Download Contaminants in the arctic marine environment

ICES Journal of Marine Science, 53: 537–563. 1996
Contaminants in the arctic marine environment: priorities for
protection
R. W. Macdonald and J. M. Bewers
Macdonald, R. W. and Bewers, J. M. 1996. Contaminants in the arctic marine
environment: priorities for protection. – ICES Journal of Marine Science, 53: 537–563.
We assess evidence for significant adverse effects on the arctic marine environment
associated with global and regional releases of chemical and radioactive contaminants
from human activities. The sources, mode of transport, and biological effects of
persistent organic compounds, lead, cadmium, mercury, artificial radionuclides, and
oil are reviewed. An outline of the physical and biogeochemical processes occurring
within the Arctic Ocean is used as a basis for identifying contaminant transport,
accumulation, and exposure pathways. Where significant anthropogenic impacts can
be inferred or suspected, we evaluate opportunities for human intervention through
the introduction of new source controls, or other restrictions, to reduce impacts. We
conclude that additional controls are justified only in the cases of persistent artificial
organic compounds and petroleum exploitation and transport. Artificial radionuclides
pose threats that are largely nominal. Protection of the arctic environment and human
health can be achieved through universal compliance with existing international
standards. Threats posed by cadmium and lead from anthropogenic sources are
relatively minor. Trends in the anthropogenic mobilization of mercury are not
currently defined with sufficient confidence to assess the associated dangers posed to
animal and human health in the Arctic. Nevertheless, the properties of mercury and its
current environmental levels suggest that the Arctic is sensitive to increased anthropogenic mobilization of this metal. There is also evidence that mercury levels in the
Arctic are increasing due to anthropogenic activities. Accordingly, priority should be
given to determining temporal trends of mercury within the Arctic.
? 1996 International Council for the Exploration of the Sea
Key words: Arctic Ocean, assessment, bioaccumulation, contaminants, controls,
effects, marine environment, metals, pathways, pollution, polycyclic aromatic hydrocarbons, organochlorines, radionuclides, regulation, risks, sources, toxicity, waste.
Previously reviewed by a wide-ranging, international panel; text accepted in its present
form 18 January 1995.
R. W. Macdonald: Institute of Ocean Sciences, P.O. Box 6000, Sidney, B.C., Canada,
V8L 4B2. J. M. Bewers: Bedford Institute of Oceanography, Dartmouth, N.S., Canada,
B2Y 4A2.
Introduction
This paper reviews the current status of the effects of
human activities on the arctic marine environment. Its
purpose is to identify legitimate concerns about the
condition of arctic coastal and ocean environments
arising from contemporary and potential disturbances
with a focus on contaminants i.e. persistent organics,
metals, radionuclides, polycyclic aromatic hydrocarbons, and oil. The approach taken is, first, to outline
the behaviour of contaminants and how they interact
with the major physical, chemical, and biological processes in the Arctic Ocean. The contemporary status of
contaminants is then reviewed to identify concerns and
priorities. The priority issues are then examined to
1054–3139/96/030537+27 $18.00/0
determine whether there are important deficiencies in
existing control measures. The paper has been prepared
using the scientific literature and the results, to date, of
scientific investigations carried out within the Arctic
Monitoring and Assessment Programme (AMAP). The
intention of this review is not to suggest legislation or
regulation; rather, it is to provide sufficient understanding to determine where additional legislation or regulatory controls might be warranted.
An explanation is first required of terms that are
commonly misinterpreted but essential to the discrimination between concerns based on evidence from the
natural sciences and those of perception, which are
based on societal factors and human behaviour. First,
we distinguish between the terms contamination and
? 1996 International Council for the Exploration of the Sea
538
R. W. Macdonald and J. M. Bewers
pollution. We use the term pollution as internationally
defined by the United Nations Joint Group of Experts
on the Scientific Aspects of Marine Environmental
Protection (GESAMP, 1986; Windom, 1991):
‘‘Pollution means the introduction by man, directly or
indirectly, of substances and energy into the marine
environment (including estuaries) resulting in such
deleterious effects as harm to living resources, hazards
to human health, hindrance to marine activities
including fishing, impairment of quality of use of sea
water and reduction of amenities.’’
This definition is used, with some minor revision, in
various international agreements including the United
Nations Convention on the Law of the Sea (UNCLOS).
Its importance is that pollution is equated to an adverse
effect on the marine environment, its resources, and/or
amenities. In contrast, the term ‘‘contamination’’ implies
that the characteristics of the marine environment have
been modified as a result of anthropogenic activities but
without inference that these modifications are in any
way deleterious. This latter term has been less widely
defined but the definition of the Advisory Committee on
Marine Pollution of the International Council for the
Exploration of the Seas (ICES, 1989) is consistent with
the use of the term in most international scientific fora,
including GESAMP:
‘‘Contamination is used to describe the situation
which exists where either the concentration of a
natural substance (e.g. a metal) is clearly above
normal, or the concentration of a purely man-made
substance (e.g. DDT) is readily detectable, but where
no judgment is passed as to the existence of pollution
(i.e. adverse effects).’’
Our ability to recognize contamination continually
improves with the evolution of technology. The sensitivity of electron-capture detectors, for example, has
assured the detection of organochlorine compounds in
virtually every environmental compartment. In contrast,
pollution depends not on our ability to measure a
contaminant but rather to identify associated harm. The
distinction between these two terms is important because
it permits discrimination between instances in which
disturbance by human activities can be detected (contamination) and those in which adverse effects occur as a
result of these disturbances (pollution). In turn, the
difference between these terms implies that there exists a
capacity for change in the marine environment without
adverse effects occurring. This concept of environmental
(or assimilative) capacity has been discussed at length by
GESAMP (1986). The terms pollution, contamination,
and environmental capacity are widely misunderstood
and misinterpreted. This often leads to inferences of
adversity resulting from anthropogenic activities based
on perception rather than science.
In general, when scientists are asked to assess an
activity and its potential effects on the environment they
will adopt what is termed a conservative approach. This
is not always the equivalent of making worst-case
assumptions for the assessment but usually makes allowance for the greater part of the uncertainty associated
with a projection or prognosis. Thus, prior assessment of
the consequences of a particular development usually
does, and must, make such allowances to ensure that
the estimate of expected damage is a pessimistic one.
Defining the scale of possible damage in this way
permits the licensing authority to understand fully the
consequences of a decision to authorize an activity. If
the actual consequences are less than those projected,
the associated damage is less than that deemed originally
tolerable when the project was authorized.
In recent years, a new term has come into prominence
– the so-called precautionary principle (‘‘precautionary
approach’’ or ‘‘principle of precautionary action’’).
Most commonly, it is couched in a context similar to the
following text derived from the Ministerial Declaration
of the Second International Conference on the North
Sea (London Conference, 1987):
‘‘Accepting that, in order to protect the North Sea
from possible damaging effects of the most dangerous
substances, a precautionary approach is necessary
which may require action to control inputs of such
substances even before a causal link is established by
absolutely clear scienific evidence.
‘‘[The Governments] therefore agree to: accept the
principle of safeguarding the marine ecosystem of the
North Sea by reducing pollution emissions of substances that are persistent, toxic and liable to bioaccumulate at source by the use of the best available
technology and other appropriate measures. This
applies especially when there is reason to assume that
certain damage or harmful effects on the living
resources of the sea are likely to be caused by such
substances, even where there is no scientific evidence
to prove a causal link between emissions and effects
(‘the principle of precautionary action’).’’
The philosophical difficulty here is that if precautionary
action is to be invoked irrespective of scientific evidence,
the principle becomes one of management or policy and
not one of science. However, if it is looked on in this
way, it in no way detracts from its effect or purpose but
becomes an additional element of precaution to that
of the conservatism commonly used in scientific evaluations. In general, it is accepted (Gray, 1994) that the
precautionary principle is indeed one of management or
policy and should not be justified on scientific grounds.
The contaminants of concern
The priority contaminants in the Arctic have been
defined under the Arctic Monitoring and Assessment
Contaminants in the arctic marine environment
Programme (AMAP) of the International Agreement on
the Arctic Environmental Protection Strategy and may
be grouped into four categories:
(1) Persistent organochlorine compounds. These include
pesticides, industrial products, byproducts of incineration, and byproducts of chemical processes.
(2) Artificial radionuclides. These include radioactive elements released to the environment by
nuclear weapons testing, nuclear fuel reprocessing,
radioactive waste disposal, and accidents.
(3) Metals. Here, we focus on the highest priority toxic
metals, lead, cadmium, and mercury.
(4) Polycyclic aromatic hydrocarbons (PAHs). These
are components of petroleum and byproducts of
combustion.
Factors controlling the behaviour of
contaminants in the Arctic Ocean
Global biogeochemical considerations
Many substances released to the environment through
human activities (e.g. metals, PAHs) enter natural cycles
that have been operating globally throughout recent
geological history. For the most part, human activities
have enhanced the mobilization of chemical compounds
rather than altered the basic constituents of the natural
environment. Human activities may have perturbed
natural cycles by, for example, mobilizing metals
through mining or producing PAHs as a byproduct of
the combustion of fossil fuels but, due to the underlying
natural cycle, it is often difficult to distinguish the
anthropogenic contribution and its effect on the ecosystem. These naturally-occurring substances may enter
the marine environment through weathering, forest fires,
or biological processes and be removed to marine
sediments by sequestering onto particles. Transport
takes place in the atmosphere (i.e. aeolian transport), in
runoff from the continents and in ocean currents. Contaminants can be in dissolved, particulate, and/or gaseous phases depending on the physical and chemical
properties of the chemical. Mercury, for example, is
transported in all three phases, gaseous transport constituting a major pathway. Less volatile elements, such
as uranium, are transported predominantly in particulate and dissolved forms. High naturally-occurring backgrounds of some substances (e.g. Hg or PAHs) may
make a particular location all the more vulnerable to
exceeding some ‘‘effects threshold’’ due to added
burdens imposed by human activities (cf. Yunker and
Macdonald, 1995).
Some contaminants such as organochlorine pesticides
and artificial radionuclides are, for all practical purposes, produced solely by human activities. Once
released to the environment these compounds enter
biogeochemical cycles according to their chemical and
539
physical properties. Finding these contaminants in the
ecosystem, therefore, unequivocally implicates humans.
These compounds, although produced in small quantities relative to naturally-occurring compounds, often
have properties that make them particularly hazardous
(e.g. toxicity, longevity).
Scientists refer to the distribution of a chemical
among phases as partitioning. Partitioning, which
depends primarily on the properties of the chemical, is
the major physical/chemical control on the transport
and fate of chemicals in the environment. For example,
elements that bind strongly to particles in the ocean (e.g.
particle-reactive elements like Pb, Pu) are susceptible to
scavenging and sedimentation and therefore tend to be
transported over comparatively short distances in the
ocean (Kershaw et al., 1995). Elements that favour the
dissolved phase (e.g. 99Tc, I) tend to remain in ocean
waters for long periods and can be transported over
greater distances. These latter elements are often
referred to as ‘‘conservative’’ because they remain in the
water and are redistributed predominantly by mixing.
Partitioning also affects the way chemicals interact
with aquatic biota. Some compounds exhibit a relationship between their concentrations in an organism and
those in the surrounding environment, e.g. water, that
can be treated as partitioning. This provides the basis for
terms such as biological concentration factor (BCF),
which reflects the equilibrium ratio of the concentration
of a substance in an organism to that in the surrounding
medium (i.e. accumulation from water). The terms bioconcentration, bioaccumulation, and biomagnification
have been used widely, but with varied meanings, to
describe processes whereby biota accumulate contaminants. Here, we use these terms as defined by Macek
et al. (1979):
‘‘Bioconcentration refers to that process whereby
chemical substances enter aquatic organisms through
the gills or epithelial tissue directly from the water;
bioaccumulation is a broader term referring to a
process which includes bioconcentration but also any
uptake of chemical residues from dietary sources.
Biomagnification refers to a process by which the
tissue concentrations of bioaccumulated chemical residues increase as these materials pass up the food chain
through two or more trophic levels.’’
The arctic marine ecosystem is particularly vulnerable
to fat-soluble contaminants because it is simple (low
diversity) and has long food chains supporting top
predators. The top predators tend to be long-lived
species with low fecundity and a high proportion of fat
in their bodies. Within the interior of the Arctic Ocean,
organic carbon is conserved near the ocean’s surface
by efficient recycling as witnessed by relatively high
bacterial production (P. Wheeler, pers. comm.). This
implies that the organochlorine compounds, which
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tend to follow carbon, will also be recycled efficiently.
Taken together, the Arctic Ocean provides an efficient
system both to capture and biomagnify many fat-soluble
compounds in top predators. Aboriginal peoples who
depend heavily on food from the sea, particularly in the
Canadian and Greenland Arctic, occupy the niche of
top-predator in the marine ecosystem.
The distinguishing characteristic of the Arctic is its
comparatively low temperature. Because low temperatures reduce volatilization, semi-volatile compounds can
be preferentially transported to the Arctic by a ‘‘global
distillation’’ process in much the same way as there
exists a net transport of heat from equatorial to polar
regions (Ottar, 1982; Mackay and Wania, 1995; Wania
and Mackay, 1995). The primary medium of such transport for semi-volatile and insoluble substances is the
atmosphere rather than the sea, whose temperature
range is comparatively narrow (Barrie et al., 1992). For
more soluble and less volatile substances, however,
oceanic transport may be more important.
The effects of substances on biological organisms
involve two factors: exposure and toxicity. Exposure
reflects the degree to which an organism is exposed to
the substance in the environment, either directly or
through the medium of food. Toxicity of a chemical
depends wholly on its physical and chemical properties.
This allows a distinction between the terms hazard and
risk, with the former circumscribing the properties,
including toxicity, of a substance and the latter depending on the combination of hazard and exposure. The
potential effects of pollutants on biological organisms
range from the biochemical level to the population level.
At the biochemical level, effects can range from the
induction of enzymes to hormone dysfunction and possible genetic damage, to death. At the population level,
effects can be observed as reproductive failure and
alteration in the ecosystem structure. In general, cause
and effect are harder to link as one moves from the
biochemical level toward the population level. In obviously polluted areas (e.g. some temperate, industrialized
harbours) it is possible to relate contaminant distribution to biochemical, reproductive, and community
changes (cf. Hansen et al., 1985; Varanasi et al., 1987;
Addison and Edwards, 1988; Warwick et al., 1990). For
the arctic marine ecosystem, however, it will be much
more difficult to show that contaminants are producing
these kinds of effect and to date no such effect has been
demonstrated of which we are aware. Risk assessments
for the arctic marine environment are really ecological
risk assessments in which the properties of a contaminant, including its chronic and acute toxicity, are
evaluated together with food web dynamics and
exposure. The focus of such assessments has been the
risks to human health associated with the consumption
of country foods rather than effects on organisms or
the ecosystem itself (e.g. Dietz et al., 1990; Hansen
et al., 1991; Kinloch et al., 1992; Wheatley and Paradis,
1995).
An interesting evaluation of the comparative risks of
fatal cancer induction in human consumers of seafood
containing radionuclides and a number of organic
chemicals suspected of being carcinogens has been carried out by the International Atomic Energy Agency
(IAEA, 1993). Its conclusions reinforce the need for
greater attention to assessing the risks and detriment
arising from potentially carcinogenic substances in seafood. It further notes that this ‘‘would enhance public
appreciation of the potential of the risks associated with
the dissemination of such chemicals into the environment and also provide an improved perspective with
which to judge the risks associated with sea disposal of
radioactive waste’’. Not all contaminants end up in the
Arctic. Reactive substances introduced to the atmosphere and the sea are continually being captured and
removed from circulation by biota and freshwater and
marine sedimentation – the latter can eventually remove
them from the biosphere by sequestering them permanently into sea-floor sediments (i.e. for geological time
scales). Neither do contaminants that reach the Arctic
Ocean simply accumulate there indefinitely. In addition
to sedimentation, other physical/chemical/biological
processes such as radioactive decay, reduction, oxidation, metabolism, hydrolysis, and photolysis remove or
transform contaminants. Lastly, contaminants may
enter the Arctic Ocean and then simply leave again in
exported ice, in outflowing water currents, and by
atmospheric exchange (outgassing).
The task here is to define, first, where human activities
have caused a major change in the incidence and distribution of chemicals in the arctic marine environment
(contamination) and, second, to assess whether these are
associated with adverse effects (i.e. pollution), giving rise
to legitimate concerns about present or future damage.
Besides the release of chemicals to the environment
there are other human activities that cause marine
environmental damage. These include physical destruction of habitat through coastal development, dam construction, dredging, mining, and the breaking of ice for
marine transport. Although not specifically covered by
the GESAMP definition of pollution, the adverse effects
of these activities produce harm to the environment and,
therefore, technically constitute ‘‘pollution’’.
Vertical structure of the Arctic Ocean
The Arctic Ocean is vertically stratified into layers that
owe their origin to water masses entering from the
Pacific and the Atlantic (Fig. 1). These layers are subsequently modified within the Arctic particularly at the
margins (i.e. on the shelves) by sea-ice formation and
runoff. Together, stratification and processes on the
shelves are the primary physical oceanographic controls
Contaminants in the arctic marine environment
72°N
75°
PACIFIC 0
WATER
200
400
ICE
Polar mixed layer
80°
85°
PACIFIC
HALOCLINE
80°
ATLANTIC HALOCLINE ~10 years
Residence time
~10 years
ATLANTIC
WATER
Residence time
~30 years
600
Depth (m)
85°
541
Residence time
~25 years
ATLANTIC LAYER
800
EURASIAN BASIN
DEEP WATER
Residence time
~75 years
1000
2000
CANADA BASIN
BOTTOM WATER
Residence time
~300 years
EURASIAN BASIN
BOTTOM WATER
Residence time
~290 years
3000
4000
CANADA
BASIN
MAKAROV
BASIN
AMUNDSEN
BASIN
BERING STRAIT
NORWEGIAN SEA
&
GREENLAND SEA
DEEP WATER
Alpha
Ridge
NANSEN
BASIN
Lomonosov Nansen Gakkel
Ridge
Ridge
FRAM STRAIT
Figure 1. The vertical structure of the Arctic Ocean showing the polar mixed layer (approximately the top 50 m), the regions of
density gradients between 50 m and 200 m (the haloclines and Pacific layer), the Atlantic Layer (approximately 200–1000 m), and
the basin waters. It should be noted that our understanding of the structure of the Arctic Ocean as shown in this figure is continuing
to undergo significant revision following recent high-quality tracer transects of the Arctic Ocean.
of contaminant distributions. Contaminant distributions
are further influenced by partitioning onto inorganic
and biological particles for which the shelves are, once
again, important locations.
The surface layer, or polar mixed layer (PML),
includes approximately the top 50 m. This layer cycles
seasonally; brine produced by ice formation in winter
tends to destabilize the water column, allowing it to mix,
while in summer, melting ice and freshwater runoff
produces stratification with a fresher surface layer
(5–10 m). Therefore, it is the PML that is in immediate
(annual) communication with the atmosphere and ice,
and it is here and within the ice that most of the
biological primary production occurs. Beneath the PML
is found a region of increasing salinity (the halocline);
the important function of this region is that it acts as a
barrier between the deeper ocean and the upper ocean.
Within the Canada Basin (Fig. 1) the halocline is
dominated by water from the Pacific Ocean which has
been modified as it passes over the Chukchi Shelf; it is
found between about 50 m and 250 m (Macdonald et al.,
1989). The Atlantic layer, which pervades both arctic
basins, is found at depths from about 200 m to greater
than 1000 m. At the top of the Atlantic layer there is a
complex region thought to be supported by insertion of
water masses produced on the shelves (cf. Aagaard et al.,
1981; Jones and Anderson, 1986). This region has been
called the lower halocline to distinguish it from the
Pacific layer, but it is probably a composite water
mass produced by brine drainage from the shelves, the
Barents Sea shelf being a key one (Schlosser et al., 1995).
The time scale of transport or replacement of water
masses varies among the layers: the surface layer has a
residence time of about 10 years in the Arctic (Ostlund,
1982). The haloclines also have residence times of this
order but, as we proceed deeper into the Atlantic layer,
the residence time increases to perhaps 30 years. The
basin waters have the longest time scale, measured in
centuries (Schlosser et al., 1994).
Horizontally, the Arctic Ocean can be subdivided into
two main basins separated by the Lomonosov Ridge
(Fig. 2). These basins are to some degree decoupled from
one another allowing distinct differences in freshwater
content (Aagaard and Carmack, 1989), water-mass
structure (Carmack, 1990; McLaughlin et al., 1995), and
current structure (Aagaard, 1989; Rudels et al., 1992).
By inference, differences in contaminant distribution
between the two major basins are also to be expected.
Modes of contaminant entry
There are five modes of contaminant entry into the
Arctic Ocean. These are: (1) inflowing ocean currents,
(2) deposition from the atmosphere, (3) northward flowing rivers, (4) direct runoff from the land, and (5) direct
disposal into the ocean.
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Figure 2. A schematic diagram showing the predominant currents entering the Arctic Ocean and their major routes around the
basin edges of the Arctic (after McLaughlin et al., 1996; Rudels et al., 1995). The Atlantic Layer circulation is shown in red; this
water mass subducts at the polar front and occupies the region between about 200–1000 m. The Pacific Layer circulation is shown
in blue; this water mass is produced by modification of water entering through Bering Strait on the broad Chukchi shelf and
occupies the region between about 50–250 m in the Canada Basin. The green arrows identify the inflow from major rivers at the
margins; this water follows the surface circulation (see Fig. 4). The numbers given are the estimated inflow or outflow in Sverdrups
(106 m3 s "1). The currents shown in this figure are based on current meter data and on tracer data, both of which are relatively
sparse for the Arctic Ocean. There is considerable variability and uncertainty in the inflows.
Ocean currents
The Arctic Ocean is often referred to as a ‘‘Mediterranean sea’’ because it is surrounded by land and
communicates with oceans to the south only through
restricted passages (Fig. 2). Large quantities of sea water
flow into the Arctic from the Atlantic via the Fram
Strait and the Barents Sea, and from the Pacific via the
Bering Strait. These water masses outcrop in the oceans
of origin, but they subduct within the Arctic Ocean or
before they reach it to produce the vertical structure
outlined above. The inflowing water potentially carries
with it those contaminants received along its flow path
before it enters the Arctic. Contaminants prone to this
mode of entry tend to fall into two categories. First,
those tending not to partition onto solid phases can be
transported over long distances and times because they
are not readily scavenged. Radiocaesium, although
somewhat particle reactive, discharged by European
reprocessing plants, is the most clearly documented
example of such a contaminant; within 5–6 years of its
disposal into the Irish Sea this radionuclide enters the
Atlantic layer of the Nansen Basin in the Arctic Ocean
(Livingston, 1984; Smith et al., 1990; Dahlgaard, 1995;
Kershaw and Baxter, 1995). Second, contaminants that
deposit from the atmosphere into the water south of the
Arctic Ocean and northward flowing currents simply
complete the journey; HCH is a good example (cf.
Hinckley et al., 1991; Gaul, 1994).
Atmosphere
Atmospheric pathways into the Arctic are undoubtedly
more complex than ocean currents. Winter and summer
flow patterns are very different (Barrie et al., 1992). In
winter, air flows mainly from the Eurasian continent
into the Arctic and then out over North America (Fig.
3). It is this flow pattern that produces arctic haze in
spring, and which has been linked to direct, rapid inputs
of contaminants from heavily industrialized areas of
eastern and northern Europe and Asia (cf. Barrie, 1986;
Heintzenberg, 1989; Pacyna, 1991; Welch et al., 1991;
Contaminants in the arctic marine environment
543
Figure 3. Arctic air masses and their transport into the Arctic in winter (after Raatz, 1991).
Klungsøyr et al., 1995). In summer, the Eurasian flow
reverses and a weak north to south transport is found.
At this time, airflow into the Arctic is frequently from
the North Pacific and North Atlantic. Although the
transport path from eastern and northern Europe is a
well-documented and dominant source of contaminants,
episodic transport can also occur from other regions
including North America, Southern Europe/Northern
Africa, and Asia (Pacyna, 1991; Raatz, 1991; Welch
et al., 1991). Unlike transport by ocean currents, which
can take years or even decades, atmospheric transport
can be very fast (days). Hence, the atmosphere can
deliver not only volatile and semi-volatile contaminants,
but also aerosols containing particle-reactive contaminants such as Pb, highly chlorinated PCBs, and PAHs
(cf. Bidleman et al., 1989; Hargrave et al., 1989; Patton
et al., 1989; Barrie et al., 1992; Akeredolu et al., 1994).
These contaminants end up in snow, on ice, in sea water,
and in northern drainage basins (Gregor and Gummer,
1989; Welch et al., 1991).
Rivers and runoff
Total runoff into the Arctic Ocean contributes
about 3300 km3 fresh water each year. This inflow
occurs mainly through four major rivers (Yenisey
(603 km3 a "1), Ob (530 km3 a "1), Lena (520 km3 a "1),
Mackenzie (340 km3 a "1); Aagaard and Carmack,
1989). The remainder of the freshwater inflow comes
from intermediate-sized rivers (totalling 590 km3 a "1)
and numerous other smaller rivers (totalling 720 km3
a "1). Contaminants discharged into northern drainage
basins or into the rivers may subsequently be delivered
to the inner shelves. During their transit, contaminants
will have the opportunity to interact with inorganic and
organic particles. Because deltas, estuaries, and large
shelves tend to trap most of the sediments brought in by
rivers (cf. GESAMP, 1987, 1992), particle-reactive contaminants entering in this manner will tend also to be
initially trapped in these places. However, dissolved
contaminants will be free to travel further, undergoing
dilution but not removal to the sediments. Many rivers
entering the Arctic Ocean form deltas on the shelf
proper but two of the largest, the Yenisey and the Ob,
discharge via long estuaries (Fig. 2; Pfirman et al.,
1995b). It is to be expected that the delivery of particles
and particle-reactive contaminants, their trapping in
coastal sediments, and their vulnerability to subsequent
resuspension and transport, will depend on exactly how
a river enters the sea. This will be particularly important
for sediment transport mechanisms involving resuspension, direct entrainment of turbid river water into ice,
and the entrainment of bottom sediments into ice either
by direct bottom contact or by suspension freezing
which will be discussed below (cf. Reimnitz et al., 1994;
Pfirman et al., 1995b).
Direct disposal
The former Soviet Union (FSU) has discharged and
dumped radionuclides onto shelves in the Barents and
Kara Seas during the past several decades (Yablokov et
al., 1993). This disposal was augmented by underwater
testing of nuclear devices near Novaya Zemlya. Direct
disposal of PCBs into soils occurred at Distant Early
Warning (DEW) line sites in the Canadian Arctic and
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Figure 4. The circulation of ice and the surface layer (compiled from Colony and Thorndike, 1985; Gloerson et al., 1992; Rigor,
1992). The top left view shows the predominant motion of the ice, the top right view shows the average thickness of ice in metres
and the bottom view shows the approximate time taken for ice at a given position to reach the Fram Strait (the zero isopleth).
These figures represent long-term averages and considerable variability can be expected in time and space.
contamination from these sites has been identified in
the nearby marine environment (Bright et al., 1995).
Chronic and acute losses from oil exploration and
production in offshore developments occurred on the
Canadian Beaufort shelf (Thomas et al., 1986). Oil
development on other shelves will similarly result in
marine disposal. Of all the modes of entry, direct
disposal is clearly the easiest to regulate.
Transport of contaminants within the
Arctic Ocean
As illustrated schematically in Figure 2, the various
water masses entering the Arctic Ocean tend to be
steered by topography into boundary currents. Replacement of water in the interior of the basins occurs too but
it is slower. The vertical layers in the ocean have motions
distinct from one another. Much of the water entering
from the Pacific escapes through the Canadian Archipelago while some of it exits at the Fram Strait.
One branch of the Atlantic layer passes over the
Barents Shelf, where it is modified, and then rejoins a
less modified branch which enters the Arctic in the
West Spitsbergen current (Fig. 2; Rudels et al., 1992;
Klungsøyr et al., 1995). The Atlantic layer circulates
counter-clockwise around the basins while being modified by shelf water and by mixing to escape eventually in
the East Greenland current.
In the context of contaminants interacting with living systems the surface currents are most important
involving water to depths of 30–50 m. The motion of
ice which, due to extensive drifter programs (cf.
Gloerson et al., 1992; Rigor, 1992) is well-understood
(Fig. 4), is the best indicator of surface currents. On a
Contaminants in the arctic marine environment
large scale, the Arctic Ocean surface circulation is
organized into two distinct zones. The Beaufort Gyre
in the Canada Basin is seen as a clockwise field of
motion connecting the Chukchi, East Siberian, and
Beaufort Seas. In the vicinity of the Lomonosov
Ridge, the organization of surface circulation undergoes a transition with the Transpolar Drift (Fig. 4).
Within the Transpolar Drift, surface water and
ice from the Kara and Barents Seas tend to cross the
central Arctic Ocean to exit at the Fram Strait. Thus,
if contaminants are introduced to the Arctic Ocean
by rivers discharging onto the Eurasian Shelf, a significant portion of them can be expected to leave by this
route.
The special role of ice
Ice acts as a surface upon which atmospheric contaminants impinge (Melnikov, 1991) and within which contaminated sediments can be incorporated from the
shelves (Pfirman et al., 1995). Therefore, ice is potentially an important vehicle for transporting contaminants over long distances with little alteration. Of equal
importance, the ice provides a vital surface for arctic
marine biota and therefore interaction with contaminants at the water-ice–air boundary is potentially a
critical step of the pathway into the food web. The
period of melting, at which time contaminants in ice are
most likely to be released, coincides with the period of
maximum primary productivity further enhancing the
coupling of ice-borne contaminants with the food web.
Ice cover also acts as a lid on the sea preventing or
slowing down the exchange between the atmosphere and
the ocean. In the event that atmospheric concentrations
of volatile contaminants are reduced, as in the case of
HCH (cf. Jantunen and Bidleman, 1995), ice cover may
slow the natural process of outgassing in the interior
ocean.
Ice tends to be produced on shelves and exported
from them to the interior ocean. Turbulence and
freezing in the water column during ice formation
(suspension freezing) can effectively entrain bottom
sediments by frazil and anchor ice formation (e.g.
Pfirman et al., 1990; Riemnitz et al., 1994 and references
therein). Ice-drift trajectories (Fig. 4) show that, on a
large scale, ice has predictable sources and lifetimes
within the Arctic. Contaminants picked up by ice in the
Kara Sea, for example, will either enter the Barents Sea
within one to two years through the straits between
Novaya Zemlya, Frans Josef Land, and Svalbard or will
exit the Arctic through the Fram Strait within a couple
of years. In contrast, contaminants in ice of the Beaufort
Gyre may have 5 or more years to accumulate as the ice
goes through annual melt/freeze cycles. As the ice drifts
within the Arctic Ocean, the melt–thaw cycle tends
to ablate the ice surface in summer and add ice to
545
the bottom of multi-year floes in winter. This process
moves particles and particle-reactive contaminants to
the ice surface within two to three years. Soluble
(conservative) contaminants, on the other hand, will be
rejected from the ice during the melt–thaw cycle along
with brine. Therefore, it is to be expected that old,
multi-year ice will contain small amounts of conservative contaminants and most of its particle-reactive
contaminants at the ice surface. These particles
will ultimately be released at locations where the ice
completely melts (i.e. the Barents and Greenland
Seas).
In the special case of oil and oilspills, ice will be the
main agent of long-distance transport. Oil will end up on
the surface of the ice because it can enter melt ponds or
leads in summer by floating up through brine channels
in the ice.
The special role of shelves
Arctic shelves process freshwater runoff and sediments
entering from rivers and produce brine in regions of
diverging ice cover in winter (Macdonald et al., 1995).
In spring and summer, rivers discharge most of their
sediment and freshwater loads, biological production
occurs, and large regions of the shelf tend to clear of ice
(Fig. 5a). In winter, freshwater inputs are vastly reduced
and constrained by ice to the nearshore (Fig. 5b). On the
middle and outer shelf, brine produced by the rapid
formation of ice in flaw leads and polynyas destabilizes
the water column and can produce dense water which
flows along the shelf bottom to enter the ocean interior
(Melling and Lewis, 1982). From the perspective of
contaminant transport, the contrast between summer
and winter may be summarized as follows. In summer,
contaminants enter the shelf from rivers, from ice melt
and from the atmosphere. The conservative contaminants will remain near the surface due to the strong
seasonal stratification which prevents mixing and may
be exported from the shelf. Bio-active or particlereactive contaminants enter the food web or become
scavenged by particles to settle toward the shelf bottom.
In winter, the supply of contaminants through ice melt
or rivers is much reduced as is the supply of particles to
scavenge them. However, convection of the water at this
time of year has the potential to move contaminants into
deeper layers of the ocean interior (e.g. into and through
the halocline) either through sinking of surface water
or through entrainment of contaminated sediments in
offshelf flowing water.
Priority pollution concerns
Persistent organic contaminants
Recent excellent reviews of these contaminants in the
Arctic have been given by Muir and Norstrom (1994),
546
R. W. Macdonald and J. M. Bewers
Figure 5. Shelf processes affecting the transport of dissolved and particulate matter in (a) summer and (b) winter.
Norstrom and Muir (1994), Muir et al. (1992), Lockhart
et al. (1992) and Barrie et al. (1992). The persistent
organic contaminants include:
(1) The chlorinated hydrocarbon pesticide family, e.g.
chlordane, toxaphene (a mixture of polychlorinated
camphenes (PCC) or bornanes), hexachlorocyclohexanes (HCH), hexachlorobenzene (HCB),
and DDT. Some of these pesticides are currently
being used in circumpolar countries (e.g. endosulfan,
methoxychlor, lindane, trifluralin, atrazine, dacthal,
and others).
(2) Chlorinated industrial products, byproducts, and
combustion products; e.g. polychlorinated biphenyls
(PCBs), polychlorinated dibenzo-p-dioxins and
dibenzofurans (PCDD, PCDF), and other chlorinated contaminants as yet unidentified (e.g. recently
Contaminants in the arctic marine environment
found
tris-(4-chlorophenyl)-methanol
(TCP)
which is now known to be a byproduct of DDT
manufacture).
(3) Polynuclear aromatic hydrocarbons (PAHs).
This list, which contains the predominant organic contaminants found in the arctic marine environment, is not
complete because new compounds are being manufactured each year. However, the environmental behavior
of new compounds can now be reasonably wellpredicted by multi-media models using only the physical
and chemical properties of the compounds (e.g. Wania
and Mackay, 1995). Therefore, in principle, there is a
basis to determine whether or not a new synthetic
compound poses a threat to the global environment.
Organochlorine compounds
Starting in the 1940s, organochlorine compounds were
produced and released predominantly in the temperate
land areas of the northern hemisphere. Because they
have given rise to evident adverse effects in the environment close to sources, many of these compounds (chlordane, DDT, PCB) have already been banned, or
restricted, in North America and western Europe
(Pacyna, 1995) and measures are presently being undertaken in some countries to reduce emissions by secondary sources such as incinerators (Anon., 1995). A
significant proportion of the releases from past production, however, continues to cycle in the global environment and contributes a large share of the organochlorine
compounds presently entering the Arctic. As pointed out
by Goldberg (1991), large quantities of organochlorine
pesticides continue to be used in less-developed
countries, particularly in the southern hemisphere.
Unfortunately, there is little quantitative information on
the amounts being used, nor are good data available for
the former Soviet Union or China, both of whom have
been major contributors to the global budget (Norstrom
and Muir, 1995). It is essential that we acquire and
document consistent and reliable statistics on the production and use of organochlorine compounds if we are
to develop regulations aimed at protecting the global
environment.
The properties that make many organochlorine
compounds particularly hazardous are their toxicity,
volatility, persistence in the environment, and affinity for
fat. Organochlorines exhibit toxicity in a variety of
ways. Some are relatively innocuous (e.g. áHCH), some
are carcinogenic (e.g. toxaphene to animals), some disrupt hormones (e.g. DDE), some are highly toxic to
mammals, fish and birds (e.g. 2,3,7,8-TCDD and planar
PCBs) while most of the organochlorine pesticides are
neurotoxins. Their mode of entry into the Arctic Ocean
has occurred predominantly via the atmosphere
although, due to their relatively high partitioning from
the gas phase into sea water, HCH and possibly some of
the components of toxaphene, may have been delivered
547
in significant amounts through ocean currents. PCBs,
DDT, and other contaminants have reportedly also
entered the Russian shelves via contaminated rivers,
particularly the Yenisey, Ob, Lena, and Kolyma rivers
(Pfirman et al., 1993). Due to the strong affinity of highly
chlorinated compounds for particles, a large proportion
of these contaminants will become trapped in shelf
and slope sediments. Some of these sediment-bound
contaminants may later enter ice through suspension
freezing or ice grounding.
The organochlorines are readily incorporated into
phytoplankton which are in direct contact with the
water and have high surface area per unit volume. Entry
of organochlorines into the food chain is most efficient
when the lipid/water solubility ratio exceeds 104 as it
does for PCBs, chlordane, and HCB. The relatively high
concentrations in marine phytoplankton (compared to
terrestrial primary producers) result in higher values
in the marine food web (Fig. 6). For example, bioconcentration factors (BCF) from water to zooplankton
range from about 7#104 for toxaphene to 2#106 for
PCBs (Muir et al., 1992); bioaccumulation in cod ranges
from about 2#105 for toxaphene to 2#108 for PCBs
and, finally, the biomagnification for top predators
relative to water is about 107 for toxaphene and 109 for
PCBs. However, this is not the entire story because
substantial changes in the composition of organochlorine compounds within organisms are mediated
by excretion and/or metabolism which varies among
species.
Our ability to assess temporal trends of organochlorine contaminants in marine biota from the Arctic is
limited by the sparseness of data; at present, usually only
two or three sampling periods are available from which
to infer trends. Efforts to detect such trends in marine
animals are further hampered by other sources of
variance including the evolution of analytical methods,
seasonal and spatial variations, difference among
species, and differences related to both the sex and age of
the animals. Organochlorine concentrations in seabird
eggs from Prince Leopold Island (Noble, 1990) and in
seabirds from the Barents Sea (Savinova et al., 1995)
suggest that burdens have declined since the mid 1970s.
Two of the more toxicologically significant organochlorine compound classes, PCBs and DDT, were
measured in ringed seals between 1972 and 1981 by
Addison et al. (1986). They found that concentrations of
both compound classes declined during that time interval and that PCBs appeared to be decreasing faster. This
finding was confirmed by Muir et al. (1988) and has been
extended with samples collected in 1989 (Addison, pers.
comm). An attractive explanation for more rapid
declines in PCBs is their complete production phase-out
by the end of the 1970s whereas DDT continues to be
used in some developing countries (cf. Pacyna, 1995;
Voldner and Li, 1995). However, a review by Kurtz
548
R. W. Macdonald and J. M. Bewers
Figure 6. A simplified Arctic food web illustrating the transfer of organochlorine contaminants from the dissolved phase into the
food web and up into the highest trophic levels. Note that the diagram is intended to be representative rather than comprehensive.
For the sake of simplicity, many links have been omitted, and most animal diets include a wider variety of items than shown here.
For example, walrus have been observed to eat seals, arctic foxes scavenge polar bear kills, and sculpins and cod provide food for
humans. For each step up the food web there is a subsequent increase (biomagnification) in fat-soluble contaminants as illustrated
by the colour scheme: green]yellow]magenta]orange]red (diagram has been modified from Welch et al., 1992; Muir and
Norstrom, 1994).
Table 1. Trends for selected organochlorine contaminants for data through 1993. Data included only
males and excluded pups and juveniles (Kurtz, 1995).
Ringed seals
Polar bears
PCBs
DDT
Chlordane
Chlorobenzenes
HCH
Drop
Rise
No change
Drop
No change
Rise
No change
Rise
No change
Rise
(1995) of the organochlorine data up to 1993 for both
ringed seals and polar bears (Table 1) clearly illustrates
the danger of generalizing contaminant trends using the
present data base for marine mammals. Well-designed
programs are urgently required to reveal how organochlorine contaminant burdens in arctic predators are
responding to the various changes in production and use
of the organochlorines.
The four major organochlorine contaminants found
in marine mammals are DDTs, PCBs, chlordanes, and
PCCs (Norstrom and Muir, 1994). A substantial, but
geographically incomplete, data set has been accumulated for marine mammals including bears, seals, and
whales (Norstrom and Muir, 1994 and references
therein). These data show variations in concentration
among species and from location to location within the
Arctic. Contamination of the Canadian arctic marine
ecosystem by DDT and PCBs is generally 10–50 times
less than the marine environment at temperate latitudes
of the northern hemisphere and well below the threshold
known to produce, for example, hyperadrenocortical
disease syndrome or reproductive failure in seals
(Norstrom and Muir, 1994). However, caution is warranted as elevated concentrations of organochlorines
have recently been found in bears and seals from the
vicinity of Svalbard (Norheim et al., 1992; Skaare et al.,
1994; Norstrom et al., 1995). Contaminant burdens
may also be very high locally near the Russian coast
(Melnikov and Vlasov, 1992) but data are lacking to
make a valid assessment of this region.
The top predators (e.g. small-toothed whales, polar
bears, birds, humans) are at the greatest risk from
organochlorines. However, precise cause and effect relationships between the organochlorine compounds and
biological disorders remain elusive due to many factors
including the complexity of the compounds themselves
Contaminants in the arctic marine environment
(e.g. PCBs alone include 209 different congeners each
with a specific toxicity). Nevertheless, Norstrom and
Muir (1994) point out that where comparisons can be
made, the levels of PCBs and DDTs in arctic marine
mammals are below the effects level (bear populations at
Svalbard may be an exception – cf. Norheim et al.,
1992). Polar bears and seals, both of which undergo
large seasonal changes in fat content, may be more at
risk than mean organochlorine burdens would suggest
due to the release of fat-associated contaminants during
periods of fasting. Cod, which undergo large changes in
fat content during spawning, may similarly be at greater
risk during that time.
High concentrations of organochlorine contaminants
in arctic marine mammals pose a risk to humans who
consume large quantities of country foods (Dewailly
et al., 1989). In a study at Broughton Island, NWT,
Kinloch et al. (1992) found that 12% of women and
22% of men exceeded the medium-term tolerable daily
intake of 1 ìg kg "1 day "1 PCBs established by the
Canadian Government. Despite this disturbing finding,
the authors concluded that altering consumption
patterns to avoid these contaminants was not warranted:
nutritional and cultural benefits of these country foods
outweighed the risks of contamination.
Polynuclear aromatic hydrocarbons
Unlike the organochlorines, PAHs are widespread products of natural cycles (forest fires, plant decomposition,
diagenesis, petrogenesis) and therefore anthropogenic
contamination is more difficult to identify. Contaminant
sources for the Arctic include petroleum exploration and
production (Thomas et al., 1986) and long-range atmospheric transport (Patton et al., 1991; Lockhart et al.,
1992; Kawamina and Suzuki, 1994; Masclet and Hoyau,
1994; Yunker and Macdonald, 1995; Klungsøyr et al.,
1995; Peters et al., 1995). PAHs, especially the more
soluble low molecular weight compounds, can bioconcentrate in fish from sediments (cf. Hellou et al., 1995
and references therein). However, PAHs are less prone
to bioaccumulation or biomagnification than the
organochlorines, partly because of metabolic degradation of PAHs in top predators and their prey. The
toxicity of PAHs derives from their potential, during
metabolism, to form carcinogenic and mutagenic diols
and epoxides that react with DNA (Zedeck, 1980). The
potential of contaminant PAHs to produce toxic effects
must be viewed in the context of a variable, naturally
occurring background (Yunker and Macdonald, 1995).
Even where high natural PAH background and concentrations have been found (the Canadian Beaufort
Sea), Yunker and Macdonald (1995) concluded that the
measured levels were below the threshold required for
observable effects on biota. Because PAHs are not prone
to biomagnification, there is less concern about dispersed, long-range inputs. PAHs are more likely to have
549
an effect in regions where direct inputs occur (e.g.
creosote applications, oil spills) especially in circumstances where there is confined, long-term contact
between the organism and PAHs (e.g. flatfish associated
with contaminated sediments; cf. Malins et al., 1984;
Hellou et al., 1995).
Radionuclides and radioactivity
This section assesses the sources and effects of radionuclides in the Arctic leading to an evaluation of the
need for additional regulatory action. Much of the
information presented is based on the review of Aarkrog
(1993).
Sources of radionuclides
The predominant radionuclides derived from anthropogenic sources measured in the arctic marine environment
comprise 99Tc, 129I, 137Cs, 90Sr, 238Pu, 239+240Pu, 241Pu,
and 241Am. To these can be added a range of short-lived
fission and activation products derived from the Chernobyl accident which are now mostly below normal
limits of detection. The important sources of radionuclides are: atmospheric fallout (both local and global)
from weapons tests and satellite power unit burnup;
nuclear fuel reprocessing installations in Europe; discharges from the Ob, Yenisey, and Lena rivers; and
releases from dumped Soviet and Russian radioactive
wastes and accidents. The only sources giving rise to
detectable quantities of radionuclides in the arctic
marine environment generally are global nuclear
weapons fallout and nuclear fuel reprocessing operations in Europe, particularly prior releases from
Sellafield. The relative importance of these sources varies among radionuclides. For example, the vast majority
of the 99Tc in the Arctic seas originates from European
reprocessing plants (Dahlgaard, 1995a) while the correlation of 90Sr with salinity suggests that the dominant
contemporary source for this nuclide is land runoff
containing global atmospheric fallout (Dahlgaard,
1995b). In contrast, 137Cs derives from atmospheric
fallout, discharges from European reprocessing plants,
re-mobilization from Irish Sea sediments, and releases
from the Chernobyl accident, partly via the outflow
from the Baltic Sea (Dahlgaard, 1995b). With limited
exceptions, such as the 238Pu from the radionuclide
thermoelectric power unit in the US Navy navigation
satellite Transit 5BN that burned up over the Indian
Ocean in 1964, introductions of artificial radionuclides
from other sources are regionally undetectable against
the background of those present from the three principal
source categories. On local scales, however, such as in
proximity to rivers delivering radionuclides from human
activities on land (e.g. the Ob and Yenisey rivers), in
locations at, or near to, nuclear weapons explosion sites
(e.g. Chernaya Bay, Novaya Zemlya) or in the immediate vicinity of dumped or lost solid radioactive waste
550
R. W. Macdonald and J. M. Bewers
137
15
Cs fallout (Bq × 10 )
6.0
(a)
5.0
4.0
3.0
2.0
1.0
No data
Sellafield
137
Cs ocean discharge
15
(Bq × 10 )
0.0
6.0
(b)
5.0
4.0
3.0
2.0
1.0
0.0
1950
1955
1960
1965
1970
1975
1980
1985
1990
1995
137
Figure 7. (a) The fallout record for Cs from atmospheric weapons testing for the latitude band from 70)N–90)N (137Cs fallout
was estimated using available 90Sr measurements from 1958–1990 given in Environmental Measurements Laboratory and the
Health and Safety Laboratory reports (provided by Dr M. Monetti) and scaling these numbers upward by a constant factor of 1.6);
and (b) the discharge of 137Cs to the ocean from the Sellafield nuclear reprocessing plant for the period 1951–1992 (data taken from
Gray et al., 1995). 137Cs was chosen because it represents the largest percentage contribution to total dose from the artificial
radionuclides (Table 2).
sources, radionuclides from other sources can be both
expected and occasionally detected. The following discussion of sources provides the basis for assessing
whether they are amenable to control – an essential
prerequisite to determining the utility of any additional
regulatory controls.
Fallout from weapons tests
Early nuclear weapons tests between 1945 and mid 1952
involved only fission weapons and there was no significant large-scale transport of radioactive contamination
to the Arctic. The fallout from the major tests at Bikini,
Nevada, and Semipalatinsk was deposited at the test
sites and within the corresponding latitude bands. Thermonuclear weapons testing began in October 1952,
when the US tested its first thermonuclear device at
Eniwetok Island, and this began the period of significant
global radionuclide fallout (Fig. 7). The total direct
fallout into the arctic marine environment up to 1993
corresponded to 9 PBq of 90Sr and 15 PBq of 137Cs
(Aarkrog, 1993) (PBq=peta Becqueral=1015 disintegrations per second). In addition, some 1.5 PBq of fallout
90
Sr and 0.5 PBq of fallout 137Cs (decay corrected to
1993) has entered the Arctic Ocean via river transport
(Aarkrog, 1993).
Between 1955 and 1962, 87 atmospheric test explosions took place at Novaya Zemlya. These explosions
released between 100 and 370 PBq of 90Sr and 155 and
560 PBq of 137Cs as fallout, predominantly global but
some of it regional. Aarkrog (1993) gives an upper
estimate for the latter as 20 PBq 90Sr and 30 PBq 137Cs.
Atmospheric nuclear weapons test explosions have
almost totally ceased as a result of the Partial Nuclear
Test Ban Treaty covering testing in the atmosphere,
space, and water that entered into force on 10 October
1963. This is reflected in declining atmospheric fallout
since that time (Fig. 7).
Nuclear fuel reprocessing in Western Europe
Releases to the ocean from nuclear fuel reprocessing
plants at Dounreay, Cap de la Hague, and Windscale/
Sellafield have all contributed to the introduction
of radionuclides by water transport into the Arctic
(Kershaw and Baxter, 1995). However, Sellafield is
historically and currently the most important source
among these. Maximum releases from Sellafield
occurred in about 1975 when approximately 5 PBq of
137
Cs were discharged to the Irish Sea (Gray et al., 1995
and see Figure 7b for the 137Cs discharge record).
Releases have continually been reduced since that time
Contaminants in the arctic marine environment
and are now of the order of 0.1 PBq per year (excluding
3
H). Marine transport of 137Cs to the Arctic Ocean takes
about 5–7 years (Aarkrog et al., 1987) but a substantial
proportion of the radiocaesium previously discharged
from Sellafield, currently about 10–15 PBq, resides in
the Arctic Ocean (Aarkrog, 1993; Kershaw and Baxter,
1995). These radionuclides are found in the Atlantic
layer (Fig. 1) and therefore are not immediately accessible to the biological community in surface waters.
River discharges
Nuclear fuel reprocessing has taken place at the Mayak
Production Association, Chelyabinsk, and at plants in
Tomsk and Krasnoyarsk. During the period 1949–1956,
liquid radioactive wastes from the Mayak reprocessing
plant were released to the Techa-Iset-Tobol rivers, which
are within the Ob drainage basin, including about
100 BPq of â-activity to the Techa River (Akleyeve and
Lyubchansky, 1944). In 1993, the remnants of this
activity amounted to 4.3 PBq of 90Sr and 4.6 PBq 137Cs.
Additional releases were made from the Tomsk plant to
a tributary of the Ob and from the Krasnoyarsk plant to
the Yenisey River. The outflows from the Ob, Yenisey,
and Lena rivers contain artificial radionuclides from
fallout and possibly activities within the respective
drainage basins. Considerable research is being expended to quantify and characterize the radionuclides
in these outflows under the United States Arctic Nuclear
Waste Assessment Program (ANWAP) that is funded by
the Office of Naval Research. Presently, there appears to
be little evidence of radionuclide sources other than
fallout to the lower reaches of the Ob and Yenisey rivers
and to the Kara Sea (Baskaran et al., 1995) or of
transport from the FSU dump sites to distant regions in
the Arctic Ocean (Baskaran and Naidu, 1995).
While the industrial sources of contemporary
discharges are amenable to control, radionuclides
previously released into the environment have created
secondary sources that are more appropriately the
subject of intervention under the terms of the System
of Radiological Protection (ICRP, 1991).
Accidents (Chernobyl, submarine losses and
aircraft accidents)
Accidents inevitably occur. Chernobyl, perhaps the best
known example, was preceded by other accidents in the
nuclear industry (Windscale, Three Mile Island). At
Chernobyl, 100 PBq of 137Cs was released to the atmosphere, most of which was deposited throughout the
northern hermisphere. Of this amount, about 1 PBq was
deposited between 70–90)N (i.e. in the Arctic). This has
been augmented by the transport of approximately a
further 5 PBq of 137Cs from the north-east Atlantic into
the Arctic through ocean advection (Aarkrog, 1993).
In 1968, a B52 aircraft carrying four nuclear weapons
crashed on the sea-ice 11 km west of Thule, Greenland.
551
The conventional explosives in the weapons detonated
and distributed the warhead plutonium (total 26 kg)
onto the wreckage and into the surrounding environment (the atmosphere, on to sea ice and into and underneath the sea ice at the point of impact). Between 2.8 and
4.2 kg of plutonium was estimated to be distributed in
and around the crash site in addition to unquantified
radionuclides in contaminated wreckage (Aarkrog,
1995). The wreckage and ice-deposited plutonium were
recovered as far as possible, leaving an estimated 0.2–
0.6 kg of plutonium in the environment. Some of this
plutonium was subsequently deposited into the underlying marine sediments. The redistribution of sedimentary plutonium at the site has been the subject of periodic
investigations and, for an assessment of the transport of
this residual plutonium and dose consequences, see
Aarkrog (1995) and Smith et al. (1994).
The major submarine loss in the Arctic Ocean was
that of the Komsomolets in the Norwegian Sea. The hull
of this submarine contains a nuclear reactor and two
torpedoes with nuclear warheads. While there is some
agreement that the total plutonium activity of the
warheads is 215 TBq this has been differently equated
to 3.5 kg (Livovsky et al., 1995) and 6 kg (NATO, 1995)
in units of mass. Because of its extreme affinity for
particulate material (that is, a high sediment–water
partition coefficient) plutonium released from torpedo
warheads is likely to be retained in sediments in the
vicinity of the wreck. The warheads probably also
contain a larger quantity of 235U (NATO, 1995). The
current status of this wreck and the consequences of
corrosion and release of radionuclides have been studied
by a NATO/CCMS Pilot Study (NATO, 1995). The low
probability of catastrophic failure and limited extent of
likely contamination of the environment make this a
source of limited environmental concern. The rather
corroded state of the submarine hull makes its recovery
impractical; however, some consideration is being given
to encapsulating the wreck in an absorptive blanket.
Dumped radioactive sources
The former Soviet Union and the Russian Federation
carried out dumping at sea of liquid and solid radioactive wastes both prior to, and after, the Soviet Union
became a contracting party to the London Convention
1972 (Yablokov, 1993). Some of these dumping activities took place in contravention of the provisions of the
London Convention (IMO, 1972). The most important
dumped wastes are those of submarines and marine
nuclear reactor assemblies containing spent fuel. Assessment of the rates of release of radionuclides and the
resultant consequences in terms of risks to the environment and human health are being carried out as part of
both the US Arctic Nuclear Waste Assessment Program
and the IAEA’s International Arctic Seas Project
(IASAP), to be completed in early 1997.
552
R. W. Macdonald and J. M. Bewers
Northern natives
Adult
External
(18%)
Cs-137
(4%)
Others
(3%)
Rn-220
(7%)
Th-232
(7%)
Child
Pb-210
(3%)
Cs-137
(1%)
Others
(4%)
External
(13%)
Rn-220
(5%)
Th-232
(2%)
Po-210
(57%)
Th-228
(1%)
Th-228
(1%)
Critical group
Pb-210
(2%)
Cs-137
(5%)
Others
External (3%)
(10%)
Pb-210
(3%)
Rn-220
(5%)
Th-232
(9%)
Po-210
(72%)
Th-228
(1%)
Figure 8. Contribution to annual radiation dose for native northern Canadian residents (reproduced from Beak, 1995). The
estimated contributions made by external exposures and by specific natural and artificial radionuclides through inhalation (mainly
222
Rn) and ingestion pathways are given for average adults and children based on pooled data for five northern communities
(Broughton Island, Baker Lake, Rae-Edzo, Makkovik, and Old Crow). The estimated contributions to total individual dose for
members of more exposed communities consuming large amounts of country foods, particularly caribou meat, are depicted under
the heading ‘‘critical group’’. With the exception of 137Cs, which is produced artificially, all of the radionuclides specifically
identified in this figure are natural.
Dose consequences and associated risks
The main criterion for assessing the severity of risks
posed by radionuclides is exposure, or dose, and the
associated risk to health. Thus, it is not the presence,
levels, or sources of radionuclides in the environment
that are of immediate importance. Rather, it is the
exposure of animals and humans to those radionuclides
that is paramount. The System of Radiological Protection (ICRP, 1991) embodies an assumption of a linear
no-threshold dose-response relationship at low doses
such as those generally received from environmental
sources. The ability to convert radiation dose to a risk of
health defects allows risks associated with low doses to
be estimated and compared with risks posed by other
activities and practices (cf. IAEA, 1993).
An example of the contributions to dose made by
radionuclides to arctic inhabitants is presented in Figure
8 (after Beak, 1995). The corresponding risk of serious
health defect (i.e. fatal cancer induction) posed by
radionuclides to arctic inhabitants is presented in Table
2 (Beak, 1995). This example is for residents of
Broughton Island consuming either country foods alone
or in combination with foods imported from the south.
Individuals in this community consume relatively large
amounts of country foods and are therefore expected to
be at greater risk than average individuals within the
Canadian Arctic. Truly incremental risks, in addition to
those posed by natural background radiation, are only
associated with exposures to artificially enhanced levels
of natural radionuclides or to artificial radionuclides. It
should be noted, as indicated in the footnote to Table 2,
that in the cases of 134Cs, 137Cs, 226Ra, 210Pb, 210Po,
232
Th, and 235U the dose associated with land mammal
consumption is the dominant contributor. Only in the
cases of the natural radionuclides 228Th, 230Th, 234U,
and 238U is the dose from aquatic foodstuff consumption dominant. Among these latter radionuclides the
major pathway for 228Th, 234U, and 238U doses is marine
mammal consumption.
As can be seen in Table 2 and Figure 8, the largest
doses and risks are associated with natural radionuclides, predominantly 210Po, 210Pb, 228Th, and 232Th.
Among the radionuclides derived from anthropogenic
sources, 137Cs is the major contributor to dose but
represents only 2–3% of the total and derives predominantly from land-mammal, and not seafood,
consumption.
To provide a context against which to consider the
risks presented in Table 2, the annual probabilities of
death from a variety of causes to individuals in the
United Kingdom and United States populations are
provided in Table 3 (after IAEA, 1993).
None of the doses and risks associated with artificial
radionuclides at the currently observed levels in the
Arctic Ocean are large enough to warrant immediate
concern. Thus, the major concerns relate to practices
and sources that threaten to give rise to exposures that
would not adequately protect potentially exposed
populations. Among these are previously dumped
high-level wastes, which are the subject of the IAEA’s
International Arctic Seas Assessment Project (IASAP).
This study will include an examination of contemporary and future risks to determine if they are high
enough to warrant intervention (i.e. remediation). The
only manner of minimizing both the probabilities
and effects of accidents is to employ the same
Po-210
(64%)
Contaminants in the arctic marine environment
553
Table 2. Doses and risks to adults in the Broughton Island population.
Radionuclide
134
Cs*
Cs*
90
Sr‡
226
Ra*
210
Pb*
210
Po*
7
Be†
228
Th§
230
Th§
232
Th*
234
U§
235
U*
238
U§
Total
137
Mixed diet
(Total dose)
(Sv a "1)
Incremental
risk of serious
health defect
(a "1)
Country
food diet
dose
(Sv a "1)
Incremental
risk of serious
health defect
(a "1)
6.14#10 "7
5.91#10 "5
2.25#10 "7
1.54#10 "5
6.78#10 "5
1.45#10 "3
2.00#10 "10
6.66#10 "5
2.36#10 "6
1.11#10 "4
1.27#10 "6
2.95#10 "6
1.02#10 "6
2.36#10 "3
3.1#10 "8
3.0#10 "6
1.1#10 "6
7.7#10 "7
3.4#10 "6
7.2#10 "5
1.0#10 "11
3.3#10 "6
1.2#10 "7
5.6#10 "6
6.4#10 "8
1.5#10 "7
5.1#10 "8
1.2#10 "4
8.55#10 "7
8.28#10 "5
2.25#10 "7
2.10#10 "5
9.58#10 "5
1.99#10 "3
2.00#10 "9
8.60#10 "5
3.01#10 "8
1.54#10 "4
1.64#10 "6
4.10#10 "6
1.30#10 "6
3.0 #10 "3
4.3#10 "8
4.1#10 "6
1.1#10 "8
1.1#10 "6
4.8#10 "6
1.0#10 "4
1.0#10 "10
4.3#10 "6
1.5#10 "7
7.7#10 "6
8.2#10 "8
2.1#10 "7
6.5#10 "8
1.5#10 "4
Percent of
total dose
(%)
0.02–0.03
2.35–2.57
0.01
0.61–0.65
2.69–2.97
57.6–61.6
0
2.64–2.66
0.09
4.39–4.76
0.05
0.12–0.13
0.04
*Dose is primarily associated with land mammal consumption.
†Inhalation pathway only.
‡Calculation for water consumption only although this may not be the predominant pathway of
exposure.
§Dose is primarily associated with aquatic foodstuff consumption.
rigorous assessments and precautions that are widely
adopted in the nuclear industry on a common and
uniform basis.
Table 3. Probability of death, per year, from selected causes in
UK and USA populations.
Cause of death
Cardiovascular disease [age 35–74] (UK)
Cancer of any type (USA)
All natural causes [age 40] (UK)
Motor vehicle accident (USA)
Air pollution [eastern] (USA)
Influenza (UK)
Accident at home (USA)
Leukaemia (UK)
Motor vehicle accident [pedestrian] (USA)
Accident at home (UK)
Accident at work [working force] (UK)
Alcohol [light drinker] (USA)
Homicide (UK)
Aflatoxin in 4 tablespoons/day of
peanut butter (USA)
Electrocution (USA)
Accident on railway (UK)
Lightning (UK)
Annual
probability
9#10 "3
3#10 "3
1#10 "3
2#10 "4
2#10 "4
2#10 "4
1#10 "4
8#10 "5
4#10 "5
4#10 "5
2#10 "5
2#10 "5
1#10 "5
8#10 "6
5#10 "6
2#10 "6
1#10 "7
[Square brackets relate to nature of sub-population to which
risk applies].
(Normal parentheses define national population to which
risk applies).
Metals (cadmium, lead and mercury)
Cadmium (Cd), lead (Pb), and mercury (Hg) are toxic at
high enough concentrations and none has any known
nutritive benefit. Metals released by anthropogenic
activities enter natural geochemical cycles, making it
difficult to assess incremental effects due to human
mobilizations. Each of these metals has unique properties that give rise to differing biogeochemical behaviour.
For example, Pb tends to become strongly attached to
particles and hence to be sedimented with little tendency
to recycle; Cd favours the dissolved phase and enters
marine biological cycles, thereby correlating strongly
with the nutrient phosphate (Boyle et al., 1976). When
sedimented, Cd tends to become remobilized (Gobeil
et al., 1987). Hg becomes attached to organically rich
particles but exhibits several forms (elemental, ionic,
methylated) each of which has different toxicity and
mobility (cf. Weiner and Spry, 1995; Linqvist and
Rodhe, 1985).
Undoubtedly, the Arctic has been contaminated
by metals as reflected in Pb and Cd distributions in
Greenland ice cores (e.g. Rosman et al., 1994; Boutron,
1995) and Hg profiles in dated sediment cores (Lockhart
et al., 1995; Lucotte et al., 1995; Slemr and Langer,
1992). The predominant inputs of anthropogenic metals
to the Arctic Ocean are from long-range transport via
the atmosphere (Rahn, 1981; Jickells et al., 1990; Barrie
et al., 1992; Akeredolu et al., 1994; Pacyna, 1995),
locally from heavily industrialized regions in the Russian
Arctic (Matishov, 1993; Klungsøyr et al., 1995; Pacyna,
554
R. W. Macdonald and J. M. Bewers
1995), or from mine sites within the Arctic (Nanisivik
in Strathcona Sound, Muir et al., 1992; Blackangel
in Greenland, Asmund, 1992). Rivers may also be
an important route of entry for contaminant metals
(Melnikov, 1991). However, recent data on the three
major Russian River estuaries (Yenisey, Ob, Lena)
suggest that they are pristine with respect to a suite of
metals including Pb, Cd (Dai and Martin, 1995; Martin
et al., 1993), and Hg (Coquery et al., 1995). The largest
North American river, the Mackenzie, is reportedly also
pristine (Thomas et al., 1986). Smaller rivers draining
heavily industrialized areas of northern Europe and Asia
may provide locally important heavy metal inputs to the
Russian shelves but there are no reliable data that would
permit an evaluation.
Sufficient data are available to make preliminary estimates of the atmospheric fluxes of
anthropogenically-derived Pb, Cd, and Hg to the Arctic
Ocean. The total annual atmospheric fluxes of contaminant Pb and Cd to the Arctic from Eurasia have been
estimated (for 1979 to 1980) at 2400 tonnes a "1 and
47 tonnes a "1, respectively (Akeredolu et al., 1994).
Based on stable-isotope data, atmospheric contaminant
Pb is derived predominantly from eastern Europe and
Russia (Sturges et al., 1993). The evidence suggests that
the fluxes of contaminant Pb are significant when compared to other sources (cf. Rahn, 1981) whereas those
of contaminant Cd are not (Rahn, 1981; Yeats and
Westerlund, 1993). Pacyna and Keeler (1995) estimate
that 60–80 tonnes a "1 of contaminant Hg is presently
deposited from the atmosphere in the Arctic. Atmospheric deposition on ice close to source regions (e.g. the
Kola Peninsula) has the greatest potential to produce
metal concentrations that affect, for example, ice biota.
This would be most likely to occur over the inner
Russian shelves (cf. Melnikov, 1991). For Hg, the predominant anthropogenic sources have been the burning
of coal, mining and extraction of Hg, smelting of other
metals, chlor-alkali plants, and municipal waste
incineration. The total anthropogenic emissions of Hg
to the atmosphere, estimated for 1983 at about
3.6#103 tonnes a "1 (Nriagu and Pacyna, 1988), exceed
those from natural sources of 2.6#103 tonnes a "1 (for
estimated Hg releases see also Linqvist and Rodhe,
1985; Fitzgerald, 1986; Nriagu and Pacyna, 1988;
Linqvist et al., 1991; Fitzgerald, 1995; Pacyna and
Keeler, 1995). Evidence from sediment cores collected in
remote northern lakes in Canada and in Hudson Bay
(Lockhart, 1995; Lucotte et al., 1995) suggests that Hg
fluxes may have increased by a factor of two or more
during the past century.
Both anthropogenically and naturally-derived metals
entering the arctic shelves will be processed according to
their biogeochemical properties. Pb in the water column
is likely to be scavenged by particles and deposited and
buried in shelf sediments. Much of the Hg will also enter
sediments attached to fine, organically-rich particles.
However, Cd will be more likely to follow the nutrient
cycle, to be taken up in plankton, released during
regeneration, exported from the shelves in brineenhanced water and, finally, to be found associated with
nutrient maxima in the ocean (see Figs 1 and 5a,b;
Yeats and Westerlund, 1991; Yeats and Bewers, 1987).
Ultimately, Cd is removed to sediments by precipitation
in sulphides (Gobeil et al., 1987; Pedersen et al., 1989;
Rosenthal et al., 1995). Away from the shelves, relatively
clean surface ocean water has been observed both for Cd
and Pb in the vicinity of Svalbard (Mart, 1993) and for
Cd in the central Arctic Ocean (Moore, 1981; Yeats and
Westerlund, 1991). These observations suggest that contaminant Pb and Cd inputs to the interior ocean water
column are either small compared to the natural budgets
(likely the case for Cd) or that residence times in the
upper ocean are short (likely to be the case for Pb; e.g.
Schaule and Patterson, 1981). In contrast to the water,
ice has a greater potential to transport over long distances contaminant metals deposited on its surface or
picked up from the bottom by suspension freezing over
the shelf (Pfirman et al., 1995). The melt–freeze cycle
may further concentrate these metals at the ice surface.
Metals deposited in ice over the Russian shelves will
then be transported predominantly in the ice drift
toward the Barents Sea or the Fram Strait to enter the
East Greenland Current (Fig. 4). When the ice eventually melts, it will release its contaminants back to the
water column.
Numerous studies have investigated the concentrations of Pb, Hg, and Cd in arctic animals (cf. Dietz
et al., 1990; Muir et al., 1992 and references therein;
Wagemann et al., in press). These studies have been
conducted primarily to evaluate sources of the metals or
the dietary intake by natives rather than to determine
any effects on the animals themselves. The organs of
marine mammals from the Canadian Arctic often contain Hg concentrations that exceed human consumption
guidelines set for fish (0.5 ìg g "1; Health Canada) and
Cd concentrations that would result in exceeding consumption guidelines (450 ìg per week; Health Canada)
for a relatively modest diet (for examples of metal
content in marine animals see Muir et al., 1992; Kingsley
et al., 1993; Wagemann et al., 1995). Hansen et al. (1991)
similarly suggest that in sealing districts of Greenland,
maximum Cd content in food as recommended by the
World Health Organization (WHO) is being exceeded.
The detection of high concentrations of these two metals
in animals, however, does not necessarily imply contamination (see for example Hansen et al., 1991). It may
simply reflect a high natural background. Regional
differences in metal burdens in marine mammals
observed in the Canadian Arctic for Pb, Cd, and Hg
(Wagemann et al., 1994, 1995, in press) imply strongly
that metal concentrations in tissues depend to some
Contaminants in the arctic marine environment
degree on regional geology or biogeochemistry. For Cd
and Pb there is no compelling evidence of serious compromise of arctic marine biological resources by anthropogenic introductions. Despite this conclusion, it should
be noted that Cd is of concern because it is observed at
high concentrations in the organs of some animals (i.e. it
bioaccumulates; cf. Ray, 1984) and is easily mobilized
from marine sediments through small changes in oxygen
content (redox conditions) or pH (cf., Gobeil et al.,
1987). Our conclusion that anthropogenic loadings of
Cd are not presently of concern rests on relatively low
contaminant fluxes versus natural fluxes and the fact that
Cd does not tend to biomagnify (Guthrie et al., 1979).
However, not enough is known about the transfer of Cd
into the food chain to predict with confidence how the
system would respond to small incremental increases
resulting from anthropogenic activities. Hg is of concern
because its environmental behaviour parallels that of the
organochlorines in two ways – it is volatile, entering
easily into long-range atmospheric transport, and it bioaccumulates and biomagnifies in long-lived top predators. Methyl mercury, the more toxic form, is fat soluble,
has an affinity for proteins, and therefore biomagnifies
more than inorganic Hg (cf. Bloom, 1992). This behaviour, together with the fact that anthropogenic Hg fluxes
are comparable to natural fluxes, suggests increased risks
over the coming decades for populations heavily dependent on marine food sources. At the very least, Hg could
be considered a pollutant if the anthropogenic fluxes are
shown to lead to a greater proportion of marine foodstuffs exceeding consumption guidelines (cf. Wagemann
et al., in press).
Oil exploration, production, and transport
Three regions of the arctic shelves have experienced or
are likely to experience offshore oil production in the
foreseeable future: the American Beaufort shelf (cf.
Jaffe, 1991); the Canadian Beaufort shelf (cf. Dome et
al., 1982; Thomas et al., 1982); and the shelves of the
Barents and Kara Seas (cf. Doré, 1995; Klungsøyr et al.,
1995). Production has occurred already at Prudhoe Bay,
Alaska, but only in very shallow water. Exploration on
the Canadian Beaufort shelf has identified recoverable
quantities of oil in the offshore but it is not likely that
this oil will be produced before another decade or two. It
appears that the first substantive offshore development
will occur in the Russian sector where large gas and oil
reservoirs have been found (Doré, 1995; Klungsøyr
et al., 1995).
Extensive experience in the Canadian Beaufort Sea
during the exploration in the 1970s and 1980s has
developed techniques to produce oil in seasonally icecovered waters. Environmental impact assessments and
estimates of risk have also been conducted (cf. Dome et
al., 1982; Thomas et al., 1982). The major risks are, of
555
course, accidents, such as well blowouts, tanker collisions, and pipeline leakages. An unconstrained blow-out
in winter would be difficult, if not impossible, to remediate by drilling relief wells or mounting cleanup efforts
until the following spring. Oil would accumulate
beneath and within ice to be transported with it and later
released into melt ponds, leads, and polynyas. The main
arctic populations at risk include migratory birds and
animals associated with sea ice (seals and polar bears). It
should be emphasized that the marginal seas, which
contain the proven oil and gas reserves, are often highly
productive for fish, mammals, and birds (cf. Klungsøyr
et al., 1995).
If large-scale oil production proceeds in the Arctic,
both chronic and acute releases of oil will inevitably
occur. Spilled oil may travel long distances within the ice
cover with little dispersion or degradation; clearly, this is
of international concern. In the Beaufort Sea, the oil
might transit the Beaufort gyre several times before
eventually escaping into the Transpolar Drift (Fig. 4).
In the case of the Russian or Norwegian shelves, it is
more likely that spilled oil entrained in the ice will transit
the Arctic Ocean to arrive in the Barents Sea or in the
Fram Strait off the east coast of Greenland in a period of
one to two years (Fig. 4). Driftwood studies provide very
clear evidence that oil spilled on the Russian shelves will,
in part, end up on beaches in Svalbard (cf. Eggertsson,
1994).
Oil spilled into northern terrestrial drainage basins
also poses a risk to the Arctic Ocean as vividly illustrated by the recent Komi spill. Some, as yet unknown,
portion of the 1.2#105 m3 of oil from leaking pipelines
in Usinsk (Komi Republic) will reach the Barents Sea
via the Kolva, Usa, and Pechora rivers. As the oil travels
down these rivers, it will be subject to weathering with
the lighter fractions, usually associated with acute
toxicity, evaporating or dissolving during transit. Upon
reaching the estuary, some of the oil will become incorporated into ice forming during winter and some will
become attached to particles and sediment on the shelf
or disperse and further weather in the water column.
Subsequent weathering and microbial attack of the oil
will, however, proceed more slowly at the reduced
temperatures.
In temperate regions, the major sources of oil to the
marine environment are thought to be transportation
and chronic municipal and industrial releases that enter
the coastal ocean through runoff (National Academy of
Science, 1985). Two of the largest rivers, the Lena and
the Mackenzie, are apparently still pristine with respect
to anthropogenic inputs of petroleum (Peulvé et al.,
1995; Yunker and Macdonald, 1995). However, opening
the Arctic to oil or other forms of development will
result in greater traffic on the Northern Sea Route and
through the Canadian Archipelago. Chronic emissions
of oil in these transport lanes, parts of which pass
556
R. W. Macdonald and J. M. Bewers
through sensitive ecozones, will have the potential to
cause greater problems than in temperate regions where
degradation and dispersal of oil occurs more quickly.
The effectiveness of existing regulatory
arrangements
When evaluating the effectiveness of existing regulatory
arrangements, it is important that a distinction is made
between the provisions of controls and the extent of
compliance with them. In the following analysis,
attempts are made to draw such distinctions to define
whether existing threats or damage exist because of the
absence of regulatory controls or a lack of compliance.
Another factor that must be considered, when judging
whether new controls are warranted, is the amenability
of the source to human intervention. In cases where a
source is predominantly natural or its nature and distribution makes it intractable to human influence, there
is little point in considering the introduction of new
control measures.
Where effective regulatory arrangements do not exist
for contaminant releases that are amenable to control,
the nature of prudent action is relatively obvious –
formulate new control measures. However, in cases
where the fault lies in the failure to implement the
provisions of existing control measures, the solution,
while equally obvious, presents a range of administrative
difficulties, especially in an international context. It is
clearly desirable that all parties to the Arctic Environmental Protection Strategy implement existing control
measures in an equally committed manner. The authors
of this paper are not able, in all cases, to judge the nature
and extent of inequities in national compliance with
control measures. Nevertheless, in some cases, it is clear
that non-compliance is a significant contributing factor
to contemporary anthropogenic effects on the arctic
environment.
Persistent organic contaminants
The foregoing discussion has not identified incidents
where organic contaminants have produced toxic effects
in arctic organisms and, indeed, we would expect the
more highly contaminated areas in the northern temperate marine environment to show such effects first.
Nevertheless, the mechanism whereby some volatile,
toxic compounds are transported rapidly on global
scales, together with factors (i.e. long, fat-dominated
food chains with long-lived top predators) that lead to
bioaccumulation in prey and biomagnification in predators, is disturbing especially since the Arctic is so remote
from major production and use regions. Contamination
of country foods by these kinds of compounds has
already produced anxiety in native northerners about
the safety of their traditional foods and of breast feeding
(cf. Kinloch et al., 1992).
For regulation, it is important to note that the production and use of many of these contaminants has
already been banned or restricted in many developed
countries (cf. Pacyna, 1995) and this may be reflected in
a decline in burdens of PCBs and DDTs in seals (but see
Table 1). Much of the volatile organochlorine flux into
the Arctic comes from recycling of old residues between
terrestrial, atmospheric, and aquatic media. These recycled components of the contaminants are not
amenable to regulatory control. The fundamental
lesson, however, is that the release in large quantities of
toxic, volatile, long-lived, fat-soluble compounds can
leave a long-term legacy that is of global concern. There
are three critical steps that need to be taken: (1) the
continued use of compounds such as DDT in many
developing countries, should be restricted and/or substituted with the use of more environmentally benign
compounds; (2) a global reporting and auditing mechanism to account for the production and use of such
compounds should be devised and implemented by all
nations; and (3) new compounds proposed for use must,
as part of their screening process, be evaluated using
multi-media models for their potential to cause global
harm. The reduction of emissions from incinerators and
power plants in northern temperate countries, especially
in northern and eastern Europe, would also accomplish
reductions in the delivery of chlorinated compounds and
PAHs to the Arctic Ocean. The above steps could be
achieved through international arrangements at regional
and global levels, such as those currently being pursued
under the auspices of the Convention on Long-Range
Transboundary Air Pollution.
Radionuclides
The foregoing discussion of sources of radionuclides in
the Arctic and the associated doses and risks, albeit
presented on exemplar grounds, provides little justification for advocating the introduction of new regulatory
mechanisms for practices involving the production use,
and disposal of radionuclides. Reprocessing operations
in Europe, and their associated radionuclide releases, are
authorized on the basis of international procedures and
standards, largely those established by the International
Atomic Energy Agency (IAEA). Such authorization
takes account of risks posed to public health and deems
them acceptable. Thus, contamination of the Arctic by
radionuclides from these sources can only constitute
contamination, not pollution, unless some fundamental
fault is found within the system of radiological protection on which such authorizations are based or its
application to specific practices. The deliberate disposal
of radionuclides into the environment through dumping
or land discharge, either within the Arctic or elsewhere,
Contaminants in the arctic marine environment
is clearly amenable to control. The basis of control is
adequately provided by the IAEA Basic Safety Standards (IAEA, 1995) and the system of radiological
protection (ICRP, 1991). However, these regulatory
processes must be implemented with uniform rigour by
all countries, especially those that are signatories to the
Arctic Environmental Protection Strategy. It should be
noted that sea dumping of radioactive waste, under the
terms of the London Dumping Convention (now the
‘‘London Convention 1972’’) prior to the ban concluded
in 1993, was restricted to deep waters greater than
4000 m between latitudes 50)N and 50)S (IAEA, 1986).
Clearly, these restrictions were not observed by the
former Soviet Union.
Some steps have already been taken to consider the
probability and the consequences of accidents at northern power stations (e.g. Rantalainen, 1995; Amosov
et al., 1995; Dubkov et al., 1995) and in the servicing and
decommissioning of military and civil marine reactors in
Russia. It is impossible to guarantee the complete prevention of accidents. Accordingly, regulatory actions
must continue to focus on reducing their probability and
consequences through the analysis of design, and
beyond-design, accident scenarios. These matters are
well subsumed within the system of radiological protection; it remains, however, to ensure that all nations
apply this system in a consistent manner.
The salient observation is that mechanisms and
approaches are already in place to provide a sound basis
for the protection of human health and the environment
on local, regional, and global scales, including the
Arctic. What is needed, therefore, is to have these
mechanisms implemented ubiquitously with the same
zeal as is presently the case in most countries exploiting
nuclear technology. This is not to say that some previous
and existing practices are not in need of enhanced
re-assessment. Specifically, decommissioning and scrapping of Russian nuclear submarines and safe disposal of
spent fuel are clearly matters of justifiable contemporary
and future concern.
In summary, for radionuclides, no new regulatory
controls are needed – existing international control
mechanisms need to be rigorously observed and applied
in all Arctic states and elsewhere.
Metals
Metal levels in the environment may be enhanced locally
by releases from mining (Greenland, Canadian
Archipelago) or nearby industrial activity (the inner
Kara and Laptev Sea shelves). Mining regulations and
techniques to reduce metal mobilization from mining are
already available; therefore, what is needed is compliance with best practice. Emission controls for
industry would reduce the fluxes of metals reaching the
Arctic Ocean from eastern Europe and Russia. Recent
557
high-quality data suggest that the largest Russian rivers
are not major sources of contaminant heavy metals
(Coquery et al., 1995; Dai and Martin, 1995; Martin
et al., 1993). It remains to be seen whether there are
important contaminant fluxes from smaller rivers resulting from localized anthropogenic activity. A comprehensive survey of contaminants in river transport would
provide confidence in the estimates of metal fluxes from
rivers to the Arctic Ocean. Should it be determined that
smaller rivers represent a significant point of entry of
heavy metals (and other contaminants) to the Arctic
Ocean, the existing UN ECE Convention on Protection
and Use of Transboundary Watercourses and International Lakes (1992) may be considered as an example of
the type of instrument with which to address associated
concerns.
Although metals are clearly reaching the circumpolar
Arctic through atmospheric transport, there is no
present evidence that anthropogenically mobilized Cd
and Pb are posing a wide-scale biological threat. Hg,
however, is probably of greater concern for two reasons.
First, there is evidence of significant anthropogenic
mobilization of this metal during the past century
(Nriagu and Pacyna, 1988; Pacyna and Keeler, 1995).
Second, due to bioaccumulation and biomagnification,
long-lived marine species in the Arctic have naturally
high concentrations of Hg. Therefore, these species, and
their consumers, may be particularly vulnerable to
added burdens. This problem is not confined to the
Arctic; it is global (cf. Nriagu, 1988) and will require
global approaches to the reduction of emissions. It is
noted that countries of the UN ECE region are considering the possibility of a protocol to control the emissions of certain heavy metals (including Hg, Cd, and Pb)
under the Convention on Long Range Transboundary
Air Pollution. Further work is urgently needed to verify
Hg trends globally and in the Arctic (e.g. measurements
in the atmosphere, in sediment cores, and in biota). This
would provide a basis to determine the amenability and
the nature of any necessary additional control measures.
Oil
Regulation and design of safe practices in the oil industry in the context of arctic exploration and development
has a long history in the North American sector of the
Arctic (Prudhoe Bay, Canadian Beaufort Sea). A decision never to produce oil from the Arctic offshore would
seem difficult to achieve and unlikely to be taken. The
most stringent practices can be expected only to reduce
the probability of accident, not eliminate it, Environmental Impact Assessments (EIAs), the assessment of
risks, and operating practices clearly need to meet
international standards and to be tailored specifically to
the Arctic. A careful evaluation of the probable effects
of increased ship traffic in the Northern Sea Route must
558
R. W. Macdonald and J. M. Bewers
also be included. Emphasis needs to be placed on
realistic assessments of risk, compliance with regulation,
conformity with acceptable design, the development
of proven emergency response and oilspill cleanup
technology, and specifications of liability in respect to
redressing damage to injured parties. The contemporary
example of pipelines on the Komi Republic that have
exceeded their design life is instructive. Regulation of the
production and transport of oil in northern land areas
must meet similarly stringent criteria as for offshore oil
production and transport.
Other disturbances
Of primary importance among other disturbances to the
marine environment are likely to be habitat destruction
in the coastal zone and alteration of hydrology in
estuaries resulting from the diversion of water for power
generation (cf. Cattle, 1985; Prinsenberg, 1991; Bodaly
and Johnston, 1992; Klungsøyr et al., 1995). Existing
damage to coastal areas through physical destruction of
habitats is likely to be significant on local scales in a
number of countries, especially from the perspective of
indigenous peoples. Furthermore, without the imposition of additional controls to minimize the destruction
of habitat and associated biological diversity, continued
development of arctic marginal seas is likely to result in
further damage. Greater attention should therefore be
given by the States that are contracting Parties to the
Arctic Environmental Protection Strategy, with respect
to small-scale physical impacts on the arctic coastal
environment, in order to determine what steps should be
taken to protect resources and amenities.
give rise to contamination but not pollution (e.g. lead,
cadmium, and radionuclides); existing control measures
are adequate (e.g. radionuclides) although, in some
cases, there has been lack of compliance with them (e.g.
disposal of radioactive wastes); or their sources are not
directly amenable to control (e.g. PAHs from various
combustion sources and organochlorines from secondary environmental reservoirs). The sole exception to the
above conclusion among the priority contaminants discussed here may be mercury which is naturally present at
high concentrations in long-lived arctic species and for
which anthropogenic mobilization rivals the natural
flux. Indigenous peoples, especially in Greenland and
Canada, depend heavily on marine food from the highest trophic levels, putting them particularly at risk from
additional Hg burdens in these species. The properties of
mercury in many ways parallel the organochlorines.
Therefore, the assessment of temporal trends of Hg in
the arctic marine ecosystem should be a priority. Greater
attention should also be paid to physical impacts on the
arctic coastal environment (see Klungsøyr et al., 1995).
We note that there exists great concern about the current
state of contaminants on the Russian shelves (e.g.
Melnikov, 1991; Green, 1993; Pfirman et al., 1993; Luzin
et al., 1994) and whether hazardous amounts of these
contaminants might be broadcast to other arctic regions
by ice (Pfirman et al., 1995a,b). Presently, the database
for the Russian shelves is simply not adequate for the
task of assessing what action is needed; it should be a
priority to develop such a database. Finally, we note
that pipelines being used beyond their designed lifetimes
clearly pose a risk of future oil spills into northern
Russian river drainage basins. In this respect, assistance
to prevent future spills by inspecting and upgrading
pipelines is required more urgently than new regulations.
Conclusion
This document analyzes the contemporary situation
concerning contaminants that have been identified as
priorities within the Arctic. The development of new
control measures to protect the arctic marine environment has been advocated only for a single class of
contaminants, the organochlorines. This is because their
properties (toxicity, longevity, susceptibility to longrange transport, bioaccumulation and biomagnification)
lead to global effects, whereas their production and use
in many countries continues with inconsistent controls
and with little or poor documentation. We strongly
advocate the consistent reporting by all countries of
production and use statistics for such compounds and
that, before new compounds are brought into production, they be evaluated for their potential to cause global
harm using multi-media models like that developed by
Wania and McKay (1995). In the cases of the other
priority contaminants, either: their effects are not of
great significance in the arctic marine environment; they
Acknowledgements
The perspective of this document owes much to David
Stone who originally requested that we prepare it and
who encouraged us during its preparation. We thank
H. E. Welch for providing useful literature and for
insightful discussions and Lyle Lockhart for providing a
helpful summary on the status of metals in the Canadian
Arctic. Finally, we thank A. Aarkrog, R. F. Addison,
E. Carmack, C., Gobeil, S. Pfirman, A. Salo, W. L.
Templeton, G. Topping, A. Walton, and H. L. Windom
for providing helpful comments on the draft document.
All figures were prepared by P. Kimber.
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