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Marine Pollution Bulletin 59 (2009) 297–310
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Monitoring strategies for re-establishment of ecological reference conditions:
Possibilities and limitations
Elisabeth Alve a,*, Aivo Lepland b, Jan Magnusson c, Kristian Backer-Owe a
a
Department of Geosciences, University of Oslo, P.O. Box 1047, Blindern, 0316 Oslo, Norway
Geological Survey of Norway (NGU), 7491 Trondheim, Norway
c
Norwegian Institute for Water Research (NIVA), Gaustadaléen 21, 0349 Oslo, Norway
b
a r t i c l e
i n f o
Keywords:
Reference conditions
Capping polluted sediment
Benthic foraminiferal recovery
Species loss
Water Framework Directive
a b s t r a c t
The ecological status of an environment should be evaluated by comparison with local ‘‘reference conditions”, here defined as the pre-industrial ecological status of the 19th century. This pilot study illustrates
how micropalaeontological monitoring, using benthic foraminifera (protists) and associated geochemical
parameters preserved in inner Oslofjord (Norway) sediments, characterise local reference conditions. In
order to optimise the usefulness of the ecological information held by foraminifera and enable characterisation of temporal changes in environmental quality beyond time intervals covered by biological timeseries, the Norwegian governmental macrofauna-based classification system is applied on fossil benthic
foraminiferal assemblages. Quantitative comparisons demonstrate deteriorating ecological status in
response to increased anthropogenic forcing (eutrophication, micropollutants), including a 73% loss in
number of foraminiferal species. Despite reduced pollution during the past decades and, at one site, capping of polluted sediments with clean clay, the reference conditions are far from re-established. Micropalaeontological monitoring requires net sediment accumulation basins and careful considerations of
taphonomic processes.
Ó 2009 Elsevier Ltd. All rights reserved.
1. Introduction
After many decades of severe pollution many countries are now
in a process of cleaning up and remediating their harbours and
other human-impacted coastal areas. According to the European
Water Framework Directive (WFD, 2000/60/EC; European Communities, 2003), the EU- and associated countries including Norway,
are supposed to restore the environment to natural background
conditions, also termed reference conditions, by 2015. The reference conditions are defined as ‘‘. . .the biological quality elements
that exist, or would exist, at high status”. For most sites, local background information is not available due to lack of conventional biological and instrumental time-series extending back to preimpacted times. In such cases, the directive states that reference
conditions should be defined based on comparisons with existing,
undisturbed, or nearly undisturbed sites. However, the optimal reference conditions to be used as a baseline for comparison with the
present-day ecologic status, must be the local, natural, pre-impacted conditions rather than comparison with supposedly similar
areas. This is particularly critical for estuarine systems which spatially are very variable and where the ecologic characteristics of
* Corresponding author. Tel.: +47 22857333; fax: +47 22854215.
E-mail address: ealve@geo.uio.no (E. Alve).
0025-326X/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved.
doi:10.1016/j.marpolbul.2009.08.011
two different areas hardly are the same (Alve, 1995). Consequently,
other methods than conventional ones must be considered for
defining reference conditions.
The fossil remains of benthic foraminifera (heterotrophic,
amoeboid protists) can provide information about long-term (decades, centuries or longer) environmental and biological changes,
whether natural- or human-induced. Benthic foraminifera are
commonly used in climate reconstructions but they also allow
reconstruction of environmental change over the past few hundred
years, including changes caused by human activity (e.g., Risdal,
1963; Alve, 1991a,b; 2000; Christiansen et al., 1996; Elberling
et al., 2003; Hayward et al., 2004; Matthews et al., 2005; Scott
et al., 2005; Tsujimoto et al., 2008). As pointed out already in
1991 analyses of sediment core data ‘‘has considerable potential
for differentiating the effects of pollution from the natural ‘‘background”” (Alve, 1991a, p. 243). The increasing international application of foraminifera in environmental studies, both in time and
space, has recently been reviewed (Nigam et al., 2006). By utilizing
a combination of micropalaeontological and geochemical information stored in sediment sequences from net accumulation basins,
temporal changes in ecologic status from pre-impacted (reference)
conditions to present-day conditions can be described not only for
a local harbour or bay but even for a particular sampling site. In
such environmental assessments it is, however, important to de-
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fine what ‘‘natural” (see e.g., discussion in Willis and Birks (2006))
or ‘‘reference” conditions really are. In marine coastal zones, most
severe industrial impact occurred after the late 1800s. Consequently, and in accordance with Dale et al. (1999), we use the environmental indicators in sediments deposited before the mid 1800s
for characterising the pre-industrial background, or ‘‘reference
conditions”.
What is the time perspective for restoring the ‘‘natural” ecological status of severely impacted coastal-marine sediments? Even if
all pollution discharges were brought to an immediate halt, the
sediments inhabited by bottom-dwelling organisms would continue to be impacted for several years. Factors acting against rapid
remediation of accumulation basins include relatively low natural
sediment accumulation rates (< about 2 mm/year), run-off from
contaminated land areas, the time scales of water renewals,
bioturbation and other physical disturbance (e.g., ship traffic remobilizing contaminated sediments), and supply of air-born contaminants. Remediation in areas with highly polluted surface
sediments can be achieved by capping (or ‘‘sealing off”) the seabed
with a layer of clean sediment (e.g., Schaanning et al., 2006; Eek
et al., 2008). This method is being used in the inner Oslofjord, Norway, today.
Overall, the inner Oslofjord has experienced a reduced pollution
load in recent decades (Konieczny, 1994; Baalsrud and Magnusson,
2002). However, due to the abovementioned reasons, the surface
sediments, particularly in the harbour areas are still impacted. In
connection with the building of a new opera and a submerged tunnel in the Oslo harbour area, substantial amounts of post glacial
clay was made available as capping material. In order to test the
clay’s behaviour and suitability for capping purposes, some hundred cubic meters were deposited at 20 m water depth just outside
Oslo harbour in September 2004 (Myhre et al., 2005). The question
is how effective this effort is in restoring the reference conditions.
The aims of the present paper are to (1) illustrate and discuss
how a monitoring strategy using micropalaeontology and geochemistry can describe temporal changes in ecological status from
reference conditions through heavily impacted to improved conditions and (2) study recovery-processes and document the ecological status in surface sediments 2 years after capping of polluted
sediments with clean clay. Our data also document anthropogenically forced foraminiferal species loss in parts of the inner
Oslofjord. This is a pilot study of benthic foraminiferal recovery
following the severe human impact that inner Oslofjord has experienced over the past century. We use examples from two sites
with different pollution and remediation histories. The only published paper using benthic foraminifera for reconstructing human-induced environmental change in the inner Oslofjord is the
semi-quantitative pioneer study by Risdal (1963).
2. Investigation area
The Oslofjord is a 100 km long extension of the NE Skagerrak
and cuts into the most densely populated parts of Norway.
Throughout its length, the fjord is characterised by several sills
dividing the deeper habitats into basins. The main sill is the one
at Drøbak (19.5 m water depth), separating the inner Oslofjord
from the more southern fjordic system (Fig. 1). Inside the main sill
there is a north–south oriented ridge (about 50 m deep) ‘‘stretching out” southwards from the city of Oslo dividing the fjord in
two major basins, Vestfjorden and Bunnefjorden with
depths > 150 m. There are no topographical restrictions separating
the stations used in the present study from the main water masses
in the Bunnefjord. The fjord is microtidal with an astronomical tidal range of about 20 cm. The water circulation is estuarine with a
halocline situated at about 20 m water depth below which the
Fig. 1. Bathymetric map of inner Oslofjord, Norway. The rectangle in the upper
right hand corner shows the location of the Bjørvika-Bekkelag area just seaward of
Oslo city.
salinity and temperature generally is in the range 32–34 and
6–9 °C, respectively. The surface salinity is mainly a function of
water imported from the outer fjord as the local fresh-water discharge is small, about 20–25 m3/s, annual average (Glenne, 1963;
Gade, 1968). In Vestfjorden, more or less complete deep-water
renewals occur annually, whereas Bunnefjord experiences major
deep-water renewals about every 3–4 years only (Gade, 1968;
Baalsrud and Magnusson, 2002). Overall, the fjord has experienced
a significant decrease in dissolved [O2] since regular measurements
started in 1936 until the middle of the1980s from which time some
improvements are recorded (Fig. 2) and water depths shallower
than 125 m have not experienced anoxia since the year 2000 (Magnusson et al., 2006).
The modern industrialisation in Oslo started in the second half
of the 19th century and was mainly concentrated along the relatively small Aker River in Bjørvika but also along some other smaller rivers further west (Baalsrud and Magnusson, 2002). Municipal
waste water from the growing population of Oslo city caused
increasing pollution and, particularly after the Second World
War, the fjord received increasing amounts of toxic pollutants such
as heavy metals, PCBs, PAHs, etc. from the sewage systems as well
as from industry, land runoff and air-transport. Discharges of
organic material, nutrients and heavy metals culminated in the
early 1970’s (Konieczny, 1994). Due to governmental regulations,
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5
Oxygen equivalents (ml/l)
4
3
2
1
0
-1
1930
1935
1940
1945
1950
1955
1960
1965
1970
1975
1980
1985
1990
1995
2000
2005
2010
Year
Fig. 2. Dissolved oxygen concentrations (ml/l) at 70–75 m water depth in Bunnefjord (station Ep1), inner Oslofjord 1933–2006. Measurements from October each year,
except 1936–37, 1948, 1952, 1962 and 1966–71 (August, September, November or December). Data from Braarud and Ruud (1937), Dannevig (1945); Institute of Marine
Research (IMR), Flødevigen, Norway 1952–61 (unpublished), Norwegian Institute for Water Research (NIVA), station Ep 1 (unpublished).
closing down of local industry, and establishment of more modern
sewage treatment plants, the fjord has experienced reduced pollution supply over the past few decades (Baalsrud and Magnusson,
2002). Of the extensive environmental monitoring which has taken
place in the Oslofjord since the late 1970s, very little is so far published internationally.
In connection with Norwegian authorities’ ongoing program of
cleaning up the country’s most polluted harbours, dredging of
the Oslo harbour sediments started in late February, 2006. By the
time the samples for the present investigation were collected,
>90,000 m3 polluted harbour sediment had been dredged and
deposited in a silled basin near Malmøykalven in the southern part
of the Bekkelag Basin that is used as a deep water confined disposal
facility (CDF) (Fig. 3).
3. Material and methods
3.1. Field work
Sediment cores were collected at two sites (Fig. 3, Table 1) in
the inner Oslofjord 8th September 2006 using a light gravity corer
(core liner inner diameter 67 mm). The northernmost, site 2 at
20 m water depth, was in the outskirts of the heavily polluted Oslo
Harbour, just outside Bjørvika. Site 3, at 70 m water depth, was on
the northern slope of the Bunnefjord basin just outside the 30 m
deep SW sills separating the Bekkelag basin from the Bunnefjord.
The southernmost site was chosen because (1) it is situated in an
area which has never been directly impacted by pollution discharge from the city, and (2) it is just outside one of the sills separating the Malmøykalv disposal facility (see above) from the rest
of the inner Oslofjord. In this way site 3 is representative of the
general depositional environment in inner Oslofjord and can serve
as a reference site for evaluating possible spread of pollutants from
the Malmøykalv disposal facility to this part of the fjord.
At site 2, the pilot capping experiment was carried out in
September 2004 when an up to 8 cm thick layer of clean clay was
placed on the sea floor. In cores from site 2, the capping clay was
overlain by approximately 1 cm loose, medium brown mud which
had accumulated during the 2 years since the capping. The surface
had a dense ‘‘lawn” of polychaete tubes and there were clear signs
of bioturbation through the clean clay down into the dark
gray/blackish mud underneath. The surface veneer (<5 mm) of sediment from 7 gravity cores was carefully removed using a pipette
and frozen for later heavy metal analyses. Four of the cores were
replicates from site 2 whereas, for comparative reasons, one core
was collected about 50 m to the south (subsite 2.4) and the last
two were collected less than 100 m to the south-west (subsite 2.1).
At site 3, the top 1 cm of sediment consisted of loose, brown
mud with lots of faecal pellets < 1 mm in size and a couple of polychaete tubes extruding up into the overlying water. The underlying
sediment was dark gray down to about 5–6 cm where there was a
transition to lighter gray mud. As for site 2, the surface veneer of
sediment from three replicate cores at site 3 were carefully pipetted off for heavy metal analyses.
One sediment core from each station (labelled 2E and 3A,
respectively), was collected for stratigraphic geochemical and micropalaeontological analyses and for determination of water content. As opposed to living organisms which commonly show a
patchy and seasonally changing distribution, dead foraminiferal
tests and the other environmental tracers in sub-samples of a sediment core characterise the average environmental conditions at
the site during the time-period represented by the subsample.
Consequently, replicate sampling, which is standard procedure in
biological studies is generally less essential when analysing timeaveraged sediments. The cores were placed on an extruder, the
sediment pushed up through the core liner and sectioned on board
the ship. The surface 1 cm was cut in two 5 mm slices, the next
slice was 1–2 cm and further down, the cores were sectioned in
2 cm slices down to 20 cm.
3.2. Laboratory procedures
All samples were freeze dried and the water content calculated
as % of wet sediment weight. The dried sediment samples from
cores 2E and 3A were gently homogenised and sub-samples
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Fig. 3. Shaded relief image of the seabed of the Bjørvika-Bekkelag area with sampling sites 2 and 3. Encircled area to the south is the Malmøykalv disposal facility.
Table 1
Site details.
Site No.
Latitude
Longitude
Water depth
(m)
Site capped with
clean clay, 2004
2 (core 2E)
Subsite 2.1
Subsite 2.4
3 (core 3A)
59
59
59
59
10
10
10
10
20.4
21.0
20.8
70.0
Yes
Yes
Yes
No
53.896N
53.840N
53.882N
51.296N
44.847E
44.812E
44.850E
42.704E
collected for the following analyses: micropalaeontology of
the > 63 lm fraction, total organic carbon (TOC), extractable organic matter (EOM), and selected heavy metals.
For micropalaeontology, 1.5–1.7 g dry sediment was sieved on a
63 lm sieve and the coarser fraction dried at 50 °C. Due to limited
available material, 1.0 g sediment was used for the two topmost
sub-samples from core 3A. When possible, at least 250 individuals
(tests) of benthic foraminifera were counted for each sample. Additionally, the number of ostracods, thecamoebians, juvenile bivalves
(standardised to single valves) and gastropods, and pieces of coal/
slag per counted number of foraminifera were recorded. All absolute abundance data refer to number per gram dry sediment. Species diversity indices were calculated using the Shannon-Wiener
index H0 (log2) (Shannon and Weaver, 1963), the alpha-index of
Fisher et al. (1943), and Hurlbert’s index ES(100) (Hurlbert, 1971).
TOC was determined using the Leco combustion method (Leco
Industrial Furnace).
Samples for EOM were extracted and analysed by ‘‘Thin layer
chromatography/flame ionization detection” (TLC/FID) according
to Weiss et al. (2000): 1–3 g of dried sediment was extracted using
a dichloromethane/methanol (93:7 v/v) mixture in a Soxtec system
(Tecator AB, Höganäs, Sweden). The extracts were analysed on an
Iatroscan MK-5 instrument (Iatron Laboratories Inc., Tokyo, Japan).
The Chromarods S-III were developed first in normal hexane and
then in toluene, producing three fractions of EOM: saturates, aromatics and polars.
For heavy metal (Cu, Zn, Pb, Cd, Hg) analyses, acidified aqueous
sample solutions were obtained by leaching 1 g of sediment in 7 N
HNO3 in an autoclave at 120 °C for 1 h (Norwegian standard NS
4770). This is the standard extraction method on which the Norwegian Pollution Control Authority’s (SFT) classification of environmental quality in fjords and coastal waters (Bakke et al., 2007) is
based. All reported elements except Hg were analysed using Thermo Jarrell Ash ICP-AES 61 with detection limits of 1 mg/kg for Zn
and Pb, 0.5 mg/kg for Cu and 0.1 mg/kg for Cd. Hg-concentrations
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were determined on a CETAC M-6000A Hg atomic absorption cold
vapour analyser with a detection limit of 0.01 mg/kg (±10 rel.%).
Additionally, for comparative purposes, the acid soluble components of Cu, Zn, and Pb in cores 2E and 3A, were analysed using
0.3 N HCl extractions run in a Perkin Elmer Model 503 atomic
absorption spectrophotometer with a detection limit of 1 mg/kg.
3.3. Classification of ecological status
Hardly any papers on benthic foraminiferal responses to heavy
metal pollution use the same methodology for determining metal
concentrations. Similarly, there have been no efforts to quantify
possible relationships between concentrations of different pollutants and effects they may have on individual taxa or higher-level
faunal parameters, i.e., standardise categories for differentiating
levels of impacts. This problem is not specific for benthic foraminiferal studies. Hopefully, international consensus concerning methodologies for defining ecological status will be achieved within
the WFD. In the mean time, and as a first step to investigate possibilities for calibrating environmental information from benthic
foraminiferal assemblages and that of macrofauna, the present paper uses the criteria set by the Norwegian Pollution Control
Authority’s (SFT’s) classification of environmental quality in fjords
and coastal waters (Table 2) to characterise ecological status. In order to incorporate information from benthic foraminifera into the
established classification system, we test a new approach by using
diversity values based on benthic foraminiferal data rather than
those of soft-bottom macrofauna (Molvr et al., 1997). Impacts
that e.g., seasonal variability have on such benthic indices and their
usefulness for assessing ecosystem quality (Reiss and Kröncke,
2005) will be considered at a later stage. For metals, we use the updated version of the Norwegian environmental quality classification (Bakke et al., 2007).
4. Results
For both cores, the water content, TOC, EOM, and heavy metal
concentrations show similar down-core distribution patterns, but
there is a clear difference between sites (Figs. 4–6). The > 63 lm
fraction consisted mainly of faecal pellets and sediment aggregates
made up of finer detrital grains. This is a characteristic feature of
recent Oslofjord sediments (Alve, 1991a; Alve and Olsgard,
1999). Due to the fine sediment texture throughout the cores, it
is unlikely that the concentrations of organic carbon and heavy
metals are significantly influenced by grain size trends.
In core 2E (capped site), maximum concentrations of all
reported parameters (water content 72%, TOC 5.7%, EOM
23.9 mg/g, HNO3-extracted Hg 7.9, Cd 24, Cu 658, Zn 1450, and
Pb 427 mg/kg, and HCl-extracted Cu 728, Zn 1097, and Pb
437 mg/kg), occur in the dark gray/blackish sediment interval at
about 13–20 cm depth. The values show a gradual decrease upcore from about 13 to 4 cm, and minima in the clean clay-interval
at 1–4 cm. The concentrations in surface sediments (TOC 1.6%,
EOM 6.9 mg/g, HNO3-extracted Hg 1.2, Cd 0.7, Cu 105, Zn 246,
and Pb 97 mg/kg, and HCl-extracted Cu 65, Zn 110, and Pb
61 mg/kg) are significantly lower than in the sediments underlying
the clean capping clay, at 4–6 cm core depth.
In core 3A, the water content, TOC, and EOM were stable at
about 54%, 1.4%, and 2–4 mg/g respectively, in the lower part up
to 12 cm core-depth, increased slowly up-core to 4 cm from where
they increased more rapidly to maxima of 74%, 4.8%, and 20.0 mg/g
in the surface samples. The concentrations of Cu, Zn, and Pb below
12 cm core depth were stable at about 15, 50, and <30 mg/kg for
the HCl-extracted, and about 20, 140, and 30 mg/kg for the
HNO3-extracted samples, respectively (Fig. 5). [Hg] and [Cd] were
<0.1 and 0.1 mg/kg, respectively. From about 12 cm, the metal
values increased up-core to culminate at 1–2 cm with maxima of
142, 399, 177 and 135, 501, 182 mg/kg for the two extraction
methods, respectively. All showed a decrease in the surface few
millimetres reaching values in the same range as those at 2–6 cm
core-depth.
Heavy metal extractions using HNO3 rather than HCl gave concentrations which were 3% and 7% higher for Pb, and 29% and 37%
higher for Zn in cores 2E and 3A, respectively. For Cu, the values
were 5% for core 2E, whereas in core 3A HCl-extractions yielded
20% higher concentrations than those using HNO3.
In core 2E from the capped site, benthic foraminiferal tests were
nearly absent below 10 cm core depth (66 tests/g dry sediment).
Above 10 cm, the figures were in the range 4–44 tests/g dry sediment (minimum values in the capping layer at 1–4 cm), except
for the surface 0.5 cm which yielded 125 tests/g dry sediment.
Stainforthia fusiformis dominated the surface sample (64%) which
had a total of 22 species and ES(100), H0 (log2), and Fisher-alpha values of 15.9, 2.2, and 6.2, respectively.
Core 3A had abundant foraminiferal assemblages throughout
(Fig. 7) and a total of 91 recorded species. Of the 78 species recorded in the four oldest (lowermost) samples, 21 were present
in the surface three samples (i.e., upper 0–2 cm). Otherwise, the
oldest samples from 20–12 cm core depth were characterised by
stable, highly diverse foraminiferal assemblages (Fig. 8) with
S. fusiformis and Cassidulina laevigata as the most common, and
Nonionella auricula, N. iridea, Bulimina marginata, Hyalinea balthica,
Nonionellina labradorica, and Miliolinella subrotunda, as subsidiary
species. From 12 to 6 cm core depth the relative abundance of
S. fusiformis, B. marginata, Elphidium albiumbilicatum, and Bolivinellina pseudopunctata increased, particularly at the expense of
Table 2
The Norwegian Pollution Control Authority’s (SFT’s) classification of environmental quality (reflecting ecological status) in fjords and coastal waters for metals (Bakke et al., 2007)
and organic carbon and species diversity for soft-bottom macrofauna (Molvr et al., 1997). Only parameters relevant to the present study are shown.
Class
Environmental quality
I
Background
II
Good
III
Moderate
IV
Bad
V
Very bad
Metals
Pb (mg/kg)
Cd (mg/kg)
Cu (mg/kg)
Hg (mg/kg)
Zn (mg/kg)
<30
<0.25
<35
<0.15
<150
30–83
0.25–2.6
35–51
0.15–0.63
150–360
83–100
2.6–15
51–55
0.63–0.86
360–590
100–720
15–140
55–220
0.86–1.6
590–4500
>720
>140
>220
>1.6
>4500
Environmental quality
Very good
Good
Less good
Bad
Very bad
TOC (%)
Species diversity
Hurlbert’s index ES(100)
Shannon-Wiener index H0 (log2)
<2.0
2.0–2.7
2.7–3.4
3.4–4.1
>4.1
>26
>4
26–18
4–3
18–11
3–2
11–6
2–1
<6
<1
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Fig. 4. Down-core distribution of data in two sediment cores, inner Oslofjord, reflecting contrasting pollution and remediation histories. (A, B) Relative values for water
content (dashed line) and TOC 10 (solid line). (C, D) Concentrations of Hg (dashed line) and Cd (solid line). Note different scales on x-axes. (E, F) Absolute abundance (No./g
dry sediment) of coal and slag fragments (E) and valves of juvenile bivalves (F). Gray rectangles along core 2E y-axes represent clay cap from 2004.
N. auricula and N. iridea, and the species diversity declined. In the
surface 6 cm, the former four plus Leptohalysis spp, and Epistominella vitrea were the most common species. Whereas the relative
abundance of C. laevigata, H. balthica, and N. labradorica started
declining at about 6–8 cm, their absolute abundance did not decrease until at about 2 cm. The species diversity reached a minimum at 1–2 cm. Overall, the abundance of benthic foraminiferal
tests was stable at about 175 tests/g dry sediment up-core to
6 cm where it started increasing and reached a maximum of
2260 at about 2 cm before it declined to 1497 tests/g dry sediment
in the surface sample.
Tests of S. fusiformis, which has a general shape of a wheat grain,
were often ‘‘hidden” in faecal pellets. This probably caused an
underestimation of its abundance. Very few deformed foraminiferal tests (<1% of the assemblages) were recorded. Most samples
from core 2E contained 1–2 thecamoebians whereas none were
recorded in 3A. Insignificant numbers of juvenile gastropods were
recorded. Juvenile bivalves were abundant in core 2E, particularly
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E. Alve et al. / Marine Pollution Bulletin 59 (2009) 297–310
303
Fig. 5. Down-core distribution of heavy metal concentrations (mg/kg) in two sediment cores from inner Oslofjord reflecting contrasting pollution and remediation histories.
Dashed line = data based on 0.3 N HCl-extractions; solid line = data based on 7 N HNO3-extractions. Additional data from the surface few millimetre of sediment from
replicate cores are also shown. (A, B) Cu, (C, D) Zn and (E, F) Pb. Gray rectangles along core 2E y-axes represent clay cap from 2004. Environmental Zones I–III are indicated for
core 3A.
below 12 cm (Fig. 4F). In core 3A their abundance peaked in the
surface 1 cm.
5. Discussion
5.1. Temporal changes in pollution and sediment accumulation rates
No datings were available for the analysed sediment cores. They
were equally long (20 cm) but the results reveal that they represent different time intervals. Based on general knowledge about
the pollution history of Oslo city and correlations with heavy metal
concentrations in dated cores from the same general area, it is possible to infer inter-core chronologies. Core 2E from the capped site
was collected at 20 m water depth in the outskirts of Oslo harbour,
Bjørvika (Fig. 3), close to where two small rivers enter the fjord.
Consequently, the site has had a higher sediment accumulation
rate than the more open water site 3, not only due to supply from
rivers (as seen through presence of thecamoebians) and land drainage but also from harbour activity (e.g., sediment supply from
years of nearby dredging, propeller erosion, snow dumping). Even
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Fig. 6. Down-core yields of extractable organic matter from Iatroscan TLC/FID analysis: (A) core 2E, (B) core 3A. ND = no data.
though the sediment accumulation rates have been different, the
general supply of heavy metals to the inner Oslofjord, and thereby
to the two sites in focus, has followed the same temporal pattern.
Heavy metal concentrations in Oslofjord sediments started
increasing in the early 1900s, reaching maximum values in the
1950s to early 1970s, and have declined since then (Konieczny,
1994; Geological Survey of Norway, unpublished 210Pb-datings).
Consequently, stratigraphic patterns of heavy metal distribution
with maxima in 1950s-early 1970s are recorded in sediment cores
throughout the Oslofjord and can be used for correlation and chronological purposes. It follows that the maximum heavy metal concentrations in the lower part of core 2E most likely represent the
heavy pollution period up until the 1970s, whereas the reduction
starting between 12 and 14 cm core depth reflects subsequent reduced supply. The abrupt reduction at about 4 cm represents the
transition to the clean capping clay from 2004. For core 2E, this
gives a sediment accumulation rate of about 3 cm/yr between
the early 1970s and 2004. For core 3A, three environmental zones
are defined based on the distribution of heavy metals. Zone I, below 12 cm core depth, reflects the pre-impact condition of the
1800s. Bioturbation may have caused some mixing of the ecological signals down into older sediment layers making changes appear
to have happened somewhat earlier. From 12 cm core depth the
concentrations increase upwards, first slowly (defined as Zone II)
up to about 6 cm and then more rapidly (defined as Zone III) until
they reach maxima at about 2 cm, reflecting increased contamination of the fjord sediments through the 20th century until culmination in the 1960s-early 1970s. This gives an average sediment
accumulation rate of about 1.4 mm/year for the 12–2 cm core
interval implying that the transition at 6 cm occurred around
1940. Restrictions on pollution discharges over the past few decades have probably reduced the sediment accumulation rate, as
illustrated by reduced concentrations of coal and slag fragments
(Fig. 4E). Overall, the relatively low sediment accumulation rate
at station 3 provides a less than optimal time resolution.
The variation in total EOM in the two cores (Fig. 6) shows
roughly the same trends as TOC and heavy metals. The Iatroscan
TLC/FID method used in this study does not allow quantification
of single compounds or compound classes like PAH, but the saturated and aromatic fractions added together roughly correspond
to what in earlier studies has been called ‘‘total hydrocarbons”
(THC) (Konieczny, 1992) or ‘‘sum mineral oil” (Kibsgaard et al.,
2005). Since the PAHs belong to the aromatic fraction, this value
corresponds to a maximum possible value for PAHs. Polar and
non-polar lipids probably also constitute more or less significant
parts of all three fractions. In core 2E, the sum of the saturate
and aromatic fractions in the lower part of the core is in the range
10–14 mg/g dry sediment, compared to 2–38 mg/g sediment earlier documented for the Bjørvika region (Konieczny, 1992; Kibsgaard et al., 2005). There is a gradual reduction in both saturates
and aromatics up-core. The main source of THC in the Oslo harbour
sediments are spillage from oil stock tanks and bunker installations
(Konieczny, 1994), and the decrease in saturates plus aromatics
upwards through the core probably reflects the reduction in number of possible oil pollution sources during the last 50–60 years. If
we assume that variations in the aromatics reflect corresponding
variations in PAH, there is also a marked reduction of these components upwards through the core, probably reflecting a reduction of
PAH sources like coal, coke, pitch, soot, tar and creosote through
time. Core 3A shows low values of polars and saturates through
Zones I and II. Aromatics are below detection limit. From 4 cm
and upwards there is a marked increase in the amount of saturates
and polars, most likely caused by the increase in primary production as a response to increasing eutrophication. The top sediment
layer shows a reduction in saturates and polars, which parallels
the reduced nutrient supply during the last years.
5.2. Temporal changes in ecological status
As for most marine coastal areas, information about the softbottom communities and environmental status in inner Oslofjord
before the 1930s is very limited. In the present retrospective study,
the stable, low concentrations of heavy metals (all within SFT’s
‘‘background” levels) indicate that human activity did not impact
the environmental quality at 70 m water depth (site 3) before
the late 1800s. This is reflected in the faunal characteristics of Zone
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305
Fig. 7. Down-core distribution of benthic foraminiferal abundance data in core 3A reflecting faunal development from reference conditions (lower part of core, Zone I)
through maximum impact (1–5 cm sediment depth) to present-day (top) conditions in inner Oslofjord. (A, C, E) absolute abundance (No. of individuals/g dry sediment).
(B, D, F) relative abundance. Note different scales on x-axes. Environmental Zones I–III are indicated.
I, showing a stable production of diverse benthic foraminiferal
assemblages. The diversity indices, ES(100) and H0 (log2), show values reflecting ‘‘very good” ecological status (Fig. 8, Table 2). It is beyond the scope of this pilot study to discuss the reliability of
different biological indices for ecosystem quality assessment (see
e.g., Labrune et al., 2006; Blanchet et al., 2008; Bouchet and
Sauriau, 2009). During environmental Zone II slightly increasing
concentrations in all environmental parameters accompanied by
a moderate change in faunal composition reflect deteriorating conditions. By the end of Zone II, the ecological status had been re-
duced to ‘‘good” for most parameters. The changes were more
severe through Zone III culminating with concentrations characterised as ‘‘very bad” for Hg and TOC, and ‘‘bad” for Cu and Pb (Table
2). Degradation of organic matter is typically limited to the uppermost layers of sediments where TOC can be reduced by as much as
20% in well-oxygenated depositional settings, but it does not
change concentrations significantly in deeper anoxic sediment layers (Hodell and Schelske, 1998). Consequently, as the values in core
3A are reduced by 70% from top to bottom, it reflects increased
Corg-flux to the sediments, particularly since the 1940s. The faunal
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0
5
10
15
20
0
25
30
ES(100) (3A)
H'(log2) (3A)
Fisher α (3A)
ES(100) (2E)
H'(log2) (2E)
Fisher α (2E)
2
4
Sediment depth (cm)
Zone III
6
~ 1940
8
Zone II
10
~ 1900
12
14
16
Zone I
18
20
Fig. 8. Down-core distribution of benthic foraminiferal species diversity indices in
core 3A reflecting faunal development from reference conditions in the 19th
century (Zone I) through increasing deterioration during the first half of the 20th
century (Zone II) to maximum impacted conditions (Zone III) in inner Oslofjord.
Solid symbols (circle, diamond, triangle) = species diversities of the surface sample
from core 2E representing foraminiferal assemblages which had recolonised the site
2 years after capping with clean clay in 2004.
change accompanying this drastic deterioration culminated with a
foraminiferal production more than one order of magnitude higher
than those of the reference conditions (Zone I) and diversity indices in the SFT’s class IV, ‘‘bad” environmental quality. It can be
speculated that the decreased pollution load, and particularly reduced nutrient supply, over the past decades (Magnusson et al.,
2006) have caused the foraminiferal diversity to improve from
‘‘bad” to ‘‘less good”. However, the TOC-values still reflect ‘‘very
bad” ecological status (Fig. 4B, Table 2).
It is reasonable to assume that bioturbation over the past decades has caused some blending of the older polluted sediments
into the successively less contaminated sediments deposited in
the area and thereby ‘‘smeared” (see Schafer, 2000) the pollution
signal up-core. However, the clearly reduced heavy metal concentrations in the surface few millimetres compared to values 1–2 cm
below (Figs. 4D and 5B, D, F) indicate that the mixing is a relatively
inefficient or slow process. The reduced concentrations in the
surface few millimetres also show that 7 months of disposal of
polluted (SFT class IV–V, Table 2) harbour sediments in the adjacent Malmøykalv basin had not impacted the sediments in the site
3-area, SW of the basin.
Core 2E from the capped site did not penetrate through to
unimpacted sediments but reflects the development following
the peak in anthropogenic forcing in the 1950s–early 1970s
(Figs. 4–6). The declining heavy metal concentrations after the
early 1970s are in accordance with previous sediment core analyses in the area (Konieczny, 1994; Geological Survey of Norway
unpublished data). However, by 2004, when the clean post glacial
clay was deposited, the concentrations in the surface sediments
(now at about 4 cm core depth) were still in the categories ‘‘bad”
to ‘‘very bad”. Slightly elevated concentrations within the clean
clay relative to background values reflect that some mixing with
the underlying, in situ, sediments had occurred due to subsequent
bioturbation (active organisms recorded below the clay layer).
Prior to deposition, the clay was mixed with sea-water (1:1) which
made it ‘‘soft” (Myhre et al., 2005) and probably facilitated bioturbation by colonizing organisms and incorporation of surrounding
sediments. The heavy metal concentrations in the post-, relative
to the pre-, 2004-sediments, reflect that the capping resulted in a
reduction in all investigated heavy metals by about one class (Table 2). This improvement occurred despite the facts that (1) the
capping layer at site 2 had a thickness of about 3 cm only, (2) some
original, polluted in situ sediment was suspended during deposition of the clay and settled as a veneer (unspecified thickness) on
top of the capping clean clay as observed three months after
deposition (Myhre et al., 2005) and (3) the clay had been exposed
to bioturbation for nearly 2 years. Similarly, controlled experiments have shown that capping of harbour sediments prevent release of dissolved bioavailable organic pollutants (PAHs, PCBs, and
DDTs) to the overlying water (Schaanning et al., 2006). On the
other hand, it can be argued that improvements would have occurred anyway due to restrictions on pollution discharges over
the past decades.
Factors which have worked against an even stronger ecological
improvement of the surface sediments since the clay was laid
down include run-off from contaminated land areas (particularly
during heavy rain), ship traffic (including about 140 cruise ships
annually) which continuously resuspend neighbouring polluted
sediments, and air-born supplies of contaminants. Improved ecological status following capping of the site was also reflected by
colonization of benthic foraminifera which had not inhabited the
site since at least the 1950s (Alve, unpublished data). The few foraminiferal tests present below the surface sediments had probably
been brought into the area in suspension together with other sediment particles and it was first when the clean clay was available
as substrate that the conditions got acceptable for colonization
and subsequent growth and reproduction (Alve and Goldstein,
2003). By 2006, the environmental quality (based on foraminiferal
diversity) had, in the same way as at site 3, improved to ‘‘less good”
(Table 2).
5.3. Cause-effect relationships
Based on the abovementioned, the question rises: what is the
cause-effect relationship between the geochemical/hydrographical
and faunal changes? Ecosystems are under pressure of complex
mixtures of contaminants whose effects are not always easy to assess based on macrofauna (e.g., Trannum et al., 2004). Our data
illustrate that the same applies to benthic foraminifera. Single
parameter-effects are difficult to distinguish in multisourcepolluted, naturally unstable, laterally variable environments like
the coastal zone and little is known about benthic foraminiferal
tolerance to specific pollutants. Several investigations have pointed
to correlations between benthic foraminiferal distributions and
concentrations of e.g., specific heavy metals. The significance of
such correlations is questionable as long as synergistic effects
and other key parameters known to affect benthic foraminifera
(e.g., organic carbon fluxes, sediment pore-water hypoxia/anoxia)
are not considered. Generally, controlled experiments are needed
to single out possible specific cause-effect relationships. For example, in a field experiment changes in living (stained) foraminiferal
community structure (e.g., increased equitability; reduced abundance) only occurred at [Cu] > 900 ppm, as compared to control
values of 70 ppm (Alve and Olsgard, 1999). Neither the foraminifera nor the macrofauna (Olsgard, 1999) showed any significant differences in the number of species between control and treatments.
This is in contrast to field observations where the number of macrofaunal species is roughly halved for each 10-fold increase in sediment Cu-concentration (Rygg and Skei, 1984; Rygg, 1985) or
where negative correlations have been found between benthic
foraminiferal diversity and e.g., Cu, Zn, Hg, and Pb (Armynot du
Châtelet et al., 2004; Carnahan et al., 2008). Similarly, exposure
to 0.02 and 2.00 nmol of Tri-n-Butyltin (TBT) per g dry sediment
did not change the number of benthic foraminiferal species signif-
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icantly relative to control levels, but the 2.00 nmol treatment
caused decreased numerical abundance of individuals relative to
the 0.02 nmol/g treatment (Gustafsson et al., 2000).
It is not the total concentration of heavy metals in sediments,
but the fraction that is bioavailable which is important for the
organisms. Consequently, the leaching method used should not
be too strong so that it extracts more than what is actually bioavailable. Discussion of this topic would go beyond the scope of
the present paper but it is interesting to note that the use of the
stronger extraction medium 7 N HNO3, which the Norwegian classification is based on, as compared to 0.3 N HCl had a pronounced
impact on the concentration of extractable Zn and less so on Pb and
Cu (Fig. 5).
Why were benthic foraminiferal tests hardly present below
the clean clay at the capped site 2? Taphonomic loss through
carbonate dissolution is not likely to have occurred because
the majority of the few individuals present were calcareous
forms and thin-shelled, well-preserved juvenile bivalves were
abundant throughout the core (Fig. 4E). According to observations at a nearby station, the water masses at this site (20 m
depth just outside Oslo harbour) have experienced occasional
anoxia between 1962 (first time monitored) and 1983, mainly
in the autumn. Although the concentrations of the measured
environmental parameters fall within SFT’s most severely polluted classes IV–V (Table 2), the values of e.g., Cu are well below
those where S. fusiformis should be significantly affected (Alve
and Olsgard, 1999). The other metal values are 1/16th–1/70th
of those of the Eggerelloieds scaber dominated assemblages in inner Sørfjord, W Norway, (Alve, 1991b). The closest we get to explain the absence of in situ benthic foraminifera during the
decades prior to capping is a combination of occasional anoxic
bottom water, anoxic sediment pore-water (TOC > 5%) and a
mixture of pollutants acting in concert.
At site 3, the main faunal changes were caused by species showing a positive or a negative response to the environmental changes.
Species belonging to what is here defined as the ‘‘positive response
group” showed a dramatic increase in production, by more than an
order of magnitude, since the 1940s (Fig. 7). S. fusiformis, B. marginata and to a lesser extent E. albiumbilicatum, B. pseudopunctata, and
Leptohalysis spp. account for most of the increase. These are all species well known to thrive in organic rich fjord sediments and tolerate hypoxia (e.g., Alve, 2000; Gustafsson and Nordberg, 2001;
Patterson et al., 2000). A similar faunal shift following development of oxygen deficiency was recorded in the Gullmarfjord,
Sweden (Nordberg et al., 2000). The opportunistic species Stainforthia fusiformis predominates benthic foraminiferal assemblages in
silled oxygen-depleted Skagerrak fjords and basins (Alve, 2003)
and has been shown to bloom in response to increased primary
production (Gustafsson and Nordberg, 2001). Also, strongly
increasing abundance of S. fusiformis (as Virgulina fusiformis), B.
marginata, and to a lesser extent E. albiumbilicatum (as E. subarcticum), was recorded just before, or at, the transition from gray to
black sediments in cores collected at 25–150 m in inner Oslofjord
in 1961/62 (Risdal, 1963). The increase was attributed to introduction of modern water toilets during the 1920s and 1930s. Consequently, the increased foraminiferal production as seen at site 3
is probably a direct effect of the 7–8 times increase in nitrogen
and phosphorous supply to the fjord between 1930 and 1970 causing extensive eutrophication (Magnusson et al., 2006), subsequent
organic enrichment at the seafloor (Fig. 4B), and severe oxygen
depletion (Fig. 2). After measurements of dissolved [O2] in inner
Oslofjord started on a regular basis in 1936, anoxia was recorded
up to 75 m water depth ( site 3) for the first time in 1950 (Beyer
and Føyn, 1951). Fig. 2 shows that either hypoxia or anoxia has occurred at this depth at irregular intervals since the early 1960s.
However, the anoxic condition at 75 m recorded by Beyer and Føyn
307
(1951) one month later than the value shown in Fig. 2, is not entirely in accordance with these observations.
When exposed to bottom water oxygen concentrations <2 ml/l,
85% of the present kind of inner Oslofjord assemblages live in the
surface 1.5 cm sediment, and 55% in the surface 0.5 cm (Alve and
Bernhard, 1995). Indeed, as site 3 (70 m) has hardly experienced
dissolved oxygen concentrations >2.0 ml/l since the early 1960’s,
subsurface microhabitats can not account for the dramatic increase
in abundance of foraminiferal tests (shells) above 6–8 cm in core
3A. Since the 1970s, the nutrient supply has been reduced to
1950-levels (Magnusson et al., 2006). However, restricted deepwater renewals in the inner fjord and the ‘‘oxygen debt” in the sediments, as reflected by the higher than normal organic load
(Fig. 4B), have delayed the positive effects of reduced organic supply in Bunnefjorden as opposed to the better ventilated Vestfjorden
(i.e., as soon as oxygenated water enters the fjord basin, degradation of organic material causes the [O2] in the bottom water to
drop). This mechanism has probably kept the sediment porewaters at a lower-than-normal oxygen-level and caused a continued dominance of the abovementioned species. Reduced nutrient
supply over the past decades may have caused the production of
benthic foraminifera to drop as illustrated by reduced abundance
in the surface sediments (Fig. 7). The overall temporal development in benthic foraminiferal production in response to variations
in food supply agrees with cultural eutrophication signals (dinoflagellate cysts) from the same general area (Dale et al., 1999).
The species belonging to the ‘‘negative response group” either
started responding already during the early 1900s (N. auricula
and N. iridea) or when the anthropogenic forcing culminated during the 1950s-early 1970s (C. laevigata, H. balthica, N. labradorica;
Fig. 7). Nonionella auricula and N. iridea showed a clear reduction
(both in absolute and relative terms) during the 20th century
reflecting an overall negative response. A similar drastic decrease
was recorded by Risdal (1963) in several inner Oslofjord basins.
Although the relative abundance of C. laevigata, N. labradorica,
and H. balthica decreased up-core from about 9 cm, the fact that
their absolute abundance is stable and only decline in the surface
2 cm shows that they were not negatively affected by the changing
environmental conditions until bottom water [O2] of <1 ml/l
started occurring about every 1–3 years from the 1960s (Fig. 2).
Prior to the onset of the most severe oxygen depletion, they simply
did not respond to the increased nutrient supply. Whereas H. balthica and C. laevigata (by some authors named C. carinata which is a
variety of C. laevigata) typically live in the surface 0–1 cm (e.g.,
Fontanier et al., 2002; Langezaal et al., 2006), the near absence of
N. labradorica in the surface 1 cm is probably due to the fact that
it lives just below that level in Oslofjord sediments (Alve and
Bernhard, 1995). To what degree the different response patterns
(positive and negative) are effects of variations in nutrient, oxygen
concentration, toxic substances or a combination of these can not
be stated with certainty. However, based on the fact that the species in the ‘‘positive” group are known to thrive in organic rich,
oxygen depleted environments and, that the concentrations of heavy metals at site 3 did not seem to hamper their reproduction, it is
likely that the organic material/oxygen factor is the most important. The ‘‘negative response group” includes more sensitive
species.
All these responses caused an overall reduction in species diversity reflecting deteriorating ecological status going from ‘‘very
good” (reference conditions) in the 19th century to ‘‘bad” during
the past decades. The reduced species diversity indices were due
to a combination of increasing populations of a few opportunistic
species and a loss of more sensitive ones. In total, 57 of the 78 species occurring in the older parts of core 3A (Zone 1) were not recorded in the surface 0–2 cm, implying a loss of 73% of the
species relative to reference conditions. The fact that decreased
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pollution load to the fjord has not yet turned the ecological conditions in the sediments in a positive direction (i.e., the ecological
status is still far from reference conditions) is probably due to a
combination of relatively low sediment accumulation rate
(<1.5 mm/year) and the ‘‘oxygen debt” stored in the surface
sediments.
5.4. Monitoring re-establishment of reference conditions
The present case study demonstrates that a combination of
micropalaeontologically- and geochemically-based stratigraphic
analyses of cores from sediment accumulations basins represent
a promising tool for defining local ecological reference conditions. The method makes it possible to evaluate whether or not
the ecological status at a site has changed following suspected
changes in the environmental conditions. It allows quantification
of such changes and the data will serve as a baseline for
evaluating to what degree the reference conditions have been
re-established subsequent to environmental remediation. However, the success of such reconstructions depends on certain criteria and an awareness of possible pitfalls which may cause
misinterpretations.
(1) In order to record temporal changes, the site in focus needs
to be a net sediment accumulation area. Ideally, this should
be confirmed by adequate dating of the sediments (e.g.,
210
Pb, 137Cs) or, as in the present study, through correlations
with known historic events.
(2) The degree of time-averaging (i.e., ‘‘. . .the process by which
fossils of different ages are mixed into a single assemblage”
Martin, 1999, p. 281), which depends on sediment accumulation rate and bioturbation activity (see below), has to be
considered (discussions in e.g., Martin (1999), Schafer
(2000)). As pointed out by Martin (citing Behrensmeyer
and Kidwell, 1985) time-averaging can be an advantage
since short-term noise is damped.
(3) In order to perform a meaningful interpretation, it is essential to know if the fossil record at the site is representative
of the assemblages which lived there during the time interval in focus (discussion in e.g., Murray, 2006). Consequently, taphonomic processes need to be considered. In
the present context the most important are physical mixing of sediments, and loss and gain of species. Because
heavy bioturbation and physical human impact on the sediments (e.g., trawling, dredging) can seriously blur or
destroy parts of the original stratigraphy, sediment cores
should ideally be x-rayed and dated prior to further analyses. Although mixing through bioturbation may seem a
problem, the successful (but so far limited) use of the
method in coastal areas all over the world and the first
author’s 20 years experience show that this is less of a
problem for the biostratigraphy than could be expected.
Dating of the sediments also show whether or not a core
is suitable for chronological, palaeoecological reconstructions, i.e., if the original chronology (and thereby the chronological succession of the fossil assemblages) of the
stratigraphic record is preserved. As for loss of species, carbonate shells may be removed through dissolution and
agglutinated forms may be lost through degradation of
their organic cement. As exclusively agglutinated foraminiferal assemblages generally are confined to intertidal
marshes and open oceanic areas below the CCD, severe
effects of carbonate dissolution can be revealed by the pure
absence of all carbonate shells. Partial dissolution may be
recognized by the presence of etched and weakened shells
or by the presence of inner organic linings (although, in
subtidal areas, the latter may have been transported out
from shallower waters). However, even if all calcareous
species are lost through dissolution the remaining agglutinated assemblages still holds useful ecolocical information
(e.g., Murray and Alve, 1999). Except for the most fragile
forms, agglutinated species seem to be well preserved in
basins with sediment accumulation rates > about 1 mm/
year and where the redox boundary is situated in the
upper few mm of the sediment (Alve, 1996). Gain of exotic
species may, in addition to natural transport by bottom
(e.g., tidal) currents, occur through human activity, e.g.,
dumping of dredged- (Alve, 2000) or ballast sediments
(e.g., Bouchet et al., 2007).
(4) Our interpretations can not be better than our understanding of the biology and ecology of the preserved organisms.
For benthic foraminifera, although our insight has improved
over the past decades (Murray, 2006) there are still major
gaps in our knowledge concerning quantitative relationships
between faunal responses and the environmental forcing
causing them.
6. Conclusions
A monitoring strategy using micropalaeontological and geochemical analyses of sediment cores is a promising tool for quantifying in situ temporal changes in ecological status from
reference conditions (here defined as pre-industrial status of the
19th century) through heavily impacted to improved conditions.
The present pilot study demonstrates temporal changes in environmental quality at two sites in inner Oslofjord, Norway, using
the Norwegian Pollution Control Authority’s (SFT’s) classification
system for fjords and coastal waters. The novel effort used here
is that the classification system based on macrofaunal species
diversity has been applied on fossil benthic foraminiferal assemblages. This is done in an effort to harmonise the ecological information held by benthic foraminifera with those of the macrofauna
which traditionally are used in environmental monitoring. Such
harmonisation makes it possible to use established classification
systems to characterise temporal changes in environmental quality beyond time intervals covered by biological time-series. At a
moderately polluted site, anthropogenic forcing through the
1900s caused a dramatic reduction in species diversity as compared to the pre-industrial reference condition of the early
1800s. Following a 7–8 times increase in nitrogen and phosphorous supply between 1930 and 1970 extensive eutrophication occurred and the absolute abundance of benthic foraminifera
increased with an order of magnitude. At a heavily impacted site,
the ecological status had improved significantly 2 years after capping with about 3 cm clean clay. A micropalaeontological monitoring strategy requires a net sediment accumulation area, that
taphonomic processes (e.g., mixing of sediment, loss and gain of
species) are considered, and optimal biological and ecological
knowledge of the microfossils used.
Acknowledgements
We are grateful to the crew on FF Trygve Braarud, and particularly the captian Sindre Holm for assistance during the cruise.
We also want to thank Mufak Naoroz, Anne Blaasvr, Lena Evensen, and Aina Marie Nordskog for assistance on the cruise and in
the lab, and Vincent M.P. Bouchet for useful discussions and constructive comments on the manuscript. Constructive comments
from the guest editors, Elena Romano and Luisa Bergamin, and
two anonymous reviewers are also appreciated.
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E. Alve et al. / Marine Pollution Bulletin 59 (2009) 297–310
Appendix. Faunal reference list
Generic classification follows Loeblich and Tappan (1987). The
original descriptions can be found in the Ellis and Messina world
catalogue of foraminiferal species on www.micropress.org. The
taxa are listed alphabetically.
Bolivinellina pseudopunctata (Höglund) = Bolivina pseudopunctata
Höglund, 1947.
Bulimina marginata d’Orbigny, 1826.
Cassidulina laevigata d’Orbigny, 1826.
Elphidium albiumbilicatum (Weiss) = Nonion pauciloculum Cushman subsp. albiumbilicatum Weiss, 1954.
Epistominella vitrea Parker, 1953.
Hyalinea balthica (Schröter) = Nautilus balthicus Schröter, 1783.
Miliolinella subrotunda (Montagu) = Vermiculum subrotundum
Montagu, 1803.
Nonionella auricula Heron-Allen and Earland, 1930.
Nonionella iridea Heron-Allen and Earland, 1932.
Nonionellina labradorica (Dawson) = Nonionina labradorica
Dawson, 1860.
Stainforthia fusiformis (Williamson) = Bulimina pupoides d’Orbigny var. fusiformis Williamson, 1858.
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