PAPER
Michael Krachler,*a James Zheng,ab Roy Koerner,b Christian Zdanowicz,b David Fisherb
and William Shotyka
www.rsc.org/jem
Increasing atmospheric antimony contamination in the northern
hemisphere: snow and ice evidence from Devon Island,
Arctic Canadaw
a
Institute of Environmental Geochemistry, University of Heidelberg, Im Neuenheimer Feld 236,
69120 Heidelberg, Germany. E-mail: [email protected];
Fax: þ49 62 21 54 52 28; Tel: þ49 62 21 54 48 48; web: http://www.uni-heidelberg.de/
institute/fak12/ugc/mkrachler/krachler.htm
b
GSC Northern Canada, Geological Survey of Canada, Ottawa, Canada
Received 1st July 2005, Accepted 4th October 2005
First published as an Advance Article on the web 19th October 2005
Adopting recently developed clean laboratory techniques, antimony (Sb) and scandium (Sc) deposition were
measured in a 63.72 m-long ice core (1842–1996) and a 5 m deep snow pit (1994–2004) collected on Devon
Island, Canadian High Arctic. Antimony concentrations ranged from 0.07 to 108 pg g1 with a median of
0.98 pg g1 (N ¼ 510). Scandium, used as a conservative reference element, revealed that dust inputs were
effectively constant during the last 160 years. The atmospheric Sb signal preserved in the ice core reflects
contamination from industrialisation, the economic boom which followed WWII, as well as the
comparatively recent introduction of flue gas filter technologies and emission reduction efforts. Natural
contributions to the total Sb inventory are negligible, meaning that anthropogenic emissions have dominated
atmospheric Sb deposition throughout the entire period. The seasonal resolution of the snow pit showed that
aerosols deposited during the Arctic winter, when air masses are derived mainly from Eurasia, show the
greatest Sb concentrations. Deposition during summer, when air masses come mainly from North America,
is still enriched in Sb, but less so. Snow and ice provide unambiguous evidence that enrichments of Sb in
Arctic air have increased 50% during the past three decades, with two-thirds being deposited during winter.
Most Sb is produced in Asia, primarily from Sb sulfides such as stibnite (Sb2S3), but also as a by-product of
lead and copper smelting. In addition there is a growing worldwide use of Sb in automobile brake pads,
plastics and flame retardants. In contrast to Pb which has gone into decline during the same interval because
of the gradual elimination of gasoline lead additives, the enrichments of Sb have been increasing and today
clearly exceed those of Pb. Given that the toxicity of Sb is comparable to that of Pb, Sb has now replaced
Pb in the rank of potentially toxic trace metals in the Arctic atmosphere.
Aim of the investigation
DOI: 10.1039/b509373b
Metals are ubiquitous in our modern world, and toxic metals
such as Cd and Pb ranking high among major pollutants
Michael Krachler was born in Austria,
in 1967. He received both his M.Sc. in
chemistry in 1993 and a PhD in analytical chemistry from the University of
Graz, Austria in 1997. In 1998 he joined
the Research Centre of Juelich, Germany as a Marie Curie Fellow funded
by the European Commission. Since
2000 he has been a Research Associate
at the University of Heidelberg and is
responsible for the new trace element and ICP-MS laboratories.
His current research interests centre on applications of ICPsector field-MS for the determination of trace elements and lead
isotopes in various environmental compartments.
w This work was presented at the First International Workshop on
Antimony in the Environment, Heidelberg, Germany, 16th to 19th
May 2005.
worldwide. In order to mitigate the ecological and toxicological impact of the metals, better knowledge is required of their
sources, pathways, mobility and fate in the surficial environment. A challenge that commonly faces environmental researchers is to distinguish between lithogenic and
anthropogenic contributions of metal accumulation. While
industrial emissions of many trace elements have declined in
recent decades (e.g. Pb), they have done so from levels which
were extremely high, relative to natural, background deposition rates. There are still major gaps in our knowledge of
natural sources of atmospheric trace metals, and their variability at different time scales. In that context, ice cores are
commonly exploited as valuable environmental archives of
atmospheric metal deposition.1–10
Because antimony (Sb) is present in the Earth’s crust at
concentrations below 1 mg kg1,11 there have been few quantitative studies of the geochemical cycle of this potentially toxic
element.12 Recognizing that Sb2O3 is a suspected carcinogen,
the European Commission and the US Environmental Protection Agency have listed Sb and its compounds as ‘‘priority
pollutants’’.13,14 With the progressive removal of asbestos from
brake pads in the 1980’s and its partial replacement by Sb2S3,
today Sb is the single most highly enriched element in urban
dusts.15 Studies have shown that Sb in plants growing beside
motorways and in urban areas is almost exclusively related to
This journal is & The Royal Society of Chemistry 2005
J. Environ. Monit., 2005, 7, 1169–1176
1169
vehicular traffic and is found mainly in the thoracic particle
fraction (i.e., 10–2.5 mm) and in the respirable fraction (i.e.,
o2.5 mm).16–18 These fine dust particles, released to the atmosphere through abrasion from Sb containing brake pads, show
the greatest Sb enrichments.16
Worldwide, two-thirds of total Sb production is used in the
manufacture of flame retardants (as Sb2O3,) for plastics and
textiles, most of which eventually ends up in a non-recyclable
waste stream. In addition, Sb2O3 is used in the catalysis of
PET, and the incineration of this material and other Sbcontaining plastics (e.g. PVC) yields aerosols with an average
particle diameter below 1 mm.19
Measurements of Sb in peat cores from rain-fed bogs in
Switzerland, Scotland, and the Shetland Islands have shown
that the chronology and intensity of atmospheric Sb contamination in Europe since the Roman period is similar to that of
Pb.15 In other words, Sb data from peat core analyses reveal
that (i) atmospheric Sb contamination has a history of at least
2000 years and (ii) Sb was subjected to long-range transport
which extended contamination even to remote regions of
Europe. However, the extent of Sb contamination remains
unclear and the possible importance of Sb as a global pollutant
is not yet established. While many of the trace metals of
contemporary environmental interest show declining concentrations and enrichments in recent layers of polar ice,1,5 there is
very little Sb data available for comparison.8 To elucidate the
potential relevance of Sb as a global pollutant, investigations
of atmospheric Sb in remote locations such as the Arctic are
needed.
Adopting clean laboratory procedures, which we have developed and continue to improve,20–22 a total of 510 age-dated ice
and snow samples from Devon Island, Arctic Canada were
analysed for Sb, scandium (Sc) and Pb using inductively
coupled plasma sector field mass spectrometry (ICP-SMS).
The detection limits which we have obtained for Sb (30 fg
g1) and Sc (5 fg g1) have allowed these elements to be
measured in polar ice for the first time. Using Sc as a
conservative, lithogenic tracer of natural mineral dusts, we
employ the Sb/Sc ratio of crustal rocks, soils, and pre-anthropogenic aerosols from a peat bog (6000 to 9000 years old) to try
to determine the extent to which Sb is enriched in the Arctic
atmosphere, and how these enrichments have evolved during
the past ca. 160 years. Our main objective is to test the
hypothesis presented earlier that Sb, like Pb, is a global atmospheric contaminant.15 In addition, using 45 samples from a pit
which has accumulated 5 m of snow during the past ten years
(1994 to 2004), we are able to examine and illustrate the
pronounced seasonal variations in atmospheric Sb deposition.
Meteorological data is then used to attribute the modern inputs
to predominant source areas. Using available production
statistics, the most likely sources of atmospheric Sb contamination are identified.
Experimental
USA), a high efficiency sample introduction system (Apex Q,
Elemental Scientific Inc., Omaha, NE, USA) and a sapphire
injector tube were employed to transport the analytes into the
plasma of the ICP-SMS. The autosampler and the Apex
sample introduction system were hosted in a class 100 laminar
flow bench. Details of all analytical procedures including ICPSMS operating conditions and the data acquisition parameters
have been reported earlier.20–22
Reagents and standards
For the preparation of all solutions, high purity water (18.2
MO cm) from a MilliQ-Element system (Millipore, Milford,
MA, USA) designed for ultra trace analysis was used. Nitric
acid (65%, analytical-reagent grade, Merck, Darmstadt, Germany) was further purified twice by distillation, using a high
purity quartz unit for sub-boiling of acids. (MLS GmbH,
Leutkirch, Germany). Both the water purification system and
the sub-boiling distillation unit were operated in clean rooms.
Calibration solutions for Sb, Sc and Pb were prepared daily
by appropriate dilution of 10 mg l1 stock standard solutions
(Merck) with 0.5% (v/v) high-purity nitric acid. Quantification
of trace element concentrations was performed by linear
regression of the calibration curves.
Collection of ice and snow samples and sample treatment
A 63.72-m long firn core (D2000) was drilled in May 2000 from
the Devon ice cap (Fig. 1, Devon Island; 751 N; 821 W; 1860 m
asl). The drilling site presently experiences a mean annual
surface temperature of 23 1C, an accumulation rate of
0.28 m year1 (ice equivalent based on 40 year mass balance
records) and 20% summer melt (mass). To minimize potential
metal contamination of the core, the electro-mechanical auger
was modified by replacing steel and aluminium components of
the sonde with titanium and polyethylene parts. The modified
auger was so designed that the core was directly drilled into an
acid-cleaned, high-density polyethylene (HDPE) tube inserted
in the drill barrel. Once a core was drilled, the tube was
removed, capped and sealed in a polyethylene bag. There
was no other contact with the cores except occasionally, using
gold-plated tongs. Detailed information on the new clean drill
made from high purity titanium can be found elsewhere.23,24
To evaluate deposition trends since the core was drilled, a
snow pit was dug on Devon Island in the spring of 2004. The
5 m snow pit representing precipitation from 1994 to 2004 was
dug by hand using stainless steel shovels. The sampling pit wall
was further cleaned using titanium chisels right before sampling. The purpose of the pit was to evaluate the recent metal
pollution trends in the Arctic since the D2000 core was drilled.
Personnel working in the pit wore Tyvek clean-room jackets
during the digging and sampling. Samples were taken from the
pit in intervals ranging from 6.4 to 19.1 cm for firn layers. The
top sample (fresh snow) was taken from outside of the pit
(about 1 cm thickness of snow collected at approximately 10 m
Laboratories and instrumentation
To minimize the potential risk of contamination, all sample
handling and the preparation of all standards were performed
in clean rooms under laminar flow clean air benches of at least
class 100. Decontamination of the ice samples was carried out
in a certified cold clean room of class 100. The ice and snow
samples were melted, acidified with high purity HNO3
(68–72%) to 0.5% (v/v) and bottled in a class 10 clean air
cabinet at the Geological Survey of Canada (GSC) or at the
University of Heidelberg.
All ICP-MS measurements were carried out with an Element2 ICP-SMS (Thermo Electron, Bremen, Germany) operated in a class 1000 clean laboratory. A micro volume
autosampler (ASX 100, Cetac Technologies, Omaha, NE,
1170
J. Environ. Monit., 2005, 7, 1169–1176
Fig. 1 Map of a selected area of the Canadian High Arctic highlighting Devon Ice Cap where the ice core and the snow pit were
collected at 751 N; 821 W; 1860 m asl in the years 2000 and 2004,
respectively.
in front of the pit). All samples were taken with a pre-cleaned
titanium chisel and a pre-cleaned HDPE scoop. Before each
sampling, the titanium chisel and scoop were further cleaned
by pushing in and out of snow on the sidewall in the pit to
remove any possible dusts on their surfaces. Samples were
directly taken into 500 ml pre-cleaned HDPE bottles, which
were kept capped all the time before and after sampling.
Cleaning procedures for bottles and utensils were as described
by Zheng et al.24 A total of 45 samples were taken from the pit
and they were shipped frozen to the University of Heidelberg
laboratories for further processing and analyses. Clean chambers used for this processing are better than class 10 and those
clean chambers are located in a metal-free class 100 clean
room. Right before analysis, the samples were acidified with
high purity nitric acid when they were half-melted and then
transferred to autosampler cups with pre-cleaned HDPE bottles. To avoid localised leaching from beaker walls during
sample acidification/melting and for convenience, a 1 þ 1
dilution of concentrated nitric acid with high purity water
was used. Therefore, the acidification of ice samples was
accomplished by adding 0.08 ml of 1 þ 1 diluted high purity
nitric acid per every 10 ml of ice sample, which resulted in a pH
of the melted ice samples of approximately 1.2. Ice samples
were decontaminated manually and processed similarly in the
clean laboratories in Ottawa.24
The cleanliness of the Ti drill and effectiveness of the
decontamination procedure of the ice core sections was investigated in detail demonstrating that the Ti drill ‘‘yields’’ ice
cores that are distinctly less contaminated than those obtained
by other types of conventional drills used worldwide. Additionally an inter-laboratory study revealed that leaching of
trace elements from the storage containers is well under control
and that procedural blanks are so low that they do not
measurably affect trace element concentrations in the ice
samples.24
The chronology for the D2000 core was derived from
another, deeper core (300-m long) drilled in 1998 (D1998)
from the same site. The depth-age relationship is based on
annual layer counting using seasonal variations in d18O and
major ions (Na1, SO42 and NO3), and on distinctive time
markers such as SO42 and acidity peaks related to the Laki
(Iceland; 1783 AD), Tambora (Indonesia; 1816 AD) and)
Katmai (Alaska; 1912 AD) volcanic eruptions. Correlation
between the 1998 and 2000 cores was made using the 1958
(16.5 m in real depth below surface) and 1963 (13.5 meters)
radioactive horizons as well as the 1998 age dating model.
Based on this correlation, the age of the ice at the bottom of
D2000 core (63.72 m) corresponds to AD 1842. The dating
error for first top 20 m is estimated to be 1 year; 2 years for
depths between 20 and 40 m and 5 years for the remainder of
the core. Depth and age relationships for the pit were developed by examination of ice layers and the pit stratigraphy. The
dating error for the pit is estimated to be less than six months.
Quality control
As no certified reference material for trace elements in polar ice
is currently available, a riverine water reference material
(SLRS-4, National Research Council Canada, Ottawa, Canada) with lowest available Sb, Sc and Pb concentrations was
used for quality control purposes. Good agreement between
the experimentally established and certified (Pb, Sb) and
reported (Sc) concentrations was established.
Results and discussion
Quality control and detection limits
In contrast to our previous work,20–22 a high efficiency sample
introduction system (Apex Q) was utilized for trace element
analyses during this study. Using the Apex, the analytes are
concentrated via desolvation of the sample aerosol which is
initially heated to 140 1C followed by cooling to 2 1C before
entering the plasma. The increase in sensitivity with the Apex
compared to our conventional sample introduction system was
approximately six-fold and helped to lower the detection limit
(LOD) for Sc from 0.026 pg g1 to 0.005 pg g1.22 Detection
limits were based on the 3s criterion from the measurement of
16 independently prepared blank solutions containing 0.5%
nitric acid and high purity water. For Sb the LOD was
improved from 0.07 to 0.03 pg g1. The higher improvement
of LODs for Sc clearly indicates that the LOD for Sc is limited
by counting statistics whereas the LOD for Sb is mainly
determined by its blank, i.e. contamination level. The LOD
for Sb is lower by one order of magnitude compared to
previous ice/snow studies25 and here, Sb concentrations were
greater than the LOD for all samples. Acid blanks containing
0.5% HNO3 and high purity water amounted to 7.0 1.6 fg
g1 and 43 10 fg g1 for Sc and Sb, respectively.
The riverine reference water material SLRS-4 with certified
concentrations for Sb (230 40 pg g1) and Pb (86 7 pg g1)
was analyzed at regular intervals during ice analysis. The
experimentally determined Sb (239 18 pg g1; N ¼ 40) and
Pb (77 10 pg g1; N ¼ 40) concentrations agreed well with
the certified values. For Sc, no adequate certified water reference material is currently available, but previously obtained
values (11.3 0.6 pg g1),22 were reproduced in this study
(10.3 0.9 pg g1, N ¼ 38).
Abundance and distribution of Sb and Sc
The concentrations of Sb in the investigated ice and snow
samples (N ¼ 510) ranged from 0.07 to 108 pg g1, with a
median of 0.98 pg g1 (Table 1). These values are similar to
those determined in 68 Greenland snow samples representing
the period of 1990–1995 (range: 0.21–4.3 pg g1, median 0.72
pg g1),26 but are at least an order of magnitude lower than Sb
concentrations found in snow and ice samples in the European
Alps.8
The ratio of maximum to minimum Sb concentrations in the
present study exceeds a factor of 1500. Even though the Sb
concentration pattern shows remarkable variation, a detailed
chronology of past atmospheric Sb could be extracted from the
entire data set either by calculating a running median or
running average and using intervals of 50 data points (Fig.
2A). On plotting ice core data for trace metals, the conventional approach has been to use the running average to smooth
such a fluctuating data set.8 The data shown in Fig. 2A
illustrates how data processing can profoundly affect the
temporal trend. The Sb concentrations obtained by averaging
are dominated by 21 samples out of 510 containing Sb concentrations in the range of 5–108 pg g1. With such ‘‘outliers’’
in the data set containing extraordinary Sb concentrations, the
running average overestimates the Sb concentration profile,
relative to the smoothed concentration pattern (Fig. 2A). The
increase in Sb concentrations obtained between 1860 and 1880
A.D. calculated using a running average, for example, is an
artefact caused by a single ice sample containing 108 pg Sb g1.
Table 1 Summary of Sb and Sc concentrations (pg g1) of ice (N ¼
465) and snow (N ¼ 45) samples from Devon Island, Canadian High
Arctic
Element
Antimony
Devon 2000
Devon 2004
Scandium
Devon 2000
Devon 2004
Minimum
Maximum
Median
firn core (1842–1996)
snow pit (1994–2004)
0.07
0.13
108
3.71
0.98
1.03
firn core (1842–1996)
snow pit (1994–2004)
0.02
0.08
8.80
1.64
0.47
0.31
J. Environ. Monit., 2005, 7, 1169–1176
1171
Fig. 2 (A) Influence of data processing on the chronology of 155 years
of atmospheric Sb and Sc as revealed by the Devon 2000 ice core and
snow pit. Solid lines represent 50 point moving averages or medians,
i.e. reflecting approximately five years of snow accumulation. (B)
Chronology of antimony (Sb) and lead (Pb) in the Canadian Arctic.
Data represent measurements on 510 individual ice and snow samples,
i.e. B3 samples year1 on average. Median concentrations of Sb and
Pb together with their estimated natural concentrations are displayed.
In contrast, the running median reveals a much smaller increase in the Sb concentrations during the same period (Fig.
2A). Regardless of whether this high concentration is related to
a natural event or was caused by the sampling or processing of
the ice sample, clearly, the running median provides a much
more representative and conservative description of the longterm trends. Here, the running median is used exclusively
throughout this study.
Except for 4% of the samples with Sb concentrations 45 pg
g1, the remaining ice samples fluctuated only within a range of
a factor of 50 and this was mainly caused by seasonal variations (see below). It is important to note here that Sb concentrations greater than 5 pg g1 were confirmed by independent
ICP-MS measurements on the same sample aliquots at the
Geological Survey of Canada in Ottawa. As contamination of
the ice samples by Sb during the entire decontamination
procedure is very unlikely, these high concentrations can be
possibly explained by the relatively high resolution in time (on
average more than three samples per year) of the ice samples.
Therefore one ice sample represents only a part of the year and
thus Sb concentrations may vary largely due to seasonal
variations (see below). During periods of extended dry deposition (mainly in the winter season), for example, dust particles
deposited onto the snow surface may be responsible for an ice
sample with a distinctly elevated Sb concentration. Assuming a
particle diameter of 1 mm, a particle density of 2.5 g cm3, and
a Sb concentration of 10 mg kg1, such a single particle, in an
ice volume of 10 ml, increases the actual Sb concentration by
B0.13 pg g1.
Interpretation of the Sb chronology
The smoothed (running median) Sb concentration profile
largely follows that of Pb (Fig. 2B), with both elements
1172
J. Environ. Monit., 2005, 7, 1169–1176
showing elevated concentrations during the 20th century. In
fact, the geochemical and mineralogical association of Sb with
Pb minerals implies that the temporal trend of both elements
should evolve more or less in parallel, largely reflecting Pb
smelting.12 As chalcophile elements, both Sb and Pb are
commonly enriched in coal, and because fossil fuel combustion
appears to be the largest single source of anthropogenic Sb
(B50% of total Sb emissions) to the global atmosphere,27 the
history and intensity of Sb and Pb emissions to the environment also are both linked to the combustion of fossil fuels in
the late 19th and early 20th century.
In the oldest ice samples (B1840–1860 AD), Sb concentrations were o0.5 pg g1 (Fig. 2B). Starting from approximately
1900 AD, Sb concentrations increased rapidly, probably reflecting the growing importance of anthropogenic activities
such as coal burning, mining and smelting of Pb and Cu ores,
reaching a plateau at B1.4 pg g1 between 1910 and 1940 AD.
In contrast to Sb, the relative increase of Pb concentrations was
much smaller during the same period. The low melting point
(550 1C) of stibnite (Sb2S3) relative to galena (PbS, 1114 1C) is
a possible explanation for these differences. In other words,
during the combustion of metal sulfides, whether from ores or
in coal, larger amounts of Sb relative to Pb are emitted to the
atmosphere. Leaded gasoline, an additional source of Pb, is
recorded by the ice core since 1930 (Fig. 2B).
Antimony concentrations reached maximum values of B2
pg g1 in the late 1950s reflecting the economic boom after
WWII. Since then Sb concentrations in the snow declined,
reaching B1.2 pg g1 during the last decade. We assume that
the pronounced decrease in Sb emissions is mainly related to
the use of dedicated filter technologies in coal fired power
plants that are still widely used today and the emission reduction technologies used by modern smelters. The decline in
gasoline lead consumption starting in the 1970s allowed Pb
concentrations to fall by 80%. In contrast, Sb concentrations
declined only by 50% during the last two to three decades. The
comparison of Sb/Sc with Pb/Sc from 1840 to 2000 shows that
Sb was more enriched than Pb during ten of the past fifteen
decades, including the last two (Fig. 3B). The decades characterised by Pb enrichments exceeding those of Sb are the ones
corresponding to the greatest use of leaded gasoline (ca. 1945
to 1990). Moreover, when both Sb/Sc and Pb/Sc went into
decline (ca. 1960), Sb/Sc declined more rapidly: this may reflect
the effectiveness of filtration technologies reducing Sb and Pb
emissions from point sources. While this probably had a
comparable effect on both metals, the ongoing use of leaded
gasoline reduced the rate of the decline in Pb/Sc.
To put the Sb concentrations obtained from the ice core into
perspective, the ‘‘natural, lithogenic’’ Sb concentration (Sbnat)
has been calculated as follows:
[Sb]nat ¼ [Sc]sample([Sb]/[Sc])UCC
(1)
where [Sc]sample refers to the corresponding concentration in
the ice or snow samples while [Sb]UCC and [Sc]UCC indicate the
average Sb and Sc concentrations of the Upper Continental
Crust (UCC).11 The [Sb]nat and [Pb]nat values (B0.02 pg g1
and B1 pg g1, respectively) are very low and this shows
clearly that the chronologies of both elements have been
profoundly impacted by anthropogenic emissions throughout
the entire record (Fig. 2B). In other words, the natural
contribution to the total Sb concentrations was negligible
during the last 160 years. Therefore, increasing Sb concentrations found during particular periods cannot be explained by
contributions from natural sources and processes.
In contrast to the highly variable Sb data, Sc concentrations—that are used as a proxy of mineral dust input28—reveal
no distinct temporal trend (Fig. 2A). Except for three periods
of slightly elevated Sc inputs, mineral dust deposition was
effectively constant over the entire period represented by the
ice core and the snow pit (ca. 160 years). Aerosols from coal
burning, but perhaps also from soil dust, erosion, deforestation, agriculture and construction, are possible sources of the
increased Sc concentrations found during these three periods.
It remains to elucidated, however, which were the predominant
causes of the increased Sc deposition. Similar to Sb, the scatter
in the Sc values (0.02–8.80 pg g1) in the ice core can be
attributed to seasonal changes (see below).
Enrichment factors for Sb
To assess the enrichment or depletion of Sb concentrations
over time relative to the UCC, Sb enrichment factors (Sb EF)
were calculated as
Sb EF ¼ ([Sb]/[Sc])sample/([Sb]/[Sc])UCC
(2)
It is important to note here that this concept to calculate
elemental EF is a good approach to highlight relative changes
in the abundance of elements derived from crustal rocks, but
the absolute value of the EF should be interpreted with
caution.29 Until site-specific natural background values are
established,30 corresponding values from the UCC can be used
as an estimate for the calculation of elemental EF. Absolute
values for EF, however, vary distinctly depending on which
reference element is used for the normalization and which
published values for the UCC are considered. In addition,
average values of the UCC are not necessarily representative of
the coring site.
Although the absolute values of the EF certainly depend on
the ‘‘background’’ Sb/Sc ratio employed, the evolution of the
temporal trend in the EF is identical regardless of the reference
level used. The pronounced impact of normalizing Sb/Sc
concentrations in the ice samples to their corresponding values
in the upper or average continental crust is shown in Fig. 3A.11
Using these values, Sb EF as great as B110 and B260,
respectively, are seen (Fig. 3A). Considering the natural background levels of Sb (8 ng g1) and Sc (80 ng g1) that have been
established in natural dusts deposited in a peat profile from the
Jura Mountains, Switzerland between ca. 9000 to 6000 calendar years BP, the maximum Sb EF amounts only to B50 (Fig.
3A).30 Another alternative would be to use the Sb/Sc ratio
typical of soils, which is more representative of soil-derived
aerosols than the corresponding values for crustal rocks
(1/7).31 This approach yields a maximum Sb EF of B40
(Fig. 3A). It is important to note here, that regardless of the
reference levels used, Sb is markedly enriched in all samples,
but especially so in those corresponding to the 20th century.
The concept for calculating elemental EF is based on the
assumption that an appropriate reference level of the considered elements is well established and that the reference element
used for normalisation is only derived either from one major
source, i.e. weathering of rocks on the continental crust or
marine aerosols, etc. The calculations made above, however,
demonstrate the weakness and limitations of this concept.
Background values for Sb and Sc on Devon Island have not
yet been established. In general we prefer to use Sc as a
conservative reference element, because in contrast to other
lithogenic elements such as Al, Ti or Zr, Sc has no industrial
use, shows no preference for specific mineral phases and is
rather uniformly distributed among the dominant minerals of
the Earth’s crust. Given the above mentioned difficulties in
obtaining reliable, absolute Sb EF, it is obvious that sitespecific background values, which can only be obtained by
analyzing older ice from the same site, and their temporal
variation are urgently needed.
For the time being, however, rather than calculate an
absolute value for the enrichment, we avoid this altogether
and simply present the Sb/Sc ratio to document its temporal
evolution (Fig. 3B). For a comparison to Sb, the Pb/Sc ratio is
also plotted in Fig. 3B against time. As the Sc deposition to the
Fig. 3 (A) Influence of background concentrations used for normalization on absolute values of Sb EF. See text for details. EGR ¼ Etang
de la Gruère, a peat bog in the Jura Mountains, Switzerland. (B)
Medians of Sb/Sc and Pb/Sc ratios, respectively, reflecting the temporal
change of the enrichment of Sb and Pb. The ellipse is included to guide
the eye to the changes seen in Sb/Sc since ca. 1970, while the horizontal
line displays the Sb/Sc and Pb/Sc ratios in the Upper Continental Crust
(UCC). The differences in scales of the y axes are proportional to the
differences (55) in crustal abundance of the two elements (Pb 17 mg
kg1, Sb 0.31 mg kg1).
snow surface was roughly constant during the last 160 yrs, the
evolution of both the Sb/Sc and Pb/Sc ratios provides a
common basis to compare Sb and Pb.
The most striking feature of this plot, however, is the 50%
increase in Sb enrichments during the last three decades. While
modern state-of-the-art waste incineration plants have reduced
atmospheric emissions of Sb, 670 tonnes Sb year1 are still
estimated to be released with stack gases to the atmosphere
world-wide.32 It is worth mentioning here, that Sb emissions
from waste incineration account for 19% of the total trace
element emissions.32 In the European Union the emission
standard for Sb is part of a sum parameter including As, Co,
Cr, Cu, Mn, Ni, Pb, Sn, and V and the total of these may not
exceed 0.5 mg m3 for municipal solid waste incineration.33
Antimony together with As, Cd and Pb belongs to the so-called
Class II elements that are vaporised during combustion but
after condensation are found mainly in the fly ashes on
particulates. We note that a significant part of these fine
particles are in the sub-micrometre size class where dust
control systems are less effective.33
In addition, the worldwide annual production of Sb has
almost doubled from B70 000 ton in the late 1960’s to
B120 000 ton in the year 2000 (J. O. Nriagu, personal communication). In the year 2003 global Sb production amounted
to 142 000 ton per year, with most produced by China (88%),
South Africa (4%), Russia (3%), Tajikistan (2%), and Bolivia
(2%).12 With the increased use of Sb in various kinds of plastic
(PVC, PET, etc.) and the increased waste incineration of plastic
during the last 30 years, aerosols with an average particle
diameter below 1 mm18 are continuously emitted and are then
subjected to long-range transport. With an atmospheric residence time (1–2 weeks) comparable to that of Pb, long-range
J. Environ. Monit., 2005, 7, 1169–1176
1173
transport (several thousands of kilometres) of Sb from industrial and urban centres has clearly made its mark on the Arctic
atmosphere (Fig. 3B). With the magnitude of Sb enrichments
found in remote regions such as the Arctic today, one can
easily imagine the extent of Sb pollution in regions where the
corresponding point sources are located. For megacities such
as Tokyo, for example, Sb EF of B21 000 have been determined in the o0.2 mm size fraction of airborne particulate
matter continuously collected between 1995 and 2004.34
Seasonal variations of Sb and mineral dust depositions
Because of the excellent temporal resolution (45 samples
representing ten years), the investigation of the snow pit makes
it possible to study changing deposition rates due to seasonal
variations in air mass sources, trajectories, and chemistry. Fig.
4A highlights the seasonal and annual changes of both Sb and
Sc concentrations in the snow pit for the last decade. Data
points in Fig. 4 marked with grey bars indicate the summer of
Fig. 4 (A) Temporal trend and seasonal variations of Sb and Sc
concentrations (pg g1) as revealed by the 5 m snow pit collected on
Devon Island in 2004. Grey bars indicate the summer of the respective
year. (B) Influence of the reference level used for normalisation on the
measured Sc data in the ice and snow samples on the absolute values of
natural Sb. (C) Predominance of excess Sb during the last ten years
revealing maximum anthropogenic Sb deposited in winter.
1174
J. Environ. Monit., 2005, 7, 1169–1176
the specific year which has been determined using major ion
chemistry and snow stratigraphy.
Median concentrations of both Sb and Sc in the 5 m snow pit
(representing deposition between ca. 1994–2004) are similar to
the corresponding values in the ice core (Table 1). Antimony
concentrations (N ¼ 45) ranged from 0.13 to 3.71 pg g1 with a
median value of 1.03 pg g1.
Numerous observations at arctic locations have shown that
pollution levels during late winter and early spring can reach
concentrations that are comparable to those observed at
polluted mid-latitude locations.35 These so-called ‘‘Arctic haze
episodes’’ occur predominantly between December and April
with haze being comprised of particles no larger than 2 mm.36,37
Meteorological studies indicate that air flow into the Arctic
during winter is predominantly from Eurasia, and rarely from
North America.35 The strong transport into the Arctic from
Eurasia during winter is caused by the presence of the climatologically persistent Siberian high pressure region resulting in
a surge of polluted European air into the Arctic.37 For the rest
of the year, particulate pollution is either absent, or present at
much lower concentrations than in winter.37
A closer look at the snow pit data reveals that Sb concentrations are lowest during summer when the air masses reaching the Canadian Arctic mainly originate from North
America.35–37 The lowest Sb concentration (0.1 pg g1) was
established in the snow sample from the summer of 2003. In
contrast, maximum Sb values are often found in snow samples
representing the winter seasons with air masses in the Canadian
Arctic predominantly arriving from Europe and northern Asia
(i.e. Eurasia).35–37 Calculating a Sb inventory for the last ten
years reveals that B2/3 of the total Sb was deposited on Devon
Island in the winter season and only B1/3 in the summer
season.
Anthropogenic emissions of Sb from the emerging countries
of Asia are believed to dominate global emissions27 and the
snow pit data supports this interpretation. For example, Asian
Sb emissions from primary Cu production (159 tonnes year1)
account for 50% of the worldwide emissions in this category.27
Similarly, Asian primary lead production emits 64% (86 tonnes
year1) and primary zinc production 41% (39 tonnes year1)
of global anthropogenic Sb in this category.27 As mentioned
earlier, municipal waste incineration of various Sb-containing
plastics also contributes to Sb emissions. A report on Sb
emissions from waste incineration from the mid-1990s estimated European (33%) and Asian (42%) contributions to
dominate worldwide Sb emissions (235 tonnes yr1); this can
at least partially explain the high Sb concentrations found in
winter snow samples in the snow pit.27 Total global anthropogenic Sb emissions (1561 tonnes yr1) are clearly dominated
by contributions from Asia (44%), followed by North America
(24%), Europe (17%) and South America (6.5%).27 Ignoring
the Sb contributions from the southern hemisphere, the winter
inventory of Sb in the snow pit (2/3 of the total) is remarkably
consistent with European and Asian contributions to total
global emissions estimated by Pacyna and Pacyna,27 i.e. 61%.
The Sc concentration profile of the snow pit also shows
greater concentrations in winter (Fig. 4A). Scandium, a proxy
for mineral dust input to the snow pit, thus highlights that the
winter season is dustier—possibly reflecting coal burning—
than the summer seasons. Like the ice core, the snow pit data
reveals no temporal trend of dust inputs.
Following eqn (1), the natural Sb concentration in the snow
pit was calculated using several reference levels (Fig. 4B). The
lowest natural Sb concentrations were obtained using the Sb/
Sc ratio of the average continental crust while the soil Sb/Sc
ratio31 yielded the greatest natural concentrations. Even
though absolute values of the naturally derived, lithogenic Sb
vary distinctly depending on the compartment used for normalization, the contribution of natural Sb to the total Sb
inventory is always negligible (Fig. 4A and B). To unequivo-
atmospheric Sb in the Arctic might have broader implications
for human and ecosystem health worldwide.
Acknowledgements
This work was supported by the European Commission
(MIF1-CT-2005-008086) and the Forschungspool of the University of Heidelberg (Project: ‘‘Trace element analyses of ice
cores for reconstruction of paleoclimate and human impact’’).
Additional financial and logistic support was received from
Terrain Science Division and Metal in the Environment Program, Geological Survey of Canada. J. Z. thanks Dr. Susan
Pullan for her advice and support.
Fig. 5 Average Sb EF and Pb EF for the period of 1994 to 2004
calculated using different reference levels. See text for details.
References
1
cally establish natural Sb concentrations, however, older, uncontaminated ice samples need to be analysed. This work will
be carried out in the near future and this will also allow the
calculation of absolute enrichment factors using site-specific
natural background values.
The excess amount of Sb ([Sb]excess) can be easily calculated as
2
3
4
5
[Sb]excess ¼ [Sb]total [Sb]nat
(3)
6
where [Sb]total refers to the total Sb concentration in the ice
sample. This excess amount of Sb cannot be explained by
mineral dust deposition related to weathering of rocks, for
example, but rather reflects anthropogenic activities. Considering the concentration values of Sb and Sc of the UCC, [Sb]excess
was plotted in Fig. 4C which highlights three important
features. First, [Sb]excess is by far the predominant Sb component in the snow pit. Second, because [Sb]nat is negligible, the
total Sb concentration is a reasonable approximation of
anthropogenic Sb contamination. Third, most anthropogenic
Sb arrives in the winter season in the Arctic, with lowest
contributions found during the summer months (Fig. 4C). It
is important to stress here once again that meteorological data
showed that air masses come from Asia and Europe during
winter months.35–37 If a remote region such as the Arctic,
thousands of kilometres away from the emission sources, is
remarkably contaminated by anthropogenic Sb, one can easily
imagine the extent of Sb contamination in the air in urban and
industrial areas in Eurasia.
To highlight the importance of the need to establish reliable
natural background concentrations and their variation with
time, the average Sb EF calculated using various reference
levels are plotted in Fig. 5 for the snow pit. The difference
between lowest and highest is B7.5, with the greatest Sb EF
(234) obtained using the average composition of the continental crust and the smallest (31) obtained using soil as reference.
Regardless of the kind of normalisation used, all calculations
demonstrate that the Arctic today is profoundly contaminated
with anthropogenic Sb. Less remote regions should be correspondingly more contaminated. Moreover, independent of the
reference level employed to calculate the EF, Sb clearly is now
more enriched in Arctic aerosols than Pb (Fig. 5).
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
Conclusions
The studies of peat bogs in Europe showed that anthropogenic
activities have impacted the geochemical cycle of Sb to the
same extent as that of Pb.12 The Arctic snow and ice data,
however, shows that human impacts on the Sb cycle are global
in extent and now exceed those of Pb. Given that Sb-bearing
aerosols from combustion processes are smaller than 1 mm,
primarily in the form of relatively soluble oxides, and that the
toxicity of Sb is comparable to that of Pb, the recent trend in
25
26
27
28
C. F. Boutron, U. Gorlach, J.-P. Candelone, M. A. Bolshov and
R. J. Delmas, Nature, 1991, 353, 153.
C. Barbante, A. Veysseyre, C. Ferrari, K. Van de Velde, C. Morel,
G. Capodaglio, P. Cescon, G. Scarponi and C. Boutron, Environ.
Sci. Technol., 2001, 35, 835.
C. Barbante, K. van de Velde, G. Cozzi, G. Capodaglio, P.
Cescon, F. Planchon, S. Hong, C. Ferrari and C. Boutron,
Environ. Sci. Technol., 2001, 35, 4026.
A. Matsumoto and T. K. Hinkley, Earth Planet. Sci. Lett., 2001,
186, 33.
J. R. McConnell, G. W. Lamorey and M. A. Hutterli, Geophys.
Res. Lett., 2002, 29, 45-1.
J. R. McConnell, G. W. Lamorey, S. W. Lambert and K. C.
Taylor, Environ. Sci. Technol., 2002, 36, 7.
C. Barbante, C. Boutron, A.-L. Moreau, C. Ferrari, K. van de
Velde, G. Cozzi, C. Turetta and P. Cescon, J. Environ. Monit.,
2002, 4, 960.
C. Barbante, M. Swikowski, T. Döring, H. W. Gäggeler, U.
Schotterer, L. Tobler, K. Van de Velde, C. Ferrari, G. Cozzi, A.
Turetta, K. Rosman, M. Bolshov, G. Capodaglio, P. Cescon and
C. Boutron, Environ. Sci. Technol., 2004, 38, 4085.
J. Zheng, C. Zdanowicz, D. Fisher, G. Hall and J. Vaive, J. Phys.
IV, 2003, 107, 1405.
P. Gabrielli, F. A. M. Planchon, S. Hong, K. H. Lee, S. D. Hur, C.
Barbante, C. P. Ferrari, J. R. Petit, V. Y. Lipenkov, P. Cescon and
C. F. Boutron, Earth Planet. Sci. Lett., 2005, 234, 249.
K. H. Wedepohl, Geochim. Cosmochim. Acta, 1995, 59, 1217.
W. Shotyk, M. Krachler and B. Chen, in Anthropogenic Impacts
on the Biogeochemistry and Cycling of Antimony in Metal Ions in
Biological Systems, ed. A. Sigel, H. Sigel and R. K. O. Sigel,
Taylor and Francis, Boca Raton, FL, vol. 44, pp.171–203.
US EPA, Water related fate of the 129 Priority Pollutants,
Washington DC, 1979, Vol. 1 EP-440/4-79-029A.
Council of European Union. Council Directive 98/83/EC of 3
November 1998, Quality of Water Intended for Human Consumption. Official Journal L 330, 05/12/1998, pp. 32–54.
W. Shotyk, M. Krachler and B. Chen, Global Biogeochem. Cycles,
2004, 18, GB1016:1.
G. Weckwerth, Atmos. Environ., 2001, 35, 5525.
C. Dietl, W. Reifenhäuser and L. Peichl, Sci. Total Environ., 1997,
205, 235.
M. Krachler, M. Burow and H. Emons, J. Environ. Monit., 1999,
1, 477.
J. O. Nriagu, Environment, 1990, 32, 8–28.
M. Krachler, J. Zheng, D. Fisher and W. Shotyk, Anal. Chem.,
2004, 76, 5510.
M. Krachler, J. Zheng, D. Fischer and W. Shotyk, J. Anal. At.
Spectrom., 2004, 19, 1017.
M. Krachler, J. Zheng, D. Fischer and W. Shotyk, Anal. Chim.
Acta, 2005, 530, 291.
http://thrust_2.tripod.com/glaciology/icedrill.html.
J. Zheng, J. Vaive, M. Krachler, J. Lam, G. Lawson, G. Hall, C.
Zdanowicz, D. Fisher and W. Shotyk, J. Environ. Monit., 2005,
submitted.
C. Barbante, G. Cozzi, G. Capodaglio, K. van de Velde, C.
Ferrari, C. Boutron and P. Cescon, J. Anal. At. Spectrom., 1999,
14, 1433.
C. Barbante, T. Bellomi, G. Mezzadri, P. Cescon, G. Scarponi,
C. Morel, S. Jay, K. Van de Velde, C. Ferrari and C. F. Boutron,
J. Anal. At. Spectrom., 1997, 12, 925.
J. M. Pacyna and E. G. Pacyna, Environ. Rev., 2001, 9, 269.
W. Shotyk, D. J. Weiss, J. D. Krames, R. Frei, A. K. Cheburbin,
M. Gloor and S. Reese, Geochim. Cosmochim. Acta, 2001, 65, 2337.
J. Environ. Monit., 2005, 7, 1169–1176
1175
29 K. A. Rahn, The Chemical Composition of the Atmospheric Aerosol, Technical report, University of Rhode Island, Kingston, RI,
1976.
30 W. Shotyk, M. Krachler, A. Martinez-Cortizas, A. K. Cheburkin
and H. Emons, Earth Planet. Sci. Lett., 2002, 199, 21.
31 H. J. M. Bowen, Environmental Chemistry of the Elements, Academic Press, New York, 1979.
32 D. Stanners and P. Bourdeau, Europe’s Environment. The
Dobris Assessment, European Environment Agency, Copenhagen,
1995.
1176
J. Environ. Monit., 2005, 7, 1169–1176
33 R. Zevenhoven and P. Kilpinen, Control of Pollutants in Flue
Gases and Fuel Gases, Helsinki University of Technology course
ENE-47.153, ch. 8, ISBN 951-22-5527-8, available at http://
users.tkk.fi/Brzevenho/gasbook.
34 N. Furuta, A. Iijima, A. Kambe, K. Sakai and K. Sato, J. Environ.
Monit., 2005DOI: 10.1039/b513988k.
35 W. T. Sturges, J. F. Hopper, L. A. Barrie and R. C. Schnell,
Atmos. Environ., Part A, 1993, 27, 2865.
36 W. T. Sturges and L. A. Barrie, Atmos. Environ., 1989, 23, 2513.
37 L. A. Barrie, Atmos. Environ., 1986, 20, 643.