable time series on the
waterborne inow of Hg into the BS, this source could not be included in
this compilation. Similarly to the terrestrial environment in Poland
(Jędruch et al., 2021), the ‘response’ of the environment of the southern
BS environment to changes in atmospheric Hg inux is evident. The Hg
level in seawater has lowered since the 1990s, while in the case of
surface sediments, the Hg concentration uctuated, peaking in the
aftermath of oods. However, these elevations were relatively short-
term and the ‘ood footprint’ was no longer visible a year or two after
the event. Despite the gaps in the time series of Hg concentrations in
organisms, a downward trend accompanying the reduction in Hg
emissions was evident. Both Hg levels in water and in organisms of the
southern BS did not show a time lag relative to the atmospheric supply of
Hg. This is a typical pattern for aquatic ecosystems, which similarly to
the BS, receive Hg primarily from the atmosphere. On the contrary,
waterbodies where rivers are the main source of Hg respond much
slower, and recovery may take decades to centuries (Harris et al., 2007;
Obrist et al., 2018). Proportional responses in Hg concentrations to
changes in Hg inow are also common for shallow or surface waters and
pelagic food webs, while sediments and benthic organisms are more
likely to exhibit a delayed response time (Harris et al., 2007; Sunderland
et al., 2010; Soerensen et al., 2016).
Despite the decrease in Hg concentrations in the southern BS in
recent decades, the observed and projected climate-driven changes are
expected to largely affect Hg uptake and trophic transfer of Hg in the
marine ecosystem (Schartup et al., 2019). After a signicant decline in
Hg emission from Polish territory in the 1990s and early 2000s, the
downward slowed and remained relatively stable for more than a
decade. This means that future changes in Hg concentrations in the
southern BS will most probably be inuenced by emission-independent
factors, such as further shifts in seawater temperature and food web
structure. An early stage of these changes, resulting in the reverse of
previous declines, was already observed in sh. However, because of
gaps in the data, we were unable to determine if there was an increase in
Hg concentrations at other trophic levels as well. Given that the BS is one
of the fastest-warming marginal seas in the world (Meier et al., 2022;
Reckermann et al., 2022), there is no indication that the changes in its
ecosystem will slow down. Although the anthropogenic load of Hg has
been reduced, the pool of Hg in the southern BS may increase due to the
climate change-induced remobilization of legacy Hg both from terres-
trial and marine deposits. Furthermore, increased temperature is one of
the most important factors inuencing changes in the structure of the
trophic web and promoting methylation and bioaccumulation (Dijkstra
et al., 2013). Consequently, the joint impact of climate evolution on the
Hg pool and its bioavailability, together with changes in the food chain,
is likely to exacerbate the vulnerability of wildlife and humans to Hg,
which means a need for strong actions to both reduce Hg emissions more
effectively and to pursue a climate change mitigation strategy.
Given the key role of sh and other aquatic wildlife in dening
human Hg exposure, monitoring biotic Hg concentrations in the marine
environment is essential to the effective implementation of the Mina-
mata Convention. Based on HELCOM (2018a) evaluation, the moni-
toring network in the area of the Polish part of the BS is relatively scarce,
e.g., for 2011–2016 period data series of three or more years were
available only for four points. A good example of this approach is the
Swedish National Monitoring Programme for Contaminants in Marine
Biota (Soerensen and Faxneld, 2022). Our study showed that herring,
which is already included in the HELCOM monitoring strategy, meets
the criteria to become a bioindicator of the Hg level in the pelagic tro-
phic chain. HELCOM suggestions to include seals in the assessment of
the state of the marine environment are only realistic based on random
dead individuals. This irregular access to biological material means that
seals cannot be considered very useful for monitoring in the southern BS.
The unique interplay of perturbations, together with their cascading
effects that create additional stresses make the BS a model region to
study the consequences of occurring and forecasted changes on Hg
pollution in the marine environment. Therefore, the results presented in
this paper can be transferred to other coastal and marginal areas. We
also believe that ndings from this study will contribute to fullling the
goals of the Minamata Convention.
Funding
Financial support for this work was provided by the National Science
Center (2015/17/B/ST10/03418, 2018/31/N/ST10/00214) and the
Ministry of Science and Higher Education (N N304 161637). Agnieszka
Grajewska received nancial support from the Foundation for Polish
Science within the START 2020 scholarship for the most talented young
researchers.
CRediT authorship contribution statement
Agnieszka Jędruch: Conceptualization, Formal analysis, Investiga-
tion, Data curation, Writing – original draft, Writing – review & editing,
Visualization, Funding acquisition. Lucyna Falkowska: Conceptuali-
zation, Investigation, Writing – original draft, Supervision, Funding
acquisition. Dominika Saniewska: Investigation, Writing – original
draft. Agnieszka Grajewska: Investigation. Magdalena Bełdowska:
Investigation, Writing – original draft. Włodzimierz Meissner: Writing
– original draft. El˙
zbieta Kalisi´
nska: Writing – original draft. Kazi-
mierz Duzinkiewicz: Formal analysis. J´
ozef M. Pacyna: Conceptuali-
zation, Writing – original draft, Supervision.
Declaration of competing interest
The authors declare that they have no known competing nancial
interests or personal relationships that could have appeared to inuence
the work reported in this paper.
A. Jędruch et al.
Marine Pollution Bulletin 186 (2023) 114426
13
Data availability
The data presented in this study are available in the Electronic
Supplementary Material
Acknowledgements
The results presented in the paper were initially discussed at the
workshop ‘Mercury in the atmosphere and biosphere: spatial and tem-
poral changes in Poland’, organized during the 14th International
Conference on Mercury as a Global Pollutant (ICMGP) in Krakow,
Poland, from 8 to 13 September 2019. The authors wish to acknowledge
Prof. Jerzy Falandysz from the Medical University in Ł´
od´
z for insightful
comments on the manuscript. The authors also thank Urszula Kwasi-
groch from the University of Gda´
nsk for help with language editing and
proofreading.
In memoriam
Lucyna Falkowska, a Professor at the Faculty of Oceanography and
Geography of the University of Gdansk in Poland, passed away on 7
April 2021. She has been a leader in studies on mercury sources, fate,
behavior, and impacts on the marine environment in Poland. Her wide
research interests included oceanography, marine and atmospheric
chemistry, marine ecology, and most recently, exposure to xenobiotics
and their accumulation in marine organisms. Internationally, her work
contributed to a better understanding of mercury cycling in the marine
environment and the risks posed by the presence of mercury in the
ecosystem. In this way, her efforts helped to develop and implement the
UN Minamata Convention on Mercury. Professor Falkowska was also an
outstanding individual, a wonderful mentor, and a caring friend. She
was loved, admired and respected, and is sorely missed by all those
whose lives she touched.
Appendix A. Supplementary data
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.marpolbul.2022.114426.
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Electronic Supplementary Material
Marine Pollution Bulletin
Mercury in the Polish part of the Baltic Sea: a response to decreased atmospheric deposition
and changing environment
Agnieszka Jędruch a, b, , Lucyna Falkowska a, † , Dominika Saniewska a, Agnieszka Grajewska c, Magdalena Bełdowska a, Włodzimierz
Meissner d, Elżbieta Kalisińska e, Kazimierz Duzinkiewicz f, Józef M. Pacyna g
a University of Gdańsk, Faculty of Oceanography and Geography, Institute of Oceanography, Marszałka Józefa Piłsudskiego 46, 81-378 Gdynia, Poland.
b Polish Academy of Sciences, Institute of Oceanology, Powstańców Warszawy 55, 81-712 Sopot, Poland.
c Institute of Meteorology and Water Management – National Research Institute, Jerzego Waszyngtona 42, 81-342 Gdynia, Poland.
d University of Gdańsk, Faculty of Biology, Wita Stwosza 59, 80-308 Gdańsk, Poland.
e Pomeranian Medical University, Faculty of Pharmacy, Medical Biotechnology and Laboratory Medicine, Powstańców Wielkopolskich 72, 70-111 Szczecin,
Poland.
f Gdańsk University of Technology, Faculty of Electrical and Control Engineering, Gabriela Narutowicza 11/12, 80-233 Gdańsk, Poland.
g AGH University of Science and Technology, Faculty of Energy and Fuels, Adama Mickiewicza 30, 30-059 Kraków, Poland.
Corresponding author: Agnieszka Jędruch (ajedruch@iopan.pl).
† Deceased.
Contents
Supporting information on data collection and treatment ......................................................................................................
II
Figure S1
Map of the southern Baltic Sea with borders of the Polish Exclusive Economic Zone (EEZ) ...................................
III
Figure S2
Temporal changes of annual waterborne input of Hg (Mg) to the Baltic Sea from Polish territory: a. data
presented in previous HELCOM factsheets and reports (Knuuttila, 2008; HELCOM, 2011, 2015, 2018, 2021),
b. data presented in latest HELCOM report (HELCOM 2021) in which results from 1994-2011 were excluded
due to their uncertainty ...........................................................................................................................................
IV
Table S1
Total Hg concentration (ng L-1) in surface water (unfiltered) of the Polish part of the southern Baltic Sea
(emboldened) and other areas of the world ............................................................................................................
V
Table S2
Temporal changes of total Hg concentration (ng L-1) in surface and sub-halocline water (unfiltered) of the Polish
part of the southern Baltic Sea (data are visualized in Fig. 2) .................................................................................
VI
Table S3
Total Hg concentration (ng g-1 dw) in surface sediments of the Polish part of the southern Baltic Sea
(emboldened) and other areas of the world ............................................................................................................
VII
Table S4
Temporal changes of total Hg concentration (ng g-1 dw) in sediments of the Polish part of the southern Baltic
Sea based on data from surface layer collected at Gdańsk Deep (ca. 100 m depth) (Fig. S1) (data are visualized
in Fig. 4a) ...............................................................................................................................................................
VIII
Table S5
Temporal changes of total Hg concentration (ng g-1 dw) in sediments of the Polish part of the southern Baltic
Sea based on data from cores collected at Gdańsk Deep (ca. 100 m depth) (Fig. S1) (data are visualized in Fig.
4b) .........................................................................................................................................................................
IX
Table S6
Total Hg concentration (ng g-1 dw) in representatives of low-trophic level organisms: a. macrophytes (sea lettuce,
seagrass) and b. invertebrates (zooplankton, soft tissues of blue mussel) from the Polish part of the southern
Baltic Sea (emboldened) and other areas of the world ...........................................................................................
X
Table S7
Total Hg concentration (μg g-1 ww) in representatives of intermediate and high-trophic level organisms: a. fish
(herring, cod) birds (gull, cormorant), and b. mammals (harbor porpoise, grey seal) from the Polish part of the
southern Baltic Sea (emboldened) and other areas of the world .............................................................................
XIII
Table S8
Total Hg concentration in the muscles and livers of marine animals with different feeding habits: a. fish, b. birds,
and c. mammals (ng g-1 ww) from the Polish part of the southern Baltic Sea (data are visualized in Fig. 5) ..........
XVI
Table S9
Temporal changes of total Hg concentration in: a. vegetation (macroalgae, vascular plants), b. invertebrates
(zooplankton, soft tissues of benthic fauna) (ng g-1 dw), and c. muscles of fish (European sprat, Baltic herring,
Baltic cod) (ng g-1 ww) from the Polish part of the southern Baltic Sea (data are visualized in Fig. 6 and Fig. 7)
XVII
References ..................................................................................................................................................................................
XIX
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
II
Supporting information on data collection and treatment
Data collection
As already stated in the article, the first studies on the mercury (Hg) level in the southern Baltic Sea ecosystem (Fig. S1) date back
to the 1970s. However, most of the results from that period are very difficult to access, as the journals and reports publishing them
have not been digitized. Without questioning the undeniable scientific value of these pioneering findings, in many cases the results
presented in older papers are uncertain, which is related to the analytical capabilities and techniques used at the time. The methods
commonly used in the 1970s and 1980s detected Hg at the part per million (ppm) level, and, as shown in this study, the Hg
concentrations in the investigated elements of the Baltic Sea environment are generally lower.
To determine the status and the temporal trends of Hg pollution in the Polish part of the Baltic Sea, a total of 100 data sources were
used in this work, including 81 peer-reviewed journal articles, 5 peer-reviewed books or book chapters, 1 peer-reviewed conference
paper, 8 reports, 2 theses, and 3 open-access web material. The discussion of data concerning Hg in the Polish part of the Baltic
Sea was based on additional 266 literature sources, including 250 peer-reviewed scientific papers, 4 peer-reviewed books or book
chapters, 1 peer-reviewed conference paper, 7 reports, 2 theses and 3 open-access web material.
Data treatment
In this review, total Hg concentrations were presented in the following units: ng per L for water; ng per g of dry weight (dw) for
suspended particulate matter (SPM), coastal and marine sediments, as well as organisms representing the initial links of the marine
trophic chain, i.e., primary producers, zooplankton and benthic invertebrates; ng per g of wet weight (ww) for tissues of organisms
form higher trophic levels, i.e., fish, birds and marine mammals. This is due to the dominant way of presenting data for individual
environmental elements in the literature. Because data on water content are rarely reported in scientific studies, converting results
based on estimations could lead to unreliable results. This is especially important for organisms at the first trophic level, in which
wetness varies within wide ranges. However, in some cases, the conversion of the results from wet weight to dry weight and vice
versa was used. The converted values are marked with asterisk (*) and the calculation methods are given under the appropriate
tables.
To determine the time trends of Hg concentration in fish, two models have been proposed: the first, a piecewise linear model, and
the second, a quadratic polynomial model. Due to the limited amount of detailed biometric data on the studied fish, the Hg
concentrations could not be length-normalized.
The risk assessment of Hg-associated health effects in wildlife of the southern Baltic Sea was carried out based on five thresholds,
resulting in five risk categories, i.e., no risk category (NRC), low risk category (LRC), moderate risk category (MRC), high risk category
(HRC), and severe risk category (SRC), proposed for fish and mammals by Dietz et al. (2021). The classification regarding birds was
prepared based on studies by Zilloux et al. (1993), Badzinski et al. (2009), Evers et al. (2011), and Kalisińska et al. (2013, 2014).
Statistical analysis was carried out using STATISTICA 12 (StatSoft). The hypotheses were tested at a statistical significance level of
p=0.05. Figures presented in this study, including Fig. 2, Fig. 4, Fig. 5, Fig. 6, Fig. 7, were created using MS Excel and CorelDRAW
X6. For the preparation of Fig. 3 and Fig. S1, ArcMap 10.4.1 (ESRI) was used, with the WGS 1984 coordination system and the
UTM zone 34N for data projection. Spatial data on the Polish Exclusive Economic Zone (EEZ) were downloaded from the Maritime
Boundaries Geodatabase by FMI (2019). Fig. 1 was reproduced from a study by Jędruch et al. (2021), without changes, under the
terms of the Creative Commons Attribution 4.0 International License (http://creativecommons.org/licenses/by/4.0/). Fig. 3 was
created based on raw data originally used in the study by Kwasigroch et al. (2021) shared by the authors under the terms of the
Creative Commons Attribution 4.0 International License (http://creativecommons.org/licenses/by/4.0/). The spatial interpolation of
point data was based on the inverse distance weighting (IDW) method (Bartier and Keller, 1996) and the geometrical intervals for
symbol classification.
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
III
Figure S1 Map of the southern Baltic Sea with borders of the Polish Exclusive Economic Zone (EEZ).
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
IV
Figure S2 Temporal changes of annual waterborne input of Hg (Mg) to the Baltic Sea from Polish territory: a. Data presented in
previous HELCOM factsheets and reports (Knuuttila, 2008; HELCOM, 2011, 2015, 2018, 2021), b. Data presented in
latest HELCOM report (HELCOM 2021) in which results from 1994-2011 were excluded due to their uncertainty.
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
V
Table S1 Total Hg concentration (ng L-1) in surface water (unfiltered) of the Polish part of the southern Baltic Sea (emboldened)
and other areas of the world.
Region
Area
Year
Mean
Range
Reference
Europe
Sevastopol Bay (Black Sea)
2017
80.0
40.0-120.0
Chelyadina et al., 2022
Vistula Mouth (S Baltic Sea)
2010
7.3
4.7-11.7
Saniewska et al., 2014
Gulf of Trieste (S Adriatic Sea)
1990-1999
6.1
0.2-20.1
Faganeli et al., 2003
1998-1999
5.9
0.5-12.6
Horvat et al., 2003
2004-2005
0.9-1.4
Kotnik et al., 2015
Puck Bay (S Baltic Sea)
2012-2013
5.8
1.4-11.0
Bełdowska, unpubl.
2006-2008
3.1
1.6-7.5
Saniewska et al., 2010
Gulf of Gdańsk (S Baltic Sea)
2012-2013
5.1
1.8-7.3
Bełdowska, unpubl.
2010-2011
5.0
Miotk, 2013
2005
4.3
0.6-19.2
Murawiec et al., 2007
Puck Lagoon (S Baltic Sea)
2011-2013
4.0
0.4-19.0
Bełdowska, unpubl.
Polish coast (S Baltic Sea)
2006-2008
1.7
0.1-3.9
Saniewska et al., 2010
Celtic Sea (NE Atlantic Ocean)
1994-1995
0.1-1.8
Cossa et al., 2004
Venice Lagoon (N Adriatic)
2004-2005
1.2-1.3
Kotnik et al., 2015
Pomeranian Bay (S Baltic Sea)
2009
0.8
Pohl and Hennings, 2009
Adriatic Sea
2004-2005
0.7
0.2-1.4
Kotnik et al., 2015
Gotland Basin (central Baltic Sea)
2014
0.7
0.1-1.3
Kuss et al., 2017
2014-2016
0.5
0.1-2.1
Soerensen et al., 2018
2015
0.2
0.1-0.6
Kuss et al., 2017
2009
0.2
0.1-0.3
Pohl and Hennings, 2009
Belt Sea (W Baltic Sea)
2014-2016
0.6
0.1-1.6
Soerensen et al., 2018
Kaštela Bay (E Adriatic Sea)
1998-1999
0.6
0.1-1.5
Horvat et al., 2003
W Black Sea coast
2005
0.6
Lamborg et al., 2007
W Black Sea
2005
0.5
0.4-0.7
Lamborg et al., 2007
S Baltic Sea
1993-1995
0.4-1.1
Pempkowiak et al., 1998
2009
0.4
0.2-0.8
Pohl and Hennings, 2009
2014-2016
0.2
0.1-0.3
Soerensen et al., 2018
Gulf of Lions (W Mediterranean Sea)
2010
0.3
0.1-0.9
Cossa et al., 2017
E Mediterranean Sea
1998-1999
0.3
0.2-0.5
Horvat et al., 2003
W Mediterranean Sea
1998-1999
0.3
0.2-0.4
Horvat et al., 2003
2010
0.2
0.1-0.3
Cossa et al., 2017
Arkona Basin (W Baltic Sea)
2014-2016
0.3
0.2-0.4
Soerensen et al., 2018
2009
0.2
0.2-0.3
Pohl and Hennings, 2009
Alboran Sea (W Mediterranean Sea)
1991
0.3
0.3-0.5
Cossa et al., 2020
2012
0.1-0.2
Cossa et al., 2020
Bothnian Sea (N Baltic Sea)
2014-2016
0.2
0.1-0.4
Soerensen et al., 2018
Bothnian Bay (N Baltic Sea)
2014-2016
0.2
0.2-0.3
Soerensen et al., 2018
NE Atlantic Ocean
2012
0.0-0.2
Cossa et al., 2020
North America
Long Island Sound (NW Atlantic Ocean)
1995
1.5
0.5-4.0
Rolfhus and Fitzgerald, 2001
Chesapeake Bay (NW Atlantic Ocean)
1997
1.3
0.0-2.6
Lawson et al., 2001
Arctic Ocean
2005
0.6
<1.2
Kirk et al., 2008
2011
0.2
0.1-1.4
Heimbürger et al., 2015
2004-2005
0.2
0.1-0.4
St. Lois et al., 2007
2016
0.2
0.1-0.3
Jonsson et al., 2022
Hudson Bay (Arctic Ocean)
2005
0.5
<1.2
Kirk et al., 2008
NW Atlantic Ocean
1993
0.5
0.2-0.8
Mason et al., 1998
NE Pacific Ocean
2006
0.3
0.1-0.5
Sunderland et al., 2009
Beaufort Sea (Arctic Ocean)
2008
0.2
0.1-0.6
Wang et al., 2012
South America
S Atlantic
0.6
0.2-0.9
Mason and Sullivan, 1999
Asia
East China Sea (W Pacific Ocean)
2014-2015
4.1
0.6-7.2
Liu et al., 2020
Amur Bay (Sea of Japan)
2005
0.5-7.4
Aksentov, 2015
Minamata Bay coast (Shiranui Sea)
2012-2013
2.7
0.7-4.6
Marumoto and Imai, 2015
Yellow Sea (W Pacific Ocean)
2010
1.7
0.9-2.3
Ci et al., 2011
South China Sea (W Pacific Ocean)
2007
1.2
0.8-2.3
Fu et al., 2010
Minamata Bay (Shiranui Sea)
2012-2013
1.1
0.5-1.7
Marumoto and Imai, 2015
NW Pacific Ocean
2002
0.1
0.1-0.8
Laurier et al., 2004
Antarctica
Weddell Sea (Southern Ocean)
2013
0.5
0.1-1.4
Nerentorp Mastromonaco et al., 2017
Southern Ocean
2008
0.2
0.2-0.4
Cossa et al., 2011
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
VI
Table S2 Temporal changes of total Hg concentration (ng L-1) in surface and sub-halocline water (unfiltered) of the Polish part
of the southern Baltic Sea (data are visualized in Fig. 2).
Depth
Year
Region
Mean
Range
Reference
Surface layer
1993
S Baltic Sea
17.6
5.0-28.0
Pohl and Hennings, 2009
1994
S Baltic Sea
6.2
0.0-14.0
Pohl and Hennings, 2009
1995
S Baltic Sea
8.5
3.4-17.0
Pohl and Hennings, 2009
1996
S Baltic Sea
11.7
0.0-36.0
Pohl and Hennings, 2009
1997
S Baltic Sea
2.1
0.0-4.0
Pohl and Hennings, 2009
1998
S Baltic Sea
2.1
0.4-4.0
Pohl and Hennings, 2009
1999
S Baltic Sea
7.8
0.0-29.0
Pohl and Hennings, 2009
2000
S Baltic Sea
2.6
0.4-5.0
Pohl and Hennings, 2009
2001
S Baltic Sea
1.0
0.0-2.0
Pohl and Hennings, 2009
2002
S Baltic Sea
3.1
0.4-7.0
Pohl and Hennings, 2009
2003
S Baltic Sea
3.1
0.4-7.0
Pohl and Hennings, 2009
2004
S Baltic Sea
0.7
0.4-1.0
Pohl and Hennings, 2009
2005
S Baltic Sea
4.8
0.4-13.0
Pohl and Hennings, 2009
Gdańsk Deep (S Baltic Sea)
4.3
0.6-19.2
Murawiec et al., 2007
2006
S Baltic Sea
1.0
1.0-1.0
Pohl and Hennings, 2009
2006-2008
Polish coast (S Baltic Sea)
1.7
0.1-3.9
Saniewska et al., 2010
2007
S Baltic Sea
2.6
0.2-5.0
Pohl and Hennings, 2009
2008
S Baltic Sea
1.0
Pohl and Hennings, 2009
2009
S Baltic Sea
1.0
Pohl and Hennings, 2009
2010
S Baltic Sea
1.0-6.0
Saniewska et al., 2013
Gdańsk Deep (S Baltic Sea)
5.0
Miotk, 2013
2011
Gdańsk Deep (S Baltic Sea)
5.0
Miotk, 2013
2012
Gulf of Gdańsk (S Baltic Sea)
4.6
1.8-7.3
Bełdowska, unpubl.
2013
Gulf of Gdańsk (S Baltic Sea)
5.4
4.6-6.6
Bełdowska, unpubl.
2014
S Gotland Basin (S Baltic Sea)
0.5
Kuss et al., 2017
2015
S Baltic Sea
0.1
0.1-0.1
Soerensen et al., 2018
2016
S Baltic Sea
0.2
0.2-0.3
Soerensen et al., 2018
Below halocline
1993
S Baltic Sea
18.9
6.0-30.0
Pohl and Hennings, 2009
1994
S Baltic Sea
7.4
1.0-16.0
Pohl and Hennings, 2009
1995
S Baltic Sea
8.3
2.0-16.0
Pohl and Hennings, 2009
1996
S Baltic Sea
17.3
1.0-40.0
Pohl and Hennings, 2009
1997
S Baltic Sea
3.0
0.2-8.0
Pohl and Hennings, 2009
1998
S Baltic Sea
3.8
1.0-8.0
Pohl and Hennings, 2009
1999
S Baltic Sea
16.5
2.0-45.0
Pohl and Hennings, 2009
1999-2002
Gdańsk Deep (S Baltic Sea)
1.8
Bełdowski et al., 2009
2000
S Baltic Sea
20.5
2.0-72.0
Pohl and Hennings, 2009
2001
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2002
S Baltic Sea
4.6
0.2-10.0
Pohl and Hennings, 2009
2003
S Baltic Sea
8.4
0.2-20.0
Pohl and Hennings, 2009
2004
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2005
S Baltic Sea
3.7
0.2-9.0
Pohl and Hennings, 2009
Gdańsk Deep (S Baltic Sea)
5.4
0.9-19.6
Murawiec et al., 2007
2006
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2007
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2008
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2009
S Baltic Sea
0.6
0.2-1.0
Pohl and Hennings, 2009
2010
S Baltic Sea
1.0-10.0
Saniewska et al., 2013
Gdańsk Deep (S Baltic Sea)
10.0
Miotk, 2013
2011
Gdańsk Deep (S Baltic Sea)
9.0
Miotk, 2013
2012
Gulf of Gdańsk (S Baltic Sea)
8.2
1.4-15.0
Bełdowska, unpubl.
2013
Gulf of Gdańsk (S Baltic Sea)
4.6
2.1-7.0
Bełdowska, unpubl.
2014
S Gotland Basin (S Baltic Sea)
0.2
0.1-0.2
Kuss et al., 2017
2015
S Baltic Sea
0.2
0.1-0.2
Soerensen et al., 2018
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
VII
Table S3 Total Hg concentration (ng g-1 dw) in surface sediments of the Polish part of the southern Baltic Sea (emboldened) and
other areas of the world.
Region
Area
Year
Mean
Range
Reference
Europe
Nemrut Bay (E Aegean Sea)
2005
5,750
1,700-9,600
Esen et al., 2008
Gulf of Trieste (N Adriatic Sea)
1995-1996
5,240
100-23,300
Covelli et al., 2001
Marano-Grado Lagoon (N Adriatic Sea)
680-9,950
Acqavita et al., 2012
Gulf of Taranto (N Adriatic Sea)
2,770
360-7,730
Spada et al., 2012
Orbetello Lagoon (Tyrrhenian Sea)
2007
1,916
149-10,180
Pepi et al., 2020
Azores (N Atlantic Ocean)
10-1,310
Vieira et al., 2021
Venice Lagoon (N Adriatic Sea)
2005
209-1,177
Han et al., 2007
Gulf of Finland (NE Baltic Sea)
<580
Emelyanov et al., 2017
2003-2010
125
Leipe et al., 2013
2001-2004
100
40-320
Vallius, 2009
2007-2010
70
40-100
Vallius, 2012
Bothnian Bay (N Baltic Sea)
2003-2010
250
Leipe et al., 2013
2014
186
77-310
Apler and Josefsson, 2016
Gdańsk Basin (S Baltic Sea)
2003-2010
150
Leipe et al., 2013
2016-2017
121
1-264
Kwasigroch et al., 2021
1999-2002
110
6-240
Bełdowski and Pempkowiak, 2007
2011-2013
61
2-260
Jędruch et al., 2015
Sado Estuary (NE Atlantic Ocean)
2007
130
8-243
Lillebø et al., 2010
Jade Bay (S North Sea)
2007
108
35-243
Jin et al., 2012
Danish Straits (W Baltic Sea)
2016-2017
106
6-341
Kwasigroch et al., 2021
2014
89
66-120
Apler and Josefsson, 2016
Gotland Basin (central Baltic Sea)
2003-2010
100
Leipe et al., 2013
2014
60
22-436
Apler and Josefsson, 2016
2016-2017
52
7-114
Kwasigroch et al., 2021
Portuguese coast (NE Atlantic Ocean)
10-170
Vieira et al., 2021
Bornholm Basin (S Baltic Sea)
2003-2010
80
Leipe et al., 2013
2014
71
54-86
Apler and Josefsson, 2016
Åland Sea (N Baltic Sea)
2014
78
66-103
Apler and Josefsson, 2016
Bothnian Sea (N Baltic Sea)
2014
74
55-96
Apler and Josefsson, 2016
2016-2017
26
3-78
Kwasigroch et al., 2021
Kaliningrad Bay (S Baltic Sea)
2000-2015
55
8-216
Bogdanov et al., 2020
Puck Bay (S Baltic Sea)
2011-2017
50
1-350
Bełdowska, unpubl.
W Barents Sea (Arctic Ocean)
2015
38
10-88
Lapteva and Plotitsyna, 2017
Svalbard fjord (Arctic Ocean)
2009
24
9-87
Liu et al., 2015
Puck Lagoon (S Baltic Sea)
2011-2013
2-45
Jędruch et al., 2015
Pomeranian Bay (S Baltic Sea)
1999-2002
9
6-13
Bełdowski et al., 2014
Puck Lagoon coast (S Baltic Sea)
2019-2020
3
1-26
Graca et al., 2022
North America
San Francisco Bay (NE Pacific Ocean)
2001-2002
170
9-560
Lu et al., 2005
Chesapeake Bay (NW Atlantic Ocean)
1995
80-180
Mason et al., 1999
Mexican coast (NW Atlantic Ocean)
20-130
Vieira et al., 2021
Beaufort Sea (Arctic Ocean)
1999-2001
41
3-97
Trefry et al., 2003
Chukchi Sea (Arctic Ocean)
2009-2010
31
5-55
Fox et al., 2014
White Sea (Arctic Ocean)
2004
23
6-95
Fedorov et al., 2019
South America
Patos Lagoon (SW Atlantic Ocean)
2,390
20-17,840
Mirlean et al., 2003
Asia
Saudi Arabia coast (E Red Sea)
2018
1,830
680-3,800
Kahal et al., 2018
Haifa Bay (E Mediterranean Sea)
2007
255
79-347
Shoham-Frider et al., 2012
Bohai Sea (Yellow Sea)
2012
210
10-1,200
Zhuang and Gao, 2015
Amur Bay (Sea of Japan)
2004
120
58-196
Polyakov et al., 2008
Sea of Okhotsk (NW Pacific Ocean)
2011-2016
76
19-158
Sattarova and Aksentov, 2018
South China Sea (W Pacific Ocean)
2000-2002
54
2-201
Shi et al., 2010
East Siberia Sea (Arctic Ocean)
36
13-92
Aksentov et al., 2021
Persian Gulf
2006
22
8-34
Hassan et al., 2019
Africa
Bizerte Lagoon (S Mediterranean Sea)
1999-2000
130
10-650
Mzoughi et al., 2002
Berg River estuary (SE Atlantic Ocean)
2007
32
15-50
Kading et al., 2009
Tadio Lagoon (Gulf of Guinea)
2019
80
1-120
Mason et al., 2021
Ebrié Lagoon (Gulf of Guinea)
2019
60
1-110
Mason et al., 2021
Aby Lagoon (Gulf of Guinea)
2019
60
50-90
Mason et al., 2021
Comoe River estuary (Gulf of Guinea)
2019
50
1-120
Mason et al., 2021
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
VIII
Table S4 Temporal changes of total Hg concentration (ng g-1 dw) in sediments of the Polish part of the southern Baltic Sea based
on data from surface layer collected at Gdańsk Deep (ca. 100 m depth) (Fig. S1) (data are visualized in Fig. 4a).
Year
Mean
Range
Notes
Reference
1991-1993
100
Data extracted from the graph
Szczepańska and Uścinowicz, 1994
1994-1997
220
Data extracted from the graph
Uścinowicz et al., 2011
1999-2002
116
81-148
Bełdowski and Pempkowiak, 2007
2000
112
Data extracted from the graph
Leipe et al., 2013
2010
195
160-230
Data extracted from the graph
Miotk, 2013; Bełdowski et al., 2014
2011
172
156-187
Jędruch et al., 2013
175
Miotk, 2013
2012
83
44-201
Jędruch et al., 2015
2016
98
89-105
Kwasigroch et al., 2021
2017
119
97-138
Kwasigroch et al., 2023
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
IX
Table S5 Temporal changes of total Hg concentration (ng g-1 dw) in sediments of the Polish part of the southern Baltic Sea based
on data from cores collected at Gdańsk Deep (ca. 100 m depth) (Fig. S1) (data are visualized in Fig. 4b).
Bełdowski and Pempkowiak, 2003
Leipe et al., 2013
Kwasigroch et al., 2023
Marker
Year
Mean
Notes
Marker
Year
Mean
Notes
Marker
Year
Mean
Notes
1998
1993
1988
1983
1979
1969
1960
1950
1940
1931
1921
1912
1899
1875
81
101
148
86
145
165
138
139
133
106
101
104
69
40
Sediment accumulation rate
used in dating calculations: 2.9
mm year-1
2007
2001
1993
1986
1979
1972
1964
1958
1951
1944
1938
1931
1924
1917
1910
1903
1896
1889
1882
135
175
210
203
193
130
155
153
150
120
105
70
60
45
40
35
33
33
30
Data extracted from the
graph; sediment accumulation
rate used in dating
calculations: 2.1 mm year-1
2013
2009
2005
2001
1997
1993
1989
1985
1981
1977
1973
1965
1957
1949
1941
1933
1925
1917
1909
1901
1887
97
113
117
128
139
143
139
146
156
153
143
116
118
102
105
92
82
76
78
74
72
Sediment accumulation rate
used in dating calculations: 2.5
mm year-1
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
X
Table S6 Total Hg concentration (ng g-1 dw) in representatives of low-trophic level organisms: a. macrophytes (sea lettuce,
seagrass) and b. invertebrates (zooplankton, soft tissues of blue mussel) from the Polish part of the southern Baltic
Sea (emboldened) and other areas of the world.
Region
Area
Year
Mean
Range
Reference
a. Macrophytes
Sea lettuce
(Ulva spp.)
Europe
Kadin Creek estuary (Aegan Sea)
2011-2012
510
120-1,090
Yozukmaz et al., 2018
Aveiro Lagoon (NE Atlantic Ocean)
248
20-2,150
Coelho et al., 2005
Canary Islands (NE Atlantic Ocean)
1994-1995
140
30-250
Hardisson et al., 1998
Aegean Sea (E Mediterranean Sea)
2006
110
48-336
Akcali and Kucuksezgin, 2011
Venice Lagoon (N Adriatic Sea)
2011
70
Dominik et al., 2014
Spanish market
18
1-57
Besada et al., 2009
Pomeranian Bay (S Baltic Sea)
2006-2012
17
5-28
Bełdowska et al., 2015
Puck Bay (S Baltic Sea)
2006-2012
9
2-37
Bełdowska et al., 2015
Gulf of Gdańsk (S Baltic Sea)
2006-2012
8
2-30
Bełdowska et al., 2015
Polish coast (S Baltic Sea)
2006-2012
7
4-15
Bełdowska et al., 2015
North America
Guaymas Bay (Gulf of California)
1998
134
95-173
Green-Ruiz et al., 2005
Aleutian Islands (N Pacific Ocean)
2004
10
15
Burger et al., 2007 *
Asia
Yellow Sea (W Pacific Ocean)
2006
5
2-9
Ok et al., 2007
Africa
S Mediterranean Sea
2002
130
110-143
Mohamed and Khaled, 2005
SW Indian Ocean
2002-2003
118
20-472
Misheer et al., 2006
E Atlantic Ocean
2013
9
5-14
Diop and Amara, 2016
Seagrass
(Zostera spp.)
Europe
W Mediterranean Sea
1991
95
45-185
Sanchiz et al., 2000
S Black Sea
2013
50
Bat et al., 2016
2015-2016
17
Arici and Bat, 2020
Aveiro Lagoon (NE Atlantic Ocean)
46
17-81
Coelho et al., 2009, 2013
Pomeranian Bay (S Baltic Sea)
2006-2012
43
35-51
Bełdowska et al., 2015
SE Black Sea
2020
<50
Neshovska et al., 2021
Harboøre Tange (E North Sea)
2014-2016
36
22-74
Bjerregaard et al., 2020
Venice Lagoon (N Adriatic Sea)
2011
30
Dominik et al., 2014
Puck Bay (S Baltic Sea)
2006-2012
12
4-19
Bełdowska et al., 2015
Gulf of Gdańsk (S Baltic Sea)
2006-2012
9
5-16
Bełdowska et al., 2015
Puck Lagoon (S Baltic Sea)
2012-2013
8
5-13
Jędruch et al., 2019b
Polish coast (S Baltic Sea)
2006-2012
7
3-22
Bełdowska et al., 2015
North America
Great Bay Estuary (NW Atlantic Ocean)
1997
28
16-43
Pannhorst and Weber, 1999
Asia
Onsan Bay (Sea of Japan)
2014
40
15-100
Lee et al., 2019
Koye Bay (Sea of Japan)
2014
10
<20
Lee et al., 2019
Jaran Bay (Sea of Japan)
2014
10
<20
Lee et al., 2019
Africa
E Atlantic Ocean
2015
24
12-37
Boutahar et al., 2019
b. Invertebrates
Zooplankton
(mixed species)
Europe
Aveiro Lagoon (NE Atlantic Ocean)
2010-2011
539
49-1176
Cardoso et al., 2013 *
Gulf of Trieste (N Adriatic Sea)
2003-2004
371
Živković et al., 2017 *
Venice Lagoon (N Adriatic Sea)
2011
330
Dominik et al., 2014
Puck Lagoon (S Baltic Sea)
2012-2013
106
16-596
Bełdowska and Mudrak-Cegiołka,
2017
Gulf of Lions (W Mediterranean Sea)
2010
101
8-570
Chouvelon et al., 2019
Gulf of Gdańsk (S Baltic Sea)
2012-2013
55
17-119
Bełdowska, unpubl.
Puck Bay (S Baltic Sea)
2012-2013
47
31-69
Bełdowska, unpubl.
Central and N Baltic Sea
1991-1993
42
21-119
Nfon et al., 2009 *
Adriatic Sea
2003-2004
25
14-42
Živković et al., 2017 *
W Mediterranean Sea
2003-2004
20
7-35
Živković et al., 2017 *
Svalbard fjord (Arctic Ocean)
2007
6
4-8
Ruus et al., 2015
2018
5
1-11
Carrasco, 2019
North America
Beaufort Sea (Arctic Ocean)
2006-2012
68
2-134
Pomerleau et al., 2016
Arctic Ocean
2005-2006
59
1-123
Pomerleau et al., 2016
Long Island Sound (NW Atlantic Ocean)
2015
48
1-161
Gosnell et al., 2017 *
30
26-35
Lee and Fisher, 2016
Gulf of Lawrence (NW Atlantic Ocean)
2006
46
13-65
Lavoie et al., 2010
Baffin Bay (Arctic Ocean)
2005-2006
42
14-70
Pomerleau et al., 2016
Hudson Bay (Arctic Ocean)
2003-2010
41
5-242
Foster et al., 2012
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XI
Table S6 Continued.
Region
Area
Year
Mean
Range
Reference
NW Atlantic Ocean
2009-2010
29
6-106
Hammerschmid et al., 2013 *
South America
Guanabara Bay (SW Atlantic Ocean)
2005
105
31-199
Kehrig et al., 2009
S Atlantic Ocean
2007-2008
36
10-70
Seco et al., 2021
2016-2017
45
7-150
Seco et al., 2021
Asia
Minamata Bay (Shiranui Sea)
1,250
156-5,400
Gajbhiye and Hirota, 1993
NW Persian Gulf
171
10-530
Al-Imarah et al., 2017
100-470
Hashemi and Safahieh, 2018
Chukchi Sea (Arctic Ocean)
1998
42
22-62
Pomerleau et al., 2016
Laptev Sea (Arctic Ocean)
2008
21
3-39
Pomerleau et al., 2016
Kuwait Bay (Persian Gulf)
1996-1998
11
4-35
Al-Majed and Preston, 2000
Malacca Straits (Andaman Sea)
2000
2
1-5
Rezai et al., 2003
Oceania
Samoa (S Pacific Ocean)
2011
57
14-112
Gosnell and Mason, 2015 *
Kiribati Islands (central Pacific Ocean)
2011
39
8-189
Gosnell and Mason, 2015 *
Hawaii Islands (N Pacific Ocean)
2011
10
7-13
Gosnell and Mason, 2015 *
Antarctica
Ross Sea (Southern Ocean)
1990-1991
77
50-105
Bargagli et al., 1998
Terra Nova Bay (Southern Ocean)
1990-1991
65
48-150
Bargagli et al., 1998
South Shetland Islands (Southern Ocean)
2011
7
3-15
Matias et al., 2022
Blue mussel
(Mytilus spp.)
Europe
Bosporus (SW Black Sea)
2002-2004
1,410
140-2,860
Kayhan, 2007
Venice Lagoon (N Adriatic Sea)
2011
545
Dominik et al., 2014
Bay of Kotor (SE Adriatic Sea)
2009
501
140-2,651
Jović and Stanković, 2014
Ionian Sea (Mediterranean Sea)
368
Di Leo et al., 2010
Harboøre Tange (E North Sea)
2014-2016
340
180-500
Bjerregaard et al., 2020
Apulia coast (SW Adriatic Sea)
2009
272
100-810
Spada et al., 2013
Biscay Bay (NE Atlantic Ocean)
1990-2008
220
20-420
Solaun et al., 2013
Balearic Sea (W Mediterranean Sea)
2005
204
158-256
Deudero et al., 2009
NE Atlantic Ocean
2010
171
100-320
Briant et al., 2017
Marmara Sea
2016
165
48-288
Türkoğlu et al., 2021
Puck Bay (S Baltic Sea)
2012-2013
148
45-671
Jędruch et al., 2019a
Sevastopol Bay (Black Sea)
2017
110
50-190
Chelyadina et al., 2022
SE Adriatic Sea
2013
108
54-216
Bilandžić et al., 2015
Gulf of Trieste (N Adriatic Sea)
2008
122
70-237
Ramšak et al., 2012
Pomeranian Bay (S Baltic Sea)
1997
106
34-186
Szefer et al., 2002
Varna Bay (W Black Sea)
2016
85
55-135
Zhelyazkov et al., 2018
Funen (W Baltic Sea)
2014-2016
81
50-120
Bjerregaard et al., 2020
North Sea
2000-2018
80
<7,600
ICES, 2019 *
Iceland fjord (North Atlantic Ocean)
2018
73
63-89
Hixon, 2019
41
32-51
Hixon, 2019
Bergen Harbor (North Sea)
2001-2002
70
54-162
Airas et al., 2004
Calich Lagoon (W Mediterranean Sea)
2017
60
52-84
Esposito et al., 2021
2019
52
44-56
Meloni et al., 2022
Central Baltic Sea
2000-2018
40
<280
ICES, 2019 *
Danish Straits (W Baltic Sea)
2000-2018
40
<5,600
ICES, 2019 *
Norwegian Sea
2000-2018
40
40-160
ICES, 2019 *
Barents Sea (Arctic Ocean)
2000-2018
40
40-80
ICES, 2019 *
Puck Lagoon coast (S Baltic Sea)
2019-2020
36
17-61
Graca et al., 2022
Puck Lagoon (S Baltic Sea)
2017
23
16-29
Jędruch and Bełdowska, 2020
North America
Gulf of Maine (NW Atlantic Ocean)
1991-1997
470
240-670
Chase et al., 2001
Baja California (E Pacific)
192
53-331
Gutiérrez-Galindo and Muñoz-
Barbosa, 2003
2007
46
11-110
Uc-Peraza et al., 2021
Bay of Fundy (NW Atlantic Ocean)
2005-2011
170
90-380
Elskus et al., 2017
Aleutian Islands (N Pacific Ocean)
2008
110
50-220
Savoy et al., 2017
2004
86
43-208
Burger and Gochfeld, 2006 *
Groswater Bay (Labrador Sea)
2013
21
1-35
Schartup et al., 2015 *
Asia
Bohai Sea (Yellow Sea)
298
59-1,047
Liang et al., 2003
Ulsan and Onsan Bays (Sea of Japan)
2003
51
9-114
Kim and Choi, 2017
Changseon (Sea of Japan)
2008-2013
38
16-81
Mok et al., 2014
Korean market
2005-2008
26
19-34
Moon et al., 2011
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XII
Table S6 Continued.
Region
Area
Year
Mean
Range
Reference
Africa
Oran Bay (SW Mediterranean Sea)
2011
895
240-2,270
Benzaoui et al., 2015
NE Atlantic
2004-2005
700
10-2,310
Maanan, 2007
S Mediterranean Sea
2003
53
45-69
Galgani et al., 2014
Oceania
Hawaii Islands (N Pacific Ocean)
2014-2015
43
13-349
Motta et al., 2019
* Values converted from wet weight.
To convert the Hg concentration in wet weight (ww) to dry weight (dw) the following conversion factors were applied: sea lettuce: 5.0 (da Silva et al., 2008),
zooplankton: 7.0 (Kiørboe, 2013), blue mussel: 4.0 (Pleissner et al., 2012).
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XIII
Table S7 Total Hg concentration (μg g-1 ww) in representatives of intermediate and high-trophic level organisms: a. fish (herring,
cod) birds (gull, cormorant), and b. mammals (harbor porpoise, grey seal) from the Polish part of the southern Baltic
Sea (emboldened) and other areas of the world.
Muscle
Liver
Region
Area
Year
Mean
Range
Mean
Range
Reference
a. Fish
Herring
(C. harengus)
Europe
S Gulf of Finland (NE Baltic Sea)
1994-2001
0.14
0.05-0.23
0.16
Voigt, 2003
N Gulf of Finland (NE Baltic Sea)
1994-2001
0.10
0.01-0.19
0.12
0.03-0.21
Voigt, 2003
English Channel (NE Atlantic Ocean)
2012
0.08
0.05-0.12
Henry et al., 2017 *
Norwegian Sea
2006-2007
0.04
<0.01-0.40
Frantzen et al., 2015
S Baltic Sea
1994-2003
0.04
0.02-0.08
Polak-Juszczak, 2009
2006-2009
0.02
<0.01-0.08
0.04
0.01-0.14
Falkowska et al., 2010;
Kwaśniak et al., 2012
Bothnian Bay (N Baltic Sea)
2000-2010
0.04
0.02-0.06
Millegård, 2003
0.03
0.01-0.39
SEPA, 2019
0.03
<0.01-0.14
ICES, 2019
Gulf of Riga (E Baltic Sea)
2002-2015
0.03
<0.01-0.10
ICES, 2019
Danish Straits (W Baltic Sea)
1997-2014
0.03
0.01-0.07
ICES, 2019
Central Baltic Sea
2000-2018
0.02
<0.01-0.33
ICES, 2019
2000-2014
0.02
0.01-0.15
SEPA, 2019
SW Baltic Sea
2014
0.02
0.01-0.05
ICES, 2019
N Baltic Sea
1991-1993
0.02
0.01-0.04
Nfon et al., 2009
Svalbard fjord (Arctic Ocean)
2006
0.01
<0.01-0.02
Jæger et al., 2009
North America
San Francisco Bay (NE Pacific Ocean)
2005-2007
0.02
Greenfeld and Jahn, 2010
Kotzebue Sound (Chukchi Sea)
2015-2016
<0.01
<0.01-0.01
Cyr, et al., 2019
Asia
Korean market
2005-2008
0.15
0.07-0.23
Moon et al., 2011
Cod
(G. morhua)
Europe
NE North Sea
2000-2016
0.11
0.02-0.42
ICES, 2019
Norwegian Sea
2000-2016
0.07
0.01-0.33
ICES, 2019
Danish Straits (W Baltic Sea)
2000-2014
0.06
0.02-0.13
ICES, 2019
S Baltic Sea
2006-2009
0.05
<0.01-0.25
0.05
<0.01-0.39
Falkowska et al., 2010;
Kwaśniak et al., 2012
1997-2003
0.04
0.03-0.05
Polak-Juszczak, 2009
W North Sea
2000-2002
0.05
0.01-0.23
ICES, 2019
S North Sea
1997-1999
0.05
0.03-0.07
Baeyens et al., 2003
Iceland (NE Atlantic Ocean)
2000-2017
0.04
0.01-0.20
ICES, 2019
Barents Sea (Arctic Ocean)
2009-2010
0.04
0.01-0.16
0.02
<0.01-0.12
Julshamn et al., 2013
Central Baltic Sea
2000-2003
0.04
0.01-0.11
SEPA, 2019
2000-2018
0.03
<0.01-0.11
ICES, 2019
NE Barents Sea (Arctic Ocean)
2019
0.03
0.02-0.44
Gopakumar et al., 2021
Celtic Sea (NE Atlantic Ocean)
1994
0.04
0.01-0.07
Nixon et al., 1995
Faroe Islands (N Atlantic Ocean)
2000-2012
0.04
<0.01-0.10
ICES, 2019
SW Baltic Sea
2011-2014
0.02
0.01-0.03
ICES, 2019
Svalbard fjord (Arctic Ocean)
2006
0.01
<0.01-0.02
<0.01-0.01
Jæger et al., 2009
NW Barents Sea (Arctic Ocean)
2019
<0.01
<0.01-0.1
Gopakumar et al., 2021
North America
Aleutian Islands (N Pacific Ocean)
2004
0.17
0.01-0.86
0.11
0.01-1.25
Burger and Gochfeld, 2007
Bay of Fundy (NW Atlantic Ocean)
2004-2005
0.05
0.03
Jardine et al., 2009 *
Asia
Korean market
2005-2008
0.13
0.06-0.19
Moon et al., 2011
b. Birds
Gull
(Larus spp.)
Europe
Svalbard fjord (Arctic Ocean)
2003-2005
1.5
0.3-4.3
Sagerup et al., 2009
Wadden Sea (S North Sea)
1990
1.0
0.3-1.6
1.3
0.4-3.4
Lewis et al., 1993 *
W Mediterranean Sea
1.3
0.3-2.5
Albertos et al., 2020
0.4
0.3-0.5
Albertos et al., 2020
Spanish coast (NE Atlantic Ocean)
2014-2016
0.8
0.1-4.6
Vizuete et al., 2022 *
Ebro River delta (W Mediterranean Sea)
1997-1999
0.8
Arcos et al., 2002 *
Gulf of Gdańsk (S Baltic Sea)
2009-2012
0.2
<0.1-1.5
0.4
<0.1-1.9
Szumiło-Pilarska et al., 2016 *
0.1
<0.1-0.6
0.3
<0.1-1.5
Szumiło-Pilarska et al., 2016 *
0.1
<0.1-0.3
0.2
<0.1-0.7
Szumiło-Pilarska et al., 2016 *
0.1
<0.1-0.3
0.2
<0.1-0.5
Szumiło-Pilarska et al., 2016 *
Barents Sea (Arctic Ocean)
1991-1992
<0.1
<0.1-0.1
0.3
0.1-0.5
Savinov et al., 2003 *
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XIV
Table S7 Continued.
Muscle
Liver
Region
Area
Year
Mean
Range
Mean
Range
Reference
North America
SE Florida (USA)
2014-2016
1.3
0.1-4.9
19.1
0.3-146.6
Nay et al., 2022 *
Labrador Sea (N Atlantic Ocean)
0.9
0.2-2.0
Mallory et al., 2018 *
Great Salt Lake (USA)
2006
1.4
0.9-1.8
Conover and Vest, 2009 *
New York Bight (NW Atlantic Ocean)
0.7
0.6-0.9
1.4
1.2-1.5
Burger et al., 2000
0.4
0.3-0.4
0.8
0.7-0.9
Burger et al., 2000
Neponeset Reservoir (USA)
2007
0.7
0.5-0.8
Conover and Vest, 2009 *
Aleutian Islands (N Pacific Ocean)
2004
0.3
<0.9
Burger et al., 2007
Asia
S Caspian Sea
2008
2.8
1.5-3.8
Rajaei et al., 2010
NW Persian Gulf
2007
0.3
0.2-0.4
1.3
0.9-1.7
Zamani-Ahmadmahmoodi
et al., 2014 *
0.1
Hashemi et al., 2015 *
East Siberian Sea (Arctic Ocean)
1993
0.2
0.2-0.3
1.2
0.5-2.0
Kim et al., 1996 *
Persian Gulf
2011
0.2
0.1-0.3
0.5
0.2-1.0
Majidi et al., 2015 *
Cormorant
(Phalacrocorax
spp.)
Europe
Ketelmeer Lake (Netherlands)
16.4
4.7-49.2
Platteeuw et al., 1995 *
W Mediterranean Sea
14.6
1.4-110.6
Albertos et al., 2020
5.6
2.3-10.8
Albertos et al., 2020
Spanish coast (NE Atlantic Ocean)
2002-2003
1.2
0.5-2.0
8.0
3.0-12.9
Sanpera, et al., 2008 *
2002-2003
4.2
0.9-12.8
Carbonell et al., 2007
Moravia (Czech Republic)
2003
1.0
0.6-1.4
10.1
3.7-16.2
Houserova et al., 2005 *
Vojvodina (Serbia)
2010
0.7
0.4-1.0
1.5
0.8-2.1
Skoric et al., 2012
S Baltic Sea
2006
0.6
0.1-1.6
4.1
0.4-22.4
Misztal-Szkudlińska et al.,
2018 *
Aegean Sea (E Mediterranean Sea)
1999-2000
2.3
1.8-2.8
Goutner et al. 2011 *
Třeboň (Czech Republic)
2015
0.6
0.5-0.7
2.1
1.9-2.3
Kral et al., 2017 *
Ebro River delta (W Mediterranean Sea)
1997-1999
1.5
Arcos et al., 2002 *
Balaton Lake (Hungary)
1.3
0.3-2.2
Lehel et al., 2013 *
Tisza River (Hungary)
2020
0.3
0.1-0.5
Lehel et al., 2022 *
North America
Florida (USA)
1994-1997
24.0
0.4-250.0
Sepulveda et al., 1998
2014-2016
3.2
0.1-25.4
14.7
0.5-125.4
Nay et al., 2022 *
Aleutian Islands (N Pacific Ocean)
2000-2001
4.3
Ricca et al., 2008 *
Winnipeg Lake (Canada)
2009-2010
0.7
0.2-2.0
Ofukany et al., 2015
Gulf of California (E Pacific Ocean)
2002
0.4
Ruelas-Inzunza et al, 2009 *
Lake Erie (Canada)
2008
2.0
1.0-4.1
Robinson et al., 2011
South America
Madeira River (W Amazon)
2017
0.2
0.2-0.2
0.6
0.4-0.7
Dias dos Santos et al., 2021
Asia
Caspian Sea
2006
2.3
Mollazadeh et al., 2011 *
S Caspian Sea
2009
0.7
0.3-2.4
1.6
1.1-1.9
Aazami et al., 2011 *
2011
0.7
0.4-0.9
1.6
1.1-2.1
Aazami and KianiMehr, 2018
2004
0.6
0.3-1.0
2.3
0.9-5.8
Mazloomi et al., 2008 *
Biwa Lake (Japan)
1993
0.5
0.3-0.7
1.7
0.9-2.5
Saeki et al., 2000
Mia Lake (Japan)
2003
0.4
0.1-0.9
2.1
0.6-9.2
Nam et al., 2005 *
Tokyo (Japan)
1993-1994
0.3
0.1-0.5
1.2
0.7-1.7
Saeki et al., 2000
2001-2002
<0.01
<0.01
0.1
<0.1-0.2
Horai et al., 2007 *
Africa
Kariba Lake (Zimbabwe)
1986
0.7
0.2-1.5
1.4
0.3-3.8
Douthwaite et al., 1992 *
C. Mammals
Harbor
porpoise
(P. phocoena)
Europe
Danish Straits (W Baltic Sea)
2011
68.4
61.3-75.5
Dietz et al., 2021
1998-1999
6.4
<26
Strand et al., 2005
1987-1988
0.9
0.3-1.7
7.1
0.2-31.0
Joiris et al., 1991
2015-2019
0.7
0.1-1.5
14.4
0.4-64.7
SEPA, 2021
1989-1990
4.0
Teigen et al., 1993
Portuguese coast (NE Atlantic Ocean)
2005-2013
1.8
0.3-5.9
20.8
0.3-102.4
Ferreira et al., 2016
S North Sea
1987-1990
1.7
0.3-6.5
39.8
0.3-132.0
Joiris et al., 1991
1997-2003
25.8
0.9-139.4
Lahaye et al., 2007
1991
19.3
Kastelein and Lavaleije,
1992
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XV
Table S7 Continued.
Muscle
Liver
Region
Area
Year
Mean
Range
Mean
Range
Reference
Irish Sea (NE Atlantic Ocean)
1997-2003
30.0
0.8-165.0
Lahaye et al., 2007
21.8
0.6-190.0
Law et al., 1992
Bay of Biscay (NE Atlantic Ocean)
1997-2003
17.9
1.0-65.5
Lahaye et al., 2007
Celtic Sea (NE Atlantic Ocean)
1997-2003
16.1
0.3-54.9
Lahaye et al., 2007
1989-1993
1.1
0.7-1.6
6.5
1.1-26.8
Das et al., 2003 *
W North Sea
1988-1990
13.5
0.6-150
Law et al., 1991
1997-2003
8.7
0.4-31.0
Lahaye et al., 2007
NE Atlantic Ocean
1997-2003
12.4
1.2-26.0
Lahaye et al., 2007
2004-2015
11.6
<27.7
Fernandez et al., 2022 *
Norwegian coast (E North Sea)
1989-1990
4.5
Teigen et al., 1993
Norwegian coast (Norwegian Sea)
1989-1990
2.3
Teigen et al., 1993
Norwegian coast (Barents Sea)
1989-1990
3.0
Teigen et al., 1993
English Channel (NE Atlantic Ocean)
1997-2003
10.7
0.8-40.0
Lahaye et al., 2007
2004-2015
5.7
<12.3
Fernandez et al., 2022 *
1989-1993
1.0
0.5-1.5
2.4
0.8-7.0
Das et al., 2003 *
E North Sea
1998-1999
8.5
<19
Strand et al., 2005
S Baltic Sea
1996-2003
5.9
0.4-58.6
Ciesielski et al., 2006 *
Norwegian Sea
2016-2017
0.1-17.1
Ciesielski, unpubl.
Black Sea
1997-1999
0.3
0.1-0.6
2.7
0.1-8.7
Joiris et al., 2001
1997-1998
2.2
0.2-8.2
Das et al., 2004 *
North America
NW Atlantic Ocean
9.9
0.6-38.6
Mackey et al., 1995
W Pacific Ocean
2002-2009
0.6
0.01-1.65
Wintle et al., 2011
Labrador Sea (N Atlantic Ocean)
1999
6.6
Strand et al., 2005
1988-1989
0.5
<0.01-1.1
4.2
0.5-20.7
Paludan-Müller et al., 1993
Asia
Japanese coast (NW Pacific Ocean)
1985-2010
2.3
<6.1
Yasuda et al., 2012 *
Grey seal
(H. grypus)
Europe
W North Sea
1988
99.6
<0.1-430.0
Law et al., 1991
Bothnian Sea (N Baltic Sea)
2017-2018
4.6-359.7
Dietz et al., 2021
2007-2010
14.4
4.0-33.5
Helsingen, 2011 *
English Channel (NE Atlantic Ocean)
1989-1993
2.2
99.5
88.1-110.5
Das et al., 2003 *
S Baltic Sea
2007-2010
102.7
6.3-247.8
Helsingen, 2011 *
2001-2007
0.5
0.1-1.5
80.5
0.4-263.4
Bełdowska and Falkowska,
2016 *
1996-2003
30.5
0.1-150.5
Ciesielski, 2006 *
Faroe Islands (N Atlantic Ocean)
1993-1995
0.7
0.2-4.6
59.7
1.1-238.0
Bustamante et al., 2004
Central Baltic Sea
2007-2010
58.7
1.7-222.4
Helsingen, 2011 *
SW Baltic Sea
2007-2010
52.2
Helsingen, 2011 *
Bothnian Bay (N Baltic Sea)
1996-1998
0.7
0.4-1.6
78.0
15.0-348.0
Nyman et al., 2002
2007-2010
27.6
Helsingen, 2011 *
Danish Straits (W Baltic Sea)
2010-2018
19.2
1.7-88.7
Dietz et al., 2021
Norwegian Sea
1989-1990
13.5
0.7-48.3
Skaare et al., 1994
Åland Sea (N Baltic Sea)
2007-2010
10.3
8.9-11.7
Helsingen, 2011 *
North America
NW Atlantic Ocean
1996-1998
0.6
0.2-1.0
109.0
27.0-278.0
Nyman et al., 2002
Gulf of Lawrence (NW Atlantic Ocean)
2015-2019
0.3
0.1-1.9
28.7
3.8-290.0
MacMillan et al., 2022
* Values have been converted from dry weight.
To convert the Hg concentration in dry weight (dw) to wet weight (ww) the following conversion factors were applied: fish: 4.0 (Nogueira et al., 2013); bird
liver: 3.6; bird muscle: 3.4 (Kalisińska and Salicki, 2010); marine mammal liver: 3.7; marine mammal muscle: 3.4 (Joiris et al., 1991).
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XVI
Table S8 Total Hg concentration in the muscles and livers of marine animals with different feeding habits: a. fish, b. birds, and
c. mammals (ng g-1 ww) from the Polish part of the southern Baltic Sea (data are visualized in Fig. 5).
* Values have been converted from dry weight.
To convert the Hg concentration in dry weight (dw) to wet weight (ww) the following conversion factors were applied: bird liver: 3.6; bird muscle: 3.4 (Kalisińska
and Salicki, 2010); marine mammal liver: 3.7; marine mammal muscle: 3.4 (Joiris et al., 1991).
Muscle
Liver
Feeding habit
Period
N
Mean
Range
N
Mean
Range
Reference
a.
Fish
European sprat (S. sprattus)
planktivorous
2015-2016
11
16
10-23
Polak-Juszczak, 2017
Baltic herring (C. harengus)
planktivorous
2007-2008
81
24
4-78
17
44
10-139
Kwaśniak et al., 2012
Common bream (A. brama)
benthivorous
2015-2016
10
34
2-102
Polak-Juszczak, 2017
European plaice (P. platessa)
benthivorous
2009-2010
51
51
33-79
51
40
10-55
Polak-Juszczak, 2012
Common roach (R. rutlius)
benthivorous
2015-2016
9
52
22-95
Polak-Juszczak, 2017
European flounder (P. flesus)
benthivorous
2009-2010
74
77
45-106
74
31
17-83
Polak-Juszczak, 2012
Turbot (S. maximus)
benthivorous
2009-2010
50
114
47-212
50
85
66-141
Polak-Juszczak, 2012
European Eel (A. anguilla)
benthivorous
2000
15
150
20-400
Wyrzykowska et al., 2012
Sea trout (S. trutta)
piscivorous
2015-2016
4
42
32-51
Polak-Juszczak, 2017
Baltic cod (G. morhua)
piscivorous
2006-2009
62
46
9-249
26
52
5-392
Kwaśniak and Falkowska, 2012
Atlantic salmon (S. salar)
piscivorous
2007-2008
6
48
15-80
3
73
70-76
Kwaśniak et al., 2012
Burbot (L. lota)
piscivorous
2015-2016
3
66
54-73
Polak-Juszczak, 2017
Pike-perch (S. lucioperca)
piscivorous
2015-2016
16
88
29-286
Polak-Juszczak, 2017
European perch (P. fluvitilis)
piscivorous
2007-2008
15
151
26-453
11
73
8-191
Kwaśniak et al., 2012
b.
Birds
Mallard (A. platyrhynchos)
omnivorous
2005-2006
50
37
2-261
50
73
<284
Kalisińska et al., 2013 *
Black-headed gull (L. ridibundus)
omnivorous
2010-2012
19
71
4-242
18
184
16-701
Szumiło-Pilarska et al., 2016 *
Common gull (L. canus)
omnivorous
2010-2012
16
103
8-536
16
317
39-1,594
Szumiło-Pilarska et al., 2016 *
Herring gull (L. argentatus)
omnivorous
2010-2012
60
98
12-299
58
188
17-519
Szumiło-Pilarska et al., 2016 *
Razorbill (A. torda)
piscivorous
2006-2011
37
144
53-294
37
347
142-713
Rutkowska et al., 2018 *
White-tailed eagle (H. albicilla)
piscivorous
1991-1995
12
202
25-630
10
894
210-2,500
Falandysz et al., 1997 *
Black-throated loon (G. arctica)
piscivorous
2006-2011
13
333
137-2,049
8
882
321-2,086
Rutkowska et al., 2018 *
Great black-backed gull (L. marinus)
piscivorous
2010-2012
9
378
91-1,424
9
827
179-2,003
Szumiło-Pilarska et al., 2016 *
Great cormorant (P. carbo)
piscivorous
2006
55
569
136-1,544
55
4,294
418-23,735
Misztal-Szkudlińska et al., 2018 *
Common merganser (M. merganser)
piscivorous
2005
16
731
91-1,097
Kalisińska et al., 2014 *
c.
Mammals
Striped dolphin (S. coeruleoalba)
piscivorous
1999-2003
2
4,378
3,703-5,054
Ciesielski et al., 2006 *
Harbour porpoise (P. phocoena)
piscivorous
1999-2003
14
5,919
414-58,649
Ciesielski et al., 2006 *
Ringed seal (P. hispida)
piscivorous
1999-2003
2
695
497-892
Ciesielski et al., 2006 *
Harbour seal (P. vitulina)
piscivorous
1999-2003
1
20,378
Ciesielski et al., 2006 *
Grey seal (H. grypus)
piscivorous
2001-2003
5
532
112-1,491
Bełdowska and Falkowska, 2016 *
2007-2010
8
102,700
6,300-247,800
Helsingen, 2011 *
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XVII
Table S9 Temporal changes of total Hg concentration in: a. vegetation (macroalgae, vascular plants), b. invertebrates
(zooplankton, soft tissues of benthic fauna) (ng g-1 dw), and c. muscles of fish (European sprat, Baltic herring, Baltic
cod) (ng g-1 ww) from the Polish part of the southern Baltic Sea (data are visualized in Fig. 6 and Fig. 7).
Year
Mean
Range
N
Reference
a.
Vegetation
Macroalgae
(mixed species)
1989
62
19-130
14
Falandysz, 1994
1998
41
18-95
58
Boszke et al., 2003 *
2006
11
2-37
48
Saniewska et al., 2010; Bełdowska et al., 2015
2007
13
1-31
74
Saniewska et al., 2010; Bełdowska et al., 2015
2008
10
2-26
33
Saniewska et al., 2010; Bełdowska et al., 2015
2009
9
2-19
10
Bełdowska et al., 2015
2010
11
1-33
43
Bełdowska et al., 2015
2011
9
2-28
34
Bełdowska et al., 2015
2013
12
3-22
106
Jędruch et al., 2019b
2014
15
6-46
53
Jędruch et al., 2019b
2016
14
1-37
45
Zalewska and Danowska, 2017
2018
13
5-21
15
Jędruch and Bełdowska, 2020
Vascular plants
(mixed species)
1989
48
15-110
6
Falandysz, 1994
1998
34
17-95
62
Boszke et al., 2003 *
2006
15
3-21
13
Bełdowska et al., 2015
2007
16
3-27
26
Bełdowska et al., 2015
2008
13
6-21
17
Bełdowska et al., 2015
2009
13
3-22
18
Bełdowska et al., 2015
2010
12
3-22
28
Bełdowska et al., 2015
2011
9
1-30
102
Bełdowska et al., 2013, 2015
2012
9
1-27
119
Bełdowska et al., 2013; Jędruch et al., 2019b
2013
9
1-31
58
Jędruch et al., 2019b
2017
6
1-11
18
Jędruch and Bełdowska, 2020
b.
Invertebrates
Zooplankton
(mixed species)
1998
240
48-380
11
Boszke et al., 2003 *
2008
121
83-140
40
Bełdowska et al., 2015
2012
100
16-204
68
Bełdowska and Mudrak-Cegiołka, 2017
2013
62
34-84
12
Bełdowska and Mudrak-Cegiołka, 2017
Zoobenthos
(mixed species)
1989
125
33-240
45
Falandysz, 1994
1998
179
35-370
423
Boszke et al., 2003 *
2010
37
16-79
40
Polak-Juszczak, 2012
2012
53
4-252
435
Jędruch et al., 2019b
2013
66
2-186
86
Jędruch et al., 2019b
2015
27
12-75
92
Wilman et al., 2019
2016
38
7-111
21
Jędruch et al., 2019a
2017
26
3-49
50
Jędruch and Bełdowska, 2020
c.
Fish
European sprat
(Sprattus sprattus)
1995
45
17-73
36
Polak-Juszczak, 2009
1996
29
20-38
42
Polak-Juszczak, 2009
1997
31
23-39
48
Polak-Juszczak, 2009
1998
18
14-22
34
Polak-Juszczak, 2009
1999
17
2-33
28
Polak-Juszczak, 2009
2000
12
1-23
35
Polak-Juszczak, 2009
2001
16
13-19
77
Polak-Juszczak, 2009
2002
19
14-24
38
Polak-Juszczak, 2009
2003
14
9-19
41
Polak-Juszczak, 2009
2009
16
12-20
20
Polak-Juszczak, 2012
2016
16
10-23
11
Polak-Juszczak, 2017
Baltic herring
(Clupea harengus)
1994
84
72-96
40
Polak-Juszczak, 2009
1995
65
11-119
63
Polak-Juszczak, 2009
1996
36
20-52
55
Polak-Juszczak, 2009
1997
32
23-41
62
Polak-Juszczak, 2009
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XVIII
Table S9 Continued.
* Values converted from wet weight.
To convert the Hg concentration in wet weight (ww) to dry weight (dw) the following conversion factors were applied: vegetation: 5.0 (da Silva et al., 2008),
zooplankton: 5.0 (McConvile et al., 2017); zoobenthos: 4.0 (Pleissner et al., 2012).
Year
Mean
Range
N
Reference
1998
29
14-44
61
Polak-Juszczak, 2009
1999
18
9-27
32
Polak-Juszczak, 2009
2000
19
1-37
44
Polak-Juszczak, 2009
2001
22
10-34
89
Polak-Juszczak, 2009
2002
31
3-58
51
Polak-Juszczak, 2009
2003
22
9-35
54
Polak-Juszczak, 2009
2005
26
GIOŚ, 2017
2006
14
GIOŚ, 2017
2007
27
4-78
>63
Falkowska et al., 2010; Kwaśniak et al., 2012; GIOŚ, 2017
2008
19
>18
Falkowska et al., 2010; Kwaśniak et al., 2012; GIOŚ, 2017
2009
26
GIOŚ, 2017
2010
22
GIOŚ, 2017
2011
27
GIOŚ, 2017
2012
31
GIOŚ, 2017
2013
26
GIOŚ, 2017
2014
30
GIOŚ, 2017
2015
32
GIOŚ, 2017
2016
23
8-45
>16
Polak-Juszczak, 2017, GIOŚ, 2017
2017
32
24-62
>12
GIOŚ, 2017; Jędruch et al., 2018
2018
34
7-127
83
Saniewska, unpubl.
Baltic cod
(Gadus morhua)
1997
46
20-72
Polak-Juszczak, 2009
1998
48
19-77
Polak-Juszczak, 2009
1999
29
11-47
Polak-Juszczak, 2009
2000
27
11-43
Polak-Juszczak, 2009
2001
31
15-47
Polak-Juszczak, 2009
2002
43
12-74
Polak-Juszczak, 2009
2003
31
9-53
Polak-Juszczak, 2009
2006
22
9-31
Kwaśniak and Falkowska, 2012
2007
30
12-65
Kwaśniak and Falkowska, 2012
2008
36
24-111
Kwaśniak and Falkowska, 2012
2009
62
17-123
Kwaśniak and Falkowska, 2012; Kwaśniak et al., 2012
2013
52
LOD-140
Polak-Juszczak, 2015
2014
71
67-75
Podolska et al., 2016
2016
92
25-132
Polak-Juszczak, 2017
Electronic Supplementary Material Jędruch et al., Mercury in the Polish part of the Baltic Sea: …
XIX
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