Spatial distribution of mercury in seawater, sediment, and seafood from the Hardangerfjord ecosystem, Norway

Abstract

Hardangerfjord is one of the longest fjords in the world and has historical mercury (Hg) contamination from a zinc plant in its inner sector. In order to investigate the extent of Hg transferred to abiotic and biotic ecosystem compartments, Hg and monomethylmercury (MeHg) concentrations were measured in seawater, sediment, and seafood commonly
consumed by humans. Although total mercury in seawater has been described previously, this investigation reports novel MeHg data for seawater from Norwegian fjords. Total Hg and MeHg concentrations in seawater, sediment, and biota increased towards the point source of pollution (PSP) and multiple lines of evidence show a clear PSP effect in seawater and sediment concentrations. In fish, however, similar high concentrations were found in the inner part of another branch adjacent to the PSP. We postulate that, in addition to PSP, atmospheric Hg, terrestrial run-off and hydroelectric power stations are also important sources of Hg in this fjord ecosystem. Hg contamination gradually increased towards the inner part of the fjord for most fish species and crustaceans. Since the PSP and the atmospheric Hg pools were greater towards the inner part of the fjord, it is not entirely possible to discriminate the full extent of the PSP and the atmospheric Hg contribution to the fjord food web. The European Union (EU) Hg maximum level for consumption was exceeded in demersal fish species including tusk (Brosme brosme), blue ling (Molva dypterygia) and common ling (Molva molva) from the inner fjord (1.08 to
1.89 mg kg-1 ww) and from the outer fjord (0.49 to 1.07 mg kg-1 ww). Crustaceans were less contaminated and only European lobster (Homarus gammarus) from inner fjord exceeded the EU limit (0.62 mg kg-1 ww). Selenium (Se) concentrations were also measured in seafood species and Se-Hg co-exposure dynamics are also discussed.

Full text

Spatial distribution of mercury in seawater, sediment, and seafood from
the Hardangerfjord ecosystem, Norway
Atabak M. Azad
a,b,
⁎, Sylvia Frantzen
a
, Michael S. Bank
a,c,
⁎, Ingrid A. Johnsen
a
,EmmanuelTessier
d
,
David Amouroux
d
, Lise Madsen
a,e
, Amund Maage
a,b
a
Institute of Marine Research, Bergen, Norway
b
Faculty of Mathematics and Natural Sciences, University of Bergen, Bergen, Norway
c
Department of Environmental Conservation, University of Massachusetts, Amherst, USA
d
CNRS/ Univ Pau & Pays Adour/ E2S UPPA, Institut des Sciences Analytiques et de Physicochimie pour l'Environnement et les Matériaux –MIRA, UMR5254, 64000 Pau, France
e
Department of Biology, University of Copenhagen, Denmark
HIGHLIGHTS
•Hardangerfjord is a mercury (Hg) con-
taminated ecosystem with a legacy
point source.
•Hg species were analyzed in seawater,
sediment and seafood.
•Hg concentrations in seawater, sedi-
ment and biota increased towards the
inner fjord.
•Demersal fish from the entire fjord
exceeded acceptable Hg limits for
human consumption.
GRAPHICAL ABSTRACT
abstractarticle info
Article history:
Received 12 February 2019
Received in revised form 22 February 2019
Accepted 22 February 2019
Available online 23 February 2019
Editor: Mae Sexauer Gustin
Hardangerfjord is one of the longest fjords in the world and has historical mercury (Hg) contamination from a
zinc plant in its inner sector. In order to investigate the extent of Hg transferred to abiotic and biotic ecosystem
compartments, Hg and monomethylmercury (MeHg) concentrations were measured in seawater, sediment,
and seafood commonly consumed by humans. Although total mercury in seawater has been described previ-
ously, this investigation reports novel MeHg data for seawater from Norwegian fjords. Total Hg and MeHg con-
centrations in seawater, sediment, and biota increased towards the point source of pollution (PSP) and
multiplelines of evidence showa clear PSP effect in seawater and sediment concentrations. In fish, however, sim-
ilar high concentrations were found in the inner part of another branch adjacent to the PSP. We postulate that, in
addition to PSP, atmospheric Hg, terrestrial run-off and hydroelectric power stations are also important sources
of Hg in this fjord ecosystem. Hg contamination gradually increased towards the inner part of the fjordfor most
fish species and crustaceans. Since the PSP and the atmospheric Hg pools were greater towardsthe inner part of
the fjord, it is not entirely possible to discriminate the full extent of the PSP and the atmospheric Hg contribution
to the fjord food web. The European Union (EU) Hg maximum level for consumption was exceeded in demersal
fish species including tusk (Brosme brosme), blue ling (Molva dypterygia) and common ling (Molva molva)from
Keywords:
Mercury
Bioaccumulation
Fjords
Seafood safety
Speciation
Norway
Science of the Total Environment 667 (2019) 622–637
⁎Corresponding authors at: Institute of Marine Research, P.O. Box 1870, Nordnes, Bergen, Norway 5005.
E-mail addresses: ata@hi.no (A.M. Azad), Michael.Bank@hi.no (M.S. Bank).
https://doi.org/10.1016/j.scitotenv.2019.02.352
0048-9697/© 2019 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Contents lists available at ScienceDirect
Science of the Total Environment
journal homepage: www.elsevier.com/locate/scitotenv

the inner fjord(1.08 to 1.89 mg kg
−1
ww) and fromthe outer fjord (0.49to 1.07 mg kg
−1
ww). Crustaceans were
less contaminated and only European lobster (Homarus gammarus) from inner fjord exceeded the EU limit
(0.62 mg kg
−1
ww). Selenium (Se) concentrations were also measured in seafood species and Se-Hg co-
exposure dynamics are also discussed.
© 2019 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://
creativecommons.org/licenses/by/4.0/).
1. Introduction
Mercury (Hg) is a widespread global pollutant with significant impli-
cations for environmental and public health. Anthropogenic activities,
such as emissions from coal-fired plants and mining have significantly in-
creased the concentrations of Hg and monomethylmercury (MeHg) in
the environment, including marine ecosystems and their inhabitants
(Lamborg et al., 2014). Increased MeHg concentrations in some Arctic
marine biota have been reported in comparison to pre-industrial times
(Braune et al., 2005), however, the ocean is not uniformly polluted
(Lamborg et al., 2014). For example, Vo et al. (2011) reported an increase
in MeHg concentrations during a 120-year period in black-footed alba-
tross museum specimens sampled from the Pacific Ocean, but recent
studies have reported small-scale temporal declines in MeHg concentra-
tions in coastal and pelagic fish species from the Atlantic Ocean (Cross
et al., 2015;Lee et al., 2016). Although air-sea exchange, terrestrial inputs
and atmospheric processes are recognized as important drivers of the Hg
cycle, numerous important processes governing marine Hg biogeochem-
ical cycling and bioaccumulation have a high degree of uncertainty and
remain poorly understood (Strode et al., 2007;Black et al., 2012).
Inorganic Hg may exist in different forms such as elemental Hg, Hg
2+
inorganic complexes, Hg
2+
organic complexes, and Hg
2+
with different
degrees of bioavailability. However, inorganic Hg can be methylated by
anaerobic, mainly sulfate reducing, bacteria in marine sediments
(Compeau and Bartha, 1987) and also in the open water column
(Topping and Davies, 1981). MeHg is highly neurotoxic and the most bio-
available form of mercury (Hong et al., 2012). Methylation dynamics and
trophic transfer are critical processes involved in MeHg bioaccumulation
in coastal and open ocean food webs (Bank et al., 2007;Senn et al., 2010).
MeHg easily biomagnifies in the marine food web, and in top predator
marine organisms 70 to 100% of the total Hg may be present in the
MeHg form (Bloom, 1992;Magalhães et al., 2007;Hong et al., 2012).
Fish may bioconcentrate MeHg as much as 10
6
-foldcomparedtolowsea-
water concentrations (Watras and Bloom, 1992).
Atmospheric deposition is considered an important source of Hg to
the marineenvironment (Driscoll et al., 2013). Hg precipitated in terres-
trial catchments and transported via run-off can be substantial for
aquatic ecosystems including streams, rivers, ponds, lakes, and coastal
zones. Although biotic methylation of inorganic Hg in the sediment
and in the water column is the primary process governing MeHg, abiotic
methylation may also occur, but at a far lower rate (Weber, 1993;Celo
et al., 2006). Hg methylation in marine sediments has been shown to
be enhancedby anaerobic conditions, increased temperature, decreased
pH, and intermediate concentrations of organic carbon (Ullrich et al.,
2001). Additionally, organic carbon composition and overall quality
(i.e., humic substances content), sulfur availability, and fraction of Hg
available for methylation have been shown to have important roles in
controlling Hg methylation (Avramescu et al., 2011;Bełdowska et al.,
2014;Schartup et al., 2014).
Seafood is the main contributor to MeHg exposure in humans
(Batista et al., 1996;Al-Majed and Preston, 2000;Olivero et al., 2002)
and the EU maximum level (EUML) of Hg (0.5 mg kg
−1
ww) applies
to most fish and fishery products for legal trade (EC, 2006). The interac-
tion between MeHg and seafood nutrients, particularly selenium (Se),
may influence the bioavailability and toxicity of MeHg (Ralston et al.,
2008), and it is advantageous to measure and evaluate these elements
simultaneously, across fish species, to make accurate decisions
pertaining to food safety and human exposure.
The Hardangerfjord ecosystem is one of the longest fjords in western
Norway (Fig. 1). The fjord is polluted by industry and other anthropo-
genic Hg pollution sources, including a zinc plant, hydroelectric power
stations, and local mining and aquaculture facilities (deBruyn et al.,
2006). The zinc plant has existed for ~100 years and produces zinc and
aluminum fluoride at a site located 4 km north of Odda in the inner sector
of Sørfjord, an arm of the Hardangerfjord (Fig. 1). Zinc ores typically con-
tain Hg and zinc plants may emit high amounts of Hg to the atmosphere.
For instance, it is estimated that approximately 107.7 tons of Hg was
emitted to the atmosphere from zinc smelting activities in 2006 in
China (Yin et al., 2012). Industrial wastes associated with zinc production
with high concentrations of toxic metals were released to Sørfjord until
1986 (Julshamn and Grahl-Nielsen, 1996) even though a mercury re-
moval system was introduced early in the 1970's. In the 1970's, it was es-
timated that an average of 1–3kgofsolidphaseHgperdaywasreleased
into the local environment (Skei et al., 1972;Melhuus et al., 1978), most
likely as metacinnabar (HgS). In 1986 the company initiated a waste
treatment and processing program storing the main tailings and effluents
from the zinc plant on land in mountain tunnels. However, the sediments
in the inner part of the Sørfjord were already highly polluted with toxic
trace metals including Hg, and today the Hardangerfjord ecosystem is
still widely considered to be one of the most trace metal polluted fjords
in the world (Skei et al., 1972;Everaert et al., 2017).
Early investigations on toxic trace metal contaminationintheareafo-
cused on zinc (Zn), arsenic (As), cadmium (Cd), lead (Pb), and Hg in ma-
rine organisms such as brown algae (Ascophyllum nodosum), blue mussel
(Mytilus edulis), flounder (Platichthys flesus)andsaithe(Gadus virens). Hg
received relatively little attention because Hg concentrations were not
very high in the investigated species which were from low positions in
the food web (Haug et al., 1974;Stenner and Nickless, 1974;Melhuus
et al., 1978;Julshamn and Grahl-Nielsen, 1996). Julshamn et al. (2001) re-
ported a significant decrease in toxic trace metals in Sørfjord following the
termination of jarosite discharge in 1986, however, the degree of Hg con-
tamination in demersal fish species was unknown. More recent investiga-
tions have reported Hg concentrations in fillets of tusk (Brosme brosme),
inhabiting the demersal habitats of Sørfjord ~3 times greater than the
EUML (Ruus and Green, 2007), and additional data (Kvangarsnes et al.,
2012) led the Norwegian Food Safety Authority (NFSA) to issue extended
consumption advisories for deep-water fish caught in the entire
Hardangerfjord ecosystem, as well as for shellfish from the Sørfjord sector.
In this investigation, we focused on evaluating the spatial extent of
Hg and MeHgconcentrationsin several Hardangerfjord ecosystem com-
partments including marine organisms consumed by humans, seawa-
ter, and sediment. We hypothesized that the zinc plant and
surrounding highly polluted sediments, as a point source of pollution
(PSP), would be an important driver of Hg contamination and spatial
distribution in seawater, sediment, and biota. This would result in
higher Hg and MeHg levels in the different ecosystem compartments
sampled from the inner sector of the fjord compared to the outer sec-
tors. Additionally, we compared our measurements in seafood to the
EUML and discuss Se-Hg co-exposure dynamics.
2. Materials and methods
2.1. Study area
The Hardangerfjord ecosystemis the second longest fjord system in
Norway, located in the western coastal region (59.4–60.6°N, 4.5–7.3°E;
623A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

Fig. 1). The water depth ranges from 120 to 800 m and thefjord has sev-
eral basins separated by shallower sills. The fjord ecosystem is con-
nected to the ocean through one main fjord mouth and three
narrower channels to the north. At the inner part, the fjord branches
into Sørfjord to the south and Eidfjord to the northeast (Fig. 1). The
Sørfjord is ~40 km long and up to 1 to 2 km wide and is substantially
shallower than the main fjord, with depths of ~100 to 350 m and only
~50 m at the head of the fjord (Fig. 1). The Opo River is the main source
of freshwater for Sørfjord. The Opo flows north at the head of the fjord
within the Odda municipality (Fig. 1), and has a catchment area of
483 km
2
(Pettersson, 2008). River Tysso, with a catchment area of
390 km
2
, is another large river which flows into the southern part of
Sørfjord close to the PSP at Tyssedal that also houses a power station
(Fig. 1). Eidfjord is the northwards fork extension of Hardangerfjord
and is ~29 km long with depths reaching ~400–600 m. The Eio and
Sima Rivers are both main sources of freshwater to the Eidfjord sector
Fig. 1. Location of different sampling sitesin Hardangerfjord.(A) Fish and crustacean species sampledin 2011 and (B) sediment and seawater sampled in 2015and 2018. The letters after
site numbers in map A represent the names of the fjords (S: Sørfjord; E: Eidfjord; OH: Outer Hardangerfjord). Details of biotic samples collected from each site are described in Table 1.
624 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

of the Hardangerfjord ecosystem (Fig. 1), with catchment areas of
1173 km
2
and 146 km
2
respectively (Pettersson, 2008). Additionally,
there is another hydroelectric power station located on the Sima River
(Fig. 1). Apart from these four major rivers, the Glacier Folgefonna,
consisting of three sub-glaciers with a total area of 200 km
2
,isanimpor-
tant source of freshwater along with several other low-order and head-
water streams within the catchment area of the fjord.
2.2. Sediment, seawater and seafood sampling and preparation
Fish were caught during cruises organized by the Institute of Marine
Research (IMR) as part of a larger Hardangerfjord study. The demersal
deep-water fishes blue ling (Molva dypterygia) (4 sites) and tusk
(Brosme brosme) (8 sites) were caught using long line fishing. Common
ling (Molva molva) (7 sites) and Atlantic wolffish (Anarhichas lupus)(2
sites) were sampled using a trammel net and European sprat (Sprattus
sprattus) (5 sites) were sampled using purse seine nets. Crustacean spe-
cies including brown crab (Cancer pagurus) (2 sites), European lobster
(Homarus gammarus) (3 sites) and Norway lobster (Nephrops
norvegicus) (1 site) were caught using lobster trap and trammel nets.
All seafood sampling was conducted during 2011 (Table 1 and
Fig. 1A). Due to a low number of samples, data for wolffish and
Norway lobster were not included in the spatial distribution analyses.
All fish and crustacean specimens were shipped whole and frozen to
the Institute of Marine Research, Bergen, Norway. Individual weights
(g)andlengths(cm)offish and crustaceans were measured and regis-
tered in the Laboratory Information Management System (LIMS). For all
fish species except sprat, skin and bone free fish fillets were dissected.
For tusk, we also analyzed liver tissue. For sprat, 25 whole fish were
composited and homogenized. For European and Norway lobster, the
tail meat was dissected, while for the brown crab, both claw meat
(both claws) and brown meat (mixture of hepatopancreas, gonads
and internal white meat), were sampled and analyzed. All biota samples
were homogenized using a food processor, and all samples, except liver
of tusk and brown meat of crab were subsequently lyophilized. After ly-
ophilization to a constant mass, the water content (% moisture) of each
sample was calculated and recorded prior to Hg and Se analyses.
Sediment samples (7 sites) were collected from the top 15 ± 2 cm of
the bottom sediment using a van Veen grab or by diving. The sediment
sampling was conducted during April –July 2015. The samples were
frozen (−30 °C) before being sent to the laboratory for analyses.
Seawater samples (9 sites) were collected during May 28–31, 2018
on the RV Hans Brattstrøm (Fig. 1B). Seawater was collected using
acid-washed Niskin-Type oceanographic general purpose, plastic
water samplers (2.5 L model; Hydro-Bios Inc.) at depths of 15, 50, and
300 m. Trace metal clean sampling techniques (Bravo et al., 2018)
were employed using acid-washed 120 mL and 250 mL Teflon bottles
(Nalgene FEP). Teflon bottles and silicon tubing were acid washed in a
Milestone acid-washer using 37% ultra-pure, trace metal grade HNO
3
and were rinsed five times using Milli-Q deionized water. The Niskin
type plastic water samples were acid washed using two consecutive
overnight, acid baths (1 HNO
3
and 1 HCL at 10% volume:volume pre-
pared with milli-Q water). Teflon bottles and Niskin type bottles were
dried in an EPA clean 100 room under a laminar flow hood. Bottles
were stored in double plastic bags before and immediately after sam-
pling seawater. Seawater was collected using a standard oceanographic
rosette (Hydro-Bios, Inc.) and samples were transferred to individually
labeled Teflon bottles using acid-washed silicon tubing that was rinsed
between samples with deionized water and stored in a sterile and clean
plastic bag. Seawater was then acidified using 0.5% ultrapure HCl (vol-
ume:volume) and placed in a dark refrigerator (4 °C) prior to laboratory
analyses.
2.3. Total mercury and selenium measurements in biota
The concentrations of Hg and Se were determined using inductively
coupled plasma-mass spectrometry (ICP-MS) after microwave diges-
tion. First, weighed samples were digested using concentrated (65%)
HNO
3
and 30% H
2
O
2
in a microwave oven (Milestone Microwave diges-
tion system: MLS-1200 MEGA Microwave Digestion Rotor - MDR 300/
10). Hg and Se concentrations were determined using quantitative
ICP-MS (Agilent 7500 with collision cell and ICP-ChemStation soft-
ware). A standard curve was used to determine the concentration of
Hg and Se. Germanium (Ge), thulium (Tm) and rhodium (Rh) were
used either individually or in combination as internal standards, and
gold (Au) was added to stabilize the Hg signals. The method is a CEN
standard and Norway accredited laboratory method (ISO 17025) for
these two elements (NMKL, 2007;CEN, 2009) and is described in detail
elsewhere (Julshamn et al., 2007). Accuracy and precision of these
methods have been tested by analyzing certified reference materials
and the recoveries of both Hg and Se ranged from 80% to 120%. Certified
reference material (CRM) 1566 (oyster tissue) from the National Insti-
tute of Standards and Technology(NIST, Gaithersburg, USA) and lobster
hepatopancreas (TORT-2, TORT-3) from the National Research Council
of Canada (Ottawa, Canada)were used for measurementquality control
by including them in each sample run. The limitsof quantification (LOQ)
of this method were 0.005 and 0.01 mg kg
−1
dry weight (dw) for Hg
and Se, respectively.
2.4. Mercury speciation in sediment samples
Methylmercury concentrations in sediment samples were measured
using EPA method 1630 (USEPA, 1998). Samples were prepared by
leaching potassium bromide and copper sulfate solution to release the
organic Hg species from inorganic complexes. MeHg was subsequently
extracted by dichloromethane. An aliquot of the dichloromethane was
then back-extracted into ultrapure deionized water by purging with
argon. Samples were treated with sodium tetraethyl borate to form
MeHg. Inorganic Hg was simultaneously converted to diethyl Hg. The
ethylated Hg species are volatile and are stripped of the solution by
purging with N
2
and then adsorbed onto Tenax traps. Hg-species were
then thermally desorbed from the Tenax traps in a stream of helium
Table 1
Number of fish and crustacean samples collected from different sites in Hardangerfjord in 2011 (S: Sørfjord; E: Eidfjord; OH: Outer Hardangerfjord). Locations are shown on the map
(Fig. 1).
Species Scientific name N Sampling stations (N)
1S 2S 3S 4E 5E 6OH 7OH 8OH 9OH 10OH 11OH
Blue ling Molva dypterygia 41 5 2 20 6 7
Common ling Molva molva 30 1 1 6 3 4 13 2
European Sprat
a
Sprattus sprattus 511 12
Tusk Brosme brosme 138 2 8 7 24 30 13 32 22
Wolffish Anarhichas lupus 431
Brown crab Cancer pagurus 20 10 10
European lobster Homarus gammarus 26 5 11 10
Norway lobster Nephrops norvegicus 10 10
a
Each sample is a composite of 25 whole specimens.
625A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

and separated by means of isothermal gas chromatography. Finally, the
methyl/ethylated Hg species are decomposed to elemental Hg and de-
tected using Cold Vapor-Atomic Fluorescence Spectroscopy (CV-AFS)
by heating a pyrolysis column to 700–800 °C. The LOQ was 0.05
μgkg
−1
dw. Total Hg in sediment was measured using laboratory
accredited methods (EN ISO12846) and Cold Vapor-Atomic Absorption
Spectrometry (CV-AAS) technique (ISO12846, 2012). The LOQ was
0.001 mg kg
−1
dw and the measurement uncertainty was 20%. The sed-
iment analyses were conducted by Eurofins Environment Testing
Norway AS, Moss, Norway.
2.5. Mercury speciation in seawater
Inorganic Hg and MeHg concentrations in unfiltered seawater sam-
ples were simultaneously measured using the species-specific isotope
dilution, and a GC-ICP-MS method developed for Hg speciation at
ultra-trace levels in seawater (Monperrus et al., 2005;Cavalheiro
et al., 2016;Bravo et al., 2018). The analyses were operated by a capil-
lary gas chromatograph (Trace GC Ultra, Thermo Fisher, equipped
with a TriPlus RSH auto-sampler) hyphenated to an inductively coupled
plasma mass spectrometer (ICP-MS Thermo X Series 2). Briefly, an ali-
quot of 100 mL of unfiltered water sample was accurately weighed
and spiked with known amounts and of isotopically enriched standards
solutions Me(201)Hg and (199)inorganic Hg (ISC Science, Spain).
Spiked samples were left overnight for equilibration in a laminar flow
hood. The pH of the solution was then adjusted to 3.9 by adding 5 mL
of sodium acetate-acetic acid 0.1 M buffer solution and about 1 mL of ul-
trapure ammonium hydroxide solution. At last, 250 μL of isooctane
(HPLC grade) and 80 μL of sodium tetra-propyl borate solution (5% w/
v, Merseburger Spezial Chemikalien, Germany) were added to achieve
the derivatization of the Hg species and its subsequent extraction into
the GC solvent. The vials were capped and shaken for 20 min at
400 rpm (orbital shaker); then the isooctane was recovered and ana-
lyzed in triplicate by GC-ICP-MS.
All materials were cleaned prior to use according to ultra-trace stan-
dard operating protocols (Bravo et al., 2018). In absence of any Certified
Reference Material available for organomercury species, quality assur-
ance and quality control (QA/QC) was based on reagent blank analyses,
replicated assays and an extensive QA/QC procedure described else-
where (Cavalheiro et al., 2016). Additionally, repeated participations
in international inter-laboratory comparison exercises (GEOTRACES in-
tercalibration cruises for Hg species in seawater) complement the QA/
QC effort.
Inorganic Hg concentrations measured in the blanks averaged 0.016
± 0.003 ng L
−1
, whereas no MeHg was observed in the blanks. The
MeHg blank equivalent concentration for the GC-ICP-MS instrument
was estimated at 0.002 ± 0.001 ng L
−1
. The detection limits of this
method were 0.03 ng L
−1
for inorganic Hg and 0.008 ng L
−1
for MeHg,
respectively. The measurement error (calculated by analyzing each
sample three times) was b2.9% and 4.9% for inorganic Hg and MeHg
concentrations, respectively. All seawater samples were analyzed at
the IPREM laboratory (CNRS/University of Pau, France) within 28 days
after sampling.
2.6. Salinity measurements and modeling
The salinity was observed in situ using a portable instrument
(SAIV A/S SD 208) measuring the conductivity, temperature, and
depth (CTD). The instrument was used in STD mode, and calculations
of salinity from the conductivity were done automatically using the
instrument's software. The accuracy of the salinity is ±0.003 with a
range from 0 to 50. The instrument also measured dissolved oxygen
(range: 0–20 mg L
−1
accuracy: ±0.2 mg L
−1
) supplied by SAIV A/S.
The instrument was sampled with a time interval of 1 s and lowered
with a speed of 0.2 ms
−1
. Data was downloaded from the instrument
for every 0.1 m in the upper 10 m and for every meter under 10 m
depth. In addition to measuring the salinity in situ water was sam-
pled using a multi water sampler slim line 6 with mounted plastic
Niskin bottles supplied by Hydro-Bios. Water samples for salinity
analyses were taken at a depth of 300 m at every site. The water sam-
ples were bottled and analyzed at the in-house salinity lab using a
Guildline 8410A portasal (range: 0.004–76, Accuracy: ±0.003). By
comparing the in-situ measurements to the salinity data from the
seawater samples it became evident that the SAIV SD 208 instrument
showed a deviation in its calibration (−0.12) and therefore we used
a correction value of +0.12.
The salinity distribution of the fjord was modeled using the Re-
gional Ocean Model System (ROMS) solving the hydrodynamic
equations (Haidvogel et al., 2000;Shchepetkin and McWilliams,
2005). The model was set up with a horizontal resolution of 160 m
× 160 m, with 35 terrain following coordinates in the vertical. 170
rivers were included with daily run-off from the Norwegian Water
Resources and Energy Directorate (NVE ) and atmospheric conditions
were provided by 2.5 km resolved AROME model provided by the
Norwegian Meteorological Institute (http://thredds.met.no). The
modelwasrunwithaninternaltime-stepof6s,writingenviron-
mental data as temperature, salinity and currents every hour. Fur-
ther details of the model setup are described in Albretsen (2011).
The model simulation was started 1st of April 2018, where the first
month is considered spin-up time. The salinity distribution at the
time of the cruise is illustrated as the mean salinity from May 28th
to May 31st in 2018, for the sea surface.
2.7. Statistical analyses
Data were log transformed to meet the assumption of normal dis-
tribution and homogeneity of variances prior to statistical analyses.
Analysis of covariance (ANCOVA) was used for comparison of Hg
concentrations across sampling sites for seafood species with length
as a covariate to remove the possible effect of length across sites.
One-way analysis of variance (ANOVA) was used for crustacean spe-
cies since length measurements and Hg concentrations were not cor-
related. In European lobster, the Hg concentrations increased with
increasing weights, however since weight was not significantly dif-
ferent between sites, ANOVA was used for comparison across sam-
pling sites. Independent Student's t-tests were used to compare
length and Hg concentration between the inner and outer sections
of Hardangerfjord. For post-hoc comparisons, unequal sample
Tukey-HSD tests were used to evaluate the effects of unequal sam-
pling efforts and unbalanced design. Only sites with two or more in-
dividuals were considered for spatial comparisons. Distance from
PSP was calculated as distance from the industrial unit close to
Odda and distance from the open ocean was calculated from the
mouth of the Hardangerfjord at Kvinnsvika (Fig. 1). Statistical signif-
icance was accepted at Pb0.05 (Zar, 2010). All statistical analyses
were performed using STATISTICA 13 (Statsoft Inc., Tulsa, USA) or
GraphPad Prism 7.02 (GraphPad software Inc., San Diego, CA, USA).
2.8. Selenium health benefit value
Selenium health benefitvalue(HBV
Se
) has been suggested as an
evaluation index showing the Se amount provided in fish after seques-
tration of Hg and was calculated using the following formula (Ralston
et al., 2016):
HBVSe ¼Se−Hg
Se Se þHg
ðÞ
Se = Selenium content in molar concentration.
Hg = Mercury content in molar concentration.
626 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

2.9. Bioconcentration Factors and Biota-Sediment Accumulation Factors
Bioconcentration Factors (BCF) and Biota-Sediment Accumulation
Factors (BSAF) for tusk were calculated for total Hg and MeHg using
the following formulas:
BCF ¼Log Hg concentration in fillet
Hg concentration in water

BSAF ¼Log Hg concentration in fillet
Hg concentration in sediment

BCF was calculated using average seawater concentration from 15,
50 and 300 m depths closest to the tusk sampling location and 100%
of Hg in tusk fillet was assumed to be MeHg.
3. Results and discussion
3.1. Hg and Se concentrations in seafood
Tusk and blue ling fillet samples collected from the inner sector of
Hardangerfjord had the highest mean Hg concentrations (1.87 and
1.44 mg kg
−1
ww, respectively) and all individual fish were above the
EUML of 0.5 mg kg
−1
ww (Table 2). In comparison tusk and blue ling
samples from outer Hardangerfjord had lower Hg concentrations, but
the mean levels were still higher than EUML (mean = 0.84 and
1.07 mg kg
−1
ww, respectively). Wolffish (0.14 mg kg
−1
ww) and
sprat (0.01 in outer and 0.03 mg kg
−1
ww in the inner Hardangerfjord)
had the lowest Hg concentrations. In a previous study, Azad et al. (2019)
showed that Hg concentrations in blue ling and tusk from the Northeast
Atlantic Ocean were similarly high, whereas common ling had lower
concentrations and wolffish had the lowest of all demersal fish species
analyzed in this study. The high concentrations of Hg in tusk, blue ling,
and common ling were likely influenced by their high trophic position,
and preference for deep-water, demersal habitats (Bergstad, 1991;
Husebø et al., 2002;McMeans et al., 2010). Atlantic wolffish feed on
molluscs, echinoderms, and other low trophic level prey species (Falk-
Petersen et al., 2010), and this may explain their lower Hg
concentrations.
Crustaceans had lower concentrations of Hgthan demersal fish spe-
cies (Table 2) likely as a result of their considerably lower trophic posi-
tion and similar observations have been reported from Spain (Olmedo
et al., 2013). European lobster tail meat sampled from inner
Hardangerfjord had the highest mean Hg concentration of all sampled
crustaceans (0.62 mg kg
−1
ww). These values are higher than those
previously reported in commercially caught European lobster from
Scotland (Barrento et al., 2008;Noël et al., 2011). European lobster
from outer Hardangerfjord had a mean Hg concentration of
Table 2
Mean, firstand third quartiles,standard deviationand standard error of Hg and Se levels(mg kg
−1
ww) in muscletissue and length (cm)of demersal fish and crustacean speciesfrom the
Hardangerfjord ecosystem, 2011. HBV
Se
are calculated from mean values.
Species Scientific
name
Area N Hg (mg kg
−1
ww) Se (mg kg
−1
ww) Length (cm) Percent with
Hg ≥0.5
(mg kg
−1
ww)
HBVSe
Mean Q25 Q75 SD SE Mean Q25 Q75 SD SE Mean Q25 Q75 SD SE
Blue ling Molva
dypterygia
Out.
Hard.
33 1.07 0.64 1.19 0.83 0.15 0.43 0.38 0.49 0.08 0.01 90.73 80.00 97.00 13.65 2.38 93.9 0.4
Inn.
Hard.
8 1.44 1.07 1.85 0.66 0.23 0.50 0.49 0.53 0.04 0.01 92.33 86.00 100.00 8.41 3.43 100 −1.7
Common
ling
Molva molva Out.
Hard.
28 0.49 0.19 0.59 0.44 0.08 0.47 0.42 0.51 0.08 0.01 73.93 61.50 83.50 18.63 3.52 35.7 5.0
Inn.
Hard.
2 1.08 0.40 1.76 0.97 0.68 0.65 0.43 0.87 0.32 0.22 78.00 72.00 84.00 8.49 6.00 50 4.7
Tusk Brosme brosme Out.
Hard.
97 0.84 0.42 1.11 0.52 0.05 0.59 0.51 0.64 0.11 0.01 64.26 55.00 75.00 13.68 1.39 64.9 5.0
Inn.
Hard.
41 1.89 1.26 2.19 0.89 0.14 0.72 0.57 0.83 0.23 0.04 62.95 55.00 69.00 9.85 1.54 100 −0.6
Sprat
a
Sprattus
sprattus
Out.
Hard.
3
a
0.01 0.43 7.27
Inn.
Hard.
2
a
0.03 0.42 8.20
Wolffish Anarhichas
Lupus
Out.
Hard.
4 0.14 0.11 0.16 0.04 0.02 0.47 0.27 0.66 0.36 0.18 79.00 74.00 84.00 6.32 3.16 0 5.8
All fishes Out.
Hard.
162 0.63 0.49 76.98 48.6 4.1
Inn.
Hard.
51 1.47 0.62 77.76 83.3 0.8
Brown crab Cancer
pagurus
Out.
Hard.
10 0.12 0.05 0.20 0.08 0.02 1.47 0.92 1.83 0.80 0.25 14.85 14.40 15.40 1.49 0.47 0 18.6
Inn.
Hard.
10 0.22 0.13 0.30 0.14 0.05 0.76 0.60 0.74 0.33 0.10 15.37 13.40 17.60 2.45 0.77 10 9.5
European
lobster
Homarus
gammarus
Out.
Hard.
21 0.19 0.16 0.23 0.07 0.01 0.61 0.49 0.65 0.19 0.04 26.55 25.50 27.00 1.48 0.32 0 7.6
Inn.
Hard.
5 0.62 0.63 0.70 0.13 0.06 0.55 0.49 0.65 0.14 0.06 27.80 27.00 30.00 2.77 1.24 80 5.5
Norway
lobster
Nephrops
Norvegicus
Out.
Hard.
10 0.20 0.19 0.22 0.03 0.01 0.99 0.87 1.05 0.21 0.07 18.27 16.90 19.30 1.63 0.51 0 12.4
All Crustaceans Out.
Hard.
41 0.17 1.02 19.89 0 12.9
Inn.
Hard.
15 0.42 0.65 21.59 45 7.5
All species Out.
Hard.
203 0.44 0.72 52.51 27.8 7.8
Inn.
Hard.
67 1.05 0.64 55.29 68 3.5
a
Each sample is a composite of 25 whole specimens and thus, percent exceeding EUML and HBV
Se
are not calculated.
627A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

0.19 mg kg
−1
ww that is consistent with their reported range. Claw
meat of brown crab had lower Hg concentrations than European lobster
with mean values of 0.22 and 0.12 mg kg
−1
ww in samples from inner
and outer Hardangerfjord, respectively(Table 2). The Hg concentrations
in brown crab samples from outer Hardangerfjord were similar to the
mean reported for this species from the Norwegian coast
(0.1 mg kg
−1
ww) (IMR, 2018), whereas Hg concentrations in samples
from the inner fjord were ~2-fold higher. Norway lobster was only sam-
pled fromouter Hardangerfjord. The mean Hg concentrationin tail meat
of Norway lobster was 0.20 mg kg
−1
ww, similar to European lobster
and within the range reported for Norway lobsters caught in other re-
gions of Norway (IMR, 20 18). The Hg levels in Norway lobster measured
in this investigation were lower than the reported levels in samples
from the Mediterranean (Cresson et al., 2014). All crustaceans in this
study are benthic carnivores (Cristo and Cartes, 1998;Meeren, 2007;
IMR, 2008), and the observed variation in Hg concentrations is likely
driven by several factors including body size, toxicokinetics, growth di-
lution, prey type, ecosystem methylation potential, and species migra-
tion patterns.
Overall, Se concentrations in all sampled taxa were less variable than
Hg concentrations. Se concentrations in fish species from outer
Hardangerfjord were ~50% lower compared to crustaceans analyzed
from the same area (mean= 0.49 vs 1.02 mg kg
−1
ww; Table 2). How-
ever, Se concentrations in fish and crustaceans sampled from the inner
part of Hardangerfjord, where Hg contamination in sediment and sea-
water was substantially higher, were similar (mean = 0.62 in fishes
vs mean = 0.65 mg kg
−1
ww in crustaceans). Fish Se concentrations
were greater in the inner sector of Hardangerfjord compared to the
less contaminated areas of the fjord, whereas crustacean Se concentra-
tions were lower in the inner sector (Table 2;Fig. 1).
The liver in fishes and the hepatopancreas in crustaceans both play
significant roles in the distribution of toxic trace metals and high con-
centrations have often been reported (Engel, 1983;Romeo et al.,
1999). Tusk liver contained higher concentrations of both Hg (6.39 vs
1.37 mg kg
−1
ww) and Se (9.95 vs 0.66 mg kg
−1
ww) in comparison
to fillet tissue (Fig. S1; Table 2). Tusk sampled from the inner sector of
Hardangerfjord had greater Hg and Se concentrations in liver compared
to the outer sector (Hg: 8.14 vs 4.63 mg kg
−1
ww; Se: 10.42 vs
9.48 mg kg
−1
ww). However, brown meat of crab (a mixture of hepato-
pancreas, gonad and internal connective tissue) sampled from the inner
sector had higher Hg concentrations (0.16 vs 0.06 mg kg
−1
ww),
whereas Se concentrations were 42% lower compared to the outer
fjord area (0.83 vs 1.42 mg kg
−1
ww).
Se concentrations increased concomitantly with Hg concentrations
in all fish species and Pearson's correlation coefficient ranged from r
=0.36incommonlingtor= 0.49 in tusk. Similar findings have been
reported in several fish species from the Northeast Atlantic Ocean
(Azad et al., 2019). Crustacean Se concentrations in muscle varied in
the opposite direction of Hg and decreased slightly with increasing Hg
concentrations and no significant correlation was observed (Fig. 3).
Similarly, hepatic Hg and Se concentrations in tusk increased concomi-
tantly (r= 0.73; Pb0.0001). However, no correlation was found be-
tween Hg and Se concentrations for brown crab hepatopancreas
(Fig. S2). Collectively, these findings suggest an organ specificdistribu-
tion pattern in fish and crustacean species that may be driven by differ-
ential uptake mechanisms and toxicokinetics of Hg and Se.
3.2. Seafood Hg concentrations and body size
Hg concentrations increased with both length and weight in all sam-
pled fish species (Fig. 2; Table S1). Length explained a larger part of the
variation in fillet Hg concentrations (r
2
between 0.18 and 0.60; Pb0.01)
than weight (r
2
between 0.20 and 0.40; Pb0.01). Over time, Hg bioac-
cumulation leads to increasing concentration with fish age (Power
et al., 2002). Our data shows that fish length is a better proxy for age
than weight, as weight can be affected by seasonal variation and food
availability, body condition and rates of gonad maturation (Table S1).
Hg concentrations were not correlated to length in crustaceans except
for Norway lobster, where a negative linear relationship was observed
(r
2
=0.42;Pb0.05). However, Hg concentrations increased with
weight in both European lobster (r
2
=0.19;Pb0.05) and Norway lob-
ster (r
2
=0.68;Pb0.01), but not in brown crab. Since crustaceans molt,
their length increases incrementally and during the periods in between
molting steps, the weight may be a better predictor of growth than
length (Cameron, 1989).
In many crustaceans, clear differences in Hg concentrations be-
tween sexes and interactions with length have been reported. For
example, female Norway lobster from the Ligurian Sea (Minganti
et al., 1990)andoutsideScotland(Canli and Furness, 1993)showed
steeper increases in Hg concentrations with length than males. This
is likely due to slower growth rates of females in comparison to
males and as a result of more energy investment related to reproduc-
tion. In this study, crustaceans were not sexed and consequently it
was not possible to make comparisons among sexes. Analyzing indi-
viduals without information on sex could also mask effects of size on
Hg concentrations in crustaceans.
3.3. Spatial variation of Hg in seafood and sediments
In most of the studied species, Hg concentrations were higher in
samples of marine organisms collected towards the inner fjord and
PSP at Odda than in samples taken in the outer fjord (Figs. 1;4). This
spatial variation was consistent across crustacean and fish species in-
cluding European lobster, crab, tusk and sprat, but not for blue ling or
common ling.
The highest mean concentration of Hg in tusk were observed in the
two Eidfjord sites 4E and 5E (2.88 and 1.78 mg kg
−1
ww), and not in the
inner part of Sørfjord where sites 1S and 2S also showed high Hg values
(1.90 and 1.36 mg kg
−1
ww). The differences between sites 4E, 5E and
1S were, however, not significant (Fig. 4) and considering the limited
number of tusk collected from 4E (n = 7) and 1S (n = 2), tusk from
both branches appear to be contaminated at similar levels. However,
the observed high Hg concentrations in tusk from Eidfjord, 47 and
59 km from the PSP, indicate that PSP may not be the only source of
Hg to the biota in Hardangerfjord. The tusk from Eidfjord may hence
have been influenced by the freshwater inputs from two large rivers
and the hydroelectric power station located upstream on the Sima
River (Fig. 1). Moreover, substantial transport of Hg from PSP in Odda
to Eidfjord does not seem very likely, basedon the sediment concentra-
tions of Hg in Sørfjord. Measured concentrations decreased rapidly from
2.26 mg kg
−1
dw at site 1 to 0.72 mg kg
−1
dw at site 2 and 0.03 mg kg
−1
dw at site 3. At site 6 in Eidfjord the sediment concentration was again a
bit higher, with 0.17 mg kg
−1
dw, but still more than an order of mag-
nitude lower than at site 1. The combination of depth (350 m) and the
Sørfjord sill may also prevent the movementof contaminants. However,
the run-off from Opo River and fjord/estuarine water circulation driven
by local tidal conditions may also redistribute and resuspend the con-
taminants to outside Sørfjord, but the majority of Hg from PSP stays
within the Sørfjord sector. If transport of Hg should take place from
PSP to Eidfjord, resuspended Hg would have to be transported in higher
water layers, over theSørfjord sill and to the right-hand side due to the
Coriolis force effect before being deposited in Eidfjord. Tusk from site
7OH close to Steinstø, had significantly lower Hg concentrations than
tusk from the Eidfjord sites, but significantly higher Hg concentrations
than tusk from the three outermost sites.
Hg concentrations in sediment increased from the outer
Hardangerfjord towards the inner fjord and PSP at Odda (Fig. 5;
Table S3) and were in good accordance with the tusk Hg data. Sedi-
ments were sampled from the top and intermediate layers (15 ±
2 cm) resulting in an integrated sample which limits our resolution of
the interpretation. However, our spatial results are consistent with
other studies and show anincreasing gradientof mercury from offshore
628 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

to the interior of the fjord. Using a meta-analysis, Everaert et al. (2017)
reported Hg concentrations in sediment samples of 0.13 mg kg
−1
dw in
Norwegian inner fjord areas and 0.02–0.03 mg kg
−1
dw in offshore
areas. Their reported levels from inner fjord areas were comparable to
Hg concentrations measured at sites 5 and 6 in Eidfjord (0.072 and
0.173 mg kg
−1
dw), whereas the reported levels in offshore areas were
comparable to the Hg concentrations in sediment from the less im-
pacted areas of outer Hardangerfjord (0.015–0.050 mg kg
−1
dw, sites
3, 4 and 7).
Concentrations of Hg in sediments were not significantly correlated
with distance from PSP (Fig. 5). On the other hand, distance from open
ocean that takes into account both input from catchment and PSP in the
same direction, showed a significant correlation with Hg concentrations
in sediment (Kendall tau 0.90; Pb0.05) (Table S4). For tusk, there was a
significant correlation between Hg concentration and distances from
both ocean and PSP, but the correlation with distance from open
ocean was stronger (Fig. 5;r
2
= 0.59 and r
2
= 0.76, respectively) and
overall Hg concentrations in both tusk and sediment were in good
accordance.
A recently published study, which included tusk specimens from
sites 1S and 7OH as well as tusk from other areas on the Norwegian
coast, showed that the Hg stable isotope values were different in
Hardangerfjord, particularly Sørfjord, compared to the open coast of
Norway (Rua-Ibarz et al., 2019). The isotopic composition changed
somewhat from Sørfjord to the outer Hardangerfjord, to a profile
more similar to that of the open coast. This indicated that in the outer
Hardangerfjord there was an influence from the zinc plant in Sørfjord,
but also from atmospheric sources.
In areas not impacted by specific sources of pollution, atmospheric
deposition of Hg is considered a major source of Hg to the ecosystems
(Mason et al., 1994), and in coastal ecosystems Hg mostly originates
from freshwater input, organic matter decomposition and erosion
(Bełdowska et al., 2014). Fjords naturally have large river inputs often
at the ends and these often drain large catchment areas. This freshwater
run-off contains Hg deposited over the entire catchment area, including
throughfall (Kahl etal., 2007). In the Hardangerfjord, there are two large
rivers at theend of Sørfjord located close to PSP and two large rivers at
the end of Eidfjord (Fig. 1). The River Eio at the end of Eidfjord has the
Fig. 2. Linear regression between length and log Hg (redcircles) and length andlog Se (black squares) in fishand crustacean species from Hardangerfjord sampled in 2011. Slope,r
2
and P
are presented. NS = not significant. (Forinterpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)
629A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

largest catchment area in the inner part of Hardangerfjord and is likely
to transport larger amounts of atmospherically deposited Hg than the
other rivers. Sima, the other mainriver in Eidfjord sector, also has a hy-
droelectric power plant that may impact the Hg load as well as methyl-
ation (Schartup et al., 2015). In hydroelectric stations, water usually
comes from the hypolimnion layer of the reservoir which often has fa-
vorable conditions for Hg methylation. Additionally, wetting and drying
from periodic flooding of the adjacent soils can also increase MeHg pro-
duction and bioavailability. These increases in MeHg are largely driven
by the timing, frequency and severity of the reservoir flooding. Water
released from the reservoir to the fjord is often enriched in MeHg
(Pestana et al., 2018) and several studies have reported increased Hg
levels in water, plankton and fish from downstream of hydroelectric
dams (Hylander et al., 2006;Kasper et al., 2014). Also, in Sørfjord
there are several hydroelectric power plants. Freshwater inputs from
the rivers is reflected in the salinity measurements and modeling that
showed a decreasing trend in surface water salinity from the outer
part of Hardangerfjord towards both Sørfjord and Eidfjord (Fig. 6;
Fig. S3). The rivers also deliver significant amounts of terrestrial organic
matter (Jassby and Cloern, 2000) that may influence Hg methylation
and bioavailability dynamics (Lambertsson and Nilsson, 2006).
3.4. Mercury methylation in sediments
Concentrations of MeHg in sediment varied from 0.12 μgkg
−1
dw at
site 4 to 8.4 μgkg
−1
dw at site 1, closest to PSP (Fig. 5). Atmospheric depo-
sition and terrestrial run-off have been suggested as significant sources of
MeHg and inorganic Hg that can be methylated, particularly in estuarine
and coastal areas (Mason et al., 2012;Schartup et al., 2015). However,
close to the PSP, a relatively high concentration of MeHg indicates that
methylation of inorganic Hg originating from the zinc plant is taking
place to some degree. High concentrations of both total Hg and MeHg
were found close to the PSP (site 1). Comparing the Hg concentrations in
sediment at the end of Sørfjord, close to PSP, with Eidfjord (2.26 vs
0.17 mg kg
−1
dw) and the MeHg concentrations in these sites (8.4 vs
0.82 μgkg
−1
dw) shows that methylation efficiency (i.e., % MeHg) from
PSP is similar (0.37% vs 0.47%) (Table S3). In general, MeHg concentrations
in sediment increased towards the inner part of the fjord (Figs. 1;5)and
Fig. 3. Relationship between log Hg and log Se (mg kg
−1
ww) in fish and crustacean species from Hardangerfjord, 2011. Slope, r and Pare presented. NS= not significant.
630 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

were well correlated with total Hg concentrations (Kendall tau 0.71; Pb
0.05), and we can conclude that Hg concentrations likely had an important
influence on MeHg production in sediments in Hardangerfjord. In a study
from Öre River estuary in Sweden, organic matter was shown to be a pri-
mary factor controlling MeHg formation in estuarine sediments, while
total Hg had little or no effect on net MeHg production (Lambertsson
and Nilsson, 2006). The main difference between the Hardangerfjord sys-
tem and the Öre River Estuary studied by Lambertsson and Nilsson (2006)
was the absence of local anthropogenic pollution in the Öre River Estuary,
and consequently much lower concentrations of THg (ca. 18 times) in
sediment samples than what we observed in inner Sørfjord. In another
study, from the estuarine environment of the Penobscot River, Maine,
USA, with high concentrations of Hg in sediment originating from indus-
trial sources, a clear positive linear relationship was observed between
Hg and MeHg concentrations (Rudd et al., 2018).
Distance from the open ocean was the best predictor for MeHg var-
iation between the sites (Kendall tau 0.81; Pb0.05), while no correla-
tion between MeHg and distance from PSP was detected since MeHg
levels were relatively high in the inner sectors of both Sørfjord and
Eidfjord (Fig. 1). Methylmercury concentrations in the sediments are
Fig. 4. Least squares means (adjusted for mean length) + standarderror of Hg and Se concentrationsin fish and crustacean species collected from different sites in Hardangerfjord, 2011.
Hg and Se concentrations are presented on the left and right Y axes, respectively. ANCOVA/ANOVA test results are presented and letters were used to showsignificant differences when
applicable. For lobster and browncrab ANOVA was used forcomparisons between areas andarithmetic meansare presented. Onlycomposite samplesof sprat and their arithmetic means
are shown. Stations are sorted according to the distance from point source of pollution (PSP) at Odda. Letters after each station numberrepresentthe location in detail; S=Sørfjord, E =
Eidfjordand OH=Outer Hardangerfjord.The dashed red linesshow the EU maximum level of Hg (0.5mg kg
−1
ww). (For interpretation of the references to color in this figurelegend, the
reader is referred to the web version of this article.)
631A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

likely governed by Hg concentrations, anaerobic microbial activity
mainly driven by sulfate reducing bacteria in the inner sector of the
fjord, and/or by organic matter quantity and composition.
3.5. Mercury speciation in seawater
Hg species and physiochemical parameters were measured in seawater
samples taken from nine sites and three depths including 15, 50, and
300 m (Table 3). Salinity and temperature measurements showed that
the three sampling depths belong to different hydrographic layers. Brack-
ish layers were restricted to the upper 7 m of the fjord at the time of mea-
surement. Water samples taken at 15 and 50 m depths were both within
the intermediate layer while samples from 300 m depths were under
the sill level in the fjord basin water. Total Hg concentrations increased
with depth (mean of all sites 0.25 ng L
−1
at 15 m; 0.43 ng L
−1
at 50 m
and 0.52 ng L
−1
at 300 m; Fig. S4), whereas the MeHg concentrations
were highest at 50 m and lowest at 15 m depth (0.02 ng L
−1
at 15 m;
0.09 ng L
−1
at 50 m and 0.04 ng L
−1
at 300 m; Table 3; Fig. S4). The
lower total Hg concentrations observed in the shallower layers may be re-
lated to the physical properties involved with water residence time in fjord
ecosystems. Internal waves generated by wind conditions creating up- and
down-welling at the coast are an important forcing mechanism for the re-
newal of the fjord water above the sill (Asplin et al., 1999). These internal
waves are shown to occur irregularly 1 to 2 times a month and are
restricted to the upper 30 m in May and June in the Hardangerfjord ecosys-
tem (Asplin et al., 2014). Therefore, the water at 15 m depth will be ex-
changed more frequently than the water at 50 m depth, despite both
depths being intermediate layers. The lower concentration of Hg found
at 15 m depth compared to 50 m depth can be explained as a mixed effect
of both different water residence times and that the deeper layers receive
Hg deposited from the upper layers. The overall highest concentration of
Hg was found at the 300 m depth level in sites 4 and 1 (1.65 and
1.55 ng L
−1
) and also at 50 m at site 9 (1.2 ng L
−1
). MeHg concentrations
at all depths were highest at site 1 close to the PSP (0.04, 0.25 and
0.11 ng L
−1
at 15, 50 and 300 m depths, respectively).
MeHg concentration at 50 and 300 m depths in seawater, as well as
total Hg and MeHg concentrations in sediment, increased gradually to-
wards the PSP indicating a possible interaction between Hg pools in sur-
face sediments and deep layers of seawater. At deep parts of the
Hardangerfjord ecosystem, below the sill, water exchange and mixing
are very limited. MeHg produced in sediments as well as biological pro-
duction of MeHg under the mixed layer that sinks as particles to deeper
water are probably the main sources of MeHg in deep-water environ-
ments (Blum et al., 2013). MeHg concentrations in seawater at 50 and
300 m depths increased from the outer fjord towards PSP and the
inner part of Hardangerfjord (Kendall tau −0.94 and −0.93 respec-
tively; Table S5). In the inner part, a higher effect of the PSP, anaerobic
conditions (i.e. lower oxygen conditions at the fjord's interior) and
Fig. 5. Mercury pollution in sediment and tusk fillet sampled from different sites in the Hardangerfjord ecosystem. A and B: Total Hg and MeHg concentrations in sediment samples
collected from different sites sorted by distance from point source of pollution (PSP) at Oddaand distance from the open ocean. Nonparametric Kendall tau correlation coefficients are
presented. NS = not significant. C and D: Least squares means Hg ± standard error of tusk fillet (adjusted for mean length) collected from different sites with varying distance from
the point source of pollution (PSP) at Odda and distance from the open ocean. Dashed red lines show the EU maximum level of Hg (0.5 mg kg
−1
ww). (For interpretation of the
references to color in this figure legend, the reader is referred to the web version of this article.)
632 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

terrestrial run-off are expected. There was no such significant trend at
15 m depth (Table S5) where MeHg concentrations were generally
low at all sites (up to 0.04 ng L
−1
).
Percent MeHg increased significantly towards PSP for the 50 m
depth samples but not at15 m nor 300 m (Table S5).Oxygen concentra-
tions at both 50 and 300 m depths decreased towards the inner part of
the fjord, in the opposite trend of MeHg concentration and percent
MeHg at 50 m depths. Lower oxygen concentrations in deep layers are
typical of fjords due to lower rates of water exchange inside fjord sills.
A combination of low oxygen concentrations and higher organic matter
bound Hg
2+
within fjords likely provided ideal conditions for biotic
methylation and higher MeHg concentrations (Soerensen et al., 2018).
3.6. Bioconcentration Factors and Biota Sediment Accumulation Factors
For each site in the Hardangerfjord, Bioconcentration Factor (BCF)
and Biota Sediment Accumulation Factor (BSAF) were calculated for
total Hg and MeHg in tusk fillet tissue (Fig. S5). These are indicators of
how much THg and MeHg are transferred to tusk fillet from water and
sediment, respectively. Tusk was chosen for this purpose as it is a ben-
thic feeder and a deep-water fish species with low vagility (Cohen
et al., 1990). Tusk samples were collected across a broad area and in
both the inner and outer sectors of the Hardangerfjord ecosystem. BCF
values varied from 6.2 to 7.0 for total Hg and from 7.0 to 7.5 for MeHg
(Fig. S5). Tusk BSAF values for total Hg was 0.2 at site 1S, and between
1.4 and 1.8 for the other sites and BSAF values for MeHg varied between
2.6 and 4.2. Site 1S closest to PSP had lower BSAF than the other sites
due to very high Hg concentration in the sediment close to the PSP
that was not reflected in the tusk fillets. Both BCF and BSAF were higher
for MeHg than total Hg at all sites due to lower MeHg concentration in
seawater and sediment compared to total Hg and the more efficient tro-
phic transfer and bioavailability of MeHg. Lower BCF values close to the
PSP at sites 1S and 3S and much lower BSAF values for both MeHg and
Hg at site 1S compared to other parts ofHardangerfjord(Fig. 5S) may in-
dicate that Hg and MeHg originating from PSP is less bioavailable com-
pared to the Hg pool in other parts of the Hardangerfjord ecosystem.
Fig. 6. Salinity level modeled for surface (A) and 50 m depth (B) in Hardangerfjord seawater sampled in May 2018. Note the different salinity scales in each map.
633A.M. Azad et al. / Science of the Total Environment 667 (2019) 622–637

3.7. Does point source of pollution drive the spatial distribution of Hg in
Hardangerfjord?
To investigate the effect of the point source, length adjusted Hg con-
centrations in tusk from different sites were analyzed as a function of
distance from the PSP at Odda and the distance from the open ocean.
The distance from PSP explained 59% of the variation of mean Hg con-
centrations in tusk fillet, but distance from the open ocean improved
the model to 76% variance explained (Fig. 5C and D). The same trend
was observed when Hg concentrations in individual tusk were used
(Fig. S6). An explanation for this may be that an increased distance
from the open ocean not only increases the effect of the PSP, but also
water residence time, freshwater run-off and terrestrial organic matter.
Hg concentrations in sediment were very high near the PSP at Odda
and decreased sharply towards the outer parts of Sørfjord (Fig. 5), indicat-
ing that Hg pollution from PSP is likely quite local. Probably a limited
amount of Hg is transported from PSP, and Hg found in the outer
Hardangerfjord and Eidfjord may originate largely from terrestrial run-
off (Rua-Ibarz et al., 2019). A study of local soils showed that Hg concen-
trations around the zinc plant are very high compared with background
concentrations, but within a 10–20 km distance from the source, Hg con-
centrations are comparable to background conditions (Svendsen et al.,
2007). Thus, air emissions and atmospheric deposition of Hg from this
point source will likely remain mostly inside the catchment area of the
Sørfjord, and the major effect of Hg emissions is concentrated at the
head of the fjord near Odda as well as thesouthernhalfofSørfjord.How-
ever, Hg can also be distributed long distances to outside catchment areas
via atmospheric transport (Fitzgerald et al., 1998).
The average MeHg concentration in the sediment in Sørfjord closest
to the PSP was approximately 10 times higher than sediments in the
inner Eidfjord sector. Even so, the mean concentration of Hg in tusk fillet
(length adjusted) in Sørfjord was approximately 30% lower than in tusk
from Eidfjord. This suggests that MeHg from different sites may not be
equally bioavailable in these branches or that other factors such as
trophic position varies across sites. In a mesocosm experiment, using
Hg isotope tracers in both inorganic and organic forms, Jonsson et al.
(2014) showed that MeHg from terrestrial and atmospheric sources
have higher bioavailability compared to MeHg formed in the sediment
and that MeHg from terrestrial run-off has a significant effect on
MeHg burdens in estuarine biota. These findings could explain the
trend in our results, where MeHg produced in sediments from inorganic
Hg originating from the PSP appear to be less bioavailable than MeHg in
Eidfjord that likely mainly originates from terrestrial run-off and atmo-
spheric deposition, although the effects from local hydropower stations
may also be substantial. High Hg concentrations in tusk have been re-
ported from inner Nordfjord (another fjord in western Norway) com-
pared to open ocean habitats (Berg et al., 2000), but further
investigations in a fjord ecosystem without a point source are required
to fully evaluate this hypothesis. However, it seems likely thatthe fjord
ecosystems favor high Hg accumulation in deep-water, demersal fish
compared to pelagic species, with some exceptions. Moreover, life his-
tory characteristics, and spatial and temporal variation in trophic com-
plexity in these ecosystems, must also be considered important
drivers of Hg in seafood species inhabiting fjords and other coastal envi-
ronments, especially considering that subtle differences in diet and food
web position may lead to substantial differences in Hg bioaccumulation
(Bank et al., 2007). The high freshwater input from large catchment
areas are believed to deliver highly bioavailable terrestrial MeHg to
the fjord. When run-off reaches the fjord, the increase in salinity in-
creases the partitioning of contaminants bound to organic matter in
the particulate phase and thus the contaminants sedimentation can be
enhanced from suspended particulate matter entering the sediments
(Turner and Millward, 2002). MeHg will be retained in the fjord due
to limited exchange of bottom water and the presence of a shallow
sill. Additionally, Wang et al. (2018) reported the importance of Hg
methylation in subsurface water in predicting Hg in marine biota from
the Arctic. Future research should evaluate the role of subsurface meth-
ylation in relation to Hg dynamics in fjord food webs.
Table 3
Mercury speciation and physiochemical properties of seawater from the different sampling sites in Hardangerfjord, May 2018.
Site Sampling
depth
(m)
Max
depth
(m)
MeHg
(ng
L
−1
)
SD iHg
(ng
L
−1
)
SD THg
(ng
L
−1
)
%
MeHg
Temperature
(°C)
Salinity
CTD
a
Salinity
(Water
sample)
Oxygen
(%)
Oxygen
(mg
L
−1
)
Latitude Longitude
1 15 380 0.04 0.0011 0.40 0.01 0.45 9.33 7.25 31.762 97.62 10.10 60°
14,656
6° 35,747
50 0.25 0.0146 0.53 0.01 0.78 32.17 8.22 34.65 50.47 5.01
300 0.11 0.0102 1.45 0.05 1.55 6.98 7.65 34.963 34.95 49.08 4.93
2 15 780 0.02 0.0011 0.24 0.01 0.25 7.17 7.00 31.656 96.48 10.05 60° 26,58 6° 34,33
50 0.15 0.0059 0.25 0.01 0.39 37.18 8.28 34.713 56.54 5.61
300 0.07 0.0014 0.17 0.01 0.24 29.45 7.68 34.98 34.98 61.30 6.15
3 15 800 0.01 0.0002 0.13 0.01 0.14 7.68 7.25 31.762 97.62 10.10 60° 23.40 6° 20,43
50 0.13 0.0030 0.31 0.02 0.45 29.39 8.22 34.65 50.47 5.01
300 0.06 0.0003 0.11 0.01 0.17 33.01 7.71 35.012 34.99 61.15 6.13
770 0.10 0.0015 0.20 0.01 0.30 32.40 7424 35.058 35.05 53.27 5.37
4 15 500 0.01 0.0001 0.13 0.00 0.15 9.46 7.04 31.630 99.43 10.35 60° 15.55 6° 11,43
50 0.07 0.0010 0.21 0.02 0.28 25.31 8.39 34.721 61.23 6.06
300 0.03 0.0011 1.63 0.04 1.65 1.53 7.75 34.999 34.99 68.17 6.83
5 15 650 0.01 0.0004 0.20 0.01 0.21 6.25 7.09 31.580 100.50 10.45 60° 09.12 6° 04,73
50 0.06 0.0046 0.13 0.01 0.19 32.46 8.47 34.684 64.91 6.41
300 0.04 0.0013 0.22 0.01 0.26 16.54 7.72 34.996 34.99 67.76 6.79
6 15 500 0.04 0.0003 0.20 0.01 0.24 16.63 7.01 31.568 98.46 10.26 60° 00.47 5° 56,15
50 0.05 0.0016 0.16 0.01 0.21 23.82 8.53 34.693 67.93 6.70
300 0.02 0.0006 0.15 0.01 0.17 9.31 7.65 34.974 34.97 75.50 7.58
7 15 500 0.04 0.0005 0.24 0.01 0.27 13.30 7.06 31.644 99.14 10.31 59° 55.07 5° 45,15
50 0.04 0.0013 0.27 0.02 0.31 12.93 8.51 34.713 68.57 6.77
300 bLOD bLOD 0.21 0.01 0.21 bLOD 7.50 34.951 34.95 79.60 8.02
8 15 330 0.02 0.0005 0.21 0.01 0.23 7.53 7.97 32.802 95.16 9.62 59° 44.45 5° 30,38
50 0.02 0.0003 0.09 0.01 0.10 14.69 7.86 34.726 77.32 7.74
300 0.01 0.0001 0.15 0.01 0.16 5.23 7.22 35.079 35.08 81.81 8.29
9 15 330 0.01 0.0002 0.25 0.01 0.26 4.21 8.74 31.792 99.73 9.98 59° 35.72 5° 15,72
50 0.04 0.0003 1.16 0.02 1.20 3.26 7.69 34.619 79.00 7.94
300 0.01 0.0002 0.28 0.01 0.29 2.44 7.24 34.9 34.91 79.81 8.09
LOD: limit of detection.
a
An offset of 0.12 was added.
634 A.M. Azad etal. / Science of the Total Environment 667 (2019) 622–637

Using a linear model, distance from the open ocean explained 76% of
the variation in Hg concentrations in tuskfillets. The linearmodel can be
used to estimate the range of Hg contamination in deep-water species
such as tusk. Therefore, tusk can be considered an important
bioindicator for Hg contamination in high trophic species inhabiting
fjord ecosystems especially since they have a very wide distribution
and low vagility.
3.8. Comparison of Hardangerfjord seafood with the EU maximum level
Hg concentrations in fish and fishery products, including claw and
tail meat of crustaceans, are regulated by the EU in different categories
and should be below the maximum level (EUML) of 0.5 mg kg
−1
ww
for all species investigated in this study (EC, 2006). Tusk and blue ling
collected from all sites in both inner and outer Hardangerfjord had
2–3 times higher average Hg concentrations than the EUML and all indi-
vidual measurements in samples collected from inner Hardangerfjord
exceeded the EUML (Table 3). Mean Hg concentrations in common
ling from the inner part of the fjord exceeded the EUML by ~2-fold,
and mean Hg concentrations in samples from the outer part were
close to EUML (0.49 mg kg
−1
ww). Hg levels in wolffish and sprat
were well below the EUML. The sampled crustacean species had aver-
age Hg concentrations below the EUML, except for European lobster
caught in Sørfjord that had 0.62 mg Hg kg
−1
ww and four out of five
specimens had concentrations above EUML. The EUML regulates com-
mercial fishery in this area, as it is illegal to sell food exceeding
EUMLs. To protect local recreational fishers and their families from
MeHg exposure, the Norwegian Food Safety Authority (NFSA) has is-
sued a consumption advisory to avoid blue ling and tusk from the
whole Hardangerfjord and common ling from Sørfjord. Further, preg-
nant and nursing women are advised by NFSA to avoid consumption
of crab, European lobster and sentinel fish species from Sørfjord
(www.miljostatus.no).
Recently, Selenium Health Benefit Value (HBV
Se
), was suggested as a
comprehensive human health index considering the Se co-exposure
that potentially reduces bioavailability, exposure and toxicity of MeHg
(Ralston et al., 2016). Negative HBV
Se
values imply higher molar con-
centration of Hg than Se, and consumption of seafood with negative
values may be more detrimental for human health than consumption
of seafood with positive values. In this investigation only blue ling and
tusk from inner Hardangerfjord had HBV
Se
with negative values of
−1.7 and −0.6, respectively.
In general, crustaceans had higher HBV
Se
values than fish species (~3
times higher in the outer part of the fjord and ~9 times higher in the
inner sectors) since they contain less Hg and more Se (Table 2). Al-
though Hg and Se concentrations were correlated in both tusk and
blue ling from the inner part of Hardangerfjord, tusk with higher Hg
concentrations had higher HBV
Se
values than blue ling. This may be
due to differences in bioaccumulation mechanisms and toxicokinetics
of Se and Hg across taxa which have important implications for seafood
safety and overall food security.
4. Conclusions
Hardangerfjord is a Hg impacted fjord with a pollution source at the
end of its inner sector and provides a unique opportunity to investigate
Hg bioavailability in seafood species commonly consumed by humans.
Although the direct release of jarosite containing contaminants from
the zinc plant into Hardangerfjord was stopped in 1986, legacy Hg is
still present in the environment and concentrations in seawater and
sediment were highest close to this point source at the inner most
part of Sørfjord (Fig. 1). Tusk, blue ling and common ling from the entire
Hardangerfjord area and European lobster from the inner part of the
Hardangerfjord are highly polluted by Hg and well above the EUML.
Concentrations of Hg in both seafood, sediment, and seawater increased
from the open ocean to the inner part of the fjord. Although sediment
concentrations were ten times higher in the inner fjord branch with a
PSP (Sørfjord) compared to an adjacent fjord branch that may have
been influenced by freshwater inputs, Hg concentrations in the demer-
sal fish species tusk sampled from each branch were similar. Although
Hg originating from the point source was methylated in sediments
and Hg contamination in both fish and crustacean species increased to-
wards the PSP, atmospheric Hg transferred by run-off and hydroelectric
power stations cannot be ruled out as important sources of Hg to biota.
The effects of the PSP, run-off and organic matter input from the
catchment, anaerobic conditions, and residence time gradually in-
creased in the same direction (towards inner parts) and therefore it is
difficult to separate the effect of these different Hg pools on biota.
Adding a study in another fjord with similar conditions, but without a
pollution point source or conducting Hg stable isotope analysis on ter-
restrial and marine ecosystem compartments from Hardangerfjord
will likely help to better understand the relationship between different
sources of Hg, local biogeochemistry patterns andoverall bioavailability,
fate, and transport of MeHg.
Acknowledgments
The authors wish to thank the technical staff in the laboratories at
the Institute of Marine Research (IMR)for sample preparation and anal-
yses. We also thank Otte Bjelland at IMR for organizing the fishing
cruises and master student Michael Lindgren for analyzing part of the
fish and crustacean samples. Photos used in thegraphical abstract were
reprinted with permission; European lobster (photo credit: E. Senneset,
IMR) and brown crab (photo credit: O. Paulsen, IMR). We thank Arne
Duinker for assistance with preparing the study area map. This study
was supported by funding from three organizations including: Norwe-
gian Ministry of Trade, Industry and Fisheries, Norwegian Food Safety
Authority and Institute of Marine Research.
Appendix A. Supplementary data
Supplementary data to this article can be found online athttps://doi.
org/10.1016/j.scitotenv.2019.02.352.
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